United States Department of Agriculture / Forest Service

Rocky Mountain Research Station

Proceedings RMRS-P-52

September 2008

Proceedings— 

Shrublands Under Fire: 

Disturbance and Recovery in a Changing World; 

2006 June 6–8; Cedar City, UT
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Abstract

The 27 papers in these proceedings are divided into five sections. The first includes 

an introduction to the symposium theme of Shrublands under Fire, along with a 

keynote address comparing ecosystem degradation in Iran with current problems 

in the semi-arid to arid Western United States. The next three sections cluster papers on invasive species, community dynamics and restoration, and fire recovery 

of shrubland ecosystems. The final section contains descriptions of two field trips; 

one focusing on the transitional zone between cold and warm desert systems, and 

the other on Great Basin shrublands and the Desert Experimental Range. 

Keywords: wildland shrubs, disturbance, recovery, fire, invasive plants, restoration, 

ecology, microorganisms

Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; Vernon, Jason, 

comps. 2008. Proceedings—Shrublands under fire: disturbance and 

recovery in a changing world; 2006 June 6–8; Cedar City, UT. Proc. RMRS-P-52. 

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

Mountain Research Station. 190 p.
Proceedings— 

Shrublands Under Fire: 

Disturbance and Recovery in a Changing World; 

2006 June 6–8; Cedar City, UT

Compilers:

Stanley G. Kitchen

Rosemary L. Pendleton

Thomas A. Monaco

Jason Vernon

i
Availability of Previous Wildland Shrub Symposia Proceedings

First: Tiedemann, A. R.; Johnson, K. L., compilers. 1983. Proceedings—research and management of bitterbrush and cliffrose 

in western North America; 1982 April 13-15; Salt Lake City, UT. General Technical Report INT-152. Ogden, UT: U.S. Department 

of Agriculture, Forest Servcie, Intermountain Forest and Range Experiment Station, 279 p. Out of print—available from National 

Technical Information Service as document PB83261537. Contact NTIS at: (800) 553-6847, or order online at: www.ntis.gov

Second: Tiedemann, A. R.; McArthur, E. D.; Stutz, H. C.; Stevens, R.; Johnson, K. L., compilers. 1984. Proceedings—symposium 

on the biology of Atriplex and related chenopods; 1983 May 2-6; Provo, UT. General Technical Report INT-172. Ogden, UT: U.S. 

Department of Agriculture, Forest Service, Intermountain Forest and Range Experiment Station. 309 p. Out of print—available 

from National Technical Information Servcie as document PB85116358. Contact NTIS at: (800) 553-6847, or order online at: www.

ntis.gov

Third: McArthur, E. D.; Welch, B. L., compilers. 1986. Proceedings—symposium on the biology and management of Artemisia and 

Chrysothamnus; 1984 July 9-13; Provo, UT. General Technical Report INT-200. Ogden, UT: U.S. Department of Agriculture, Forest 

Service, Intermountain Research Station. 398 p. Out of print—available from National Technical Information Service as document 

PB86182318. Contact NTIS at: (800) 553-6847, or order online at: www.ntis.gov

Fourth: Provenza, F. D.; Flinders, J. T.; McArthur, E. D., compilers. 1987. Proceedings—symposium on plant herbivore interactions; 

1985 August 7-9; Snowbird, UT. General Technical Report INT-222. Ogden, UT: U.S. Department of Agriculture, Forest Service, 

Intermountain Research Station. 179 p. A few copies are available from the Rocky Mountain Research Station; otherwise available from National Technical Information Service as document PB90228578. Contact NTIS at: (800) 553-6847, or order online at: 

www.ntis.gov

Fifth: Wallace, A.; McArthur, E. D.; Haferkamp, M. R., compilers. 1989. Proceedings—symposium on shrub ecophysiology and 

biotechnology; 1987 June 30-July 2; Logan, UT. General Technical Report INT-256. Ogden, UT: U.S. Department of Agriculture, 

Forest Service, Intermountain Research Station. 183 p. Available from Rocky Mountain Research Station: http://www.fs.fed.us/

rm/publications/

Sixth: McArthur, E. D.; Romney, E. M.; Smith, S. D.; Tueller, P. T., compilers. 1990. Proceedings—symposium on cheatgrass 

invasion, shrub die-off, and other aspects of shrub biology and management; 1989 April 5-7; Las Vegas, NV. General Technical 

Report INT-276. Ogden, UT: U.S. Department of Agriculture, Forest Service, Intermountain Research Station. 351 p. Out of print—

available from National Technical Information Service as document PB91117275. Contact NTIS at: (800) 553-6847, or order online 

at: www.ntis.gov

Seventh: Clary, W. P.; McArthur, E. D.; Bedunah, D.; Wambolt, C. L., compilers. 1992. Proceedings—symposium on ecology and 

management of riparian shrub communities; 1991 May 29-31; Sun Valley, ID. General Technical Report INT-289. Ogden, UT: U.S. 

Department of Agriculture, Forest Service, Intermountain Research Station. 232 p. Out of print—available from National Technical 

Information Service as document PB92227784. Contact NTIS at: (800) 553-6847, or order online at: www.ntis.gov

Eighth: Roundy, B. A.; McArthur, E. D.; Haley, J. S.; Mann, D. K., compilers. 1995. Proceedings: wildland shrub and arid land restoration symposium; 1993 October 19-21; Las Vegas, NV. General Technical Report INT-GTR-315. Ogden, UT: U.S. Department 

of Agriculture, Forest Service, Intermountain Research Station. 384 p. Available from the Rocky Mountain Research Station: http://

www.fs.fed.us/rm/publications/ or you can download the publication at: http://www.fs.fed.us/rm/pubs_int/int_gtr315.html

Ninth: Barrow, J. R.; McArthur, E. D.; Sosebee, R. E.; Tausch, R. J., compilers. 1996. Proceedings: shrubland ecosystem dynamics 

in a changing environment; 1995 May 23-25; Las Cruces, NM. General Technical Report INT-GTR-338. Ogden, UT: U.S. Department of Agriculture, Forest Service, Intermountain Research Station. 275 p. Available from the Rocky Mountain Research Station: 

http://www.fs.fed.us/rm/publications/

Tenth: McArthur, E. D.; Ostler, W. K.; Wambolt, C. L., compilers. 1999. Proceedings: shrubland ecosystem ecotones; 1998 August 

12-14; Ephraim, UT. Proceedings RMRS-P-11. Ogden, UT: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 299 p. Available from the Rocky Mountain Research Station: http://www.fs.fed.us/rm/publications/

Eleventh: McArthur, E. D.; Fairbanks, D. J, compilers. 2001. Shrubland ecosystem genetics and biodiversity: proceedings; 2000 

June 13-15; Provo, UT. Proceedings RMRS-P-21. Ogden, UT: U.S. Department of Agriculture, Forest Service, Rocky Mountain 

Research Station. 365 p. Available from the Rocky Mountain Research Station: http://www.fs.fed.us/rm/publications/ or you can 

download the publication at: http://www.fs.fed.us/rm/pubs/rmrs_p021.html

Twelfth: Hild, A. L.; Shaw, N. L.; Meyer, S. E.; Booth, D. T.; McArthur, E. D., compilers. 2004. Seed and soil dynamics in shrubland 

ecosystems: proceedings; 2002 August 12-16; Laramie, WY. Proceedings RMRS-P-31. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 216 p. Available from the Rocky Mountain Research Station: http://

www.fs.fed.us/rm/publications/ or you can download the publication at: http://www.fs.fed.us/rm/pubs/rmrs_p031.html

Thirteenth: Sosebee, R. E.; Wester, D. B.; Britton, C. M.; McArthur, E. D.; Kitchen, S.G., comps. 2007. Proceedings:shrubland 

dynamics--fire and water; 2004 August 10-12; Lubbock, TX. Proceedings RMRW-P-47. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 173 p. Available from the Rocky Mountain Research Station: http://

www.fs.fed.us/rm/pubs/rmrs_p047.html

ii
Shrub Research Consortium

USDA Forest Service, Rocky Mountain Research Station, Shrub Sciences Laboratory*, Provo, Utah,

E. Durant McArthur (chair)

Brigham Young University*, Provo, Utah, Daniel J. Fairbanks

USDA Agricultural Research Service, Renewable Resource Center*, Reno, Nevada, James A. Young

Utah State University*, Logan, Eugene W. Schupp

State of Utah, Department of Natural Resources, Division of Wildlife Resources*, Ephraim, Jason L. Vernon

University of California, Los Angeles, Philip W. Rundel

Colorado State University, Fort Collins, William K. Lauenroth

University of Idaho, Moscow, Steven J. Brunsfeld

University of Montana, Missoula, Donald J. Bedunah

Montana State University, Bozeman, Carl L. Wambolt

University of Nevada, Reno, Barry L. Perryman

University of Nevada, Las Vegas, Stanley D. Smith

Oregon State University, Corvallis, Lee E. Eddleman

New Mexico State University, Las Cruces, Kelly W. Allred

Texas A & M System, Texas Agricultural Experiment Station, Vernon, R. J. (Jim) Ansley

Texas Tech University, Lubbock, David B. Wester

USDA Agricultural Research Service, High Plains Grassland Research Station, Cheyenne, Wyoming,

D. Terrance Booth

USDA Agricultural Research Service, Jornada Experimental Range, Las Cruces, New Mexico, Mary E. Lucero

USDA Agricultural Research Service, Forage and Range Laboratory, Logan, Utah, Thomas A. Monaco

USDA Forest Service, Rocky Mountain Research Station, Forestry Sciences Laboratory, Albuquerque,

New Mexico, Rosemary L. Pendleton

USDA Forest Service, Rocky Mountain Research Station, Great Basin Ecology Laboratory, Reno, Nevada,

Robin J. Tausch

University of Utah, Salt Lake City, James R. Ehleringer

Weber State University, Ogden, Utah, Barbara A. Wachocki

Battelle Pacific Northwest Laboratories, Richland, Washington, Janelle L. Downs

Bechtel Nevada, Las Vegas, Nevada, W. Kent Ostler

University of Wyoming, Laramie, Ann L. Hild

*Charter members

The Shrub Research Consortium (SRC) was formed in 1983 with five charter members (see list). Over time, SRC 

has grown to its present size of 25 institutional members. The SRC charter had three principal objectives: (1) developing plant materials for shrubland rehabilitation; (2) developing methods of establishing, renewing, and managing shrublands in natural settings; and (3) assisting with publication and dissemination of research results. These 

objectives have been met by a series of symposia sponsored by the Consortium and partners. This publication 

is the 14th in the series. The 13 previous symposia proceedings are listed on the previous page. Proceedings of 

all publications to date have been published by the U.S. Department of Agriculture, Forest Service, Intermountain 

Research Station and Rocky Mountain Research Station. Each symposium has had a theme, but the executive 

committee has encouraged attendance and participation by shrubland ecosystem biologists and managers with 

wider interests than any particular symposium theme. The heart of the Consortium’s programs are wildland shrub 

ecosystem biology, research, and management.

iii

Contents

Introduction and Overview ..................................................................................... 1

Shrublands Under Fire: Disturbance and Recovery in a Changing World.................................. 3

Stanley G. Kitchen

Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure ...................... 5

David A. Charlet

Invasive Species and Shrublands........................................................................ 25

Cheatgrass and Red Brome; the History and Biology of Two Invaders.................................... 27

Chad R. Reid, Sherel Goodrich, and James E. Bowns

Invasion of the Exotic Grasses: Mapping their Progression via Satellite.................................. 33

Eric B. Peterson

The Indirect Effects of Cheatgrass Invasion: Grasshopper Herbivory on Native Grasses 

Determined by Neighboring Cheatgrass Abundance................................................................41

Julie Beckstead, Susan E. Meyer, and Carol K. Augsperger

Effects of Herbaceous Species Removal, Fire and Cheatgrass (Bromus tectorum) 

on Soil Water Availability in Sagebrush Steppe........................................................................49

Alison Whittaker, Bruce Roundy, Jeanne Chambers, Susan Meyer, Robert Blank, 

Stanley Kitchen, and John Korfmacher

Controlling Cheatgrass in Winter Range to Restore Habitat and Endemic Fire ....................... 57

Jennifer L. Vollmer and Joseph G. Vollmer 

Cheatgrass (Bromus tectorum) Biocontrol Using Indigenous Fungal Pathogens..................... 61

Susan E. Meyer, David L. Nelson, Suzette Clement, and Julie Beckstead

Spatial and Temporal Seed Dispersal of Squarrose Knapweed (Centaurea

virgata Lam. ssp. squarrosa (Willd.) Gugler) in West Central Utah, a Case Study .................. 69

Scott L. Jensen, Stephen B. Monsen, and Pat Fosse

Medusahead: Available Soil N and Microbial Communities in Native and Invasive Soils......... 73

Robert R. Blank, René Sforza, and Tye Morgan

Effects of Fire and Restoration Seeding on Establishment of Squarrose Knapweed 

(Centaurea virgata var. squarrosa)...........................................................................................77

Alison Whittaker and Scott L. Jensen

Shrubland Community Dynamics and Restoration............................................ 81

Symbiotic Fungi that Influence Vigor, Biomass and Reproductive Potential of Native 

Bunch Grasses for Remediation of Degraded Semiarid Rangelands....................................... 83

Jerry R. Barrow, Mary E. Lucero, and Isaac Reyes-Vera

Genetic Characterization of Uncultured Fungal Endophytes from Bouteloua eriopoda 

and Atriplex canescens.............................................................................................................87

Mary E. Lucero, Jerry R. Barrow, Ruth Sedillo, Pedro Osuna-Avila, Isaac Reyes-Vera

Wildfire Rehabilitation Success with and without Chaining on the Henry Mountains, Utah...... 91

Cristina Juran, Bruce A. Roundy, and James N. Davis

v
Vole-Driven Restoration of a Parariparian Meadow Complex on the Colorado Plateau 

(South-Central Utah)...............................................................................................................107

Dennis M. Bramble and Jean C. Bramble

Browse Evaluation of Tall Shrubs Based on the Direct Measurement of a 

Management Objective........................................................................................................... 115

Richard B. Keigley and Michael R. Frisina 

Mule Deer and Elk Winter Diet as an Indicator of Habitat Competition .................................. 123

Michael R. Frisina, Carl L. Wambolt, W. Wyatt Fraas, and Glen Guenther

Radial Symmetry in the White Bursage (Ambrosia dumosa) in the Mojave Desert................ 127

Steven A. Lei and Simon A. Lei

Reproductive Biology of Larrea tridentata: A Comparison Between Core Shrubland and 

Isolated Grassland Plants at the Sevilleta National Wildlife Refuge, New Mexico ................. 131

Rosemary L. Pendleton, Burton K. Pendleton, Karen R. Wetherill, and Terry Griswold

Fire Recovery in Shrublands ............................................................................. 137

A Survey of Vegetation and Wildland Fire Hazards on the Nevada Test Site......................... 139

Dennis J. Hansen and W. Kent Ostler

Trout Creek 1999 Burn ...........................................................................................................147

Sherel Goodrich

Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites, 

Uinta Mountains and West Tavaputs Plateau, Utah................................................................151

Sherel Goodrich, Allen Huber, and Brian Monroe

Very Large Scale Aerial (VLSA) Imagery for Assessing Postfire Bitterbrush Recovery.......... 161

Corey A. Moffet, J. Bret Taylor, and D. Terrance Booth

20 Years of Natural Recovery after Wildfire on Northern Nevada Rangelands ...................... 169

Ann P. Bollinger and Barry L. Perryman

Microsite and Time Since Prescribed Fire’s Influence on Soil Microbiology in a 

Pinyon Woodland....................................................................................................................175

Benjamin M. Rau, Robert R. Blank, and Tye Morgan

Field Trips............................................................................................................. 179

Great Basin Cold Desert Shrublands and the Desert Experimental Range ........................... 181

Stanley G. Kitchen and Stephanie L. Carlson

Disturbance and Rehabilitation of Cold to Warm Desert Transitional Shrublands in 

Southwestern Utah .................................................................................................................187

Chad Reid and James Bowns

vi
Introduction and Overview

Invasive Species and Shrublands

Fire Recovery in Shrublands

Field Trips

Shrubland Community 

Dynamics and Restoration

USDA Forest Service Proceedings RMRS-P-52. 2008 1

Introduction and Overview

USDA Forest Service Proceedings RMRS-P-52. 2008 3

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas 

A.; Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Stanley G. Kitchen is a Research Botanist, U.S. Forest Service, 

Rocky Mountain Research Station, Shrub Sciences Laboratory, 

Provo, UT.

Shrublands Under Fire: Disturbance and 

Recovery in a Changing World

Stanley G. Kitchen

These proceedings are part of a series corresponding 

to fourteen symposia sponsored by the Shrub Research 

Consortium and published by the U.S. Forest Service, 

Intermountain Research Station (1983–1996) and Rocky 

Mountain Research Station (1999 to present). Themes for 

each symposium were selected to focus on various aspects 

of the biology, ecology, and management of wildland shrubs 

and shrubland ecosystems. The Fourteenth Wildland Shrub 

Symposium was held at Southern Utah University in Cedar 

City, Utah, on June 6–8, 2006.

Patterns of disturbance and recovery function to shape 

the composition, structure, and integrity of all ecosystems, 

including shrublands, and are strongly influenced by both 

natural (for example, climate) and anthropogenic (for 

example, livestock husbandry) controls. Although these 

processes vary naturally, rapid or persistent change in 

either disturbance regimes or successional pathways may 

result in a loss of ecosystem functionality or resiliency and 

subsequently result in ecosystem degradation.

Humans affect shrublands by altering natural disturbance regimes, creating new disturbances processes, and by 

altering succession. Anthropogenic activities produce both 

intended and unintended consequences. For example, fire 

might be used as a tool to favorably modify heavily stocked 

forest environments for hunting. Unintended consequences 

of widespread burning would include altered habitat for 

numerous wildlife species and decreased availability of fuel 

wood. Over time, the nature and magnitude of human-caused 

impacts on shrublands reflect changes in the values associated with shrub-dominated landscapes and the capacity 

of evolving cultural systems and associated technologies 

to enact change. Impacts have been significant and often 

severe in regions of the world where human activities have 

been both intense and persistent over long time periods. In a 

plenary address, David Charlet (this proceedings) describes 

the results of long-term degradation of arid and semiarid 

shrublands, woodlands, and forests in Iran, and contrasts 

that to the recent and relatively low levels of degradation 

in the arid and semiarid ecosystems of western North 

America’s Great Basin. He warned, however, of the potential 

for severe degradation not unlike that found in Iran and in 

similar old-world environments if management and use of 

this region do not reflect a stronger conservation ethic than 

that observed from extraction-based economies introduced 

by Euro-American settlers in the 19th century.

Worldwide, shrub-dominated ecosystems are increasingly 

at risk of degradation and loss as human-related impacts 

increase and natural patterns of disturbance and recovery 

are truncated due to the cumulative and interacting impacts 

of changing climate, CO2

 fertilization, altered fire regimes, 

invasive species introductions, fragmentation and conversion. Indeed, it is not overstated to say that shrublands are 

under fire. The shrublands of western North America are no 

exception. Proactive measures to curtail losses of functional 

shrubland ecosystems must include the development of scientifically sound strategies and practices for restoration and 

rehabilitation. Of necessity, the development and application 

of these strategies and practices will require adaptive approaches as paradigms shift in an ever changing world.

Sessions of the symposium were organized based upon 

three sub-themes: (1) Invasive Species and Shrublands, 

(2) Shrubland Community Dynamics and Restoration, and 

(3) Fire Recovery in Shrublands. Additional papers addressing these themes and related topics were presented 

in a poster session. Twenty-four papers were prepared for 

publication in these proceedings.

Two mid-symposia tours were offered to participants. A 

Great Basin cold desert tour included stops to observe a pigmy 

rabbit colony in basin big sagebrush, post-burn recovery in 

a mountain big sagebrush community, 20th century juniper 

expansion in mixed desert shrub-grassland, various aspects 

of the ecology and management of salt-desert shrublands, 

and the effectiveness of mechanical thinning (bull hog) and 

reseeding treatments (native and introduced species mixes) 

in restoring Wyoming big sagebrush communities. The tour 

included stops at the historic U.S. Forest Service, Desert 

Experimental Range headquarters and the abandoned mining community of Frisco, Utah. A second tour featured the 

shrublands that occupy the transitional zone between cold 

and warm desert ecosystems with particular emphasis on 

the impacts of fire and introduced annual brome grasses 

on blackbrush communities. Stops were made to examine 

short and long-term results of seeding disturbed sites in 

this transitional zone to introduced perennial grasses and 
4 USDA Forest Service Proceedings RMRS-P-52. 2008

Kitchen Shrublands Under Fire: Disturbance and Recovery in a Changing World: Introduction

to the Eurasian sub-shrub, forage kochia. Additional stops 

were made to examine the effects of drought and a severe 

wind event on erosion in a Wyoming sagebrush community 

and adjacent subdivision, the ecological transformation 

(degradation) that has occurred to the wet meadows at 

the Mountain Meadows historic site, and the fire-adapted, 

interior chaparral communities on the east flank of the Pine 

Valley Mountains.

The fifteenth Wildland Shrub Symposium will be held 

June 17–19, 2008, in Bozeman, Montana. The theme will 

be, “Shrublands: Wildlands and Wildlife Habitat.”

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein.
USDA Forest Service Proceedings RMRS-P-52. 2008 5

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas 

A.; Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

David A. Charlet is Professor, Department of Biology, Community College of Southern Nevada, Henderson, NV 89002, U.S.A.; 

email: david.charlet@csn.edu. 

Abstract—Semi-arid shrublands and woodlands of Nevada are 

changing in the face of many daunting challenges. I compare Nevada 

to Iran to understand these challenges better, because Iran and 

Nevada have similar climate, geology, physiography, and latitude. 

Floristically, Iran and Nevada share many dominant genera, and 

many of Nevada’s troubling invasive species are native to Iran. Yet, 

Iran is different in its long history of civilization and concomitant 

human exploitation of the landscape. Thus, we can look to Iran to 

gain insight into possible outcomes of our management actions in 

our remaining wild shrublands. The structure of Iranian vegetation 

is simple compared to that of Nevada. It usually possesses only a 

single canopy layer and has low shrub species diversity, producing 

one-dimensional vegetation of low value for wildlife. The Iranian flora 

shows the mark of long-term grazing and fire disturbance. In spite of 

enormous mountains, woodlands and montane forests are virtually 

non-existent. Nevada’s vegetation, in spite of the challenges it faces, 

is in far better condition than the vegetation of Iran. However, if 

we allow the processes that simplify vegetation to gain momentum, 

then we can look to Iran to see our monotonic future.

Threats to Our Nation, 

Threats to Our Shrublands __________

Today in the news we hear about many political challenges 

to our nation. At this time, three nations appear boldly in 

our newspapers. China’s rapid economic development poses a 

great challenge to our economy. Conversely, a lack of economic 

opportunity in Mexico and other Latin American countries 

has led to a steady and substantial inflow of undocumented 

immigrants across our southern border. Prominent in the 

news is Iran. We read that our government is suspicious of 

Iran’s nuclear ambitions, despite the stated intention of its 

leaders to harness the atom strictly for the peaceful production of energy. Moreover, our government has accused Iran’s 

government of aiding the insurgency in Iraq and Hezbollah 

in Lebanon. Coincidentally, each of these nations poses 

not only political challenges to the United States but also 

introduced elements of their floras threaten the ecosystems 

of our semi-arid West.

Shah-Kan-Daw: Anthropogenic Simplification 

of Semi-arid Vegetation Structure 

David A. Charlet

Our arid and semi-arid shrublands are also in the news. 

Alien species have invaded our shrublands and woodlands. 

These species are affecting fundamental changes in basic 

ecosystem processes. Wildfires burn everywhere with 

dramatically increasing frequency (Whisenant 1990). The 

changes in the vegetation are reflected in the status of formerly abundant vertebrates, and so the public learns from 

the news the threats to sage grouse in Artemisia shrublands 

and the desert tortoise in Larrea shrublands. Misinformation 

abounds; we read in scientific literature that native juniper 

trees are invasive (Ansley and Rasmussen 2006), pinyons 

are often regarded as such, and both are treated as if they 

were undocumented immigrants.

A bright young Ph.D. student at Beijing Forestry University 

challenged me after my talk on semi-arid forestry conservation when he asked: “Why do you want to save the deserts?” 

He clarified himself, “We want to improve the deserts.” I was 

not yet fully aware of what their concerns were, but I quickly 

learned. His concern was desertification in China. He was 

interested in shrubs and trees from Nevada that could hold 

back expanding sand dunes and so conserve the soil and the 

productivity of the land. I suggested he try species native to 

China, such as Tamarix.

Why Conserve Deserts?

My idea of deserts was different from that fine Chinese 

student’s idea of deserts. I have spent all my life in the 

American West, most of it in Nevada. I am used to shrublands 

in my deserts and woodlands and forests in my mountains, 

with sand dunes and spring-fed wetlands as occasional fascinating features of a landscape already rich in species. It is no 

wonder that we had difficulty understanding one another, as 

definitions of deserts vary greatly. For instance, McGinnies 

and others (1968) report that definitions of deserts based 

on precipitation alone vary from a maximum mean annual 

precipitation of 5 cm to as much as 38 cm! Whatever we call 

them, in much of China—and as I was soon to find out in 

Iran as well—basic ecosystem processes in semi-arid regions 

are failing. By comparison, in Nevada, these processes are 

merely threatened. But if human activities further destabilize Nevada’s ecosystems, the ecosystems may collapse. 

Such a collapse inevitably leaves a burdensome economic 

wake, and potentially leads to tragic human consequences 

(Diamond 2005). Humans deliver a one-two punch to wild 

ecosystems: they disturb the ground and bring seeds that 

flourish in the disturbance.

Human activities in Nevada during the last 150 years have 

caused countless disturbances. They have accelerated the 

spread of alien weeds, such as cheatgrass (Bromus tectorum), 
6 USDA Forest Service Proceedings RMRS-P-52. 2008

Charlet Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure

that may lead to their dominance throughout much of the 

region (Billings 1990; Bradley and Mustard 2006; Knapp 

1996). Groundwater pumping constitutes a major threat to 

valley bottom phreatophytic vegetation and unique spring 

ecosystems (Charlet 2006). Utility corridors cut large swaths 

throughout the state, fragmenting formerly continuous 

ecosystems, creating convenient migration corridors, and 

providing suitable ground for the establishment of these 

invasive species (Bradley and Mustard 2006; Lathrop and 

Archbold 1980). In Nevada, alien species are notoriously 

flammable and invasive. Cheatgrass provides unbroken 

patches between shrubs, dries early in the growing season 

of most native plants, and so provides fine fuels early in the 

season that carry fires throughout large areas (Young and 

Evans 1978). Once burned, the likelihood of another catastrophic fire returning is high, as fire return intervals are 

now shorter by as much as an order of magnitude from what 

they were before settlement (Whisenant 1990). Life history 

characteristics conspire to accelerate the ascendancy of these 

species, which can prevent the establishment of shrubs even 

in the absence of returning fires (Billings 1990). The prospect of increasing atmospheric CO2 levels almost ensures 

that this and other opportunistic aliens such as Bromus 

madritensis (Smith and others 2000) will come to dominate 

the region. The net effect of alien brome grasses in Nevada 

is to impoverish biodiversity (Billings 1990) by converting 

shrublands and semi-arid woodlands to annual grasslands 

(Bradley and others 2006). Moreover, the consequences of 

grassland replacement of shrublands and woodlands in the 

global carbon budget are that Nevada semi-arid vegetation 

rapidly transforms from a carbon sink to a carbon source 

(Bradley and others 2006).

Twenty years ago, Young and Sparks (1985) made an 

ominous, and hopefully not prophetic, prediction about the 

sagebrush shrublands of the Great Basin:

If the burned sagebrush ranges are not restored, the 

alien weeds will inherit the sagebrush/grasslands. 

The way is thus paved for repeated burnings and a 

continuing downward spiral of degradation.

Over the 35 years I have lived in Nevada, spending much of 

that time in its wild mountains and basins, I have witnessed 

a continuing and expanding degradation of its vegetation 

formations. I was troubled by these changes, and wanted 

to get closer to the source of these problems. So, when an 

opportunity presented itself for me to go to Iran, I pursued 

it with enthusiasm and diligence.

Why Iran?

While they may have arrived here via Europe, most of the 

invasive species that are destabilizing Nevada ecosystems 

are native to Iran or elsewhere in the Middle East. Cheatgrass (Bromus tectorum) is aggressively invading upland 

shrublands and semi-arid woodland formations. Although 

not spreading as rapidly, saltlove (Halogeton glomeratus) 

continues to expand in the halophytic zone; occasionally 

forming pure stands over large areas. Russian-olive (Elaeagnus angustifolium) is mainly a problem in riparian areas 

in central Nevada, but salt-cedar (Tamarix ramosissima) is 

rapidly gaining the upper hand along streams and at springs 

throughout Nevada. It makes sense to see the landscapes 

from where these alien elements are native, in order that 

we may better understand the threats to our region.

Latitude, physiography, and climate are shared by Iran 

and Nevada. Major genera, both in number of taxa and their 

importance on landscapes, are shared by Iran and Nevada 

(such as Acer, Acacia, Amelanchier, Artemisia, Astragalus, 

Atriplex, Ephedra, Fraxinus, Juniperus, Prosopis, Prunus, 

Quercus, Salvia, and Suaeda).

Interior Iran and Nevada 

Physiographic Settings ______________

Iran is considerably larger than Nevada, occupying 

1,648,000 km2. This is roughly equivalent to the area of 

California, Nevada, Utah, Arizona, Colorado, and New 

Mexico combined (1,483,637 km2). Because of its size, Iran 

occupies a wider latitudinal range than Nevada, but virtually all of Nevada lies within the latitudinal range of Iran. 

Iran lies from 39° 46’ N to 25° N (fig.1), while Nevada occurs 

from 42° N to 35° N. A considerable amount of Iranian 

territory occupies coastal regions along the Persian Gulf, 

Indian Ocean, and Caspian Sea, environments that have no 

equivalent in Nevada. Nevertheless, the great interior of Iran 

is much larger than the Great Basin and Mojave combined, 

about half of which is in Nevada. My expedition to Iran 

traversed very nearly the latitudinal range of Nevada, 

from 32° 30’ N near Esfahan in the south, to 39° 30’ N at the 

Iranian-Turkish frontier northwest of Urumieh.

The physiography of Iran is similar to Nevada and the 

whole of the Basin and Range physiographic province (Hunt 

1967), which includes the Great Basin, and the Mojave, 

Sonoran, and Chihuahuan Deserts . Iran is a land of large, 

parallel mountain ranges with large intermountain, internally drained basins (Zohary 1973). While the western 

boundary of the Basin and Range province drains to the 

Pacific Ocean, the western slopes of the Zagros Mountains 

of western Iran drain into the Persian Gulf. The northern 

boundary of the Great Basin drains into the Columbia River 

and thence to the Pacific, much as the northern slopes of the 

Alborz Mountains of Iran drain into the Caspian Sea.

The Alborz Mountains are large, equivalent in length and 

breadth to the Sierra Nevada (both are 650 km long), but 

considerably higher than the Sierra Nevada and the Zagros 

Mountains. Although not taller than the Alborz Mountains, 

the Zagros Mountains are more massive, spanning 1,630 km 

in length and 400 km in width. Mount Dena is the highest point in the Zagros Mountains at 5,098 m (16,998 ft), 

compared to the highest point in the Sierra Nevada, Mount 

Whitney (4,417 m; 14,494 ft). Nevada has two small, active 

“rock” glaciers in cirque basins below Wheeler Peak in the 

Snake Range (Orndorff and Van Hoesen 2001) and Cougar 

Peak in the Jarbidge Mountains (Coats 1964). Iran has four 

clusters of small glaciers: at Mount Damāvand and 

Takht-e-Suleiman in the Alborz, and at Kūhhā-ye Sabālan 

and Zard Kūh in the Zagros Mountains (Ferrigno 1990).

The non-coastal climates of Iran and Nevada are remarkably similar. Both have interior climates in the rain-shadow 

of enormous mountain ranges that intercept storms coming 

ashore and strip them of their moisture. Not only are the 

regions semiarid to arid, but they are also characterized 
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Figure 1—Iran, its neighboring nations, and geographic features named in the text. 

The Alborz Mountains extend across northern Iran from Turkey to Afghanistan, 

connecting the mountains of eastern Europe with the mountains of Central Asia. The 

Zagros Mountains run nearly the entire north-south dimension of Iran along its 

western border. 

by mainly winter precipitation and have cold winters 

in the north and hot summers in the south. For instance, 

the northern Nevada city of Reno occurs at the base of the 

Sierra Nevada, much like Tehran sits at the base of the 

Alborz Mountains. Las Vegas and Esfahan are southern 

cities in hot deserts amidst treeless, arid mountains. Paradise Valley and Urumieh are both high agricultural valleys 

situated between moderately sized mountains that develop 

Table 1—Annual average temperature (T) and precipitation (P) for selected cities 

in Iran and Nevada. Iran data from the Iran Meteorological Organization 

(2006). Nevada data from the Western Regional Climate Center (2006).

Nevada Iran

City/Town T (°C) P (mm) City/Town T (°C) P (mm)

Paradise Valley 8.6 238.1 Urumieh 11.2 346.3

Reno 9.9 189.7 Tehran 9.5 240.7

Virginia City 8.8 337.0 Tabriz 13.7 348.0

Las Vegas 19.5 107.4 Esfahan 15.9 113.3

large snow packs. Virginia City and Tabriz are located at 

about 39°N and 38°N, respectively. Both cities occur near 

high mountains and at moderately high elevations, with 

Virginia City at about 1890 m A.S.L while Tabriz is situated at 1430 m A.S.L. In each case, not only are average 

annual temperature and precipitation close (table 1), but 

the monthly distribution of precipitation and temperature 

is also comparable (fig. 2).
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Charlet Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure

Figure 2—Climagraphs for selected city pairs from Nevada (left) and Iran (right), indicating monthly precipitation (mm) 

as bars and monthly average temperature (°C) as a line. From top to bottom the city pairs are Paradise Valley and 

Urumieh, Reno and Tehran, Virginia City and Tabriz, and Las Vegas and Esfahan.
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The elevation range of Nevada is large, from 146 to 

4005 m A.S.L (479 to 13,140 ft), but Iran has greater relief, 

rising from the Caspian Sea (–28 m A.S.L; [–92 ft]) to Mount 

Damāvand (5671 m A.S.L [18,600 ft]) in the Alborz Mountains. Two vast interior basins in Iran, the Dasht-e-Kavir 

and the Dasht-e-Lut are reminiscent of the Bonneville 

Basin of Utah, although far larger than Bonneville and not 

as dissected by mountain ranges as is the Lahontan Basin 

of Nevada, nor do they possess permanent lakes. However, 

Lake Urumieh in northwestern Iran, situated between the 

Alborz and Zagros Mountains, is very similar to the Great 

Salt Lake and the lakes even possess at the base of their 

food webs closely related brine shrimp (Artemia urmiana in 

Urumieh and Artemia franciscana in the Great Salt Lake) 

(D. Christopher Rogers, written communication 2006).

Today both Iran and Nevada are experiencing dramatic human population growth (fig. 3). And so at this time in both places, 

demands on water, minerals, and other natural resources, 

including the land itself, are escalating at an ever-increasing 

pace. Thus, comparing Nevada and Iran is like a “Natural 

Experiment” of Diamond (2005). The main difference 

between the two regions is that in Iran during the Holocene, 

the vegetation evolved in a context of agriculture, pastoralism, metallurgy, and denser human populations. Therefore, 

an investigation of the flora and vegetation of Iran should 

provide insights on management and conservation of the 

Great Basin and Mojave Desert of Nevada.

Figure 3—Human population growth in Nevada (top) and Iran (bottom), 1900 

to 2000. Both experienced geometric population increases during the 20th

century. The rate of growth slowed in Iran during 1990–2000 due to the IranIraq war. Data for Nevada from US Census Bureau (2006), data for Iran from 

Lahmeyer (2006).
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Charlet Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure

Vegetation in Iran __________________

Hot, non-saline deserts are occupied by a Zygophyllum

desert shrubland. These shrublands are reminiscent of the 

Mojave Desert’s Larrea shrublands, our only native member 

of the Zygophyllaceae. However, unlike Nevada’s Larrea

shrublands, Iran’s Zygophyllum is rarely accompanied by 

other shrub species. Instead, these monotonic shrublands 

extend across vast areas of the Turan Biosphere Preserve 

and vegetated portions of the Dasht-e-Kavir (Moore and 

Bhadresa 1978). Widely spaced chenopodiaceous shrubs, in 

genera such as Atriplex, Suaeda, and Haloxylon, dominate 

salt deserts in terminal basins. Upland xeric vegetation is 

largely dominated by short Artemisia shrublands, again with 

few other shrub species. Shrublands are omnipresent in the 

montane zone; these dominated largely by thorny members of 

the fabaceous genera Astragalus and Onobrychis. Similarly, 

subalpine habitats do not have multi-story, highly diverse 

shrublands. Instead, these are invariably dominated by a 

single low, shrubby Astragalus species. Inter-shrub spaces 

are dominated by a wide array of geophytes largely in the 

genera Lilium, Allium, and Tulipa whether in the subalpine, 

the montane, or the low deserts. Desert shrublands sometimes have grasses such as Stipa barbata, while montane 

shrublands are sometimes accompanied by bunchgrasses 

such as Agropyron spp.

As the expedition’s conifer expert, I expected to find some 

of the 10 species that were on my list of native woody species of Iran. Curiously, most of the conifer specimens in the 

five herbaria we visited were not part of the native flora 

at all, but instead were collections made from the respective University’s arboretum of non-native species such as 

Sequoiadendron giganticum.

As the expedition’s vegetation expert, I was stunned: here 

were familiar landforms, familiar conditions, but major 

vegetation elements were missing. While the landscapes 

before me were familiar, they also were eerie in what was 

absent.

Questions arose like warriors from the dragon’s teeth 

Jason planted in the Field of Ares. Where are the forests? 

Why is neither Pinus nor Picea even on my list? Where are 

the woodlands? I saw on my list that there were Quercus, 

Juniperus, Cupressus, Acer, and Pistacia. Of these, I ultimately found all but Cupressus, but the find was difficult 

in every case, and I never saw more than 10 individuals of 

any of them. Our planned trip to a relic Quercus forest in 

Kurdistan was turned away by military personnel, incredulous, as they were to find our permits and papers in order. 

Still, we went 100s of km through what appeared to be fine 

habitat for woodlands and montane forests, yet woodlands 

and forests were not there. Mountains like these abound 

in Nevada, yet Nevada’s mountain bases are cloaked with 

pinyon-juniper woodlands and their highlands are sprinkled 

with subalpine forests.

Of the American invasives that I was seeking to find, I 

saw Halogeton glomeratus, Bromus tectorum Elaeagnus 

angustifolium, and Tamarix ramosissima in Iran, but were 

dominant nowhere. I found Halogeton glomeratus once, as a 

rare, diminutive plant in a fertile river valley in Kurdistan. 

In only one place did I see wild Elaeagnus angustifolium, as a 

stand of about six trees along a mountain stream. I observed 

a full, healthy stand of Tamarix only once, northwest of 

Esfahan. Otherwise, throughout Iran I encountered only 

a few small Tamarix trees that rarely occurred along 

streams.

Similarly, I found Bromus tectorum at many locations, 

but nowhere was it dominant. North of Lake Urumieh was 

a nearly shrubless piedmont slope leading up to tall cliffs. 

It was a familiar landscape. If in Nevada, I would have 

known that I was looking at a recent burn in sagebrush 

that had been replaced by cheatgrass. Here was the largest 

patch dominated by cheatgrass that I saw in Iran, but it 

was barely 1 m2, and the reproductive individuals were no 

more than 10 cm tall. The annual grass that dominated the 

landscape was a long-awned, unpalatable Aegilops species. 

Cheatgrass does not dominate these landscapes. Other, 

more frightening weeds out compete the Iranian natives 

that wreak havoc in Nevada.

At the base of most mountains one can see a fascinating 

engineering feat (the Qanât ) that allowed for the expansion 

of agriculture and urban areas into places otherwise able to 

sustain humans only in a nomadic lifestyle. Persian engineers invented the Qanât during the Achaemenid Empire

by at least 2000 BP to extend agriculture into desert valleys 

to support the growing cities. A Qanât is a series of wells 

extending in a line from the base of the mountain out into 

the valley. The wells are all connected with a subterranean 

tunnel that collects the water and is tipped slightly out into 

the valley so that the water descends out near the piedmont 

base (Yazd Regional Water Authority 2003). From that point 

forward to the present, the phreatophytic halophytes such 

as Tamarix and the chenopodiaceous Suaeda aegyptiaca, 

Haloxylon persicum, and H. recurvum that occupy suitable 

areas of high water table must have been severely impacted 

over more than two millennia of this practice.

I left Iran richer, having gained many new colleagues, 

students, friends, and magnificent vistas permanently etched 

in my neurons. But I also left Iran with five disconcerting 

impressions concerning its vegetation: (1) the lack of trees on 

landscapes otherwise apparently well suited for them; (2)the 

“flat” nature of the native shrublands was puzzling; that is, 

the vertical structure of the shrublands was simplistic, typically dominated by a single shrub species at a single height, 

with very few other shrub species in the formations; (3) the 

replacement of shrublands by annual grasslands dominated 

by species that out compete cheatgrass; (4) the multitude 

of large, fresh, and active erosion features wherever we 

went; and (5) I observed the appearance of a heavy imprint 

of human use.

Archaeobotany of Iran______________

Upon my return to Nevada, I was compelled to find out 

how the vegetation came to be as it is. I examined the literature concerning the vegetation history of Iran. Here I found 

a wealth of information from archaeobotanical research 

supported by William Sumner and currently led by Naomi 

Miller (such as Miller 1996, 2001, 2002a,b, 2003a,b) at the 

University of Pennsylvania Museum.

Although the record is not geographically or temporally 

complete, the evidence allows a general picture of the development of civilization in a semi-arid, mountainous region. 
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Several critical developments occurred throughout the 

Middle East, including Iran, that led to one of the earliest 

developments of urbanized culture. The first permanent 

settlements appear in Iran at about 12000 BP (Miller 2004). 

By 10000 BP, the first crops were cultivated, and almost 

immediately the record shows the increase of weedy species. 

At about this same time, the ancient Persians began herding 

sheep, goats, and cattle. The introduction of pastoralism was 

coincident with the decline of palatable vegetation, and the 

ascendancy of armed and poisonous species. By 8000 BP, 

irrigation was applied to croplands in the Zagros Mountains 

(Miller 1992a). This allowed these early agricultural people 

some protection from annual variations in precipitation, 

providing communities with some insurance from droughtinduced crop failure. This security allowed the successful 

communities to grow in size, requiring more land clearance 

for the structures, more land clearance for more agricultural fields, and more sheep and goats in the mountains 

to support them (Miller 1992b). By around 4500 BP, major 

erosional events appear in the record (Wilkinson 1990, cited 

in Miller 1992a).

One site where extensive work has been done is Malyan, 

located at about 1,700 m A.S.L. in the southern Zagros 

Mountains, about 50 km from Shiraz (Miller 1985). Although 

it is not the earliest of sites nor was it the first to employ 

these technologies, the archaeobotanical evidence is complete 

enough to suggest a sequence of vegetation change that was 

likely repeated at numerous locations in the Zagros Mountains. People lived in settlements on the plain near Malyan 

since about 9000 BP. Malyan was settled at about 5400 to 

4800BP and at its height (ca. 4200 to 3600 BP) the urban area 

of Malyan had from between 13,000 and 26,000 people, with 

a permanent population associated with the city of between 

30,000 and 60,000 (Zohary 1973). Settlements like Malyan 

were usually associated with semi-arid woodlands ranging 

from Pistacia woodland-steppe at low elevations, through 

Pistacia-Prunus woodlands and forest, to Quercus forests, 

with Juniperus species occurring throughout (Miller 1985). 

Riparian areas were rich and possessed Ficus, Fraxinus,

Populus, Tamarix, and Elaeagnus (Miller 1985).

Throughout this time, coastal forest resources were exploited for buildings, ships, and weapons of war (Kuniholm 

1997; Wallinga 1993). About 5000 BP, the Persians used 

plaster and lined their homes and other buildings with it. 

Ancient production technology required burning two tons of 

limestone with two tons of wood to produce 1 ton of plaster 

(Miller 1992a). The development of metallurgy required 

fires much hotter than are produced by raw wood, straw, 

or dung. Charcoal was developed because it produces fires 

with sufficient heat and because direct contact of the ore 

with charcoal is required to produce pure copper (Rostoker 

and others 1988, cited in Miller 1992a). However, charcoal 

production is an extremely inefficient process, as most of the 

energy in the parent wood is lost in its burning to produce 

the charcoal (Horne 1982). Hence, the demands on the surrounding vegetation for the fuel necessary to support large 

urban areas were intensive and widespread (Miller 1992a). 

More rural landscapes were developed with agricultural, 

pastoral, mining, and other resource extraction activities 

to support the cities, casting a long shadow of human use 

on the landscape.

By 4500 BP, the woodlands were gone, Juniperus was 

virtually extinct, and the people were going high into the 

mountains to harvest oak at great cost in order to produce 

charcoal (Miller 1985). Residents of Malyan were burning 

dung as their domestic fuel (Miller 1985, 1996), as the high 

costs to ship wood from ever-farther distances reduced the 

use of wood for fuel (Miller 2002). Ultimately, Malyan was 

abandoned (Miller 1985). This pattern of settlement, development, and abandonment of small cities in the Zagros 

Mountains appears to be widespread throughout the region 

(Miller 2003c).

The Insurgency

We hear on the news that the government of Iran may be 

supporting insurgent activities in Iraq. I was interested to 

read the definition of this word:

Insurgent: [...present participle of insurgere to rise 

up, insurge] A person who rises in revolt against civil 

authority or constituted government: rebel; especially; 

a rebel who is not recognized as a belligerent. (Gove 

1961:1173).

In a different context, then, we can consider “insurgents” 

to be people who wish to improve their living conditions 

in the natural world by creating a civilization; people who 

rise up against the authority of nature and the limitations 

it imposes. They do this by developing and producing the 

tools of civilization from the natural resources about them. 

In this sense, then, it is the “insurgents” who cut forests to 

build a civilization. Trees are cut and converted to charcoal to 

smelt metals and make plaster. The land cleared of forests is 

plowed and irrigated for agriculture. Forests and shrublands 

are burned to increase grazing productivity. Alien species 

are introduced and spread through the modified landscapes. 

Surface and ground water is moved for increased agriculture 

productivity that supports urbanization. An empire is made 

from wood.

The result of this “insurgent” activity is a destabilization 

of ecosystems. This destabilization ultimately disrupts 

ecosystem processes and leads to a rapid simplification 

of vegetation structure. This simplification is characterized by a type conversion that transforms multi-layered 

shrublands to annual grasslands and converts woodlands 

to monotypic, one-layered shrublands. The biodiversity is 

reduced, and conditions ultimately promote a reduction 

of the flora to geophytes and armed and poisonous plants 

that can withstand heavy grazing and fire. Ultimately, the 

shrublands are converted to even simpler annual grasslands. 

The simplification of the vegetation results in widespread 

and severe erosion that may forever prevent the return to 

the pre-civilization state. To understand what forces were 

behind the creation of the contemporary Iranian flora and 

its dominant associations, we have to look no further than 

ourselves as human beings creating civilization.

Nevada Prehistory and History_______

Humans have occupied Nevada for millennia. The earliest 

human records in Nevada are from Paleoindian populations 

termed “Paleoarchaic” (sensu Beck and Jones 1997) and are 
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Charlet Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure

among the oldest in the continental United States. These 

earliest records date back to the Pleistocene/Holocene Epoch 

boundary about 11500 yr BP (Beck and Jones 1997), and 

Nevada has been inhabited ever since. The evidence indicates 

a rather abrupt change in tools around 7500 yr BP, suggesting a cultural response of the Paleoindians to climatically 

induced environmental change in the region (Grayson 1993). 

Perhaps as late as 1000 yr BP, the Numic-speaking Western Shoshone and Northern and Southern Paiute arrived, 

and successfully hunted and foraged in a nomadic lifestyle 

until the late 1800s (Elston 1986). Pinyon pine seeds were 

of particular importance for the survival of the Shoshone 

and Paiute (Fowler 1986).

While Native Americans elsewhere in the Southwest 

developed agriculture, pastoralism did not develop until after 

the arrival of the Europeans and their domestic herd animals. 

The oldest agriculture in Nevada is in the Moapa Valley in 

Clark County and Snake Valley in White Pine County, both 

near the eastern border, and both sites apparently developed 

by the Fremont culture around 1500 yr BP and occupied 

until about 900 yr BP (Fowler and Madsen 1986).

Shah-Kan-Daw in Nevada Pinyon-Juniper 

Woodlands: 1859-1899

Nevada changed forever in 1859 when gold was discovered 

near Virginia City (Angel 1881). The discovery helped bolster 

the western migration of people from the east to California 

and Nevada, while those already in California were also 

drawn to Nevada. Major anthropogenic changes to Nevada’s 

environment were inevitable and imminent. Not long after its 

founding in 1860, Virginia City became the largest city west 

of the Mississippi, and the first with an elevator, electricity, 

and an opera house. Development in Virginia City and the 

other Nevada towns that were born suddenly (fig. 4) led to 

intense use of natural resources throughout the state.

Figure 4—Thirty-five towns dispersed throughout Nevada were established at these locations from 1859–1869 

(Angel 1881). 
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The early Nevada settlements and towns used singleleaf 

pinyon (Pinus monophylla) and Utah juniper (Juniperus 

osteosperma) woodlands extensively. The trees were used 

for structures and fuel, both as firewood and processed as 

charcoal. Firewood was employed as fuel for all Nevada 

railroads until 1885, except in the case of the Virginia and 

Truckee Railroad that burned firewood into the early 20th

century (Wendell Huffman, Nevada State Museum, personal 

communication 2006). Firewood applications included heating all homes and businesses.

Virginia City required wood for structural lumber, both in 

the mines and for the city, especially in 1875 when the city 

needed rebuilding after the Great Fire destroyed most of it 

(Angel 1881). From 1874 to 1879, Virginia City consumed 

more than 300 million board feet of lumber from nearby 

Sierra Nevada old-growth forests that were fully stocked 

with yellow pines (Pinus jeffreyi and P. ponderosa) and 

western white pine (Pinus monticola) (Galloway 1947). 

This is a significant amount, considering that it is 

nearly equal to the total production of the Sierra Nevada 

(814,000 cubic m or 354 million bf) from 2000 through 2002 

(USDA Forest Service 2004). The city and its industries 

could afford the high cost of transportation to bring lumber 

from the Lake Tahoe area, first by wagon and later by 

train (Galloway 1947). Towns that were located farther 

from the Sierra Nevada had no access to its timber, so 

they had to make due with what was available locally. 

For instance, 24 sawmills were operating in Nevada in 

1868 (Lanner 1981). The locations of six sawmills are 

indicated on the 1908 USGS topographic map of the Las 

Vegas quadrangle alone.

Before trains bearing Sierra Nevada lumber could reach 

central Nevada, pinyon was milled producing “Reese River 

Lumber” (Angel 1881). Not until 1880 did the Nevada Central 

reach Austin, but even then, Reese River Lumber was used 

and sold for about half the price of Sierra Nevada lumber 

(Lanner 1981). The amount of pinyon that was produced was 

not trivial; in 1865, 5,660 cubic m (2.4 million bf) of Reese 

River Lumber was used in Austin alone (Lanner 1981). I am 

not certain how much Reese River Lumber was produced in 

the 1900s, nor do I know how to estimate how many board 

feet of pinyon can be produced in a single km2, but these 

lumber operations clearly had a profound ecological impact 

on the native landscape.

Virginia City mining operations used pinyon and juniper fuel relatively sparingly because they had smelting 

options that other Nevada towns did not. Still, its 30,000 

industrious residents had many needs that were met by 

the surrounding woodlands. The primary use of trees 

was for home heating, and this is likely to have taken a 

great toll on the surrounding vegetation in which they 

were immersed. From 1874 to 1879 Virginia City used 

1,140,000 cords of wood for fuel (Galloway 1947). A high 

estimate of cordage per acre of old-growth pinyon-juniper 

woodlands in Nevada is 3,000 cords/km2 (12 cords/acre) 

(Young and Budy 1979). Thus in these 6 years alone, 

Virginia City consumed 380 km2 of pinyon-juniper woodland for fuel.

Railroads

Railroads were vital for the development of Nevada 

and the exploitation of its natural resources. The Central 

Pacific railroad was begun through Nevada in late 1867 

after the line was built through the Sierra Nevada. The 

Atlantic-Pacific link was completed in May 1869, allowing other railroads to be constructed throughout the state 

before 1900. The most significant of these lines were the 

Eureka & Palisade, Nevada Central, Pioche-Bullionville, 

and the Virginia & Truckee. Nevada had no source of coal, 

and so Nevada railroads used wood-burning engines exclusively from 1868–1885, and the Virginia and Truckee did 

so into the 20th Century (Wendell Huffman, Nevada State 

Railroad Museum, written communication, 2006). Wood 

burning engines are easy to recognize due to their wide 

smokestacks fitted with spark arrestors. These railroads 

used a great deal of wood for fuel, and the ready source of 

fuel in Nevada was found in pinyon-juniper woodlands. A 

cord of wood was enough for powering a train an average 

of 6.4 km (4 mi) (Wendell Huffman, Nevada State Railroad Museum, written communication, 2006). By 1870, 

wood was so expensive that the Central Pacific began to 

buy coal from its archrival Union Pacific to supplement 

its wood use. Central Pacific paid $4.75/cord in Nevada 

(under $3/cord elsewhere in the West), but the Eureka & 

Palisade was paying $5.50/cord and Nevada Central was 

paying $7.00/cord (Angel 1881).

Summing the total length of track for each of the railroads, 

we find that in the 1800s there were 1,100 km (683 mi) of 

railroad in Nevada. If we assume that each railroad ran one 

train per day the length of its tracks in Nevada, then through 

the lifetime of wood burning for these railroads, 5,795,980 

km (3,601,455 mi) were traveled over the 33 years (table 2). 

At 6.4 km/cord and 3000 cords/ km2, we find that the total 

amount of pinyon-juniper woodlands consumed for railroad 

fuel was about 300 km2.

Charcoal

Nevada residents needed tools and services dependent on 

charcoal-fired forges. As a result, an impressive amount of 

pinyon-juniper wood was used to produce the charcoal. In 

1865, one Virginia City mill used 946 kg (21,500 bushels) of 

charcoal for fuel (Browne 1869). It took 100 cords of pinyonjuniper wood to make between 123,300–145,200 kg (2,800 to 

3,300 bushels) of charcoal (Young and Budy 1979). Thus, 

it required about 715 cords of wood, or about 0.24 km2

(60 acres) of pinyon-juniper forest to produce the charcoal 

for this one mill in Virginia City in one year. Virginia City’s 

ore was more easily processed than other Nevada mines, 

because water was more available and the ore was better on 

the Comstock, allowing the Washoe Pan Process (Oberbillig 

1967) to refine the ore. Thus, these Virginia City mills avoided 

the extreme charcoal needs of other early Nevada mining 

settlements, because the others lacked sufficient water and 

the nature of their “rebellious” ore (Oberbillig 1967) did not 

allow the pan process. Charcoal was the largest production 

expense (Lanner 1981).
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Charlet Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure

Outside of Virginia City, smelting required enormous 

amounts of charcoal, and this was produced locally from 

the pinyon-juniper forests. For instance, the central Nevada 

town of Eureka produced $60,000,000 of gold and silver 

and 200,000 metric tons of lead from 1869 to 1883 (Young 

and Budy 1979). Eureka was spectacular in its use of pinyon 

and juniper for charcoal and domestic and industrial fuel 

uses, ultimately sparking a “Charcoal Burners War,” as the 

price and availability of fuel wood became a serious economic liability (Earl 1979). In 1880 alone, 55,000 metric tons 

(1.25 million bushels) of charcoal was consumed in Eureka. 

At 30 bushels of charcoal from one cord, 41,666 cords were 

used to produce the 1880 charcoal. At 3,000 cords per km2

(Young and Budy 1979), the firewood used to make the 

charcoal resulted in 14 km2 (3,472 acres) of pinyon-juniper 

woodland cut in this one year for Eureka alone. As early as 

1874, a 32 km (20 mi) radius was cut, by 1878 the hauling 

distance was about 56 km (35 mi), and by 1880 this distance 

may have expanded to 80.5 km (50 mi) (Earl 1979). By 

1900, 582 km2 (600,000 acres) surrounding Eureka were 

denuded, of which about 24% was pinyon-juniper (Young 

and Budy 1979).

All other Nevada towns in the late 1800s had the same 

needs, but archaeologists have worked only in a few to 

establish the extent of deforestation. Using tree-ring crossdating of living pinyon pines, stumps, and archaeological 

features, Hattori and Thompson (1987a,b) found evidence of 

three episodes of intense woodcutting in the Cortez Mining 

District in north-central Nevada from 1863 to 1904. Hattori 

and Thompson (1987a,b) do not provide an estimate of charcoal used or total area deforested, but they do have data on 

monthly wood consumption for the district. They concluded 

that the claim that a team of 200 woodcutters worked the 

area (Ashbaugh 1963) is an overstatement. Hattori and 

Thompson (1987a,b) found that a few trees were left standing, 

and so proposed that the area was not entirely deforested, 

but nevertheless seriously impacted.

Nathan D. Thomas, an archaeologist, has examined the 

record at Ward, in White Pine County (Thomas 2006). This 

short-lived mining town had 1,500 people, a post office, two 

newspapers, and even two breweries in 1877 (Thomas 2006). 

Thomas worked diligently to estimate how much wooded area 

was impacted to supply both the charcoal ovens and domestic 

uses. Thomas (2006), like Hattori and Thompson (1987a,b) in 

the Cortez Mining District, doubts that complete deforestation 

took place at Ward, and contends that modern White Pine 

County locals are incorrect in their claims of the woodlands 

being completely clear-cut in a 32 km (20 mi) radius around 

Ward. Notwithstanding the overstatement, Thomas (2006) 

concludes that woodcutting had a highly significant impact 

on the landscape, and provides data that allow an estimate of 

the areal extent of the wood removal operations.

Likewise, Zeier (1987) examined the archaeological record 

of charcoal production near Mt. Hope, about 32 km (20 mi) 

northwest of Eureka. Zeier (1987) studied temporary surface 

ovens, habitation and other working sites and concluded that 

the area was probably not completely deforested. The archaeologists all acknowledge the possibility of a known practice 

of stump removal for fuel, complicating interpretations of 

the available record. From the archaeological record so far 

assembled, it appears that the complete removal of trees 

in all areas where charcoal was produced probably did not 

occur. It is also clear that the woodlands that were originally 

cut consisted of much larger trees and were more widely 

spaced than the stands that have replaced them (Hattori 

and Thompson 1987a,b; Thomas 2006). We also know that 

a staggering amount of wood was used for a great many 

purposes. If the pinyon-juniper woodlands were not entirely 

cut, then I underestimate the area impacted by the woodcutting operations because the operations had to be carried out 

over a larger area to obtain the same amount of wood.

Ranches

Hundreds of ranches sprang up around Nevada to support 

the new towns that arose around the mines. Ranches also 

needed wood. Not only were pinyon and juniper wood the 

primary sources of household fuel, but also they were used 

to build fences, corrals, and homes (Young and Budy 1979). 

The impact on woodlands due to juniper cutting for fence 

posts was profound. For example, the corrals at Walti Hot 

Springs Ranch in central Nevada are made of 3,000 juniper 

poles. Most ranches built fences from juniper posts at 161 

posts per km (260 posts per mi) (Young and Budy 1979). 

Railroads and ranching activities both contributed to an 

extremely high incidence of both intentional and accidental 

wildfires during the late 1800s, further shocking the vegetation formations (Lanner 1981).

Table 2—Nevada railroads, their length, and the years they operated while using wood for fuel. Distance traveled calculated by assuming one train travels the full length of the railroad every 

day during the life of the railroad. Fuel efficiency average was 2.48 cords per km (4 cords 

per mile) (Wendell Huffman, Nevada State Museum, personal communication).

 Distance

 Length Wood as fuel Number traveled if one Cords

Railroads (km) (dates) years train/day (km) wood used

Central Pacific 723 1868-1880 13 3,428,715 532,626

Eureka & Palisade 135 1874-1889 16 789,479 122,640

Nevada Central 150 1880-1889 10 546,292 84,863

Pioche-Bullionville 8 1873-1881 8 23,496 3,650

Virginia and Truckee 84 1868-1900 33 1,007,997 156,585

 Total 900,364
USDA Forest Service Proceedings RMRS-P-52. 2008 15

Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure Charlet

Cumulative Impact on 

Pinyon-Juniper Woodlands

From the earliest settlement days, incoming Europeans 

valued the pinyon and juniper of Nevada only as fuel. To 

livestock operations, these woodlands had less available 

forage than the surrounding shrublands, and so were seen 

as worthless. These forests and woodlands had no value 

to either timber companies or the government, so if they 

were necessary to use for shafts in the mines, or reduced 

to charcoal for the smelting of ore, then they should be so 

used. This right was affirmed by the “Act of July 26, 1866” 

(Hattori and others 1984). The U.S. Senate (Committee on 

Mines and Mining 1889:I-II, cited in Hattori and others 1984) 

stated that “depriving or charging miners for worthless trees 

would, in effect, be un-American” (paraphrased by Hattori 

and others 1984:7).

Sargent (1879) and others opposed the wanton destruction of the pinyon-juniper forests and woodlands, but the 

Congressional Act of 1878 permitted the felling of any tree 

on public land with minerals for useful purposes by any U.S. 

citizen (Hattori and others 1984). This law was applied to 

stop Italian and Mexican woodcutters in 1888 (Hattori and 

others 1984), and numerous lawsuits were swiftly brought 

forward against non-Americans. American miners and charcoal makers had no such restrictions imposed on them. But 

miners and mills were in direct conflict once the “worthless” 

trees were in short supply. The miners’ use of wood to shore 

up shafts and build structures depleted the forests gradually, but the cutting of trees for fuel had an immediate and 

devastating effect (W.A.J. Sparks correspondence, Committee 

on Mines and Mining 1889:31, cited in Hattori and others 

1984). Not until 1907 was there any protection offered to 

pinyon and juniper on public land in Nevada, but this protection amounted to little more than a permit requirement, the 

marking of a U.S. stamp on the stump by federal rangers, 

and a fee of $1 per cord (Hattori and others 1984).

The problem of determining the areal extent of pinyonjuniper woodlands in 1859 is daunting. The current sum 

of all juniper, pinyon-juniper, and pinyon woodlands as 

mapped by Nevada GAP (Edwards and others 1996) is 

about 30,000 km2. I have pinyon and juniper wood use data 

only for the towns of Virginia City (Galloway 1947), Eureka 

(Young and Budy 1979), Cortez (Hattori and others 1984), 

and Ward (Thomas 2006). Only the record for the smallest 

of these towns, Ward, is complete. I have derived an (under) 

estimate of total pinyon and juniper wood used by Nevada 

railroads by assuming that only one train ran the length of 

track each day of operation. Still, in compiling these scant 

data we find that more than 1,200 km2 were clear-cut, or 

nearly so, which is more than 4% of the entire area covered 

by pinyon-juniper woodlands in the state today (table 3). If 

we extrapolate these uses out to the rest of the at least 35 

towns in the state that arose from 1860–1869—and what 

was likely to be 100 by 1900—it is clear that while the 

pinyon-juniper forests were not wiped out, their range was 

significantly decreased, and the shock wave of the disturbance sharply resounded throughout the formation. Clearly, 

the view of the Western Shoshone was that this impact was 

devastating, as recounted here:

The trees that bore pine-nuts were all cut down 

and burned in the quartz mills and other places... 

(Tsa-wie, Western Shoshone Headman, quoted by 

Indian Agent Levi Gheen [1876:521] in Hattori and 

others 1984).

Moreover, the impact was significant in the view of prominent foresters of the day:

...in view of the vast importance of their [central 

Nevada mountain ranges] remaining wooded to serve 

as reservoirs of moisture..., it would seem wise ... to 

check, or at least to regulate, that terrible destruction of forest, which follows..., every new discovery 

of the precious metals (Charles S. Sargent 1879, The 

Forests of Central Nevada, quoted in Hattori and 

others 1984).

Shah-Kan-Daw in the Shrublands: 

Early Nevada Ranching

The towns needed to be fed, and much of the food was 

produced locally. The conversion of valley meadows and 

sagebrush shrublands to agricultural production devastated 

Table 3—Pinyon-juniper woodlands deforested in late 1800s to provide fuel for railroads and to support economic growth in the 

Nevada towns of Virginia City, Eureka, and Ward. The conservative estimates here derived from the railroad use and 

these four towns and years alone, deforested about 1218 km2, or more than 4% of the current land occupied by pinyonjuniper woodlands in Nevada.

Area deforested Deforestation Years of record

(km2) purpose Location or estimate Reference

304.0 Railroad fuel Throughout Nevada 1865-1899 Huffman, pers. comm. (2006)

768.9 Firewood & charcoal Virginia City 1860-1885 Young and Svejcar (1999)

139.7 Charcoal Eureka 1869-1889 Young and Budy (1979)

0.6 Charcoal Cortez 1897 Hattori et al. (1984)

3.5 Charcoal Ward 1872-1888 Thomas (2006)

1.3 Domestic use Ward 1872-1888 Thomas (2006)

Total deforested: 1,218
16 USDA Forest Service Proceedings RMRS-P-52. 2008

Charlet Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure

both plant formations wherever this occurred. However, a 

much greater areal-extent of Nevada landscapes were affected by beef and sheep production.

...the Great Basin was the last natural grazing land 

to be exploited in western North America and, most 

probably, the world. (Young and Sparks 1985)

Nevada cattle and sheep ranches in the late 1800s were 

unregulated. They used no supplemental feeding of their 

herds, and so operated on enormous open ranges that possessed vegetation that could provide both winter and summer feed (Young and Sparks 1985). The operators rapidly 

increased their herds to supply the growing population. 

By 1888, there were more than 300,000 cattle on the 

range. This initial shock to the shrublands was turned back 

when two severe winters (1888–1889, 1889–1890) caused 

tens of thousands of cattle to freeze to death (Young and 

Sparks 1985). The ranchers responded well to the challenge 

by changing breeds, and by learning to grow alfalfa hay in 

the summer to provide over-wintering feed. Range practices 

evolved in the 1900s, allowing Nevada ranchers to maintain 

more than 500,000 cattle every year from 1951 through 2002 

(fig. 5) (USDA Nevada Agricultural Statistics 2006).

Restoration in Iran Today ___________

The problem facing Iranian resource scientists today is 

not advancing pygmy conifers replacing their shrublands. 

Instead, Iranian resource scientists and land managers are 

dealing with advancing sand deserts replacing shrublands. 

Iranian restoration biologists are working in collaboration 

with scientific teams from Germany to stabilize sand dunes 

in many parts of the nation. They irrigate Tamarix in plantations, and then out-plant them to sand dunes. The dunes 

are pre-treated by spraying them with oil to temporarily 

stabilize them until the Tamarix can hold them in place 

(Geisbauer 2004a,b; Khosroshahi and others 2006; Morell 

2004; Pakparvar 2004).

Iranian soil scientists make recommendations for proper 

plowing and soil conservation, but the farmers and shepherds treat the recommendations as mere suggestions. 

I saw Iranian farmers in Kurdistan plowing native steppe 

on steep slopes for the first time with their tractors going 

up and down the slopes instead of following the contour. I 

queried the soil scientist with whom I witnessed this. Why 

were they doing that, and did they not know of the potential 

for an erosion disaster this could cause? He said they did tell 

them, but that the farmers explained it was easier on the 

tractors this way. Immediate practical matters overwhelm 

long-term concerns. That evening powerful thunderstorms 

caused massive flooding, nearly preventing our return to 

Urumieh because of several bridge washouts. The immediate 

practical matters suddenly became the crisis caused by that 

morning’s practical, but ill-advised, decision.

Afforestation projects consist of monoculture plantations, 

generally of non-native Austrian pine (Pinus nigra) or Arizona 

cypress (Cupressus arizonica). They are seen as greenbelt 

production forests, evenly spaced in rows and with no understory. In the valley floor south of Tehran, these plantations 

are irrigated by wastewaters that flow through the city of 

Figure 5—Number of cattle in Nevada, 1867–2002. Nevada Agricultural Statistics (2006). 

Note two periods of sharp decline, one following massive die-offs in the latter 1800s, 

and another prior to the Great Depression. Total number of cattle in Nevada has been 

at levels above 500,000 every year since 1951.
USDA Forest Service Proceedings RMRS-P-52. 2008 17

Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure Charlet

12 million people. Significant forest conservation efforts are 

rightfully placed in the Caspian Basin, with climate and 

vegetation that has no analogue in Nevada (Mooney 1956; 

Zohary 1973). Here, Hyrcanian forests, relics of the Tertiary, 

persist in a few locations (Zohary 1973). These forests are 

under tremendous pressure and are in dire need of conservation (Caspian Environment Program 2005). However, there 

seems to be no effort to restore woodlands or forests in the 

semi-arid interior of Iran. More than 30 years ago, Zohary 

(1973: 652) recognized that many areas formerly covered 

with woodlands had lost so much soil that they no longer 

even had the potential to become woodlands again:

Large areas of the Middle East have lost their arboreal 

vegetation forever, especially those areas of woodland 

situated on the borders of steppes and deserts. Here 

conditions for tree growth are marginal and moisture 

deficiency and soil erosion are impeding regeneration 

of the destroyed vegetation. Living testimony in the 

form of relics, remnants, single shade trees, sacred 

forests, cemetery trees, etc., located amidst timberless lands tell the story of the lost forests and of the 

climax vegetation that did not come back centuries 

after destruction.

Nevada Today _____________________

What are we creating in Nevada? It seems that by following 

practices initiated in the Nineteenth Century ecologically 

we are turning Nevada into Iran. Who are the insurgents 

responsible for this change? They may be difficult to recognize, but we have met them, and they are we. As we take up 

arms (plows, drill rigs, bullhogs, and backhoes) against the 

established natural order of our landscapes, we fall into the 

same dilemmas that faced the amazing people that created 

the great Persian civilization.

When I moved to Nevada in 1971, the population of the 

state was nearly 500,000; now it exceeds 2,000,000, a fourfold increase in 35 years. The growth is not going to stop 

anytime soon because Clark County leaders are convinced 

that any cessation in growth will have immediate and dire 

consequences for the local economy (Hobbs and others 

2004). Obviously, urbanization destroys and re-creates 

valley vegetation. Urban centers incidentally increase the 

fire frequency in a halo on their periphery, a periphery that 

is rapidly moving upslope out of the valleys and into the 

mountains (Brussard and others 1999).

To support our rapidly growing communities, we construct 

more utility corridors, further fragmenting the region and 

providing habitat and migration corridors for alien invasive 

species. These corridors, if left undisturbed after construction, 

will take centuries to recover (Lathrop and Archbold 1980). For 

instance, the Southern Nevada Water Authority (SNWA) is 

aggressively pursuing water rights from throughout Clark, 

Lincoln, Nye, and White Pine counties, in order to build a 

2.13 m (84 in) pipeline, estimated in 1989 to cost $2 billion 

for a 1.52 m (60 in) pipe, to bring enough water for 425,000 

new homes in the Las Vegas and neighboring valleys (BLM 

2004, 2005a).

Our activities replace forests with shrublands, and simplify 

the vertical dimension of these shrublands. Our activities 

replace shrublands with grasslands dominated by annual 

alien species from the Middle East. We are degrading 

riparian areas with alien species from the Middle East, and 

are reducing biodiversity of fauna and flora everywhere. 

Historically, mining activities laid waste to large tracts of 

pinyon-juniper woodlands for charcoal production, and the 

railroads contributed their own assault on these woodlands for 

the cordage they provided to the early trains. Today, mining 

destroys the land itself by crushing whole mountains and 

spraying the resulting piles of ore with cyanide solution.

We have had many conferences on the topic of our deteriorating shrublands and woodlands of the interior West. For 

instance, in my personal library, I have the following volumes 

of USDA Forest Service conferences on Western Range Restoration: Pinyon-Juniper Conference (1987), Desired Future 

Conditions for Piñon-Juniper Ecosystems (1994), Wildland 

Shrub and Arid Land Restoration Symposium (1995), 

Ecology and Management of Pinyon-Juniper Communities 

Within the Interior West (1999), Restoring Western Ranges 

and Wildlands (2004), and now I will soon be able to add 

Shrublands Under Fire (2006). Throughout these volumes, 

we find many ideas about how to control the expansion of 

pinyon and juniper woodlands. A lone voice of the Native 

Americans was present in these proceedings. Wassen (1987) 

felt that there was much talk about removing the pinyon 

and juniper, but little consideration for what was to replace 

them. Wassen (1987) eloquently expressed the view that all 

living things in the forest are “intertwined in the chain of 

life” and that to the Indian the pinyon is the grocery tree and 

the juniper the medicine tree. “Therefore, the cutting down 

of a single living tree is sacrilegious—the cutting down of a 

forest—UNTHINKABLE” (Wassen 1987:39).

Now we are looking to the pinyon and juniper to provide fuel 

again, calling it renewable energy (Fadali and others 2005; 

Harris and Dick 2005). We continue to manipulate the vegetation for preferred species. Instead of “range improvement” 

projects of the 1950s through 1970s that involved chaining 

of pinyon and juniper and planting with crested wheatgrass 

(Lanner 1977), we are now conducting “habitat restoration” 

projects—presumably for mule deer and sage grouse—that 

involve chaining of pinyon-juniper and planting of wheatgrass 

(BLM 2005b). Upon closer inspection, we find that this project 

(BLM 2005b) is designed to also provide benefit for the grazing permittee, as well as for the non-native wild horse herds. 

We also have a new mechanical “treatment” technology—the 

“bullhog”—that can chew up pinyon trees and spit out landscape 

mulch. This is seen as progress because the old chainings left 

many supple young pinyon and juniper saplings that actually 

increased their local dominance on the site over time. Now we 

can get rid of them once and for all.

We use the herbicide Spike 20PTM to kill big sagebrush 

(Artemisia tridentata ssp. tridentata), while Tordon 22K TM

is used to kill rubber rabbitbrush (Chrysothamnus nauseosus) 

and greasewood (Sarcobatus vermiculatus) (Williamson and 

Parker 1996). I found the rationale fascinating: “Although 

livestock and wildlife will browse it, the plant [rabbitbrush] 

rates low in palatability. It is considered a problem because 

it tends to increase where soils have been disturbed, such 

as on overgrazed rangelands, at the expense of more desirable plants” (Williamson and Parker 1996:1). We need 

to poison the native plants that respond to overgrazing 

disturbance because those that overgraze will not eat it? 
18 USDA Forest Service Proceedings RMRS-P-52. 2008

Charlet Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure

Similarly, greasewood is “undesirable” because “if eaten in 

large quantities by livestock it may be toxic” (Williamson and 

Parker 1996). If cattle overgraze to the point where there 

are only toxic plants left, they will eat the toxic plants. How 

do we conclude from this that we must poison the dominant 

native shrub? If the real interest was a healthy ecosystem, 

we could simply remove the source of the disturbance, which 

is overgrazing. Today we are using a new tractor-pulled 

mower to destroy sagebrush in eastern Nevada in the name 

of restoration (Eastern Nevada Landscape Coalition 2006). 

It seems to me that something is fundamentally wrong with 

activities that are designed to destroy the native, dominant 

species in the region.

Considerations of and 

For the Long Term__________________

We know that the ranges of both pinyon and juniper 

have expanded and contracted greatly in the past (Miller 

and Wigand 1994; Tausch 1999a), and there is no reason to 

believe that this will change in the future. Moreover, while 

we have a good idea of the long-term changes throughout the 

past 50,000 years or so, we do not know exactly what those 

changes look like. I imagine that range shifts in pinyon and 

juniper did not occur slowly and evenly, but instead were 

dynamic; that is, there are pulses of recruitment and pulses 

of dieback. The organisms we are considering here live much 

longer than we do. Tausch and others (1993) remind us that 

it is extremely difficult to recognize natural or climatic vegetation change going on around us, much less know what 

direction it will be taking in the future.

It is difficult to interpret the snapshots of vegetation 

communities that we see on the ground. That is, what is 

the landscape like when pinyon and/or juniper woodlands 

expand their range? We should see recruitment both in the 

understory and beyond the limits of the existing forest or 

woodland. For a moment, there may be excellent establishment success of pinyons and junipers. That pulse may last 

for 5–10 years or 50 years, and at any time could be halted 

by a killing freeze that nails all the saplings. The climatic 

factor of cold winter low temperatures induced by cold-air 

drainage to valley bottoms, not repeating fires, is what appeared to Billings (1954) to prevent pinyon-juniper expansion 

onto valley floors, while a warm thermal belt allowed them 

to dominate lower mountain slopes.

Pinyon-juniper woodlands with closed canopies and no 

understories also represent a type of vegetation simplification. I know there is genuine concern over pinyon-juniper 

stands where canopy closure is complete and understory 

vegetation is absent or nearly so (Jameson 1987; Pellant 

1999; Tausch 1999b). However, in Nevada this condition 

is the exception, not the norm. In the past 18 years, I have 

visited 285 of Nevada’s 314 named mountain ranges (Charlet 

1996 2007). I visited most of these ranges multiple times, 

often in the same year, and many for extended periods of 

time. Counting all visits to a mountain range in any year 

as one visit, I have averaged 5 visits per mountain range 

(n = 285) over these 18 years. I have made more than 1,300 

collections of conifers in 183 of these ranges and keep track 

of vegetation wherever I go. I have seen canopy-closure with 

no understory only rarely in Nevada’s pinyon and/or juniper 

woodlands. Stands that are approaching canopy closure I 

judge as being uncommon.

Are pinyon-juniper communities with canopy closure and 

no understory steady-state “pseudo-climax” communities 

with no change in their future, or, will they also change? I 

submit that these stands are already changing. The canopy 

closure we see in many stands of pinyon-juniper may simply 

be an early succession phenomenon that will be arrested by 

self-thinning services provided by the innumerable organisms 

of our biodiversity that are waiting in the wings to capitalize on the abundance of pinyons and junipers. Weber and 

others (1999) report that there are 16 pathogenic organisms 

responsible for mortality of pinyon, and 15 for Utah juniper. 

Most recently in my surveys, in closed stands and in stands 

approaching closure, I observe self-thinning in the form of 

beetle-kill, mistletoe-kill, and other pathogens. These events 

create the openings in the pinyon-juniper formation that the 

land managers seek. This is one way that old-growth pinyonjuniper forests are created. Alternatively, the pinyon-juniper 

expansion we think we see may simply be a minor interval 

of recruitment that will be quelled by a catastrophic event 

such as a deep freeze event that occurred in western juniper 

(Juniperus occidentalis) woodlands recently in eastern Oregon (Knapp and Soulé 2005). I am certain that events like 

this are commonplace in the timeline of pinyon and juniper 

woodland range shifts.

Jameson (1987) elegantly proposed that there are multiple 

pathways and potential states possible in vegetation change 

in pinyon-juniper woodlands. Tausch (1999b) expanded on 

transition and steady-state models, and discussed critical 

thresholds in the Great Basin pinyon-juniper woodlands, 

and how consideration of these models can inform our 

management decisions. How well do we understand these 

concepts, and how well have we tested the propositions and 

the unintended consequences of our actions? It seems that 

whenever pinyon and juniper are seen that they are now 

perceived as invading. We are over-generalizing and overreacting by taking drastic actions.

I am concerned that as we proceed with our plans that the 

projects themselves will serve to introduce more disturbances 

and further destabilize our systems to the point where they 

may cross a threshold (sensu Tausch 1999b) that prevents 

the return to woodland formations. The power that we can 

wield on the region is awesome, and it may be that we effect 

a change as great as Zohary (1973: 654) senses has occurred 

in the Middle East.

This change of the flora, vegetation and the animal 

world, has already reached dimensions never paralleled in the history of the region. In other words, 

man has introduced here floristic and distributional 

changes that in their magnitude exceed by far those 

caused by Pleistocene climatic changes.

Management Implications ___________

I think we need to slow down with our restoration efforts. 

Eddleman (1999) offered five guidelines for the ecological 

management of pinyon-juniper woodlands. As enumerated by 

Pellant (1999), the first guideline for ecological management 

of pinyon-juniper woodlands is to establish clear goals and 

objectives. I am concerned that we have not yet accomplished 
USDA Forest Service Proceedings RMRS-P-52. 2008 19

Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure Charlet

this. We need to decide what we want our landscapes to 

be in accordance to their potential. We need to develop an 

honest, realistic vision of what our region will look like in 

5, 20, 100, 1,000 years into the future. Are we managing for 

healthy ecosystems, or are we managing for the artifacts of 

our destruction of the system in the first place: the mines, 

cattle, and horses? Are we managing for fire because of the 

health of the ecosystem, or because of the danger such a fire 

may pose to utility corridors and expanding town limits? It 

is difficult to be aware of one’s own bias, but it affects how 

we interpret what we see. A shepherd sees young pines 

on a shrub-steppe and says the pines are encroaching and 

should be removed. A forester looking upon the same scene 

may say instead that the pines are insurance for the future 

of his forester sons and should be pruned and thinned and 

allowed to grow to large size. A fireman sees a fuel model 

and wants to remove fuels. An ornithologist sees resource 

partitioning, and wants to construct a blind. A soil scientist 

sees a soil type and wants to dig a hole. A miner sees supports 

for shafts, while a charcoal maker sees starting material. 

A developer sees a new subdivision, and I see biodiversity 

(fig. 6). How much longer will our population increase? Will 

we occupy every valley with human development? The habitat conversion from shrubland to houses is going to have a 

much greater impact on sage grouse than does the current 

pinyon-juniper “encroachment.” We need to engage in an open 

discussion about the future of the valleys and mountains of 

Nevada and the life that is teeming there, and see if we can 

hold a shared vision of our future in the midst of a landscape 

whose vegetation will constantly change.

Eddleman’s (1999) third guideline for ecological management of pinyon-juniper woodlands is to develop baseline 

data by conducting an inventory to determine understory 

vegetation and functional status of stands. As far as I 

can tell, the baseline data of pinyon-juniper communities required to inform wise management has not been 

assembled. Since Eddleman (1999), other than my own 

cursory surveys (Charlet 1996, 2007), only Tausch (work in 

progress), McArthur and Sanderson (2006), and Greenwood 

and Weisberg (2006) have endeavored to obtain a sense of 

the breadth of variability in pinyon-juniper communities 

across the region, each using different methods. More 

efforts of this kind are desperately needed to inform us 

as to the real dynamics on the ground, rather than our 

impressions from a limited area of operations. Moreover, 

it would be superior to integrate the different approaches 

and develop a statewide vegetation inventory database. 

Analyses can be performed on the database to better 

understand the controlling factors of the variations in 

vegetation. Monitoring will be more informative so that 

statewide trends can be identified and quantified. But at 

present, our knowledge of the distribution and condition 

of shrublands and woodlands in Nevada is good in general 

but poor in the particulars.

For as much as we know, we still lack a great deal of basic 

research on these ecosystems. Our ignorance of the basic 

processes that are operating in the ecosystems of Nevada 

is so great that I am concerned when we take large-scale 

management actions based on what we assume we know. 

Instead, these actions exacerbate the problems by introducing more disturbances, creating more opportunities for alien 

weeds to colonize and ultimately dominate the landscapes 

of Nevada. We are reminded in our Environmental Assessment forms that we should consider the consequences of 

both action and no action. Tausch and others (1993) pointed 

out that “no action” is still a management action. I agree, 

Figure 6—Biologically diverse shrublands with structural complexity are rare in Iran 

and are rapidly disappearing in Nevada. This complex Mojave shrubland remnant was 

spared from flattening at the new Coyote Springs Valley housing development, Clark 

County, Nevada, May 2006. Photograph by the author.
20 USDA Forest Service Proceedings RMRS-P-52. 2008

Charlet Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure

and I propose that “no action,” in most cases, is superior to 

aggressive action in the absence of baseline data and experimental design. “No action” is superior to “experiments” 

with no replication and insufficient controls. It is the “no 

action” option that is enthusiastically embraced by Romme 

and others (2006) concerning the natural, “self-thinning” 

fire protection service provided by forest insect outbreaks 

in Colorado forests.

Currently state and federal agencies are conducting and 

planning dozens of high-impact manipulative projects on 

vegetation throughout Nevada. Each of these projects represents an opportunity for experimentation and thus learning 

from our actions. Instead, virtually none of the projects are 

designed with replication or even proper controls. Lacking 

such design, the projects keep us busy, but we learn nothing 

from them. Worse, we may create more harm than good. We 

act as if we know the role and history of fire in pinyon-juniper 

woodlands, when in fact our knowledge base of fire history 

in Nevada is very poor (Baker and Shinneman 2004).

For example, the ambitious restoration plan for Spruce 

Mountain (BLM 2005b:3) states in its introduction that “a study 

published in 1976 identified ... improper livestock grazing, wild 

horse use and abuse, and pinyon-juniper encroachment into 

sagebrush/perennial grass communities” as the three major 

factors causing decline of mule deer winter habitat on Spruce 

Mountain. The plan seeks to remove pinyon and juniper, but 

addresses the other two causes of degradation (wild horses 

and cattle) only by changing the management levels in order 

to “attain multiple use objectives” for the allotments (BLM 

2005b:3). It may be that the grazing enhances pinyon-juniper 

“encroachment” (Miller and Wigand 1994). Nevertheless, the 

pinyon-juniper woodland is demonized and removed but the 

cattle and horses will remain.

Later, the proposal continues by stating, as before without 

citation or reference, that “Studies show that the expansion of 

pinyon-juniper has more than tripled in the areas dominated 

by pinyon-juniper woodlands within the last 150 years” (BLM 

2005b:5). I am unaware of any study that proposes a threefold expansion of the range of pinyon-juniper woodlands in 

the past 150 years in Nevada, much less multiple studies 

as the proposal states. The proposal states that one of the 

contributing factors to this shock wave of pinyon-juniper 

expanding throughout Nevada was that the forests were cut 

for the mines. But how is that possible? Mining in Nevada 

began in 1860, more than 150 years ago, and the industry 

consumed large amounts of the woodlands for fuel. Surely 

the trend from 1860 to 1910 was a drastic decline in the 

distribution of the forests. I submit that we are, in most 

cases, witnessing a reoccupying of pinyon-juniper range as a 

response to prior human actions that included cutting many 

of the forests down to bare ground, or nearly so.

Every summer we have large-scale fires consuming 

pinyon-juniper forests and woodlands in Nevada. When 

the woodlands regrow, do we call this an “invasion”? It is 

a reoccupation. When they are burned, do we count it as 

a loss? Photographic time series studies such as Rogers 

(1982), fascinating and informative as they are, usually 

have photographs available at the earliest in the early1900s, 

long after most of the initial damage was done. If an area 

is clear-cut in 1880 and photographed in 1900, the site will 

not appear to be woodland. If the woodland recovers then 

we will see an increasing amount of trees and they will be 

getting bigger throughout the photographic history. The 

statement that pinyon-juniper woodlands have increased 

threefold in 150 years seems like an example of data distortion that occurs when one whispers something to a person 

who whispers it to the next and on the story goes around 

the room until at the end it is unrecognizable as the original statement. It is possible that the authors of this report 

had read this passage in Tausch (1999b:362) and misread 

“third threshold” as three-fold: “The outcome of this [third] 

threshold has been the dramatic increase in the area and 

dominance of pinyon-juniper woodlands that has been progressing largely unrestricted over the last 150 years.” This 

“third threshold” for vegetation change that Tausch (1999b) 

proposed was crossed as we passed out of the Little Ice Age. 

This released certain potentials that before were inaccessible. 

“Abiotic conditions and associated patterns of disturbance and 

succession that prevented this [pinyon-juniper expansion] 

in the past are gone” (Tausch 1999b:362). But potential is 

one thing and reality is another. The potential for increase 

of pinyon-juniper woodlands to expand may have turned 

150 years ago, but have they really been expanding all this 

time? I submit that, in light of the above discussion, in the 

first 50 of the past 150 years there was a dramatic decline 

in the extent and density of the pinyon-juniper woodlands. 

It may be that this is a phase of pinyon-juniper expansion 

because of a peculiar confluence of variables that turned 

to favor woodland development at the close of the Little 

Ice Age, but the climatic variables at least will once again 

turn to favor shrublands over woodlands. Given that both 

the potential and the reality of pinyon-juniper expansions 

and contractions have occurred many times in the past, 

we should have no expectation that their distributions and 

abundances should remain the same. I submit that a phase 

of expansion—if that is what we are seeing—is not necessarily a “bad” thing; it simply is what is happening. Getting 

in the way of the natural trend will be painful, expensive, 

and frustrating for us.

Likewise, successful management will work with, instead 

of against, the native elements of biodiversity within the 

system. As Tausch (1999b:363) stated (emphasis mine):

These treatments should be done based on the conditions existing on the entire associated landscape to 

maintain the diversity of the community, its successional stages, and their interconnectedness, and 

to help avoid the establishment of new, unwanted 

thresholds. The treatments used must incorporate 

the biological, topographic, and edaphic heterogeneity of the sites involved into their application. This 

is to preserve, and to take advantage of, the existing 

diversity— both biotic and abiotic.

For instance, land managers should take advantage of 

self-thinning processes instead of fighting them. The Spruce 

Mountain restoration proposal expressed concern that bark 

beetles and mistletoe are killing pinyon and juniper trees, 

and so the work must go forward (BLM 2005b:6, 22, 23, 

43). If the “No Action” alternative were taken, then “Dwarf 

mistle toe [sic] and bark beetle infestations would continue 

to spread” (BLM 2005b:6). The decision to take “Action” thus 

makes no sense to me. The fact is that these are but two of 

the at least 31 native organisms in our ecosystems that are 
USDA Forest Service Proceedings RMRS-P-52. 2008 21

Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure Charlet

taking action now to thin closing stands of pinyon-juniper. 

There is no reason to fight those elements of biodiversity in 

the systems that are actively creating the solution for the 

perceived problem, instead we should either work with them 

or get out of their way and watch them create the openings 

that we have as our objective.

Additionally, areas chained in Spruce Mountain in 1970 

“are now in need of maintenance” because “young trees are 

threatening to compromise the value of the original projects 

for mule deer” and “the previously chained areas ... most 

beneficial to mule deer will be retreated to restore the original value” (BLM 2005b:5). Treatments include prescribed 

fires, with or without seeding, but there is no experimental 

design for the treatments, so little can be learned from 

this exercise. The times for these fires will be from March 

through mid-May and October through November, times 

that historically and pre-historically were at low risk for 

fire. Naturally, if the federal agency wishes to set fires on 

public land, the fires are easier to manage during these 

months. The agency considers that there may be negative 

impacts for resident vertebrate populations, but does not 

consider the impacts of these out-of-season fires on resident 

invertebrate populations, many of which are crucial to the 

healthy functioning of the ecosystem. How many elements 

of biodiversity in this system will we unintentionally knock 

out with our actions?

I submit that the goals and objectives of our management 

actions should be designed to restore habitat for all the native species in the area. They have maintained a dynamic 

equilibrium with one another for millennia before we arrived with our giant footprint on the land. If management 

is conducted to promote species such as mule deer that in 

the pre-disturbance state were of incidental importance 

(Tanner and others 2003), or for wild horses that we protect 

because of sentiment and law, or for cattle that we protect 

because of tradition and short-term economics, then we 

are neglecting our responsibility of maintaining wild and 

semi-wild, fully functioning ecosystems, with all their native 

parts. For all of our knowledge, we may lack the wisdom to 

intervene on the ecosystem’s behalf. I trust the wisdom of 

the ecosystems themselves rather than our own biased and 

short-term impressions. Management actions that oppose 

native dominants and that require constant vigilance on our 

part in order to sustain the original efforts are fundamentally flawed in their approach. Management actions that 

require constant or oft-repeated interventions by humans 

are doomed to fail, because they run contrary to the natural 

trend on the ground, whatever that may be.

We must incorporate long-term experimental approaches to 

vegetation management activities, with adequate replication 

and controls. We should experimentally assess specific treatments before we apply them at the landscape or regional scale. 

Toward this end, the work of Chambers and others (2006) is 

an excellent example of the experimental approach testing 

certain management actions that perhaps should be taken. 

The results can be assessed and analyzed, and compared to a 

control. Unfortunately, such studies are rarely undertaken, 

and more often we have large-scale management actions 

producing major effects that we cannot adequately evaluate. 

For short-term needs, we can consider low-intensity activities that minimize disturbance, such as thinning techniques 

suggested by Goodwin and Murchie (1980).

We need to establish more collaboration with archaeologists and historians. The historical investigations of Creque 

and others (1999) are excellent. The authors point out the 

method’s shortcomings and provide fine recommendations 

for improvement and future studies. In Nevada we also 

need to conduct studies that reconstruct the recent fire and 

vegetation history, especially for the past 500 years such as 

that conducted by Kitchen and others (2006) in Utah. Such 

research can help us more precisely see the transition from 

the pre-settlement past through the present state of instability. We must collaborate with historians and archaeologists 

to develop the means whereby we can actually estimate 

the extent of the pinyon and juniper woodland formations 

throughout Nevada at settlement time. Only then can we get 

an accurate picture of the trend in pinyon-juniper distribution and abundance in the past 150 years.

Conclusions _______________________

The climate and physiography of Nevada and much 

of Iran are comparable. However, the vegetation of Iran 

is now greatly simplified from its pre-civilization state. 

Complex shrublands and woodlands have been reduced to 

simple shrublands and annual grasslands, dominated by 

weedy, armed, and poisonous species. The type conversion 

sequence during the Holocene in Iran was from forest and 

semi-arid woodlands, to shrublands, and ultimately annual 

grasslands.

In contrast, Great Basin vegetation remains complex 

and the Mojave Desert is exquisitely complex in its multidimensional structure. There are many types of shrub communities, the different species occupying different canopy 

levels, providing multi-canopy shrublands of high diversity 

and high wildlife value. Nevada’s Great Basin and Mojave 

Desert semi-arid woodlands appear to be advancing at this 

time, but it is extremely difficult to determine how much 

is expansion beyond the historic range prior to widespread 

deforestation in the late 1800s. As a result, we tend to 

overestimate the expansion of the woodlands today. This 

could lead to overreaction in land management whereby we 

use invasive techniques that cause much disturbance and 

destabilize the communities even more. It may be that the 

greatest danger of type conversion to our shrublands is not 

from shrublands to woodlands, but rather from shrublands 

to annual grasslands of extremely low biodiversity and ecological value, as is the case in Iran. Nevada’s Great Basin 

and Mojave Desert native vegetation are global treasures, 

and I believe it is our obligation to conserve them.

Acknowledgments__________________

This research was funded in part by the National Geographic Society, grant #7452-03, Collecting Iranian Botanical 

Specimens for Comparative Research Collaborations. The 

generosity of my host institutions (the University of Tehran, 

The University of Tehran at Karaj, The University of 

Esfahan, The University of Urumieh, and Guilan University) 

was essential for the completion of this project. Thanks are 

due Community College Southern Nevada (CCSN) for supporting my travel costs to Cedar City for the symposium. 
22 USDA Forest Service Proceedings RMRS-P-52. 2008

Charlet Shah-Kan-Daw: Anthropogenic Simplification of Semi-arid Vegetation Structure

Cleo Wilson and Marion Martin, librarians at CCSN, were 

very helpful in securing hard to find papers. I am grateful 

to Michael Green for helping answer my questions concerning Nevada history. Nathan Thomas, Archaeologist at the 

BLM Ely District, was generous in providing me with his 

unpublished manuscript and several archaeological papers 

on early Nevada history. Wendell Huffman, at the Nevada 

State Museum, was enormously helpful in providing me 

with data concerning early Nevada railroads. My students 

Angela Pitt and Saeed Reezi assisted me in many ways to 

complete this paper, as did my colleagues Mehdi Zarrei and 

Ali Reza Naqinezhad. I owe special thanks to my wife Dene 

for the final preparation of figure 2, and to Lisa Smith for 

the preparation of figure 4. My reviewers Stan Smith, Rich 

Rust, and Naomi Miller made useful comments that improved 

the paper. I especially am thankful to Naomi Miller and our 

stimulating conversations, as well as her assistance with 

figure 1. My special thanks extend to my friend and mentor 

Barbara Ertter, who enabled me to go around the world.

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The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

The use of trade, firm, or corporation names in this publication is 

for the information and convenience of the reader. Such use does not 

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USDA Forest Service Proceedings RMRS-P-52. 2008 25

Invasive Species and Shrublands

USDA Forest Service Proceedings RMRS-P-52. 2008 27

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Chad R. Reid is an Extension Agent, Utah State University, 

Cedar City, Utah; email: chadr@ext.usu.edu.

Sherel Goodrich is a Forest Ecologist, Ashley National Forest, 

Vernal, Utah.

James E. Bowns is an Extension Specialist, Utah State University, Cedar City, Utah.

Abstract—In recent history, there has not been a more ecologically important event than the introduction of cheatgrass (Bromus 

tectorum) and red brome (Bromus rubens) into the Intermountain 

West. These grasses are very similar in ecology and history and are 

separated mostly by function of elevation. Both species are from the 

Mediterranean region, and both arrived in the Western United States 

about the same time (1880). Cheatgrass and red brome have greatly 

affected fire frequency and intensity, which has been detrimental 

to native shrubs and other perennials in these systems. Red brome 

may have had an even greater impact, in that it has readily invaded 

non-disturbed areas, has had great impact on fire sensitive shrub 

species, and, to this point, we have not identified adapted species 

native or non-native for rehabilitating burned areas. Introduction 

of cheatgrass and red brome in the West has wreaked ecological 

havoc on the areas they have invaded and will continue to affect 

structure, function, and management of these areas well into the 

future. This paper will detail the history and ecology of these two 

highly invasive species.

Introduction_______________________

In recent history, there has not been a more ecologically 

important event than the introduction of cheatgrass (Bromus 

tectorum) and red brome (Bromus rubens) into the Intermountain West. This paper details the history and biology 

of these two highly invasive species.

The reality of these annual grasses is well summed up 

by Peters and Bunting (1994) with the suggestion that the 

introduction of exotic annual grasses, including cheatgrass, 

into the Snake River Plain may have been the most important event in the natural history of that region since the 

last glacial period. Catastrophic ecosystem change for the 

western Great Basin has been suggested as a function of 

cheatgrass by Billings (1994).

Cheatgrass and Red Brome: 

History and Biology of Two Invaders

Chad R. Reid, Sherel Goodrich, James E. Bowns

These grasses are very similar in biology, ecology, and 

history and are separated mostly by function of elevation; 

we will review cheatgrass first and then follow with red 

brome.

Cheatgrass ________________________

Cheatgrass—also called June grass, bronco grass, downy 

chess, and downy brome—is a winter annual that was introduced from the Mediterranean region in packing material 

and first found near Denver, Colorado, (Whitson and others 

1991) or perhaps in the eastern coastal states prior to its 

entry into the West (Monsen 1994).

Colonization of the West by Cheatgrass

Railways, roads, and contaminated grain seed are indicated 

as the principal means of initial spread of cheatgrass after 

which it was advantageous with heavy grazing and other 

disturbance (Billings 1994). One of the earliest reported collections of cheatgrass in the West was made in Washington 

by Sandberg and Lieberg in 1883, and a year later it was 

collected near Provo, Utah, by M. E. Jones (Billings 1994). 

In the course of extensive field surveys of about 1900, it was 

not reported for northern Nevada, and the first report for 

Elko County was in 1906 (Young and others 1987).

By 1946 it occupied at least 10 million acres in eastern 

Oregon (Monsen 1994). By the 1980s, Kunzler and others 

(1981) found it among the most abundant species in Gamble 

oak communities of central and northern Utah, and by the 

1990s it was common to dominant over hundreds of thousands 

of acres in the Great Basin of Utah and Nevada. Roberts 

(1991) reported 900,000 acres (about 30%) of the Salt Lake 

District of the Bureau of Land Management was producing 

cheatgrass. It is present in all counties of Utah, where it is 

an integral part of the flora (Arnow 1987).

Cheatgrass is now widespread in North and South America 

in many plant communities of plains, deserts, foothills, and 

lower montane areas and especially where snow depth and 

temperatures allow for fall germination and some winter 

development or at least persistence of seedlings through 

winter.

It is most abundant in areas of between 6–16 inches of 

annual precipitation (Monsen 1994). It does not do well on 

saline soils, but its ability for rapid growth might allow it to 

make considerable growth from shallow, less saline moisture 

early in the season (Rasmuson and Anderson 2002).

Cheatgrass is an explosive invader in the Great Basin, 

Snake River Plain, Columbia Plateau, and other areas of the 
28 USDA Forest Service Proceedings RMRS-P-52. 2008

Reid, Goodrich, and Bowns Cheatgrass and Red Brome: History and Biology of Two Invaders

West where it proliferates with fire and other disturbance 

including roads and associated traffic, off road vehicle use, 

construction of recreation facilities, and livestock grazing.

It is highly flammable when dry, and with relatively little 

moisture it produces enough biomass to create continuous 

fine fuel that leads to high frequency and increasing size of 

subsequent fire, which perpetuates this plant and excludes 

many others including sagebrush.

Although some native and introduced grasses compete well 

with cheatgrass when the grasses are mature, the seedlings 

of few species can compete with cheatgrass. Francis and 

Pyke (1996) found that cheatgrass seedlings were superior 

competitors compared with seedlings of two cultivars of 

crested wheatgrass. However, they found that increasing 

densities of Hycrest cultivar of crested wheatgrass reduced 

cheatgrass biomass and tiller production.

Near complete control of cheatgrass might be necessary 

before seedlings of some perennial grasses, including desert 

needlegrass (Achnatherum speciosum), can be established 

(Rafferty and Young 2002).

In number of seeds produced per plant per unit of area, 

cheatgrass has the capacity to overwhelm native perennials 

at the seedling level even if the starting density of cheatgrass 

seeds is low (Young and Allen 1997).

Management Implications

With dynamic expansion of cheatgrass with disturbance, it 

is desirable to promptly rehabilitate burned sagebrush and 

pinyon-juniper communities before cheatgrass has a chance 

to establish dominance of the site (Evans and Young 1978). In 

some cases, cheatgrass has been found to make an initial rapid 

increase and then greatly decline (Barney and Frischknecht 

1974; Erdman 1970; Davis and Harper 1990).

A fire return interval of 3–6 years fueled by cheatgrass 

tends to wear down perennials. Regardless of some perennial 

plants being able to compete with cheatgrass at one point in 

time, the ability of cheatgrass to drive ecosystem dynamics 

over time is a function of high fire frequency as well as its 

aggressive growth features.

To beat cheatgrass in communities where sprouting 

perennial species have been depleted often requires prompt 

seeding of adapted perennials that are able to establish uniform stands with a single seeding. Few species are known 

to be able to establish stands with a single seeding in areas 

of less than 10–12 inches annual precipitation. Most of 

these, such as crested wheatgrass, have been introduced 

from Eurasia.

In recent years, great emphasis has been placed on seeding only natives after fire or other disturbance to maintain 

native plant communities. However, unless seedings are 

successful in keeping cheatgrass from dominating, the goal 

for natives is not achieved, and cheatgrass drives a departure 

from native ecosystems that exceeds the departure induced 

by crested wheatgrass.

Until native plant materials with the capability of competing with cheatgrass in low precipitation areas are available in 

large quantities, “pick your alien” (annual or perennial) will 

remain a dilemma for the native only concept. The option to 

not seed exotic perennials has and will likely continue to favor 

cheatgrass. Indeed, the concept of pure native communities 

has become not only problematic, but it is presented with 

catastrophic challenges by cheatgrass.

A growing body of evidence strongly suggests that cheatgrass presents a potential to turn the pure native concept 

into romantic fantasy. That body of evidence includes the 

presence, the abundance, and even dominance of cheatgrass 

in areas where native plant communities have been protected 

from disturbance.

Kindschy (1994) reported the presence and increase of 

cheatgrass in southeastern Oregon’s Jordan Crater Research 

Natural Area that has been protected from human activities 

including livestock grazing.

On Anaho Island in Nevada, Tausch and others (1994) 

found cheatgrass has displaced native perennials despite 

a general absence of human-caused disturbance and fire. 

They attributed the increase to the competitive ability of 

cheatgrass.

In Red Canyon of the Green River, cheatgrass has been 

found as the most frequent species where livestock use and 

other post European related disturbance have been minimal 

(Goodrich and Gale 1999).

Young and Clements (1999) reported invasion of cheatgrass into ecologically high condition shadscale/greasewood 

communities in Nevada despite apparent lack of livestock 

grazing.

Young and Tipton (1990) cited two works from southeastern Washington that documented observations of cheatgrass 

successfully inserting itself into climax perennial grass/shrub 

communities that had been protected from fire and grazing 

for as long as 50 years. They proposed that the idea of cheatgrass spreading in a biological vacuum created by excessive 

grazing may be somewhat misleading or overstated.

Young and Allen (1997) have emphasized that site degradation is not necessary for cheatgrass invasion.

In western Utah, Harper and others (1996) found cheatgrass able to establish in ungrazed areas in desert shrub 

communities where, although native perennials were able 

to greatly suppress the size of cheatgrass plants, cheatgrass 

was able to maintain a presence by which it could expand 

upon disturbance including gopher mounds.

Austin and others (1986) found cheatgrass present in Red 

Butte Canyon of the Wasatch Mountains where livestock 

grazing was discontinued in 1905, which was essentially prior 

to cheatgrass reaching that area. Austin and others (1986) 

also found cheatgrass in Emigration Canyon of the Wasatch 

Mountains where livestock grazing was discontinued in 1957. 

They reported higher cover values for cheatgrass in 1983 

than for 1935 in Red Butte Canyon and higher values in Red 

Butte Canyon than for Emigration Canyon in 1983. These 

values for cheatgrass are contrary to the concept of cheatgrass 

only increasing as a function of livestock grazing.

Knight (1994) reported that the cheatgrass problem is 

not restricted to land grazed by livestock, and he gave an 

example of an increase of cheatgrass following fire in Little 

Bighorn Battlefield National Monument in southern Montana. He suggested that managing vegetation of a National 

Monument so that it reflects presettlement conditions is a 

goal that may be impossible once certain introduced species 

become established.

Although some of these examples deal with areas that have 

been relatively little affected by human activities, nowhere 
USDA Forest Service Proceedings RMRS-P-52. 2008 29

Cheatgrass and Red Brome: History and Biology of Two Invaders Reid, Goodrich, and Bowns

is the climatic zone of cheatgrass in North America wholly 

isolated from the modern world. The rapidity, volume, and 

distance of transport of people and goods across the globe by 

air, water, and ground strongly suggest additional introductions of cheatgrass and other aggressive species from around 

the world will not only continue but will increase.

That livestock grazing has been a factor in the spread and 

abundance of cheatgrass is not disputed here. However, the 

suggestion by Young and Tipton (1990) that this factor has 

been overstated seems appropriate in that other factors of 

spread have perhaps received less attention. Disturbance 

of roadsides and water runoff from roads creates favorable 

habitat for several weedy species. Highway and off-road 

vehicles are highly efficient seed catching and dispersing 

agents. These features would have resulted in the spread of 

cheatgrass even in the absence of livestock grazing.

The reality of modern life and the aggressive nature of 

cheatgrass present challenges for managing wildland resources that will not be well addressed by clinging to concepts 

based on conditions that no longer exist. The world is not 

what it was prior to European settlement of the Americas. 

Air traffic, super highways, railways, and roads of high density were not part of the environment prior to 1492. Reality 

of today includes not only vehicles that travel hundreds of 

miles in a day with the potential to carry seeds not only 

across major drainages, but also across oceans.

A highly mechanized and highly mobile human population 

contributes to a high fire frequency that favors cheatgrass. 

The competitive nature of cheatgrass will not be reduced by 

the concept that native communities that are well managed 

or even untouched will keep it out. In some environments, it 

has demonstrated that it is a better competitor than native 

species. Within the ecological range of cheatgrass, basing 

potential natural plant communities of today on the environment prior to 1492 makes little sense.

Dealing with this force might require seeding some of the 

most aggressive and less fire prone perennials the world has to 

offer, regardless of origin. Although this concept is laced with 

the problem of the cure being worse than the malady where 

native communities are desired, such a desire is laced with 

a serious problem or dilemma of its own. The replacement 

of native ecosystems and their function by cheatgrass driven 

systems indicates a departure from “native” that exceeds 

that associated with seeding selected perennials.

Each fire on the Snake River Plain, valleys and foothills 

of the Great Basin, and other cheatgrass prone areas of the 

West tightens the grip cheatgrass has on these ecosystems. 

Opposition to seeding highly competitive perennials in 

cheatgrass prone areas is indicated to be a demonstration 

of values that are no longer a potential. Billings (1994) indicated that the potential has changed with the catastrophic 

ecosystem change induced by cheatgrass as indicated by 

Billings (1994).

The retort that cheatgrass is a function of past mismanagement of livestock on rangelands will do nothing 

to improve the condition, and it conveniently ignores the 

high likelihood that cheatgrass spread and dominance 

was inevitable with European settlement with or without 

livestock. Although the early advance of cheatgrass in 

the West was facilitated by livestock grazing, the ultimate spread of this species is a function of a number of 

factors, some of which could have advanced cheatgrass 

in the absence of livestock grazing. The dense network of 

roads, off-road vehicle use, and frequent use of these by 

an expanding population would have been the means of 

spread throughout the potential range of cheatgrass. Fire 

would have done the rest. It is quite likely that livestock 

simply set forward or accelerated the inevitable by a few 

decades.

The long-term trend in numerous crested seedings indicates that these seedings could be managed for the return 

and maintenance of at least native sagebrush (Huber and 

Goodrich 1999). In contrast, cheatgrass and its shortened fire 

cycles excludes sagebrush. Although crested wheatgrass is 

introduced, it presents an opportunity for greater diversity 

than does cheatgrass. Where stands of crested wheatgrass 

are managed for return of sagebrush, it can facilitate development of much greater structural diversity than does 

cheatgrass.

The use of natives at the present appears problematic; 

Britton and others (1999) evaluated performance of 24 taxa 

at a sagebrush site and 20 taxa at a greasewood site. The 

top performing 9 taxa at the sagebrush site and 10 taxa at 

the greasewood site were introduced.

Hull (1974) evaluated the performance of 90 plant taxa 

including many natives in rangelands of southern Idaho. 

Where annual precipitation was less than 25 cm (10 in), only 

17 of the 90 taxa rated over 1 on a relative scale of 1–10. 

Of these 17 only 6 were natives and none of these natives 

rated over 2.1. Phases of crested wheatgrass rated from 7.7 to 

9.5. Pubescent wheatgrass rated at 6.2 and intermediate 

wheatgrass rated at 5.1. No other taxon rated over 5.

In general, it seems that the expansion and dominance 

of cheatgrass has been more dramatic in the inherently 

grass-poor regions of the sagebrush ecosystem than in the 

inherently grass-rich regions. As indicated by Tausch and 

others (1994), the boundary between Wyoming big sagebrush 

and mountain big sagebrush in western Nevada represents 

a boundary below which moisture and other conditions 

favor annual grasses and above which perennial grasses 

are favored.

However, the Wyoming big sagebrush region of Wyoming 

compared to that of the Great Basin seems to be relatively 

rich in native grasses. In this grass-rich region, the invasion 

of cheatgrass has been comparatively mild. This contrast 

demonstrates that features other than livestock grazing are 

important in abundance of cheatgrass. Intensity of livestock 

grazing on sagebrush areas of Wyoming has probably been 

equal to, if not greater than, that in the Great Basin.

Forage Value

Cheatgrass is nutritious when young and palatable to a 

wide range of ungulates and is highly preferred by mule deer 

during spring and fall (Austin and others 1994). Bighorn 

sheep have been observed using cheatgrass on the steep, 

southerly facing slopes of Red Canyon of the Uinta Mountains 

in winter when this winter annual is one of the few green 

herbaceous plants. It is of great economic importance to the 

domestic livestock industry in some places. Emmerich and 

others (1993) and DeFlon (1986) reported on range operations where cheatgrass is a major part of the winter and 
30 USDA Forest Service Proceedings RMRS-P-52. 2008

Reid, Goodrich, and Bowns Cheatgrass and Red Brome: History and Biology of Two Invaders

spring forage base. Kufeld (1973) rated it as valuable to elk 

in winter. This value is likely—at least in part—a function 

of the southerly exposures or other warm places that are 

often open in winter where this plant grows best.

On the Mojave Desert, Phillips and others (1996) found 

that domestic sheep showed high preference for this plant 

in spring. Cheatgrass seeds and new growth are valuable 

forage for chuckars (Alectoris graeca), Gamble quail (Lophortyx gambelii), and mourning doves (Zenaidura macroura) 

(Plummer and others 1968).

Disadvantages of cheatgrass include (1) high fire frequency, 

which greatly alters forage supplies and maintains rangelands in an annual condition frequently exposed to wind 

and water erosion; (2) a short green-feed period; (3) great 

variability of herbage production between moist and dry 

years; and (4) mechanical damage to mouth parts of grazing animals after drying and hardening of sharp seed-parts 

(Young and Allen 1997). Also cheatgrass has been associated 

with reduced nutritive quality of other species (Haferkamp 

and others 2001). DeFlon (1986) reported winter use of 

cheatgrass ranges was more favorable to achieving a stable 

forage base than was spring use. He explained that spring 

use lead to an increase in halogeton in his study area.

It appears that the consequences of cheatgrass remain 

little understood. Preservation of native plant communities 

based on hands-off management, seeding only locally collected 

native seed to avoid genetic contamination, and rangeland 

evaluation criteria that ignore cheatgrass in site potential 

are ideals difficult to implement where cheatgrass is well 

adapted. However, these and other preservation based ideals 

often prevail in planning, management, and legal maneuvering dealing with cheatgrass prone rangelands. Yet there 

seems to be little in the literature dealing with cheatgrass 

to support ideals of preservation on lands where cheatgrass 

is highly competitive.

The departure from native ecosystems inflicted by cheatgrass exceeds that of seeding selected, exotic perennial plants 

and other cultural practices that foster perennial plants that 

tend to reduce influence of cheatgrass. Ironically, advocates 

of preservation often support actions that favor cheatgrass 

over establishment of adapted perennial plants.

Red Brome ________________________

As stated earlier, red brome is very similar to cheatgrass, thus most, if not all of the proceeding discussion on 

management and ecological implication also applies to red 

brome. The following is specific to the history and biology 

of red brome.

Introduction and History

Red brome, often called an ecological equivalent of cheatgrass, is another Mediterranean winter annual that has invaded disturbed and undisturbed areas of western Northern 

America, especially the desert southwest. Red brome was 

brought to North America from the Mediterranean before 

1880 (Watson 1880). Three possible scenarios have been 

proposed for introduction into this area: (1) California Gold 

Rush and Central Valley Wheat, (2) Southern California 

Shipping, and (3) Northern California Sheep.

The period of most rapid spread was from 1930 to 1942. 

The greatest spread into new regions during the past 50years 

coincides with “warm” Pacific Decadal Oscillations regimes 

(El Nino) (Salo 2005). El Nino southern oscillations that 

result in consecutive years of above-average winter precipitation provides red brome 1st year germination and 2nd year 

high biomass. Increased CO2

 and N deposition also may be 

contributing to red brome’s success.

The early history of California includes the Mexican Period 

(1822–1846). It appears unlikely that red brome was introduced from Mexico because red brome was first reported in 

Mexico in 1931–1932 (Howell 1942), 50 years after its first 

collection in the United States.

In contrast to accidental introductions, red brome was 

seeded near the University of Arizona at Tucson from 1906 

to 1908 for evaluation as a forage plant; this grass soon escaped and became established along the Santa Cruz River 

(Thornber 1909)

Red brome was also reported in northern Arizona in 1911 

and collected near St. George, Utah, in 1926, where it increased in collections from this area for the next 35 years. 

In addition, it was also becoming common in waste places 

and cultivated areas around Las Vegas, Nevada, during this 

time (Maguire 1935).

Red brome was found throughout northeastern Nevada 

by the 1940s and continued to spread in designated natural 

areas of Arizona during this time. By the 1960s, red brome 

dominated even relatively undisturbed areas of Nevada, 

Utah, and Arizona and was also reported in New Mexico. 

Collections of red brome in south-central Utah increased 

dramatically after the construction of Glen Canyon Dam 

in 1963.

Red brome now occurs from British Columbia to northwestern Mexico and coastal California to western Texas; continued 

introductions may have provided new genotypes.

Biology

The available literature suggests that red brome does not 

maintain a soil seed bank but exhibits early and uniform 

germination. In contrast, native annuals depend on soil seed 

banks, hedging in time; however, at least one researcher 

disputes this (J.A.Young, personal communication). Red 

brome has nearly uniform germination under cool, moist 

conditions typical of this region and can germinate with 

0.5 inch of precipitation, Mojave native annuals appear 

to require twice that amount (Beatley 1966). One author 

contends this characteristic can lead to population crashes 

during drought, and winter droughts dramatically reducing 

red brome densities (Salo 2004). However, Beatley (1974) 

states that although its numbers vary from season to season 

where established, red brome has never been observed to 

miss a growing season.

Fire Ecology

Red brome may be even more problematic than cheatgrass 

from a fire standpoint. Low humidity in its range leads to 

slower decomposition than cheatgrass, increasing fuel 

loading for a longer period of time. Red brome has been 

particularly troubling in this aspect because most of the 
USDA Forest Service Proceedings RMRS-P-52. 2008 31

Cheatgrass and Red Brome: History and Biology of Two Invaders Reid, Goodrich, and Bowns

systems it occurs in are not fire adapted, nor have we identified suitable native or non-native species for reseeding into 

these areas. These characteristics along with its ability to 

occupy non-disturbed blackbrush (Coleogyne ramosissima) 

sites, the shrub type most susceptible to fire in the region 

(Beatley 1966), is of great concern.

Conclusion________________________

The introduction of cheatgrass and red brome in the West 

has degraded invaded ecosystems and will continue to affect 

structure, function, and management of these areas well 

into the future. New land management paradigms will be 

required to manage these ecosystems.

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USDA Forest Service Proceedings RMRS-P-52. 2008 33

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Eric B. Peterson is a plant ecologist, Nevada Natural Heritage 

Program, 901 South Stewart Street, #5002, Carson City, NV 89701; 

email: peterson@heritage.nv.gov.

Abstract—Several exotic annual grass species are invading the 

Intermountain West. After disturbances including wildfire, these 

grasses can form dense stands with fine fuels that then shorten fire 

intervals. Thus invasive annual grasses and wildfire form a positive feedback mechanism that threatens native ecosystems. Chief 

among these within Nevada are Bromus tectorum (cheatgrass), 

Bromus rubens (red brome), and Schismus barbatus (Mediterranean 

grass). These grasses have an early phenology for both green-up and 

senescence that can be detected from the difference in greenness 

between two well-timed satellite images, allowing grass cover to be 

geographically modeled. The Nevada Natural Heritage Program is 

using imagery from Landsat and MODIS satellites to map annual 

grass invasion, and has completed a map of annual grass cover for 

the State of Nevada. The models and final map were developed from 

806 training sites, remotely sensed data for two seasons from Landsat 

5 TM and MODIS satellite sensors, and accessory geographic data. 

Accuracy of the final map was analyzed from two independent data 

sets: Southwest Regional Gap Analysis Project (REGAP) training 

data (15,318 plots) and California Department of Fish and Game 

vegetation plots from the Mojave region (939 plots). Root-MeanSquare-Error (RMSE) from the REGAP data was 10.33 percent; 

75 percent of predictions for gap plots were off by 5 percent or less; 

and 95 percent of predictions were off by 21 percent or less. For the 

Mojave data set, RMSE was 7.48 percent; 75 percent of predictions 

were off by 9 percent or less; and 95 percent of predictions are off by 

15 percent or less. Accuracy assessment on REGAP data suggests 

that annual grass cover is generally underestimated for sites with 

high cover, thus the map should be interpreted as an index of cover 

rather than an estimate of actual cover. Nevertheless, the map 

reveals the pattern of annual grass invasion across Nevada.

Introduction_______________________

Invasive exotic annual grasses are an ecological catastrophe that we must contend with. What has brought on this 

problem is uncertain, and what the solutions will be are 

even less known. Herbicides have been successfully used to 

treat limited areas. However, treating tens to hundreds of 

Invasion of the Exotic Grasses: 

Mapping Their Progression Via Satellite

Eric B. Peterson

millions of infested acres throughout the West with chemicals is unrealistic. Herbicides might be useful in maintaining fire breaks in the ‘cheatgrass sea,’ or southward in the 

‘Schismus sea,’ but are not a solution to the invasion. That 

does not mean that we should give up! California’s Central 

Valley was almost completely converted to invasive annual 

grasses nearly a century ago, yet people maintain hope for 

small pieces of remnant native grasslands. The Intermountain West still retains vast areas of native communities. 

Furthermore, the region has vast areas maintained by public 

agencies, providing potential for landscape-level strategies 

against complete transformation of our ecosystems. There 

is room for optimism, if regional strategies can be sought, 

found, and employed.

Regional strategies for dealing with invasive annual 

grasses will require a geographic understanding of the 

problem—knowing the pattern of invasion across the landscape. For this reason, the Nevada Natural Heritage Program (NNHP), which monitors biodiversity in the state, has 

taken an interest in invasive species. The goal of the work 

presented here is to map the current status of invasive exotic 

annual grasses in Nevada, through statistical modeling of 

annual grass cover as detected by satellite sensors. This is 

one step toward a more comprehensive understanding and 

monitoring of native ecosystems, their biodiversity, and 

their biogeography.

Methods and Materials______________

Annual grasses tend to have short life-spans and senesce 

earlier in the season than perennials. Satellite sensor 

data (imagery) can detect chlorophyll concentration over 

the landscape, primarily from its reflectance in the nearinfrared (Jensen 1996). Locations with large concentrations 

of annual grasses show a marked drop-off in chlorophyll 

as annual grasses senesce. Thus the change in chlorophyll 

concentrations as measured by satellite can be correlated 

with the actual ground cover of annual grasses at training 

sites and can be used to create predictive models of annual 

grass cover (Peterson 2005).

The methods used here are very similar to those previously used to map Bromus tectorum (Peterson 2003, 2005) 

and further details than are given below were reported 

online (Peterson 2006). Briefly: training data were collected 

from the field; satellite sensor data were obtained for appropriate times during the annual grass growing season 

and senescence season; statistical models were developed 

and expressed through geographic mapping algorithms 

for visual evaluation; and independent ground data were 
34 USDA Forest Service Proceedings RMRS-P-52. 2008

Peterson Invasion of the Exotic Grasses: Mapping Their Progression Via Satellite

obtained for post-modeling accuracy assessment. During 

data processing and analysis, all raster data mosaics and 

reprojections to match the Landsat data set (below) were 

calculated in ENVI 4.2 (Research Systems Inc. 2005) with 

nearest neighbor resampling and a 100 X 100 matrix of triangulation points for reprojections. All statistical analyses 

were performed with the R statistical package (R Development Core Team 2005).

Training Data

Ocular estimates of annual grass cover were made at 

training sites over 0.1 ha plots (see Peterson 2005 and 2006 

for details).

Satellite Data

Satellite sensor data (imagery) for this project needed to 

provide measures of green vegetation during the growing 

and the senescence periods for annual grasses. These periods 

can be short, particularly for the senescence period, as the 

sensor data must be collected when annual grasses have 

mostly senesced yet perennial vegetation and most forbs 

remain photosynthetically active. Target dates for growing 

season were the month of March in the south (roughly the 

Mojave ecoregion), and mid-April to early May in the north. 

Dates for senescence were mid-April to early May in the 

south, and mid- to late-June in the North.

Annual grasses have substantial inter-annual variation 

in production, particularly between wet and dry years as 

measured by total precipitation (Bradley and Mustard 

2005). Fieldwork on this project in the years 2002–2005 

suggested that phenology of annual grasses is more geographically variable in wet years than in dry years, so a dry 

year with relatively low cloud cover (2004) was targeted for 

the northern portion of Nevada. However, during dry years 

in the southern portion of the state, the low annual grass 

production might be difficult to detect reliably. Thus, a wet 

year (2005) was targeted for the southern portion of Nevada 

despite substantial obscuring of the land surface from satellite sensors by clouds.

Landsat 5 TM sensor data were purchased for dates within, 

or close to, the ideal target dates, from the USGS. EROS data 

center using UTM zone 11, NAD 83 projection with terraincorrection and a spatial resolution of 28.5 meters. Landsat 

satellites capture a given path on 16 day intervals, forcing 

the use of several clouded scenes. To fill in the holes left by 

clouds (referred to hereafter as ‘cloud-holes’), data composites over 16-day periods were obtained from the MODIS 

satellite sensor (MOD 13 product; NASA 2005).

The satellite data were analyzed in four sets: (1) the 2004 

data in the north where Landsat data exists for both seasons 

(LL04), (2) the 2004 MODIS data to fill in cloud-holes with the 

prior (MM04), (3) the 2005 data in the south where Landsat 

data exists for both seasons (LL05), and (4) the 2005 MODIS 

data to fill in cloud-holes (MM05). Raw spectral band values 

were analyzed, along with derived values: NDVI for each 

season (Normalized Difference Vegetation Index; Jensen 

1996), ΔNDVI, a Greenness Ratio (for each season), and 

ΔGreenness. ΔNDVI was calculated as early season NDVI 

minus late season NDVI and forms the main measure of 

phenology. The Greenness Ratio was constructed to represent 

a greenness of annual grass infestations visible in false-color 

maps using Landsat bands 7, 4, 2 for red, green, and blue, 

respectively. The Greenness Ratio was calculated as:

GreennessRation  B4

( ) B B 2 7  * . 0 5

where Bx refers to the spectral band number within the 

Landsat 5 TM data (B4 is the same near-infrared band as 

used in calculating NDVI). The ΔGreenness was calculated 

in similar fashion to ΔNDVI using the Greenness Ratios.

Accessory Data

Statistical modeling of vegetation features from satellite 

imagery is generally enhanced by the use of accessory data 

that relate to climate and land features. Data gathered and 

tested for use in models included simple geographic coordinates, elevation models and derived land features, estimated 

precipitation pattern, and ecoregional variation.

Elevation and derived land features—A digital elevation model (DEM) covering the entire extent of all Landsat 

data scenes with 1 arc-second resolution (ca. 30 m) was extracted from the National Elevation Dataset (USGS. 2005) 

and reprojected to match the Landsat data. From elevation 

data, we calculated slope, aspect, heat index, exposure, and 

aridity indices as described in Peterson (2006).

Precipitation—Total annual precipitation corresponds 

strongly to elevation within the State of Nevada, so total 

precipitation itself was not analyzed. However, substantial 

variation exists in the timing of precipitation across the 

state. Two datasets were constructed to represent normalized seasonality of precipitation, as derived from the PRISM 

precipitation models (Spatial Climate Analysis Service, 

Oregon State University 2003). The first contrasted only 

two months, January versus August (JA), while the second 

contrasted the total of January to March versus July to 

September (JFMJAS) (see Peterson 2006 for details).

Ecoregions—Maps produced by early models indicated a 

strong geographic trend in error; most noticeably predicted 

annual grass values were low on the eastern side of the 

state. To allow for ecoregional variation, an ecoregion map 

was constructed (fig. 1). The regions are largely based on 

agglomerations of sub-regions defined by Bryce and others 

(2003), but with substantial editing based on field experience and with the intension of mapping boundaries in the 

behavior of annual grasses. The map construction did not 

directly utilize a geographic analysis of error from annual 

grass modeling. Ecoregions were then entered into models 

as binary factors.

Modeling Process

The statistical modeling procedure followed Peterson 

(2005). In short, a form of survival analysis, specifically 

Tobit regression (Tobin 1958; Austin and others 2000), was 

used to construct equations that predict ground cover from 

satellite and accessory data. Separate models were developed 

for each of the four satellite sensor data sets (LL04, LL05, 

MM04, and MM05). Once models were chosen, they were 
USDA Forest Service Proceedings RMRS-P-52. 2008 35

Invasion of the Exotic Grasses: Mapping Their Progression Via Satellite Peterson

used geographically to calculate rough maps of estimated 

annual grass ground-cover. Maps from LL04 and LL05 were 

then filtered with spectral algorithms described in Peterson 

(2006) to reduce error in playas, forests, wetlands and lakes 

where false positives may result from land reflection characteristics. The maps from separate models were assembled 

into a final map by first mosaicking models by satellite using 

cut-lines based on the ecoregion data, then by overlaying 

MODIS-based mosaic with Landsat-based mosaic. Lastly, 

a conservative smoothing kernel was applied.

Map Validation

Independent data sets were utilized for map validation. 

These were the data collected for the Southwest Regional Gap 

Analysis Project (REGAP) (Lowry and others 2005) and the 

Mojave Desert Ecosystem Program (MOJAVE) (Thomas and 

others 2002, 2004). Although the MOJAVE data lie entirely 

outside of Nevada, sufficient numbers of plots are covered 

by our satellite data and are in appropriate ecoregions for 

assessing map quality. Actual annual grass cover for each 

plot was summed from the field data. Estimated annual 

grass cover was extracted for each plot from the final map 

produced herein. Measures of accuracy were calculated 

based on these actual and estimated values, using Pearson 

Correlation, Root-Mean-Squared-Error, and 75th and 95th

percentile errors.

Results and Discussion______________

Field Work

Training data were collected by the Nevada Natural Heritage Program (NNHP) at 806 plots from 2002–2005 (fig. 1). 

Average ground cover of annual grasses was 9.3 percent, but 

was skewed (fig. 2) and 64.7 percent of plots had no annual 

grasses. Predominant annual grasses were Bromus tectorum

to the north, and Bromus rubens and Schismus barbatus

to the south. Bromus arvensis, Poa bulbosa, Taeniatherum 

caput-medusae, Vulpia microstachys, and Vulpia octoflora 

were observed infrequently. Of the annual grass species 

encountered, only the Vulpia species are native to Nevada, 

but their abundance is minimal. Thus the map produced 

herein effectively targets exotic annual grasses.

Field sampling in 2002–2004 followed relatively dry winters, while sampling in 2005 followed a relatively wet winter. 

Of the 806 plots, 30 plots in northern Nevada were repeats, 

visited first in 2002 or 2003 then again in 2005, to assess 

the affect of total seasonal precipitation on annual grass 

cover. In 2005 mean cover of annual grasses among repeat 

plots was 8.5 percent lower while mean cover of the native 

Poa secunda was just 0.9 percent lower and biological soil 

crusts was 1.1 percent lower. Although cover estimates are 

vulnerable to observer bias, which may change over time, 

the lack of substantial reduction in the later two measures 

(both perennial vegetation components) suggests that the 

reduction in annual grass cover in the 2005 season is real. My 

personal observation was that although the annual grasses 

were taller, their cover had not increased. My subsequent 

Figure 1—State of Nevada, showing county boundaries (black 

lines), training data plots (black dots), and ecoregions for 

modeling accessory data (variously gray backgrounds).

Total Ground Cover of Annual Grasses

Frequency

0 20 40 60 80 100

500

400

300

200

100

0

Figure 2—Histogram of percent ground-cover for annual 

grasses in training data plots. Each bar covers a 5 percent 

range.
36 USDA Forest Service Proceedings RMRS-P-52. 2008

Peterson Invasion of the Exotic Grasses: Mapping Their Progression Via Satellite

personal observations during the 2006 field season, also 

following a wet winter, were different in that annual grass 

cover did appear to increase. One might hypothesize that 

after a prolonged drought, the first year of heavy rains does 

not increase cover, but instead increases seed production. 

A second year of heavy rains then allows the seed production to manifest into greater cover of annual grasses. These 

observations are tentative and limited to northern Nevada. 

Inter-annual variation in annual grass production has been 

used as an alternate method for mapping (Bradley and 

Mustard 2005), but patterns may be more complex than 

typically presumed. Careful study of these patterns would 

be welcomed.

Statistical Modeling

Among the four modeling data sets, hundreds of models 

were tested with more than fifty expressed geographically 

for evaluation. The final models are provided in table 1. In 

general, models included early season NDVI, ΔNDVI, an 

exponential of ΔNDVI, elevation, and ecoregions. The values 

for ΔNDVI were low, including negative values, and covered 

only a short range, so to calculate an exponential of ΔNDVI, 

the variable was rescaled by adding two and then a power of 

four was used. Ecoregions were included as binary variables 

and were frequently retained even when not statistically 

significant because regional patterns in error were more 

visually obvious without them. Models for MODIS data 

were difficult to develop independently of the Landsat based 

models, possibly due to the low spatial resolution of MODIS 

data resulting in noisy data. However, when the variables 

from the Landsat-based models were applied to the MODIS 

data with newly calculated coefficients, the resultant models 

performed well despite the loss of statistical significance 

for some coefficients. Greenness, aridity, and precipitation 

indices were found to be statistically significant for some 

models, but were dropped for the final models as elevation 

and multiple NDVI measures accounted for some of the 

Table 1—Final models for each dataset. Ecoregion was included as a set of binary 

variables.

Data Set Variable Coefficient p-value

LL04 (n = 689) Intercept –612.9975 0.0203

NDVIearly 362.2552 0.000000293

ΔNDVI + 2	 414.7017	 0.0177

(ΔNDVI + 2)4	 –11.3159	 0.0362

ELEV –0.0222 0.000000187

Region2 –0.3103 0.00143

Region3 1.5568 0.483

Region4&5 –4.4817 0.0549

Region6 –10.5149 0.0107

NDVIearly*ELEV –0.1603 0.0000831

 

LL05 (n = 130) Intercept 58.4205 0.00000000000787

NDVIearly 79.7398 0.000230

ΔNDVI	 –69.0320	 0.00165

ELEV –0.0163 0.000115

LateSeasonDate –0.2737 0.00156

DaysApart 0.0532 0.390

Region3&7 12.8757 0.00397

Region6 4.4046 0.335

 

MM04 (n = 688) Intercept –524.1190 0.107

NDVIearly 545.6410 0.000000000782

ΔNDVI + 2	 372.9796 	 0.0853

(ΔNDVI + 2)4	 –12.0503 	 0.0764

ELEV –0.0193 0.000160

Region3 3.3588 0.137

Region4&5 –5.3061 0.0209

NDVIearly*ELEV –0.2488 0.00000261

 

MM05 (n = 118) Intercept 22.1276 0.182

NDVIearly 76.2708 0.138

ΔNDVI	 1.4205	 0.988

ELEV –0.0230 0.0000000834

Region6 5.0345 0.409
USDA Forest Service Proceedings RMRS-P-52. 2008 37

Invasion of the Exotic Grasses: Mapping Their Progression Via Satellite Peterson

same variation in the data. The individual model maps were 

assembled and smoothed to form the final Annual Grass 

Index map (ANGRIN) (fig. 3). Data for this map are available online at http://heritage.nv.gov/reports.

Accuracy Assessment

Use of REGAP data for assessment of the ANGRIN map 

involved 15,318 plots (fig. 4). Actual cover according to 

REGAP data compared to ANGRIN modeled values show 

RMSE =10.33 percent; 75 percent of modeled values differ by 5 percent or less ground cover; and 95 percent of 

modeled values differ by 21 percent or less. Of the total 

REGAP dataset, 8,909 plots are identical in value between 

ANGRIN and measured ground cover, of those 8,848 

are zero in both datasets. However, correlation between 

ANGRIN and REGAP data showed weak correspondence 

(R = 0.24). The discrepancy between the estimated values 

in ANGRIN and the measured values suggests frequent 

underestimation at high cover sites in the ANGRIN map. 

For this reason, as well as inter-annual variation in annual grass cover (Bradley and Mustard 2005), I suggest 

that the map be interpreted as an index of cover rather 

than an estimate of actual cover.

Although REGAP plots were distributed across most of 

the state, the number of plots from the northern Great 

Basin region was much greater than in the Mojave region 

(fig. 4) so assessment by REGAP data applies mainly to 

the Great Basin. Despite underestimation of high cover 

sites in the Great Basin, detection accuracy of no infestation or of low cover infestation was strong and modeling 

performed quite well for annual grasses specifically. This 

is an improvement over previous mapping (Peterson 2005) 

with less area inappropriately estimated to have annual 

grasses. For example, considerable cover of Bromus tectorum had been estimated for the shores of Walker Lake 

in the previous map whereas appropriately little area of 

annual grasses are mapped there in the present effort.

Figure 4—Distribution of plots from REGAP data set showing 

discrepancy between REGAP plots and ANGRIN estimates. 

Overlapping points are displayed with greater discrepancies 

on top.

Figure 3—Annual grass index (ANGRIN), as expressed 

geographically over the satellite sensor dataset. Dark green 

represents areas with no detectable annual grasses; 1-100 

percent ground cover of annual grasses is represented with 

the color gradient from light green through yellow and red to 

purple. County boundaries (white lines) and roads (black lines) 

are provided for reference.
38 USDA Forest Service Proceedings RMRS-P-52. 2008

Peterson Invasion of the Exotic Grasses: Mapping Their Progression Via Satellite

The California Department of Fish and Game collected 

939 plots in the MOJAVE dataset that were useful to this 

project. Comparing the measured annual grass cover at those 

plots to the ANGRIN estimates shows RMSE = 7.48 percent; 

75 percent of predictions are off by 9 percent or less; and 

95 percent of predictions are off by 15 percent or less. The 

discrepancy between measured and estimated values shows 

that in most cases, the estimated value is higher than the 

measured value. Two factors likely contributed to this appearance of over-estimation in the Mojave region. First, the 

MOJAVE dataset was collected in drier years than the 2005 

growing season, thus it could be expected that 2005 satellite 

data would show greater amounts of annual grasses. Second, 

the phenology of this warmer desert region is much more 

temporally compressed than in the Great Basin, leading to 

greater contamination of the satellite measured phenology 

signal by other annual and deciduous perennial species. Although many of the species contaminating the signal may be 

other invasives, particularly mustards (Brassicaceae), many 

native annual wildflowers are probably included.

Sources of Error

There are countless sources for error in any remote 

sensing project; any imaginable natural or human-caused 

variation in the spectral reflectance or accessory modeling 

can reduce the quality of a map. Lengthy discussions of error with these mapping methods are provided in Peterson 

(2003, 2006), including observer bias, inter-annual variation in growth and senescence, signal contamination from 

other early-season species, timing of satellite data, data 

contamination by clouds or other atmospheric phenomena, 

and regional variations.

Another source of error can be the modeling method. Tobit 

regression has a strong advantage over many statistical 

modeling methods in its handling of censored data (limited 

to zero and positive values). However, it is a linear modeling method, while responses of ecological variables such 

as annual grass cover are often non-linear with respect to 

predictor variables. This source of error was discussed by 

Peterson (2003) so I will not dwell on it here, but alternative methods do need to be explored. Generalized additive 

models (Yee and others 1991) and multivariate windowing 

techniques (Peterson 2000), particularly Nonparametric 

Multiplicative Regression (McCune 2006), have potential 

to provide more realistic models of the data. However, software choices for those techniques are limited in the size of 

geographic data files that they can manipulate within a 

reasonable amount of time. Considerable effort may need to 

be devoted to reprogramming their data handling capabilities, or perhaps the continual increase in computing power 

may overcome this issue.

Summary of Map

Various authors have attempted to provide statistics on the 

degree of annual grass invasion in the Intermountain West. 

West (1999) estimated that half of the sagebrush steppe had 

been invaded to some degree and that a quarter had been 

converted to annual grass dominated vegetation systems. 

The USDI Bureau of Land Management (2002) estimated 

that as of 2000 in the Great Basin, 14 million acres of public 

lands were at risk of conversion and three million acres had 

already been converted. The politically defined boundaries 

of Nevada used herein do not match the regions covered by 

West and Wisdom, so direct comparisons of my models to 

their estimations cannot be made. However, by clipping the 

ANGRIN map to the border of Nevada and masking out open 

water, urban, and cultivated areas using data landcover data 

from REGAP (Lowry and others 2005), my models suggest 

that 41 percent of the state has been invaded to a detectable 

degree (ANGRIN value of 1 or greater). Although ANGRIN 

does not distinguish converted vegetation from heavily invaded vegetation, a value of 10 or greater might reasonably 

represent heavy invasion, a value met across 11 percent of 

the state of Nevada.

High degrees of invasion appear to follow the transportation 

corridor of Highway 80 and the historic Central Pacific, and 

northward from that along Highways 95 and 140. High degrees of invasion are also prevalent in many valleys of southern Nevada, though interference from other short-phenology 

plants inhibits any strong statements about invasion in the 

Mojave. Perhaps of most interest are the areas that are not 

strongly invaded. Most obvious is the large swath in the 

middle of the state running from southwest to northeast. 

Much of this area is of relatively high elevation with some 

basins at over 2,000 m (6,500 ft). However, invasion is also 

low in the Tonopah area where elevation and total annual 

precipitation is comparable to the Lovelock area, which is 

highly invaded. Another notable area is the western Owyhee 

Desert, along the northern edge of the state. This is a region 

of sagebrush at 1,500–1,700 m (4,900–5,600 ft) elevation and 

thus within the vulnerable elevation range, though perhaps 

the poor soils in the area may help impede invasion.

Management Implications

Peterson (2005) demonstrated that mapping of annual 

grasses over large areas is possible. This project further 

demonstrates that annual grasses can be mapped on the 

scale of states and to some extent across ecoregional boundaries. Any mapping projects with large geographic scope will 

contain a degree of error and land management proposals 

should be verified on-site. However, maps of large geographic 

scope are useful in evaluating and prioritizing management 

actions. The annual grass index (ANGRIN) map reveals the 

pattern of invasion and infestation by these grasses across 

the landscape of Nevada. It thus provides an assessment 

of a major component of landscape condition, or ecological integrity. The map indicates areas that may warrant 

restoration or rehabilitation. Areas that do not show much 

infestation may be important for protective actions. Since, 

annual grasses form fine, often continuous fuels, the map is 

also useful for wildfire planning and management.

Acknowledgments__________________

I thank the Bureau of Land Management for funding the 

project; Dianne Osborn and Bill Draught were of particular 

help with the project proposal and administration. I greatly 

thank the Carson Valley Chucker Club, the Nevada Department of Wildlife, and the California Department of Fish and 
USDA Forest Service Proceedings RMRS-P-52. 2008 39

Invasion of the Exotic Grasses: Mapping Their Progression Via Satellite Peterson

Game for help with the non-federal match; Dave Pulliam 

and Todd Keeler-Wolf were instrumental in that effort. 

I’d also like to thank the entire Nevada Natural Heritage 

Program for assistance both in the field and in the office. 

And thanks go to Peter Weisberg for external review of the 

manuscript.

References ________________________

Austin P.C.; Escobar, M.; Kopec, J.A. 2000. The use of the Tobit 

model for analyzing measures of health status. Quality of Life 

Research. 9(8): 901-910.

Bradley, B.A.; Mustard, J.F. 2005. Identifying land cover variability 

distinct from land cover change: cheatgrass in the Great Basin. 

Remote Sensing of Environment. 95(2): 204-213.

Bryce, S.A.; Woods, A.J.; Morefield, J.D.; Omernik, J.M.; McKay, 

T.R.; Brackley, G.K.; Hall, R.K.; Higgins, D.K.; McMorran, D.C.; 

Vargas, K.E.; Peterson, E.B.; Zamudio, D.C.; Comstock. J.A. 2003. 

Ecoregions of Nevada [color poster with map, descriptive text, 

summary tables, and photographs]. Reston, VA: U.S. Geological 

Survey.

Jensen, J.R. 1996. Introductory Digital Image Processing: a Remote 

Sensing Perspective. Second Edition. New Jersey: Prentice Hall.

Lowry, J.H, Jr.; Ramsey, R.D.; Boykin, K.; Bradford, D.; Comer, 

P.; Falzarano, S.; Kepner, W.; Kirby, J.; Langs, L.; Prior-Magee, 

J.; Manis, G.; O’Brien, L.; Sajwaj, T.; Thomas, K.A.; Rieth, W.; 

Schrader, S.; Schrupp, D.; Schulz, K.; Thompson, B.; Velasquez, C.; 

Wallace, C.; Waller E.; and Wolk B. 2005. Southwest Regional Gap 

Analysis Project: Final Report on Land Cover Mapping Methods. 

Logan, UT: RS/GIS Laboratory, Utah State University.

McCune, B. 2006. Nonparametric multiplicative regression for 

habitat modeling. Online at http://www.pcord.com/NPMRintro.

pdf. Accessed 15 November 2006.

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MODIS Web. Online at http://modis.gsfc.nasa.gov/index.php. 

Accessed various times from August 2005 through March 2006.

Peterson, E.B. 2000. Analysis and prediction of Patterns in lichen 

communities over the western Oregon landscape. Ph.D. Dissertation, Oregon State University.

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Bromus tectorum (cheatgrass) over a large portion of Nevada from 

satellite imagery. Report for the U.S. Fish and Wildlife Service, 

Nevada State Office, Reno. Carson City, NV: Nevada Natural 

Heritage Program.

Peterson, E.B. 2005. Estimating cover of an invasive grass (Bromus 

tectorum) using tobit regression and phenology derived from two 

dates of Landsat ETM+ data. International Journal of Remote 

Sensing. 26(12): 2491-2507.

Peterson, E.B. 2006. A map of invasive annual grasses in Nevada 

derived from multitemporal Landsat 5 TM imagery. Report for the 

USDI Bureau of Land Management, Nevada State Office, Reno. 

Carson City, NV: Nevada Natural Heritage Program.

R Development Core Team. 2005. R 2.2.0—A language and environment. [Online]. Available: http://www.rproject.org. [15 

November 2005].

Research Systems Inc. 2005. ENVI Version 4.2—The environment 

for visualizing images. Boulder, CO.

Spatial Climate Analysis Service, Oregon State University. 2003. 

PRISM products. [Online]. Available: http://www.ocs.oregonstate.

edu/prism/. Data created in 2003. [9 March 2006].

Thomas, K., Keeler-Wolf, T.; Thorne, J. 2002. Central Mojave field 

data. A digital database (MS Access). US Geological Survey and 

California Department of Fish and Game (available digitally 

from second author).

Thomas, K.; Keeler-Wolf, T.; Franklin, J.; Stine, P. 2004. Mojave 

Desert Ecosystem Program: Central Mojave Vegetation Database. 

Sacramento, CA: USGS Western Ecological Research Center and 

Southwest Biological Science Center.

Tobin, J. 1958. Estimation of relationships for limited dependent 

variables. Econometrica. 26: 24-36.

USDI Bureau of Land Management. 2002. Management Considerations for Sagebrush (Artemisia) in the Western United States: a 

Selective summary of current information about the ecology and 

biology of woody North American sagebrush taxa. Washington, 

DC: USDI, Bureau of Land Management.

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ned.usgs.gov/. [August through December 2005].

West, N.E. 1999. Synecology and disturbance regimes of sagebrush 

steppe ecosystems. In: Entwistle, P.G.; DeBolt, A.M.; Kaltenecker, 

J.H.; Steenhof, K., comps. Sagebrush Steppe Ecosystems Symposium. BLM Pub. No: BLM/ID/PT-001001+1150. Boise, ID: 

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The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

USDA Forest Service Proceedings RMRS-P-52. 2008 41

The Indirect Effects of Cheatgrass Invasion: 

Grasshopper Herbivory on Native Grasses 

Determined by Neighboring Cheatgrass

Julie Beckstead, Susan E. Meyer, and Carol K. Augsperger

Abstract—Invasion biology has focused on the direct effects of 

plant invasion and has generally overlooked indirect interactions. 

Here we link theories of invasion biology and herbivory to explore 

an indirect effect of one invading species on associational herbivory 

(the effect of neighboring plants on herbivory) of native species. We 

studied a Great Basin shadscale (Atriplex confertifolia)-bunchgrass 

community in western Utah, U.S.A. This community is dominated by 

native bunchgrasses, invaded by cheatgrass (Bromus tectorum), and 

home to large populations of grasshoppers (Xanthippus corallipes 

and Melanoplus confusus). Observations of associational herbivory 

provided support for the attractant-source hypothesis for one of 

two native species studied. Elymus elymoides experienced 43% 

greater cumulative herbivory of vegetative structures and produced 

11X fewer reproductive structures in naturally high- compared to 

naturally low-density cheatgrass, thus demonstrating associational 

susceptibility. In contrast, Poa secunda showed no indication of 

associational herbivory; its amount of herbivory did not differ in 

high- compared to low-density cheatgrass patches. Elymus elymoides 

remained a viable food source when cheatgrass senesced, whereas 

P. secunda entered early summer dormancy. Through indirect effects, invading species may generate important shifts in herbivory 

on native species.

Introduction_______________________

The amount of herbivory a plant experiences depends 

not only on its own chemical and physical traits, but also 

on the characteristics of its neighbors (Root 1973; Atsatt 

and O’Dowd 1976). This concept of neighborhood effect, also 

known as associational herbivory, arose from observations 

that plants in natural polycultures suffer less herbivory than 

plants in agricultural or natural monocultures (Pimentel 

1961). Associational herbivory has two contrasting outcomes: 

associational resistance, in which a focal plant species 

experiences less herbivory when grown in the presence of 

a neighbor (Tahvanainen and Root 1972; Root 1973; Bach 

1980), and associational susceptibility, in which a focal 

plant experiences increased herbivory (Brown and Ewel 

1987)—also called “associational damage” (Thomas 1986) 

and “shared-doom” (Wahl and Hay 1995).

The introduction of a “new” neighbor, such as an invasive 

plant species, can alter old and/or create new patterns of associational herbivory by affecting the herbivore’s behavior and/

or population density. The effect could occur directly as a new 

food source and/or indirectly by influencing the herbivore’s 

ability to locate native plants. Associational herbivory theories 

have not been applied to the interactions of invasive species in 

natural systems; this study explores that connection.

Three alternative hypotheses regarding associational 

herbivory emerge from previous research (for review, see 

Huntly 1991). We incorporate invasive species into these 

hypotheses. Predicted outcomes for the native species depend on whether the herbivore is a generalist or a specialist, 

whether the invasive species also is utilized by the herbivore, 

and the relative feeding preference among plant species. If 

an invasive species experiences herbivory, it is predicted to 

come from a generalist that consumes native species of a 

similar phylogeny (Strong and others 1984; Mack 1996).

1. Repellent-plant hypothesis: If the invader acts as a 

repellent-plant to the herbivore (Root 1973; Atsatt and 

O’Dowd 1976) and indirectly affects the herbivore’s 

behavior, then a native species located next to an invader will experience less herbivory than in its absence 

(associational resistance).

2. Attractant-sink hypothesis: If the invader acts as an 

attractant-sink (decoy) to the herbivore (Atsatt and 

O’Dowd 1976) and is directly utilized by the herbivore, 

then a native species located next to an invader will 

experience less herbivory than in its absence (associational resistance).

3. Attractant-source hypothesis: If the invader acts as 

an attractant-source to the herbivore (Wahl and Hay 

1995) and is directly utilized by the herbivore, then a 

native species located next to an invader will experience 

greater herbivory than in its absence (associational 

susceptibility).

To determine which hypothesis is supported for an invasive 

species, we studied a cold desert shadscale (Atriplex confertifolia Wats.)-bunchgrass community in western Utah, U.S.A. 

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Julie Beckstead is an Associate Professor, Department of Biology, Gonzaga University, Spokane, WA 99258; email: beckstead@

gonzaga.edu.

Susan E. Meyer is a Research Ecologist, U.S. Forest Service, 

Shrub Sciences Laboratory, Provo, Utah 84606.

Carol K. Augsperger is a Professor, Department of Plant 

Biology, University of Illinois, Urbana-Champaign, IL.
42 USDA Forest Service Proceedings RMRS-P-52. 2008

Beckstead, Meyer, and Augsperger The Indirect Effects of Cheatgrass Invasion: Grasshopper Herbivory on Native Grasses....

This system has been invaded by cheatgrass (Bromus tectorum L.) and is home to large populations of grasshoppers, the 

primarily grass-feeding redshanked grasshopper (Xanthippus corallipes Haldeman) and the mixed-feeder (grasses and 

forbs) pasture grasshopper (Melanoplus confusus Scudder). 

Previous studies found that cheatgrass is a highly utilized 

food source within this system (Beckstead 2001). Given 

that the herbivores in this system are generalists and that 

cheatgrass is eaten by the herbivores, the repellent-plant 

hypothesis can be eliminated. However, this hypothesis may 

be applicable to other invaded systems in which herbivore 

specialists have a strong influence. To explore the remaining 

two hypotheses, we obtained data from naturally occurring 

patches of cheatgrass. At the study site, cheatgrass exists in 

a mosaic of low-, intermediate-, and high-density cheatgrass 

patches on a scale of <10 m (a distance that grasshoppers 

can traverse in minutes), intermixed with the native plant 

community (Beckstead 2001). Previous field experiments at 

the study site found that differential resource availability 

explains the observed variation in cheatgrass density, and 

variation in natural resistance to invasion by cheatgrass 

(Beckstead and Augspurger 2004). This spatial pattern of 

contrasting densities at this scale provides an opportunity 

to study the indirect effects of cheatgrass on native plant 

species through associational herbivory.

The primary objective of this study is to document whether 

associational herbivory is occurring for a natural system 

involving an invasive species. First, we investigated the 

associational herbivory hypotheses by comparing the amount 

of herbivory on two dominant native grasses surrounded by 

low vs. high densities of cheatgrass (range naturally available 

to grasshoppers). Secondarily, to interpret the associational 

herbivory results, we investigated feeding preferences of the 

grasshoppers among six common native species, including the 

two dominant grasses in the associational herbivory study, 

and developed a mechanistic explanation for the pattern 

based on tissue chemistry and phenology.

Materials and Methods______________

Study Sites and Study Species

The two study sites were located within the same valley in 

western Utah, U.S.A., on Bureau of Land Management land. 

Mean monthly temperature is 18.3 °C and mean annual precipitation is 176 mm (Stevens and others 1983). The observations 

addressing the associational herbivory hypotheses occurred at 

a cheatgrass "mosaic" site (4 ha; 14 km S of Dugway; 40°7’N, 

112°40’W; 1,550 m elevation). Currently, the dominant native 

species are the shadscale shrub and two perennial bunchgrasses: 

sandberg bluegrass (Poa secunda Presl.) and squirreltail (Elymus elymoides Raf.). Experiments and data utilized to develop 

the predictive mechanism (i.e., grasshopper feeding preferences, 

tissue chemistry, and phenology) were conducted at a nearby 

site. Both study sites have been grazed by cattle, sheep, wild 

horses, and antelope (J. Beckstead, personal observation). Prior 

to data collection, the study sites were enclosed by a wire fence 

to exclude large grazing ungulates.

Cheatgrass, a winter annual of European origin (Mack 

1981), is the most common and widespread invasive species 

at both study sites. In western Utah, its seeds usually 

germinate in October and November in response to autumn 

rains (Beckstead and others 1995) and seeds ripen in June 

and July, usually before most native grass species.

Two common grasshopper herbivores of the community are 

Xanthippus corallipes and Melanoplus confusus. Xanthippus corallipes is a large grasshopper (female body length 

35–41mm; male body length 24–30 mm) that feeds exclusively 

on grasses and sedges (Pfadt 1994). It has a wide distribution 

in western North America, where it inhabits grassland and 

shrub-grass communities. Melanoplus confusus is a smaller 

grasshopper (female body length 23–24 mm; male body length 

18–19 mm) that feeds on both grasses and forbs (Pfadt 1994). 

It is common in grassland habitats of the West and meadows 

and pastures of the Midwest and eastern United States. 

Both grasshopper species have an early season phenology: 

eggs hatch in late February–March; nymph stages progress 

quickly in April; adult size is attained in May; and egg-laying 

and death occur in June–July (Pfadt 1994).

Associational Herbivory Study

To investigate associational herbivory on native species in 

the presence of neighboring cheatgrass, we utilized naturally 

occurring low- and high-density cheatgrass patches interspersed among a native shadscale-bunchgrass community at 

the “mosaic” site. We arbitrarily set limits for the low- (<15% 

cover) and high- (>85% cover) density cheatgrass patches. In 

a mapped area (5,000 m2) within the study area, low-density 

patches covered 25% and were intermixed with high-density 

patches that covered 30% of area (data obtained from aerial 

photographs taken in 1999). Native species are found within 

the mosaic of cheatgrass density patches, most commonly 

P. secunda and E. elymoides. Frequencies for these two native species were obtained for 15 low- and 15 high-density 

cheatgrass patches in 1997. Frequency data were analyzed 

using Yates corrected chi-square test. The total percent 

plant cover was slightly lower in low- than in high-density 

cheatgrass patches (Beckstead 2001), although the two patch 

types had similar species richness (J. Beckstead, unpublished 

data on file at Gonzaga University).

Grasshoppers were quantified in early June 1999, when 

grasshoppers had achieved adult size. We randomly selected 

12 low- and 12 high-density patches, each with two 30 x 30 

cm plots, in which to quantify grasshopper densities. Then, 

in the early morning as the observer walked towards each 

plot within each patch, the number of grasshoppers jumping 

out was counted (modified USDA Animal and Plant Health 

Inspection Service (APHIS) method; Onsager and Henry 

1977). Grasshopper densities were analyzed using a one-way 

analysis of variance (ANOVA) and a mixed model procedure 

(PROC MIXED) ( SAS 1997) with subsampling; data met 

the normality assumption. Herbivory effects on cheatgrass in 

low- and high-density cheatgrass patches were measured in 

a concurrent study in many of the same cheatgrass patches 

assessed for grasshopper density (Beckstead and Augspurger 

2004); results are summarized in the discussion.

The Mormon cricket (Anabrus simplex Haldeman), a shieldbacked katydid and not a true cricket (Pfadt 1994), migrated 

to the site in 1998. To isolate the herbivore effects of resident 

grasshoppers from that of migrating Mormon crickets, we 

baited the study area with 3 L of carbaryl (active ingredient in SevinTM) wheatbran from March to June 1999. This 
USDA Forest Service Proceedings RMRS-P-52. 2008 43

The Indirect Effects of Cheatgrass Invasion: Grasshopper Herbivory on Native Grasses... Beckstead, Meyer, and Augsperger

selective insecticidal bait differentially kills Mormon crickets 

but not grasshoppers (Quinn and others 1989; confirmed by 

counting carcasses following bait application).

In late June 1999, we measured cumulative grasshopper 

herbivory on two native grasses in low- and high-density patches 

of neighboring cheatgrass. The target species in this natural 

experiment were E. elymoides and P. secunda, the two most 

common native grass species. We randomly selected 17 lowand 17 high-density cheatgrass patches between 5–15 m wide 

within a 5,000 m2 area at the study site. Within each patch, 

two individuals of each species were randomly selected for a 

total of 34 E. elymoides plants and 34 P. secunda plants. Given 

that plants in this desert system have a very short growing 

season, we chose to measure the herbivory only once at the end 

of the growing season. At the time of this cumulative herbivory 

measurement, cheatgrass was senescing, E. elymoides was at 

the seed-ripening stage, P. secunda had just gone dormant, 

and grasshoppers were nearing the end of their life cycle.

We measured the amount of herbivory on above-ground 

vegetative structures (leaves and stems combined) and 

reproductive structures (seed-bearing stems) of E. elymoides

and P. secunda. For vegetative structures, the percent 

removed by herbivores (eaten) was determined by visually 

estimating the damage on the individual and placing each 

into one of five classes (0, 25, 50, 75, and 100% removed; 

similar to White and Whitham 2000). Although this estimation has its limitations, it is reasonably accurate, given the 

large measurement increments (25%).

We measured herbivory on reproductive structures for 

each plant by counting the number of stems supporting seeds 

that remained following herbivory. The number of grazed 

seed-bearing stems was not counted due to the difficulty of 

distinguishing grazed reproductive stems from the narrow, 

round grazed leaves for P. secunda. Assuming that plant 

size is the best predictor for seed-bearing stem production 

(i.e., number of stems present in the absence of herbivory), 

we measured plant basal diameter (plant width at ground 

level) to control for variation in reproduction due to plant size. 

Basal diameter is a good indicator of reproductive potential 

for bunchgrasses (S. Monsen, personal observation; and supported by significant correlations in these results).

We analyzed herbivory of vegetative structures (percent 

leaf/stem eaten) as a mixed model ANOVA and herbivory of 

reproductive structures (number of seed-bearing stems remaining) as a mixed model analysis of covariance (ANCOVA; 

PROC MIXED) ( SAS 1997). This two-factor nested design 

with subsampling included fixed effects (patch, species 

nested within patch, and basal diameter as a covariate) 

and random effects (replication and subsampling). PROC 

MIXED is recommended for mixed models and is based on 

restricted maximum likelihood estimation (REML) of linear 

statistical models involving both fixed and random effects (for 

further discussion see Steel and others 1997). Assumptions 

of ANCOVA were met following model selection procedures 

recommended by Littell and others (1996). Differences 

between plant basal diameter for each species in low- and 

high-density patches were determined by a similar mixed 

model ANOVA (SAS 1997). Appropriate transformations 

were performed when needed to meet the assumption of 

normality. To compare differences between the two species 

within each patch type, we used the protected Fisher’s LSD 

using the least square means (to account for the appropriate 

standard errors) (SAS Institute 1997).

Comparative Herbivory Experiment

To compare the amount of herbivory among common native 

species, we planted six species in an experimental array in a 

common garden and subjected them to the natural density of 

grasshoppers. The species were the five dominant perennial 

grasses and one forb found in this shadscale-bunchgrass community (table 1). Two of the grass species were E. elymoides

and P. secunda, the target species for the associational herbivory study. To create the experimental arrays, seeds were 

collected from the study site when available; otherwise, they 

were obtained from desert communities with similar arid 

environments. Seeds were germinated and seedlings grown 

for 5 months in a greenhouse. In March 1997, one seedling 

of each species was transplanted 15 cm apart into each of 

six corners of nine adjacent hexagons in a 1.44-m2 plot. Each 

plot was spaced 1 m apart and replicated 12 times. Prior to 

transplanting the seedlings, the pre-existing seed bank of 

Table 1—Descriptive traits of six native plant species common in the shadscale-bunchgrass community 

and the invasive cheatgrass. Nomenclature is from Welsh and others (1987).

Species Growth form Flowering Seed-ripening

Elymus elymoides Raf.

Squirreltail Perennial bunchgrass June July

Elymus smithii Gould

Western wheatgrass Perennial rhizomatous grass June-July July-Aug

Stipa hymenoides R. & S.

Indian ricegrass Perennial bunchgrass June July

Poa secunda Presl.

Sandberg bluegrass Perennial bunchgrass May June

Stipa comata Trin. & Rupr.

Needle-and-thread Perennial bunchgrass June July

Sphaeralcea munroana Gray

Munroe globemallow Perennial forb July-Aug Aug-Sept

Bromus tectorum L.

Cheatgrass Annual grass May June
44 USDA Forest Service Proceedings RMRS-P-52. 2008

Beckstead, Meyer, and Augsperger The Indirect Effects of Cheatgrass Invasion: Grasshopper Herbivory on Native Grasses....

cheatgrass was removed in 1996 by raking the litter and seeds 

off each plot and then watering in August 1996 to promote 

precocious germination of any remaining seeds (Beckstead 

and others 1995). At the time of measuring herbivory, the 

natural densities of X. corallipes and M. confusus were 

15-20 grasshoppers per m2 (J. Beckstead, personal observation). The arrays covered an area of 1,024 m2 and were 

surrounded by cheatgrass densities of 11,834/m2 ± 3,056 

(mean ± SD; n = 17).

In June 1998, cumulative herbivory of the six native 

species was measured once during flowering, which occurred 

near the end of the grasshopper life cycle. We estimated 

the percent eaten (herbivory) for each species collectively 

within a plot using five classes (as described above). We 

analyzed species differences in percent eaten (herbivory) with 

analysis of variance (ANOVA; PROC GLM) (SAS 1997). The 

assumption of normality was met by a square root (y + 3/8) 

transformation. Multiple comparisons between the six native 

species were performed with Scheffe’s test.

To assess the relationship between herbivory and nitrogen and silica content, plant tissue was collected in June 

1999 from a natural community near the experimental site. 

Samples were taken from 15 individuals for each of the six 

native plant species at the flowering stage and at three 

developmental stages of cheatgrass (bolting, flowering, and 

seed-ripening). Above-ground tissues (leaves and stems, 

excluding inflorescences, except for cheatgrass) were dried 

at 60 °C for 48 hours, ground to a uniform powder, randomly 

combined into three or five samples (due to limited amount of 

tissue collected from each individual), and analyzed for total 

nitrogen content using the Association of Official Analytical 

Chemists (AOAC) semi-automated method (Horwitz 1980). 

Due to a limited amount of tissue, samples of P. secunda, 

S. munroana, and bolting and seed-ripening cheatgrass 

samples were analyzed for total nitrogen using the AOAC 

micro-Kjeldahl method (Horwitz 1980). Nitrogen content was 

expressed as total percent nitrogen. Percent silica content 

was analyzed using a muffle furnace for complete combustion followed by a hydrochloric acid extraction (Allen and 

others 1974).

Species differences in percent nitrogen and silica were 

analyzed with ANOVA by a general linear model procedure 

(PROC GLM) (SAS 1997). Values for both traits were arcsine 

transformed to meet the assumption of normality. Multiple 

pairwise comparisons for the six native species were performed with protected Fisher’s least significance difference 

(LSD) tests. Pairwise comparisons were also made for each 

of the six native species and the three developmental stages 

of cheatgrass.

Descriptive traits, such as plant growth form and phenology, were noted at the time of tissue collection and confirmed 

by species descriptions in Welsh and others (1987; table 1). 

This information was used for data interpretation.

Results ___________________________

Associational Herbivory Study

The low-density patches and high-density patches differed 

in frequency for the two native grass species, P. secunda and 

E. elymoides, and the density of grasshoppers. Although 

P. secunda was found more frequently in the low-density 

patches than the high patches, E. elymoides did not differ in 

its frequency between the two patch types (table 2). Grasshopper densities were twice as high in the high-density than 

in the low-density cheatgrass patches (ANOVA, F = 14.74, 

df = 1, 11, P = 0.0028) (table 2).

Mean cumulative herbivory on vegetative (leaves and 

stems) and reproductive (seed-bearing stems) structures 

was greater in the high-density cheatgrass patches in comparison to the low-density patches (patch effect, ANOVA, 

F = 24.96, df = 1,16, P < 0.001 and ANCOVA, F = 61.10, 

df = 1, 16, P < 0.0001, respectively). However, the two native grasses differed in their response to the associational 

herbivory on both vegetative structures (species effect nested 

within patch, ANOVA, F = 100.51, df = 2,16, P < 0.0001) 

and reproductive structures (species effect nested within 

patch, ANCOVA, F = 80.50, df = 2,32, P < 0.0001). For E. elymoides, the mean percent eaten per individual (leaf/stem) 

was higher in high- (98.3%) than in low-density cheatgrass 

patches (55.2%; Fisher’s LSD, t = 5.37, P < 0.0001) (fig. 1). 

The high herbivory in high-density patches resulted in a 

lower mean number of seed-bearing stems per individual in 

comparison to low-density cheatgrass patches: 1.4 vs. 15.7, 

respectively (Fisher’s LSD, t = 10.56, P < 0.0001) (fig. 1). 

Table 2—Native plant (frequency and basal diameter) and grasshopper (density) descriptive data for low- vs. 

high-density cheatgrass patches at the “mosaic” study site. Data are means ± 1 SD. Different small letters indicate differences between low- vs. high-density cheatgrass patches significant at P < 0.05 from 

Protected Fisher’s LSD test following Analysis of Variance for basal diameter and grasshopper density 

and chi-square analysis for native plant frequency.

Native plants Grasshoppers

Frequency Frequency Basal dia (cm) Basal dia (cm ) Density

Cheatgrass patch E. elymoides P. secunda E. elymoides P. secunda Number m–2

- - - - - - - - (n = 15) - - - - - - - - - - - - - - - - - (n = 34) - - - - - - - - - - (n = 24)

Low density 27% a 100%a 14.2 a ± 3.8 8.2 a ± 1.9 11.6 b ± 8.0

(≤ 15% cover)

High density 22% a 67%b 14.9a ± 4.4 9.1a ± 2.4 22.7 a ± 6.5

(≥ 85% cover)
USDA Forest Service Proceedings RMRS-P-52. 2008 45

The Indirect Effects of Cheatgrass Invasion: Grasshopper Herbivory on Native Grasses... Beckstead, Meyer, and Augsperger

E. elymoides experienced associational susceptibility when 

located in high-density cheatgrass patches as a result of 

the higher concentration of grasshoppers in high- compared 

to low-density patches. In contrast, for P. secunda neither 

herbivory on vegetative or reproductive structures differed 

between patch types (Fisher’s LSD, t = 1.69, P = 0.10 and t = 

0.54, P = 0.59, respectively); thus there was no associational 

herbivory (fig. 1).

Plant basal diameter was a significant covariate for 

herbivory in the reproductive structures model (ANCOVA, 

F = 47.00, df = 1, 67, P < 0.0001). It was also significantly 

correlated with the number of seed-bearing stems for 

E. elymoides (Pearson’s r = 0.31, P < 0.01) and P. secunda

(Pearson’s r = 0.56, P < 0.0001). The basal diameters of 

E. elymoides and P. secunda did not differ significantly 

between low- and high-density patches (ANOVA, F = 1.44, 

df = 1, 16, P = 0.25) (table 2), indicating that differences in 

number of seed-bearing stems could not be attributed to 

differences in plant size.

Figure 1—Cumulative herbivory on vegetative (left 

axis; mean % eaten/individual) and reproductive (right 

axis; mean number of seed-bearing stems remaining/

individual) structures for E. elymoides, a highly preferred 

species, and P. secunda, a less-preferred species, in 

high- vs. low-density cheatgrass patches (means + 1

SE; n = 26 and n = 34, respectively).

Figure 2—Amounts of grasshopper cumulative herbivory 

(mean % eaten/individual) on six native plant species grown 

in common garden plots (means + 1 SE; n = 12). Different

small letters indicate significant differences between species 

(P < 0.05; Scheffe’s multiple comparison test).

Comparative Herbivory Experiment

Cumulative herbivory varied significantly among the six 

native species (fig. 2; one-way ANOVA, F = 69.29, df = 5, 

P< 0.0001) and resulted in three distinct groups: (1) a highly 

preferred group (>70% plant cover removed; E. elymoides, 

E. smithii, and S. hymenoides), (2) a less-preferred group 

(<30% plant cover removed; P. secunda and S. comata), and 

(3) one unpalatable species (0% plant cover removed; S. munroana). Thus, the two target species for the associational 

herbivory study were categorized as a highly preferred species 

(E. elymoides) and a less-preferred species (P. secunda).

The six native species differed significantly in total percent 

nitrogen (one-way ANOVA, F = 54.34, df = 5, P < 0.0001) and 

silica content (one-way ANOVA, F = 74.12, df= 5, P<0.0001). 

However, only total percent nitrogen was separable into 

three distinct groups that corresponded to the herbivore 

preference groups of the palatable species (table 3; fig. 2). 

The less-preferred species, P. secunda and S. comata, contained significantly lower nitrogen than two highly-preferred 

species, E. elymoides and S. hymenoides. E. smithii, one of 

three highly preferred species, did not differ from the lesspreferred species. Although S. munroana, the only forb, had 

the highest total percent nitrogen, its leaves were not eaten. 

It did receive some damage from the grasshoppers, which 

congregated around the base of the plants and girdled some 

of the lower branches resulting in branch death.

Difference in total percent nitrogen between cheatgrass 

and the native species depended on the developmental 

stage of cheatgrass (one-way ANOVA, F = 73.51, df = 8, 

P < 0.0001) (table 3). Young cheatgrass plants at the bolting 
46 USDA Forest Service Proceedings RMRS-P-52. 2008

Beckstead, Meyer, and Augsperger The Indirect Effects of Cheatgrass Invasion: Grasshopper Herbivory on Native Grasses....

stage contained significantly higher levels of nitrogen 

than the five palatable species. At its flowering stage, 

total percent nitrogen did not differ significantly from 

E. smithii, P. secunda, or S. comata. Both E. elymoides

and S. hymenoides contained higher total percent nitrogen 

than flowering cheatgrass. By the seed-ripening stage, 

cheatgrass contained significantly lower levels of nitrogen 

in comparison to all native species.

Discussion ________________________

This study offers a unique example of an invasive species 

acting as an attractant-source resulting in associational 

susceptibility for a native species. Native E. elymoides experienced greater herbivory to both reproductive and vegetative structures in high- compared to low-density cheatgrass 

patches; thus demonstrating associational susceptibility. 

In contrast, native P. secunda showed no indication of associational herbivory; its amount of herbivory did not differ 

in low- compared to high-density cheatgrass patches. These 

contrasting associational herbivory patterns can be explained 

in part by the herbivores’ feeding preference, which is driven, 

in turn, by the tissue chemistry and phenology of cheatgrass 

in relation to the two native species.

Associational Herbivory Study

First, we found that high-density cheatgrass patches attracted greater numbers of generalist grasshoppers than 

did low-density cheatgrass patches, presumably by providing a preferred feeding area. According to the resource 

concentration hypothesis (Root 1973), herbivores should 

accumulate where their food resources are most concentrated. 

Supporting this, measurements taken concurrently with this 

study found that, in both low- and high-density cheatgrass 

patches, herbivory on cheatgrass was higher with greater 

cheatgrass biomass and density (Beckstead and Augspurger 

2004). We propose that cheatgrass acts as an attractantsource by its influence on the grasshopper population density 

(a numerical response via immigration). Grasshopper densities were twice as high in high-density cheatgrass patches 

as they were in low-density patches. The high nitrogen 

levels found in bolting and flowering cheatgrass plants 

may attract and/or retain more grasshoppers in the highdensity compared to the low-density cheatgrass patches. It 

is unknown whether cheatgrass densities may also influence 

grasshopper survival or how frequently these patterns occur 

across years. Future studies are necessary to decipher the 

mechanism(s) by which cheatgrass acts as an attractantsource and the thresholds of cheatgrass and grasshopper 

densities required for associational herbivory. Given the 

physical uniformity of the site, the magnitude of the effect we measured, and the relatively small spatial scale 

of low- vs. high-density cheatgrass patches (<10 m), it 

seems unlikely that grasshopper density and consequent 

differences in herbivory levels on associated native grasses 

were greatly influenced by factors other than cheatgrass 

density itself.

In this study using a natural range of cheatgrass densities, the attractant-source hypothesis was supported by 

E. elymoides, but not by P. secunda. Using a marine system, 

Wahl and Hay (1995) found that the strength of associational 

herbivory patterns varied with the relative preference of 

the omnivorous sea urchin for pairwise combinations of 

host seaweed and epibionts. It is possible that the relative 

preference in our plant-grasshopper system is a mechanism 

underlying our observed associational herbivory pattern.

Table 3—Total percent nitrogen and percent silica for plant tissue of six native species at their flowering stage and 

for cheatgrass at three developmental stages. Values are means (± 1 SE). Different small letters indicate 

significant differences among species at P < 0.05 from Protected Fisher’s LSD test following Analysis 

of Variance. Asterisks indicate differences in percent nitrogen between selected species pairs (a given 

native species and a cheatgrass stage) significant at P < 0.05 from Protected Fisher’s LSD test following 

analysis of variance (* P < 0.05; ** P < 0.01; *** P < 0.001; ns = nonsignificant).

Percent nitrogen at 

cheatgrass developmental stages

 % % Bolting, n = 3 Flowering, n = 5 Seed-ripening, n = 3

Species Nitrogen Silica 2.35 (0.04) 1.42 (0.07) 0.37 (0.01)

E. elymoides, n = 5 1.64 bc 4.54 a *** * ***

(0.06) (0.20)

E. smithii, n = 5 1.49 cd 2.71 b *** ns ***

(0.10) (0.03)

S. hymenoides, n = 5 1.74 b 1.76 e *** ** ***

(0.08) (0.10)

P. secunda, n = 3 1.29 d 2.52 bc *** ns ***

(0.11) (0.25)

S. comata, n = 5 1.40 d 2.50 cd *** ns ***

(0.04) (0.16)

S. munroana, n=3 3.26 a 0.51 f *** *** ***

(0.13) (0.03)
USDA Forest Service Proceedings RMRS-P-52. 2008 47

The Indirect Effects of Cheatgrass Invasion: Grasshopper Herbivory on Native Grasses... Beckstead, Meyer, and Augsperger

A Mechanistic Explanation of Associational 

Herbivory

Among the five common grass species in the shadscalebunchgrass community, the amount of herbivory corresponded to the total percent nitrogen and not percent silica. 

The only forb in the study, S. munroana, which had the 

highest level of nitrogen and was not consumed by grasshoppers, was the exception. The plant chemistry of forbs, in 

general, is distinct from that of grasses. The unpalatability of 

S. munroana could be due to some unmeasured secondary 

compound; Bernays and Chapman (1970) found that forbs 

lack a chemical substance necessary to induce biting by many 

grass-feeding grasshoppers. Among the five grass species, 

the highly preferred species, E. elymoides, E. smithii, and 

S.hymenoides, had higher nitrogen levels than less preferred 

species. High nitrogen or amino acid content has been associated with increases in grasshopper herbivory (for review see 

Chapman 1990). Although high silica content in grasses has 

been proposed to be an important deterrent for grasshopper 

herbivory (Joern 1979), no such pattern was found in this 

study. Water content (Lewis and Bernays 1985) and sucrose 

concentration (Bernays and Chapman 1978) are potentially 

important tissue traits affecting herbivory of grasses, but 

were not measured in this study.

In addition to explaining feeding preferences, we propose 

that phenology is also a key factor driving the contrasting 

associational herbivory patterns for the highly-preferred, 

E. elymoides (associational susceptibility) and the lesspreferred, P. secunda (no associational herbivory). As noted 

in table 1, the early summer dormancy of P. secunda may 

allow it to escape herbivory temporally; it becomes dormant 

before the grasshoppers reach adult size. Although this study 

did not include pairwise preference experiments between 

cheatgrass at different phenological stages and each of 

the native species, we did compare total percent nitrogen 

of cheatgrass at three developmental stages with nitrogen 

levels of native species at the flowering stage. Cheatgrass, 

an annual, undergoes these stages in a very short period 

of time (Young and others 1969) and each perennial native 

species at its flowering stage will encounter more than one 

cheatgrass stage. Based on these comparisons, we predict 

that: (1) young cheatgrass plants will be preferred over all 

native grasses, (2) E. elymoides and S. hymenoides will be 

preferred over flowering cheatgrass plants, and (3) all native grasses in the flowering stage will be preferred over 

cheatgrass at the seed-ripening stage. Direct feeding trials 

would be a means of testing these predictions.

Our findings for the associational herbivory study support 

our predictions of pairwise comparisons of cheatgrass with 

the common native species. The herbivore effects in this 

experiment were measured at seed production for the native 

grasses, such that the native species had experienced herbivory during all cheatgrass stages. E. elymoides, a predicted 

preferred food source over cheatgrass at two of three stages, 

had lower final reproductive output in high- compared with 

low-density cheatgrass patches (i.e., associational susceptibility). In contrast, P. secunda is predicted to be preferred 

over cheatgrass only during cheatgrass’ seed-ripening stage; 

however, it escapes herbivory at this time through early 

summer dormancy (table 1). Associational susceptibility 

appeared to result when grasshoppers shifted food preference 

to E. elymoides, driven by the early phenology of cheatgrass. 

White and Whitham (2000) found associational susceptibility resulted when a generalist herbivore reached outbreak 

proportions and consumed their preferred host; the herbivore 

then moved to nearby less-preferred hosts to complete their 

life cycle. In this study, cheatgrass was not completely eliminated by grasshoppers. Instead, it appears that cheatgrass 

became less preferred near the end of its life cycle.

This study indicates that indirect interactions between 

invasive and native species may have large consequences 

for native species success. The indirect effect of cheatgrass 

on native species via associational herbivory indicates 

the potential for a shift in species composition within the 

shadscale-bunchgrass community, specifically, a decrease 

in the E. elymoides population relative to no change for 

P. secunda populations. Although these shifts in species 

composition are possible, the long-term effects of associational 

herbivory and the frequency of these events on community 

composition are unknown.

Acknowledgments__________________

We thank the Bureau of Land Management and Gary Kidd 

for their cooperation; Stephanie Carlson, Leila Ellsworth, 

and Jennifer Schmidt for field assistance; Greg Abbot and 

Tim Graham for grasshopper identification; and Bruce Webb 

(Brigham Young University), for assistance with tissue 

analysis. Statistical advice was provided by Susanne Aref. 

The manuscript was improved by insightful comments from 

Catherine Bach, George Batzli, Evan DeLucia, Tim Graham, 

Ken Paige, and discussion with the Ingrid Parker lab group. 

Logistic support was provided by USDA, Forest Service, 

Shrub Sciences Laboratory in Provo, Utah. The study was 

financially supported by research grants from the Graduate College and the Program in Ecology and Evolutionary 

Biology at the University of Illinois.

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for the facts and accuracy of the information presented herein.
USDA Forest Service Proceedings RMRS-P-52. 2008 49

Abstract—Over the past several decades, cheatgrass (Bromus 

tectorum) has been continually expanding in the sagebrush steppe 

ecosystem. There has been very little research that examines why 

cheatgrass is able to invade these communities. To determine the 

effects of herbaceous vegetation removal and fire on available water 

for cheatgrass invasion, as well as effects of cheatgrass itself, we 

measured spring soil moisture and vegetation cover on nine sites in 

Nevada and Utah. Total time of soil water availability was calculated 

for 1 April through 30 June in 2002 and 2003 on plots that were left 

intact, sprayed with an herbicide to kill all herbaceous vegetation, 

burned, or both burned and sprayed in 2001. Soil water was also 

measured on subplots either left unseeded or seeded to cheatgrass 

in fall 2001. The number of days that soil moisture was available 

increased with site elevation and annual precipitation. Time of 

surface soil (1 to 3 cm) water availability was increased by 12 days 

in a wetter year (2003) compared to a drier year (2002) and was 

not significantly affected by disturbance. Time of subsurface soil 

water availability (13 to 15 and 28 to 30 cm) was increased more by 

disturbance than seasonal precipitation. Vegetation removal plots 

had 7 days longer subsurface soil water availability than intact plots 

the first year after removal. Burned plots had 4 to 6 days longer 

subsurface soil water availability than unburned plots. Cheatgrass 

establishment was limited, especially at the Utah sites, and did not 

significantly affect soil water availability. Soil water resources may 

be increased by disturbance, higher annual precipitation, or both 

and thereby facilitate weed invasion.

The Effect of Herbaceous Species Removal, Fire, 

and Cheatgrass (Bromus tectorum) on Soil Water 

Availability in Sagebrush Steppe

Alison Whittaker, Bruce Roundy, Jeanne Chambers, Susan Meyer, 

Robert R. Blank, Stanley G. Kitchen, and John Korfmacher

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Alison Whittaker is a Habitat Biologist, Utah Division of Wildlife 

Resources, Northeastern Region, Vernal, UT; email: alisonwhittaker@utah.gov.

Bruce Roundy is a Professor, Department of Plant and Wildlife 

Sciences, Brigham Young University, Provo, UT.

Jeanne Chambers is a Research Ecologist, U.S. Forest Service, 

Rocky Mountain Research Station, Reno, NV.

Susan Meyer is a Research Ecologist, U.S. Forest Service, Rocky 

Mountain Research Station, Shrub Sciences Laboratory, Provo, UT.

Robert R. Blank is a Soil Scientist, USDA-ARS, Reno, NV.

Stanley G. Kitchen is a Research Botanist, U.S. Forest Service, 

Rocky Mountain Research Station, Shrub Sciences Laboratory, 

Provo, UT.

John Korfmacher is a Physical Scientist, U.S. Forest Service, 

Rocky Mountain Research Station, Fort Collins, CO.

Introduction_______________________

The spread of cheatgrass (Bromus tectorum) has displaced 

much of the native vegetation throughout the west. Cheatgrass was introduced to the Great Basin region at the end 

of the 19th century and in just 20 to 30 years had spread 

throughout most of the western United States (Knapp 

1996; Mack 1981). Today it is hard to find a big sagebrush 

(Artemisia tridentata) community that has not been invaded 

by cheatgrass (Knapp 1996). Cheatgrass invasion in the 

Great Basin has changed the fire frequency from 50 to 200 

years to 3 to 5 years, resulting in cheatgrass-dominated 

communities (Whisenant 1990). But how does this plant or 

any other weed invade a community?

Elton (1958) argued that communities with lower plant 

diversity were most susceptible to invasion. Levine and 

D’Antonio (1999) suggested that this is not necessarily the 

case. Their review of community diversity and invasibility 

indicated that more diverse communities were often more 

susceptible to invasion than less diverse communities. Stohlgren and others (1999) also found that communities with 

higher diversity were more invasible and that invasion is 

more closely related to the amount of resources that were 

available in the community. Other studies have shown that 

nutrient-enriched plots were dominated by weedy annuals 

(Carson and Barrett 1988; Foster and Dickson 2004). 

Davis and others (2000) observed that any factor leading 

to increased availability of limiting resources increases 

susceptibility of a community to invasion.

Potentially limiting resources to plant growth such as 

light, water, and nutrients plays a large part in community 

composition (Tilman 1986). Resources fluctuate annually, 

seasonally, and spatially. Susceptibility of a community to 

invasion varies with these fluctuations in resource availability (Davis and others 1998; Davis and others 2000). 

Disturbance may increase resources by reducing resource 

use by pre-disturbance vegetation. Disturbance also leaves 

an open space for invasion (Burke and Grime 1996).

In the Great Basin, water is a major limiting resource to 

plant growth and development, and greatly influences plant 

community composition (Ceballos and others 2002; Coronato 

and Bertiller 1996; Daubenmire 1972; Fowler 1986). Soil 

moisture changes over time according to the amount of 

precipitation and evapotranspiration, which varies depending on the season, topography, vegetation, soil depth and 

texture, and land use history (Coronato and Bertiller 1996; 

Wythers and others 1999). In semi-arid climates like the 
50 USDA Forest Service Proceedings RMRS-P-52. 2008

Whittaker, Roundy, Chambers, Meyer, Blank, Kitchen, and Korfmacher The Effect of Herbaceous Species Removal, Fire, and Cheatgrass (Bromus tectorum)...

Great Basin, soil water recharge occurs during the winter 

months usually in the form of snow. Precipitation in the late 

spring and summer is very sporadic with long dry periods 

between precipitation events (Cable 1980). This leads to a 

wide variation in the amount of soil moisture that is available over time. Other variations in soil moisture availability 

come from the type and amount of vegetation that is present. 

Vegetation affects infiltration as well as evapotranspiration. 

The loss of vegetation from grazing, fire, or other disturbances may decrease infiltration and transpiration and 

increase evaporation (Cable 1980). Growth form also affects 

transpiration. Woody species use water from throughout the 

soil profile, whereas annual plants only use water from the 

upper most portions of the soil profile (Daubenmire 1968). 

Soil moisture loss through transpiration is also highest at 

the time that the vegetation is actively growing. Changes in 

precipitation, vegetation cover, or disturbance can increase 

the availability of soil moisture and may play a large role in 

the susceptibility of a community to invasion.

Several wet years in the early 1980s were associated with 

an increase of cheatgrass invasion in salt desert communities 

(Meyer and others 2001). Soto and Diaz-Fierros (1997) found 

that after a fire burns gorse (Ulex europaeus, L.) shrubland, 

more precipitation actually hit the ground, but lack of vegetation resulted in increased evaporation and decreased 

transpiration. This changed how the soil profile dried out. 

The upper portion of the soil profile dried very quickly while 

the lower portion took longer to dry out. Other studies have 

shown that burning actually decreases the amount of soil 

moisture that is available (Daubenmire 1968; Melgoza and 

others 1990; Redmann 1978). Grazing can also change the 

vegetation in a community; however, one study showed that 

grazing did not have an effect on soil moisture (Coronato 

and Bertiller 1996). Melgoza and others(1990) also found 

that the upper soil profile dries out more rapidly on sites 

dominated by cheatgrass.

Our objectives were to determine the effects of vegetation 

disturbance on the number of days that water was available 

in the soil (matric potential > –1.5 MPa) across an elevational gradient for big sagebrush communities in the Great 

Basin. We hypothesized that the time of available water 

should be extended as elevation increases with increasing 

precipitation, and the loss of vegetation associated with 

disturbance should result in a longer period of available 

water as a result of decreased transpiration. The presence of cheatgrass after disturbance may also decrease soil 

moisture availability. To address these questions, available 

soil moisture was measured at three different depths under 

three different treatments for a period of 2 years, focusing 

on the April through June growing period.

Methods __________________________

Study Sites

There were five study sites selected in Utah and four selected in Nevada. We selected sites that had an intact native 

perennial understory. In Nevada, one site was located in 

a crested wheatgrass (Agropyron cristatum (L) Gaertn.) 

community and three sites were located in big sagebrush communities along an elevational gradient. A crested wheatgrass 

site and three big sagebrush sites similar to the Nevada sites 

were also selected in Utah. The fifth Utah site was located 

in a grass community dominated by bottlebrush squirreltail 

(Elymus elymoides (Raf.) Swezey) (table 1).

The four Nevada study sites are located on the HumboltToiyabe National Forest at 39o north latitude, 117° 30’ west 

longitude. The crested wheatgrass site was located in an area 

that was cleared in the early 1960s and was reseeded with 

crested wheatgrass. The five Utah study sites are located 

in Tooele County on land managed by the Bureau of Land 

Management at 40° north latitude, 112° west longitude.

The squirreltail and crested wheatgrass sites in Utah and 

the crested wheatgrass site in Nevada have a few scattered 

Wyoming big sagebrush plants. The dominant plants on the 

lower elevation site in Nevada are Wyoming big sagebrush, 

Sandberg bluegrass (Poa secunda J. Presl.), and needle 

and threadgrass (Hesperostipa comata (Trin. & Rupr.) 

Barkworth). The middle elevation site in Nevada is dominated 

by mountain big sagebrush and Sandberg bluegrass. The 

upper elevation site in Nevada is dominated by mountain 

big sagebrush, Idaho fescue (Festuca idahoensis Elmer), and 

western wheatgrass (Pascopyrum smithii (Rybd.) A. Love). 

In Utah, the lower elevation site is dominated by Wyoming big 

sagebrush, bluebunch wheatgrass (Pseudoroegneria spicata

(Pursh) A. Love) and galleta (Pleuraphis jamesii Torr.). The 

middle elevation site in Utah is dominated by mountain big 

sagebrush, bluebunch wheatgrass, and Fendler’s bluegrass 

Table 1—Description of study sites in Utah and Nevada.

State Site Elevation (m) Soil texture Soil classification

Nevada Crested Wheatgrass 2065 sandy loam Xereptic Haplodurid

Lower Wyoming big sagebrush 1960 sandy loam Xeric Haplocambid

Middle mountain big sagebrush 2190 sandy loam Typic Haploxeroll

Upper mountain big sagebrush 2380 loam Xeric haplocryoll

Utah Bottlebrush squirreltail 1608 loam Xeric Haplocalcid

Crested Wheatgrass 1628 gravelly sandy loam Xeric Haplocalcid

Lower Wyoming big sagebrush 1710 gravelly sandy loam Petrocalcic Palexeroll

Middle mountain big sagebrush 2085 gravelly loam Lithic Calcixeroll

Upper mountain big sagebrush 2274 gravelly loam Lithic Calcixeroll
USDA Forest Service Proceedings RMRS-P-52. 2008 51

The Effect of Herbaceous Species Removal, Fire, and Cheatgrass (Bromus tectorum)... Whittaker, Roundy, Chambers, Meyer, Blank, Kitchen, and Korfmacher

(Poa fendleriana (Steud.) Vasey). The dominant plants of the 

upper elevation site in Utah are mountain big sagebrush, 

Fendler’s bluegrass, spike fescue (Leucopoa kingii (S. Wats.) 

W. A. Weber), and Columbia needlegrass (Achnatherum 

nelsonii (Scribn.) Barkworth.

Experimental Design and Treatments

We imposed three disturbance treatments including chemical removal of understory vegetation, burning, and seeding 

of cheatgrass. Treatments were imposed on each of the four 

sites in Nevada and five sites in Utah. All combinations of 

understory vegetation removal (0 and 100 percent) and fire 

(unburned and burned) treatments were applied to determine 

combined and separate effects of vegetation removal and 

fire on soil water availability. These treatments resulted in 

a potential matrix of vegetation components (table 2). On 

control plots, both the understory and shrub components 

are still present. The herbaceous vegetation removal plots 

only have the shrub component, the burn plots lose the 

shrub component, and the burned plots that have also had 

the herbaceous vegetation removed lose both the shrub and 

understory components. Treatments were applied on three 

replicates of each treatment combination in a randomized 

block design at each site. The cheatgrass seeding treatment 

was applied as a subplot within each treatment combination. 

Each main plot was circular and had an area of 36.30 m2 

(r=1.7m). Understory vegetation removal plots were treated 

with 170.5 ml of Roundup® mixed in 4.5 L of water during 

active spring growth in 2001. The plots were then either left 

unburned or burned in early November 2001. Plots were 

burned in 4-m diameter burn barrels after adding 4.5 kg of 

weed-free straw for consistent fuel loading (Korfmacher and 

others 2003). Subplots 1.5 cm x 0.75 cm were overseeded with 

cheatgrass at a rate of 625 seeds/m2 within each treatment 

plot. Soil moisture sensors were installed inside and outside 

of these seeded subplots in summer 2001.

Soil Moisture

Delmhorst, Inc. GB-1 gypsum blocks were buried at 

three depths, 1 to 3, 13 to 15, and 28 to 30 cm to measure 

soil water matric potential on each unseeded and seeded 

subplot within each vegetation removal-burn treatment plot 

combination. Measurements were taken every minute and 

hourly averages recorded by CR-10 and CR10X Campbell 

Scientific, Inc. microloggers. Resistance data were then 

converted into MPa. Then every hour that had a matric 

potential >–1.5 MPa was summed and divided by 24 to obtain the number of days that water was available. Hourly 

precipitation totals and air temperature averages were also 

recorded by microloggers at each site. Data were collected 

for 2 years after treatment.

Vegetation Sampling

Vegetation was sampled during the summers of 2002 and 

2003. Each treatment plot was sampled with eight, 0.25 m2

quadrats. A modified Duabenmire (1959) cover class method 

with eight cover classes was used to estimate cover. The 

eight cover classes are: (1) 0.01 to 1 percent, (2) 1.1 to 5 percent, (3) 5.1 to 15 percent, (4) 15.1 to 25 percent, (5) 25.2 to 

50 percent, (6) 50.1 to 75 percent, (7) 75.1 to 95 percent, and 

(8) 95.1 to 100 percent. Cover was estimated by species and 

by type of ground cover (bare ground, litter, rock, pavement, 

and cryptogamic crusts). Cover was estimated by selecting 

the smallest cover class in which all of one species or cover 

type would fit. Then, for each species or cover type, the midpoint for that cover class was taken and averaged with the 

midpoints from the seven other quadrat subsamples from 

that plot. Individuals of each species within the 0.25 m2

quadrats were counted to determine density.

Analysis

Mixed model analysis (Littell and others 1996) was used 

to compare vegetation variables and time of available 

water for 1 April through 30 June of 2002 and 2003. Year 

and vegetation removal, fire, and seeding treatments were 

considered fixed variables, while site and replication were 

considered random. Two separate sets of analyses were done 

for both the vegetation data and the available water data. 

Vegetation was then split into three different categories: 

(1)total vegetation, (2) perennial herbaceous vegetation, and 

(3) annual vegetation. The first analysis compared the six 

big sagebrush sites that were along the elevation gradient, 

while the second analysis compared the lowest elevation big 

sagebrush sites in Nevada and Utah and the crested wheatgrass and bottlebrush squirreltail sites. These five lower sites 

were similar in elevation but varied by vegetation. Total 

time of soil water availability was analyzed separately for 

each soil depth. Best linear unbiased predictions (BLUPS) 

were calculated to compare treatment effects by site (Littell 

and others 1996).

Table 2—Presence or absence of vegetation after treatment.

Vegetation Vegetation

Control removal Fire removal + fire

Understory vegetation	 +	 –	 +/– a –

Sagebrush	 +	 +	 –	 –

a Depends on particular species response to fire. Most herbaceous species survive coolseason, low intensity fire.
52 USDA Forest Service Proceedings RMRS-P-52. 2008

Whittaker, Roundy, Chambers, Meyer, Blank, Kitchen, and Korfmacher The Effect of Herbaceous Species Removal, Fire, and Cheatgrass (Bromus tectorum)...

Figure 1—Precipitation measured between 1 October and 

30 June at 9 big sagebrush sites in (A) Nevada and (B) Utah. 

Reese River, Smokey Valley, Fairfield, and Vernon precipitation 

data were obtained from long-term climate stations (Western 

Regional Climate Center).

Results ___________________________

Precipitation

Overall precipitation from 1 October to 30 June increased 

as elevation increased (figs. 1a and 1b). The exception to 

this occurred in 2002 to 2003 on the middle site in Nevada, 

which had higher precipitation than the upper Nevada site. 

Precipitation was lower in 2001 to 2002 than 2002 to 2003 

with the exception of the Nevada crested wheatgrass site, 

which was drier in 2002 to 2003. When comparing the precipitation that fell in these two years to the 30-year average 

from nearby weather stations, 2001 to 2002 was well below 

average and 2002 to 2003 was a little below average.

Figure 2—Vegetation cover measured in 2002 and 2003 at 

six sagebrush sites. Data categories are annual vegetation, 

perennial herbaceous vegetation, and total woody plus 

herbaceous vegetation.

Big Sagebrush Sites

Vegetation—Total vegetation, perennial vegetation, and 

annual vegetation cover all varied significantly (p < 0.01) 

by year and by removal treatment. Total vegetation also 

varied significantly by burn treatment (p < 0.01) and by the 

interaction of year by burn treatment (p < 0.01). Perennial 

vegetation cover also varied with the interaction of year by 

removal treatment (p < 0.01). Vegetation cover was higher 

in 2003 than 2002 for all three vegetation groups (fig. 2). 

Total and perennial vegetation cover was significantly lower 

on removal plots than intact plots. Annual vegetation cover 

however, was significantly higher on removal than intact plots 

(fig. 3). Total vegetation cover was the only variable that was 

significantly affected by burning. In 2002 total vegetation 

cover was similar on unburned and burned plots. In 2003, 

however, the unburned plots had higher total vegetation 

cover than the burned plots (fig. 4).

Figure 3—Cover of annual, perennial, and total woody and 

herbaceous vegetation on plots where herbaceous vegetation 

was left intact or removed at six big sagebrush sites.

A

B

Nevada

350

300

250

200

150

100

50

0

Crested Lower Middle Upper Reese R. Smokey V.

Precipitation (mm)

2001-2002 2002-2003 30-Y Mean

Utah 

Squirreltail Crested Lower Middle Upper Fairfield Vernon

350

300

250

200

150

100

50

0

Precipitation (mm)

Vegetation Cover 25

20

15

10

5

0

Annual Perennial Total

Cover (%)

2002 2003

25

20

15

10

5

0

Cover (%)

Vegetation Cover

Intact Removal

Annual Perennial Total
USDA Forest Service Proceedings RMRS-P-52. 2008 53

The Effect of Herbaceous Species Removal, Fire, and Cheatgrass (Bromus tectorum)... Whittaker, Roundy, Chambers, Meyer, Blank, Kitchen, and Korfmacher

Figure 4—Total vegetation (herbaceous and woody) 

was measured in 2002 and 2003 on plots that were not 

burned or burned at six big sagebrush sites. Treatments 

with means not significantly different (p < 0.05) are labeled 

with the same letter.

Available Water—Total days of available water (estimated using BLUPs) increased as elevation increased 

(fig. 5). This was mainly true for all three depths. The only 

exception was the lowest elevation big sagebrush site in Utah 

had slightly more days of available water than the middle 

sagebrush site at 1 to 3 cm.

At 1 to 3 cm, the only significant (p < 0.01) was year. In 

2002, the total time of available water was 42 days and in 

2003 it was 53 days. No other treatment effects were significant at this depth.

At the 13 to 15 cm depth, total days of available water 

varied significantly by year, removal treatment, burn treatment, and by the interaction of year by removal treatment, 

and year by burn treatment. In 2002, the first growing season after the treatment, the intact plots had 7 fewer days 

of available water than the removal plots (fig. 6a). In 2003 

there was not a significant difference between the removal 

and intact plots. Similar results occurred on the burned and 

unburned plots. The burned plots in 2002 had 6 more days 

of available water than the unburned plots, and in 2003 

there was no significant difference between the burned and 

unburned plots (fig. 6b).

At 28 to 30 cm, the significant (p < 0.01) effects were removal 

and burn. The intact plots had 62 days of available water and 

the removal plots had 68 days of available water. The unburned 

plots had 63 days of available water and the burned plots had 

67 days of available water.

Low Elevation Sites

Vegetation—Total and annual vegetation cover varied 

significantly both by year and removal (p < 0.01). Total 

vegetation cover also varied significantly (p < 0.01) by 

burn treatment. Perennial vegetation cover only varied 

significantly (p < 0.01) by removal. In 2002 both total and 

Figure 5—Total days of available soil moisture (soil matric 

potential > –1.5 MPa) measured at three soil depths between 

1 April and 30 June at six big sagebrush sites.

A

B

Figure 6—Total days of available soil moisture (soil matric 

potential > –1.5 MPa) on six big sagebrush sites at 13–15 cm 

between 1 April and 30 June 2002 and 2003 on plots where 

(A) herbaceous vegetation was left intact or removed and 

(B) vegetation was either not burned or burned. *Treatment 

was only significantly different (p < 0.05) in 2002.

Total Vegetation 35

30

25

20

15

10

5

0 2002 2003

Cover (%)

Unburned

Burned

c

c

a

a

90

80

60

50

40

30

20

10

0

70

Nv Lower Nv Middle Nv Upper Ut Lower Ut Middle Ut Upper

1-3 cm 13-15 cm 28-30 cm

Time of Available Water (days)

70

60

50

40

30

20

10

0 2002* 2003

Intact Removal

Time of Available Water (days)

70

60

50

40

30

20

10

0

Time of Available Water (days)

2002* 2003

Unburned Burned
54 USDA Forest Service Proceedings RMRS-P-52. 2008

Whittaker, Roundy, Chambers, Meyer, Blank, Kitchen, and Korfmacher The Effect of Herbaceous Species Removal, Fire, and Cheatgrass (Bromus tectorum)...

Discussion ________________________

Goldberg (1990) describes how the environment and 

plants both affect resources, which in turn affect associated 

plant growth and development. Soil water is affected by the 

environment through precipitation and evaporation and is 

also affected by the vegetation through transpiration and 

shading. If any of those factors change, soil moisture will 

change also. The soil water then affects the vegetation that 

is still present. The input of water through precipitation 

was much higher in 2003 than 2002, which resulted in an 

increase in both available water and vegetation cover. As 

elevation increased so did precipitation and the time of 

available water. When vegetation cover was decreased by 

disturbance, soil water was available for a longer period, 

annual vegetation cover was lower than 2003; however, 

perennial cover didn’t differ significantly between years 

(fig. 7). Total and perennial vegetation cover was significantly higher on intact plots than removal plots, but annual 

vegetation cover was significantly lower on the intact than 

removal plots. The burn treatment only had a significant 

effect on the total vegetation, lowering vegetation cover 

on the burned plots by 3 percent. The unburned plots had 

a total vegetation cover of 10.4 percent and the burned 

plots had vegetation cover of 7.5 percent.

Available Water—The total days of available water for 

each of the low elevation sites differed little among sites 

(fig. 8). The total days of available water on the squirreltail 

site (at 1 to 3 cm) is the only measurement that was significantly higher than the others.

At 1 to 3 cm, the total days of available water were significantly different by year and by removal treatment. In 

2002, the low elevation sites had 23 fewer days of available 

water than in 2003 (fig. 9). The total time of available water 

was also significantly shorter on intact plots compared to 

removal plots. Intact plots had 38 days of available water, 

whereas the vegetation removal plots had 40 days of available water.

Total days of available water at the 13 to 15 cm depth 

varied significantly by year, removal treatment, burn treatment, and by the interaction of year by removal treatment, 

and year by burn treatment. In 2002, intact plots had 10 

fewer days of available water than removal plots; whereas, 

in 2003 the difference between intact and removal plots 

was not significant. Likewise, in 2002 the unburned plots 

had 6 fewer days of available water than the burned plots 

and there was no significant difference between the two 

treatments in 2003.

At the third depth, 28 to 30 cm, only year and removal 

were significant (p < 0.01) for total days of available water. 

There were 11 more days of available water in 2003 than in 

2002 (fig. 9). The intact plots had 52 days of available water 

and the vegetation removal plots had 63 days of available 

water.

Figure 9—Total days of available soil moisture (soil matric 

potential > –1.5 MPa) measured at three soil depths in 2002 

and 2003 between 1 April and 30 June at five Wyoming big 

sagebrush zone sites. *Depth with means significantly different 

(p < 0.05)

Figure 7—Annual, perennial, and total (herbaceous and woody) 

vegetation cover measured in 2002 and 2003 at five Wyoming 

big sagebrush zone sites.

Figure 8—Total days of available soil moisture (soil matric 

potential > –1.5 MPa) measured at three soil depths between 

1 April and 30 June at five Wyoming big sagebrush zone 

sites.

14

12

10

8

6

4

2

0

Annual Perennial Total

Cover (%)

2002 2003

Vegetation Cover

70

60

50

40

30

20

10

0 Crested

Nevada

Crested

Utah

Lower Lower Squirreltail

1-3 cm 13-15 cm 28-30 cm

Time of Available Water (days)

70

60

50

40

30

20

10

0

1-3 cm 13-15 cm 28-30 cm

2002 2003

Available Water (days)
USDA Forest Service Proceedings RMRS-P-52. 2008 55

The Effect of Herbaceous Species Removal, Fire, and Cheatgrass (Bromus tectorum)... Whittaker, Roundy, Chambers, Meyer, Blank, Kitchen, and Korfmacher

Removal and burn treatments had significant effects on 

soil moisture only during the first year after disturbance. 

However, disturbance resulted in lower vegetation cover 

both years after treatment. Both seasonal precipitation and 

vegetation cover were higher in 2003 than 2002. By 2003, 

understory vegetation that had been burned had grown 

back, shrubs that had not been burned were still present, 

and cover of annual grasses and forbs was higher on removal 

plots. The response of the vegetation to the treatments and 

precipitation in 2003 resulted in only a short-term effect of 

disturbance on time of available soil moisture.

In this study cheatgrass did not affect the time of available water. One possible reason for this is that cheatgrass 

was not seeded until late November 2001 and so cheatgrass 

establishment was relatively low especially on the Utah 

plots. The cheatgrass that did establish did not germinate 

until spring. Fall germination of cheatgrass might have had 

a greater affect on spring soil moisture by allowing greater 

development of roots and increased transpiration. Melgoza 

and others (1990) conducted a large-scale study that examined the effect of cheatgrass on soil moisture and found 

that the presence of cheatgrass did decrease the amount of 

soil moisture. Cheatgrass competition for soil moisture is a 

major factor in the failure of establishment for many desirable grasses, forbs, and shrubs (Young and others 1987). If 

germinating species do not have a long enough wet period 

for emergence and root elongation, seedlings will not survive 

(Abbot and Roundy 2003).

In this study, disturbance evidently decreased the amount 

of resource use resulting in a net increase in time of available soil moisture. As observed by Chambers and others 

(2007), the net increase in resources resulted in an increase 

of cheatgrass establishment and seed production. It also 

increased the cover of other annual grasses and forbs. The 

effects of disturbance on soil water resources in our study 

were highly dependent on seasonal precipitation inputs. 

Seasonal precipitation and weather conditions may be more 

important to cheatgrass invasion than disturbance. In our 

study annual weather fluctuations affected the time of water 

availability in the surface soil where germination occurs more 

than did disturbance. Roundy and others (in progress) found 

that sufficient wet thermal time accumulated for successful 

cheatgrass germination on these sites in spring and often in 

fall, with or without disturbance. The effect of disturbance 

and annual fluctuation on soil moisture in our study was only 

temporary, but it left the community vulnerable to invasion 

as resources fluctuated and increased.

Acknowledgments__________________

This project was funded by the Joint Fire Science Program 

and Federal Aide Pittman-Robertson Wildlife Restoration 

Project W-82-R.

References ________________________

Abbott, L.B.; Roundy, B.A. 2003. Available water influences field 

germination and recruitment of seeded grasses. Journal of Range 

Management. 56: 56-64.

Beckstead, J.; Augspurger, C.K. 2004. An experimental test of 

resistance to cheatgrass invasion: limiting resources at different 

life stages. Biological Invasions. 6:417-432.

probably due to decreased transpiration. Schlesinger and 

others (1987) found plots with vegetation harvested had 

higher available soil water than plots of intact vegetation. 

Coronato and Bertiller (1996) examined effects of grazing 

on soil moisture and found that soil moisture did not change 

significantly. They concluded that the loss of vegetation 

decreased transpiration but also increased evaporation, 

resulting in no net change in soil moisture.

Davis and others (1998) hypothesized that it is the net 

change in resources that affects the competition and the 

invasibility of a community. Foster and Dickson (2004) found 

that increasing the available water made the community 

more invasible. Increased resources over time increase the 

opportunity for the invader to establish and set seed (Booth 

2003). In our study, annual vegetation cover was significantly higher in 2003 than 2002 on removal plots compared 

to intact plots. Both year and removal were important to 

increasing time of available water. Chambers and others 

(2007) examined cheatgrass establishment and production 

on these same plots. They found that cheatgrass establishment, biomass, and seed production were higher on removal 

plots and burned plots than on intact plots. Beckstead and 

Augspurger (2004) did a study where they removed vegetation cover in a shadscale (Atriplex confertifolia)-bunchgrass 

community and found that the vegetation removal plots had 

higher cheatgrass density and biomass.

At the 1 to 3 cm depth, the removal and fire treatments did 

not have a significant effect on the time of available water, 

however the variation by year was significant. In 2003, there 

were 12 more days of available water than 2002. During this 

same year, the burn treatment and the vegetation removal 

treatment only caused a 2-day increase in available water. 

Roundy and others (in progress) found that variation in 

annual weather affected wet thermal-time for cheatgrass 

germination more than vegetation removal or fire on these 

same plots. At the soil surface, seasonal precipitation had 

a greater effect on the time of available water than did 

disturbance treatments. Soil water inputs and outputs by 

precipitation and evaporation would be expected to have a 

greater effect on surface soil moisture than the presence or 

absence of vegetation as affected by disturbance.

Subsurface soil moisture is lost more slowly than that 

at the surface, with transpiration being the major process 

of water loss (Schwinning and Ehleringer 2001). At these 

lower depths, the effects of vegetation loss and fire were 

more evident. At the 13 to 15 cm depth, removal plots had 

7 more days of available water than intact plots, while in 

2003 there were only 3 more days of available water than 

in 2002. At the 28 to 30 cm depth, removal plots had 7 more 

days of available water than intact plots, burned plots had 4 

more days of available water than unburned plots, and time 

of available water did not vary by year. Increase in surface 

soil moisture by increased precipitation provides more opportunity for cheatgrass to germinate, and the increase in 

subsurface soil moisture after disturbance allows cheatgrass 

to establish and set seed. Meyer and others (2001) found that 

irrigating near shrubs, where the ground is shaded, resulted 

in higher cheatgrass plant biomass than near non-irrigated 

shrubs. This provided the cheatgrass that germinated enough 

soil moisture to establish and set seed similar to increasing 

water through disturbance.
56 USDA Forest Service Proceedings RMRS-P-52. 2008

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USDA Forest Service Proceedings RMRS-P-52. 2008 57

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas 

A.; Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Jennifer L. Vollmer and Joseph G. Vollmer are Environmental Resource Specialist, and Market Development Specialist, 

respectively, BASF Corporation, Laramie, WY; email: jennifer.

vollmer@basf.com.

Controlling Cheatgrass in Winter Range 

to Restore Habitat and Endemic Fire

Jennifer L. Vollmer and Joseph G. Vollmer

Abstract—Habitat managers can better prepare a program for 

prescribed burns, wildfire management, and maximum forage 

biomass by understanding the response of key shrubs to the tools 

utilized to reduce cheatgrass (Bromus spp.) competition. Application 

of Plateau® herbicide, prior to annual brome germination, at rates 

up to 8 oz/acre with or without surfactant gave acceptable annual 

brome reduction without adverse injury to true mountain mahogany 

(Cercocarpus montanus), antelope bitterbrush (Purshia tridentata.), 

and sagebrush (Artemisia spp.). By removing cheatgrass prior to or 

after a burn, shrub winter range species can more efficiently and 

vigorously re-occupy the site or survive the burn.

Introduction_______________________

Fine fire fuel management programs are being implemented by Game and Fish to reduce loss of big game winter 

range. Prescribed fire has been an important tool to regenerate brush, improving winter browse. However, severe drought 

during the late 1990s and early 2000s have favored annual 

brome species (Bromus spp.) and allowed this invasive weed 

to influence burn area recovery. Wyoming critical winter 

habitat areas have had an increase of annual brome, such 

as cheatgrass (Bromus tectorum), after wildfire, resulting in 

decreased desirable vegetation including grass, shrubs and 

forbs. These results have alerted Wyoming Game and Fish 

to evaluate areas prior to a prescribed burn to determine 

if an annual brome component is present and if likelihood 

of habitat degradation, rather than improvement, may occur. If the site has potential of degradation due to annual 

brome release, Plateau® herbicide, imazapic, is incorporated 

into the winter range improvement plan for pre-emergence 

control of cheatgrass.

Increasing use of the herbicide Plateau, imazapic, for 

selective control of annual brome in Western wildlands has 

dictated the need for increased knowledge of tolerant brush 

species. Western bunchgrass and forb tolerance trials have 

shown Plateau to be an acceptable tool for release of desirable plant species and renovation of annual brome infested 

areas1 (Foy 2003, Rayda 2003). In general, Plateau is not 

effective at control of brush; however, some brush species 

exhibit unacceptable injury. Brush tolerance to Plateau is key 

when considering use of this herbicide for selective control of 

annual brome prior to a prescribed burn for critical winter 

range brush regeneration.

In addition, Plateau is gaining recognition and use as a 

tool to produce aesthetically acceptable fuel breaks and green 

strips. Plateau can selectively remove the fine annual brome 

fuel from more fire resistant bunch grasses and shrubs. Removal of the annual brome helps eliminate an ignition fuel 

as well as eliminating the main fire carrier. Fire modeling 

of Plateau treated areas utilizing the BehavePlus fire 

model has shown significant reduction of flammable biomass 

as well as decreasing flame height and length (Kury 2003). 

Applications of Plateau are typically broadcast, applied over 

the top of brush remaining in the green strip for aesthetic, 

moisture catching, or soil stabilizing purposes. Brush tolerance is an important aspect when considering the use of 

Plateau for enhancing green strips and fuel breaks, as well 

as an additional tool for habitat improvement.

Tolerance Mechanism

Plateau herbicide, imazapic, is a member of the imidazolinone family. The active ingredient of an imidazolinone 

herbicide controls susceptible plants by binding to the acetohydroxyacid synthase (AHAS) enzyme and preventing 

production of three essential amino acids. Plant tolerance 

to imidazolinones can be due to inherent differences in the 

AHAS enzyme itself and/or differences in the stability of the 

enzyme. For some species, such as legumes, tolerance to imidazolinones is contributed to their ability to metabolize the 

herbicide active ingredient. Mature tissues in plants appear 

to be relatively unaffected by inhibition of the AHAS enzyme 

(Shaner 1991). This accounts for the higher susceptibility 

of annual versus perennial plants, since perennial plants 

1Data on file with BASF Corporation, J.L. Vollmer, Laramie, 

WY, as “GRASS AND FORB TOLERANCE TO PLATEAU® 

® Registered Trademark of BASF Corporation HERBICIDE – Update July 26, 2006.
58 USDA Forest Service Proceedings RMRS-P-52. 2008

Vollmer and Vollmer Controlling Cheatgrass in Winter Range to Restore Habitat and Endemic Fire

would have a higher percentage of mature tissue. After direct 

treatment with an imidazolinone herbicide, mature leaves 

of perennial susceptible plants will remain green for a long 

period of time, several months, prior to desiccation. Leaves 

continue to photosynthesis, although amino acid production 

is arrested. In treated susceptible species, photosynthesis 

translocation can be disrupted, depriving roots of an energy 

supply (Shaner 1991). Susceptibility of well-established 

shrubs may take up to two years to determine.

Results and Discussion______________

True Mountain Mahogany Tolerance

True mountain mahogany (Cercocarpus montanus) trials 

were conducted on a post burn site in Douglas, WY. At one 

year after a wildfire, further loss of mountain mahogany was 

threatened by competition and additional fine fuel buildup of 

cheatgrass, tumble mustard (Sisymbrium altissimum) and 

thistle (Cirsium spp.) invasion. Plateau treatments were 

broadcast applied September 4, 2002, prior to cheatgrass 

emergence. The trial had seven treatments: 6, 9 and 12 oz 

of Plateau per acre, with and without methylated seed oil 

(MSO) surfactant at 1 qt/acre, compared to a non-treated 

plot. Plot size included seven to 10 bushes in a 10 x 50 foot 

area replicated three times. The same treatments were conducted on an adjacent area in spring 2003. Treatment goals 

were to reduce the fine fuel load to prevent further loss of 

the remaining mountain mahogany population in the event 

of a wildfire. Data were to be used to aid in plans to prepare 

similar sites for a prescribed burn.

The spring after application, all fall 2002 treatments 

showed delayed leaf expansion with some yellowing during 

the first growing season. This response increased as the 

Plateau rate increased, with a greater negative response from 

treatments that included the MSO surfactant. All plots were 

evaluated the first full growing season after applications on 

August 10, 2004. Fall treated plots at the typical cheatgrass 

recommended application rates of Plateau at 6 or 9 oz/acre 

and Plateau at 6 oz + MSO had no adverse effects on the 

true mountain mahogany. Addition of MSO to the Plateau 

at 9 oz/acre simulated a burn response by stimulating new 

shoot growth from the base of the plant. The Plateau at 9 oz/acre 

plus MSO treatment could be used to simulate a prescribed 

burn for habitat enhancement when the cheatgrass population is high enough to threaten mahogany mortality during 

a wildfire or prescribed burn.

Spring applications showed greater variability in response 

between plants within a treatment. Plateau at 6 oz/acre was 

the only treatment resulting in mahogany growth similar to 

the non-treated plot. Plateau at 12 oz/acre + MSO applied in 

the spring resulted in initial unacceptable stunting of new 

growth. The observed stunting was viewed as unacceptable 

due to the decreased amount of vegetation that could be 

utilized as browse.

New basal growth was evaluated for all treatments as 

an important source of mahogany regeneration. New basal 

growth was acceptable for all fall treatments except the 

Plateau 12 oz/acre + MSO. Results suggest that this treatment affected the overall plant system, inhibiting the ability 

of the plant to recover. For spring-applied treatments, new 

basal growth, at 16 months after treatment, was not affected 

by Plateau at the 6 or 12 oz rate or MSO plus Plateau at 9 or 

12 oz rate. All other treatments had individual plants that 

elicited variable shoot growth. Differences in basal growth 

may be due to individual plant genetics or microclimate 

including soil type and/or depth.

Fall pre-emergence Plateau treatments without surfactant 

provided the needed cheatgrass control. The addition of 

MSO was not needed to achieve adequate cheatgrass control 

to reduce competition and fire hazard. Spring treatments 

required the addition of MSO to control the cheatgrass post 

emergence, but spring treatment is not recommended due to 

variation in brush response and annual brome efficacy.

Antelope Bitterbrush Tolerance

The antelope bitterbrush (Purshia tridentate) tolerance 

research was conducted on a site prior to a prescribed burn, 

at 8,000 feet east of Laramie, WY. Plateau treatments were 

broadcast applied September 30, 2003, prior to cheatgrass 

emergence. The trial had nine treatments: Plateau at 6, 8, 10 

and 12 oz/acre, Plateau at 6 and 8 oz/acre plus non-ionic surfactant (NIS) at 0.25 percent v/v, and Plateau at 10 and 12 oz/acre 

with MSO surfactant at 1 qt/acre, all compared to a non-treated 

plot. Plot size included 10 bitterbrush plants in a 10 x 50 foot 

area replicated three times. Treatment goals were to prevent 

cheatgrass domination after a prescribed burn.

At the beginning of the first growing season after application, June 24, 2004, bitterbrush showed little response from 

most treatments. Exceptions were Plateau at 12 oz/acre 

alone and Plateau 10 and 12 oz/acre plus MSO. The response 

elicited by these herbicide treatments was a delay in leaf 

expansion, smaller mature leaves and shortened internodes of 

new stems (typical imidazolinone symptomology). First year 

results indicated that later ratings were needed to evaluate 

long-term herbicide effect on bitterbrush.

At 2.5 years after treatment, May 23, 2006, bitterbrush 

mortality was evaluated. The 6 and 8 oz/acre rates of Plateau with and without surfactant had no mortality and no 

evidence of treatment effect (table 1). The two high rates 

of Plateau with surfactant resulted in 28 to 43 percent mortality. Surviving plants had no new stem growth and spring 

leaf growth was severely delayed with the few new leaves 

displaying typical imidazolinone symptomology, indicating 

these plants were still under severe stress from the Plateau 

herbicide. The second and third replications of 10 and 

12 oz/acre rate of Plateau without surfactant, showed little 

to no injury on recovered plants. However, the first replication, located on a drought prone slope, had mortality of 

80 percent for both treatments. This response to adverse 

environmental factors in combination with a Plateau treatment indicates marginal tolerance of bitterbrush to Plateau 

at high rates. Of the recovering plants in the second and 

third replications, first year growth after application was 

a fifth of the second year growth. Wildlife managers would 

need to assess bitterbrush recovery potential and browsing demands on these plants to determine if high rates of 

Plateau were acceptable for their program goals. Rates of 

10 and 12 oz of Plateau per acre is rarely needed to achieve 

acceptable cheatgrass control, allowing managers to adjust 

rates to achieve antelope bitterbrush selectivity.
USDA Forest Service Proceedings RMRS-P-52. 2008 59

Controlling Cheatgrass in Winter Range to Restore Habitat and Endemic Fire Vollmer and Vollmer

Sagebrush Species Tolerance

Sagebrush (Artemisia spp.) steppe communities have been 

severely impacted by fire carried by annual brome species. 

Restoration of fire-scarred land can be unsuccessful due 

to competition from annual brome; therefore, a selective 

herbicide that can be used to preserve remaining sagebrush 

steppes as well as aid in restoration is very important. 

Plateau herbicide has been applied over the top of several 

sage species through research and commercial applications. 

Table 2 is a summary of tolerance research results and 

commercial observations made across the Western United 

States sage steppe areas.

Sagebrush Case Study—The Johnson Creek Unit of 

Sybille Canyon, WY, suffered the loss of critical bighorn 

sheep winter habitat in August of 2001. An escaped campfire 

resulted in a 448-acre wildfire. During the fall of 2001 cheatgrass dominated the area, out competing the native vegetation. A rescue/release treatment of Plateau at 8 oz/acre plus 

MSO was applied in August of 2002. Prior to treatment, 

100 foot transects were installed on the Wyoming Game 

and Fish treated area and on an adjacent non-herbicidetreated, burned Bureau of Land Management section. In 

2003, post application, belt density transects and nested 

frequency quadrants were added. Measuring relative 

cover at one year after treatment, cheatgrass increased 

by 8 percent in the non-treated area to 75 percent, while 

native vegetation decreased a corresponding amount to 

25 percent of the cover (table 3). In the Plateau treated 

area, cheatgrass decreased from 84 percent to 0 percent 

with a corresponding increase in native vegetation to 

100 percent.

Table 2—Tolerance summary of sage species to Plateau herbicide at 2 to 12 oz/acre with or without MSO.

Silver sagebrush (A. cana)a No injury

Fringed sagebrush (A frigida)a No injury, new growth greater than in non-treated areas, 

 possibly due to elimination of annual brome competition

Wyoming big sage (A. tridentate)

spring applied No injury

 fall applied No injury, new leader growth often increased compared 

 to non-treated areas, possibly due to elimination of 

 annual brome competition

Seedling Wyoming Big Sagea No injury

a Fall applied Plateau herbicide treatment.

Table 1—Antelope Bitterbrush tolerance, leaf expansion and mortality.

 Evaluation June 24, 2004 Evaluation May 23, 2006

Treatment % Injury by leaf expansion reductiona % Mortalitya

Plateau 6 oz + NIS	 7	 0

Plateau 8 oz + NIS	 8	 0

Plateau 10 oz + MSO	 90	 28

Plateau 12 oz + MSO	 90	 43

Plateau 6 oz 0 0

Plateau 8 oz 0 0

Plateau 10 oz 0 27b

Plateau 12 oz 22 33b

Non-treated 0 0

a Average over three replications.

b First replication, located on slope with drought tendency had 80% mortality.

Table 3—Percent relative cover of the Johnson Creek Unit at Sybille Canyon, WYa.

Bromus tectorum Native flora

Treatment Pre-treatment Post treatment Pre-treatment Post treatment

Non-treated area 67% 75% 33% 25%

Plateau 8 oz/acre + MSO	 84%	 0%	 16%	 100%

aPreliminary data compiled by Wyoming Game and Fish (data on file with the Wyoming Game and Fish, 

Casper, WY).
60 USDA Forest Service Proceedings RMRS-P-52. 2008

Vollmer and Vollmer Controlling Cheatgrass in Winter Range to Restore Habitat and Endemic Fire

Conclusion________________________

Fall Plateau treatment, prior to cheatgrass emergence, 

remains the best program for true mountain mahogany 

tolerance in addition to best cheatgrass control at the lowest 

herbicide rates. Results indicated that an alternative to fire 

for mahogany regeneration is Plateau at 9 oz/acre plus MSO. 

This treatment would give wildlife managers a treatment 

option when annual brome populations prohibit burning due 

to the increased fire temperatures threatening mahogany 

survival. A cheatgrass control program in an antelope bitterbrush community should not exceed Plateau at 8 oz/acre 

with or without surfactant. Higher rates can increase the 

possibility of unacceptable injury to bitterbrush. Artemisia

spp. exhibited the greatest tolerance with no negative treatment response to the highest label rate of Plateau with or 

without surfactant.

Plateau has proven to be an effective fire mitigation, release 

and restoration tool for grass/shrub landscapes. The selective ability of the product gives wildlife and land managers 

options for improving shrub communities. Specific species 

tolerance to Plateau is important when choosing rate, timing, and additive.

Acknowledgments__________________

BASF wishes to acknowledge and thank Ryan Amundson 

and Keith Schoup, biologists for Wyoming Game and Fish, 

for bringing the research need of Plateau brush tolerance to 

our attention, and helping us determine injury acceptability 

limits dependant on anticipated wildlife utilization.

References ________________________

Foy, Lise P.; Whitson, Thomas D.; Stevens, Bart W. 2003. Establishment of native wildflowers and grasses for roadside re-vegetation 

using imazapic. Western Society of Weed Science Proceedings. 

56: 25.

Kury, Brenda; Vollmer, Jennifer L. 2003. Data collection and fire 

modeling determine the potential for the use of Plateau ® herbicide 

to establish fuel breaks in Bromus tectorum-dominated rangelands. 

Western Society of Weed Science Proceedings. 56: 24.

Rayda, Margaret M.; Miller, Stephen D. 2003. Effect of imazapic on 

crop, weed and native legumes. Western Society of Weed Science 

Proceedings. 56: 77.

Shaner, Dale L.; O’Conner, Susan L., eds. 1991. The imidazolinone 

herbicides. Boca Raton, FL: CRC Press, Inc. 135 p.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein
USDA Forest Service Proceedings RMRS-P-52. 2008 61

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Susan E. Meyer is Research Ecologist, David L. Nelson is 

Research Plant Pathologist (Retired), and Suzette Clement is 

Microbiological Technician at the USFS Rocky Mountain Research 

Station, Shrub Sciences Laboratory, Provo, Utah; e-mail: smeyer@

fs.fed.us

Julie Beckstead is Associate Professor in the Department of 

Biology, Gonzaga University, Spokane, Washington.

Abstract—Cheatgrass (Bromus tectorum) is an exotic winter 

annual grass weed that has invaded millions of hectares in the 

Intermountain West. Restoration of cheatgrass-invaded wildlands 

is generally impractical without some form of cheatgrass control. 

We are investigating the possibility of manipulating indigenous 

fungal pathogens that already occur on cheatgrass for short-term 

biocontrol in conjunction with restoration seedings. Three potential 

biocontrol organisms have been identified. The head smut pathogen 

(Ustilago bullata) and the chestnut bunt pathogen (Tilletia fusca) 

infect at the seedling stage and prevent seed set, while the blackfingers-of-death pathogen (Pyrenophora semeniperda) kills seeds in 

the seed bank. Both head smut and chestnut bunt pathogen races 

on cheatgrass are host-specific, whereas black-fingers-of-death is a 

generalist grass seed pathogen that does not appear to form hostspecific races. Inoculation trials with the head smut pathogen yielded 

high levels of disease only when seedlings emerged at moderate 

temperatures in fall, whereas the chestnut bunt pathogen infects 

at near-freezing winter temperatures but requires persistent snow 

cover for successful infection. The black-fingers-of-death pathogen 

is most effective at destroying seeds in the carryover seed bank. A 

combined approach using all three pathogens shows some promise 

for biocontrol of this troublesome weed.

Introduction_______________________

The invasion of cheatgrass (Bromus tectorum) into the Intermountain West has been called the most significant plant 

invasion in the modern history of North America (D’Antonio 

and Vitousek 1992). Cheatgrass creates the disturbance it 

needs to perpetuate itself by producing a continuous, fine 

fuel that is associated with increases in the frequency and 

size of wildfires (Whisenant 1990). With repeated burning, 

vast areas are converted to cheatgrass monocultures that 

are extremely difficult to restore to native plant communities 

or even to rehabilitate with introduced forage grasses. For 

Cheatgrass (Bromus tectorum) Biocontrol 

Using Indigenous Fungal Pathogens

Susan E. Meyer, David L. Nelson, Suzette Clement, and Julie Beckstead

a seeding to be successful, some form of cheatgrass control 

is necessary, especially in the arid and semiarid habitats 

where this plant is most problematic (Monsen 1994).

Traditional methods of cheatgrass control include early 

season burning to prevent seed dispersal, tillage after 

emergence in the fall, and the use of various herbicides. 

These methods tend to be risky, expensive, or damaging 

to remnant perennial vegetation, as well as resulting in 

unpredictable levels of control. We have initiated research 

on possible biocontrol organisms for cheatgrass because of 

the need for targeted, environmentally benign, and effective 

control methods to be used in conjunction with restoration 

seeding. At least one biocontrol organism has already been 

patented for use against cheatgrass in winter cereal crops 

(Kennedy and others 1991, 2001). This organism is a specific rhizobacterial strain that targets the cheatgrass root 

system and greatly reduces its growth, thereby reducing 

grain yield losses to this weed. There is a good possibility 

that this biocontrol agent could also be used on wildlands, 

possibly in conjunction with the biocontrol agents that are 

currently under study at our laboratory (Ann Kennedy, 

personal communication).

In our cheatgrass biocontrol studies, we are investigating three fungal pathogens that target different stages of 

the life history of cheatgrass, but each of which has the net 

effect of reducing the size of the seed bank (fig. 1). Two of 

these pathogens (Ustilago bullata and Tilletia fusca) are 

smut fungi that cause systemic diseases resulting in prevention of seed production in infected plants, while the third 

(Pyrenophora semeniperda) is an ascomycete fungus that 

kills seeds directly in the seed bank.

The Head Smut Pathogen 

(Ustilago Bullata) __________________

We began our biocontrol investigations with work on the 

head smut pathogen U. bullata. This organism has a wide 

host range, infecting several genera of cool season grasses, 

but Fischer (1940) determined that this pathogen is characterized by a high degree of host specificity, with different 

races infecting different grass host species. This has been 

confirmed in our own studies; races of the pathogen from 

cheatgrass are unable to cause significant levels of disease 

on any of the native perennial grass or introduced forage 

grass species that commonly co-occur with cheatgrass (Meyer 

and others, in review).

Almost every cheatgrass population contains endemic 

levels of head smut disease, and the disease can sometimes 

reach epidemic levels. The focus of our research has been 
62 USDA Forest Service Proceedings RMRS-P-52. 2008

Meyer, Nelson, Clement, Beckstead Cheatgrass (Bromus tectorum) Biocontrol Using Indigenous Fungal Pathogens

to study the genetic and environmental factors that control 

disease levels in natural populations. We wanted to know 

how to cause head smut epidemics through artificial inoculation, so that seed production would be drastically reduced, 

facilitating the success of a restoration seeding. Our studies 

on the genetics of the cheatgrass-head smut pathosystem 

have revealed a complex array of resistance phenotypes in 

cheatgrass populations and corresponding virulence races in 

co-occurring populations of the pathogen (Meyer and others 

2001, 2005). But these patterns of co-evolutionary response 

control disease levels only when environmental factors are 

highly conducive to disease development (Meyer and others, 

in review). Otherwise, many susceptible host individuals 

fail to develop the disease even in the presence of pathogen 

races that can successfully attack them. In order to create 

epidemics artificially, we have to make sure that our inoculum includes all the necessary pathogen races, but we also 

have to understand how environmental factors influence 

disease development.

The life cycle of the head smut pathogen starts with germination of diploid teliospores to produce haploid gametes 

through meiosis. These gametes, called sporidia, are capable 

of saprophytic proliferation in the yeast-like haploid state, 

in effect making many hundreds of copies of themselves, 

increasing the chances of encountering an infection site. 

When sporidia of opposite mating types fuse, they form a 

dikaryotic infection hypha. This hypha is not saprophytic; 

it has only a short time to encounter an infection site before 

it spends its limited resources. If penetration at an infection 

site on a young cheatgrass seedling coleoptile successfully 

occurs, then the fungus grows systemically inside its host, 

overwintering in vegetative tissues and growing upward 

during bolting in spring to take over the floral meristems 

for teliospore production.

Because the head smut pathogen infects at the seedling 

stage, conditions during seed germination and seedling 

emergence mediate infection levels. As with many biological 

processes, temperature is an important factor in infection 

success. We made a detailed study of the effect of temperature 

on different phases of the infection process, including eight 

different pathogen populations (Boguena and others 2007). 

Teliospore germination rate increased with temperature, 

as did cheatgrass seed germination rate. At temperatures 

of 10–25 °C, teliospore germination was faster than seed 

germination, while at a cold temperature, below 5 °C, teliospore germination lagged behind seed germination (fig. 2a). 

Sporidial proliferation rate was also strongly influenced by 

temperature, with an exponential increase in rate as a function of temperature over the range 2–25 °C (fig. 2b). One 

outcome of these temperature relationships is that disease 

incidence was drastically reduced at low temperatures in 

growth chamber inoculation trials (fig. 2c). Not only did 

teliospores germinate very slowly at 2 °C, they germinated 

directly to the dikaryotic state, greatly reducing chances 

of successful infection and precluding survival as sporidia 

until temperatures became more favorable (Boguena and 

others 2007).

Under field conditions, these temperature effects translate 

to successful infection by this pathogen only when cheatgrass 

seeds germinate at moderate temperatures in the fall. When 

inoculated seeds are planted in early fall, disease levels can 

be very high, but when inoculated seeds are prevented from 

fall-emerging by sowing on a late planting date, disease 

levels are very low (fig. 3). This is also the pattern we have 

observed with naturally occurring levels of this disease. 

Epidemic levels (>50 percent disease incidence) are rarely 

seen except in mesic environments with predictable autumn 

precipitation (Meyer and Smith, unpublished report on 

file at the USFS Shrub Sciences Laboratory, Provo, Utah). 

This will limit effective use of the head smut pathogen as a 

biocontrol agent for cheatgrass to environments with reliable fall precipitation and near-complete fall germination 

of cheatgrass seeds.

For commercial development of a biocontrol organism, 

it is necessary to find a way to produce inoculum in bulk, 

preferably in vitro, for example, as yeast is produced in 

industrial facilities. For an obligate biotroph like the head 

smut pathogen, this presents problems, because direct 

production of teliospores would be possible only through 

‘farming’ cheatgrass for the teliospore crop. Fortunately, 

the saprophytic sporidial stage of the life cycle is more amenable to mass production techniques. We have developed the 

technology to produce sporidial inoculum and to dehydrate 

it onto a carrier that can be used in a field setting, though 

Figure 1—Schematic diagram showing how three proposed 

cheatgrass biocontrol organisms affect the target host at 

different stages of its life cycle to negatively impact the seed 

bank. Ustilago bullata infects fall-germinating cheatgrass 

seedlings at relatively warm temperatures, grows systemically in 

the plant and prevents seed set. Tilletia fusca infects cheatgrass 

seedlings that emerge at cold temperatures in winter and early 

spring. It also grows systemically in the plant and prevents seed 

set. Pyrenophora semeniperda attacks and kills ungerminated 

cheatgrass seeds in the seed bank. 
USDA Forest Service Proceedings RMRS-P-52. 2008 63

Cheatgrass (Bromus tectorum) Biocontrol Using Indigenous Fungal Pathogens Meyer, Nelson, Clement, Beckstead

our studies on inoculum production are still in the preliminary stages. We have obtained high infection using liquid 

sporidial inoculum, but we still need to optimize dehydration 

and storage protocols to obtain these high infection levels 

with dehydrated sporidial inoculum.

The Chestnut Bunt Pathogen 

(Tilletia Fusca) _____________________

The discovery of environmental limitations on the use 

of the head smut pathogen for biocontrol of cheatgrass 

motivated us to consider other pathogens that might act in 

complementary roles. The chestnut bunt pathogen is related 

to the head smut pathogen and has a similar life history, 

infecting at the seedling stage, growing systemically, and 

preempting the floral structures for teliospore production. 

Figure 2—The effect of temperature on: (A) Germination rate of 

Bromus tectorum (host) seeds and Ustilago bullata (pathogen) 

teliospores, (B) Proliferation rate of Ustilago bullata sporidial 

cells in liquid culture, and (C) Ustilago bullata disease incidence 

on Bromus tectorum plants after inoculation at two densities 

at a range of temperatures. Data represent the means of 8 

pathogen populations (from Boguena and others 2007).

Figure 3—Disease incidence on cheatgrass plants after seeds 

were inoculated with teliospores from eight different populations 

of the head smut pathogen and planted early in the fall (midSeptember) or late in the fall (mid-November). Within the early 

planting, bars headed by the same letter are not significantly 

different at P < 0.05; there were no significant differences among 

pathogen populations within the late fall planting (from Boguena 

and others 2007).

It also forms host-specific pathogen races (Hoffman and 

Meiners 1971). There are some important differences, however (table 1). Teliospores of the chestnut bunt pathogen 

germinate only at cold temperatures (fig. 4) (Meiners and 

Waldher 1959). Teliospores of both pathogens are dormant 

at dispersal and lose dormancy under dry conditions as a 

function of temperature in a manner parallel to dormancy 

loss in seeds of the host (Bauer and others 1998; Meyer and 

Clement, unpublished data). But nondormant teliospores of 

the head smut pathogen germinate very slowly if at all in the 

cold, whereas nondormant teliospores of the chestnut bunt 

pathogen germinate only in the cold, with no germination 

at all at temperatures of 10 °C or higher. This pathogen is 

adapted for infection of emerged cheatgrass coleoptiles under 

snow cover (Meiners 1958).

Another difference between the two organisms is teliospore 

longevity. We have data to suggest that head smut teliospores rarely if ever live for more than a year in the field; 

recolonization through teliospore dispersal from existing 

populations reestablishes populations after local extinction 

events. Teliospores are dispersed by wind after the bullae are 

ruptured by swelling when wet. Chestnut bunt teliospores, 

on the other hand, are adapted for long term persistence as 

a soil spore bank. They are produced in ‘bunts,’ modified 

cheatgrass florets that have no dehiscence mechanism. 

The spikelets containing the bunts are bent to the ground 

with the first winter snows, and the teliospores are released 

slowly from these bunts by weathering over time. We have 

some data to indicate that chestnut bunt teliospores undergo 
64 USDA Forest Service Proceedings RMRS-P-52. 2008

Meyer, Nelson, Clement, Beckstead Cheatgrass (Bromus tectorum) Biocontrol Using Indigenous Fungal Pathogens

cyclic dormancy changes, becoming less dormant as winter 

approaches and more dormant with the onset of warmer 

weather (Meyer and Clement, unpublished data). This would 

also be an adaptation for long term persistence.

Our inoculation trials with the chestnut bunt organism 

have met with limited success. This organism is not as tractable as the head smut pathogen in culture or in inoculation 

trials (table 1). Direct inoculation onto the seeds prior to 

planting does not result in infection; the teliospores must 

be placed on the soil surface in order for infection to occur, 

but conditions at the surface must be carefully controlled, 

for example, through the use of snow-simulating insulating 

materials. In a survey of levels of head smut and chestnut 

bunt disease in cheatgrass populations over a range of habitats, we determined that chestnut bunt disease incidence 

was closely tied to the presence of persistent snow cover in 

winter (Meyer and Smith, unpublished report on file at the 

USFS Shrub Sciences Laboratory, Provo, Utah). Use of this 

organism for cheatgrass biocontrol would thus present many 

of the same environmental constraints as use of the head 

smut pathogen, even though they infect under contrasting 

conditions. This is because mesic habitats, where reliable 

autumn precipitation occurs, are also the habitats that are 

most likely to have persistent snow cover in winter. Xeric 

sites with low autumn precipitation probability also have 

low probability of persistent winter snow cover and negligible levels of chestnut bunt disease. Higher elevation sites 

sometimes have epidemic levels of this disease. But these 

sites are in habitats where cheatgrass is usually not a major 

problem and where natural succession to native perennial 

communities can readily occur. This new information has 

led us to place reduced emphasis on developing the chestnut 

bunt pathogen as a cheatgrass biocontrol organism.

The Black-Fingers-of-Death Pathogen 

(Pyrenophora Semeniperda) _________

The third organism that we have investigated as a possible 

biocontrol agent for cheatgrass is the black-fingers-of-death 

pathogen, Pyrenophora semeniperda, which infects mature 

grass seeds in a wide range of genera (Medd 1992; Medd and 

others 2005). This ascomycete fungus is usually seen as its 

Table 1—A comparison of life history attributes of the head smut pathogen (Ustilago bullata) and the chestnut 

bunt pathogen (Tilletia fusca).

Ustilago bullata Tilletia fusca

Infects at warm temperatures in autumn. Infects at cold temperatures under the snow.

Spores short-lived in soil. Spores form persistent banks in soil.

Infects barely-emerging coleoptile; seed Infects coleoptile after emergence from soil; seed

inoculation effective. inoculation not effective.

Easy to grow in culture. Hard to grow in culture.

Easy to work with in greenhouse. Hard to work with in greenhouse.

Common, occurs at some level in almost every Less common, restricted to populations in places with

cheatgrass population. frequent heavy snow cover.

Figure 4—The effect of storage time, storage 

temperature, and incubation temperature on 

germination of Tilletia fusca teliospores. Data 

represent the means of teliospore collections from 

six populations. 
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Cheatgrass (Bromus tectorum) Biocontrol Using Indigenous Fungal Pathogens Meyer, Nelson, Clement, Beckstead

anamorph (asexual state), Drechslera campanulata, which 

forms characteristic stromatal fruiting bodies that resemble 

black fingers, hence the common name. These fruiting bodies are readily visible and very distinctive, making this seed 

bank pathogen relatively easy to detect and to study. Most 

of the work with this pathogen has centered on nondormant cereal crop seeds (Medd and Campbell 2003). Direct 

inoculation of mature seeds that are nondormant and that 

germinate quickly rarely results in seed mortality, though 

the fungus can sporulate as a weak pathogen on germinated 

seeds. These infected germinated seeds are usually little 

impacted by the pathogen. This led most early workers to 

assume that Pyrenophora semeniperda would always act 

as a weak pathogen, although Kreitlow and Bleak (1964) 

showed conclusively that this pathogen could cause major 

mortality in the transient seed banks of native perennial and 

introduced forage grasses. We have shown that the ability of 

this pathogen to cause cheatgrass seed mortality is directly 

related to seed germination rate (Beckstead and others, in 

press). Nondormant cheatgrass seeds in the transient seed 

bank in autumn can germinate very quickly and are rarely 

killed at naturally occurring inoculum levels, while seeds in 

secondary dormancy in the carryover seed bank can suffer 

high mortality. We demonstrated this clearly in laboratory 

inoculation experiments with cheatgrass seeds that differed 

in dormancy status (fig. 5). Seeds in primary dormancy germinated very slowly if at all, with low percentages even after 

28 days. Seeds that were fully after-ripened germinated to 

high percentages in less than two days, with partially afterripened seeds germinating somewhat more slowly than fully 

after-ripened seeds. When the seeds were inoculated with 

conidia (spores) of the pathogen and incubated at laboratory 

temperature, 100 percent of the dormant seeds eventually 

succumbed, while only 8 percent of the fully after-ripened 

Figure 5—Mortality caused by the pathogen Pyrenophora 

semeniperda in laboratory inoculation experiments with Bromus 

tectorum seeds of differing dormancy status (seeds dormant, 

partially after-ripened, or fully after-ripened). Data represent 

means of two seed collections. Error bars shown are standard 

errors of the mean.

seeds and 13 percent of the partially after-ripened seeds 

were killed by the pathogen. Uninoculated controls showed 

a similar pattern but at much reduced absolute levels, with 

6 percent of the dormant seeds, 2 percent of the partially 

after-ripened seeds, and 0.5 percent of the fully after-ripened 

seeds killed by the pathogen. This mortality was caused by 

conidia that dispersed to the seeds prior to collection in the 

field. All seeds inoculated with pathogen conidia eventually 

exhibited the characteristic black stromata, showing that 

all seeds were infected. But only seeds that exhibited black 

fingers prior to germinating were killed by the pathogen.

The black-fingers-of-death pathogen has been demonstrated to cause high mortality in cheatgrass seed banks 

under natural conditions (Meyer and others 2007). In order to 

evaluate temporal patterns of seed mortality caused by this 

pathogen, we collected monthly seed bank samples during 

winter and spring 2006 at the Whiterocks study site in Skull 

Valley, Utah. The 2004 to 2005 growing season was favorable for seed production, with over 50,000 seeds m-2 on the 

ground at the end of summer. The autumn of 2005 was dry at 

the site, and germination did not occur until late December, 

when a relatively warm winter storm triggered germination 

of 52 percent of the seed bank. Most of the remaining seeds 

entered secondary dormancy by early February. Each month, 

we determined the density of field-killed seeds, the density 

of ungerminated seeds that exhibited stromata during the 

first 7 days of incubation and were assumed to be infected 

prior to collection, and the total density of ungerminated 

seeds that exhibited stromata during the 28-day incubation 

period (fig. 6). In January, the density of field-killed plus 

field-infected seeds was low, but each month during the 

spring these numbers increased, showing that mortality 

was taking place on dormant seeds during the long, wet 

spring. By the end of spring (mid-May), 86 percent of the 

Figure 6—Number of seeds per square meter killed by 

Pyrenophora semeniperda at natural inoculum levels in the 

field: prior to sample collection (day 0), in the first 7 days of 

laboratory incubation, and after 28 days of laboratory incubation. 

Plotted values represent means for twenty samples collected 

at monthly intervals during spring 2006 at the Whiterocks study 

site in Skull Valley, Utah. 
66 USDA Forest Service Proceedings RMRS-P-52. 2008

Meyer, Nelson, Clement, Beckstead Cheatgrass (Bromus tectorum) Biocontrol Using Indigenous Fungal Pathogens

potential carryover seed bank had been infected and killed 

by the pathogen. The carryover seed bank was reduced from 

a potential 48 percent of total seeds present in the autumn 

to only 7 percent. This study shows that this pathogen can 

have a major negative impact on cheatgrass seed banks even 

at naturally occurring inoculum loads.

There are two principal problems associated with the 

use of the black-fingers-of-death pathogen as a cheatgrass 

biocontrol agent. First, in the field its impact seems to be 

limited to causing mortality in the carryover seed bank. 

Because most cheatgrass plants that successfully produce 

seed establish from the transient (current-year) seed bank, 

negative impacts to the carryover seed bank, even if large, 

may have little effect on population dynamics. It is only in 

years when establishment and seed production from the 

transient seed bank fails, due to fire, grasshopper herbivory, 

a head smut epidemic, or some other catastrophe, that the 

fate of the carryover seed bank becomes pivotal. But if 

other forms of control can target the plants produced from 

the transient seed bank, this pathogen could be useful in 

eliminating the carryover seed bank that remains. Very 

few control methods impact ungerminated seeds, and this 

is a potential strength of this particular biocontrol agent. 

Cheatgrass establishment from the carryover seed bank can 

be a substantial impediment to seeding success even when 

control of actively growing plants or of seed production is 

successful.

Another approach to overcoming the obstacle of low 

impact of this pathogen on the transient cheatgrass seed 

bank would be to select for pathogen strains that can kill 

rapidly germinating cheatgrass seeds. There is evidence that 

different strains of this pathogen possess different degrees 

of virulence (Capio and others 2004; Campbell and others 

2003), so the potential for artificial selection for increased 

virulence probably exists. We are currently developing a 

protocol for screening multiple isolates of the pathogen for 

ability to kill nondormant cheatgrass seeds.

A second problem with use of the black-fingers-of -death 

pathogen for cheatgrass biocontrol is the fact that it is a 

generalist pathogen that apparently lacks host-specific 

races (Beckstead, unpublished data). If this organism were 

successfully used for biocontrol of cheatgrass by eliminating the seed bank, the inoculum produced on these killed 

seeds could pose a threat to the seeds of planted species. The 

host range of this pathogen is not completely known, but it 

seems to attack mainly cool season grasses of the Hordeae 

and Festuceae. It may be possible to seed with species that 

have low susceptibility, or to develop fungicidal seed dressings to protect seeded species. The risk to seeded species 

depends on several factors, including the density of target 

seeds and resulting inoculum production as well as the 

ability of inoculum to persist in the absence of host seeds. 

We have data to suggest that inoculum does not persist for 

more than a year without a host seed. One strategy would 

be to carry out biocontrol and seeding on burns, where the 

cheatgrass seed density is already much reduced, so that 

inoculum production from these seeds would be low. This 

would also increase the chances of achieving complete or 

near-complete cheatgrass control.

Conclusions _______________________

In summary, our work on indigenous fungal pathogens as 

potential biocontrol organisms for cheatgrass has given us 

a sense of guarded optimism. These organisms pose none of 

the threats posed by classical biocontrol organisms imported 

from the Old World range of cheatgrass, and we know that 

there are scenarios where each of them has resulted in 

local extinction or near-extinction of the cheatgrass host 

population. By investigating the ecological requirements of 

these organisms and understanding how they interact with 

cheatgrass seeds to cause either endemic or epidemic levels of 

disease, we may be able to combine them with each other and 

with other methods of control to achieve the near-complete 

control required to permit successful seeding of native species 

into cheatgrass monocultures on arid and semiarid sites. As 

more and more plant communities in the Intermountain West 

are invaded and supplanted by this highly successful weed, 

the imperative to find ways to effectively restore cheatgrass 

monocultures can only become stronger.

Acknowledgments__________________

This research was supported in part by grants from 

CSREES NRI (Grant #98-35303-6957 from the Plant 

Pathology Program and Grant #2000-35319-09920 from the 

Biology of Plant-Microbe Interactions Program) and by longterm funding from the Idaho Army National Guard. Toupta 

Boguena, Alisa Ramakrishnan, Sam Inouye, Duane Smith, 

Bettina Schultz, Kirsten Ward, Megan Ferguson, Katie 

Temus, Thom Stewart, Joybeth Stewart, Jay Weaver, and 

Dana Quinney helped with field data and sample collection 

and sample processing.

References ________________________

Bauer, M.C.; Meyer, S.E.; Allen, P.S. 1998. A simulation model 

for predicting dormancy loss in the field for seeds of Bromus 

tectorum L. Journal of Experimental Botany. 49: 1235-1244.

Beckstead, J.; Meyer, S.E.; Molder, C.; Smith, C. In press. A race 

for survival: Can Bromus tectorum seeds escape Pyrenophora 

semeniperda-caused mortality by germinating quickly? Annals 

of Botany.

Boguena, T.; Meyer, S.E.; Nelson, D.L. 2007. Low temperature during 

infection limits Ustilago bullata (Ustilaginaceae, Ustilaginales) 

disease incidence on Bromus tectorum (Poaceae, Cyperales). 

Biocontrol Science and Technology. 17: 33-52.

Campbell, M.A.; Medd, R.W. 2003. Leaf, floret and seed infection 

of wheat by Pyrenophora semeniperda. Plant Pathology. 52: 

437-447.

Campbell, M.A.; Medd, R.W.; Brown, J.F. 2003. Phytotoxicity of 

metabolites produced by Pyrenophora semeniperda in liquid 

culture. Australian Journal of Experimental Agriculture. 43: 

1237-1244.

Capio, E.R.; Tate, M.E.; Wallwork, H. 2004. Phytotoxic metabolites 

from Drechslera wirreganensis and D. campanulata. Australasian 

Plant Pathology. 33: 23-28.

D’Antonio, C.M.; Vitousek, P.M. 1992. Biological invasions by exotic 

grasses, the grass fire cycle, and global change. Annual Review 

of Ecology and Systematics. 23: 63-88.

Fischer, G.W. 1940. Host specialization in the head smut of grasses, 

Ustilago bullata. Phytopathology. 30: 991-1017.

Hoffmann, J.A.; Meiners, J.P. 1971. Host specialization in the complex species Tilletia fusca. Phytopathology. 61: 225-227.
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Cheatgrass (Bromus tectorum) Biocontrol Using Indigenous Fungal Pathogens Meyer, Nelson, Clement, Beckstead

Kennedy, A.C.; Elliot, L.F.; Young, F.L.; Douglas, C.L. 1991. 

Rhizobacteria suppressive to the weed downy brome. Soil Science 

Society of America Journal. 55: 722-727.

Kennedy, A.C.; Johnson, B.N.; Stubbs, T.L. 2001. Host range of a 

deleterious rhizobacterium for biological control of downy brome. 

Weed Science. 49: 792-797.

Kreitlow, K.W.; Bleak, A.T. 1964. Podosporiella verticillata, a soilborne pathogen of some western Gramineae. Phytopathology. 

54: 353-357.

Medd, R.W. 1992. A review of the world distribution and host 

range of Pyrenophora semeniperda. Review of Plant Pathology.

71: 891-901.

Medd, R.W.; Campbell, M.A. 2005. Grass seed infection following 

inundation with Pyrenophora semeniperda. Biocontrol Science 

and Technology. 15: 21-36.

Medd, R.W.; Murray, G.M. Murray; Pickering, D.I. 2003. Review 

of the epidemiology and economic importance of Pyrenophora 

semeniperda. Australasian Plant Pathology. 32: 539-550.

Meiners, J.P. 1958. Studies on the biology of Tilletia bromi-tectorum.

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Meiners, J.P.; Waldher, J.T. 1959. Factors affecting spore germination of twelve species of Tilletia from cereals and grasses. 

Phytopathology. 49: 724-728.

Meyer, S.E.; Nelson, D.L.; Clement, S. 2001. Evidence for resistance 

polymorphism in the Bromus tectorum-Ustilago bullata pathosystem: Implications for biocontrol. Canadian Journal of Plant 

Pathology. 23: 19-27.

Meyer, S.E.; Quinney, D.; Nelson, D.L.; Weaver, J. 2007. Impact 

of the pathogen Pyrenophora semeniperda on Bromus tectorum

seedbank dynamics in North American cold deserts. Weed Research. 47: 54-62.

Meyer, S.E.; Nelson, D.L.; Clement, S.; Waters, J.; Stevens, M.; Fairbanks, D. 2005. Genetic variation in Ustilago bullata: Molecular 

genetic markers and virulence on Bromus tectorum host lines. 

International Journal of Plant Sciences. 166: 105-115.

Meyer, S.E.; Nelson, D.L.; Clement, S. In review. Host specificity of 

Ustilago bullata races from Bromus tectorum: Assessing the risk 

to non-target perennial grass species. Phytopathology.

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review. Ecological genetics of the Bromus tectorum-Ustilago 

bullata pathosystem: A role for frequency-dependent selection? 

New Phytologist.

Monsen, S.B. 1994. The competitive influences of cheatgrass (Bromus tectorum) on site restoration. In: Monsen, S.B.; Kitchen, 

S.G., comps. Proceedings—Ecology and management of annual 

rangelands. Gen. Tech. Rep. INT-GTR-313. Ogden UT: U.S. Department of Agriculture, Forest Service, Intermountain Research 

Station: 43-52.

Whisenant, S.G. 1990. Changing fire frequencies on Idaho’s Snake 

River Plains: ecological and management implications. In: 

McArthur, E.D.; Romney, E.M.; Smith, S.D.; Tueller, P.T., comps. 

Proceedings—Symposium on cheatgrass invasion, shrub die-off, 

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Rep. INT-276. Ogden, UT: U.S. Department of Agriculture, Forest 

Service, Intermountain Research Station: 4-10.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

USDA Forest Service Proceedings RMRS-P-52. 2008 69

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Scott L. Jensen is a Botanist, U.S. Forest Service, Rocky 

Mountain Research Station, Shrub Sciences Laboratory, Provo, 

UT; email: sljensen@fs.fed.us.

Stephen B. Monsen is a Research Botanist (retired), U.S. 

Forest Service, Rocky Mountain Research Station, Shrub Sciences 

Laboratory, Provo, UT.

Pat Fosse is Supervisory Rangeland Management Specialist, 

Bureau of Land Management, Dillon, MT.

Abstract—This case study documents temporal and spatial squarrose knapweed (Centaurea virgata Lam. spp. squarrosa (Willd.) 

Gugler) seed dispersal from study sites in central Utah. Felt seed 

traps were placed along transects extending from sites occupied 

by squarrose knapweed into adjacent plant communities. Seed 

dispersal from knapweed plants occurred over an 8-month period. 

By 1 November 1999, 50 and 53 percent of dispersed seed had been 

recovered at the Knight and Cherry Creek locations respectively. 

By May, nearly all seed has dispersed from the producing plant. 

Most seed was deposited and remained within 2 m of the parental 

plant.

Introduction_______________________

Identifying the mechanisms, timing and extent of seed 

dispersal is critical to understanding potential spread of 

invasive species. Squarrose knapweed (Centaurea virgata

Lam. spp squarrosa (Willd.) Gugler) was first documented 

in Juab County Utah in 1954. At that time it was reported 

to occur along approximately 80 km of highway in the Tintic 

valley, on 162–202 hectares on a ranch in Juab County and 

in small patches at other locations (Tingey 1960). Presently 

it is found on tens of thousands of hectares in Juab, Utah, 

Tooele, Millard, and Sanpete counties. Its occurrence has 

been linked to a variety of dispersal mechanisms including 

livestock grazing and trailing, wildlife use, as a contaminant in crop seed and hay, gravel moving activities, vehicle 

traffic, railroad activity, and recreation traffic (Roche and 

others 1992). We have seen it proliferate along ephemeral 

streams and drainages and increase in abundance dramatically following fire. Once established, knapweed gradually 

increases in density. Few plant communities appear resilient 

to its invasion.

Spatial and Temporal Seed Dispersal of Squarrose 

Knapweed (Centaurea virgata Lam. spp. squarrosa (Willd.) 

Gugler) in West Central Utah, a Case Study

Scott L. Jensen, Stephen. B. Monsen, and Pat Fosse

In Utah, squarrose knapweed flowering typically begins 

the first week of July and continues through the first week 

of August with about 20 percent of the population flowering 

per week during that interval (Jensen, unpublished data). 

A flower progresses from bud to senescence in 7 days or less 

(Jensen, unpublished data). Most seed is mature by the end 

of August. Seed production can vary greatly from year to year 

with almost no production during extended drought periods 

to over 4,000 seeds/m2 during good moisture years (Jensen, 

unpublished data). Achenes develop within an upright encapsulated seedhead (capitula). Achene bodies are 3 to 3.8 mm 

long and have a pappus up to 3.5 mm long (Roche and Roche 

1989). The small pappus is not capable of providing lift for 

airborne dispersal common to other Compositae seeds. A 

capitula typically contains 1–5 viable achenes (Roche and 

Roche 1989), and often as many nonviable achenes. Upon 

maturation the capitula opens partially allowing some seed 

dispersal if the plant is aggressively brushed or wind blown. 

Typical dispersion occurs several months after ripening as 

the capitula is separated from the plant. An abscission layer 

develops at the base of the capitula allowing the structure 

to break free from the stem (Roche and Roche 1989). Seed 

viability is typically greater than 85 percent with germination occurring within 7 days in warm moist conditions 

(Jensen, unpublished data).

Roche and others (1992) conducted a study to determine 

if plant distribution was due to seed carried in the wool 

of sheep. These authors concluded that while sheep carry 

and disperse seed, they are by no means the only dispersal 

mechanism. A plethora of unexplored mechanisms remain. 

In this study we document the persistent contiguous population expansion by monitoring seed dispersal from dense 

knapweed infestations into adjacent plant communities. This 

study reports the timing and volume of seed movement at 

two scales, <15 m and at 15 m intervals to 60 m, regardless 

of responsible mechanism.

This case study is the first to attempt to quantify squarrose knapweed seed dispersal by distance and time. Methods 

developed as the study progressed resulting in data that 

were not as quantitative as we would have liked. Nonetheless, this study provides useful data to land managers that 

is otherwise not documented.

Methods __________________________

Separate studies addressed short (<15 m) and medium 

(<60 m) distance seed dispersal at two study sites in the 

Tintic valley of Juab County, Utah. Both the Cherry Creek 
70 USDA Forest Service Proceedings RMRS-P-52. 2008

Jensen, Monsen, and Fosse Spatial and Temporal Seed Dispersal of Squarrose Knapweed ...

and Knight sites are located on previously dry farmed 

agricultural fields dominated by squarrose knapweed to 

the exclusion of almost all other species. A wildfire burned 

the Cherry Creek site several years prior to the initiation of 

the study. Rangeland communities of pinyon-juniper (Pinus 

edulis, Juniperus osteosperma), big sagebrush (Artemesia 

tridentata) rabbitbrush (Ericameria nauseosa) and perennial grass are adjacent to this site. The Knight site adjoins 

crested wheatgrass (Agropyron cristatum) seedings and 

pinyon-juniper and big sagebrush communities.

The short distance study focused on seed dispersal from an 

individual plant and was conducted only at the Knight study 

site between August 1999 and May 2000. Medium distance 

dispersal studies focused on distance, direction, timing, and 

volume of seed dispersed from dense squarrose knapweed 

populations. Medium distance studies were conducted at 

the Knight and Cherry Creek study sites between August 

1999 and March 2001

Four knapweed plants were selected to document short 

distance dispersal. Surrounding vegetation was mowed 

to 15 m beyond each plant to facilitate locating dispersed 

capitulas. Upon cessation of flowering in August 1999, plants 

were painted using fluorescent orange spray paint. In the 

spring of 2000, the number of painted capitulas occurring 

within 1⁄2 m concentric circles was recorded.

To document medium distance dispersal, ten transects 

were established at irregular intervals along the periphery 

of both the Knight and Cherry Creek sites. Two transects 

were oriented roughly perpendicular to each field edge in the 

following directions: 0°, 72°, 144°, 216°, and 288°. Transects 

began 15 m inside the knapweed infested field and extended 

60 m into the adjoining plant communities creating data 

points at –15 m, 0 m , 15 m, 30 m, 45 m and 60 m. At each 

data point, 6 seed traps were equally spaced along a 10-m 

line lying perpendicular to the transect. Each seed trap 

consisted of a 232 cm2 piece of plush felt material held to 

the ground with 16 penny nails. Knapweed abundance drops 

off dramatically at the ecotone of the knapweed field and 

the adjacent plant community. Seed producing knapweed 

plants, occurring along transects in adjoining communities, 

were removed to eliminate onsite seed production.

For the 1999 seed cohort distance, volume, direction, and 

timing data were recorded in September and October of 

1999 and April of 2000. Sampling continued at the Cherry 

Creek site through early July of 2000. For the 2000 seed 

cohort distance, volume and direction data were collected 

in mid-March 2001. The timing study was not conducted 

during the second year.

Results and Discussion______________

Seed Dispersal: Distance, 

Volume and Direction

The short distance seed dispersal study documented an 

abbreviated natural dispersal distance. Seventy eight percent (206) of the capitulas collected (total sample = 264) fell 

within 0.5 m of the producing plant (fig. 1); an additional 15 

percent (40) traveled between 0.5 and 1 m, 5 percent (14) 

between 1 and 1.5 m, and just 2 percent (4) between 1.5 and 

2 m. This dramatic decline in capitula numbers as distance 

0

10

20

30

40

50

60

70

80

0 - 0.5 m 0.5 - 1 m 1 - 1.5 m 1.5 - 2 m

Distance from Maternal Plant Capitulas were Recovered

Number of Capitulas Recovered

Plant 1 Plant 2 Plant 3 Plant 4

Figure 1—The number of orange painted capitulas recovered from the four painted plants is shown at 

1⁄2 m increasing distance intervals.
USDA Forest Service Proceedings RMRS-P-52. 2008 71

Spatial and Temporal Seed Dispersal of Squarrose Knapweed ... Jensen, Monsen, and Fosse

0

20

40

60

80

100

120

140

160

-15 m 0 m 15 m 30 m 45 m 60 m

Data Points

Number of Seeds Collected

Knight 1999 Cohort

Cherry Creek 1999 Cohort

Knight 2000 Cohort

Cherry Creek 2000 Cohort

Figure 2—The number of seeds recovered at increasing distances from the knapweed field is shown 

for the 1999 and 2000 seed cohorts at the Knight and Cherry Creek sites.

increases suggests the majority of natural dispersion occurs 

within 2 m of the maternal plant.

Medium distance data that is intended to assess dispersal 

from the population as a whole illustrate a similar occurrence. Combining data from the 1999 and 2000 seed cohorts 

at both sites, 73.6 percent of recovered seed accumulated at 

the –15-m data points, which lie within the knapweed infested 

fields (fig. 2). An additional 24.4 percent of the collected 

seed accumulated at the 0-m data points, which occur at 

the ecotone of the knapweed fields and adjacent plant communities. Only at the Knight site during 1999 were achenes 

collected beyond the 0-m point. This amounted to 2.0 percent 

of the total recovered seed. The combined trap area at each 

distance interval was 1.39 m2. Standardizing to a m2 unit 

the equivalent of 2.9 achenes were recovered at the 15 m 

data point, 2.2 at 30 m, 2.9 at 45 m and 1.4 at 60 m.

We hypothesized wind transported achenes would accumulate in larger quantities along transects directed 

inline with predominant winds, which generally blow from 

the south or southwest. Because little seed was collected 

outside the boundary of the knapweed stands our data are 

inconclusive.

Seed Dispersal: Timing and 

Subsequent Movement

With seed maturation in August, the stage was set 

for propagule dissemination. By 1 September 1999 the 

abscission layer subtending capitulas had formed permitting 

detachment when heads were brushed but holding them 

through the strongest winds. From August to mid-October 

1999, foraging grasshoppers caused considerable quantities 

of capitulas to accumulate on the ground. By mid-September 

15.5 percent of the total seed collected at the Knight site and 

25.8 percent of the total seed collected at the Cherry Creek 

site was recovered in traps. This early seed fall is an anomaly 

that will only occur during grasshopper outbreaks. Not until 

mid-October were capitulas observed accumulating on the 

ground as a result of natural fall. By the first of November 

1999, 50 percent of the total seed collected at the Knight 

site and 53 percent at the Cherry Creek site was recovered. 

Winter weather conditions provide the appropriate environment for remaining capitulas to dehisce. At the Knight site, 

the remaining 50 percent of seed was collected from traps in 

April. At Cherry Creek, 47 percent of the seed was collected 

in April, with the final 10 percent in July (fig. 3).

A small amount of seed fall and or movement occurs over 

the summer months accounting for the July accumulation. 

We suspect this is due to several conditions. First, plants 

retain trace amounts of seed in non-dehisced capitulas. 

Secondly, soils in the Tintic valley are high in clay content. 

Capitulas coming in contact with the soil during wet winter and spring months adhere to the surface. As soils dry 

throughout the summer, seed may be released allowing 

further movement.

Management Implications ___________

All data and observations indicate knapweed has not developed mechanisms for rapid and massive seed dispersal 
72 USDA Forest Service Proceedings RMRS-P-52. 2008

Jensen, Monsen, and Fosse Spatial and Temporal Seed Dispersal of Squarrose Knapweed ...

over expansive areas. But the tendency to retain seed for a 

long duration combined with our cultural use of the lands 

accounts for a high probability of seed attaching to animals or 

vehicles for long distance dispersal. This, aided by abundant 

short distance dispersal and the competitive nature of the 

plant, results in dense stands producing large quantities of 

seed that advance at a relatively slow but consistent rate 

into new landscapes.

For nearly 8 months beginning in August, squarrose 

knapweed plants retain some amount of seed on their stems 

and large amounts of seed near their base. This prolonged 

retention period facilitates seed transfer by wind, animals, 

vehicles, water and other carriers. Many of these vectors are 

not readily manageable. Of those we have influence over, 

grazing and travel activities have the greatest potential 

for spreading seed and therefore the greatest opportunity 

to minimize expansion through management. Reducing or 

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

Figure 3—The percent of seed recovered by date and site is shown for the 1999 seed cohort at the Knight 

and Cherry Creek sites. For the 2000 cohort seed was collected from traps only in March. 

0.0%

20.0%

40.0%

60.0%

80.0%

100.0%

120.0%

16-Sep-99 19-Oct-99 24-Apr-00 3-Jul-00 21-Mar-01

Date

Percent of Seed Recovered

Knight Cherry Creek

excluding activities in knapweed infested sites from August 

to April will reduce the potential for seed transfer.

References ________________________

Jensen, S.L. Unpublished data collected in 2006. Available on file 

at the U.S.DA Forest Service RMRS, Shrub Sciences Laboratory, 

Provo, UT.

Roche, C.T.; Roche, B.F. 1989. Introductory notes on squarrose knapweed (Centaurea virgata Lam. ssp. squarrosa Gugl.). Northwest 

Science. 63(5): 246-252.

Roche C. T.; Roche, B.F.; Rasmussen, G.A. 1992. Dispersal of squarrose knapweed (Centaurea virgata ssp. squarrosa) capitula by 

sheep on rangeland in Juab County, Utah. Great Basin Naturalist. 52(2): 185-188.

Tingey, D.C. 1960. Control of squarrose knapweed. Utah State 

University Experiment Station Bulletin No. 432. Logan, UT: 

Utah State University. 11 p.
USDA Forest Service Proceedings RMRS-P-52. 2008 73

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Robert R. Blank is a Soil Scientist, USDA-ARS, Exotic and Invasive 

Weed Research Unit, Reno, NV; email: blank@unr.nevada.edu.

René Sforza is a Biocontrol of Weeds Specialist, USDA-ARS, European Biological Control Lab, Montpellier, France; email: rsforza@

ars-ebcl.org.

Tye Morgan is a Research Technician, USDA-ARS, Reno, NV.

Abstract—To better understand why medusahead (Taeniatherum 

caput-medusae) is invasive, we quantified soil N availability and 

characterized soil microbial communities between native and invasive populations. No consistent differences in soil N mineralization 

potentials were noted between native medusahead sites in Spain, 

Turkey, France, and Greece and two invaded sites on the volcanic 

tablelands of northeastern, California, U.S.A. The proportional 

makeup of the microbial community, as quantified by phospholipid 

fatty acid analysis (PLFA), did not differ appreciably between a native 

site in southern France and one site on the tablelands of northeastern, California sites. Microbial markers indicative of growth phase 

suggest soils of invasive populations have higher turnover rates 

than native soil. No useable DNA could be extracted from the native 

medusahead soil in southern France using bacterial S16 ribosomal 

DNA. Tablelands soil contained distinct bacterial DNA bands for 

bacteria that utilize methane and methanol, anaerobically reduce 

sulfur, catabolize aromatics, for symbiotic relationships with root 

nodules of legumes. In summary, there is no definitive evidence in 

the studies carried out to indicate why medusahead is invasive in 

the Western United States.

Introduction_______________________

Why do some exotic plants become invasive? The question has been and is being vigorously researched. Myriad 

mechanisms have been proposed whereby an exotic plant 

can gain a competitive advantage over native species. An 

influential hypothesis posits that exotic species, set free 

from native biological enemies, from microbes through arthropods, now flourish in their new environment (Keane and 

Crawley 2002; Torchin and others 2003). This hypothesis is 

the underpinning for introducing biological control agents 

from the exotic’s home of origin. It has become clear, however, that the ultimate answer to how some exotic species 

become invasive is incredibly complex. Some invasive plants 

are allelopathic, capable of exuding powerful phytoxins, 

Medusahead: Available Soil N and Microbial 

Communities in Native and Invasive Soils 

Robert R. Blank, René Sforza, and Tye Morgan

which displace competing native vegetation (Bais and others 

2003). Other invasive weeds alter or “engineer” the soil to 

increase nutrient availability or are able to access a previously under-utilized nutrient niche and thereby increase 

their competitive profile (Blank 2002). Soil food webs can 

be changed by exotic plants to gain a competitive advantage 

(Belnap and Phillips 2001). Confounding explanations of 

why some exotic plants become invasive is the finding that 

anthropogenic elevated atmospheric carbon dioxide confers 

greater advantage to faster growing weedy species many of 

which are exotic (Ziska 2003).

A robust factor that explains invasive success is soil nutrient availability, particularly availability of N and P (Milbau 

and Nijs 2004). Indeed, increased atmospheric deposition of 

nitrogen increases competitive ability of some weeds (Fenn 

and others 2003). Two of the dominant invasive annual 

grasses in the Intermountain West, cheatgrass (Bromus 

tectorum) and medusahead (Taeniatherum caput-medusae) 

greatly increase their competitive stature upon soil additions 

of N (Paschke and others 2000, Monaco and others 2003). 

The present study on medusahead is linked to a larger 

program on invasion processes and biological control using 

natural enemies from Eurasia (Widmer and Sforza 2003, 

Sforza and others 2004).

Our purpose here was to compare soil N availability and 

the soil microbial community in native and exotic soils 

occupied by medusahead. Our working hypotheses were: 

1) soil N availability is similar in native and exotic soils 

and 2) the soil microbial community is similar in native 

and exotic soils.

Materials and Methods______________

Hypothesis testing was accomplished by collecting soil, 

0 to 20 cm, directly beneath native and exotic populations of 

medusahead (see fig. 1 for photographs of native and exotic 

landscapes). We collected native soil from Spain (3 sites), 

Turkey (4), Greece (2), and France (4) with 4 replications 

at each site. These soils were contrasted with exotic populations in two locations of northeastern California (fig. 1). 

The California sites were near Alturas and on the volcanic 

tablelands just north of the Honey Lake Valley. The European sites were in Andalousia, near Huelva, Cordoba, and 

Granada for Spain, in Central and Eastern Anatolia, near 

Göreme, Erzurum, Gaziantep, and Erzinçan for Turkey, in 

Northern Greece, near Thessaloniki, and in Southern France, 

near Montpellier. On the less than 2 mm soil fraction, available soil N was quantified using a 30 day aerobic incubation 

procedure (Hart and others 1994). In another experiment, 

freshly-collected soil (two replicates) from southern France 
74 USDA Forest Service Proceedings RMRS-P-52. 2008

Blank, Sforza, and Morgan Medusahead: Available Soil N and Microbial Communities in Native and Invasive Soils

(native) and from the volcanic tablelands of northeastern 

California (invasive) were sent in iced coolers to a commercial laboratory for phospholipid fatty acid analysis (PLFA) 

(Zelles and Bai 1993) and S-16 ribosomal DNA analysis using 

denaturing gradient gel electrophoresis (DGGE) (Stephen 

and others 1999).

Results and Discussion______________

Nitrogen Availability

There is no consistent trend in soil N availability after 30 

day aerobic incubation between native and exotic populations of medusahead (fig. 2). Nitrogen availability is a robust 

predictor of ecosystem invisibility of exotic plant species 

(Huenneke and others 1990). Based on this relationship, 

one would suspect that the soil from Spain and three of the 

sites in Turkey would be at similar risk of invasion as the 

sites in the United States that are actively being invaded by 

medusahead. We are unaware, however, of any threshold of 

N availability that fosters plant invasion. In all likelihood, 

the growth of weedy species such as medusahead would respond differently to pulses of N availability through fertilizer 

addition versus the slow increases due to mineralization. 

None-the-less, similar soil N availability between native and 

exotic populations suggests that it is not the controlling factor 

explaining its invasiveness in western North America.

PLFA Analyses

Native medusahead soils from southern France have far 

greater total microbial, bacterial, and eukaryotic biomass 

than the invaded sites on the tablelands of northern California 

(table 1). These findings seem logical given that native soils 

of southern France are wetter and contain more organic matter than the tableland soils. Both invaded and non-invaded 

soils are bacteria-dominated and their ratios of bacterial to 

eukaryotic biomass are similar. Bacterial dominance may 

be indicative of a soil system that is not self-regulating — 

out of equilibrium (Bardgett and McAlister 1999). Ratios of 

fatty acid biomarkers, the growth phase (table 1), suggest 

that the tableland soils have higher microbial turnover 

rates than the native soils in France, which is logical given 

the greater climatic variability and harshness compared to 

the French sites. Despite the large differences in microbial 

biomass and site climatic condition between the two sites, 

the soil microbial community structure has only minor 

differences between the native and invasive sites (fig. 3). 

Given the marked difference in soils and climate, this finding seems remarkable. Moreover, invasive plants can alter 

microbial community structure promoting different classes 

of organisms presumably to increase plant fitness (Batten 

and others 2006).

Figure 1—Top photo is from a native medusahead area in 

Nemrut national park in southwestern Turkey. Bottom photo is 

from a medusahead-invaded area in the volcanic tablelands of 

northeastern, California where the native low sage community 

has largely been extirpated.

Spain

France

Greece 1Greece 2

Turkey 1Turkey 2Turkey 3Turkey 4

Alturas 1Alturas 2

Tablelands 1Tablelands 2

0

1

2

3

4

5

6

7

30 day N mineralization potential (mmol / kg soil)

 Native Populations Invasive Populations

Figure 2—Total available N (NH4

+ + NO3–

) in soil following 

30 day aerobic incubation.
USDA Forest Service Proceedings RMRS-P-52. 2008 75

Medusahead: Available Soil N and Microbial Communities in Native and Invasive Soils Blank, Sforza, and Morgan

DNA Analyses

We were unable to obtain bacterial DNA from the native 

medusahead site in France. For the invaded tablelands site, 

six bands with useable sequence information were obtained 

(table 2). Methylomonas and Sulfuricurvum seem out of 

place in the semi-arid environment of the tablelands. In 

these lake-laid, clayey soils, however, there is the potential 

Figure 3—Microbial community structure of native (France) 

and invasive (tablelands) soils as discerned by phospholipid 

fatty acid analyses (PLFA).

Table 1—Results of PLFA analyses for native medusahead site in 

France and invasive site on the volcanic tablelands of 

northeastern, CA. Standard deviations in parentheses.

Attribute France Tablelands

- - - pico moles PLFA g–1 soil- - -

Microbial biomass 51996 (5206) 22995 (4773)

Bacteria biomass 48077 (5120) 21344 (4354)

Eukaryotic biomass 3916 (1017) 1654 (459)

Ratio Bacteria/Eukaryotes 12.3 12.9

Growth phase 0.635 (0.103) 0.363 (0.023)

Gram+/Anaerobic gram -

Gram -

Anaerobic metal reducers

SRB/Actinomycetes

Genera

Eukaryotes

0

5

10

15

20

25

30

35

40

45

50

Community Structure (% of total PLFA)

France

Tablelands

Table 2—Bacterial DNA analyses of medusahead invaded areas on the tablelands. 

Genus Function % Match

Methylomonas utilize methane and methanol 97

Sulfuricurvum anaerobe – sulfur reducing 100

Eubacerium biphenyl-contaminated site 96

Sphingomonas aerobe – broad catabolic potential (aromatics) 100

Caulobacter specialized for oligotrophic environments 100

Bradyrhizobium root nodules legumes 97

at depth and in reduced microsites for methane production 

and sulfur reduction. Caulobacter generally live in dilute 

aquatic environments where the most common limiting 

nutrient is phosphorus. Certainly, availability of P in 

these soils is reduced by sorption to clay, but it is hardly an 

aquatic environment? The genus Sphingomonas is known 

for its ability to catabolize complex organics and is used to 

remediate contaminated sites. Its presence in tableland 

soil may be a reflection of the complex biochemistry of 

Artemisia? The tablelands have legumes, thus the presence 

of Bradyrhizobium is logical.

Conclusions _______________________

Available N does not appear to be the controlling factor in 

medusahead invasion. Soil microbial community structure 

does not appreciably differ between native and invasive medusahead populations. Does this fact imply that community 

structure cannot explain the invasive nature of medusahead 

in the Western United States? Alternatively, do microbial 

populations between native and invasive sites differ at the 

genus and species level and does this difference explain 

invasibility? We would additionally suggest testing these 

hypotheses with closely-related cheatgrass because it invades 

some of the same habitats in the Western United States, and 

it originates from similar habitats in Eurasia.

References ________________________

Bais, H.P.; Vepachedu, R.; Gilroy, S.; Callaway, R.M.; Vivanco J.M. 

2003. Allelopathy and exotic plant invasion: From molecules and 

genes to species interactions. Science. 301: 1377-1380.
76 USDA Forest Service Proceedings RMRS-P-52. 2008

Blank, Sforza, and Morgan Medusahead: Available Soil N and Microbial Communities in Native and Invasive Soils

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

Bardgett, R.D.; McAlister, E. 1999. The measurement of soil 

fungal:bacterial biomass ratios as an indicator of ecosystem 

self-regulation in temperate meadow grasslands. Biology and 

Fertility of Soils. 29: 282-290.

Batten, K.M.; Scow, K.M.; Davies, K.F.; Harrison, S.P. 2006. Two 

invasive plants alter soil microbial community composition in 

serpentine grasslands. Biological Invasions. 8: 217-230.

Belnap J.; Phillips, S.L. 2001. Soil Biota in an ungrazed grassland: 

Response to annual grass (Bromus tectorum) invasion. Ecological 

Applications. 11: 1261-1275.

Blank, R.R. 2002. Amidohydrolase activity, soil N status, and 

the invasive crucifer Lepidium latifolium. Plant and Soil. 239: 

155-163.

Fenn, M.E.; Baron, J.S.; Allen, E.B.; Rueth, H.M.; Nydick, K.R.; 

Gieser, L.; Bowman, W.D.; Sickman, J.O.; Meixner, T.; Johnson, 

D.W.; Neitlich, P. 2003. Ecological effects of nitrogen deposition 

in the Western United States. BioScience. 53: 404-420.

Hart, S.C.; Stark, J.M.; Davidson, E.A.; Firestone, M.K. 1994. Nitrogen mineralization, immobilization and nitrification. In: Weaver, 

R.W., chief ed. Methods of Soil Analysis, Part 2. Microbiological 

and Biochemical properties. SSSA Book Series, no. 5. Madison, 

WI: Soil Science Society of America: 985-1013.

Huenneke, L.F.; Hamburg, S.P.; Koide, R.; Mooney, H.A.; Vitousek, 

P.M. 1990. Effects of soil resources on plant invasion and community structure in Californian serpentine grassland. Ecology. 

71: 478-491.

Keane, R.M.; Crawley, M.J. 2002. Exotic plant invasions and the 

enemy release hypothesis. Trends in Ecology and Evolution. 17: 

164-170.

Milbau, A.; Nijs, I. 2004. The role of species traits (invasiveness) and 

ecosystem characteristics (invasibility) in grassland invasions: a 

framework. Weed Technology. 18: 1301-1304.

Monaco, T.A.; Johnson, D.A.; Norton, J.M.; Jones, T.A.; Connors, 

K.J.; Norton, J.B.; Redinbaugh, M.B. 2003. Contrasting responses 

of Intermountain West grasses to soil nitrogen. Journal of Range 

Management. 56: 282-290.

Paschke, M.; McLendon, W.T.; Redente, E.F. 2000. Nitrogen availability and old-field succession in a shortgrass steppe. Ecosystems. 

3: 144-158.

Sforza, R.; Eken, C.; Hayat, R.; WidmerT.L. 2004. First evaluation 

of Ustilago phrygica for the biological control of taeniatherum 

caput-medusae (Triticeae). In: Proceedings of the XIIth International Workshop on Weed Biology, Dijon, France, 31 Aug.– 2 

Sept. 2004: 407-412.

Stephen, J.R.; Chang, J.; Gan, Y.D.; Peacock, A.; Pfiffner, S.; Barcelona, M.; White, D.C.; MacNaughton, J. 1999. Microbial characterization of a JP-4 fuel-contaminated site using a combined 

lipid biomarker/polymerase chain reaction-denaturing gradient 

gel electrophoresis (PCR-DGGE) based approach. Environmental 

Microbiology. 1: 231-241.

Torchin, M.E.; Lafferty, K.D.; Dobson, A.P.; McKenzie, V.J.; Kuris, 

A.M. 2003. Introduced species and their missing parasites. Nature. 421: 628-630.

Widmer T.L.; Sforza R. 2004. Exploration for plant pathogens against 

Taeniatherum caput-medusae (medusahead). In: Cullen, J.M.; 

Briese, D.T.; Kriticos, D.J.; Lonsdale, W.M.; Morin, L.; Scott, J.K., 

eds. Proceedings of the XI International Symposium on Biological 

Control of Weeds. Canberra, Australia: CSIRO: 193-197.

Zelles, L.; Bai, Q.Y. 1993. Fractionation of fatty acids derived from 

soil lipids by solid phase extraction and their quantitative analysis 

by GC-MS. Soil Biology and Biochemistry. 25: 495-507.

Ziska, L.H. 2003. Evaluation of the growth response of six invasive 

species to past, present and future atmospheric carbon dioxide. 

Journal of Experimental Botany. 54: 395-404.
USDA Forest Service Proceedings RMRS-P-52. 2008 77

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Alison Whittaker is a Habitat Biologist, Utah Division of Wildlife 

Resources, Northeastern Region, Vernal, UT.

Scott L. Jensen is a Botanist, U.S. Forest Service, Rocky Mountain Research Station, Shrub Sciences Laboratory, Provo, UT.

Abstract—Squarrose knapweed (Centaurea virgata var. squarrosa), herein referred to simply as knapweed, is a noxious weed that 

invades both disturbed and healthy sagebrush communities. Fire, 

grazing, mining, recreation, and farming have all played a large 

part in the establishment of knapweed in Tintic Valley, Utah. This 

study was designed to look at the effect of fire on the establishment 

of knapweed in four community types, namely cheatgrass, crested 

wheatgrass, degraded sagebrush, and healthy sagebrush. Plots 

were either burned or left unburned after which subplots were 

seeded with knapweed. Results show that fire plays an important 

role in knapweed invasion especially in sagebrush communities. 

Knapweed establishment on cheatgrass plots did not vary between 

burn treatments, and knapweed establishment was lowest on crested 

wheatgrass sites.

Introduction_______________________

Squarrose knapweed (Centaurea virgata var. squarrosa), 

is a secondary weed invading the cold desert in central Utah. 

This perennial, native to the Middle East (Roche and Roche 

1989), is well adapted to the climate of the Great Basin. 

Knapweed was recognized in Utah over 50 years ago, and 

initial research on life history and control measures began 

in the 1960s (Tingey 1960). By the late 1980s, knapweed was 

estimated to cover 40,468 ha, and organized control efforts 

were materializing (Fosse 98). Knapweed is currently found 

in high densities in five Utah counties (Roche and Roche 

1989) and is increasing in acreage at a rate of approximately 

10 to 15 percent per year (Fosse, unpublished data on file at 

the BLM Dillon, MT Field Office). It has invaded communities throughout the sagebrush-steppe ecosystem, including 

functional and at risk big sagebrush (Artemisia tridentata) communities, pinyon/juniper (Pinus edulis/ Juniperus

osteosperma), primary weed communities, and communities 

reseeded with introduced perennial grasses.

Effects of Fire and Restoration Seeding 

on Establishment of Squarrose Knapweed 

(Centaurea virgata var. squarrosa)

Alison Whittaker and Scott L. Jensen

Tintic Valley is located in central Utah and has a recent 

history of high fire frequency, due in part to an increase in cheatgrass (Bromus tectorum) abundance. Prior to the introduction of 

domestic livestock, valley ecosystems were dominated by big 

sagebrush and associated perennial herbaceous species, particularly grasses. Current big sagebrush community condition 

is a good indicator of disturbance (fire, grazing) history. Healthy 

sagebrush sites have an abundant herbaceous understory while 

degraded sagebrush sites lack this component.

Much of the land is currently managed by the Bureau of 

Land Management and, following fire, was reseeded with 

introduced grasses, primarily crested wheatgrass (Agropyron cristatum). Other major land tracts in the valley are 

privately owned. Some of the private holdings were seeded 

with crested wheatgrass following fire; however, unseeded 

portions are now dominated by the introduced annual cheatgrass. This has created a patchwork of sagebrush, crested 

wheatgrass, and cheatgrass-dominated communities across 

the landscape. All of these communities appear to have the 

potential to be invaded by knapweed. This project addressed 

the question of how fire affects the invasibility of functional 

and degraded sagebrush, cheatgrass-dominated, and revegetated communities by knapweed. We hypothesized that all 

community types would be more susceptible to knapweed 

invasion following fire.

Methods __________________________

Three study areas, or experimental blocks, were selected 

in Tintic Valley. Each block contained sites from the four 

community types: (1) healthy Wyoming big sagebrush 

(A. tridentata wyomingensis), (2) degraded Wyoming big 

sagebrush, (3) crested wheatgrass, and (4) cheatgrass. Community types within each block were located as close together 

as possible to minimize environmental differences.

Two 20 m x 20 m plots were located on each site, one of 

which was designated for burning and one to be left nonburned. Within each plot, 40 1-m2 permanent subplots were 

randomly located. Twenty of these subplots were seeded 

with knapweed and the other 20 were left unseeded. While 

knapweed was not present on the selected plots prior to 

seeding, knapweed occurred in varying densities adjacent 

to each block.

In May 2002, basal cover was inventoried with a 1-m2

modified 20-point frame in all subplots prior to burning and 

seeding treatments. This method estimates basal cover for 

litter, rock, microbial crusts, bare ground, and vegetation by 

species at a 5 percent level of precision (Bonham 1989).
78 USDA Forest Service Proceedings RMRS-P-52. 2008

Whittaker and Jensen Effects of Fire and Restoration Seedingon Establishment of Squarrose Knapweed (Centaurea virgata var. squarrosa)

Burning was completed in June 2002 with the help of 

BLM, USFS, and Utah State fire crews. Drip torches were 

used to ignite the fires. Fall of 2002, subplots in both the 

burned and unburned plots were broadcast seeded with 

knapweed at a rate of 2.17 kg/ ha (12 lbs/ acre). To eliminate 

edge effect, a 4-m2 area was seeded with the 1-m2 subplot 

positioned in the middle. Post-treatment assessment of basal 

cover was conducted in May 2003 and 2004 using the same 

protocols as were used prior to treatments. Knapweed basal 

cover and density were also monitored.

Data were analyzed using Systat General Linear Model and 

Analysis of Variance with Tukey’s pairwise comparison.

Results ___________________________

Spring and summer of 2003 experienced a severe 

drought, which resulted in low knapweed germination. 

Most seed remained in the seedbed until 2004 when a flush 

of knapweed seedlings were observed. As a result, only data 

from 2004 are presented here. Seeded knapweed density was 

dependant (p = 0.010) on the interaction between community type, burn treatment, and seed treatment. Knapweed 

density was significantly higher on unburned cheatgrass 

plots than on burned cheatgrass plots (fig. 1). Conversely, 

knapweed density was higher after burning compared to 

unburned plots for both healthy and degraded sagebrush 

communities. In contrast, there was not a significant difference in knapweed density between treatments on the 

crested wheatgrass plots.

Basal cover of the vegetation was significantly different 

among community types. In pairwise comparisons, perennial 

grass cover was significantly higher (p < 0.0001) on crested 

wheatgrass sites than on all other sites. Native forb cover 

was significantly higher on healthy sagebrush sites than 

on degraded sagebrush (p = 0.004) and crested wheatgrass 

(p=0.028) sites. Also in pairwise comparisons, the cheatgrass 

sites had higher (p = 0.017) annual grass and annual forb 

cover than did crested wheatgrass sites.

Discussion ________________________

Disturbances such as fire can cause an increase of soil 

nitrogen. Huenneke and others (1990) found that increasing the nitrogen in a system increases the invasibility of a 

community. Disturbance can also cause an increase in the 

duration of available soil moisture by decreasing vegetation. 

The loss of vegetation decreases the amount of transpiration occurring in that area, which results in greater water 

availability (Coronato and Bertiller 1996; Schlesinger and 

others 1987). With some exceptions, notably cheatgrass, 

early successional species such as primary and secondary 

weeds are typically not good competitors, but they are good 

colonizers and take advantage of available resources (Tilman 

and Wedin 1991). Fire increases both the nitrogen and water 

availability and creates an opening for plants to colonize. 

Given squarrose knapweed’s invasive tendencies and the 

presence of these conditions, we hypothesized knapweed 

densities would be greatest in all community types under 

burned and seeded conditions. This hypothesis proved 

0

2

4

6

8

Cheatgrass Crested

Wheatgrass

Degraded

Sagebrush

Intact

Sagebrush

Density per sq. meter

Nonburn Nonseeded Nonburn Seeded

Burn Nonseeded Burn Seeded

Figure 1—Mean squarrose knapweed density 2 years post treatment (May 2004) in 

four community types where burned and unburned plots were unseeded or seeded 

with knapweed propagules. 
USDA Forest Service Proceedings RMRS-P-52. 2008 79

Effects of Fire and Restoration Seedingon Establishment of Squarrose Knapweed (Centaurea virgata var. squarrosa) Whittaker and Jensen

correct among healthy and degraded sagebrush sites, but 

was not the case on crested wheatgrass and cheatgrass sites. 

The crested wheatgrass sites didn’t have a lot of knapweed 

establishment with or without fire. This is in contrast to the 

other three community types. The fire may not have been 

an intense enough disturbance to counter the competitive 

nature of crested wheatgrass and permit knapweed invasion, 

resulting in little contrast between burned and unburned 

seeded plots.

Invasion may also be facilitated by low biodiversity (Naeem 

and others 2000). The relatively low levels of biodiversity 

of cheatgrass, crested wheatgrass, and degraded sagebrush 

communities may make them more susceptible to invasion 

without fire. Knapweed invasion without fire, however, was 

only observed on the cheatgrass sites. Perennial species were 

not present on the cheatgrass site but dominated the other 

two sites. The presence of perennial species may contribute 

to poor knapweed establishment on these two low-diversity 

sites. It wasn’t until after the degraded big sagebrush sites 

were burned and the shrub component lost that squarrose 

knapweed was able to establish.

Another important factor to the invasibility of a community 

is the presence of weed propagules (Davis and others 2000). 

Knapweed can be likened to a slow moving army that simply 

keeps plodding along. It’s not necessarily an abundant seed 

producer, nor is it adapted to either a long-lived seed bank or 

excellent seed dispersion. Rather, most seed falls adjacent to 

the maternal plant, germinates at first opportunity, grows 

to maturity, and annually adds propagules that follow the 

same pattern. It seems to establish readily, is moderately 

long-lived and quite hearty. These data suggest that knapweed is capable of obtaining an initial foothold in each of 

these community types following fire, and most successfully 

in cheatgrass communities.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

Acknowledgments__________________

This project was funded by the Joint Fire Sciences Program 

and the Federal Pittman Robertson W-82-R for wildlife 

habitat restoration.

References ________________________

Bonham, C.D. 1989. Measurements for Terrestrial Vegetation. New 

York: John Wiley and Sons, Inc.

Coronato, F.R.; Bertiller, M.B. 1996. Precipitation and landscape 

related effects on soil moisture in semi-arid rangelands of Patagonia. Journal of Arid Environments. 34: 1-9.

Davis, M.A.; Grime, J.P.; Thompson, K. 2000. Fluctuating resources 

in plant communities: A general theory of invasibility. Journal 

of Ecology. 88: 528-534.

Fosse, P. 1998. Squarrose knapweed demonstration weed management area. National Weed Symposium. April 8-10. [Online] 

Available: http://www.blm.gov/weeds/sympos98/fosse.html

Huenneke, L.F.; Hamburg, S.P.; Koide, R.; Mooney, H.A.; Vitousek, 

P.M. 1990. Effects of soil resources on community structure in 

Californian serpentine grassland. Ecology. 71(2): 478-491.

Naeem, S.; Knops, J.M.H.; Tilman, D.; Howe, K.M.; Kennedy, K.M.; 

Gale, S. 2000. Plant diversity increases resistance to invasion in 

the absence of covarying extrinsic factors. Oikos. 91: 97-108.

Roche, C.T.; Roche, B.F. 1989. Introductory notes on squarrose 

knapweed (Centaurea virgata ssp.squarrosa Gugl.). Northwest 

Science. 63(5): 246-252.

Schlesinger, W.H.; Fonteyn, P.J.; Marion, G.M. 1987. Soil moisture 

content and plant transpiration in the Chihuahuan Desert of New 

Mexico. Journal of Arid Environments. 12: 119-126.

Tilman, D.; Wedin, D. 1991. Plant traits and resource reduction 

for five grasses growing on a nitrogen gradient. Ecology. 72(2): 

685-700.

Tingey, D.C. 1960. Control of squarrose knapweed. Utah State 

University Experiment Station Bulletin No. 432. Logan, UT: 

Utah State University. 11 p.

USDA Forest Service Proceedings RMRS-P-52. 2008 81

Shrubland Community 

Dynamics and Restoration

USDA Forest Service Proceedings RMRS-P-52. 2008 83

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Jerry R. Barrow and Mary E. Lucero are Research Scientist 

and Research Molecular Biologist, respectively, USDA-ARS, Jornada 

Experimental Range, New Mexico State University, Las Cruces, 

NM; email: jbarrow@nmsu.edu.

Isaac Reyes-Vera is a Graduate Student, Department of 

Agronomy and Horticulture, New Mexico State University, Las 

Cruces, NM.

Abstract—A steady decline of perennial bunch grasses in arid 

rangelands has resulted in losses of productivity and germplasm. 

Remediation is costly and rarely successful. Cryptic symbiotic 

fungi, structurally integrated with cells and organs of native plants 

cannot be separated from host plant tissue. However, they were 

successfully transferred from cell cultures of native plants to other 

native grasses and tomato. Fungi in recipient plants were heritable and substantially enhanced vigor, biomass and reproductive 

potential. The transfer of novel fungal symbionts to other native 

or crop plants may offer an alternative method of improving plant 

performance without modifying native plant germplasm.

Introduction_______________________

Native perennial bunch grasses have steadily declined 

in arid southwestern USA rangelands in the past century 

as a result of complex biotic and abiotic stress. Attempts to 

re-establish them have been costly and rarely successful. 

On the other hand, invasive woody shrubs and grasses that 

produce seed and establish seedlings, particularly during 

years of below normal precipitation, have effectively invaded these ecosystems. Disturbed native rangelands have 

suffered substantial losses in productivity, soil and native 

plant germplasms.

Mycorrhizal fungi and fungal endophytes of cool season 

grasses have demonstrated roles in regulating ecosystem 

structure, growth and tolerance to biotic and abiotic stress 

(Smith and Read 1997; Clay and Schardl 2002). Recent 

studies suggest that symbiotic microbes are important 

components that have regulated the sustainability of native 

ecosystems. Bouteloua eriopoda Torr. (black grama) and 

Symbiotic Fungi that Influence Vigor, Biomass and 

Reproductive Potential of Native Bunch Grasses for 

Remediation of Degraded Semiarid Rangelands

Jerry R. Barrow, Mary E. Lucero, and Isaac Reyes-Vera

Atriplex canescens (Pursh) Nutt. (fourwing saltbush), a native 

grass and shrub of the northern Chihuahuan Desert, were 

extensively colonized by non-destructive endophytic fungi 

(Barrow and others 1997; Barrow and Aaltonen 2001; 

Barrow 2003). Specific fungal staining revealed a novel group 

of fungi that differed morphologically from traditionally 

recognized hyphae, spores, and fruiting bodies in all leaf 

and root cells (Barrow and Aaltonen 2001; Barrow 2003). 

Fungal interfaces with physiologically important cells, such 

as photosynthetic, vascular, and stomata suggests that these 

fungi may significantly regulate physiological processes to 

insure ecological success in this water and nutrient stressed 

desert ecosystem.

Traditional experimental protocol to test the efficacy of 

a plant associated microbe requires comparing plants with 

and without the microbe. Plant associated microbes could 

not be removed from germinating seedlings using standard 

disinfestation methods. Therefore we attempted to regenerate 

microbe free B. eriopoda plants from embryonic meristem 

cells of disinfested germinating seedlings (Osuna and Barrow 

2004). Unexpectedly, regenerated plants were intrinsically 

integrated with different fungal symbionts (Barrow and 

others 2004; Lucero and others 2006). Light and electron 

microscope studies of cells and tissues of all phases of the 

regeneration process revealed a continuous fungal presence 

from precursor embryonic meristem cells to regenerated 

plants. After thousands of observations of native plants over 

years and seasonal variation and following their transfer 

from single cells to regenerated plants, key endosymbionts 

were determined to be structural components of all B. eriopoda cells, tissues, and organs suggesting co-evolutionary 

development. These symptomless endophytes have escaped 

detection because they are not always expressed as hyphae 

or commonly recognized fungal structures. Scanning electron 

microscopy revealed a fungal enmeshed biofilm that protectively encapsulated axenically grown callus tissue. Pirttila 

and others (2002) reported similar fungal biofilms in callus 

cultures derived from Pinus bud cultures. This structural 

integration of fungi with cells and tissues prevented their 

removal by traditional sterilization or separation methods. 

To test their influence in plant performance, fungi were 

transferred from cell cultures of native plants to roots of 

germinating seedlings of non-host native grasses and crop 

plants. Here we report a general response of fungi integrated 

with cell cultures of native desert plants when transferred 

to other native grasses and tomato that resulted in a phenomenal enhancement of vigor, biomass and reproductive 

potential.
84 USDA Forest Service Proceedings RMRS-P-52. 2008

Barrow, Lucero, and Reyes-Vera Symbiotic Fungi that Influence Vigor, Biomass and Reproductive Potential of Native Bunch Grasses...

Materials and Methods______________

Seed Sources

The following plant materials were used in this experiment. Black grama, Bouteloua eriopoda Torr, (BOER) and 

sand dropseed, Sporobolus cryptandrus (Torr.) Gray (SPCR) 

were obtained from Plants of the Southwest, Santa Fe, NM. 

Seed of fourwing saltbush, Atriplex canescens (Pursh) Nutt., 

(ATCA), alkali sacaton, Sporobolus airoides Torr. (SPAI) 

and creosote bush, Larrea tridenta (DC.) Coville (LATR) 

were collected from native populations on the USDA-ARS 

Jornada Experimental Range in south-central New Mexico. 

Tomato seed, Lycopersicon esculentum Mill. var. Bradley was 

purchased at a local farm and ranch supply.

Callus Initiation

Callus was initiated from BOER, SPCR, SPAI, ATCA and 

LATR, all dominant grasses and shrubs of the northern 

Chihuahuan Desert (Osuna and Barrow 2004; Barrow and 

others 2004). Seeds of each species were surface disinfested by 

soaking for 10 min in 50 percent ethanol, followed by 25 min 

in 50 percent bleach. Seeds were rinsed three times in sterile 

distilled water, and allowed to germinate on Murashigie 

and Skoogs medium (MS) (Murashige and Skoog 1962). 

Shoot apexes were dissected from seedlings (15 d old) and 

transferred to fresh plates of MS supplemented with 1 mg/l 

2,4-D for callus induction. The calluses were maintained 

with monthly subcultures to fresh MS medium with the 

same growth regulator concentration. Fungal presence 

was detected in all cultures initiated from all native plant 

sources by either microscopy or by isolation. Endophytes 

isolated from some callus lines are described in Lucero and 

others (2006).

Fungal Transfer to Non-Host Plants

Systemic fungal endophytes structurally integrated 

with cell, tissues and organs of native plants could not be 

separated from host plants. Therefore, to determine their 

function, they were transferred to non-host BOER and SPCR 

and tomato. Seeds of each non-host species were surface 

disinfested as described above and were germinated on MS 

media and incubated at 26 °C in a growth chamber receiving light (151μmol m-2s-2) over a 12-hour/day photoperiod. 

Seedlings with emerging radicles were aseptically transferred 

to callus cultures of each donor species, with the radicle in 

direct contact with the callus for 20 days. Fungi from BOER, 

SPAI, LATR and ATCA were transferred to tomato. Fungi 

from ATCA, BOER, and SPAI were transferred to SPCR and 

reciprocal fungal transfers were made from callus tissue of 

ATCA, SPCR and SPAI to non-host BOER seedlings.

Seedling Establishment

After 20 days in contact with fungal integrated callus, 

healthy grass seedlings of non-host recipient plants were 

transplanted to 2-inch peat pots containing a mixture of 

50 percent commercial potting soil and 50 percent soil collected from the field site located on the Jornada Experimental 

Range. Tomato seedlings were transplanted to commercial 

potting soil. Plants were grown for 20 days in an environmental chamber described above, then transferred to a 

heated greenhouse for 45 days. During this time, seedlings 

were irrigated with tap water as needed.

Plant Culture

Tomato plants were transferred to 30 x 38 cm pots with 

commercial potting soil and grown to maturity in the greenhouse. Native grasses were transplanted to a field site on 

the Jornada Experimental Range. The field site (32 x 50 ft) 

was fenced to exclude large animals and rodents, and tilled 

to a depth of 6 inches. Drip irrigation lines were installed, 

18 inches apart, with emitters spaced 18 inches apart. A bag 

of steer manure (1.415 m3) (“Grow King,” Western Organics) 

was uniformly incorporated into the soil, and RoundupTM 

herbicide was applied according to manufacturer’s instructions 2 weeks prior to planting to reduce weeds. Seedling flats 

were moved to the field site and maintained in peat pots for 

1 week for acclimation prior to transplanting. Seedlings were 

planted by hand on 18-inch centers at each water emitter. 

BOER and SPCR plants were planted in a randomized block 

design with 16 plants per replicate and three replicates per 

endophyte treatment. Edge effects were reduced by placing 

two additional rows of grasses around the entire plot. Water 

was applied once a week for 45 min (approximately 1 liter/

plant). Weeds were removed manually with a scuffle hoe or 

by hand as required throughout the growth season.

Results and Discussion______________

Callus tissues derived from native plants, examined by 

selective staining and light microscopy and also scanning 

electron microscopy, revealed that all callus tissues were 

encapsulated with a protective fungal biofilm. Immediately 

upon contact with seedling germinants, fungal hyphae was 

observed with a stereo microscope growing from the callus 

cultures to the seedling radical, followed by swelling of the 

root tissue and induction of lateral rootlets. Within a week, 

substantial increases in rooting and shooting were observed 

in both tomato and seedling grasses.

Phenomenal responses were observed in tomato. Figure1a 

compares tomato control plants (the top row) with plants with 

fungi transferred from SPAI, BOER, LATR and ATCA. In 

each of these cases plants receiving fungi were more vigorous and ranged from 3- (BOER) to 8- (LATR) fold increases 

in root and shoot biomass. Substantial increases in root 

and shoot branching were also observed. Plants receiving 

fungi had more and larger stems and leaves with greater 

chlorophyll and tissue phosphorous in the leaves. It can be 

seen from the photos that there is good uniformity within 

each transfer, yet there are differences between fungal treatments. This suggests that each native plant species transfers 

a different fungal symbiont that differentially influences the 

performance of the recipient host.

Figure 1b shows an initial response of fungal transfers 

from ATCA and SPAI to SPCR, where greater root and shoot 

biomass was observed within 3 weeks of fungal transfer. 

This response was generally observed in both grasses, but 

was greatest in the SPCR recipient plants. Plants in the 
USDA Forest Service Proceedings RMRS-P-52. 2008 85

Symbiotic Fungi that Influence Vigor, Biomass and Reproductive Potential of Native Bunch Grasses... Barrow, Lucero, and Reyes-Vera

Figure 1—(A) Tomato plants var. 

Bradley with fungi transferred from 

sand dropseed (SPAI), black grama 

(BOER), creosote bush (LATR) and 

fourwing saltbush (ATCA) on bottom 

row compared to control plants without 

fungi in top row of each photo. All 

plants with transferred fungi are larger 

than control plants. (B) Sand dropseed 

plants with fungi transferred from 

fourwing saltbush (ATCA) and alkali 

sacaton (SPAI). Substantially greater 

root and shoot biomass is observed 

in each transfer. (C) Comparison of 

seed harvested from sand dropseed 

plants with fungi transferred from black 

grama (1.25 times more seed); alkali 

sacaton (4.5 times more seed) and 

fourwing saltbush (6 times more seed. 

In each transfer, emerging shoots 

were larger than control germinants 

in the far right photo.

A

B

C
86 USDA Forest Service Proceedings RMRS-P-52. 2008

Barrow, Lucero, and Reyes-Vera Symbiotic Fungi that Influence Vigor, Biomass and Reproductive Potential of Native Bunch Grasses...

field receiving symbiotic fungi expressed greater vigor, 

crown diameters, and seed production, and less mortality 

than control plants receiving no fungi. The experimental 

procedure and statistical analysis of these traits are being 

prepared for publication elsewhere.

One significant example, reported here, is seed production 

of SPCR receiving fungi from ATCA, SPAI and BOER compared to the controls (fig. 1c). Seed production from plants 

with fungi from ATCA was 6 times greater, fungi from SPAI 

was 4.5 times greater and fungi from BOER was 1.25 times 

greater than control plants. Germinants from each fungal 

transfer were more vigorous, with larger and greener shoots 

than control germinants.

The multiple benefits of these indigenous fungal symbionts 

confer similar benefits as are reported for mycorrhizal fungi 

(Smith and Read 1997) and endophytes of cool season grasses 

(Clay and Schardl 2002), and contribute multiple benefits 

to their host plants. These include enhanced nutrition and 

water economy, and tolerance to abiotic stress, pathogens 

and herbivores. Redman and others (2002) demonstrated 

phenomenal increases in heat tolerance. Waller and others 

(2005) showed that barley inoculated with an endophyte, 

Piriformospora indica, discovered in plants of the India 

Thar desert, exhibited enhanced productivity and tolerance 

to salt and pathogens.

Multiple benefits conferred by symbiotic microbes in 

agriculture are well documented, but their incorporation 

into mainstream agriculture has generally met with difficulties. The transfer of structurally integrated symbiotic 

fungi indigenous to native grasses and shrubs generally, but 

not uniformly, induced substantial and multiple benefits 

to recipient native grasses. These included enhanced vigor, 

biomass, seed production, and seedling vigor in the second 

generation. Indigenous fungi apparently have a major role 

in regulating performance of native host plants. Increased 

vigor, biomass, and reproductive potential conferred by these 

novel fungi provide attractive and alternative approaches 

to improving native grasses for the remediation of degraded 

arid rangelands.

Our findings indicate that once transferred to other nonhost species, such as other native grasses or tomato, they 

become vertically transmitted or heritable units in their 

new hosts. Thus the benefits are transferred to succeeding 

generations. The conferral of multiple benefits described 

in this paper far exceeds any expectations resulting from 

gene transfers by conventional breeding or recombinant 

DNA. Fungal transfer via cell cultures is simple and the 

results are immediate in the first generation, compared to 

high inputs of capital, time, and technological skill required 

for breeding or recombinant DNA. Another advantage is 

that time tested plant and fungal materials adapted to 

arid rangelands are recombined, resulting in synergistic 

plant performance. Plant genomes are not modified, for 

example BOER or SPAI remain genetically unchanged, but 

are enhanced by the integration of unique symbiotic fungi. 

This method may allow the preservation of these valuable, 

yet declining germplasms. Cell cultures of nine different 

native grasses and shrubs (Reyes-Vera, unpublished data) 

have been generated and each has transferable symbiotic 

fungi. Hundreds of native plant species adapted to these 

arid rangelands represent a vast untapped reservoir of 

genetic variability for improvement of both native and crop 

plants. This research has only scratched the surface of the 

potential of symbiotic microbes. More research is required to 

understand their value in improving native and crop plants 

in dynamic stressed ecosystems.

References

Barrow, J.R.; Havstad, K.M.; McCaslin, B.D. 1997. Fungal root 

endophytes in fourwing saltbush, Atriplex canescens, on arid 

rangelands of southwestern USA. Arid Soil Research Rehabilitation. 11: 177-185.

Barrow, J.R.; Aaltonen, R.E. 2001. A method of evaluating internal 

colonization of Atriplex canescens (Pursh) Nutt. roots by dark 

septate fungi and how they are influenced by host physiological 

activity. Mycorrhiza. 11: 199-205.

Barrow, J.R. 2003. Atypical morphology of dark septate fungal 

root endophytes of Bouteloua in southwestern USA rangelands. 

Mycorrhiza. 13: 239-247.

Barrow, J.R.; Osuna-Avila, P.; Reyes-Vera, I. 2004. Fungal endophytes intrinsically associated with micropropagated plants 

regenerated from native Bouteloua eriopoda Torr. and Atriplex 

canescens (Pursh) Nutt. In Vitro-Cell Developmental Biology—

Plant. 40: 608-612.

Clay, K.; Schardl. C. 2002. Evolutionary origins of ecological consequences of endophyte symbiosis with grasses. The American 

Nataturalist. 160: s99-s127.

Lucero, M.E.; Barrow, J.R.; Osuna, P.; Reyes I. 2006. Plant-fungal 

interactions in arid and semi-arid ecosystems: Large-scale impacts from microscale processes. Journal Arid Environments. 

65: 608-612.

Osuna-Avila P.; Barrow, J.R. 2004. Regeneration of black grama 

(Bouteloua eriopoda Torr.) via somatic embryogenesis. In VitroCell Developmental Biology—Plant. 40(3): 303-310.

Pirttila, A.M.; Laukkanen, H.; Hohtola. A. 2002. Chitinase production in pine callus (Pinus sylvestris L.): A defense reaction against 

endophytes? Planta. 214: 848-852.

Redman, R.S.; Sheeran, K.B.; Stout, R.G.; Rodriguez, R.J.; Henson, 

J.M. 2002. Thermotolerance generated by plant/fungal symbiosis. 

Science. 298: 1581.

Smith, S.E.; D.J. Read, eds. 1997. Mycorrhizal Symbiosis, Ed 2. San 

Diego, CA: Academic Press: 59-60.

Waller, F.; Achatz, B.; Baltruschat, H.; Fodor, J.; Becker, K.; Fischer, 

M.; Heler, T.; Huckelhoven, R.; Neumann, C.; von Wettstein, D.; 

Franken P.; Kogel, K. 2005. The endophytic fungus Piriformospora 

indica reprograms barley to salt-stress tolerance, disease resistance, and higher yield. Proceedings of the National Academy of 

Science. 102: 13386-13391.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein
USDA Forest Service Proceedings RMRS-P-52. 2008 87

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas 

A.; Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Mary Lucero is Research Molecular Biologist, USDA-ARS Jornada 

Experiment Range, Las Cruces, NM; email: malucero@nmsu.edu.

Jerry R. Barrow is Research Scientist, USDA-ARS Jornada 

Experimental Range, New Mexico State University, Las Cruces, NM; 

email: jbarrow@nmsu.edu.

Ruth Sedillo is Biological Science Technician, USDA-ARS 

Jornada Experimental Range, New Mexico State University, Las 

Cruces, NM; email: rusedill@nmsu.edu.

Pedro Osuna-Avila is Professor-Investigado, Universidad Autonoma de Cd. Juarez, Juarez, Mexico; email: posuna@uacj.mx.

Issac Reyes-Vera is PhD Candidate, Department of Plant and 

Environmental Sciences, New Mexico State University, Las Cruces, 

NM; email: isareyes@jornada.nmsu.edu.

Abstract—Obligate fungal endophytes form cryptic communities in 

vascular plants that can defy detection and isolation by microscopic 

examination of reproductive structures. Molecular detection by PCR 

amplification of fungal DNA sequences alone is insufficient, since 

target endophyte sequences are unknown and difficult to distinguish 

from sequences already characterized as plant DNA. We have successfully separated fungal and plant ribosomal DNA sequences by 

amplifying plant-extracted DNA with polymerase chain reaction, and 

separating sequences with denaturing gradient gel electrophoresis 

(DGGE). The resulting electrophoregrams theoretically produce specific bands unique for each organism present in a plant-endophyte 

community. This method has successfully identified endophyte 

sequences in B. eriopoda and A. canescens, and has tracked these 

endophytes as they are transferred to novel hosts.

Introduction_______________________

The ability to transfer uncultured fungal endophytes to 

novel hosts provides us with unprecedented opportunities 

for developing plant material capable of restoring disturbed 

landscapes (Barrow and others, this proceedings). The potential for heritable endophytes to dramatically modify the 

competitive ability of a plant and its progeny also has implications for traditional agriculture and plant biotechnology. 

However, detection and characterization of endophytes in 

plant material can be a formidable task. This is attributed to 

the obligate nature of many endophytes, the diverse nature 

Genetic Characterization of Uncultured 

Fungal Endophytes from Bouteloua eriopoda 

and Atriplex canescens

Mary Lucero, Jerry R. Barrow, Ruth Sedillo, Pedro Osuna-Avila, and Issac Reyes-Vera

of plant-fungal communities (Ganley and others 2004; 

Lucero and others, in review; Vandenkoornhuyse and others 2002), and the tendency of fungi to propagate in planta 

without forming telltale structures such as fruiting bodies 

(Barrow 2003). For this reason, only a handful of studies 

examining the potential long-term ecological impacts of 

endophyte modified plants are available (Clay and Holah 

1999; Clay and others 2005).

In vitro propagation of regenerated plants restricts the 

complexity of the endophyte community by eliminating 

those microbes that are strictly associated with the surface 

layer of plant tissues. In previous reports, light microscopy 

combined with staining has been utilized to examine fungal 

colonization of A. canescens and B. eriopoda (Barrow and 

Aaltonen 2003). While staining methods reveal the presence 

of fungi, they fail to distinguish between fungal species. When 

fruiting bodies and cell walls are not present (as is often 

the case in planta), molecular tools are required to evaluate the nature and complexity of endophyte communities. 

However, care must be taken when probing plant tissues for 

endophyte DNA. Because endophytes are often undetected 

in plant tissues, it is possible that multiple sequences previously characterized as plant genes are actually coming from 

endophytes (Camacho and others 1997).

To truly distinguish plant from endophyte DNA, it is often 

necessary to separate endophytes from their plant hosts. 

However, obligate endophytes do not survive apart from 

host plants. For this reason, we are using combinations of 

denaturing gradient gel electrophoresis and an in-vitro system for transferring uncultured endophytes to foreign host 

plants (Barrow and others, this proceedings) to isolate and 

identify putative fungal DNA sequences. Efforts reported 

herein focus on optimization of methods for separation of 

unique, potentially fungal rDNA sequences extracted from 

regenerated plants. Efforts to amplify fungal sequences 

directly from in vitro propagated plants have focused on 

B. eriopoda. Efforts to separate fungi from host plants 

through endophyte transfer, followed by identification of 

novel bands in recipient plant fingerprints have utilized 

both A. canescens and B. eriopoda.

Methods __________________________

DNA Isolation and Amplification

DNA was isolated from shoot or callus of regenerated B. eriopoda and Atriplex canescens using a MoBio 

UltraCleanTM Plant DNA Kit. PCR was carried out using 
88 USDA Forest Service Proceedings RMRS-P-52. 2008

Lucero, Barrow, Sedillo, Osuna-Avila, and Reyes-Vera Genetic Characterization of Uncultured Fungal Endophytes ...

HotStarTaq Master Mix (Qiagen, Inc.). Primer pairs and 

annealing temperatures are described in table 1. All 

primers were purchased as custom oligos from Operon 

Technologies.

Cloning and Sequencing

The amplified PCR products were cloned into pCR 2.1TM

plasmids using an Invitrogen TA Cloning KitTM. Individual 

clones were sequenced using universal M13 forward and 

reverse primers and a BigDyeTerminator v3.1 Cycle Sequencing Kit (Applied Biosystems) in conjunction with an 

Applied Biosystems 3100 Genetic Analyzer.

Sequence Analysis

Sequences obtained from clones were evaluated for 

chromatogram purity, then subjected to BLASTn analysis 

(Altschul and others 1997) against the non-redundant GenBank nucleotide database using the BLASTn interface at the 

National Center for Biotechnology Information (NCBI).

Endophyte Transfer

Tomato seeds (Lycopersicon esculentum var. Bradley) 

were surface disinfested to remove competing microbes by 

slowly vortexing in 95% ETOH for 1 min, followed by soaking for 7 min in a 5 percent solution of commercial bleach 

(ChloroxTM) and 0.01 percent Tween 20. Seeds were rinsed 

thoroughly in sterile water before plating on Murashige and 

Skoog’s (MS) medium (Murashige and Skoog 1962). Plants 

were allowed to germinate for 7 days.

Callus cultures of B. eriopoda and A. canescens were 

prepared as previously described (Lucero and others 2006), 

then transferred to hormone free MS media, where they 

were maintained for 30 days prior to inoculating tomato 

seedlings. This was done to minimize effects of residual 

growth regulators utilized to maintain callus.

After this time, callus was transferred to new plates of 

MS medium, along with germinated tomato seedlings. Care 

was taken to place seedling radicles in physical contact 

with callus. Co-culturing of tomato seedlings with callus 

continued for 1 week, after which endophytes from callus 

were presumed to have transferred to the tomato seedlings. 

Tomato plantlets were transferred to magenta boxes containing MS medium, where they were maintained for 4 weeks. 

During this time, young leaves were removed and DNA was 

extracted as described above.

Denaturing Gradient Gel 

Electrophoresis (DGGE)

DNA extracted from callus cultures of B. eriopoda and 

A. canescens, from uninoculated tomato plants, and from 

tomato plants inoculated with uncultured endophytes from 

the callus cultures was amplified using the primer sequences 

5’CGCCCGGGGCGCGCCCCGGGCGGGGCGGGGGCCCTTMTCATYTAGAGGAAGGAG3’ and 5’CCGCTTATTKATATGCTTAAA3’. PCR was carried out as described above, 

except that an initial annealing temperature of 57 °C was 

used for the first 10 cycles, followed by 52 °C for 20 cycles. 

Products resulting from this amplification were separated 

by denaturing gradient gel electrophoresis (DGGE) using 

a BioRadTM DCode Universal Mutation Detection System. 

Manufacturer’s protocols were modified by pouring a 

40-60 percent gradient. Amplicons were loaded onto the 

gradient gel and separated at 60 °C, with 75 volts for 16.25 h. 

Gels were stained with Sybr Green and photographed. Other 

GC clamped primers shown in table 1 have also been utilized 

for DGGE; however, less complex banding patterns were 

obtained with other primers.

Results ___________________________

DNA Sequence Comparisons

In a recent manuscript, we described four endophytes 

isolated or identified in Bouteloua eriopoda (Lucero and 

others 2006). None of these isolates were detected using 

the universal ITS primers shown in table 1. This is not 

surprising, since PCR is highly biased towards amplification 

of those sequences most similar to the primers. Sequences 

Table 1—GenBank Accessions of putative ITS sequences amplified from aseptically propagated, regenerated B. eriopoda. Accession numbers 

are accompanied by PCR primers utilized for amplification, annealing temperature utilized in PCR, and GenBank Accessions of sequences that produced the closest alignments to amplifed sequences when compared against GenBank using BLASTn. Sequences 

that align with B. gracilis most likely represent the B. eriopoda ITS rDNA gene.

GenBank Forward Primer Reverse Primer

Accession sequence or reference sequence or reference Tz Closest BLAST match

DQ497245 ITS1/ (White et al., 1990) ITS4/ (White et al., 1990) 53 AF019835.1 Bouteloua gracilis (grass)

DQ497244 ITS1/ (White et al., 1990) ITS4/ (White et al., 1990) 53 AF019835.1 Bouteloua gracilis (grass)

DQ645748 CCTTMTCATYTAGAGGAAGGAG CCGCTTATTKATATGCTTAAA 55 AF019835.1 Bouteloua gracilis (grass)

DQ649069 ITS3/ (White et al., 1990) CCTGTTTGAGTGTCATGAAA CC 55 AY079521.1 Calectasia cyanea (non-

 grass monocot)

DQ649074 ITS3/ (White et al., 1990) CCTGTTTGAGGTCATGAAACC 55 AY753717.1 Begonia alpina (dicot)

DQ649068 ITS5/ (White et al., 1990) ITS4/ (White et al., 1990) 53 AF019835.1 Bouteloua gracilis(grass) 

DQ649075 ITS5/ (White et al., 1990) ITS4/ (White et al., 1990) 53 AF378603.1 Moringa longituba (dicot)

DQ645749 CCTTMTCATYTAGAGGAAGGAG CCGCTTATTKATATGCTTAAA 55 AY640053.1 Corydalis saxicola (dicot)
USDA Forest Service Proceedings RMRS-P-52. 2008 89

Genetic Characterization of Uncultured Fungal Endophytes ... Lucero, Barrow, Sedillo, Osuna-Avila, and Reyes-Vera

with BLASTn homology to grasses, particularly to Bouteloua 

sp., were expected, since similarities exist between fungal 

and plant ITS regions. At present, more specific PCR primers designed to uniquely target the isolated species are 

being used in combination with microscopy (Barrow and 

others, this proceedings) to validate the persistence of these 

endophytes.

The fact that more than one ITS sequence resembled 

Bouteloua is easily explained by the highly repetitive nature 

of ITS DNA. When multiple copies of a gene are present, 

numerous alleles are no surprise.

Two notable features observed herein were that half of the 

sequences obtained from PCR exhibited BLAST homology to 

non-grass monocot and dicot plants, and that sequences of 

endophytes previously isolated from these cultures (Lucero 

and others 2006) were not detected.

Sequences resembling plant species that have never been 

handled in our laboratory were unlikely to represent. We 

hypothesize that these sequences represent uncultured, obligate fungal endophytes, and that ITS rDNA sequences from 

related endophytes, present in the plant species Calectasia 

cyanea, Begonia alpina, Moringa longituba, and Corydalis 

saxicola have been mistaken for ITS sequences from the 

host. Additional support for this hypothesis is being pursued 

through DGGE analysis.

Meanwhile, the absence of sequences from endophytes 

believed to be present in the regenerated plants (Lucero 

and others, in review) may indicate bias inherent in PCR. 

Sequence specific probes have since been developed and 

are being optimized to selectively amplify Moniliophthora, 

Aspergillus ustus, Engyodonium album, and an unknown 

endophyte with sequence homology to a Mycospharella species, respectively.

DGGE

Figure 1 reveals the potential diversity of PCR products 

amplified from aseptically cultured plant materials. To identify the source material from which each band is derived, 

bands may be excised from the gel, reamplified, cloned, and 

sequenced. In this image, the majority of the bands shown 

in lanes containing tomato DNA have exhibited sequence 

homology to tomato ITS sequences and probably originate 

in the plant genome. However, the arrows reveal two novel 

bands in tomato inoculated with B. eriopoda callus that 

co-migrate with bands amplified from B. eriopoda DNA 

(Lanes 3 and 5). These bands are expected to reveal the identity of uncultured endophytes. They have been excised from 

the gels, reamplified, and cloned, but await sequencing.

Many bands that appear on DGGE gels are too faint 

to pick up in photographed images. Such bands, present 

in tomato inoculated with A. canescens, co-migrated with 

bands from A. canescens callus (not shown). These bands 

also appeared in the offspring of tomato plants inoculated 

with A. canescens.

Discussion ________________________

Barrow and others (this proceedings) have discussed 

beneficial impacts transferred fungal endophytes may 

have on host plant survival and fitness. Implications for 

restoration of shrub and grasslands are profound. Yet direct 

evidence demonstrating which endophytes have migrated to 

inoculated plants has been difficult to obtain. The atypical 

morphology of plant endophytes has made them difficult to 

observe microscopically, so it is not yet known what species 

of endophytes have migrated from native plant callus to 

foreign hosts.

Initially, efforts to identify endophytes present in native 

plants utilized in vitro propagation and isolation of fungi 

from internal plant tissues (Barrow and Aaltonen 2001; 

Osuna and Barrow 2004). More recently, PCR, cloning, and 

sequencing of ITS rDNA from both the isolated endophytes 

and the source plant tissues have suggested that both 

Figure 1—DGGE permits separation of highly similar ITS 

rDNA amplified from the following plant and endophyte sources 

(lanes left to right): (1) tomato, (2) tomato inoculated with 

A. canescens callus, (3) tomato inoculated with B. eriopoda

callus, (4) A. canescens callus, (5) B. eriopoda callus, and 

(6) offspring of tomato inoculated with A. canescens callus. 

The high resolution of DGGE permits detection of DNA bands 

in inoculated tomato plants that appear to have come from 

the callus source. Dark bands indicate DNA present in tomato 

plants that appear to have come from B. eriopoda. Light 

bands indicate DNA. Alternate primer sets reveal different 

bands (not shown).
90 USDA Forest Service Proceedings RMRS-P-52. 2008

Lucero, Barrow, Sedillo, Osuna-Avila, and Reyes-Vera Genetic Characterization of Uncultured Fungal Endophytes ...

B. eriopoda and A. canescens host not one, but many endophytes (Barrow and others, this proceedings; Lucero and 

others 2006; Lucero and others, in review). The complexity 

of these systems impedes identification of individual fungi, 

particularly when they cannot be reproducibly isolated from 

the plant tissue. Without sequence data to guide development 

of selective primers, it has been necessary to use universal 

primers that amplify both plant and fungal DNA.

The transfer of endophytes to foreign hosts, followed by 

comparison of DGGE profiles between plants pre-and postinoculation has provided a powerful, albeit tedious, tool for 

endophyte isolation. We now have the ability to determine 

which endophyte(s) are enhancing plant growth, and to isolate 

DNA specific for such endophytes. As we acquire sequence 

data describing these endophytes, we will be able to develop 

PCR assays that can monitor the presence of transferred endophytes across generations. Such assays will be crucial for 

evaluating the persistence and long term ecological impacts 

of endophyte-modified plants in field settings.

References ________________________

Altschul, S.F.; Madden, T.L.; Schäffer, A.A.; Zhang, J.; Zhang, Z.; 

Miller, W.; Lipman, D.J. 1997. Gapped BLAST and PSI-BLAST: 

A new generation of protein database search programs. Nucleic 

Acids Research. 25: 3389-3402.

Barrow, J.; Aaltonen, R. 2003. A staining method for systemic 

endophytic fungi in plants. In: Lartey, R.T.; Caesar A.J., eds. 

Emerging concepts in plant health management: Kerala, India: 

Research Signpost: 61-67.

Barrow, J.; Lucero, M.; Reyes-Vera, I.; Havstad, K. 2007. Endosymbiotic fungi structurally integrated with leaves reveals a lichenous condition of C4 grasses. In Vitro Cellular & Developmental 

Biology—Plant. 43: 65-70.

Barrow, J.R. 2003. Atypical morphology of dark septate fungal root 

endophytes of Bouteloua in arid southwestern USA rangelands. 

Mycorrhiza. 13(5): 239.

Barrow, J.R.; Aaltonen, R.E. 2001. Evaluation of the internal 

colonization of Atriplex canescens (pursh) nutt. roots by dark 

septate fungi and the influence of host physiological activity. 

Mycorrhiza. 11(4): 199.

Camacho, F.J.; Gernandt, D.S.; Liston, A.; Stone, J.K.; Klein, A.S. 

1997. Endophytic fungal DNA, the source of contamination in 

spruce needle DNA: Molecular Ecology. 6(10): 983-987.

Clay, K.; Holah, J. 1999. Fungal endophyte symbiosis and plant 

diversity in successional fields: Science. 285(5434): 1742-1744.

Clay, K.; Holah, J.; Rudgers, J.A. 2005. Herbivores cause a rapid 

increase in hereditary symbiosis and alter plant community 

composition. Proceedings of the National Academy of Sciences 

of the United States of America. 102(35): 12465.

Ganley, R.J.; Burunsfeld, S.J.; Newcombe, G. 2004. A community 

of unknown, endophytic fungi in western white pine. Proceedings of the National Academy of Sciences of the United States of 

America. 101(27): 10107-10112.

Lucero, M.E.; Barrow, J.R.; Osuna, P.; Reyes, I. 2006. Plant-fungal 

interactions in arid and semiarid ecosystems: Large scale impacts 

from microscale processes. Journal of Arid Environments. 65(2): 

276-284.

Lucero, M.E.; Barrow, J.R.; Osuna-Avila, P.; Reyes-Vera, I. [Under 

revision]. A cryptic microbial community persists within micropropagated Bouteloa eriopoda (Torr.) Torr. Plant Science.

Murashige, T.; Skoog, F. 1962. A revised medium for rapid growth 

and bioassays with tobacco tissue cultures. Physiologia Plantarum. 15: 473-497.

Osuna, P.; Barrow, J.R. 2004. Regeneration of black grama (Bouteloua 

eriopoda Torr. Torr) plants via somatic embryogenesis. In Vitro 

Cellular and Developmental Biology—Plant. 40(3): 303-310.

Vandenkoornhuyse, P.; Baldauf, S.L.; Leyval, C.; Straczek, J.; Young, 

J.P.W. 2002. Evolution: extensive fungal diversity in plant roots. 

Science. 295(5562): 2051.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein
USDA Forest Service Proceedings RMRS-P-52. 2008 91

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas 

A.; Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Cristina Juran (at the time of this report) was a Graduate Student, Department of Plant and Animal Sciences, Brigham Young 

University, Provo, Utah; email:livingwild@gmail.com.

Bruce A. Roundy is Professor, Department of Plant and 

Animal Sciences, Brigham Young University, Provo, Utah; email: 

bruce_roundy@byu.edu.

James N. Davis is Project Leader, Utah Division of Wildlife 

Resources, Shrub Sciences Laboratory, Provo, Utah; email: jimdavis@

utah.gov.

Abstract—We sampled unchained and chained areas in 2004 and 

2005 on the Henry Mountains that had been aerially seeded after 

the Bulldog Fire of 2003. Establishment of seeded grasses was high 

on unchained and chained areas although chaining increased seeded 

grass establishment on some sites. Western yarrow established well 

on unchained areas. Initially, high seedling emergence in 2004, 

followed by high precipitation in late winter 2005, produced 15 to 

25 percent total seeded plant cover on both unchained and chained 

areas by summer of 2005. Aerial seeding at lower elevations and 

with less precipitation usually requires some form of seed coverage 

to be successful.

Introduction_______________________

Wildfire is a natural disturbance in the Western United 

States that has been actively suppressed and also reduced 

by removal of fine fuels by grazing of domestic livestock 

(Blackburn and Tueller 1970; Burkhardt and Tisdale 

1976; Gruell 1986; Horman and Anderson 1999; Miller and 

Wigand 1994; Naillon and others 1999; Young and Evans 

1981). Fire suppression increased the spread and density 

of pinyon pine (Pinus edulis and P. monophylla) and Utah 

juniper (Juniperus osteosperma). Under more frequent fire 

return intervals, pinyon and juniper trees were mainly found 

in dense stands on areas of dissected topography or had a 

savanna-like distribution (Burkhardt and Tisdale 1969; 

Miller and Rose 1994; Miller and Tausch 2001; Miller and 

Wigand 1994). These trees are now encroaching into big 

sagebrush (Artemisia tridentata) communities on alluvial 

fans (Naillon and others 1999).

Wildfire Rehabilitation Success With and 

Without Chaining on the Henry Mountains, Utah

Cristina Juran, Bruce A. Roundy, and James N. Davis

As pinyon and juniper density increases in an area, biomass of understory perennial vegetation is greatly reduced 

(Arnold and others 1964; Barney and Frischnecht 1974; Clary 

1971; Dye and others 1995; Miller and others 1995; Tausch 

and Tueller 1990; Tress and Klopatek 1987; Vaitkus and 

Eddleman 1991; Wilson and Schmidt 1990); the native seed 

bank is also r educed (Koniak and Everett 1982; MacDonald 

1999; Poulsen and others 1999; West and Van Pelt 1987). 

Along with a reduction of perennial understory biomass, there 

is also a change in the species composition from perennials to 

annuals, with cheatgrass (Bromus tectorum) as a prominent 

species (Armentrout and Pieper 1988; Bunting and others 

1999; Everett and Koniak 1981; Poulsen and others 1999; 

Schott and Pieper 1985).

Dense pinyon-juniper forests eventually burn (Goodrich 

and Rooks 1999), leaving little native understory to reestablish on the site. Crown fires, in these dense stands, 

occur under very hot and dry conditions (Bunting and others 

1999; Erskine and Goodrich 1999; Gruell 1999; Swetnam 

1988), and kill rhizomes, roots, and seeds of other species 

(Zwolinski 1988). If cheatgrass is present, it will use soil 

water and nutrient resources made available by the fire and 

dominate the disturbed site (Evans and Young 1978; Harris 

1967; Hironaka 1994; Nasri and Doescher 1995a,b; Pyke and 

Novak 1994; Young and Evans 1978). Thus, after fire in a 

dense pinyon-juniper stand, recovery of native understory is 

slow and the area is often “dominated by introduced annuals” 

(Erskine and Goodrich 1999). The lack of sufficient perennial understory to recover after fire indicates that a biotic 

threshold has been crossed (Tausch 1999a, b; Whisenant 

1999). Restoration of such areas may require substantial 

investment in weed control and revegetation.

To prevent weed dominance, soil erosion, and costly restoration, burned areas are rehabilitated by seeding as soon 

as possible after fire, using funds from the Emergency Fire 

Rehabilitation (EFR) program of the Federal Government. 

The goal of the EFR program is to prevent soil erosion and 

weed establishment after a fire (Utah BLM 2001). Burned 

areas are preferably seeded with a rangeland drill to best 

control distribution, rate, and depth of seeds. But drills 

cannot be used on steep slopes, rocky soils, or those with 

obstacles such as tree skeletons (MacDonald 1999). Burned 

forests are usually aerially seeded. An anchor chain may be 

dragged one-way over the ground after seeding (Thompson 

and others 2006) to help cover the seed and increase the number of seeds in “safe sites” (Harper and others 1965; Young 

1988). Chaining has been shown to increase establishment 

of seeded species and reduce weed dominance after aerial 

seeding burned areas between 1,615 and 1,810 m elevation 

(MacDonald 1999; Ott and others 2003).
92 USDA Forest Service Proceedings RMRS-P-52. 2008

Juran, Roundy, and Davis Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah

Burned areas of the Henry Mountains were seeded after 

the Bulldog fire burned more than 12,000 ha in the summer 

of 2003. The BLM administers more than 11,000 ha of the 

burned area with 7,700 ha designated as Wilderness Study 

Area (WSA). Only native species are permitted to be seeded 

and no mechanical treatments are allowed on WSA sites. 

The Utah Division of Wildlife Resources (UDWR) also administers sections of the burned area owned by the State of 

Utah. Since much of the burned area had moderate to steep 

slopes and was rugged, it was aerially seeded. Some sites 

were chained, while steep slopes, archaeological sites, and 

the WSA were left unchained. Chaining costs may vary and 

equipment transport costs were high for this remote site. 

Chaining for this project cost $141/ha. The cost of chaining 

and difficulty in its implementation led to the question of 

whether it was really necessary to chain after aerial seeding 

at relatively higher elevation (>1,900 m) burned sites such 

as those on the Henry Mountains. The objective of this study 

was to compare seeded species establishment on unchained 

and chained sites on the Henry Mountains and generally 

evaluate the success of the Bulldog fire rehabilitation on 

seeded areas above 1,900 m elevation.

Study Sites and Methods ____________

The Henry Mountains are located in the southeast quarter 

of Utah. The Bulldog fire burned between the elevations of 

1,800 and 3,400 m (Leroy Mead, Personal communication, 

Utah Division of Wildlife Resources, Price, Utah). Ignited 

by a four-wheeler, the fire burned from the second week of 

July and was not considered out until the end of August 2003 

(Brant Hallows, personal communication, Richfield Field 

Office, Bureau of Land Management). Plant communities 

before the fire were mostly pinyon pine (Pinus edulis) and 

Utah juniper (Juniperus osteosperma), with understory 

shrubs that included Wyoming big sagebrush (Artemisia 

tridentata var. wyomingensis), mountain big sagebrush 

(Artemisia tridentata var. vaseyana), or black sagebrush 

(Artemisia nova), or mixed conifer woodlands with little 

understory and cheatgrass present on some sites. The burned 

area has moderate to steep slopes, with sandy loam dioritic 

and sedimentary soils that are locally highly erosive.

Burned areas were aerially seeded using fixed-wing aircraft in November and December 2003. Seeding included 

four different seed mixes (table 1). The WSA mix consisted 

of only native species while the other three mixes had both 

native and introduced species. Designated seeded areas were 

chained from the end of March to the end of April 2004. Normally, chaining after fall seeding is completed by December. 

An ely chain was pulled by two D8 and D9 bulldozers 30 m 

apart and in a J or U shape. An ely chain is an anchor chain 

with railroad segments welded perpendicular to some of 

the links to provide additional weight and disturbance of 

the soil (Vallentine 1989). Different sites were seeded with 

different seed mixes depending on whether they were on 

UDWR or BLM land, and whether or not they were in the 

WSA. The Aspen, Benson A, and Coyote sites were seeded 

with the same seed mix on both unchained and chained areas, 

but those mixes were different for each site. The Benson B 

and Cass Creek sites were seeded with the same mix at 

each site, but those mixes were different for unchained and 

chained areas. The Aspen Creek site was seeded with the 

UDWR low elevation mix on both unchained and chained 

areas. The Coyote Creek site was seeded with the UDWR 

high elevation mix on both unchained and chained areas. 

The Benson A site was seeded with the BLM non-WSA mix 

on both unchained and chained areas. The Benson B and 

Cass Creek sites were seeded with the BLM-WSA mix on 

unchained areas in the Wilderness Study area and with the 

BLM non-WSA mix on the chained area.

Vegetation Sampling

In the spring of 2004 five paired study sites were chosen 

for intensive sampling. Four of these sites were intensively 

sampled August 2004 and 2005. One of the sites was omitted 

after the first year because of inability to distinguish between 

chained and unchained areas. Intensively sampled study 

sites were located on the southern half of the mountains 

between 2,100 and 2,700 m elevation (table 2). Study sites 

included adjacent unchained and chained transects with 

similar aspect, slope, and elevation.

Intensive Sampling—On each site, two or three permanently marked 120–150 m transects were laid out contiguously. Perpendicular to each transect, five 30-m cross-belts 

were centered at random distances from the end. On each 

cross-belt, 20 0.25 m2 quadrats were placed every 1.5 m. 

In each quadrat, canopy cover and density were recorded. 

Canopy cover was estimated using a modified Daubenmire 

(1959) cover class method (Bailey and Poulton 1968) with 

seven cover classes: (1) 0.01 to 1 percent, (2) 1.1 to 5 percent, 

(3) 5.1 to 25 percent, (4) 25.1 to 50 percent, (5) 50.1 to 75 

percent, (6) 75.1 to 95 percent, (7) 95.1 to 100 percent. Cover 

was calculated by averaging the midpoints of each cover class 

across all quadrats for a site. Cover was recorded for vegetation (total and individual species), litter, rock, pavement, 

cryptograms, and bare ground. Density was only recorded 

for perennial species.

Shrub cover was measured using the line intercept method 

on all the cross-belts (Bonham 1989). Shrub density was 

measured with belt transects centered over the length of 

each cross-belt. Each shrub belt transect was 30 m in length 

and 1.34 m wide, for a total of 0.004 ha. Burned tree density 

was also measured on the unchained sites at points spaced 

30 m apart along the baseline using a point-quarter method. 

At each point, the distances to the center of the nearest tree 

skeleton base was measured for each quarter of the area 

centered on the point. A BLM rain gage at Coyote Bench 

provided precipitation data (data on file Henry Mountains 

BLM office, Hanksville, Utah).

Extensive Sampling—Seven additional WSA-unchained 

and non-WSA chained paired sites, as well as 14 non-paired 

WSA-unchained sites were extensively sampled. The 

paired WSA-unchained transects, seeded with the WSA 

mix, were compared to chained transects seeded with the 

BLM non-WSA mix. Density of seeded shrubs, grasses, and 

forbs, as well as that of residual perennial plants and annual plants, were recorded in 40 0.25 m2 quadrats placed 

1.5 m apart along a 60-m transect in each treatment. Forb 

and shrub data were collected by species, but seeded perennial grass species were counted in total and not by species. 
USDA Forest Service Proceedings RMRS-P-52. 2008 93

Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah Juran, Roundy, and Davis

Table 1—Bulk seeding rates (kg/ha) for species seeded after the 2003 Bulldog fire on the Henry Mountains, Utah. (UDWR = Utah 

Division of Wildlife Resources; high = high elevation mix; low = low elevation mix; BLM = Bureau of Land Management; 

WSA = Wilderness Study Area mix; non = non-WSA mix; N = species native to the United States; I = species introduced 

to the United States.)

 UDWR UDWR BLM BLM 

Seeded species Accession Scientific name N or I high low non WSA

Grasses

Sandberg bluegrass MH Poa secunda N 1.1

Mountain brome Bromar Bromus marginatus N 1.1 3.5

Arizona fescue Redondo Festuca arizonica N 0.6 

Hard fescue Festuca brevipila I 0.9 

Sheep fescue Festuca ovina I 0.7 

Prairie junegrass Koeleria macrantha N 0.2 

Spike muhly El Vado Muhlenbergia wrightii N 0.07 

Orchardgrass Paiute Dactylis glomerata N 0.9 0.6 

Indian ricegrass Rimrock Achnatherum hymenoides N 0.5

Crested wheatgrass Hycrest Agropyron cristatum x I 0.6 1.1 2.8

 Agropyron desertorum 

Intermediate wheatgrass Thinopyrum intermedium I 1.1 

Pubescent wheatgrass Thinopyrum intermedium I 1.1 

Pubescent wheatgrass Luna Thinopyrum intermedium I 2.7

Slender wheatgrass Pryor Elymus trachycaulus N 1.1 

Slender wheatgrass SanLuis Elymus trachycaulus N 3.8

Tall wheatgrass Alkar Thinopyrum ponticum I 2.1

Thickspike wheatgrass Critana Elymus lanceolatus N 1.1 2.1

Blue wildrye Arlington Elymus glaucus N 1.1 

Russian wildrye Bozoisky Psathyrostachys juncea I 1.1 2.7

Forbs

Alfalfa Ladak Medicago sativa I 1 0.5

Alfalfa Ranger Medicago sativa I 1.1 0.1 

Arrowleaf Balsamroot Balsamorhiza sagittata N 0.03 

Rocky Mountain Beeplant Cleome serrulata N 0.1 0.6 

Small Burnet Delar Sanguisorba minor I 0.8 1.3 

Small Burnet VNS Sanguisorba minor I 1.4 

Blue Flax Appar Linum perenne spp. lewisii I 0.2 0.3 0.2 0.2

Cicer Milkvetch Windsor Astragalus cicer I 0.7 

Sainfoin Onobrychis viciifolia I 1.2 

Bird’s-foot Trefoil Lotus corniculatus I 0.6 

Western Yarrow Achillea millefolium N 0.02 0.2

Shrubs

Bitterbrush Purshia tridentata N a a 

Mountain Big Sagebrush Artemisia tridentata var. vaseyana N 0.6 0.3 

Wyoming Big Sagebrush Artemisia tridentata var. wyomingensis N 0.6 

Fourwing Saltbush Atriplex canescens N a a 0.5

Utah Serviceberry Amelanchier utahensis N a a 

Total rate excluding dribbler mix 12.1 12.3 12.4 10.8

a included in dribbler mix and seeded at 1.1 kg/ha over part of the area.
94 USDA Forest Service Proceedings RMRS-P-52. 2008

Juran, Roundy, and Davis Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah

Table 2—Locations of intensively-sampled fire rehabilitation sites, associated soil series and classifications (Downs 1990) on the Henry Mountains, Utah.

NAD 27 UTM coordinates

Sites Northing Easting Soil series Classification

Aspen Creek 4198420 525550 Circleville-Blazon complex Aridic Argiborolls

 Ustic Torriorthents

Benson 4199230 527690 Montosa Aridic Argiustolls

Cass Creek 4194930 523980 Duneland-Mido complex Ustic Torripsamments

Coyote Creek 4202750 521860 Duneland-Mido complex Ustic Torripsamments

Cover of rock, litter, and bare ground was occularly estimated 

in cover classes in these quadrats in 2004 and 2005. Cover of 

seeded grasses, shrubs, and forbs, along with that of annual 

and residual perennial plants, was also occularly estimated 

in cover classes in 2005. Extensive sampling was done in 

late June 2004 and 2005.

Analysis

Second year seeding success was categorized for all intensively and extensively sampled WSA-unchained, unchained, 

and chained transects according to plant density: <0.5 seeded 

plants/m2 = failure, 0.5 to 2 seeded plants/m2 = moderately 

successful, and >2 seeded plants/m2 = successful. Density 

means for 0.25 m2 quadrats were converted to plants/m2 by 

multiplying by 4.

Because most sites were seeded with different seed 

mixes, initial analysis for both the intensively sampled and 

extensively sampled data used different sampling sites as 

random blocks and compared categories of seeded species: 

total seeded grasses, total seeded forbs, and the sum of 

seeded grasses and forbs as herbaceous vegetation. For the 

intensively sampled data, two initial analyses were conducted in this manner: (1) using the Aspen, Benson A, and 

Coyote sites as blocks, and (2) a separate analysis using the 

Benson B and Cass Creek sites as blocks. To better sort out 

responses to chaining, analysis of sites and treatments was 

subsequently conducted for categorical vegetation responses 

using transects as pseudoreplicates for the intensively 

sampled data, and quadrats as pseudoreplicates for the 

extensively sampled data (Hurlbert 1984). This allowed us 

to determine site by treatment interactions. Reasonable 

inference using pseudoreplication is dependent on having 

similar site characteristics on both treated and untreated 

areas. Our sampled areas were selected to be similar in site 

characteristics. Finally, density and cover of dominant seeded 

species were analyzed for unchained and chained areas for 

each site separately, using transects as pseudoreplicates 

for the intensively sampled data and quadrats as replicates 

for the extensively sampled data. Mixed model analysis 

(Littell and others 1996) was used for analysis of all extensively and intensively sampled data. Sites or transects 

were considered random for the intensively sampled data 

and sites or quadrats were considered random for the 

extensively sampled data. Data for the 2004 and 2005 

sampling years were analyzed separately, and intensively 

and extensively sampled data were analyzed separately. All 

individual quadrats were considered subsamples for treatment analysis, except when considered as pseudoreplications 

for the extensively sampled data.

Although seed mix was confounded with chaining treatment or among different sites or blocks, comparisons are 

useful because land management agencies will continue to 

seed different mixes on different sites. Native seed mixes will 

be sown on wilderness study areas that cannot be chained, 

while native and introduced species will be sown on nonwilderness study areas that can be chained.

Results and Discussion______________

The relationship of increasing pinyon-juniper density 

and cover, and correspondingly decreasing understory species density and cover varies with soil depth and climate 

(Miller and Tausch 2001). A post-fire stand reconnaissance 

showed that the burned pinyon and juniper communities 

on the Henry Mountains generally had limited residual 

understory plants with a relatively high tree density of 

912 trees/ha. Median rock cover for the intensively sampled 

sites was 23.6 percent. From January 2004 to August 2004, 

74 mm of precipitation helped seeded plants to establish 

and reach an average of 6.5 percent cover by August 2004 

on the intensively sampled sites. Between September 2004 

and August 2005, the sites received 662 mm of precipitation, which supported high plant survival and significantly 

increased cover the second year after seeding (fig. 1). Overall, 

plant establishment from the fire rehabilitation seeding was 

highly successful.

WSA General Success

The two WSA sites that were intensively sampled 

(Benson B and Cass Creek) had >2 seeded plants/m2

on every transect the second year after seeding and were 

therefore considered to be successfully seeded. Total seeded 

species cover was over 25 percent on the intensively 

sampled sites (fig. 2).

Of the 21 extensively sampled WSA transects, 19 had 

>2 seeded plants/m2 by the second year and were considered successfully seeded. These transects averaged about 

30 percent cover of all seeded species (fig. 3). Because the 

two failed transects were on similar soils and at similar 
USDA Forest Service Proceedings RMRS-P-52. 2008 95

Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah Juran, Roundy, and Davis

A

A

A

A

B 

A A

0 

5 

10 

15 

20 

25 

30 

Grass Forb Herbaceous Total 

Cover (%) 

WSA Unchained 

Chained 

A 

A

A 

A

B 

A 

0 

5 

10 

15 

20 

25 

Grass Forb Herbaceous

Plants/m2 

WSA Unchained 

Chained 

A

Figure 1—Monthly precipitation at BLM 

weather station at Coyote Bench (data 

on file at Henry Mountains BLM office, 

Hanksville, Utah) and estimated from 

PRISM (Parameter-Elevation Regression on 

Independent Slopes Model available at http://

forest.moscowfsl.wsu.edu/engr/climatesw.

html) for the Aspen Creek sampling site, 

Henry Mountains, Utah. Annual amounts 

in mm were 254 for 2003–2004, 484 for 

2004–2005, 273 for the 26-year average, 

and 304 for the PRISM average.

Figure 2—Density and cover of seeded species on intensively 

sampled WSA unchained and chained fire rehabilitation sites 

in the Henry Mountains, Utah in 2005. Total is all herbaceous 

seeded vegetation plus the shrubs on the sites. Different letters 

for pairs of bars indicate a significant difference (p < 0.05).

350

300

250

200

150

100

50

0

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

Precipitation (mm)

2003-2004

2004-2005

26 year average

PRISM average

elevations as the successful transects (fig. 4), we conclude 

that these transects were either missed or on the margin 

of the aerial seeding operation. Davis and Harper (1990), 

Goodrich and Rooks (1999), and Thompson and others (2006) 

observed a reduction of weedy species with successful establishment of seeded species, and increased weed cover with 

seeding failure. We did not observe high densities or cover 

of cheatgrass on failed (evidently unseeded) extensively 

sampled sites, but found Coyote tobacco (Nicotiana attenuata) 

and pinyon groundsmoke (Gayophytum ramossisimum) to 

be dominant plants there. Native winter annuals are easily 

replaced within a short time by cheatgrass. Bare ground 

and available resources on some of these sites could result 

in cheatgrass invasion and dominance in the future.

Density of seeded species was much higher in 2004, the 

first year after seeding, than in 2005 (fig. 5). Seedings always decrease in density as established plants mature and 

use more resources per plant. Density of seeded grasses in 

2005 was significantly (p > 0.01) correlated with 2004 density (r2 = 0.41), which emphasizes the importance of initial 

seedling establishment.

Unchained Versus Chained Transects: 

Extensively Sampled Sites

Chained treatments had significantly (p < 0.05) greater 

grass density and cover than unchained treatments in 

both 2004 and 2005 (fig. 6). Unchained treatments had 

significantly higher seeded forb cover (p < 0.05) and density 

(p < 0.0561) than chained treatments in 2005. For the pseudoreplication analysis of 2005, data, site, chaining treatment, 

and the site by chaining treatment interaction was significant 

(p < 0.001) for seeded grass cover and density and seeded forb 

cover and density. Seeded grass cover percentages were 

greater on chained than unchained treatments on three sites 

(Stanton Pass = 20.5 ± 1.7 chained, 7.8 ± 1.7 unchained; Dark 

Canyon = 21.6 ± 1.7 chained, 6.4 ± 1.7 unchained; and Cass 

Creek = 10.5 ± 1.7 chained, 1.1 ± 1.7 unchained).
96 USDA Forest Service Proceedings RMRS-P-52. 2008

Juran, Roundy, and Davis Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah

Figure 3—Number of sites out of 21 that had different densities of seeded 

grasses, forbs, and total plants (above) and percent cover (below) associated 

with those density categories for unchained Wilderness Study Area fire 

rehabilitation sites in the Henry Mountains, Utah, in 2005.

2005

0

5

10

15

20

25

30

35

1900 2000 2100 2200 2300 2400 2500

Elevation (m)

Plants/m2

Grasses

Forbs

Total

Figure 4—Density of seeded species on 21 extensively sampled fire rehabilitation sites 

in the Henry Mountains, Utah, in relation to site elevation. Each point represents 2004 

and 2005 density data at a site.
USDA Forest Service Proceedings RMRS-P-52. 2008 97

Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah Juran, Roundy, and Davis

0

0.5

1

1.5

2

2.5

3

0 5 10 15 20 25 30 35 40

Plants/m2 2004

Plants/m2 2005

Grasses

Forbs

Total

Figure 5—Density in 2004 and 2005 on 21 extensively sampled fire rehabilitation 

sites in the Henry Mountains, Utah.

Figure 6—Density and cover of seeded grasses and forbs in 2005 on 

seven extensively sampled unchained and chained paired sites in the Henry 

Mountains, Utah. Paired bars with different letters are significantly different 

(p < 0.05 for all but forb density where p < 0.0561).

B 

A 

A 

A

B 

A 

0 

2 

4 

6 

8 

10 

12 

14 

16 

18 

Grasses Forbs Total

Plants/m2 

Unchained 

Chained 

B 

A 

A 

A 

A 

B 

A 

A 

0 

5 

10 

15 

20 

25 

Grasses Forbs Total Cheatgrass

Cover (%) 

Unchained 

Chained 
98 USDA Forest Service Proceedings RMRS-P-52. 2008

Juran, Roundy, and Davis Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah

Seeded grass density (plants/m2) was greater on chained than 

unchained treatments at four sites (Aspen Creek = 11.6±1.0 

chained, 6.0 ± 1.0 unchained; Dark Canyon = 14.8 ± 1.0 chained, 

3.6 ± 1.0 unchained; Corral Canyon = 15.6 ± 1.0 chained, 

10.0 ± 1.0 unchained; and Cass Creek = 10.8 ± 1.0 chained, 

1.6 ± 1.0 unchained) of the seven site-pairs. Because the 

chained areas were seeded with introduced and native 

grasses (BLM non-WSA mix), and the unchained areas were 

seeded with only native grasses (BLM WSA mix), it is not 

possible to determine whether greater density and cover 

of seeded grasses on chained than unchained treatments 

for the extensively sampled sites was a result of chaining, 

better establishment of introduced species than native species, or both.

Seeded forb cover and density were also significant 

(p<0.001) for site, chaining treatment, and the interaction of 

site and chaining treatment for the pseudoreplication analysis. Seeded forb cover percentages were greater on unchained 

than chained treatments for two of the seven sites sampled 

(Stanton Pass = 2.1 ± 1.9 chained, 11.9 ± 1.9 unchained; Cass 

Creek = 15.4 ± 1.9 chained, 27.1 ± 1.9 unchained). Seeded 

forb density (plants/ m2) was greater for unchained than 

chained treatments on three of the seven site-pairs sampled 

(Corral Canyon = 4.0 ± 1.2 chained, 11.6 ± 1.2 unchained; 

Cass Creek = 9.2 ± 1.2 chained, 17.2 ± 1.2 unchained; and 

Bulldog Peak = 16.8 ± 1.2 chained, 24.0 ± 1.2 unchained). All 

extensively sampled unchained treatments were WSA sites. 

Western yarrow (Achilleamillefolium) was only seeded on the 

WSA unchained areas (BLM-WSA mix, seeded at 0.2 kg/ha) 

and was not designated to be seeded on the unchained areas 

that were extensively sampled (BLM non-WSA mix). It was 

the dominant forb encountered in sampling both unchained 

and chained areas and averaged 10.4 percent cover across all 

WSA-unchained sites. The range of yarrow relative cover on 

unchained sites was 0–96 percent with a mean of 45 percent. 

Yarrow cover on chained sites was 6.8 percent. The range 

of yarrow relative cover on chained sites was 0–100 percent 

with a mean of 39 percent. Western yarrow on many chained 

extensively sampled sites must have drifted onto these sites 

from aerial seeding on the adjacent WSA. Higher densities 

and cover of western yarrow on unchained than chained 

extensively sampled sites may be because chaining reduced 

yarrow establishment; but is much more likely because the 

unchained area had many more seeds than did the chained 

area where seeds landed by drift or misapplication.

Where chained areas on some sites had greater density 

and cover of seeded grasses than on unchained sites, seeded 

forb cover and density (mainly yarrow) was greater on some 

unchained than chained sites. Even with differences in cover 

and density of seeded grasses and forbs among sites, most 

of the extensively sampled sites were considered successful 

with the establishment of >2 seeded plants/m2. Seeded 

plant density was not correlated with elevation (fig. 4).

Unchained Versus Chained Transects: 

Intensively Sampled Sites

Sites as Blocks—Total seeded plant cover and density 

for both 2004 and 2005 were similar between unchained and 

chained treatments when analyzing the Aspen, Benson A, 

and Coyote sites containing the same seed mix on each 

treatment of the site, but different seed mixes between 

sites (table 3). Seeded perennial plant density in 2005 was 

>2 plants/m2 on almost every belt of these transects. Most 

belts averaged ≥9 perennial plants/m2. An exception was 

one belt that averaged two perennial plants/m2 on the Aspen 

Creek site. Other belts on the unchained treatment at that 

site also had lower densities of perennial plants than the 

rest of the belts on the site. We conclude that parts of this 

site were either missed or marginally seeded. There was also 

a higher density of annual weeds on these belts. Average 

cheatgrass cover on the Aspen site was 1.9 ± 0.3 percent on 

the unchained transects and 5.2 ± 0.4 percent on the chained 

Table 3—Significance of sites and chaining treatments on density and cover of seeded perennial grasses and forbs in 2005 

after the Bulldog fire on the Henry Mountains, Utah.

Sites sampled Density Cover

and analysis Effects Grass Forb Total Grass Forb Total

Sites as blocks

Aspen, Benson A, Coyote Chaining NS NS NS NS NS NS

Cass Creek, Benson B Chaining NS 0.04 NS NS NS NS

Transects as 

psuedoreplicates

Aspen, Benson A, Coyote Site NS NS NS NS NS NS

Chaining NS NS NS NS NS NS

Site×chaining 0.01 0.01 0.01 NS 0.01 0.01

Cass Creek, Benson B Site NS NS NS 0.05 NS NS

Chaining 0.05 0.05 NS NS 0.05 NS

Site×chaining 0.05 NS 0.05 0.01 NS 0.01
USDA Forest Service Proceedings RMRS-P-52. 2008 99

Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah Juran, Roundy, and Davis

transects, suggesting that more of the chained than unchained 

area was missed during the seeding. Second year data on 

these sites showed no significant differences (p > 0.05) in 

density and cover between unchained and chained treatments when native and introduced species where analyzed 

separately (fig. 7).

Analysis of the Benson B and Cass Creek sites, which were 

seeded with different mixes on each treatment but the same 

seed mixes between sites, showed that seeded forb cover 

and density were higher on WSA-unchained than chained 

treatments in 2005 (fig. 2). The higher forb density on the 

WSA transects is a result of high establishment of western 

yarrow (table 4), which was not designated to be seeded on 

the chained areas at these two sites.

Overall, established species included a high proportion of 

seeded species on each site whether unchained, chained, all 

native seed mix, or a native and introduced seed mix (table5). 

All sites included perennial species that were not in the seed 

mix to be seeded on that specific site, but evidently received 

seeds from another mix seeded in an adjacent area, due to 

seeding overlap, drift, or both.

Site by Treatment Interactions—For the Aspen, 

Benson A, and Coyote sites, there was a significant site 

by chaining treatment interaction for all variables tested, 

except for seeded perennial grass cover (table 3). Perennial 

grass cover was similar on unchained and chained treatments for all sites, but perennial grass density was higher 

on chained than unchained treatments at the Benson A 

site (fig. 8). Crested wheatgrass (Agropyron cristatum or

desertorum) had higher cover on chained than unchained 

areas at the Benson A site, but the reverse was true at the 

Aspen site (table 4). Higher density of crested wheatgrass 

on the unchained Aspen site may have been due to a small 

residual population of crested wheatgrass observed in early 

spring 2004 at this site. Chaining may have reduced the 

residual crested wheatgrass density at the Aspen site. Very 

few residual perennial grasses were observed in early spring 

2004 after the fire on most areas of this study.

Perennial forb cover and density were similar for unchained and chained treatments at the Aspen site (fig. 8). 

The unchained treatment had higher cover and density than 

the chained treatment at the Benson A site, but the reverse 

A

A

A

A

A

A

A

A

0

2

4

6

8

10

12

14

16

18

20

Native Grass Introduced

Grass

Native Forb Introduced

Forb

Cover (%)

Unchained

Chained

A A

A

A

A

A

A

A

0

2

4

6

8

10

12

Native Grass Introduced

Grass

Native Forb Introduced

Forb

Plants/m2

Unchained

Chained

Figure 7—Effects of chaining on density and cover in 2005 of native and 

introduced species seeded for fire rehabilitation on Aspen, Benson A, and 

Coyote sites on the Henry Mountains, Utah. Native forb data includes blue 

flax as a naturalized forb. Different letters above pairs of bars indicates a 

significant difference (p < 0.05).
100 USDA Forest Service Proceedings RMRS-P-52. 2008

Juran, Roundy, and Davis Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah

Table 4—Cover and density in 2005 of dominant and other selected species seeded on unchained and chained sites on the Henry 

Mountains, Utah. UC = unchained, C = chained.

Aspen- UDWR Benson A- BLM Coyote- UDWR 

low elevation mix non-WSA mix high elevation mix

Species UC C UC C UC C UC C UC C UC C

Cover (%) Density Cover (%) Density Cover (%) Density

 (plants/m2) (plants/m2) (plants/m2)

Grasses

Mountain brome 0b 0.9b 0b 0.4b 0b 0b 0b 0b 0.4 1.4 0.2 2.1

Slender wheatgrass 1.0 0.6 1.0 0.6 2.7a 0 2.1a 0 0.3 2.4 0.9 4.9

Crested wheatgrass 7.2a 3.2 6.4 3.2 2.7 4.3 3.1 7.4 5.7 2.6 5.9 2.3

Intermediate wheatgrass 2.4 4.3 2.4 3.6 1.6 1.9 1.2 1.4 2.5b 2.1b 2.0b 1.4b

Orchardgrass 0.9 0.9 0.7 0.7 0b 0b 0b 0b 1.5 2.7 1.8 3.4

Russian wildrye 0.5 0.2 0.4 0.2 1.1 3.3 0.8 4.6 0.7b 0.3b 0.4b 0.3b

Thickspike wheatgrass 0 0 0 0 0b 0b 0b 0b 0b 0b 0b 0b

Forbs

Alfalfa 2.5 2.9 1.4 1.8 0.1 0.2 0.2 0.1 1.1 0.6 0.6 0.6

Blue flax 0.2 0.7 0.4 0.8 0.7 3.5a 0.5 1.1 0.3 0.2 0.3 0.3

Small burnet 3.5 2.9 2.0 1.8 0b 0b 0b 0b 2.2 7.7 2.6 22.4

Western yarrow 0.4b 0.5b 0.3b 0.4b 14.1ab 2.4b 8.4ab 1.0b 1.7b 0.6b 3.0b 1.1b

Benson B Cass Creek

 UC C UC C UC C UC C

Seed mix BLM BLM BLM BLM BLM BLM BLM BLM

Species WSA Non-WSA WSA Non-WSA WSA Non-WSA WSA Non-WSA

- - - Cover (%) - - - Density (plants/m2) - - - Cover (%) - - - Density (plants/m2)

Grasses

Mountain brome 0.1 0b 0.2 0b 0.4 0.8b 0.4 1.4

Slender wheatgrass 7.0a 0b 5.6 0b 0.7 1.7b 0.7 2.0b

Crested wheatgrass 0.1b 4.3a 0.1b 7.2a 4.2b 5.1 4.4b 6.4

Intermediate wheatgrass 0b 1.9a 0b 1.4a 1.3b 3.1 1.0b 2.4

Orchardgrass 0b 0b 0b 0b 0b 0b 0b 0.1b

Russian wildrye 0b 3.3a 0b 4.6a 0.7b 2.2a 0.1b 0.3a

Thickspike wheatgrass 0 0b 0 0b 0 0b 0 0b

Forbs

Alfalfa 0b 0.2 0b 0.1 0.9b 1.2 0.4b 0.5

Blue flax 0.2 3.5a 0.2 1.1a 0.3 0.2 0.3 0.2

Small burnet 0b 0b 0b 0b 0b 1.1b 0b 0.6b

Western yarrow 14.6a 2.4b 10.0a 1.0 15.7a 3.8b 12.0a 2.0b

a Significantly higher than other treatment (p<0.05)

b Not designated to be seeded on that treatment at that site

was true at the Coyote site. At the Benson A site, western 

yarrow was the dominant forb and was not designated to be 

seeded on either the unchained or chained areas (table 4). 

Higher establishment in the unchained than chained area 

may be a result of a greater number of seeds that were 

applied or drifted onto the unchained than chained area, 

a result of excessive seed burial on the chained area, or a 

result of both factors. At the Coyote site, the dominant forb 

was small burnet (Sanguisorba minor) and its establishment was apparently favored by chaining (table 4). Small 

burnet had similar density and cover on both chained and 

unchained areas at the Aspen site (table 4). Two accessions 

of small burnet were seeded for a total of 2.2 kg/ha at the 

Coyote site (UDWR high elevation mix) compared to only 

one accession seeded at 1.3 kg/ha (UDWR low elevation mix) 

at the Aspen site (table 1).

For the Cass Creek and Benson B sites seeded with the 

BLM WSA mix on unchained areas and seeded with the 

BLM non-WSA mix on chained areas, there were significant site, chaining treatment, and interaction effects (table 

3). Perennial grass density was higher on chained than 

unchained areas at both sites, but cover was greater on 

chained areas only at the Cass Creek site (fig. 9). Both 

unchained and chained areas at these sites were seeded 

at a similar rate for perennial grasses (table 1) (10.5 kg/ha 

for unchained and 10.8 kg/ha for chained areas), but were 

seeded with different species. Therefore, we cannot determine 

if greater grass establishment on chained areas was a result 
USDA Forest Service Proceedings RMRS-P-52. 2008 101

Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah Juran, Roundy, and Davis

Table 5—Relative cover (percent) on seeded fire rehabilitation sites on the Henry Mountains, Utah.

 Aspen Cass Coyote 

Creek Benson Benson Creek Creek

Sites UDWR BLM BLM WSA-unchained UDWR

Seeding mix low non-WSA BLM non-WSA-chained high

Unchained

Dominant perennial grasses

Crested Wheatgrass 7.2 2.7 0.1 4.2 5.7

Intermediate Wheatgrass 2.4 1.6 0.01 1.3 2.5

Slender Wheatgrass 1 2.7 7 0.7 0.3

Total seeded grass cover 12.3 8.1 7.4 7.5 13.3

Dominant perennial forbs

Western Yarrow 0.4 14.1 14.6 13.7 1.7

Blue Flax 0.2 0.7 0.2 0.3 0.3

Alfalfa 2.5 0.1 0.01 0.9 1.1

Small Burnet 3.5 0 0 0 2.2

Total seeded forb cover 7.4 15.2 14.9 16.9 5.5

Total seeded shrub cover 2.1 0.1 0.02 0.1 0.1

Chained

Dominant perennial grasses

Crested Wheatgrass 3.2 4.3 4.3 5.1 2.6

Intermediate Wheatgrass 4.3 1.9 1.9 3.1 2.1

Slender Wheatgrass 0.6 0 0 1.7 2.5

Total seeded grass cover 10.9 9.4 9.4 17.7 13.5

Dominant perennial forbs

Western Yarrow 0.5 2.4 2.4 3.8 0.6

Blue Flax 0.7 3.5 3.5 0.2 0.2

Alfalfa 2.9 0.2 0.2 1.2 0.6

Small Burnet 3 0.01 0.01 1.1 7.7

Total seeded forb cover 8.2 6.2 6.2 6.4 9.4

Total seeded shrub cover 2.8 0.02 0.02 0.3 0.1

of chaining, associated with use of more successful species, 

or both (table 4).

On the Cass Creek and Benson B sites, seeded perennial 

forb density and cover were greater on unchained than chained 

areas (fig. 9). Western yarrow was the dominant forb encountered on these sites on both unchained and chained areas, even 

though it was not designated to be seeded on the chained areas 

(table 4). Greater establishment of yarrow on the unchained 

areas at these sites may be a result of chaining decreasing 

its establishment, but is more likely because the number of 

seeds that landed there by drift or overlap was much less than 

landed on the unchained area where it was designated to be 

seeded. Chaining could have excessively buried some of the 

yarrow seeds that landed in the chained area.

Several researchers have reported that predation (Nelson 

and others 1970), desiccation, insufficient soil contact, and 

lack of root penetration reduce establishment of plants from 

uncovered seeds (Vallentine 1989; Winkel and others 1991). 

Chaining may help bury seeds and increase water intake by 

loosening the soil surface and increasing porosity, aeration, 

and water infiltration (Cluff and others 1983; Wood and others 

1982). Increased water infiltration may provide an extended 

period of available water for the seeds. Conversely, chaining 

covers seeds at various depths, and emergence of seedlings 

can be inhibited by excessive burial (Lawrence and others 

1991; Vallentine 1989). Although chaining may be beneficial 

or detrimental at times, the amount of precipitation during 

critical growing periods is what most determines the success 

of revegetation on semi-arid zones (Call and Roundy 1991). 

High post-seeding spring precipitation resulted in successful plant establishment for fire rehabilitation on unchained 

and chained areas in this study. MacDonald (1999) also 

found chaining to be unnecessary for successful seedling 

establishment at elevations higher than 2,000 m. However, Ott 

and others (2003) found chaining was critical to success of 

aerially seeded fire rehabilitation sites at elevations under 

1,800 m in central Utah.

Species Comparisons

Cover and relative cover of different species varied 

among sites and treatments in 2005, and some species 

were evident on sites and treatments where they were not 

included in the seed mix (tables 4 and 5). For example, at the 

Cass Creek site, Hycrest crested wheatgrass (Agropyron

cristatum x Agropyron desertorum) had 4.2 percent cover 

on the WSA unchained treatment and western yarrow 

showed 3.8 percent cover on the chained treatment even 

though these species were not included in the seed mixes for 

these respective treatments at this site (tables 1 and 4). 
102 USDA Forest Service Proceedings RMRS-P-52. 2008

Juran, Roundy, and Davis Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah

B

BC

AB

B

AB

C

B

C

BC

BC BC

B

AB

C

B

A

A

A

0

5

10

15

20

25

30

35

40

45

50

Perennial

grass

Aspen Creek

Total Forbs Perennial

grass

Coyote Creek

Total

Plants/m2

Unchained

Chained

AB

C

A

A

AB

BC A

A

A

A

B

B

BC

AB

BC A

A

A

0

5

10

15

20

25

30

Perennial

grass

Aspen Creek

Total Forbs Perennial

grass

Coyote Creek

Total

Cover (%)

Unchained

Chained

A

A

C

A

A

BC

A

B

A

A

AB

B

0

5

10

15

20

25

Perennial

grass

Cass Creek

Total Forbs

Plants/m2

Unchained

Chained

B

A

A

B

A

A

A

B

A

B

B

B

0

5

10

15

20

25

30

Perennial

grass

Cass Creek

Total Forbs

Cover (%)

Unchained

Chained

Figure 9—Density and cover of species seeded 

on intensively-sampled unchained and chained fire 

rehabilitation sites on the Henry Mountains, Utah, in 

2005. Pairs of bars with different letters are significantly 

different (p < 0.05).

Figure 8—Density and cover of species seeded 

on intensively-sampled unchained and chained fire 

rehabilitation sites on the Henry Mountains, Utah, in 

2005. Pairs of bars with different letters are significantly 

different (p < 0.05).
USDA Forest Service Proceedings RMRS-P-52. 2008 103

Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah Juran, Roundy, and Davis

Table 6—Numbers of species seeded and encountered on fire rehabilitation sites on the Henry Mountains, Utah. Some species seeded in adjacent areas were encountered.

 Aspen Cass Coyote 

Creek Benson Benson Creek Creek

Sites UDWR BLM BLM WSA-unchained UDWR

Seeding mix low non-WSA BLM non-WSA-chained high

Perennial grasses

Seeded in mix

Unchained 8 5 4 4 8

Chained 8 5 5 5 8

Seeded in mix and encountered 

Unchained 6 4 3 3 6

Chained 6 4 4 4 6

Not seeded in mix but encountered 

Unchained 2 5 4 7 4

Chained 7 1 1 7 3

Total encountered

Unchained 8 9 7 10 10

Chained 13 5 5 11 9

Perennial forbs

Seeded in mix

Unchained 6 2 2 2 9

Chained 6 2 2 2 9

Seeded in mix and encountered 

Unchained 4 2 2 2 6

Chained 4 2 2 2 6

Not seeded in mix but encountered 

Unchained 1 2 2 2 0

Chained 1 3 3 2 0

Total encountered

Unchained 5 4 4 4 6

Chained 5 5 5 4 6

Shrubs

Seeded in mix

Unchained 2 1 0 0 1

Chained 5 1 1 1 4

Seeded in mix and encountered 

Unchained 1 0 0 0 1

Chained 2 1 1 1 2

Not seeded in mix but encountered 

Unchained 1 1 2 1 1

Chained 0 0 0 2 0

Total encountered

Unchained 2 1 2 1 2

Chained 2 1 1 3 2
104 USDA Forest Service Proceedings RMRS-P-52. 2008

Juran, Roundy, and Davis Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah

Occurrence of seeded species on sites or treatments where 

they were not included in the designated seed mix may be 

explained by close proximity of the different sites or treatments and overlap or drift of the aerial-seeded mixes.

Where crested wheatgrass was included in both the 

chained and unchained seed mixes, its cover on these treatments varied by site. Crested wheatgrass had lower cover 

percentages on chained than on unchained treatments at 

the Aspen and Coyote Creek sites, but greater cover on the 

chained than unchained treatment at the Benson A site 

(table 4). Chaining may have removed some residual plants 

at the Aspen Creek sites as already mentioned. The seeding rate of crested wheatgrass was higher at the Benson A 

site (BLM non-WSA mix, 2.8 kg/ha) than that of both the 

Aspen site (UDWR low elevation mix, 0.6 kg/ha) and the 

Coyote Creek site (UDWR low elevation mix, 1.1 kg/ha). 

Densities on both unchained and chained treatments at all 

sites where crested wheatgrass was seeded were high, and 

this species is considered to be successfully established on 

both treatments.

Density and cover of intermediate, pubescent (Thinopyrum 

intermedium), slender (Elymus trachycaulus) wheatgrasses, 

and Russian wildrye (Psathyrostachys juncea), though not 

significantly different between treatments when analyzed 

separately, all had higher mean density and cover for chained 

than unchained treatments when they were seeded at the 

same rate on both treatments (table 4). Blue flax (Linum 

lewisii) had higher cover on chained than unchained treatments at the Benson A site where it was seeded at the same 

rate on both treatments (table 4). Blue flax has a much larger 

seed than yarrow and might have greater seed coverage with 

than without chaining. At this same site, slender wheatgrass 

had higher cover and density on the unchained than chained 

treatment (table 4). It is not clear whether chaining directly 

reduced its establishment, or if greater establishment and 

possible interference of other seeded grasses on the chained 

area reduced its establishment there. Blue flax and small 

burnet appear to have benefited from chaining on some sites, 

while western yarrow successfully established without chaining. Yarrow had the highest cover of any seeded species on 

the WSA unchained sites, and produced 2.4 to 3.8 percent 

cover on the chained treatments at the Benson B and Cass 

Creek sites, even though it was not included in the seeding 

mix. During aerial seeding, its small seeds could easily have 

blown across the road that formed the boundary between the 

WSA unchained and non-WSA chained treatments. Its small 

seeds are also well-suited to broadcasting on sites where 

rocks and the loose, dioritic surface soil of the Henry Mountains provide safe sites or shallow burial for germination. 

Light is recommended to germinate yarrow for propagation 

purposes and it is recommended for shallow seeding <6 mm 

(Stevens and Monsen 2004). Establishment of small-seeded 

species like yarrow from aerial broadcasting is usually only 

successful when precipitation is consistent enough to provide 

available soil moisture for germination near the surface soil 

for a week or more. Little or no seed burial and a wet spring 

following seeding must have encouraged high germination 

of yarrow on unchained treatments.

Blue flax was seeded on both unchained and chained 

areas at all extensively and intensively sampled sites. On 

the extensively samples sites, it had slightly greater density 

(plants/m2) on chained (0.54 ± 0.1) than unchained (0.29 ± 0.1) 

areas (p < 0.0511). On two of the intensively sampled sites 

(Benson A and Benson B), it had higher cover on chained 

than unchained areas (table 4). Blue flax establishment 

appears to be benefited by chaining.

Mountain big sagebrush was included in the UDWR low 

and high elevation mixes, but only a few seedlings where 

found on both the intensively and extensively sampled 

sites. Other seeded shrubs also had limited establishment. 

Establishment of shrubs in large-scale direct seedings on 

semi-arid rangelands continues to be a challenge.

Conclusions _______________________

Aerial seeding was successful in this study without chaining, but these results may not apply on lower elevation 

sites or on similar sites in years with lower post-seeding 

precipitation. On rehabilitation sites that are at high elevations (>1900 m) and receive high post-seeding precipitation, 

land management agencies may save the cost of chaining. 

Yet, precipitation is not predictable. Ott and others (2003) 

found that aerially seeded fire rehabilitation sites at lower 

elevations (<1,900 m) and in lower precipitation zones had 

limited seeded plant cover and were dominated by annual 

weeds. Seed burial and the microtopographic depressions 

resulting from chaining should increase time of available 

water, enhance germination and seedling survival under 

marginal precipitation. Seedbed modifications in semi-arid 

environments do not ensure revegetation success on dry 

years, may not be necessary on wet years, but could make 

the difference between success and failure in intermediate 

years (Winkel and Roundy 1991).

Confounding of seed mix with chaining treatment in this 

study limits the making of strong conclusions about the 

need for chaining to successfully establish certain species or 

functional groups. Although chaining increased grass density 

on the Benson A site, it did not significantly affect cover on 

this site, nor did it significantly increase seeded grass cover 

or density on the Aspen and Coyote Creek sites, which were 

also seeded with the same grass species on both unchained 

and chained treatments. Density is usually a more important 

measure than cover for determining seeding success, because 

if enough plants are established, they can grow and increase 

cover over time. Grass establishment is especially important 

to the Henry Mountains for forage and soil conservation. 

Chaining was associated with greater seeded grass cover 

and density on seven extensively sampled sites, but this 

may have been more a result of the species seeded on the 

chained areas than a result of chaining itself. Both chained 

and unchained treatments seeded with introduced grasses 

had successful plant establishment. Chained areas at the 

Benson B and Cass Creek sites had greater seeded grass 

densities and cover than unchained WSA areas, but the 

chained areas were seeded with introduced grasses known 

to establish well, while the unchained WSA areas were 

seeded with native grasses. Blue flax was seeded on both 

unchained and chained areas at all extensively sampled and 

intensively sampled sites. Generally higher density of blue 

flax on chained than unchained treatments suggests that 

its establishment was benefited by chaining. Western yarrow was the dominant forb encountered on unchained and 
USDA Forest Service Proceedings RMRS-P-52. 2008 105

Wildfire Rehabilitation Success With and Without Chaining on the Henry Mountains, Utah Juran, Roundy, and Davis

chained treatments and established much better than blue 

flax, even though it was only seeded on the unchained WSA 

area. Except for establishment of western yarrow and blue 

flax, establishment of other forbs and shrubs was limited.

Comparison of unchained and chained plant establishment 

using the same seed mixes on each site would allow stronger 

conclusions about the effects of chaining on revegetation success. 

Also, monitoring seeding success on unchained and chained sites 

using the same seed mix would allow determination of which 

species establish best in the seeded environment.

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USDA Forest Service Proceedings RMRS-P-52. 2008 107

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Dennis M. Bramble is Professor, Department of Biology, University of Utah, Salt Lake City, UT 84112.

Jean C. Bramble, M.S., P.O. Box 554, Escalante UT 84726; 

email: bramble@bioscience.utah.edu

Abstract—Rapid and substantial reductions in the local density of 

invasive rubber rabbitbrush (Chrysothamnus nauseosus) have been 

achieved on a shrub-infested meadow complex solely by manipulating grazing so as to benefit the native meadow vole, Microtus 

montanus. The key adjustment has been a shift from spring-summer 

to late season grazing, thereby allowing sufficient growth of herbaceous vegetation to offer suitable vole habitat. We hypothesize 

that voles were instrumental in defending many historic western 

meadowlands against woody shrub encroachment, but have now 

been widely displaced by livestock competition. This suggests that 

voles could prove an important restoration tool on certain shrubdominated rangelands.

Introduction_______________________

Woody shrub encroachment into grassland is a worldwide 

phenomenon and often attributed to poor land management practices (Archer 1994). In the arid rangelands of 

the Western United States, the transition from grassland 

to shrub-dominated vegetation often began soon after the 

introduction of livestock (McPherson and others 1988). Excessive grazing together with a reduction in fire frequency 

is thought to have been a primary driver of the grassland to 

shrubland transition in many parts of the region. Among the 

most heavily impacted areas have been the grass and sedge 

dominated meadow complexes that flank the riparian zones 

in many parts of the Intermountain West. These “Parariparian Meadow Complexes” (PMCs) vary considerably in size, 

extending from just a few tens of meters to over a kilometer 

to either side of the riparian zone. They also vary in vegetation type and according to water availability. Thus, PMCs 

range from wet through mesic to dry meadow complexes, 

Vole-Driven Restoration of a Parariparian 

Meadow Complex on the 

Colorado Plateau, South-Central Utah

Dennis M. Bramble and Jean C. Bramble

with the gradients between them usually correlated with 

distance and elevation away from the riparian zone. The 

extent of a PMC is strongly influenced by the local hydrology, 

geomorphology and soil characteristics (Jewett and others 

2004). Regardless of size, these areas are typically among 

the most heavily disturbed of all rangeland habitats, owing 

to their ability to provide reliable sources of forage and water, both in close proximity. For the same reason, the PMCs 

were the first to feel the pressure of livestock introductions 

in the mid 1800s and most have continued to be impacted 

by livestock to some degree ever since (Belsky and others 

1999). At the same time, the PMCs are among the most 

valuable of all rangeland environments in the arid West 

because they typically harbor the greatest biotic diversity 

and function as important dispersal and breeding corridors 

for a host of animals, including migrating neotropical birds 

(Martin and Finch 1996). Restoration of degraded PMCs, 

therefore, deserves to be among the highest priorities on 

western rangelands.

Woody shrub invasion into western PMCs is often so 

extensive that most of the original cover of grasses, sedges, 

and forbs has been lost through shrub competition and persistent grazing pressure. Among the most common woody 

shrubs to invade these environments are big sage (Artemisia 

tridentata), rubber rabbitbrush (Chrysothamnus nauseosus) 

and, where soils are alkaline, greasewood (Sarcobatus vermiculatus). Whereas sage is readily controlled by a variety 

of common brush treatment strategies (for example; burning, cutting, herbicides), this is not true of rabbitbrush and 

greasewood. Both of these shrubs are very resistant to most 

standard brush control measures due to their capacity to 

re-grow from the root crowns (Whitson and others 1996). 

Such resistance generally assures that, in the absence of 

considerable effort and expense, a PMC dominated by these 

shrubs will likely remain so. Moreover, as with some other 

shrub-invaded rangelands in the West (Archer 1994), many 

transformed PMCs have probably reached a new alternate 

stable state with respect to vegetation structure and would 

not return to their original condition even if grazing were 

discontinued.

In this preliminary study, we report extraordinary success 

in reducing brush domination in a former meadow complex 

without resort to conventional control techniques. Rather, we 

have relied exclusively on biological control using native small 

mammals as the primary agents of our restoration program. 

In a surprisingly short time, this strategy has substantially 

shifted vegetative structure away from a brush-dominated 

complex to one that is beginning to resemble the original 
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Bramble and Bramble Vole-Driven Restoration of a Parariparian Meadow Complex on the Colorado Plateau, South-Central Utah

mesic and dry meadow to savannah-like vegetation that 

appears to have occupied the site prior to the introduction 

of livestock.

Background

Site History and Characteristics—The South Hollow 

study site occupies approximately 200 acres in the Upper 

Valley region of Garfield County in south-central Utah. It 

has a mean elevation of approximately 7,100 ft (2,165 m) and 

an average annual precipitation of between 12 to14 inches 

(30.5 to 35.6 cm). Roughly 30 to 40 percent of the precipitation arrives as monsoonal rains between mid July and mid 

September. The site consists of an original 160-acre (~ 65-ha) 

homestead patented in 1911 plus roughly 60 acres (~24 ha) 

of rangeland belonging to the Dixie National Forest. The 

rangeland acres are included within the homestead fences 

due to water rights on nearby Upper Valley Creek. The site 

is almost entirely surrounded by an active grazing allotment 

administered by the Escalante District of the Dixie National 

Forest. The whole of the Upper Valley region was subjected 

to intense unregulated livestock (cattle and sheep) grazing 

beginning in the early 1880s and continuing until about 

1910 when a USFS grazing permit system was established 

(Hall 1954). According to historic records, the Upper Valley had no defined creek, but instead displayed a system 

of meadows dotted by numerous seeps and short spring 

runs (Webb 1985). Present day Upper Valley Creek, which 

resides within a deeply entrenched arroyo (fig. 1), appears 

to have been the result of major erosion linked to a series of 

catastrophic floods between 1909 and 1932 (Webb and Baker 

1987). The depth of the arroyo system on and adjacent to the 

study site varies from approximately 6 to 30 feet and has 

resulted in a substantial drop in the original water table. In 

the West, stream incision and the resulting changes in geohydrological processes can have profound effects on riparian 

corridors and their associated meadow complexes, chiefly by 

reducing water availability through lowering the water table 

(Jewett and others 2004). Though several factors, including 

climatic change, may have contributed to arroyo formation 

in the Upper Valley, severe overgrazing of the surrounding 

rangelands and their watersheds was likely an important 

ingredient. Local soil types within the South Hollow study 

area are highly varied and range from deep sandy loams to 

heavy alkaline clays.

Historic Land Office reports indicate that by the 1890s 

greasewood was already beginning to take hold on eroded 

soils in the Upper Valley (Webb 1985). Sage dendrochronology 

(D. Bramble unpublished data) confirms that the invasion 

of South Hollow by big sage was already underway by at 

least 1910. Prior to a 1988 prescribed burn by the former 

owner, the vegetation on the study site consisted chiefly of 

dense stands of mature to senescent basin big sage (A. tridentata ssp. tridentata); the graminoid community was 

overwhelmingly represented by a single introduced bunch 

grass, crested wheatgrass (Agropyron cristatum). The fire, 

which was patchy, succeeded in killing virtually 100 percent 

of the sage where it burned, but left many areas untouched. 

Figure 1—View of north end of South Hollow study site showing extent of most recent shrub kills in this 

area as well as untouched rabbitbrush (RB) on Forest Serve range on the opposite side of the riparian 

zone. The red square marks the approximate location of study Plot II (photographed on July 4, 2005).
USDA Forest Service Proceedings RMRS-P-52. 2008 109

Vole-Driven Restoration of a Parariparian Meadow Complex on the Colorado Plateau, South-Central Utah Bramble and Bramble

The northern portion of the study site, and that containing 

our study plots, was the most thoroughly burned. Following 

the fire, and depending on local soil characteristics, either 

rubber rabbitbrush (RB) or greasewood, quickly invaded 

this area (fig. 1). By the time the property was acquired in 

the fall of 1992, both shrubs were already established in 

dense stands.

Methods __________________________

Grazing

Our restoration project was initiated in 1993 and continues 

at this time. It has utilized only a single restoration tool—the 

manipulation of grazing. Our grazing program departs from 

that traditionally employed on the South Hollow study site 

in three important ways:

1. A reduction in grazing intensity and duration. We 

estimate that our AUMs (Animal-Unit Months), which have 

ranged between 22 and 60, are at least 50 percent lower (and 

very likely more) than that of the historic grazing pattern 

on our study site. There are no actual records of the stocking rate on the property. Our cattle have remained on the 

property from 4 to 8 weeks depending on range conditions. 

Traditionally, the property appears to have been continuously 

grazed from late spring or early summer until fall.

2. Deferred grazing. With a single exception (1999), 

we have elected to shift to late season grazing. Under our 

plan, cows are not allowed onto the property until late 

August or September, and sometimes not until the end of 

September.

3. No grazing in drought years. Thus, the property was 

not grazed in 1996 and 2000. Limited grazing took place 

(late September to early November) in 2002, after summer 

rains broke severe drought conditions.1

Vegetation Monitoring

We have periodically monitored the vegetation through 

field observation and repeat photography since 1993. Our 

only quantitative measurements have come from two study 

plots established in the summer of 2002. Each plot is a cattle 

exclosure 22.9 m (75 ft) square. Plot II is the lower of the 

two and closest (~83 m or 275 ft) to the Upper Valley Creek 

riparian zone. It is also the more mesic plot. Sedge (Carex sp.) 

is a common and increasing element of the herbaceous vegetation within Plot II. Sedge appears to be a good indicator 

of reliable soil moisture at the study site. Plot I is located 

approximately 210 m (700 ft) south of Plot II and thus farther 

(~210 m or 1,000 ft) from Upper Valley Creek. It is slightly 

higher in elevation (~ 25 feet) and drier; Carex has not 

been recorded from this plot. Since plot establishment in 

2002, there have been several brief instances of tresspass 

grazing on Plot II, but Plot I has remained free of cows. 

From 2002 to 2005 the study plots were protected from 

cattle by portable electric fencing. In 2006 permanent post 

and wire fencing was installed around each site.

The shrub composition of the plots has been surveyed at 

three points in time: July 2002, September 2005. and October 

2006. Each plot was divided into 10 equal sections within 

which all established shrubs (>15 cm or 6 inches in height) 

were counted and then added to obtain the total for the 

plot. The species identity of each shrub was noted, but only 

three species of woody shrub occur within these plots (RB, 

big sage, mountain snowberry). Both plots were thoroughly 

burned in the 1988 fire, resulting in 100 percent mortality 

of the existing sage population. In the initial 2002 census, 

we counted all burned sage stumps to determine pre-fire 

shrub density. We also scored all established shrubs (RB) 

with respect to the following categories: (1) live or killed, 

(2) rodent damaged or not, and (3) if damaged, then by 

girdling or burrowing (tunneling beneath the shrub). In the 

2005 census all established shrubs were scored according to 

two categories: (1) killed or alive, and (2) rodent damaged 

or untouched. In the 2006 survey, shrubs were scored only 

as to killed or alive; at this time all dead shrubs exhibited 

clear evidence of rodent damage, including burrowing around 

their bases. Moreover, all RB within the plots were counted 

because by 2006 all plants exceeded 15 cm in height.

The first indication of widespread shrub mortality associated with rodent activity occurred during the period of 

1997 to 1999. The 2002 census reflects this event, because 

there was no detectable vole herbivory on shrubs during the 

dry period of late 1999 to 2003. The second episode of vole 

predation on woody shrubs began with the exceptionally wet 

winter of 2004 to 2005 and continued into 2006 (fig. 2).

1 Additionally, the property was not grazed in 2005 owing to an 

exceptionally wet winter and spring that eliminated the need for 

our lessee to rent our pasture. 

Figure 2—Closer view of vole-killed and damaged RB in 

June 2005. Shrub skeletons from first shrub kill (1997-99) are 

weathered and lack terminal branches. Most of the skeletons 

in this picture represent shrubs killed by voles during the winter 

of 2004 to 2005. Notice re-growth on some severely damaged 

but surviving plants in the background. With the removal of 

the shrub canopy, herbaceous vegetation is expanding in this 

area. The backpack rests on freshly disturbed soil from a vole 

burrow system. 
110 USDA Forest Service Proceedings RMRS-P-52. 2008

Bramble and Bramble Vole-Driven Restoration of a Parariparian Meadow Complex on the Colorado Plateau, South-Central Utah

Results ___________________________

The 2002 census shows that prior to the 1988 fire both 

plots had been dominated by dense, pure stands of basin 

big sage. The fire killed all sage shrubs within the plots, but 

their burned stumps remained and were readily identifiable 

in 2002. By 1993 a large percentage of these stumps had 

become sites of RB establishment and dense stands of this 

shrub, most of which exceeded 3 feet in height (~ 0.9 m), were 

already beginning to form a substantially closed canopy.

The overall trend in shrub numbers on the two study 

plots is illustrated in figure 3. Pre-fire sage densities, as 

reflected in burned sage stumps, differed on the two sites, 

being greater on the drier site, Plot I (fig. 3A). However, 

the replacement of sage by RB following the 1988 fire 

resulted in a larger fraction of stumps being occupied by 

RB in Plot II (76.77 percent) than in Plot I (62.06 percent). 

There was no significant difference, however, in the absolute 

number of established (>15 cm height) RB shrubs on the 

two plots by the time of the first survey in 2002 (table 1). 

Figure 3—Changes in total number of 

established shrubs (>15 cm in height) on 

two study plots over a 19-year interval. 

Big sage (hatched) originally occupied the 

sites, but was removed by fire in 1988 and 

was then replaced by rubber rabbitbrush 

(stippled). Rabbitbrush has been reduced 

during two episodes of shrub killing by voles. 

The number of live shrubs at the beginning 

(1997) and end (1999) of the first episode 

are reconstructed from a survey of live and 

skeletal shrub remains conducted in 2002. 

Recruitment through growth of numerous 

young plants (<15 cm) accounts for the 

increase in shrub density on Plot II between 

2002 and 2005, but no such recruitment 

occurred on Plot I. There was no vole activity, 

shrub death, and little shrub growth during 

the drought period beginning in late 1999 

and ending in late 2002. 
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Vole-Driven Restoration of a Parariparian Meadow Complex on the Colorado Plateau, South-Central Utah Bramble and Bramble

The impact of the first episode of rodent predation on RB 

produced strikingly different levels of shrub mortality on the 

two sites, with nearly 86 percent of the established shrubs 

killed on Plot II compared to just 35 percent on Plot I. Beginning with the second episode of rodent predation in the fall 

of 2004, shrub density on Plot I continued to decline during 

the 2005 and 2006 surveys. In Plot II, however, there was 

actually a large increase in RB density between 2002 and 

2005 (fig. 3B). This was almost entirely attributable to the 

growth of a large cohort of very young plants that were too 

small to be counted as “established” in 2002. These shrubs 

grew rapidly afterward, having been “released” by the near 

elimination of large canopy forming RB during the initial 

round of rodent activity. Compared to 2005, shrub numbers 

had once more declined on Plot II in 2006, and again due 

entirely to rodent predation. Thus, despite the temporary 

spike in shrub numbers in 2005, the total reduction of established RB between the initial (1997) and 2006 density 

is still impressive (47.18 percent). That on Plot I was even 

more dramatic (64.57 percent) (fig. 3A).

Closer examination of the 2002 data offers additional insight into rodent-shrub interactions (table 1). When scored 

for status, the shrubs on our plots were again significantly 

different in several categories: percent killed, damaged by 

girdling, and burrowed. Vole-driven mortality of established 

plants was much higher in Plot II. However, in both plots 

the number of killed shrubs was not significantly different 

from the number having rodent burrows beneath their 

crowns (table 1). This statistic simply confirms what has 

been obvious in our field observations: RB damaged by the 

girdling of their stems and base, even when extensively so, 

are rarely killed. But once serious digging begins under a 

shrub, its eventual death is nearly certain. While burrowing 

was specifically scored only in 2002, it is clear that essentially 

all subsequent RB mortality, both within and external to our 

study plots, has been associated with rodent burrowing.

Discussion ________________________

Our restoration project has produced large reductions in 

RB density in a short period of time (Bramble and Bramble 

2003). Moreover, this success has been achieved without 

resort to any of the traditional methods of shrub control 

(such as burning, disking, herbicide application). Our data 

suggest that local densities of woody shrubs may now be 

approaching pre-grazing levels. Nearly all the reductions 

in invasive woody shrubs, and especially RB, appear to be 

directly linked to the activities of a single species of native 

rodent, the montane meadow mouse or vole (Microtus montanus). This phenomenon, in turn, appears to be a response 

to the manipulation of a single variable, the pattern of 

livestock grazing.

Our grazing protocol compared to the historic pattern at 

the site involves both a reduction in grazing intensity as 

well as a shift in timing. Of the two, timing seems the most 

important. Grazing on the study site has been deferred 

until the fall. Traditionally our pastures, like elsewhere 

in the Upper Valley, have been grazed from late spring/

early summer until fall. With deferred grazing, both cool 

and warm season components of the grass flora have ample 

time to grow, mature, and set seed. In the process, the herbaceous cover attains densities and heights that would not 

otherwise be realized. This change in vegetative structure 

is critical to montane voles. During the warm season this 

Table 1—Results of surveys of established (> 15 cm ht.) rubber rabbitbrush (RB) on 

two 22.9 m (75 ft) square plots conducted in July 2002 (top), together with 

observed shrub densities in September 2005 and October 2006 and reconstructed earlier densities of big sage (1988) and RB (1997).

PLOT I Percent PLOT II Percent

Total Rabbitbrush (RB)

(2002) 175 100.00 195 100.00f

Killeda 61 34.86 167 85.64***

Damaged 64 36.57 17 8.72***

Untouched 50 28.57 11 5.64***

Girdled 124 70.86 159 81.54

Burrowedb 73 41.71 160 82.05***

RB density (1997)c,d 3350 — 3732 —

RB density (2002) 2182 — 563 —

RB density (2005) 1608 — 2699 —

RB density (2006) 1187 — 1971 —

Sage density (1988)e 5397 — 4862 —

a All “killed” shrubs exhibited positive signs of rodent attack. Girdled = stripping of bark from lower 

stems; Burrowed = excavation and gnawing of root crowns. b There is no significant difference ( P >.05 ; χ2 test) between the number of RB burrowed and 

killed within the same plot, reflecting the fact that the two are correlated. c Densities are given as shrubs per hectare.

d Density of RB in 1997 reconstructed from census of live and dead shrubs in 2002. 

e Sage densities in 1988 reconstructed from burned stumps recorded in 2002.

f The total number RB on the two plots was not significantly different in 2002. 

*** Values for which the differences between plots are highly significant (P< .001; χ2 test). 
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Bramble and Bramble Vole-Driven Restoration of a Parariparian Meadow Complex on the Colorado Plateau, South-Central Utah

rodent appears to be limited by environmental conditions, 

especially adequate food, cover, and humidity close to the 

ground. Grazing related reductions in grass density and 

height on the pastures of the Upper Valley during the spring 

and summer generally preclude occupation by voles. There 

has, for example, been little indication of vole activity on the 

Forest Service grazing allotment immediately adjacent to our 

property (fig. 1). This range, which is primarily dense sage 

with crested wheat in the interspaces but also containing 

locally dense stands of RB in close proximity to the north 

end of the study site, continues to be grazed in late spring 

and summer. Montane meadow voles are chiefly grazers and 

hence dependent on the availability of grasses, sedges, and 

forbs. Their food requirements overlap those of livestock to 

a considerable degree. Unlike livestock, however, individual 

voles have a restricted foraging range and are consequently 

more dependent on the local availability of plant resources. 

For these reasons, and despite immense differences in size, 

voles and livestock actually compete for much the same 

vegetation during the warm season. Such ecological overlap 

may help to explain the endangered or threatened status of 

several species of Microtus (including races of M. montanus) 

in various parts of the Western United States, most of which 

appear to be associated with rangelands altered by livestock 

utilization (Nowak 1999).

The exact mechanism by which voles kill otherwise healthy, 

mature RB is unclear. Girdling of the basal stems is almost 

always the first outward signal that voles have begun to 

target a particular shrub. But, as previously mentioned, 

RB is noted for its ability to re-sprout from the root crowns 

and thus girdling alone is unlikely to prove fatal. Burrowing 

around the base of the shrub, which is generally initiated 

only after girdling has started, virtually assures the death 

of the shrub (table 1). Our initial observations indicate that 

voles feed on the rootstock, which may contribute directly 

to a shrub’s death. However, burrowing through or near 

the rootstock may promote shrub mortality in other ways, 

including the introduction of fungal pathogens or increasing 

the plant’s vulnerability to water stress. Voles clearly prefer 

larger shrubs and generally ignore small, young RB plants 

(<15 cm in height). Small shrubs are sometimes girdled 

but rarely burrowed. This suggests that beyond providing 

access to underground food, the burrows built beneath an 

established shrub have other important functions, including 

shelter and protection from aerial predators. Burrowing also 

creates a disturbed area surrounding each dead RB that 

frequently becomes a germination site for native and exotic 

annual weeds (such as tansymustard, lambsquarters, and 

cheat grass). However, our observations suggest that this is 

a transitory condition and that native graminoids and forbs 

typically replace the weedy annuals within a few years. Importantly, these disturbed sites have not been re-populated 

by either RB or sage despite ample seed availability and the 

sorts of climatic perturbation (severe drought followed by an 

exceptionally wet period) that favor woody shrub establishment (Archer and Smeins 1991).

The two major episodes of vole predation on RB are related to climatic conditions. Both the 1997 to 1999 and the 

2004 to present intervals are associated with above average 

precipitation in south-central Utah (Region 4). During both 

episodes, positive indicators of vole activity (such as surface 

runways, burrow systems, fecal pellets, grass cuttings) were 

extremely abundant, and daytime sightings of these normally 

nocturnal to crepuscular rodents became commonplace. The 

greatest damage to shrubs, especially from stem girdling, corresponded to periods of deep winter snow pack (for example, 

2004 to 2005). At these times the rodents took advantage of 

the snow to strip bark as far up the plant as the surface of 

the snow allowed. (When feeding, voles remain beneath the 

snow surface.) There was no apparent sign of vole activity or 

shrub predation during the dry period lasting from fall 1999 

to late summer 2002. Rodent numbers declined dramatically 

and the animals became either very rare or altogether absent 

from the shrub-dominated PMCs in South Hollow, including 

our study plots. Small populations of voles did persist in 

the riparian zones along Upper Valley Creek (fig. 1) where 

locally dense stands of grass, sedge, and rush (Juncus spp.) 

remained despite drought conditions. Riparian areas thus 

appear to act as refugia, serving to buffer vole populations 

against local extinction when unfavorable conditions render 

adjacent environments uninhabitable. The much higher rates 

of vole-killed RB on Plot II (table 1) may be related to the 

fact that it is located closer to the riparian staging area from 

which voles reentered the South Hollow meadows following 

a return to wetter conditions.

The influence of small mammals on rangeland vegetation structure and health is still poorly understood (Archer 

and Smeins 1991); however, recent evidence suggests that 

it might well be more profound than previously assumed. 

The systematic elimination of prairie dogs, for example, is 

thought to have altered the local vegetation composition on 

short-grass prairie, including an increase in woody shrubs 

(Archer and others 1987). Landscape level impacts by rodents 

and lagomorphs may have helped produce the vegetation 

contrasts that mark the United States-Mexican border in 

parts of the Southwest (Curtin and Brown 2001). Recent 

experiments have established the importance of voles in 

shaping vegetation structure and succession in the Midwest. 

There meadow voles (Microtus pennsylvanicus) have been 

shown to influence the herbaceous composition of synthetic 

tallgrass prairie (Howe and Lane 2004) as well as to bias the 

rate and species composition of trees that successfully invade 

old-field habitats (Mason and others 2001). In California, 

voles (Microtus californicus) likewise alter the structure of 

grasslands (Batzli and Pitelka 1970) and significantly impact 

the survival of young redwood trees through their girdling 

behavior (Guisti 2004).

Our study on the Colorado Plateau extends these findings 

by offering evidence that small mammal activity may actually reverse the pattern of woody shrub encroachment into 

grassland and meadow habitats. Net reductions of established RB within our study plots are as high as 68 percent, 

but some areas outside the plots have realized reductions of 

nearly 100 percent over the same period of time. Preliminary 

calculations suggest that more than 10,000 shrubs (mostly 

RB, but a growing number of big sage and greasewood) have 

already been killed by voles on the South Hollow study site. 

Moreover, all documented reductions in invasive woody vegetation have occurred within less than a decade of the first 

unambiguous signs (spring of 1997) of vole activity within 

these former meadows. Vole-driven shrub mortality has also 

been accompanied by a parallel expansion of open habitat 
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Vole-Driven Restoration of a Parariparian Meadow Complex on the Colorado Plateau, South-Central Utah Bramble and Bramble

that is increasingly occupied by native grasses, sedges and 

forbs (fig. 2). The composition of this vegetation is assumed 

to be similar to that originally occupying the study site, although no direct data bearing on this matter are presently 

available. Voles appear to shape the composition of vegetation on mid-western prairies over periods of years through 

selective herbivory and seed predation, the results of which 

may exhibit a considerable time-lag (Howe and Brown 2001). 

It will probably be years, therefore, before the exact nature 

of a stable herbaceous plant assemblage becomes apparent 

on those areas of South Hollow recently cleared of shrubs. 

Regardless, we are witnessing a surprisingly rapid shift from 

a closed-canopy shrub land to an open, graminoid dominated 

vegetation complex. In effect, Microtus montanus appears to 

be forcing vegetation structure back toward an earlier (pregrazing?) state on this site and one that, not incidentally, is 

much closer to its preferred natural habitat.

It is important to note that vole predation on woody shrubs 

in South Hollow has thus far been highly selective. The 

greatest shrub losses have been to mature RB. Increasingly, 

however, basin big sage (A. tridentata ssp. tridentata) is being 

targeted, especially on the higher and drier portions of the site 

where dense stands of this shrub still exist. Girdling readily 

kills sage, but voles frequently establish burrows under this 

shrub as well. Greasewood (Sarcobatus vermiculatus), which 

following the 1988 fire, developed very dense stands on exposed alkaline soil, is likewise coming under more frequent 

vole predation. Like RB, this shrub is very difficult to control 

due to its propensity to re-sprout from underground rootstock, 

but it may likewise be vulnerable to burrowing. Since vole 

activity resumed in late 2004, numerous examples of girdled 

and burrowed greasewood have appeared on the study site, 

although to date only a handful of these shrubs have actually 

died. The voles have, at the same time, ignored other native 

shrubs including mountain snowberry (Symphoricarpos 

oreophilus), Wood’s rose (Rosa woodsii), squawbush (Rhus 

trilobata) and bitterbrush (Purshia tridentata).2 All these 

shrubs are increasing to varying degrees, but snowberry most 

obviously. Snowberry, in fact, may have special resistance 

to voles. This shrub is routinely burrowed without apparent negative effect, and in some instances it has replaced 

vole-killed RB by establishing directly on the disturbed soil 

underneath the dead shrub. We think it is significant that 

the woody shrubs targeted by vole predation (RB, big sage, 

greasewood) in South Hollow are among the most common, 

widespread, and aggressive invaders of western grasslands 

and meadows. Among the shrubs avoided by voles are species that are commonly reduced or eliminated on heavily 

grazed rangelands. Thus, vole activity is not only reducing 

the overall presence of woody shrubs on the study site but 

is also shifting its composition.

2 Interestingly, voles also ignore “yellowbrush,” another native 

rabbitbrush (C. viscidifloris lanceolatus), even when very close to 

RB shrubs under active predation. However, yellowbrush is a much 

smaller shrub and is not an aggressive invader of PMC habitats 

in our area.

We believe that what has occurred on our study site could 

be reflective of a more general phenomenon affecting the vegetation and ecological health of many similar environments 

on western rangelands. Specifically, we suggest that small 

mammals, and especially grazers, may have had a significant 

but largely unrecognized role in creating and maintaining 

the types of open meadow complexes and grassy savannahs 

that historically flanked most riparian zones across the arid 

West. If so, these species would qualify as keystone elements 

in such ecosystems, just as heteromyid rodents have been 

for other western ecosystems (Brown and Heske 1990). Because the western PMCs were among the first to be severely 

degraded by livestock introductions, and because voles are 

small and relatively cryptic, these rodents may well have 

disappeared as functional components of these ecosystems 

without notice. In this context it is appropriate to raise an 

important question: What kept aggressive and prolific woody 

shrubs like RB from dominating western PMCs prior to the 

onset of grazing? RB, in particular, invades disturbed mesic 

sites more readily than sage but, unlike sage, will not be 

removed by fire once it has become established.

Voles and perhaps other herbivorous small mammals 

provide a possible answer to this question. Through their 

proclivity to kill invasive woody shrubs these small grazers 

appear to both defend and promote their preferred natural 

habitat—open, sunny expanses dominated by a graminoidrich, herbaceous vegetation. This is in contrast to the apparent 

role of voles in mid-western and Pacific Coast grasslands, 

where the rodents act to shape the existing vegetation 

(Batzli and Pitelka 1970; Howe and Lane 2004) but do not 

create (or recreate) habitat of this type. It is reasonable to 

assume that the exact relationship of small mammal activity 

to vegetation structure and evolution varies considerably 

from region to region and is influenced by many variables, 

including the abiotic and biotic history of the area and the 

specific ecological requirements of the mammalian species 

involved. It is also possible that the historic absence of 

large grazing mammals in much of the Intermountain West 

(Mack and Thompson 1982) may have conferred on small 

mammalian grazers (such as rodents and lagomorphs) a 

larger role in structuring rangeland vegetation than in other 

regions where large grazers were historically present.3 Our 

findings, therefore, raise the possibility that one of the most 

biologically important classes of western ecosystems (PMCs) 

may now be compromised, and widely so, by the absence of 

a key functional element, microtine rodents. If correct, our 

success in quickly triggering the initial transformation of a 

shrub-infested site into one probably more like the historic 

pre-grazing meadow and grassland vegetation is unlikely 

to prove an isolated anomaly. More likely, in our opinion, it 

is the predictable outcome of having reactivated a vital but 

missing dynamic in a local ecosystem.

3 Elk, while common at the South Hollow study site, appear to 

have been absent from this region in pre-settlement times (Bryant 

and Maser 1982). Elk are lacking in faunal collections from local 

archaeological sites (such as Fremont and Anasazi) and there is no 

mention of this large native grazer by the early European colonizers 

of the Escalante region. Present elk populations apparently stem 

from introductions beginning in the 1970s.
114 USDA Forest Service Proceedings RMRS-P-52. 2008

Bramble and Bramble Vole-Driven Restoration of a Parariparian Meadow Complex on the Colorado Plateau, South-Central Utah

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

Within the Intermountain West, Microtus montanus 

is most often recognized for its ability to damage or kill 

young trees and shrubs in agricultural and suburban settings (Askham 1988). But our data clearly imply that such 

destructive behavior, which involves both girdling and 

burrowing, is merely a negative manifestation of the same 

natural tendencies that, until quite recently, assigned these 

rodents a central role in maintaining some of the West’s 

most biologically rich ecosystems. This rodent is very widely 

distributed in the West (Zeveloff 1988) and small populations likely persist along most riparian zones where water 

is permanent and degradation from livestock has not been 

too severe. Therefore, it is probably safe to assume that 

many shrub-infested PMCs in the interior West are in 

close proximity to relict populations of M. montanus within 

nearby riparian corridors. During both major episodes of 

vole population expansion in South Hollow, vole dispersal 

and shrub-killing activities have extended to the fence-line 

marking the extreme southern end of our property, at least 

1.64 km from their ultimate dispersal source on Upper 

Valley Creek. This portion of the study site is nearly 200 

feet higher in elevation and distinctly drier than either 

of our study plots. The prevailing vegetation is similar to 

that of sage-steppe and related temperate savannah types 

(McPherson 1997), consisting of mature stands of big sage 

with bunch grasses (Agropyron cristatum, Stipa comata, 

Sporobolus cryptandrus) and blue grama (Bouteloua gracilis)

in the interspaces. This fact demonstrates that the shrub 

control effects of these voles are not necessarily restricted 

to the mesic meadow environments close to riparian zones. 

All of these observations, in our view, strongly suggest that 

the montane meadow vole could be a new and potentially 

important restoration tool on western rangelands and one 

certainly worthy of additional examination.

Acknowledgments__________________

We are grateful to Drs. Phyllis Coley, Denise Dearing, and 

Fred Adler for their helpful comments on the manuscript

References ________________________

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USDA Forest Service Proceedings RMRS-P-52. 2008 115

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Richard B. Keigley is a Research Ecologist, USGS Northern 

Rocky Mountain Science Center, Bozeman, MT; email: richard_

keigley@usgs.gov.

Michael R. Frisina is a Range-Habitat Coordinator, Montana 

Fish Wildlife & Parks, Butte, MT; email: mfrisina@mt.gov.

Abstract—The monitoring of Geyer willow was based on the following management objective: Browsing will prevent fewer than 50 

percent of Geyer willow shrubs from growing taller than 3 m. Three 

questions were addressed: (1) Is browsing a potential factor? (2) If 

so, can young plants grow taller than 3 meters? (3) If not, is browsing the dominant factor? All shrubs were intensely browsed. With 

a post-browsing growth rate of 5.0 cm per yr, no shrub could grow 

3 m tall. Analyses of stem growth rate excluded dominant roles for 

climate and plant vigor. Browsing and stem age were the dominant 

factors that limited growth to 3 m tall.

Introduction_______________________

The stature and density of trees and shrubs across the 

landscape is a major influence on the potential diversity and 

abundance of wildlife. On winter range, wild ungulates require access to the high quality forage (Cooke 2002; Renecker 

and Schwartz 1997; Schmidt and Gilbert 1978; Schwartz and 

Renecker 1997; Short 1981). Shrubs are among the most 

nutritious food sources on winter ranges (Peterson 1995; 

Young and Clements 2002). Shrubs and trees provide nesting sites and other habitat components for both resident and 

migratory birds, with many species having particular plant 

height and cover requirements (DeGraaf and others 1991). 

To maintain desired wildlife populations, land managers 

must provide habitat of appropriate quality, and structural 

diversity is an important aspect of habitat quality.

Ungulate browsing can potentially alter the structural 

diversity of a plant community (Kay 1990; Lovaas 1970; 

Rasmussen 1941; Wagner 2006). By preventing young plants 

from attaining full stature, protracted heavy browsing can 

move a community of trees and shrubs toward an herbaceous 

meadow-like condition that is less suitable as winter range 

(Kay and Walker 1997). Livestock grazing and ungulate 

Browse Evaluation of Tall Shrubs Based on Direct 

Measurement of a Management Objective

Richard B. Keigley and Michael R. Frisina

browsing can reduce the nesting success and abundance 

of neotropical migratory birds (Ammon and Stacey 1997; 

Bowman and Harris 1980; Jackson 1992, 1993; Knopf and 

others 1988). When the maintenance of ungulate winter 

range and bird-nesting opportunities are management 

objectives, browsing must be limited to a level at which 

structural diversity is preserved. A problem that managers 

face is that of distinguishing between harmful and benign 

levels of browsing (Keigley and Frisina 1998).

In addition to the potential influence of herbivory, climate 

and other environmental factors affect the height to which 

plants can grow (Harper 1977; Kozlowski 1991). While the 

level of herbivory may be regulated by reduction of ungulate 

population size, environmental factors are generally not 

under management control. For that reason, the development of an appropriate management strategy requires that 

environmental effects be distinguished from the effects of 

ungulate use. Big game and livestock exclosures are an 

established method for distinguishing between the effects 

of ungulates and the environment (Kay 1990; Thompson 

2002). However exclosures have often not been constructed 

at areas of interest. Once an exclosure is constructed, a number of years is required before the differences inside and out 

become apparent. Therefore, managers ideally require an 

approach that distinguishes between the effects of browsing 

and environment without the use of exclosures.

We use a case study to describe a method for determining 

if browsing pressure is having an adverse effect on structural diversity. The method can be applied in areas where 

exclosures have not been constructed. The evaluation process 

begins with the definition of a management objective and 

the selection of an indicator species. Although we believe 

that the method can be applied to short-growing shrubs, the 

results are clearest when applied to trees or shrubs capable 

of growing at least 3 m tall.

Methods __________________________

Study Area

The study took place on a portion of the 22,743 ha 

Mt. Haggin Wildlife Management Area located in southwest Montana about 19 km south of the town of Anaconda 

(45°58’

N, 113° 4’W). The study area consists of a combination 

of wet and dry meadows. Willows (Salix spp.) are common 

along numerous stream courses and in wide riparian created by beaver dams. Elevation of the study area is about 

1,900 meters; annual precipitation is about 51 cm (Frisina 

1992).
116 USDA Forest Service Proceedings RMRS-P-52. 2008

Keigley and Frisina Browse Evaluation of Tall Shrubs Based on Direct Measurement of a Management Objective

Prior to 1976, the area was privately owned (Newell and 

Ellis 1982). Domestic livestock grazed pastures continuously 

and intensively; willow was sprayed, dozerpiled, and burned 

to create more grassland for livestock. In 1976, the Montana 

Department of Fish, Wildlife and Parks (MFWP) purchased 

the land and established Mt. Haggin Wildlife Management 

Area to provide public recreational opportunities and habitat 

for wildlife (Frisina 1982). Efforts to reduce willow stopped. 

Populations of moose (Alces alces), elk (Cervus elaphus), 

mule deer (Odocoileus hemionus), and pronghorn antelope 

(Antilocapra americana) increased (Frisina 1982; Kuntz 

1993). Recently whitetail deer (Odocoileus virginianus) 

began using the study area during the summer.

Moose are the only large ungulates present on the study 

area during winter; they concentrate in willow bottoms as 

snow accumulates. Browsing of willows by moose has been 

a concern of MFWP for several years (Frisina and Keigley 

2004; Keigley and others 2002a,b; Keigley and Fager, in 

press). Deer, elk, and antelope inhabit the area during snowfree months when herbaceous plants are generally preferred 

and all habitat types are available for foraging. After establishment of the wildlife management area, livestock numbers 

were reduced (Frisina 1992). A rest-rotation grazing system 

was established in 1984, in which cattle are allowed access 

to a pasture during the growing season once every third year 

(Frisina 1992; Frisina and Keigley 2004).

Management Objective

A measurable management objective provides the basis 

for the design of a monitoring program (Elzinga and others 

1998). We adopted the following as a management goal: 

Browsing will permit some young plants to attain their potential stature, their growth being primarily limited by local 

environmental conditions. The management goal was refined 

to a measurable management objective by including: (1) the 

browse species to be monitored, (2) the potential height of 

those plants, and (3) the number that must be capable of 

growing to that potential height. The purpose of testing 

for the attainment of the objective is to either eliminate 

or isolate browsing as the dominant factor that influences 

habitat quality.

Indicator Species

We selected Geyer willow (Salix geyerianna) as the species 

to monitor. The species was widely distributed throughout the 

study area. A study that reconstructed the browsing history 

of Geyer willow at Mt. Haggin found that, at about 1985, 

browsing pressure increased from a light-to-moderate level 

to an intense level (Keigley and others 2003). Some plants 

encountered in this study were more than 5 m tall (fig. 1). 

Shorter plants were heavily browsed and had a hedge-like 

appearance. The tall shrubs attained their height when 

browsing pressure was less intense. From these observations we concluded that lightly browsed Geyer willow had 

the potential for growing to about 5 m tall.

Although Geyer willow has the potential for growth to 5 m 

at Mt. Haggin WMA, we used 3 m as the benchmark height 

in the management objective. We assumed that at 3 m tall, 

a stem had grown above the browse height of moose and its 

growth would no longer be subject to influence by browsing. 

Finally, to make the objective measurable, we defined the 

required number to be 50 percent; individual management 

agencies may wish to set different thresholds. The complete 

management objective was: Browsing will prevent fewer than 

50 percent of Geyer willow shrubs from growing taller than 

3 m; the growth of those shrubs will be primarily limited 

by local environmental conditions. It is important to note 

that this objective focuses on the effect of browsing. Other 

factors, such as site condition and inter- and intraspecific 

competition could prevent growth of Geyer willow shrubs 

to 3 m tall.

Having defined a management objective, selected an indicator species, and determined that the species was capable of 

growth to 3 m tall, we proceeded to monitor for attainment 

of the management objective.

Monitoring Method

General Approach—Monitoring was accomplished by 

addressing three questions in the following order:

1. Is browsing a potential influence that could prevent 

young plants from growing through the browse zone? If 

browsing is not a potential influence on height growth, 

the management objective was attained and further 

analysis was not necessary.

2. If browsing is a potential influence, are 50 percent of 

the sampled plants likely to grow to 3 m tall? If so, the 

management objective was attained.

Figure 1—Over the past four decades, the moose population at 

Mt. Haggin WMA increased, causing a corresponding increase 

in browsing pressure. The taller Geyer willow shrubs in the 

background grew to about 4.5 m tall when browsing pressure 

was low. Shrubs of intermediate height grew for a time when 

browsing pressure was low; their growth stopped when browsing 

pressure increased. The short willow shrubs in the foreground 

were established after browsing pressure increased to an intense 

level. Site conditions on which these plants grow are similar. 

The tall shrubs indicate the height potential for this area.
USDA Forest Service Proceedings RMRS-P-52. 2008 117

Browse Evaluation of Tall Shrubs Based on Direct Measurement of a Management Objective Keigley and Frisina

3. If 50 percent of the sampled plants are unlikely to grow 

to 3 m tall, was browsing the dominant factor that 

limited height growth? If browsing was found to be the 

dominant factor, the management objective would not 

be attained.

Site and Sample Selection—During the summer of 

2005 we sampled 20 Geyer willow shrubs at each of 10 study 

sites. Three sites were located within the Sullivan Creek 

corridor and seven sites within the Deep Creek corridor. 

Sampling was restricted to shrubs 50 to 150 cm tall as measured to the base of current-year-growth. This height-range 

criterion was selected because shrubs less than 50 cm tall 

may be protected from browsing by snowcover, while shrubs 

taller than 150 cm may escape browsing by livestock and 

deer. To grow taller than 3 m, young shrubs must grow 

through the 50 to 150 cm browse zone. Individual shrubs 

were selected for measurement by walking a transect line, 

stopping at five-pace intervals, and finding the closest shrub 

meeting the above criteria.

Within each site, conditions such as soil type and moisture 

appeared homogeneous. Stem growth rates were compared 

by ANOVA to assess variation in growing conditions across 

sites.

Question 1: Is Browsing a Potential Influence?—This

question was addressed by: (1) classifying the browsing level 

of sampled shrubs into two categories (light-to-moderate 

and intense), and (2) by determining the percentage of 

shrubs that were intensely browsed. The browsing level 

of a shrub was determined from the examination of stems. 

An individual stem was classified as light-to-moderately 

browsed if, throughout the life of the primary stem, a given 

year’s annual growth segment developed from the segment 

produced the preceding year (fig. 2a). Two situations led to 

a stem being classified as intensely browsed. First, a stem 

was classified as intensely browsed if at some point in the 

life of the stem a complete annual segment was browsed 

and had died, thus causing the next year’s terminal leader 

to develop from an annual segment older than the preceding 

year (fig. 2b). Secondly, a stem was classified as intensely 

browsed if: (1) the initial primary stem was browsed, and 

(2) a complete annual increment within that segment was 

dead, thus causing a lateral branch to be responsible for 

subsequent height growth (fig. 2c). In this case it is possible 

for a stem to be classified as intensely browsed yet meet the 

condition that each year’s annual growth segment developed 

from the segment produced the preceding year.

A shrub was classified as intensely browsed if at least 

one stem comprising the shrub was intensely browsed. 

The shrub was classified as light-to-moderately browsed 

if all stems were light-to-moderately browsed. At a given 

site, browsing was assumed to be a potential influence if 

at least 50 percent of the shrubs were intensely browsed. 

If browsing was determined to be a potential influence at 

a site, additional samples and measurements were taken 

to address questions 2 and 3 above. It is important to note 

that shrub condition is not based solely on the assessment 

of percent intensely browsed.

Collection of Samples and Measurements for 

Questions 2 and 3—Using the protocol described above, 

20 shrubs were selected for measurement and sampling at 

each site. From each shrub, the following live stems were 

selected for measurement: (1) the oldest, and (2) the tallest. 

Heights of those stems were measured to the base of current 

year growth (HBCYG). The selection criteria could be met by 

a single stem or by two separate stems.

Figure 2—Criteria for classifying the browsing level of individual 

stems into two categories: intense or light-to-moderate. A 

shrub was classified as intensely browsed if a single stem was 

intensely browsed.
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Keigley and Frisina Browse Evaluation of Tall Shrubs Based on Direct Measurement of a Management Objective

The oldest live stem was selected using girth at the base 

as a surrogate for age. Older stems were selected because 

young stems growing within the protective canopy may not 

have been exposed to browsing; because of this protective 

effect, the behavior of the shrub over the long term is best 

characterized by the condition of older stems. The measurement of the tallest height was used to assess recent changes 

in height that may have occurred.

The oldest stem was examined from tip to base to determine 

if it was browsed at any point in its life. If the oldest stem 

was not browsed, the shrub was recorded as unbrowsed. If 

the selected stem was browsed, the length from the base of 

the stem to the browsed tip was measured (HBB) (fig. 3). The 

browsed stem was then cut at ground level. The effective 

height of the stem that grew after the first browsing event 

(HAB) was calculated as:

HAB = HBCYG – HBB

Sections were cut from the base of the stem and from 

the highest living part of the unbrowsed length of stem 

(see fig. 3). The annual rings in these sections were counted; 

the ages of those sections were designated AS and AL respectively. Typically, above the point where AL was cut was a 

section of stem that had died after browsing. That section 

was examined for bud scars to determine the number of 

years over which the now-dead stem segment had grown; 

that period was designated AD.

The number of growing seasons over which the unbrowsed 

stem grew (YBB) was calculated as:

YBB = (AS – AL + 1) + AD

The number of growing seasons that occurred after the 

first browsing event (YAB) was calculated as:

YAB = AS – YBB

The net annual growth rate that occurred before the 

browsing event (NAGRBB) was calculated as:

NAGRBB = HBB / YBB

The net annual growth rate that occurred after the first 

browsing event (NAGRAB) was calculated as:

NAGRAB = (HBCYG – HBB) / (YAB – 1)

One was subtracted from YAB because height measurements were made to the base of current year growth.

To determine the lifespan of browsed stems, each shrub 

was examined for the presence of dead stems that had been 

browsed. Where present, the oldest dead stem was selected 

for sampling, using girth at the base as a surrogate for age. 

The height of each dead stem was measured to its tip (HD) 

and a section taken from its base; the age of the dead stem 

was determined from a count of annual rings (AD). The ages 

of these stems were used as an estimate for the lifespan of 

the live stems sampled. Linear and curvilinear correlation 

procedures were run to determine the model that best fit 

the height on age relationship.

Question 2: Are 50 Percent of the Sampled Plants 

Likely to Grow to 3 m Tall?—This question was first addressed by examining the long-term trend of the shrub based 

on the growth characteristics of the oldest stem. Because 

browsing was not a growth-limiting factor for unbrowsed 

stems, such stems were treated as capable of growing to 

3 m tall. In the case of browsed stems, this question was 

addressed by determining, for each stem, the net annual 

growth that occurred after the first browsing event (NAGRAB) 

and an estimated lifespan (YL). Using those values we then 

determined if the respective stems would live long enough 

to grow from HBCYG to 3-m (fig. 4).

The number of years required to grow to 3 m (Y300) was 

calculated as:

Y300 = (300 – HBCYG) / NAGRAB

Based on the correlation model that best fit the age and 

height relationship of dead stems, the lifespan of the browsed 

live stems (YL) was estimated based on their respective 

heights. The estimated remaining-years-of-life (YR) was 

calculated for each browsed stem as: YR = YL – AS. The surplus or deficit in lifespan was calculated by subtracting YR

from Y300. If YR was greater than Y300, the stem was deemed 

likely to grow to 3 m tall. If fewer than 10 shrubs at a site 

were found to be unlikely to grow to 3 m tall, the analysis 

proceeded to Question 3. In addition, we conducted a similar 

test in which it was assumed that all stems could live to the 

age of the oldest dead stem found during the study. The test 

using maximum age was used as further clarification of the 

ability of plants to attain 3 m tall.

The growth of the oldest live stem characterizes a long-term 

history of the shrub. If growing conditions remained constant 

over the life of the shrub, older stems would generally be 

taller than younger stems, and the history documented by 

the older stems would accurately characterize the growth 

Figure 3—Samples and measurements taken to determine the 

pre- and post-browsing growth rates (NAGRBB and NAGRAB, 

respectively). 
USDA Forest Service Proceedings RMRS-P-52. 2008 119

Browse Evaluation of Tall Shrubs Based on Direct Measurement of a Management Objective Keigley and Frisina

of the shrub within the 50 to 150-cm zone. But if growing 

conditions markedly improved from some adverse state, 

young vigorous stems, may become taller than the older, 

physiologically stressed stems, and the growth of older stems 

may not accurately characterize potential growth to 3 m tall. 

We tested for changes in growing conditions by comparing 

the height of the tallest stem on each shrub to the height 

of the oldest stem (HTALL – HOLD). A positive value would 

indicate an improvement in growing conditions. If there was 

a positive difference that exceeded 2 years of growth at the 

pre-browsing growth rate, we assumed that the shrub would 

be able to attain 3 m tall (HTALL – HOLD > 2 * NAGRBB).

Question 3: Is Browsing the Dominant Factor That 

Limits Height Growth?—Many factors can potentially 

affect stem growth rate, including browsing, climate, plant 

vigor, and stem age. We began the process of assessing 

the potential dominance of the browsing factor by comparing NAGRBB to NAGRAB; the comparison was done by the 

Mann-Whitney test. We then evaluated the potential roles 

of climate, plant vigor, and stem age.

To assess the potential relationship between stem growth 

rate and climate, we first determined the mean number of 

years in which stems grew before the first browsing event 

(YBB). Based on that mean, we constructed a running average annual precipitation using Wise River 3WNW climate 

data (station 249082). Mean NAGRBB values were calculated 

for stems established during a given year; only years having 

three or more values were used. The mean NAGRBB values 

and the running precipitation averages were then plotted 

to determine if a correspondence existed. NAGRBB values 

were plotted on establishment year; running averages were 

plotted on initial year over which the average was taken. 

If there was no correspondence in trends, we assumed that 

climate was not a dominant factor in determining stem 

growth rate.

A regression of NAGRAB on AS was used to test for a 

possible age-related reduction in the post-browsing growth 

rate. A negative correlation would indicate an age-related 

influence.

In evaluating the potential role of vigor, we assumed that 

vigorous plants would have high values of NAGRBB compared 

to less vigorous plants. If vigor played a dominant role in the 

post-browsing growth rate, then there should be a positive 

correlation between NAGRAB and NAGRBB. We tested for 

that correlation by regression.

We examined the role of browsing: (1) by comparing mean 

pre- and post-browsing growth rates, and (2) by comparing 

the pre- and post-browsing variances of those growth rates. 

If browsing played a dominant role, the mean NAGRAB

should be less than NAGRBB, and dominant roles for climate, 

age, and vigor would have been excluded. Our rationale 

for comparing variances as a test for browsing dominance 

was as follows. Prior to browsing, stem growth rate would 

be influenced by plant vigor and environmental factors. If 

none of these factors were dominant across the landscape, 

growing conditions would be diverse, thus resulting in large 

variation among NAGRBB values. The introduction of a 

landscape-wide dominant-factor (that is, browsing) would 

cause growth rate to decline to a relatively constant value 

compared to the pre-browsing rates. We compared variances 

by the Levene test; a difference in variance would suggest 

the effect of a landscape-wide dominant factor.

Results and Discussion______________

The first step of data analysis was to address Question 1: 

Is browsing a potential factor that could prevent young 

plants from growing through the browse zone? All shrubs 

at all sites were intensely browsed, thus exceeding the 50 

percent threshold set prior to the survey.

Having deemed that browsing was a potential factor, we 

proceeded to address Question 2: Are young plants likely 

to grow through the browse zone? The grand mean net annual growth rate after the first browsing event (NAGRAB) 

was 5.0 ± 0.3 cm / year (± SE). The means and standard 

errors for individual sites ranged from 2.2 ± 0.9 cm / year 

to 9.4 ± 1.6 cm / year. The maximum observed NAGRAB for 

an individual stem was 26 cm/year. Sixteen stems had negative or zero growth rates, the base of current year growth 

respectively being lower than, or equal to, the height that 

the stem was when first browsed. Because these stems did 

not have a positive growth rate, they were excluded from 

consideration when determining if the browsed stem would 

grow taller than 3 m.

Dead stems were found in 195 out of 200 sampled shrubs. 

The grand mean lifespan of the browsed dead stems was 

11.6 ± 0.3 years (± SE). The mean lifespan at different sites 

ranged from 9.4 ± 1.0 years to 14.0 ± 1.0 years. In some cases, 

the age of the old live stem of a shrub exceeded the age of 

the old dead stem. This finding indicates that the estimated 

lifespan of live stems was conservative. The maximum observed lifespan was 22 years.

Figure 4—The rationale for assessing whether a stem could 

grow to 3 m tall.
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Keigley and Frisina Browse Evaluation of Tall Shrubs Based on Direct Measurement of a Management Objective

A logarithmic model provided the best fit for correlating 

lifespan with height (N = 195, Adjusted R2 = 0.11, P < 0.0001). 

Mechanical protection is likely responsible for the positive 

correlation between lifespan and stem height. Stems growing 

in the interior of taller shrubs would enjoy a longer period of 

protection from browsing, during which they could branch 

and develop greater photosynthetic potential compared to 

shorter, younger stems that are killed by browsing. The 

variation of observed values from predicted values could be 

due to many factors including disease, frost damage, plant 

vigor, and local site conditions.

We found that by growing at a rate of its measured NAGRAB, no stem would grow out of the browse zone within 

the respective lifespan estimated by the logarithmic model 

(fig. 5a). If it were assumed that all shrubs could live to the 

maximum observed lifespan of 22 years, 14 (7.0 percent) of 

the shrubs would live long enough to grow to 3 m tall (fig. 5b). 

At no site were 50 percent of the plants likely to grow to 3 m 

tall using lifespan based on either the correlation model or 

on the maximum observed age. Based on the characteristics 

of the older stems, growth to 3 m was deemed unlikely.

The comparison of the height of young tall stems with the 

height of old stems suggests a recent slight improvement 

in growing conditions. Younger stems had overtopped older 

stems in 65.5 percent of the shrubs. But only 8.5 percent of 

the stems had exceeded the height of their respective older 

stems by two annual growth increments (that is, equal to 

twice the respective NAGRBB). The slight improvement in 

growth was consistent with monitoring conducted elsewhere 

on the Mt. Haggin Wildlife Management Area (Keigley and 

Fager, in press). Based on both a long-term and short-term 

perspective, no site met the 50 percent threshold of growth 

to 3 m tall.

Having determined that shrubs are unlikely to grow 

through the browse zone, we addressed Question 3: Is browsing the dominant factor that limits height growth? The 

mean NAGRBB for all sites was19.0 ± 0.3 cm / year (± SE). 

The variance of NAGRBB across sites was homogeneous; 

there was no significant difference in NAGRBB between sites 

(P = 0.149). We concluded that between-site environmental 

condition was not a significant factor in stem growth rate.

In an analysis of pooled data, the variances of NAGRBB and 

NAGRAB were heterogeneous (Levene test, P < 0.001). The 

mean NAGRAB of all sites (5.0 cm/year) differed significantly 

from the mean NAGRBB (Mann-Whitney test, P < 0.001). We 

examined how climate plant vigor, stem age, and browsing 

might have contributed to the reduction in growth rate.

On average, stems grew 2.9 ± 0.1 years before being 

browsed. Based on this period, we constructed a 3-year 

running average precipitation. A plot of NAGRBB and precipitation showed no similarity, suggesting that betweenyear differences in precipitation had little effect on stem 

growth (fig. 6). Given the proximity to flowing surface water, 

it is likely that all stands received subsurface moisture 

throughout the growing season, thus moderating any effect 

that precipitation might have.

The stem growth rate of plants diminishes as stems reach 

maturity and cease growing taller, so an age-related decline 

in growth rate was expected to occur in our sample. The question was: Was an age-related decline the dominant influence 

on the post-browsing growth rate? There was a negative 

correlation between stem age and NAGRAB indicating an 

age-related effect (Adjusted R2 = 0.03, P = 0.009) (fig. 7). 

If we examine the maximum growth rate that occurred 

within each year, NAGRAB diminishes with age; no stem 

older than 14 years of age had a post-browsing growth rate 

that exceeded 6.8 cm/year. In contrast, the post-browsing 

growth rates of some young stems were extremely low, some 

having negative values. For these stems, age could not have 

been the cause of low initial growth rates. We concluded 

that age was a significant factor for stems having high 

initial post-browsing growth rates but age did not exercise 

landscape-level dominance.

We anticipated that stems that grew vigorously before 

browsing would continue to grow vigorously after browsing, 

although perhaps at a reduced rate. This relationship would 

be documented by a positive correlation between and NAGRAB

and NAGRBB. In fact there was a negative correlation 

Figure 5—Frequency diagram showing the deficit number of 

years required to grow to 3 m tall. The upper graph depicts the 

deficit using estimated lifespan based on stem height. Over 60 

percent of the sample would require more than 40 years to attain 

3 meters. The lower graph depicts the deficit assuming that all 

shrubs could attain the age of the oldest dead stem sampled 

during the study (22 y); seven percent (14 plants) would live 

long enough to grow to 3 meters, while over 50 percent would 

not. In either case, less than 50 percent of the browsed shrubs 

are likely to grow to 3 m tall.
USDA Forest Service Proceedings RMRS-P-52. 2008 121

Browse Evaluation of Tall Shrubs Based on Direct Measurement of a Management Objective Keigley and Frisina

Figure 8—Relationship between pre- and post-browsing growth 

rates. If plant vigor exerted a dominant influence on growth 

rate over the life of the plant, one would expect a positive 

correlation between growth rates. There was a significant 

negative correlation (P = 0.001).

Figure 6—Three-year running average annual precipitation 

and pre-browsing growth rate. Annual precipitation values are 

plotted on the initial year of the 3-year-mean; the pre-browsing 

growth rates are plotted on the establishment year of the stem. 

The trends do not correspond with one another, indicating 

that climate was not a dominant factor. Proximity to flowing 

water likely reduces the effect of between-year variation in 

precipitation.

Figure 7—Relationship between stem age and post-browsing 

growth rate. There was a significant negative correlation 

(P<0.01). Although some young plants had a low post-browsing 

growth rate, the maximum post-browsing growth rate within 

each year diminishes with age.

between NAGRAB and NAGRBB (b = –0.14, Adjusted R2 = 

0.05, P = 0.001) (fig. 8). We concluded that pre-browsing 

vigor did not result in an enhanced post-browsing growth 

rate compared to less vigorous plants, and that vigor was 

not a dominant factor.

The significant difference between pre-and post-browsing 

mean growth rates (19.0 versus 5.0) was consistent with a 

browsing effect. The significant reduction in post-browsing 

variance (93.1 versus 34.5, P < 0.001) was consistent with 

the onset of a landscape-level dominant factor, namely 

browsing.

Having excluded climate and vigor as dominant factors, 

we concluded that browsing, in combination with age, dominated the growth of Geyer willow at the Mt. Haggin WMA. 

Because of the age-related decline in post-browsing growth 

rates, the respective values of NAGRAB that were measured 

at the time our study will decline over the remaining life of 

the stems, further supporting our interpretation that stems 

will not live long enough to grow to 3 m tall.

In summary, we conducted a four-part process that consisted of (1) defining a management objective, (2) selecting 

an indicator species, (3) determining its potential stature, 

and (4) monitoring for the attainment of the objective. We 

concluded that the management objective was not attained: 

browsing will prevent more than 50 percent of the plants 

from growing taller than 3 m.

Our approach is distinctive in how monitoring relates 

to the management objective. Over the past decades, land 

managers have monitored browsing level using form class 

(Cole 1958, 1959; Dasmann 1951; Wallmo and others 1977) 

and percent twigs browsed (Aldous 1945; Julander 1937; 

Nelson 1930; Stickney 1966). While form class and utilization 

can indicate that browsing has a minor or major effect on 

plant growth, neither form class nor utilization has intrinsic 

ecologic value. It is the condition of plants that is of intrinsic value. Form class and utilization are indirect measures 

of current and future plant condition. The assessment of 

habitat variables using indirect measurement is one of the 

primary weaknesses in wildlife-habitat research (Morrison 

and others 1998).

In contrast, the management objective we described has 

intrinsic ecologic value. If browsing pressure is maintained 

at a level that allows some plants to attain full stature, then 

it can be assumed that browsing does not have a significant 

negative effect on characteristics such as structural diversity 

and fruit production. The process described above directly 

tests for the attainment of a management objective that has 

intrinsic ecologic value.
122 USDA Forest Service Proceedings RMRS-P-52. 2008

Keigley and Frisina Browse Evaluation of Tall Shrubs Based on Direct Measurement of a Management Objective

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

Our approach is also distinctive in that the effect of browsing can be isolated without the use of big game and livestock 

exclosures. In the absence of exclosures, land managers have 

had to accept the possibility that climate and other factors 

played a dominant role in controlling the condition of browse 

plants. The analysis of stem growth rate provides a means 

of assessing the role of those factors without experimentally 

isolating the plants from browsing.

Acknowledgments__________________

This project was partially funded by Montana Department of Fish, Wildlife and Parks under Federal Aid Project 

W-154-R.

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USDA Forest Service Proceedings RMRS-P-52. 2008 123

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Michael R. Frisina is Range-Habitat Coordinator, Montana 

Department of Fish, Wildlife & Parks, Butte, MT 59701; email: 

mfrisina@mt.gov.

Carl L. Wambolt is Professor, Montana State University, Department of Animal and Range Sciences, Bozeman, MT 59717; email: 

cwambolt@montana.edu.

W. Wyatt Fraas and Glen Guenther are former Research 

Associates, Montana State University, Department of Animal and 

Range Sciences, Bozeman, MT 59717.

Abstract—Mule deer (Odocoileus hemionus) populations have 

gradually declined in recent decades, while elk (Cervus elaphus) 

have often increased throughout their common ranges. The cause is 

uncertain and a source of debate. Increasing elk numbers on these 

ungulate winter ranges may be causing competition for resources. 

We contrast winter diets of mule deer and elk and relate them to 

population trends of both species on the Mount Haggin Wildlife 

Management Area in southwest Montana. Elk increased from 

an observed number of 172 in 1978 to 700 in 1996 and slightly 

declined in 1997. Likewise, mule deer increased from 202 in 1978 

to 586 in 1989, but declined to 288 between 1990 and 1997. The 

same five browse species comprised 95 percent and 52 percent of 

the mule deer and elk winter diets, respectively. These data indicate there is a potential for forage competition between mule deer 

and elk sharing winter ranges. Because elk have a more varied 

diet (55 percent browse, 32 percent grass, 12 percent forbs) than 

mule deer (98 percent browse, 2 percent grass, 0.5 percent forbs) 

on this winter range, it is likely that mule deer will be the most 

negatively impacted.

Introduction_______________________

During the past several decades elk (Cervus elaphus) 

populations have increased across the Western United 

States (O’Gara and Dundas 2002; Peek and others 1982). 

Concurrently, concern has been expressed that mule deer 

(Odocoileus hemionus) populations are on a gradual, but 

steady decline (Connolly 1981; Wallmo 1978). Much conjecture has been put forth regarding the causes of the mule 

deer decline; there are probably many factors involved. Our 

Mule Deer and Elk Winter Diet as an 

Indicator of Habitat Competition

Michael R. Frisina, Carl L. Wambolt, W. Wyatt Fraas, and Glen Guenther

objective was to determine mule deer and elk winter diets 

and consider the possibility that overlapping diets impact 

one or both ungulates.

Study Area ________________________

The study area within the Mount Haggin Wildlife Management Area (MHWMA) is approximately 16 km southeast 

of Anaconda in southwest Montana and was described in 

detail by Guenther (1989). The MHWMA was purchased 

in 1976 by the Montana Department of Fish, Wildlife and 

Parks (MFWP) to conserve habitat for a variety of wildlife 

species including elk and mule deer. All 18 sampling locations 

studied were on ungulate winter range within the northern 

portion of the MHWMA (Guenther 1989).

The topography is characterized by mountainous slopes. 

Sampling locations are near the Continental Divide at 

elevations between 1,577 m and 1,943 m. Average annual 

precipitation is about 340 mm, with 47 percent falling from 

April through July. June and February are the wettest and 

driest months averaging 70 and 16 mm of precipitation, 

respectively.

Although vegetation on the study area is diverse, the 

bitterbrush (Purshia tridentata)-grass type predominates 

(Guenther 1989). Many of the sample locations show the 

impacts of historic intensive livestock grazing and air-born 

pollutants from the nearby Anaconda smelter. Guenther 

(1989) identified 12 shrub, 25 graminoid, and 44 forb species that occur on the 18 sampling locations. Other common 

shrubs in the area include snowberry (Symphoricarpos albus) and Oregon grape (Berberis repens). The most common 

grasses are Kentucky bluegrass (Poa pratensis), basin wildrye 

(Elymus cinerus), needleandthread (Stipa comata), green 

needlegrass (S. viridula), rough fescue (Festuca scabrella), 

Idaho Fescue (F. idahoensis), and bluebunch wheatgrass 

(Agropyron spicatum). Among the most abundant forbs are 

the exotic whitetop (Cardaria draba) and spotted knapweed 

(Centaurea maculosa), which have invaded nine of the study 

sites. Native long-leaved aster (Aster chilensis) is also common. Rocky Mountain Juniper (Juniperus scopulorum) and 

lodgepole pine (Pinus contorta) are present and Douglas-fir 

(Pseudotsuga menziesii) appears to be increasing on some 

sites (Guenther 1989). Small stands of aspen (Populus 

tremuloides) are scattered across the winter range.

The MHWMA provides year-long range for elk, mule deer, 

moose (Alces alces), white-tailed deer (Odocoileus virginianus), and pronghorn (Antilocapra americana)) (Frisina 

1982, 1992). The portion of the MHWMA studied is an important winter range for elk, mule deer, and moose.
124 USDA Forest Service Proceedings RMRS-P-52. 2008

Frisina, Wambolt, Fraas, and Guenther Mule Deer and Elk Winter Diet as an Indicator of Habitat Competition

Methods __________________________

Mule deer feces were analyzed from 18 sampling locations 

during 1991 and elk feces were analyzed from 13 of the 

sampling locations. No elk feces were found at five of the 18 

sampling locations. The samples were sent to the Composition 

Analyses Laboratory at Colorado State University for diet 

determination using microhistological techniques (Sparks 

and Malecheck 1968). Each sample was a composite from 

10 pellet groups. From each composite sample, the mean 

composition of plants in the diet was determined from five 

slides of 20 fields each.

Mule deer and elk population trends were determined 

from aerial surveys flown during winter when clear skies, 

cold temperatures, and snow cover provided ideal conditions 

for observing animals. During the years of our study, 1978 

through 1997, the entire winter range was flown each winter 

to obtain trend counts. The first author was the observer 

on all flights. Separate flights were made for deer and elk. 

A mule deer trend count was obtained every winter and 

an elk trend count was completed for 16 of the 20 winters. 

During four years, weather conditions were unusually mild 

and elk too widely scattered for a comparable survey to be 

accomplished. When weather conditions are unusually mild, 

not all elk migrate to the winter range. A fixed-wing aircraft 

was used for all elk counts. A small helicopter was used to 

conduct nine of the mule deer counts (1978 through 1984, 

and 1997) and a fixed wing aircraft for 11 of the counts (1986 

through 1996). The open nature of the winter range (lack 

of dense forest) made deer similarly observable from both 

types of aircraft (fig. 1).

Results ___________________________

Mule Deer Diet

Twenty-five taxa or plant groups were identified in the 

winter diet of mule deer (table 1). These consisted of 14, 5, 

and 6 browse, grass, and forb taxa, respectively. Browse was 

the most important component with five species comprising 

95 percent of the diet: antelope bitterbrush, Oregon grape, 

Rocky Mountain juniper, Douglas-fir, and, lodgepole pine. 

Antelope bitterbrush provided over half of the mule deer 

winter diet (table 1). Antelope bitterbrush is often very important to the nutrition of mule deer (Young and Clements 

2002). In a synthesis of mule deer food habits studies for 

the Western United States and Canada, Kufeld and others 

(1973) rated antelope bitterbrush as a preferred browse 

during winter. Kufeld and others (1973) considered Rocky 

Mountain juniper, Oregon grape, and Douglas-fir to be of 

moderate value, and lodgepole pine of low preference. Grasses 

and forbs combined contributed only about 2.5 percent to the 

winter diet of mule deer (table 1). Idaho fescue and rough 

fescue combined were the only herbs >1 percent of the winter 

diet. Idaho fescue was considered moderately preferred in 

the winter by mule deer and rough fescue of low preference 

by Kufeld and others (1973).

Elk Diet

Twenty-four plant species were identified in the winter 

diet of elk (table 1). These consisted of 9, 10, and 5 browse, 

grass, and forb taxa, respectively. Browse was the most 

Figure 1—Population trend for mule deer and elk wintering on the Mount Haggin 

WMA, 1978 to 1997.
USDA Forest Service Proceedings RMRS-P-52. 2008 125

Mule Deer and Elk Winter Diet as an Indicator of Habitat Competition Frisina, Wambolt, Fraas, and Guenther

significant portion at 55 percent of the diet. Only three 

browse species, antelope bitterbrush, Oregon grape, and 

Douglas-fir comprised 53 percent of the diet. Kufeld (1973) 

and Cook (2002) considered antelope bitterbrush a highly 

preferred winter forage plant, Oregon grape a valuable 

plant, and Douglas-fir to be of low preference to wintering 

elk. Grasses and forbs combined comprised about 44 percent 

of elk winter diets. Grasses contributed 32 percent with 

fescues the most important at 17 percent. Previous studies 

found both rough fescue and Idaho fescue to be preferred 

winter forage for elk (Kufeld 1973; Cook 2002). Forbs were 

least eaten by elk during winter at 12 percent of the diet, 

Table 1—Mean percent plant content in feces found within the 18 and 13 sampling sites 

for mule deer and elk, respectively, on the Mount Haggin Wildlife Management 

Area, 1991 (T = species cover <0.1 percent).

Species Mule Deer Elk

Grasses 2.0 31.9

Wheatgrasses Agropyron spp. 0.1 0.2

Bentgrasses Agrostis spp. – 0.2

Brome grasses Bromus spp. 0.1 0.3

Sedges Carex spp. 0.1 4.0

Fescue grasses Festuca spp. 1.1 17.4

Rushes Juncus spp. – 0.3

Bluegrasses Poa spp. 0.6 3.3

Needlegrasses Stipa spp. – 5.7

Prairie junegrass Koeleria macrantha – 0.4

Unknown grasses – 0.1

Forbs 0.5 12.0

Pussytoes Antennaria rosea T –

Sunflower family Asteraceae spp. 0.1 0.2

Milkvetch & Locoweeds Astragulus–Oxytropis spp. – 0.5

Arrowleaf balsamroot Balsamorhiza sagittata 0.1 –

Common horsetail Equisetum arvense – 2.0

Wyeth lupine Lupinus wyethii 0.3 9.3

Phloxes Phlox spp. T T

Flannelleaf mullein Verbascum thapsus T –

Shrubs/Trees 97.4 55.3

Kinnikinnick Arctostaphylos uva-ursi T –

Serviceberry Amelanchier alnifolia 0.1 –

Fringed sagewort Artemisia frigida 0.1 –

Oregon grape Berberis repens 22.4 10.9

Winterfat Ceratoides lanata – T

Rubber rabbitbrush Chrysothamnus nauseosus 1.6 0.1

Green rabbitbrush Chrysothamnus viscidifloris 0.1 –

Rocky Mountain juniper Juniperus scopulorum 12.0 0.1

Douglas-fir Pseudotsuga menziesii 5.0 7.6

Antelope bitterbrush Purshia tridentata 52.8 34.4

Lodgepole pine Pinus contorta 3.0 2.0

Quaking aspen Populus tremuloides – 0.1

Raspberry Rubus spp. T –

Russet buffaloberry Shepherdia canadensis 0.2 0.1

Common snowberry Symphoricarpus albus 0.1 –

Gray horsebrush Tetradymia canescens 0.2 –

but lupine did comprise 9 percent of the diet. The only 

documented species of lupine at the study sites was Wyeth 

lupine (Lupinus wyethii). Kufeld (1973) considered lupines 

of moderate preference to wintering elk.

Conclusions and Management 

Implications_______________________

The winter diets of mule deer and elk were similar with 

the same five browse plants (antelope bitterbrush, 

Oregon grape, Rocky Mountain juniper, Douglas-fir, and 
126 USDA Forest Service Proceedings RMRS-P-52. 2008

Frisina, Wambolt, Fraas, and Guenther Mule Deer and Elk Winter Diet as an Indicator of Habitat Competition

lodgepole pine) comprising 95 percent of the deer diet and 

55 percent of the elk diet (table 1). However, the elk diet was 

more diverse than deer with grasses and forbs contributing 

32 percent and 12 percent of the diet, respectively. The deer 

diet was 98 percent browse while this forage class contributed 

56 percent to that of elk.

Population trend data indicate elk increased from an observed number of 172 in 1978 to 700 in 1996 and began to 

decline in 1997 (fig. 1). Likewise, mule deer increased from 

202 in 1978 to 586 in 1989, but declined to 288 between 

1990 and 1997 (fig. 1).

Guenther and others (1993) found utilization on MHWMA 

bitterbrush ranged from 57 percent to 96 percent, averaging 80 percent at the 18 sampling locations during 

the winter of 1988 to 1989. They noted that the majority of 

bitterbrush plants appeared heavily hedged due to historical browsing. Fraas (1992) found lower browsing rates at 

the same sites during the winter of 1990 to 1991 (0 to 60 

percent). Fraas (1992) attributed the difference in browsing 

rate to differences in sampling methods. While Guenther 

(1989) chose twigs on the outermost portion of each plant, 

Fraas (1992) randomly chose branches, some deep within 

the canopy and thus not as susceptible to browsing. The 

diet similarity between elk and mule deer indicates there is 

potential for competition between these species. The trend 

in declining mule deer between 1990 and 1996 with a concomitant upward trend of elk through 1996 also indicates 

competition may be occurring. Elk diets at MHWMA are 

more varied providing them a survival advantage over mule 

deer when difficult winter conditions occur. Douglas-fir 

and lodgepole pine are both considered a low value browse 

for elk and deer (Kufeld 1973; Kufeld and others 1973), yet 

contributed 8 percent and 3 percent, respectively, to the diet. 

This relatively high contribution of these low value browse 

species to the winter diets is a further indication that the 

combined populations of mule deer and elk may be exceeding 

habitat carrying capacity. Guenther and others (1993) found 

browsing to be intense enough to suggest a management 

strategy directed toward reducing the number of wintering 

deer. Our data indicate it maybe more important to control 

the size of the wintering elk population.

Acknowledgments__________________

Montana Department of Fish, Wildlife and Parks funded 

the project through Federal Aid Project W-154-R4. The 

Department of Animal and Range Sciences, Montana 

State University, Bozeman also supported the project. The 

authors thank Stephen J. Knapp for his encouragement 

and support.

References ________________________

Connolly, Guy. E. 1981. Trends in populations and harvests. In: 

Wallmo, Olof C., ed. Mule and black-tailed deer of North 

America. Lincoln, NB: University of Nebraska Press: 225-243.

Cook, John G. 2002. Nutrition and food. In: North American Elk: 

Ecology and Management. Washington, DC: Smithsonian Institution Press: 259-349.

Fraas, William W. 1992. Bitterbrush growth and reproductive characters in relation to browsing in southwest Montana. Bozeman, 

MT: Montana State University. 137 p. Thesis.

Frisina, Michael R. 1982. Management plans. Montana Outdoors. 

13(3): 31-33.

Frisina, Michael R. 1992. Elk habitat use within a rest-rotation 

grazing system. Rangelands. 14(2): 93-96.

Guenther, Glen. E. 1989. Ecological relationships of bitterbrush plant 

communities on the Mount Haggin Wildlife Management Area. 

Bozeman, MT: Montana State University. 73 p. Thesis.

Guenther, Glen. E.; Wambolt, Carl L; Frisina, Michael R. 1993. 

Characteristics of bitterbrush habitats that influence canopy 

cover and mule deer browsing. Journal of Environmental Management. 36: 175-181.

Kufeld, Roland C. 1973. Foods eaten by the Rocky Mountain elk. 

Journal of Range Management. 26(2): 106-113.

Kufeld, Rolan C.; Wallmo, Olof C.; Feddema, Charles. 1973. Foods 

of the Rocky Mountain mule deer. Res Pap RM-111. Fort Collins, 

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

Mountain Forest and Range Experiment Station [now Rocky 

Mountain Research Station]. 31 p.

O’Gara, Bart W.; Dundas, Robert, G. 2002. Distribution: Past and 

present. In: Toweill, Dale E.; Thomas, Jack W., eds. North 

American elk: Ecology and management, Washington, DC: 

Smithsonian Institution Press: 67-119.

Peek, James; Pederson, Richard J.; Thomas, Jack W. 1982. The future 

of elk and elk hunting. In: Thomas, Jack W.; Toweill, Dale E., eds. 

Elk of North America, Ecology and Management. Harrisburg, 

PA: Stackpole Books: 599-625.

Sparks, Donnie R.; Malecheck, John C. 1968. Estimating percentage dry weights in diets using microscopic techniques. Journal 

of Range Management. 21(4): 264-265.

Young, James A; Clements, Charlie D. 2002. Purshia: the wild and 

bitter roses. Reno, NV: Univeristy of Nevada Press. 266 p.

Wallmo, Olof C. 1978. Mule and black-tailed deer. In: Schmidt, John L.; 

Gilbert, Douglas L., eds. Big game of North America, Ecology and 

Management. Harrisburg, PA: Stackpole Books: 31-34.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein
USDA Forest Service Proceedings RMRS-P-52. 2008 127

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Steven A. Lei (at the time of this report) was a graduate student, 

Department of Electrical Engineering, Arizona State University, 

Tempe, AZ 85488.

Simon A. Lei is Instructor of Statistics and Research Methods, 

Department of Educational Psychology, University of Nevada, 

Las Vegas, NV 89154; email: leis2@unlv.nevada.edu.

Abstract—Radial symmetry of white bursage (Ambrosia dumosa) 

canopies was quantitatively investigated in four isolated mountain 

ranges in the Mojave Desert. In contrast to what has been observed 

for the co-dominant shrub, creosote bush (Larrea tridentata), no 

significant differences were detected among canopy quadrants in 

leaf biomass, relative leaf tissue percentage, or canopy height. A 

slight tendency for directional asymmetry was observed for all study 

sites. However, the cardinal direction of the canopy asymmetry 

varied among study sites, suggesting that no ecological selection for 

asymmetry is in force. Three possible explanations for the nearly 

symmetrical arrangement of white bursage canopy foliage that we 

observed are offered.

Introduction_______________________

Creosote bush (Larrea tridentata) and white bursage 

(Ambrosia dumosa) are co-dominant or dominant shrub 

species occurring at valley floors and at lower mountain 

foothills and slopes in the warm Mojave Desert. These species dominate different canopy layers when they co-occur. 

Creosote bush is dominant because of biomass or productivity. In contrast, white bursage dominates because of higher 

density (Lei 1995). The current distributions of both species 

probably developed after the last pluvial period (20000 to 

10000 B.P.) (Bradley 1964).

Creosote bush orients foliage clusters (canopy foliage) predominantly toward the southeast in the Mojave Desert and 

the northern Chihuahuan Desert (Neufeld 1988). However, 

canopy foliage of creosote bush at the southernmost distribution extreme in the Chihuahuan Desert of Mexico shows no 

predominant orientation (Neufeld 1988). Canopy orientation 

is expressed as shorter branches in the southeastern sectors of crowns, reducing self-shading early in the morning 

(Neufeld 1988). A study of white bursage crown symmetry 

has not been quantitatively investigated. Thus, the objective 

of this study was to determine if canopy foliage asymmetry 

exists in Mojave Desert white bursage.

Radial Symmetry in the White Bursage 

(Ambrosia dumosa) in the Mojave Desert

Steven A. Lei and Simon A. Lei

Methods __________________________

Field Surveys

Field studies were conducted in May 2005 in the Spring and 

Newberry Mountains of southern Nevada, Clark Mountain of 

southern California, and Virgin Mountains of northwestern 

Arizona. Foothills of each of the four mountain ranges possess isolated white bursage populations (table 1).

To ensure comparability, study sites were selected at 900 

m elevation above sea level. Linear transects 200 m in length 

were located in areas where creosote bush and white bursage 

were co-dominant. A total of eight transects, two per mountain 

range, were positioned near, but at least 200 m away from 

roadways to eliminate possible road effects. Each transect 

was subdivided into one of two geomorphic surface types: 

terrace or wash. Due to high abundance of white bursage, 

40 individual plants were randomly selected within each 

geomorphic surface type on each transect. Plants located at 

close proximity to other plants, rocks, and other objects that 

might physically interfere with a plant’s canopy development 

were excluded from the study. The canopy of each plant was 

divided into four, 90° quadrants based upon cardinal direction divisions after correction for true north (0°). Quadrant 

divisions occurred at 45°; 135°; 225°; and 315°.

For each study plant, dry leaf tissue mass (biomass), relative leaf tissue percentage, and height of the tallest branch 

were determined. All fresh leaves from the four quadrants 

of individual plants were collected, and were oven-dried at 

50 °C for 72 hours before weighing. Relative quadrant leaf 

tissue percentage was computed for each white bursage 

by dividing the leaf biomass per quadrant by the total leaf 

biomass.

Statistical Analyses

Two-way Analysis of Variance (ANOVA) (Analytical Software 1994) was used to detect differences in mean values 

among the four quadrants, with site and geomorphic surface 

as main variables. Mean values were presented with standard errors, and statistical significance was determined at 

p ≤ 0.05 level.

Results and Discussion______________

Mean quadrant leaf biomass of white bursage varied 

slightly, ranging from 3.8 g in the SE quadrant of the 

Clark Mountain to 5.4 g in the NE section of the Newberry 

Mountains (table 2). Mean percentage of white bursage 

leaves ranged from 23.7 percent to 27.4 percent in the Clark 
128 USDA Forest Service Proceedings RMRS-P-52. 2008

Lei and Lei Radial Symmetry in the White Bursage (Ambrosia dumosa) in the Mojave Desert

Table 1—Geographical characteristics of four white bursage study sites. Location, 

as well as approximate latitude (°N) and longitude (°W) are shown. 

Mountain ranges are arranged alphabetically in the table. All study 

sites were located at 900 m in elevation to ensure comparability.

Location County, State Latitude (°N) Longitude (°W)

Clark Mountain San Bernadino, CA 35°32’ 115°31’

Newberry Mountains Clark, NV 35°17’ 115°43’ 

Spring Mountains Clark, NV 36°05’ 115°38’

Virgin Mountains Mohave, AZ 36°38’ 114°06’

Table 2—Mean quadrant leaf biomass, relative leaf tissue percentage, and shrub height of white bursage in four isolated 

mountain ranges of the Mojave Desert. n = 10 per quadrant per geomorphic surface per study site.

Clark Mountain Newberry Mountains Spring Mountains Virgin Mountains

Variable Terrace Wash Terrace Wash Terrace Wash Terrace Wash

Leaf biomass (g)

NE 4.6 4.7 5.4 5.2 5.0 5.1 4.4 4.5

SE 4.8 5.0 5.2 5.5 4.0 4.4 4.3 4.6

SW 4.2 4.3 3.9 4.7 4.9 5.3 4.4 4.3

NW 3.8 4.0 4.2 4.4 4.6 4.8 4.0 4.3

Leaf tissue (percent)

NE 26.8 24.9 27.4 26.8 24.6 25.7 26.8 27.3

SE 25.3 25.1 26.7 25.9 25.9 25.7 27.4 27.0

SW 24.9 23.9 24.6 26.7 25.5 26.8 25.1 26.2

NW 23.7 24.3 25.8 26.9 26.3 26.1 24.2 23.8

Shrub height (cm)

NE 36.7 37.1 40.6 39.0 36.2 36.4 39.5 40.1

SE 37.9 37.6 36.9 39.4 38.4 38.9 39.7 40.3

SW 37.4 37.2 37.8 36.7 38.8 38.3 40.9 40.4

NW 38.4 37.8 38.2 37.1 39.7 38.2 39.5 39.9

(SE quadrant) and Newberry (NE quadrant) Mountains, 

respectively. Mean quadrant shrub height also varied 

slightly, ranging from 36.4 cm in the NE quadrant of the 

Spring Mountains to 40.9 cm in the SW quadrant of the 

Virgin Mountains (table 2).

Mean quadrant leaf biomass, relative leaf tissue percentage, and shrub height of white bursage did not differ 

significantly (ANOVA, p > 0.05, table 3) among four sites, 

four quadrants, or between two geomorphic surfaces. This 

is different than what has been observed for creosote bush, 

which often co-occurs with white bursage (Neufeld 1988). 

Although slight (but not significant) tendencies in orientation were noted among sites in this study, the direction of 

those asymmetries differed among sites, suggesting that no 

ecological selection for asymmetry is in force.

There are at least three possible explanations for the apparent symmetrical arrangement of white bursage canopy 

foliage observed in the natural field. First, it is theorized 

that the SE asymmetry of creosote bush is a mechanism 

to maximize dawn and dusk photosynthetic rates and to 

minimize mid-day photosynthetic rates; such a mechanism 

would help increase daily water use efficiencies (WUE) 

(Neufeld 1988). White bursage, however, may not have the 

same adaptive impetus for heliotropic asymmetry, owing to 

the facts that white bursage becomes deciduous in periods 

of intense heat and water (drought) stress during summer 

months (Barbour and others 1999). Aridity is a permanent 

feature in the warm Mojave Desert. Drought, however, is a 

temporary feature, occurring when precipitation falls below 

normal. Most drought-deciduous species, including white 

bursage, produce only one crop of leaves a year, and enter 

a long summer dormancy following leaf drop (Barbour and 

others 1999). Their leaves are energetically inexpensive to 

manufacture compared to conifer and other evergreen leaves 

such as creosote bush.

A second explanation for the apparent symmetry might 

be that a tendency for canopy foliage asymmetry exists 

but is masked by the effects of herbivory, stem breakage, 
USDA Forest Service Proceedings RMRS-P-52. 2008 129

Radial Symmetry in the White Bursage (Ambrosia dumosa) in the Mojave Desert Lei and Lei

and other uncontrolled factors in the natural environment 

(Swaddle and others 1994). This possibility could be tested 

by growing plants under more controlled conditions such as 

are found in greenhouse studies. Finally, potential canopy 

foliage asymmetries might be too slight relative to plant size 

making them difficult to detect using our methods.

Acknowledgments__________________

We gratefully acknowledge David Valenzuela and Shevaun 

Valenzuela for providing valuable field and laboratory assistance. Stanley Kitchen and Stephen Ferris provided careful, 

thorough review on earlier versions of this manuscript.

References ________________________

Analytical Software. 1994. Statistix 4.1, an interactive statistical 

program for microcomputers. St. Paul, MN: Analytical Software. 

329 p.

Barbour, M.G.; Burk, J.H.; Pitts, W.D.; Gilliam, F.S.; Schwartz, M.W. 

1999. Terrestrial plant ecology. 3rd edition. Menlo Park, CA: The 

Benjamin/Cummings Publishing Company, Inc. 649 p.

Bradley, W.G. 1964. The vegetation of the Desert Game Range with 

special reference to the desert bighorn. TransAmerican Desert 

Bighorn Council 8: 43-67.

Lei, S.A. 1995. A gradient analysis of Coleogyne communities in 

southern Nevada. Las Vegas, NV: University of Nevada. 117 p. 

Thesis.

Neufeld, H.W. 1988. Canopy architecture of Larrea tridentate (DC.) 

Cov., a desert shrub: foliage orientation and direct beam radiation 

interception. Oecologia. 75: 54-60.

Swaddle, P.J.; M.S. Witter.; I.C. Cuthill. 1994. The analysis of 

fluctuating asymmetry. Animal Behavior. 48: 986-989.

Table 3—Summary table of two-way ANOVA showing effects of study site, geomorphic surface, 

and their interactions on leaf biomass, relative leaf tissue percentage, and shrub height. 

n = 40 per geomorphic surface per study site. df = 1 for geomorphic surface, df = 3 for 

study site and for study site * geomorphic surface combination. Statistical significance 

is determined at p ≤ 0.05.

Site Geomorphic surface Site * surface

Variable F P F P F P

Leaf biomass 1.18 0.3832 5.28 0.055 0.49 0.7010

Leaf tissue 0.68 0.5906 2.64 0.1480 0.17 0.9110

Shrub height 0.80 0.5298 1.79 0.222 1.31 0.3434

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

USDA Forest Service Proceedings RMRS-P-52. 2008 131

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Rosemary L. Pendleton and Burton K. Pendleton are 

Research Ecologists, Rocky Mountain Research Station, USDA 

Forest Service, Albuquerque, NM, 87102-3497; email: rpendleton@

fs.fed.us.

Karen R. Wetherill is Research Scientist, Sevilleta LTER, Department of Biology, University of New Mexico, Albuquerque, NM.

Terry Griswold is Research Entomologist, USDA-ARS Bee 

Biology and Systematics Laboratory, Logan, UT.

Abstract—Expansion of diploid creosote shrubs (Larrea tridentata 

(Sessé & Moc. ex DC.) Coville)) into grassland sites occurs exclusively 

through seed production. We compared the reproductive biology 

of Larrea shrubs located in a Chihuahuan desert shrubland with 

isolated shrubs well-dispersed into the semiarid grasslands at the 

Sevilleta National Wildlife Refuge. Specifically, we examined (1) reproductive success on open-pollinated branches, (2) the potential 

of individual shrubs to self-pollinate, and (3) bee pollinator guild 

composition at shrubland and grassland sites. Sampling of the bee 

guild suggests that there are adequate numbers of pollinators at 

both locations; however, the community composition differs between 

shrub and grassland sites. More Larrea specialist bee species were 

found at the shrubland site as compared with the isolated shrubs. 

Large numbers of generalist bees were found on isolated grassland 

bushes, but their efficiency in pollinating Larrea is currently unknown. Higher percent seed fill of unbagged, open-pollinated shrubs 

at the shrubland site, compared with isolated grassland shrubs (76 

versus 57 percent) suggests that bee specialists may increase plant 

pollination success. Isolated grassland shrubs varied greatly in the 

number of seeds produced in pollinator-exclusion bags, whereas the 

number of self-pollinated seeds produced by shrubland plants was 

more uniform. Overall, the difference in seed produced by bagged 

and unbagged branches of isolated shrubs was much less than the 

difference produced by plants located at the shrubland site. These 

trends will be explored in greater detail in future years.

Reproductive Biology of Larrea tridentata: 

A Preliminary Comparison Between Core 

Shrubland and Isolated Grassland Plants at the 

Sevilleta National Wildlife Refuge, New Mexico

Rosemary L. Pendleton, Burton K. Pendleton, Karen R. Wetherill, and Terry Griswold

Introduction_______________________

Chihuahuan Desert shrubland is expanding into semiarid 

grasslands of the Southwest. Creosote (Larrea tridentata) 

seedling establishment in grasslands is a key factor in this 

conversion. Diploid Larrea plants of the Chihuahuan Desert are not clonal as has been reported for some hexaploid 

Mojave populations (Vasek 1980). Consequently, Larrea

establishment in semiarid grasslands of New Mexico must 

occur exclusively through seed. At McKenzie Flats in the 

Sevilleta National Wildlife Refuge, there exists a gradient 

in Larrea density stretching from dense Larrea shrubland 

(4,000 to 6,000 plants per hectare) to semiarid desert grassland with only a few scattered shrubs. This study investigated 

the effects of spatial isolation on Larrea pollination and 

seed production. We compared the reproductive biology of 

Larrea shrubs located in the Larrea shrubland community 

with isolated shrubs well dispersed into the grasslands of 

McKenzie Flats. Specifically, we compared (1) reproductive 

success on open-pollinated branches, (2) the potential of 

individual shrubs to self-pollinate, and (3) bee pollinator 

guilds at shrub and grassland sites.

Methods __________________________

Flower and seed production of Larrea shrubs occurs most 

reliably in spring, but can also occur in late summer or early 

fall depending on the timing and amount of monsoonal precipitation. This study took place in the spring of 2005. We 

used mesh bags to exclude pollinators from four branches 

per plant of six isolated and six shrubland Larrea plants. 

After flowering was complete, we bagged four additional 

branches per shrub with mesh bags to prevent insect damage and facilitate collection of all flowers and fruits. Flowers 

that did not develop fruit could still be counted because the 

bag mesh size was small enough to retain remnant gynoecia. Once the fruits had ripened on the branch, all bags 

were collected and returned to the lab for processing. For 

each sampled branch, we counted total numbers of flowers, 

developed fruits, and filled seeds to determine percent fruit 

set, percent seed fill, and return on investment calculated as 

the average number of filled seeds per flower (five possible). 

Seed fill was determined using a cut test.
132 USDA Forest Service Proceedings RMRS-P-52. 2008

Pendleton, Pendleton, Wetherill, and Griswold Reproductive Biology of Larrea tridentata: A Preliminary Comparison Between Core Shrubland...

Bee pollinator guilds at shrubland and grassland sites 

were sampled twice weekly from May 21 to June 17, once in 

the morning and once in the afternoon. Five shrubs at each 

site were sampled for 15 minutes per bush per visit. Bee 

collections were later pinned and provisionally identified to 

species using available literature. Determinations were later 

confirmed at the USDA-ARS Bee Biology and Systematics 

Laboratory in Logan, Utah. Classification as generalist or 

specialist pollinators followed Minckley and others (1999). 

Bee species not specifically identified in that paper as Larrea

specialists were considered to be generalists. Identifications 

have so far been performed only for female specimens. We 

assume females comprise the primary pollinating agents, 

because they collect pollen with which to provision their 

nests (Simpson and others 1977). Males will be included in 

the analysis when identifications are completed.

Results and Discussion______________

Larrea Reproductive Biology

We were able to recover open-pollinated seed from only 

two of the six isolated shrubs. Because of unusually high 

numbers of grasshoppers present in the grassland in 2005, 

flowers not protected by pollinator-exclusion bags were subject to herbivory, the severity of which varied with shrub. 

Consequently, seed and fruit production numbers for openpollinated branches of isolated shrubs are based on two 

shrubs, resulting in a high standard deviation.

Average fruit set (percentage of flowers developing fruit) for 

open-pollinated branches of shrub (91 percent) and grassland 

(93 percent) Larrea was nearly identical (fig. 1A). Average 

percent seed fill for shrubland plants was substantially 

higher than for isolated shrubs (76 compared to 57 percent), 

resulting in a higher number of filled seed produced per 

flower (fig. 1B, 1C). Although the difference in seed fill was 

not significant due to the low number of sampled shrubs (two 

isolated shrubs), it does suggest greater pollination success 

for shrubs located within the shrubland population.

Individual grassland shrubs varied greatly in the number 

of seeds produced in pollinator-exclusion bags, whereas the 

number produced by shrubland plants was more uniform (see 

error bars in fig. 1). Percentage fruit set of self-pollinated 

branches ranged from 15 to 83 percent for isolated shrubs, 

and from 26 to 51 percent for shrubland plants. Overall, 

however, isolated shrubs had higher fruit set and seed fill 

on self-pollinated branches than did shrubland plants. The 

reason for this difference is not clear, but may be due to a 

redistribution of resources within the shrub. Shrubland 

Larrea plants had greater seed fill for open-pollinated 

branches and may have directed resources to those branches 

and away from selfed branches (Knight and others 2006). Conversely, isolated shrubs that lost potential open-pollinated 

flowers to herbivory may have directed available resources 

into self-pollinated fruit production. Mixed pollination systems, in which plants produce seed through outcrossing or 

by selfing, constitute a bet-hedging strategy that allows for 

seed production in the absence of successful cross-pollination 

(Kalisz and others 1999).

Our results on seed production of self- and open-pollinated 

flowers are comparable to those reported previously. 

Figure 1—Means and standard errors for (A) percent fruit 

set, B) percent seed fill, and (C) number of seeds per flower 

of selfed and open-pollinated branches of isolated grassland 

and shrubland Larrea. For open-pollinated branches of isolated 

shrubs, n = 2. For all other groups, n = 6.
USDA Forest Service Proceedings RMRS-P-52. 2008 133

Reproductive Biology of Larrea tridentata: A Preliminary Comparison Between Core Shrubland... Pendleton, Pendleton, Wetherill, and Griswold

Tetraploid plants growing near Tucson, Arizona, produced 

an average of 0.41 filled seeds per flower for mesh-bagged 

branches (Simpson and others 1977). This is roughly 

equivalent to the average we found of 0.37 seeds per flower 

for isolated shrubs, but higher than the mean of 0.20 seeds 

per flower in the shrubland. Open-pollination of the Tucson 

shrubs resulted in 0.79 seeds per flower, compared to our 0.76 

seeds per flower for shrubland plants. Again, the greater disparity in seed production of selfed and out-crossed branches 

in shrubland as opposed to isolated grassland shrubs may 

be due to differential resource allocation.

In creosote shrublands, Larrea is likely to be an overabundant pollen source for visiting bees in years with 

prolific flowering, resulting in proportionately fewer successful cross-pollinations (Simpson and others 1977). The 

relatively low number of filled seeds per flower reported both 

here and by Simpson and others (1977) for open-pollinated 

shrubs occurred in years with abundant spring moisture. 

Higher seed fill is likely to occur in years of low flowering. 

High temporal variation in precipitation patterns associated 

with desert ecosystems may promote the retention of selfpollination as a back-up reproductive mechanism in Larrea

(Simpson and others 1977).

Bee Pollination Guilds

Bees are by far the most important group of Larrea

pollinators (Simpson and others 1977). Over 120 species 

of native bees visit Larrea flowers (Minckley and others 

1999). Several species are oligolectic, collecting pollen 

exclusively from Larrea even in the presence of alternate 

potential pollen sources. Others are polylectic, collecting 

pollen from multiple sources. Larrea provides both pollen 

and nectar resources to visiting insects. Nectar feeders 

generally do not restrict their visitation to a particular 

plant species. Conversely, some pollen foragers specialize 

on productive, dependable hosts such as Larrea (Minckley 

and others 1999).

Results from our bee guild sampling suggest that there 

are adequate numbers of pollinators at both locations; 

however, the community composition differs between 

shrubland and grassland sites (table 1). More than 

three times as many bees were collected from isolated 

shrubs within the grassland. This may suggest resource 

limitation for generalist bees in the grassland. The large 

numbers collected from isolated bushes primarily reflect 

the abundance of three species of presumed generalists, 

Colletes salicicola, Halictus tripartitus, and Lasioglossum 

pruinosiformis. Simpson and others (1977) report that 

C. salicicola, while considered a generalist, appeared to 

prefer Larrea and Prosopis as pollen sources, which may 

explain the relatively high numbers also collected from 

plants in the shrubland. The relatively low numbers of 

bees collected from the shrubland site is not unusual. 

Simpson and others (1977) reported that pure stands of 

Larrea were consistently characterized by low pollinator 

densities. However, we found a notable difference in pollinator composition between sites. Five of the six Larrea

specialist bee species were found at the shrubland site, 

whereas only three of the six were found on the isolated 

bushes. Also, the number of individuals of specialist 

species was more than five times the number on isolated 

bushes, even though the total number of bees collected 

from isolated bushes was much higher (table 1).

Recent research by Cane and others (2006) investigated 

the effects of fragmentation of Larrea populations on the bee 

pollination guild. They reported differences in bee functional 

guilds based on dietary breadth (generalist or specialist 

feeders) and nesting substrate. Population fragmentation 

had a negative effect on Larrea specialists and on groundnesting bees. Our study investigated the extreme in fragmentation of a Larrea population, namely isolated shrubs 

within a semiarid grassland. We found a similar negative 

effect on Larrea specialists, but no effect on ground-nesting 

bees. Ground-nesting bees were found in abundance at 

both locations, whereas stem and cavity-nesting bees (four 

species) were found only in the shrubland site. The lack of 

stem and cavity-nesting bees in the grassland site reflects 

a lack of nesting substrate, a resource well-represented in 

the fragmented urban Larrea populations studied by Cane 

and coworkers (2006).

Bee numbers may be limited by appropriate nesting sites, as 

well as by the abundance of floral resources (Cane 2001). For 

ground nesting bees, suitable nest sites should be relatively 

constant between shrubland and grassland, and therefore, 

the spatial density of bees is likely to be constant between 

the two sites. The apparent higher abundance of bees that 

we report for grassland shrubs likely reflects attraction to 

an abundant concentrated food resource in what is otherwise a relatively homogeneous grass-dominated landscape. 

Conversely, the superabundant floral resources present in 

the shrubland would create an apparently lower density 

per bush because individuals would visit shrubs closest to 

their nest site.

The effects of fragmentation on Larrea reproduction have 

not previously been examined. While some studies report 

reduced pollination success with fragmentation, others 

argue that an abundance of floral generalists may contribute 

more to successful seed production than infrequent visits by 

specialists (Cane and others 2005; Minckley and Roulston 

2006). In this study, large numbers of bees were collected 

from isolated shrubs, but their pollination efficiency is 

unknown. The higher seed fill of open-pollinated branches 

in the shrubland compared to isolated grassland shrubs 

suggests that bee specialists may indeed increase plant 

pollination success.

Our results are from one flowering season. The amount 

of floral resources provided by Larrea can vary greatly 

among years and between spring and fall growing seasons. 

The number of bee species collected in Larrea shrublands 

is also quite variable seasonally and annually (Cane and 

others 2005; Minckley and others 1999; Simpson and others 1977). The majority of bee individuals collected come 

from a few common species, with the remainder comprised 

of a small number of individuals from multiple species. A 

high degree of variation in bee abundance and composition has often been reported in other studies (Williams 

and others 2001). Additional research is now underway 

to determine if the results reported here are consistent 

among years.
134 USDA Forest Service Proceedings RMRS-P-52. 2008

Pendleton, Pendleton, Wetherill, and Griswold Reproductive Biology of Larrea tridentata: A Preliminary Comparison Between Core Shrubland...

Table 1—Numbers of female bees collected from isolated grassland and core shrubland 

Larrea plants. Bee species considered to be Larrea specialists are in bold.

# Collected from # Collected from 

Family Scientific name isolated bushes core bushes

Andrenidae Andrena prunorum 1 0

Perdita semicaerula 1 28

Apidae Anthophora sp. 1 1 0

Apis mellifera 1 3

Bombus morrisoni 2 0

Centris caesalpiniae 4 4

C. cockerelli 2 0

C. ferrisi 0 1

Diadasia rinconis 0 1

Epeolus mesillae 0 2

Melissodes sp. 1 10 2

Triepeolus sp. 1 0 1

Xylocopa californica 0 2

Colletidae Colletes clypeonitens 6 0

C. covillae 2 9

C. hyalinus 4 0

C. louisae 4 4

C. salicicola 66 18

C. sphaeralceae 1 1

C. wootoni 3 0

Halictidae Agapostemon angelicus 4 2

Halictus ligatus 1 0

H. tripartitus 248 6

Lasioglossum sp. 3 8 2

L. sp. 5 5 2

L. morrilli 7 1

L. pruinosiformis 72 46

L. sisymbrii 2 3

Nomia sp. 1 1 0

Megachilidae Anthidium cockerelli 0 1

Ashmeadiella bigeloveae 0 1

A. meliloti 0 1

Hoplitis biscutellae 0 2

Megachile spinotulata 7 0

M. sp. 1 7 0

M. sp. 2 1 0

Trachusa larreae 0 7

Mellitidae Hesperapis sp. 1 0 1

 Total number of individuals 473 151

 Total number of species 27 26

 Total number of specialist species 3 5

 Total number of specialist individuals 9 47

 Number of site specific species 12 11
USDA Forest Service Proceedings RMRS-P-52. 2008 135

Reproductive Biology of Larrea tridentata: A Preliminary Comparison Between Core Shrubland... Pendleton, Pendleton, Wetherill, and Griswold

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

Acknowledgments__________________

We thank Julieta Betinelli, Steve Bodio, Justyn Bell, 

Joanna Redfern, and Carol Jones for assistance in bagging, 

seed processing, bee collection, and pinning of specimens.

References ________________________

Cane, J.H. 2001. Habitat fragmentation and native bees: a premature 

verdict? Conservation Ecology. 5(1): 3. [online] Available: http://

www.consecol.org/vol5/iss1/art3 [11/20/06].

Cane, J.H.; Minckley, R.L.; Kervin, L.J.; Roulston, T.H. 2005. Temporally persistent patterns of incidence and abundance in a pollinator 

guild at annual and decadal scales: the bees of larrea tridentata. 

Biological Journal of the Linnean Society. 85: 319-329.

Cane, J.H.; Minckley, R.L.; Kervin, L. J.; Roulston, T.H.; Williams, 

N.M. 2006. Complex responses within a desert bee guild (Hymenoptera: Apiformes) to urban habitat fragmentation. Ecological 

Applications. 16: 632-644.

Kalisz, S.; Vogler, D.; Fails, B.; Finer, M.; Shepard, E.; Herman, 

T.; Gonzales, R. 1999. The mechanism of delayed slefing in Collinsia verna (Scrophulariaceae). American Journal of Botany. 

86: 1239-1247.

Knight, T.M.; Steets, J.A.; Ashman, T.-L. 2006. A quantitative 

synthesis of pollen supplementation experiments highlights the 

contribution of resource reallocation to estimates of pollen limitation. American Journal of Botany. 93: 271-277.

Minckley, R.L.; Cane, J.H.; Kervin, L.; Roulston, T.H. 1999. Spatial 

predictability and resource specialization of bees (Hymenoptera: 

Apoidea) at a superabundant, widespread resource. Biological 

Journal of the Linnean Society. 67: 119-147.

Minckley, R.L.; Roulston, T.H. 2006. Incidental mutualisms and 

pollen specialization among bees. In: Waser, N.M.; Ollerton, J., 

eds. Plant-pollinator interactions: from specialization to generalization. Chicago, IL: University of Chicago Press: 69-98.

Simpson, B.B.; Neff, J.L.; Moldenke, A.R. 1977. Reproductive systems 

of Larrea. In: Mabry, T.J.; Hunziker, J.H.; DiFeo, D.R., Jr., eds. 

Creosote bush: biology and chemistry of Larrea in New World 

deserts. US/IBP Synthesis Series 6. Stroudsburg, PA: Dowden, 

Hutchinson & Ross, Inc.: 92-114.

Vasek, F.C. 1980. Creosote bush: long-lived clones in the Mojave 

Desert. American Journal of Botany. 67: 246-255.

Williams, N.M.; Minckley, R.L.; Silveira, F.A. 2001. Variation in 

native bee faunas and its implications for detecting community 

changes. Conservation Ecology 5(1): 7. [online] Available: http://

www.consecol.org/vol5/iss1/art7 [11/20/06].

USDA Forest Service Proceedings RMRS-P-52. 2008 137

Fire Recovery in Shrublands

USDA Forest Service Proceedings RMRS-P-52. 2008 139

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Dennis J. Hansen and W. Kent Ostler are Senior Scientist and 

Science Supervisor, respectively, National Security TechnologiesLLC, 

Las Vegas, Nevada 89193-8521; email: hansendj@nv.doe.gov.

Abstract—In the springs of 2004, 2005, and 2006, surveys were 

conducted on the Nevada Test Site (NTS) to characterize vegetation 

resources and climatic components of the environment that contribute 

to wildland fires. The NTS includes both Great Basin Desert and 

Mojave Desert ecosystems and a transitional zone between these 

two deserts. The field surveys assessed 211 sites along major NTS 

corridors for the abundance of native perennial and annual species 

and invasive weeds. The abundance of fine-textured (grasses and 

herbs) and coarse-textured (woody) biomass was visually estimated 

on numerical scales ranging from zero to five. Distribution of biomass is shown in Geographic Information System maps by NTS 

operational area. Precipitation on the NTS from January through 

April of 2004, 2005, and 2006 was above average. There has been 

an average of 11 wildland fires per year on the NTS over the past 

28 years with an average of about 239 acres (97 hectares) per fire. 

A map showing the location and description of historic fires is presented. The three most commonly observed invasive annual plants 

to colonize burned areas are Arabian schizmus (Schizmus arabicus) 

at low elevations, red brome (Bromus rubens) at lower to middle 

elevations, and cheatgrass (Bromus tectorum) at middle to higher 

elevations. Colonization by invasive species increases the likelihood 

of future wildland fires because they provide abundant fine fuels 

that are more closely spaced than native vegetation. Blackbrush 

(Coleogyne rammosissima) vegetation types appear to be the most 

vulnerable plant communities to fire followed by pinyon juniper/

sagebrush vegetation types. Wildland fires are costly to control 

and to mitigate once they occur. Recovery of burned areas is very 

slow without reseeding or transplanting with native species and 

other rehabilitation efforts. Untreated areas become much more 

vulnerable to future fires once invasive species, rather than native 

species, colonize a burned area.

Introduction_______________________

The Nevada Test Site (NTS) is operated by the U.S. Department of Energy, National Nuclear Security Administration Nevada Operations Office. It occupies approximately 

A Survey of Vegetation and Wildland 

Fire Hazards on the Nevada Test Site

Dennis J. Hansen and W. Kent Ostler

1,375 square miles (356,123 ha) of desert and woodland 

terrain 70 mi (113 km) northwest of Las Vegas, Nevada. The 

NTS lies on the transition between the Mojave and Great 

Basin deserts; consequently, the flora and fauna consist of 

species characteristic of both deserts. Ten major vegetation 

alliances have been described (Ostler and others 2000) on 

the NTS. The ecology of the site has been studied extensively 

for more than 40 years (Webb and others 2003; Wills and 

Ostler 2001). The presence of wildland fires at the site have 

also been previously described (Hansen and Ostler 2004). 

The importance of wildland fires in shaping the ecology of 

the Mojave and Great Basin deserts is reported in numerous 

papers in these proceedings.

Wildland fires on the NTS are costly to control and to 

mitigate once they occur. Recovery of burned areas is very 

slow without reseeding or transplanting with native species and other rehabilitation efforts. Untreated areas may 

become much more vulnerable to future fires once invasive 

species, rather than native species, colonize a burned area. 

A need was recognized for a relatively rapid method of 

identifying where wildland fire hazards (in other words, 

the amount of fine-textured fuels such as grasses and forbs, 

and coarse-textured fuels such as woody shrubs and trees) 

occur on the NTS and for quantifying such hazards spatially and temporally. It was recognized that hazards are 

significantly influenced by the fuels that are produced by 

winter and spring precipitation (Hansen and Ostler 2004). 

In 2004 a road survey was designed and implemented to 

assess wildland fire fuel hazards and to summarize data 

for management of wildland fires and natural resources of 

the site. This paper describes the results of roads surveys 

conducted during 2004, 2005, and 2006.

Methods __________________________

Records describing wildland fires for the period of record 

from 1978 to 2006 were collected, reviewed, and summarized 

to determine the location, cause, size, and frequency of 

wildland fires on the NTS. Geographic Information System 

maps were then created to show the distribution of historic 

recorded wildland fires and major corridors on the NTS. 

From this information, road routes were established with 

sampling intervals varying from 1 to 2 mi (1.6 to 3.2 km) 

between stations. Exact distance between stations was 

adjusted to avoid roadside disturbances or where highway 

conditions, such as curves in the road or lack of sufficient 

road shoulder width to pull off, created safety hazards. As 

many of the NTS operating areas as possible were included 

and sampled in the route, consistent with the size of the 
140 USDA Forest Service Proceedings RMRS-P-52. 2008

Hansen and Ostler A Survey of Vegetation and Wildland Fire Hazards on the Nevada Test Site

area (in other words, smaller areas had less sampling stations than larger areas). A total of 211 sampling stations 

were selected.

Scale values for fine fuels and woody fuels were then 

defined, based on a knowledge of the range of biomass conditions present on the NTS (Ostler and others 2000). The fuel 

scale values used were

0 = Little or no biomass present,

1 = Very little biomass present,

2 = Little biomass present,

3 = Moderate biomass present,

4 = High biomass present, and

5 = Very high biomass present.

Digital pictures were then taken to document this range 

of biomass conditions for both fine fuels and woody fuels 

at several sites. Fuel ratings for fine fuels and woody fuels 

were added together for each site to give a combined fuel 

rating that ranged from 0 to 10. Additionally, a data sheet 

was designed to record the presence of the top three species 

of woody plants (listed in order of apparent biomass from 

greatest to least), top four species of herbaceous plants (from 

greatest to least), and top three species of invasive annuals 

(listed from greatest to least).

The survey was conducted during the spring of each year 

from mid-April to mid-May when fuel biomass was near 

maximum, starting at the lower elevations and working into 

the higher elevations. Survey stations were located 164 to 

328 ft (50 to 100 m) from the roadside to avoid disturbances 

associated with maintenance of the roads and the abundance 

of invasive species frequently found along the road edges. 

Geographic Positioning Satellite (GPS) coordinates were 

recorded to document the location of sampling points and a 

digital photograph was taken to document site conditions.

Precipitation data were summarized from seventeen 

meteorological recording stations, National Oceanic and 

Atmospheric Administration, Air Resources Laboratory/

Special Operations and Research Division (NOAA ARL/SORD 

stations as reported for May 22, 2006), for the NTS for the 

period of record for each station for the months of January, 

February, March, and April. Precipitation at each recording 

station was given equal weighting and averaged for all 17 

stations. Percent increase or decrease from the site average 

was used to indicate a measure of above or below average 

precipitation for the year compared to the long-term average, 

based on the mean over the period of record (approximately 

46 years). Lightning data were based on 8 years of warm 

season data as reported by NOAA ARL/SORD in North Las 

Vegas, Nevada (Randerson and Sanders 2006).

Results and Discussion______________

Based on interpretation of aerial photographs, fire scars, 

species composition, and other evidence observed in the 

field, it appears that wildland fires have been common and 

have contributed to the patterns of vegetation observed at 

the NTS. Historical records (table 1) suggest that there 

have been an average of approximately 11 fires per year on 

the NTS since 1978, and that the average area burned per 

fire was about 239 ac (97 ha). A chart of the distribution of 

wildfires on the NTS since 1978 is shown in figure 1. The 

distribution of fires by months (fig. 2) indicates a fire season 

of June, July, and August.

Most of the historical wildland fires during the past 

40 years were located at mid-elevation ranges and mostly 

within foothills in the blackbrush (Coleogyne rammosissima) 

plant communities (fig. 3). These plant communities are 

well adapted to drought and are often found with few green 

leaves during periods of drought. The fine diffuse branching 

structure of blackbrush makes it particularly vulnerable to 

fire. These shrub communities have seen increases in red 

brome (Bromus rubens, drier sites) and cheatgrass (Bromus 

tectorum, wetter sites) during the past half century, which 

contribute to their increased susceptibility to wildland fire. 

Once burned, these communities that were once dominated by 

blackbrush are colonized by more fire-resistant shrubs such 

as joint fir (Ephedra spp.) and rabbitbrush (Ericameria spp.). 

Recovery of blackbrush communities burned by wildland fires 

Table 1—Number and acreage of wildland fires on the NTS.

Year Fires Acres Hectares

1978 10 7,901 3,197

1979 6 2 1

1980 26 13,504 5,465

1981 13 7 3

1982 6 2 1

1983 16 18,291 7,402

1984 17 1,132 458

1985 11 1,609 651

1986 12 236 96

1987 14 213 86

1988 23 821 332

1989 15 323 131

1990 7 7 3

1991 4 4 2

1992 12 239 97

1993 7 7 3

1994 8 15 6

1995 8 4,605 1,864

1996 2 1,700 688

1997 6 15 6

1998 9 2,580 1,044

1999 7 50 20

2000 11 151 61

2001 8 490 198

2002 7 360 146

2003 4 4 2

2004 8 8 3

2005 31 13,000 5,261

2006a 16 9,000 3,642

 

28-year total 324 76,276 30,869

Average per year 11 2,630 1,064

Average per fire 239 97

aNumber of fires and area through August 7, 2006. (Source: NTS Fire 

Incident Reports and personal communication with James A. Brown, 

Fire Marshal, NTS, August 7, 2006.)
USDA Forest Service Proceedings RMRS-P-52. 2008 141

A Survey of Vegetation and Wildland Fire Hazards on the Nevada Test Site Hansen and Ostler

Figure 1—The number of wildland fires on the NTS by year since 1978.

Figure 2—The total number of wildland fires on the NTS by month for the period from 1978 to 2002, 

indicating an active fire season of June, July, and August.
142 USDA Forest Service Proceedings RMRS-P-52. 2008

Hansen and Ostler A Survey of Vegetation and Wildland Fire Hazards on the Nevada Test Site

is very slow and rare. Wildland fires are more frequently 

found in this vegetation than in woodland vegetation such 

as pinion juniper/sagebrush types, despite the fact that 

these latter vegetation types receive more lightning strikes 

during the warm season. The difference in susceptibility is 

probably due to the relative lack of understory grasses in 

pinion juniper types compared to blackbrush types and the 

greater water content in foliage of these trees compared to 

the foliage of blackbrush shrubs.

Lightning is responsible for the majority of wildland fires 

on the NTS. Large acreages of land burned in 2005 and 2006 

(more than 8,903 ha; 22,000 ac) due to lightning, make this 

cause the source of the largest loss of land area burned by 

wildland fires on the NTS. While other causes of wildland 

fire are reported as occurring less frequently, these causes 

are still responsible for large portions of burned area. For 

example, during the period of 1978 to 1987 it can be calculated that 41.2 percent of the area burned was from vehicle 

exhaust, 39.3 percent from lightning, 15.2 percent from 

military and security training, 4.1 percent from unknown 

causes, and less than 0.2 percent from smoking, construction, and other human-related activities. Data were still not 

available for the years 1994, 1996, 1997, 1998, 1999, 2003, 

2004, and 2005, because recording practices switched from 

Figure 3—Location of historic wildland fires on the NTS in predominantly middle 

elevations.
USDA Forest Service Proceedings RMRS-P-52. 2008 143

A Survey of Vegetation and Wildland Fire Hazards on the Nevada Test Site Hansen and Ostler

a paper log to a digital log that did not specifically identify 

wildland fires. The switch in recording practices made it 

difficult to identify and separate such fires and their causes 

from other emergency call outs, which were numerous.

Eight years (1993 through 2000) of warm-season (June 

through September) lightning data on the NTS have been 

summarized by Randerson and Sanders (2006). For the 

eight warm seasons, a total of 9,596 cloud-to-ground (CG) 

lightning flashes (fl) were detected. For the approximately 

1,375 mi2 (356,123 ha) of NTS, this total yields an average 

NTS warm-season flash density of 0.137 fl/mi2 (0.356 fl/km2). 

Climatologically, CG lightning activity begins to develop 

rapidly after 1100 Pacific Daylight Time (PDT), reaching 

a peak between 1400 and 1459 PDT over the mesas and 

between 1500 and 1559 PDT over the southern half of the 

NTS. Although CG lightning has occurred during every 

hour of the day, minimum CG lightning occurs between 

0600 and 1100 PDT, with 0800 to 0859 PDT being the hour 

of least CG activity. Maximum daily warm season flash 

densities ranged from 1.2 to 2.7 fl/mi2 (3.0 to 7.0 fl/km2). 

The most active thunderstorms generated CG flash rates 

of 50 to 75 fl/hr. Average spacing between successive CG 

flashes was 3.4 mi (5.5 km), and 77 percent of these flashes 

were 5.0 mi (8 km) apart. Maximum separation detected 

was 11.9 mi (19.2 km).

Historically, most large burns on the NTS occurred during 

late summer (late July or early August) as compared to early 

spring, correlating with monsoonal storm patterns that were 

accompanied with lightning. Areas of maximum lightning 

activity are typically located in mountainous areas in the 

northern, upper elevation ranges of the NTS. Although these 

areas receive more lightning, they do not necessarily result 

in more wildland fires; either the vegetation in these areas 

does not catch fire as readily or the fire does not spread when 

ignited as it does in the blackbrush communities. It appears 

that understory invasive grasses contribute significantly to 

increased hazards of wildland fires, especially when located 

in blackbrush communities.

Late winter and early spring precipitation (January 

through May) at the NTS during the first year of the 

Table 2—Precipitation history and percent presence in surveyed sites (top species contributing to fine fuels).

Precipitation history 2004 2005 2006

- - percent above average - -

Precipitation (January–April) 4 67 20

Invasive introduced species 2004 2005 2006

- - - - percent presence - - - -

Bromus rubens (red brome) 51.7 64.4 67.8

Bromus tectorum (cheatgrass) 40.3 54.0 60.7

Erodium cicutarium (redstem stork’s bill) 5.2 6.2 24.6

Schismus arabicus (Arabian schismus) 4.7 2.8 5.2

Native species 2004 2005 2006

- - - - percent presence - - - -

Amsinkia tessellata (bristly fiddleneck) 34.0 62.0 16.1

Mentzelia albicaulis (whitestem blazingstar) 49.8 8.1 0.0

Chaenactis fremontii (pincushion flower) 27.0 8.0 0.0

wildland fire survey (2004) was 4 percent above average 

and was considered typical of recent precipitation at the 

NTS during the past two decades (table 2). However, the 

late winter and early spring precipitation at the NTS during 

2005, the second year of the survey, was 67 percent above 

average and was distributed early in the year and more 

or less evenly across most of the NTS. This precipitation 

initiated germination of many invasive species of grasses 

and native annual species (table 2). Biomass of annual species was very apparent by the end of May 2005. Maximum 

growth of above-ground biomass of woody species in 2005 

occurred later in the summer and early fall. Increases in 

shrub diameter were not very noticeable; however, height 

and foliar density of individual shrubs or trees was very 

apparent based on comparisons of shrubs and trees in site 

photographs. Late winter and early spring precipitation in 

2006 was less than in 2005, but was still about 20 percent 

above average. This precipitation resulted in only modest 

amounts of annual and perennial biomass; however, the fuels 

persisted from the 2005 growing season and added to the 

total 2006 fuel loading with levels approaching, or in many 

cases exceeding, those observed during 2005 (fig. 4).

Scalar values for fine fuels and woody fuels were higher 

at most of the 211 sites. Sites with thin, rocky soils (for example, sites with desert pavement or in exposed bedrock) 

had very little change in the magnitude of response due to 

changes in precipitation from dry to wet years. Moisture 

apparently moved too quickly through the thin soils by percolation or runoff thereby limiting plant-available moisture 

and subsequently plant growth. Other sites that exhibited 

little change in scalar values were sites located on deeper 

soils where soil moisture was high, thereby always encouraging maximum fine and woody fuels. Such sites were often 

typical of drainages where soil fines, moisture, and plant 

nutrients tended to accumulate. Other variances in scalar 

values of fuels could also be observed each year among a 

few of the 211 sites regardless of average sitewide trends 

because of the nonuniform distribution of precipitation where 

thundershowers would deliver rain to some watersheds, 

but not others.
144 USDA Forest Service Proceedings RMRS-P-52. 2008

Hansen and Ostler A Survey of Vegetation and Wildland Fire Hazards on the Nevada Test Site

Figure 4—Presence of fine fuels at Site 99 in Yucca Flat on the 

NTS over three years (2004, 2005, and 2006 had about 4, 67, 

and 20 percent above average precipitation, respectively).

Scalar values for fine fuels and woody fuels were each plotted separately to provide GIS maps showing the color-coded 

values and locations of the 211 sites. Maps of woody fuels 

did not show appreciable changes in scalar values over the 

three-year period, however, an evaluation of maps of fine 

fuels showed that scalar values for sites were more responsive to increases in precipitation than woody fuels. When the 

six-point scalar values of fine fuels were equally weighted 

and added together to create an 11-point scalar index for 

each sampling stations, the resulting map is considered more 

indicative of wildland fire hazards (fig. 5). This combined 

index reflects the fine fuels needed to start a wildland fire 

(for example, catching sparks from a lightning-struck shrub 

or tree) and the woody fuels needed to carry a fire (such as 

raising the quantity of fuels and elevating the flame height). 

An evaluation of old wildland fire sites on the NTS indicated 

that sites that were exclusively fine fuels or woody fuels did 

not maintain wildland fires as well as sites with a combination of fine fuels and woody fuels. This observation was also 

apparent in the field where recently burned areas that were 

plentiful with invasive grasses did not carry a fire until the 

sites had aged sufficiently to produce some woody biomass. 

The presence of strong wind was also observed as a modifying 

condition that could cause wildland fires to burn through 

areas that contained predominantly one type of fuel.

Conclusions _______________________

Ecological field evidence suggests that wildland fires have 

been historically present on the NTS, but are increasing in 

area and importance as invasive grasses provide more fine 

fuels to carry fires ignited by lightning from shrub to shrub 

and tree to tree. Species composition data suggest that plant 

communities like blackbrush provide both woody fuels and 

fine fuels needed for ignition and the spread of wildland 

fires. The majority of areas burned on the NTS during the 

past 50 years have been in the blackbrush plant community 

type, which is recovering very slowly. Most wildland fires are 

started by lightning that tends to occur late in the summer 

when monsoonal storms entering southern Nevada create 

an abundance of dry lightning from convectional local storm 

clouds. It was observed during this study that fine fuels 

that result from high winter and early spring precipitation 

create a major hazard for wildland fires, especially when 

followed by drought conditions and high temperature in the 

summer months. High amounts of fine fuels may also carry 

over into subsequent years. The cost of fighting wildland 

fires is high and the loss of infrastructure, such as power 

lines and communication structures, provides an incentive 

to gain a better understanding of the science and ecology of 

wildland fires on the NTS, an understanding that will be 

complicated by the spread of invasive species and higher 

temperatures that have been observed to increase on the 

NTS in recent years.

Reference herein to any specific commercial product, process, or service by trade name, trademark, manufacturer, 

or otherwise, does not necessarily constitute or imply its 

endorsement, recommendation, or favoring by the United 

States Government or any agency thereof or its contractors 

or subcontractors.
USDA Forest Service Proceedings RMRS-P-52. 2008 145

A Survey of Vegetation and Wildland Fire Hazards on the Nevada Test Site Hansen and Ostler

Figure 5—Combined Fuels Index (fine fuels + woody fuels) during 2006 on

the NTS (darker colors indicate higher combined fuels, while lighter colors 

indicate lower combined fuels).

Acknowledgments__________________

Work performed by National Security Technologies, LLC, 

under Contract No. AC52-06NA25946 with the U.S. Department of Energy.

References ________________________

Hansen, D.J.; Ostler, W.K. 2004. A survey of vegetation and wildland fire hazards on the Nevada Test Site. DOE/NV/11718-981. 

September 2004. Las Vegas, NV: U.S. Department of Energy, 

National Nuclear Security Administration, Nevada Site Office.

Ostler, W.K.; Hansen, D.J.; Anderson, D.C.; Hall, D.B. 2000. Classification of vegetation on the Nevada Test Site. December 6, 

2000. DOE/NV/11718-477. Las Vegas, NV: U.S. Department of 

Energy, Nevada Operations Office.

Randerson, D.; Sanders, J.B. 2006. Characterization of cloud-toground lightning flashes on the Nevada Test Site, 2006. Las 

Vegas, NV: National Oceanic and Atmospheric Administration, 

Air Resources Laboratory/Special Operations and Research 

Division.

Webb, R.H.; Muronv, M.B.; Esque, T.C.; Boyer, D.E.; DeFalco, L.A.; 

Haines, D.F.; Oldershaw, D.; Scoles, S.J.; Thomas, K.A.; Blainey, 

J.B.; Medica, P.A. 2003. Perennial vegetation data from permanent 

plots on the Nevada Test Site, Nye County, Nevada. Open-File 

Report 03-336. Tucson, AZ: U.S. Geological Survey.

Wills, C.A.; Ostler, W.K. 2001. Ecology of the Nevada Test Site: 

An annotated bibliography with narrative summary, keyword 

index, and species lists. December 2001. DOE/NV/11718--594. 

Las Vegas, NV: U.S. Department of Energy, National Nuclear 

Security Administration, Nevada Operations Office.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

USDA Forest Service Proceedings RMRS-P-52. 2008 147

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas 

A.; Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Sherel Goodrich is an Ecologist at the Supervisors Office, Ashley 

National Forest, Vernal, UT.

Abstract—A small prescribed fire near the mouth of Trout Creek 

in Strawberry Valley, Wasatch County, Utah, on the Uinta National Forest provided an opportunity to compare production and 

vascular plant composition in unburned and burned areas. At four 

years post burn, production of herbaceous plants was about four 

times greater in the burned area than in the unburned area. Most 

plants found in the unburned area were found in the burned area 

at four years post burn.

Study Area and Methods ____________

In 1999 a prescribed fire near the mouth of Trout Creek, 

Strawberry Valley on the Uinta National Forest was controlled after it burned about 20 acres. This fire burned in a 

mountain big sagebrush (Artemisia tridentata ssp. vaseyana) 

community. The fire consumed fuels including sagebrush 

stems to ground level. At the south end of the fire, an abrupt 

boundary between the burned and unburned areas (fig. 1) 

provided an opportunity to measure production and determine composition in similar sites in the unburned area (fig.2) 

and in the burned area (figs. 3 and 4). Areas measured in 

and out of the burn were of similar gradient, exposure, and 

landform. They are adjacent to each other and are reasonably 

expected to have a similar history of ungulate use.

Production was measured by clipping and weighing all 

herbaceous growth of the season in transects inside the burn 

and outside the burn on 14 October 2003 or at four years 

post-burn. Clipped plots were taken along parallel transects 

with transects located about 50 ft inside the burn (fig. 3) and 

about 50 ft away from burn in the unburned sagebrush community (fig. 2). Plot size was 9.6 ft2. Ten plots were clipped 

and weighed in each of the burned and unburned areas. 

Clipping took place after growth of herbaceous plants had 

mostly ceased. Thus the standing crop at the time of clipping likely represented most of the graminoid production 

of the year, but it likely under represented forb production. 

Production was measured in grams. Measurements in grams 

from 10 plots of 9.6 ft2 convert directly to pounds per acre 

as reported in table 1.

Trout Creek 1999 Burn

Sherel Goodrich

Figure 1—Abrupt boundary between unburned site 

dominated by sagebrush (foreground) and burned site 

with high production of graminoids (photo taken on 

14 October 2003).

Figure 2—Clip-plot transect in unburned area with 35 

percent crown cover of mountain big sagebrush and 

low production of herbaceous plants (photo taken 

14 October 2003).

All standing crops of herbaceous species were clipped to 

ground level. However, sagebrush production was clipped and 

weighed from parts of some of the plots and the remainder 

was estimated based on the percent clipped. Thus production 

of herbaceous species is based on 100 percent clipping, and 

sagebrush production is based on partial clipping.

Samples of clipped material of various species were placed 

in paper bags. These samples were weighed in the field 

and left in the cab of a vehicle for a few days to dry where 

temperatures went well over 90 degrees Fahrenheit in the 

day. After a few warm days, they were weighed again to 

determine air-dry weight. Production values listed in table 1 

are based on air-dry weight.
148 USDA Forest Service Proceedings RMRS-P-52. 2008

Goodrich Trout Creek 1999 Burn

Table 1—Production (lbs/acre air dry weight) at 

4 years post burn.

Not burned Burned

Graminoid 193 1,088

Forb 59 22

 Total herbaceous 252 1,110

Sagebrush 787 0

Yellowbrush 3 43

 Total shrub 790 43

Total production 1,042 1,153

Figure 4—Burned site on 6 August 2003 at 4 years 

post burn. Native, perennial graminoids dominate the 

burned area.

Figure 3—Clip-plot transect inside the burn with high 

production of herbaceous plants. The mountain big 

sagebrush plant just to the right of the hoop established 

from seed after the fire. By year 4 post burn it is mature 

and producing abundant seed. Area of the hoop is 

9.6 ft2. It was used to delineate clip-plots (photo taken 

14 October 2003).

Crown cover of shrubs was determined by measuring 

crowns of sagebrush along 500 ft of line intercept in both the 

burned and unburned areas following procedures outlined 

in the Rangeland Ecosystem Analysis and Management 

Handbook (USDA FS 1993). Crown cover of all species was 

estimated in a 1/10 acre macro plot in each of the burned 

and unburned areas. Data from the macro plots provides a 

species list or composition of plants present in the burned 

and unburned areas.

Results ___________________________

Total production was similar in the burned (1,153 lbs/acre) 

and the unburned (1,042 lbs/acre) areas (table 1). A range 

of annual production 350 to 1,261 pounds/acre) is indicated 

for mountain big sagebrush communities (Goodrich and 

Huber 2001; Harniss and Murray 1973; Jensen 1989; Tart 

1996; Tew 1988). The trout creek sites are near the upper 

end of this range.

Herbaceous production in the burn was 1,110 lbs/acre 

compared to 252 lbs/acre in the unburned area. There was 

858 lbs/acre (4.4 times) more herbaceous production in 

the burned area. Graminoid production was 895 lbs/acre 

(5.6 times) greater in the burn than in the unburned area. 

Forb production was 37 lbs/acre less in the burned area than 

in the unburned area. However, nearly all of this reduction 

is attributed to a single forb species. Ballhead sandwort 

(Arenaria congesta) contributed 55 of the 59 lbs/acre of forb 

production in the unburned area. Production of this species was 27 lbs/acre in the burned area. The list of species 

made at these sites in August 2003 (table 2) indicates that 

sampling earlier in the summer would likely improve the 

production data for forbs.

Although forb production measured in October was 37 lbs/

acre less in the burned area, 17 forb species were found in 

the burned area in August compared to six forb species in the 

unburned area. As might be expected, a greater number of 

annual forb species was found in the burned area. However, 

there were also more perennial forb species in the burned 

area (table 2). Twelve graminoid species were found in the 

burned area compared to eight in the unburned area. Foxtail 

wheatgrass (Elymus saxicola), recognized as a hybrid cross of 

squirrel tail (Elymus elymoides) and bluebunch wheatgrass 

(Elymus spicatus), was one of the grasses of much greater 

cover in the burned area compared to the unburned area 

(table2). I have noted plants of this taxon to be comparatively 

common in other recent burns.

Four forbs of the Chicory Tribe of the sunflower (Asteraceae) 

family that are likely of high value for sage grouse were found 

in the burn. Only one of these was found in the unburned 

area (table 2). All of these are well adapted to fire.

Crown cover of mountain big sagebrush was measured at 

35.5 percent in the unburned area and zero percent in the 

burned area. Crown cover of yellowbrush (Chrysothamnus 

viscidiflorus var. lanceolatus) was 2.8 and 2.2 percent, respectively, in and out of the burn (table 3).
USDA Forest Service Proceedings RMRS-P-52. 2008 149

Trout Creek 1999 Burn Goodrich

Table 2—Estimated crown cover of species found in 1/10 acre plots on 6 August 2003.

% Crown cover

Common name Scientific name Burned Unburned

Shrubs

Mountain big sagebrush Artemisia tridentata vaseyana 0.2 38.0

Yellowbrush Chrysothamnus viscidiflorus lanceolatus 2.0 3.0

Oregon grape Mahonia repens 0.1 –

Graminoids 

Liddon sedge Carex petasata 0.8 –

Ross sedge Carex rossii 0.2 1

Squirreltail Elymus elymoides 14.0 0.1

Foxtail wheatgrass Elymus saxicola 6.0 0.2

Bluebunch wheatgrass Elymus spicatus 10.0 5

Thickspike wheatgrass Elymus lanceolatus 6.0 –

Junegrass Koeleria macrantha – 0.1

Muttongrass Poa fendleriana 1.0 –

Kentucky bluegrass Poa pratensis 0.1 1

Sandberg bluegrass Poa secunda 0.2 2

Needle-and-thread Stipa comata 1.0 –

Letterman needlegrass Stipa lettermanii 1.0 1.0

Columbia needlegrass Stipa nelsonii 0.5 –

Perennial forbs

Mountain dandelion Agoseris glaucaa 0.1 –

Pussytoes Antennaria microphylla – 0.1

Ballhead sandwort Arenaria congesta 2.0 4

Cainville thistle Cirsium calcareum 0.5 0.1

Whorled buckwheat Eriogonum heracleoides 0.7 0.1

Pale stickseed Hackelia patens 0.4 –

Hoary aster Machaeranthera canescens 0.5 –

Watson penstemon Penstemon watsonii – 0.2

Longleaf phlox Phlox longifolia 0.1 –

Munro globemallow Sphaeralcea munroana 2.0 –

Common dandelion Taraxacum officinalea 0.2 –

Annual and biennial forbs

Musk thistle Carduus nutans 0.1 –

Mountain goosefoot Chenopodium atrovirens 0.2 –

Autumn willowherb Epilobium brachycarpum 0.1 –

Groundsmoke Gayophytum sp. 0.2 –

Western stickseed Lappula occidentalis 0.1 –

Prickly lettuce Lactuca serriolaa 0.1 –

Douglas knotweed Polygonum douglasii 0.1 –

Yellow salsify Tragopogon dubiusa 0.2 0.1

aThese four species of the Chicory Tribe of Asteraceae with milky juice are likely highly selected by sage grouse. 

All of these appear to be well adapted to fire.

Table 3—Percent crown cover of shrubs based on 500 ft of 

line intercept.

Not burned Burned

Sagebrush 35.5 0

Yellowbrush 2.2 2.8

 Total shrub crown cover 37.7 2.8
150 USDA Forest Service Proceedings RMRS-P-52. 2008

Goodrich Trout Creek 1999 Burn

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

Management Implications ___________

The strong contrast of herbaceous production between 

the burned and unburned sites is a clear demonstration 

of increase of herbaceous production that can be achieved 

by burning in mountain big sagebrush communities. The 

contrast strongly supports a concept of competitive relationships between sagebrush and its herbaceous understory as 

suggested by Winward (1991).

However, the mountain big sagebrush ecosystem is extensive and highly variable. Results achieved at Trout Creek 

are likely typical of many areas, but perhaps not typical of 

others.

Work in Idaho (Jensen 1984) and Utah (Woodward 1981; 

Woodward and others 1984) strongly suggests that sites with 

high K-Mg ratios in the soil favor the growth of shrubs and 

sites with low K-Mg ratios favor grass production. On sites 

deficient in magnesium but having adequate potassium, plant 

species having large root cation exchange capacity may be 

at a competitive advantage. In the Utah study dicots were 

found with high root cation exchange capacity and monocots 

(grasses) were found with low cation exchange capacity. Big 

sagebrush was found with about 2.6 times greater root cation 

exchange capacity than bluebunch wheatgrass and other 

perennial grasses included in the Utah study (Woodward 

and others 1984).

These relationships are likely factors in varied responses 

to burning that might be found within sagebrush areas. 

McArthur and Plummer (1978) suggested range managers 

will benefit by learning as much as they can about variation 

in sagebrush. Such differences as adaptation, palatability, 

height, and other features should be considered in management decisions. The same can be said of learning more about 

sites on which sagebrush grows. The more known about 

specific sagebrush sites, the more likely desired outcomes 

will result from management actions.

Winward (1991) suggested release of understory is related 

to crown cover of sagebrush prior to burning. He suggested 

comparatively little release of the herbaceous understory 

where crown cover of mountain big sagebrush was less than 

about 20 percent. Crown cover of mountain big sagebrush 

in the burned area was 0 percent. Crown cover of this shrub 

in the unburned area was 35.5 percent. This is well within 

the range of sagebrush cover suggested by Winward (1991) 

where release of understory species might be expected.

The fourth year post burn production measurements taken 

at this site likely represent the peak in contrast between the 

burned and unburned areas. Measurements of production 

in additional years would give a more complete history of 

response of production to burning.

As reported by Goodrich and others (this proceedings), 

fire can be used as a tool to create diversity in crown cover 

of shrubs. The Trout Creek burn demonstrates diversity 

in herbaceous production and species composition where 

burned areas are adjacent to unburned areas.

All but three species found in the unburned area were 

found in the burned area at four years post burn. This 

indicates resilience and fire adaptability of plants found in 

the mountain big sagebrush communities of this area. Although mountain big sagebrush was killed across all of the 

burn, sagebrush reestablished in the burn from seed, and 

by 2003 (four years post burn) mature plants of sagebrush 

were producing seed (fig. 3).

The Trout Creek Fire burned in a large sagebrush area 

where potential for fire was in the thousands of acres. That 

this burn was controlled after it burned only about 20 acres 

is evidence that small burns are possible within mountain big 

sagebrush areas. This control was achieved by pre-ignition 

preparation and heavy presence of suppression equipment 

and personnel during the burning (Schoppe, personal communication). This example is not suggested as reason to ignore 

the potential for prescribed burns to grow beyond planned 

size. However, the Trout Creek Burn stands as evidence 

that small burns can be achieved in this type.

References ________________________

Goodrich, S.; Huber, A. 2001. Mountain big sagebrush communities 

on the Bishop Conglomerate in the Eastern Uinta Mountains. In: 

McArthur, E.D.; Ostler, K.W.; Wambolt, C.L., comps. Proceedings: shrubland ecotones. Proceedings RMRS-P-11. Ogden, UT: 

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

Research Station: 336-343.

Harniss, R.O.; Murray, R.B. 1973. 30 years of vegetal change 

following burning of sagebrush-grass range. Journal of Range 

Management. 26: 322-325.

Jensen, M.E. 1984. Some edaphic relations of southeastern Idaho 

wildlands. Great Basin Naturalist. 44: 265-271.

Jensen, M.E. 1989. Soil climate and plant community relationships 

on some rangelands of northeastern Nevada. Journal of Range 

Management. 42: 275-280.

McArthur, E.D.; Plummer, A.P. 1978. Biogeography and management of native western shrubs: a case study, section Tridentatae 

of Artemisia. Great Basin Naturalist Memoirs. 2: 229-243.

Schoppe, J. 2006. [Personal communication]. November 20. Panquitch, UT: U.S. Department of Agriculture, Forest Service, 

Intermountain Region, Dixie National Forest.

Tart. D.L. 1996. Big sagebrush plant associations of the Pinedale 

Ranger District. Pinedale, WY: Bridger-Teton National Forest. 

97 p.

Tew, R.K. 1988. Ecosystem stratification on the Fishlake National 

Forest. Richfield, UT: U.S. Department of Agriculture, Forest 

Service, Fish Lake National Forest. 177 p.

U.S. Department of Agriculture, Forest Service. 1993. Rangeland 

ecosystem analysis and management handbook. FSH 2209-21. 

Ogden, UT: U.S. Department of Agriculture, Forest Service, 

Intermountain Region.

Winward, A.H. 1991. A renewed commitment to management 

of sagebrush grasslands. In: Miller, R.F., ed. Management of 

sagebrush steppe. Special Rep. 880. Corvallis, OR: Oregon State 

University Agricultural Experiment Station: 2-7.

Woodward, R.A. 1981. An ecological consideration of the significance 

of cation exchange capacity of roots of some Utah range plants. 

Provo, UT: Brigham Young University. 29 p. Thesis.

Woodward, R.A.; Harper, K.T.; Tiedemann, A.R. 1994. An ecological 

consideration of the significance of cation-exchange capacity of 

roots of some Utah Range Plants. Plant and Soil. 79: 169-180.
USDA Forest Service Proceedings RMRS-P-52. 2008 151

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas 

A.; Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Sherel Goodrich is Ecologist, Supervisors Office, Ashley National 

Forest, Vernal, Utah; email: skgoodrich@fs.fed.us.

Allen Huber is Ecologist, Duchesne-Roosevelt District, Ashley 

National Forest, Roosevelt, Utah.

Brian Monroe is Rangeland Specialist, Cedar City District, 

Dixie National Forest, Cedar City, Utah, and recently Rangeland 

Specialist at Vernal District, Ashley National Forest.

Abstract—Photography and notes on file at the Supervisors Office, Ashley National Forest make it possible to date many fires 

in mountain big sagebrush (Artemisia tridentata ssp. vaseyana) 

communities on this National Forest. Crown cover of mountain big 

sagebrush and other shrubs was measured in repeat visits to many 

burned sites. Burned areas studied varied in age from 1 year to 

42 years. Crown cover measurements in these burns demonstrate 

high capability of mountain big sagebrush to return to burned sites. 

Crown cover of mountain big sagebrush was highly variable in post 

burn environments. After 15 years post burn, crown cover of mountain big sagebrush varied from 4 to 46 percent at the various study 

sites. This variability indicates highly diverse structure and cover 

of mountain big sagebrush in post burn environments. In addition 

to crown cover, ground cover was also measured. These measurements demonstrate rapid return of ground cover in mountain big 

sagebrush communities. Most burned sites had greater than 80 

percent ground cover after 5 years post burn.

Introduction_______________________

Mountain big sagebrush (Artemisia tridentata ssp. 

vaseyana) communities are extensive in montane areas of 

the West. They have high value for sagebrush obligate species such as sage grouse (Centrocercus urophasianus). They 

are a major source of browse and herbaceous forage for wild 

ungulates, and they are among the most highly selected lands 

for livestock grazing in the Intermountain West.

Fire appears to have been a major factor in the ecological 

history of these communities on the Ashley National Forest. 

Evidence of a strong fire history is provided by the understory species and associated shrub species. Nearly all of the 

Trend of Mountain Big Sagebrush Crown 

Cover and Ground Cover on Burned Sites, 

Uinta Mountains and West Tavaputs Plateau, Utah

Sherel Goodrich, Allen Huber, and Brian Monroe

perennial understory species are highly capable of sprouting 

after fire. Most of the associated shrubs sprout vigorously 

after fire. Mountain big sagebrush does not sprout after fire, 

and it is one of the slower species to return in abundance 

to burned areas. Although it is slower to recover, mountain 

big sagebrush generally returns to burned areas.

Capability to produce fuel is another indicator of fire potential. A range of annual production of 392 to 1,413 kg/ha (350 to 

1261 lbs/acre) is indicated for mountain big sagebrush communities (Goodrich and Huber 2001; Harniss and Murray 

1973; Jensen 1989; Tart 1996; Tew 1988). A value near the 

low end of the range (418 kg/ha or 373 lbs/acre) was reported 

from near Dubois, Idaho, (Harniss and Murray 1973) where 

sample sites were likely near the ecotone with Wyoming big 

sagebrush (Artemisia tridentata var. wyomingensis). Production was generally lower where bluebunch wheatgrass 

(Elymus spicatus) and blue gramma (Bouteloua gracilis) were 

indicators of community type than where mountain snowberry (Symphoricarpos oreophilus) and slender wheatgrass 

(Elymus trachycaulus) were indicators of community type.

Litter and vegetation often cover over 85 percent of the 

ground surface. Biomass production and continuous fuels of 

these communities have high potential to carry fire. These 

factors are major differences of fire potential between mountain big sagebrush communities and Wyoming big sagebrush 

communities. Both productivity and ground cover are lower 

in Wyoming big sagebrush communities than in mountain 

big sagebrush communities (Goodrich and Huber 2001).

History of fire is also indicated for mountain big sagebrush 

communities by the relative ease with which fire is started 

and carried in these communities. In the area of this study, 

mountain big sagebrush communities have been prescribedburned under weather conditions that did not support fire 

in adjacent communities.

At mountain big sagebrush sites of the Uinta Mountains 

with a history of about 100 years of livestock grazing, ground 

cover has been found at 87 to 95 percent with an average 

of 92 percent (Goodrich and Huber 2001). For the Fishlake 

National Forest, Tew (1988) considered mountain big sagebrush communities at potential to have between 7 and 22 

percent bare soil. This indicates ground cover potential of 

between 78 and 93 percent for that National Forest. Tart 

(1996) listed percent bare soil at 8 to 13 percent for two 

mountain big sagebrush plant associations at late seral 

condition on the West Flank of the Wind River Mountains. 

This indicates potential for ground cover of about 87 to 92 

percent or about the same as data from the Uinta Mountains 

indicates (Goodrich and Huber 2001).
152 USDA Forest Service Proceedings RMRS-P-52. 2008

Goodrich, Huber, and Monroe Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites,...

Ground cover is greatly reduced by fire. Plants and litter are consumed leaving ash and rock for ground cover. 

Ash is readily removed by wind and by overland flow of 

water. Although ash might provide some protection against 

raindrop-splash, its overall value for watershed protection is 

relatively low. The most fire-resistant ground cover is rock. 

However, some big sagebrush communities have low percent 

rock cover. Depending on rock cover, mountain big sagebrush 

communities can be left with essentially no resistant ground 

cover immediately following fire.

Return intervals for sagebrush crown cover and for ground 

cover after fire are important to the management of these 

plant communities. Considerations for rehabilitation following fire should be based on the inherent capability of 

plant communities to provide ground cover following fire. 

Also considerations for grazing and other management 

practices need to be coordinated with return of ground cover 

and development of the flora following fire. Dynamics of 

sagebrush crown cover and ground cover following fire are 

the focus of this study.

Study Area and Site Selection ________

This study was conducted in the Uinta Mountains and 

on the Tavaputs Plateau on the Ashley National Forest in 

northeastern Utah. Crown cover data (table 1) and ground 

cover data (table 2) were taken from many sites. Some sites 

were visited more than once, and these sites provide multiple 

reference points for comparing time with return of crown 

cover of shrubs and ground cover. Each site represents an 

area of about 0.28 ha (0.72 acre). All sites selected for study 

were in mountain big sagebrush (Artemisia tridentata ssp. 

vaseyana)/grass communities with a known history of fire. 

Burned areas of this study varied in age from 1 to 40 years 

(table 1).

Yellowbrush (Chrysothamnus viscidiflorus var. lanceolatus) and mountain snowberry (Symphoricarpos oreophilus) 

were present in 75 percent and 65 percent of the readings 

respectively. At the majority of sites, mountain needle-andthread (Stipa comata var. intermedia) was the understory 

plant that indicated community type. On some sites of warm 

aspect, bluebunch wheatgrass (Elymus spicatus) indicated 

community type. At a few sites on the Tavaputs Plateau, 

Salina wildrye (Elymus salinus) was the understory-indicator 

of community type.

Livestock grazing has been a part of the history of most 

of the study sites for over 100 years. Livestock grazing has 

been discontinued for a decade or more on a few of the sites. 

We did not find a sufficient number of sites without current 

livestock grazing to make a comparison between livestockgrazed and ungrazed sites. All of the study sites have a history 

of use by elk (Cervus elaphus) and mule deer (Odocoileus 

hemionus). However, elk and mule deer have used the study 

sites only in light winters, and use of sagebrush by these 

animals has been light.

Methods __________________________

Burned sites and years since each site was burned was 

determined from monitoring studies filed at the Supervisor’s 

Office and District Offices of the Ashley National Forest. Sites 

with a known history of fire were visited, and permanent 

belt-line transects were established in these burned areas. 

In some cases permanent transects had been established 

prior to burns. In these cases, data was taken along existing transects.

At most sites, crown cover of shrubs was determined from 

152 m (500 ft) of line intercept. At a few sites, line intercept 

measurements were limited to 91 m (300 ft). Protocol found 

in the Rangeland Ecosystem Analysis and Management 

Handbook (USDA FS 1993) was used for taking line intercept data. This included measuring the intercept of each 

shrub along the transect line by species. Gaps in crowns of 

shrubs of 15 cm (0.5 ft) or greater were omitted from crown 

cover measurements.

Ground cover was determined from 400 points along the 

same belt transects used to determine crown cover of shrubs. 

Point size was the sharp end of a spike or about 1 mm. Basal 

area of vegetation, plant litter, and rock 2 cm (0.75 in) or 

greater in diameter were considered ground cover. Plant 

litter in contact with the soil surface was considered ground 

cover. Standing litter was not included. Bare soil and rock 

or pavement less than 2 cm diameter were not included in 

ground cover.

Results ___________________________

Return of crown cover of mountain big sagebrush and other 

shrubs was highly variable as demonstrated in table 1 and 

in the scatter diagram of figure 1. Crown cover of mountain 

big sagebrush varied from 4 percent to 46 percent after 15 

years post burn. After 15 years post burn, crown cover of 

mountain big sagebrush was greater than 15 percent at most 

of the burned sites. Shrubs other than sagebrush included 

bitterbrush (Purshia tridentata), mountain snowberry and 

yellowbrush. Mountain snowberry is particularly capable of 

rapid increase in cover after fire. At some sites and especially 

on the Tavaputs Plateau, yellowbrush increased rapidly after 

fire. Crown cover of other shrubs appeared to decrease as 

mountain big sagebrush increased in cover (fig. 1).

Ground cover increased rapidly following fire (table 2 and 

fig. 2). After 5 years, ground cover was 80 percent or greater 

on most burned sites.

Discussion ________________________

Return of sagebrush crown cover was highly variable. 

Some of the variation was likely site related. Work in Idaho 

(Jensen 1984) and Utah (Woodward 1981; Woodward and 

others 1984) strongly suggests that sites with high K-Mg 

ratios in the soil favor the growth of shrubs and sites with 

low K-Mg ratios favor grass production. On sites deficient 

in magnesium but having adequate potassium, plant species having large root-cation exchange capacity may be at 

a competitive advantage. In the Utah study dicots were 

found with high root-cation exchange capacity and monocots 

(grasses) were found with low cation exchange capacity. Big 

sagebrush was found with about 2.6 times greater root-cation 

exchange capacity than bluebunch wheatgrass and other 

perennial grasses included in the study (Woodward and 

others 1984). Recovery of sagebrush crown cover might be 

related to elemental ratios in the soil.
USDA Forest Service Proceedings RMRS-P-52. 2008 153

Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites,... Goodrich, Huber, and Monroe Table 1—Postburn crown cover of shrubs (percent) (ARTRV: mountain big sagebrush. CHVIL4: yellowbrush. PUTR2: bitterbrush. SYOR: mountain snowberry. ARSP8: spiked big sagebrush, CHNA2: rubber rabbitbrush). Study # Year burned Year read # YEARS ARTRV CHVIL4 PUTR2 SYOR AMAL2 ARSP8 CHNA2 Total 37-4B 2002 2003 01 4.7 0.3 0.0 0.0 0.0 0.0 0.0 5.0 37-22A 1999 2000 01 0.0 4.0 0.0 2.4 0.0 0.0 0.0 6.4 37-22A 1999 2000 01 0.0 4.0 0.0 2.4 0.0 0.0 0.0 6.4 37-29A 2002 2003 01 0.1 0.4 0.0 0.8 0.0 0.0 0.0 1.3 39-48A 1998 1998 01 0.0 0.0 0.1 2.7 1.3 0.0 0.0 4.1 42-8A 1988 1989 01 0.7 0.5 0.0 0.8 0.0 0.0 0.0 2.0 42-8B 1988 1989 01 0.0 0.3 0.0 0.3 0.0 0.0 0.0 0.6 42-11 1990 1991 01 0.1 0.3 0.9 0.1 0.2 0.0 0.0 1.6 44-3B 2002 2002 01 0.9 0.5 0.0 0.0 0.0 0.0 0.0 1.4 65-21 2002 2003 01 0.0 1.4 0.1 2.4 0.1 0.0 0.0 4.0 67-83C 1999 2000 01 0.0 0.1 0.0 1.6 0.0 0.0 0.0 1.7 67-93 2000 2001 01 0.0 0.0 0.0 0.5 0.0 0.0 4.7 5.2 67-93 2000 2005 01 0.0 0.0 0.0 1.0 0.0 0.0 2.3 3.3 68-62A 1998 1999 01 0.0 8.7 4.1 0.0 0.0 0.0 0.0 12.8 19-30 1993 1995 02 0.0 1.1 0.3 0.0 0.0 0.0 0.0 1.4 32-78 1993 1995 02 0.2 0.7 4.8 0.0 0.0 0.0 0.0 5.7 32-85A 1998 2000 02 0.0 0.0 0.0 0.2 0.0 0.0 0.0 0.2 37-22A 1999 2001 02 0.0 10.2 0.0 2.1 0.0 0.0 0.0 12.3 37-24C 1999 2001 02 0.0 2.6 1.7 4.4 0.0 0.0 0.0 8.7 37-28A 1999 2001 02 0.0 3.1 0.0 9.4 0.3 0.0 0.0 12.8 39-28 1999 2001 02 0.1 0.0 0.0 4.3 0.0 0.0 0.0 4.4 42-10 1991 1993 02 0.1 1.0 0.0 30.2 0.0 0.0 0.0 31.3 43-24 1988 1990 02 0.1 0.3 0.0 0.0 0.0 0.0 0.0 0.4 64-1 1987 1989 02 0.3 0.0 0.0 0.0 0.0 0.0 0.0 0.3 65-21 2002 2004 02 0.0 7.3 0.3 7.3 0.5 0.0 0.0 15.4 67-86A 1999 2001 02 0.0 4.2 0.0 7.4 0.0 0.0 0.0 11.6 32-85A 1998 2001 03 0.3 0.1 0.0 0.3 0.0 0.0 0.0 0.7 32-85W 1998 2001 03 0.0 0.3 0.0 1.3 0.0 0.0 0.0 1.6 37-19A 1998 2001 03 0.0 0.6 0.0 13.9 0.6 0.0 0.0 15.1 37-19B 1998 2001 03 0.0 0.4 6.0 4.2 0.1 0.0 0.0 10.7 39-22B 1999 2002 03 0.0 0.0 0.0 0.4 0.0 0.0 0.0 0.4 39-48A 1998 2001 03 0.0 0.0 0.2 3.5 3.2 0.0 0.0 6.9 40-5C 1999 2002 03 3.7 10.9 1.0 0.0 4.4 0.0 0.0 20.0 43-28 1988 1991 03 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 44-26 2000 2003 03 0.1 0.6 0.0 1.0 0.0 0.0 0.0 1.7
154 USDA Forest Service Proceedings RMRS-P-52. 2008

Goodrich, Huber, and Monroe Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites,... 51-1 1998 2001 03 8.1 1.7 2.3 23.5 0.9 0.0 0.0 36.5 51-6A 1998 2001 03 2.7 2.6 0.0 8.0 0.0 0.0 0.0 13.3 51-15B 1999 2002 03 0.1 0.8 0.0 3.1 0.1 0.0 0.0 4.1 67-68 1999 2002 03 0.0 1.2 0.0 0.6 0.0 0.0 0.0 1.8 67-76 1999 2002 03 0.4 4.5 0.0 0.6 0.0 0.0 0.0 5.5 67-83B 1999 2002 03 0.3 0.4 0.0 7.7 0.0 0.0 0.0 8.4 31-10A 1998 2002 04 0.0 0.0 0.0 0.1 0.0 0.0 0.0 0.1 31-10H2 1998 2002 04 0.0 0.7 0.0 15.2 0.0 0.0 0.0 15.9 39-34A4 1999 2003 04 0.0 0.0 0.0 7.6 0.0 0.0 0.0 7.6 43-17 1988 1992 04 0.1 0.0 0.0 0.0 0.0 0.0 0.0 0.1 51-9B 1998 2002 04 0.0 4.6 0.0 1.1 0.0 0.0 0.0 5.7 68-67A 1998 2002 04 0.0 1.9 0.0 0.9 0.0 0.0 0.0 2.8 68-67B 1998 2002 04 0.3 2.3 0.0 0.0 0.0 0.0 0.0 2.6 68-67C 1998 2002 04 0.0 7.8 0.0 0.7 0.0 0.0 0.0 8.5 19-30 1993 1998 05 0.0 2.5 0.9 0.0 0.0 0.0 0.0 3.4 32-70 1993 1998 05 1.5 2.7 0.0 0.0 0.0 0.0 0.0 4.2 64-1 1987 1992 05 11.8 0.1 0.0 0.0 0.0 0.0 0.0 11.9 37-19B 1998 2004 06 0.1 3.6 7.6 3.9 0.0 0.0 0.0 15.2 42-11 1990 1996 06 1.5 1.2 8.1 0.0 0.1 0.0 0.0 10.9 43-24 1988 1994 06 0.7 1.2 0.0 0.1 0.0 0.0 0.0 2.0 43-46B 1994 2000 06 0.0 2.4 0.0 20.0 0.0 0.0 0.0 22.4 58-7 1992 1998 06 0.6 40.3 0.8 4.6 0.8 0.0 0.0 47.1 68-8 1996 2002 06 0.0 0.1 0.0 0.0 0.0 0.0 0.0 0.1 68-51 1996 2002 06 0.0 0.3 0.0 0.2 0.0 0.0 0.0 0.5 68-54 1996 2002 06 4.6 0.7 0.0 0.4 0.0 0.0 0.0 5.7 32-78 1993 2000 07 1.2 1.8 8.8 0.7 0.0 0.0 0.0 12.5 41-8A 1989 1996 07 0.0 0.0 0.0 2.3 0.0 0.0 0.0 2.3 42-4B 1988 1995 07 3.4 0.3 0.1 0.0 0.0 0.0 0.0 3.8 42-15 1988 1995 07 0.1 0.0 0.6 0.0 0.6 0.0 0.0 1.3 19-30 1993 2001 08 0.2 3.5 2.3 0.0 0.0 0.0 0.0 6.0 32-83B 1993 2001 08 2.4 0.6 0.0 0.0 0.0 0.0 0.0 3.0 32-83C 1993 2001 08 0.1 5.7 0.0 0.0 0.0 0.0 0.0 5.8 42-8A 1988 1996 08 2.8 11.3 0.0 0.7 0.0 0.0 0.0 14.8 42-8B 1988 1996 08 3.5 3.4 0.0 0.0 0.0 0.0 0.0 6.9 43-28 1988 1996 08 0.7 0.7 0.7 0.0 0.0 0.0 0.0 2.1 19-30 1993 2002 09 0.2 4.1 2.2 0.0 0.0 0.0 0.0 6.5

Table 1—(Continued)

Study # Year burned Year read # YEARS ARTRV CHVIL4 PUTR2 SYOR AMAL2 ARSP8 CHNA2 Total
USDA Forest Service Proceedings RMRS-P-52. 2008 155

Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites,... Goodrich, Huber, and Monroe 32-70 1993 2002 09 4.0 2.6 0.0 0.5 0.0 0.0 0.0 7.1 32-78 1993 2002 09 2.1 3.8 9.8 0.7 0.0 0.0 0.0 16.4 38-14 1990 2000 10 0.2 3.5 0.0 0.0 0.0 0.0 0.0 3.7 38-41B 1990 2000 10 0.0 0.1 0.0 1.0 0.0 0.0 0.0 1.1 42-4B 1988 1998 10 6.3 0.7 0.1 0.0 0.0 0.0 0.0 7.1 43-14C 1988 1998 10 8.8 0.4 1.3 0.0 0.0 0.0 0.0 10.5 43-46B 1994 2004 10 0.8 1.2 0.0 24.0 0.0 0.0 0.0 26.0 68-60A 1992 2002 10 3.9 0.1 0.0 3.5 0.0 0.0 5.6 13.1 68-60B 1992 2002 10 18.4 0.3 0.0 0.2 0.0 0.0 0.0 18.9 42-10 1991 2002 11 0.5 1.4 0.0 30.8 2.8 0.0 0.0 35.5 42-11 1990 2001 11 4.1 0.9 12.6 0.1 0.7 0.0 0.0 18.4 64-1 1987 1998 11 34.8 0.0 0.0 0.0 0.0 0.0 0.0 34.8 68-104 1992 2003 11 5.1 10.0 0.0 0.9 0.0 0.0 0.0 16.0 42-4B 1988 2000 12 8.2 0.3 0.5 0.0 0.0 0.0 0.0 9.0 42-6 1988 2000 12 1.7 0.8 9.8 0.0 0.0 0.0 0.0 12.3 41-8A 1989 2002 13 0.4 0.0 0.0 2.4 0.0 0.0 0.0 2.8 41-10A 1989 2002 13 0.9 0.0 0.0 3.3 0.0 0.0 0.0 4.2 42-8B 1988 2001 13 7.4 3.8 0.0 0.1 0.0 0.0 0.0 11.3 43-24 1988 2001 13 1.8 0.5 0.0 0.0 0.0 0.0 0.0 2.3 43-24B 1988 2001 13 0.2 0.8 0.0 0.0 0.0 0.0 0.0 1.0 64-1 1987 2001 13 30.9 0.0 0.0 0.0 0.0 0.0 0.0 30.9 42-4B 1988 2002 14 10.3 0.3 0.0 0.0 0.0 0.0 0.9 11.5 43-14C 1988 2002 14 11.3 0.5 2.7 0.1 0.0 0.0 0.0 14.6 43-28 1988 2002 14 2.5 0.9 0.6 0.5 0.0 0.0 0.0 4.5 41-6H 1985 2000 15 0.9 0.8 21.5 6.0 0.0 0.0 0.0 29.2 44-9 1976 1991 15 5.6 7.5 0.0 1.3 0.0 0.0 0.0 14.4 41-6A3B 1985 2002 17 19.8 0.1 0.4 0.6 0.0 0.0 0.0 20.9 41-6C 1985 2002 17 1.4 0.0 0.0 0.0 0.0 0.0 0.0 1.4 42-15 1988 2005 17 1.6 0.0 1.3 0.0 1.8 0.0 0.0 4.7 43-54 1980 1997 17 6.2 4.3 0.0 13.5 0.0 0.0 0.0 24.0 39-10 1961 1979 18 18.6 0.0 0.9 4.5 0.0 0.0 0.0 24.0 39-28 1978 1996 18 22.4 0.0 0.0 11.6 0.1 0.0 0.0 34.1 38-9 1978 1997 19 32.3 0.1 0.0 15.3 0.0 0.0 0.0 47.7 38-33 1978 1997 19 46.0 0.0 0.0 18 0.0 0.0 0.0 64.0 42-39 1978 1997 19 6.3 1.5 16.7 0.0 0.0 0.0 0.2 24.7 43-8A1 1978 1997 19 22.7 0.0 0.0 2.6 0.0 0.0 0.0 25.3

Table 1—(Continued)

Study # Year burned Year read # YEARS ARTRV CHVIL4 PUTR2 SYOR AMAL2 ARSP8 CHNA2 Total
156 USDA Forest Service Proceedings RMRS-P-52. 2008

Goodrich, Huber, and Monroe Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites,... Table 1—(Continued) 64-11 1982 2001 19 28.2 6.6 0.0 1.0 0.0 0.0 0.0 35.8 68-10 1977 1996 19 21.4 2.8 0.0 9.1 0.0 0.0 0.0 33.3 68-9 1976 1996 20 33.5 0.7 0.0 0.3 0.0 0.0 0.0 34.5 42-39 1978 1999 21 6.6 1.3 20.7 0.0 0.0 0.0 0.0 28.6 44-1A 1976 1997 21 26.0 0.6 0.0 7.3 0.0 0.0 0.0 33.9 44-1B1 1976 1998 21 38.0 0.0 0.0 1.5 0.0 0.0 0.0 39.5 44-9 1976 1997 21 10.0 10.6 0.0 0.1 0.0 0.0 0.0 20.7 43-54 1980 2002 22 7.1 3.4 0.0 14.0 0.0 0.0 0.0 24.5 38-35A 1978 2001 23 20.1 0.0 0.0 6.6 0.0 0.0 0.0 26.7 39-15U2 1979 2002 23 29.1 0.0 1.0 0.0 0.0 0.0 0.0 30.1 38-9 1978 2002 24 32.6 0.0 0.0 13.4 0.0 0.0 0.0 46.0 39-15U 1979 2003 24 43.7 0.0 0.0 0.1 0.0 0.0 0.0 43.8 42-39 1978 2002 24 7.6 1.2 18.5 0.0 0.0 0.0 0.0 27.3 43-8A1 1978 2002 24 31.0 0.1 0.0 2.3 0.0 0.0 0.0 33.4 64-6A 1978 2002 24 7.5 20.7 0.0 0.5 0.0 16.8 0.0 45.5 64-18A 1978 2002 24 8.8 1.2 0.0 0.0 0.0 0.0 0.0 10.0 68-32 1978 2002 24 26.7 0.1 0.0 0.0 0.0 0.0 0.3 27.1 68-10 1977 2002 25 38.2 2.0 0.0 4.4 0.0 0.0 0.0 44.6 68-11B 1977 2002 25 12.3 0.8 0.0 4.8 0.0 0.0 0.0 17.9 68-9 1976 2002 26 33.0 0.1 0.0 0.5 0.0 0.0 0.0 33.6 68-13A 1976 2002 26 29.0 0.0 0.0 0.0 0.0 0.0 0.0 29.0 68-18B 1976 2002 26 33.0 0.0 0.0 0.2 0.0 0.0 0.0 33.2 41-6A5B 1978 2005 27 29.5 0.2 0.4 0.0 0.0 0.0 0.0 30.1 41-6A6B 1978 2005 27 28.7 0.5 2.2 0.1 0.1 0.0 0.0 31.6 44-1A 1976 2003 27 33.1 0.2 0.0 7.8 0.0 0.0 0.0 41.1 44-9 1976 2003 27 14.4 7.5 0.0 0.0 0.0 0.0 0.0 21.9 39-10 1961 1996 35 17.8 0.0 0.6 0.7 0.0 0.0 0.0 19.1 39-10 1961 2001 40 36.6 0.0 0.0 0.0 0.0 0.0 0.0 36.6 39-10 1961 2003 42 32.3 0.0 0.3 0.0 0.0 0.0 0.0 32.6 Study # Year burned Year read # YEARS ARTRV CHVIL4 PUTR2 SYOR AMAL2 ARSP8 CHNA2 Total
USDA Forest Service Proceedings RMRS-P-52. 2008 157

Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites,... Goodrich, Huber, and Monroe

Table 2—Trend of ground cover following fire

Study #

 Year

burned

Year

read

Years

postfirePercent

Cover

5-11 1976 1991 15 91

5-11 1992 1997 5 91

5-61A 1999 2002 3 63

19-30 1993 2002 9 nd

31-10A 1998 2002 4 76

31-10H2 1998 2002 4 84

32-70 1993 2002 9 82

32-78 1993 2002 9 87

32-83B 1993 2001 8 90

32-83C 1993 2001 8 nd

32-85A 1998 2001 3 81

32-85W 1998 2001 3 82

37-4B 2002 2003 1 92

37-12A 1998 2002 4 92

37-19A 1998 2001 3 81

37-19B 1998 2004 6 80

37-22A 1999 2001 2 81.5

37-24C 1999 2001 2 70

37-28A 1999 2001 2 14

37-29A 2002 2003 1 85

38-9 1978 2002 24 86

38-14 1990 2000 10 97

39-10 1961 2003 42 87

39-15U 1979 2003 24 93

39-15U2 1979 2002 23 83

39-22B 1999 2002 3 92

39-27C2 2000 2002 2 30

39-34A4 1999 2003 4 97

40-5C 1999 2002 3 75

41-8A 1989 2002 13 93

41-10A 1989 2002 13 81

42-8A 1988 1989 1 90

42-8A 1988 1996 8 83

42-8B 1988 1989 1 45

42-10 1991 1993 2 92

42-10 1991 2002 11 87

42-11 1990 1991 1 59

42-39 1978 2002 24 77

43-8A1 1978 2002 24 94

43-14C 1988 1998 10 87

43-14C 1988 2002 14 87

43-17 1988 1992 4 76

43-24 1988 1990 2 83

43-24 1988 1994 6 96

43-24 1988 2001 13 94

43-24B 1988 2001 13 85

43-28 1988 1991 3 76

43-28 1988 1996 8 88

43-28 1988 2002 14 84

44-1A 1976 2003 27 95

44-3B 2002 2002 0 61

44-9 1976 1991 15 84

44-9 1976 2003 27 92

44-26 2000 2003 3 85

51-6A 1998 2001 3 90

51-9B 1998 2002 4 84

51-15B 1999 2002 3 78

64-1 1987 1989 2 31

64-1 1987 1992 5 47

64-1 1987 2001 14 59

64-6 1978 1980 2 71

64-6A 1978 2002 24 94

64-18A 1978 2002 24 91

65-21 2002 2003 1 11

65-21 2002 2004 2 46

67-68 1999 2000 1 11

67-68 1999 2002 3 63

67-76 1999 2002 3 64

67-83B 1999 2002 3 56

67-86A 1999 2001 2 48

68-8 1996 2002 6 95

68-9 1976 2002 26 78

68-10 1977 2002 25 89

68-11B 1977 2002 25 97

68-13A 1976 2002 26 79

68-18B 1976 2002 26 81

68-32 1978 2002 24 80

68-51 1996 2002 6 82

68-54 1996 2002 6 91

68-60A 1992 2002 10 77

68-60B 1992 2002 10 77

68-67A 1998 2002 4 58

68-67B 1998 2002 4 70

68-67C 1998 2002 4 71

68-72A 1999 2002 3 74

Study #

 Year

burned

Year

read

Years

postfirePercent

Cover
158 USDA Forest Service Proceedings RMRS-P-52. 2008

Goodrich, Huber, and Monroe Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites,...

Figure 1—Trend of shrub crown cover following fire.

Figure 2—Trend of gound cover following fire.
USDA Forest Service Proceedings RMRS-P-52. 2008 159

Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites,... Goodrich, Huber, and Monroe

Twenty percent or greater crown cover of mature sagebrush capable of high seed production in pre-burn communities appears to be important to recruitment of sagebrush 

seedlings in post-burn communities. This seemed to be a 

dominant feature of stands in which there was rapid return 

of sagebrush cover.

Weather patterns could also be a factor. Low recruitment 

of sagebrush seedlings in post-burn communities might 

be expected following fire in years of low seed production. 

Seedling establishment could also be influenced by season 

and amount of precipitation following fire. At several of the 

monitoring sites of this study, high numbers of sagebrush 

seedlings were observed in the favorable years of 2004 and 

2005. It seems noteworthy to point out this high recruitment 

of sagebrush seedlings followed a drought in 2002.

Sites of this study include various intensities of use by 

livestock. This could also be a factor in the variable rate of 

return of sagebrush at our sites. Return of sagebrush might 

be expected to be accelerated by increasing levels of livestock 

grazing. As stated above, there were too few sites without a 

current history of livestock use to test this concept.

In addition to measurements of crown cover of shrubs, 

observations were made of herbaceous species at burned 

sites of this study. The herbaceous layer was dominated by 

native species at most sites. Native species seen in adjacent 

unburned sites and pre-burn sites were generally present in 

post burn sites. Essentially all native perennial herbaceous 

species sprouted after fire. This is consistent with return of 

species recorded for a burned site in Strawberry Valley, Utah 

(Goodrich, this proceedings). These observations indicate the 

indigenous herbaceous species are highly adapted to fire.

Shrubs except sagebrush sprouted after fire. Bitterbrush 

showed various capacities to sprout, and the other shrub 

species showed vigorous sprouting. Rubber rabbitbrush 

(Chrysothamnus nauseosus) was found at only one site, 

and, as expected, it sprouted after fire. However, it did not 

show increase in cover between the first and fifth years post 

burn. Spiked big sagebrush was found at one high elevation 

site where it provided nearly 17 percent crown cover at 24 

years post burn while mountain big sagebrush had returned 

to about 8 percent. Higher crown cover value for spiked big 

sagebrush likely reflects its ability to sprout following fire. 

Serviceberry (Amelanchier alnifolia) was found with minor 

cover at several sites. This reflects influence from mountain 

brush communities that are often found near mountain big 

sagebrush communities of this study area.

Crown cover of other shrubs appeared to decrease as 

mountain big sagebrush increased in cover (fig. 1). Measurements of several other sites without a history of burning 

(not included in this study) indicated lower cover values for 

other shrubs where crown cover of mountain big sagebrush 

exceeded 30 percent.

Management Implications ___________

The following management implications are intended for 

the area of this study only. They are not intended for broad 

extrapolation to other areas. Sites of this study are typically 

not winter range for deer and elk, and use of sagebrush 

by these animals has been light. Implications of this paper 

should not be applied to areas where wild ungulates concentrate in winter.

High variability of shrub cover in post burn environments 

demonstrates diversity in cover and structure achieved by 

fire. Although the data indicate low predictability for precise 

return intervals of sagebrush at a given site, the variability 

is a positive factor where diversity of habitat is desired. 

The data show sagebrush strongly trending toward greater 

than 25 percent crown cover in the absence of fire or other 

disturbance. The data strongly indicate that without some 

disturbance that reduces sagebrush cover, the great majority 

of the mountain big sagebrush type on the Ashley National 

Forest will persist in high percent cover of sagebrush. Where 

a diversity of sagebrush cover is desired, fire is demonstrated 

as an effective tool to achieve diversity.

The data strongly indicate many stands of mountain big 

sagebrush are sustainable under a fire regime that allows 

sagebrush to mature to 20 percent or more crown cover with 

a high level of seed production. Greater than 20 percent 

crown cover is documented for many stands of mountain 

big sagebrush within 25 years of burning. The data suggest 

sustainability of many stands of mountain big sagebrush at 

a fire interval of 25 to 30 years. However, we recommend 

that the application of fire be based on actual recovery of 

sagebrush crown cover at individual sites rather than on 

modeled fire intervals.

A 30-year interval would indicate that burning an average 

of about 3.3 percent per year of a mountain big sagebrush 

landscape would maintain highly diverse cover of sagebrush 

with mature stands well represented on the landscape. 

Where sagebrush was able to fully recover from fire within 

20 years, about 5 percent of the landscape per year might 

be burned and still maintain high percent canopy cover of 

sagebrush over most of the landscape.

Unless the understory has been displaced by cheatgrass 

(Bromus tectorum) or other highly invasive species, there 

appears to be no need for seeding mountain big sagebrush 

communities following fire in the area of this study. The 

indigenous shrubs and associated understory species have 

demonstrated high resilience and adaptability to fire, and 

they can be expected to provide cover for watershed protection and wildlife habitat.

Ground cover returned to greater than 80 percent within 

5 years after fire at most burned sites. At many sites this 

recovery was concurrent with light to moderate livestock 

grazing with some rest from livestock grazing for 1 and 

sometimes 2 growing seasons after fire.

Litter was the major component in the increase of ground 

cover. Basal area of live-plant cover was comparatively 

minor. However, litter production is dependent on plant 

production. Grazing intensities that remove a major portion 

of production in the early post-fire years can be expected 

to prolong recovery of ground cover. However, our studies 

indicate rapid recovery of ground cover can be concurrent 

with managed livestock grazing in mountain big sagebrush/

grass communities. In some cases rapid recovery was achieved 

where light use (up to about 30 percent use by weight) was 

allowed the first year after fire where this use was delayed 

until fall after herbaceous species were mature.
160 USDA Forest Service Proceedings RMRS-P-52. 2008

Goodrich, Huber, and Monroe Trend of Mountain Big Sagebrush Crown Cover and Ground Cover on Burned Sites,...

References ________________________

Goodrich, S.; Huber, A. 2001. Mountain big sagebrush communities on the Bishop Conglomerate in the Eastern Uinta 

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comps. Proceedings: shrubland ecotones. Proceedings RMRS-P-11. 

Ogden, UT: U.S. Department of Agriculture, Forest Service, 

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following burning of sagebrush-grass range. Journal of Range 

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Jensen, M.E. 1984. Some edaphic relations of southeastern Idaho 

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Tart, D.L. 1996. Big sagebrush plant associations of the Pinedale 

Ranger District. Pinedale, WY: Bridger-Teton National Forest. 

97 p.

Tew, R.K. 1988. Ecosystem stratification on the Fishlake National 

Forest. Richfield, UT: U.S. Department of Agriculture, Forest 

Service, Fish Lake National Forest. 177 p.

U.S. Department of Agriculture, Forest Service. 1993. Rangeland 

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Ogden, UT: U.S. Department of Agriculture, Forest Service, 

Intermountain Region.

Woodward, R.A. 1981. An ecological consideration of the significance 

of cation exchange capacity of roots of some Utah range plants. 

Master Thesis. Provo, UT: Brigham Young University. 29 p.

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The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein
USDA Forest Service Proceedings RMRS-P-52. 2008 161

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas 

A.; Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Corey A. Moffet and J. Bret Taylor are Rangeland Scientist and 

Animal Scientist, respectively, USDA-ARS U.S. Sheep Experiment 

Station, Dubois, ID; email: cmoffet@pw.ars.usda.gov.

D. Terrance Booth is Rangeland Scientist, USDA-ARS High 

Plains Grasslands Research Station, Cheyenne, WY; email: Terry.

Booth@ars.usda.gov.

Abstract—Very large scale aerial (VLSA) imagery is an efficient 

tool for monitoring bare ground and cover on extensive rangelands. 

This study was conducted to determine whether VLSA images could 

be used to detect differences in antelope bitterbrush (Purshia tridentata Pursh DC) cover and density among similar ecological sites 

with varying postfire recovery periods. In 2005, VLSA images were 

acquired at 253 points from high bitterbrush potential ecological 

sites at the USDA-ARS, U.S. Sheep Experiment Station, Dubois, 

Idaho. For each image, fire history was classified and bitterbrush 

density and cover were measured. Bitterbrush cover in images with 

no recorded history of fire during the previous 68 yr (22 percent of 

all images) was 1.71 percent and density was 875 plant ha–1. Areas 

with postfire recovery interval between 10 and 68 yr (60 percentage 

of all images) had bitterbrush density (587 plant ha–1) and cover 

(1.23 percent) that were not different (α = 0.05) from areas with no 

fire history. Images with postfire recovery interval less than 7 yr 

(18 percent of all images) exhibited less bitterbrush cover (0.49 

percent) and density (263 plant ha–1). These results are consistent 

with other studies of postfire bitterbrush recovery in eastern Idaho 

and indicate that analysis of VLSA imagery is an effective method 

for evaluating the impact of fire history on bitterbrush recovery.

Introduction_______________________

Natural resource assessments have been limited to 

detailed measurements on small areas (usually by people 

working on the ground) or to inexact measurements from 

space or high-altitude aerial photography. Ground-based 

measurements are expensive and usually impractical for 

landscape-scale assessments. Conventional aerial and space 

imagery are suitable for obtaining a large-area perspective, 

but inadequate for accurate measures of plant and soil parameters supporting land-management decisions. However, 

innovative technology is opening new avenues of resource 

Very Large Scale Aerial (VLSA) Imagery for 

Assessing Postfire Bitterbrush Recovery

Corey A. Moffet, J. Bret Taylor, and D. Terrance Booth

assessment. The recently developed very large scale aerial 

(VLSA) imagery has high resolution, a narrow field of view, 

and is generally acquired as intermittent samples across the 

landscape, permitting detailed measurements on a landscape 

scale (Booth and Cox 2006; Booth and others 2006). VLSA 

images are acquired using a high-resolution (>10 MP) digital 

camera mounted on a light, piloted aircraft. Typically, VLSA 

images are acquired from altitudes between 100 and 300 m 

above ground level (AGL) and the camera lens focal length 

ranges between 100 and 840 mm. The pixel ground sample 

distance (GSD), a measure of resolution, and the image onground footprint depend on the combination of AGL altitude, 

lens focal length, and camera imaging chip characteristics 

(Booth and Cox 2006; Booth and others 2006).

Antelope bitterbrush (Purshia tridentata Pursh DC) is easily distinguished from other plant species of the sagebrush 

steppe in 1 mm GSD VLSA imagery. Bitterbrush is a widely 

distributed, long-lived (90 to 162 yr) western shrub that is 

important for wildlife and domestic livestock particularly on 

winter ranges (Ganskopp and others 1999, 2004; Kituku and 

others 1992; Lesperance and others 1970; Nord 1965). Annual production increases with plant age until a maximum 

is reached between 60 and 70 years (McConnell and Smith 

1977). In native stands, bitterbrush plants may require 10 

or more years to produce seed (Nord 1965).

The response of bitterbrush to fire varies among populations. Bitterbrush populations in portions of California, 

Oregon, Montana, and much of the Great Basin, rarely resprout and are nearly eradicated by fire (Ayers and others 

1999; Clark and others 1982; Clements and Young 2002; 

Nord 1965). In eastern Idaho, parts of Oregon, and some 

areas in California; however, populations of bitterbrush 

re-sprouted frequently after fire (Blaisdell and Mueggler 

1956; Busse and others 2000; Nord 1965). For those populations that have high bitterbrush re-sprouting potential, 

postfire response will depend on a number of other factors. 

Re-sprouting may occur as soon as 3 weeks after fire or may 

be delayed as much as 13 months (Blaisdell and Mueggler 

1956). Furthermore, re-sprout mortality may be delayed 

after fire. In central Oregon, 39 percent of bitterbrush plants 

re-sprouted after fire, but by the fifth growing season only 

25 percent of burned re-sprouts remained alive (Busse and 

others 2000). Similar mortality of re-sprouts was measured 

in eastern Idaho (Blaisdell and Mueggler 1956). Greater resprouting was observed following a 1953 fire in plots that 

also burned in 1939 compared to plots with no evidence of 

prior burning (Blaisdell and Mueggler 1956). Greatest resprouting success in central Oregon was found for 20-yr old 

and younger plants (Busse and others 2000).
162 USDA Forest Service Proceedings RMRS-P-52. 2008

Moffet, Taylor, and Booth Very Large Scale Aerial (VLSA) Imagery for Assessing Postfire Bitterbrush Recovery

Sites in the northern portion of the USDA-ARS, U.S. Sheep 

Experiment Station (USSES), Dubois, Idaho, headquarters 

property have potential for a significant amount of bitterbrush 

in the plant composition. A long history of fire research at 

the USSES coupled with good geographic records of fire 

history provided good opportunities to study bitterbrush 

recovery after fire at the landscape scale. The objective of 

this study was to determine whether VLSA imagery could be 

used successfully to assess differences in bitterbrush cover 

and density due to past fire history at the landscape scale 

and whether these results were consistent with published 

postfire bitterbrush response at the USSES.

Materials and Methods______________

Site Description

This study was conducted on a 4,400 ha site at the USSES 

headquarters property near Dubois, Idaho (44° 17’ N, 112° 

7’ W, 1800 elevation m). From 1971 to 2000, median annual 

precipitation at the nearest climate station (44° 14’ 37” 

N, 112° 12’ 2” W, 1661 m elevation) was 326 mm (NOAA 

2003). Interannual variation in precipitation is large, 

one tenth of years had less than 234 mm and one tenth 

had greater than 437 mm (NOAA 2003). Mean January 

temperature was –7 °C, and mean July temperature was 

20 °C during this same 30-yr interval (NOAA 2003). The 

study site, northeast of the climate station, typically receives greater annual precipitation. Estimated median 

annual precipitation at the study area for 1971 to 2000 

is 400 mm (PRISM Group 2006).

The study site was chosen to correspond with 4 soil mapping units (the map unit symbols LAC, LMC, LME, and 

LVC) at the USSES (fig. 1). These soil mapping units were 

selected because their component soils were associated with 

high bitterbrush potential ecological sites (USDA-NRCS 

1995, unpublished report). For this study, high bitterbrush 

potential sites are those sites allowing greater than 5 percent annual bitterbrush production in the historical climax 

plant community. The four soil mapping units were each 

complexes of two or more of the following tentative soil series: 

Akbash (fine-loamy, mixed, superactive, frigid Calcic Pachic 

Argixerolls), Maremma (fine-loamy, mixed, superactive, 

Figure 1—Soils map of the U.S. Sheep Experiment Station headquarters property near Dubois, Idaho. Map background 

is Quickbird imagery of October, 2002. High bitterbrush potential soils are LAC, LMC, LME, and LVC.
USDA Forest Service Proceedings RMRS-P-52. 2008 163

Very Large Scale Aerial (VLSA) Imagery for Assessing Postfire Bitterbrush Recovery Moffet, Taylor, and Booth

frigid Calcic Pachic Haploxerolls), and Pyrenees (loamyskeletal, mixed, superactive, frigid Typic Calcixerolls). 

These soil map units only occur in the northern portion of 

the property (fig. 1).

Fire History

Researchers have been conducting rangeland fire research 

at the USSES since 1936 and fire history on the headquarters 

property is known. Fire polygons digitized in a geographic 

information system (GIS) prior to 1995 were derived from 

hand-drawn maps. From 1995 to the present, fire polygons 

were digitized using GPS or satellite imagery. The 1995 and 

newer fire polygons exclude unburned islands within fire 

boundaries. In 2005, when images were acquired, about 22 

percent of the study site had no history of fire since 1936. 

The remainder of the study site had burned at least once and 

few small areas had burned as much as five times since 1936 

(fig. 2). Recovery interval since last fire was well distributed 

(table 1 and fig. 3). There was, however, one large gap in 

recovery interval—between 32 and 57 years. Most areas 

burned between 1948 and 1973 have burned subsequently. 

Additionally, a disproportionate number of samples were 

from the 24-yr recovery interval, which corresponds to a 1981 

wildfire (table 1 and fig. 3). About 30 percent of the study 

site was last burned in a 1981 wildfire (fig. 3). In 2005, the 

shortest recovery interval on the study site was 2 yr.

Imagery

VLSA images were acquired July 12, 2005 with a Canon 

EOS 1Ds (Canon USA, Inc., Lake Success, NY) 11MP digital 

SLR camera mounted in a light, piloted, fixed-wing airplane 

flown at 27 m s–1 average ground speed 100 m AGL. The 

camera was equipped with a Canon 600 mm, F/2.8 EF lens 

plus a 1.4x teleconverter to yield the equivalent of an 840 

mm, F/4.0 lens. Images acquired with this configuration 

were 4064 x 2704 pixels and covered approximately 14 m2

on the ground (4.54 x 3.02m) with a mean resolution of 1.118 

mm GSD. The images were acquired as part of a USSES 

rangeland survey. A total of 703 VLSA images were collected 

in accordance with the flight plan (400 x 400 m point grid 

overlay of the property) and of those, 253 were within the 

study site. An aerial survey system (Track’ Air, Hengelo, The 

Figure 2—Map of the number of fires that have occurred since 1936 in the northern portion of the U.S. Sheep Experiment 

Station headquarters property.
164 USDA Forest Service Proceedings RMRS-P-52. 2008

Moffet, Taylor, and Booth Very Large Scale Aerial (VLSA) Imagery for Assessing Postfire Bitterbrush Recovery

Figure 3—Map of fire recovery interval classes as defined in table 1 for the high bitterbrush potential sites at the 

U.S. Sheep Experiment Station.

Table 1—Image counts for each fire recovery interval and the groupings of images 

into recovery interval classes.

Recovery interval 

class

Images in 

class

Year of 

last fire

Recovery 

interval (yr)

Images in 

interval

2-3 17 2003 2 9

2002 3 8

6-7 28 1999 6 14

1998 7 14

10-15 18

1995 10 15

1993 12 1

1990 15 2

24 77 1981 24 77

26-31 32

1979 26 4

1977 28 14

1974 31 14

58-68 25

1947 58 7

1942 63 1

1939 66 4

1938 67 5

1937 68 8

69+ 56 –a 69+ 56

aYear of last fire is uncertain because fire history records prior to 1936 are not available.
USDA Forest Service Proceedings RMRS-P-52. 2008 165

Very Large Scale Aerial (VLSA) Imagery for Assessing Postfire Bitterbrush Recovery Moffet, Taylor, and Booth

Netherlands) was used to automatically trigger the camera 

as the pilot maneuvered the airplane over each point in the 

flight plan grid. A laser range finder used as an altimeter 

recorded the airplane’s altitude AGL at high frequency. A 

geographic positioning system (GPS) receiver recorded the 

coordinates of the airplane. Time stamps logged from the 

altimeter and GPS were matched with the trigger time 

logged by the camera to determine the altitude and location 

of the airplane at the time the image was captured (Booth 

and Cox 2006; Booth and others 2006).

The soil map unit and fire history data were associated 

with each image by intersecting the image coordinates 

with soil and fire history layers in a GIS (fig. 4). Images 

with coordinates intersecting the four soil map units that 

define the study site were analyzed for bitterbrush cover 

and density.

Images were converted from a RAW file format to a TIF file 

format so that it could be loaded into image analysis software. 

In this study, we used the ImageJ software for image analysis 

(Rasband 2006). Images were analyzed by one technician experienced in bitterbrush identification. Aside from hints due 

to obvious difference in total shrub cover between images of 

recent and older burns, the technician did not know the fire 

history of an image at the time the images were analyzed. The 

polygon selection tool was used to delineate the drip lines of 

each bitterbrush shrub in each image (fig. 5). The polygons 

selected with this tool are known in ImageJ as regions of 

interest (ROI). The ROI manager tool was used to save each 

bitterbrush plant ROI and measure the size (pixels in the 

ROI) of each bitterbrush ROI. The pixel GSD, calculated from 

the laser altimeter measurement for the image, was used to 

convert ROI and total image pixel area to on the ground area 

(m2). Unique ROIs were counted as single bitterbrush plants 

for density calculations.

Analysis

The images were stratified into 7 recovery interval classes 

according to when the image point was most recently burned. 

Since there are no fire records at the USSES for the period 

prior to 1936, any image point not intersecting a fire polygon 

was grouped into the 69+ yr recovery interval class. Other 

interval classes were designed to include between 15 and 

30 images, where possible (table 1).

Figure 4—Map of the location and fire history classification for all VLSA images acquired on high bitterbrush potential 

sites at the U.S Sheep Experiment Station.
166 USDA Forest Service Proceedings RMRS-P-52. 2008

Moffet, Taylor, and Booth Very Large Scale Aerial (VLSA) Imagery for Assessing Postfire Bitterbrush Recovery

Figure 5—Example of a VLSA image with bitterbrush canopies delineated. Inset is a 3x zoom in to one of the bitterbrush plants. The 

labels are the plant symbol (USDA, NRCS 2006) and unique plant identification number for each plant delineated in the image. 

One-way analysis of variance (ANOVA) was used to test 

whether cover and density were different among fire recovery 

interval classes. Homogeneity of error variance was tested 

using Levene’s test (Levene 1960). Normality of the errors was 

tested with the Shapiro-Wilk test (Shapiro and Wilk 1965). 

When the Shapiro-Wilk test was significant, the data were 

log transformed and reanalyzed. All results reported are 

back transformed to the original units. When Levene’s test 

was significant for an ANOVA, Welch’s modified F-test was 

used to test recovery interval effects (Welch 1951). Further, 

when Levene’s test was significant and Welch’s F-test was 

significant, a modified t-test using Satterthwaite’s denominator degrees of freedom was used to separate means 

(Satterthwaite 1946). If the homogeneous error variance 

assumption was satisfied and the conventional F-test was 

significant, then we proceeded to separate means by least 

significant difference (LSD). Due to unequal replication 

among recovery interval classes, LSDs were computed using 

the pooled mean square error as the estimate of variance 

and standard errors were computed specific to each contrast 

(Damon and Harvey 1987).

Results and Discussion______________

Back-transformed mean bitterbrush density was not 

different among 10-yr and longer recovery interval classes 

(table 2). Density in 7-yr and shorter recovery interval class 

images was less than in 58-yr and longer recovery interval 

class images (table 2). Blaisdell and Mueggler (1956) reported 

a 50 percent mean bitterbrush sprouting rate after fire with a 

33 percent subsequent mortality rate. Using 965 plants ha–1

(mean 58- to 68-yr recovery interval class density) as a base 

density we expect 280 plants ha–1 density in the 2- to 3-yr 

class from re-sprouting alone. This is near the measured 

densities in the 7-yr and shorter recovery interval classes. 

It is likely that seedlings established in these intervals. Bitterbrush seedling recruitment is greatest in years soon after 

fire. Up to 80 percent of bitterbrush seedlings established 

in the first 5 years after fire (Nord 1965). Some seedlings 

may have escaped detection in the short recovery interval 

class images. The minimum size bitterbrush plant measured 

from the VLSA images was 0.06 m2 and 75 percent of plants 

were greater than 0.22 m2. Density may be over estimated 
USDA Forest Service Proceedings RMRS-P-52. 2008 167

Very Large Scale Aerial (VLSA) Imagery for Assessing Postfire Bitterbrush Recovery Moffet, Taylor, and Booth

Table 2—Mean bitterbrush canopy cover and density for seven classes of postfire recovery interval 

at the U.S. Sheep Experiment Station, Dubois, Idaho, measured from VLSA imagery.

Recovery interval Bitterbrush canopy Bitterbrush Mean canopy

class (yr) N cover (percent) density (#/ha) size (m2)

2–3 17 0.30 c 262 bc 0.142 b

6–7 28 0.61 bc 264 b 0.355 ab

10–15 18 1.33 ab 460 ab 0.549 a

24 77 0.94 ab 534 ac 0.185 b

26–31 32 1.19 ab 493 ab 0.192 b

58–68 25 2.06 a 965 a 0.337 a

	 69+	 56	 1.71 a	 875 a	 0.313 a

in all VLSA images (perhaps more so for the longer recovery 

interval classes) due to uncertainty about where a plant was 

rooted when the whole canopy was not contained in the image. For density calculations, all bitterbrush canopies were 

counted as a plant inside the image area. The branching 

pattern found in older bitterbrush plants often makes it difficult to determine whether adjacent canopies are part of the 

same plant or separate. The mean size of a canopy increased 

with increasing recovery interval for the first 15 years of 

recovery, but then decreased in the 24 to 31 yr recovery 

interval. Ganskopp and others (1999) reported increased 

bitterbrush canopy diameter with increasing recovery time 

throughout 6 years of recovery. Nord (1965) reported that 

stem layering is common in bitterbrush, but that it should 

not be expected on plants less than 12 yrs old. The young 

bitterbrush branches are erect and have little opportunity to 

layer, but as the plants get older and larger, the likelihood 

of a branch reaching the ground and rooting is increased. It 

may be that what we detected with the dip in canopy size of 

bitterbrush plants in the 24-yr to 31-yr recovery interval was 

due to a change in their branching habit, from more erect 

to more decumbent layered branches. In this analysis, each 

discontinuous region of bitterbrush canopy was counted as 

a unique plant.

Canopy cover is not subject to the edge effect errors that 

the measure of density has with this method. Back transformed mean bitterbrush canopy cover was low regardless 

of recovery interval class (table 2). Bitterbrush cover was 

less in 7-yr and shorter recovery interval class images than 

58-yr and longer recovery interval class images (table 2). 

Sampling cover suffers from problems of appropriate scale 

due to the aggregation of cover within the canopy and within 

shrub clusters. The distribution of the data is effected by the 

selection of plot size. If plot size is too small, then plots may 

have no bitterbrush to measure in the plot when there is bitterbrush in the community. Regardless of recovery interval 

class, greater than 48 percent of images had no bitterbrush 

plants. This could be remedied by acquiring images with a 

larger footprint. Increased footprint reduces GSD resolution. 

During image analysis the technician typically viewed the 

images at 33 percent magnification (9 pixels rendered as 1) 

with good differentiation between bitterbrush and other species. By acquiring images with a larger footprint and lower 

GSD resolution, we can reduce variability and mitigate some 

of the error distribution problems.

Conclusions _______________________

The very large scale aerial technology has been demonstrated as an effective tool to monitor important rangeland 

indices such as bare ground and total cover. In this work, we 

demonstrate the utility of this technology for measuring the 

response of antelope bitterbrush to fire recovery interval at 

the landscape scale. The measurement of increases in cover 

over the first 7 yr after fire was consistent with eastern 

Idaho postfire bitterbrush recovery literature. Difficulty in 

determining whether adjacent canopies belong to the same 

plant makes it difficult to accurately measure plant density 

from VLSA images. Analysis of species with a more compact 

canopy may not have this issue. The small size of the image 

footprint used in this study resulted in large experimental 

errors and significantly non-normal data that required 

transformation. These data characteristics are likely to be 

improved with larger footprint images. Larger footprint 

images, given the same acquisition equipment, will result 

in greater GSD and, if the images become very large, could 

make identification of the species of interest difficult. Images 

with up to three times the GSD should achieve improvement 

in the data characteristics without jeopardizing the ability 

to differentiate bitterbrush from other plant species.

References ________________________

Ayers, D.M.; Bedunah, D.J.; Harrington, M.G. 1999. Antelope bitterbrush and Scouler’s willow response to a shelterwood harvest 

and prescribed burn in western Montana. Western Journal of 

Applied Forestry. 14(3): 137-143.

Blaisdell, J.P.; Mueggler, W.F. 1956. Sprouting of bitterbrush 

(Purshia tridentata) following burning or top removal. Ecology. 

37: 365-370.

Booth, D.T.; Cox, S. E. 2006. Very large scale aerial photography for 

rangeland monitoring. Geocarto International. 21(3): 27-34.

Booth, D.T.; Cox, S.E.; Berryman, R.D. 2006. Precision measurements from very-large scale aerial digital imagery. Environmental 

Monitoring and Assessment. 112: 293-307.

Busse, M.D.; Simon, S.A.; Riegel, G.M. 2000. Tree-growth and understory responses to low-severity prescribed burning in thinned 

Pinus ponderosa forests of central Oregon. Forest Science. 46(2): 

258-268.

Clark, R.G.; Britton, C.M.; Sneva, F.A. 1982. Mortality of bitterbrush after burning and clipping in eastern Oregon. Journal 

of Range Management. 35: 711-714.

Clements, C.D.; Young, J.A. 2002. Restoring antelope bitterbrush: 

management guidelines for overcoming the challenges of establishing antelope bitterbrush after a wildfire. Rangelands. 24: 3-6.
168 USDA Forest Service Proceedings RMRS-P-52. 2008

Moffet, Taylor, and Booth Very Large Scale Aerial (VLSA) Imagery for Assessing Postfire Bitterbrush Recovery

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

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and regression. New York: Harper & Row.

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cattle grazing among young bitterbrush stimulate shrub growth? 

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1999. Seasonal cattle management in 3 to 5 year old bitterbrush 

stands. Journal of Range Management. 52: 166-173.

Kituku, V.M.; Powell, J.; Smith, M.A.; Olson, R.A. 1992. Increasing 

bitterbrush nutrient quality with 2,4-D, mowing, and burning 

in southcentral Wyoming. Journal of Range Management. 45: 

488-493.

Lesperance, A.L.; Tueller, T.T.; Bohman, V.R. 1970. Symposium 

on pasture methods for maximum production in beef cattle: 

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I., ed. Contributions to probability and statistics. Palo Alton, CA: 

Stanford University Press: 278 -292.

McConnell, B.R.; Smith, J.G. 1977. Influence of grazing on age 

yield interactions in bitterbrush. Journal of Range Management. 

30: 91-93.

The use of trade, firm, or corporation names in this publication is 

for the information and convenience of the reader. Such use does not 

constitute an official endorsement or approval by the United States 

Department of Agriculture or the Agricultural Research Service of any 

product or service to the exclusion of others that may be suitable.

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2003. Climatography of the United States NO. 84, 1971-2000. 

[Online]. Available: http://www.ncdc.noaa.gov/DLYNRMS/

dnrm?coopid=102707 [October 28, 2006].

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Monographs. 35: 307-334.

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org [October 28, 2006].

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nih.gov/ij/ [October 28, 2006].

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normality (complete samples). Biometrika. 52: 591-611.

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Service (USDA NRCS). 2006. The PLANTS Database. [Online]. 

Available: http://plants.usda.gov [October 28, 2006]. Baton Rouge, 

LA: National Plant Data Center.

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USDA Forest Service Proceedings RMRS-P-52. 2008 169

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Ann P. Bollinger is a Graduate Assistant, Department of Animal 

Biotechnology, College of Agriculture, Biotechnology, and Natural 

Resources, Mailstop 202, University of Nevada, Reno, NV 89557; 

email: annboll@yahoo.com.

Barry L. Perryman is an Assistant Professor, Department of 

Animal Biotechnology, College of Agriculture, Biotechnology, and 

Natural Resources, Mailstop 202, University of Nevada, Reno, NV 

89557; email: bperryman@cabnr.unr.edu.

Abstract: In recent decades Northern Nevada has experienced a 

dramatic increase in cheatgrass (Bromus tectorum). As a result, 

wildfire frequency and size has increased. The objective of this 

project was to examine natural vegetation recovery and trend following 1985 wildfires. Density and cover measurements determined 

a high but fluctuating occurrence of cheatgrass and minimal change 

of perennial grass and shrubs. Loamy 8-10 ecological sites have a 

greater potential for consistently producing more cheatgrass plants 

and cover than loamy 5-8 and droughty loam sites.

Introduction_______________________

The Great Basin is experiencing accelerated rates of ecological change. Across large expanses of sagebrush-grass 

rangelands, the introduction and proliferation of cheatgrass 

(Bromus tectorum) has changed vegetative composition and 

plant community dynamics. Since the discovery of cheatgrass 

in the Intermountain region early in the 1900s, estimates of 

cheatgrass dominance on 19 million acres of big sagebrush 

communities in Nevada have increased from 1 percent in 

the 1960s to 25 percent in the 1970s and even higher today 

(Young and Blank 1995). In addition, cheatgrass has expanded and adapted to drier conditions found in salt desert 

shrub plant communities.

Though drought and poor livestock management have 

influenced vegetative composition, cheatgrass has possibly 

caused the most dramatic changes on Great Basin landscapes 

in recent decades. With cheatgrass comes a change in the 

fire cycle (D’Antonio and Vitousek 1992). Once established, 

biomass production and high seed viability provide a fuel 

20 Years of Natural Recovery After 

Wildfire on Northern Nevada Rangelands

Ann P. Bollinger and Barry L. Perryman

load with greater continuity than native plant communities. Wildfires occur more often and burn more acres at 

any one time.

Several studies throughout the Great Basin describe 

vegetation history and dynamics. Robertson (1971) examined vegetation cover within and outside an exclosure on 

sagebrush-grass range in Humboldt County, Nevada. The 

Singas Creek exclosure was established as a site for range 

improvement studies in 1939, and eliminated further livestock grazing. Robertson, comparing vegetation data from 

1940 and 1970, observed an increase in total vegetation cover 

including a 38 percent increase in cheatgrass. Wright and 

Klemmedson (1965) conducted an experiment to examine 

the effect of fire on four species of bunchgrasses: Sandberg 

bluegrass (Poa secunda), squirreltail (Elymus elymoides), 

needle-and-thread (Hesperostipa comata), and Thurber 

needlegrass (Stipa thurberiana). Sandberg bluegrass was 

least affected by burning, followed by squirreltail, and 

then both Stipa species. The season of burn, predicated by 

air temperature and humidity, was the greatest factor in 

determining damage or death to these grass species.

Today, rangeland science recognizes and is attempting 

to further define the concept of state and transition models 

(Westoby and others 1989) and thresholds (Laycock 1991). 

In the state and transition model, vegetation dynamics 

in sagebrush-grass communities recognizes that a site is 

first dominated by (perennial) grasses and later by shrubs. 

Harniss and Murray (1973) found that after a prescribed 

fire near Dubois, Idaho, total shrub cover, and particularly 

sagebrush, did not begin to increase until 12 years postfire. 

Young and Evans (1978) found that succession in sagebrush 

grasslands is now determined by cheatgrass. Given sufficient 

litter (Young and others 1976), cheatgrass seeds germinate 

in the first year following wildfire, outcompeting native annual and perennial species. Sagebrush is unable to compete 

with other native shrub species such as green rabbitbrush 

(Chrysothamnus viscidiflorus) and horsebrush (Tetradymia

canescens) that resprout from the roots and crown. Due to the 

new, frequent wildfire cycle created by cheatgrass, formally 

abundant big sagebrush/Thurber needlegrass communities 

of northern Nevada have declined.

In 1985, wildfires burned approximately 600,000 acres 

around Winnemucca, Nevada (Corliss 1992). Wildfires occurred in subsequent years, but Nevada experienced a dramatic fire year in 1999 with 1.8 million acres burned across 

the state. The BLM Winnemucca Field Office experienced 

640,000 acres alone (Western Great Basin Coordination 

Center 1999), with fires occurring on some of the same areas 

burned in 1985.
170 USDA Forest Service Proceedings RMRS-P-52. 2008

Bollinger and Perryman 20 Years of Natural Recovery After Wildfire on Northern Nevada Rangelands

In this semi-arid region, vegetation often requires long 

periods of time to recover. We expect that frequent fires have 

changed vegetation composition to cheatgrass dominance and 

a threshold has been crossed. The overall goal of this study 

was to determine ecological trend based on perennial grass 

and shrub composition after wildfire on public rangelands 

of north-central Nevada. Specifically, the project analyzed 

vegetation response within and among ecological sites as 

well as change through time on areas of natural recovery. 

This report summarizes post-burn central tendencies for 

density and cover of functional plant groups from 1986 to 

2005. Sites included in this report burned in 1985 and did 

not re-burn during the period of record.

Materials and Methods______________

All monitoring sites were located on land managed 

by the Bureau of Land Management (BLM) in Humboldt 

County, Nevada. They were within a 30-mile radius of 

Orovada, Nevada, but all were located west of the Santa Rosa 

Range of the Humboldt-Toiyabe National Forest. The area 

has a semi-arid climate characterized by long, cold winters 

and warm summers. Annual average minimum and maximum temperatures are 34 °F (1 °C) and 64 °F (18 °C), respectively. Annual average precipitation for the three ecological 

sites is: loamy 5-8, 8.3 inches (211 mm); and 10.6 inches 

(269 mm) for both the loamy 8-10 and droughty loam, with 

the majority falling in winter and spring (Western Regional 

Climate Center, 2005). Historically, vegetation types were 

shadscale and big sagebrush-grass, but many areas are 

now dominated with annual forbs and cheatgrass. Elevation on these ecological sites varies from 4,000 ft to 6,500 ft 

(1,220 m to 1,980 m). Aspect was varying directions. Slope 

was generally less than 5 percent; however, two transects 

were greater than 15 percent.

Permanent transects were established in 1986, the first 

year following the 1985 wildfire, by the BLM in an area 

representative of fire impact and vegetation composition. 

Each transect was monitored for the first 2 years, occasionally thereafter, and during the summers of 2004 and 2005. 

Following the 1999 wildfires, monitoring continued at the 

permanent transects established in 1986. A long-time BLM 

employee assisted with monitoring seven of the nine transects 

in the past and all transects in 2004 and 2005. This fortunate 

opportunity added consistency to data collection and reduced 

error in observer bias.

For this study, data from sites with natural recovery were 

examined. Three ecological sites were identified based on 

information in NRCS technical guides and included: loamy 

5-8, loamy 8-10, and droughty loam 8-10. Sampling methods 

utilized in 2004 and 2005 remained consistent with those of 

previous years. Density was measured with a 9.6 ft2 round 

quadrat (0.89 m2) for herbaceous plants and 96 ft2 round 

quadrat (8.9 m2) for shrubs. Transects were 200 ft (60 m) 

in length and measurements taken at 20 ft (6 m) intervals. 

Canopy cover for both herbaceous and shrub species was 

estimated with Daubenmire quadrats (Daubenmire 1959). 

Transects were 1,000 ft (300 m) in length and measurement 

taken at 100 ft (30 m) intervals. Central tendencies of density and cover were determined for each sample period and 

plotted by ecological site. Functional groups included annual 

grass, perennial grass, and shrubs. We also include a brief 

discussion of annual precipitation and potential effects on 

density and cover.

Results and Discussion______________

Annual Grass

Cheatgrass density and cover on loamy 5-8 and droughty 

loam sites fluctuated throughout the period of record; 

whereas on loamy 8-10 sites, cheatgrass density and cover 

exhibited an increasing trend (fig. 1). Cheatgrass cover on 

loamy 5-8 and droughty loam sites closely followed annual 

precipitation patterns (fig. 2); whereas on loamy 8-10 sites, 

cheatgrass cover increased to nearly 20 percent despite a 

precipitation decrease in 1992. This suggests that the loamy 

5-8 and droughty loam sites are more dynamic with respect 

to these response variables, while loamy 8-10 sites provided 

conditions that allowed cheatgrass density and cover to 

steadily increase.

Perennial Grass and Shrubs

Perennial grass and shrubs exhibited few changes 

throughout the period of record. Density and cover trends 

are displayed in figure 3. On loamy 5-8 sites, perennial 

grass never exceeded two plants m–2 or 1 percent cover. 

Shrub density fluctuated, whereas cover was low for the first 

3 years followed by a substantial increase (0 to 7 percent) 

by 1995, perhaps related to an above average precipitation 

year in 1994/1995 of 10.7 inches (272 mm). The dominant 

shrub was shadscale (Atriplex confertifolia). As Sharp (1990) 

demonstrates, shadscale communities fluctuate with time 

(for example, precipitation and insect outbreaks).

Loamy 8-10 sites experienced minimal change in perennial 

grass and shrub density and cover throughout the period 

of record. The decrease in precipitation from 11.9 inches 

(302 mm) in 1985/1986 to 7.5 inches (190 mm) in 1986/1987 

but slight increase in cover of both perennial grasses and 

shrubs suggest that short-term change in precipitation did not 

immediately affect vegetative cover. Between 1987 and 1992, 

precipitation was variable; however, 5.8 inches (147 mm) 

in 1991/1992 was considerably below-average year and the 

lowest during the 20 years of this study. Between 1992 and 

2005, precipitation was also variable with 4 years of aboveaverage precipitation including the year of data collection in 

2004/2005. However, the previous 6 years from 1998/1999 to 

2003/2004 were below-average. The increase in shrub cover 

from 0 percent in 1992 to 6.0 percent in 2005 may reflect a 

steady but undocumented increase throughout the 14 years 

as well as the sudden increase in annual precipitation.

On droughty loam sites, a long-term trend of perennial 

grass and shrub composition was static. Perennial grass 

density ranged from less than one plant m–2 to 6.9 and 

cover ranged from less than 1 percent to 7.4 percent, with 

high density and cover percentages correlating to aboveaverage precipitation years. Shrub density remained below 

3.2 plants 10 m–2 throughout the period of record. Cover 

did not exceed 3 percent until 2005.
USDA Forest Service Proceedings RMRS-P-52. 2008 171

20 Years of Natural Recovery After Wildfire on Northern Nevada Rangelands Bollinger and Perryman 

Figure 1—Cheatgrass density and cover by ecological site across the period of record, 1986 to 2005, in northern Nevada.

Figure 2—Precipitation data from the Western Regional Climate 

Center for the Orovada 4 WSW (solid line) and Winnemucca 

WSO Airport (dotted line) weather stations. Horizontal lines 

show water year precipitation (October to September) and 

long-term mean. The Orovada mean includes data from July 

1948 through December 2005, and the Winnemucca mean 

includes data from January 1897 through December 2005. 

Orovada weather data represents precipitation for loamy 8-10 

and droughty loam 8-10 sites. Winnemucca weather data 

represents precipitation for loamy 5-8 sites.
172 USDA Forest Service Proceedings RMRS-P-52. 2008

Bollinger and Perryman 20 Years of Natural Recovery After Wildfire on Northern Nevada Rangelands

Figure 3—Perennial grass and shrub density and cover by ecological site across the period of record, 1986 to 2005, in northern 

Nevada. Number of plants equals per m–2 for perennial grass and per 10 m–2 for shrubs.
USDA Forest Service Proceedings RMRS-P-52. 2008 173

20 Years of Natural Recovery After Wildfire on Northern Nevada Rangelands Bollinger and Perryman 

Conclusions _______________________

Vegetation communities are dynamic and changes are 

caused by any number and combination of biotic and abiotic 

factors. However, the introduction of cheatgrass has changed 

fire cycles and vegetative communities. By documenting 

long-term trend, we can evaluate changes on the landscape, 

and the new frequent fire cycle may provide us with preand post-burn data. In examining 20 years of postfire data, 

our results indicate that there is greater potential for high 

cheatgrass density and cover on loamy 8-10 sites whereas 

cheatgrass density and cover is more variable on loamy 5-8 

and droughty loam sites. While this study is most relevant 

to areas with similar soil and precipitation characteristics, 

results provide documented insight to natural recovery on 

the semi-arid rangelands of northern Nevada.

References ________________________

Corliss, J. 1992. Getting the jump on grass fires—range grass 

research. Agricultural Research. September.

D’Antonio, C.M.; Vitousek, P.M. 1992. Biological invasions by exotic 

grasses, the grass/fire cycle, and global change. Annual Review 

of Ecology and Systematics. 23: 63-87.

Daubenmire, R. 1959. A canopy coverage method of vegetational 

analysis. Northwest Science. 33: 43-66.

Harniss, R.O.; Murray, R.B. 1973. 30 years of vegetal change 

following burning of sagebrush-grass range. Journal of Range 

Management. 26: 322-325.

Laycock, W.A. 1991. Stable states and thresholds of range condition 

on North American rangelands: A viewpoint. Journal of Range 

Management. 44: 427-433.

Roberson, J.H. 1971. Changes on a sagebrush-grass range in 

Nevada ungrazed for 30 years. Journal of Range Management.

24: 397-400.

Sharp, L.A.; Sanders, K.; Rimbey N. 1990. Forty years of change in 

a shadscale stand in Idaho. Rangelands. 12: 313-328.

Western Great Basin Coordination Center. 1999. Season Overview. 

[Online]. 18 p. Available: http://gacc.nifc.gov/wgbc/predictive/

intelligence/paststatistics/EOY99.pdf. [October 2006].

Western Regional Climate Center. n.d. Historical climate information. [Online]. Available: http://www.wrcc.dri.edu/. [October 

2005]

Westoby, M.; Walker, B.; Noy-Meir, I. 1989. Opportunistic management for rangelands not at equilibrium. Journal of Range 

Management. 42: 266-274.

Wright, H.A.; Klemmedson, J.O. 1965. Effect of fire on bunchgrasses 

of the sagebrush-grass region in southern Idaho. Ecology. 46(5): 

680-688.

Young, J.A.; Blank, R.R. 1995. Cheatgrass and wildfires in the 

Intermountain West. In: Lovich, J.E.; Randall, J.; Kelly, M.D., 

eds. Proceedings of the California Exotic Pest Plant Council, 

Symposium ‘95: 6-8.

Young, J.A.; Evans, R.A. 1978. Population dynamics after wildfires 

in sagebrush grasslands. Journal of Range Management. 31: 

283-289.

Young, J.A.; Evans, R.A.; Tueller, P.T. 1976. Great Basin plant 

communities—pristine and grazed. In: Elston, R., ed. Holocene 

environmental changes in the Great Basin. Nevada Archeology 

Survey, Research Paper 6. Reno, NV: University of Nevada: 

187-215.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

USDA Forest Service Proceedings RMRS-P-52. 2008 175

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Benjamin M. Rau is an M.S. Graduate Research Assistant, 

Hydrologic Sciences Program, University of Nevada, Reno, NV.

Robert R. Blank is a Soil Scientist, USDA-ARS, Reno, NV; 

email: blank@unr.nevada.edu.

Tye Morgan is a Research Assistant, USDA-ARS, Reno, NV.

Abstract—Pinyon-juniper (Pinus monophylla Torr. & Frém.— 

Juniperus osteosperma Torr.) encroachment into sagebrush 

grasslands is a continuing problem in the Western United States. 

Prescribed burning has been suggested to slow woodland encroachment. We examined surface soil microbial community structure 

using Phospholipid Fatty Acid (PLFA) analyses to determine differences between burned and unburned woodlands at two microsites. 

Tree canopy interspace microsites had a greater total PLFA and 

percentage Eukaryotes. Conversely, under tree canopy microsites 

had a greater percentage of fermicutes, anaerobic metal reducers, 

and higher PLFA Cis/Trans fatty acid ratio. Time since burning 

increased Eukaryote PLFA in both microsites.

Introduction_______________________

Pinyon-juniper (Pinus monophylla Torr. & Frém.— 

Juniperus osteosperma Torr.) woodland encroachment into 

sagebrush-grasslands is a large scale ongoing problem in the 

Great Basin and other arid regions in the Western United 

States (Miller and Tausch 2001). Woodland encroachment 

can lead to exotic invasion by competition with perennial 

understory vegetation for resources, elimination of perennial 

understory seedbanks over time, and increasing the risk of 

catastrophic wildfire due to woody fuel accumulation (Tausch 

1999a,b). Also litter produced by pinyon pine and juniper 

contains high concentrations of monoterpenes and could be 

allelopathic to native perennial vegetation (Everett 1987; 

Wilt and others 1993). We hypothesize that pinyon litter 

and burning influences the microbial community possibly 

through the changes in carbon substrate, pH, monoterpene 

concentration, or other chemical parameters.

Methods __________________________

We collected soils from a Great Basin pinyon woodland 

in Nevada, U.S.A. (39°15’11” N, 117°35’83” W). Soils in 

this study are classified as coarse loamy mixed frigid Typic 

Microsite and Time Since Prescribed Fire’s 

Influence on Soil Microbiology in a Pinyon Woodland

Benjamin M. Rau, Robert R. Blank, and Tye Morgan

Haploxerolls derived from silica ash flow tuffs (Rau 2005). 

Surface (0 to 15 cm) soil was collected from under pinyon 

tree canopies after removing the O horizon (litter) and from 

between tree canopies where healthy perennial understory 

vegetation occurred. Soils were collected from three adjacent tree plots in May 2005. The first plot had not been 

burned in greater than 80 years, the second was prescription burned in May of 2002, and the third was prescription 

burned in May of 2004. Three sub-samples were collected 

from three individual tree canopies and canopy interspaces 

at all three burn treatments using a 10 cm diameter sample 

auger. Sub-samples were homogenized in a bucket in the 

field then placed on ice and returned to the lab (N = 3 burn 

treatments x 2 microsites x 3 replicates = 18). Samples were 

packed in ice and sent to Microbial insights® (Rockford, TN, 

U.S.A.) for PLFA analyses. Microbial Insights calculates 

microbial biomass as cells g –1 based on a conversion factor 

(20,000 cells mol–1 PLFA). Results of the PLFA analyses 

were reported as total cells g –1 soil, PLFA Cis/Trans fatty 

acid ratio, percent fermicutes, percent anaerobic and metal 

reducing bacteria, percent proteobacteria, percent actinomycetes, percent eukaryotes, and the fungus to bacteria 

ratio. A two way ANOVA was used to compare the results 

from the PLFA analyses with burn treatment and microsite 

considered as main effects, and group means were compared 

using Duncan’s test (SAS Institute 2004).

Results and Discussion______________

Results from the ANOVA are presented in table 1. Canopy 

interspaces had a greater total concentration of PLFA than 

soil from beneath tree canopies, and the PLFA beneath tree 

canopies had a higher Cis/Trans fatty acid ratio (fig. 1). The 

Cis/Trans fatty acid ratio is a measure of cell wall permeability and cell growth rate. A higher Cis/Trans fatty acid 

ratio indicates that the microbial community was possibly 

under more environmental stress and its cell membrane is less 

permeable and growth rate has slowed (Guckert and others 

1985). Microbial stress could be related to carbon substrate 

quality, low pH, and monoterpene content of pinyon litter 

and the soil beneath it (Guckert and others 1985; Pietikainen 

and others 2000; Rau 2005; Wilt and others 1993) Fermicutes 

and anaerobic reducers were a higher proportion of the microbial community under tree canopies (fig. 2). Soil beneath 

pinyon canopies is typically much higher in organic carbon 

than canopy interspace soil which may facilitate reducing 

conditions (McDaniel and Graham 1992). Anaerobic reducers 

were generally a higher portion of the community in areas 

burned in 2002 than in unburned areas (fig. 2). This seems 

counter intuitive since some organic carbon was inevitably 
176 USDA Forest Service Proceedings RMRS-P-52. 2008

Rau, Blank, and Morgan Microsite and Time Since Prescribed Fire’s Influence on Soil Microbiology in a Pinyon Woodland

Table 1—Result for the ANOVA for burn treatment and microsite differences.

DF # Cells/g Soil Cis/Trans FA Actinomycetes Proteobacteria

F P F P F P F P

Treatment (T) 2, 12 0.3 0.7488 0.14 0.8712 3.29 0.0725 3.49 0.0638

Vegetation type (V) 1, 12 28.97 0.0002 18.74 0.001 1.85 0.1986 0.1 0.7609

T x V 2, 12 0.95 0.4159 0.83 0.4614 0.85 0.4507 0.6 0.5619

 DF Fermicutes Reducing 

 bacteria Eukaryotes Fungi/Bacteria

F P F P F P F P

Treatment (T) 2, 12 3.87 0.0504 7.17 0.0089 52.39 <0.0001 51 <0.0001

Vegetation type (V) 1, 12 5.66 0.0348 26.91 0.0002 42.27 <0.0001 36.7 <0.0001

T x V 2, 12 0.06 0.9467 9.86 0.0029 5.53 0.0198 5.26 0.0229

Figure 1—Mean number of microbes per gram of soil, Cis/Trans fatty acid ratio of PLFA, 

and standard errors for burn treatment and microsite. Upper case letters indicate microsite 

differences. IS = canopy interspace, UT = Under tree canopy.

consumed by combustion. Likewise the results for treatment 

microsite interactions do not provide a logical pattern of 

anaerobic reducer distribution (fig. 2). Eukaryotes were a 

significantly greater portion of the microbial community in 

the canopy interspace microsites than under tree canopies 

(fig. 3). Eukaryotes increased their proportion steadily with 

time following fire in canopy interspaces, but immediate 

recovery was delayed under tree canopies (fig. 3). Eukaryote PLFA is indicative of fungi and may represent not only 

decomposers but also mycorrhizae, which are essential for 

many perennials in the Great Basin (Grayston and others 

2001; Wicklow 1994). Identical results were observed for the 

fungi to bacteria ratio because no differences in bacterial 

PLFA were observed during this study (fig. 3).

The results of our limited study contrast and confirm 

results observed by other researchers. In a study from 

pinyon-juniper woodlands in Arizona the authors suggest that 

bacterial populations are more diverse under tree canopies 

than inter-canopy, although relative biomass and fungi 

estimates were not compared (Dunbar and others 1999). 

Results from a lab study where soil and litter from pinyonjuniper woodlands was burned in microcosms showed soils 
USDA Forest Service Proceedings RMRS-P-52. 2008 177

Microsite and Time Since Prescribed Fire’s Influence on Soil Microbiology in a Pinyon Woodland Rau, Blank, and Morgan

Figure 3—Mean percent of Eukaryote PLFA, fungus to bacteria ratio, and standard errors for 

burn treatment and microsite. Upper case letters indicate microsite differences, lower case 

letters above each bar indicate burn treatment differences within microsite, and lowercase 

letters within bars indicate differences between all burn treatments and microsites. IS = canopy 

interspace, UT = Under tree canopy.

Figure 2—Mean percent of fermicute PLFA, anaerobic reducing bacteria PLFA, and standard 

errors for burn treatment and microsite. Upper case letters indicate microsite differences, 

lower case letters above each bar indicate burn treatment differences within microsite, and 

lowercase letters within bars indicate differences between all burn treatments and microsites. 

IS = canopy interspace, UT = Under tree canopy.
178 USDA Forest Service Proceedings RMRS-P-52. 2008

Rau, Blank, and Morgan Microsite and Time Since Prescribed Fire’s Influence on Soil Microbiology in a Pinyon Woodland

beneath burned pinyon litter supported less mycorrhizal 

colonization than unburned controls (Klopatek and others 

1988). A prescribed burn and wood ash treatment study in 

Scots pine (Pinus sylvestris L.) indicate that microbial biomass particularly fungi are reduced by burning and raising 

pH (Baath and others 1995). Prieto and others (1998) also 

observed that although microbial biomass decreases immediately after fire, it rebounds, but may take up to 4 years to 

reach pre-burn levels. Pietikainen and others (2000) found 

that heating forest floor material to 230 °C changed the 

carbon substrate and raised the pH in Scots pine litter and 

subsequently altered the microbial community.

The changes observed in the microbial community caused 

by pinyon encroachment and burning may have significant 

effects on perennial understory vegetation recovery potential. Most native perennial vegetation in the Great Basin is 

dependent on mycorrhizal associations, and although the 

relationships may not be species specific, each plant species 

responds differently to various mycorrhizae species (Bever 

2002; Wicklow 1994). It has also been observed that plants 

have positive feedback loops with certain mycorrhizae. 

A plant alters its soil environment after establishment 

through litter chemistry or root exudates so that particular 

mycorrhizae are favored; the preferred micorrhizae then 

optimally assists the plant with nutrient uptake (Bever 

2002). It is known that pinyon pines are one of the only 

species in arid woodlands to host ectomycorrhizal species 

(Haskins and Gehring 2004). Most other perennial shrubs, 

herbaceous, and tree species, including Juniperus spp., are 

arbuscular mycorrhizal hosts (Haskins and Gehring 2004). 

If pinyons have altered the microbial communities under 

their canopies (especially mycorrhizae) and fire has reduced 

microbial numbers, then it may be difficult for native perennial understory vegetation to recover following fire on sites 

previously occupied by pinyon trees. Klironomos (2002) 

showed that native vegetation grows significantly better on 

soil that was previously occupied by the same species than 

on soil previously occupied by a different species.

Conclusions

From our study we cannot say with certainty which species 

of fungus or mycorrhizae inhabit the under tree canopy and 

tree canopy interspace microsites. However, we know that 

these microsites differ in there microbiological properties. 

Changes in soil biota caused by woodland encroachment 

along with monoterpene production by pinyon, and the harsh 

physical environment produced after burning may inhibit 

native perennial understory vegetation recovery in pinyon 

canopy microsites following burning and lead to exotic invasion by cheatgrass (Bromus tectorum). Studies designed 

to identify the magnitude of impact each of these factors 

has on perennial vegetation establishment are needed to 

improve the likelihood of successful prescribed burning in 

pinyon woodlands.

References ________________________

Baath, E.; Frostegard, A.; Pennanen, T.; Fritze, H. 1995. Microbial 

community structure and pH response in relation to soil organic 

matter quality in wood-ash fertilized, clear-cut or burned coniferous forest soils. Soil Biology and Biochemistry. 27(2): 229-240.

Dunbarr, J.; Taklia, S.; Barns, S.M.; Davis, J.A.; Kuske, C.R. 1999. 

Levels of bacterial community diversity in four arid soils compared by cultivation and 16S rRNA gene cloning. Applied and 

Environmental Microbiology. 65(4): 1662-1669.

Everett, R.L. 1987. Allelopathic effects of pinyon and juniper litter on 

emergence and growth of herbaceous species. In: Frasier, G.W.; Evans, R.A., eds. Proceedings, Symposium on Seed and Seedbed Ecology 

of Rangeland Plants. Washington, DC: USDA-ARS: 62-67.

Grayston, S.J.; Griffith, G.S.; Mawdsley, J.L.; Campbell, R.D.; 

Bardgett, R.D. 2001. Accounting for variability in soil microbial 

communities of temperate upland grassland ecosystems. Soil 

Biology and Biochemistry. 33: 533-551.

Guckert, J.B.; Hood, M.A.; White, D.C. 1985. Phospholipid esterlinked fatty acid profiles from Desulfovibrio species. Journal of 

Lipid Research. 26: 982-988.

Haskins, K.E.; Ghering, C.A. 2004. Interactions with juniper alter pinyon 

pine ectomyccohizal fungal communities. Ecology. 85(10): 2687-2692.

Klironomos, J.N. 2002. Feedback with soil biota contributes to plant 

rarity and invasiveness in communities. Nature. 417: 67-70.

Klopatek, K.C.; DeBano L.F.; Klopatek, J.M. 1988. Effects of simulated fire on vesicular-arbuscular mycorrhizae in pinyon-juniper 

woodland soil. Plant and Soil. 109: 245-249.

McDanial, P.A.; Graham, R.C. 1992. Organic carbon distributions 

in shallow soils of pinyon-juniper woodlands. Soil Science Society 

of America Journal. 56: 499-504

Miller, R.F.; Tausch, R.J. 2001. The role of fire in juniper and pinyon 

woodlands: a descriptive analysis. In: Gallet, K.E.M.; Wilson, T.P., 

eds. Proceedings of the invasive species workshop: The role of 

fire in the control and spread of invasive species. Miscellaneous 

Publications No. 11. Tallahassee, FL: Tall Timbers Research 

Station: 15-30.

Pietikainen, J.; Hiukka, R.; Fritze, H. 2000. Does short-term heating 

of forest humus change its properties as a substrate for microbes? 

Soil Biology and Biochemistry. 32:2 77-288.

Prieto-Fernandez, A; Acea, M.J.; Carballes, T. 1998. Soil microbial 

and extractable C and N after wildfire. Biology and Fertility of 

Soils. 27: 132-142.

Rau, B.M. 2005. Soil physical, chemical and understory plant 

nutritional response to pinyon-juniper encroachment and prescribed fire in a central Nevada woodland. Reno, NV: University 

of Nevada. Thesis.

SAS Institute. 2004. SAS 9.1 user’s guide. Copyright 2004. Cary, 

NC: SAS Institute Inc.

Tausch, R. J. 1999a. Historic woodland development. In: Monsen, 

S.B.; Stevens, R., comps. Proceedings: Ecology and management 

of pinyon-juniper communities in the Interior West. Proceedings 

RMRS-P-9. Ogden, UT: USDA Forest Service, Rocky Mountain 

Research Station: 12-19.

Tausch, R. J. 1999b. Transitions and thresholds: influences and 

implications for management in pinyon and Utah juniper woodlands. In: Monsen, S.B.; Stevens, R. comps. Proceedings: Ecology 

and management of pinyon-juniper communities in the Interior 

West. Proceedings RMRS-P-9. Ogden, UT: USDA Forest Service 

Rocky Mountain Research Station: 61-65.

Wicklow, H.M. 1994. Mycorrhizal ecology of shrub-steppe habitat. 

In: Monsen, S.B.; Kitchen, S.G., comps. Proceedings: Ecology 

and management of annual rangelands. Gen. Tech. Rep. INTGTR-313. Ogden, UT: U.S. Department of Agriculture, Forest 

Service, Intermountain Research Station: 207-210.

Wilt, M.F.; Miller, G.C.; Everett, R.L.; Hackett,M. 1993. Monoterpene 

concentrations in fresh senescent, and decaying foliage of singleleaf pinyon (Pinus monophylla TORR. & FREM.: PINACEAE) 

from the western Great Basin. Journal of Chemical Ecology. 

19(2): 185-194.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein
USDA Forest Service Proceedings RMRS-P-52. 2008 179

Field Trips

USDA Forest Service Proceedings RMRS-P-52. 2008 181

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

The Great Basin____________________

The Great Basin is a vast, internally drained region of 

the Western United States, bounded by the Sierra Nevada 

and southern Cascade Mountain ranges to the west and the 

Wasatch Mountains and western rim of the Colorado Plateau 

to the east. Although less discrete, northern and southern 

boundaries are generally defined by the drainages of the 

Columbia and Colorado Rivers. Included are most of Nevada 

and major portions of Utah, California, and Oregon. Contrary 

to what the name implies, the Great Basin is composed of not 

one, but many closed basins or valleys separated by over 100 

relatively narrow mountain ranges arranged with a more or 

less north-south orientation. This geography is characteristic of the larger Basin and Range Province extending from 

southern Oregon and Idaho to Sonora, Mexico, of which the 

Great Basin occupies the northern and most elevated third. 

Valley floor elevations range from 86 m below sea-level in 

Death Valley, California, to approximately 2,000 m above 

sea-level in central Nevada. Mountain elevations also 

vary with the summits of 33 mountain ranges exceeding 

3,000 m. The mountains of the Great Basin began forming approximately 17 million years ago when a complex 

series of normal faults began forming in the earth’s crust 

in response to continental stretching (Morris and Stubben 

1994). Mountains ranges (horsts) and valleys (grabens) 

formed as blocks of crust were vertically displaced along 

those faults. As the mountains eroded, the surrounding 

valleys filled with their debris. The cores of many eastern 

Great Basin ranges are composed of limestone, dolomite, and 

quartzite; rocks derived from Paleozoic marine sediments. 

Younger Mesozoic and Cenozoic volcanics such as granite, 

tuff, basalt, and breccia are found throughout the Great 

Basin but dominate in the west and north. The physical and 

chemical characteristics of valley soils from massive sand 

dunes to fine-textured playas reveal the diversity of these 

parent materials.

The climate of the Great Basin is arid to semiarid with 

cold winters and warm summers except in the southern, 

low-elevation valleys where winters are mild and summers 

hot. Seasonality of precipitation varies along a geographic 

Great Basin Cold Desert Shrublands and 

the Desert Experimental Range

Stanley G. Kitchen and Stephanie L. Carlson

gradient with the importance of winter and spring Pacific 

frontal storms decreasing and summer monsoons increasing as one travels from north to south and from west to east 

(Peterson 1994). Precipitation increases and temperature 

decreases with increasing elevation resulting in fairly distinct vegetation zones. Higher elevation forest and alpine 

landscapes are relatively small and isolated; subsequently, 

these mountain ecosystems function as ecological islands 

separated by a sea of desert (Billings 1978; Wells 1983).

The distribution of plant species in the Great Basin has 

changed substantially over time with changes in climate. 

During the late Pleistocene (20,000 to 16,000 years b.p.), over 

100 lakes filled valley bottoms indicative of the much colder 

and wetter climate than today (Wells 1983; Peterson 1994). 

Plant macrofossils found in preserved woodrat (Neotoma

spp.) middens and pollen sediment records suggest that 

subalpine coniferous forests and sagebrush-dominated 

shrub-grass steppes covered foothill and valley locations 

surrounding these Pleistocene lakes (Wells 1983; Thompson 

1990). Tundra occupied upper elevations and northern 

latitudes. With Holocene warming, alpine communities 

and subalpine conifers retreated to high elevation refugia 

on the tallest mountains, and more thermophilic conifers 

migrated from the south and formed mid-elevation forests 

and woodlands. While sagebrush communities remained 

dominant on many low to mid-elevation landscapes, saltdesert shrubland-grassland mosaics now fill the driest valley 

locations especially on the fine-textured, salty soils that were 

once lake bed sediments.

Stop 1: Pigmy Rabbit Colony 

and Habitat _______________________

Pygmy rabbits (Brachylagus idahoensis) are the smallest known leporids, living or extinct (Oliver 2004) and are 

unique in the practice of digging their own burrows though 

they will use burrows dug by other animals. Pigmy rabbits 

are almost always associated with dense stands of tall big 

sagebrush (Artemisia tridentata), which they rely on for 

both food (especially during winter) and concealment from 

predators. The scampering or scurrying mode of travel used 

by pygmy rabbits differs from the bounding and leaping 

characteristic of other members of the rabbit family and is 

well suited for the complex structure of the preferred habitat. Although members of the rabbit family are generally 

silent, except when captured, pygmy rabbits demonstrate a 

unique capacity to produce warning calls, a trait well suited 

for a species with populations clustered in dense shrubby 
182 USDA Forest Service Proceedings RMRS-P-52. 2008

Kitchen and Carlson Great Basin Cold Desert Shrublands and the Desert Experimental Range

habitat with poor visibility (Green and Flinders 1981). High 

densities of characteristically small (diameter 2-5 mm) fecal 

pellets concentrated near burrow openings are diognostic of 

pigmy rabbit colonies. Although appropriate legal status for 

the species across its range is currently unresolved, concern 

for the species is growing due to extensive loss of suitable 

sagebrush habitat on western North American landscapes 

(Oliver 2004).

The tour stopped near a known pigmy rabbit colony located 

a short distance from the Parowan Gap turnoff from State 

Highway 130. At this, site pigmy rabbit burrows appeared to 

be restricted to the sides of a shallow ravine where soils were 

loose and sagebrush plants relatively large. Tour participants 

located several burrows and about six pigmy rabbits were 

identified moving through the sagebrush cover.

Big and Black Sagebrush 

Habitat Comparisons _______________

North of the Parowan Gap turnoff, Highway 130 begins 

a gradual climb up a broad bajada on the south side of the 

Black Mountains. Several distinct examples of ecotones 

between big and black sagebrush (A. nova) communities 

are apparent near the road. Black sagebrush has darker 

foliage and generally lower stature than big sagebrush and 

transitional boundaries between stands of the two species are 

generally easy to distinguish (McArthur and Stevens 2004). 

Solid stands of black sagebrush indicate shallow or rocky 

soils; thus, the pattern of these two sagebrush communities 

on this slope are an indication of variability in the development and structure of the alluvial soils that support them.

Stop 2: Mountain Big Sagebrush 

Burn Site/Utah Prairie Dog 

Rehabilitation Site__________________

Scattered Utah juniper (Juniperus osteosperma) trees occur 

imbedded within the sagebrush matrix on the upper reaches 

of the Black Mountain bajada. Tree density increases with 

elevation and trees dominate much of the steeper slopes and 

ridge-tops on the rolling hills above the bajada. Just south 

of the summit, the road passes through the burn area of the 

1988 Coyote Pond Fire. Here a large transplanted colony of 

the threatened, Utah prairie dog (Cynomys parvidens) can be 

observed along a dirt road extending east from the highway. 

The numerous animals and burrows encountered along this 

road suggest that this colony is doing well.

Most sagebrush species are easily killed by fire. Recovery 

is from seeds that either survive the fire or are dispersed 

from plants located on nearby unburned areas. Post-burn 

sagebrush recovery occurs more rapidly on mesic sites (such 

as those occupied by mountain big sagebrush (spp. vaseyana) 

than on xeric sites (Kitchen and McArthur 2007). This stop 

(1,980 m) was near the lower elevational limit of mountain 

big sagebrush (spp. vaseyana) in this region and could be 

considered a moderately mesic site with mean annual precipitation of approximately 330 mm. Differences in sagebrush 

and understory density and cover as affected by slope position, 

aspect, and post-burn seeding treatments were observed at 

this site 18 years after disturbance. Distinct size classes of 

big sagebrush plants suggest that establishment has been 

episodic, likely related to sequential climatic patterns favorable for seed production and seedling survival.

The tour turned northwest on state road 21 at Milford, 

Utah. On this route Wyoming big sagebrush (spp. wyomingensis) soon gives way to black sagebrush, juniper woodland, 

and salt-desert shrubland and grassland vegetation. Near 

the southern end of the San Francisco Range the highway 

passes a part of the old San Francisco Mining District and 

the remnants of the once booming town of Frisco—more 

about that later—and quickly descends towards the Wah 

Wah Valley floor dissecting an atypically large stand of 

Nevada ephedra (Ephedra nevadensis) that occupies several hundred hectares of the west-facing bajada. The salt 

desert communities below the Lake Bonneville shoreline 

in Wah Wah Valley (1,600 m) are highly degraded as a 

result of heavy overstocking with domestic sheep in the 

late 19th and early 20th centuries (Murdock and Welsh 

1971). Introduced annuals dominate where shrublands of 

shadscale (Atriplex confertifolia) and winterfat (Ceratoides 

lanata) were once abundant. Pedestaled plants are an 

indication that wind erosion is still a problem.

Stop 3: Wah Wah Summit— 

Juniper Expansion into Desert 

Grasslands and Shrublands __________

On the western side of the valley the highway passes 

the irrigated circles of the Wah Wah Ranch and begins an 

ascent from the valley floor, up an alluvial slope to Wah 

Wah Mountain Pass. Relatively young Utah juniper trees 

are scattered within the shorter shrub and grass vegetation growing a short distance up this slope, increasing in 

density with elevation. South of the highway about 2 km 

from the summit (1,800 m elev.) is a good place to stop for a 

closer examination. Here, understory species included the 

shrubs black sagebrush, Nevada ephedra, broom snakeweed 

(Gutierrezia sarothrae), winterfat, and the perennial grasses 

needle-and-thread (Hesperostipa comata), Indian ricegrass 

(Achnatherum hymenoides), and bluegrass (Poa spp.). 

The introduced annual cheatgrass (Bromus tectorum) is 

ubiquitous. Aging of several small to moderate sized juniper 

trees from this location reveals that most are young, having 

established during the latter half of the 20th century. Trees 

that appear to be several centuries old are restricted to the 

shallow ravines and rocky ridge-lines that dissect the alluvial 

slope. Charred remnants provide evidence that fire played a 

role in shaping this plant community in the past, including 

the suppression of tree expansion.

Stop 4: Desert Experimental Range 

Headquarters______________________

The Desert Experimental Range, consisting of 22,530 ha, 

was set aside in 1933 by President Herbert Hoover “as an 

agricultural range experiment station.” Events that led to 

its designation began some 60 years prior to official establishment. In the 1860s, Euro-American settlers began using 
USDA Forest Service Proceedings RMRS-P-52. 2008 183

Great Basin Cold Desert Shrublands and the Desert Experimental Range Kitchen and Carlson

the desert valleys of the Great Basin as pasturelands for 

domestic livestock. They soon discovered that many grasses 

and shrubs that dominate the landscape make excellent 

livestock forage, even in winter. Use of this resource during 

warmer months was restricted due to a scarcity of surface 

water; however, during winter and early spring, livestock 

often met their water requirements from unmelted snow 

allowing for widespread herd distribution.

The public rangelands of the Great Basin presented an open 

invitation to livestock owners; they were viewed as commons 

where grazing was free on a ‘first come first serve’ basis with 

no bothersome regulation. Cattle came first, then, beginning in the 1880s, large numbers of sheep followed. Sheep 

herds spent summers in well-watered mountain pastures 

that surround the Great Basin. Several weeks in spring and 

autumn were spent trailing up to 300 miles between mountain 

and desert ranges. Livestock numbers soon exceeded what the 

land could bear by several times, resulting in serious erosion 

and long-term damage to preferred plant species. Livestock 

numbers remained high and desert rangelands continued to 

degrade for several decades. Finally, during the drought years 

of the 1930s, the nation began to address the need for change. 

With passage of the Taylor Grazing Act in 1934 the era of unrestricted grazing on public lands came to an end. By this time 

it was clear that knowledge necessary for wise management of 

these desert shrublands was lacking. The Desert Experimental 

Range was created to help fill that knowledge gap.

In 1933 a large camp of the Civilian Conservation Corps., 

or CCC’s, was established a few hundred yards from the 

current site of the DER headquarters. Under the direction 

of federal researchers, their mission was to establish and 

begin operation of the Desert Experimental Range. By 1935 

they had completed construction of the headquarters facility with an office, three dwellings, shop, ice house, horse 

barn, garages, a 600-foot well, tennis court, and 45 miles 

of telephone line from Milford. Initial plant surveys were 

conducted and vegetation maps completed. Crews also 

constructed 118 miles of fence using hundreds of rolls of 

woven field fence and barbed-wire and over 40,000 ‘cedar’ 

posts cut by hand from Utah juniper woodlands located 

in the foothills to the south and west. The straightness of 

the fence-lines and uniformity of post top-cuts attest to the 

precision employed in this work. Most of the original posts 

still stand, evidence of their longevity and of the dryness of 

the climate. Fencing was located primarily along the outside 

boundary and on the perimeters of experimental pastures 

and ungrazed exclosures.

Long-term grazing experiments were designed to examine 

the economic and ecological impacts of sheep grazing on 

salt-desert shrublands using different grazing practices, and 

have continued with only minor alteration to the present. 

Eventually research broadened to include the evolution and 

ecology of several of the plant species adapted to the harsh 

environment. Desert animals have also been subjects of 

research with studies focused on pronghorn antelope (Antilocapra americana), kit fox (Vulpes macrotis), American 

kestrels (Falco sparverius) and various rodents. Additional 

studies have addressed ecosystem processes and mechanisms 

of ecosystem change through time. The Desert Experimental 

Range remains a valuable site for research and education in 

the ecology and management of salt-desert shrublands.

Stop 5: Desert Research 

Natural Area ______________________

A 740-ha parcel of land was set aside for protection from 

livestock grazing in 1934. This parcel was officially designated 

a Forest Service Research Natural Area in 1972. Although 

several smaller 0.4 and 1.6 ha fenced exclosures are scattered 

within and near the Desert Experimental Range, they are 

too small for investigations in salt-desert shrubland ecology 

requiring protection from livestock grazing at larger spatial 

scales. The size of the Research Natural Area and its history 

of continuous protection from livestock use make it unique 

in the Great Basin. The soils of the Research Natural Area 

are a loose mix of gravel, sand, and clay particles to a depth 

of 20 to 40 cm (Tew and others 1997). Below this, a 20- to 

40-cm thick cemented calcite layer acts as a barrier to root 

development and water penetration below this point. This 

semi-impermeable layer is a typical feature for stable alluvial 

soils with a calcareous parent material and dry climate.

Salt-Desert Shrublands

Plant species that inhabit salt-desert shrubland communities must be adapted to long periods of drought, moderate to 

high levels of soil-salt content, summer heat, and extreme 

winter cold (Blaisdell and Holmgren 1984). Mean annual 

precipitation for the Desert Research Natural Area is 165 

mm. Dominant plant forms include short-statured shrubs 

such as shadscale, winterfat, budsage (Artemisia spinescens), 

and low rabbitbrush (Chrysothamnus greenei), and perennial grasses such as Indian ricegrass, needle-and-thread, 

squirreltail (Elymus elymoides), purple three-awn (Aristida 

purpurea), sand drop-seed (Sporobolus cryptandrus), and 

blue grama (Bouteloua gracilis). Numerous short-lived forbs 

add floristic diversity and are periodically abundant (Harper 

and others 1996). Gooseberry-leaf globemallow (Sphaeralcea 

grossulariifolia) is the most common of these forbs at the 

Desert Experimental Range. Representatives of these plant 

forms are blended in an ever-changing mosaic across the 

landscape. The short plants help to hold water and nutrients 

in the soil and protect the soil surface from erosion caused by 

wind and pounding summer rains. Bare soil between plants 

is protected by tightly-fitted surface gravel, or ‘desert pavement’ and by an erosion-resistant soil crust glued together 

by fungi and algae living near the surface. Improper grazing 

practices destroy these microbiotic crusts and disturb desert 

pavement resulting in increased erosion.

Burrow Island Ecology

A variety of animal inhabitants of salt-desert shrublands 

require burrows for nesting, hunting, hiding from predators, 

and thermoregulation in summer and winter. Desert animals that make or use burrows in this desert include pocket 

gopher (Thomomys bottae), kangaroo rat (Dipodomys ordii), 

deer mouse (Peromyscus maniculatus), skunk (Mephitis 

mephitis), badger (Taxidea taxus), kit fox, burrowing owl 

(Speotyto cunicularia), Great Basin rattlesnake (Crotolus 

viridus lutosus), and a variety of lizards. On landscapes with 

well developed calcite layers, burrows are restricted primarily 

to scattered patches or ‘islands’ where somewhat continuous 
184 USDA Forest Service Proceedings RMRS-P-52. 2008

Kitchen and Carlson Great Basin Cold Desert Shrublands and the Desert Experimental Range

burrowing has prevented the formation of a calcified barrier. 

These burrow islands occupy about 10 to 15 percent of the 

landscape and are somewhat evenly scattered across the 

landscape varying from 3 to 15 m in diameter (Kitchen and 

Jorgensen 1999). They are easily identified by an abundance 

of holes and soil mounds and by a marked change in plants 

species present when compared to surrounding vegetation. 

Vegetation on the burrow mounds differs from that of the 

surrounding landscape because of frequent soil disturbance, 

higher soil fertility, and deeper water penetration. Burrow 

mounds in Pine Valley, as in other parts of the Great Basin, 

have been invaded by Russian thistle (Salsola spp.), cheatgrass, and the poisonous succulent halogeton (Halogeton 

glomeratus) (weedy herbs introduced from Europe and Asia) 

(Kitchen and Jorgensen 1999, 2001).

Sheep Grazing Studies

The effects of domestic sheep grazing on salt-desert 

shrubland ecosystems have been under study at the Desert 

Experimental Range since 1934. In that year, specific grazing treatments were assigned to 20 paddocks, each 97 to 

130 ha in size. Treatments combined one of three levels of 

grazing intensity with one or more seasons of use. Originally, 

seasonal treatments were autumn, winter, spring, autumn/

winter, autumn/spring, and winter/spring. Later, autumn and 

winter grazing were combined into a single mid-December 

to mid-January dormant-season treatment. Spring grazing 

occurs in late March and early April when many plants are 

actively growing.

Two 0.4-ha exclosures are located inside each of 16 of the 

paddocks and serve as ungrazed reference areas. Vegetation 

changes are periodically measured using permanent plots 

located as paired series inside and outside each exclosure. 

The study design allows scientists to evaluate long-term 

patterns of salt-desert shrubland succession and response to 

climate variability with, and without, the different grazing 

treatments (Chambers and Norton 1993; Kitchen and Hall 

1996; Alzerreca and others 1998). Research suggests that 

the salt-desert shrubland ecosystem is relatively insensitive to grazing intensity during winter months when plants 

are dormant. Conversely, spring grazing at all intensities 

causes ecosystem instability and the loss of important shrub 

species and a proliferation of annual weeds (Harper and 

others 1996).

Stop 6: Near Pine Valley Playa: 

Halogeton Invasion of Winterfat 

Shrublands ________________________

Winterfat also known as whitesage, is a dominant, evergreen shrub in Great Basin salt-desert shrublands. It 

is generally long-lived and extremely tolerant of drought. 

Winterfat is readily eaten by livestock and wildlife. The high 

nutritive value of this plant during winter was recognized 

by early livestock operators and is reflected in the common 

name. Though common throughout salt-desert communities, 

winterfat often occurs as near monocultures in the deep 

heavy-textured soils common to valley bottoms (Blaisdell 

and Holmgren 1984).

Numerous species of Eurasian plants are disruptive to 

North American ecosystems. Several of these aliens have successfully invaded shrublands of the Great Basin. Halogeton, 

a salt-adapted annual forb, is the most disruptive of these 

species on sites dominated by winterfat (Harper and others 

1996; Kitchen and Jorgensen 2001). Although halogeton is a 

poor competitor in healthy perennial communities, it quickly 

occupies and dominates disturbances such as roadways, 

livestock bedding grounds, and rodent burrows. In good 

water years, large quantities of long-lived seeds are produced 

and dispersed by wind. Over time, halogeton can change the 

physical, chemical, and biological properties of soils making 

them unsuitable for most plants species (Harper and others 

1996). The result is a form of desertification where winterfatdominated shrublands are replaced by near pure stands of 

halogeton (Kitchen and Jorgensen 2001). Management of 

infested landscapes is further complicated by halogeton’s 

toxic effects on livestock, especially sheep.

The first halogeton plants observed on the Desert Experimental Range were found near the playa in 1952. By 1958 

it was widespread. Halogeton has increased in dominance 

with time as new disturbances presented opportunities for 

expansion. For example, spring mortality of winterfat was 

observed here at this stop west of the Pine Valley Playa in 

the early 1970s after intense thunderstorms caused late 

summer flooding. In the absence of perennial competition, 

halogeton quickly occupied the affected landscape. The halogeton stand has expanded episodically as more winterfat 

plants have died, effectively moving the winterfat-halogeton 

ecotone. Winterfat has failed to reoccupy halogeton-infested 

sites. Soils tests have quantified changes in soils 30 years 

after halogeton infestation (Duda and others 2003; Harper 

and others 1996; Kitchen and Jorgensen 2001). Current 

investigations focus on how halogeton-caused changes in 

the soil environment affect winterfat survival and seedling establishment. Strategies for restoring shrublands to 

halogeton-infested landscapes are also being tested. Studies 

evaluating the establishment and persistence of a variety 

of desirable shrubs on halogeton-infested landscapes are 

part of these tests.

Stop 7: Frisco Mine Town Site________

The Frisco town site is located within the San Francisco 

Mining District near a low summit of the San Francisco 

Mountains approximately 27 km west of present-day 

Milford. It was established in 1876 after two miners 

discovered a large outcrop of ore on their way to work in 

September of the previous year. Believing the ore deposit 

was small, these miners sold their claim for $25,000. Over 

the next 10 years the Horn Silver Mine would yield some 

$60,000,000 silver, gold, zinc, copper, and lead (Anon. 

2005). Reaching a population of approximately 6,000, 

Frisco emerged as the major residential and commercial 

center supporting the Horn Silver and smaller mines 

in the district. With no local source, water had to be 

freighted at least 20 km to supply the needs of the town. 

With over 20 saloons, brothels, and gambling halls, and 

frequent murders, Frisco gained a reputation of being a 

wild town even by mining town standards. A make-noarrest lawman hired from Pioche, Nevada, is given credit 
USDA Forest Service Proceedings RMRS-P-52. 2008 185

Great Basin Cold Desert Shrublands and the Desert Experimental Range Kitchen and Carlson

for creating a tamer environment by driving the lawless 

element from town.

Five large beehive ovens were constructed in 1877 to supply 

charcoal for the new Frisco Mining and Smelting Company 

smelters. Wood for the charcoal was primarily from pinyon 

pine (Pinus monophyla), Utah juniper, and curlleaf mountain 

mahogany (Cercocarpus ledifolius) growing on the slopes of 

the San Francisco Mountains. These ovens still stand and 

are in reasonably good condition. Other charcoal ovens were 

located elsewhere on the mountain at close proximity to good 

wood sources. The Utah Southern Railroad completed a line 

from Milford to Frisco in 1880. From the completion of the 

line, the ore was shipped to the Salt Lake Valley for smelting 

and the need for charcoal from the large smelters largely 

disappeared. A major cave-in at the Horn Silver Mine on 

February 12, 1885, marked the end of the mine and the town 

connected to it, although a small population persisted for a 

few decades. By 1920, 35 years after ore was first discovered, 

the mine and town were completely abandoned.

Stop 8: Greenville Bench Pinyon-Juniper 

Treatment and Reseeding Site ________

The final stop of the tour was at a pinyon-juniper treament site, a few miles southwest of Beaver. This project 

is a cooperative effort between the USDI Bureau of Land 

Management, USDA Natural Resource Conservation 

Service, Utah Division of Wildlife Resources, and the nongovernmental organizations, Beaver Conservation District 

and Sportsmen for Fish and Wildlife. The objectives of the 

project were to reduce sediment yields into the Beaver River 

and Minersville Reservoir and to improve wildlife habitat by 

reducing tree cover and increasing cover and production of 

preferred shrubs and perennial herbaceous species. Trees on 

treatment areas were individually reduced to piles of wood 

chips and shredded bark by a piece of equipment known as 

a bull hog. The distribution of these mulch piles provided 

an indication of pre-treatment tree density. Two separate 

seeding treatments were applied to adjoining areas using 

diverse seed mixes composed of either native or introduced 

species. Good herbaceous cover values were associated with 

both seed mixes. Tour participants were also able to compare 

treatment effects to untreated areas and ask questions of 

BLM project managers.

References ________________________

Alzerreca-Angelo, H.; Schupp, E.W.; Kitchen, S.G. 1998. Sheep 

grazing and plant cover dynamics of a shadscale community. 

Journal of Range Management 51: 214-222.

Anonymous. 2005. Utah legends: Frisco—a ten year high. Legends 

of America. Available at URL: http:// www.legendsofamerica.com/

UT-Frisco.html (last accessed 16 April 2008).

Billings, W.D. 1978. Alpine phytogeography across the Great Basin. 

Great Basin Naturalist Memoirs 2: 105-117.

Blaisdell, J.P.; Holmgren, R.C. 1984. Managing Intermountain 

rangelands: salt-desert shrub ranges. Gen Tech. Rep. INT-163. 

Ogden, UT: U.S. Department of Agriculture, Forest Service, 

Intermountain Research Station. 52 p.

Chambers, J.C.; Norton, B.E. 1993. Effects of grazing and drought on population dynamics of salt desert shrub species on the Desert Experimental 

Range, Utah. Journal of Arid Environments 24: 261-275.

Duda, J.J.; Freeman, D.C.; Emlen, J.M.; Belnap, J.; Kitchen, S.G.; 

Zak, J.C.; Sobek, E.; Tracey, M.; Montante, J. 2003. Changes in 

native soil ecology associated with invasion of the exotic annual 

chenopod, Halogeton glomeratus.Biol Fertil Soils 38: 72-77.

Green, J.S.; Flinders, J.T. 1981. Alarm call of the pygmy rabbit 

(Brachylagus idahoensis). Great Basin Naturalist 41: 158-160.

Harper, K.T.; Van Buren, R.; Kitchen, S.G. 1996. Invasion of alien annuals and ecological consequences in salt desert shrublands of western 

Utah. In: Barrow, J. R., McArthur, E.D., Sosebee, R.E., Tausch, R.J., 

comps. Proceedings: Shrubland ecosystem dynamics in a changing 

environment; 1995 May 23-25; Las Cruces, NM. Gen. Tech. Rep. 

INT-GTR-338. Ogden, UT: U.S. Department of Agriculture, Forest 

Service, Intermountain Research Station: 58-65.

Kitchen, S.G.; Hall, D.B. 1996. Community stability in a salt-desert 

shrubland grazed by sheep: the Desert Experimental Range story. 

In: Barrow, J.R., McArthur, E.D., Sosebee, R.E., Tausch, R.J., 

comps. Proceedings: Shrubland ecosystem dynamics in a changing 

environment; 1995 May 23-25; Las Cruces, NM. Gen. Tech. Rep. 

INT-GTR-338. Ogden, UT: U.S. Department of Agriculture, Forest 

Service, Intermountain Research Station: 102-107.

Kitchen, S.G; Jorgensen, G.L. 2001. Winterfat decline and halogeton 

spread in the Great Basin. In: McArthur, E.D.; Fairbanks, D.J., 

comps. Shrubland ecosystem genetics and biodiversity: Proceedings; 2000 June 13-15; Provo, UT. Proc. RMRS-P-21. Ogden, UT: 

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

Research Station: 200-203.

Kitchen, S.G.; Jorgensen, G. L. 1999. Annualization of rodent burrow clusters and winterfat decline in a salt-desert community. In: 

McArthur, E.D.; Ostler, W.K.; Wambolt, C.L., comps. Proceedings: 

Shrubland ecotones; 1998 August 12-14; Ephraim, UT. Proc. 

RMRS-P-11. Ogden, UT: U.S. Department of Agriculture, Forest 

Service, Rocky Mountain Research Station: 175-180.

Kitchen, S.G.; McArthur, E.D. 2008. Big and black sagebrush landscapes. In: Hood, S.M.; Miller, M., tech. eds. Fire ecology and management of the major ecosystems of southern Utah; Gen. Tech. Rep. 

RMRS-GTR-212. Fort Collins, CO: U.S. Department of Agriculture, 

Forest Service, Rocky Mountain Research Station: 73-95.

McArthur, E.D.; Stevens, R. 2004. Composite shrubs. In: Monsen, 

S.B.; Stevens, R.; Shaw, N.L., comps. Restoring western ranges 

and wildlands. Gen. Tech. Rep. RMRS-GTR-136-vol-2. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky 

Mountain Research Station: 493-537.

Morris, T.H.; Stubben, M.A. 1994. Geologic contrasts of the Great Basin 

and Colorado Plateau. In: Harper, K.T.; St. Clair, L.L.; Thorne, K.H.; 

Hess, W.M., eds. Natural history of the Colorado Plateau and Great 

Basin. Niwot, CO: University Press of Colorado: 9-25.

Murdock, J.R.; Welsh, S.L. 1971. Land use in Wah Wah and Pine 

Valleys, western Utah. Science Bulletin, Biological Series – Vol. 

XII, No. 4. Brigham Young University, Provo, UT. 25 p.

Oliver, G.V. 2004. Status of the pygmy rabbit (Brachylagus idahoensis) in Utah. Utah Natural Heritage Program, Utah Division of 

Wildlife Resources. 45 p. Available at URL: http://greatbasin.

wr.usgs.gov/docs/PYRA_Oliver_rpt2004.pdf (last accessed 16 

April 2008).

Peterson, K.L. 1994. Modern and Pleistocene climatic patterns in 

the West. In: Harper, K.T.; St. Clair, L.L.; Thorne, K.H.; Hess, 

W.M., eds. Natural history of the Colorado Plateau and Great 

Basin. Niwot, CO: University Press of Colorado: 27-53.

Tew, R.K.; Kitchen, S.G.; Holmgren, R.C. 1997. Soil survey—desert 

Experimental Range. Gen. Tech. rep. INT-GTR-347. Ogden, UT: 

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

Research Station. 22 p.

Thompson, R.S. 1990. Late quarternary vegetation and climate in 

the Great Basin. In: Betancourt, J.L.; Van Devender, T.R.; Martin, 

P.S., eds. Packrat middens: The last 40,000 years of biotic change. 

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since the last glaciopluvial. Ecological Monographs 53: 341-382.

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein

USDA Forest Service Proceedings RMRS-P-52. 2008 187

In: Kitchen, Stanley G.; Pendleton, Rosemary L.; Monaco, Thomas A.; 

Vernon, Jason, comps. 2008. Proceedings—Shrublands under fire: 

disturbance and recovery in a changing world; 2006 June 6–8; Cedar 

City, UT. Proc. RMRS-P-52. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Stop 1: Mile Square Subdivision, 

Escalante Valley Near 

Beryl Junction _________________

Extensive drought during the years of 2002, 2003, 

and 2004 removed most plant cover. On May 10, 2004, 

a cold front to the north resulted in weather stations 

in the area recording 600+ miles of wind travel (comparable to 60 mile hour winds for 10 hours). The effect of these two climatic events was to bury the Mile 

Square subdivision in wind-blown sand. Sand filled 

homes, garages, barrow pits, and closed roads. Even 

though the drought decimated perennial vegetation, 

including killing most of the shrubs in the area, the 

only areas that did not blow were the areas occupied 

by dead shrubs.

Stop 2: Mountain Meadow _____

This stop was the scene of the tragic Mountain 

Meadow Massacre of September 11, 1857, where 120 

men, women, and children were massacred. History 

of the incident was discussed. This site was a resting 

and staging area on the Old Spanish Trail circa 1800 

to 1850 and later a wagon road to southern California. 

The original vegetation was wet meadows near the 

numerous springs and grasslands on the drier sites. 

Overgrazing in the mid 1800s caused arroyos to cut 

north and south resulting in the lowering of the water 

table and the displacement of the original grassland 

by xeriphytic sagebrush. Presently, private lands at 

the site are seeded rangelands or are under cultivation. A photo taken on March 23, 1877, reveals few 

pinyon or juniper trees at that time (fig. 1). Abundant 

pinyon and juniper at the present time is consistent 

with evidence observed from throughout the West that 

pinyon-juniper woodlands have increased dramatically 

since the mid 1800s (fig. 2).

Disturbance and Rehabilitation of 

Cold to Warm Desert Transitional 

Shrublands in Southwestern Utah

Chad Reid and James Bowns

Figure 1—Mountain Meadows, March 1877.

Figure 2—Mountain Meadows, July 2005.
188 USDA Forest Service Proceedings RMRS-P-52. 2008

Reid and Bowns Disturbance and Rehabilitation of Cold to Warm Desert Transitional Shrublands in Southwestern Utah

Stop3: Pinyon-Juniper Blackbrush 

Chaining Near Jackson Spring________

Chaining was conducted on deep soils in 1964 near Jackson 

Spring. The BLM seeded the area to crested wheatgrass and 

intermediate wheatgrass (fig. 3). These introduced grasses 

were abundant when data was first collected in 1984. Small 

amounts of Indian ricegrass, sand dropseed, purple threeawn, and bottlebrush squirreltail were also present. By 2001 

most of the wheatgrass plants had disappeared. Sagebrush 

and juniper dominate the site at the present time (fig. 4).

Figure 4—Same seeding as figure 3 except a slightly different 

view. Previously seeded area as it looks today (2007).

Figure 3—Excellent seeding following chaining of blackbrush 

and pinyon-juniper, approximately 6 years after the seeding 

(photo taken in 1970).

Stop 4: Blackbrush _________________

This site was formerly a dense, essentially monotypic stand 

of blackbrush (Coleogyne ramosissima) (such as fig. 5). From 

this vantage point you can see extensive burned areas in 

Utah and Nevada that were previously stands of blackbrush 

(fig. 6). Once burned, backbrush sites such as this do not 

return to blackbrush. We are currently seeing extensive areas 

of red brome (Bromus rubens) and cheatgrass (B. tectorum) 

taking the place of blackbrush. Seedings in this area using 

introduced forage grasses are largely unsuccessful and are 

instead occupied by cheatgrass, red brome, and sprouting 

shrubs such as desert almond (Prunus fasciculata), desert 

bitterbrush (Purshia glandulosa), turpentine broom (Thamnosma montana) and to a lesser extent creosote bush (Larrea 

tridentata). Joshua tree (Yucca brevifolia) plants are often 

able to survive a single fire, but appear to be susceptible to 

repeated burning. Some preliminary success in seeding the 

introduced half-shrub, forage Kochia (Kockia prostrata), 

offers some hope that this drought tolerant, fire resistant 

species may successfully establish and persist on these harsh 

sites following fires.

Figure 5—Tour participants examining undisturbed stand of 

blackbrush.

Figure 6—View of burned-over blackbrush (stop 4) looking 

northwest (photo taken in 1981).
USDA Forest Service Proceedings RMRS-P-52. 2008 189

Disturbance and Rehabilitation of Cold to Warm Desert Transitional Shrublands in Southwestern Utah Reid and Bowns

Stop 5: 1979 Prescribed Burn in 

Blackbrush, Seeded to Introduced 

Wheatgrass _______________________

This burn was seeded to intermediate and crested wheatgrasses. Favorable moisture conditions at that time resulted 

in a good stand of these introduced perennial grasses (fig. 7). 

An exclosure was constructed on this site to permit evaluation 

of any effects of livestock grazing on the seeding. Like most 

seedings on these monotypic blackbrush sites, the seeding 

has failed both inside the exclosure and in the surrounding 

area. Herbaceous vegetation is predominantly, cheatgrass 

and red brome with a few native annuals. Desert bitterbrush 

and desert almond have sprouted but there is no evidence 

of blackbrush recovery on the site.

Figure 7—Former blackbrush area subjected to controlled burn 

and seeding (photo take in 1980).

Stop 6: 1945 Blackbrush Burn _______

This stop is a burn that occurred in 1945 adjacent to the 

Desert Inn (DI) landing strip (fig. 8). This is the oldest burn 

that we have a record of and it is unclear whether this was 

an intentional or natural burn. There is no evidence that 

this burn was seeded to perennial grasses. A desirable 

stand of native shrubs including desert bitterbrush, desert 

almond, and big sagebrush have subsequently occupied 

and stabilized the site. To the best of our knowledge these 

were not seeded.

Figure 8—Photo of 1945 blackbrush burn (stop 6). Burned 

area on the right; unburned blackbrush on left.

Stop 7: 1980 Unauthorized 

Intentional Burn ___________________

This stop was seeded and chained to cover seed (fig. 9). A 

good stand of introduced perennial wheatgrasses resulted. 

However, as with most seedings in this area, the seeded species did not persist and are now absent. A recent (2 years ago) 

seeding of forage Kochia may ultimately provide persistent 

perennial cover. The encroachment of yerba santa (Eriodictyon angustifolium) was apparent at a later date.

Figure 9—View of unauthorized 1981 burn and subsequent 

seeding.

Figure 10—Blackbrush burn near Jackson Spring. Photo shows 

forage Kochia as small half shrub on the near left side of the 

fence. Large plants on near slope are mostly sideoats grama, 

intermediate wheatgrass, and Kochia.

Stop 8: Burn Near Jackson Spring ____

This blackbrush site is located near the Pahcoon Spring 

south of the Motoqua road (fig. 10). The area burned in 1995 

and again in 1999. This site was seeded to intermediate 

and crested wheatgrass, forage Kochia, and sideoats grama 

(Bouteloua curtipendula). The wheatgrasses are still occupying the site but this has happened on previous burns only 

to disappear over time. This was one of the first sites to be 

seeded to forage Kochia and sideoats grama; however, it is 

too early to know if these species will persist or reduce the 

biomass of the annual bromes.
190 USDA Forest Service Proceedings RMRS-P-52. 2008

Reid and Bowns Disturbance and Rehabilitation of Cold to Warm Desert Transitional Shrublands in Southwestern Utah

Figure 11—True interior chaparral type east of Pine Valley 

Mountains.

Stop 9: True Interior Chaparral 

on Oak Grove Road ________________

This stop was in true interior chaparral, a vegetative community clearly adapted to fire in contrast to what had been 

observed at previous stops (fig. 11). To us, this is the most 

unique vegetation type in Utah. This chaparral, which is 

characterized by low growing evergreen shrubs, lies between 

the pinyon-juniper and gambel oak/ponderosa pine zones. 

Dominant shrubs are desert lilac (Ceanothus greggii), Silktassel (Garrya flavescens), manzanita, and live oak (Quercus 

turbinella). Non-sprouting shrubs such as snakeweed and 

sagebrush as well as small amounts of pinyon and juniper 

trees were also observed. Vegetation response to a 1986 

burn was examined.

Return to Cedar City _______________

The content of this paper reflects the views of the author(s), who are responsible 

for the facts and accuracy of the information presented herein
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