United States

Department of

Agriculture

Forest Service

Rocky Mountain

Research Station

General Technical Report 

RMRS-GTR-135-vol. 1.

September 2004

Assessment of Grassland 

Ecosystem Conditions 

in the Southwestern 

United States

Volume 1

Editor

Deborah M. Finch
Abstract___________________________________________________________

Finch, Deborah M., Editor. 2004. Assessment of grassland ecosystem conditions in the Southwestern United States. Volume 1. Gen. 

Tech. Rep. RMRS-GTR-135-vol. 1. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research 

Station. 167 p.

This report is volume 1 of a two-volume ecological assessment of grassland ecosystems in the Southwestern United States. Broadscale assessments are syntheses of current scientific knowledge, including a description of uncertainties and assumptions, to provide a 

characterization and comprehensive description of ecological, social, and economic components within an assessment area. Volume 1 

of this assessment focuses on the ecology, types, conditions, and management practices of Southwestern grasslands. The second volume, 

due to be published in 2005, describes wildlife and fish species, their habitat requirements, and species-specific management concerns, 

in Southwestern grasslands. This assessment is regional in scale and pertains primarily to lands administered by the Southwestern Region 

of the USDA Forest Service (Arizona, New Mexico, western Texas, and western Oklahoma). A primary purpose of volume 1 is to provide 

information to employees of the National Forest System for managing grassland ecosystems and landscapes, both at the Forest Plan level 

for Plan amendments and revisions, and at the project level to place site-specific activities within the larger framework. This volume should 

also be useful to State, municipal, and other Federal agencies, and to private landowners who manage grasslands in the Southwestern 

United States.

Key words: grasslands, ecological assessment, Southwestern United States, ecosystem conditions, Arizona, New Mexico, Texas, Oklahoma

Acknowledgments__________________________________________________

This project is a collaborative effort between the Rocky Mountain Research Station and the Southwestern Region of the U.S. Department 

of Agriculture, Forest Service. In early 2000, a Southwestern Grassland Ecosystem Sustainability Team was formed to determine the assessment approach and general content. By 2001, the team had evolved into a core group that finalized the assessment topics and authored 

the report. This current work is volume 1 of a two-volume report.

The core team members are:

Robert Calamusso, Southwestern Region, Tonto National Forest

Cathy Dahms, Southwestern Region, Ecosystem Analysis and Planning (retired)

Deborah Finch, Rocky Mountain Research Station

Reggie Fletcher, Southwestern Region, Ecosystem Analysis and Planning (retired)

Paulette Ford, Rocky Mountain Research Station

Gerald Gottfried, Rocky Mountain Research Station

Burt Pendleton, Rocky Mountain Research Station

Rosemary Pendleton, Rocky Mountain Research Station

Deborah Potter, Southwestern Region, Ecosystem Analysis and Planning

Carol Raish, Rocky Mountain Research Station

Bryce Rickel, Southwestern Region, Wildlife, Fish and Rare Plants

Wayne Robbie, Southwestern Region, Ecosystem Analysis and Planning

Michele Zwartjes, Rocky Mountain Research Station

We acknowledge Art Briggs (Southwestern Region), Jim Saveland (Rocky Mountain Research Station), Dave Stewart (Southwestern 

Region), and George Martinez (Southwestern Region) for their early role in the team’s development, and Alison Hill (Rocky Mountain 

Research Station), Bob Davis (Southwestern Region), Amy Unthank (Southwestern Region), and Don DeLorenzo (Southwestern Region) 

for assisting in seeing both volumes through.

In addition, the following individuals contributed sections or information for the report:

Rick Atwell, Southwestern Region, Cibola National Forest

Brent Botts, Southwestern Region, Recreation Management

Curt Flather, Rocky Mountain Research Station

Pat Frieberg, Southwestern Region, Engineering

William Krausmann, Southwestern Region, Engineering

Penny Luehring, Southwestern Region, Watershed and Air Management

Gene Onken, Southwestern Region, Rangeland Management

Jean Szymanski, Southwestern Region, Public Affairs Office

Reuben Weisz, Southwestern Region, Ecosystem Analysis and Planning

Bill Woodward, Southwestern Region, Engineering

Patrick Zwartjes, Rocky Mountain Research Station

We thank Louise Kingsbury and her staff for assistance in editorial work and layout of chapters in preparation for publication. We thank 

Nora Altamirano for help in formatting the document. The team also recognizes the valuable input provided by the following individuals: 

Reggie Blackwell, Sara Chubb, Larry Cosper, Carl Edminster, Vicki Estrada, Steve Libby, Doug MacPhee, Mitch McClaran, Darrell Musick, 

Dan Rael, George Ruyle, Nick Schmal, David Seesholtz, John Sidle, Rory Steinke, Nancy Walls, and Mitchel White. During the review 

process, valuable comments and useful critiques were provided by all members of the team, by reviewers of individual chapters, as well 

as by Will Moir and Rex Peiper.

Support for this project was provided by USDA Forest Service’s Southwestern Region, and by the USDA Forest Service’s Rocky Mountain 

Research Station (Research Work Units are Ecology, Recovery, and Sustainability of Grassland and Riparian Ecosystems in the Southwest, 

Cultural Heritage Research, and Southwestern Borderlands Ecosystem Management).

Editor’s Note_______________________________________________________

Although this report may be used as input in processes initiated under the National Environmental Policy Act (NEPA), National Forest 

Management Act (NFMA), and other applicable laws, it is not a decision document, does not allocate resources on public lands, and does 

not make recommendations to that effect. The information in this report is general in nature rather than site-specific. The opinions expressed 

by the authors do not necessarily represent the policy or position of the U.S. Department of Agriculture and the Forest Service.
Assessment of Grassland Ecosystem 

Conditions in the Southwestern 

United States

Volume 1

Editor

Deborah M. Finch

USDA Forest Service

Rocky Mountain Research Station

333 Broadway SE, Suite 115

Albuquerque, New Mexico 87102

Cover photos clockwise from upper left: A grassland-juniper ecotone, Bernalillo watershed, Cibola National Forest, New 

Mexico (photo by Rosemary Pendleton). Urban development of native grassland near Flagstaff, Arizona (photo by John 

Yazzie). Prescribed fire on the Sevilleta National Wildlife Refuge, New Mexico (photo by Burt Pendleton). Introduced bison 

coexist with reintroduced black-tailed prairie dogs on the Armendaris Ranch, New Mexico (photo by Paulette Ford).
The Authors_______________________________________________________

Cathy Dahms worked as a Resource Analyst in the Ecosystem Analysis and Planning Section of the Southwestern Region of the USDA 

Forest Service from about 1990 until retirement in 2001. There she led the Regional Forest Health Assessment and the Regional Strategic 

Monitoring Plan, coordinated the Continuing Education in Ecosystem Management Program, and was a member of the Scientific Work 

Group for the Southwest Strategy. Prior to that Dahms was the Region’s Geographic Information Systems Coordinator in the Information 

Systems and Telecommunications Staff, a computer programmer analyst for the Computer Systems Staff, and a computer systems analyst 

with the Federal Highway Administration.

Deborah M. Finch is a Research Wildlife Biologist and Project Leader with the USDA Forest Service, Rocky Mountain Research Station 

in Albuquerque, NM. She is also Research Associate Professor in the Department of Biology at University of New Mexico. She received 

her B.S. degree in wildlife management from Humboldt State University, California, her M.S. degree in zoology from Arizona State University, and her Ph.D. degree in zoology from the University of Wyoming. Finch’s research interests include neotropical migratory bird 

ecology and reproduction, bird migration, endangered species, riparian and grassland ecosystem ecology and conservation, and relationships between commodity use and natural resources. Finch has published more than 125 articles and journal papers and is an active 

member of several professional societies. Since 1998, she has served as Leader of the Technical Subgroup of the Recovery Team for the 

Southwestern Willow Flycatcher.

Reggie Fletcher is a consultant with his consultant firm, Ecological Interpretations, specializing in promoting management toward sustainability of Southwestern ecosystems. Projects for State and Federal government, nonprofit organizations, and private concerns have included 

inventory, monitoring, assessment, and training in ecosystems ranging from alpine to the Sonoran desert. He retired from the Forest Service Region 3 in 2000 as Regional Botanist and Regional Ecologist. He was actively involved in all natural resource issues undertaken by 

the Forest Service in Arizona and New Mexico for nearly two and a half decades including rare species, range management, watershed, 

riparian (including serving as a member of the Regional Riparian Committee for more than 10 years and as Riparian Coordinator for more 

than 5 years), fire, timber and wildlife, with tasks ranging from policy design to field projects. He has an M.S. degree in plant taxonomy 

from the University of New Mexico.

Paulette L. Ford is a Research Ecologist with the Rocky Mountain Research Station in Albuquerque. She received a B.S. degree in biology 

and psychology, and an M.S. degree in biology from the University of New Mexico, in addition to a Ph.D. degree in renewable natural 

resources from the University of Arizona. Her interests include the role of disturbance in structuring grassland and desert communities, 

and scale and ecosystem resilience.

Gerald J. Gottfried is a Research Forester with the Rocky Mountain Research Station assigned to the Southwestern Borderlands Ecosystem 

Management Project in Phoenix, AZ. His current research centers on rangeland restoration in southeastern Arizona and southwestern 

New Mexico using mechanical and prescribed fire treatments. He and his collaborators are conducting a study of the impacts of cool and 

warm season prescribed fire in oak savannas on hydrologic, soil, overstory and understory vegetation, and wildlife ecosystem components. 

Gottfried also is conducting research on the ecology and silviculture of southwestern pinyon-juniper woodlands. Gottfried has a Ph.D. 

degree from the University of Arizona in watershed management with an emphasis on dryland forestry.

Michele Merola-Zwartjes is a Fish and Wildlife Biologist and is working for the U.S. Fish and Wildlife Service, Division of Endangered 

Species Recovery, Region 1, Portland, OR. She was a postdoctoral Research Wildlife Biologist with the Rocky Mountain Research Station 

in Albuquerque, NM. She holds a B.A. degree in communications from the University of California, San Diego, and a Ph.D. degree in 

biology from the University of New Mexico. Among her varied work experiences, Merola-Zwarthes has served as a conservation planner for The Nature Conservancy and taught in the Wildlife Department of Humboldt State University. Her interests include ornithology, 

conservation of threatened, endangered and sensitive vertebrate species, animal behavior, wildlife management, and natural resource 

management issues.

Burton K. Pendleton is a Research Ecologist with the Rocky Mountain Research Station in Albuquerque. He received a B.S. degree in 

botany and an M.S. degree in ecology from Brigham Young University in Provo, UT. He received his Ph.D. degree in biological sciences 

from Wayne State University in Detroit, MI. His research interests include plant recruitment, reproductive biology, and establishment 

ecology.

Rosemary L. Pendleton is a Research Ecologist with the Rocky Mountain Research Station in Albuquerque. She received a B.S. degree in 

botany and an M.S. degree in range science from Brigham Young University, and a Ph.D. degree in biological sciences from Wayne State 

University. She has been a scientist with the Rocky Mountain Research Station and former Intermountain Research Station since 1986. 

Her research interests include plant reproductive biology and mycorrhizal fungi.

Deborah Ulinski Potter is a Physical Scientist with the USDA Forest Service, Region 3, Watershed and Air Management since 1989, including a 2-year assignment to the Rocky Mountain Research Station in Albuquerque. Program responsibilities include Ecology, Air Resource 

Management, Riparian Ecosystems, and Research Liaison. She received a B.S. degree in biology with distinction from the University of 

New Mexico, an M.S. degree in zoology, specializing in ecology and limnology, from the University of Georgia, and a Ph.D. degree from 

the University of New Mexico, specializing in ecosystem ecology. She has two faculty appointments, Research Assistant Professor of Biology and Research Assistant Professor of Earth and Planetary Sciences at the University of New Mexico.

Carol Raish is a Research Social Scientist with Cultural Heritage Research, Rocky Mountain Research Station, Albuquerque Lab. Her study 

interests include understanding the role of traditional economic practices, such as ranching, in maintaining cultural identification and 

traditional lifeways among American Indians, Hispanic Americans, and Anglo Americans in northern and central New Mexico. She has 

the B.A. degree in Spanish and M.A. and Ph.D. degrees in anthropology/archeology.

Wayne A. Robbie is a Supervisory Soil Scientist within the Ecosystem Analysis and Planning and Watershed and Air staff, USDA Forest Service, Southwestern Region, Albuquerque. He has worked on five National Forests in the Southwest and is the ecological inventory program 

manager for Region 3. He received a B.S. degree in agronomy and an M.S. degree in forest soils from New Mexico State University.
Contents__________________________________________________________

Page

Chapter 1: Purpose and Need for a Grassland Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

Deborah M. Finch

Cathy W. Dahms

Chapter 2: Grassland Assessment Categories and Extent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .11

Wayne A. Robbie

Chapter 3: Southwestern Grassland Ecology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .18

Paulette L. Ford

Deborah U. Potter

Rosemary Pendleton

Burton Pendleton

Wayne A. Robbie

Gerald J. Gottfried

Chapter 4: Biodiversity, Functional Processes, and the Ecological Consequences of

 Fragmentation in Southwestern Grasslands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .49

Michele Merola-Zwartjes

Chapter 5: Historic and Contemporary Land Use in Southwestern Grassland Ecosystems. . . . . . . . .86

Carol Raish

Chapter 6: Historic and Current Conditions of Southwestern Grasslands . . . . . . . . . . . . . . . . . .120

Reggie Fletcher

Wayne A. Robbie

Chapter 7: Grassland Sustainability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .130

Deborah U. Potter

Paulette L. Ford

Chapter 8: Tools for Management for Grassland Ecosystem Sustainability: 

 Thinking “Outside the Box” . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .142

Gerald J. Gottfried

Appendix: Discussion Question Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .164
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 1

Purpose___________________________

This report is volume 1 of an ecological assessment 

of grassland ecosystems in the Southwestern United 

States, and it is one of a series of planned publications 

addressing major ecosystems of the Southwest. The 

first assessment, General Technical Report RM-GTR295, An Assessment of Forest Ecosystem Health in the

Southwest (by Dahms and Geils, technical editors, 

published July 1997), covered forested ecosystems. 

Given the complexities of grassland ecology and the 

increasing number of challenges facing grassland 

managers, the USDA Forest Service Southwestern 

Region, in partnership with the agency’s Rocky 

Mountain Research Station, focused on grasslands 

in its second assessment. The assessment is regional 

in scale and pertains primarily to lands administered 

by the Southwestern Region (Arizona, New Mexico, 

Texas, and Oklahoma).

Broad-scale assessments are syntheses of current 

scientific knowledge, including a description of uncertainties and assumptions, to provide a characterization 

and comprehensive description of ecological, social, 

and economic components within an assessment area 

(USDA Forest Service 1999b). A primary purpose of 

this assessment is to provide context to National 

Forest System land management planning efforts 

involving grasslands, both at the Forest Plan level 

for Plan amendments and revisions, and at the 

project level to place site-specific activities within 

the larger framework. The assessment is not a decision document because it surfaces issues and risks 

to grassland ecosystems that provide the foundation 

for future changes to Forest Plans or project activities but does not make any site-specific decisions or 

recommendations. The report also provides a scientific 

basis for conducting ecosystem restoration projects, 

provides a starting point for public discussion on 

desired conditions for the future, and contributes to 

the overall understanding of the physical, biological, 

and human dimensions of grassland ecosystems in 

the Southwest.

The report is divided into two volumes. The first 

volume (herein) focuses on the ecology, types, conditions, and management practices of Southwestern 

grasslands. The second volume emphasizes wildlife 

and fish species and their habitat requirements in 

Southwestern grasslands.

To prepare this document, we assembled a team of 

authors from the Southwestern Region and the Rocky 

Mountain Research Station whose expertise focused 

on or included grassland ecosystems. An outline of 

chapter titles and chapter contents was prepared using a group consensus process. Authors volunteered 

to write specific chapters that were then reviewed 

by the team. Following team review, each individual 

chapter was sent to a minimum of two peer reviewers 

for critique, and in addition, the entire revised volume 

was sent to two reviewers. Also, the team interviewed 

Forest Service employees (see appendix).

Deborah M. Finch

Cathy W. Dahms

Chapter 1:

Purpose and Need for a Grassland 

Assessment
2 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

We thank all the authors for writing and rewriting 

their chapters. We are grateful to Art Briggs and 

Bob Davis from the Regional Office for supporting this project. We appreciate helpful reviews on 

the entire document by Will Moir and Rex Peiper. 

Reviews provided on individual chapters are also 

much appreciated. This project was financially supported by the Regional Office of USDA Forest Service 

Southwestern Region and by the USDA Forest Service 

Rocky Mountain Research Station. We thank Paulette 

Ford, Carol Raish and Rose Pendleton for helpful comments on chapter 1. We thank the Station’s Publishing 

Services staff for helpful editing and layout.

Southwestern Grassland 

Ecosystems________________________

In the Southwestern Region, the Forest Service 

has adopted the Soil Conservation Society (SCS) of 

America’s definition of grasslands, that is, “lands 

on which the existing plant cover is dominated by 

grasses” (SCS 1982). Risser (1995) defined grasslands 

as “biological communities that contain few trees 

or shrubs, are characterized by mixed herbaceous 

vegetation, and are usually dominated by grasses.” 

Supported by the National Science Foundation, the 

U.S. International Biological Program (IBP) characterized natural grasslands as climatically determined 

by soil water availability and precipitation volume 

and seasonality, successional grasslands where forest vegetation has been removed, and agricultural

grasslands where a few native or introduced species are maintained. This report addresses natural 

grasslands.

This assessment includes the following Southwestern grassland types:

• Montane grassland

• Colorado Plateau

• Desert grassland

• Great Basin grassland

• Plains grassland

Subalpine grasslands are discussed within the 

montane grassland category. Alpine grasslands are 

not discussed as a separate category because they 

have a limited extent in the Southwest. Where they 

are mentioned, they are discussed in conjunction with 

montane grasslands, although they occur on a different 

mountain gradient. Riparian and/or wetland inclusions occur in all grassland types and are discussed 

separately where appropriate.

Ecologists and geographers identify broad-scale 

Southwestern grasslands (that is, biome level) as 

temperate grassland. These biome classifications 

are according to macroclimate conditions defined 

by Köppen, Threwartha and others (Coupland 

1992, French 1979). Further subdivisions according 

to physiographic province include Great Basin 

grassland and Plains grassland. In general, the term 

plains refers to grasslands in areas of flat topography 

(Coupland 1992). Plains grassland is then subdivided 

into short, tall, and mixed grasslands. At a finer scale, 

grasslands are designated by vegetation community/plant association as classified by Küchler (1964), 

Clements (1920), and others.

Temperate grasslands are areas at mid-latitude that 

are dominated by perennial grasses and forbs. Climate 

is moderately dry (semiarid) with discrete wet/dry 

seasons and temperature and precipitation extremes 

(Sims 1988). Soils are predominantly characterized 

as Aridisol or Mollisol with large amounts of humus 

(Aber and Melillo 1991, Sims 1988, Whittaker 1975). 

Temperate grasslands include tall, mid and short 

grasses (Odum 1971, Whittaker 1975). Tall grasses 

are about 150 to 245 cm (5 to 8 feet) high, mid grasses 

approximately 60 to 120 cm (2 to 4 feet) high, and short 

grasses 15 to 45 cm (0.5 to 1.5 feet) in height. Short 

grasses include buffalograss (Buchloe dactyloides),

blue grama (Bouteloua gracilis ), and other gramas. 

Mid grasses include little bluestem (Schizachyrium 

scoparium), needlegrass (Stipa spp.), western wheatgrass (Pascopyrum smithii), and Indian rice grass 

(Achnatherum hymenoides) (Odum 1971).

Temperate grassland biomes include prairie and

steppe (French 1979, McKnight 1993, Odum 1971, 

Whittaker 1975). Prairie, including the true tall-grass 

prairie, mixed-grass prairie, and short-grass prairie, 

is dominated by grasses and forbs, has a scarcity 

of shrubs, and has no trees. Mixed-grass prairie 

is an ecotone between tall and short-grass prairie. 

The term steppe refers to a temperate biome that 

is dominated by short grasses and bunchgrasses 

(McKnight 1993) and is dryer than prairie. Steppes 

receive approximately 25 to 50 cm (10 to 20 inches) 

of rain per year and experience hot summers and cold 

winters; these climatic conditions support plants such 

as blue grama, buffalograss, big bluestem (Andropogon

gerardi), cacti, and sagebrush.

Grasslands can be subdivided using the U.S. National 

Vegetation Classification system (Federal Geographic 

Data Committee 1997) and other methods according to 

class, subclass, group, formation, regional biome type, 

alliance, plant associations, or habitat types. Grassland 

categories for the Southwestern United States include 

the Plains grassland, Great Basin grassland, and 

the Colorado Plateau as discussed above, as well as 

montane grassland and desert grassland. Montane 

grassland can be found in small patches within the 

mixed conifer and ponderosa pine forests. Montane or 

high-mountain grasslands consist of meadows below 

timberline (French 1979), while alpine grassland is 

located above timberline (Whittaker 1975). Desert 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 3

grassland occurs in arid and semi-arid climates. It 

varies in composition from mixed herbaceous species 

with few shrubs to primarily a combination of shrub 

species (French 1979) such as mesquite (Prosopis spp.) 

and creosote bush (Larrea tridentata). Subregions include the Chihuahuan Desert grasslands of southern 

New Mexico, characterized by black grama (Bouteloua 

eriopoda), and the Sonoran Desert grasslands of southeastern Arizona.

Biome classification systems use the term savanna

to describe tropical grasslands that are primarily 

located in Africa and Australia, South America and 

southern Asia/India (Whittaker 1975, McKnight 

1993). Similarly, Bailey’s ecosystem classification 

uses the term savanna to describe a Division within 

the Humid Tropical Domain. However, some people 

use the term temperate savanna to describe areas in 

the Southwestern United States. McPherson (1997) 

defines North American savannas as “ecosystems 

with a continuous grass layer and scattered trees or 

shrubs.” The woody plant overstory has approximately 

30 percent cover or less with a grass understory. He 

further defines and maps areas of the Southwest as 

Piñon-Juniper Savanna, Southwestern Oak Savanna, 

Ponderosa Pine Savanna, and Mesquite Savanna. 

Although this terminology is sometimes used, a 

savanna biome does not occur in the Southwestern 

United States according to broad-scale vegetation 

classification systems based on climate. The Forest 

Service also uses a classification system based on a 

geographic approach, also referred to as regionalization, which is a process of classification and mapping 

to identify homogeneous map units at various scales. 

The National Hierarchical Framework of Ecological 

Units adopts Bailey’s classification by ecoregions 

(Bailey 1995); the hierarchy consists of domain, division, province, section, subsection, landtype, landtype 

association, and landtype phase. At the regional scale, 

provinces and sections are the most useful units for 

assessments.

Relationship of Assessment to 

Ecosystem Management____________

Ecosystem management is an evolving philosophy 

that has been adopted by many government agencies including the Forest Service. The Forest Service 

has defined ecosystem management as “a concept 

of natural resource management wherein national 

forest activities are considered within the context of 

economic, ecological, and social interactions within a 

defined area or region over both short and long term” 

View of Animas Valley, looking east toward Animas Mountains, New Mexico. Mesquite in perennial grassland. 

(Photo by Ronald Bemis)
4 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

(Thomas and Huke 1996). National Forest activities 

in this context are all activities occurring on National 

Forest System lands, including grassland ecosystems. 

Ecosystem management is sometimes referred to as 

ecology-based multiple-use management in that there 

is a shift from focusing exclusively on sustaining 

production of goods and services to sustaining the 

viability of ecological, social, and economic systems. 

While other agencies and organizations have developed their own definitions of ecosystem management 

reflecting their differing missions, they typically have 

a goal of ecosystem sustainability or maintaining 

ecological integrity (Grumbine 1994, Kaufmann and 

others 1994) while recognizing that people are part 

of the ecosystem and that human needs should be 

reflected in ecosystem sustainability (Keystone Center 

1996, USDA Forest Service 1994).

A significant difference among various ecosystem 

management philosophies may be the degree in which 

people are included within the sustainability concept. 

Former Forest Service Chief Dombeck stated, “We will 

still track traditional outputs of goods and services 

but they will be accomplished within the ecological 

sideboards imposed by land health” (Dombeck 1999). 

The Keystone Center’s National Policy Dialogue 

Group on Ecosystem Management placed the goals 

of sustaining vibrant, livable, and economically diverse human communities and the involvement of 

stakeholders on a par with the goals of maintaining 

ecosystem integrity and sustaining biodiversity and 

ecosystem processes at a regional scale (Keystone 

Center 1996). The Southwestern Region’s philosophy 

embraces all these goals as well. In developing the 

human dimension principles and strategies for the 

Southwestern Region, the Human Dimensions Team 

recognized that human needs and wants must be 

balanced with ecosystem capabilities. However, on a 

practical level, they also recognized that the goal of 

meeting human needs frequently conflicts with the 

goal of sustaining natural ecosystems (USDA Forest 

Service 1994). When this is the case, ecosystem management may be considered an optimal integration 

of ecological sustainability and human dimensions 

(including both economic considerations and societal 

needs and desires) (Jensen and others 1996).

Assessments are a tool in ecosystem management to 

develop a holistic understanding of ecological sustainability as well as the human dimension of ecosystems. 

As an introductory chapter, chapter 1 not only describes 

the purpose and need for assessments, it also gives a 

brief overview of Southwestern grassland types and 

defines two terms in frequent use in later chapters: 

ecosystem sustainability and adaptive management. 

Chapter 1 also explores and emphasizes the role of, 

and need for, monitoring of grassland conditions and 

trends, a topic not covered in detail in later chapters. In 

this report, chapter 2 describes the extent and types 

of grasslands in the Southwest. The general ecology 

of Southwestern grasslands is evaluated in chapter 

3. A discussion of the biological diversity, functional 

processes, and consequences of grassland fragmentation is provided in chapter 4. Cultural dimensions 

of grassland management, both from a historic and 

contemporary perspective, are covered in chapter 

5. Historic and current conditions of southwestern 

grasslands in relation to land management are covered in chapter 6. Chapter 7 discusses the concept 

of grassland sustainability and why understanding 

sustainability is critical for managers and stakeholders to collaboratively develop desired conditions for 

grassland areas. Management decisions will need to 

be site-specific based on the unique characteristics of 

the area. Because there is no one-size fits-all management strategy for an area, chapter 8 discusses 

a wide range of tools available for use by grassland 

managers.This chapter has an eye toward highlighting some of the more innovative work being done in 

grasslands rather than attempting to document all 

possible tools. Chapter 8 also covers research needs, 

since adaptive management and the utilization of the 

best scientific knowledge are important components 

of ecosystem management.

Ecosystem Sustainability____________

Ecosystem sustainability is the ability of an 

ecosystem to maintain ecological processes and functions, biological diversity, and productivity over time 

(Kaufmann and others 1994). It was the subject of 

the 1992 Earth Summit/United Nations Conference 

on Environment and Development in Rio de Janeiro 

and a 1987 World Commission on Environment and 

Development report called Our Common Future (also 

known as the Bruntland Report; Bruntland 1987). 

Sustainable ecosystems are able to maintain their 

ecological integrity.

Ecological integrity is achieved when ecosystem 

structure, function, processes, and services are preserved over space and time (Grumbine 1994). Ecosystem 

structure is the spatial arrangement of the living and 

nonliving elements of an ecosystem, for example, abiotic 

elements (temperature, light, wind, relative humidity, 

rainfall) and community structure (species richness 

and the distribution of heterotrophs, autotrophs, and 

consumers). Ecosystem function refers to the processes 

whereby the living and nonliving elements of ecosystems change and interact, such as biogeochemical 

processes and succession. Ecological processes are the 

actions or events that link organisms and their environment. Ecosystem processes include disturbance, 

succession, evolution, adaptation, natural extinction 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 5

rates, colonization, dispersion, fluxes of materials, and 

decomposition (Kaufman and others 1994).

Sustainable ecosystems continue to provide essential services. The Ecological Society of America 

(1997) has identified the following services provided 

by ecosystems:

• moderating weather extremes and their impacts

• seed dispersal

• drought and flood control

• protection from ultraviolet rays of the sun

• nutrient cycling

• protection from channel and coastal erosion

• waste decomposition and breakdown

• agricultural pest control

• maintaining biodiversity

• generating and preserving soils; renewing soil 

fertility

• helping to stabilize climate

• cleaning the air and water

• regulating organisms that carry disease

• pollination

Sustainable grasslands can be described as productive grasslands with erosion rates that do not 

exceed soil tolerance for sediment and nutrient loss; 

having natural fire frequencies; biologically diverse; 

not overgrazed; and containing important social and 

aesthetic values (Mac and others 1998).

Adaptive Management______________

This assessment utilizes existing information, and 

the assessment team did not initiate any new data 

collection activities. However, the assessment team 

synthesized information on grasslands that had 

been available only in a piece-meal fashion up to 

this point, and the presentation of that information 

at the regional scale provides new insights that can 

be used in adaptive management.

Adaptive management is a proactive approach to 

implement ecosystem management. The theory and 

practice of adaptive management has evolved over 

the past two decades through the works of Holling 

(1978) and Lee (1993). One of the fundamental tenets 

of adaptive management is that ecosystems and people 

are unpredictable as they evolve together. Ecological 

conditions change as do societal values and economic 

developments. In addition, our understanding of ecosystem behavior is imperfect, and managers will never 

be able to completely predict responses to management 

activities. The purpose of adaptive management is to 

rapidly increase our level of knowledge of ecosystem 

dynamics and the effects of management. In the 

adaptive management approach, management plans 

and activities are continually altered in response to 

changing societal values and goals, emerging issues, 

and new scientific understanding (Haynes and others 

1996).

The dynamic nature of grasslands requires an 

adaptive management approach. Change has been 

and continues to be rapid in grassland ecosystems. 

The intensity of grazing has fluctuated through time, 

new uses such as off-road vehicle recreation and 

energy development have emerged, and development 

pressures continue to escalate. Some wildlife species 

have declined or disappeared altogether, while other 

species have increased in population. These changes, 

along with others such as climatic variation, introduced 

plants, and changing fire regimes, have all shaped 

grassland ecosystems, sometimes to the point where 

the land cannot be returned to a previous state. 

Adaptive management experiments would be useful 

in understanding where these thresholds occur and 

management’s ability to reverse changes as the system 

approaches these thresholds.

Adaptive management encompasses both deliberate experimentation to gain new knowledge (active 

adaptive management) as well as the ongoing process 

of monitoring and inventorying to assess the effects 

of management actions (passive adaptive management). Passive adaptive management may seem a 

misnomer because it requires an active program for 

the monitoring and evaluation of project activities. 

Active adaptive management is a departure from 

traditional management in that it views management actions as experiments from which to learn. 

Conducting adaptive management experiments 

involves being explicit about expected outcomes, 

designing methods to measure responses, collecting 

and analyzing information to compare expectations to 

actual outcomes, learning from the comparisons, and 

changing actions and plans accordingly. In both forms 

of adaptive management, monitoring plays a crucial 

role in surfacing any needed changes to management 

plans and activities, and monitoring can lead to the 

need for new assessments.

Ecological Monitoring of 

Grasslands________________________

The basis of ecosystem management is management driven by explicit goals and made adaptable by 

monitoring and research based on ecological interactions and processes necessary to sustain ecosystems 

(Gordon et al. 1992). Monitoring, as an overall process, 

is a measurement process that establishes a baseline 

with periodic measurements for the purposes of change 

detection and adaptive management. Monitoring is 

defined as the systematic observation of parameters 

related to a specific problem, designed to provide 

information on the characteristics of the problems 
6 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

and their changes with time (Spellerberg 1991). In a 

natural resources context, ecological monitoring assesses the status and trends of ecological, social, or 

economic outcomes (Powell, unpublished paper 2000). 

The information being collected depends on the purpose 

of the monitoring—that is, which questions are to be 

addressed—and the scale at which a question needs to 

be answered. The results of monitoring are expected 

to generate an action of some kind, even if the action 

is to maintain current management (Johnson and 

others 1999).

Monitoring is a step-wise process that involves:

1. Framing a question(s) and developing a study 

plan to address the question(s) using a standard 

protocol.

2. Collecting data according to the monitoring 

plan.

3. Storing the data for retrieval.

4. Evaluating the results.

Monitoring should include goals, thresholds for 

change, and remedial actions that occur when thresholds are met or exceeded. An ecological systems 

approach to monitoring ensures a strong foundation in ecological theory, adequate consideration and 

understanding of cause-and-effect relationships, and 

a systematic approach to selecting and evaluating 

parameters that are monitored (USDA Forest Service, 

unpublished paper 1999a). Furthermore, the questions should focus on key ecological processes and 

interactions, rather than on individual parts of the 

system. Powell (unpublished papers 2000, 2001) 

listed the following characteristics for successful 

monitoring:

• Purposeful and conducted to answer specific questions. Conducted at the appropriate spatial and 

temporal scale to answer the question.

• Conducted in collaboration with others (for example, agencies, interested publics, researchers, 

and nongovernmental organizations) to share the 

workload (including obtaining data from other 

sources), gain expertise, and build credibility 

and trust.

• Conducted using the best available science and 

established protocols to collect and evaluate the 

data.

• Conducted using modern information management techniques and tools.

• Stringent selection criteria applied so that a 

monitoring activity is only conducted if it is 

feasible, realistic, and affordable.

• Evaluation emphasized as much as the collection 

of the data.

Protocols for monitoring should include standard 

sampling and analytical methods that determine 

the precision and accuracy of measurements. Proper 

training and supervision of field and laboratory staff 

is necessary to ensure adherence to the protocol and 

the success of the monitoring program.

Based on these criteria, one of the activities in the 

interagency Southwestern Strategy effort is to develop 

a unified set of tools and techniques, approved by 

Federal and State agencies, for rangeland monitoring training in New Mexico (de la Torre, personal 

communication).

Monitoring addresses both management activities 

on the land and conditions of the ecosystem being 

monitored. The monitoring of management activities 

can further be subdivided into three arenas: implementation, effectiveness, and validation monitoring. 

Implementation monitoring addresses, evaluates, and 

determines whether plans, projects, and activities 

were implemented as designed and in compliance 

with Forest Plan objectives, standards, and guidelines. 

Effectiveness monitoring addresses, evaluates, and 

determines whether plans, projects, and activities 

met Forest Plan management direction, objectives, 

standards, and guidelines. Validation monitoring determines whether the assumptions and relationships 

between the activity and the expected results were 

valid. In 2000, the Southwestern Region undertook 

an analysis of the Forest Plan Monitoring Reports 

for fiscal years 1998 and 1999. Most of the forest 

level monitoring in the Region was implementation 

monitoring (65 percent). About 32 percent of the 

reported monitoring was evaluation monitoring, and 

the remaining 3 percent was validation monitoring. 

The information was not detailed enough to know 

whether grassland monitoring followed this general 

trend (USDA Forest Service, unpublished data, 

2000). To monitor ecological conditions, quantitative or qualitative parameters are monitored; these 

parameters are typically referred to as indicators, 

attributes, variables, or monitoring elements. In connection with condition and trend monitoring, it is 

useful to establish baseline reference conditions to 

serve as a context for determining desired conditions 

and interpreting existing conditions.

In the Forest Service, monitoring and evaluation 

of management actions is one step in the overall 

process for proposed action identification, design, 

and monitoring. In outlining the process for Forest 

Plan implementation, the Southwestern Region 

has developed a desk guide, Integrated Resource

Management: The Road to Ecosystem Management. 

Monitoring forms the base of the Southwestern Region’s 

Integrated Resource Management Process triangle 

(USDA Forest Service 1993). In an effort to capture 

monitoring activities on National Forest System 

Lands, since 1997 an annual “Forest Monitoring and 

Evaluation Report and State of the Region Evaluation” 

has been included in the Forest Service’s performance 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 7

reporting system (MAR—Management Attainment 

Report). Despite being integrated into Forest Service 

processes, monitoring within the agency tends to be 

uncoordinated and inconsistent (Powell 2000). There 

is a perception in the Southwestern Region that 

monitoring activities chronically suffer from a lack 

of funding and personnel (see appendix, question 1). 

These problems may be even greater in grassland areas 

than in forested ecosystems, since progress has been 

slower at developing an agreed-upon set of ecological 

indicators to monitor.

Ecological Indicators of 

Sustainability______________________

Numerous ecological indicators have been proposed 

or are being used to address grassland sustainability. 

Indicators can be used to define any expression of the 

environment that estimates the condition of ecological 

resources, magnitude of stress, exposure of a biological 

component to stress, or the amount of change in a 

condition (Breckenridge and others 1994). Indicators, 

by their very nature, will vary depending on the 

scale to which they are applied. At the project-level 

and landscape-level scales, a commonly agreed-upon 

set of indicators for grasslands is lacking. Selecting 

indicators at these scales is an enormous challenge 

because of the complexity of ecosystems. There are 

potentially hundreds if not thousands of ecosystem 

characteristics that could be measured in response to 

environmental or management-induced changes, and 

it is difficult to separate responses occurring at the 

fine scale from potential influences occurring at the 

broader scale. One approach used at the Sevilleta Long 

Term Ecological Research (LTER) site is to identify 

a matrix of indicators at many levels and scales to 

sample the diversity of responses across spatial and 

temporal scales. Identifying the key indicators and 

the scales they reflect is an ongoing process (Gosz 

and others 1992).

On a national scale, there have been several major 

efforts to identify ecological indicators, though none 

specifically address grasslands. The National Academy 

of Science’s National Research Council recommended 

the following indicators to portray ecological condition: land cover and land use; total species diversity; 

native species diversity; nutrient runoff estimates, 

especially for nitrogen and phosphorus; soil organic 

matter; productivity (including carbon storage, net 

primary production, and production capacity); lake 

trophic status; stream oxygen concentration; and nutrient-use efficiency and nutrient balance (Committee 

to Evaluate Indicators for Monitoring Aquatic and 

Terrestrial Environments and others 2000). Land 

cover—including cropland, forest, nature reserves, 

rivers, wetlands, and riparian zones—is usually 

detected and monitored using remote sensing, while 

land use is typically classified by measurements 

from the ground. Soil organic matter is important 

as a nutrient and energy source for soil biota. It 

improves soil structure by strengthening soil aggregates, increases water retention and available 

water capacity, reduces the sealing of soil surfaces 

thereby promoting infiltration and reducing erosion, 

increases cation exchange capacity, chelates metals, 

and influences the fate of pesticides. Nutrient-use 

efficiency is a good indicator for crops, industries, 

counties, or watersheds.

A multiscale effort to develop indicators has 

sprung from the Montreal Process, which initially 

identified international criteria and indicators for 

sustainability of forest ecosystems. The United States 

has endorsed the seven criteria and 67 indicators, 

and the Forest Service is in the process of applying 

them within the agency. The criteria that have been 

identified are:

• Conservation of biological diversity.

• Maintenance of productive capability of forest 

ecosystems.

• Maintenance of forest ecosystem health and 

vitality.

• Conservation and maintenance of soil and water 

resources.

• Maintenance of forest contribution to global 

carbon cycles.

• Maintenance and enhancement of long-term 

multiple socio-economic benefits to meet the 

needs of societies.

• Legal, institutional, and economic framework 

for forest conservation and sustainable management.

The Forest Service has been conducting a series of 

pilot projects in an effort to identify indicators that 

would be appropriate at a Forest Management Unit 

scale, including those appropriate for rangelands 

(USDA Forest Service Inventory and Monitoring 

Institute 2001). Concurrently, a national team reviewed the Montreal indicators for their applicability 

to rangelands at a national scale. While the results 

should be considered preliminary, for the first five 

criteria, the review concluded that almost all the 

indicators were as relevant for rangeland ecosystems 

as for forested ecosystems, although the lack of a 

generally accepted classification system, consistent 

definitions, data, and/or mechanisms to measure most 

of the indicators makes quantification at a national 

scale problematic. Also, it was noted that some of the 

indicators may be more effective if they were refined 

to capture aspects of the criteria goals that are more 

relevant to rangeland dynamics (Flather and Sieg 

2000, Joyce 2000, Joyce and others 2000, McArthur 

and others 2000, Neary and others 2000).
8 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Why Assess Grassland Ecosystems?__

The selection of grassland ecosystems as the second 

major Southwestern ecosystem type to be assessed was 

based on several factors. A huge amount of grasslands 

acreage has been lost to development, agriculture, 

and other uses across the country, yet it is likely that 

the average person is unaware that grasslands are 

endangered. Over half of the critically endangered 

ecosystems in the United States (those with over 98 

percent of area either lost or degraded) are grasslands (Noss and others 1995). In the Southwest as 

elsewhere, grasslands are often lumped together with 

other grazed ecosystems and analyzed as rangelands 

or grazing lands. Between 1982 and 1992, Arizona 

had a net loss of 382,000 acres of non-Federal grazing 

land; New Mexico, 351,000 acres. During this period 

in New Mexico, 23 percent of non-Federal rangeland 

that was converted to other uses, was converted to 

developed land, and 16 percent was developed for agricultural uses (Goddard and others 1999). The amount 

of National Forest lands, including rangelands, has not 

increased significantly in the past 20 years (Mitchell 

2000) and the amount of Federal grasslands may even 

be lower in the Southwest due to the loss of montane 

meadows and the expansion of woodlands. The loss of 

grazing land in the non-Federal sector increases the 

importance of the remaining Federal lands.

The effects of grazing on grasslands have always been a contentious issue, particularly in the 

Southwest, since domestic livestock was introduced 

to the Southwest much earlier than other Western 

States. Sheep were brought to the Rio Grande pueblos 

in 1598, and other livestock soon followed. By 1890, it 

is estimated that over 1.5 million cattle were grazing 

in the Southwest, and range deterioration was being 

reported (Baker and others 1988). In a document 

submitted to the Senate in 1936, the Secretary of 

Agriculture indicated that the Southwest contained 

the most severe range depletion in the Western United 

States. Much of the Southwest was in the “extreme” 

range depletion class, where 76 to 100 percent of the 

area was considered depleted (Secretary of Agriculture 

1936). Determining the condition of rangeland today 

is complicated by not having a standard process in 

place across all ownerships. As assessment of nonFederal lands was made in 1989 from data collected in 

1982 during the National Resource Inventory (USDA 

Soil Conservation Service 1990) and supplemented 

in 1992. Conditions were determined based on how 

closely the species composition of a range site met 

those projected for the climax plant community for 

that site. For both 1982 and 1992, only 2 percent 

of non-Federal rangelands were considered to be in 

excellent condition in Arizona and New Mexico. During 

this period, percentages of good condition rangeland 

increased from 16 to 27 percent in Arizona, and from 

30 to 36 percent in New Mexico.

The USDI Bureau of Land Management (BLM) 

is a major Federal land manager of rangelands in 

the Southwest. The agency’s assessments of range 

condition are by ecological status categories: potential 

natural community (PNC), late seral, mid seral, and 

early seral. In 1996, while only 5,607,000 acres had 

been classified out of the 11,643,000 acres of BLM 

managed Arizona rangelands, the results were 9 

percent PNC, 40 percent late seral, 40 percent mid 

seral, and 12 percent early seral. In New Mexico, 

the percentage of acres classified is much greater. 

Out of 12,597,000 acres of BLM rangelands in New 

Mexico, 9,752,000 have been classified. These fall into 

the following categories: PNC 1 percent, late seral 

36 percent, mid seral 48 percent, and early seral 15 

percent (USDI Bureau of Land Management 1997). 

Forest Service assessments have also been in upland 

rangelands rather than specifically grasslands. These 

are not assessed by range condition categories, but 

by whether the vegetation meets or is moving toward 

Forest Plan Management Objectives (FPMO). Of all 

the Regions assessed during the 1995 to 1997 period, 

the Southwestern Region had the largest percentage 

of rangelands not meeting FPMO objectives and not 

moving toward FPMO, roughly 25 percent (Mitchell 

2000).

Another primary factor for selecting grasslands for 

an assessment effort is the complexity of grassland 

ecology combined with the variety of anthropogenic 

influences on grassland ecosystems. As was mentioned 

previously in the adaptive management section, 

changes have occurred rapidly in grassland ecosystems, 

and some areas may have passed a threshold where 

they cannot be returned to an earlier condition.

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 11

Introduction_______________________

This chapter establishes a general framework for 

describing the various kinds of grasslands outlined 

in subsequent chapters. This framework outlines the 

major categories or classes of grasslands that occur 

as part of Southwestern terrestrial ecosystems within 

National Forest System lands and provides an ecological and environmental context in regards to how they 

differ in their floristic, geographic, spatial, and climatic 

settings. More detailed information about these grassland systems is also presented in chapter 6.

Grasslands of the Southwest vary according to 

vegetation, climate, soils, and topography and disturbance regimes. They are distinctly different from 

other vegetation assemblages in that the dominant 

and codominant plants are graminoid species. For 

example, other forbs and shrub plant species occur 

within the grasslands but are subordinate to grass in 

the total cover and composition.

The major grassland categories used in this assessment—that is, those categories that represent the major 

grasslands formations in the Southwestern Region on 

National Forest System lands (Carleton and others 

1991)—are Desert, Plains, Great Basin, Montane, 

and Colorado Plateau grasslands. These generalized 

groupings reflect the geographic and ecological differences that are determined by unique floristic, edaphic, 

physiographic, and climatic characteristics. Although 

not taxonomic with respect to any vegetation hierarchy, these categories are intended to aid the reader 

in understanding the uniqueness, distribution, and 

extent of these systems. Other classification systems of 

Southwestern grasslands exist (Barbour and Billings 

2000, Brown 1994, Dick-Peddie 1993, Küchler 1964) and 

emphasize biogeographic, ecological, and biophysical 

features that are consistent with the scale and level of 

generalization being used here. The general distribution of grasslands for this assessment is located on the 

National Forest System lands in Arizona, New Mexico, 

Texas, and Oklahoma (fig. 2-1).

Grassland Categories_______________

Descriptions of each grassland assessment category 

follow.

The Desert Grassland encompasses annual and 

perennial graminoid and adjacent shrub communities 

at low elevations adjacent to the Chihuahuan, Mohave, 

and Sonoran deserts. These grasslands occur between 

the Great Basin grasslands, chaparral, and woodland 

ecosystems and have been commonly referred to as 

semidesert grasslands by Brown (1994). The distribution of these grasslands are mainly within the Basin 

and Range, Sonoran-Mohave Desert, Tonto Transition 

ecoregion sections, and limited areas within the White 

Mountain-San Francisco Peaks, Northern Rio Grande 

Intermontane, and Sacramento-Monzano Mountain 

ecoregion sections (McNab and Avers 1994). Desert 

grasslands intermingle with desert scrub communities 

(Dick-Peddie 1993) and have evolved through natural 

and anthropogenic successional disturbance processes. 

Grass species that are diagnostic to this category include 

Chapter 2:

Grassland Assessment Categories 

and Extent

Wayne A. Robbie
12 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 Figure 2-1. Grasslands of National Forest System lands in Arizona, New Mexico, Texas, and Oklahoma.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 13

black grama (Bouteloua eriopoda), tobosa (Pleuraphis 

mutica), and curly mesquite (Hilaria belangeri). Other 

key graminoid species that occur within this formation include bushmuhly (Muhlenbergia porteri) and 

burrograss (Scleropogon brevifolius). Major shrubs 

that occur in association with these species include 

creosote bush (Larrea tridentata), velvet mesquite 

(Prosopis velutina) in Arizona, western honey mesquite 

(Prosopis glandulosa var. torreyana) in southern New 

Mexico, tarbush (Flourensia cernua), turpentine bush 

(Ericameria laricifolia), desert ceanothus (Ceanothus 

greggii), and soaptree yucca (Yucca elata).

The Great Basin Grassland occurs within the White 

Mountain-San Francisco Peaks, Saramento-Manzano 

Mountains, Central Rio Grande Intermontane, and 

higher elevations of Basin and Range and Sonoran 

Desert ecoregion sections (McNab and Avers 1994) of 

the Southwestern region. These grasslands are higher 

in elevation and climatically cooler and moister than 

desert grasslands and are adjacent to and intermingle 

with juniper (Juniperus spp.) savanna ecosystems. The 

Great Basin grasslands are similar to Brown’s (1994) 

Plains and Great Basin grasslands and Dick-Peddie’s 

(1993) Plains–Mesa grasslands except the geographic 

range of this category for this assessment is restricted 

to the Basin and Range Physiographic province 

(Fenneman 1928). Diagnostic plant species include 

blue grama (Bouteloua gracilis), galleta (Pleuraphis 

jamesii), Indian ricegrass (Achnatherum hymenoides), 

and sideoats grama (Bouteloua curtipendula). Some 

dropseeds, (Sporobolus spp.) and wolftail (Lycurus 

phleoides) are codominant and add to the diversity 

of this category. The Great Basin grasslands tend to 

be drier than the Shortgrass Steppe grasslands and 

have a blend of warm and cool season graminoid and 

forb species. Shrubs that are present in association 

with grassland vegetation of this category include 

fourwing saltbush (Atriplex canescens), sacahuista 

(Nolina microcarpa), small soapweed yucca (Yucca 

glauca), skunkbush sumac (Rhus trilobata), and catclaw mimosa (Mimosa biuncifera). As this grassland 

integrades with savanna ecosystems, minor amounts 

of trees such as emory oak (Quercus emoryi), alligator juniper (Juniperus deppeana), and Utah juniper 

(Juniperus osteosperma) dominated woodlands are 

evident.

The Colorado Plateau Grassland is located in 

northern Arizona above the Mogollon Rim and northern 

New Mexico in association with the Colorado Plateau 

and adjacent to small areas of the Rocky Mountain 

physiographic provinces (Fenneman 1928). It occurs 

within the Grand Canyonlands, Painted Desert, Tonto 

Transition, White Mountains-San Francisco Peaks, 

Navajo Canyonlands, Southcentral Highlands, SouthCentral Highlands, Southern Parks and Ranges, and 

Upper Rio Grande Basin ecoregion sections (McNab 

and Avers 1994). Colorado Plateau Grasslands—a new 

category of Southwestern grassland—primarily splits 

the expansive Great Basin Grassland category based 

upon recent ecological mapping (Laing and others 1986, 

Miller and others 1995, Robertson and others 2000) 

and what Kuchler (1970) referred to as the GalletaThreeawn Shrub Steppe. These grasslands occur on 

nearly level, wind-desiccated geomorphic surfaces of 

sedimentary and igneous origin. Grass species that 

characterize this category include western wheatgrass 

(Pascopyrum smithii), needle and thread (Hesperostipa 

comata), blue grama (Bouteloua gracilis), galleta 

(Pleuraphis jamesii), and New Mexico feathergrass 

(Hesperostipa neomexicana), and various species of 

three-awn (Aristida spp). Common shrubs include 

big sagebrush (Artemisia tridentata), black sagebrush 

(Artemisia nova), fourwing saltbush (Atriplex canescens), and Mormon tea (Ephedra trifurca). Oneseed 

juniper (Juniperus monosperma) and Utah juniper 

(Juniperus osteosperma) woodlands and savannas are 

adjacent to Colorado Plateau grasslands.

The Plains Grasslands consist of the shortgrass, 

midgrass, and tallgrass prairies of the National 

Grasslands. These grasslands extend throughout 

the Great Plains physiographic province (Fenneman 

1928) and occur within the Southern High Plains, 

Pecos Valley, Redbed Plains, and Texas High Plains 

ecoregion sections (McNab and Avers 1994). Climate 

ranges from subhumid to semiarid as these grasslands 

extend from east to west. The characteristic plant 

species that are abundant throughout the shortgrass 

prairie include blue grama (Bouteloua gracilis) and 

buffalo grass (Buchloe dactyloides). The midgrass 

prairie ecosystem is codominated by little bluestem 

(Schizachyrium scoparium), blue grama (Bouteloua 

gracilis), and plains bristlegrass (Setaria vulpiseta). 

The tallgrass prairie is dominated by big bluestem 

(Andropogon girardii). These different prairie ecosystems are aggregated and reduced to one category for 

this assessment and reflects a wide range of ecological 

properties and processes.

The Montane Grasslands category includes the 

montane, subalpine and alpine meadows, valleys, and 

high elevation grasslands that occur throughout the 

Grand Canyonlands, Painted Desert, Tonto Transition, 

White Mountain-San Francisco Peaks, Basin and 

Range, Central Rio Grande Intermontane, SouthCentral Highlands, Sacramento-Manzano Mountain, 

Southern Parks and Ranges, and Upper Rio Grande 

Basin ecoregion sections (McNab and Avers 1994). 

These grasslands are similar to Subalpine-Montane 

Grasslands described by Dick-Peddie (1993) and the 

Alpine and Subalpine and Montane Meadow grasslands 

of Brown (1994). Carleton and others (1991) classified 

montane, subalpine and alpine terrestrial ecosystems 

as edaphic-fire and topo-edaphic-zootic disclimaxes 

with temperate continental climates. Diagnostic plant 

species that characterize these ecosystems include 

Arizona fescue (Festuca arizonica), mountain muhly 

(Muhlenbergia montanus), Kentucky bluegrass (Poa 
14 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

pratensis), timber oatgrass (Danthonia intermedia), 

Thurber fescue (Festuca thurberii), tufted hairgrass 

(Deschampsia caespitosa), alpine avens (Geum rossii), 

and Bellardi bog sedge (Kobresia myosuroides).

Mapping__________________________

The delineation of grasslands for this assessment 

involved integrating and cross-walking the categories 

of vegetation types within existing land cover classes 

and ecological units from five main sources: (1) General 

Terrestrial Ecosystem Survey (GTES) (Carleton and 

others 1991), (2) New Mexico Gap Analysis Project 

(Thompson and others 1996), (3) Texas Gap Analysis 

Project (Parker 2001), (4) Oklahoma Gap Analysis 

Project (Fisher 2001), and (5) Arizona GAP Analysis 

Project (Thomas 2001). These five primary sources 

were used for assessing distribution and extent of the 

five grasslands assessment categories.

The University of New Mexico, Earth Data Analysis 

Center, Albuquerque, performed data processing and 

geographic information system analysis.

The grassland assessment categories were nested 

within the Ecoregion and Subregions map of ecological 

units (Bailey and others 1994, McNab and Avers 1994). 

The Ecoregion and Subregions map and descriptions 

contain integrated biophysical information about 

broadscale ecological characteristics including climate, 

soils, geomorphology, potential natural vegetation, 

surface water characteristics, disturbance regimes, and 

land use. This integrated approach to regionalization of 

ecosystems allows managers, planners, and scientists 

to study management issues on a multi-Forest and 

Statewide basis. More mapping particulars are given 

in figures 2-2 and 2-3.

GAP land cover classes and GTES vegetation taxa 

(series) were combined through a process of correlation 

(table 2-1). This process involved aggregating categories 

with similar physiognomic, floristic, and geographic 

ranges into the five assessment classes. Differences occur between nomenclature and image resolution of land 

cover classes for each State GAP product. Furthermore, 

some States had broader land cover classes that include 

plant communities of adjacent vegetation formations. 

Consequently, the spatial resolution as predicted by 

the map may depict grasslands to be of more variable 

extent than what would be evident at finer scales 

with higher resolution. This is particularly true for 

the Desert and Great Basin grasslands where these 

communities integrade and commingle with adjacent 

shrubland steppe communities. Conversely, some 

areas of known grasslands on National Forest lands 

in Arizona and New Mexico failed to be recognized and 

delineated because of map scale limitations based upon 

a 200-ha threshold that excluded these smaller isolated 

areas that were eliminated to maintain cartographic 

integrity and utility of the map product. These areas 

typically occurred at the edges of the National Forest 

System boundary.

Practical Application_______________

The categorization of grasslands into generalized 

vegetation types assists natural resource managers 

in understanding the geographic variability and 

spatial distribution across National Forest Lands in 

the Southwestern Region. This understanding will 

potentially lead to progressive management actions 

to maintain and restore these grasslands to ensure 

their ecological sustainability. 

References________________________

Bailey, R.G.; Avers, P.E.; McNabb, H.W. 1994. Ecoregions and subregions of the United States. Washington, DC: U.S. Department 

of Agriculture, Forest Service.

Barbour, M.G.; Billings, W.D. 2000. North American Terrestrial 

Vegetation. 2nd edition. Cambridge University Press. 708 p.

Brown, D.E. 1994. Biotic Communities: Southwestern United States 

and Northwestern Mexico. Salt Lake City, UT: University of 

Utah Press. 342 p.

Carleton, J.O.; Robbie, W.A.; Shaw, D.; Gass, J.; Brown, H.G. III; 

Spann, C.; Robertson, G.; Miller, G.; Popp-Galeano, R.; Fletcher, 

R.; Potter, D.U.; Moir, W.H.; Robison, T. 1991. General ecosystem 

survey. Albuquerque, NM: U.S. Department of Agriculture, Forest 

Service, Southwestern Region.

Dick-Peddie, W.A. 1993. New Mexico Vegetation: Past, Present 

and Future. Albuquerque, NM: University of New Mexico Press. 

244 p.

Fenneman, N.M. 1928. Physiographic divisions of the United States. 

Assoc. Amer. Geog. Annals. 18: 261-353.

Fisher, W. 2001. Oklahoma GAP analysis land cover map. Stillwater, 

OK: Oklahoma Cooperative Fish and Wildlife Research Unit.

Kuchler, A.W. 1964. Potential natural vegetation of the conterminous 

United States. Special Publication No. 36. American Geographical 

Society. 116 p.

Kuchler, A.W. 1970. Potential natural vegetation (map at scale 

1:7,500,000), pp. 90-91. In: The National Atlas of the USA. 

Washington, DC: U.S. Government Printing Office.

Laing, L.; Ambos, N.; Subirge, T.; McDonald, C.; Nelson, C.; 

Robbie, W. 1986. Terrestrial ecosystem survey of the ApacheSitgreaves National Forest. Albuquerque, NM: U.S. Department 

of Agriculture, Forest Service, Southwestern Region. 452 pp plus 

maps.

McNab, H.; Avers, P.A. 1994. Ecological subregions of the United 

States. Section descriptions. Washington, DC: U.S. Department 

of Agriculture, Forest Service.

Miller, G.; Ambos, N.; Boness, P.; Reyher, D.; Robertson, G.; Scalzone, 

K.; Steinke, R.; Subirge, T. 1995. Terrestrial ecosystem survey of 

the Coconino National Forest. Albuquerque, NM: U.S. Department 

of Agriculture, Forest Service, Southwestern Region. 405 pp 

plus maps.

Parker, N. 2001. Texas GAP analysis land cover map. Lubbock, TX: 

Texas Tech University.

Robertson, G.; Boness, P.; Gallegos, J.; Hurja, J.; Leahy, S.; Miller, 

G.; Robbie, W.; Scalzone, K.; Stein, R.; Steinke, R. 2000. Terrestrial 

ecosystem survey of the Prescott National Forest. Albuquerque, 

NM: USDA Forest Service, Southwestern Region. 724 pp plus 

maps.

Thomas, K. 2001 Arizona GAP analysis land cover map. Flagstaff, 

AZ: Northern Arizona University, Colorado Plateau Research 

Station.

Thompson, B.C.; Crist, P.; Prior-Magee, J.S; Deitner, R.A; Garber, 

D.L; Hughes, M.A. 1996. Gap analysis of biological diversity 

conservation in New Mexico using Geographic Information 

Systems. Las Cruces, NM: New Mexico Cooperative Fish and 

Wildlife Research Unit. 93 pp plus appendices.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 15

Figure 2-2.—Grasslands assessment category by National Forest and ecosection for Arizona.
16 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Figure 2-3.—Grasslands assessment category by National Forest and ecosection for New Mexico, Oklahoma, and Texas.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 17

Table 2-1. Crosswalk of grassland assessment categories, Arizona and New Mexico GAP landcover classes and general terrestrial 

ecosystem survey vegetation categories.

Grassland General terrestrial

 assessment Arizona GAP New Mexico GAP ecosystem survey

 categories landcover classes landcover classes vegetation classes

Montane Rocky Mountain-Great Basin Rocky Mountain Alpine Forb Tundra Kobresia myosuroides

 Dry Meadow Grasslands Festuca thurberi

Rocky Mountain Alpine Graminoid Tundra Festuca arizonica

 Grasslands Bromus anomalus

Rocky Mountain Subalpine and Montane Poa pratensis

 Grasslands

Colorado Plateau Great Basin Mixed Grass Great Basin Foothill-Piedmont Grassland Artemisia tridentata

Great Basin Grass-Mixed shrub Great Basin Lowland/Swale Grassland Bouteloua gracilis

Great Basin Grass-Mormon tea Shortgrass Steppe Hesperostipa neomexicana

Great Basin Grass- Saltbush Pleuraphis jamesii

Great Basin Riparian/Sacaton

 Grass scrub

Great Basin Riparian/Wet

 Mountain Meadow

Great Basin Sagebrush-Mixed

 Grass- Mixed Scrub

Great Basin Shadscale-Mixed 

 Grass-Mixed Scrub

Great Basin Semidesert Mixed Grass- Great Basin Foothills-Piedmont Grassland Bouteloua curtipendula

 mesquite Great Basin Lowland/ Swale Grassland

Semidesert Mixed Grass- mixed 

 scrub

Semidesert Mixed Grass- Yucca-

 Agave

Semidesert Tobosa Grass Scrub

Plains Not described Midgrass prairie Not described

Shortgrass steppe

Desert Sonoran Paloverde-Mixed Cacti Chihuahuan Foothill-Piedmont Desert Prosopis glandulosa

Semidesert Grassland Grassland Prosopis velutina

Chihuahuan Lowland/ Swale Desert

 Grassland
18 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Introduction_______________________

This chapter provides a brief overview, and selected 

in-depth coverage, of the factors and processes that 

have formed, and continue to shape, our Southwestern 

grasslands. In general, this chapter looks at how distributions of grasslands are regulated by soils and 

climate, and modified by disturbance (natural and/or 

anthropogenic). The attendant ecological components 

of grasslands will vary according to climate, soil, and 

other biotic factors including competition, predation, 

and mutualism. A shifting equilibrium typically exists between grasslands, deserts, and shrublands in 

the Southwest, such that changes in the severity or 

frequency of disturbance events (such as grazing, fire, 

or drought) can cause a change from one community 

type to another.

A problem of increasing importance to Southwestern land managers is that of exotic weed invasions. 

This chapter highlights the often-overlooked importance of fungi and soil crusts to grassland ecosystem 

function, and points out possible linkages to exotic 

weed invasions.

Grassland Evolution and 

Distribution_______________________

Humans have manipulated grassland vegetation for 

thousands of years through the use of fire, livestock 

grazing, and other means. Therefore, it can be difficult at times to separate our influence from that of 

nature in the formation and maintenance of grassland 

ecosystems. The extensive North American grasslands 

evolved in the Miocene and Pliocene, during a period 

of global climate change (Axelrod 1958). In the late 

Miocene, C4 dominated Southwestern grasslands 

expanded at the expense of C3 vegetation. CO2 levels 

had decreased prior to the Miocene, and C4 plants 

were more tolerant of the lower CO2 than C3 plants. 

Periods of increased aridity, caused in part by the 

uplift of the Asian land mass and changes in seasonal 

precipitation patterns, facilitated the rapid expansion 

of drought-tolerant grasses and forbs, while restricting 

the growth of forests and woodlands (Jacobs and others 

1999, Pagani and others 1999). Subsequently, natural 

and aborigine-caused fires swept across the grasslands 

at sufficient frequency to restrict the occurrence of 

trees and shrubs (Axelrod 1985, Dix 1964, Erickson 

2001). Regional climate change and the Industrial 

Era increase in atmospheric CO2 might have played 

a role in the current expansion of woody ecosystems 

into grasslands (Brown and others 1997).

Currently, fire helps to maintain the stability of 

grasslands by reducing the establishment of trees 

and shrubs. It also facilitates nutrient cycling by 

releasing nutrients from litter (Odum 1997) and accelerating the rate of decomposition in the soil. The 

reestablishment of periodic fire as a disturbance agent 

Paulette L. Ford

Deborah U. Potter

Rosemary Pendleton

Burton Pendleton

Wayne A. Robbie

Gerald J. Gottfried

Chapter 3:

Southwestern Grassland Ecology
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 19

can be fundamental to the ecological restoration of 

Southwestern grasslands. However, prior to proceeding with large-scale fire reintroduction, appropriate 

fire frequencies and season need to be determined 

for each grassland type. In general, the response 

of grasslands to fire seems to depend primarily on 

pre- and postfire levels of precipitation (Ford 1999). 

Therefore, the use of fire as a management tool in 

a drought year should be carefully considered and 

aligned with management goals.

An excellent resource to find out more information about fire effects on vegetation and wildlife 

in Southwestern grasslands is the Fire Effects 

Information System (FEIS) http://www.fs.fed.us/

database/feis. The FEIS database contains literature 

reviews taken from current English-language literature of almost 900 plant species, about 100 animal 

species, and 16 Kuchler plant communities found 

on the North American continent. The emphasis of 

each review is on fire and how it affects each species 

(FEIS 2004).

Central North American grasslands, including 

mixed-grass prairie, shortgrass steppe, and desert 

grasslands of the Southwestern United States 

and Mexico, are considered temperate grasslands. 

Temperate grasslands, broadly distributed between 

30° and 60° latitude, are generally characterized by 

(1) rainfall intermediate between temperate forest and 

desert, (2) a long dry season, (3) seasonal extremes of 

temperature (alternating long warm summers and 

short cold winters), (4) dominance of grasses, and (5) 

large grazing mammals and burrowing animals (Brown 

and Lomolino 1998, Lincoln and others 1998).

The periodic precipitation of temperate grasslands 

varies seasonally and annually. However, they generally average 250 to 750 mm of rain each year, though 

periods of drought are common and often prolonged. 

The rate of precipitation in temperate grasslands allows 

plants to release nutrients slowly into the ground over 

long periods. Dry temperate grasslands have maximum precipitation in the summer, and soil moisture 

is recharged by snowmelt in the spring. Precipitation 

in the arid and semiarid Southwestern grasslands is 

characterized by low rainfall, high evapotranspiration 

potential, low water yields (Branson and others 1981), 

and intermittent stream flow.

North American Plains grassland is represented 

mainly by mixed-grass prairie and shortgrass steppe 

in the Southwest. This largely midsummer flowering 

grassland extends from approximately 55° latitude in 

the Canadian Provinces of Alberta and Saskatchewan 

southward to below 30° latitude in Mexico, and once 

covered most of the American “Midwest” from the 

Eastern deciduous forest westward to the Rocky 

Mountains and beyond. More than 70 percent of the 

Plains grassland is now under cultivation (Garrison 

and others 1977). The plains grasslands developed 

under grazing by large herbivores and are generally 

tolerant of grazing (Engle and Bidwell 2000, Mack and 

Thompson 1982, Milchunas and others 1988).

Great Basin grassland merges with Plains grassland over a large transition area adjacent to the Rocky 

Plains Mesa/Desert grassland ecotone, Sandoval County, New Mexico. (Photo by Rosemary Pendleton)
20 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Mountains in Montana, Wyoming, Colorado, New 

Mexico, and Arizona. Much of this grassland has been 

converted to cultivated cropland through irrigation, 

and most of the remainder has experienced a degree 

of shrub expansion due to grazing and fire suppression 

(Brown 1994). The spring-flowering Great Basin or 

Intermountain grassland is restricted to those areas 

west of the Rockies and east of the Sierra-Cascades 

that possess favorable soils, climate, and grazing 

history. Much of this grassland has been appropriately described as a shrub-steppe in that pure-grass 

landscapes without shrubs are limited (Franklin and 

Dyrness 1973).

In the Southwest, warm temperate grasslands 

are represented by a semidesert grassland with a 

more or less biseasonal to summer precipitation 

pattern. Since the 1970s, populations of woody 

plants, leaf succulents, and cacti have expanded, 

replacing perennial grass cover (Brown 1994). Factors 

attributed to changes in woody plant cover include 

regional climate shifts, increases in CO2 concentrations, 

changes in fire frequency, and herbivory (Brown and 

others 1997, Detling 1988, Pagani and others 1999). 

Semidesert grassland adjoins and largely surrounds 

the Chihuahuan Desert, and with the possible exception of some Sonoran Desert areas in west central 

Arizona, it is largely a Chihuahuan, semidesert 

grassland. Extensive areas of this grassland occur 

in the Southwest in Chihuahua, western Coahuila, 

Trans-Pecos Texas, the southern half of New Mexico, 

southeast Arizona, and extreme northeastern Sonora 

(Brown 1994). Unlike the plains grasslands, desert 

grasslands were without megafaunal grazers for the 

last 10,000 to 11,000 years (Haynes 1991). Therefore the 

plant communities currently in place in the Southwest 

may be more susceptible to livestock grazing disturbance than other grasslands (Bock and Bock 1993, 

Loftin and others 2000, Mack and Thompson 1982). (For 

a comprehensive review of Southwestern grassland 

history and evolution see Van Devender 1995).

Soils and Climate__________________

Precipitation and temperature are the main parameters of climate and are important properties 

that strongly influence both soil function and plant 

growth. Soil moisture and temperature directly affect the nature and development of soils. Grassland 

ecosystems are generally considered highly productive 

due to the hydrological, biological, and geochemical 

cycling between soil properties, and the resulting outputs of mass and energy. Soils vary considerably for 

grasslands in the Southwestern Region. The different 

grassland types in the Southwest result from unique 

combinations of climate, soil, topography, and parent 

materials. Given the wide geographic range of these 

ecosystems, a high degree of inherent variability of 

climates, geology, landforms, and plant communities 

exists that directly influences rates of weathering, 

degree of stability, and site productivity.

Additions to, removals from, and vertical transfers 

and transformations within the soil are all basic kinds 

of soil-forming processes. These are, in turn, influenced 

by natural and anthropogenic disturbances, both at 

the soil surface and within the soil profile (Hendricks 

1985). Historic and current disturbances have had an 

effect on the rates to which soil-forming processes have 

taken place. They also affect the degree to which the 

grassland system maintains its resiliency and ability to sustain soil functions that reflect stability and 

productivity.

Desert Grasslands

At lower elevations of the Southwest, grasslands 

encompass the Chihuahuan and Sonoran Deserts and 

are characterized by an arid climate with limited precipitation. The seasonal distribution of rainfall differs 

between these two desert environments. Chihuahuan 

Desert rainfall is unimodal and receives the majority 

of moisture within the summer months during the 

monsoon season. Sonoran Desert rainfall is bimodal 

with two distinct periods of precipitation. One period 

is during the winter months in which half or more of 

the annual precipitation is received. The remaining 

rainfall is received during the summer months.

Soil temperature is an important property that has 

a strong influence on plant growth and soil formation. 

Two soil temperature regimes are recognized in desert 

grasslands that include thermic and hyperthermic 

Tulip prickly pear (Opuntia phaeacantha). (Photo by Rosemary 

Pendleton)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 21

classes. Generally the hyperthermic temperature 

regime occurs in Arizona where precipitation is 250 

mm or less, whereas the thermic temperature regime 

occurs in both Arizona and New Mexico desert grasslands where precipitation ranges from 250 to 410 mm 

(Hendricks 1985).

Landforms and parent materials of desert grasslands 

vary according to the degree of weathering, slope, relief, and mode of transportation of geologic materials. 

Soils of these landforms have evolved under paleoclimatic conditions and continue to change through 

the influences of today’s climate and disturbances. 

Fluvial erosion, deposition, and volcanic activity are 

the primary geomorphic processes responsible for the 

origin and development of landforms supporting desert 

grasslands.

Desert grasslands characteristically occur on gently 

sloping floodplains, lower alluvial fans, concave playas, 

bajadas, and nearly level valley plains. Some remnant 

desert grasslands occur on isolated terraces, mesas, and 

sideslopes of moderately steep and steep hills. Minor 

patches of desert riparian vegetation with grassland 

affinities occur along drainage ways and dissected 

alluvial fans. Typically, these landforms contain both 

igneous and sedimentary lithologies of alluvial origin, 

although areas within desert grasslands and adjacent 

desert shrublands have areas of eolian, wind-deposited 

features that yield undulating dunes.

Dominant soils in desert grasslands are classified 

as Aridisols. Vertisols are of minor extent for those 

desert grasslands occurring within playas and closed 

basin topography. Entisols are characteristic of valley 

plains and drainages where active fluvial processes are 

taking place. A number of diagnostic soil properties 

help differentiate these soil orders and influence the 

kind and amount of desert grassland vegetation and 

its location (Robertson and others 2000). The physical and chemical weathering of soils along with the 

atmospheric deposition of dust yields soluble salts and 

carbonates. Reaction (pH) of soils generally ranges 

from neutral to alkaline.

The interconnections of soil, climate, and vegetation 

relationships in an ecological framework are described 

by Carleton and others (1991). For instance, it is 

recognized that mesquite (Prospsis spp.) will occur 

as a shrub throughout the Chihuahuan Desert but is 

confined to the upper Sonoran Desert. Mesquite does 

occur in the lower Sonoran Desert as a tree on certain 

alluvial soils as a result of fluctuating water tables. 

The hyperthermic/aridic soil climate causes a lack of 

perennial grasses in the lower Sonoran Desert. The 

high concentration of calcium carbonate favors calciphillic plants such as creosotebush (Larrea tridentata

(Sessé & Moc. ex DC.) Coville) and crucifixion-thorn 

(Canotia holacantha Torr.) (Hendricks 1985). The presence of heavy clay surface horizons in playas supports 

rhizomotous plants such as tobosa (Pleuraphis mutica

Buckl.), whereas loamy soil textures are suitable for 

black grama (Bouteloua eriopoda (Torr.) Torr.), and 

sand soil textures are dominated by sandhill muhly 

(Muhlenbergia pungens Thurb.).

Major disturbances include fire, livestock grazing, 

recreational activities, and the introduction of exotic 

species. Composition and successional sequence of some 

desert grassland plant communities have changed due 

to the introduction of livestock grazing and exotic, 

predominantly annual, plant species. Recreational 

activities such as off-road vehicle use have altered 

the soils hydrologic function in certain areas. Climateinduced disturbances include lengthy droughts and 

flash floods during torrential summer monsoons.

Great Basin Grasslands

Soils of Great Basin grasslands differ from other 

grassland soils because of their unique combination of 

climate, landform, and vegetation. The soil climate of 

Great Basin grasslands ranges from an aridic to ustic 

moisture regime, and a thermic to mesic temperature 

regime; with combinations of ustic/thermic classes in 

central and southern Arizona (Robertson and others 

2000), and aridic/mesic classes in central and southern 

New Mexico.

The region typically receives less than 250 mm of 

precipitation per year. Mean monthly precipitation 

Black grama (Bouteloua eriopoda). (Photo by Rosemary 

Pendleton)
22 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

shows a strong winter-dominated pattern in the west, with a gradual shift 

eastward to more summer moisture, 

with less distinct wet and dry seasons 

compared with other deserts (Turner 

1994). Air temperatures are cool in the 

winter months and hot during the summer months. The soil climate is quite 

variable and is dependent upon the fluctuations in 

weather patterns throughout the Southwest. With this 

degree of variability, the resulting genesis of geology, 

landform, and soils is complex. Yet, for this assessment, 

it is recognized that the geographic location of these 

grasslands is confined to the higher elevations of the 

Basin and Range physiographic province (Fenneman 

1928): above desert grasslands and below or adjacent 

to pygmy conifer and evergreen oak woodlands, and 

therefore considered semiarid.

Volcanic and fluvial events are the primary geomorphic processes responsible for the origin and 

development of landforms supporting Great Basin 

grasslands. Landforms are typically nearly level 

elevated and lowland plains, gently sloping piedmont 

plains, and moderately steep uplands. These landforms 

vary in age and morphometric features, and they have 

experienced differing rates of erosion through anthropogenic and natural disturbances. These landforms 

also create a rainshadow effect along the frontal ranges 

of mountainous areas bordering basins and valley 

floors (Dick-Peddie 1993). 

Parent materials are derived 

from igneous to sedimentary 

sources. Mixed alluvium 

occurs in fan and piedmont 

positions and valley plains, 

whereas colluvium and residuum parent materials are 

dominant on elevated plains and upland landforms.

Soils are classified as Typic Haplustalfs, mesic or 

Typic Argiustolls, mesic for uplands (Carleton and 

others 1991). These soils are moderately deep to 

deep, with loam surfaces, and support blue grama 

(Bouteloua gracilis (Willd. ex Kunth) Lag. ex Griffiths), 

New Mexico Feathergrass (Hesperostipa neomexicana

(Thurb. ex Coult.) Barkworth), and Needle-and-Thread 

(Hesperostipa comata (Trin. & Rupr.) Barkworth). 

Aridic Haplustalfs, thermic, and Aridic Argiustolls, 

thermic, are sightly drier and warmer soils that support blue grama (Bouteloua gracilis), black grama 

(Bouteloua eriopoda), and Threeawns (Aristida spp.). 

Sideoats grama (Bouteloua curtipendula (Michx.) 

Torr.), which occur extensively throughout this grassland and can be in association with emory oak (Quercus

emoryi Torr.) at adjacent ecotones in southern Arizona 

and gray oak (Quercus grisea Liebm.) in southern 

New Mexico. Some inclusions of Typic Ustipsamments 

with sandy surface textures support little bluestem 

(Schizachyrium scoparium (Michx.) Nash). Another 

Indian ricegrass (Achnatherum hymenoides). 

(Photo by Rosemary Pendleton)

Needle-and-thread grass (Hesperostipa comata). (Photo by Rosemary 

Pendleton)

Blue grama (Bouteloua 

gracilis). (Photo by Rosemary Pendleton)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 23

Great Basin grassland indicator is Soaptree yucca 

(Yucca elata (Engelm.) Engelm.) (Dick-Peddie 1993). 

Western wheatgrass (Pascopyrum smithii (Rydb.) 

A. Löve) can occur on Vertic Haplustalfs and Vertic 

Argiustolls on nearly level lowlands and concave 

swales (Laing and others 1987).

The effects of grazing, in combination with changes 

in climate over time, and the absence of natural fire, 

have resulted in encroachment of woody shrubs and 

trees into the Great Basin grasslands. Other disturbances include the increase in abundance of exotic weed 

species and the resultant change in fire regimes.

Colorado Plateau Grassland

Soils of Colorado Plateau grasslands of this assessment are characterized by an ustic bordering on xeric 

soil moisture regime and a mesic soil temperature 

regime. Typically, over half the precipitation occurs 

during the winter months in the form of snow and 

rainfall. However, some areas receive more summer 

than winter moisture. Air temperature is generally cold 

during the winter and hot during the summer.

The geographic range of these grasslands is limited 

to the lower elevations of the Colorado Plateau and 

Southern Rocky Mountain physiographic provinces 

(Fenneman 1928, Laing and others 1987, Miller and 

others 1995). These grasslands are an integrade to 

adjacent cold desert sagebrush (Artemisia spp.) or 

four-wing saltbush (Atriplex canescens (Pursh) Nutt.) 

steppe communities and higher pinyon (Pinus edulis

Engelm.) -juniper (Juniperus spp.) woodland ecosystems (Brewer and others 1991).

The primary geomorphic processes responsible 

for the origin and development of landforms of the 

Colorado Plateau grasslands are fluvial and volcanic. 

Landforms are nearly level to flat mesas, plateaus, 

and rolling elevated plains. Parent materials are dominated by basaltic igneous and limestone or sandstone 

sedimentary sources. Mixed alluvium parent materials 

support deep soils in narrow valley plains, whereas 

residuum from both igneous and sedimentary sources 

support shallow to moderately deep soils in steeper 

landscape positions.

Aridic Ustochrepts, mesic, calcareous are moderately deep soils with sandy loam surface horizons. 

Subsurface horizons are fine-loamy textured or skeletal 

and are generally calcareous in the lower part of the 

profile. These soils occur on upland plains derived 

from limestone and support stands of big sagebrush 

(Artemisia tridentata Nutt.), fourwing saltbush 

(Atriplex canescens), winterfat (Krascheninnikovia 

lanata (Pursh) A.D.J. Meeuse & Smit), Needle-andthread (Hesperostipa comata), and Indian ricegrass 

(Achnatherum hymenoides (Roemer & J.A. Schultes) 

Barkworth). Typic and Pachic Argiustolls, mesic are 

deep soils with loam to clay loam surfaces that occur 

on valley plains and lowlands derived from mixed 

alluvium. Dominant vegetation for these areas includes 

western wheatgrass (Pascopyrum smithii) and blue 

grama (Bouteloua gracilis).

The major disturbance of the Colorado Plateau 

grasslands is ungulate grazing. Composition and 

successional sequence of grassland plant communities 

have changed as a result of grazing. Climate induced 

disturbances include drought and flash floods during 

summer monsoon storms. Strong winds are common 

during the spring.

Plains Grasslands

The Plains Grasslands are within the Great Plains 

physiographic province (Fenneman 1928). Soils have 

ustic soil moisture regimes for shortgrass and midgrass 

communities and ustic bordering udic soil moisture 

regimes for tallgrass communities. Soil temperature 

is mesic for those areas of the plains that encompass 

the National Grasslands in eastern New Mexico, 

Oklahoma, and Texas. Climate ranges significantly 

in this region from semiarid to subhumid, which has 

a direct effect on the distribution of vegetation, along 

with the origin and genesis of the soils.

Pachic Argiustolls and Haplustolls occur on nearly 

level to gently sloping uplands and plains that are 

formed in aluvium, sandstone, and shale. These soils 

support mid- and tallgrass species. Typic Ustochrepts, 

Calciorthidic Paleustolls, and Psammentic Haplustalfs 

occur on sandy and loamy, calcareous recent alluvium 

along stream terraces, outwash plains and eolian deposits (Burgess and others 1963, Murphy and others 1960). 

The depth to a calcareous substrate varies according 

to the degree of weathering and precipitation. Stands 

of tall- and midgrasses, shin oak (Quercus havardii

Rydb.), and sand sagebrush (Artemisia filifolia Torr.) 

frequently occur on these soils. Typic Ustifluvents 

are located along drainage ways, flood plains, and 

stream courses that are derived from recent alluvium. 

Fluventic Haplustolls that support tallgrass species 

occur in association with riparian hardwoods and 

wetland plants. Historically, the plains grasslands have 

evolved over time with repeated ungulate grazing in 

combination with natural fire.

Montane Grasslands

Soils of Montane Grasslands have ustic to udic 

soil moisture regimes, and temperature regimes 

that include frigid, cryic and pergelic. The Montane 

Grassland assessment category includes those environments characterized by montane, subalpine, and 

alpine ecosystems. Climates are extreme with mean 

annual air temperature ranging from –3 to 7 °C; mean 

annual precipitation ranges from 56 to 76 cm, of which 

over 50 percent is received during the months October 
24 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

through March. A significant portion of precipitation 

is snowfall.

The origin and development of landforms supporting 

montane grasslands are the result of glacial, fluvial, 

and volcanic geomorphic processes. Landforms associated with these grasslands include lowlands, upland 

plains, mountain slopes, and summit plains.

Typical Montane grasslands are characterized by 

Arizona fescue (Festuca arizonica Vasey) meadows on 

elevated plains of basaltic and sandstone residual, and 

alluvial parent materials that yield deep, clay loam, 

Typic, and Pachic Argiborolls (Brewer and others 1991, 

Laing and others 1987, Miller and others 1993, 1995). 

Other landform positions within this zone include 

valley plains of Kentucky bluegrass (Poa pratensis

L.) supported by deep, loamy, Fluventic Haploborolls 

adjacent to riparian corridors of narrowleaf cottonwoods (Populus angustifolia James) growing in deep, 

sandy loam, Aquic Ustifluvents, frigid.

The Thurber fescue (Festuca thurberi Vasey) 

subalpine grasslands are higher in elevation than 

typical montane grasslands (Moir 1967). This environment is somewhat colder and wetter with a 

slightly shorter growing season. These grasslands are 

treeless expanses that border subalpine forests that 

are characterized by Subalpine fir (Abies lasiocarpa

(Hook.) Nutt.), Engelmann spruce (Picea engelmanii

Parry ex Engelm.), and aspen (Populous tremuloides

Michx.). These meadows and grasslands occur on 

mountain sideslopes, elevated plains and valley 

plains derived from dominantly igneous with inclusions of sedimentary parent materials. Argic Pachic 

Cryoborolls and Haploborolls are soils that contain 

deep, dark loamy surface horizons (Miller and others 

1993). Surface rock varies; however, internal coarse 

fragments are numerous, thereby classifying these 

soils as skeletal.

Alpine tundra ecosystems are located above timberline, where a harsh climate of extreme cold and wet 

conditions creates wind-swept treeless expanses along 

mountaintop positions. These plant communities are 

limited in the Southwest and occur at elevations above 

11,500 feet. The growing season for this environment 

is less than 90 days, approximately June through 

August. Strong winds, high intensity solar radiation 

and severe evapotranspiration are characteristic of 

alpine tundra environments. Soils are dominantly 

Pergelic Cryumbrepts, which contain dark surface 

layers that have a low base saturation that does not 

meet the mollic epipedon criteria (Miller and others 

1993, 1995). Shallow to moderately deep with very 

cobbly sandy loam surface horizons, these soils are 

susceptible to erosion from wind and water, and human trampling. The landforms include nearly level 

to steep summit plains and mountain slopes.

Soil Biota__________________________

A variety of small organisms exist in grassland soils 

including bacteria, actinomycetes, fungi, algae, nematodes, micro- and macroarthropod invertebrates, and 

larvae. These soil organisms play a vital role in maintaining Southwestern grassland function. Although the 

majority of species involved remain undescribed, soil 

microflora and fauna have profound effects on essential 

ecosystem processes such as decomposition, nutrient 

cycling, and the maintenance of soil fertility (Adams 

and Wall 2000).

Soil microorganisms are essential components of the 

biogeochemical cycles that sustain life (Wolters and 

others 2000). Below- and aboveground communities 

are inextricably linked through complex interactions and feedback mechanisms. Any disturbance or 

change in the environment that affects aboveground 

vegetation will also affect the soil biota. The reverse 

is equally true. For example, an increase in CO2

concentrations above North American grasslands 

could disproportionately increase photosynthesis 

relative to decomposition, resulting in a buildup of 

organic matter in grassland soils (Swift and others 

1998). This buildup would, in turn, cause functional 

shifts in the soil community as nutrient turnover 

rates declined.

Changes in climate (precipitation patterns or 

temperature) directly affect soil processes. An environmentally induced decline in the types of organisms 

capable of degrading complex organic molecules such 

as lignin or chitin could result in a buildup of surface 

litter and a reduction in the nutrients available for 

plant growth, thus effecting changes in the surface 

vegetation (Schimel and Gulledge 1998). Fortunately, 

soil organisms have a high level of functional redundancy that could help buffer many of the effects of 

global change (Andren and others 1995). However, some 

key processes—including nitrogen transformations, 

nitrogen fixation, and the breakdown of recalcitrant 

molecules—are controlled by unique or specialized organisms (Wolters and others 2000). Changes affecting 

these taxa may have profound effects on ecosystem 

functioning.

Above- and belowground organisms interact in a 

variety of ways. For example, (1) microorganisms and 

soil invertebrates are responsible for the breakdown 

of complex organic material into plant-available nutrients; (2) cyanobacteria, actinomycetes, and other 

rhizobacteria fix atmospheric nitrogen, thereby increasing soil fertility; (3) cell material and excretions 

from soil microorganisms affect the formation and 

stability of soil aggregates that, in turn, affect water 

and air movement within the soil; and(4) mutualistic 

relationships formed with mycorrhizal fungi affect a 

plant’s ability to grow and compete. Consequently, 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 25

changes that affect composition of the aboveground 

grassland community (that is, changes in climate, land 

use, atmospheric composition, or the introduction of 

invasive species) will likely affect the belowground 

community as well (Wolters and others 2000).

Bacteria, Actinomycetes, Fungi, and Algae

Bacteria are by far the most abundant group of soil 

microorganisms in terms of numbers, yet comprise 

less than half the microbial biomass because of their 

small size (Alexander 1977). Bacteria function in the 

decomposition of organic matter and in transformation 

and availability of many essential minerals. Several 

genera of free-living bacteria, such as Azotobacter, 

are capable of fixing atmospheric nitrogen. Vast 

communities of bacteria live on or near plant root 

surfaces where they feed on root secretions and dead 

cell material (Alexander 1977). Members of the genus 

Rhizobium form symbiotic relationships with roots of 

leguminous plants. Colonization by rhizobia results 

in the formation of root nodules, where they fix appreciable amounts of atmospheric nitrogen that is 

then available to the plant. Common legumes found on 

arid grasslands include the lupines (Lupinus spp.) and 

milkvetch or locoweed (Astragalus spp.). The woody 

mesquites (Prosopis spp.) also form nitrogen-fixing 

nodules (Geesing and others 2000).

Actinomycetes comprise the second most abundant 

class of soil microorganisms. They are extremely tolerant of desiccation and are found in large numbers in 

the grassland and steppe vegetation types common to 

the Southwest (Alexander 1977). Actinomycetes can 

utilize a variety of complex organic molecules including 

chitin and cellulose as energy sources. Many produce 

antibiotics and may be important in regulating the 

composition of the soil community. Members of the 

genus Frankia are capable of fixing nitrogen and form 

symbiotic root nodules with certain non-leguminous 

woody plants. Many of these species also form mycorrhizal associations (Rose 1980). These species are not 

common to grasslands, however.

Fungi account for large proportion of soil microbial 

biomass due to the extensive network of filaments 

(Alexander 1977). They function largely in the breakdown of complex organic molecules including lignin, a 

compound that is resistant to bacterial degradation. 

Some classes of fungi form specialized associations 

with plant roots. In addition to arbuscular mycorrhizal fungi, discussed in detail later in this chapter, 

Barrow and McCaslin (1996) observed two other major 

classes of fungi in roots of Southwestern grasses. 

Roots were commonly colonized by septate fungi that 

formed nondestructive interfaces within and among 

cortical cells, forming mycorrhizal-like associations. 

Chytridiomycete fungi were also commonly observed 

in plant roots. The function of chytrid fungi in the 

soil ecosystem is unknown, but they may play a role 

in regulating mycorrhizal colonization and nutrient 

uptake (Barrow and McCaslin 1996).

Soil algae are found in every conceivable location 

throughout the world, from the arctic tundra to thermal 

springs to deep within commercial caves, anywhere that 

light is present. Algae are often the primary colonizers 

following volcanic eruptions (Alexander 1977). The role 

of green algae and cyanobacteria (blue-green algae) 

in fixing carbon compounds and stabilizing soils is 

addressed later in this chapter. Another algal group 

commonly found in soils is the diatoms. These algae 

have an outer wall that is highly silicified. Although 

small, the striking beauty of their regular geometric 

shapes makes them a fascinating group to study. As 

with other kinds of algae, most diatoms are obligate 

photoautotrophs (requiring light) and live in the upper soil surface where light can penetrate. They are 

most prevalent in neutral to alkaline soils (Alexander 

1977).

Soil Invertebrates

Many macroarthropods spend part of their life 

cycle, usually as larvae, in the soil, and many microarthropods, such as mites (Acari) and springtails 

(Collembola), spend virtually all their lives there. 

Soil- and litter-inhabiting mites and nematodes (Zak 

and Freckman 1991) and collembolans (springtails) 

(Crawford 1990) occur in vast numbers and are species-rich in nearly all Southwestern habitats. The soil 

mites (Acari) are the smallest of the arachnids, the 

group that contains spiders and their relatives. Some 

mites are large enough to be visible to the unaided eye 

(about the size of grains of ground pepper), but not large 

enough to allow determination of their morphological 

characteristics. The mite fauna in desert grassland 

soils include more than 30 families and 100 species 

(Cepeda-Pizarro and Whitford 1989). The most abundant are the generalist microbe feeders such as the 

nanorchestid and tydeid mites. Some mite species feed 

on nematodes as well as fungi, yeasts, and bacteria. 

Included in this fauna are a variety of predators that 

capture and eat other mites and nematodes (Whitford 

and others 1995).

Nematodes represent another group of extremely 

abundant microscopic soil animals. There are approximately 100,000 bacteria feeders, 1,000 fungus 

feeders, 10,000 omnivore-predators, and 10 to 1,000 

plant root feeders (root parasites) per square meter 

of soil (Freckman and others 1987). Nematodes are 

aquatic animals confined to single-molecule-thick 

water films surrounding soil particles, although they 

are well adapted to deal with dehydration. As the 

water films disappear in a drying soil, the nematodes 

enter an inactive state called anhydrobiosis, which 

is immediately reversible when the soil is wetted 
26 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

(Whitford and others 1995). Studies have indicated 

that nematodes play a substantial role in the cycling of 

carbon and nitrogen in the soil environment (Bongers 

and Ferris 1999).

Protists (single-celled organisms that feed on bacteria and algae) populate the soil of all desert grasslands, 

but practically nothing is known of their distribution 

and biology. Naked amoebae predominate, along with 

smaller numbers of other orders. The abundance of 

protists in desert soils is astounding: there are 25,000 

naked amoebae, 4,900 flagellates, and 700 ciliates in 

every gram of dry soil (Parker and others 1984). These 

numbers are deceiving, however, because most of the 

protists in dry soil are encysted (in an inactive physiological state; Whitford 1989). After a rain they quickly 

resume activity and remain active until the soil dries 

and they once again encyst (Whitford and others 1995). 

Grazing by protists on microbes stimulates the rate of 

decomposition of organic matter (Finlay 2001).

Grassland Invertebrates_____________

Invertebrates perform vital ecological functions in 

grassland ecosystems. The scope of their contributions 

includes soil aeration, seed dispersal, and plant pollination and consumption, in addition to facilitating 

the decomposition of organic debris (such as dung 

and animal remains). Invertebrates also provide an 

important prey base for grassland fish and wildlife. (See 

volume 2 for more information on grassland wildlife, 

small mammals, birds, and fish.) This section gives 

a brief overview of the roles of some of the common 

invertebrates in Southwestern grasslands.

The importance of invertebrate animals to ecosystem 

properties and processes is frequently underemphasized (Whitford and others 1995). When the indirect 

effects of invertebrates on nutrient cycling processes 

and their direct and indirect effects on soil heterogeneity are considered, however, their importance becomes 

more evident. In fact, ecosystem properties such as 

resilience (the ability to recover following disturbance) 

are directly affected by the activities of key invertebrate 

species (Whitford and others 1995).

The invertebrate fauna of desert grasslands, and 

Southwestern rangelands in general, is incredibly 

diverse and includes several phyla (Parmenter and others 1995, Whitford and others 1995). While mammal, 

bird, reptile, and vascular plant species occur in the 

tens to hundreds, invertebrate species in desert grasslands number in the thousands or tens of thousands. 

Many of the less conspicuous species have never been 

described by taxonomists (Whitford and others 1995), 

and little is known about the diversity of arthropods on 

Southwestern rangelands. The available data indicate 

that species diversity for most groups of rangeland 

arthropods is higher in the Southwest than in other 

parts of the country (Parmenter and others 1995).

Most of what we know about desert grassland invertebrates and their general life history characteristics 

is based on data from economically important species 

(Crawford 1981, Whitford and others 1995). Insects of 

the Southwestern rangelands are often thought of as 

agricultural pests because of the economically costly 

forage consumption by some species. Good reviews of 

important rangeland insect pests and research on those 

insects are found in Capinera (1987), Watts and others 

(1982, 1989), and Parmenter and others (1995). The pest 

species that are included in the above-cited literature 

represent only a small fraction of the insects and other 

arthropods that occur on Southwestern rangelands. 

Most species are not agricultural pests, many are rare, 

and many perform important functions in rangeland 

ecosystems (Crawford 1981, 1986, Lightfoot and 

Whitford 1990, MacKay 1991, Parmenter and others 

1995, Walter 1987, Zak and Freckman 1991).

Grasshoppers (Orthoptera)

Many different species of plant-feeding insects occur 

on Southwestern rangelands (Crawford 1981, Watts 

and others 1989, Wisdom 1991). Of these, grasshoppers are among the most prevalent and conspicuous. 

A considerable amount of research has been conducted 

on grasshoppers throughout the Southwest, and more 

is known about the diversity and biology of grasshoppers than about other rangeland plant-feeding insects 

(Parmenter and others 1995). In North America, grasshopper species diversity is highest in the Southwest. 

Otte (1981) demonstrates that species densities of 

slant-faced grasshoppers (Gomphocerinae, primarily 

grass-inhabiting and -feeding grasshoppers) average 

Darkling beetle (Eleodes longicollis). (Photo by David C. 

Lightfoot)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 27

around 30 species for locations in the Southwest, 

compared to five to 20 species for most of the rest of 

North America (Parmenter and others 1995).

Hewitt (1977) and Hewitt and others (1976) stated 

that grasshoppers compete with cattle for forage and 

that both herbivores have similar preferences in 

grass species. Estimates of the biomass consumption 

on rangelands in the Western United States by grasshopper herbivory have been difficult to calculate, but 

typically range from 6 to 12 percent of available forage, 

to as much as 50 percent in certain areas (Loftin and 

others 2000). In grassland at the San Carlos Indian 

Reservation in Arizona, Nerney (1961) found that 

grasshopper consumption ranged from 8 to 63 percent of the vegetation. However, Hewitt (1977) also 

reported a study from southeastern Arizona showing 

no correlation between grasshopper density and loss 

of forage. Quinn and others (1993) noted that as yet 

there is no measure of economic injury from grasshoppers that can be applied across all species or habitats 

(Loftin and others 2000).

Beetles (Coleoptera), Ants (Hymenoptera), 

Termites (Isoptera), and Spiders (Araneida)

The order Coleoptera contains a third of all known 

insects—approximately 300,000 species worldwide and 

approximately 30,000 species in North America (Milne 

and Milne 1996). Beetles are conspicuous components of 

the terrestrial invertebrate fauna of arid and semiarid 

ecosystems of North America (Allsopp 1980). They have 

an important role in the functioning of Southwestern 

grasslands, serving as predators, prey, scavengers, and 

parasites. In addition, many plant-feeding species attack plants and stored foods, while others pollinate 

flowers and eat plant pests.

Beetles appear to be relatively diverse in Southwestern grasslands. For example, the tiger beetle 

(Cicindelidae) is found worldwide, but in North America, 

the Rocky Mountains and the Great Plains contain the 

highest numbers of species, 15 to 20, compared with 10 

for New England and 15 for the Middle Atlantic States 

(Parmenter and others 1995, Pearson and Cassola 1992). 

Ford (2001) found an average of eight beetle species per 

hectare on unburned, ungrazed shortgrass steppe in 

the southern Great Plains of New Mexico. Burned areas 

contained an average of 11 beetle species per hectare 

(Ford 2001). Darkling beetles (Tenebrionidae) are more 

diverse in Western arid lands than elsewhere in North 

America and are major detritivores in the Southwest 

(Crawford 1990).

African dung beetles have been recently introduced 

into grasslands throughout the United States to speed 

the decomposition of livestock fecal pats, primarily 

with the goal of reducing numbers of nuisance and 

disease-transmitting insects that breed in the dung 

(Dymock 1993; Hoebeke and Beucke 1997). While the 

grasslands of the Central and Southwestern United 

States contain fecal pats, the overall habitat conditions 

are too hot and dry to support these species (Loftin 

and others 2000).

Ants originated in tropical areas and spread into 

temperate habitats. Many of the species found in the 

Western United States are not unique to the region 

(Holldobler and Wilson 1990, Parmenter and others 

1995). Ants are the dominant arthropod predators (on 

other arthropods and on plant seeds) in some ecosystems (Holldobler and Wilson 1990). In some areas of 

the Chihuahuan Desert there may be as many as 4,000 

ant colonies per hectare. Arid regions of the Southwest 

Regal grasshopper (Melanoplus regalus). (Photo by David C. 

Lightfoot)

Tarantula (Aphonopelma sp.). (Photo by David C. Lightfoot)
28 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

often contain a diverse ant fauna, with 23 to 60 species 

(MacKay 1991). For example, 23 to 50 species occur 

in the Chihuahuan Desert (Loftin and others 2000, 

MacKay 1991). Among the more conspicuous species 

are harvester ants in the genera Pogonomyrmex and 

Aphaenogaster. These species have large long-lived 

colonies (a decade or more) and contribute to soil 

mixing and aeration (MacKay 1991) and to seed dispersal (Loftin and others 2000).

Although termites occur primarily in the tropics, the 

species that are successful in temperate regions are of 

great importance in the recycling of nutrients in dead 

grass and wood (Loftin and others 2000, MacKay 1991, 

Schaefer and Whitford 1981). To avoid desiccation, 

temperate species are primarily subterranean, and 

their activity periods coincide with warm temperatures 

and an area’s rainy season (Loftin and others 2000, 

MacKay and others 1989, ).

Termites have low species richness (up to a dozen 

species in the Southwestern United States), but may 

be the greatest regional consumers of net primary production (MacKay 1991, Parmenter and others 1995). 

Bodine and Ueckert (1975) estimated that termites 

removed 20 to 50 percent of grass and plant litter 

from a grassland in Texas (Loftin and others 2000). 

Termites’ positive contributions include nutrient cycling and nitrogen enrichment of the soils in arid and 

semiarid regions (Loftin and others 2000). Termites 

are known to consume cattle fecal pats (MacKay and 

others 1989).

Spiders form a major part of the arthropod fauna 

of the Southwest, but, as is the case with many other 

arthropod groups, the total number of species in the 

region is still unknown (Gertsh 1979). In a recent review 

of the status of arthropod systematics, Schaefer and 

Kosztarab (1991) estimate that most of the United 

States species of arachnids (and insects) that are 

still undescribed occur in the desert and montane 

Southwest and Great Basin areas (Parmenter and 

others 1995). Spiders serve as both predator and prey 

in Southwestern grassland ecosystems.

Mycorrhizal Fungi__________________

The majority of plants in grassland ecosystems form 

some kind of mutualistic relationship with mycorrhizal 

fungi (Allen 1991, Miller 1987). In the relationship, soil 

resources are provided to the host plant root in exchange 

for energy-containing fixed-carbon compounds for the 

fungus. The fungus acts as an extension of the host 

root system. For example, 1 ml of soil may contain 

2 to 4 cm of root, 1 to 2 cm of root hair, and 50 m of 

mycorrhizal fungal filaments, or hyphae (Allen 1991). 

Most external hyphae are concentrated near the plant 

root, but may extend 4 to 7 cm from the root surface 

(Read 1992, Rhodes and Gerdemann 1975). The small 

diameter of the hyphae allows the fungus to penetrate 

small pores in the soil and efficiently extract mineral 

nutrients such as nitrogen, phosphorus, potassium, 

calcium, sulfur, copper, and zinc (Allen 1991, Stribley 

1987).

In addition to uptake of minerals, mycorrhizae 

have been shown to improve plant water relations 

and soil structure (Allen and Allen 1986, Ames and 

Bethlenfalvay 1987, Mathur and Vyas 2000, Miller and 

Jastrow 1994, Thomas and others 1993), and to reduce 

susceptibility to pathogens and nematodes (Habte and 

others 1999, Newsham and others 1995). The result is 

an increased ability for the host plant to survive and 

grow under stressful environmental conditions.

Mycorrhizal relationships are classified according 

to the morphology of the root/fungus interface and 

vary depending on the species of plants and fungi 

involved. In grasslands, the vast majority of mycorrhizas are of the arbuscular mycorrhizal type, formed 

with fungi of the order Glomales (Allen 1991, Miller 

1987). Arbuscular mycorrhizae are so named for the 

internal fungal structure where nutrient exchange 

takes place. Initial fungal colonization can occur 

through germination of soil-borne spores, or through 

pieces of hyphae or infected root fragments. A hyphal 

network then develops between cells of the root cortex. 

Hyphae enter the cell wall of some cortical cells and 

branch dichotomously to form a profusely branched 

surface, known as an arbuscule, surrounded by the 

host cell membrane. Nutrient exchange takes place 

across this interface (Allen 1991, Bonfante-Fasolo and 

Scannerini 1992, Bowen 1987).

Other types of mycorrhizal associations may be 

found in isolated patches within the grasslands, or at 

the ecotone between grasslands and other vegetation 

types (Allen 1991, Trappe 1987). Ectomycorrhizal 

associations, formed between coniferous plants and 

basidiomycete or ascomycete fungi, are found at the 

periphery of montane grasslands. Ericaceous plants 

(for example, Vaccinium and Gaultheria), which form 

their own kind of ectendomycorrhizae, are found in 

the understory of coniferous forests and on rocky 

uplands within or adjacent to montane meadows. 

Willows present within the meadow may form either 

ecto- or arbuscular mycorrhizas. Orchids and saprophytes, found in a wide variety of habitats, also form 

a unique type of mycorrhizae.

The boundary between arid grasslands and other 

vegetation types is less distinct in terms of mycorrhizal associations. Desert shrubland plants, including 

Larrea and Prosopis, are also primarily arbuscular 

mycorrhizal (Staffeldt and Vogt 1975, Titus and others 2001, Virginia and others 1986). Juniper species, 

found at the upper boundary of arid grasslands, also 

form arbuscular mycorrhizae (Klopatek and Klopatek 

1997, Lindsey 1984), facilitating their expansion into 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 29

grass-dominated areas. Expansion of pinyon, which is 

ectomycorrhizal (Acsai 1989, Klopatek and Klopatek 

1997), most likely depends upon the availability and 

dispersal of ectomycorrhizal inoculum.

Distribution and Occurrence

Arbuscular mycorrhizal fungi are found throughout 

the world, the same species often occurring on multiple 

continents. The fungi are grouped into three families 

and either five or six genera, based on wall characteristics of the soil-borne spores (Morton and Benny 

1990). Associations between fungus and host plant 

appear to be nonspecific; a single root system may host 

associations with multiple species of fungi, and fungal 

networks may extend between adjacent mycotrophic 

plants. The fungi themselves are obligately mutualistic, 

and attempts to culture them separately have failed. 

Some species of fungi do appear to be habitat-specific, 

being associated with certain soil textures, nutrient 

levels, or extremes in pH (for example, Glomus diaphanum Morton & Walker and G. spurcum Walker 

& Pfeiffer), while others are common in a variety of 

habitats throughout the world (for example, Glomus 

etunicatum Becker & Gerdemann) (J. Morton, personal communication, Tatsuhiro and others 2000). 

Widespread or common fungal species may comprise 

different physiological ecotypes, however (Allen and 

others 1995).

Plant species differ in their ability to form mycorrhizal associations and in the amount of benefit 

derived. The degree of dependence on mycorrhizal 

fungi is correlated with the fineness of the plant root 

system and the number and length of root hairs per 

unit length (Baylis 1975, Hetrick and others 1992). 

Colonizing annuals in advanced families, such as the 

chenopod, mustard, and amaranth families, rarely 

form mycorrhizal associations and are referred to as 

“nonmycotrophic” (Allen 1991). Reeds and sedges in 

inundated soils also rarely form mycorrhizae; however, 

associations may form later in the season as soils dry 

out (Allen and others 1987, Miller and Sharitz 2000, 

Rickerl and others 1994).

Most grassland species are facultative mycotrophs 

and can survive with or without the association. They 

range from species that show no positive response to 

inoculation with mycorrhizal fungi, to those that show 

dramatic increases in growth and reproduction. The 

degree of response for any given plant depends on 

a number of factors, including the fungal species 

involved (fungal symbionts differ in their ability to 

acquire resources), soil texture and nutrient levels, 

environmental growing conditions such as temperature, precipitation, and light, and biotic factors such 

as competition and herbivory (Abbott and Gazey 1994, 

Cade-Menun and others 1991, Frey and Ellis 1997, 

Hetrick and Bloom 1984, Hetrick and others 1986, 

Koide 1991, van der Heijden and others 1998).

In general, C4 or warm-season grasses are more 

dependent on mycorrhizae than are C3 or cool-season 

grasses (Hetrick and others 1990, Van Auken and Bloom 

1998, Wilson and Hartnett 1998). Warm-season grasses 

alter their root morphology in response to colonization 

(Hetrick and others 1991, 1994), a trait associated with 

mycorrhizal dependency. Some grass species (such as 

Andropogon gerardii Vitman) are highly dependent 

on mycorrhizae for normal growth and reproduction 

(Hetrick and others 1989). There are some indications 

that perennial grass species may be more responsive to 

mycorrhizae than annuals (Boerner 1992). Grassland 

forbs and shrubs also vary in mycorrhizal dependence, 

with genera such as Linum, Sphaeralcea, and Artemisia

consistently showing a positive response to mycorrhizal 

fungi (Lindsey 1984, R. Pendleton, data on file, Rocky 

Mountain Research Station, Albuquerque, NM). Few 

Southwestern grassland species could be considered 

truly obligately mycorrhizal, that is, unable to grow 

and reproduce without the symbiosis.

Community Interactions

Considerable recent work supports the idea that mycorrhizae are involved in regulating plant community 

interactions by (1) increasing the fitness of mycotrophic 

plant species, (2) affecting the outcome of competitive 

interactions among species, and (3) connecting plant 

root systems via hyphal networks.

Most studies on the beneficial effects of mycorrhizae 

measure plant growth or biomass as the dependent 

variable. Plant fitness, however, is measured by the 

number and quality of offspring produced. Production 

of biomass is important only as it contributes to the 

plant’s ability to survive and reproduce. Therefore, 

although plant growth effects in field studies are often 

negligible, the true benefit of mycorrhizal fungi may rest 

in the enhanced ability of a plant to survive “ecological 

crunch” periods of acute stress (Allen and Allen 1986, 

Fitter 1986, 1989, Trent and others 1993). Such periods 

may include the vulnerable seedling stage, periods of 

short-term drought, or episodic outbreaks of insects or 

disease. The presence of mycorrhizal fungi has also been 

shown to affect the reproductive capacity of a plant, 

influencing the timing of reproduction, quantity of seed 

produced, and quality (size and competitive ability) of 

the resultant offspring (Koide and Lu 1992, Lewis and 

Koide 1990, Shumway and Koide 1995).

The presence of mycorrhizal fungi can alter the 

outcome of interspecies competition among plants. 

This has been demonstrated most conclusively in 

competition experiments between mycotrophic and 

nonmycotrophic plants. In the presence of mycorrhizal 

fungi, the mycotrophic grasses Pascopyrum smithii
30 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

and Bouteloua gracilis were able to outcompete nonmycotrophic Salsola tragus L., an exotic invasive of 

disturbed soils. In their absence, the reverse was true 

(Allen and Allen 1984). Similar changes in competitive 

outcome have been obtained using other grass-exotic 

weed, grass-grass, and grass-legume combinations 

that differed in their dependence on mycorrhizae 

(Benjamin and Allen 1987, Crush 1974, Fitter 1977, 

Goodwin 1992, Hall 1978, Hartnett and others 1993, 

Hetrick and others 1989, West 1996). Multispecies 

experiments have shown that the presence or absence 

of mycorrhizal fungi can alter species composition and 

diversity in grassland ecosystems (Grime and others 

1987, Hartnett and Wilson 1999, van der Heijden and 

others 1998, Wilson and Hartnett 1997).

Species composition and diversity of the fungal community itself may affect plant community structure 

(van der Heijden and others 1998). Mycorrhizal fungal 

species differ in their ability to take up nutrients and 

promote growth (Haas and Krikun 1985, Jakobsen and 

others 1992, Stahl and Smith 1984, van der Heijden 

and others 1998, Wilson and Tommerup 1992). Fungal 

species are known to have a patchy distribution 

and may occur in specific microhabitats (Allen and 

MacMahon 1985, Johnson 1993). Therefore, individual 

plants may be colonized by the same fungal species, 

by different fungal species, or by multiple species of 

fungi (Rosendahl and others 1990, van der Heijden 

and others 1998).

The number and proportion of fungal symbionts 

can vary from plant to plant even within a relatively 

homogenous vegetation type. Differences among plants 

in their response to colonization, as well as in the 

complement of mycorrhizal fungi present on the root 

system, will differentially affect their ability to compete 

for soil resources (Streitwolf-Engel and others 1997, 

van der Heijden and others 1998).

Another factor affecting plant-plant interaction 

is that of interplant connections through a shared 

mycelial network. Studies have documented carbon 

and phosphorus transfer between plants connected 

by arbuscular mycorrhizal hyphae (Chiariello and 

others 1982, Francis and Read 1984, Newman and 

Ritz 1986, Whittingham and Read 1982). The ecological significance of these fungal connections is not well 

understood, however. Interplant connections may 

promote the transfer of nutrients from larger “source” 

plants to subordinate “sink” plants, thus allowing less 

competitive species to coexist with dominants (Allen 

1991, Grime and others 1987). This scenario was invoked by Grime and others (1987) to explain why the 

addition of mycorrhizal fungi increased plant diversity 

in a microcosm experiment. Source to sink resource 

transfer may be of particular benefit to seedlings establishing in existing vegetation (Fitter 1989, Francis 

and Read 1984) and to plants growing in nutrient-poor 

soil patches (Allen and Allen 1990).

In contrast to the above theory, existing data from 

greenhouse competition experiments using mycotrophic species show an increase in competition in the 

presence of mycorrhizal fungi, rather than a cooperative interaction (Caldwell and others 1985, Hetrick 

and others 1989, Zobel and Moora 1995). There is 

great difficulty, however, in extrapolating data from 

greenhouse experiments with two species to actual 

field conditions, where the complexity of interaction 

is much greater. Zobel and Moora (1995) suggested 

that competition experiments might not include the 

critical stages of plant development (such as seedling 

and flowering stages) where source-sink transfer is 

thought to be most important.

Variation in precipitation patterns or patchiness 

in resource availability, factors rarely included in 

greenhouse experiments, can greatly affect mycorrhizal functioning and plant community composition 

(Allen 1991, Duke and others 1994, Hartnett and 

Wilson 1999, Pendleton and Smith 1983). Temporal 

and spatial niche partitioning due to differences in 

microhabitat requirements and in plant phenology 

limit the amount of actual competition experienced by 

a plant (Hetrick and others 1989). Certainly, extensive 

research is required before the role of mycorrhizae 

in regulating natural grassland communities can be 

well understood.

Disturbance and Succession

In grasslands, most arbuscular mycorrhizal (AM) 

hyphae and spores occur in the top 20 cm of the soil 

profile where the concentration of plant roots is the 

greatest (Allen 1991). Any disturbance resulting in the 

redistribution of these soils can affect the number and 

dispersion of mycorrhizal roots, hyphae, and spores. In 

most cases, soil disturbance leads to a reduction in the 

number of mycorrhizal propagules (Allen and others 

1987, Jasper and others 1989, Moorman and Reeves 

1979, Powell 1980). Small-scale natural disturbance 

such as gopher activity can create patches of low inoculum density as subsurface soils are brought to the 

surface (Koide and Mooney 1987). Inoculum density 

increases as hyphae from adjacent mycorrhizal plants 

slowly expand into sterile areas (Allen 1991). Spores 

may also be carried into these areas by crickets, grasshoppers, and other insects (Trappe 1981).

Large-scale disturbances such as surface mining, 

road construction, and watershed erosion can create 

extensive areas in which the amount of mycorrhizal 

inoculum is greatly reduced or eliminated (see for 

example Allen and Allen 1980). Mycorrhizal inoculum 

is introduced into these areas largely through windblown spores (Allen 1991). Consequently, the unaided 

restoration of mycorrhizal-dependent vegetation may 

require a long time (Miller 1987).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 31

Changes in soil inoculum potential can affect a 

plant’s ability to colonize disturbed areas. Many 

researchers have noted an increase in the abundance 

of mycotrophic species along a successional sequence 

(Allen 1991, Janos 1980, Miller 1987). Allen and Allen 

(1990) have proposed a successional model in which 

biomes are classified according to nutrient and moisture 

gradients. The degree of dependence on mycorrhizae 

increases with precipitation and decreases with enhanced soil fertility.

In arid grasslands, as in arid shrubland, early seral 

plant species are predicted to be largely nonmycotrophic, whereas late seral species are facultative in 

their dependence. These predictions are supported by 

numerous research reports. For example, Allen and 

Allen (1980) found five of seven annuals growing on 

strip-mined and disked prairie sites were nonmycorrhizal. Nonmycotrophic annuals predominated for up 

to 10 years on mined sites composed of sterile subsoil. 

Similarly, Pendleton and Smith (1983) found flat semiarid disturbed sites to be dominated by nonmycorrhizal 

species. Cover of mycorrhizal species increased with 

water availability. Soil fumigation can delay succession, suggesting that soil biotic communities, including 

mycorrhizal fungi, have a significant impact on successional dynamics (Stevenson and others 2000).

Little is known about successional patterns in mesic 

and dry alpine grasslands. The successional model of 

Allen and Allen (1980) would predict that all seral 

stages comprise facultative species, with perhaps some 

obligate species in late seral stages. Allen and others 

(1987) found all surveyed species colonizing disturbed 

alpine ecosystems to be mycotrophic. No relationship 

between age since disturbance and degree of colonization or spore number was apparent.

Species of the fungal symbiont also appear to change 

with succession (Allen 1991). Fungal communities of 

late successional grasses apparently differ from those 

of early successional grass species (Johnson and others 

1992). Others report an increase in diversity of VA 

fungal species with increasing seral stage (Allen and 

others 1987, Brundrett 1996, Nicolson and Johnston 

1979).

Grazing

The response of mycorrhizae to aboveground herbivore grazing varies widely depending on the plant 

and fungal species involved, on the type and intensity of grazing, and on a number of abiotic factors. 

Some studies have shown a decrease in mycorrhizal 

colonization of forage grasses as a result of livestock 

grazing or defoliation (Allsopp 1998, Bethlenfalvay 

and Dakessian 1984, Bethlenfalvay and others 1985, 

Wallace 1981). Others report either no effect or an 

increase in mycorrhizal activity (Allen and others 1989, 

Davidson and Christensen 1977, Tisdall and Oades 

1979, Wallace 1987). In some instances, colonization by 

mycorrhizal fungi has been shown to enhance tillering 

and a prostrate growth habit, thereby increasing the 

plant’s ability to tolerate grazing (Bethlenfalvay and 

Dakessian 1984, Miller 1987, Wallace and others 1982). 

Allsopp (1998) found that the response to defoliation 

by three mycorrhizal forage grasses depended on the 

grazing tolerance of the host plant species.

The response of the mycorrhiza to grazing is likely 

related to the carbohydrate storage and photosynthetic 

capacity of the host plant at the time defoliation occurs. 

Regulation of the association seems to be a function of 

the host plant (Koide 1993). And, although generally 

considered beneficial to the host plant, the fungi can 

be parasitic if the net cost of the symbiosis exceeds the 

net benefit (Johnson and others 1997, Koide 1993). 

Plants benefit from the association when photosynthate 

is readily available or when the association increases 

the plant’s ability to photosynthesize (Bethlenfalvay 

and Pacovsky 1983). Therefore, when soil resources 

are more limiting than photosynthetic leaf area, herbivory may have little effect on mycorrhizal functioning 

(Allen 1991).

Under conditions that severely limit the amount 

of carbon a plant is able to fix and store—such as 

extremely low nutrient availability, insufficient light, 

or prolonged drought—the fungus may constitute an 

excessive carbon drain on the plant (Johnson and 

others 1997). Herbivory under these conditions would 

likely be accompanied by a reduction in mycorrhizal 

function. For example, in a pinyon-juniper ecosystem, 

defoliation caused a decrease in mycorrhizal function 

for plants growing in nutrient- and water-stressed 

volcanic fields, but not for plants growing in a nearby 

sandy loam (Gehring and Whitham 1995).

Prescribed burn on the Kiowa National Grassland, Union County, 

New Mexico. (Photo by Mike Friggens)
32 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Fire

Response of mycorrhizal fungi to fire is variable, 

with some studies reporting a decrease in mycorrhizal 

function, while others report no effect or an increase in 

function (Allen 1991, Eom and others 1999). Of those 

studies reporting a temporary decrease in mycorrhizal 

function, whether in soil propagule numbers or percent 

root colonization, most had recovered within 1 to 2 years 

(Dhillion and others 1988, Gurr and Wicklow-Howard 

1994, Pendleton and Smith 1983, Rashid and others 

1997). Recovery rates depended in part on fire intensity 

and on soil conditions at the time of burning (Klopatek 

and others 1987,Wicklow-Howard 1986). Klopatek and 

others (1987) reported moderate decreases in propagule numbers when soil temperatures reached 50 to 

60 oC. Significantly greater propagule loss occurred 

at soil temperatures greater than 60 oC. Colonization 

of plant roots also decreased following fire; however, 

the decrease was less in wetter soils.

The observed response of mycorrhizal fungi to 

fire may be a reflection of fire effects on the plants 

themselves, rather than a direct effect upon the fungi 

(Dhillion and others 1988). Soil provides good insulation. Riechert and Reeder (1971) reported no increase 

in soil temperature at a depth of 1 cm, despite surface 

temperatures of 200 oC. Dhillion and co-workers (1988) 

attributed temporary decrease in root colonization to a 

fire-induced stimulation of root production that temporarily outstripped colonization by the fungi. Much of the 

research on fire effects on mycorrhizal fungi has been 

done in tallgrass prairie, however. Additional research 

on arid grasslands is needed. Recovery of vegetation 

in arid grasslands is highly dependent on postfire precipitation patterns and generally occurs at a slower 

rate than in more mesic grasslands. The response of 

mycorrhizal fungi to fire is most likely reflective of the 

vegetation response, generally proceeding at a slower 

pace in arid areas and depending on the same suite of 

factors that affect vegetation response.

The interaction between fire, watershed stability, 

and mycorrhizal fungi needs further research. However, 

one of the most important factors affecting mycorrhizal 

recovery following fire is that of soil erosion. In areas 

where postfire erosion is significant (such as steep 

slopes or high-intensity fires), loss of topsoil may delay 

the recovery of mycorrhizal function (O’Dea and others 

2000). Eroding topsoil carries with it a large proportion 

of mycorrhizal propagules. Seedlings planted on eroded 

soils had significantly less colonization, growth, and 

survival than those with additions of captured eroded 

soil (Amaranthus and Trappe 1993).

Atmospheric Change

Disturbance in the form of increasing atmospheric 

CO2 concentrations and anthropogenic nitrogen 

deposition may also significantly affect mycorrhizal 

functioning within grassland communities (EgertonWarburton and Allen 2000, Rillig and others 1999). 

Researchers are just beginning to explore these possibilities. McLendon and Redente (1991) reported a delay 

in successional change of plant species on fertilized 

plots. Early seral colonizers, nonmycotrophic Salsola 

tragus and Bassia scoparia (L.) A.J. Scott, still retained 

dominance of N-fertilized plots 2 years after control 

plots had changed to perennial (and mycotrophic) 

grasses, shrubs, and forbs.

Exotic Weeds

Noxious weeds compose a serious threat to the 

structure, organization, and function of ecological 

systems (Olsen 1999). Weeds prefer disturbed areas 

where resource availability is increased (Davis and 

others 2000). Of course, these are also areas where 

the mycorrhizal inoculum potential of the soil has 

been reduced. A healthy soil microflora can, in some 

cases, differentially enhance survival and production 

of native species over that of exotic weeds.

As mentioned previously, mycorrhizae generally enhance the competitive ability of those plants that are 

most responsive to the fungi. Many of the world’s most 

aggressive weeds are either nonmycotrophic or facultatively mycotrophic (Trappe 1987). Invading annuals 

on arid grasslands are often nonmycotrophic weeds of 

the chenopod, mustard, and amaranth families (Allen 

1991). Competition experiments between these species 

and native grasses have demonstrated enhanced competitive ability of native grass species in the presence of 

mycorrhizal fungi. Mycorrhizae contribute, therefore, 

to the resistance of these native communities to invasions of exotic weeds (Goodwin 1992).

Exotic grasses, on the other hand, are facultatively 

mycotrophic. Large areas in the Western United 

States are now dominated by exotic grasses such as 

cheatgrass (Bromus tectorum L.) in the Great Basin, 

Lehmann lovegrass (Eragrostis lehmanniana Nees) 

in the Southwest, and Mediterranean annual grasses 

in southern California (Allen 1995, McClaran 1995, 

Stylinski and Allen 1999). The role of mycorrhizal fungi 

in regulating competitive interactions between these 

exotic species and native vegetation is uncertain. It is 

likely, however, that competition between exotic and 

native grass species will be little altered by mycorrhizal 

fungi (Goodwin 1992).

In monoculture, cheatgrass shows no positive 

response to mycorrhizal fungi, although colonization 

levels can be quite high (Allen 1984, Benjamin and 

Allen 1987, Schwab and Loomis 1987). Cheatgrass may 

benefit, however, when the relative density of mycorrhizal competitors is high (Goodwin 1992, Schwab 

and Loomis 1987). Similarly, Nelson and Allen (1993) 

reported that the addition of mycorrhizae did not 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 33

enhance competitive ability of purple nodding tussock 

grass (Nassella pulchra (A.S. Hitchc.) Barkworth) in 

competition with annual slender oats (Avena barbata

Pott ex Link).

Unfortunately, in some cases, invasive nonnative 

woody or herbaceous species are more responsive to 

mycorrhizae than are native grasses. In these instances, 

the presence of mycorrhizal fungi may actually enhance the competitive ability of the invading species. 

For example, Marler and others (1999) report that 

AM fungi strongly enhance the competitive ability 

of spotted knapweed (Centaurea biebersteinii DC.) 

grown in competition with Idaho fescue (Festuca

idahoensis Elmer). Clearly, the problems stemming 

from invasive weeds will increasingly challenge the 

resourcefulness, skill, and ecological knowledge of all 

managers and researchers concerned with the health 

of our Southwestern landscapes.

Ecosystem Restoration

The establishment of mycorrhizal associations is 

an important consideration in the design of successful revegetation and restoration efforts of grasslands. 

Miller and Jastrow (1992) list eight site conditions 

under which management for mycorrhizae may be 

particularly important, nearly all of which are applicable to Southwestern grasslands. Mycorrhizae 

may aid in restoration attempts by enhancing plant 

survival, increasing soil stability, and through their 

effect on competitive interactions of successional plant 

species. Mycorrhizae are also important contributors to 

stable soil structure and the redevelopment of nutrient 

cycles (Miller and Jastrow 1992).

As discussed above, anthropogenic activities that 

remove, compact, or otherwise disturb soil can severely 

reduce or eliminate mycorrhizal propagules. This, in 

turn, affects the ability of mycotrophic plant species to 

colonize the site. In addition to aiding plant survival and 

establishment, the reintroduction of mycorrhizae may 

allow late or mid-seral species to perform better than 

nonmycotrophic early seral species, thereby speeding 

up the rate of succession (Allen and Allen 1988, Reeves 

and others 1979). Because of this, mycotrophy could 

be an important factor in determining of seed mixes 

for grassland rehabilitation.

Severely disturbed sites such as mine spoils are 

particularly difficult sites for plant establishment, 

combining adverse growing conditions with a lack 

of viable mycorrhizal inoculum. Inoculation with VA 

fungi can greatly improve survival and growth of 

desirable reclamation plants, particularly shrubs, 

grown on mine spoil material and other severely 

disturbed soils (Aldon 1975, Lindsey and others 

1977, Smith and others 1998, Stahl and others 1988, 

Williams and Allen 1983). Transplants may then form 

islands of inoculum that can spread through root and 

hyphal growth, as well as wind and animal dispersal 

(E.B. Allen 1984, M.F. Allen 1991). Reintroduction 

of mycorrhizae to severely disturbed sites may be 

enhanced through the use of soil amendments or 

redistribution of stockpiled topsoil, provided topsoil 

has been managed properly (Allen and Allen 1980, 

Allen 1984, Johnson 1998, Miller 1987, Pendleton 

1981, Zak and Parkinson 1983).

Where severe soil disturbance makes it impractical 

to seed with mycorrhizal-dependent late-seral species, 

seeding with facultative midseral species such as 

cool-season grasses may help to increase soil inoculum to the point where more dependent warm-season 

grasses and shrubs can establish successfully (Noyd 

and others 1995). Johnson (1998) suggests that seeding 

of facultative mycotrophic species and manipulation 

of edaphic conditions to optimize mycotrophy may be 

a more cost effective method of rehabilitation than 

large-scale inoculation. The precise combination of 

techniques needed for successful rehabilitation will 

depend on soil conditions and the severity and scale 

of the disturbance, as well as management objectives 

(Allen 1995).

Biological Soil Crusts_______________

Biological soil crusts are known by many names, 

including microphytic crusts, microfloral crusts, 

cryptogamic crusts, and cryptobiotic crusts (St. Clair 

and Johansen 1993). As the name implies, they are 

biological in nature, being composed of cyanobacteria, 

green algae, lichens, and mosses. Other bacteria and 

fungi (including mycorrhizal fungi) may also be present 

(Belnap and others 2001). It is important to distinguish biological soil crusts from inorganic soil crusts 

formed by chemical and physical means. Inorganic 

soil crusts reduce water infiltration and may hinder 

seedling emergence. In contrast, biological soil crusts 

play an important role in arid and semiarid lands by 

stabilizing soil surfaces, increasing soil fertility through 

the fixation of atmospheric nitrogen, and improving 

seedling establishment (Belnap and others 2001, 

Johnston 1997, St. Clair and Johansen 1993).

Biological soil crusts are formed by the interweaving of cyanobacterial and fungal strands within the 

upper soil surface. Gelatinous sheath material and 

other polysaccharides extruded by these organisms 

further help bind the soil particles together. The 

resulting crust may be anywhere from 1 mm to 10 cm 

thick (Belnap and Gardner 1993, Johnston 1997). The 

dominant photosynthetic organisms—cyanobacteria 

and green algae—require sunlight, and most of the 

living cells are found 0.2 to 0.5 mm below the surface 

(Belnap and others 2001). The polysaccharide sheaths 

of cyanobacteria expand when wet, pushing the interior 

bacterial filaments out across the soil surface. New 
34 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

sheath material is produced by 

these filaments as they dry out. 

Old sheath material gradually 

becomes buried and, although no 

longer photosynthetic, continues 

to contribute to the thickness and 

water-holding capacity of the crust 

(Belnap and Gardner 1993).

Biological soil crusts are highly 

variable in appearance and may 

be grouped according to their morphology. They may be smooth or 

rough, flat or pedicelled, greenish 

or black, depending on the temperature regime of the area and 

the types of organisms involved 

(Belnap and others 2001). Simple 

algal and cyanobacterial crusts 

may be flat. Topography of these 

crusts increases as the more complex lichens and mosses colonize 

the site. Increased topography of 

crusts is also associated with frost 

heaving (Johnston 1997). Crust 

morphology in hot desert regions, 

including the Chihuahuan and Sonoran Deserts, is 

flat to slightly rough or rugose. Cool desert regions 

such as the Great Basin and Colorado Plateau are 

dominated by rolling or pinnacled crusts (Belnap and 

others 2001). Different morphotypes differ in their 

ecological function with regard to water retention and 

erodability (Eldridge and Rosentreter 1999).

Distribution and Occurrence

Biological soil crusts are found throughout the 

world, occurring on every continent and in multiple 

vegetation types. Although most well known from arid 

and semiarid regions, they also occur on shallow alpine 

soils and other areas where vascular plant cover is low. 

In the Southwestern United States, cyanobacteria, 

lichens, and mosses are the most important components 

of soil crusts (Ladyman and Muldavin 1996). Many of 

the most common organisms, such as Microcoleus spp., 

Nostoc spp., Collema spp., Psora decipiens (Hedwig) 

Hoffm., and Catapyrenium lachneum (Ach.) R. Sant., 

are cosmopolitan in nature, occurring in many geographic areas. Other species are endemic or have a 

much narrower geographic range (Belnap and others 

2001, St. Clair and others 1993).

Regional trends in the timing and amount of precipitation greatly affect species composition of the 

crustal community. Regions in which precipitation 

falls mainly during the winter months have a diverse 

lichen community and a cyanobacterial community 

dominated primarily by Microcoleus (Belnap and 

others 2001). Great Basin areas are dominated by 

low-growing soil lichens such as Collematenax (Swartz) 

Ach., Catapyrenium lachneum, and Caloplaca tominii

(Savicz) Ahlner (St. Clair and others 1993). In contrast, 

crusts of the Great Plains are dominated by detached 

foliose (leaflike) or fruiticose lichens (Belnap and others 

2001). Crusts of the Colorado Plateau are dominated 

by the cyanobacteria Microcoleus, along with the lichen 

Collema. Upland areas of the Colorado Plateau have 

many species of the genus Psora (St. Clair and others 

1993). In regions characterized by summer rainfall, 

the lichen community is small and the cyanobacterial 

community diverse. Common cyanobacterial genera in 

the Sonoran Desert include Nostoc, Schizothrix, and 

Scytonema (Belnap and others 2001).

On a local level, species composition is largely influenced by soil texture, soil chemistry, slope, and aspect 

(Belnap and others 2001, Ladyman and Muldavin 

1996). In general, silty loams support a greater diversity of crust organisms than do sandy soils (Belnap 

and others 2001). Green algae are more common on 

acidic soils, while blue-green algae (cyanobacteria) 

are more common on alkaline soils (Ladyman and 

Muldavin 1996). Gypsiferous and calcareous soils 

often have extensive crusts with a high diversity of 

organisms (St. Clair and others 1993). Alpine sites 

have many unique lichen taxa, as well as some that 

are common to the region (Belnap and others 2001). 

In the arid Southwest, crust development is often 

noticeably greater on north-facing slopes and near 

ridge tops (Brotherson and Masslich 1985, Pendleton, 

personal observation).

Soil crust composed of algae, lichens, and mosses. (Photo by Rosemary Pendleton)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 35

Ecological Function

Biological soil crusts contribute to a variety of 

ecological functions, including soil stabilization, 

nitrogen fixation, nutrient availability, and vascular 

plant establishment. Of primary importance is the 

ability of biological crusts to reduce wind and water 

erosion of soil surfaces (Belnap 1993, Belnap and 

others 2001, Eldridge 1993, Johnston 1997, Ladyman 

and Muldavin 1996). Within the crust, filamentous 

cyanobacteria and green algae intertwine with soil 

particles, stabilizing soil surfaces. Fungal strands, or 

hyphae, further bind soil particles, as do the rhizoids 

of mosses. Polysaccharides exuded by cyanobacteria, 

algae, and some fungi (see mycorrhizal section) cement soil particles together into stable aggregates. 

The increased surface topography of later successional 

crusts also serves to protect the soil surface from wind 

and water erosion (Belnap and others 2001, Johnston 

1997).

Several studies have shown that crusted soils require much higher wind velocities before soil movement 

will occur (Belnap and Gillette 1998, Mackenzie and 

Pearson 1979, Williams and others 1995). Eldridge 

(1993) found a significant increase in splash erosion 

when biological soil crust cover was less than 50 

percent. Loss of soil fines—particles associated with 

soil nutrients—also increased with decreasing crust 

cover. Moss-dominated and some lichen-dominated 

crusts increase infiltration of surface water, thereby 

decreasing erosion potential (Ladyman and Muldavin 

1996 and references therein). Other studies report reduced infiltration, particularly in crusts dominated 

by cyanobacteria (Ladyman and Muldavin 1996 and 

references therein). However, even when infiltration 

is not increased, sediment loss is still reduced when 

compared with disturbed soils where the crust layer 

has been completely lost.

Biological soil crusts contribute to the fertility of 

grassland soils by increasing soil nitrogen and carbon 

content. This may be particularly important in arid 

areas of the Southwest where vegetation is sparse. 

Carbon contributions are greater where crusts are 

dominated by lichens and mosses, whereas nitrogen 

gains are greater where cyanobacteria and cyanolichens predominate (Belnap and others 2001). Carbon 

inputs to the soil come from active secretion of fixedcarbon compounds, as well as through the destruction 

of cell membranes during wet-dry cycles and upon 

death of the organism (Ladyman and Muldavin 1996). 

The amount of biomass contributed may be substantial. Beymer and Klopatek (1991) estimated carbon 

contributions of up to 350 kg/ha/year by undisturbed 

soil crusts in northern Arizona. Harper and Pendleton 

(1993) reported higher levels of soil organic matter in 

crusted areas of southern Utah.

Crust organisms are metabolically active only when 

wet. However, even small amounts of liquid moisture 

can trigger a rapid induction of metabolic processes. 

Available evidence suggests that positive carbon 

gains require a prolonged wet period (Jeffries and 

others 1993). In the Colorado Plateau, most growth of 

biological soil crusts occurs during the spring (Belnap 

and others 2001). Similar studies in areas of high 

summer rainfall would help define seasonal growth 

patterns for these regions. Because organisms differ 

in the temperature and moisture contents needed for 

maximum photosynthetic efficiency, regional shifts 

in the timing and amount of precipitation received 

may result in a compositional change in the crust 

community.

The cyanobacterial component of soil crusts, whether 

free-living or part of a lichen symbiosis, is capable of fixing atmospheric nitrogen. This process takes place only 

under anaerobic conditions. Many cyanobacteria have 

specialized cells for this purpose, called heterocysts. 

Nonheterocystic cyanobacteria, such as Microcoleus

create anaerobic conditions by the layering of their 

filaments beneath the soil surface (Belnap and others 

2001). Estimates of the amount of nitrogen fixed by 

soil crusts range from 2 to 365 kg/ha/year (Belnap and 

others 2001, Johnston 1997, Ladyman and Muldavin 

1996, Rychert and others 1978). Fixation rates depend 

on temperature and precipitation patterns, as well as 

the species composition of the soil crust.

Arid regions are generally low in nitrogen content 

compared to other regions, and arid regions have 

few nitrogen-fixing plant species (Farnsworth and 

others 1976, Wullstein 1989). Studies using stable 

isotopes demonstrate that much of the nitrogen used 

by higher plants in arid and semiarid regions was 

originally fixed by soil crust organisms (Belnap 1995, 

Evans and Ehleringer 1993). Other studies report 

that plants growing on crusted soils have higher tissue concentrations of nitrogen than plants growing 

in nearby disturbed areas (Belnap and Harper 1995, 

Harper and Pendleton 1993).

The presence of biological soil crusts can significantly affect germination and growth of vascular 

plant species. Small seeds can lodge in the cracks of 

the roughened crust surface. The dark surface of the 

soil crust can raise soil temperatures 5 o

C or more, 

promoting earlier spring germination and growth 

(Harper and Marble 1988, Harper and Pendleton 

1993). The greatest benefit may be to shallow-rooted 

annuals and to deeper-rooted perennials during the 

critical establishment phase following germination. 

Many studies have reported increased survival and 

nutrient content of young seedlings growing in crusted 

soils (Belnap and Harper 1995, Harper and Marble 

1988, Harper and Pendleton 1993, Pendleton and 

Warren 1995).
36 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Nutrients enhanced by the presence of soil crusts 

include nitrogen, mentioned previously, and also P, 

K, Ca, Mg, and Zn. Enhanced nutrient content may 

be due to the tendency of biological crusts to trap 

nutrient-rich soil fines (Belnap and Gardner 1993). 

Cyanobacterial sheaths also secrete chelating agents 

that increase the availability of essential nutrients 

(see references in Belnap and others 2001, Harper 

and Pendleton 1993, Ladyman and Muldavin 1996). 

Biological crusts interact with other soil microorganisms in promoting the establishment and growth of 

vascular plant species. Harper and Pendleton (1993) 

found increased colonization by several rhizosymbionts (mycorrhizal fungi, Rhizobium, and rhizosheath 

organisms) in plants growing on crusted soils.

There is no credible evidence that cover of vascular 

plant species and biological soil crusts are negatively 

related, as was suggested by Savory (1988). Cover of 

biological soil crusts has been positively correlated with 

diversity of plant species (Beymer and Klopatek 1992, 

Harper and Marble 1988). Numerous other studies 

report either a positive correlation or no correlation 

between plant cover and the cover of biological soil 

crusts (Belnap and others 2001).

Ladyman and others (1994) examined vegetative 

cover on three undisturbed mesas in New Mexico 

and found that the mesa with the highest crust cover 

also had the highest grass cover. Specifically, blue 

grama grass (Bouteloua gracilis) was positively associated with soil crust cover. In another comparison 

between two grassland communities, Kleiner and 

Harper (1977) found no competition between soil crust 

organisms and vascular plant species. Rather, grass 

cover and crust cover appeared to be independent of 

one another. Ladyman and Muldavin (1996) suggest 

that specific associations of vascular plants and crust 

organisms that appear to be negative, positive, or 

neutral, may be more a reflection of their particular 

habitat requirements rather than actual competition. 

Greenhouse experiments, however, seem to indicate 

some competition between soil microorganisms and 

plants, particularly when crust organisms are themselves becoming established (Harper and Pendleton 

1993, Pendleton, unpublished data).

Disturbance and Succession

Disturbance of the soil crust through trampling, 

vehicular travel, or soil movement affects the functioning of biological soil crusts through changes in 

cover, composition, or by affecting rates of carbon and 

nitrogen fixation (Belnap and others 2001, Ladyman 

and Muldavin 1996). The impact of the disturbance 

depends on the severity and type of disturbance, the 

frequency of disturbance, soil texture, and climatic 

conditions. Disturbance that compacts the soil crust 

but leaves crust material in place is less severe than 

disturbance that removes or kills crust organisms. 

Where the crust is destroyed, the loss of soil structure 

results in an unstable surface that is highly susceptible 

to erosion. Crusts formed on sandy soils are more 

easily damaged when dry, whereas crusts formed on 

clay soils are more vulnerable when wet. Lichens and 

mosses are more susceptible to disturbance than are 

cyanobacteria and green algae (Belnap and others 

2001).

Repeated disturbance can result in a less complex 

(early successional) crust structure, or in the complete 

loss of the soil crust (Belnap and others 2001, Harper 

and Marble 1988). Estimates of recovery rates for 

soil crusts vary widely. Under optimal conditions 

visual recovery may occur within 1 to 5 years (Belnap 

1993, Cole 1990, Johansen and Rushforth 1985). 

Recovery in terms of chlorophyll content, community composition, and organism density takes much 

longer. Belnap (1993) estimated recovery rates for 

Colorado Plateau crusts of 40 years for chlorophyll 

a content, 45 to 80 years for the lichen component, 

and 250 years or more for the moss component. 

Clearly, full recovery from severe disturbance can 

take a long time.

Crust development follows a definite successional 

pattern, with various organisms classed as either early 

or late successional species. Initial colonization of a 

site is usually accomplished by either cyanobacteria 

or green algae, depending on the acidity of the soils 

(Belnap and others 2001). A common colonizer of 

Southwestern soils is the large filamentous cyanobacteria, Microcoleus. These filaments help stabilize the 

soil surface, a function especially important for sandy 

soils, and allow colonization by other cyanobacteria 

and algae.

In hot deserts with summer monsoonal rainfall, 

different cyanobacteria, such as Schizothrix and Nostoc

spp., may be more common (Johnston 1997). Once 

soils are stabilized, colonization by gelatinous nitrogen-fixing lichens, such as Collema tenax, can occur 

(Belnap and others 2001). Other early successional 

Southwestern species include the lichens Cladonia

chlorophaea (Flk.) Spreng., Endocarpon pusillum

Hedwig, and Peltigera refuscens (Weiss) Humb. Mid 

to late successional species include other lichen taxa 

(such as Psora spp.), mosses (Grimmia and Tortula

spp.), and liverworts (Riccia spp.; Ladyman and 

Muldavin 1996). Because growth of crustal organisms 

can take place only when wet, succession occurs at a 

faster rate in years of high precipitation, on northfacing slopes, and on fine-textured soils that retain 

moisture longer (Belnap and others 2001, Ladyman 

and Muldavin 1996).

Monitoring of biological soil crusts has been suggested as a means of assessing ecosystem health. More 

recently, Eldridge and Rosentreter (1999) proposed 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 37

a system of monitoring using morphological groups 

rather than species. Crust morphology determines its 

functioning in terms of water retention, erodability, 

and resistance to disturbance (Belnap and others 2001, 

Eldridge and Rosentreter 1999). This system provides 

an easier and more accurate method for nonspecialists 

to use in monitoring the effects of different management 

practices on ecosystem health. Specific techniques for 

monitoring biological soil crusts are covered in Belnap 

and others (2001).

Grazing

Different regions of the Southwest have different 

grazing histories, soils, and precipitation patterns. 

With the exception of the short-grass steppe, most 

regions historically experienced minimal disturbance 

by large herds of grazing mammals and may therefore 

be more susceptible to damage from grazing animals 

(Belnap and others 2001, St. Clair and others 1993). 

The effect of livestock grazing on biological soil crusts 

is due primarily to trampling, or hoof action. Trampling 

breaks up the sheath and filament structure that binds 

the soil particles, leading to increased erosion and loss 

of soil fertility.

The destruction of the crust through hoof action, 

while advocated by some (Savory 1988), does not 

result in increased plant cover as had been surmised 

(Ladyman and Muldavin 1996 and references therein). 

The magnitude of the disturbance is proportional to 

the grazing pressure and is also dependent on the 

seasonality of grazing. As discussed previously, biological crusts on most soils of the Southwest are more 

tolerant of disturbance when soils are moist. On these 

soils, late fall/early winter grazing is least likely to 

cause damage to crusts (Anderson and others 1982, 

Belnap and others 2001, Marble and Harper 1989). 

Several researchers have suggested that periodic 

rest from grazing would benefit crust organisms and 

grasses alike (Brotherson and others 1983, Johansen 

1986). Belnap and others (2001) recommend light to 

moderate grazing in the early to mid wet season. 

Low-elevation grazing during the winter months followed by grazing of high-elevation sites during the 

summer would more closely mimic grazing patterns 

of native ungulates.

Fire

Fire has historically been a common occurrence 

in grassland ecosystems. Evidence from other vegetation types suggests that fire can damage soil 

crusts, depending on the fire’s intensity, frequency, 

and timing, and depending on soil moisture content 

during and immediately following the fire (Belnap 

and others 2001, Ladyman and Muldavin 1996). High 

intensity fires, followed by drought, can result in 

substantial loss of the crust community, particularly 

lichens and mosses. Recovery may take many years. 

Unburned patches within the burned matrix can, 

however, provide propagules for reestablishment 

of soil crust organisms (Belnap and others 2001). 

Low intensity fires may allow for rapid regrowth 

of algae and cyanobacteria (Johansen and others 

1993, Johnston 1997). Johansen and others (1993) 

theorized that recovery of the crust might be partially 

dependent on recovery of vascular plant cover. This 

suggests that recovery in fire-adapted grasslands may 

proceed at a faster pace. Adequate moisture during 

the postfire recovery period will further speed the 

recovery process.

Ford (2000) studied the effect of fire season on 

biological crusts in a short grass steppe. In the 

short term, growing-season fire appeared to reduce 

the impact of fire on microbiotic crusts ability to fix 

nitrogen, compared to dormant-season fire, due to the 

differences in the nature of the fire. Growing-season 

fires occurred when vegetation was green and moisture was high, whereas dormant-season fire occurred 

when moisture was low and fine fuel (vegetation) was 

dry. Again, adequate moisture appears to be the key 

component in crust recovery following fire and abundant precipitation will override any fire effects (Ford, 

personal communication).

Pollution

Many lichens are sensitive to common air pollutants and have been used as indicator species in other 

regions (Nash and Wirth 1988, St. Clair and others 

1993). However, studies on Southwestern soil lichens 

found near pollution sources such as power plants have 

found no change in crust composition or cover (Belnap 

and others 2001, Ladyman and Muldavin 1996). It is 

thought that the low stature of soil lichens and the 

alkalinity of Western soils may ameliorate the effects of 

acidic sulfur emissions. Human-induced increases in atmospheric nitrogen deposition do have a negative effect 

on soil crusts, inhibiting natural fixation rates by soil 

microorganisms (Harper and Marble 1988). Fixation 

rates of cyanobacteria and cyanolichen crusts declined 

in the presence of power plant effluents (Belnap and 

others 2001). The effects of increasing atmospheric 

CO2 levels have not yet been studied.

Exotic weeds

Little is known about the effects of exotic weed 

introductions on soil crust functioning. In areas 

where native bunchgrasses have been replaced by 

annual brome (such as Bromus tectorum), the soil 

crust consists of only a few species of annual mosses 

and cyanobacteria instead of the complex perennial 

crust community characteristic of undisturbed sites 
38 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

(Kaltenecker 1997). This change in crust composition 

may be at least partially due to the drastic change in 

fire frequency that accompanies annual brome invasions. Similar studies are needed to determine the 

effects of lovegrass and other exotic weed invasions on 

composition and functioning of biological soil crusts 

in the Southwest.

Restoration

The use of soil inoculants may hasten recovery of 

soil crusts on severely disturbed sites. This would 

be particularly helpful for large-scale disturbances 

where nearby sources of naturally dispersed inocula 

are not available. Initial trials have been promising. 

St. Clair and others (1986) used a slurry of soil crust 

and water to inoculate pinyon-juniper and salt desert shrub communities following severe fire. Within 

6 months, inoculated soils had greater numbers of 

algae and lichens. Belnap (1993) used crumbled dry 

crust material to inoculate scalped plots on four sites 

in the Colorado Plateau. Recovery of inoculated sites 

was significantly greater in terms of chlorophyll a

content, crust cover, and species diversity 2 to 5 years 

following disturbance. These experiments relied on 

preexisting crust material taken from other locations. 

Johansen and others (1994) have developed techniques 

to culture nitrogen-fixing cyanobacterial species. 

Initial tests used intact or ground alginate pellets as 

the inoculum source (Buttars and others 1994, 1998, 

Johansen and others 1994). Currently, development of 

a more economically feasible delivery system is under 

way (Johansen, personal communication).

Invasive Weeds____________________

For centuries weeds have been the focus of study for 

farmers and agronomists. Alien weeds have arrived 

in North American from a number of sources. Many 

species were introduced as contaminants of agricultural 

seed or as unnoticed hitchhikers that immigrated with 

people, household goods, and livestock to North America 

and subsequently moved with settlement across the 

continent. Other species were originally introduced as 

ornamental plants and subsequently escaped cultivation to become established in the native landscape. A 

third group of species are those intentionally introduced 

with the hope that they would fulfill a specific management objective, but that have since proliferated to the 

point where they now pose a serious problem (Brock 

1998, Cousens and Mortimer 1995).

The invasion of alien species has been likened to a 

biological wildfire that is rapidly spreading at a rate of 

200 acres/hour in the West (Lee 1999, Mitchell 2000). 

As the number of exotic species and acreage affected 

has risen, so has the concern of environmentalists and 

managers. Exotic weed invasions now command the attention of scientists and conservationists from a variety 

of backgrounds (Cousens and Mortimer 1995).

At the Federal level, President Clinton (1999) 

established a cabinet-level Invasive Species Council, 

which is charged with providing leadership in the management of invasive species. Executive Order 13112 

outlines a far-reaching program designed to “prevent 

the introduction of invasive species and provide for 

their control and to minimize the economic, ecological, and human health impacts that invasive species 

cause.” The executive order was the impetus for the 

current Forest Service weed management program. 

As Sheley and Petroff (1999) stated, our “commitment 

to addressing the rapid invasion of noxious weeds has 

been revived and intensified.”

Among the definitions of a weed are the triedand-true “a weed is a plant out of place” and “a weed 

is any plant that interferes with the management 

objectives of a site” (Lee 1999). Lee (1999) defines an 

invasive as a plant that is not native to a particular 

ecosystem, and an invasive weed as one that has or 

will have a negative impact on the environment. The 

terms “invasive” and “weed” are used with both native 

and nonnative species. While a species may be native 

to North America, it may become invasive in regions 

where it was not historically present.

The definitions used in this section are as follows. 

A noxious weed is a legal term that identifies an undesirable plant as one that is regulated in some way by 

law (Sheley and Petroff 1999). Executive Order 13112 

(Clinton 1999) standardizes the terminology used by 

Federal land managers. Alien species are species not 

native to a particular ecosystem. Invasive species are 

alien species whose introduction causes or is likely 

to cause economic or environmental harm or harm to 

human health.

Problems stemming from invasive plant species are 

many. Biological invasions disrupt natural ecosystems, 

posing negative consequences for both ecological and 

socioeconomic systems (Davis and others 2000, Mitchell 

2000). Invasive species displace desirable native species, reduce the quality of wildlife habitat, damage 

sensitive riparian and watershed areas, and increase 

erosion (Lee 1999, Masters and Sheley 2001). As native 

vegetation becomes displaced, opportunities for land 

use decline and land values drop. Improvement requires 

enormous inputs of time and money from managers 

and management agencies (Lee 1999). Less visible, 

but perhaps more important, are the alterations in 

ecological processes that accompany vegetation change 

(Masters and Sheley 2001). Changes in vegetation result in changes in the soil microflora, affecting nutrient 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 39

cycling and decomposition rates (Adams and Wall 

2000) as discussed in a previous section.

One of the greatest challenges facing scientists and 

managers is the unpredictable nature of biological 

invasions—why certain species become problems and 

where these problems are likely to occur (Peterson 

and Vieglais 2001). Invasive species have a number 

of characteristic traits that have been identified 

(Mitchell 2000). They are generally fast growing, 

produce numerous small seeds, and are adapted to 

disturbance (Grime 1988). Some, however, appear 

to be extremely competitive and can invade under 

conditions of low stress and low disturbance (Mitchell 

2000). The presence and spread of nonnative species 

is thought to be an indicator of ecosystem health 

(Mitchell 2000), but even communities in good condition can be susceptible to weed invasions (Sheley 

and Petroff 1999).

Several hypotheses have been proposed in an 

attempt to explain why invasions take place (see 

Masters and Sheley 2001 and references therein for a 

review). Invasions seldom occur as a moving front, but 

rather by the establishment of small satellite populations that are often some distance from previously 

known populations (Cousens and Mortimer 1995). 

Noble (1989) theorized that displacement of native 

vegetation could occur either because the invader 

is a superior competitor, or because the invader is 

adapted to novel conditions present at the time of 

the invasion. Climatic change (for example, global 

warming or increased CO2 levels) may constitute 

such novel conditions and contribute to the spread of 

some invasive species (Brock 1998). Lonsdale (1999) 

lists three factors that affect the ability of an alien 

species to invade a new environment: (1) the number 

of propagules entering the new environment, (2) 

the physiology of the invading species, and (3) the 

invasibility of the environment.

Invasibility of a given environment depends on a 

multitude of factors, including degree of disturbance 

and the health and makeup of the resident community 

(Lonsdale 1999). Each community appears to have 

its own invasibility criteria. Davis and others (2000) 

recently proposed a general theory of invasibility. They 

hypothesize that resources required by an invading 

species are not uniformly available through time, but 

rather fluctuate. These fluctuations may be caused 

by meteorological events and/or site disturbances 

such as grazing, pests, or mechanical disturbance. 

Therefore, the degree to which a community is 

susceptible to invasion varies through time, and 

invasion events are episodic. A number of modeling 

studies have attempted to predict where invasions 

will occur geographically, some with a fair degree of 

success (Peterson and Vieglais 2001).

Management

No universal prescription exists for managing invasive weeds that grow on Forest Service grasslands. 

Management plans take into consideration the biology 

of the weed, the number of plants involved, and the 

condition of the plant community. A single weed that 

has not yet flowered can be eradicated using several 

methods. But if that individual plant is allowed to set 

seed, eradication becomes more complex and will likely 

require more than one treatment. Without treatment, 

those seeds will produce a population, and as the seed 

bank and number of plants increases, eradication 

becomes exponentially more difficult. At some point 

in the growth of the population, eradication becomes 

impossible and alternative management methods 

become necessary (Brock 1998).

To provide guidance for appropriate management 

and control strategies, weeds are assigned an A, B, or 

C classification (Lee 1999). These class assignments 

are based on the distribution and population size of an 

invasive weed within a State or other land management area. Class A species have limited distribution 

within a management unit. Preventing new outbreaks 

and eliminating existing populations is the primary 

focus of management plans for class A weeds. Invasive 

weeds that are not present within a management unit 

but are found in adjacent areas, and therefore pose 

an invasive threat, are also listed as class A weeds. 

Class B weeds have well-established populations, but 

these populations are found only in limited areas. 

Containment within the current population location 

and preventing the establishment of new populations 

is the management focus for class B weeds. Class C 

weeds are widespread throughout a State or management area. Class C weeds are candidates for long-term 

management and suppression programs (Lee 1999).

The 1970s saw an increase in concern about the presence, expansion, and difficulty in controlling noxious 

and invasive weeds. These concerns—coupled with 

the increased costs of weed control, concerns about 

herbicide use, and general complexity involved in trying 

to control weeds on public and private lands—led to 

a new concept for the management of weeds. Termed 

Integrated Weed Management (IWM), this concept 

incorporates multiple management techniques into an 

integrated and well-planned strategy for management 

and control of weeds (Sheley and others 1999).

Walker and Buchanan (1982) define IWM as the 

application of technologies in a mutually supportive 

manner and selected, integrated, and implemented 

with consideration of economic, ecological, and sociological consequences. IWM involves several components: 

(1) prevention, (2) early detection and eradication of 

new weed populations, (3) containment and treatment 
40 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

of established invasive weed populations, and (4) revegetation and site rehabilitation. A successful IWM 

plan includes education, a constant sustained effort, 

evaluation of results, and improvement of management strategies as the plan is implemented (Lee 1999, 

Masters and Sheley 2001, Sheley and others 1999). 

Prior to the development of an IWM plan, survey and 

mapping of the management area must be done to 

identify existing invasive weed populations.

Prevention has two aspects. The first is to limit the 

movement of plant propagules to new locations. Some 

measures that limit propagule migration include the 

cleaning of seeds from equipment and vehicles, use of 

weed-free livestock feed, and efforts to prevent seed 

movement by other human activities. The second 

is the ongoing monitoring of the landscape so that 

new satellite populations are identified quickly and 

eradicated.

Management options for the treatment of existing 

weed populations fall into four basic categories: (1) 

chemical, (2) biological, (3) mechanical, and (4) cultural (Lee 1999, Masters and Sheley 2001, Sheley and 

others 1999). Chemical treatment involves the use of 

herbicides. A broad range of herbicides—registered for 

use on rangeland and grasslands—act upon various 

parts of the plant and at different life cycle stages of 

growth. It is extremely important to use the appropriate 

herbicide, one that has the maximum impact on the 

target species and minimizes impact on nontarget species (see Bussan and others 2001, Masters and Sheley 

2001 for a review of herbicides and their uses).

Biological treatment uses living organisms to reduce 

weed populations. Alien plant species generally lack 

their complement of population-controlling insects and 

pathogens (Mitchell 2000). Biological control specialists 

collect potential control agents (insects or pathogens) 

from the country where the weed originated. After 

extensive testing, these are released into infested areas 

with the goal of suppressing the alien weed population. Eradication is not the goal of biological control, 

but rather the reduction of the population to more 

acceptable levels (Masters and Sheley 2001).

Mechanical (also called physical) treatment involves 

removal of aboveground plant parts and/or disruption of 

the root system such that the plant is killed or severely 

injured (Masters and Sheley 2001). Methods include 

hand pulling, tilling, plowing, and mowing. As with all 

forms of treatment, the method used must be appropriate to the biology of the target species. For example, 

tilling a weed that reproduces vegetatively will spread 

the infestation further, whereas tilling an annual weed 

prior to seed set may provide good control.

Cultural practices include all methods to promote 

the growth of desired vegetation. Healthy native plant 

communities are less susceptible to weed invasion. 

Cultural practices include fertilization, alteration of 

grazing practices, reseeding, revegetation, and other 

practices that promote the growth of desirable plant 

species (Masters and Sheley 2001). Cultural practices 

go hand in hand with other forms of treatment. When 

invasive weeds are removed from a community, steps 

can be taken to facilitate the establishment of desirable 

vegetation and prevent weeds from filling the recently 

vacated niche in the plant community (Masters and 

Sheley 2001).

Fire, both wildfire and prescribed, can impact weed 

populations either positively or negatively. Although 

fire can reduce the population of some weed species, 

the disturbance caused by fire provides opportunities 

for weed population expansion (Crawford and others 

2001). The effect of fire on invasive species populations 

in the Southwest is the subject of ongoing research.

USDA Forest Service Southwestern Region

The Southwestern Region (Region 3) of the USDA 

Forest Service has a variety of grassland communities—from National Grasslands to high elevation 

montane grassland and meadow communities, with 

a number of grassland community types in between. 

All grassland communities in the Region are subject 

to invasion by weeds. Presently, the primary weed 

management focus of Region 3 is the preparation of 

National Environmental Policy Act (NEPA) documents, 

which are required prior to the formulation of weed 

management plans (D. Parker, personal communication).

The Region currently lists 38 species on its noxious 

weed list (table 3-1); 23 are classified as class A, five 

as class B, and 10 as class C. The total of 15 B and 

C species is a relatively low number compared to the 

number of well-established invasive weed species in 

States adjacent to the Region (R. Lee, personal communication). This fact, coupled with the list of 23 class 

A species, illustrates that there is still an opportunity 

to prevent many species of weeds from becoming well 

established in the Region. Prevention is the least expensive and most efficient form of weed management. Cox 

(2001, also available at http://web.nmsu.edu/~kallred/

herbweb/newpage3.htm) provides a detailed list of the 

alien plant species of New Mexico. In addition, the 

Cooperative Extension Services of Arizona and New 

Mexico maintain current State lists of invasive species 

that can be accessed online.

Summary__________________________

Distributions of grasslands are regulated by soils 

and climate and modified by disturbance. A shifting 

equilibrium typically exists between grasslands, 

deserts, and shrublands of the Southwest, such that 

changes in the severity or frequency of disturbance 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 41

Table 3-1. Noxious weed list, USDA Forest Service Southwestern Region (Region 3). Life cycle classification; A=annual, 

B=biennial, P=perennial. Class refers to Region 3 management priority classification (see text).

Scientific Name Common Name Origin Life Cycle Class

Aegilops cylindrica Host Jointed Goatgrass So. Europe A C

Alhagi pseudalhagi Medicus. Camelthorn Asia A C

Asphodelous fistulosus L. Onion Weed Mexico P A

Cannabis sativa L. Marijuana Asia A C

Cardaria chalepensis (L.) Hand. –Maz. Lens-podded Hoary Cress Eurasia P A

Cardaria draba (L.) Desv. Whitetop/Hoary Cress Eurasia P A

Cardaria pubescens (C.A. Mey.) Jarmolenko Globe-potted Hoary Cress Eurasia P A

Carduus nutans L. Musk Thistle So. Europe B B

Centaurea calcitrapa L. Purple Starthistle Europe A A

Centaurea diffusa Lam. Diffused Knapweed Eurasia P A

Centaurea maculosa auct. non Lam.1 Spotted Knapweed Eurasia P A

Centaurea melitensis L. Malta Starthistle Europe A B

Centaurea repens L. 2 Russian Knapweed Eurasia P A

Centaurea solstitialis L. Yellow Starthistle Europe A A

Cirsium arvense (L.) Scop. Canada Thistle Eurasia P A

Cirsium vulgare (Savi) Ten. Bull Thistle Eurasia B C

Convolvulus arvensis L. Field Bindweed Europe P C

Dipsacus sylvestris Huds.3 Teasel Eurasia B A

Drymaria arenarioides Will. Alfombrilla Mexico P A

Eichhornia azurea (Sw.) Kunth Anchored Waterhyacinth Brazil P A

Euphorbia esula L. Leafy Spurge Eurasia P A

Halogeton glomeratus (Bieb.) C.A. Mey. Halogeton Asia A B

Hydrilla verticillata (L.F.) Royle Hydrilla So. Africa P A

Hyoscyamus niger L. Black Henbane Europe B A

Isatis tinctoria L. Dyer’s Woad Europe B A

Kochia scoparia (L.) Schrad.4 Kochia Asia A A

Linaria dalmatica (L.) P. Mill. Dalmation Toadflax Europe P A

Linaria vulgaris P. Mill. Yellow Toadflax Eurasia P A

Lythrum salicaria L. Purple Loosestrife Europe P A

Onopordum acanthium L. Scotch Thistle Europe B A

Peganum harmala L. African Rue No. Africa P B

Rorippa austriaca (Crantz) Bess. Austrian Field Cress Eurasia P B

Salsola iberica (Sennen & Pau) Botsch. Russian Thistle Russia A C

ex Czerepanov5

Sonchus arvensis L. Perennial Sowthistle Eurasia P A

Sorghum halepense (L.) Pers. Johnson Grass Mediterranean P C

Tamarix ramosissima Salt Cedar Eurasia P C

Tribulus terrestris L. Puncture-vine So. Europe A C

Verbascum thapsus L. Mullein Asia B C

1

= C. biebersteinii DC. 2

= Acroptilon repens (L.) DC. 3

= D. fullonum L. 4

= Bassia scoparia (L.) A.J. Scott 5

= S. tragus L.
42 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

events (such as grazing, fire, or drought) can cause a 

change from one vegetation type to another. Humans 

have manipulated grassland vegetation for thousands 

of years through the use of fire, livestock grazing, and 

other means. It is therefore difficult to separate human 

influence from that of climate and other factors in the 

formation and maintenance of these ecosystems.

Southwestern grasslands are generally characterized by low to intermediate rainfall, a long dry season, 

seasonal extremes of temperature, dominance of 

grasses, and large grazing mammals and burrowing 

animals. Because some Southwestern grasslands 

developed under grazing by large herbivores, they are 

generally tolerant of grazing. Other Southwestern 

grasslands are thought not to have had a long evolutionary history of grazing and might be more susceptible 

to grazing disturbance.

Soils vary considerably for grasslands in the 

Southwestern Region because of different climate, 

vegetation, topography, and parent materials. Small 

organisms that exist in grassland soils include bacteria, 

fungi, algae, and nematodes. These soil organisms have 

a profound effect on essential ecosystem processes such 

as decomposition, nutrient cycling, and maintenance 

of soil fertility. Arthropods also perform vital ecological functions in grassland ecosystems. Some of their 

contributions include soil aeration, seed dispersal, 

and plant pollination, in addition to facilitating the 

decomposition of organic debris. Invertebrates also 

provide an important prey base for wildlife.

The majority of plants in Southwestern grasslands 

form some kind of mutualistic relationship with mycorrhizal fungi. These range from plant species that show 

a negative response to inoculation with mycorrhizal 

fungi, to those that show dramatic increases in growth 

and reproduction. Because of this differential response, 

the quantity and type of mycorrhizal fungi present in 

the soil can affect plant community interactions by 

influencing the relative fitness and competitive ability of mycotrophic plant species. Biological soil crusts 

also play an important role in the Southwest. They 

stabilize soil surfaces, increase soil fertility through the 

fixation of atmospheric nitrogen, and improve seedling 

establishment. The cover of biological soil crusts has 

been positively correlated with diversity of plant species. There is no credible evidence that plant cover and 

biological soil crusts are negatively related.

Biological invasions disrupt natural ecosystems, 

posing negative consequences for both ecological and 

socioeconomic systems. They displace desirable native 

species, reduce the quality of wildlife habitat, damage 

sensitive riparian and watershed areas, and increase 

erosion. As native vegetation becomes displaced, alterations in ecological processes occur including changes 

in vegetation, soil microflora, nutrient cycling, and 

decomposition rates.

No universal prescription exists for the management of invasive weeds growing on Forest Service 

grasslands. Weed management plans currently being 

written by the Forests of the Southwestern Region will 

provide guidelines for addressing the complex problem 

of invasive weeds.

Acknowledgments_________________

We gratefully acknowledge Rex D. Pieper, Will 

L. Moir, Douglas L. Parker, Sandra L. Brantley, 

Carleton S.White, Roy Jemison, and Mary E. O’Dea 

for contributing comments, reviews, and information 

to this chapter. 

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 49

What is Biodiversity?_______________

Concern over accelerating extinction rates and loss 

of species diversity on a global scale was the subject 

of E.O. Wilson’s seminal volume Biodiversity (Wilson 

1988). This work essentially transformed the term 

“biodiversity” into a household word as a short-hand 

for species diversity—or more simply, the full array 

and variety of living organisms on Earth. But the term 

biodiversity means much more than the complement 

of plant and animal species that one expects to find in 

some given area. The term now encompasses not only 

the diversity of species, but their genetic structure, the 

interaction of the biotic and abiotic components of the 

environment at the ecosystem level, and at an even 

higher level the array of communities and ecosystem 

processes and functions that make up the landscape 

or regional level of biological diversity. Such a detailed 

exploration of biodiversity is beyond the scope of this 

section.

In the interest of (relative) brevity, the following 

discussion will focus on the more limited definition 

of biodiversity, concentrating on the variety of plant 

and animal species of the New Mexico and Arizona 

grasslands.

A few points of clarification on terminology: The 

most common indices of species diversity (such as 

Simpson index, Shannon-Weiner index) are a function 

of two parameters: (1) species richness, or quite simply 

the number of species that occur in a defined area; 

and (2) species evenness, a measure of the extent to 

which the individuals of the different species present 

are equally abundant. Diversity is thus a measure 

of species richness weighted by relative abundance. 

Technically the most highly diverse communities 

are those with the greatest species richness, each 

component species being equally abundant within the 

community. In reality, such a situation is unlikely 

to occur, as many species (such as top predators) 

are naturally less numerous in the community. 

Furthermore, from a management standpoint it is 

often the more rare species in the community that 

are of interest, rather than the common species that 

tend to dominate weighted indices of diversity. The 

term “biodiversity,” then, as used by many biologists, 

most often refers to the more simple measure of 

species richness, and that is how the term is used 

in this discussion. Biodiversity as used here refers 

to natural or native biodiversity. The distinction is 

important because the introduction of exotic plants or 

animals may technically increase diversity (at least 

in the short term) by adding to the overall species 

richness of a given area. However, as discussed later, 

introduced species usually interfere with normal community and ecosystem functions at some level and 

often eventually replace the native species, thereby 

leading to a net loss of native biodiversity.

Chapter 4:

Biodiversity, Functional Processes, and the 

Ecological Consequences of Fragmentation 

in Southwestern Grasslands

Michele Merola-Zwartjes
50 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Why is Biodiversity Important?

There are two main schools of thought on this 

subject. One theory is commonly referred to as the 

“rivet hypothesis” (Ehrlich and Ehrlich 1981). This 

hypothesis proposes that each species plays some small 

but important role in the ecosystem – like the rivets 

that collectively hold an airplane together. After some 

number of species (rivets) are lost, a critical threshold 

is crossed and the system (plane) rapidly begins to 

disintegrate. The fundamental assumption is that 

greater diversity (more rivets) results in increased 

stability of the system.

The other theory may be called that of “functional 

redundancy.” This idea holds that communities comprise a few functional groups of ecologically equivalent 

species, to the effect that several species may be lost 

with little impact because there are several “back-up” 

species in place that are capable of carrying out the 

same function in the ecosystem (Walker 1992).

Although sources of great debate, there is not 

a great deal of hard data to back up either theory. 

Studies in grasslands to date, however, tend to support the rivet hypothesis. Grassland plots with greater 

species diversity were found to be more resistant to 

drought and to recover more quickly than less diverse 

plots (Tilman and Downing 1994). Another grassland 

experiment showed that plant productivity and soil 

nitrogen utilization both increased significantly as a 

function of plant species diversity, leading the authors 

to conclude that “the establishment and functioning of 

these grassland ecosystems depended on their species 

richness” (Tilman and others 1996:720).

How, then, to manage for biodiversity in grassland 

ecosystems? With so little evidence, the assumption of 

functional redundancy seems somewhat reckless. Even 

if there is some degree of ecological overlap between 

species, we have no idea if the “backup” system is as 

efficient as the primary one (Odum 1992). The most 

prudent course of action in Southwestern grasslands 

– or any ecosystem—is to ensure the ecological integrity of the system by managing for the conservation of 

maximal native biodiversity, or, in the oft-quoted words 

of Aldo Leopold, “to save every cog and wheel”—or, to 

see the context of his remarks (Leopold 1953):

The last word in ignorance is the man who says 

of an animal or plant: “What good is it?” If the land 

mechanism as a whole is good, then every part is good, 

whether we understand it or not. If the biota, in the 

course of aeons, has built something we like but do 

not understand, then who but a fool would discard 

seemingly useless parts? To keep every cog and wheel 

is the first precaution of intelligent tinkering.

Biodiversity in the Southwest________

Although at first glance it might seem that the 

climatic extremes of the desert Southwest would 

prevent many species from being successful in this 

environment, in fact the opposite is true. The highly 

variable precipitation, temperature extremes, and 

elevational gradients created by the basin and range 

topography have provided numerous opportunities 

for adaptation and evolution. Small-scale variations 

in soils, aspect, and moisture all affect microclimate 

and resource conditions, and this variety of conditions 

in turn offers a diverse array of niches available for 

exploitation, resulting in increased floral and faunal 

species richness. Species diversity in the Southwest 

has been further enhanced by its biogeographic history. 

The flora and fauna of distinctive regions historically 

isolated during the Pleistocene (for example, Great 

Plains, Chihuahuan Desert, Sonoran Desert, Mojave 

Desert, and Great Basin Desert) have converged following glacial retreats, leading to the characterization 

of the desert Southwest as a “biological melting pot” 

(Parmenter and others 1995). For Southwestern grasslands in particular, biodiversity is further enhanced 

by the complex, mosaic nature of their distribution. 

Grasslands in this region are interspersed with shrubs, 

woodlands, and riparian areas, leading to high species 

diversity due to the presence not only of grassland 

specialists, but also more generalist species from 

adjacent habitats that may utilize the grasslands for 

other purposes (Parmenter and Van Devender 1995). 

Highly patchy interspersed habitats, such as sand 

dunes, playa lakes, and lava flows, add to biodiversity 

by the presence of specialized and/or endemic plant 

and animal species distinctive to these areas (such 

as fringe-toed lizards in sand dunes) (Parmenter and 

others 1995). Brown and Kodric-Brown (1995) argue 

that biodiversity along the Arizona and New Mexico 

borderlands may be among the greatest on the continent 

due to the biogeographic confluence in this region.

Underlying the considerable biodiversity of the 

Southwest at the level of species richness is a high 

degree of genetic variability. Rangeland plants as a 

group exhibit high levels of genetic diversity, probably 

due to a combination of adaptation to diverse ecological conditions and frequent hybridization between 

interfertile species (Nevo and Beiles 1989, Wayne and 

Bazzaz 1991). Modern grassland plants have survived 

historical cycles of great climatic variation; high levels 

of genetic variability have given these plants the 

ability to adapt and persist throughout such oscillations (Tausch and others 1993). Genetic diversity is 

also high in Southwestern animals, both vertebrates 

and invertebrates. Pocket gophers (Geomydiae), for 

example, are common mammals in Southwestern 

grasslands. Although the species often are difficult to 

distinguish visually, they are some of the most genetically variable mammals known; species may differ not 

only in terms of the alleles represented, but in the 

numbers of chromosomes carried as well (Parmenter 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 51

and others 1995, Patton and Sherwood 1983). Despite 

the fact that Southwestern mammals are relatively 

well known, modern laboratory techniques continue to 

identify new species on the basis of genetic differences 

between populations. Within the last few decades 

several new species have been identified in New 

Mexico, including a grasshopper mouse (Onychomys

arenicola), a meadow-jumping mouse (Zapus hudsonius), and a deer mouse (Peromyscus gratis) (Hafner 

and others 1981, Modi and Lee 1984). Similar patterns 

of high genetic variability have been documented in 

Southwestern invertebrate species as well, including 

fruit flies and grasshoppers (Dobzhansky 1944, Rentz 

and Weissman 1980).

Plant Diversity

Plant diversity in the Southwest is high relative 

to that elsewhere in the country (Brown 1982), 

due in large part to the factors discussed above 

(environmental variability, biogeographic history, 

and genetic mixing), as well as the convergence of 

temperate and subtropical species along the border 

with Mexico. In addition to representation of plants 

from different regions, the Southwest supports a high 

number of endemic species. In the grasslands of the 

Chihuahuan Desert, for example, two-thirds of the 

grass species may be considered endemics (Burquez 

and others 1998). The presence of not only grasses, 

but forbs, shrubs, and occasional trees as well, adds 

to the diversity of the grassland community (Burgess 

1995, McClaran 1995). The richness of plant species 

present is considered to be one of the most important 

indicators of overall rangeland health (West 1993). 

Diversity in plant communities reflects much more 

than just the variety of species present; the array of 

cover values, patchiness, and densities of plants all 

contribute to vegetative diversity (Moir and Bonham 

1995). Species composition is affected not only by 

factors such as soil characteristics, precipitation, and 

topography, but also by the frequency and intensity of 

disturbances such as fire or livestock (Burgess 1995). 

In general the identity of the plant species present 

tends to remain constant over time (not accounting for 

introductions of exotics), but the relative abundances 

of these component species may change dramatically, 

thereby leading to alterations in the physical structure of the grassland system (Westoby and others 

1989). The vegetative architecture of the ecosystem 

is a particularly important consideration, as it is one 

of the key characteristics influencing the use of the 

habitat by both invertebrate and vertebrate animals 

(Parmenter and others 1995).

About 10,000 species of grasses exist worldwide, 

making the grass family Poaceae the fourth most speciose family in the plant kingdom (behind the asters, 

legumes, and orchids; Smith 1993, Watson 1990). If 

one considers grasses in terms of their contribution 

to range quality, Stubbendieck and others (1986) 

suggest 94 species of grasses in their list of the 200 

most important range plants in North America. Some 

of the more common native grasses of the Southwest 

include several species of gramas (such as black 

grama [Bouteloua eriopoda], blue grama [B. gracilis], 

hairy grama [B. hirsuta]), bush muhly (Muhlenbergia

porteri), mesa dropseed (Sporobolus flexuosus), tobosa 

(Hilaria mutica), and Arizona cottontop (Digitaria

californica). Some native grasses have nearly been 

extirpated from their former ranges, such as the giant sacaton (Sporobolus wrightii), a bunchgrass that 

once grew 1 to 2 m high across the floodplains of 

the Southwest. In this climate characterized by low 

rainfall, high rates of evapotranspiration and shallow 

soils, more than 95 percent of the grass production in 

desert grasslands is from C4 species (Sims and others 1978). Desert grasslands are typically composed 

of a mixture of perennial and some annual grasses, 

most of the dominants being perennial caespitose 

bunchgrasses such as blue grama interspersed with 

suffretescent grasses such as bush muhly. Some of 

the common desert grasses are sod forming, such as 

curly mesquite grass (Hilaria belangeri) (Burgess 

1995). These may be interspersed with small trees 

or shrubs such as mesquite (Prosopis spp.), creosote 

(Larrea tridentata), sages (Artemisia spp.), saltbush 

(Atriplex spp.), or rabbitbrush (Chrysothamnus spp.), 

subshrubs such as snakeweed (Gutierrezia spp.), and 

succulents or cacti such as sotol (Dasylirion spp.) or 

prickly pear (Opuntia spp.). The diversity of plant life 

forms that typify desert grasslands today is different 

from the vast monotypic “grasslands” one associates 

with the prairies of the Midwest—so much so, that 

Burgess (1995:58) proposes that a more appropriate 

name for Southwestern desert grasslands would be 

“Apacherian mixed shrub savanna” (historical conditions may have differed, however, as will be discussed 

below).

Herbaceous plants provide much of the ecological 

and botanical diversity in Southwestern grasslands; 

legumes and asters are particularly prominent members of many grassland communities. Members of the 

legume genus Astragalus are common, amongst others, and numerous members of the sunflower family 

may be present (such as Aster, Antennaria, Wyethia, 

Chrysopsis spp.). Other familiar forbs in grassland 

communities include Rocky mountain beeplant 

(Cleome serrulata), various species of flax (Linum

spp.), and penstemons (Penstemon spp.). Cacti and 

succulents are particularly distinctive features of desert 

grasslands. In New Mexico and Arizona, nearly 150 

species of cacti contribute to the plant diversity of the 

Southwestern region. The prickly pear (Opuntia spp.) 

has been proposed as a “keystone resource species” 
52 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

in Southwestern grasslands, a species that provides 

resources during bottlenecks of availability, due to 

the dependence of a variety of animals on prickly 

pears for both food and water during times of drought 

(West and Whitford 1995). Cacti also demonstrate the 

importance of microclimate variation in providing for 

increased species diversity. The rare grama grass cactus 

(Toumeya papyracantha), for example, depends upon 

the favorable microclimate provided by black grama 

grass to survive (Fletcher and Moir 1992, as cited in 

Moir and Bonham 1995).

The variable climate and topography of the 

Southwest contribute to the plant diversity in the 

grasslands of this region. The plant species composition 

and distribution of Southwestern grasslands depends 

heavily upon soil depth and texture, which in turn 

control the water retention potential of the soil (Burgess 

1995). The conservation of desert soils is particularly 

critical, as soils in arid Southwestern grasslands tend 

to be shallow, and more than 30 percent of the available nitrogen and organic matter is in the top 10 cm 

of soil (Charley 1977). This top layer is also highly 

susceptible to erosion, and loss of these soils may lead 

to a decrease in floral diversity and associated faunal 

diversity (Noss and Cooperrider 1994). Soil characteristics are also greatly influenced by the activities of 

invertebrate animals. Soil invertebrates play a critical 

role in decomposition of organic matter, soil development, and alteration of the physical characteristics of 

the soil leading to increased water holding capacity, 

thereby influencing the associated plant community as 

well (Abbott 1989, Hutson 1989, Whitford and others 

1995). The importance of subterranean termites, for 

example, has been demonstrated by the experimental 

elimination of this group on study plots, resulting in 

dramatic changes in both the species composition and 

productivity of grasses, forbs, and shrubs (Whitford 

1991). Herbivorous vertebrates also have a strong influence on plant community composition, structure, 

and productivity, particularly fossorial rodents such as 

kangaroo rats and prairie dogs. The activities of these 

burrowing rodents aerate the soil, modify soil nutrient 

levels, and enhance moisture retention (Parmenter 

and Van Devender 1995). As moisture availability 

is probably the single greatest factor limiting plant 

diversity in the Southwest (Burgess 1995), many of the 

vertebrate and invertebrate animals of grasslands are 

making a critical ecological contribution to maintaining 

the overall biodiversity of these systems through their 

impacts on soil characteristics.

Beyond aerating the soil, cycling nutrients, and 

creating pockets of moisture retention through their 

subterranean activities, grassland rodents such as 

kangaroo rats have yet another significant impact on 

plant diversity. Herbivores generally promote plant 

diversity by suppressing more vigorous species that 

might otherwise exclude other members of the community, thereby allowing less competitive species to persist 

in the system (Burgess 1995). In desert grasslands, 

the abundance of annual grasses and forbs increases 

in proximity to kangaroo rat mounds (Andersen and 

Kay 1999). Furthermore, kangaroo rats have been 

demonstrated to actually control vegetative species 

composition and structure through selective seed 

predation and soil disturbance. A Chihuahuan desert 

shrubland was converted to grassland following exclusion of kangaroo rats from the site; both annual and 

perennial grasses increased in density up to three-fold 

in the absence of these animals (Brown and Heske 

1990).

Plant diversity is also strongly affected by the 

presence or absence of mycorrhizal fungi. Mycorrhizal 

fungi have been proposed as “critical link species” in 

Western grasslands (West and Whitford 1995). Critical 

link species are species that play an important role in 

ecosystem function, but are not necessarily considered 

keystone species (Westman 1990). Approximately 90 

percent of vascular plants are believed to depend upon a 

mutualistic association with mycorrhizae for enhanced 

phosphorous uptake; plant establishment or growth 

may be severely inhibited in the absence of the appropriate fungus. Studies of sagebrush steppe invaded 

by the exotic cheatgrass (Bromus tectorum) found that 

the repeated fires carried by the cheatgrass led to the 

widespread elimination of soil mycorrhizae, thereby 

inhibiting the reestablishment of perennial grasses and 

shrubs (Wicklow-Howard 1989). Granivorous rodents 

again play an important role by dispersing mycorrhizal 

spores throughout the grassland system through seed 

transport (Parmenter and others 1995).

Invertebrate Diversity

When most people think of animal biodiversity, it 

is the larger, more conspicuous animals characteristic 

of the terrestrial ecosystem that come to mind—birds, 

mammals, reptiles, or amphibians. In virtually all 

ecological systems, however, it is the invertebrate 

animals that not only account for the vast majority of 

species diversity and animal biomass, but also make 

the greatest contribution to key ecosystem processes 

such as nutrient cycling. Grasslands are no exception 

to this rule. Referring to shortgrass prairie systems, 

Arenz and Joern (1996:91) dub invertebrates as “the 

most significant contributor to the diversity of the 

prairie system.” The contribution of invertebrates 

to biodiversity is hardly surprising, as insects alone 

compose approximately 90 percent of all terrestrial 

animal species, and fewer than 10 percent of these have 

even been identified and named (Gaston 1991). What 

invertebrates lack in size they make up for in sheer 

numbers; Lauenroth and Milchunas (1992) estimate 

the total biomass of arthropods on North American 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 53

grasslands exceeds that of vertebrates if domestic 

livestock are excluded.

The invertebrate inhabitants of grasslands begin 

at the microscopic level in the soils with protozoans 

and nematodes. Protozoans are found in most any 

grassland with sufficient soil moisture. Although this 

requirement might seem to exclude them from the 

grasslands of the desert Southwest, this is hardly 

the case. Just a tiny amount of moisture trapped between soil particles is sufficient to sustain them. One 

study of the semiarid shortgrass prairie of Colorado 

found more than 20,000 protozoans in each gram of 

dry soil (Elliott and Coleman 1977). Soil protozoans 

feed on bacteria, yeasts, algae, and fungal mycelia 

(Curry 1994), and play an important role in the 

transformation of soil organic matter (Elliott and 

Coleman 1977). Nematodes are a highly diverse group 

of roundworms that also play a key ecological role in 

grasslands. Although their distribution and activity 

is also restricted to some degree by moisture availability, at temperate grassland sites a square meter 

of soil will yield an average of 9 million individuals of 

various nematode species (Sohlenius 1980). In terms 

of ecosystem function, nematodes are considered by 

some to be the most important consumers of energy 

(Scott and others 1979), and their habit of feeding on 

plant roots strongly affects net primary productivity 

in grassland systems (Smolik and Lewis 1982).

Arthropods form a more familiar component of the 

grassland invertebrate fauna and may include whip 

scorpions, crickets, grasshoppers, beetles, flies, bees, 

wasps, cicadas, centipedes, spiders, ants, and termites. 

Some grassland arthropods are less conspicuous due 

to their nocturnal or subterranean habits, a common 

strategy for coping with the heat stress of life in the 

desert (examples include sun spiders, scorpions, termites) (Whitford and others 1995). About 90 percent 

of grassland arthropods reside in either the soil or 

litter, including those that nest belowground or spend 

a significant portion of their life stage belowground 

(Arenz and Joern 1996). The presence of ephemeral 

wetlands in grassland systems provides for a more 

diverse array of arthropods, including mosquitoes, 

freshwater shrimp, and water fleas. If water is present 

long enough, dragonflies or damselflies may be added 

to the invertebrate fauna (Loring and others 1988).

The species diversity of most groups of grassland 

arthropods appears to be higher in the Southwestern 

United States than in the rest of the country (Danks 

1994, Parmenter and others 1995), although data on 

arthropod diversity is relatively limited (Kosztarab and 

Schaefer 1990). Grasshoppers are one of the best known 

and highly diverse groups of grassland arthropods. 

The species diversity of grasshoppers in the Southwest 

is greater than all other Western States, with the 

exception of California (Parmenter and others 1995). 

For example, 30 species of the subfamily Gomphocerinae 

occur in the Southwest, as opposed to only five to 20 

species in the rest of North America (Otte 1981, as 

cited in Parmenter and others 1995). Darkling beetles 

(Tenebrionidae) also reach the peak of species diversity 

in Western arid lands and are one of the major detritivores in Southwestern ecosystems (Crawford 1990). 

Other species rich arthropod groups in the Southwest 

include soil-dwelling mites and collembolans (Crawford 

1990, Zak and Freckman 1991).

Termites (Isoptera) are considered keystone species 

in Southwestern grasslands (West and Whitford 1995). 

A keystone species is one whose ecological impact on 

the community is disproportionately large relative to its 

abundance (Power and others 1996). Although termites 

contribute relatively little to invertebrate diversity in 

the Southwestern grasslands (there are only about 12 

species), they are widely considered to be one of the most 

important invertebrates in these grassland systems 

in terms of their contribution to ecosystem function. 

The biomass of subterranean termites in desert grasslands is estimated to exceed that of domestic livestock 

(Whitford and others 1995). The presence of termites 

has a strong influence on the abundance and species 

composition of soil microfauna, and they are important 

consumers of dead plant material and dung. This latter point is not to be taken lightly; dung decomposes 

so slowly in desert systems that without termites to 

decompose the dung from livestock and incorporate it 

into the soil, grazed desert grasslands would eventually 

be covered with dry dung, leading to reduced plant 

productivity and leading to an overall decrease in the 

carrying capacity of the grassland system (Whitford 

and others 1995). Termites also consume approximately 

50 percent of all photosynthetically fixed carbon in 

desert grassland systems (Whitford and others 1995). 

Termites therefore play a critical role in carbon and 

nutrient cycling in desert soils, and their subterranean 

activities also have a strong effect on soil aeration 

and water infiltration, increasing the water storage 

capacity of the soil. The cumulative ecological impact 

of subterranean termites in grassland systems led 

Whitford and others (1995:181) to declare this group 

the “most abundant and functionally most important 

arthropods in desert grasslands.”

Ants (Formicidae) are one of the more familiar 

arthropods of Southwestern grasslands. Ants play an 

important role in maintaining the plant diversity of 

desert grassland systems by preferentially harvesting 

the seeds of dominant plant species (Whitford and 

others 1995). The subterranean nests of ants are important in grassland systems for concentrating nutrients 

and allowing for increased water infiltration; these 

properties lead to high density vegetation surrounding ant nests in the desert grassland environment 

(Whitford and others 1995). Ant nests may persist in 
54 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

grassland habitats for up to 80 years (Whitford and 

others 1995).

The local species richness of invertebrates in 

grasslands is largely dependent upon the species 

composition, productivity, and habitat structure of 

the plant community (Arenz and Joern 1996, Lawton 

1983, Strong and others 1984). Herbivorous arthropods tend to dominate the invertebrate community 

in grasslands (French 1979). Each plant species tends 

to have its own specialized set of invertebrate herbivores, which in turn support an array of invertebrate 

predators and parasites, and so on. Herbivores make 

up approximately 85 percent of the arthropod biomass 

in shortgrass steppe (Lauenroth and Milchunas 1992), 

and spiders are significant secondary consumers in 

these systems (Schmidt and Kucera 1975). Even though 

most of the invertebrates in grasslands are herbivorous, 

invertebrates are estimated to consume less than 10 

percent of the live biomass in grassland systems (Chew 

1974). As many arthropods are associated with the 

early and midstages of vegetational succession (Usher 

and Jefferson 1991), the diversity of invertebrates is 

often greatest in areas with a diverse mixture of plant 

species and physiognomies, in conjunction with natural 

disturbances (Samways 1994).

Despite their largely inconspicuous nature, invertebrates form a critical component of Southwestern 

grasslands through their contribution to decomposition and nutrient cycling, increasing soil porosity and 

water infiltration, regulating the growth of soil bacteria 

and fungi, and controlling the availability of mineral 

nutrients for plants (Whitford and others 1995).

Vertebrate Diversity

The American Southwest has been called “one of the 

most biologically diverse regions in the United States” 

when it comes to vertebrate animals (Parmenter and 

Van Devender 1995:196). Many groups of animals reach 

their highest levels of species richness in the country 

along the border with Mexico in southern Arizona, 

New Mexico, and Western Texas. The species richness 

of mammals in the Southwest is rivaled only by that 

of central California; bird species richness reaches its 

peak in the Southwest, southern Texas, and California, 

and the numbers of species of reptiles are higher only 

in eastern Texas (Parmenter and others 1995 and references therein). The only group of vertebrates that 

are not richly represented in the arid Southwest is, 

not surprisingly, aquatic organisms; the species richness of amphibians and fishes reaches its peak in the 

Southeastern States (Parmenter and others 1995).

Once again it is environmental heterogeneity created through a combination of elevational variability, 

climate dynamics, and the natural mosaic pattern of 

desert grassland habitats, in conjunction with the 

biogeographic history of the region, that accounts for 

this faunal richness. The greatest diversity of animals 

is usually within the more temperate conditions found 

at intermediate elevations. Animals at high elevations 

tend to be limited by cold temperatures, while those 

at low elevations are limited by aridity (Parmenter 

and others 1995). Although how the merging of formerly isolated faunal regions has contributed to the 

present diversity of Southwestern animal species was 

mentioned briefly above, a more concrete example of 

this blending offered by Parmenter and others (1995) 

may help to clarify how this process has acted to enhance 

vertebrate diversity in the Southwest (table 4-1).

The vertebrate diversity of grasslands is enhanced 

by the collective representation of taxa from many 

convergent habitats. Not only are grassland specialists 

represented, but numerous other animals that may 

Table 4-1. An example of how the rich assemblage of terrestrial 

vertebrate species present today in Southwestern grasslands 

is in part the result of biogeographic history. Formerly distinct 

faunas derived from several major geographic regions now 

coexist in this area due to range expansions and the removal 

of physical barriers following glacial retreats at the end of the 

Pleistocene. A few of these species are listed below (adapted 

from Parmenter and others 1995).

Biogeographic region Vertebrate species contributed

 to Southwestern grasslands

 Great Plains Western box turtle

 Great Plains skink

 Black-tailed prairie dog

 Northern grasshopper mouse

 Swainson’s hawk

 Lark sparrow

 Sierra Madre (Mexico) Yarrow’s spiny lizard 

 Rock rattlesnake

 Pygmy mouse

 Montezuma quail

 Chihuahuan Desert Texas horned lizard

 Trans-Pecos rat snake

 Silky pocket mouse

 Banner-tailed kangaroo rat

 Scaled quail

 Cassin’s sparrow

 Sonoran Desert Collared lizard

 Sidewinder

 Desert kangaroo rat

 Southern grasshopper mouse

 Gila woodpecker

 Bendire’s thrasher

 Mojave/Great Basin Desert Short-horned lizard

 Chisel-toothed kangaroo rat

 Sagebrush vole

 Sage thrasher

 Sage sparrow
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 55

dwell primarily in the surrounding mosaic of desert 

scrub, pinyon juniper, or riparian areas, and that use 

the grasslands for foraging, may add to the species 

richness; this is particularly true of more mobile, 

generalized species of mammals or birds (Parmenter 

and others 1995). Up to 18 species of bats, for example, 

may be found in Southwestern grasslands. While six 

of these species are commonly found in grasslands, 

most of them will utilize grasslands only if their other 

habitat requirements may be met within a reasonable distance, specifically the presence of appropriate 

roost sites and water (Chung-MacCoubrey 1996). 

The interspersion of Southwestern grasslands with 

habitats such as pinyon-juniper woodlands meets 

these requirements.

Reptiles and Amphibians

Herptiles—reptiles and amphibians—are important 

components of the grassland vertebrate community. In 

the Southwest, reptiles in particular make a significant 

contribution to overall diversity. Approximately 44 species of reptiles are associated with desert grasslands, 

with fewer (18 species) present in high elevation 

mountain meadows (Parmenter and Van Devender 

1995). Although the arid Southwest is a challenging 

environment for many amphibians, certain toads are 

relatively common in desert grasslands (such as the 

Western spadefoot toad [Scaphiopus hammondi]), and 

true frogs (such as the Chiricahua leopard frog [Rana 

chiricahuensis]) and tiger salamanders (Ambystoma 

tigrinum) may be found near permanent water 

sources. Common grassland/desert specialists include 

various species of box turtles, spadefoot toads, earless 

lizards, whiptails, horned lizards, bullsnakes, and 

rattlesnakes. Many species associated with grassland 

habitats require specific habitat features, such as rock 

outcroppings to serve as dens for wintering snakes 

(Collins 1982, Hammerson 1986). Many reptiles and 

amphibians join a variety of other animals in taking 

advantage of the beneficial microclimate provided by 

prairie dog burrows. At least 12 species of amphibians and 17 species of reptiles have been reported as 

regularly associated with black-tailed prairie dog towns 

(Reading and others 1989, Sharps and Uresk 1990). 

Spadefoot toads (family Pelobatidae) require playas, 

temporary pools that fill with water during the summer monsoons, for breeding (Corn and Peterson 1996). 

Most Southwestern amphibians have short aquatic 

larval stages to take advantage of the ephemeral water 

sources for breeding. In fact, permanent water sources 

may be detrimental to amphibian populations by attracting predators such as raccoons (Procyon lotor), 

or by encouraging the establishment of exotic species 

such as bullfrogs (Rana catesbiana) and centrarchid 

fishes. These predators have been implicated in the 

disappearance of several native species of amphibians 

in various areas of the country (Collins and others 

1989, Hayes and Jennings 1986).

Both reptiles and amphibians play functional roles in 

food webs both as predators on invertebrates and small 

vertebrates and as prey for larger animals (although 

some herps, such as the desert tortoise, are herbivores). 

Amphibians in particular are an important avenue for 

nutrient transport between aquatic and terrestrial 

systems. Amphibian and reptile populations are also 

particularly sensitive indicators of environmental 

stresses and may thus serve as a warning signal of 

problems such as pesticide contamination (Beiswenger 

1986, Blaustein 1994). Like other grassland animals, 

herps are sensitive to changes in habitat composition 

and structure, and herp species richness is generally greatest in relatively heterogeneous habitats. In 

dense grasslands, for example, moderate grazing that 

increases the patchiness of the grass density and adds 

some variety to its structure enhances the habitat for 

a variety of snakes, lizards, and toads.

One of the more fascinating groups of grassland 

reptiles is the genus Cnemidophorus, the whiptail 

lizards. There are 10 unisexual species of whiptails 

in the Southwest that are made up entirely of female 

individuals. These lizards reproduce by an asexual 

autofertilization method known as parthenogenesis. 

All-female whiptail lines are polyploid (have more 

than one set of chromosomes), indicating that they 

were originally formed through the hybridization of 

two sexual species. Each individual in the species 

is genetically identical to the original hybrid (Cole 

1984, Parmenter and Van Devender 1995). The desert 

grasslands of the Southwestern United States are 

the evolutionary center for this highly unusual group 

of vertebrates; seven of the 10 unisexual species are 

restricted almost entirely to this habitat. The desert 

grasslands of Texas and New Mexico are home to the 

New Mexican (Cnemidophorus neomexicanus), checkered (C. tesselatus), and Chihuahuan (C. exanguis) 

whiptails; desert grassland (C. uniparens), Sonoran (C. 

sonorae), and Gila spotted (C. flagellicaudus) whiptails 

are found in Arizona; and the plateau whiptail (C. 

velox) is found in the grasslands of the Great Basin. 

Furthermore, several whiptail species (New Mexican 

whiptail, desert grassland whiptail, and the little 

striped whiptail C. inornatus—a nonparthenogenetic 

species) are believed to be dependent on native stands 

of grasses within these habitats.

Birds

Although birds are some of the most abundant vertebrates found in Southwestern grasslands (Parmenter 

and Van Devender 1995), the bird community tends 

to be the most simplistic in terms of species richness 

(Knopf 1996). As one example, Knopf (1996) cites 

his results from a series of 112 point count surveys 
56 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

in the Pawnee National Grasslands in Colorado, in 

which just three species (horned lark [Eremophila

alpestris], McCown’s longspur [Calcarius mccownii], 

and lark bunting [Calamospiza melanocorys]) accounted for 87 percent of all individuals recorded; 

only 14 species of native birds were recorded in total. 

Bird species commonly found in the grasslands of 

the Southwest include the horned lark, lark bunting, 

meadowlarks (both Eastern [Sturnella magna] and 

Western [Sturnella neglecta]), scaled quail (Callipepla

squamata), mountain plover (Charadrius montanus), 

burrowing owl (Athene cunicularia), short-eared 

owl (Asio flammeus), prairie falcon (Falco mexicanus), ferruginous hawk (Buteo regalis), and various 

sparrows (such as vesper [Pooecetes gramineus], 

lark [Chondestes grammacus], Cassin’s [Aimophila

cassinii], and Botteri’s [A. botterii]) (Knopf 1996, 

Parmenter and Van Devender 1995). Even some 

shorebirds, such as long-billed curlews (Numenius

americanus), utilize Southwestern grasslands for 

breeding. At high elevations, rosy finches (Leucosticte

spp.) and white-tailed ptarmigan (Lagopus leucurus) 

may be found in alpine meadows of the Southwest 

(Parmenter and others 1995).

Many of the typical avian residents of Southwestern 

grasslands nest on the ground, as the lack of vertical 

structure in grasslands offers little other choice. The 

burrowing owl, a common denizen of Southwest grasslands, nests below the ground in abandoned prairie 

dog burrows. Most members of the grassland bird 

community are granivores, and the dynamics of this 

community are closely tied to levels of seed production. 

As most desert grasses set seed in late summer or early 

fall following the monsoons (McClaran 1995), the peak 

in resident grassland bird numbers usually occurs in 

late summer in coincidence with maximum seed production (Maurer 1985). Avian species richness increases 

during the winter months, when the grasslands of the 

Southwest support a great concentration of migratory 

species such as Baird’s sparrow (Ammodramus bairdii), white-crowned sparrow (Zonotrichia leucophrys), 

grasshopper sparrow (Ammodramus savannarum), 

sage sparrow (Amphispiza belli), Sprague’s pipit 

(Anthus spragueii), McCown’s longspur, and chestnutcollared longspur (Calcarius ornatus).

Birds respond strongly to changes in habitat architecture, and the diversity of the grassland bird 

community will increase in the presence of shrubs, 

trees, cacti, or even human structures (Grinnell 1922, 

Knopf and Scott 1990, Parmenter and others 1995, 

Szaro 1981). The addition of vertical structure provides 

a far greater range of avian habitats and thereby adds 

a whole new component to the bird community. The 

maximum species richness is probably found along 

riparian corridors or near permanent wetlands in 

grasslands, where large numbers of migrants and 

transients concentrate (Parmenter and Van Devender 

1995); here, various species of warblers, vireos, and 

other decidedly nongrassland species may be found. 

It is important to recognize, however, that even if species diversity is technically enhanced by the addition 

of vertical structure such as woody plants or human 

developments, grassland specialists are usually lost 

in the process. The enhanced bird diversity witnessed 

in such cases is most likely provided by an increase in 

relatively common generalist species and might mask 

any concomitant population declines or extirpations of 

narrow endemics that may occur (Knopf 1992).

In pure grasslands without significant vertical structure provided by shrubs or trees, different bird species 

demonstrate preferences for an array of grass heights 

and various patterns of patchiness. Mountain plovers 

and McCown’s longspurs, for example, occur in short 

grasslands, often those that have been subjected to 

“heavy grazing pressure to the point of excessive surface 

disturbance” (Knopf 1996:141 and references therein). 

Lark buntings will use areas of shortgrass prairie but 

require that tufts of taller grasses be interspersed in 

the landscape to provide nest concealment (Finch and 

others 1987). Baird’s sparrow can be found across a 

wide range of grassland types and grazing intensities 

(Kantrud 1981), and Cassin’s sparrow requires grasslands that provide at least 6 percent shrub cover and 

may be lightly grazed (Bock and Webb 1984).

Grassland birds are a source of great conservation 

concern, as this group of birds has shown consistently 

steep population declines over the past few decades, 

on the order of 25 to 65 percent—more than any other 

guild of North American bird species (Askins 1993, 

Knopf 1992, 1996). Formerly widespread and common 

species such as the lark bunting and Cassin’s sparrow 

are showing statistically significant declines, and the 

mountain plover has been proposed for Federal listing 

as a threatened species. Although some theorize that 

declines in populations of neotropical migratory birds 

are due to loss of wintering habitat in the tropics (see 

Briggs and Criswell 1979, Lovejoy 1983, Terborgh 

1980), most of the grassland birds in question are 

short-distance migrants that spend their winters in the 

grasslands of the Southwest and Mexico, suggesting 

that alteration of Southwestern grassland habitats may 

be contributing to the decreases witnessed (DeSante 

and George 1994).

Mammals

The Southwestern grasslands owe much of their high 

vertebrate diversity to mammals, and more particularly to rodents. In a comparison of mammals found 

in six habitat types in a single region of New Mexico, 

desert grasslands had the greatest species richness 

of any major ecosystem type with 56 species, ahead 

of desert scrub, pinyon-juniper woodland, montane 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 57

forest, montane meadow, and riparian zone (Parmenter 

and Van Devender 1995). The high species richness of 

the desert grassland is primarily due to the diversity 

of rodents in this system, especially ground squirrels 

(Sciuridae), kangaroo rats (Dipodomys spp.), and mice 

(Muridae). Rodents tend to be the dominant mammals 

in all desert grasslands, and are well represented by 

grassland specialists, including the bannertail (D. 

spectabilis) and Ord (D. ordii) kangaroo rats, blacktailed prairie dog (Cynomys ludovicianus), and spotted 

ground squirrel (Spermophilus spilosoma) (Parmenter 

and Van Devender 1995). As one indication of just how 

diverse the rodent community of the Southwestern 

grasslands is, Parmenter and others (1995) point out 

that just one 20 ha area of Chihuahuan Desert has 

the same number of native rodent species as the entire 

States of Michigan and Pennsylvania combined, and 

that’s allowing those States two introduced species 

and two semiaquatic species.

Many other mammal species are also characteristic 

of Southwestern grasslands, including, amongst others, pronghorn (Antilocapra americana), white-sided 

jackrabbits (Lepus callotis), swift fox (Vulpes velox), 

badgers (Taxidea taxus), coyotes (Canis latrans), mule 

deer (Odocoileus hemionus), several species of bats, 

and—in high elevation grasslands—pikas (Ochotona 

princeps). Historically, bison (Bison bison) were found 

in the shortgrass prairie regions of eastern New 

Mexico, but probably did not occur regularly in the 

arid grasslands farther west (Berger and Cunningham 

1994, Kay 1994, Mack and Thompson 1982). Many of 

these mammals have a strong impact on the overall 

diversity of the grassland system through various types 

of disturbance. Vegetation structure and species composition, for example, are affected by selective feeding 

of herbivores and by soil disturbance. By selectively 

grazing on dominant species, herbivorous mammals 

allow subdominant plant species to compete and persist 

in the community (Risser and others 1981). Digging 

by badgers, prairie dogs, kangaroo rats, and gophers 

creates soil disturbances that allow for the establishment of annual forbs and grasses, and also increases 

the porosity and water-holding capacity of the soil 

(Benedict and others 1996 and references therein). 

Wallowing by bison in areas where they formerly occurred, and small scrapes created by pronghorn, serve a 

similar function (Benedict and others 1996, Parmenter 

and Van Devender 1995). These small-scale natural 

disturbances add unique microhabitats available for 

colonization by other species, increasing vegetative 

diversity, enhancing the mosaic nature of the habitat, 

and leading to increased faunal diversity as well (both 

invertebrate and vertebrate) (Benedict and others 1996, 

Collins and Barber 1985). Overall, Collins and Barber 

(1985) found that diversity in a mixed-grass system was 

enhanced by moderate levels of natural disturbance 

(sensu the “intermediate disturbance hypothesis” of 

Connell 1978) and that small-scale disturbances have 

an additive effect that further enhances diversity.

Grassland biodiversity is also strongly impacted 

by the presence of keystone species. The power that 

kangaroo rats exert over the structure and dynamics 

of their habitat has led to their designation as a keystone species in the grassland systems of the Southwest 

(Brown and Heske 1990, West and Whitford 1995). 

Kangaroo rats exert their influence largely through 

selective seed predation and soil disturbance. In a longterm study on an Arizona desert shrubland, Brown 

and Heske (1990) demonstrated that the removal of 

kangaroo rats resulted in dramatic increases in grass 

densities, as well as a shift toward large-seeded winter 

annual plant species. Small-seeded winter annuals 

decreased, herbaceous vegetation increased (including 

both grass and forbs), litter accumulation increased, 

seed-eating birds decreased, and several new species 

of rodents colonized the plots where kangaroo rats 

were absent. The other native rodents on the plots 

where kangaroo rats were removed were not able to 

prevent the conversion of the habitat from shrubland 

to grassland, along with the associated changes in the 

resident fauna, thus supporting the keystone role of 

the kangaroo rats in this system. Furthermore, kangaroo rat burrows provide favorable microclimates for 

a diverse array of both invertebrate and vertebrate 

animals. Western box turtles, Great Plains skinks, 

and massasaugas use kangaroo rat mounds for shelter, 

and several species of roaches, crickets, and beetles are 

found almost exclusively in these mounds (Hawkins 

and Nicoletto 1992).

The black-tailed prairie dog is a critically important 

keystone species in Southwestern grassland systems 

whose presence greatly enhances local biodiversity 

(Kotliar and others 1999, Miller and others 1994, 

Whicker and Detling 1988). The burrowing and feeding 

behaviors of prairie dogs have drastic effects on the 

structure, species composition, and nutritive value of 

surrounding vegetation, create open areas to add to 

the heterogeneity of the habitat, modify the physical 

characteristics of soils, affect energy and nutrient 

cycles, and provide valuable microclimates utilized 

as shelters by a multitude of both invertebrate and 

vertebrate animals (Benedict and others 1996, Miller 

and others 1990, 1994, Whicker and Detling 1988 

and references therein). The activities of prairie dogs, 

feeding on and clipping vegetation in the area of their 

colony, stimulates fresh plant growth and enhances the 

nutritional content of the vegetation, leading herbivores 

such as pronghorn to preferentially feed on prairie dog 

towns (Coppock and others 1983). Prairie dog towns are 

considered to be centers of animal diversity due to the 

great numbers of species that converge on the colonies 

to either use the old burrows, forage on the surrounding 
58 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

vegetation, or feed on the prairie dogs themselves 

(Miller and others 1990, 1994). Nearly 170 species of 

vertebrates have been reported using prairie dog towns, 

although this number is undoubtedly excessive due to 

the inclusion of birds flying over (Benedict and others 

1996). A more critical recent review suggests that there 

is sufficient evidence for the strong dependence of 

nine vertebrates upon prairie dog colonies, 20 species 

appear to use the colonies opportunistically, and 117 

species may have some relationship with the colonies, 

but data to support any solid conclusions are lacking 

(Kotliar and others 1999). However, of those animals 

that are closely associated with prairie dogs, several 

are of great conservation concern, including the blackfooted ferret (Mustela nigripes), swift fox, ferruginous 

hawk, burrowing owl, and mountain plover (Kotliar 

and others 1999, Samson and Knopf 1994). This association does not bode well as eradication programs 

have resulted in eliminating the black-tailed prairie 

dog from 98 percent of its former range, reducing its 

numbers to the point that the species is now under 

consideration to be listed as endangered (Miller and 

others 1994, U.S. Fish and Wildlife Service 2001).

Threats to Grassland Biodiversity____

The native biodiversity of Southwestern grasslands 

has been greatly altered through human activities. 

Numerous animal species have been extirpated or 

greatly reduced through direct persecution, including 

the black-tailed prairie dog, Mexican wolf, bison, and 

grizzly bear (Benedict and others 1996; table 4-2). 

Others have been reduced presumably due to their 

dependence on a keystone species that has been removed from the system. For example, the black-footed 

ferret (endangered) and mountain plover (proposed 

for listing) are strongly dependent upon prairie dog 

colonies for survival (Kotliar and others 1999), and the 

Table 4-2. Terrestrial vertebrates of Southwestern grasslands that are now extinct or have been extirpated from the region. The following 

species are not necessarily restricted to grassland habitats; although many are grassland specialists, this list also includes those species 

that rely heavily on grasslands in some parts of their range or as one component of a mosaic of habitats utilized. Sources for the information presented here include Arizona Game and Fish Department (1988), Association for Biodiversity Information (2001), and New Mexico 

Department of Game and Fish (2000).

 Federal Extirpated Primary cause of 

Common name Scientific name status or extinct extinction or extirpation Notes

Western boreal toad Bufo boreas boreas CW Extirpated Unknown Believed extirpated; formerly 

 occurred in alpine meadows

New Mexico Tympanuchus Extinct Habitat loss or degradation 

 sharp-tailed grouse phasianellus hueyi due to overgrazing, agriculture, 

 succession

Sage grouse Centrocercus Extirpated from Overhunting, habitat loss or 

urophasianus NM & AZ degradation from overgrazing

New Mexican Dipodomys Extirpated Habitat degradation from Inhabited Great Basin desertscrub

 banner-tailed spectabilis baileyi from AZ overgrazing 

 kangaroo rat

Black-tailed Cynomys ludovicianus CW Extirpated Direct human persecution Some small populations persist 

 prairie dog from AZ in NM

Black-footed ferret Mustela nigripes E Extirpated Elimination of prairie dogs Currently attempting reintroduction 

 (primary prey) in NM

Bison Bison bison Extirpated from Overhunting Now exist on private ranches

 NM and AZ 

Merriam’s elk Cervus elaphus Extinct Overhunting Native AZ elk

 merriami 

Mexican wolf Canis lupus baileyi E Extirpated Direct human persecution Experimental populations 

 reintroduced in NM and AZ

Intermountain wolf Canis lupus youngi Extinct Direct human persecution

Grizzly bear Ursus arctos T Extirpated from Direct human persecution Persist in Northwestern States (for 

 NM and AZ example, Montana), Canada, and 

 Alaska

E = endangered, T = threatened, CW = candidate with “warranted but precluded” determination
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 59

declines in their populations have been linked to the 

extirpation of prairie dogs (Miller and others 1994). 

Not only does the disappearance of the prairie dog have 

dire consequences for the species dependent upon it, 

but the prairie dogs themselves are now threatened 

with deleterious genetic consequences as a result of 

the fragmentation and isolation of their remaining 

populations (Pizzimenti 1981).

A far more subtle factor has been responsible for 

most extinctions at the local level in Southwestern 

grasslands. Changes in the structure and function of 

grassland habitats have probably been responsible for 

more losses of native diversity than any other cause 

(Stacey 1995). “While losses of biological diversity 

at the local level are often the least noticed,” Stacey 

(1995:34) points out, “they are extremely important 

because they change the functional dynamics of the 

local community and because if local extinctions continue long enough the species will be lost over wide 

areas and may not recover without human intervention.” Changes in grassland habitat structure and 

function may come about in many ways, but some 

of the most important sources of these changes in 

Southwestern grasslands have been the loss of fire 

as a natural cyclical event, the elimination of prairie 

dog colonies, heavy grazing by livestock, the introduction of nonnative grasses, and shrub encroachment 

(Parmenter and Van Devender 1995, Risser 1988). 

The division of formerly expansive rural landscapes 

into increasingly fragmented “ranchettes” is the latest 

recognized threat to grassland biodiversity (Brown and 

McDonald 1995, Maestas and others 2002), and will 

be discussed separately under the section on habitat 

fragmentation.

Loss of Natural Fire Cycles

Fire plays a key role in the maintenance of most 

grassland systems. Without periodic fires, woody 

plants begin to encroach into grasslands, converting 

them to shrublands or woodlands. The grasslands of 

the Southwest are no exception. Many researchers 

agree that historically fires were both common and 

extensive in the desert grasslands, and that these 

fires were instrumental in maintaining the integrity 

of these systems (Bahre 1991, Humphrey 1958, 

McPherson 1995, McPherson and Weltzin 2000).

The exception to this rule may be grasslands dominated by black grama. The extreme difficulty black 

grama exhibits in recovering from a burn indicates that 

this species is not fire-adapted and probably did not 

evolve under a history of frequent burning (Buffington 

and Herbel 1965, Dick-Peddie 1993). More recently, 

however, it has been proposed that the negative effects 

witnessed may have been attributable to a coincident 

period of drought rather than to fire (Curtin and others 2002). Precipitation has a considerable impact on 

grassland productivity following fire, both in terms of 

timing and quantity.

The natural frequency and extent of grassland 

fires in the Southwest are believed to have declined 

dramatically since Euro-American settlement of the 

region in the late 1800s (Bahre 1991, 1995, Humphrey 

1958). A review of the role of fire in desert grasslands 

reveals that the natural frequency of fire in these 

systems was probably on the order of every 7 to 10 

years (McPherson 1995 and references therein). Fires 

occurring on this cycle are believed to be sufficient to 

prevent the establishment of woody plants, by killing 

seeds on the surface and preventing woody plants 

from reaching the age where resprouting is possible 

(McPherson 1995). Although fires eliminate grass 

cover in the short term, in the long term, grasses are 

rejuvenated by the occurrence of fire and benefit from 

the elimination of woody plants. The timing of fires 

is also important. Fire in the early summer, when 

the growth of many perennials is just beginning, can 

negatively impact warm season grasses, whereas these 

same grasses are tolerant of fire during the dormant 

season (McPherson 1995). The level of soil moisture 

at the time of ignition is also a consideration; for some 

plant species, burning on dry soils may be damaging 

(W. Moir, personal communication 2003).

Although many factors contribute to fire regimes, 

perhaps the most important change that has resulted in 

decreased fire frequency and intensity in the Southwest 

is the lack of fine fuels to carry the fires (Humphrey 

1958, McPherson 1995). Historically, the timing of this 

change corresponded with the widespread increase 

in livestock grazing in the Southwest after 1880. At 

this time, stocking rates reached record levels, and 

overgrazing was actually encouraged to reduce the fire 

hazard and encourage the growth of trees (Bahre 1991, 

Leopold 1924 as cited in McPherson 1995). In addition, 

the use of wooden posts for livestock fencing provided 

the incentive for quickly suppressing rangeland wildfires that would compromise the integrity of the fences 

(Sayre 2002). Today, fragmentation from roads and 

suburban developments serve as a kind of artificial 

firebreak to contain the spread of extensive wildfires 

(Bahre 1995, McPherson 1995). The continuing growth 

of residences on formerly undeveloped lands has also 

led to a demand for active fire suppression in these 

areas (Hansen and others 2002). These changes in the 

frequency and intensity of natural fire regimes have 

doubtless contributed to the widespread conversion of 

Southwestern grasslands to shrublands (Archer 1989, 

Brown 1982, Humphrey 1958), thereby radically altering the nature of the habitat for native grassland 

species.
60 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Prairie Dog Eradication

In the Southwestern grasslands, the prairie dog is 

considered a “keystone” species—a species that has 

a large overall effect on a community or ecosystem 

disproportionate to its abundance (Kotliar and others 

1999, Power and others 1996). The activities of these 

burrowing animals have a dramatic impact on the patch 

dynamics and ecosystem function of the grasslands 

that they inhabit. Prairie dog disturbances impact 

the physical and chemical properties of the soil, alter 

vegetational structure, affect plant species composition, 

and improve the nutrient value of plants growing in 

the vicinity of their colonies (O’Meilia and others 1982 

and references therein, Whicker and Detling 1988). 

The increased nutritional value of forage on colonies 

may act to offset any decrease in biomass as a result 

of clipping by prairie dogs (Holland and Detling 1990, 

O’Meilia and others 1982 and references therein). 

Bison, elk, pronghorn, and livestock all preferentially 

graze on prairie dog colonies, presumably because of the 

increased value and palatability of the herbage there 

(Coppock and others 1983, Knowles 1986, Krueger 

1986, Wydeven and Dahlgren 1985).

Active prairie dog towns contribute to increased biological diversity by supporting a different complement 

of species compared to areas unoccupied by prairie dogs 

(Agnew and others 1986, Mellink and Madrigal 1993, 

O’Meilia and others 1982). Furthermore, several vertebrate species are considered highly dependent upon 

prairie dogs either as prey or for the habitat provided 

by their colonies, including the endangered blackfooted ferret. Other animals considered true prairie 

dog associates are the mountain plover, burrowing owl, 

ferruginous hawk, golden eagle (Aquila chrysateos), 

horned lark, swift fox, deer mouse (Peromyscus maniculatus), and northern grasshopper mouse (Onychomys 

leucogaster) (Kotliar and others 1999).

Once a dominant force in the grasslands of the 

Western United States, the ecological impact of the 

prairie dog on these systems has nearly been extinguished. Up until the early 1900s, prairie dog colonies 

were estimated to cover hundreds of millions of acres 

of shortgrass prairie and desert grasslands west of 

the Great Plains (Anderson and others 1986). Today 

prairie dogs are estimated to persist on a mere 2 

percent of their former range (Anderson and others 

1986, Miller and others 1994). One species, the Utah 

prairie dog (Cynomys parvidens), is endangered, and 

the black-tailed prairie dog, formerly the most abundant and widespread of the five species of prairie dogs 

in North America, is a candidate for listing (USFWS 

2000). The population numbers of the black-tailed 

prairie dog are estimated to have been reduced by 98 

percent, and the species may occupy as little as 0.5 

percent of its original range (Mac and others 1998 

as cited in USFWS 2000). Although the conversion 

of native prairie habitat to other land uses may have 

contributed to some degree, undoubtedly the greatest 

single factor in the loss of prairie dogs has been the 

concerted effort by both Federal and State government 

agencies to exterminate these animals for the benefit 

of the livestock industry (Mulhern and Knowles 1996, 

Parmenter and Van Devender 1995).

The campaign to eradicate prairie dogs from Western 

grasslands began in earnest following the release of a 

Department of Agriculture report suggesting that the 

presence of prairie dogs may reduce range productivity 

by 50 to 75 percent (Merriam 1902). The U.S. Biological 

Survey responded with a massive poisoning campaign 

under the auspices of its Predator and Rodent Control 

program. Aiming to reduce competition with livestock, 

millions of acres of prairie dog colonies were poisoned, 

and shooting of prairie dogs was encouraged across 

their range (Bell 1921, Mulhern and Knowles 1996, 

Parmenter and Van Devender 1995, Van Pelt 1999). 

Fear of sylvatic plague buoyed these efforts after the 

bacterium was discovered in black-tailed prairie dogs in 

Texas in the 1940s (Cully 1989, Mulhern and Knowles 

1996). In some States, annual extermination of prairie 

dogs on State and privately owned lands was a legal 

requirement. Nebraska, for example, only recently 

repealed this mandate in 1995 (Van Pelt 1999).

The black-tailed prairie dog and the Gunnison’s 

prairie dog (C. gunnisoni) are the two species that 

inhabit the grasslands of Arizona and New Mexico. 

Described as occurring in “immense colonies” in Arizona 

in 1885 (Mearns 1907), the Arizona prairie dog (C.l. 

arizonensis), a subspecies of the black-tailed prairie 

dog, was largely extirpated from that State by 1938; 

one small single colony survived until 1960 (Van Pelt 

1999). In New Mexico, the range of the black-tailed 

prairie dog has been reduced by at least 25 percent 

(Hubbard and Schmitt 1984). In the Animas Valley, for 

example, biologists from the Museum of Southwestern 

Biology did not observe one single prairie dog between 

the years 1955 and 1972 (Findley 1987). Yet in 1908 

Vernon Bailey had described this same area as an 

almost continuous prairie dog town for its length and 

breadth, estimating that over 6 million prairie dogs 

inhabited the valley (Bailey 1932).

Notwithstanding the drastic declines already witnessed in prairie dog numbers and the evidence of a 

cascade effect on other species, prairie dogs today are 

still widely considered to be vermin and enjoy little in 

the way of legal safeguards from any states (Van Pelt 

1999). This is in spite of more recent evidence that the 

level of competition between prairie dogs and livestock 

is more likely on the order of 4 to 7 percent (Uresk and 

Paulson 1988 as cited in Miller and others 1994) and 

that there is no significant difference in the market 

weight of steers whether they graze in conjunction 

with prairie dogs or not (O’Meilia and others 1982; 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 61

although it should be noted that the statistically insignificant weight difference did result in an economic 

loss). Following their comprehensive review, Kotliar 

and others (1999:186) concluded that prairie dogs are 

“crucial to the structure and function of native prairie 

systems.” Not only are scientists today stressing the 

importance of preserving remaining prairie dog colonies to maintain biodiversity (for example, Miller and 

others 1994), some are going further and calling for 

the reintroduction of prairie dogs to restore ecosystem 

function (for example, Manzano-Fischer and others 

1999).

Overgrazing by Livestock

Livestock grazing is the predominant land use in the 

Western States. More than 70 percent of the land area in 

the West (11 states, from Montana, Wyoming, Colorado, 

and New Mexico westward) is grazed by livestock, predominantly cattle, including wilderness areas, wildlife 

refuges, National Forests, and some National Parks 

(Fleischner 1994 and references therein). Some argue 

that overly heavy levels of livestock grazing are one 

of the greatest sources of habitat degradation in the 

West (for example, Noss and Cooperrider 1994), leading 

to widespread declines in the native wildlife of North 

American grasslands (Fleischner 1994). Others point 

out that most studies of grazing effects have suffered 

from poor experimental design (for example, Brown 

and McDonald 1995, Jones 2000), or have found the 

impact of grazing to be relatively negligible on factors 

such as native species richness (Stohlgren and others 1999). The issue of livestock grazing in the West 

is highly contentious. In all cases, it is important to 

remember that the impact of grazing will vary greatly 

depending upon any number of variables, including 

the season of use, stocking rate, environmental conditions, and the evolutionary history of grazing in the 

region (Fleischner 1994, Jones 2000, Milchunas and 

Lauenroth 1993), and may also differ according to 

geographic scale (Stohlgren and others 1999).

This discussion will focus primarily on how poorly 

managed grazing of livestock may impact grassland 

habitats in the Southwest.

Grazing by livestock has the potential to alter grassland habitats in many ways. Depending on the intensity 

and length of the grazing regime and environmental 

conditions, livestock activities may significantly alter 

plant species composition, extent of vegetative cover, 

and physical structure of the habitat (Bock and others 1984). As discussed above, any changes in these 

vegetative parameters exert a strong influence on the 

associated fauna, so that changes in plant diversity 

and structure result in changes in animal diversity. 

Most frequently, overgrazed sites result in a loss of 

specialized native fauna and may or may not exhibit 

an increase in more widespread, generalist species 

(Bock and others 1984, Bock and Webb 1984, Jones 

1981). In an Arizona grassland, for example, heavily 

grazed pastures had an abundance of birds that are 

commonly found in disturbed areas, such as horned 

larks and scaled quail, while grassland specialists 

such as Cassin’s, Botteri’s, and grasshopper sparrows 

were the dominant species in ungrazed plots (Bock 

and Bock 1988). In an extensive review of the grazing 

literature, Jones (2000) found that a majority of the 

studies for which there were sufficient data reported a 

decrease in both rodent species richness and diversity 

in response to grazing.

Livestock grazing can have more indirect effects on 

the environment as well. Soil disturbances created by 

trampling and digging produce microsites ripe for the 

invasion of weedy plant species, and cattle tend to import propagules of nonnative plant species on their coats 

or through their feces (Hobbs and Huenneke 1992). The 

combination of selective grazing by livestock on more 

palatable species and the opportunities for invasion by 

exotic species through soil disturbance and increased 

nutrient input from dung results in the decline of native perennial grasses and an increase in nonnative 

annuals (Mack 1981, 1989, Moore 1970). Livestock 

grazing may also contribute to shrub encroachment 

by eliminating the grasses and reducing competition 

for the seedlings of woody plants (Humphrey 1958). 

Grazing had largely negative impacts on numerous 

soil and vegetation variables examined in the review 

by Jones (2000), including increased soil loss to erosion, decreased infiltration rates, and decreased litter 

cover. Although the results of her review suggest that 

grazing has an overall negative impact on arid ecosystems in North America, Jones also points out that it 

was not possible to control for important factors such 

as stocking rates, grazing intensity, or timing in her 

comparison.

Some would argue that moderate levels of grazing 

may benefit Southwestern grasslands because maximum biodiversity is achieved under intermediate levels 

of disturbance (Connell 1978). In a test of the intermediate disturbance hypothesis, Collins and Barber 

(1985) found that grassland vegetation diversity was 

high on light to moderately grazed mixed-grass prairie 

(as opposed to undisturbed or most severely disturbed 

treatments). They concluded that diversity in such 

systems may be increased by moderate levels of natural 

disturbance, and furthermore that such disturbances 

have additive effects that further increase diversity. 

Today, some practitioners promote the use of properly controlled livestock grazing as a key component 

of sustainable ecosystem management in Southwestern 

grasslands (Savory and Butterfield 1999).

The question of whether grazing is a natural disturbance in the grasslands of the Southwest has been 

the subject of some debate. In the shortgrass prairie, 
62 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

blue grama and buffalo grass Buchloe dactyloides 

coevolved with the bison and are apparently adapted 

to heavy grazing pressure; these grasses thrive under 

such conditions by reproducing both sexually and by 

tillering (Knopf 1994). By contrast, in more recent 

history the desert grasslands of the Southwest have 

been devoid of large herds of grazing ungulates. Most 

evidence points to an absence of large herds of bison 

west of the Rockies (Berger and Cunningham 1994, 

Durrant 1970, Gustafson 1972 as cited in Mack and 

Thompson 1982, Kay 1994). Although Southwestern 

grasses undoubtedly coevolved with grazing due to the 

presence of herbivorous megafauna in the Pleistocene, 

these grasses have now been released from selection 

for grazing defenses for at least 10,000 to 12,000 years 

(Jones 2000 and references therein). In their extensive 

review, Milchunas and Lauenroth (1993) report that 

the sensitivity of grasslands to grazing increases with 

increased aridity and/or the lack of an evolutionary 

history of grazing. If one accepts their results, it is 

hard to avoid the conclusion that the grasslands of the 

Southwest must be especially sensitive to potential 

grazing impacts.

The bunchgrasses of the arid Southwest are indeed 

highly susceptible to grazing by ungulates and respond 

in a manner quite distinctive from the grasses of the 

shortgrass prairie (Daubenmire 1970, Dyer 1979, 

Tisdale 1961). Whereas grazed areas in the shortgrass 

prairie tend to be recolonized by predominantly native plants (Mack and Thompson 1982 and references 

therein), the morphological and physiological features 

of bunch grasses render them incapable of recovering 

quickly from grazing. Continuous grazing in desert 

grasslands leads to changes in species composition, 

where bunch grasses are replaced by sod-forming 

grasses or annuals (Brown 1982), or invaded by 

Eurasian weeds (Mack and Thompson 1982; see also 

Milchunas and others 1988). Furthermore, the soils 

of these grasslands that evolved with few native grazers are protected by a cryptogamic crust of mosses, 

lichens, and liverworts; this crust can be permanently 

destroyed by the trampling of large ungulates, producing sites for the establishment of exotic species 

(Daubenmire 1970, Jones 2000, Mack and Thompson 

1982). Uncontrolled livestock grazing also endangers 

riparian systems in grasslands, one of the greatest 

sources of local diversity, because livestock will eat 

the palatable woody species such as cottonwoods and 

willows, not only removing the bulk of the riparian 

plant community but also destabilizing the banks and 

potentially leading to a lowering of the water table 

(Kovalchik and Elmore 1992).

Grazing impacts in the Southwest remain a highly 

controversial and confusing issue. As Jones (2000) 

points out, the poor experimental design employed 

in the majority of grazing studies has left us with 

a dearth of solid information about the impacts of 

grazing on arid rangelands. Furthermore, results are 

contradictory. Although Jones (2000) found that grazing had negative impacts in the majority of studies 

reviewed, Stohlgren and others (1999) suggest that 

factors such as soil fertility or water availability may 

overshadow the impacts of grazing on variables such 

as native plant species richness. There is little question that the astronomical stocking rates of livestock 

in the late 1800s did great environmental damage to 

the grasslands of the American Southwest; what is 

not well known is how current grazing practices are 

impacting the system (Curtin and others 2002). From 

a management standpoint, one important point to keep 

in mind is that restoration of degraded rangelands 

will require much more than merely removing cattle. 

Ecosystem function must be restored, which demands 

the incorporation of dynamic processes such as fire 

and precipitation to affect grassland condition, and 

furthermore may require mechanical removal or chemical treatment to turn the clock back on woody invaders 

(Curtin and others 2002, McPherson and Weltzin 

2000). Today there is a generally greater awareness of 

the importance of maintaining ecosystem function for 

long-term sustainability, and a growing emphasis on 

the proper management of livestock grazing to ensure 

the ecological integrity of Southwestern grasslands 

(Brown and McDonald 1995, Curtin 2002, Savory and 

Butterfield 1999, Sayre and Ruyle 2001), as witnessed 

by the recent evolution of sustainable ranching organizations such as the Malpai Borderlands Group and 

The Quivira Coalition. More well-designed scientific 

studies of various grazing practices and their effects 

on the biodiversity of Southwestern grasslands are 

clearly needed to eliminate the confusion surrounding this issue and to develop sound management 

guidelines.

Exotic Grasses

A mounting problem in the West is the spread of 

exotic grasses. Grasses such as cheatgrass and crested 

wheatgrass (Agropyron cristatum) may be intentionally introduced as livestock forage or invade following 

disturbance, soon displacing native grasses (Mack 1981, 

Marlette and Anderson 1986). Although exotics such 

as lovegrasses (Eragrostis spp.) are planted as cattle 

forage, these grasses actually increase in response to 

grazing, as the livestock tend to preferentially forage 

on the native grasses and reduce competition for the 

lovegrasses (Bahre 1995 and references therein).

From a biodiversity standpoint, one of the problems 

with at least some exotic grasses is that they do not 

appear to provide adequate habitat for native grassland species. In Arizona, grasslands that have been 

seeded with Lehmann and Boer lovegrass (Eragrostis

lehmanniana and E. curvula var. conferta) have been 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 63

described as “biologically sterile” (Bock and others 

1986:462). Twenty-six native species (10 plants, five 

birds, three rodents, and eight grasshoppers) were 

found to be significantly more abundant in native 

grasslands; only three native species (one bird, one 

rodent, and one grasshopper) were more common in 

the grasslands dominated by the African lovegrasses. 

Bock and others (1986:462) explain: “Indigenous 

animals appear to have evolved specific dependencies on the native flora and/or its associated fauna, 

insofar as most find the exotic grasslands far less 

inhabitable.”

The increase of exotic grasses in the Southwest 

may have further ramifications as well, as they alter 

the natural fire regimes and lead to further ecological changes in the system (Anable and others 1992, 

Cox and others 1990). Some exotics such as Lehmann 

lovegrass increase after fire, and such grasses provide 

more fine fuel to carry fires than native species of 

grasses (Cox and others 1984). This may result in a 

positive feedback loop, in which introduced grasses play 

a beneficial role in terms of increasing fire frequency, 

yet simultaneously have the negative consequences 

of extending the coverage of the less-desirable exotic 

grasses as well as increasing the intensity of fire (Anable 

and others 1992).

Shrub Encroachment

The landscape of the Southwest has been inexorably 

altered over the past century by an extensive invasion of woody plants into areas that were formerly 

grasslands. Numerous authors have documented this 

transition from grassland to shrubland (for example, 

Bahre 1991, Buffington and Herbel 1965, Glendening 

1952, Humphrey 1987) using a variety of techniques 

including early survey records (York and Dick-Peddie 

1969) and photo points (Hastings and Turner 1965). 

The evidence suggests that although about 75 percent 

of southern New Mexico was covered in grasslands 

prior to the late 1800s, by the late 1960s only 5 percent 

grassland coverage remained (York and Dick-Peddie 

1969). In another study of the Chihuahuan Desert, an 

estimated 25 to 50 percent of the area that is currently 

covered by shrublands was actually grassland less 

than 200 years ago (Dinerstein and others 2000). Such 

a conversion represents a significant loss of habitat 

for both plant and animal species that are grassland 

specialists.

Many factors appear to have played a role in this 

transformation. The conversion of grasslands to shrublands is a common result of overgrazing (Risser 1988). 

As livestock preferentially consume the more palatable 

species, initially the perennial grasses, competition 

is reduced and unpalatable woody species have the 

opportunity to become established (Humphrey 1958). 

Furthermore, heavy grazing reduces the fuel loads 

provided by grasses to the point that fire frequency 

and intensity become reduced, thereby removing the 

natural source of control for woody shrubs in grassland 

systems (Archer 1989).

Increases in woody plants such as mesquite following active fire suppression were recognized early on 

by Griffiths (1910). Since that time, the critical role of 

periodic fires in restricting woody plant establishment 

has been clearly demonstrated (for example, McPherson 

1995 and references therein). Humphrey (1958:37) 

argued that the grasslands of the Southwest are a 

“fire-caused subclimax,” but many other factors—such 

as soil type and herbivory by native animals—are now 

believed to interact with fire to maintain the grasslands 

of the Southwest (Curtin and others 2002). Although 

fire alone is not considered sufficient to prevent shrub 

encroachment and maintain the grassland condition indefinitely (McPherson 1995), it is a critically 

important element, and human alteration of natural 

fire cycles through suppression efforts has undoubtedly facilitated the spread of woody plants into these 

grassland systems.

Another explanation for the shift from grassland 

to shrubland in the Southwest is climate change. It 

has long been recognized that shrubs will increase in 

grassland systems in response to drought (for example, 

Schlesinger and others 1990). However, Brown and 

others (1997) found evidence for shrub increases not in 

response to drought, but rather in response to increased 

levels of winter rainfall in recent years. Furthermore, 

through the observation of livestock exclosures they 

were able to document that these increases in woody 

plants occurred in spite of protection from grazing. The 

authors argue that under conditions of high winter precipitation, the establishment of cool-season C3 woody 

shrubs is favored over that of the warm-season C4

grasses that normally dominate the landscape (Brown 

and others 1997 and references therein).

Prairie dogs are believed to be instrumental in 

retarding the growth of woody invaders such as 

mesquite (Koford 1958, Weltzin and others 1997), and 

some authors have suggested that the elimination of 

prairie dogs may be partially responsible for the widespread encroachment of mesquite into Southwestern 

grasslands observed in recent years (Parmenter 

and Van Devender 1995). Other authors have also 

found that small mammals play an important role 

in maintaining grassland systems by restricting the 

establishment of woody plants (for example, Curtin 

and others 2000).

Whatever the mechanism(s), there is little doubt 

that the continuing expansion of woody plants and 

cacti into Southwestern grasslands is one of the greatest sources of habitat degradation or loss threatening 

grassland specialists today.
64 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 Table 4-3. Threatened and endangered terrestrial animal species of Southwestern grasslands. The following species are not necessarily restricted to grassland habitats; although many are grassland spe- cialists, this list also includes those species that rely heavily on grasslands in some parts of their range or as one component of a mosaic of habitats utilized. Sources for the information presented here include Arizona Game and Fish Department (1988), Association for Biodiversity Information (2001), New Mexico Department of Game and Fish (2000), and U.S. Fish and Wildlife Service (2001). New Common name/ Federal Mexico Arizona Scientific name status status status NM AZ Primary threats Types of grassland habitat utilized Note Amphibians Great Plains narrowmouth toad E C • • Water developments, water table Flooded desert grasslands Gastrophryne olivacea draw-down, overgrazing, road development Northern leopard frog C • unknown High elevation wet meadows Rana pipiens Plains leopard frog E • • Habitat loss, bullfrog predation Low elevation wetlands Rana blairi Northern casque-headed frog C • Vegetation clearing, overgrazing, Desert grasslands and scrub AKA lowland burrowing Pternohyla fodiens water table draw-down, roads treefrog Reptiles Bunchgrass lizard T • • Habitat loss and degradation due Dense grass cover at mid-elevations; Now restricted primarily Sceloporus slevini to overgrazing, drought, shrub alkali sacaton in NM to montane grasslands invasion Desert tortoise T C • • Habitat fragmentation, habitat loss, Semi-desert grassland Gopherus agassizii overgrazing, off-road vehicles Massasauga E • • Agricultural development, road kills Desert grasslands Sistrurus catenatus Arizona ridgenose rattlesnake C • Mining, woodcutting, road Ecotonal montane grasslands Crotalus willardi willardi development, collecting New Mexico ridgenose rattlesnake T E • Collecting, habitat degradation Ecotonal montane grasslands Crotalus willardi obscurus due to overgrazing Birds Crested caracara T C • Human disturbance at nest sites Desert grasslands Small breeding population Caracara cheriway in Arizona (Pima Co.) Northern aplomado falcon E E E • Collecting, pesticides, loss of Desert grasslands, savannah Now primarily in Mexico Falco femoralis septentrionalis grasslands to shrub invasion Ferruginous hawk T • • Loss of prairie dogs, human Desert, plains grasslands & Buteo regalis disturbance at nest sites shrubsteppe California condor E E • Direct and indirect human Forages in grasslands Experimental population Gymnogyps californianus persecution (shooting, lead reintroduced in Arizona poisoning, and so forth) Lesser prairie chicken C • Habitat loss, fragmentation due to Shortgrass steppe (in NM, Tympanuchus pallidicinctus overgrazing, agriculture shinnery oak-bluestem association) Masked bobwhite E • Habitat loss and degradation due Desert grasslands and scrub Colinus virginianus ridgwayi to overgrazing Mountain plover PT • • Loss of prairie dog towns (early Shortgrass steppe

Charadrius montanus declines due to market hunting)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 65

White-tailed ptarmigan E • Grazing of tundra habitats, human Alpine tundra, montane grassland

Lagopus leucuras altipetens disturbance

Gray vireo T • • Unknown – possibly clearing of Juniper grassland

Vireo vicinior shrubs

(Arizona) grasshopper sparrow T • • Habitat loss and degradation due Arid grasslands, desert scrub

Ammodramus savannarum ammolegus to overgrazing

Baird’s sparrow T T • • Urban development, overgrazing Shortgrass steppe, plains, desert Wintering migrant in 

Ammodramus bairdii on wintering grounds; loss of native grasslands Southwest

 prairies on breeding grounds

Bobolink E • • Agricultural and urban development, Plains grasslands, agricultural fields Local breeder in Arizona 

Dolichonyx oryzivorus overgrazing (Apache and 

 Navajo counties)

Sprague’s pipit C • • Agricultural and urban development, Winters in Sonoita and San Rafael Wintering migrant in 

Anthus spragueii overgrazing grasslands (Cochise and Santa Cruz Southwest

 Co., Arizona)

MammalsBlack-tailed prairie dog C E • Direct human persecution, canine Shortgrass steppe, desert and plains Extirpated in AZ; remnant 

Cynomys ludovicianus distemper grasslands populations in NM

Hualapai Mogollon vole E E • Loss of habitat due to overgrazing, High elevation grassy areas near springs Listed as Hualapai mexican 

Microtus mogollonensis hualpaiensis recreational development assoc. with ponderosa pine/mixed vole; distribution highly 

 conifer forest restricted

Navajo Mexican vole T • Loss of habitat due to overgrazing, Mid-elevation montane grasslands

Microtus mogollensis navaho mining, recreational development

(Arizona) montane vole E • • Loss of habitat due to water High elevation mesic grasslands and 

Microtus montanus arizonensis diversions, livestock impacts marshes

Least shrew T • Loss of habitat due to water Mesic grasslands and marshes

Cryptotis parva diversions, livestock impacts, 

 agriculture

Meadow jumping mouse T T • • Loss of habitat due to overgrazing, Montane meadows

Zapus hudsonius urban encroachment

White-sided jackrabbit T • Loss and degradation of habitat Desert grasslands Found in Hidalgo County, 

Lepus callotis due to overgrazing, agriculture NM (and Mexico)

Black-footed ferret E E • • Elimination of prairie dogs Desert and plains grasslands in Extirpated; experimental 

Mustela nigripes (primary prey source) association with prairie dog towns reintroductions 

 in NM and AZ

Mexican gray wolf E E E • • Direct human persecution Desert grasslands Extirpated; experimental 

Canis lupus baileyi reintroductions 

 in NM & AZ

Sonoran pronghorn E E • Loss of habitat due to grazing, Desert grasslands, galleta grasslands 

Antilocapra americana sonoriensis agriculture; in Mexico, poaching of SW Arizona

Chihuahuan pronghorn T • Overhunting, habitat loss and Southern Arizona grasslands Extirpated; present 

Antilocapra americana mexicana fragmentation populations reintroduced

(Desert) bighorn sheep E • • Overhunting, habitat loss and Desert grasslands, montane Reintroduction programs 

Ovis canadensis mexicana fragmentation grasslands under way

E = endangered, T = threatened, P = proposed, C = candidate
66 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Summary of Threats to Biodiversity in 

Southwest Grasslands

Unfortunately the native biodiversity of 

Southwestern grasslands is threatened by multiple 

sources, as are the native flora and fauna of all 

major ecosystems today. The alteration of natural 

fire cycles, inappropriate grazing regimes, eradication of keystone species, exotic grasses, habitat loss 

to shrub encroachment—these are just a few of the 

many factors believed responsible for the plant and 

animal species of Southwestern grasslands that have 

declined to the point of being listed as threatened or 

endangered (tables 4-3 and 4-4). Human activities 

such as urban development, mining, water diversions, and collecting have all contributed to declines 

in biodiversity, as has the purposeful elimination 

of certain species in several cases. Given the vast 

array of potential factors impacting the biodiversity 

of Southwestern grasslands, the discussion here of 

threats is not meant to be comprehensive, but only 

to touch on some of the major sources of declines in 

native species richness. Habitat fragmentation, a 

major potential threat to myriad grassland species 

in the Southwest, is discussed separately in the following section.

Ecological Consequences of Habitat 

Fragmentation_____________________

Introduction and Theoretical Background

Human use of the environment has led to a condition 

in which large areas of formerly continuous landscapes 

Table 4-4. Threatened and endangered plants of Southwestern grasslands. The following species are not necessarily restricted to grassland 

habitats; although many occur primarily in grasslands, this list also includes those species that are found in grasslands in some parts 

of their range or as one component of a mosaic of habitats utilized. Sources for the information presented here include Association for 

Biodiversity Information (2001), New Mexico Native Plants Protection Advisory Committee (1984), New Mexico Rare Plant Technical 

Council (2001), and U.S. Fish and Wildlife Service (2001).

Common name/ Federal Primary Grassland 

Scientific name status NM AZ threats habitat Notes

Arizona agave E • Collecting, overgrazing Juniper grasslands 1100-2750 m Endemic to central AZ

Agave arizonica

Cochise pincushion cactus T • Collecting, pesticides, mining Limestone hills in desert Only two populations, 

Coryphantha robbinsorum grasslands 1280 m one in SE AZ and 

 one in Mexico

Pima pineapple cactus E • Collecting, livestock impacts, Open, flat alluvial basins of semi- Pima and Santa Cruz 

Coryphantha scheeri ORVs, habitat loss due to desert grasslands and Sonoran Co., Arizona and 

 var. robustispina development desert-scrub 700-1400 m Sonora, Mexico

Kuenzler’s hedgehog cactus E • Collecting, livestock impacts Great Plains grassland Southcentral New 

Echinocereus fendleri 1600 – 2000 m Mexico

 var. kuenzleri

Fickeisen pincushion cactus C • Collecting, livestock impacts, Limestone soils in Great Plains Coconino and Mohave 

Pediocactus peeblesianus ORVs grasslands ~1500m Co., Arizona

 var. fickeiseniae

San Francisco Peaks T • Recreational: off-trail hiking Alpine tundra 3350-3750 m Isolated mountain 

 groundsel and climbing endemic

Senecio franciscanus

Sacramento Mountains T • Water development, livestock Wet meadows ~2440 m Endemic; persists only 

 Thistle impacts in areas too steep for 

Cirsium vinaceum livestock

Canelo hills ladies’ tresses E • Water diversions, livestock Permanently wet meadows Limited to four 

Spiranthes delitescens impacts, exotic species, (cienegas) ~1525 m cienegas in southern 

 mining Arizona

Gypsum wild-buckwheat T • Gypsum mining, recreational Open gypsum in grama Isolated population in 

Eriogonum gypsophilum development grasslands ~1500m Eddy County, NM

E = endangered, T = threatened, P = proposed, C = candidate
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 67

have become increasingly fragmented and isolated. 

Urban development, agriculture, power lines, road 

construction, and other such activities have accelerated 

over the past century, subdividing the natural world 

into disjunct remnants of native ecosystems embedded 

in a matrix of anthropogenic land uses (Saunders and 

others 1991). The negative ecological impacts of such 

fragmentation on natural systems has led many conservation biologists to identify habitat fragmentation 

as one of the greatest threats to biodiversity today 

(Harris 1984, Noss and Cooperrider 1994, Wilcox and 

Murphy 1985).

Under the traditional definition, there are two 

fundamental components to habitat fragmentation. 

First, the activity that leads to fragmentation usually 

leads to an outright loss of some area of the original 

habitat; this component can be considered habitat loss 

or destruction. The second component is fragmentation per se, in which the remaining natural areas are 

relegated to patches of reduced size isolated from one 

another across the landscape (Wilcove and others 

1986). A common analogy is that these fragments now 

exist essentially as habitat islands in a sea of altered 

or degraded lands; thus, this effect is also referred 

to as insularization (Wilcox 1980). The introduction 

of “edge effects” might be considered a third fundamental component of fragmentation. Edge effects are 

manifested in the form of altered physical processes 

and biotic interactions along habitat edges. The amount 

of edge habitat may increase dramatically through the 

process of fragmentation because reducing the size of 

the habitat patches results in a proportional increase 

in the amount of edge (Janzen 1983, Williamson 

1975, Yahner 1988); altering the shape of fragments 

may also have this effect (Diamond 1975, Wilson and 

Willis 1975).

Many of the negative impacts of fragmentation 

are based on the principles of island biogeography, 

a classic model in conservation biology that predicts 

the number of species that will be found on an island 

as a function of species colonization and extinction 

rates, the size of the island, and its degree of isolation 

(MacArthur and Wilson 1967; fig. 4-1). Large islands 

near a potential source of immigrants have the greatest 

rates of colonization because individuals can traverse 

a short distance more easily, and the size of the island 

makes it more likely that dispersing individuals will 

happen upon it. Small islands far from the source have 

the least chance of intercepting potential colonists; it is 

less likely that individuals will be capable of traveling 

the greater distance required, and the small size of the 

island makes it less likely to be discovered. Balanced 

against the effect of colonization is that of extinction. 

Large islands support large populations of different 

Figure 4-1. Graphic representation of island biogeography theory. Large islands located near a potential source of colonists should 

support a greater equilibrium number of species due to high immigration rates and low extinction rates. Small, isolated islands are 

predicted to have the least number of species due to lower colonization rates and greater extinction rates (after MacArthur and 

Wilson 1967).
68 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

species, and as large populations are more resilient 

in the face of potential extinction events, extinction 

rates should be relatively low. Small islands would 

support small populations of different species, and 

as small populations are particularly vulnerable to 

extinction (Harris 1984, Saunders and others 1991), 

extinction rates would be high. Based on the interaction 

between extinction rates and colonization rates, the 

model thus predicts that large islands located near a 

potential source of immigrants will support the greatest 

equilibrium number of species, while small, isolated 

islands will support the lowest number of species.

Although the theory of island biogeography was 

based on the species-area relationships observed on 

oceanic islands, this model has been widely applied 

to habitat fragments (“islands”) in continental terrestrial systems as the basis for conservation planning 

(for example, Shafer 1990) and has largely formed 

the foundation of scientific inquiry into the effects of 

habitat fragmentation. Closely aligned with the theory 

of island biogeography and often applied to fragmented 

systems as well is the idea of metapopulation biology. 

The metapopulation concept holds that a population 

may be composed of a number of scattered subpopulations that are subject to repeated colonizations and 

extinctions, but that as a whole generally persists 

at some equilibrium level of abundance over time 

(Levins 1969, 1970). Metapopulations are typically 

characterized by one or more core or source populations and several fluctuating satellite populations. 

Satellite populations may occasionally go extinct 

when conditions are not favorable, but are replaced 

by new colonists dispersing from the core population 

when conditions improve (Bleich and others 1990). 

The viability of a metapopulation thus depends on the 

persistence of the core subpopulation and the ability 

of dispersing individuals to balance local extinction 

events by successfully recolonizing vacant patches 

(Gilpin and Hanski 1991); such recolonization events 

have been deemed the “rescue effect” (Brown and 

Kodric-Brown 1977).

Much of applied conservation biology rests on the 

precepts of island biogeography theory, the metapopulation concept, and the avoidance of edge effects. Basic 

principles of conservation design in fragmented landscapes include maximizing the size of habitat fragments 

to preserve species diversity and reduce extinction 

risk, minimizing the distance between fragments to 

facilitate dispersal, and controlling the shape of fragments to minimize the amount of edge relative to core 

habitat (Diamond 1975, Shafer 1990). In recent years, 

the major hypotheses stemming from the application 

of these theories (for example, that small fragments 

will support fewer species than large fragments) have 

been tested repeatedly and with largely variable results 

(discussed below; also see Bierregaard and others 

1992, Debinski and Holt 2000 for a review). The vast 

majority of studies on the effects of fragmentation 

have centered on forested landscapes, particularly 

those in the tropics (for example, Lovejoy and others 1984) and the Central or Eastern United States 

(for example, Askins and others 1990, Robbins 1980, 

Robinson and others 1995, Whitcomb and others 1981, 

Wilcove and Robinson 1990). Few studies have focused 

on the impacts of fragmentation in North American 

grasslands (for example, Collinge 1998, 2000, Quinn 

and Robinson 1987, Robinson and Quinn 1988), and 

many of these have concentrated primarily on birds 

of the tallgrass prairie in the Midwest (for example, 

Johnson and Temple 1990, Winter and Faaborg 1999). 

As the effects of fragmentation in Western grasslands 

have gone largely unstudied, the discussion of the ecological impacts of habitat fragmentation that follows 

here is of necessity derived from studies conducted 

primarily in other ecosystems. On an ecological level 

grasslands suffer many of the same consequences of 

habitat fragmentation as do forested areas, although 

the contrast between the natural and altered conditions may not appear as abrupt. Data from grassland 

systems are utilized whenever they are available.

Edge Effects

Habitat edges exhibit a marked contrast in the 

structure and species composition of the vegetation 

between two adjacent elements in the landscape. 

Although natural edges occur in nature, as when blowdowns create openings in forests, the most common 

edge in a fragmented landscape is the product of human 

activity—an “induced” edge (Yahner 1988). Such edges 

have been associated with numerous negative impacts 

on the organisms originally inhabiting the remaining 

fragment, including increased levels of parasitism and 

predation, changes in species composition, and physical 

alterations in environmental conditions; these impacts 

are collectively known as edge effects (Lovejoy and 

others 1986, Yahner 1988). In forest systems, it is the 

removal of trees that results in fragmentation and 

the creation of habitat edges. Clearing for logging, 

development, agriculture, road construction, and other 

purposes all contribute to this process. In grassland 

ecosystems, it is just the opposite: the introduction of 

trees or shrubs is one of the primary causes of fragmentation and edge effects. Such seemingly innocuous 

human constructs as treelines planted for windbreaks 

or fencerows, stringers of trees along irrigation ditches, 

and trees along roadsides create long, linear stretches 

of edge habitat that can negatively impact the native 

species of the surrounding grasslands (O’Leary and 

Nyberg 2000).

Trees and shrubs create vertical structure in the 

grassland landscape, providing cover and perches for 

predators and leading to increased levels of predation 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 69

along the edge habitat created by the interface between 

the grassland and the treeline (Burger and others 1994, 

Gates and Gysel 1978, Johnson and Temple 1990, 

Møller 1989, Ratti and Reese 1988, Winter and others 

2000). Nest predators such as jays, raccoons, skunks 

and opossums hunt preferentially along the perimeter 

of agricultural fields or not far from a wooded edge, 

and prairie raccoons are known to use shelterbelts 

as travel lanes (Bider 1968, Fritzell 1978, Gates and 

Gysel 1978, Wilcove 1985). Structures such as fences, 

telephone poles, or rooftops provide perches for predators as well, and any human developments also tend 

to serve as a source of “urban predators” such as cats 

(Wilcove 1985). Perches that provide a good view to 

scan for potential host nests are also considered a critical habitat feature for brown-headed cowbirds, a nest 

parasite (Norman and Robertson 1975). In grassland 

systems, the introduction of trees, shrubs, or human 

structures provides these lookout sites, leading to 

significant increases in parasitism levels along such 

edges and resulting in reduced nest productivity or 

even nest failure in grassland breeding birds (Best 

1978, Johnson and Temple 1986, 1990, Wray and others 

1982). These effects have been found to extend into 

grasslands up to 75 m in from a woody edge (Burger 

and others 1994, Helzer 1996, Paton 1994), and many 

grassland breeding birds appear to avoid nesting or 

foraging within this zone (Delisle and Savidge 1996, 

Johnson and Temple 1990). In addition, grassland birds 

that do not tend to fly toward shrubs for cover when 

disturbed have been found to actively avoid woody 

edges, and the density of these birds tends to decrease 

as the amount of woody cover increases (Lima and 

Valone 1991). Such impacts are of particular concern 

because most species of grassland breeding birds have 

been exhibiting consistent and often striking population declines over the past few decades (Herkert 1994, 

Peterjohn and Sauer 1999, Samson and Knopf 1994), 

and many of these declines are believed to be linked 

with the loss and fragmentation of native grassland 

habitats (Herkert 1994, Johnson and Temple 1986, 

1990, Peterjohn and Sauer 1999, Samson 1980, Vickery 

and others 1994).

The creation of edges, whether through increased 

woody vegetation or clearing, opens up avenues for 

incursion by opportunistic “edge” species and invasive 

exotics. The disturbance of native plant communities 

facilitates invasion by weedy and/or exotic plants, 

and such disturbance events typically accompany the 

activities that lead to fragmentation, such as road 

construction (Ewel 1986, Hobbs 1989, 1991, Rejmanek 

1989, Saunders and others 1991, Schowalter 1988). 

Such increases in edge species or habitat generalists 

have also been found in such diverse taxa as insects 

(for example, Suarez and others 1998, Webb and 

Hopkins 1984), frogs (Laurance and Bierregaard 1996) 

and birds (for example, Herkert 1994, Samson 1980). 

Fragmentation and edge effects have been found to 

have a dramatic impact on the diversity of native ant 

species, for example. Suarez and others (1998) found 

that habitat fragments were characterized by high 

numbers of introduced ant species along the edges, 

and that unfragmented control plots supported three 

times as many native ant species as did habitat fragments. Furthermore, there was a negative correlation 

between the number of native ant species and time since 

fragmentation, suggesting that the native ants were 

incapable of recolonizing patches in the fragmented 

landscape once local extinctions had occurred.

Changes in the vegetative structure or species 

composition of the plant community may effect some 

changes in the animal community as well, typically 

leading to increased numbers of opportunistic species or habitat generalists (Saunders and others 

1991). In grassland communities, the introduction of 

woody vegetation is correlated with increased species 

diversity of birds and lizards due to greater representation by generalists or species that normally utilize 

shrubby vegetation, while grassland specialists that 

formerly occupied the area tend to be lost (Germano 

and Hungerford 1981, Saunders and others 1991, 

Schmiegelow and others 1997). Changes in the faunal 

composition of habitat fragments may also impact the 

remainder of the community. In California grasslands, 

there was a significant correlation between the loss 

of native mammal species richness and the numbers 

of exotic birds and mammals occupying habitat fragments (Smallwood 1994). The loss of native species, it 

is proposed, leads to unstable population dynamics and 

lowers the “biotic resistance” (Simberloff 1986) of the 

community, leaving it vulnerable to invasion by exotics. 

Edges allow for the infiltration of formerly inaccessible 

interior habitats by a diverse array of invasives, but 

while fragments tend to support increased numbers 

of exotic or opportunistic species, habitat specialists 

tend to consistently decline within these patches (for 

example, Harris and Scheck 1991, Herkert 1994, 

Robinson and Quinn 1988, Samson 1980, Suarez and 

others 1998, Verner 1986, Webb 1989).

Impacts on the physical environment and

ecological processes—The reduction in area of the 

original habitat and concurrent increase in the amount 

of edge can provoke physical changes in the fragment 

microclimate. Studies of forest systems have found 

that habitat fragments experience increased solar 

radiation along edges, altering plant species composition and leading to higher soil temperatures, in turn 

potentially affecting nutrient cycling (Lovejoy and 

others 1986, Saunders and others 1991). Increased soil 

temperatures may impact the numbers and activities 

of soil-dwelling organisms involved in decomposition 

as well as decrease the moisture retention capacity of 
70 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

the soil (Klein 1989, Parker 1989, Saunders and others 

1991). Whether increased solar radiation effects such 

changes in grassland systems is largely unknown, although one study reports that nutrient cycling was not 

affected in studies of fragmented old fields (Debinski 

and Holt 2000). Another physical edge effect in forest fragments is the increased penetration of wind, 

which may result in direct physical damage to the 

vegetation or act to increase evapotranspiration and 

hence desiccation (Lovejoy and others 1986). Although 

grassland systems probably do not experience the same 

impacts due to their naturally short stature, winds 

do have increased accessibility to grasslands along 

cleared edges, resulting in the potential for increased 

transport of seeds, insects, and disease organisms 

into grassland fragments (Hobbs and Atkins 1988, 

Saunders and others 1991). Fragmentation can lead to 

changes in water regimes, as cleared areas contribute 

to increased runoff and erosion and lowered absorption of water into the soil (Kapos 1989, Saunders 

and others 1991). Replacement of deep-rooted native perennial grasses with introduced annuals can 

also contribute to reduced evapotranspiration rates, 

increased runoff, and increased temperatures at the 

soil surface. Such changes in the moisture levels of 

soils and runoff patterns can lead to the creation of 

new substrates for invasion by exotic or weedy plant 

species, impact seedbed characteristics, and result 

in the displacement of organisms that are unable to 

survive the altered environmental conditions (Hobbs 

and Huenneke 1992, Jones 1981, Saunders and others 1991).

Fire regimes may also be affected by habitat 

fragmentation. Most grasslands are considered firedependent ecosystems, requiring frequent fires to set 

back succession and maintain the natural distribution, productivity, and diversity of the grassland 

(McPherson 1995). As discussed above, habitat fragments are vulnerable to invasion by exotic species, 

and increased numbers of exotic grasses may seriously 

disrupt normal fire cycles. Introduced species such as 

lovegrasses (genus Eragrostis) are common throughout 

Southwestern rangelands (Loftin and others 2000). 

Sites dominated by lovegrasses may exhibit biomasses 

up to four times that of native grasslands, resulting 

in abnormally high fire frequencies and intensities 

that tend to kill the native plants but that lead to 

even greater abundances of the lovegrass (Anable 

and others 1992, Cox and others 1990). Fragments of 

native grasslands, on the other hand, may face the 

problem of decreased fire frequency. As fragments 

diminish in size, it becomes increasingly unlikely that 

they will be struck by lightning frequently enough 

to maintain the grasslands. A study of small prairie 

fragments in Wisconsin showed that the absence of 

fire over 32 to 52 years resulted in a loss of between 

8 and 60 percent of the original plant species (Leach 

and Givnish 1996). Rare plants showed the greatest 

losses from these grassland fragments in the absence 

of fire. Roads and other agents of fragmentation 

may also act as firebreaks, restricting the spread 

of what would otherwise be extensive range fires. 

Finally, fragmentation due to human habitation also 

provides an incentive for active suppression of fires 

that could potentially threaten structures (Hansen 

and others 2002).

Area-Sensitive Species: Interaction of Edge 

Effects and Habitat Reduction

That species richness will decrease as a function of 

reduced geographic area is the most basic prediction of 

island biogeography theory. Studies of grassland birds 

show that this guild closely follows this prediction, 

as species richness is significantly correlated with 

the size of habitat fragments (Herkert 1994, Samson 

1980). However, several species of grassland birds drop 

out of the community even in fragments that appear 

large enough to support them. These species simply 

will not utilize habitat fragments below a certain 

threshold size for nesting, even if the fragment is 

large enough to hold several average-sized territories 

and the habitat appears to be suitable; such species 

have been termed “area-sensitive” (see for example 

Herkert 1994, O’Leary and Nyberg 2000, Samson 

1980, Vickery and others 1994; for forest birds, see 

Blake and Karr 1987, Robbins 1980, Robbins and 

others 1989, Winter and Faaborg 1999). The minimum area required by area-sensitive species varies 

widely: Eastern meadowlarks require only 5 ha, 

whereas Henslow’s sparrow will not nest in a fragment of less than 55 ha. Grasshopper sparrows and 

savannah sparrows fall toward the larger end of the 

range at 30 and 40 ha, respectively (Herkert 1994). 

Greater prairie chickens and upland sandpipers are 

well known for their avoidance of small grassland 

fragments and are found regularly breeding in fragments of 160 ha or more (Cannon and Christisen 

1984, Samson 1980, Westemeier 1985). Furthermore, 

not only does nest density and nest success decrease 

with fragment size (Burger and others 1994, Johnson 

and Temple 1986, 1990, Samson 1980, Winter and 

Faaborg 1999; but see Delisle and Savidge 1996), but 

the simple distribution and density of several species 

of grassland birds is also positively correlated with 

fragment size (Helzer 1996, Herkert 1994, Winter 

and Faaborg 1999).

Exactly why these birds avoid small fragments, 

even when adequate suitable habitat appears to be 

available, is unclear. Most likely it is not the amount 

of available habitat per se that is important, but 

rather the amount of core habitat—that is, the amount 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 71

of habitat that is far enough into the interior of the 

fragment to escape the edge effects of increased 

predation and parasitism—that is the critical factor 

(Brittingham and Temple 1983, Burger and others 

1994, Gates and Gysel 1978, Johnson and Temple 

1986, Winter and others 2000; see “Edge Effects” 

section above). Grassland birds clearly avoid nesting close to edges in habitat fragments (Delisle and 

Savidge 1996; Johnson and Temple 1990; O’Leary and 

Nyberg 2000; Warner 1994; Winter and others 2000); 

small fragments, particularly if they are somewhat 

linear in shape, may simply not provide any core 

area for nesting (fig. 4-2). Avoidance of edge for other 

activities, such as foraging, may also be a reaction 

to increased predator activity along edges (Andrén 

and Angelstam 1988, Fritzell 1978, Gates and Gysel 

1978, Johnson and Temple 1986, 1990, Wilcove 1985, 

Yahner and Scott 1988). Several studies have found 

that the edge:area ratio of fragments has a greater 

influence on the presence and richness of grassland 

birds, and on the presence and success of nesting 

species, than does fragment area (Burger and others 1994, Helzer and Jelinski 1999, Temple 1986, 

Winter and others 2000). Area-sensitivity has been 

attributed only to birds thus far, but may possibly 

occur in other taxa as well.

Loss of Grassland Habitats and 

Fragmentation in the Southwest_____

In the Midwest, as little as 4 percent of the original 

native tallgrass prairie is estimated to remain; in 

some states, that figure may drop as low as 1 percent 

(Samson and Knopf 1994). Most of the Midwestern 

grasslands have been cleared for agriculture, particularly rowcropping for products such as wheat and 

corn. Destruction of Southwestern grasslands due to 

clearing for rowcrop agriculture has been relatively 

minimal, since such crops are few (for example, chile, 

cotton) and cover only a small portion of the land area 

of New Mexico and Arizona. Nonetheless, clearing 

for such purposes has contributed to the outright 

loss of native Southwestern grasslands, as has clearing for urban development (Bahre 1995). Grazing is 

the predominant use of Southwestern rangelands, 

and improper grazing practices can lead to loss of 

grasslands not through clearing per se, but through 

degradation of the grasslands to the point that they 

no longer function as suitable habitat for native species (for example, Bahre 1995, Bock and others 1984, 

Bock and Webb 1984, Noss and Cooperrider 1994). 

Uncontrolled heavy use of native arid grasslands by 

domestic livestock can lead to the loss of native grasses, 

the introduction of invasive exotic grasses and other 

weedy species, the destruction of cryptogamic crusts, 

altered grassland structure, and contribute to the conversion of grasslands to shrub-dominated desert scrub 

or pinyon-juniper (Bahre and Shelton 1993, Hobbs and 

Huenneke 1992, Humphrey 1958, Mack 1981, 1989, 

Martin 1975, Moore 1970, Wright and others 1979). 

Most grassland systems are maintained by periodic 

fires that set back succession, but a history of fire suppression has allowed the widespread encroachment of 

shrubs and trees into such systems (Humphrey 1958, 

McPherson 1995). The increased invasion of grasslands 

by exotic plants facilitated by grazing, road construction, and other forms of disturbance also contribute to 

altered fuel structure and fire regimes, leading to the 

eventual conversion of the native grassland to some 

other habitat type (Loftin and others 2000, MacDonald 

and others 1989, Panetta and Hopkins 1991, Saunders 

and others 1991). Whether lost through outright 

Figure 4-2. An example of how fragmentation and edge 

effects may render what appears to be an adequate area 

of quality habitat unsuitable for nesting or other activities 

for grassland birds. (A) Assuming edge effects extend a 

distance of 75 m into the interior from any edge (the shaded 

area), an area of 30 ha would offer 12.75 ha of potential 

core habitat. (B) Bisecting this area with a treeline (an 

induced edge) effectively places the entire area within 

the zone of edge effects and eliminates all potential core 

habitat, even though the total area is essentially unchanged.
72 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

clearing, degradation, or gradual conversion, the end 

result of such habitat loss is twofold: first, there is an 

overall reduction in the area of extant native grassland, 

and second, those grasslands that do remain are relegated to disjunct fragments of relatively small size.

A new term for one particular source of fragmentation is becoming increasingly prevalent in the Western 

United States: exurban development (Knight 1999). 

Exurban development refers to low-density residential 

development that occurs beyond the limits of incorporated towns and cities. Expanses of land that were 

once devoted to agriculture or ranching are subdivided 

and sold for the development of “ranchettes” (single 

houses generally situated on from 10 to 40 acres of 

land) which contribute to this new trend of rural sprawl 

(Brown and McDonald 1995, Hansen and others 2002, 

Theobald 2000). Between 1994 and 1997, nearly 80 

percent of the new home construction in the United 

States was in nonmetropolitan areas, and 57 percent 

of the houses were built on lots equal to or greater 

than 10 acres (Heimlich and Anderson 2001). This 

conversion of private ranching and farming lands to 

rural residential developments has been called “the 

most profound land use change in the New West” 

(Maestas and others 2002).

Although much of the land surrounding these homes 

remains in a relatively natural state, these low-density rural developments still introduce the negative 

effects of fragmentation into the environment with 

the associated predictable negative impacts on native 

biodiversity. The construction of buildings, roads, 

fences, and other structures associated with these rural 

subdivisions result in a dramatic increase in habitat 

fragmentation (Knight and others 1995). Knight (2003) 

reports that approximately one-fifth of the land area 

of a subdivided ranch is affected by houses and roads. 

The native species community composition changes in a 

predictable fashion, as specialized native species, such 

as dusky flycatchers, tend to be replaced by generalist, 

human-adapted species, such as black-billed magpies 

(Maestas and others 2002; see also Hansen and others 

2002, Odell and Knight 2001). These changes are apparently little affected by the density of the housing 

development; that is, these effects are seen whether 

houses are densely clustered or spread more widely 

across the landscape (Odell and Knight 2001).

Predation and parasitism on native birds and mammals increases as residential development brings a 

concurrent increase in predators, both in the form of 

family pets and through associated increases in human-adapted species such as brown-headed cowbirds 

or jays (Hansen and others 2002, Maestas and others 

2002). Nonnative plant species also tend to increase in 

association with exurban developments (Knight and 

others 1995, Maestas and others 2002), and natural 

disturbance regimes (such as fire) are disrupted (Bahre 

1995, Hansen and others 2002). Furthermore, people 

tend to settle in the same areas that are most favored 

by wildlife, and outdoor recreationists moving into 

these rural areas both disturb and displace native 

wildlife (Hansen and others 2002). Although there 

has been little discussion of this issue until relatively 

recently, conservation biologists and land managers 

are becoming increasingly concerned about this newest threat to the biodiversity of the Western United 

States, apparently with good reason (see for example, 

Bahre 1991, 1995, Brown and McDonald 1995, Hansen 

and others 2002, Knight and others 2002, Odell and 

Knight 2001).

Changes in Species Richness and 

Species Composition_______________

The reduced size and increased isolation of areas 

of native habitat have numerous theoretical repercussions for the native species that depend upon them, 

such as reduced species richness in the remaining 

fragments, interference with dispersal and colonization 

abilities, interruption of metapopulation dynamics, and 

increased risk of extinction (Meffe and others 1997, 

Noss and Cooperrider 1994, Wilcove and others 1986). 

Examination of these hypotheses in studies of habitat 

fragmentation has yielded mixed results. Although 

some habitats reduced in size do exhibit decreased 

species richness as predicted by island biogeography 

theory (Baur and Erhardt 1995, Bierregaard and 

others 1992, Collinge and Forman 1998), many either 

maintain the same number of species as prior to fragmentation, or actually exhibit an increase in species 

richness (Quinn and Robinson 1987, Simberloff and 

Abele 1982, Simberloff and Gottelli 1984). This is one 

of the key problems in applying island biogeography 

theory to continental systems: real islands are surrounded by a habitat matrix that is truly inhospitable to 

terrestrial species, whereas habitat “islands” are often 

encompassed by a matrix of habitats that, although 

hostile, may be habitable to some extent (Andrén 1994). 

In continental systems this matrix may serve as a 

source of potential colonists, allowing for the invasion 

of habitat fragments by weedy edge species, habitat 

generalists, or exotics (Doak and Mills 1994, Noss 

and Cooperrider 1994, Zimmerman and Bierregaard 

1986). In such cases, fragmentation may actually result 

in an increase in species richness. However, the key 

point that is often overlooked is that while the overall 

number of species may rise, the species composition of 

the fragment may be irretrievably altered. Sensitive 

species of habitat interiors, endemic species, or habitat 

specialists may be lost, while numbers of common opportunistic species increase (for example, Harris and 

Scheck 1991, Lynch 1987, Noss 1983, Samson 1980, 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 73

Verner 1986, Webb 1989, Webb and Hopkins 1984; see 

also discussion in the “Edge Effects” section above).

Although the number of species found in a habitat 

fragment may initially be high, one theory holds that 

over time the number of species the reduced fragment 

can now support will eventually drop—a process known 

as “relaxation” (Diamond 1972). A common criticism 

of contemporary studies of habitat fragmentation is 

that the time frame is too short—often on the order 

of just a few years—to document the relatively slow 

process of extinction following fragmentation (for 

example, Andrén 1994, Gonzalez 2000, Schmiegelow 

and others 1997, Tilman and others 1994). The process 

of species relaxation has been observed to some degree 

in birds (Schmiegelow and others 1997), small mammals, and insects (Debinski and Holt 2000), but perhaps 

the most thorough documentation of this phenomenon 

comes from a study of a microarthropod community 

in a bryophyte-based microlandscape (Gonzalez 2000). 

The reduced spatial and temporal scale of the dynamics 

in this community enabled the investigator to observe 

the effects of habitat fragmentation over many generations, which did in fact finally result in substantial 

numbers of local extinctions. Fragmentation thus has 

two effects on species richness operating on two time 

scales: first, immediately following fragmentation there 

is an “instantaneous sampling effect,” in which species 

richness is a sample of the richness at a larger scale; 

second, there is the long-term process of community 

relaxation, resulting in a decrease in species richness 

to a new steady state value (Gonzalez 2000). This difference between the initial postfragmentation level 

of species richness and the eventual lowered steady 

state value has been termed “the extinction debt,” 

because although the extinctions do not occur until 

many generations following fragmentation, they are 

bound to occur and are thus a debt that will come due 

in future years (Tilman and others 1994). The model 

upon which the extinction debt is based showed a 50 to 

400 year or more time lag between habitat destruction 

and species extinctions; it also predicted that even 

those species initially most abundant in undisturbed 

habitat fragments can be the same species that are 

most susceptible to eventual extinction (Tilman and 

others 1994).

Vulnerability to Local Extinction_____

Habitat fragments may lose species for many 

reasons. Those species generally considered most susceptible to local extinctions are naturally rare species, 

species of habitat interiors or “area-sensitive species” 

(see discussion above), sedentary species, species with 

limited dispersal capabilities, species with specialized habitat requirements (especially if the resources 

required are patchy or unpredictable in occurrence) 

and animals with large home ranges or wide-ranging 

animals (Meffe and others 1997, Saunders and others 

1991, Wilcove and others 1986). For naturally rare 

species, or those that occur at low densities in the 

environment, extinction due to fragmentation is largely 

a matter of chance. Being widely distributed across 

the landscape, the initial persistence of such a species 

would depend upon the likelihood that any remnant 

habitat patches just happen to capture some individuals of the population. The long-term maintenance of 

the larger population would depend upon the ability 

of these surviving individuals to successfully interact 

and reproduce in the fragmented landscape.

For other organisms, survival in a fragmented 

landscape may depend on the size of the remaining 

fragments. For each species, there is some “critical 

threshold” size of habitat area below which the species cannot persist. A generic threshold of 10 to 30 

percent of the remaining habitat has been reported 

for birds and mammals (Andrén 1994), but the exact 

value of any such threshold ultimately depends upon 

the scale at which an organism interacts with its environment. In other words, it depends upon whether 

or not individuals of the species in question perceive 

the landscape as connected or fragmented (With and 

Crist 1995). For example, a wide-ranging species that 

is a habitat generalist, such as a robin, might essentially be able to experience a fragmented landscape as 

functionally connected, because the robin can easily 

utilize several disjunct fragments by flying between 

them and would be able to make use of the resources 

in most any fragment it happens upon. However, an 

animal with limited mobility and specialized habitat 

requirements such as a frog, might be incapable of 

crossing the surrounding landscape matrix and would 

therefore experience the same landscape as fragmented 

and restrictive. Furthermore, even if the frog managed 

to travel to another habitat patch, it would have to 

depend upon the presence of water in the new patch 

to persist there. The degree of fragmentation, then, as 

well as the value of a critical threshold, is a matter not 

only of the area of habitat remaining and its spatial 

arrangement, but also the habitat requirements and 

dispersal ability of the species in question (O’Neill 

and others 1988, Plotnik and Gardner 1993, With 

and Crist 1995). Even for relatively wide-ranging 

species, however, fragmentation can have significant 

impacts. Grassland raptors such as prairie falcons, 

ferruginous hawks, and rough-legged hawks have 

been found to decline in numbers if as little as 5 to 7 

percent of the landscape becomes urbanized (Berry 

and others 1998).

Despite their excellent dispersal abilities, large 

animals are often the first to be lost from small fragments. For many of these, the remnant habitat patches 

may simply be smaller than their minimum home 
74 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

range or territory sizes. Some species of Midwestern 

raptors are thought to be declining because there are 

few tracts of habitat left that are extensive enough to 

meet their needs during the breeding season (Robinson 

1991). Large carnivores typically maintain extensive 

home ranges; the home ranges of male mountain lions 

may exceed 400 km2 (Seidensticker and others 1973). 

Mountain lions and other large carnivores such as 

grizzly bears are decreasing in numbers as the large 

tracts of habitat they require continue to shrink in size 

and become increasingly isolated from one another 

(Picton 1979, Wilcove and others 1986). Even many 

of our National Parks do not provide areas of habitat 

extensive enough to sustain populations of large animals over time without active management (Meffe and 

others 1997). Although larger parks are more likely to 

maintain their native animal communities, nonetheless nearly 30 species of mammals have experienced 

local extinctions from National Parks, including many 

smaller species such as rabbits (Newmark 1987, 1995). 

Overall, the current system of nature reserves in the 

world is considered to be too small to support viable 

populations of large carnivores and herbivores over 

the long term (Belovsky 1987, Grumbine 1990).

Animals with large area requirements face the problem of inadequate fragment size, but in addition these 

and all species face the problem of barriers to dispersal. 

Even for animals that have the ability to travel long 

distances, the terrain that must be traversed to move 

from one fragment to another is often so vast and hostile 

in nature that they stand little chance of surviving 

the trip. Roads are one potential barrier to dispersal 

and are a major cause of habitat fragmentation. One 

obvious consequence of roads is direct mortality. It 

is estimated that one million vertebrate animals are 

killed on roads in the United States every day (Lalo 

1987). For the Florida panther, a wide-ranging species 

whose endangered status stems largely from habitat 

fragmentation, roadkill is the single greatest source 

of mortality (Meffe and others 1997). Roads also serve 

to block the movement of animals, both small and 

large, effectively isolating populations within habitat 

fragments. Many species of small mammals have been 

found to cross roadways only rarely, if ever (Adams and 

Geis 1983, Garland and Bradley 1984, Mader 1984, 

Oxley and others 1974). The same has been found 

for some carabid beetles (Mader 1984), and animals 

as large as black bears may find roads a barrier to 

movement (Brody and Pelton 1989). Even a 3-m dirt 

track was found to deter the movement of prairie voles 

and cotton rats in a Kansas grassland (Swihart and 

Slade 1984). Fencing of rangelands may also serve 

as a barrier to movement for large grassland species. 

Pronghorn, for example, normally travel across wide 

ranges but are restricted in their movements by their 

inability or reluctance to jump fences, potentially leading to death in cases where the animals are unable 

to escape particularly severe winter weather (White 

1969, Wilson and Ruff 1999, Yoakum 1978). Although 

fences have now been designed to allow passage of 

pronghorn (Yoakum and others 1996 and references 

therein), they are not widely used, and recent studies 

demonstrate that fencing still serves as a barrier to 

natural pronghorn movements in the Southwest (van 

Riper and others 2001).

The inability of individuals to move freely between habitat patches may interrupt the stability of 

metapopulations, leading to their eventual decline 

and local extinctions. Key source populations may be 

eradicated in the process of fragmentation, or barriers 

such as roads, agricultural fields, or other inhospitable 

altered habitat may simply impede the dispersal of individuals to the point that the potential colonists required 

to shore up satellite populations are eliminated. Real 

world metapopulations in fragmented landscapes, such 

as that of the endangered bay checkerspot butterfly, 

closely follow the predictions of the theory of island 

biogeography: the probability of extinction of satellite 

populations increases with isolation from the source 

population and declines with increasing patch area 

(Thomas and Jones 1993).

Problem of Small, Isolated 

Populations_______________________

Long-lived species in particular may persist for 

many years following fragmentation due simply to the 

longevity of the individuals making up the population. 

Unless successful reproduction and recruitment is taking place, however, this species will disappear from the 

fragment as these individuals die out. For small populations in a fragmented landscape, the impediments 

to reproduction and recruitment are many. Simply by 

chance, the demographics of the population may not be 

conducive to successful reproduction; the age structure 

and sex ratio of the remaining few individuals are 

critical. A classic example of demographic misfortune 

is the dusky seaside sparrow: this endangered species 

was eventually reduced to a population of only six 

individuals, all of whom were male, thus dooming the 

species to extinction (Kale 1983). Successfully locating 

a mate is key to reproduction for most species, but 

fragmentation of the habitat may make it difficult for 

potential mates to find each other. Several studies 

have found a greater percentage of unmated male birds 

in small habitat fragments, indicating that females 

may not be able to locate them in isolated patches 

(Gibbs and Faaborg 1990, Robinson 1988, Simberloff 

and Gotelli 1984, Van Horn and others 1995, Villard 

and others 1993). Predation or parasitism may occur 

at greater levels in habitat fragments, thus reducing 

the reproductive success of individuals residing there 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 75

(see discussion in the “Edge Effects” section). This 

last point underscores the importance of productivity data for estimating the viability of populations; 

numerous studies have shown that the abundance 

and/or density of individuals or nests are not reliable 

indicators of habitat quality (Maurer 1986, Van Horne 

1983, Vickery and others 1992, Zimmerman 1992) or 

of nest success (Johnson and Temple 1990, Vickery 

and others 1992, Zimmerman 1984). Simply because 

large numbers of individuals of a particular species 

are found in a habitat fragment does not necessarily 

mean that the fragment is capable of supporting that 

species over the long term.

In addition to these problems, the small size and 

isolated nature of fragmented populations makes 

them vulnerable to other random processes. Natural 

catastrophes such as floods or fires may eliminate 

the few remaining individuals of a small population 

purely by chance. Random environmental changes may 

prove disastrous for such a population; a prolonged 

drought, for example, might lead to the extinction of 

a population of pupfish when the spring that they live 

in dries up. The genetic structure of populations may 

be profoundly affected by isolation. The interruption 

of gene flow among individuals in subpopulations may 

result in increased genetic drift, population bottlenecks, 

and inbreeding, all of which could lead to the fixation 

of deleterious alleles and decreased genetic diversity 

(Falconer 1981, Lerner 1954, Ralls and Ballou 1983, 

Wright 1977). Any of these factors—catastrophes, environmental variations, altered gene flow—as well as 

changes in demographic structure, might potentially 

lead to the extinction of a small, isolated population 

(Shaffer 1981). In reality, however, it is more likely a 

synergistic interaction between two or more of these 

factors that ultimately leads to the extinction of such 

populations in a process that is called an “extinction 

vortex” (Gilpin and Soulé 1986).

Of all these processes, the genetic consequences of 

isolation and interrupted gene flow in particular have 

received a great deal of attention by conservation biologists. In general, population bottlenecks, inbreeding, 

and the loss of genetic diversity are all believed to 

have a negative impact on the fitness of individuals 

through decreased fecundity and survivorship, a condition known as inbreeding depression (Falconer 1981, 

Lerner 1954, Ralls and Ballou 1983). The negative 

effects of inbreeding depression have been witnessed 

primarily in captive animal populations, but such 

impacts have also been documented in small wild 

populations that have become isolated, such as the 

lions of the Ngorongoro Crater in Africa that exhibit 

high levels of sperm abnormalities and low reproductive 

success (Packer and others 1991). However, there are 

also examples of small, isolated populations that have 

either retained relatively high levels of genetic variability (for example, one-horned rhinos; Dinerstein and 

McCracken 1990) and/or have survived severe population bottlenecks with no apparent problems stemming 

from inbreeding (for example, elephant seals; Bonnell 

and Selander 1974). Plants in particular seem to be 

resistant to the negative effects of inbreeding, most 

likely an adaptation to the limited dispersal ability of 

many species and self-fertilization (Barrett and Kohn 

1991), although reductions in genetic diversity have 

been correlated with decreased fecundity in some plants 

found in isolated patches (Baur and Erhardt 1995). 

Although the impacts may be variable, the changes in 

gene flow and reduced number of individuals resulting from habitat fragmentation have the potential to 

significantly impact both the demographic and genetic 

structure of remnant populations (Fahrig and Merriam 

1994). In the short-term, such alterations may be 

reflected in the reduced reproductive capacity and 

survivorship of individuals, possibly leading to localized 

extinctions for some species. On an evolutionary time 

scale, the reduced genetic variability stemming from 

processes such as fragmentation impedes the ability 

of individuals to respond to selection pressures, possibly leading to the extinction of the species (Frankel 

and Soulé 1981).

Corridors and Connectivity in 

Fragmented Landscapes____________

A central tenet of applied conservation biology has 

been the maintenance or reconstruction of habitat 

corridors to achieve connectivity between fragments 

(Meffe and others 1997, Preston 1962, Saunders and 

others 1991, Shafer 1990). In theory such corridors 

would essentially reconnect an otherwise fragmented 

landscape, facilitating the movement of individuals between patches, enabling continued gene flow, 

maintaining metapopulation dynamics, and reducing 

mortality of animals attempting to disperse through 

hostile terrain. In practice such corridors have produced 

mixed results. Most experiments have found that corridors do enhance movement, although this has been 

true primarily for small, less-mobile animals (Debinski 

and Holt 2000; but see Haas 1995). Small mammals 

such as chipmunks may use treelines to successfully 

colonize wooded patches in a fragmented forest landscape (Henderson and others 1985), and a few species of 

invertebrates have been found to preferentially utilize 

corridors in grassland fragments (Collinge 1998, 2000). 

In arid grasslands, movements of beetles in the genus 

Eleodes are strongly affected by vegetation structure 

(Crist and others 1992, Wiens and Milne 1989), and 

these beetles are one of the few species known to use 

corridors (Collinge 2000). Neotropical migratory birds, 

on the other hand, decrease in habitat fragments 

regardless of the degree of connectivity (Debinski and 
76 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Holt 2000). In the case of highly mobile organisms, the 

ability to move between fragments is most likely not 

the factor limiting populations. Corridors also do not 

appear to assist organisms with limited mobility; the 

habitat corridors provided by roadsides or ditches did 

not prove effective for colonization of habitat fragments 

by plants with short-range dispersal mechanisms (Van 

Dorp and others 1997). In general, many questions 

remain regarding the efficacy of corridors, particularly 

as they can serve not only for dispersal but also as 

conduits for predators, parasites, and disease transmission (see Noss 1987).

Impacts of Fragmentation on 

Grassland Plants___________________

Due to their small area requirements, plants are 

often proposed to be relatively immune to habitat 

fragmentation (Noss and Cooperrider 1994). Some 

short-term studies have found that small fragments 

support just as many plant species as large fragments 

and that rare species appear to persist in such small 

fragments (Simberloff and Gottelli 1984). However, 

studies of fragmented grasslands in Europe using 

historical records have documented high levels of 

extinctions of plants when followed over the long 

term (Fischer and Stöcklin 1997). Although the 

total numbers of species at each site were essentially unchanged, there were significant increases 

in habitat generalists, while habitat specialists 

that formerly occupied the sites had disappeared. 

The interruption of plant-pollinator interactions, 

leading to reduced viability of plant populations in 

fragments, may be one factor contributing to local 

extinctions. Isolation of patches has been found to 

diminish both the abundance and species richness 

of bees, butterflies, and other pollinators (Debinski 

and Holt 2000, Jennersten 1988, Steffan-Dewenter 

and Tscharntke 1999). Fewer visits by pollinators 

can lead to reduced fecundity, viability, and decreased 

genetic diversity in plants isolated in habitat remnants (Aizen and Feinsinger 1994, Baur and Erhardt 

1995, Jennersten 1988). The small population size 

and isolation of plants in habitat fragments may 

also result in inbreeding depression and loss of 

genetic diversity through founder effects, random 

genetic drift, and inbreeding (Rajmann and others 

1994, Templeton and others 1990, van Treuren and 

others 1991, Young and others 1996, 1999). In some 

cases genetic variation continues to remain high in 

isolated plant populations, although even in these 

cases rare alleles may be lost from smaller fragments 

(Young and others 1999). In terms of negative genetic 

impacts, wind-pollinated species appear to be much 

more resilient to the effects of fragmentation (Fore 

and others 1992, Young and others 1993).

Summary of Habitat Fragmentation 

Effects____________________________

Not all organisms respond in the same manner 

to habitat fragmentation; persistence in habitat 

fragments, impacts on dispersal abilities, and use of 

corridors are highly species specific (Debinski and Holt 

2000). Nor are all habitat fragments created equal; 

the size of the fragment, shape, amount of edge, and 

nature of the surrounding matrix will all influence the 

nature of the impacts on the individuals residing in the 

remnant patch (Helzer and Jelinski 1999). Taken as a 

whole, however, the majority of evidence from studies 

of habitat fragmentation indicates that the loss and 

isolation of natural habitats pose a strong threat to 

both regional and global biodiversity. Fragmentation 

greatly increases the risk of extinction for the native 

species of the original habitat through several mechanisms, including:

• Loss of habitat area for interior species.

• Barriers to dispersal, colonization, and maintenance of metapopulation dynamics.

• Random alteration of demographic and genetic 

structure resulting from isolation and small 

population size.

• “Edge effects” such as increased predation and 

parasitism, and invasion by exotic species or 

habitat generalists.

• Interference with biotic relationships, such as 

plant-pollinator interactions.

• Alteration of the physical environment, ecological 

processes, and natural disturbance regimes.

Strategies for counteracting the effects of habitat 

fragmentation include:

• Preventing or minimizing further fragmentation.

• Managing lands to restore natural disturbance 

regimes.

• Maintaining or restoring large natural areas to 

act as avenues for dispersal and genetic mixing 

of populations.

• Restoring habitat to increase the size of remaining 

habitat patches or buffer existing patches (Meffe 

and others 1997, Shaffer 1990).

Some basic design principles to counteract the effects 

of habitat fragmentation are summarized in figure 4-3. 

In an already fragmented landscape, it may be necessary to strive for protection of the largest possible 

area through strategies such as creating corridors by 

restoring connections between natural areas (Gatewood 

1998). However, the importance of preserving extensive 

landscapes in order to maintain ecosystem processes 

cannot be overemphasized; wherever possible, such a 

strategy should be the first line of defense.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 77

The vast majority of information that we have on 

the consequences of habitat fragmentation is based on 

studies of forested ecosystems or tallgrass prairie. One 

of the greatest needs for research in the grasslands 

of the Southwest is the need to study the impacts of 

fragmentation in these arid grassland systems.

Acknowledgments_________________

Conrad Bahre, Dave Gori, Nathan Sayre, and Peter 

Warren were all kind enough to provide comments 

on the draft manuscript. Their constructive criticism 

was extremely useful in helping me to present a 

more balanced discussion of the threats to grassland 

biodiversity in the Southwest. Pam Stoleson provided 

helpful editorial suggestions to improve the manuscript 

as well. Any errors, omissions, or biases remaining 

in this manuscript are the sole responsibility of the 

author.

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86 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Introduction_______________________

This chapter encompasses the lands of the Southwest 

as defined by Region 3 of the USDA Forest Service 

(USFS): Arizona, New Mexico, and portions of western 

Oklahoma and the Texas Panhandle. I examine human 

use and modification of the grasslands/rangelands of 

this region, with an emphasis on those areas managed by the Forest Service. Because the majority of 

publications serving as sources for this chapter use 

both “rangelands” and “grasslands” when referring to 

a variety of different grassland and rangeland vegetation types, this section does not distinguish between 

the terms. An exhaustive examination of all human 

uses and related topics and issues in Southwestern 

rangelands is well beyond the scope of this discussion; 

thus, a selective review of historical and contemporary 

topics is presented. The section begins with a review of 

continuous regional land use and modification from preEuro-American settlement (that is, American Indian 

times) to the present to serve as a background for 

understanding current land uses and land-use-related 

problems and issues. I then examine contemporary 

rangeland condition and the major human uses and 

activities affecting these lands, focusing on domestic 

livestock grazing, mineral extraction, and recreation. I 

conclude the chapter with a brief discussion of continuing and future trends in Southwestern rangelands and 

rangeland management.

American Indian (Pre-Euro-American 

Contact) Land Use_________________

Paleoindian Period

The earliest undisputed evidence of human occupation of the Southwest began between 11,650 and 10,250 

years ago. Archeological complexes from this period are 

referred to as Paleoindian and represent the remains 

of ancestral Native Americans from late Pleistocene 

times. These people were hunters and gatherers using 

both plant and animal food sources, under climatic conditions and vegetational patterns that were markedly 

different from those of today. Areas from which data 

have been obtained indicate that climatic conditions 

were generally wetter than present with less seasonal 

variation resulting in relatively mild winters and cool 

summers (discussed in Cordell 1997: 67-100).

Owing to the paucity of archeological remains from 

this period and to the fact that such remains are often 

deeply buried, interpretations of Paleoindian economic 

practices and resource use are sketchier and less well 

documented than those from later times. According to 

the most commonly accepted scenarios, relatively small 

groups of highly mobile hunters and gatherers exploited 

a variety of plants and animals, including Pleistocene 

megafauna such as mammoth (Mammuthus columbi), 

at least during the early portion of the period. In later 

Paleoindian times there was considerable reliance on 

Chapter 5:

Historic and Contemporary Land Use in 

Southwestern Grassland Ecosystems

Carol Raish
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 87

bison (Bison spp.), including a now extinct larger form 

(Bison antiquus), as well as on modern fauna. In the 

eastern Southern Plains portions of the study area 

extending from present-day eastern New Mexico into 

present-day Texas and Oklahoma, human groups relied 

heavily on hunting bison supplemented by smaller 

game and plant foods. Studies have shown that bison 

numbers and range expanded and contracted throughout the Holocene with expansion and contraction of the 

grasslands, tied to climatic fluctuations and available 

moisture. Because the Southern Plains and Southwest 

are generally drier than more northern areas of the 

Plains, reliance on bison hunting may not have been 

a productive strategy at all times during this period. 

Sufficient numbers of animals may not have been present during periods of grassland contraction (Cordell 

1997: 95; Reher 1977).

On the western margins of the Plains and in the 

foothills and mountains adjacent to the Plains, there 

was apparently a fairly equal reliance on hunting and 

gathering. The role and importance of bison hunting 

in these areas fluctuated to an even greater degree as 

bison range expanded and contracted. Mountain sheep 

(Ovis canadensis) and mule deer (Odocoileus hemionus) 

were also present and hunted in about equal numbers 

to bison. In the more western portions of the region, 

human groups were dependent upon plant foods and 

relatively small, nonmigratory game (Cordell 1997: 

90-100).

Especially on the Plains and in the areas adjacent to 

the Plains, human population growth was tied to the 

availability of bison and the grasslands to support them. 

Thus, to understand important questions concerning 

the timing and effects of increasing regional populations, additional paleoenvironmental reconstruction 

studies are needed to determine the expansion, contraction, and spread of grasslands during this period. Little 

is also known concerning human manipulation of the 

physical environment in Paleoindian times (Cordell 

1997: 90-100), although use of fire has been mentioned 

as a means of moving game during hunting drives 

(Bahre 1995: 232-235). Clearly, additional archeological and archeoenvironmental research is needed for 

this period to clarify Paleoindinan resource use and 

environmental manipulation strategies.

Archaic Period

Around 7,500 years ago (5500 B.C.) the Paleoindian 

period gave way to what is referred to as the Archaic, 

lasting until about A.D. 200. From about A.D. 200 

until Euro-American contact, agricultural groups of 

varying intensity occupied the diverse environments 

of the Southwest. During this period from approximately A.D. 200 until the mid-1500s, Southwestern 

climate and vegetation assumed essentially modern 

forms (Cordell 1997: 101-102). There was a gradual 

climatic warming and drying from around 7,000 to 

5,000 years ago (5000 to 3000 B.C.). This warm and dry 

period was followed by a gradual increase in moisture 

with a corresponding increase in woodland and forest 

vegetation (Hall 1985, Periman and Kelly 2000). A 

cooler period followed from approximately A.D. 1450 

until 1900, which is referred to as the Little Ice Age 

(Kreutz and others 1997, Periman and Kelly 2000). As 

a whole, this entire period has been marked by climatic 

variability with both broad-scale and periodic climate 

fluctuations (Periman and Kelly 2000).

During the Archaic, human groups continued to rely 

on hunting and gathering but with a greater focus on 

food plants and locally available resources. Hunted 

game were all modern forms (Cordell 1997: 101-102). 

Although subsistence patterns varied throughout 

the Southwest, some generalizations can be made. In 

many areas, people relied on uplands with high topographic and vegetational diversity, which allowed use 

of a wider range of resources in a smaller geographic 

area (Gunnerson 1987, Tainter and Tainter 1996). 

Grasslands were also favored resource procurement 

locales of Archaic foragers.

Archeological information from various locations, 

such as from sand dune sites in the San Juan Basin 

and from grassland sites in southeastern Arizona, 

indicates heavy reliance on the seeds of weedy plants 

and grasses. These include Indian ricegrass (Oryzopsis 

hymenoides), dropseed (Sporobolus spp.), goosefoot 

(Chenopodium spp.), pigweed (Amaranthus spp.), mustard (Descurainia spp.), tickseed (Corispermum spp.), 

and mallow (Sphaeralcea spp.). Both goosefoot and 

pigweed favor disturbed ground and may have thrived 

and/or been semicultivated in areas where humans 

camped repeatedly (Bahre 1995: 233; Cordell 1997: 

120-121). Small, medium, and large-sized game were 

taken. Faunal remains from cave sites in the highlands 

along the central New Mexico-Arizona border include 

bison (Bison bison), pronghorn antelope (Antilocapra 

americana), mountain sheep (Ovis canadensis), and 

deer (Odocoileus spp.). Rabbit (Leporidae) and deer 

were found at Ventana Cave in the low Sonoran Desert 

of southern Arizona (Cordell 1997: 122).

Southern Plains groups were apparently not heavily dependent upon bison during the Early Archaic 

but were concentrating on deer, small mammals, and 

plants. By later portions of the Archaic, however, bison 

increase in the area’s archeological sites (Cojeen 2000, 

Drass and Turner 1989: 20-22; Simpson and others 

1998). Bison may become a more abundant and predictable resource in the later Archaic owing to improved 

shortgrass range brought about by wetter conditions 

(Cojeen 2000, Creel and others 1990).

Ranging from about 1500 B.C. to 1000 B.C. in the 

low deserts, central mountains, and northern plateaus 

of the Southwest, human groups began to cultivate 
88 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

corn (Zea mays). However, they remained hunters and 

foragers with a mobile lifestyle for many hundreds of 

years thereafter, using corn and later beans (Phaseolus

spp.) and squash (Cucurbita spp.) to varying degrees 

in their diets. Even with the considerable emphasis on 

agriculture found among many Southwestern groups 

immediately prior to Euro-American contact, wild foods 

always retained an important role in diet and survival 

(Cordell 1997: 147-151).

Puebloan and Other Agricultural Groups

Beginning around A.D. 200 and continuing until 

European contact and colonization in the mid-to-late 

16th century, people became more sedentary and more 

committed to the production of agricultural crops. 

The widespread appearance of permanent dwellings, increasingly larger settlements, and ceramics 

are indicators of these trends. Subsistence patterns 

varied throughout the region depending upon both 

agricultural productivity and locally available wild 

resources (Cordell 1997: 221-224). Settlement locations also varied, most probably conditioned by both 

economic and defensive factors.

As dependence on agriculture increased, settlement 

location moved away from high eminences and upland 

areas to alluvial terraces and benches associated with 

major rivers and their tributaries, and to arroyos and 

intermittent watercourses in the Four Corners area, 

southern New Mexico, and southeastern Arizona. 

Agricultural occupations in south-central Arizona, 

however, were located on open flatlands without apparent defensive considerations from the beginning 

(Cordell 1997: 241). The Southwestern farming groups 

are ancestral to contemporary Native Americans in 

the region and included the Anasazi or Ancestral 

Puebloan of the Four Corners and Rio Grande areas; 

the Mogollon of southern New Mexico, southeastern 

Arizona, and northern Mexico; the Hohokam of southcentral Arizona; and the Patayan of western Arizona 

(discussed in Periman and Kelly 2000).

In northeastern New Mexico, western Oklahoma, 

and the Texas Panhandle, groups of the Plains Village 

Tradition practiced hunting and gathering and grew 

corn along the creeks and watercourses of the region. 

Although agriculturally based, these groups are considered to have had a greater reliance upon hunting 

deer and bison than the Puebloan groups to the west 

(Fredine and others 1999). A mixed dependence on 

agriculture and bison hunting continued until around 

A.D. 1500 when agricultural groups were pushed south 

and west by incursions of nomadic Plains Apachean 

groups who were both raiding and trading with the 

Pueblos bordering the Plains (Fredine and others 1999, 

Glassow 1972). Migration of Plains agricultural groups 

as a result of increasingly drier climates has also been 

suggested, as has coalescing or amalgamating into 

preexisting nomadic cultures such as the Kiowa Apache 

(Baugh 1984, Cojeen 2000, Hughes 1991: 43; Simpson 

and others 1998). Continuity between prehistoric and 

protohistoric groups in the area continues to be a matter of debate (Cojeen 2000, Drass and Turner 1989: 

28; Hughes 1991: 43).

As in earlier times, grasslands were used as 

sources of wild plant and animal foods. Bahre (1995: 

234) concludes that although there is not a great 

deal of information concerning American Indian use 

of southeastern Arizona’s grasslands, groups were 

known to collect vegetal foods from arborescent and 

succulent taxa in the Desert Grasslands and to harvest grass seeds in the Plains Grasslands (Huckell 

1995). Especially favored were the seeds of big sacaton (Sporobolus wrightii) and browntop panicgrass 

(Brachiaria fasciculate) (Bahre 1995: 234; Doebley 

1984). In describing animals hunted by Puebloan 

and other groups of the Middle Rio Grande, Scurlock 

indicates that bands of hunters journeyed to the Llano 

Estacado (plains of eastern New Mexico and western 

Texas) in the fall to hunt bison for both meat and hides 

(discussed in Scurlock 1998: 98).

In addition, bison ranged seasonally into the San 

Augustín Plains and the grasslands of northeastern 

Arizona in late prehistoric times. A herd was reported 

in the Chama Valley as late as 1690 (Callenbach 1996: 

17-18; Scurlock 1998: 209). There was also considerable trade in raw materials and food stuffs among 

regional groups occupying the Puebloan areas and 

those areas farther to the east on the Plains proper 

(Scurlock 1998: 104-105). As mentioned previously, 

the resident Southern Plains groups focused heavily 

on both hunted and gathered grassland resources 

(Cojeen 2000, Fredine and others 1999, Simpson and 

others 1998).

Understanding Pre-Euro-American Contact 

Land Use

Understanding the role, extent, and importance of 

American Indian use and manipulation of Southwestern 

environments prior to Euro-American contact is critical 

for understanding contemporary land conditions, as 

well as the past “reference conditions” called for in studies of ecosystem management and restoration. Tainter 

and Tainter (1996: 28-29) recommend a combination of 

information derived from contemporary environmental 

sciences and social sciences with information derived 

from the historical sciences to elucidate past land 

use and management practices. Both ethnohistoric 

and archeological research into past environmental 

conditions and land use practices can provide valuable 

information. Periman and Kelly (2000: 27-28) discuss 

the role of archeological data in describing the impact of 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 89

Native American use on certain Southwestern riparian 

ecosystems in terms of reductions in tree cover from 

cutting roof timbers and firewood and from tree clearance for agricultural fields (Le Blanc 1985). Ongoing 

studies of changing landscape use in the Rio del Oso 

valley of northern New Mexico are providing both new 

information and new methodologies for determining 

past use and manipulation of riparian, woodland, and 

grassland ecosystems (Periman 2001).

These studies will make a timely contribution as 

information on prehistoric human use and manipulation of grassland ecosystems is sorely needed. Much of 

the available information (and more is needed) focuses 

on the use of fire by aboriginal peoples to manipulate 

their environment. According to Scurlock (1998: 201, 

268-269), Native American groups in the Rio Grande 

Basin intentionally burned the grasslands periodically, which, along with lightning-caused fires, may 

have killed encroaching half shrubs and shrubs and 

stimulated vigorous growth of grasses. Although Baisan 

and Swetnam (1997) generally conclude that natural 

processes account for fire frequencies at most of their 

60 study sites in Arizona and New Mexico, they suggest that fire frequencies at sites in the Manzanita 

and Sandia Mountains of the Rio Grande Valley during pre-Euro-American settlement times may have 

been influenced by human-caused ignitions. High 

fire frequencies prior to Euro-American settlement 

in these areas where lightning occurrence is low led 

them to the conclusion that American Indian groups 

were igniting fires for resource manipulation (Baisan 

and Swetnam 1997).

In discussions of southeastern Arizona, Bahre (1995: 

234-235) concludes that the extent to which American 

Indian groups influenced grassland ecology is unknown 

but that the accidental or intentional introduction of 

fire to the grasslands may have contributed to their 

largely brush-free state at the time of Euro-American 

contact. However, the full role of fire in maintaining 

desert grasslands is unknown (Bahre 1991: 138-141; 

Dick-Peddie 1993: 106-107).

American Indians used fire to drive game, stimulate the growth of understory and food plants such 

as berries, clear areas for campsites and agriculture, 

and produce nutrient-rich ashes in fields (taken from 

a discussion in Scurlock 1998: 269). Indeed, Sullivan 

(1996: 145-156) argues that Western Puebloan groups 

on the Colorado Plateaus actively managed their habitats to increase production of wild resources such as 

Indian ricegrass, sunflower (Helianthus sp.), and 

goosefoot by controlled burning of the pinyon-juniper 

woodland. We look forward in the coming years to 

archeoenvironmental research that will help to clarify 

the past human role in maintenance and manipulation of pre-Euro-American contact North American 

ecosystems.

Contemporary American Indian Uses 

of National Forest Grassland Areas___

Contemporary American Indian groups use Western 

grasslands, many of which occur on public land, in a 

variety of ways, including hunting, and gathering plant 

materials for food, medicines, and basketry. They also 

use these lands for grazing domesticated animals. The 

following brief review cannot do justice to the complex 

topic of Native American wild and domesticated plant 

use and its body of ethnobotanical literature. Especially 

important, detailed information on this topic can be 

found in the work and publications of Gary Nabhan 

(1985, 1989), founder of Native Seeds/Search. Nabhan 

has worked to gather and preserve seeds used in aboriginal agriculture and to document Native American 

gathering and farming practices in the Southwest in 

grasslands, as well as in the Sonoran Desert region.

In recent years, the Forest Service has collaborated 

with indigenous groups to learn from their traditional 

practices and to assist in assuring continued supplies 

of materials needed for economic pursuits and craft 

production. As examples, several forests in California 

have cooperated with Native American groups to identify and assure the continued propagation of plant 

materials such as the bear grass (Nolina spp.) used in 

basket making. Information on the use of bear grass 

and other grassland plants in the Southwest can be 

found in Bell and Castetter (1941).

The Hopi have worked with the Kaibab National 

Forest in northern Arizona to develop a comprehensive plan to manage pinyon-juniper woodlands and 

rangelands to protect cultural and spiritual values, 

while addressing concerns related to watersheds, 

soils, wildlife, recreation, range, and other resources 

(Thakali and Lesko 1998). The Apache, Navajo, and 

Hopi still collect plants from various meadows on 

the San Francisco Peaks, Coconino National Forest, 

Arizona, and the Hopi consider Bonito Park, by Sunset 

Crater, New Mexico, a Traditional Cultural Property 

(Pilles personal communication 2002).

The Tohono O’odham collect bear grass for basketry, 

and the San Carlos and White Mountain Apache gather 

medicinal herbs in grassland and mountain meadow 

areas of the Coronado National Forest in southern 

Arizona. The Mescalero Apache report collecting small 

amounts of coral beans (Erythrina flabelliformis) 

from the Chiricahua and Dragoon Mountains. The 

Coronado is currently working on draft memoranda 

of understanding with the Apache Tribes that would 

authorize collecting of small amounts of plants and 

acorns by Tribal members for personal use without 

permit. A potential conflict arises with the desire of 

some Tribal members to collect agave (Agave spp.) from 

the grasslands around Patagonia because agave is an 

important food source for the endangered long-nosed 
90 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

bat (Leptonycteris sanborni) in some areas (Farrell 

personal communication 2002). Discussions of the role 

and importance of agave as a food and fiber source to 

Native American groups are presented in Castetter and 

others (1938) and Gentry (1998), among others.

Prehistorically the Wichita and Affiliated Tribes occupied areas of western Oklahoma, northern Texas, 

and northeastern New Mexico that now comprise 

the Kiowa and Rita Blanca National Grasslands 

and the Black Kettle and McClellan Creek National 

Grasslands. In the 1700s and 1800s, horse-using groups 

such as the Cheyenne, Arapaho, Apache, Comanche, 

and Kiowa entered the area as a result of territorial 

dislocations caused by encroaching Anglo-American 

settlement. There is, however, little information concerning whether these groups still use the area for 

traditional practices, although they may consider it a 

traditional use area and/or important in their Tribal 

history (Black Kettle and McClellan Creek National 

Grasslands Geographic Area Assessments 2000, Kiowa 

and Rita Blanca National Grasslands Geographic Area 

Assessments 1999). Research to address these issues 

is planned for the near future (Benedict personal communication 2001).

Hispanic and Post-European Contact 

American Indian Land Use__________

The Spanish were the first Europeans to enter the 

Southwestern areas of present-day Texas, New Mexico, 

and Arizona beginning in the late 1520s with the shipwrecked Álvar Núñez Cabeza de Vaca. Over a period of 

about 6 years, Cabeza de Vaca and several companions 

made their way from the Texas coast through parts 

of New Mexico and probably Arizona to finally rejoin 

their countrymen on the west coast of Mexico (Stephens 

and Holmes 1989, Stout and Faulk 1974, Udall 1995, 

Weber 1992). Further expeditions followed in search 

of the fabled wealth reported by Cabeza de Vaca. In 

1540, Francisco Vásquez de Coronado led an expedition that covered portions of New Mexico, Arizona, 

and the Southern Plains, probably passing through 

parts of the Texas Panhandle and into present-day 

western Kansas. However, the expedition route onto 

the Plains remains the subject of considerable debate 

(Bahre 1995, Bolton 1949, Fredine and others 1999, 

Trimble 1989, Wildeman and Brock 2000).

New Mexico

Significant Spanish influence began with colonization of the Rio Grande valley in 1598 by Juan de Oñate, 

initiating both the political and biological conquest of 

the region. Although the exact numbers are debated, 

Oñate brought with him soldiers, colonists, priests, 

and Mexican Indian servants, along with cattle, sheep, 

goats, and horses (Baxter 1987, Hammond and Rey 

1953(1), Wildeman and Brock 2000). The settlers introduced domesticated plants and animals, as well as 

new technologies, which altered the flora, fauna, and 

landscape of the Southwest as has occurred throughout 

the New World (Crosby 1972, Melville 1994).

Throughout the 1600s, the Pueblo Indian populations of the Rio Grande declined as a result of diseases 

introduced by Euro-American contact, warfare, famine 

caused by a series of severe droughts, and destruction 

of food stores by raids from nomadic Indian groups. As 

the Native American population declined, the tribute 

and labor requirements of the Spanish colonists became more onerous. These conditions, along with forced 

relocations and intensive religious mission programs, 

led to the Pueblo Revolt of 1680. During this rebellion, 

the majority of Spanish were forced out of the Upper 

Rio Grande for 12 years. They returned in the period 

from 1692 through 1696 when Diego de Vargas initiated and completed the reconquest of New Mexico for 

the Spanish Crown (Simmons 1979).

Even though Hispano populations rose throughout 

the 1700s, the significant population declines of the 

Puebloan groups left a sufficient amount of land for 

both groups to farm and ranch along the rivers of the 

region. After the reconquest, the economic, political, 

and religious systems of New Mexico were significantly 

different from the earlier systems. The new generation 

of Spanish colonists were accomplished agriculturalists and stock raisers who worked their own land and 

maintained relatively cordial relations with the Pueblo 

Indian groups as both used the land in similar ways 

(Simmons 1979).

During both the Spanish Colonial and Mexican 

periods (1598 through 1846), land use and ownership 

throughout the Southwest were confirmed by land 

grants from the Spanish Crown or Mexican government. Although land grants were of several types, 

community grants to a group of settlers in common 

were of particular importance for later resource conflicts in the area (Eastman and others 1971, Harper 

and others 1943). Within community grants, settlers 

received individually owned building sites and agricultural plots of irrigated land near the ditch or 

stream. The irrigated plots were often quite small, 

averaging from 5 to 10 acres (Van Ness 1987). The 

villagers also used the grant grazing lands, timber 

lands, and pastures communally (Eastman and others 1971). Because kinsmen often worked their fields 

cooperatively and herded their animals together, they 

were able to manage on the small-sized, scattered 

agricultural plots.

Throughout the Colonial period, a subsistence 

economy based in small villages prevailed along the 

Rio Grande and its tributaries. Information on the 

community of Cañones (Kutsche and Van Ness 1981, 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 91

Van Ness 1987) provides a good description of farming 

and ranching in the Hispano villages. Both animal 

and plant production were critical for the success of 

the mixed farming system, with sheep and goats the 

most important stock for food. Cattle were used to plow 

fields, thresh grain, transport produce, and manure 

fields. The community stock were individually owned 

but cooperatively grazed. They were moved into the 

higher elevation pastures during the spring and summer and returned to the village after the harvest to 

graze and manure the stubble fields (Kutsche and Van 

Ness 1981, Van Ness 1987).

Livestock numbers were modest for the first two 

centuries after colonization. Raids by nomadic Apache, 

Navajo, Ute, and Comanche limited range expansion, 

commerce, and trade (Clark 1987: 19-23; Van Ness 

1987). Thus, economic production was for subsistence, 

not for competition in a commercial market. Sheep 

were more numerous than cattle and horses in the 

early years, primarily because of sale and loss of the 

latter to surrounding nomadic Indian groups (Gonzales 

1969). By the early 1800s, the number of sheep began to 

increase as the Spanish population expanded eastward 

onto the plains around present-day Las Vegas, across 

the Sandia and Manzano Mountains, and westward 

from the Rio Grande Valley. This movement resulted 

from growing human and animal populations, a decline 

in raiding by the nomadic groups, and an expanded 

trade in wool and sheep during the Mexican period 

(1821 to 1846).

Sheep were the mainstay of the livestock economy 

of New Mexico until after U.S. takeover in 1848 (Beck 

1962). Scurlock (1998: 116-117) reports 1 million sheep 

in New Mexico during the 1820s rising to 3 million 

by the mid-1800s with a reduction to approximately 

377,000 by 1850. The reduction occurred as a result 

of Apache and Navajo raids and losses from drought, 

blizzards, and predators. During the first half of the 

1800s, New Mexico supplied sheep to the West and to 

extensive mining operations in Mexico. Beck (1962) 

and Schickedanz (1980) state that annual sheep drives 

to Mexico often averaged 200,000 head between 1815 

and 1830. Baxter (1987: 90-95) offers more conservative 

figures, listing some 240,000 sheep in the Albuquerque 

and Santa Fe areas in 1827, with a total of approximately 200,000 driven south to Mexico between 1835 

and 1840. For example, about 80,000 were exported 

in 1835. In spite of the contradictory figures, sheep 

were an extremely important part of the economy 

throughout the period.

Sheep numbers also increased in the Navajo areas of 

northwestern New Mexico during the 1800s, so much 

so that some Navajos were described as wealthy by 

the time of the Mexican-American War (Brugge and 

Gerow 2000: 449-451). Scurlock (1998: 117) suggests 

that rising sheep numbers and Navajo practices of 

grazing outward from the hogan during the day and 

returning the animals to associated corrals at night may 

have initiated the first regional overgrazing west and 

north of the Puebloan and Hispanic settlements.

Herding sheep near Los Ojos, NM, in the 1990s. (Photo by Anne R. Baldwin)
92 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Although concentrations of sheep and cattle near 

Hispano and Puebloan settlements created areas of 

resource overuse during Spanish Colonial times, herds 

were generally small, and there was abundant land 

(Baxter 1987: 23-24; Rothman 1989: 196-197; Scurlock 

1995a: 15). Thus, relatively small populations of subsistence farmers successfully used the region’s resources 

over the long period of Spanish control (Raish 2000a: 

495). Overuse of favored areas intensified during the 

Mexican period as commercial sheep production increased (Scurlock 1995a: 15). By the end of the Mexican 

era, range deterioration was noted by American military personnel in some areas of the Rio Grande Valley, 

the Rio Puerco watershed, and on the Navajo lands 

higher up the Rio Puerco (discussed in Elson 1992, 

Scurlock 1998: 117). Nonetheless, the majority of farms 

and ranches remained small and subsistence oriented 

prior to U.S. conquest. On several visits to New Mexico 

in the 1830s, Josiah Gregg described the irrigation 

systems and subsistence orientation of agriculture 

in the small Hispano settlements, among his many 

other observations on the life and people of the area. 

He especially noted the fertility of the bottomlands 

in contrast to the barren condition of the unirrigated 

uplands (Gregg 1954: 104; Wozniak 1995: 34).

Arizona

To the west, Father Eusebio Kino is credited with 

establishing missions and bringing domesticated livestock and crops into southern Arizona beginning in the 

later 1600s (Bahre 1995: 235; Bolton 1919, Hadley and 

Sheridan 1995: 7-11), although domesticates may have 

reached the area earlier through trade with Spanish 

colonists to the east in the Rio Grande Valley (Bahre 

1995: 235-236; Bolton 1952). At about this time, 

Apaches also entered the area warring with resident 

Piman groups and later with the Spanish. By 1740 

Spanish settlers had begun ranching in the area, and 

by 1775 Tucson had been established.

Decreased hostilities with the Apaches after 1786 

led more Hispanic ranchers and miners to enter the 

region, but their numbers remained small. According 

to Bahre (1995: 236-237), their activities appeared to 

have had little impact on the grasslands. Although 

livestock may have degraded rangelands adjacent to 

settlements as in the Rio Grande Valley, their numbers 

were small and their effect on regional grasslands was 

probably minimal. In 1804, the two largest settlements, 

Tucson and Tubac, reported a total of 4,500 cattle and 

7,600 sheep. In addition, Tucson reported 1,200 horses 

(Bahre 1995: 236-237; McCarty 1976).

Although land grants have never played the role in 

rangeland use and management in Arizona that they 

have in New Mexico, there were a few Spanish land 

grants in southern Arizona, and Mexican stock-raising land grants were established in the grasslands of 

the San Pedro, San Rafael, and Santa Cruz Valleys of 

southeastern Arizona. Large numbers of cattle were 

reported on these ranches, but increased Apache raiding during the Mexican period drove off the majority of 

the ranchers by the mid-1800s (Bahre 1995: 237-240; 

Mattison 1946, Officer 1987, Ruyle and others 2000, discussed in Wildeman and Brock 2000: 5). Considerable 

numbers of livestock were also reported in the area of 

the Hopi villages in northeastern Arizona with one having 30,000 sheep in 1776 (Schickedanz 1980). Overall, 

both human populations and livestock numbers were 

low prior to the American period. According to Bahre 

(1995: 240), both Spanish and Mexican populations 

were concentrated in the upper Santa Cruz Valley, and 

the most significant human impact on the grasslands 

was probably the establishment of the large-scale 

Mexican ranches.

Bahre (1995: 238-240) provides a good listing of the 

several American military and scientific expeditions 

that passed through the region in the mid-1800s, 

and provides descriptions of the grasslands. He notes 

that many of them mentioned wild cattle on the 

abandoned Mexican land grants, but none provided 

written evidence of overgrazing. He says that the 

numbers of animals and their effects are not known 

and further states: “No doubt some grasslands were 

severely affected, but, given the descriptions of the lush 

grasslands in southeastern Arizona in the 1850s and 

1860s, ranching during the Mexican occupation seems 

to have caused little long-lasting and extensive disturbance” (Bahre 1995: 240). Many Spanish and Mexican 

archival records remain to be examined for information 

on the southeastern Arizona grasslands.

Southern Plains

The Early Historic period on the Plains of eastern 

New Mexico was also characterized by continuing 

conflicts between nomadic American Indian groups 

and Hispanic settlers. Highly mobile Comanche and 

Apache groups following the bison herds raided and 

traded with more sedentary Pueblos and Hispanic villages to the west. Actual Spanish settlement in the area 

began in 1794 with the establishment of San Miguel del 

Bado, the first Hispanic village east of Pecos Pueblo. 

Throughout the Spanish and Mexican periods, various 

attempts were made to settle and use the Plains grasslands, but these were generally not long lived owing to 

raiding by the nomadic groups (Baxter 1987, Fredine 

and others1999, Sebastian and Levine 1989a,b).

Farther to the east on the Plains of western 

Oklahoma, Spanish explorers out of New Mexico 

encountered nomadic hunters and traders, as well as 

the ancestors of the historic Wichita with whom they 

fought a battle on the Red River in 1759 (Hays and 

others 1989, Hofman 1989, Simpson and others 1998). 

By 1800 the area was controlled by France and became 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 93

part of the United States with the Louisiana Purchase 

in 1803 (except for the Oklahoma Panhandle, which 

became part of Oklahoma Territory in 1890). The land 

was designated Indian Territory in 1804 (Burgess and 

others 1963, Cojeen 2000).

Other Resource Use

Other resource use activities during the Spanish 

and Mexican periods, such as trapping, mining, and 

logging, played less of a role in the grasslands than in 

other areas of the Southwest. Extensive beaver (Castor

canadensis) trapping, first in New Mexico and then 

extending into Arizona, reached its peak in the 1820s 

and concluded by the late 1830s with the decline in 

popularity of the beaver hat. The large numbers of 

animals taken by the trappers caused severe reductions and extirpation of some local beaver populations 

leading to habitat degradation and decreasing water 

retention in streams and valleys (Scurlock 1998: 119-

121, 155-158; Wildeman and Brock 2000: 5-6, 11).

Small-scale mining occurred during the Spanish 

period in both Arizona and New Mexico with the first 

large-scale mining venture in New Mexico beginning 

with the copper mines at Santa Rita in 1804 (Gregg 

1954, Scurlock 1998: 118-119; Wildeman and Brock 

2000: 6). The impacts of mining did not become significant until the American period, however. Extraction 

of mineral resources was slow in the area owing to 

the nature of the ores, the lack of transportation, 

and raiding by the Navajo, Apache, Comanche, and 

Ute (Hadley and Sheridan 1995: 46-47; Paul 1963, 

Wildeman and Brock 2000: 6).

Although large-scale commercial logging did not 

occur during the Spanish and Mexican periods, there 

was considerable use of pinyon (Pinus edulis), juniper 

(Juniperus spp.), ponderosa pine (Pinus ponderosa), 

and fir (Abies spp.) for construction and for fuel wood. 

Pinyon and juniper were the preferred fuel woods for 

both Hispanos and Native Americans leading to intensive exploitation of the woodlands around settlements. 

Scurlock (1998: 118) states that as more land grants 

were established in the plains and mesas of eastern 

New Mexico, local residents packed fuel wood back to 

population centers in the Rio Grande valley for sale 

(discussed in Scurlock 1998: 117-118). This practice 

may have had considerable effect on local vegetation. 

Research in southwestern New Mexico has shown that 

precontact Native American groups profoundly affected 

local vegetation, greatly reducing tree cover in riparian 

and other areas. Areas were cleared for agriculture, and 

wood was routinely harvested for fuel and construction 

timbers. As an example, before A.D. 1000 cottonwood 

was a common source of fuel wood and construction materials. This source of wood became depleted, however, 

and wood charcoal remains show that groups were being 

forced to go farther from communities and use upland 

trees as substitutes because cottonwoods were no longer 

available (Le Blanc 1985, Periman and Kelly 2000).

Environmental Manipulation

As in earlier times, both the American Indian and 

Hispano populations of the Southwest manipulated 

the environment to enhance production of desired resources. Fire was often the management tool of choice 

in these endeavors. In New Mexico, Hispano ranchers 

and farmers reportedly burned forests to create grazing lands and drive game. They also burned livestock 

pastures to stimulate new grass growth and burned 

sheep rangelands to kill encroaching woody species 

(Allen 1984, Scurlock 1998: 269).

Wildfires were relatively common in the grasslands of southeastern Arizona during Spanish and 

Mexican times, based on historical records and fire 

scar data (Bahre 1985, 1991, Dobyns 1981, Pyne 

1982). Bahre (1995: 240) suggests that American 

Indian groups, mainly the Apache, were responsible 

for the burning but also states that the incidence 

of lightning-caused fires is so high in the area that 

the cause of the burning is difficult to determine. 

Continued research, both archival and archeological, 

is needed to understand the incidence and scope of 

these practices and their full effects on both past 

and present landscapes.

Anglo-American Land Use__________

The majority of the lands in the study area 

became part of the United States at the conclusion 

of the Mexican-American War (1846 through 1848). 

Western Oklahoma and the extreme southern portions 

of Arizona and New Mexico were acquired with the 

Louisiana (1803) and Gadsden (1853) Purchases, respectively. The Oklahoma Panhandle and the eastern 

portion of New Mexico were part of disputed lands 

claimed by both Texas and Mexico. With the end of 

the war, these lands went to Texas and were turned 

over to the Federal government in 1850.

Changing Patterns of Land Ownership 

and Use

Annexation of the region ultimately led to changes 

in both landownership and patterns of range use especially in New Mexico. Differences in American and 

Spanish land laws combined with unscrupulous land 

speculation resulted in the loss of over 80 percent of 

the Spanish and Mexican land grants by their original 

owners (Eastman and others 1971, Harper and others 

1943, Westphall 1965). Under the Treaty of Guadalupe 

Hidalgo, the United States was supposed to recognize 

and respect the property rights of the landowners of 

the region.
94 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

To obtain sound titles according to U.S. law, however, 

land grantees had to petition for title confirmation, at 

first through the Surveyor General to the Congress 

and after 1891 to the Court of Private Land Claims 

(Eastman and others 1971, Eastman and Gray 1987). 

To accomplish this, claimants usually had to hire 

an attorney, file their claim, and obtain necessary 

supporting documents. Eastman (1991: 103) notes, 

“...landholders were turned into claimants who had 

to incur a substantial expense to have their property respected.” Money was scarce in the subsistence 

economy of New Mexico, so many claimants signed 

over portions of their land as payment for legal fees. 

Because of this, even successful claimants lost land 

since the legal fees often amounted to from one-third 

to one-half of the land involved (Eastman 1991: 103). 

Additionally, many land claims were rejected; about 

24 percent of the acres claimed in New Mexico were 

confirmed compared to about 73 percent in California 

(Ebright 1987: 33). Lands from unconfirmed claims 

became part of the public domain.

The Surveyor General and the Court of Private 

Land Claims did not confirm grants for a variety of 

reasons. Boundaries were sometimes vague, original 

titles may have been lost, and communal ownership of 

pasture and woodlands did not conform to 19th century 

American ideas of private ownership (Eastman and 

others 1971). The court often confirmed house lands 

and irrigated plots but did not confirm community 

pastures, surrounding rangelands, and woodlands that 

had always provided the Hispano villagers with their 

primary grazing and fuel wood resources. Villagers 

also lost confirmed land because they could not pay 

property taxes under the American system of monetary 

tax payments.

There was also during this time considerable unscrupulous land speculation, which took advantage 

of many Hispanic farmers who did not speak English 

and did not understand the American legal system 

(de Buys 1985 personal communiction). Large parcels 

of land were bought and sold by outside investors, 

often depriving families of land they had occupied for 

generations. These losses were a source of bitterness 

throughout the region—a bitterness that has surfaced 

periodically since U.S. annexation and came to a head 

in the land grant movement of the1960s (Gardner 

1970, Knowlton 1967, 1980, 1985, Rosenbaum 1981, 

Schlesinger 1971).

Large-Scale Commercial Livestock Ranching 

of the Mid to Late 1800s

Other forces of change were also at work during 

this time as the economy changed from subsistence 

to commercial orientation, and the population of the 

territory grew tremendously with in-migration from 

the United States. Many of these immigrants brought 

considerable amounts of capital for investment in largescale operations and a 19th century, entrepreneurial 

resource utilization ethic focused on maximum harvest 

for maximum profit (Scurlock 1995b: 2). To add to 

the climate of growth and development, Federal and 

Territorial legislation promoted intensive use of the 

environment (McCormick 1865, Scurlock 1995b: 2). 

Detailed discussions of this legislation are beyond the 

scope of this chapter but can be found in Clark (1987), 

Scurlock (1998), and Donahue (1999), for example.

These growth and development factors, combined 

with expanding markets opened by the development 

of military bases and the entrance of the railroad into 

Arizona and New Mexico in the 1870s and early 1880s, 

as well as the final subjugation of the nomadic Indian 

groups, led to rapid increases in large, commercial 

ranching operations (Bailey 1994, 1998). These operations often displaced the older, subsistence farms and 

ranches of the Hispano villagers. Commercial farming, 

timbering, and mining also flourished (Harper and 

others 1943: 48; Rothman 1989: 192-204; Wildeman 

and Brock 2000: 17).

Prior to the Civil War, substantial numbers of sheep 

and cattle were driven to the gold mining camps of 

California from the Southwest. Between 1852 and 1860 

more than 550,000 sheep were trailed from New Mexico 

(Bailey and Bailey 1986), with an estimated 15,000 

to 20,000 head of cattle driven from Texas through 

southern Arizona in the peak year of 1854 (Hadley and 

Sheridan 1995, Wildeman and Brock 2000: 13). The 

Civil War disrupted ranching operations and put an 

end to these drives, as ranchers left to fight in the war 

and Native American raiding increased with soldiers 

occupied elsewhere. By the end of the war, there were 

extremely large herds of free-ranging cattle in Texas 

from ranches that had been abandoned during the 

war. There were markets for these animals both in 

the North and in the West with the reestablishment 

of military posts and American Indian reservations 

(Wildeman and Brock 2000: 15-16).

Thus began the period of cattle drives out of Texas 

lasting from 1866 to 1880. During this time an estimated 10 million to 12 million animals were driven 

out of Texas, often through New Mexico and Arizona. 

This period also saw the decimation of the Plains bison 

herds between 1868 and 1881 (Schickedanz 1980). The 

herds were slaughtered for both economic and political 

reasons. Their decimation caused intense retaliation 

by the Plains Indian Tribes, who were ultimately subjugated by the U.S. Army. What had been bison range 

became cattle range, leading to the westward spread of 

the livestock industry and its full development in the 

Southwest (Wildeman and Brock 2000: 16-17).

Although there are conflicting livestock numbers, the 

cattle industry in Arizona grew from an estimated 5,000 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 95

animals in 1870 to around 200,000 in 1872. By 1891 

various authors report cattle numbers ranging from 

over 980,000 to 1.5 million (Antle 1992, Ferguson and 

Ferguson 1983, Peplow 1958, discussed in Wildeman 

and Brock 2000: 18), in addition to an estimated 700,000 

sheep brought into the State from New Mexico and 

California. In 1870, some 17,000 sheep were reported 

for the State. This growth was fueled by an extremely 

large flow of capital into the area, primarily from nonlocal investors (Bailey and Bailey 1986, Wildeman and 

Brock 2000: 18).

The livestock industry in New Mexico grew in 

the same way, with the entry of large numbers of 

animals into the area fueled by financial investment 

from outside the region. Extremely large numbers 

of both sheep and cattle were put on the rangelands 

with more than 1.5 million sheep in 1870 rising to 4 

million or 5 million in 1883. In addition to the sheep, 

there were an estimated 250,000 cattle (Bailey and 

Bailey 1986, Schickedanz 1980). Rangeland use in 

both Arizona and New Mexico reached its peak in 

the late 1880s to early 1890s with almost 9 million 

animal units (AU) in New Mexico and 4.5 million 

in Arizona (Bailey and Bailey 1986, Wildeman and 

Brock 2000: 19).

Land Degradation and Reform 

Legislation________________________

The land could not sustain the large number of animals being grazed in the attempt to achieve maximum 

economic gain. In the Southwest, the cattle population 

crashed after severe drought in the summers of 1891 

and 1892. Severe ecosystem degradation resulted 

from several interacting factors including overstocking of rangelands, decrease of herbaceous plant cover, 

drought, suppression of natural fires, and removal of 

beaver along streams by trapping earlier in the century 

(Tellman and others 1997, Wildeman and Brock 2000: 

19-20). Heavy stocking of the rangelands stressed native grasses, leading to decreased cover and root depth. 

In addition, some of the Southwestern grasses may have 

had less resistance to grazing than plains grasses that 

evolved with grazing by bison (Hyder 1972, Loftin and 

others 2000). Fire suppression, which became a major 

factor early in the 20th century, in combination with 

reduced plant cover from grazing, allowed woody shrubs 

and plants with low grazing preference to increase 

across the landscape. The combination of drought and 

overgrazing led to soil cover loss from wind and water 

erosion (Wildeman and Brock 2000: 20).

Corral on the road to Gallina, NM, 1993. (Photo by Alice M. McSweeney)
96 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

An example from the Pajarito Plateau west of 

Santa Fe discussed in Raish (2000b: 287) can help 

to illustrate the environmental and social consequences of commercial development that resulted in 

overstocking and subsequent environmental damage. 

Throughout the 1800s, local Hispanic and Pueblo 

residents of the area used the plateau as common 

property. They brought small herds to the plateau 

for summer grazing, harvested timber for personal 

use and small-scale ventures, and planted occasional 

summer crops. They were not attempting to maximize 

profit by developing large herds because for much of 

the period there was limited access to commercial 

markets. The small size and noncommercial nature 

of these operations ensured that sufficient grass and 

fuel wood were available for those who needed them 

(Rothman 1989: 192-194).

During the livestock boom period in 1885, a commercial Texas cattle operation leased a large amount 

of land on the plateau and brought in 3,000 head of 

cattle. The area could probably support around 300, 

according to modern calculations. This large operation drove off the small-scale Hispanic and Puebloan 

operations that had used the area for many years, as 

the resources were simply not sufficient to provide for 

both commercial and subsistence economies simultaneously. The large Texas enterprise failed after the severe 

winter of 1886 to 1887, and the small, local operators 

returned with their herds. However, the considerable 

overstocking had caused long-lasting damage to the 

resources of the plateau (Rothman 1989: 200-204). 

Rothman (1989: 202) notes:

In an arid marginal region, the impact of comprehensive overgrazing persists for generations. 

Ecological climax communities in arid areas like the 

Pajarito Plateau take hundreds of years to mature. 

Because its soils were fragile, thin, and highly erodible, the removal of first-growth cover by overstocking 

precluded the slow process of natural recovery.

This damage was increased by commercial timber 

operations that began after the failed cattle business. 

Small-scale timber and personal-use fuel wood harvest 

was replaced by clear-cutting on large areas of the 

plateau. The combined effects of the commercial cattle 

and timber enterprises damaged the native grasses and 

removed much of the old-growth timber. Destruction 

of the grass cover and the spread of less palatable 

plants affected the subsistence of the local people, 

forcing them to pasture their animals farther away 

and compete among themselves for increasingly poor 

range (Rothman 1989: 203-204). This small vignette 

was repeated throughout the Southwest with effects 

lasting to the present day.

To deal with such problems of land degradation and 

resource overexploitation throughout the West, the 

direction of Federal legislation changed from promotion of intensive resource use to promotion of resource 

conservation. Scurlock (1998: 331-384) provides a detailed review of the development and implementation 

of conservation legislation and the resulting Federal 

land management agencies and programs. In brief, first 

attempts to deal with resource conditions occurred with 

creation of Forest Reserves from public domain lands 

in 1891 during the presidency of Benjamin Harrison. 

He set aside more than 13 million acres, and President 

Grover Cleveland reserved another 21 million acres 

during his time in office. President Theodore Roosevelt 

added over 16 million acres (Ferguson and Ferguson 

1983, Steen 1976: 22-46). In 1905 President Roosevelt 

signed the bill transferring the Forest Reserves to the 

Department of Agriculture, and in July of that year 

the Bureau of Forestry was renamed the United States 

Forest Service (Steen 1976: 74-75). In 1906 fees were 

charged for grazing on Forest Service land. By 1908, 

there were 21 Forests in Forest Service District 3, which 

would become the 11 National Forests of contemporary 

Region 3 (Tucker 1989: 1-2, 68). Tucker (1989, 1992) 

provides an excellent discussion, personal interviews 

with early day Forest Service employees, and general 

sources of information on the development and early 

days of the Forests of Region 3.

The National Grasslands administered by Region 3 

in eastern New Mexico, the Panhandles of Texas and 

Oklahoma, and western Oklahoma had a somewhat 

different history. Between 1900 and 1930 increasing 

numbers of farmers settled in the area as higher than 

normal rainfall levels and high wheat prices encouraged local farmers to move from stock raising and 

planting a variety of crops to a single cropping system. 

As grasslands were plowed under, the ground surface 

became destabilized and susceptible to erosion. In the 

1930s, drought and dust storms hit the area prompting 

the Federal government to attempt to restore Dust 

Bowl areas to grassland (Bonnifield 1979, Cojeen 

2000, Fredine and others 1999, Lewis 1989, Simpson 

and others 1998).

Much land reverted to the government during this 

period owing to foreclosures and sale incentives. In 

1934, Congress passed the Taylor Grazing Act, which 

established the U.S. Grazing Service within the 

Department of the Interior to rehabilitate overgrazed 

and eroded lands and manage grazing on public lands. 

Under this act the bulk of unappropriated grassland 

was closed to further settlement. In 1946 the Grazing 

Service merged with the General Land Office to form 

the Bureau of Land Management. From 1938 until the 

early 1950s the Soil Conservation Service assumed 

responsibility for restoring grasslands in the areas 

under discussion. In 1953, the Forest Service took over 

responsibility for managing what became the Kiowa, 

Rita Blanca, and Black Kettle National Grasslands 

(Bonnifield 1979, Cojeen 2000, Fredine and others 

1999, Lewis 1989, Simpson and others 1998).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 97

Forest Service Response to Rangeland 

Degradation

The 20th century Forest Service response to issues 

of grassland condition and overutilization has been 

discussed in various publications and reports going 

back to and preceding Aldo Leopold in 1924 and The 

Story of the Range in 1926 (see for example Alexander 

1997, Barnes 1926, Leopold 1924, Roberts 1963, 

Robinson 1975, Rowley 1985, Scurlock 1998, Steen 

1976). Concerning the condition of lands that became 

part of the National Forest system, Robinnson (1975: 

199) states:

So long as the forests were part of the public 

domain, their access and use [were] virtually unrestricted. As a consequence, many of the lands were 

heavily overgrazed...When management responsibility was transferred to the Forest Service, regulation 

of grazing, through a system of permits and grazing 

fees, was established as one of the first and foremost 

management tasks.

In the West, where much rangeland is Federally 

owned and ranchers are heavily dependent upon public 

land for grazing, most grazing permits are issued for 

10 years. This term provides greater security for the 

rancher than do annual permits, which are more common in the East and South where much more land is 

privately owned (Robinson 1975: 200). Since agency establishment, regulation of livestock grazing on National 

Forest land has been a compromise, although an often 

contentious one, between the interests of the livestock 

industry, the range management goals of the Forest 

Service, and, more recently, the preservation/protection 

concerns of the environmental movement.

Research on rangeland condition has been ongoing in the Forest Service since the early years of the 

agency and its predecessors with the development of 

the section of Special Investigations in the first year 

of Gifford Pinchot’s tenure at the Bureau of Forestry. 

By 1902, the section had become a division, and by 

1908 the Fort Valley Experiment Station, with Gus 

Pearson as Director, was established in northern 

Arizona (Steen 1998: 5-8). In 1915, the Branch of 

Research was officially established, and by the end 

of the 1920s, 12 regional Research Stations were 

operating (Steen 1976: 131-137). In 1915, the Santa 

Rita and Jornada Experimental Ranges in Arizona and 

New Mexico were transferred to the Forest Service 

from the Bureau of Plant Industry, with the research 

goals of restoring, improving, and maintaining basic 

range resources and obtaining the greatest returns on 

livestock (Steen 1998: 17). Medina (1996) has provided 

a comprehensive annotated bibliography of the work 

of the Santa Rita Experimental Range.

Early and continuing issues of major concern between the Forest Service and the livestock industry 

included grazing fees and regulations surrounding 

issuance of grazing permits, including their duration, 

use limits, and transferability (Robinson 1975: 209-

214). These issues are discussed in detail by Robinson 

(1975), Rowley (1985), and Steen (1976), among others, 

and will not be reviewed here. Of great concern to the 

agency (throughout the early years of the 20th century 

and continuing to the present) have been improving 

the condition of and/or restoring Western grasslands 

and rangelands. As Robinson (1975: 202) notes: “The 

task of surveying the range and adjudicating individual permits to conform with range capacity began 

in 1910. Sixty years later it is still unfinished, and 

range improvement continues to be a major concern 

of management.”

Suffice it to say that the condition and sustainability of the region’s rangelands remains an issue of 

concern and considerable philosophical debate today 

among not only land managers but also among many 

concerned groups.

Over the years, the Forest Service and other land 

management agencies have undertaken range improvement practices using a variety of techniques including 

reducing stocking levels, shortening grazing seasons 

(postponing entry into allotments), implementing specialized grazing systems (discussed in greater detail in 

following sections), and emphasizing range improvements by permittees such as development of waters and 

fencing (Alexander 1997: 179-194; Robinson 1975: 202). 

Additional programs, some of which are considered 

controversial by today’s standards, were also implemented in the region. According to Bahre (1995: 247), 

the contemporary grassland landscape of southeastern 

Arizona resulted from many Forest Service and BLM 

practices designed to protect watersheds and improve 

the livestock industry. Some of these included contour 

plowing, fencing, using specialized grazing systems, 

prescribed burning, suppressing fire, controlling woody 

plants with herbicides, introducing nonnative forage 

plants, constructing check and spreader dams, and 

controlling weeds, as well as chaining and bulldozing 

pinyon (Pinus edulis), juniper (Juniperus spp.), mesquite (Prosopsis spp.), and oak (Quercus spp.).

Often, agency-mandated improvement programs 

met with varying degrees of resistance from user 

groups and the livestock industry. An example from 

the forests of northern New Mexico is instructive in this 

regard, highlighting the tension between traditional 

use practices and Federally mandated management 

programs. In this part of the region, loss of land grant 

lands limits the grazing areas open to small, local 

communities, many of which are now surrounded 

by National Forest (Van Ness 1987: 201). In 1938 a 

Forest Service report estimated that demand for grazing on portions of the Carson and Santa Fe National 

Forests exceeded potential by 111 percent (Hassell 

1968: 12). In the late 1960s, estimates showed grazing 
98 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

obligations on the two Forests for 21,637 cattle and 

32,203 sheep, compared to an estimated capacity of 

14,370 cattle and 25,237 sheep. Over the years the 

agency attempted to deal with these discrepancies 

using a variety of techniques.

Beginning in the 1920s and continuing into the 

1960s, livestock ranching on the two Forests changed 

as the economy changed and the Forest Service 

implemented range improvement programs (de Buys 

1985: 247-249). There was a steady decline in both the 

number of permittees and the number of animals, from 

2,200 permits in 1940 to fewer than 1,000 in 1970. 

For example, the permittees of the village of Canjilon 

who grazed animals on the Carson lost permits for 

1,000 cattle over a period of a few years (de Buys 1985: 

247-259). Free-use permits, issued for animals used 

in household operation such as milk cows and draft 

horses, were completely phased out by 1980.

Also during this period, there was a major change in 

the kinds of animals being grazed, with large declines 

in sheep and goats under permit. By 1980, there were 

no goats on either Forest and no sheep on the Santa Fe 

(de Buys 1985: 247-248; Van Ness 1987: 202). These 

changes occurred both as a result of Forest Service 

direction and as a result of changes brought about by 

the switch from a subsistence-based to a cash-based 

economy. Land losses and cutbacks in herd size undoubtedly pushed many people into the cash-based 

economy of wage work. Although there were important 

rangeland health issues that required treatment, these 

changes seriously impacted the livelihoods of many 

villagers who had ranched and farmed in northern 

New Mexico for generations prior to U.S. conquest. 

These losses have contributed to resentment of and 

protest against the Forest Service that continues to 

the present day.

Today, Federal agency grassland/rangeland management is critiqued by both user groups and the organized 

environmental community (for example, see Donahue 

1999 for an argument against livestock grazing on arid 

and semiarid Western public lands). In recent years, 

Region 3 has been increasingly involved in appeals and 

lawsuits related to efforts to align livestock grazing 

activities with Federal environmental statutes such 

as the National Environmental Policy Act (NEPA), 

National Forest Management Act (NFMA), and the 

Endangered Species Act (ESA). Under NEPA, environmental assessments are required for issuing and 

reissuing grazing permits to determine if livestock 

grazing is an appropriate and suitable use of the 

land (Recision Act of 1995, PL 104, Section 504). In 

an example of this type of analysis, the Black Kettle 

National Grassland has produced an environmental 

assessment of livestock grazing and associated 

vegetation management, which includes all 114 

rangeland units of the grassland (Black Kettle and 

McClellan Creek National Grasslands Geographic Area 

Assessments 2000). The role and impact of environmental legislation and efforts at mediation and conflict 

resolution on the management and sustainability of 

grassland areas will be discussed in greater detail in 

following sections of this chapter.

The following portions of this chapter also briefly 

discuss the current condition of rangeland vegetation resulting from human uses, which are primarily 

ranching/grazing by domesticated livestock followed 

by recreation and mineral extraction. Certain effects 

of recreation and mineral extraction on wildlife are 

also presented. Rangeland/grassland condition and the 

effects of grazing on wildlife in general and on threatened, endangered, and sensitive species in particular 

are examined in detail in volume 2 of this report.

The effects of livestock grazing and recreational use 

on archeological sites of varying types are taken into 

consideration in environmental assessments and must 

be considered in issuing and reissuing grazing permits. 

Several studies have examined the damaging effects of 

trampling and rubbing (against the walls of standing 

structures) by both domesticated livestock and wild 

ungulates such as elk; other ongoing problems are human visitors’ trampling, leaning, or sitting on walls, or 

removal of artifacts (Gifford-Gonzalez and others 1985, 

Knudson 1979, Osborn and others1987). Rubbing or 

leaning against fragile walls can lead to their collapse. 

Fencing off standing structures in grazing allotments 

has been used as a means of excluding domesticated 

animals from the structures. Trampling can damage 

sites and interfere with their correct interpretation by 

reducing artifact size through breakage and by displacing artifacts from their original positions. Adding 

additional fractures to chipped stone tools can obscure 

interpretations of functionality and production technique. This type of damage primarily affects surface 

artifact scatters of ceramics and chipped stone but can 

also affect buried deposits near the surface (Osborn 

and others1987).

Contemporary Condition of Federal 

and Non-Federal Rangelands________

Despite improvements in the condition of many 

grasslands since the early years of the 20th century, 

significant issues remain to be addressed. Land use 

and management activities such as domesticated 

livestock grazing, fire suppression, agriculture, and 

urban/suburban development impact the grasslands 

of the region as do the expansion of woodlands and 

the introduction of nonnative plant species (Bahre 

1995: 243-255; Mitchell 2000). Mitchell (2000) reviews 

these issues for the entire United States, as well as 

for the region. The following discussion is primarily 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 99

drawn from his work, with other sources cited where 

appropriate.

Information from the National Resource Inventory 

(NRI) conducted on non-Federal lands by the Natural 

Resources Conservation Service (NRCS) in 1982 and 

1992, including an unpublished supplemental study 

conducted in 1992, is used to characterize rangeland 

condition on non-Federal lands during the late 20th

century (USDA Natural Resources Conservation 

Service 1995, USDA Soil Conservation Service 1987, 

1990). The assessments use a species composition 

model (estimated on a biomass basis and compared 

to a typical “climax” plant community for the site) to 

evaluate range condition (Dyksterhuis 1949, Mitchell 

2000: 28).

Mitchell (2000: 27-28) discusses problems with 

this classificatory scheme and reviews the scientific 

advances in understanding rangeland condition of the 

past 25 years as follows. Earlier concepts of range 

condition were based on the Clementsian equilibrium 

theory of retrogression produced by overgrazing, and by 

secondary succession to a stable climax after removing 

the disturbance caused by grazing (Dyksterhuis 1949 

discussed in Mitchell 2000). Rangeland condition was 

divided into four classes: excellent (rangeland in a 

near climax condition with 76 to 100 percent remaining climax), good (rangeland in a late successional 

stage with 51 to 75 percent remaining climax), fair 

(rangeland in a midsuccessional stage with 26 to 50 

percent remaining climax), and poor (rangeland in 

an early seral stage with 0 to 25 percent remaining 

climax) (discussed in Holechek and others 1998). Many 

range managers now use the terms climax (or Potential 

Natural Community, PNC), late seral, mid-seral, and 

early seral to replace the terms excellent, good, fair, 

and poor, distinguishing range ecological condition 

from how well existing vegetation may be suited for 

specific uses such as grazing by domesticated animals 

(Holechek and others 1998). Tables 5-1 and 5-2 show 

use of these classifications.

U.S. land management agencies have maintained 

these classification schemes in inventory and assessment programs although scientific evidence has 

challenged the Clementsian theory of succession 

(discussed in Mitchell 2000). Work by Westoby and 

others (1989) set forth an alternate hypothesis for 

nonequilibrium vegetation dynamics on disturbed 

rangelands. In addition, the view that ecosystem 

behavior under stress should be described only in 

terms of vegetation responses has been questioned by 

ecologists from various disciplines (Rapport and others 1985). Joyce (1993) has summarized the changes 

in the conception of range condition from a prior 

simplistic model to the current changing, complex 

situation (discussed in Mitchell 2000).

In 1994, the Committee on Rangeland Classification 

proposed a new paradigm for assessing rangeland 

conditions based on several years of review and study. 

This assessment paradigm is based on nonequilibrium, 

state-and-transition models of succession, focusing 

on ecosystem function rather than state or plant 

community composition. Three major criteria of 

ecosystem function are soil stability and watershed 

function, distribution of nutrient cycling and energy flow, and recovery mechanisms (Committee on 

Rangeland Classification 1994). The Society for Range 

Management also assembled a Task Group on Unity 

in Concepts and Terminology (1995) that called for 

making sustainability—defined in terms of maintaining soil productivity—the fundamental goal of range 

management. Unfortunately, these advances have not 

yet been incorporated into national data sets. Thus, 

the older classificatory system is used in Mitchell’s 

review and discussion (2000).

Table 5-1 shows the condition of non-Federal rangelands in Arizona, New Mexico, Oklahoma, and Texas 

in both 1982 and 1992. Arizona and New Mexico have 

slight improvements (defined as increases in excellent 

and good condition classes) over the 10-year period, 

while Oklahoma shows greater improvement. Texas, 

on the other hand, has virtually no improvement, 

and according to Mitchell (2000: 30), the non-Federal 

rangelands of Texas are more degraded than those of 

any other Great Plains State. Since less than 2 percent 

of Texas ranges are Federal, these condition estimates 

essentially cover the entire State (USDI Bureau of 

Land Management 1997).

Federal ranges of the area include lands managed by 

the BLM and the Forest Service. Discussing their condition is complicated because the two agencies report 

condition in different terms. The BLM currently uses 

the Ecological Site Inventory (ESI) procedure (USDA 

Soil Conservation Service 1976, USDI Bureau of Land 

Management 1984), and the Forest Service assesses 

the condition of rangeland vegetation in terms of 

established Forest Plan Management Objectives 

(FPMO). Uplands and riparian range areas are described separately by both agencies. Mitchell (2000: 

33) is critical of the Forest Service evaluation system, 

observing that many National Forests are operating 

under plans approved in the 1980s, “...and the elements 

relating to rangeland health are not well correlated 

to trends in indicators judged to be relevant by contemporary standards” (Committee on Rangeland 

Classification 1994).

Table 5-2 gives information on the condition of 

upland BLM rangelands in Arizona and New Mexico 

in 1986 and 1996. There are but a few hectares in 

Oklahoma and none in Texas. Both Arizona and 

New Mexico have increases in areas with late seral 
100 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Table 5-1. Condition of non-Federal rangelands in Arizona, New Mexico, Oklahoma, and 

Texas, 1982 and 1992, from the National Resource Inventory (adapted from Mitchell 

2000: 29, 31, table 3.2 and 3.6). 

Year

State Condition class 1982 1992

Arizona Excellent 5181 2% 650 2%

Good 4,923 16% 8,446 27%

Fair 16,574 54% 15,886 50%

Poor 8,832 28% 6,661 21%

New Mexico Excellent 659 2% 591 2%

Good 12,262 30% 14,314 36%

Fair 22,617 55% 21,227 53%

Poor 5,422 13% 3,645 9%

Oklahoma Excellent 907 6% 1,749 12%

Good 3,601 24% 4,492 32%

Fair 7,639 51% 5,835 42%

Poor 2,904 19% 1,951 14%

Texas Excellent 480 1% 174 <1%

Good 13,546 15% 16,324 18%

Fair 53,543 57% 49,899 55%

Poor 25,681 27% 24,922 27%

1

Acres x 103 and percent of total area for which condition ratings were applied.

Table 5-2. Condition of rangelands managed by the Bureau of Land Management (BLM) 

in Arizona and New Mexico, 1986 and 1996 (adapted from Mitchell 2000: 32, table 

3.7).

Year

State Condition class 1986 19961

Arizona PNC2 4673 4% 521 9%

Late seral 2,801 24% 2,217 40%

Mid seral 6,068 52% 2,217 40%

Early seral 2,334 20% 652 12%

Unclassified 0 913

Not inventoried 5,123

New Mexico PNC 125 1% 102 1%

Late seral 3,002 25% 3,555 36%

Mid seral 6,003 50% 4,673 48%

Early seral 2,877 24% 1,422 15%

Unclassified 500 305

Not inventoried 2,540

1

The acres and percentages by condition class for 1996 are based only on those acres inventoried 

using the Soil-Vegetation Inventory Method (used for 5 years prior to the ESI) or the Ecological Site 

Inventory (ESI) and classified by condition.

2

Potential Natural Community (Kuchler 1964)

3

Acres x 103 and percent of total.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 101

communities (developmental stages of an ecological succession) and decreases in the areal extent of 

early seral communities, distinct from other BLM 

rangelands in Western States, which show little 

change during the period (Mitchell 2000:31). Table 

5-3 presents upland range condition information for 

Forest Service grazing allotments in Arizona and New 

Mexico, showing virtually no change over the 3 years 

of assessment used in the study (1995, 1996, 1997). 

Slightly over 25 percent of the ranges under discussion are verified or estimated as neither meeting nor 

moving toward FPMO, as compared to 15 percent in 

the Rocky Mountain Assessment Region as a whole. 

This assessment region encompasses Forest Service 

grazing allotments in Montana, Idaho, Wyoming, 

North Dakota, South Dakota, Nevada, Utah, Colorado, 

Kansas, Nebraska, Arizona, and New Mexico (Mitchell 

2000: 35-37). Mitchell (2000: 36) notes:

Vegetation in the mostly dry Southwestern Region 

has been subjected to a history of fire suppression since 

the late 19th century and improper grazing, primarily between the 1880’s and World War I (Secretary 

of Agriculture 1936, Rasmussen 1941, Cooper 1960, 

Buffington and Herbel 1965, Mortensen 1978). These 

factors and others have caused upland vegetation 

and soil changes that are slow to improve, some of 

which probably will not ever recover to pre-existing 

conditions (Schlesinger et al. 1990, Wang and Hacker 

1997). As a result, the higher percentage of rangeland 

in R-3 (USFS Southwestern Region 3) that is not 

meeting or progressing towards FPMO’s, roughly 25 

percent, should not be surprising...

Conditions of riparian rangeland areas in Arizona 

and New Mexico managed by the BLM are reported in 

terms of Proper Functioning Condition (PFC) (Barrett 

and others 1995, USDI Bureau of Land Management 

1991), while those managed by the Forest Service are 

reported in terms of FPMOs, as are upland ranges 

(Robertson 1992). In Arizona, 39 percent of BLM-managed riparian areas are reported in Proper Functioning 

Condition, while 33 percent of those in New Mexico are 

considered to be in PFC. Arizona reports 58 percent 

Functioning at Risk and 3 percent Nonfunctional. New 

Mexico has 45 percent Functioning at Risk and 22 

percent Nonfunctional (Mitchell 2000: 41, table 3.20). 

These BLM data are for 1997.

The Forest Service assessment of riparian area 

condition within grazing allotments is reported for 

the Rocky Mountain Assessment Region as a whole 

for the years 1995, 1996, and 1997. Table 5-4 gives 

this information and shows that about 1 out of 6 acres 

of riparian land does not meet and is not moving towards FPMO; this is roughly the same as reported for 

the entire BLM Rocky Mountain Assessment Region 

(Mitchell 2000: 41). As discussed for the upland areas 

and reviewed in following paragraphs, various historical and contemporary factors and activities, as well 

as site-specific attributes of different locales, must be 

taken into consideration when assessing the condition 

of riparian areas throughout the region.

Table 5-3. Condition of rangelands in grazing allotments managed by the Forest Service in 

Arizona and New Mexico (combined) as related to Forest Plan Management Objectives 

(FPMO), 1995, 1996, 1997 (adapted from Mitchell 2000: 37, table 3.14).

Year

Land category 1995 1996 1997

Verified meeting FPMO 1,6761

 1,859 1,945

Estimated meeting FPMO 2,541 2,595 2,376

Total 4,217 4,454 4,321

(26%) (27%) (27%)

Verified moving toward FPMO 1,486 1,456 1,455

Estimated moving toward FPMO 6,227 6,208 6,105

Total 7,713 7,664 7,560

(47%) (46%) (47%)

Verified not meeting or moving toward FPMO 788 673 741

Estimated not meeting or moving toward FPMO 3,663 3,702 3,548

Total 4,451 4,375 4,289

(27%) (27%) (26%)

Undetermined status 1,938 1,839 1,849

Total 18,319 18,322 18,019

(Lands with range vegetation management 

objectives)

1

 Acres x 103 and percent of total lands with range vegetation management objectives.
102 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Specific information on the rangeland condition of 

the National Grasslands in New Mexico, Texas, and 

Oklahoma is presented in a somewhat different manner in the Geographic Area Assessments prepared by 

grasslands management during 1999 and 2000. The 

Black Kettle and McClellan Creek National Grasslands 

are located in Oklahoma and Texas. Two of their four 

geographic areas (High Plains and Redbed Plains) 

are used for regular livestock grazing under permit. 

The Lake Marvin Area is not currently grazed, and 

the Lake McClellan Area has occasional grazing as 

a management tool. All rangelands within the four 

geographic areas are described as in satisfactory 

range condition, which is defined as follows (Black 

Kettle and McClellan Creek National Grasslands 

Geographic Area Assessments 2000: appendix 38): 

“Satisfactory Rangelands—Those that have sufficient vegetative cover to protect the soil resource and 

provide a sustainable base for desired plant/animal 

communities.” Satisfactory rangelands are those 

where current soil loss (the amount of soil that would 

be lost under the current or existing vegetative ground 

cover) is less than tolerance soil loss (the amount of 

soil that would be lost under conditions of the potential natural vegetative ground cover and where soil 

development equals soil loss) (TES Handbook 1986). 

Further information on rangeland condition can be 

found in the Geographic Area Assessments and in 

the Environmental Assessment for Livestock Grazing 

and Associated Vegetation Management, Black Kettle 

National Grassland—Roger Mills County, Oklahoma 

(1999).

The Kiowa National Grassland is in New Mexico 

and contains the Mills Canyon, Mills Upland, and the 

western third of the Southern Prairie Geographic Areas. 

The Rita Blanca, in Oklahoma and Texas, comprises 

the remainder of the Southern Prairie Geographic 

Area. The three areas are described as in satisfactory 

rangeland condition with the exception of small areas 

occupied by playa lakes, which contain substantial 

portions of bare ground. The floodplain of the Mills 

Canyon Area is also described as in unsatisfactory 

condition resulting from severe flood events (Kiowa 

and Rita Blanca National Grasslands Geographic Area 

Assessments 1999). Again, more detailed information 

can be found in the Geographic Area Assessment.

As the previous discussion indicates, regional rangeland condition and the ways in which it is measured 

and assessed are topics of both concern and debate. The 

Southwestern rangelands/grasslands are used for and 

impacted by many activities, which are also topics of 

debate. Discussions of these activities, such as livestock 

grazing, mining/energy extraction, recreation/tourism, 

and other special uses, follow.

Certainly the most widespread and increasingly 

scrutinized use, especially on public lands, is livestock 

grazing. Mitchell (2000: 31) summarizes:

The two predominant opposing viewpoints (on 

grazing) are epitomized by Fleischner (1994) and Box 

(1990). Fleischner believes grazing has caused a loss 

of biodiversity, disruption of ecosystem function, and 

Table 5-4. Area of riparian range vegetation within grazing allotments on National Forest System 

lands in relation to Forest Plan Management Objectives (FPMO), Rocky Mountain Assessment Region (adapted from Mitchell 2000: 42, table 3.22).

Year

Land category 1995 1996 1997

Verified meeting FPMO 1,3911

 1,450 1,777

Estimated meeting FPMO 4,549 4,421 4,662

Total 5,940 5,871 6,439

(42%) (45%) (47%)

Verified moving toward FPMO 922 858 875

Estimated moving toward FPMO 4,519 3,973 4,092

Total 5,441 4,831 4,967

(39%) (37%) (37%)

Verified not meeting or moving toward FPMO 402 266 208

Estimated not meeting or moving toward FPMO 2,226 2,059 1,905

Total 2,628 2,325 2,113

(19%) (18%) (16%)

Undetermined status 2,331 2,290 2,241

Total 16,340 15,317 15,760

(Lands with range vegetation management

objectives)

1

 Acres x 102 and percent of total.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 103

irreversible changes in ecosystem structure, while Box 

concludes that the trend of U.S. public rangelands, 

on the average, has been upwards over a number of 

decades and the land is in the best ecological condition of this century.

For a detailed argument supporting the view 

expressed by Fleischner, see Donahue (1999). For 

a rebuttal to the Fleischner paper, see Brown and 

McDonald (1995).

Mitchell (2000: 42) states further:

Some reviewers conclude that grazing by domestic 

livestock is not compatible with restoring watersheds 

and water quality, at least in the short term (Belsky 

et al.1999). However, comparative [sic] research and 

case studies show that improved rangeland and 

livestock management practices are compatible 

with watershed and water quality improvement when 

designed to address the attributes of each individual 

site (Elmore and Kauffman 1994). Ultimately, social 

and political values, along with scientific knowledge, 

will drive future laws and regulations affecting the 

grazing use of riparian areas, just as they will for 

uplands (Lee 1993).

Livestock Ranching on Federal and 

Non-Federal Lands_________________

The section describes the form and occurrence of 

ranching in the Southwest, as well as its socioeconomic 

and cultural contributions to local communities and 

the region. The focus is on livestock ranching on public 

lands. Ranching and range condition on Tribal lands 

are beyond the scope of this review, but information on 

this topic can be found in Brugge and Gerow (2000). 

Ranches in Arizona and New Mexico are generally composed of a base of private land augmented by Federal 

and/or State grazing allotments. Ranchers do not hold 

fee-simple property rights to grazing allotments, but 

these public lands are traditionally seen as part of the 

ranch and are considered in determining ranch sale 

prices and appraisals for Federal estate tax purposes 

(Ruyle and others 2000: 380).

The number of cattle ranches in Arizona varies 

depending on the definition of a ranch. Ruyle and 

others (2000: 387) note that during 1995 there were 

about 2,500 farms and ranches reporting at least one 

beef cow. Most livestock operations in the State have 

fewer than 50 cows, which is consistent with the national average. The majority of Arizona ranches are 

cow-calf operations, consisting of a base cow herd and 

the animals needed to support them.

Most of the ranches in New Mexico are also small, 

cow-calf operations with from one to 99 head. This 

size class comprised 70 percent of the State’s 8,313 

ranches in 1996. In the north-central mountain area 

of the State, the small operations made up 82 percent 

of the listed 1,804 ranches. This area also has fewer 

large (greater than 500 head) ranches—3 percent of 

the total; Statewide, they account for 7 percent of the 

total (Torell and others 1998: table 1). In both New 

Mexico and Arizona, 30 percent of ranches range from 

medium to large with at least 100 head (Ruyle and others 2000: 387-388). In northern New Mexico, however, 

only 17.5 percent fall within this size range (Torell 

and others 1998: table 1). Thus, northern New Mexico 

has considerably fewer medium-to-large ranches than 

either Arizona or New Mexico as a whole.

New Mexico has over twice as many grazing permittees on National Forests as Arizona, but New Mexico 

has 35 percent fewer animals (Raish and others 1997: 

28-35). These figures result in large part from the small 

ranches and small herd sizes of northern New Mexico, 

as well as from the tendency to have multiple-permittee 

grazing allotments on National Forests in the area. 

These figures indicate the continuity of the long-standing tradition of small-sized operations and communal 

herding prevalent in the area since Spanish colonial 

times (Raish and others 1997: 28-35).

Because many ranching operations in the region 

rely to some degree on public land, regulations and 

management decisions affecting these lands can impact 

the operation and future of ranching throughout the 

area. Ruyle and others (2000: 380-383) found that for 

Arizona, but also applicable to New Mexico, many 

operators rely on a combination of privately leased 

land as well as State and Federal (USFS and BLM) 

grazing allotments. In Arizona, public and State grazing 

permits and leases account for roughly 85 percent of the 

State’s grazing land, excluding American Indian lands. 

In New Mexico, 53 percent of the land is nonprivate 

(Fowler 2000: 423).

The degree to which a ranch relies on leased and 

permitted land affects the complexity of ranch management, with regulations, fees, and enforcement often 

varying between agencies and within the same agency 

from location to location. Because ranchers rely on 

government grazing permits, they are affected by the 

permitting agency’s regulations. The managing agency 

defines grazing seasons and stocking rates, which are 

often limited by competing uses and values such as recreation or riparian restoration. Restrictions stemming 

from the Endangered Species Act (ESA), the National 

Environmental Policy Act (NEPA), and the National 

Historic Preservation Act (NHPA), for example, can 

affect the timing and construction of range improvements such as water developments and fencing (Ruyle 

and others 2000: 382-383).

The grassland areas of northeastern New Mexico, 

Texas, and Oklahoma comprise significantly greater 

amounts of private land than do other portions of the 

Southwest (Fowler 2000: 423-427; USDI Bureau of 

Land Management 1997). Fowler (2000: 423-427: table 

2) notes that northeastern New Mexico has the most 

productive rangeland, supports the greatest number 

of ranches (27 percent), and the greatest number of 
104 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

large ranches (35 percent) in the State. These figures 

come from the total of 8,313 ranches identified in 1996 

(Torell and others 1998: 4, table 1). Both cow-calf and 

the majority of the State’s yearling operations are located in the area. Yearling operations—not including 

cow-calf ranches that purchase weaned calves when 

forage is available—purchase calves to put on leased 

pasture. They are generally grazed on winter wheat 

until they reach sufficient size to be sent to the feedlot 

(Fowler 2000: 426).

The grasslands of the Great Plains States (North 

Dakota, South Dakota, Nebraska, Kansas, Oklahoma, 

and Texas) contained almost 40 percent of the nation’s 

beef cow herd according to 1996 statistics (USDA 

National Agricultural Statistics Service 1997). In 

these six States income from livestock generally 

exceeds income from other agricultural commodities, 

and cattle grazing is the predominant land use. There 

are twice the numbers of stocker cattle in these States, 

especially in Kansas, Oklahoma, and Texas, than in the 

Western States (including Arizona and New Mexico). 

Engle and Bidwell (2000: 105) found:

For a variety of reasons, including closer proximity 

to feedlots and lower winter feed costs for cow-calf 

enterprises, overall costs of cattle production in the 

Great Plains are lower than in the regions to the 

west. The sum of these factors provides a competitive 

edge for Great Plains grassland cattle over cattle 

production elsewhere in the western United States 

(Cheeke and Davis 1997).

Livestock Numbers on Forest Service Land

Nationally, Region 3 ranks second to Region 4 (southern Idaho, Nevada, Utah, and western Wyoming) in 

permitted grazing use, according to the Forest Service 

Grazing Statistical Summary available at the time of 

this report, which reports information for the 1998 

grazing season (USDA Forest Service 1999). Table 5-5 

gives figures on the numbers of permittees, animals, 

animal unit months (AUMs), and head months (HMs) 

both permitted and authorized to graze on National 

Forest system lands (including the grasslands) in 

Region 3 in 1998. An AUM is the amount of forage 

required to support a mature, 1,000-lb cow and calf 

or its equivalent for 1 month, while an HM is the 

time in months that livestock spend on National 

Forest system land (used for billing purposes) (USDA 

Forest Service 1999). “Permitted to graze” indicates 

“Livestock permitted by a grazing permit, grazing 

agreement, livestock use permit, or other permitting 

document.” “Authorized to graze” indicates “The 

number of livestock that are authorized and billed for 

grazing on National Forest System land” (USDA Forest 

Service 1999: 96). Authorized livestock are those that 

are actually grazing on the Forest and being paid for. 

Tables 5-6, 5-7, and 5-8 present the same information 

broken down for the National Forests in Arizona and 

New Mexico (excluding the National Grasslands), and 

for the National Grasslands of the region located in 

New Mexico, Texas, and Oklahoma.

The figures from tables 5-6, 5-7, and 5-8show that the 

average permitted number of animals per permittee is 

around 318 for the National Forests in Arizona, 102 for 

the Forests in New Mexico, and 65 for the Grasslands. 

There are 2.5 times as many permittees in New Mexico 

(1,082) as in Arizona (425) with the Arizona permittees 

averaging roughly three times as many permitted 

animals. The larger numbers of permittees and smaller 

numbers of animals in New Mexico in many cases reflect 

the persistence of traditionally small operations with 

multiple permittees per allotment, especially in the 

Table 5-6. Livestock grazing on National Forests in Arizona 

in 1998.

Number of

 animals AUMs HMs

Permitted 135,188 1,327,115 1,113,027

Authorized 107,081 916,392 779,235

 Total of 425 permittees.

Table 5-7. Livestock grazing on National Forests in New 

Mexico in 1998.

Number of

 animals AUMs HMs

Permitted 110,393 1,044,416 928,407

Authorized 98,466 904,940 798,381

 Total of 1082 permittees.

Table 5-8. Livestock grazing on National Grasslands in 

Region 3 in 1998.

Number of

 animals AUMs HMs

Permitted 12,698 80,722 74,181

Authorized 12,423 79,213 72,720

 Total of 196 permittees.

Table 5-5. Livestock grazing on National Forest System 

lands in Region 3 in 1998.

 Number of animals AUMs HMs

Permitted 258,279 2,452,253 2,115,615

Authorized 217,970 1,900,545 1,650,336

 Total of 1703 permittees.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 105

north-central portion of the State. Figures from the 

Grasslands tend to be somewhat misleading because 

livestock operations in the Plains areas generally have 

access to considerably greater amounts of private land 

on which to graze other portions of their herds. There 

are also a fair number of part-time farmers and ranchers with small herds using the Grasslands (Dickerson 

personal communication 2001) as there are throughout 

the Southwestern Region.

Region 3 rangeland use statistics show a fluctuating but primarily downward trend from 1982 to 1995 

in numbers of permittees, animals permitted and 

authorized to graze, and AUMs. During this period, 

permittee numbers in both Arizona and New Mexico 

dropped by about 25 percent, and animals authorized to 

graze dropped by approximately 25 percent in Arizona 

and 20 percent in New Mexico (Raish and others 1997: 

36). Declines are apparently related to climatic and 

market fluctuations, permit consolidation, and growing urbanization in the area (Raish and others 1997: 

36). From 1996 through 1998, regional permittee 

numbers rose from 1663 to 1703, an increase of 2.4 

percent. Permitted animals declined from 270,068 to 

258,279, or a 4.4 percent decrease. AUMs increased 

by 1.9 percent from 2,406,708 in 1996 to 2,452,253 in 

1998 (USDA Forest Service 1997, 1998, 1999). These 

are relatively slight fluctuations when compared to 

the declines of the period from 1982 to 1995.

Grazing Management Systems on Forest 

Service Land

A considerable body of information exists concerning grazing systems in use on the arid and semiarid 

rangelands of the Southwestern and Western United 

States. This information is ably defined, reviewed, and 

discussed elsewhere, for example in Holechek and others (1998) and Kruse and Jemison (2000), to name only 

two. From a relatively rough assessment and review 

of grazing systems used in the Southwestern Region 

(USFS), it is apparent that systems incorporating 

deferment, rest, and rotation are the most common. 

(All regional rangeland information used in this discussion and in table 9 was provided by Gene Onken, 

Rangeland Management Staff, USFS Southwestern 

Region). According to Holechek and others (1998: 

228-229):

Deferment involves delay of grazing in a pasture 

until the seed maturity of the key forage species. 

This permits the better forage plants to gain vigor 

and reproduce. Rest is distinguished from deferment in that the range receives nonuse for a full 

year rather than just during the growth period. 

This gives plants a longer period to recover from 

past grazing influences and provides wildlife with 

a pasture free from livestock use during the critical 

dormant period...Rotation involves the movement of 

livestock from one pasture to another on a scheduled 

basis. It is the critical feature of all specialized 

grazing systems.

Out of approximately 1,009 allotments that I was 

able to determine for the Forests of the Region (excluding the National Grasslands), 330 (33 percent) report 

using deferred-rotation, 315 (31 percent) rest-rotation, 

146 (14 percent) deferred, and 142 (14 percent) continuous systems (table 5.9). Another 55 (5 percent) use 

rotation or alternate forms (table 5-9). Thus, around 

83 percent are using some form of deferment, rest, or 

rotation.

Deferred-rotation “discontinues grazing on various 

parts of a range, allowing each part to rest successively 

during the growing season...two, but usually three or 

more, separate units, or pastures, are required” (Kruse 

and Jemison 2000: 41). In rest-rotation, grazing is 

deferred for a complete year on various portions of 

the range during succeeding years. Two or more units 

are required but most rest-rotation systems involve 

three or four pastures (Holechek and others 1998: 

244; Kruse and Jemison 2000: 42). Deferred systems 

discontinue grazing on an area for a specified period 

(during the growing season, for example) to promote 

plant reproduction, establishment of new plants, or 

restoration of vigor to older plants (Kruse and Jemison 

2000: 40). Continuous grazing, on the other hand, refers 

to grazing the same area throughout a year or that 

portion of the year when grazing is feasible. It may be 

yearlong or shorter depending upon environmental or 

other restrictions (Kruse and Jemison 2000: 37). Table 

9 presents information on the primary systems in use 

on the National Forests in Arizona and New Mexico 

broken down by State. It should be noted that many 

of the continuous systems reported for New Mexico 

are seasonal in that grazing is not feasible in high 

altitude units during portions of the year.

As with the Forests of the Region, a rough count 

from the available information was used to examine 

grazing systems in use on the National Grasslands. 

The Geographic Areas Assessments for the Kiowa 

and Rita Blanca (1999) show deferred-rotation as 

Table 5-9. Grazing systems in use on the range management units of the National Forests in Arizona and New 

Mexico.

Number and percentage

Grazing system Arizona New Mexico

Continuous 33 (6%) 109 (23%)

Deferred 95 (18%) 51 (11%)

Deferred-Rotation 200 (37%) 130 (28%)

Rest Rotation 183 (34%) 132 (28%)

Rotation or Alternate 13 (2%) 42 (9%)

Other/Unlisted 18 (3%) 3 (.6%)

Total 542 (100%) 467 (100%)
106 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

the system in use for the range management units 

with available information. These comprise approximately 80 units in the Mills Canyon and Mills Upland 

Geographic Areas. The Black Kettle and McClellan 

Creek National Grasslands systems are described as 

variable season grazing and variable season rotational 

grazing. Grazing units occur in the High Plains and 

Redbed Plains Geographic Areas. The vast majority of 

the 114 rangeland units are described as being grazed 

in a rest-rotation type of system with variable livestock 

numbers and/or seasons. According to the description 

for the High Plains Geographic Area (Black Kettle 

and McClellan Creek National Grasslands Geographic 

Area Assessments 2000: 27-30), a maximum number of 

AUMs is set forth according to season and vegetative 

conditions with a minimum 30-day recovery period 

during the growing season.

Sociocultural and Economic Role of 

Livestock Ranching_________________

Ranching in New Mexico and Arizona, as well as 

on the Federally managed grasslands of eastern New 

Mexico, Oklahoma, and Texas, provides both economic 

and sociocultural contributions to families, communities, and the region. As discussed by Ruyle and others 

(2000: 379):

Despite major shifts toward urbanization in the 

state, livestock grazing remains the most widespread 

use of Arizona rangelands. Through the sale of calves, 

yearlings, stocker cattle, and culled cows and bulls, 

cattle ranching accounts for nearly 25% of the agricultural economy in the state. However, based on 

profit alone, the economic viability of ranching in 

Arizona is questionable. Because of economic factors 

related to income production, decisions to remain 

in the ranching business are not entirely financial. 

Lifeway considerations have long played an important 

role in the process.

Economic status of ranches is determined by the 

larger factors of productivity, market prices, and 

production costs. More specific factors include range 

conditions, weather, the number of AUMs assigned to 

a grazing permit, management decisions, and ranch 

size (Fowler and others 1994, Ruyle and others 2000). 

Ranch size can be extremely important. Fixed costs, 

which remain the same whether animal numbers rise 

or fall, drive the move toward larger ranches. Seperich 

and others (1995) show that large-sized ranches do 

cover the fixed cost of operator salary better than do 

medium or small-sized ranches. A ranch should exceed 

150 AUs to meet minimal family living expenses, a 

greater number than the one to 99 AUs of the smallsized ranches. Ruyle and others (2000: 393) estimate 

that small-sized operations would need to bring in about 

50 percent of their income from nonlivestock sources 

to meet living expenses (for Yavapai County, Arizona). 

Fowler and others (1994: 1, appendix 4) discuss offranch income in New Mexico as follows:

Seventy-five percent of the small ranches (1-99 

A.U.) had people employed off the ranch contributing 

44 percent of the family’s income. Fifty-five percent 

of the medium size (100-350 A.U.) ranches had family members working off the ranch also earning 20 

percent of the income. Forty-three percent and 36 

percent of the large and extra large ranches had 

people working off the ranch, respectively, earning 13 

percent and 6 percent of the family’s income.

In addition to economic considerations, culture, tradition, and quality of life seem to become increasingly 

important in the decision to continue ranching among 

owners of the smaller ranches. Of course, these qualities are also valued by the owners of larger ranches. In 

addition, other factors influence the decisionmaking 

process among the small ranch operators. Eastman 

and others (2000: 543) found that small ranches in 

northern New Mexico, but equally applicable to other 

areas of the Southwest, are often viewed as an investment and a form of savings:

While the ranch may produce little or even a 

negative operating income, the assets have a high 

value, which is expected to increase. Most northern 

ranchers own their homes, land, and cattle, and 

these constitute a significant investment and form 

of savings, which often has very high value. Managed 

properly, operating losses often provide income tax 

write-offs against other income. Thus, small operators 

stand to benefit from a reduced tax burden while 

their assets increase in value.

The ranchers often view their animals as bankson-the-hoof, which can be used...for emergencies, for 

periods of unemployment, or for special needs such as 

college tuition for the children...(Eastman and Gray 

1987; de Buys personal communication 1995).

In addition to providing family income at various 

levels, ranching operations pay a variety of taxes that 

contribute to community, State, and Federal resources. 

At the State level, average ranches in Arizona paid 

$10,468 in combined property, livestock, and other 

taxes in 1991 (Fowler and others 1994). In New Mexico 

average ranches paid $4,479 in 1991 (Fowler and others 1994). Of course, property taxes vary according to 

millage rates and the amount of private land owned 

by the ranch. Property and other taxes (workmen’s 

compensation, unemployment, and sales tax) are 

considerably higher in Arizona than in New Mexico 

(Fowler and others 1994).

Local communities also benefit from ranch-related 

expenditures from all size classes of livestock operation. Fowler and others (1994) estimate that ranch 

expenditures in local communities averaged $20,680 in 

Arizona and $16,529 in New Mexico in 1991. Although 

smaller ranches spend less in their local communities 

than larger ranches, basic family expenses, such as 

food, clothing, medical expenses, gasoline, and vehicle 

repairs, must be met to maintain a viable operation. 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 107

Ranch business also contributes to veterinary and 

livestock supply businesses as well as to restaurants 

and movie theaters. “If these ranches were no longer 

viable the local economies could expect a sharp decrease 

in their business, as many of these (businesses) are 

very dependent upon the ranches of the rural regions 

of the state (New Mexico)” (Fowler and others 1994: 

2, appendix 4).

As discussed for the Kiowa and Rita Blanca National 

Grasslands but applicable to the Black Kettle and 

McClellan Creek as well, agriculture and the rural 

lifestyle are strongly predominant, almost to the 

exclusion of other forms, in these areas. Grazing is 

the primary agricultural use, with few if any other 

industries present or predicted to develop. The local 

communities are almost entirely supported by agricultural income, and taxes on rangeland and livestock 

contribute the majority of the county’s tax revenue. 

Many of the ranchers in the immediate area are dependent upon the National Grasslands because of their 

extent and the pattern of intermingled private and 

Federal lands. Most could not sustain their operations 

without grazing permits on the Grasslands (Kiowa 

and Rita Blanca National Grasslands Geographic 

Area Assessments 1999).

In addition to economic studies, various authors have 

examined the sociocultural contributions of the rural 

ranching lifeway to the continuity of culture and tradition in the Southwest. For example, recent studies by 

Atencio (2001), Eastman and others (2000), McSweeney 

(1995), and Raish and McSweeney (2001), among others, examine the role of ranching in maintaining the 

culture, heritage, and ties to ancestral lands of the 

Hispanic villages of northern New Mexico. In many 

cases in this area public land serves as a replacement 

for grazing lands and resource collection areas lost by 

local communities in the period following U.S. conquest. 

Thus, their continued use is considered to be critically 

important in maintaining community solidarity and 

traditional lifeways (Eastman and others 2000). These 

cultural values often outweigh the purely economic 

contributions of agricultural operations (Hess 1990; 

also discussed in Raish and McSweeney 2001).

This is the case in many other areas of the Southwest 

as well, especially with respect to family-owned farms 

and ranches. Research conducted on rancher motivations in Arizona showed that approximately 72 percent 

of the 89 ranchers interviewed indicated that profits 

were not a primary motivating goal for them (Smith 

and Martin 1972). Other research between 1970 and 

1991 showed that ranchers had noneconomic values, 

obtaining social and psychological benefits from ranch 

ownership that offset economic problems (discussed 

in Ruyle and others 2000). According to Fowler (2000: 

440):

Ranchers in New Mexico and Arizona were 

similar...Arizona ranches had been in the business 

for 36 years and New Mexico 35 years; the average 

(rancher’s) age was 56 years. Arizona ranch families 

had been in their respective states for over 64 years. 

This information strongly suggests longevity in the 

industry (and) is usually associated with long-term 

commitment and stewardship. Ranching as a “way 

of life” was also supported by such demographic 

information.

Writings by ranch owners express strong attachment 

to one particular place with enduring ties to the local 

community and its economy and considerable relucFeeding on the mesa, northern New Mexico, 1993. (Photo by Alice M. McSweeney)
108 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

tance to relocate (discussed in Ruyle and others 2000). 

Although more recent writings express concern over 

low income returns from ranch operations, ranchers 

place tremendous value on their operations and want 

their children to carry on the family tradition even if 

it provides lower income than other professions. These 

writers express pride in agricultural occupations, feeling that they are superior to urban occupations. They 

respect the knowledge of the older generation, which 

is often based in experience rather than formal education. They believe that ranching allows them to feel 

closer to the earth and provides a sound place to rear 

children. There is high value placed on rural life with 

a strong attachment to the land and land stewardship 

and a deeply felt sense of place (Cofer n.d., Duncklee 

1994, Flieger 1991, Hughes 1980).

Mineral Resource Extraction on 

Forest Service Lands________________

Without doubt, livestock grazing is an important 

use of Forest Service grasslands in the Southwest; 

however, other uses of these lands merit attention. 

This section focuses on mineral/energy extraction, 

and the next section focuses recreation—two other 

major land uses.

Both surface and hard rock mining and oil and gas 

extraction occur on National Forest System lands, 

although the proportion of grasslands that support 

these activities is not known. The mining laws of 

1866, 1870, and 1872 set the stage for mineral extraction activities; under the Mining Act of 1872 and 

the Mineral Leasing Act of 1920, mining and energy 

concerns are allowed to explore, stake claims, and apply for leases on the public domain (Robinson 1975: 

4; Roth 1997: 242). Rules and regulations for mineral 

extraction on Forest Service land are presented in 36 

Code of Federal Regulations (CFR) Chapter 11 (7-1-

00 Edition), Part 228—Minerals. Standards are also 

contained in Surface Operating Standards for Oil and 

Gas Exploration and Development (USDI Bureau of 

Land Management and USDA Forest Service Rocky 

Mountain Regional Coordinating Committee 1989) 

and in the Uniform Format for Oil and Gas Lease 

Stipulations (USDI Bureau of Land Management 

and USDA Forest Service Rocky Mountain Regional 

Coordinating Committee 1989).

Although mining has a long history in the region 

going back to American Indian and Spanish Colonial 

times, large-scale mining efforts primarily developed 

after U.S. takeover. In addition to collecting and 

quarrying many types of rocks, minerals, and clays, 

American Indian groups mined turquoise for ornamentation and religious purposes with major quarries 

and mines located near present-day Cerrillos, NM. 

Galena, or lead ore, was also mined in the Cerrillos 

area, as well as in the San Pedro, Sandia, and Sangre 

de Cristo Mountains in New Mexico. Much of the lead 

was used in producing lead glaze paint for Puebloan 

pottery (Scurlock 1998: 100-103).

Mining became progressively more developed under 

the Spanish, Mexican, and American governments. 

Although exploration for mineral wealth was a major 

impetus for Spanish exploration, these riches were 

never realized in the region. In the Middle Rio Grande, 

the Spanish colonists did mine turquoise and lead, as 

the American Indian groups had done, as well as copper, silver, and gold. They also extracted nonprecious 

minerals and rocks, such as mica, which was used to 

cover window openings (discussed in Scurlock 1998: 

118-119). Silver and gold were sought during the 

Spanish and Mexican periods in southern Arizona, 

although mining in the area was limited. There is 

some evidence of mining for gold, silver, and lead in the 

region, but since later operations often used traditional 

Spanish and Mexican techniques, it becomes difficult to 

determine whether the mining occurred before or after 

the United States assumed control of the area (Hadley 

and Sheridan 1995: 46-47). According to Whittlesey 

and others (1997: 288), there were no mining centers 

north of the Gila River until 1860.

Anglo-American takeover led to major changes in the 

mining industry in technology, extent, and intensity. 

Transportation was limited until entry of the railroad 

from 1879 into the 1890s, with construction of both 

major lines and spurs into specific mining locations. 

After railroad entry, serious commercial markets 

developed. In the Rio Grande, Mexican period gold 

mines in the Ortiz and San Pedro Mountains continued to be productive after U.S. conquest. Other 

productive locations for gold, silver, lead, and other 

minerals included Elizabethtown, Cerrillos, BlandAlbemarle (which remain as ghost towns within the 

Santa Fe National Forest), and Socorro-Magdalena 

(Scurlock 1998: 129-134). Scurlock (1998: 130-132) 

gives a detailed listing of mining locations with types 

of minerals and dates of production for the Rio Grande 

area. Most of these locations are found in the forested, 

montane regions, however; Scurlock does not detail 

the extent of grasslands or meadows that might also 

be present.

American period mining in southern Arizona is 

described by Hadley and Sheridan (1995: 47-65) as 

producing several mining booms after the Gadsden 

Purchase (1854) and into the 1880s despite threats 

from Apache raiding. Lead and silver were produced 

in the Mowry area from the Mowry and adjacent mines 

near Patagonia into the 1950s. Other mines produced 

silver, lead, zinc, copper, and gold ores. Mining occurred 

at Washington/Duquesne, Sunnyside, Harshaw, and 

Meadow Valley. Mines in central and western Arizona 

increased during the 1860s as a means of financing the 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 109

Civil War (Whittlesey and others 1997: 288). Central 

Arizona mines produced gold, silver, and copper and 

were dependent to a considerable degree on the state’s 

network of railroads. Copper mines along the Verde 

River led to the boomtown of Jerome, and copper and 

other minerals drew settlement to the Tonto Basin 

and surrounding areas (Whittlesey and others 1997: 

290).

As in New Mexico, many of the historic hardrock 

mining locations are situated in or adjacent to contemporary National Forests and are generally located in 

the more mountainous, forested zones as opposed to the 

grasslands. Nonetheless, growth of the mining industry 

affected livestock production and associated grasslands 

as demands for meat, milk, hides, and tallow from the 

mining camps increased, encouraging greater livestock 

production and ranch development, producing what 

has been discussed by West (1993) as a mine-ranch 

complex (Hadley and Sheridan 1995). Contemporary 

mining activities include those for both locatable (for 

example, gold and silver) and common variety minerals (for example, common variety pumice). Some 

examples include gold, silver, and copper mines on 

the Kaibab, Tonto, Lincoln, and Gila National Forests 

and common variety mineral and rock extraction on 

virtually all Forests in Region 3. Common variety 

pumice, caliche, sand, gravel, landscape/building stone, 

and cinders are all extracted and, collectively, have 

considerable value to forest users (Linden personal 

communication 2001).

The most prevalent mineral extraction activities on 

the grasslands of Region 3 consist of oil and gas leases, 

which occur on several Ranger Districts of the Santa 

Fe and Carson National Forests and on the National 

Grasslands, as well (Linden personal communication 

2001). In some cases, these leases have been in effect 

for 50 to some 100 years. The government receives a 

12.5 percent royalty from producing oil and gas wells, 

with approximately 50 percent of revenue from public 

domain lands returned to the county and 25 percent 

of revenue from acquired lands (such as the majority 

of the National Grasslands) going back to the county. 

The Jicarilla Ranger District of the Carson National 

Forest, primarily grazed woodland/canyon country, is 

located within a major oil and gas field in the San Juan 

Basin with an estimated 630 producing wells (Linden 

personal communication 2001). Coal bed methane is 

also extracted from some 150 to 200 of the producing 

wells. In fiscal year (FY) 2000 oil and gas production 

on the Carson generated $9,168,935.69 in rents and 

royalties. The Cuba and Coyote Ranger Districts of 

the Santa Fe have about 60 oil and gas wells located 

mainly in sagebrush mesa country. The forest as a 

whole reported $287,905.71 in rents, bonuses, and 

royalties (USDA Forest Service 2000, Linden personal 

communication 2001).

Concerning the National Grasslands themselves, 

the Kiowa and Rita Blanca report no drilling on the 

Kiowa portion in New Mexico with minimal drilling 

(about two wells) on the Rita Blanca in Texas (Kiowa 

and Rita Blanca National Grasslands Geographic 

Area Assessments 1999, Linden personal communication 2001). The Black Kettle and McClellan Creek, 

on the other hand, report that most of the land 

tracts within the High Plains and Redbed Plains 

Geographic Areas (approximately 21,500 acres) are 

under oil and gas leases effective for 10 years unless 

a producing well is drilled. If a producing well is 

drilled, the lease remains effective as long as the well 

produces. The lands in the Lake Marvin (577 acres) 

and Lake McClellan (955 acres) Geographic Areas 

are also under oil and gas leases with producing 

wells. Both Marvin and McClellan are under private 

development; the Forest Service does not own the 

mineral rights, although the facilities are located 

on Forest Service land. Overall, the Black Kettle 

and McClellan Creek listed $1,425,083.82 in rents, 

bonuses, and royalties from oil and gas operations 

in FY 2000 (Black Kettle and McClellan Creek 

National Grasslands Geographic Area Assessments 

2000, USDA Forest Service 2000, Linden personal 

communication 2001).

Issues and controversies related to mining on public 

land vary according to the type, extent, location, and 

management of the extraction activity. Extraction and 

processing can cause environmental impacts through 

earth removal and the distribution of metals and 

chemicals. Removed earth can obstruct watercourses 

resulting in redirection and/or downstream erosion. 

Both groundwater and soil may be impacted by acid 

tailings and processing chemicals percolating downward into the water table (Dean 1982:1-10, discussed 

in Hadley and Sheridan 1995: 62-65). Indirect impacts 

come from increased settlement and mine working 

activity including building roads, hauling ore, and clearing and leveling home sites. Historically, woodcutting 

for fuel, construction, and ore processing had significant 

impacts on forested lands surrounding mine sites in 

the region. In many cases, impacts from historic mine 

sites are still apparent today. Hadley and Sheridan 

(1995: 64-65) describe the effects of mining at some 

of the historic southern Arizona locations in terms 

of seriously altered topography from earth removal, 

denuded lands on which trees have not regenerated, 

and eroded, downcut watercourses.

Although recent years have seen growing opposition 

to mining on public land, the Forest Service has little 

authority to deny mining on its lands. Congressionally 

designated Wilderness areas, however, have been closed 

to mining exploration since 1984 under provisions 

of the Wilderness Act of 1964 (Robinson 1975: 158-

160; Roth 1997: 238-242). States hold the permitting 
110 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

regulations for mining, and the New Mexico regulations are stringent in terms of bonding, close-out plans, 

and engineering plans (Linden personal communication 2001). Many of the small operators cannot meet 

the standards and are forced out of business. Those 

that remain can cause difficulties for Forest Service 

administrators in terms of obtaining sound operating 

and reclamation plans and proper environmental 

clearances.

Although the larger mining concerns generally 

contract for their own environmental work, Forest 

Service administration and review of these efforts 

can still be costly in terms of time and personnel. As 

an example, the proposed Carlota Copper Mine on the 

Tonto National Forest, with an environmental impact 

statement prepared by a private company, still required 

the time and effort of considerable numbers of Forest 

Service employees for review and comment. There are 

also ongoing monitoring and clean-up programs for 

abandoned mines that are on Forest land or on lands 

bordering National Forests. These can be both costly 

and time consuming (Linden personal communication 

2001).

Oil and gas leasing follows the standards set 

out in the previously mentioned Surface Operating 

Standards and Uniform Format for Oil and Gas 

Lease stipulations developed by the BLM/Forest 

Service Rocky Mountain Regional Coordinating 

Committee (USDI Bureau of Land Management 

and USDA Forest Service Rocky Mountain Regional 

Coordinating Committee 1989). Stipulations cover 

surface uses such as road construction, drilling and 

producing operations, reclamation, and abandonment. 

The density of wells, referred to as well spacing, is 

regulated by the State based on efficient draining of 

underground reservoirs. Spacing varies by formation, 

depth, and the kind of resource being exploited and 

is generally one well per 80, 160, or 320 acres. Many 

of the older leases, especially those on the Jicarilla 

Ranger District, were approved over 50 years ago 

and do not fall under the contemporary standards. In 

these cases, operators are “encouraged” to conform to 

the current guidelines, with greater success reported 

for the larger, richer operators than for the smaller 

ones (Linden personal communication 2001).

The timing and location of drilling must be consistent 

with cultural resources, wildlife, and other resource 

values in the vicinity, as is the case with other projects 

on Federal lands. Drilling takes 2 weeks to a month 

with the resulting well in production for many years. 

Impacts from drilling and production can include 

habitat fragmentation from road density, as well as 

disturbance from truck traffic hauling out water and 

oil and servicing the well. Noise from both road traffic and compressors used to pump is an additional 

disturbance (Linden personal communication 2001). 

Research is still needed to assess the standards for 

density of surface mining structures (well spacing) 

and the impacts of general operations on the varying 

grassland ecosystems and human cultural values of 

the area (Seescholtz personal communication 2000).

Recreation on Forest Service Lands__

Recreation in the National Forests of the Southwest 

has a long history and includes many varied activities such as camping, hiking, picnicking, trail biking, 

hunting, fishing, water sports, and viewing wildlife 

and historic sites. In recent years, participating in 

volunteer archeology projects such as Passport in 

Time and heritage tourism activities, such as Heritage 

Expeditions, have increased in popularity. Since World 

War II, recreation has played an extremely significant 

role in National Forest System management by virtue 

of the large numbers of people engaged in recreational 

activities on public lands. Indeed, the economic impacts 

of recreation outweigh those from traditional resource 

extraction activities (discussed in Menning and Raish 

2000, among others).

The national trend toward increased recreation 

(Boyle and Samson 1985, Flather and Cordell 1995) 

is evident in the Southwest in the rising numbers of 

recreationists in the Forests in Arizona and New Mexico 

(Raish and others 1997: 38). Although a great amount 

of recreational activity occurs in the ponderosa pine 

forests of the region, grassland areas also contribute 

to the trend in rising recreational use. Holechek and 

others (1998: 15) note:

The large human population increases in the 

United States since the 1940s have made rangelands 

increasingly important as places for people to engage 

in outdoor recreational pursuits. Hiking, camping, 

trail biking, picnicking, hunting, fishing, and rock 

hounding are some of the important recreational 

uses of rangelands. The importance of open space, 

scenery, and aesthetic values from rangelands in the 

United States is difficult to quantify.

Recreation use increased in Region 3 (including the 

National Grasslands) from 1992 to 1995 with over 40 

million visitor days recorded in 1995 (Raish and others 

1997: 38, fig. 8). Most visitors were viewing scenery, 

camping, picnicking, or swimming. Hiking, horseback riding, and river rafting have increased rapidly 

(Flather and Cordell 1995), while hunting, fishing, 

winter sports, and resort camping have remained 

fairly stable. Nonconsumptive wildlife recreation, 

such as bird watching, has also increased (Raish and 

others 1997: 38, fig. 8), as has visiting historic sites 

and heritage tourism in general.

In 1996, wildlife-related recreation expenditures tied 

to National Forests totaled $6.8 billion, with anglers 

spending $2.7 billion, participants in nature-related 

activities spending $2 billion, and hunters spending 

$2 billion (NatureWatch 1996). In the same year, 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 111

wildlife watching expenditures/sales totaled $426.9 

million in Arizona and $223.2 million in New Mexico 

(NatureWatch 1996). With respect to fishing in the 

Southwest, the following information was drawn from 

the Region 3 Forest Service Web site (www.fs.fed.us/r3, 

select Recreation, FishUS, September 2001):

The uniqueness of the Southwestern Region’s fisheries is a valued resource. Fishing...is big business, 

critical to the rural economies of the arid Southwest. 

The streams and lakes of the Southwest provide 10 

percent of all angling (3.9 million angler visits per 

year) in national forests and grasslands across the 

nation. Fishing is so popular in the Southwest that no 

other Forest Service region has the fishing pressure 

per surface acre of water (more than 300 hours of 

fishing per acre annually on 48,735 surface acres). 

According to the latest United States Fish and Wildlife 

Service survey (1996), fishing in the Southwest brings 

in more than $550,000,000 annually. More than 4,000 

jobs are directly related to fishing enterprises, with 

another 10 to 15 thousand jobs indirectly associated 

with fishing in Arizona, New Mexico, and Texas...In 

Arizona 54 percent of the fish habitat is in the national 

forests; in New Mexico, 25 percent.

This section covers fishing, hunting, and wildlife 

viewing—all major recreational activities along the rivers and lakes in the grassland areas of the region.

Examination of recreation trends on all the forested 

lands of the region is well beyond the scope of this 

discussion, so the focus here is on recreation in the 

grasslands. Regionally, grassland recreation occurs on 

the Apache-Sitgreaves National Forest in the montane 

grasslands around Big Lake and within the Escudilla 

Wilderness, as well as at artificially created lakes in the 

Tonto Basin (Tonto National Forest), although much 

of this latter area is Sonoran Desert. The Valle Vidal 

(Carson National Forest) and the Pecos and San Pedro 

Parks Wilderness areas (Santa Fe National Forest) 

have considerable acreages of montane grassland that 

provide recreation, elk hunting, and livestock grazing (Moir personal communication 2003). According 

to Brent Botts, Recreation Assistant Director for 

the Southwestern Region, a request was sent to the 

Forests of the Region for information on recreational 

activities and issues on their grassland area, but little 

additional information was received except from the 

officially designated National Grasslands—Kiowa/Rita 

Blanca and Black Kettle/McClellan Creek (Botts personal communication 2001). Hence, these areas form 

the basis of the present review.

Visitor use figures are reported by Forest and are not 

separated out by grassland portions of specific National 

Forests. National Visitor Use Monitoring reports for 

2000, 2001, and 2002 were sought for the Kiowa/Rita 

Blanca and Black Kettle/McClellan Creek National 

Grasslands to determine visitor days for these units, 

but figures for these areas are included with figures 

for the Cibola National Forest (which manages the 

Grasslands) as a whole and are currently not broken 

out for the National Grasslands themselves.

Recreation use on the three geographic areas of 

the Kiowa/Rita Blanca includes both developed and 

dispersed sites. There is one developed campground in 

Mills Canyon located adjacent to the Canadian River, 

and three developed day-use areas on the Southern 

Prairie, including two picnic grounds and an interpretive

hiking trail associated with the Santa Fe Trail. 

There are no developed facilities in the Mills Upland 

Geographic Area (Kiowa and Rita Blanca National 

Grasslands Geographic Area Assessments 1999).

The majority of recreation use on the Kiowa/Rita 

Blanca is dispersed. In the Mills Canyon Geographic 

Area hunting and fishing are the two primary uses, 

which include hunting bear (Ursus americanus), 

Barbary sheep (Ammotragus lervia), mountain lion 

(Felis concolor), deer (Odocoileus hemionus), (scaled) 

quail (Callipepla squamata), dove (Zenaidura macroura), and pronghorn antelope (Antilocapra americana), 

as well as warm-water fishing in the Canadian River. 

There is also dispersed camping, hiking, picnicking, water play, and scenery and wildlife viewing. Mills Canyon 

has been designated a New Mexico Wildlife Viewing 

site in cooperation with the New Mexico Department 

of Game and Fish. Mills Upland has only dispersed 

recreation with hunting deer, quail, dove, pronghorn 

antelope, and coyote (Canis latrans) the main use. 

Other recreational uses of the area include camping, hiking, picnicking, viewing scenery, and wildlife 

watching. Southern Prairie also has a predominance of 

dispersed recreation featuring the same types of uses 

as the other areas (Kiowa and Rita Blanca National 

Grasslands Geographic Area Assessments 1999).

Although three developed camping sites exist within 

the High Plains and Redbed Plains Geographic Areas of 

the Black Kettle/McClellan Creek National Grasslands, 

the majority of recreation use is dispersed, with an 

emphasis on hunting and fishing as described for the 

Kiowa/Rita Blanca. Other uses include developed 

and dispersed camping, hiking, bird watching, and 

horseback riding. The Black Kettle is considered one 

of the best public hunting locations in the country for 

Rio Grande turkey (Meleagris gallopavo intermedia) 

and bobwhite quail (Colinus virginianus) and is the 

largest public hunting area in the western half of 

Oklahoma. It is also the only public hunting unit in 

this region where an over-the-counter permit can be 

purchased. These qualities have led to problems with 

hunting pressure during some seasons. Surveys found 

one hunter per 54 acres on the Black Kettle during the 

opening day of the 1999 deer rifle season, leading to a 

potentially unsafe situation for the public (Black Kettle 

and McClellan Creek National Grasslands Geographic 

Area Assessments 2000).

Another problem related to dispersed recreation use 

has been indiscriminate driving on interior (unpaved, 

undeveloped) roads and indiscriminate parking in 
112 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

undeveloped areas. These activities have impacted 

soil, water, and wildlife resources. Holechek and others 

(1998: 446-447) and Payne and others (1983) discuss 

the damage that unregulated, off-road vehicle travel 

can cause to soils and vegetation under different 

moisture regimes and during different cycles of the 

growing season. They conclude that unregulated offroad vehicle travel can be as destructive as uncontrolled 

grazing. They recommend regulation of recreation 

use on rangelands, just as other uses are regulated 

(Holechek and others 1998: 447). The High Plains and 

Redbed Plains Geographic Areas have moved in this 

direction by creating designated parking locations for 

dispersed recreation, which are designed to reduce 

vehicle access and encourage walk-in use. The areas 

have also instituted road closures in conjunction with 

the parking spaces for the same purpose (Black Kettle 

and McClellan Creek National Grasslands Geographic 

Area Assessments 2000).

Recreation on the Lake McClellan and Lake Marvin 

Geographic Areas (in Texas) tends to be considerably 

more developed than on the High Plains and Redbed 

Plains, although dispersed recreation is also present. 

Lake Marvin offers warm-water fishing, developed 

camping (two campgrounds and a recreation building), 

picnicking, hiking, and bird watching. The primary uses 

are fishing and camping, but hiking the trails and viewing birds and other wildlife are also popular. Hunting is 

prohibited owing to the area’s small size (Black Kettle 

and McClellan Creek National Grasslands Geographic 

Area Assessments 2000).

Developed recreation facilities in the Lake 

McClellan Area are managed under a concessionaire 

permit and consist of several campgrounds and picnic 

grounds, as well as a store offering rentals such as 

paddleboats and lifejackets. Activities at the lake 

include camping, picnicking, fishing, hiking, and 

wildlife viewing, as well as boating, jet skiing, bicycle 

riding, and motorcycle riding. The north end of the 

area is open all year, while the south end is closed 

from November 1 to March 31 to protect eagle roost 

habitat. Dispersed recreation throughout the area 

consists of fishing, boating, water play, camping, and 

motorcycle riding; fishing is the primary dispersed use. 

Hunting is generally not allowed. Motorcycle riding 

is permitted by the concessionaire and is popular on 

the motorized trail system. It is, in fact, the primary 

use of the trail (Black Kettle and McClellan Creek 

National Grasslands Geographic Area Assessments 

2000). No problems were mentioned concerning the 

impact of motorcycle use, which is apparently being controlled and regulated by designation of the 

motorized trail.

As discussed previously, however, off-road vehicle 

use by recreationists is growing in extent and numbers of occurrences and has considerable potential for 

negative environmental impacts. All-terrain vehicles 

(ATVs), in particular, are growing in popularity, and 

their use is on the rise in many areas, with the potential 

for serious environmental damage. The growing use 

of off-road vehicles has prompted a study by the San 

Dimas Technology and Development Center (USFS) 

on the resource impacts of off-road vehicles on soil, 

water, air, vegetation, wildlife, and aquatic species of 

National Forests throughout the country. The ongoing 

project is a survey examining types of uses, equipment, 

resource impacts, and use regulations.

A Nevada study examined by Holechek and others 

(1998: 446-447) compared the effect of motorcycle, 

four-wheel-drive truck, and no traffic on infiltration 

rate and sediment production. In the study, infiltration 

rates were decreased and sediment production was 

increased by both forms of traffic, with the four-wheeldrive truck having considerably greater impact than 

the motorcycle (Eckert and others 1979). As populations and urbanization increase in the Southwest, all 

forms of recreational use will increase on the region’s 

rangelands. Assessing, regulating, and controlling the 

effects of these uses will become increasingly important. 

The population growth, with its attendant urbanization and suburbanization, are the topics of the final 

section of this chapter.

Continuing/Future Trends in 

Southwestern Rangelands and 

Rangeland Management____________

Conversion of rangeland to nonagricultural uses 

during the past 20 years is becoming an important 

concern for public land managers and private ranchers 

and farmers. In the United States, this trend primarily 

results from human population increases and a demographic shift from the eastern to the western portion 

of the country (Holechek 2001: 39). World population 

projections show the North American population 

rising from 172 million in 1950 to 297 million in 

1995 with a projected increase to 384 million in 2050 

(Holechek 2001: 39, table 1; McQueen 2000, United 

Nations 1998). In a listing of States with the greatest population growth rate from 1990 through 1994, 

Arizona ranks third and New Mexico ranks eighth 

(Mitchell 2000: 25, table 2.9; Riebsame and others 

1997). Increasing urbanization and suburbanization 

result from both population growth and inmigration 

from other areas. For example, Bahre (1995: 243-244) 

describes southern Arizona:

The rapidly growing population of Arizona, especially since the 1940s, has led to expanded urban 

and rural development on privately owned lands and 

is threatening federal and state trust lands, which 

are continually being sold to the public for development. Since 1950 the number of rural sub-divisions 

being built in southeastern Arizona has exploded, 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 113

especially in Cochise County, which has both the 

largest area of grassland and the largest amount 

of private land (41 percent)—almost all of it being 

used for agriculture, ranching, or subdivision development (Hecht and Reeves 1981). Cattle numbers in 

the county have declined, in part because of tract 

development in former rangelands, and much private 

grazing and agricultural land has been purchased 

by people seeking rural retirement or investment 

opportunities.

It often seems that the urban residents, as well as 

those newly arrived in the region, have different land 

attitudes and values from many of the long-time rural 

residents. These differing views can lead to conflict over 

land use and land management goals and strategies. 

Conflicts over environmental protection legislation 

and the role of organized environmental groups are 

examples. Many present-day environmental issues and 

concerns in the Southwest are aligned with a national 

shift in public attitudes and values concerning the 

appropriate use of public land, primarily in the West 

(Macon 1998). Many urban Americans now hold environmental protection oriented public land values, in 

contrast to the commodity and community economic 

development orientation of the earlier conservation 

era (from approximately 1900 to 1969) (Kennedy and 

others 1995, Macon 1998). As Kennedy and others 

(1995) point out, rural agricultural communities use 

and view land differently from urban groups, thus 

producing contrasting perceptions and values. A large 

amount of contemporary conflict over timber, wildlife, 

and rangeland issues stems from disjunct rural and 

urban values concerning human relationships with 

nature and its uses.

Nature values are not intrinsic but are human 

creations originating in the minds of individuals 

and groups as their changing perceptions and needs 

interact with environmental, political, and economic 

systems (Kennedy and others 1995: 128). For example, 

the conservation movement of the early to mid-20th

century fit well with the needs of an industrialized 

nation for sustained-yield timber and forage production to provide commodities for growing factories 

and cities. Even though recreational, biocentric, and 

esthetic values were a part of some early conservation 

visions, these views did not become a dominant force 

in public attitudes until the 1960s, with the advent of 

an urban, postindustrial society and the environmental 

movement (Kennedy and others 1995).

The later 1960s and 1970s saw passage of the main 

body of environmental protection legislation. Mitchell 

(2000: 7) in a discussion of legislation affecting rangelands stated:

From a legislative context, the 1970’s could be considered the decade of the environmental movement. 

The Wilderness Act was enacted in 1964. Then, starting in January 1970 with the Environmental Policy 

Act of 1969, no fewer than 12 major environmental 

laws affecting the conservation and management of 

U.S. rangelands were signed into law during the following 10 years. Among such laws were the Resources 

Planning Act of 1974 (RPA) and the National Forest 

Management Act of 1976 (NFMA), which called for 

a recurring assessment of America’s forest and 

rangeland situation.

Other pertinent legislation includes the landmark 

Endangered Species Act of 1973, the Federal Land 

Policy and Management Act of 1976 (FLPMA), the 

Forest and Rangeland Renewable Resources Research 

Act of 1978, the Public Rangelands Improvement Act 

of 1978, and the Archeological Resources Protection 

Act of 1979.

The orientation of range research has also changed 

over the years from a more traditional range/livestock grazing system approach to a more integrated 

ecosystem approach. Contemporary research now 

includes more ecologically based subjects such as 

watershed and riparian area management, biodiversity, disturbed site reclamation, and wildlife/livestock 

interactions, to name a few (Evans 1990, Everett 1992). 

Additionally, studies concerning grazing on public land 

now include information on visitor attitudes about grazing (Mitchell and Fletcher 1996), which indicates the 

increasing role of public concern and input into land 

management agency decisionmaking.

One result of the growing role of both public and 

interest group involvement in Federal agency planning and decisionmaking is increased litigation over 

agency projects, programs, and plans. A rising number 

of appeals and lawsuits from both environmental and 

industry groups have resulted from the agency’s efforts 

to bring livestock grazing into compliance with Federal 

environmental statutes. For example, during 2001, 

there were 11 grazing-related lawsuits against the 

Southwestern Region of the Forest Service (Gonzales 

personal communication 2002).

The increasingly litigious environment surrounding rangeland management has spurred the rise of 

regional conflict resolution groups such as the Malpai 

Borderlands Group in southeastern Arizona and southwestern New Mexico, the Diablo Trust in northern 

Arizona, the Arizona Common Ground Roundtable, 

and the Quivira Coalition in New Mexico, to name a 

few (Raish 2000c, Ruyle and others 2000, Sheridan 

2001). These groups attempt to protect both the environment and traditional lifeways and to diminish conflict, 

controversy, and litigation. They often include environmentalists, ranchers, and public land managers and 

espouse collaborative stewardship, mutual education, 

and citizen participation in agency decisionmaking. 

They emphasize that sound land management and 

healthy ecosystems can coexist with traditional rural 

economic practices such as ranching (Guess 1999).

Ruyle and others (2000) describe the work of the 

Malpai Borderlands Group and the Diablo Trust to 

coordinate management of multiagency lands as a 
114 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

promising means of maintaining the viability of ranching on public land, protecting open space, and reducing 

habitat fragmentation. The Malpai Borderlands Group 

has worked with the Nature Conservancy to protect 

and restore wildlife habitat and with public land 

managers in projects to restore fire to the ecosystems 

of the area. They have also developed a program of 

conservation easements to protect open space from 

future development and subdivision (Clifford 1998, 

Guess 1999).

These groups focus on the threat of growing development and suburbanization to Western lands and 

landscapes, as escalating rates of development are 

impacting both private lands and adjacent public 

lands. U.S. Department of Agriculture statistics show 

that the nation lost almost 1.4 million acres a year to 

development from 1982 to 1992. In the mid-1990s the 

rate more than doubled to almost 3.2 million acres per 

year (Rome 2001: 264). Land-use change from population growth is greater in rural areas than urban ones 

because of the dispersed nature of exurban development. Exurban development refers to rural residential 

development that occurs beyond incorporated city limits 

and often results from the subdivision of ranches into 

smaller parcels for home sites (“ranchettes”) generally 

ranging from 1 to 20 acres or so (Sullins and others 

2002, Theobold 2000). Almost 80 percent of the land 

used for houses built between 1994 and 1997 was in 

nonmetropolitan areas (Heimlich and Anderson 2001). 

In addition, locations in proximity to public land are 

particularly desirable for development (Riebsame and 

others 1996, Swanson 2001).

Sheridan (2001: 146-147) describes development, which has escalated since the late 1960s, in 

the Sonoita-Elgin area of southern Arizona plains 

grassland near the Coronado National Forest. In 1989 

Sonoita had approximately 400 homes, which had 

risen to 707 by 1995, a 76 percent increase. Numerous 

ranches have been subdivided in the region in the 

past 20 years, leading to segmentation of the open 

grassland into ever smaller parcels (the minimum lot 

size is currently 1 to 3 acres). Sheridan (2001: 146-

147) discusses the ways in which the resulting land 

use and human occupation can affect wildlife, native 

vegetation, and soil cover. Problems can include house 

and outbuilding construction, fencing (which inhibits 

the movement of large wildlife), removal of native 

vegetation, introduction of exotics, and introduction 

of domestic pets that prey on or harass wildlife. This 

type of land fragmentation acts against maintaining 

the biodiversity that stems from large, unfragmented 

ecosystems, as wildlife habitat and corridors that link 

grasslands with montane use areas are fragmented 

or destroyed. There is also a growing constraint on 

large-scale ecosystem management, such as the reintroduction of fire as a natural process or management 

tool, which is generally resisted as a threat to homes 

and property.

According to James H. Brown, a University of New 

Mexico biologist and past president of the Ecological 

Society of America, “Far more habitat has been destroyed to provide water for cities, subdivisions, and 

irrigated (commercial) agriculture than by even the 

heaviest grazing pressure” (quoted in Clifford 1998). 

In an effort to explore this notion of the negative 

effects of development, Maestas and others (2002) 

conducted a study designed to examine wildlife 

species and plant diversity on ranches, protected 

areas (Colorado Division of Wildlife’s State Wildlife 

Areas), and exurban developments in the foothills of 

the Colorado Front Range of the Rocky Mountains 

northwest of Fort Collins. The wildlife species studied 

were mammalian carnivores and songbirds. The study 

found that biodiversity differed among the different 

land use categories, with wildlife species occurrence 

and densities more similar between ranches and 

protected areas than on exurban developments. In 

terms of plant communities, the most nonnative species were found on exurban developments, with more 

species of native plants found on ranches than on the 

other two land uses. The most common nonnative was 

cheatgrass. However, the cover of native species did 

not differ statistically across the land uses (Maestas 

and others 2002). The information derived from this 

study has considerable implications for conservation 

efforts. Maestas and others (2002: 27) stated:

A generalization from our study is that there is an 

increase in human-adapted wildlife and non-native 

plant species with exurban development. Interactions 

among native, non-native, and human-adapted species could result in the simplification of the Mountain 

West’s natural heritage favoring species whose 

evolutionary life histories allow them to exist with 

humans. This change has negative implications for 

the maintenance of biodiversity at both the site and 

landscape scales and its consequences are increased 

with increasing development (Knight 2002).

This chapter has shown how Southwestern grasslands have been impacted by human activity for 

thousands of years. Native American groups used 

and modified the area for hunting, gathering, and 

agriculture before and after the arrival of Spanish 

colonists in the 1500s. European settlement brought 

a wide array of new domesticated plants and animals, 

as well as new technologies and land-use strategies. 

However, changes and impacts to the area’s land base 

accelerated more rapidly with commercialization of 

both farming and ranching after U.S. takeover, with 

development of railroad transport and wider national 

markets. In the years since the beginning of the 20th

century and especially after World War II, conservation and preservation efforts have come to the fore as 

concerns for ecosystem health and sustainability have 

increased as national priorities. New challenges now 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 115

face the region in terms of population growth and inmigration resulting in a proliferation of urban, suburban, 

and exurban developments, affecting rural land uses, 

water supplies, and wildlife habitat. Sheridan (2001: 

147) notes: “Some human impacts can be reversed but 

subdivisions are forever.” Managing this new challenge 

of human population growth and development will 

undoubtedly be a major effort of public land managers, 

environmental groups, and farmers and ranchers into 

the 21st century.

Acknowledgments_________________

For their well-thought-out, helpful reviews of this 

chapter, I thank John Brock, Arizona State University; 

Clyde Eastman, New Mexico State University (retired); 

Will Moir, Rocky Mountain Research Station (retired); 

Rex Pieper, New Mexico State University; and George 

Ruyle, University of Arizona. Their efforts have improved the final product considerably.

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120 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Introduction_______________________

Southwestern grasslands today share general differences from their pre-Euro-American settlement 

conditions. With few exceptions, grasslands—whether 

in the desert, prairie, or mountains—were, prior to 

non-Indian settlement, more diverse in plant and 

animal species composition, more productive, more 

resilient, and better able to absorb the impact of disturbances. Southwestern grasslands today are missing 

the elements of disturbance regimes that kept them 

functioning in the prelivestock period. Such interruptions include human activities leading to loss of 

palatable plant species and keystone grazers such as 

the buffalo and prairie dog. The loss of most fire events 

resulted in the accelerated loss of soil from wind and 

water erosion, including loss of the highest productivity 

sites from gully formation. Soil changes also include 

loss of soil pore space through compaction, decrease 

of soil organic matter, retention of surface water, and 

nutrient cycling processes whether permanent or 

temporary. The result is a loss of large-scale connectivity by isolating grasslands through urbanization and 

other anthropogenic influences.

The frequency, magnitude, and extent of these 

changes vary among major kinds of grasslands and 

among landscapes and plant communities occurring 

within them. But almost everywhere in the Southwest, 

the amount of change is vast. Perhaps if early alarms 

voiced by Wooton in 1908 in New Mexico, Griffiths 

in 1901, and Thornber (1910) and Jardine and Hurt 

(1917) a few years later had been heeded, much of 

the subsequent disruptive change could have been 

averted—at least in those places not already degraded 

too far to recover and resemble their former state.

For example, Humphrey (1987) and Turner and 

others (2003) photographed landscapes along the 

Southwestern United States-Mexico boundary to document changes in vegetation as seen in prior historical 

photos. From the severely degraded conditions in the 

1890s and the limited amount of recovery since, many 

irreversible changes had already taken place in the 

unforgiving climate along the border. Farther north, 

Aldo Leopold wrote in 1924 (p. 8), “[In] northern 

Arizona there are great areas where removal of grass 

by grazing has caused spectacular encroachment of 

juniper on park areas. But here again both grass 

competition and fire evidently created the original 

park, and both were removed before reproduction 

came in.” Many thousands of pages have been written since that time, with sound recommendations 

for moving Southwestern grasslands back toward 

ecosystem sustainability. Much progress has been 

made toward recovery, but still the gap widens across 

tens of millions of acres.

With the exception of older Euro-American occupations such as in the Rio Grande Valley, most change 

has taken place since the end of the Civil War and the 

Chapter 6:

Historic and Current Conditions of 

Southwestern Grasslands

Reggie Fletcher

Wayne A. Robbie
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 121

beginning of the cattle era. The cattle boom began in 

1865, reached its height about 1885, and gradually 

adjusted to an environment that no longer included a 

frontier. Between 1875 and 1883, much of the grassland 

was fenced with barbed-wire (Hollon 1961). The advent 

of the railroads increased the extent, area, and rate 

of change, shrinking grasslands, as they once were, 

into smaller, relictual fragments, and relegating them 

to remote locations, or at least locations remote from 

water. The 1936 Department of Agriculture report to 

the Senate titled “The Western Range” documented 

the widespread degradation of the grasslands. Some 

553 million acres of Western rangeland, 76 percent of 

the total, were continuing to degrade at the time of 

the report (USDA 1936). Much of the upland deterioration has been halted, with restoration of adequate 

vegetation cover to stem upland erosion over at least 

a majority of the grasslands. However, the creation 

of gullies had effectively and almost universally 

dewatered the grasslands with the exception of the 

flatter portions of the shortgrass prairie. The gullies 

reduced the effectiveness of the precipitation across 

the grasslands with some loss of cienega and sacaton 

communities. Recovery of the grassland cover in the 

uplands has occurred and continues, but without 

parallel restoration of the incised valley floors and 

basins.

Pre-Euro-American settlement grasslands are difficult to describe accurately. Qualitative and descriptive 

information must be pieced together and inferences 

made from inventories of isolated and protected 

landscape fragments, pastures managed for good 

and excellent rangeland health, properly functioning 

watersheds, historical documents, life-history studies 

of plants and animals, professional experiences of those 

working and doing research in grassland ecosystems, 

and documented accounts of early explorers and land 

survey records.

Reference conditions that describe what any sustainable historic grassland community should ecologically 

represent are difficult to achieve. Even the most homogeneous grassland has extremely dynamic ecological 

processes. Plant species readily vary with time on 

any given site according to rainfall patterns and the 

latest disturbance process. Vegetation dynamics in 

grasslands can be characterized by the shifting patterns of abundance and sequential species replacement 

over time, or succession. The complex patterning of 

small-scale partitioning by grassland species in less 

degraded conditions makes consistent classification 

challenging at a finer scale.

For all except the montane grasslands, increased 

vegetative cover and productivity of pre-EuroAmerican settlement grasslands are hypothesized to 

have increased convective warm season precipitation 

over today’s rainfall amounts (Grissino-Mayer 1996). 

However, drought in earlier times was just as important 

a disturbance factor as it is today. Short-term deviations in rainfall amount and timing influence species 

frequency and occurrence, but long-term changes have 

an effect on grassland boundaries and extent.

The lower elevation grasslands have numerous 

plant species in common. Long-term climate changes 

enable some species to spread long distances while 

others retract in ebb and flow across time and space. 

The result leaves relictual pockets of species far from 

their normal associates, with big bluestem (Andropogon 

gerardii) migrating from the tallgrass prairie to the 

Zuni mountains of New Mexico, and black grama 

(Bouteloua eriopoda) from the desert grassland to 

Kansas and Oklahoma. While today’s conditions favor 

movement of more invasive species, the more diverse 

and resilient systems of yesterday would have favored 

periodic expansion of a wider variety of species.

Because they were more diverse than the grasslands 

of today, all grasslands in the Southwest were historically able to withstand drought. Tilman and Downing 

(1994) found grassland species richness led to greater 

drought resistance because some grasses within the 

complex of species were more drought resistant than 

others and partially compensated for the decreased 

growth of the less drought resistant species. This 

appears to have been most prominent in the prairie 

grasslands, although their topographic homogeneity 

may have made such relationships more noticeable. 

The higher diversity in all grasslands would have had 

the similar benefits of fire, hail, insect outbreaks, and 

other disturbances. Even with the benefit of increased 

diversity, there were periodic events of sufficient magnitude to overcome the resiliency of any grassland. The 

common situation in grassland ecosystems, both then 

and now, is to be in a state of transition from some 

type of disturbance.

Wooton (1908) noted the New Mexico Territory included approximately 300 grasses and sedges, of which 

only 25 or 30 furnished the “great bulk” of livestock 

forage. Livestock homogenized the grasslands, masking 

or eliminating both local and regional differences. The 

ecotones of the nonmontane grasslands were naturally 

broad. Tolerance limits of species such as blue grama 

(Bouteloua gracilis) and sideoats grama (Bouteloua 

curtipendula) were broader than the limits of the 

major grasslands.

It is useful to look at the distribution of accompanying shrubs to set a practical boundary for the various 

grasslands. The limits of big sagebrush (Artemisia

tridentata var. tridentata) help set a climatic and physiographic boundary for Colorado Plateau and Great 

Basin grasslands. Creosote bush (Larrea tridentata) 

likewise is an aid in approximating the boundary for 

desert grassland and shrubland. The zone of integration 

between Yucca glauca of the shortgrass prairie and 
122 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

the taller Yucca elata of the desert grassland provides 

a good boundary. The ecotone between the two is the 

home of Yucca intermedia, an intermediate between 

the two yuccas.

Montane Grasslands_______________

Montane grasslands ranged from the alpine and 

subalpine regions at high elevations through the 

ponderosa pine (Pinus ponderosa) ecosystem at lower 

elevations. These grasslands were the most naturally 

fragmented and ranged in size from thousands of 

acres, such as those in the White Mountains of Arizona 

where grasslands covered about 80,000 acres above 

9,000 feet (Baker 1983), to only a few acres, limited by 

topography or the surrounding forest. The origin and 

maintenance of these grasslands, except for those few 

acres in the alpine, are a subject of debate, perhaps 

because the causative factors not only vary from place 

to place but also usually combine to create a variety 

of effects. Below the alpine zone, fire and climate 

appear to have been the major factors in both the 

creation and maintenance of these grasslands. For wet 

meadows, fire may have been the ultimate creator of 

many openings; with soil moisture levels beyond what 

trees could tolerate maintaining them. Wind desiccation, and snow and ice abrasion readily maintained 

at least the larger meadows and may have interacted 

with drought and/or insect and disease outbreaks in 

the creation of some meadows. The greater level of soil 

organic material and higher inherent productivity of 

meadow communities helped grasses and grasslike 

plants play a major, competitve role in maintaining 

these grasslands as tree free.

In pre-Euro-American settlement times, most 

montane grasslands would have been less fragmented 

with a greater degree of connectivity, and their total 

acreage would have been greater. Almost all of these 

grasslands, either large or small, have either yielded 

acreage to forest, or disappeared altogether. Inhibition 

of water movement through the soil—due to compaction by livestock and ungulate wildlife—decreased 

soil moisture availability because of increased tree 

density in surrounding forests. Water diversions and 

road development combined to reduce the condition 

and extent of wet meadows.

The alpine ecosystem in the Southwest has few 

gentle slopes conducive to development of grasslands. 

Most alpine is dominated by the forb-rich fellfield community, but where the topography permits, Kobresia 

(Kobresia myosuroides) dominated communities of 

sedges and grasses are considered cushion plants and 

develop what can be termed a grassland (Andrews 

1983). Baker (1983), in his study on Wheeler Peak, 

identified 10 alpine communities in a complex mosaic 

defined mainly by slight variations in topography, 

exposure to wind and sun, and snow accumulation. In 

the harsh elements above timberline, “grassland” turf 

is compact and complete in cover except where broken 

by surface rock or pockets of gopher activity. Where the 

sod has been broken, as in the case of heavy sheep use, 

or misplaced recreation trails, wind erosion unravels 

the turf to the rocky substrate below. The degraded 

area expands until a change in aspect or a surface rock 

boundary is reached. The loss is permanent. Old soil 

level marks on rocks near Santa Fe Baldy reveal near 

complete loss of the Kobresia community in that area. 

Unlike other grassland communities, fire is not likely 

to have played a role in either the creation or maintenance of the alpine grasslands. Pre-Euro-American 

settlement alpine grasslands would have covered a 

few thousand acres.

As with subalpine Thurber fescue (Festuca thuberi) 

grassland communities, there is considerable intermixing with other grasses such as oatgrass (Danthonia

spp.), which tends to favor slightly less productive 

sites than Thurber fescue. Thurber fescue is often 

well in excess of a meter tall and is much sought after 

by large herbivores. Thurber fescue, was more of a 

community dominant and was more widespread in 

pre-Euro-American settlement times than in today’s 

subalpine grasslands. The result has been an expansion in dominance of sheep fescue (Festuca ovina) from 

above, and Arizona fescue (Festuca arizonica) from 

below. These high elevation (more than 8,500 feet) 

grasslands were, and continue to be, diverse in both 

grasses and forbs.

Allen (1984) documented extensive reductions of the 

Thurber fescue grasslands in the Jemez Mountains 

within the previous hundred years as both the conifer 

and aspen forests expanded. An interrupted fire regime 

and a decreased grassland competitive ability due to 

overutilization by large herbivores have contributed to 

a widespread reduction in Thurber fescue grasslands. 

These high elevation grasslands are well adapted to 

periodic fire.

The Arizona fescue (Festuca arizonica)/mountain 

muhly (Muhlenbergia montanus) grassland community 

in the mixed-conifer and ponderosa pine forests were 

naturally less productive and less diverse than the 

Thurber fescue community, but far more widespread. 

These grasslands and their variations, such as the 

screwleaf muhly (Muhlenbergia virescens) community, 

were the most extensive montane grasslands of the 

Southwest. A few relictual areas still have mountain 

muhly as a significant component in lower elevation 

woodland openings, but how common this once was 

is unknown. Because these grasses were the primary 

grasses in the understory of the open pine forests, even 

the smaller forest openings had a greater connectivity 

than in the understory-deficient, dense pine forests 

of today. In 1911, Woolsey (quoted in Fletcher 1998, 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 123

p. 86) described the typical yellow pine forest of the 

Southwest as a “pure park-like stand made up of scattered groups of from 2 to 20 trees, usually connected 

by scattering individuals. Openings are frequent, and 

vary greatly in size. Within the type are open parks 

of large extent...”

Arizona fescue and screwleaf muhly are cool season 

species, and mountain muhly is a warm season grass. 

Between the two variations is a natural ebb and flow in 

composition in step with periodic variations in seasonal 

precipitation. Arizona fescue is more tolerant of fire than 

mountain muhly. Large Arizona fescue grasslands in 

what would otherwise be the upper extent of ponderosa 

pine forest, such as Hart Prairie near Flagstaff, appear 

to have burned with a higher frequency. This favored 

Arizona fescue over mountain muhly and limited the 

density of surrounding ponderosa pine, resulting in 

an extremely open forest. The 1904 inventory of forest 

conditions on what is now the Coconino National Forest 

(Leiberg and others 1904, p. 35) noted “large parks occur 

in townships 21 and 22 north, ranges 4 and 5 east, one 

of them containing 16,000 acres.”

The cool season growth of Arizona fescue played 

a large role in maintenance of the larger parks and 

smaller openings by directly competing with ponderosa 

pine seedlings. Maximum growth of ponderosa pine 

coincides with maximum growth of Arizona fescue. 

Pearson (1949) reported that numerous ponderosa pine 

seedlings had started in an Arizona fescue opening 

in the prolific seedling year of 1919, but all had died 

within a year. Where Arizona fescue was suppressed or 

reduced by cattle grazing, the survival of ponderosa pine 

seedlings increased, and a point was quickly reached 

where ponderosa pine dominated. Today, Arizona 

fescue still continues to decrease in the dense forest 

understory across much of the Southwest.

Before the replacement of most wet meadow plant 

diversity with Kentucky bluegrass (Poa pratensis), wet 

meadows were typical for all but the narrowest and 

steepest drainages in Southwestern forested communities. The expansion of Kentucky bluegrass was made 

possible by heavy livestock use in the wet meadows. 

Compaction of wet meadow soils, particularly during 

dry years, inhibited water movement below ground, 

effectively shrinking the wet meadow and resulting 

in the elimination of riparian vegetation and favoring 

expansion of Kentucky bluegrass. Kentucky bluegrass 

has also replaced native grasses in some of the upland 

meadows (Dick-Peddie 1993).

Desert Grasslands__________________

The climatic potential for desert grasslands has 

been more or less similar to that of today for the past 

4,000 years (Van Devender 1995). Desert grassland 

in the pre-Euro-American settlement era occupied 

extensive acreage across the Southwest between 

3,000 and 5,000 feet elevation. (Martin 1975). The 

grasslands were largely free of “brush,” and streams 

and rivers dissecting the grasslands were lined with 

galeria forests and marshes (cienegas) (Bahre 1995). 

Just how much some portions of the desert grassland 

have changed can be visualized from a description of 

southeastern Arizona in the mid-19th century:

The valley bottoms were covered by a dense growth 

of perennial sacaton grass, oftentimes as high as the 

head of horseman and so thick and tall that cattle, 

horses, and men were easily concealed by it. The 

uplands were well covered with a variety of nutritious 

grasses, such as the perennial black grama, and the 

many annuals that spring into growth during the 

summer rainy season. The abundant vegetation, both 

on highlands and in valley bottoms, restrained the 

torrential storms of the region so that there was no 

erosion in valley bottoms. Instead the rainfall soaked 

into the soil and grew grass.

Sloughs and marshy places were common along 

the San Simon, the San Pedro, the Santa Cruz and 

other streams, and even beaver were abundant in 

places where it would now (1919) be impossible for 

them to live. (USDA 1937, p. 23)

Another reference from the same era noted beaver 

on the Rillito and on the Santa Cruz where they had 

“many dams that backed up water and made marshy 

ground” as far down as Tucson (USDA 1937, p. 11).

Conditions began to change rapidly with the 

buildup of the livestock industry. Harvest of native 

grasses as hay contributed to the decline. Bush muhly 

(Muhlenbergia porteri) was sufficiently abundant that 

in 1879 and 1880 hundreds of tons were delivered as 

hay to military posts in Arizona (USDA 1936). The 

summer of 1885 was unusually dry, and 1886 had half 

the normal rainfall (Hastings 1959). The major portion 

of the desert grassland received between 8 and 14 

inches of precipitation annually with a small segment 

in southeastern Arizona and a portion of southwestern 

New Mexico’s bootheel, receiving approximately 14 to 

20 inches.

Typically, the uplands within the grassland yielded 

runoff in summer thunderstorms, irrigating the alkali 

sacaton (Sporobolus airoides) and tobosa (Pleuraphis

mutica) in floodplains below. Except during major 

thunderstorm events, the desert grassland likely offered only a slight contribution of water to perennial 

streams and rivers. Vegetation along most drainages 

and in the floodplains would have been sufficient to 

reduce peak flows originating in the uplands.

Much of the historic literature addressing the desert 

grassland expresses some concern about the expansion 

of desert shrubs. Pre-Euro-American settlement desert 

grasslands occupied large areas today covered at least 

in part by creosote, burroweed (Isocoma tenuisecta), 

mesquite, and other shrubs undesired by the livestock 

manager.
124 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Many plains and uplands of the Chihuahuan 

Desert in southern New Mexico are in transition from 

a perennial desert grassland to a mesquite dominated 

shrubland. Shrubs encroach into grasslands by 

individual plants becoming established into islands 

of mesquite with a greater biodiversity than the 

surrounding grasslands. The mesquite islands eventually coalesce into mesquite “front” and biodiversity 

decreases. (Beck and others 1999, p. 84).

Similar increases in shrubs were seen 3,900, 2,500, 

and 990 years ago. Shrub increases late in the 19th

century were a natural response to drought. This 

response differed from earlier episodes due to the 

additive factor of massive numbers of livestock (Van 

Devender 1995).

York and Dick-Peddie (1969) investigated survey 

records beginning in 1858 for 31 townships in southern New Mexico and found they had changed from 

approximately 75 percent covered with grass to less 

than 5 percent grass cover at the time of the study in 

the late 1960s. One of their findings yielded a clue on 

the origins of the spread for mesquite in the uplands. 

They found a detailed 1882 map of the vegetation west 

of Las Cruces. Scattered across the mesa were scattered 

pockets of mesquite. Each of those “small locations” 

was found to have a Native American campsite near 

its center. The Native Americans used mesquite extensively for food and later as food for their horses. Other 

sites such as breaks on the edges of watercourses also 

were found to have a mesquite component.

The drier portions of desert grassland in New 

Mexico were apparently less adapted to fire than desert 

grasslands of southeastern Arizona where the higher 

precipitation provided greater fuels and connectivity 

for fire to spread (Bahre 1985). Tobosa and the sacatons are well adapted to fire, but these grasses favor 

topography where moisture accumulates. The typically 

patchy vegetation of drier, upland black grama sites 

would have restricted fire spread. In periods of greater 

than average rainfall, the connectivity of vegetation 

would have expanded, increasing the likelihood of fire 

in black grama dominated communities. Black grama 

does not fare well after fire with rainfall amounts typical 

of most of the desert grassland. A prescribed burn in 

black grama was conducted on the Bernalillo Watershed 

(White and Loftin 2000) near the northern limits of the 

desert grassland in November 1995 and January 1998. 

In May of 2001, marked differences were still evident 

between burned areas previously dominated by black 

grama and adjacent unburned black grama stands. 

Neither area had been more than minimally grazed by 

trespass livestock in several decades. At a landscape 

scale, infrequent and patchy burns in drier portions of 

the desert grassland would have enhanced diversity. 

Desert grassland developed under conditions of lower 

fire frequencies than other Southwestern grasslands. 

Fire events followed by wet years would have little 

lasting impact on community structure, but when 

followed by drought, the result could be a long-term 

change in community structure (McPherson 1995). 

The greatest threats to existing Southwestern desert 

grasslands include uncontrolled grazing, desertification, introduction of exotic and invasive species, and 

urban development (Havsted 1996).

Unique and isolated sites such as Dutchwoman 

Butte, and Research Natural Areas such as Otero 

Mesa, Bernalillo Watershed on the Cibola National 

Forest, and Rabbit Trap on the Gila National Forest, 

would be useful in establishing reference conditions for 

some of the communities within the desert grassland. 

Of greater importance are larger areas such as the 

Appleton-Whittell Research Ranch in southeastern 

Arizona where fire has been maintained. The most 

severely degraded elements of desert grasslands, such 

as the giant sacaton community, are hardest to find in 

suitable condition for use in characterizing reference 

conditions. The extreme variability of desert grasslands 

across its range makes application of any standardized 

reference condition problematic at best.

Great Basin Grasslands_____________

Brown (1982) recognized grasslands above the 

Mogollon Rim in Arizona as having an affinity with 

those in the Great Basin. However, in our (this current) 

assessment they are being referred to as Colorado 

Plateau grasslands. Therefore, in the context of this 

current assessment, the Great Basin grasslands ranged 

southeast to the Rio Grande Valley from just south of 

Albuquerque, north to the Colorado border. They also 

extended at least as far south as the San Augustin 

Plains in New Mexico. Brown (1994) called them Great 

Basin Grasslands to emphasize that influence, but 

believed they were transitional between his Plains 

grassland to the east and the true Basin and Range 

communities to the west and northwest.

Weddell (1996) reported that summer drought was 

the primary factor excluding large herds of bison from 

the steppe of the Intermountain West. As a result, 

there was a lack of significant selective pressure from 

large herbivores, limiting the co evolution between 

ungulates and Intermountain grasses. This left native 

grasslands vulnerable to invasion by exotic grasses 

(particularly cheatgrass, Bromus tectorum) when 

livestock grazing degraded these grasslands. “The 

Western Range” (USDA 1936) cited a Forest Service 

study in western Utah where during the 1931 to 1934 

drought a 20 percent decrease in forage plants occurred 

on ungrazed plots, but the reduction was 60 percent 

on nearby overgrazed areas.

One situation that makes grassland particularly 

vulnerable to degradation is a higher complement 

of cool season grasses, which are more sought after 

by livestock and are less resistant to grazing than 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 125

warm season grasses, coupled with the soils prone 

to erosion across the majority of the Great Basin 

grassland. The complement of factors degrading other 

grasslands—mainly interruption of the fire regime 

and heavy use by livestock during dry years—may 

have caused greater change here than even in the 

desert grassland. Acre for acre, the Great Basin in 

pre-Euro-American settlement condition had greater 

hydrologic function (infiltration) and higher primary 

productivity than the desert grassland.

The spread and increased density of junipers by 

expansion onto deeper, formerly grassland soils has 

long been a concern (Johnsen 1962, Miller and Wigand 

1994, Wright and others 1979), and the condition 

continues today. Even if it became politically acceptable to reintroduce fire to manage juniper (Juniperus

spp.), sagebrush (Artemisia spp.), and other shrubs, the 

break in fire connectivity due to the myriad of incised 

watersheds (gullies) together with the lack of grasses 

productive enough to provide needed fire intensities, 

make restoration efforts challenging.

With the exception of some more mesic portions of 

the Great Basin grassland, defining reference conditions is challenging because of a reduced productivity 

and species diversity. The Rio Grande watershed’s 

Rio Puerco is one of the most infamous and degraded 

watersheds in the Southwest (Dortignac 1960). Portions 

of the Rio Puerco have been grazed at least 100 years 

longer than most of the other Great Basin grasslands. 

Current trends could lead to similar conditions of other 

watersheds within the Great Basin grassland.

Plains Grasslands__________________

The pre-Euro-American settlement Plains 

Grasslands are better understood when placed in 

context of the tallgrass and shortgrass components 

of the midcontinental expanse of grasslands. Dr. John 

Weaver, student of the American Prairie, shared more 

than 40 years of study in his 1954 book North American 

Prairie. He explained:

When the white man came to North America 

magnificent grassland occupied the central part 

of the continent. From Texas it extended north to 

Manitoba, where it gave way to boreal forest. From the 

forest margins of Indiana and Wisconsin it extended 

westward far into the Dakotas and half way across 

Kansas. Early settlers designated this great area 

of waving grasses bedecked with wonderful flowers 

as Prairie. Beyond, even more extensive but drier 

and sparser grassland stretched away to the Rocky 

Mountains. This was early designated as the Great 

Plains. (Weaver 1954, p. 2).

The Prairie and Great Plains were separated according to the height of dominant grasses—in essence, 

their productivity.

Prairie grasses were classified into three groups 

according to the height they attained. Any grass 

that normally attained a height of 5 to 8 feet, or 

more, belonged to the tallgrass group. Big bluestem 

(Andropogon gerardii), switchgrass (Panicum virgatum), and sloughgrass (Beckmannia syzigachne) fell 

into this category. Midgrasses were 2 to 4 feet tall and 

included species such as little bluestem. On dry ridges 

and crests of hills, especially in Western prairie, were 

the shortgrasses, 0.5 to 1.5 feet tall. These grasses 

included blue grama, hairy grama, and buffalo grass. 

Species with “great” drought resistance, such as sideoats grama, blue grama, and buffalo grass (Buchloe 

dactyloides), or those adapted to evade drought, such 

as western wheatgrass (Pascopyrum smithii), usually 

occurred in most of the prairie only in small amounts 

because they were not normally able to compete with 

tallgrasses. These grasses increased rapidly and occupied large areas in intense drought (Weaver 1954).

Dr. F.W. Albertson described the typical situation 

in the mixed-grass prairie just before the turn of the 

20th century.

The vast majority of the land was native prairie. It 

was neither broken for cultivation nor overgrazed by 

livestock. Short grasses and low growing broadleafed 

herbaceous plants occupied the hilltops. Many of the 

hills were dotted with bunches of little bluestem and 

in the favored areas such as buffalo wallows, side oats 

grama, and big bluestem were common. The hillsides 

were occupied primarily by big and little bluestem, 

side oats grama, Indian grass (Dichanthelium acuminatum) and panic grass (Sorghastrum nutans). 

All but the little bluestem and side oats grama were 

dominant on the lowlands” (Albertson 1949, p. 10).

Dr. Albertson also noted that the dust storms of 

severe droughts in the 20th century, magnified by 

cultivating lands that could not support sustained 

agriculture, also occurred as natural events in severe 

droughts before 1900.

The higher productivity of the mixed grass prairie 

brought with it more frequent and higher intensity 

fires than fires in the shortgrass prairie. In 1924, 

Shantz noted tallgrass prairies were often burned in 

late summer or winter, and that early settlers and 

travelers could find safety only by starting backfires. 

The flames were impossible to pass through to the 

safety of the burned areas behind (Weaver 1954). In 

1825, Joseph C. Brown was surveyor in a party marking what was to become the Santa Fe Trail. Brown 

noted that Cow or Cold Water Creek, near the Great 

Bend on the Arkansas River in Kansas, marked the 

beginning of the shortgrass prairie, and the shortgrass 

was the boundary of the annual burning of the prairie 

(Hollon 1961).

Risser and others (1981) included as mixed grass 

prairie, the strip of sandy soils with Havard shin oak 

(Quercus havardii) and midgrasses running across 

the Texas Panhandle into eastern New Mexico about 

Tucumcari. This community is also common along the 

southeastern border of New Mexico. In 1844, Josiah 

Gregg described a portion of the community as a 
126 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

region covered with sandy hillocks entirely barren of 

vegetation in places while others were covered with 

shin oak and plum. In 1845, Lt. J. W. Albert found the 

community to be a continuous succession of sandhills 

with sand sagebrush and Havard shin oak. Big bluestem, sand bluestem, Indiangrass, switchgrass, giant 

dropseed (Sporobolus spp.), plains bristlegrass (Setaria 

leucopila), cane bluestem (Bothrichloa barbinodis), and 

sideoats grama are among the grasses. The range of this 

community has not increased since the mid-19th century 

(Peterson and Boyd 1988). However, the composition 

appears to have less of the component of taller grasses 

and more sand sagebrush and mesquite than before 

the livestock era. The oak component grows slowly and 

is long lived. It is estimated to have made up only 5 to 

25 percent of the prelivestock era community. While 

in some areas the oak component has not expanded 

onto former grasslands (Peterson and Boyd 1988), in 

other areas at least the individual oak motts now occupy what was once a mixed grass prairie community. 

Both midgrass and tallgrass components and oak are 

well adapted to most fire regime variations (Peterson 

and Boyd 1988) while mesquite is not. Simpson (1977) 

contended the range of mesquite had not spread, just 

the densities within the range. Mesquite densities, 

within the strip identified by Risser and others (1981) 

as mixed-grass, continue to increase, but others found 

mesquite present in the community in 1845 (Abert 

and Carroll 1999).

Even though more acreage remains for the mixed 

grass than the tallgrass prairie, the patchiness of the 

grasslands means the ecosystem operates on a much 

smaller scale than in the prelivestock era. Fire is 

implemented as a management activity by individual 

landowners rather than occurring at landscape scales 

(Weaver and others 1996). Herbivory by livestock has 

different impacts on prairie forbs and grasses than by 

buffalo. Even where a high diversity of forbs remains, 

fragmentation favors a continued reduction in associated animal species. Homogenization of large or small 

pastures by livestock reduces the inherent ability of the 

ecosystem to withstand drought. Intense development 

of stock tanks across native prairie pastures interrupts 

the functioning of watersheds, limiting free flow of 

water with a coincident loss of biological diversity. 

Interruption of the natural fire regime likely had as 

much to do with expansion of oak and sand sagebrush 

in Oklahoma and mesquite in Texas and New Mexico 

as did overgrazing by livestock.

Establishing reference conditions for the mixedgrass prairie is even more difficult than for the other 

grasslands because of the diversity needed to provide 

the ecosystem’s response to climatic variability. 

Fragmentation limits the value of potential reference 

sites, excluding smaller areas, even those in excellent 

condition. Allowances must be made for both fire, and 

the suite of short-, mid-, and tallgrasses necessary to 

withstand drought events. The buffalo, hugely reduced 

in number, is a missing keystone link across most of 

the mixed-grass prairie.

Colorado Plateau Grasslands________

Castetter (1956) lumped the shortgrass prairie, or 

steppe, and the Great Basin grassland into a mixed 

prairie “association” because the grasses were also 

those of the true steppe and occurred as a “climax 

composition” in more or less equal amounts of mid- and 

shortgrasses. Throughout the association, blue grama 

was felt to be the climax dominant. Sideoats grama was 

an important component as was hairy grama. Mostly in 

lower spots topographically, western wheatgrass was 

“rather common” in association with blue grama. On 

sandy soils, little bluestem was a common associate 

often found with sand bluestem and Indian grass. In 

low saline areas, alkali sacaton stands were common. 

On the elevated plains, the cane cholla (Opuntia spp.) 

and soapweed (Yucca elata) were uncommon due to 

grass competition.

Blue grama is the common species among 

Southwestern grasslands, being even more abundant 

today in degraded conditions of some grasslands than 

in the prelivestock era. The resistance of blue grama 

to drought and heavy grazing played a major role in 

the sustainability of the shortgrass prairie. Buffalo 

benefit from this survival mechanism because the 

grazed lawn has a denser concentration of younger, 

higher quality forage. Having many small shoots also 

facilitates rapid response to ephemeral water availability limiting loss to drought and speeding recovery 

after defoliation (Coughenour 1985). Blue grama is 

one of the grasses best adapted to periodic water 

stress. Blue grama has a high root/shoot ratio because 

a high proportion of carbohydrates are translocated 

below ground instead of being used to grow shoots. 

This increases survivability and keeps carbohydrate 

resources higher than in less drought tolerant grasses 

(Detling 1979). Blue grama was the ideal grass on 

the flat expanses of shortgrass steppe where winds 

limited snow deposition and increased desiccation in 

the spring, and where growing season precipitation 

was scattered in time and space, requiring frequent 

dormancy and regreening in a single growing season. 

Blue grama was the ideal grass in higher use areas 

where herbivores large and small would have eaten 

and trampled a less-adaptive species to oblivion.

Fire in the Colorado Plateau grasslands was of low 

intensity but adequate to keep woody shrubs from 

expanding. At the edge of sandier soils, wildfire would 

have restricted sand sagebrush. Fires could reduce 

pricklypear (Opuntia spp.) by either killing the plant 

in a hotter fire or by burning the spines, making the 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 127

Arizona vistas: Contrasts of exurban development, near Flagstaff (above), with undeveloped land, farther 

south near Naco and the U.S./Mexican border (below). (Photos by John Yazzie)
128 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

cactus more available to a wide range of herbivores. 

Abert and Carroll (1999) noted the short flame lengths 

of the typical wildfire, saying it was not difficult to 

pass through the flame front unharmed.

The prickly pear was widespread and likely increased during drought periods just as it does today. 

Pricklypear can withstand drought better than most 

steppe plants, and its shallow root system utilizes 

moisture from even light rainfalls (Branson 1985). 

Pricklypear increases in overgrazed areas, and it can 

increase during drought even in grasslands protected 

from livestock.

The best sites for developing reference conditions 

in the Colorado Plateau grasslands are on large private ranches because they are less patchy than areas 

administered by the Federal government. A variety 

of grazing intensities would best depict the variation 

found in the pre-Euro-American settlement era.

References________________________

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Allen, C.D. 1984. Mountain grasslands in the landscape of the Jemez 

Mountains, New Mexico. Unpublished M.S. thesis.

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Bahre, C.J. 1985. Wildfire in southeast Arizona between 1859 and 

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Baker, W.L. 1983. Alpine vegetation of Wheeler Peak, New Mexico, 

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Branson, F.A. 1985. Vegetation Changes on Western Rangelands. 

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Brown, D.E. 1982. Plains and Great Basin grasslands. Desert 

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herbivores: adaptations, expectations, and interacting processes. 

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Dick-Peddie, W.A. 1993. New Mexico Vegetation, Past, Present and 

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Bulletin No. 4.

Grissino-Mayer, H.D. 1996. A 2129-year reconstruction of precipitation for northwestern New Mexico, USA. Radiocarbon 1996: 

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Hollon, W.E. 1961. The Southwest Old and New. Lincoln: University 

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Along the Mexican Border. Albuquerque: University of New 

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grasslands. Ecological Monographs 32(3): 187-207.

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San Francisco Mountains Forest Reserve, Arizona. Professional 

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130-151. In: M.P. McClaran and T.R. Van Devender (eds.), The 

Desert Grassland. Tucson: University of Arizona Press.

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pinyon-juniper woodlands. Bioscience 44(7): 465-474.

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Wiens. 1981. The True Prairie Ecosystem. Stroudsburg, PA: 

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describing the early conditions regarding groundcover, flood flows, 

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(eds.), Prairie Conservation. Washington, DC: Island Press.

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disturbance regimes in steppe and desert communities. Pp 3-12. 

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to cool-season prescribed fire. Journal of Range Management. 

53(1): 52-61.

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and use of fire in sagebrush-grass and pinyon-juniper plant 

communities, a state-of-the-art review. Gen. Tech. Rep. INT-58. 

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130 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Introduction_______________________

In this chapter we discuss grassland sustainability 

in the Southwest, grassland management for sustainability, national and local criteria and indicators of 

sustainable grassland ecosystems, and monitoring for 

sustainability at various scales. Ecological sustainability is defined as:

[T]he maintenance or restoration of the composition, structure, and processes of ecosystems over time 

and space. This includes the diversity of plant and 

animal communities, and the productive capacity of 

ecological systems and species diversity, ecosystem 

diversity, disturbance processes, soil productivity, 

water quality and quantity, and air quality. (USDA 

Forest Service 2000: Glossary)

Sustainability is measured over various spatial 

scales, often including a nested hierarchy of smaller 

and larger spatial scales. For example, if the habitat 

of a rare species was studied at only a fine spatial 

scale, the impact of disturbances and larger landscape 

patterns that affect species distribution and viability 

would not be considered.

Ecosystem integrity has been variously defined and 

incorporates the concepts of ecosystem functioning 

and resilience. Grumbine (1994) described five goals 

of ecosystem integrity:

• Maintaining viable populations (biodiversity)

• Ecosystem representation

• Maintaining ecological processes

• Protecting evolutionary potential

• Accommodating human use

Definitions and measures of integrity are discussed 

by De Leo and Levin (1997).

Nearly 75 percent of all threatened ecosystems in 

the United States are either grasslands or shrublands 

(Mitchell and Joyce 2000). Grasslands are home to 

more than 7,500 plant and animal species in the 

United States. Many grassland species are now either 

threatened or endangered, and more than 700 species are candidates for listing under the Endangered 

Species Act.

The majority of Southwestern rangelands are grasslands. Rangelands are defined as those areas where the 

potential natural vegetation predominately comprises 

grasses, grasslike plants, forbs, and shrubs, and where 

herbivory is an important ecological process (Anderson 

and others 1976, Frank and others 1998, Mitchell and 

Joyce 2000). Rangelands affect the quality of life of 

every person in the United States. This land accounts 

for approximately 706 million acres, or 40 percent of 

the lands in the United States, including grasslands, 

shrublands, tundra, alpine meadows, Southwestern 

deserts, and wetlands across the country (Colorado 

State University 2001). The U.S. Federal government 

manages over 21 million acres of prairie grasslands, 

including the shortgrass prairie of eastern New Mexico 

(National Wildlife Federation 2001).

Chapter 7:

Grassland Sustainability

Deborah U. Potter

Paulette L. Ford
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 131

We discuss grassland and rangeland management in 

terms of maintaining ecosystem processes, ecological 

integrity, and ecological sustainability. Sustainable 

rangeland management is defined as management of 

rangeland ecosystems to provide a desired mix of benefits to the present generation without compromising 

their ability to provide benefits for future generations 

(Colorado State University 2001). Managing grassland 

resources, including rangelands, for sustainability 

will help ensure that USDA Forest Service meets its 

stewardship responsibility of passing the nation’s resources on to future generations in improved condition 

(Kaufmann and others 1994).

Without an effective way to accurately monitor 

social, ecological, and economic aspects of rangeland, 

and therefore grassland sustainability, it is difficult to 

measure progress toward sustainability. In recent years 

Federal land management agencies have been criticized for a lack of consistent, standardized indicators 

for reporting the status of rangelands. In response to 

this need, local, national, and international criteria for 

grassland sustainability are being developed. Efforts 

include the Sustainable Rangelands Roundtable (SRR), 

the Local Unit Criteria and Indicator Development 

(LUCID) project, and the Montreal Process. The purpose of grassland monitoring can vary, but an important 

aspect is to determine whether management activities 

have affected ecosystem sustainability and integrity.

Grassland Management for 

Sustainability______________________

Southwestern grasslands are managed for a 

variety of uses, including livestock grazing, wildlife 

habitat, protecting water quality, and for recreation 

(National Research Council 1994, USDA Forest 

Service Kiowa and Rita Blanca National Grasslands 

2001). Grasslands in the West and Southwest have 

been degraded by overgrazing and intensive use of 

riparian areas (USDA Forest Service Research and 

Development 2001). Forest Service research and 

management priorities have addressed stream bank 

erosion, sedimentation and erosion rates, restoring 

riparian vegetation, providing improved habitat for 

native species, and the cumulative impact of wildlife 

and livestock grazing.

Management actions must be consistent with 

Federal and State regulations, including the Clean 

Water Act, the Multiple Use and Sustained Yield Act of 

1960, the Resources Planning Act of 1974, the National 

Forest Management Act of 1976, the Federal Land 

Policy and Management Act of 1976, and the Soil and 

Water Resources Conservation Act of 1977.

An important principle of all grassland management is to plan for drought conditions. Hydrological 

extremes are associated with climatological events, 

especially the El Niño Southern Oscillation (ENSO). 

The Southwestern United States has a strong teleconnection to the ENSO cycle. In general, El Niño brings 

periods of heavier winter precipitation in areas that 

are affected, while La Niña events are associated with 

drought. Depending on location, December through 

March precipitation in New Mexico is 138 to 214 

percent that of average conditions in response to El 

Niño events (see National Weather Service map, NOAA 

2001). February through April temperatures are lower 

in response to an El Niño event.

For management purposes, the frequency and severity of drought are more important than long-term 

average climate conditions. Too often land managers 

plan for average climate conditions, rather than the 

climatic extremes that can be expected. Drought conditions and low spring runoff occur in response to La 

Niña events at New Mexico sites (Dahm and Molles 

1992, Molles and Dahm 1990, Molles and others 1992), 

while in response to an El Niño, winter/spring precipitation nearly equals summer precipitation. Drought 

conditions, including response to La Niña events, also 

correspond to an increase in fire frequency that is detectable in histories reconstructed from fire-scar data.

Grazing intensity can be reduced to prevent excessive vegetation and ecosystem damage during drought. 

In addition, grassbanks are being used in northern 

New Mexico and other areas as one part of a wider 

solution to address drought conditions. Grassbanks 

are rangeland management systems that provide 

alternative forage for livestock displaced from their 

regularly permitted allotments, which are undergoing 

restoration (USDA Forest Service Manual, Interim 

Directive 2001). Grassbanks can be useful during 

drought conditions to alleviate the stress on rangelands 

that do not have the capacity for higher grazing intensity due to climatic conditions. See chapter 8, “Tools 

for Grassland Management,” for further discussion of 

grazing management.

Adaptive Ecosystem Management____

Forests in the Southwestern Region utilize adaptive 

management, including management of grassland 

ecosystems (Kaufmann and others 1994). Adaptive 

management acknowledges the complexity of ecosystems and the uncertainty involved in predicting 

ecosystem responses to management actions. It 

ensures that land managers incrementally assess 

whether goals are met and whether interim results 

are acceptable. This means that decisions affecting 

ecosystem sustainability are incrementally reevaluated as monitoring data are collected, summarized, 

and evaluated. Therefore, adaptive management 

establishes a pathway to alter the course of management action as new knowledge is acquired.
132 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

For example, improvements in grazing management 

can result in reduced erosion and soil compaction, 

promote increased infiltration of precipitation into the 

soil (Roberson 1996), and increase biodiversity. When 

monitoring data show that resource conditions are 

below thresholds for ecological factors, then grazing 

management can be altered immediately or prior to 

the next grazing season. Steps can be taken to prevent 

overutilization of forage, provide rest to disturbed 

areas, alter the season of use, and improve overall 

management of livestock. Restoration programs to 

increase soil and streambank stability or to revegetate 

denuded areas are also part of adaptive management. 

To be effective, plans for adaptive management could 

include specific actions in response to observed or 

measured conditions, a timeline, and specific thresholds that trigger the management actions (Roberson 

1996).

Criteria and Indicators for Sustainable 

Grassland Ecosystems______________

Developing indicators and monitoring them over 

time can help to determine whether problems are 

emerging, whether any action is desirable or necessary, what action might yield the best results, and 

how successful past actions have been. To develop 

and implement sound environmental policies, data 

are needed that capture the essence of the dynamics of environmental systems and changes in their 

functioning. These kinds of data then need to be 

incorporated into indicators (National Research 

Council 2000).

A criterion is a category of conditions or processes 

that is an explicit goal of sustainable development or 

by which sustainable development can be assessed. 

A criterion is too general in scope to monitor directly 

but can be characterized by a set of indicators that 

can be monitored over time. An indicator is a variable that can be assessed in relation to a criterion. 

It should describe attributes of the criterion in as 

objective, verifiable, and unambiguous manner as 

practicable, and it should be capable of being estimated 

periodically in order to detect trends (Colorado State 

University 2001). Indicators are designed to inform 

us quickly and easily about something of interest. 

They communicate information about conditions and, 

over time, about changes and trends. Like economic 

indicators, environmental indicators are needed 

A grassbank has been established in Chihuahuan Desert grassland on the Diamond A Ranch (Eastern Division) in southwestern 

New Mexico. (Photo by Charles Curtin)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 133

because it is not possible to measure everything 

(National Research Council 2000). For example, a 

criterion on the conservation of biological diversity 

includes indicators related to ecosystem diversity, 

species diversity, and genetic diversity (Colorado 

State University 2001).

National Indicators of Sustainability__

One effort for identifying criteria and indicators 

(C&I) for the sustainable management of temperate 

and boreal forests at a national scale, the Montreal 

Process, has become widely recognized. Moreover, 

the concept of using C&I as factors for evaluating all 

facets of sustainability, including resource supplies, is 

receiving increasing acceptance (Corson 1996, Mitchell 

2000). In 1995, the United States agreed to use the 

Montreal Process Criteria and Indicators to measure 

national progress in achieving the goals of sustainable 

forest management. This in turn generated the need for 

sustainability C&I for grass and shrubland ecosystems, 

as well as for energy and minerals (USDA Forest Service 

Inventory and Monitoring Institute 2001). There is a 

need for consistent, national baseline information to 

provide a common language and standards for assessment and planning that will lead to proper and effective 

decisionmaking. A comprehensive set of criteria and 

indicators should provide this tool (Colorado State 

University 2001).

The Montreal Process lists seven criteria and 67 

indicators for the conservation and sustainable management of temperate and boreal forests. The C&I 

encompass a set of interrelated ecological, economic, 

social, and institutional factors. The seven criteria 

are:

• Conservation of biological diversity.

• Maintenance of productive capacity of forest 

ecosystems.

• Maintenance of forest ecosystem health and 

vitality.

• Conservation and maintenance of soil and water 

resources.

• Maintenance of forest contribution to global 

carbon cycles.

• Maintenance and enhancement of long-term 

multiple socio-economic benefits to meet the 

needs of societies.

• Legal, institutional, and economic framework 

for forest conservation and sustainable management.

The first five of the seven criteria, along with 28 

indicators, have been used to assess the applicability 

of Montreal Process biological and abiotic indicators to 

rangeland sustainability at a national scale (Mitchell 

and Joyce 2000). See The International Journal of

Sustainable Development and World Ecology, volume 7 

(2), June 2000, for a detailed analysis of the challenge 

and promise of developing C&I for rangelands, and 

for implementation at the national scale. So far 16 

indicators have tentatively been identified for rangelands, including indicators for landscape diversity, 

community diversity, and population diversity. Genetic 

diversity has been particularly difficult to define due 

to a lack of baseline data. Some of the indicators lack 

standardized protocols. More work needs to be done 

on refining definitions, designing monitoring systems, 

and testing critical assumptions.

Other efforts to develop national sustainability 

indicators for rangelands include the formation of 

the Sustainable Rangelands Roundtable, which is 

sponsored by Colorado State University, the USDA 

Forest Service, USDI Bureau of Land Management, 

and USDA Agricultural Research Service. The roundtable includes representatives from nongovernmental 

organizations, public and private land management 

professionals, rangeland scientists, and university 

professionals (Colorado State University 2001). The 

roundtable will identify indicators of sustainability 

based on social, economic, and ecological factors, 

in the effort to provide a framework for national 

assessments of rangelands and rangeland use. For 

more information, access http://sustainablerangelan

ds.cnr.colostate.edu/Roundtable_description.htm.

Local Indicators of Sustainability_____

Criteria and indicators for sustainability are 

being developed for specific scales. The intended 

scale for the Montreal Process was for all forests 

of a country, regardless of land ownership. The 

Local Unit Criteria and Indicator Development 

Project (LUCID) was initiated by the Inventory and 

Monitoring Institute to test criteria and indicators 

of sustainability at local levels. LUCID targets a 

local scale, such as the Tongass National Forest 

in Alaska or the Blue Mountain Province National 

Forests in Oregon. The local criteria and indicators 

should be revised periodically to incorporate new 

research results, technological advances, and new 

methods of measurement.

LUCID sites include forest, rangeland, and 

shrubland ecosystems. This process was developed 

to identify conditions that are needed to sustain 

ecological, economic, and social systems, and to determine the criteria and indicators for assessing how 

resource management influences sustainability. The 

Blue Mountains LUCID pilot test addressed C&I with 

verifiers (measurement protocols) and standards for 

shrublands and grasslands. The three Blue Mountains 

forests located in Oregon and Washington use the same 

ecological definitions, and the C&I are intended to 

apply to the entire province. Although it is too soon to 
134 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

draw conclusions from C&I testing at Blue Mountains, 

future reports will address water quality, forest health, 

and community growth.

Here is an example of a local grassland criteria 

and indicator that is based on the LUCID criteria 

and indicators for maintenance of ecosystem integrity. The degree of fragmentation is an indicator of 

landscape structure (USDA Forest Service Inventory 

and Monitoring Institute 2001). Accelerated habitat 

fragmentation can be a detriment to species viability 

and adversely affect ecological patterns and processes 

such as disturbance (Forman 1995). One measure of 

fragmentation is the ratio of edge to interior habitat, 

with higher ratios indicating increasing fragmentation. Corresponding landscape metrics based on the 

LUCID criteria and indicators are shown in figure 

7-1.

questions should focus on key ecological processes 

and interactions, rather than individual parts of 

the system.

Conversely, inventory involves gathering data that 

are needed to analyze and evaluate the status or condition of resources (Powell 2000). Examples include the 

Forest Service programs entitled Terrestrial Ecological 

Unit Inventory (TEUI), Biological Forest Resource 

Inventory, Rangeland Resource Inventory, Wildlife 

Habitat Inventory, Threatened and Endangered 

Species Habitat Inventory, and the Human Dimension 

Heritage Inventory. Corporate strategies, goals and 

approaches for integrated natural resource inventories 

were developed for national (strategic), forest (tactical), and project (operational) scales (USDA Forest 

Service 1999).

The Forest Service recognizes three types of 

monitoring related to Land Management Planning: 

implementation, effectiveness, and validation monitoring. Implementation monitoring tracks compliance 

with Forest Plan standards and guidelines and helps 

determine whether planned management activities 

were completed. Effectiveness monitoring helps determine whether desired outcomes were achieved 

by management actions. Validation monitoring tests 

the assumptions and models of Forest Plan implementation. In addition, monitoring associated with 

applied research can supply critical information for 

decisionmaking. The scope of ecological monitoring

needs to include programmatic monitoring that tracks 

and evaluates trends of ecological, social, or economic 

outcomes (Powell 2000).

Despite the development of protocols and indicators 

for sustainability, the criteria of sustainable grasslands 

are not easily measured. Some helpful references for 

grassland monitoring include Elzinga and others 

(1998), Gibbs and others (1998), and Munn (1988). 

Examples of existing grassland monitoring programs 

can be found in Wondzell and Ludwig (1995). A successful monitoring program for grasslands could 

include the following considerations.

• A description of each step of the monitoring 

system and the quality assurance/quality control 

protocol (Bormann and others1994, Everett and 

others 1994, Moir and Block 2001).

• The statistical parameters to be measured.

• Trend detection.

• A funding commitment from management, 

including maintenance of field equipment or 

repeated remote-sensing applications.

• Early determination of who monitors and a 

commitment of adequate human resources to 

accomplish the tasks over time.

• Public stakeholders’ involvement in the monitoring program (Cortner and Moore 1999).

Maintenance of Ecosystem Integrity

Criteria: landscape structure

Indicator: fragmentation and connectedness

Verifier: ratio of edge to interior habitat area

Figure 7-1. Criteria, indicator, and verifier for fragmentation: a 

sample metric for monitoring sustainability at landscape scale 

that is based on the LUCID criteria and indicators for maintenance of ecosystem integrity (USDA Forest Service Inventory 

and Monitoring Institute 2001). An edge is the boundary between 

habitats that protects the core. Higher ratios of edge to interior 

habitat area indicate increasing landscape fragmentation (Forman 1995, Prendergast 2000). Note that verifiers should have 

specific target values.

Edge 

Interior

Monitoring for Sustainability in 

Grasslands________________________

Monitoring is a step-wise process that involves: 

framing a question(s) and developing a study plan 

to address the question(s) using a standard protocol; 

collecting data according to the monitoring plan; 

storing the data for retrieval, and evaluating the 

results. Monitoring should include goals, thresholds 

for change, and remedial actions that occur when 

thresholds are met or exceeded. An ecological systems 

approach to monitoring ensures a strong foundation 

in ecological theory, adequate consideration and understanding of cause-and-effect relationships, and a 

systematic approach to select and evaluate parameters 

that are monitored (Sieg 1999). Furthermore, the 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 135

Monitoring for grassland sustainability should address current practices for managing grazing animals 

and grazing herds, especially conservative stocking 

rates. A summary of grazing systems is beyond the 

scope of this chapter. However, an analysis of the 

importance of grazing intensity versus grazing system 

is provided by Holechek and others (1999).

Forest Monitoring Reports__________

According to Land and Resource Management Plan 

monitoring reports for U.S. Federal government fiscal 

years 1998 and 1999, most National Forest monitoring 

in the Southwestern Region was implementation monitoring—that is, 65 percent. (Data are only available 

through the USDA Forest Service Intranet site.) About 

32 percent of the reported monitoring was evaluation 

monitoring, and the remaining 3 percent was validation 

monitoring. The information is not detailed enough 

to know whether grassland monitoring followed this 

general trend.

Study Design______________________

Successful monitoring programs should have the 

following characteristics (Powell 2000, 2001):

• Be purposeful and conducted to answer specific 

questions.

• Be done at the appropriate spatial and temporal 

scale to answer the question.

• Be done in collaboration with others (for example, 

agencies, interested publics, researchers, and 

nongovernmental organizations). Collaboration 

results in sharing the workload (including obtaining data from other sources), gaining expertise, 

and building credibility and trust.

• Use the best available science and established 

protocols to collect and evaluate the data.

• Use modern information management techniques 

and tools.

• Apply stringent selection criteria so that a monitoring activity is only conducted if it is feasible, 

realistic, and affordable.

• Emphasize evaluation as much as the collection 

of the data.

The established protocol identified in a monitoring plan should include standard sampling and 

analytical methods that determine the precision and 

accuracy of measurements. These are the procedures 

for quality assurance and quality control. Accuracy 

refers to the degree of agreement between a calculated or measured quantity and the true value 

of the parameter. Precision refers to the degree 

of agreement between replicate measurements of 

the same parameter, and includes the concepts of 

duplicability, repeatability, and reproducibility. For 

example, in the USDA Forest Service Global Change

Research Program, quality management establishes 

programwide policies and procedures that ensure 

adequate documentation and data quality for all 

field, analytical, and modeling activities. Quality 

assurance implements these policies by establishing and monitoring quality control (QC) procedures 

including the identification of variability and followup 

control recommendations to improve the accuracy 

and precision of measurements. QC procedures are 

implemented by scientists within each project and are 

designed to produce a sustained reduction of error 

and document systematic error within statistically 

defined limits (USDA Forest Service Global Change 

Research Program 2001).

Proper training and supervision of field and 

laboratory staff is necessary to ensure adherence 

to the protocol and the success of the monitoring 

program.

Scale_____________________________

The temporal and spatial scales chosen depend 

on the question or problem that monitoring will address. Spatial scales for assessing sustainability can 

be viewed, for example, as ecological units (Bailey 

1995); a nested watershed hierarchy; or other appropriate units such as community, ecosystem, and 

landscape scales. Temporal scales can range from 

the time it takes bacteria to reproduce or molecular 

processes to occur through evolutionary and geological time scales.

It is necessary to conduct ecological studies at 

the appropriate temporal and spatial scales because 

changes occur at many scales at the same time, 

and different processes are likely to be important 

at different scales. Depending on the scale chosen, 

two species can appear to be highly interrelated or 

completely independent. For example, Levin (1992), 

when considering the reasons for shrimp distribution in the ocean, found that krill distribution at 

small spatial scale was a function of the behavior of 

individuals. However, at large scale, krill distribution 

was a function of oceanographic processes. Thus, the 

conclusions about the factors that affect the distribution of these organisms varied according to the scale 

where measurements were made.

Processes and functions that are appropriate to 

monitor at watershed scale include succession, biogeochemical cycles, energy transfer through the food 

web, disturbance, and competition.

Because studies at fine spatial resolution have 

greater detail, the study results can detect heterogeneity and other ecological patterns. However, those 

same study results also have a low potential for 
136 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Table 7-1. Criteria and indicator linkage between Montreal Process and Local Unit Criteria and Indicator Development 

(LUCID). Reproduced from LUCID homepage, http://www.fs.fed.us/institute/lucid/.

MONTREAL PROCESS CRITERIA (1-7)

 LUCID Indicators

1. CONSERVATION OF BIOLOGICAL DIVERSITY

I.2.2.1 Vegetation types and structural classes

I.2.4.5 Ecologically sensitive areas, e.g., riparian areas are retained

I.2.5.1 Populations of indigenous species

I.2.6.1 Exotic species

I.2.6.2 Community guild structure

I.2.6.3 Species at risk

I.2.7.1 Gene frequencies change

2. MAINTENANCE OF PRODUCTIVE CAPACITY OF FOREST ECOSYSTEMS

I.2.2.3 Fragmentation and connectedness

I.2.2.2 Linear features

I.2.3.2 Primary productivity

I.3.1.2 Land base available for production

3. MAINTENANCE OF FOREST ECOSYSTEM HEALTH AND VITALITY

I.2.1.2 Disturbance processes

4.CONSERVATION AND MAINTENANCE OF SOIL AND WATER RESOURCES

I.2.1.1 Hydrologic condition

I.2.4.1 Pollutants

I.2.4.2 Soil quality e.g., soil compaction, displacement, erosion, puddling, loss of organic material

I.2.4.3 Soil nutrients

I.2.2.5 Water quality e.g., dissolved oxygen, suspended sediments and water nutrients

I.2.4.6 Morphology and function of stream channels

5. MAINTENANCE OF FOREST CONTRIBUTION TO GLOBAL CARBON CYCLES

I.2.3.1 Nutrient cycling

I.2.4.4 Ecological legacies and structural elements

6. MAINTENANCE AND ENHANCEMENT OF LONG-TERM MULTIPLE SOCIO-ECONOMIC BENEFITS TO MEET THE 

NEEDS OF SOCIETIES

1.1.1 Wilderness

1.1.2 Aboriginal and non-Aboriginal cultural, spiritual, social sites/values

1.2.1 Scenery

1.3.1 Recreational, tourism and education opportunities (by activity)

1.4.1 Access to forest resources

1.6.1 Worker health and safety

1.6.2 Public health and safety

1.7.1 Subsistence and non-subsistence gathering

I.3.1.1 Community economic trade balance (imports and exports)

I.3.2.1 Annual and periodic removals of products (timber and non-timber)

I.3.2.3 Money spent by visitors in local communities (by activity)

I.3.2.4 Value to products including value-added through downstream processing

I.3.2.5 Resource production component of economy

I.3.2.6 Income from National Forest activities

I.3.2.7 Employment of local population in resource management

I.3.3.1 Rent capture

I.3.3.3 Community economic diversity

7. LEGAL, INSTITUTIONAL AND ECONOMIC FRAMEWORK FOR FOREST CONSERVATION AND SUSTAINABLE 

MANAGEMENT

1.4.2 Ownership and use rights

1.5.1 Participation/involvement in decision-making

I.3.3.2 Mechanisms for economic benefits sharing
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 137

making generalizations. Studies at broad scale show 

the dynamics of a system, detect slower rates of process 

or system change, and have a high potential to derive 

generalizations. Studies at multiple or nested scales 

can be used to aggregate and extrapolate fine-scale 

results to larger scales. Examples of how to identify 

questions at the appropriate scale are provided by 

Powell (2001).

Criteria and Indicators for 

Monitoring________________________

For each question that is raised, criteria and indicators to answer that question should be identified and 

monitored. Because an essential goal of ecosystem 

management is ecosystem sustainability (Funston 

1995, Rauscher 1999), questions associated with 

ecological monitoring might address whether ecosystem conditions, or management activities affecting 

ecosystem conditions, are sustainable. The international Montreal Process was developed to identify 

criteria and indicators for forest sustainability at a 

national scale. Subsequently, the applicability of the 

criteria and indicators from the Montreal Process to 

grasslands has been assessed (Flather and Sieg 2000, 

Joyce 2000, Joyce and others 2000, McArthur and 

others 2000, Neary and others 2000). The Local Unit 

Criteria and Indicator Development (LUCID) Project 

addressed sustainability at a forest management scale. 

Monitoring for these criteria and indicators will help 

determine whether ecological conditions are sustainable (table 7-1).

Tools_____________________________

There is no standard “tool kit” for ecological 

monitoring in grasslands. The scientific methods and 

processes used to answer ecological questions are the 

monitoring tools. The methods and processes chosen 

will depend on the question(s), the ecological indicators that are selected, temporal and spatial scales 

of concern, degree of detail, precision and accuracy 

that is required, and financial budget. Therefore, 

the concept of monitoring tools is a broad one that 

can include:

• Methods for framing your questions and multiple 

working hypotheses.

• Instrumentation for quantitative or qualitative 

measurements.

• Analytical tools to detect patterns and change 

across time and space (such as GIS, remote sensing, fractal analysis).

• Other methods for data analysis such as statistics, conceptual and mathematical models, 

fuzzy logic.

• Decision-support systems or other expert systems.

Vegetation sampling methods for rangelands are 

discussed by Stohlgren and others (1998). Other 

grassland monitoring methods are discussed in 

Jacobsen and others (1998).

Evaluating and Interpreting Results__

Much of the Forest Service monitoring data are being 

stored in a corporate database, the Natural Resource 

Information System. These data require scientific 

evaluation considering current scientific knowledge 

and research results. One step in evaluating monitoring data includes a comparison to critical values for 

“keystone” indicators of change. When critical values 

are approached or exceeded, then further actions are 

likely warranted to ensure sustainability. For example, 

the Air Resource Management Program has identified 

wilderness values that can be affected by air pollution, 

and their associated sensitive receptors and concern 

thresholds. These indicators and thresholds reflect 

pollution transport and interactions between the 

atmosphere, geosphere, and biosphere, and these 

are contained in the Natural Resource Information 

System, NRIS-Air module. They provide management 

guidelines for protecting Air Quality Related Values 

in Class I Wilderness Areas.

Evaluation also includes summarizing data in statistical and graphical formats and identifying changes 

or trends. A variety of methods exist for detecting 

ecosystem changes and trends including statistical 

methods (analysis of variance, cluster analysis, and 

so forth), comparisons of current conditions with past 

reconstructions (dendrochronology, paleolimnology, 

pack-rat middens, and so forth), modeling, and spatial 

analysis using GIS, satellite imagery, and remote sensing. A common approach for trend detection is analyzing 

data from repeat sampling over time, including data 

from long-term monitoring networks. In some cases, 

a qualitative assessment and professional judgment 

may be adequate evaluation tools.

Condition Classes__________________

One way to assess and categorize monitoring data 

is according to condition classes. A good example is the 

Soil Condition Rating Guide to determine whether soil 

quality objectives are met (table 7-2). The guide incorporates multiple indicators for three soil ecosystem 

functions—hydrologic function, stability, and nutrient 

cycling—into ratings of either satisfactory, impaired, 

or unsatisfactory.

Other monitoring systems that use multiple indicators and classification include Thalweg-Watershed 
138 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Table 7-2. Soil Condition Rating Criteria. Reproduced from FSH 2509.19, Soil Management Handbook, R3 supplement 2509.18-99-1, effective 10/20/1999. Also see Doran and others (1994).

SOIL CONDITION RATING GUIDE

CONDITION CATEGORY

Indicator Satisfactory Impaired Unsatisfactory

Surface Moderate/strong granular Sub-angular blocky or weak Massive or platy.

Structure 1 or single grained. granular.

Surface Many/common tubular pores, Common/few tubular pores. Few tubular pores, low vertical

Pore Space 1 high vertical continuity. continuity.

Rupture Loose to slightly hard (dry) ---- Very hard to very rigid (dry),

Resistance 1 Loose to friable (moist). Extr. firm to very rigid (moist).

Near Surface No surface crusting or Water compacted or non-biotic Mechanically compacted.

Subzones 1 subsurface compaction. surface crust present.

Bulk Density Bulk density not increased. Moderate bulk density increases Significant increase in bulk density

(5-15%). (>15%).

Infiltration No decrease in infiltration. Moderate decrease in infiltration. Significant decrease in infiltration

(10-50%). (>50%).

Penetration No increase in resistance. Moderate increase in resistance Significant increase in resistance

Resistance (10-50%). (>50%).

Modeled Current soil loss < tolerance. Current soil loss > tolerance.

Soil Loss

Visible Sheet Sheets/rills/gullies not evident. Rills/gullies are small, discontin- Rills/gullies actively expanding,

Rill & Gully uous, poorly defined & not well-defined, continuous &

Erosion connected into any pattern. connected into a definite pattern.

Pedestaling No/slight pedestaling of plant, Grasses, forbs and rock fragments Trees and shrubs are pedeslitter and rocks. No evidence are pedestaled. Small, fibrous root taled and may be hummocked.

of exposed roots. strands of forbs & grasses are Shallow, lateral roots of trees and

exposed on the soil surface. shrubs are exposed.

Erosion None to slight. If erosion Erosion pavement is continuous or

Pavement 2 pavement exists it is exists in interspaces between

discontinuous or localized. canopy cover of trees & shrubs.

Soil Not unusual or excessive. Soil and/or litter deposition is Soil and/or litter is deposited on

Deposition present. Fine litter may be patterned the uphill side of logs, brushpiles,

as small debris accumulations. etc. Soil may be moving offsite.

Surface “A” horizon is present, well “A” horizon is present, but not “A” horizon is absent or present in

(“A”) distributed, not fragmented. evenly distributed. Changes in association with prominent plants.

Horizon physical properties exist. Properties are similar to those of the

underlying subsoil.

Vegetative Distribution of desirable, Changes in vegetation composition The perennial forb and/or graminoid 

Community perennial plants reflects indicate a shift towards a drier, less vegetative layers are absent 

Composition species by vegetative layer productive plant community. or sparse.

(i.e. trees, shrubs, forbs and There may also be an increase in

graminoids) as identified in annual plants, shallow-rooted

the potential plant community. grasses, taprooted woody

perennials or invasive plants.

Litter Litter is distributed evenly Litter is either absent or is

across the soil surface and is ----- associated only with prominent

associated with all vegetative plants and not evenly distributed

layers across the soil surface.

Coarse Pipos/Quga-----5-10 t/ac.

Woody Pipos/Fear2-----7-14 t/ac. ----- Pipos/Quga-----<5 t/ac.

Material Abco/Fear2-----8-16 t/ac. Pipos/Fear2-----<7 t/ac.

Abco/Fear2-----<8 t/ac. 

Root Many/common roots in Moderately few roots in Few/very few roots in surface

Distribution1 surface horizons. surface horizons. horizons.

1/ Categories and/or descriptions defined in USDA Handbook No. 18, Soil Survey Manual, October, 1993.

2/ Certain soils within desert ecosystems inherently contain erosion pavement (desert pavement) surfaces. Desert pavements are 

not used to indicate soil condition.

 Function

H

Y

D

R

O

L

O

G

I

C

S

T

A

B

I

L

I

T

Y

N C

U Y

T C

R L

I I

E N

N G

T 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 139

Link (T-Walk) and Proper Functioning Condition 

(PFC) for riparian areas. T-Walk classifies stream 

conditions as either robust, adequate, diminished, 

impaired, precarious, or catastrophic (Ohlander 

1998). PFC protocol uses the categories Proper 

Functioning Condition, Functional - At Risk, and 

Nonfunctional to describe riparian conditions 

(USDI Bureau of Land Management 1993). Further 

examples of rangeland classification are provided 

by the National Academy of Sciences (National 

Research Council 1994).

BLM has also recently developed procedures for 

rangeland health assessment (Pellant and others 2000). 

These are based on three attributes of rangeland health: 

soil/site stability, hydrologic function, and integrity of 

the biotic community. The procedures use 17 indicators 

that focus on vegetation or soil stability to assess the 

function of the three attributes of rangeland health. 

The indicators are:

• Rills

• Water flow patterns

• Pedestals and/or terracettes

• Bare ground

• Gullies

• Wind-scoured, blowouts, and/or deposition 

areas

• Litter movement

• Surface soil resistance to erosion

• Soil surface loss or degradation

• Plant community composition and distribution 

relative to infiltration and runoff

• Compaction layer

• Functional/structural groups

• Plant mortality/decadence

• Litter amount

• Annual production

• Invasive plants

• Reproductive capability of perennial plants

There are also additional optional indicators.

Attributes are rated qualitatively according to 

condition classes that compare current conditions 

to departure from reference areas or ecological site 

description. Condition classes include extreme departure, moderate to extreme, slight to moderate, and 

none to slight.

Modeling_________________________

State-of-the-art modeling tools can be used to examine changes in grassland processes. For example, 

the Century Model simulates the dynamics of carbon, 

nitrogen, phosphorus, and sulfur over time. It can be 

used as a tool for ecosystem analysis, to evaluate the 

effects of management changes, and to analyze changes 

in the biogeochemistry of grasslands (Parton and others 

1996). The model is based on net primary productivity 

as a function of water availability, nitrogen availability, 

and temperature; and soil carbon and nitrogen pools 

that are active, slow, or passive in their turnover times. 

Century includes a grassland/crop submodel that is 

linked to a soil organic submodel simulating the flow of 

nutrients through plant litter and soil. The model has 

recently been used to assess the response of temperate 

and tropical grasslands to climate change (Pace and 

Groffman 1998).

HilleRisLambers and others (2001) used mathematical modeling to interpret vegetation patterns (patches of 

vegetation and bare soil) in semiarid grazing systems. 

They noted that there is a prevalent, positive relationship between plant density and water infiltration that 

results in formation of vegetation patterns in semiarid 

areas throughout the world. Modeling results showed 

that other factors such as herbivory (grazing) are not 

as important in generating the patterns. Where plant 

dispersal is low, increased herbivory is predicted to lead 

to a transition from closed vegetation cover to spatial 

patterns in vegetation to bare soil. This same transition 

is not likely to occur, however, where plant dispersal is 

high. The model also predicts that vegetation changes 

are reversible if grazing is reduced.

Summary__________________________

Grasslands in the Southwest have a semiarid to 

arid climate and a patchy distribution of precipitation. However, land managers often plan for average 

climate conditions and distribution, rather than the 

extremes that can be expected. For management 

purposes, the frequency, severity and distribution of 

drought are more important than long-term average 

climate conditions.

The USDA Forest Service continues to address 

various management concerns in grasslands including 

streambank erosion, sedimentation and erosion rates, 

restoring riparian vegetation, providing improved 

habitat for native species, and the cumulative impact 

of wildlife and livestock grazing. It is important to 

determine whether management activities impact 

ecosystem integrity and sustainability. Ecological 

sustainability can be defined as “the maintenance 

or restoration of the composition, structure, and 

processes of ecosystems over time and space.” One 

way to measure sustainability is using criteria and 

indicators of sustainability at local levels; an example 

is the Forest Service’s LUCID program, which was 

designed to test appropriate monitoring parameters at 

the local level. While there is no standard “tool kit” for 

assessing ecological sustainability, a variety of useful 

tools are available, including the Forest Service’s Soil 

Condition Rating Guide and modeling tools such as 

the Century model.
140 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Acknowledgments_________________

We gratefully acknowledge the reviewers for 

this chapter: Doug Powell of USDA Forest Service, 

Ecosystem Management Coordination in Washington, 

DC; Linda Joyce, USDA Forest Service Rocky Mountain 

Research Station in Fort Collins, CO; Rex D. Pieper, 

Professor Emeritus, New Mexico State University 

Department of Animal and Range Science; and Will 

Moir (retired), USDA Forest Service Rocky Mountain 

Research Station in Flagstaff, AZ.

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142 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Introduction_______________________

Grassland ecosystem management is dynamic and 

has adapted to the development of new tools and ideas. 

Our ancestors were indirectly managing grasslands 

when they learned to move livestock to take advantage 

of better water and greener forage. One could argue 

that even their hunting of grassland wildlife, especially 

the use of fire to drive animals to waiting hunters, had 

an influence on local grassland ecology.

The science of range management is relatively new 

and is linked to the realization that land is limited and 

that managers must do their best to make the available resources last forever (Stoddart and Smith 1955). 

Poor land stewardship degrades the land’s multiple 

resources and results in poor livestock production, 

while good management will provide healthy grassland ecosystems so that managers and landowners 

will have high livestock production and the potential 

for a viable economic livelihood. Owners, especially 

private owners, probably have different goals for their 

lands; traditional livestock production is not always 

the primary or even a secondary objective, but proper 

land stewardship should be the ultimate goal regardless of other objectives.

New challenges are impacting the management 

and health of all natural ecosystems, including 

Southwestern grasslands, and these forces are affecting land stewardship efforts. One of the greatest 

challenges is the result of the phenomenal growth of 

the region’s human population, primarily urban and 

suburban—people with few direct connections to the 

land. Large expanses of privately owned grassland, 

such as some near Sonoita, AZ, have been subdivided 

into small units, fragmenting a relatively homogeneous landscape with potentially adverse impacts on 

many wildlife species, on regional hydrology, and on 

the traditional social fabric of the area. The new residents often introduce nonnative plants through their 

landscaping activities, miles of fences and new roads, 

and concentrations of predatory pets and domestic 

livestock. Even without fragmentation, the pursuit of 

recreational opportunities by this growing population is 

putting pressure on all open landscapes, impacting the 

vegetation, wildlife populations, and watershed condition. Other challenges are the growing concerns about 

the loss of native plant and animal species because of 

habitat loss, human disturbances, and the introduction of numerous nonnative species that successfully 

occupy areas of bare soil or out-compete native plants. 

Many grasslands have been adversely affected by past 

land management and need to be restored to a more 

functional ecological condition. But how can this be 

accomplished ecologically and economically?

Ecological and range management sciences continue 

to provide managers with new information and tools. 

Descriptions of many of the common techniques are 

found in standard range management textbooks dating 

from the historical coverage by Stoddart and Smith 

(1955), and from books such as Savory (1988), Holechek 

Chapter 8:

Tools for Management for Grassland 

Ecosystem Sustainability: Thinking 

“Outside the Box”

Gerald J. Gottfried
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 143

and others (1998), and Jemison and Raish (2000) 

that describe more recent developments. Important 

information has been gathered and disseminated by 

Federal, university, and State research, management, 

and extension agencies.

This chapter covers some new and current tools, 

issues, and approaches to grassland management. It 

does not attempt to be all inclusive but does reflect 

the opinions and expertise of the contributors. The 

first section includes some programs and tools and 

principles related to grassland management and restoration, and the second section covers new remote 

sensing technologies to evaluate grasslands and assist 

in management planning at a landscape level. The 

adaptation and use of remote sensing technologies 

and geographical information system databases to 

grassland management are just developing. The third 

section deals with education, a vital aspect of range 

management if current and future grassland managers 

are to do their best and if the general public is to be a 

true shareholder in these efforts.

The chapter has benefited from the written contributions of several USDA Forest Service employees from 

the Southwestern Regional Office in Albuquerque, NM, 

or from the Rocky Mountain Research Station. The 

authors and their sections are indicated throughout 

the text. Their contributions are greatly appreciated. 

Special thanks are due to Cathy Dahms, formerly of the 

Southwestern Region, who recruited the contributors 

and assisted with the editorial process.

Grassland Resources Management___

New Programs to Sustain Southwestern 

Grasslands

Attitudes and philosophies toward grassland 

management have shifted over time since the first 

herds of livestock were introduced into the Southwest. 

However, many changes have occurred more recently 

in response to the growing pressure on grasslands and 

their resources from traditional producers, conservation and environmental organizations, government 

entities, and especially from the growing regional 

population. Ranchers, conservation groups, and many 

members of the general public have a common interest in maintaining open landscapes because of the 

ecological importance of these systems, recreational 

and aesthetic concerns, and the desire to preserve the 

rural livelihood and life style (Raish, this volume). 

These groups, often working together, have developed 

innovative approaches to achieve their goals.

Grassbanks—Grassbanks are becoming popular 

methods for allowing land and grass resources to rest 

and recover during dry periods while continuing to 

provide the rancher with an economic return. The 

Malpai Borderlands Group (MBG), located in southeastern Arizona and southwestern New Mexico, was 

the first organization to initiate the Grassbank system 

to restore native grasslands and protect the open 

spaces of the region. “Grassbank” is a legally registered 

trademark. Basically, a Grassbank is grass made 

available from one ranch to another rancher’s cattle 

in return for conservation value equal to the value of 

the grass (McDonald 1995). The MBG is a nonprofit 

organization, composed mainly of ranchers, that works 

to encourage profitable ranching and traditional livelihoods that sustain the region’s open spaces (McDonald 

1995). A specific goal is: “...to restore and maintain 

the natural processes, including fire, that create and 

protect a healthy, unfragmented landscape to support 

a diverse, flourishing community of human, plant and 

animal life in our Borderlands Region” (McDonald 

1995: 483). Drum Hadley, of the Animas Foundation 

is credited with the idea of a Grassbank as a method 

of creating more grass, grass seed, and protection of 

open spaces. The Animas Foundation’s Gray Ranch 

(also known as the Diamond A Ranch) provided the 

first Grassbank.

Grassbankers receive access for their cattle to the 

Grassbank in return for conveying a conservation 

easement to the MBG. The MBG pays the owner of 

the Grassbank for the use of the grass. The value of 

the easement is determined by an appraisal of the 

development rights on the Grassbanker’s ranch. 

The Grassbanker gets access to the grass for a term 

that monetarily equals the value of the easement. 

A conservation easement prevents the sale of land 

for subdivisions and, thus, prevents landscape fragmentation. Easements can be released back to the 

landowner under two conditions. The first condition is 

the dissolution of the MBG, when no acceptable alternatives are available. This condition is grandfathered into 

original Grassbank agreements but is no longer part of 

easements now being offered. The second condition is 

the loss of access to grazing on nonprivate lands that 

are part of the landowner’s ranch when this occurs 

through no fault of the rancher. Resting grasslands 

is a recognized method of improving range conditions 

and ultimately to improve the economics of ranching. 

The MBG and some government agencies, notably, 

the USDA Natural Resources Conservation Service, 

also may cooperate with Grassbankers to conduct improvement projects and monitoring to help the land 

return to a more productive condition before cattle are 

returned to the rested ranges. Fencing projects are 

one example of range improvements. Ranchers on the 

rested ranges agree to maintain their water systems 

for wildlife use during periods of rest. Stocking rates 

have to be maintained at lower levels than normal 

within the Grassbank so that the land will be able 
144 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

to provide sufficient forage for the additional cattle. 

Some lands within the Grassbank may be rested or 

deferred in anticipation of receiving a Grassbanker’s 

livestock. Grassbank strategies need to be flexible to 

allow for variable climatic conditions.

Grassbanking also is being practiced in the Valle 

Grande Grass Bank Demonstration Program in 

northern New Mexico (Valle Grande Grass Bank 

Briefing Information, Anonymous, undated). The 

area is east of Santa Fe at the east end of Glorieta 

Mesa. The program is a cooperative effort involving 

the USDA Forest Service, the grazing permittees, 

the Conservation Fund, and the New Mexico State 

University Cooperative Extension Service. The basic 

goals of creating sustainable ranching are similar to 

those of the Malpai Borderlands Group, but this association consists of Forest Service permittees rather 

than private landowners. The Conservation Fund, a 

nonprofit organization, which holds the Valle Grande 

allotment and owns associated private lands, also hopes 

to reduce conflicts between grazing and other land uses 

and to demonstrate positive livestock and range management. Education for permittees and the public are 

important aspects of the program. Grassbankers under 

the program join the Valle Grande Grazing Association 

and are assessed a fee to cover operations; however, 

the Conservation Fund could provide supplemental 

funding. While their ranges are being rested, members 

are expected to conduct range improvements, such as 

riparian fencing and brush removal operations, which 

the Forest Service will fund. Grassbank arrangements 

can extend from 3 months to 3 years.

Other ranching organizations that foster proper land 

stewardship, such as the Quivira Coalition of Santa 

Fe, NM, also indicate that grassbanks are one tool 

for fostering ecological, economic, and social health 

of Western landscapes (Quivira Coalition undated). 

The USDA Forest Service is using a similar system 

in a coordinated and cooperative approach with forest 

partners when opportunities arise following a range 

assessment. The Tonto National Forest in central 

Arizona has policies and procedures in place that 

consider allowing one rancher to use another’s unused 

allotment if all parties are in agreement (B. McKinney 

2004, personal communication).

Conservation Easements—Conservation easements are voluntary legal instruments by which 

development rights or the rights to conduct other 

activities are conveyed to a qualified conservation 

entity (Society of American Foresters 2001, Stein and 

others 2001). The Malpai Borderlands Group requires 

conservation easements for ranchers to participate 

in its Grassbank program. Ranchers, conservation 

organizations, and many government agencies are 

concerned about the subdivision of large areas of private 

agricultural, forest, or range lands to create parcels of 

land for “40-acre ranchettes” or other small home site 

ownerships. Subdivision can create a personal loss to 

the original owner because many generations of effort 

are lost and the land will not be passed to heirs, even 

though the land sale may be financially profitable. 

Some landowners may consider selling their ranches for 

subdivision to provide for retirement or inheritance for 

their children. Society then loses valuable, scenic open 

spaces and the viability of local rural economies that 

depend on traditional agricultural activities (Rosan 

and others 1997). Subdivided landscapes can threaten 

the viability of dynamic ecosystems by obstructing traditional migration routes, destroying wildlife habitat, 

introducing nonnative plant and animal species, and 

increasing the demands on limited water supplies. 

Many wildlife species are sensitive to disturbances 

attributed to increased human activities (Mitchell 

and others 2002). One study in Colorado compared 

subdivided and intact ranches and found that subdivision resulted in an eightfold increase in road densities 

and, on one ranch, in a fourfold increase in number of 

landscape patches (Mitchell and others 2002). Another 

Colorado study found native plant communities were 

maintained better with less bare soil on ranches than 

in subdivided lands or in nature reserves (Maestas and 

others 2002). Maestas and his associates (2002) also 

cite evidence that there is change in animal biological 

diversity as lands are subdivided, with an increase of 

human-adapted species and a decrease in sensitive 

species. The alternative of conservation easements 

can be an effective tool for maintaining working 

landscapes, preserving environmental values, and 

protecting communities from excessive development 

(Society of American Foresters 2001).

Among the several approaches to conservation 

easements is the purchase of development rights 

(PDR). This allows the landowner to conserve working 

landscapes using market and incentive based, nonregulatory techniques (Stein and others 2001). PDR 

programs began in the Eastern United States in the 

1970s to protect open spaces from suburban sprawl. 

Currently, more than 1,200 nonprofit land trusts in 

the United States (Hocker 2002) have protected about 

2.6 million acres from future development through 

conservation easements (Maestas and others 2002). A 

local government or private organization, such as The 

Nature Conservancy and various nonprofit land trusts, 

purchases the real estate development rights from the 

landowner; however, the owner still owns the land 

and can use it for range, forestry, or agriculture and 

still obtains a financial benefit from the management 

of these resources. However, the activities may not 

affect the land’s conservation value (Rosan and others 

1997). The owner may sell the rights to all or part of 

the land and still retain part for limited development, 

for example for future home sites for children or to 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 145

meet cash needs. These lots should not interfere with 

the ranching operation or detract from the scenery 

or natural resources (Rosan and others 1997). The 

land remains in the family and can be passed on to 

the owner’s heirs, and a PDR may reduce inheritance 

taxes because the land’s value has been reduced by 40 

to 75 percent (Stein and others 2001). The purchaser 

of an easement is responsible for monitoring to ensure 

that the terms of the agreement are followed. The easement usually stays with the land for perpetuity and 

cannot be cancelled if the property is sold. However, 

some easements specify shorter durations (Society of 

American Foresters 2001). The PDR is a particularly 

important tool in the West because few governments 

or private organizations can afford to directly purchase 

large tracts of land. State governments have raised 

money for PDR programs, for example, by using parts 

of sales tax levies or by State lottery profits. Funding 

also can come from programs administered by Federal 

agencies such as the USDA Forest Service and USDI 

Bureau of Land Management.

There often are more requests by landowners to 

enter into a conservation easement than there is funding available. A landowner may sell the easement at 

less than its full value as a “bargain sale” and use the 

difference as a tax-deductible charitable donation.

The landowner may donate the entire easement to a 

nonprofit land trust or government entity. The amount 

of the charitable deduction is based on the impact of 

the easement on the value of the property (Rosan and 

others 1997). The value of an easement of more than 

$5,000 must be verified by a qualified appraisal if a 

contribution is claimed against Federal taxes.

Grassland Restoration—Grassland restoration 

is needed to return the ecosystems to productive and 

healthy states by enhancing the herbaceous cover 

that will increase productivity and biological diversity 

and also contribute to reduced erosion and to the 

reintroduction of fire into grassland ecosystems. Box 

(2002) says that land has been abused by people, the 

economy, and the weather. Many grasslands in the 

Southwest are in need of restoration because of past 

heavy grazing by herds owned by early ranchers and 

corporations who were trying to make a profit or just 

trying to make a living, combined with periodic drought 

that often extended for several consecutive years. This 

resulted in declines in herbaceous cover and species 

diversity, and increases in woody species. The changes 

in vegetation characteristics produced modifications 

of the hydrologic cycle, in particular, less infiltration 

of precipitation and increased surface runoff, and accelerated soil erosion and sedimentation.

Southwestern grasslands have been used and often 

abused for more than a century in parts of New Mexico. 

Sheep were an important exportable item during the 

Spanish Colonial Period in New Mexico, and large 

bands were maintained and herded down the Rio 

Grande Valley to Mexico (Gottfried and Pieper 2000). 

Large herds of cattle were imported into New Mexico 

and Arizona Territories after the suppression of the 

Apache Indians in the 1880s and the construction of 

the transcontinental railroad (Schickendanz 1980). 

In the early 1880s, three large cattle companies ran 

more than 60,000 head of cattle in Cochise County, 

Arizona, and Hidalgo County, New Mexico (Hadley and 

others 1999). Grazing continued in the area through 

the droughts of 1885, 1892 to 1893, and 1902 to 1903, 

when up to half of the cattle died of starvation and range 

resources were rapidly depleted (Hadley and others 

1999). Heavy grazing also occurred in northern Arizona, 

and probably in other areas of the Southwest, during 

World War I to meet the demand for meat (Schubert 

1974). Grazing during these earlier periods exceeded 

the carrying capacity of the land and resulted in a 

rapid decline in range vegetation resources, and in 

accelerated erosion, and channel downcutting. This 

combination of factors initiated desertification along 

the United States/Mexican border region (Hadley 

and others 1999) and throughout the Southwest. The 

impacts on watershed condition during this time still 

are apparent and can be identified by erosion rates, 

gully erosion, and soil compaction (USDA Forest 

Service 1993).

Two primary impacts of the loss of the herbaceous 

cover have been accelerated erosion and the encroachment of woody species, such as mesquite (Prosopis 

spp.) or juniper species (Juniperus spp.). The increased 

density of woody species has been linked to past heavy 

overgrazing by livestock and the consequential removal 

of fuels for wildfires as a natural control mechanism 

in many grassland communities (see the review by 

McPherson and Weltzin 2000). Exceptions are some 

drier desert grasslands where a continuous cover of 

fine fuels that would support large-scale fires did not 

occur (Buffington and Herbel 1965). Heavy overgrazing 

reduced grasses and fine fuels that were ignited by 

lightning during the late spring and early summer. Past 

fires eliminated or reduced the population of small trees 

and shrubs, maintaining low tree densities. Species 

that sprout after fires were kept in a subdominant 

position by repeated fires even if the plant survived. 

Under current conditions, many grasslands do not have 

the continuity of fuels to allow the uninhibited spread 

of fires. Without fire, trees that are adapted to a site 

will become established and regenerate successfully. 

However, the continued spread of invasive species 

such as Lehmann’s lovegrass (Eragrostis lehmanniana), which is adapted to fire, may alter the present 

fire frequencies (McPherson 1995), and because of 

competition, these invasive species will impede the 

establishment of preferred native perennial grasses. 

The loss of ground cover and the concentrations of 
146 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

livestock in stream channels and wet meadows also 

resulted in severe erosion and loss of watershed condition on many grassland ranges.

The increase in woody species also has been linked to 

increased changes in the atmosphere or in the proportion of summer and winter precipitation. McPherson 

and Weltzin (2000) linked the increase in woody species to increased concentrations of carbon dioxide and 

other greenhouse gases in the atmosphere. Ranchers 

and researchers in southeastern Arizona, where most 

vegetation is dependent on summer rains, indicate 

that the proportion of annual precipitation has shifted 

to more winter precipitation when compared to the 

century-long average (Brown 1999, Valone and others 

1998). This would favor the C3 plants such as mesquite 

over the native C4 grasses (Brown 1999); it also has 

been linked to declines in some key small mammal 

populations and increases in other species (Valone and 

others 1998). Aggressive fire suppression policies of 

Federal and State land management agencies have 

contributed to the situation by preventing potentially 

beneficial fires from burning. However, Allen (1998) 

indicates that fire suppression was not as effective as 

some maintain.

Many range restoration techniques call for the 

removal of trees. Restoration aims can range from 

complete removal of trees, although this is difficult, 

to the creation of savannas that retain a proportion 

of the trees as groups or individuals, to the creation 

of mosaics of grass openings and tree-covered zones. 

The retention of some residual trees and groups of 

trees would be beneficial for certain guilds of birds 

and as hiding and thermal cover for larger wildlife 

and livestock. Some prescriptions hope to establish a 

continuous and relatively dense stand of grasses and 

forbs so that prescribed fire eventually can be used 

to maintain the tree component in a less dominant 

position relative to the herbaceous species (Gottfried 

and others 1999).

Restoration efforts must be conducted with caution; 

one treatment will not be appropriate for all sites. 

Grasslands are complex ecosystems with different 

mixes of herbaceous and nonherbaceous species, site 

characteristics, and climatic conditions. Historical land 

uses have altered most areas, and the amount of change 

will affect present conditions and management options. Many areas that are perceived as grasslands are 

actually woodlands and never supported the grass communities that are projected to have been present. This 

includes areas adjacent to prehistorical and historical 

American Indian population centers where woodlands 

were heavily cut for domestic and agricultural purposes, 

and to areas surrounding Spanish-Mexican settlements 

and United States military posts in New Mexico. Many 

of the grasslands were actually savannas that contained 

scattered trees that provided the seed sources for an 

increase in tree density once fire events became less 

frequent. The presence of old or large trees would be 

a sign that caution is warranted. Some sites also have 

been so eroded and changed since the grasses were 

lost, that it may not be ecologically or economically 

possible to return to the original grassland communities. The introduction of invasive nonnative species, 

such as Lehmann’s lovegrass and buffelgrass (Cenchrus 

ciliaris), has made restoration of native grasses more 

difficult (Weltzin and McPherson 1995). Long-term 

changes in greenhouse gases may make restoration 

difficult. Managers should recognize that even the best 

planned and executed treatments will not produce the 

desired results if weather conditions are not favorable. Except for ecologically critical areas, economical 

considerations may be the ultimate factor determining 

if, when, and what techniques should be employed to 

restore a grassland in a given location.

A wide variety of mechanical methods and prescriptions exist for use in reducing the woody species 

cover on grasslands (Vallentine 1971). The successes 

of these treatments depend on the characteristics of 

the woody species, including age, reproductive strategies, and stand densities; site characteristics such as 

slope, rockiness, terrain, and soil conditions; desired 

replacement vegetation; and soil seed bank availability. 

Herbicides also have been applied to mesquite and 

other woody species, but primarily on private lands. 

Climatic conditions before, during, and after an operation may determine if even the best planned and 

executed restoration treatment is successful. Forage 

increases after mesquite control, for example, will 

only occur if work is done during years of average or 

above average precipitation (Scifres and Polk 1974). 

It is recommended that livestock grazing be deferred 

from a site for a prescribed time after treatment so that 

new grasses can become established. Grazing by large 

and small wildlife species is more difficult to control. 

Deferral for 2 to 4 years has been recommended for 

many pinyon-juniper ranges (Gottfried and Severson 

1993), but the amount of time will depend on the amount 

and condition of residual species, site potential, and 

weather conditions (Gottfried and Pieper 2000). Many 

of the failures of past pinyon-juniper treatments can 

be linked to premature grazing by livestock.

Unless satisfactory seed from residual grasses and 

forbs or from the soil seed bank are present, restoration 

will require seeding of native species. Forbs and shrub 

species could be seeded on some locations depending 

on site conditions and management objectives. It may 

be necessary and desirable to reestablish some shrubs, 

such as Purshia spp., which are important browse 

species. Although seed is usually not collected locally 

before a treatment, seed should come from a source as 

near to the project site as possible. The same principles 

that apply to tree seed provenance should apply to 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 147

grass and associated species seed. Grassland species 

seed should, as closely as possible, be adapted to the 

site—for example, elevation and soils. Some desirable native species, such as black grama (Bouteloua 

eriopoda), may not be available from commercial 

seed companies, and adapted species may have to be 

substituted. The seeding of adapted nonnative species 

such as Lehmann lovegrass was common at one time 

but currently is discouraged.

Managers must consider the proper method of applying seed, whether seed should be drilled into the 

soil surface with a rangeland drill or if seed can be 

broadcast directly on the ground without cover. The 

creation of depressions in the soil surface in conjunction 

with seeding will enhance grass stand establishment 

in semidesert grasslands (Gottfried and others 1999). 

Each species has specific requirements. Timing of a 

treatment is important and should be done prior to or 

during the precipitation season for which the species 

is adapted. Optimum season for seeding will vary by 

site. Obviously, unprotected seed would be subject to 

granivory by insects and rodents or be blown or washed 

from the site. An established tree or shrub canopy that 

creates a moderate microclimate or reduces herbivore 

activity could result in reduced losses of viable seed and 

new seedlings. Restoration prescriptions should include 

monitoring before and for a period after treatment to 

ascertain if satisfactory results were achieved or if 

procedures should be modified for future success.

Research on the value of inoculating restoration 

sites with appropriate arbuscular mycorrhizal fungi to 

promote the survival and growth of the grasses could 

Weather stations, such as this 

one on the Cascabel Watershed 

Area of southwestern New Mexico, 

provide data that are used to plan 

and evaluate grassland management and restoration treatments. 

(Photo by John Yazzie)

A grassland restoration treatment designed 

to crush mesquite and seed native perennial 

grasses at the George Wright Pasture of the 

Diamond A Ranch Central Division in southwestern New Mexico. The crusher created 

depressions in the soil that enhance seeding 

success. (Photo by Ronald Bemis)
148 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

enhance restoration successes. At present, there are 

no practical and affordable methods of reintroducing 

mycorrhizal plants to disturbed grasslands (Caplan 

and others 1999). Some work has been started on the 

relation between giant sacaton (Sporobolus wrightii) 

and mycorrhizae from a southeastern Arizona site 

(Elliott and others 1998). Seedballs consisting of 

clay, soil humus, grass seed, water, and a source of 

mycorrhizae may be an excellent way to disperse seed 

in disturbed areas (Caplan and others 1999). The ball 

should contain root segments of the native culture host 

plant with hyphae.

Watershed Condition—Many degraded grasslands have severe erosion problems; surface soils have 

been lost, and gullies and degraded stream channels 

are present. Most Southwestern grasslands are influenced by convectional, high-intensity thunderstorms 

during the summer monsoon. These thunderstorms 

have been linked to most surface runoff and erosion 

events, especially on degraded watersheds (Rich 

1961). A change in surface conditions that increases 

infiltration capacity, whether by increasing the plant 

cover or by mechanical stabilization, should result 

in decreased surface runoff and soil erosion (Rich 

1961). Increasing the herbaceous cover, especially in 

interspace areas between tree-influence zones will 

protect the soil surface from raindrop compaction and 

consequential soil pore sealing. Plant roots, especially 

the fibrous roots of monocotyledons, help buffer the 

soil from sealing, hold soil particles in place, improve 

the soil porosity and, thus, water movement in and 

through the soil, and increase water holding capacity 

because of organic additions to the soil. Plants also 

act as microbarriers to the free movement of surface 

runoff, allowing for greater infiltration and 

less soil movement within and from the 

larger watershed and the resulting losses in 

stream water quality. Increased infiltration 

provides more water for plant survival and 

growth. In many cases, changes in grazing 

season and intensities can reduce sediment 

loads from some grassland sites (Holechek 

and others 1998). Vegetation manipulations, by themselves, will not work on all 

sites; more expensive mechanical methods 

also may be necessary to prepare a site 

for restoration. Contour furrowing and 

trenching, ripping, or pitting has been used 

to retain water and reduce sedimentation 

(Brooks and others 2002). An early study 

in central Arizona found that surface runoff 

and erosion were reduced by a treatment 

that combined brush control, sloping of 

steep gully sides, placing cut brush in 

gully channels, and grass seeding (Rich 

1961). Gullies develop when surface runoff 

is concentrated at a point where there is 

an abrupt change in elevation and slope and a lack 

of vegetative cover (Brooks and others 2002). Check 

dams and other barriers may be needed to retard head 

and down cutting and stabilize gullies until vegetation becomes established. Ranchers in Arizona’s San 

Bernardino Valley are using local rocks to construct 

dams and structures to limit headcutting and to create 

meanders in relatively straight gullies. The meanders 

slow streamflow velocities, reducing potential bank 

cutting and increasing sediment deposition creating 

sites for the reestablishment of vegetation. Engineered 

structures may be necessary on larger streams to raise 

their base levels, thus reducing channel gradients and 

cross sections and streamflow velocities (Brooks and 

others 2002). Practices should not result in increased 

water pollution and should be conducted under existing 

Best Management Practice guidelines.

Prescribed Fire—The introduction of prescribed 

fire and the management of natural fire in grassland 

ecosystems are of major interest today. There is wide 

acceptance of fire as a paramount factor in maintaining native grass plant communities by reducing woody 

plants and removing the buildup of grass litter prior to 

Euro-American settlement. Landscape-level prescribed 

fires are usually designed to create mosaics of wooded 

and grass areas in addition to reducing general tree 

densities. In addition to reducing the cover of woody 

species, periodic fires may keep ungrazed grasslands 

vigorous by killing forbs and removing grass litter 

(Robinett and Barker 1996). However, some forbs may 

be desirable because they are high in the protein and 

moisture that are important for a number of animals 

(C.H. Sieg 2002, personal correspondence).

The sediment dam at Cascabel Watershed, New Mexico, is used to measure 

sediment production related to land management activities. (Photo by Gerald 

Gottfried)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 149

Fire must be used judiciously after a thorough 

analysis of benefits and risks. A single prescribed 

fire may not achieve all of the ecosystem objectives 

for an area, and multiple burns may be necessary. 

Information is often lacking concerning pretreatment 

fuel loadings on an area dominated by trees. The Fuel 

Characteristic Classification System (FCC) (Ottmar 

and others 2003) has coverage for many pinyon-juniper 

sites and is currently being expanded to cover oak and 

juniper woodlands and savannas that are common 

in the Southwest. Several fire prediction models and 

systems may be appropriate in planning prescribed 

burns on sites with significant tree cover. A version 

of BEHAVE (BEHAVE PLUS), which was originally 

developed by Burgan and Rothermel (1984), was used 

in preparation for the 46,000-acre Baker II Burn in 

the southern Peloncillo Mountains of Arizona and 

New Mexico in June 2003 (P.A. Gordon 2004, personal 

correspondence). FCC information eventually will 

be linked to models such as BEHAVE and FARSITE 

(Ottmar and others 2003). Monitoring of fire effects 

is needed if a program is to be biologically and economically successful. Monitoring should be designed 

to include soil and hydrological responses as well as 

the common vegetation and wildlife inventories. The 

Revised Universal Soil Loss Equation (RUSLE) has 

been used to evaluate soil erosion from grasslands, but 

it tends to underestimate soil losses (Spaeth and others 

2003). There is a shift to the Water Erosion Prediction 

Program (WEPP), especially for postfire predictions. A 

current effort in southeastern Arizona is attempting to 

define parameters for the Disturbed WEPP that could 

be used to predict soil losses following fires in semiarid 

grasslands (Paige and others 2003).

Wildfires can be beneficial, and each ignition should 

be evaluated prior to suppression actions. Fires obviously can be attacked if they are a potential danger to 

structures and improvements or if the landowner favors 

suppression; they can be monitored before a decision 

is made, or they can be allowed to burn. However, 

the last option must be taken with an understanding 

of the characteristics and needs of the specific area 

and of the actual and perceived impacts on encroaching population areas (Allen 1998). The Peloncillo 

Programmatic Fire Plan is an attempt to identify the 

prescribed fire and wildfire suppression philosophies 

of ranchers and landowners in this mountain range 

so that land managers may anticipate the appropriate actions (Allen 1999). Potential fire intensities 

may have changed since settlement times because of 

The prescribed burn at Baker Canyon in the Peloncillo Mountains, which straddle the border between Arizona and New Mexico, 

June 1995. The objective was to reduce tree densities and to create mosaics of grassland and tree dominated areas. (Photo 

by Gerald Gottfried)
150 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

local buildups of larger woody fuels in savannas and 

riparian corridors. Current and desired fuel loadings 

are considerations, as are the habitat requirements of 

threatened, endangered, and sensitive (TES) species. 

The U.S. Fish and Wildlife Service and appropriate 

State agencies must be consulted on these issues.

Understanding the role of fire in maintaining habitats for rare species and the dynamics of fire in riparian 

areas will lead to more informed decisions (C.H. Sieg 

2002, personal correspondence). The proper time to 

ignite prescribed fires is a concern where TES species 

habitats are present or where riparian areas could be 

adversely impacted. In semidesert grasslands, species 

are more influenced by fire season and frequency than 

by fire behavior (Steuter and McPherson 1995). Most 

wildfires in southeastern Arizona and southwestern 

New Mexico occur in the warm period as the annual 

monsoon develops, and most species are adapted to 

the relatively hot fires that occur prior to the growing 

season. Some authorities prefer cool season fires in 

the fall, winter, or early spring because they feel that 

less damage is done to the ecosystem. Others believe 

that cool season burning will cause a shift in species 

composition to favor shrubs and half-shrubs or will 

leave the soil surface open to accelerated erosion for 

a longer period. However, scientific data and documented observations are unavailable, and research 

is needed to determine if one season is better than 

the other. The research must address the impacts of 

the two burning prescriptions on other physical and 

biological ecosystem components, such as hydrology, 

sedimentation, arthropod populations, common animal species, and vegetation, as well as TES species. 

Ford and McPherson (1998) and Gottfried and others 

(2000) are studying some of these questions on the 

shortgrass prairies in eastern New Mexico and the 

oak (Quercus emoryi) savannas of southwestern New 

Mexico, respectively.

Fire frequency is a consideration in any prescribed 

burning program because frequent repeated fires 

may not allow for the recovery of many grass species. 

Pase and Granfelt (1977) recommended that at least 

5 years separate fires so that herbaceous plants have 

an adequate period to recover and set seed. Kaib and 

others (1999) indicate that the grasslands of southeastern Arizona had low intensity fires every 4 to 8 years 

prior to the introduction of large herds of livestock. 

Research in southern Arizona has shown that grasslands in good condition can sustain a fire interval of 

between 5 and 10 years without a loss of productivity 

(Robinett and Barker 1996). Grass recovery depends 

on precipitation after treatment and the amount of 

herbivory (McPherson 1995).

Many prescribed fires in the Southwest occur in the 

early summer prior to the monsoon period. This also is 

a period when fire suppression and control resources 

for prescribed fires may be scarce because of wildfires. 

Prescribed fires are cancelled because of the lack of 

personnel and equipment during this period even when 

fuel and weather conditions are satisfactory and the 

risks to control are low. One suggestion is that the 

land management agencies consider creating a fire 

management organization, including personnel and 

equipment, to plan and conduct prescribed fires that 

is separate from the fire suppression organization 

(Bemis 2003).

Improving Soil Fertility—Heavy grazing and 

accelerated soil erosion have been linked to reductions in soil organic matter and soil fertility on many 

rangelands (Aguilar 1993). Any attempt to improve 

site potential and grass productivity must restore soil 

organic matter and associated nutrients; one recent 

method is to apply treated municipal sewage sludge 

to the soil surface. The Rocky Mountain Forest and 

Range Experiment Station conducted several experiments on grasslands in the Albuquerque area. Aguilar 

(1993), citing a number of studies by Fresquez and his 

associates, indicates that a one-time sludge application 

of 10 to 20 tons per acre increased plant production 

and ground cover without producing unsatisfactory 

levels of potentially hazardous constituents, such 

as heavy metals, in either soils or plant tissue. Blue 

grama (Bouteloua gracilis) yielded 1.5 to 2.7 times 

more production on treated compared to control plots. 

Total nitrogen, phosphorus, potassium, and electrical 

conductivity increased during the first year, and soil 

organic matter increased by the fifth year after application. Aguilar (1993) concluded that many areas with 

depleted soil nutrients would benefit from prescribed 

amendments of sewage sludge. A pot study in Texas 

found that applications of biosolids increased shoot 

growth of blue grama and tobosagrass (Hilariamutica) 

because of increased soil NO3-N (Mata-Gonzalez and 

others 2002). Low levels of biosolids and irrigation (40 

percent of field capacity) resulted in increased root 

biomass, while high levels of both (80 percent of field 

capacity) resulted in a greater allocation of resources 

to the grass shoots. High applications of biosolids in 

the spring (34 and 90 Mgha-1) produced greater soil 

nitrogen concentrations and shoot growth than did 

summer applications.

Another alternative is to use local materials as 

mulch. A recent study in the semidesert grasslands 

on the Santa Rita Experimental Range in Arizona 

examined the effects of combined mesquite overstory 

treatments and soil surface mulch treatments on 

herbaceous production (Pease 2000, Pease and others 

2003). Overstory treatments consisted of removing the 

mesquite with or without sprout control and leaving 

the mesquite trees. These were combined with soil 

surface treatments consisting of mulches of lopped 

and scattered mesquite slash, commercial compost, 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 151

and mesquite wood chips. The mesquite control treatment with no sprout controls resulted in a 79 percent 

increase in annual and fall herbaceous production of 

native species relative to the control plots; Lehmann 

lovegrass was not affected by the treatments. The 

mulching treatments had no effect on total annual, 

spring, or fall herbaceous production. The lack of response was attributed to low precipitation during the 

4-year study, and to levels of mulch that were too low 

to impede evaporation (Pease 2000, Pease and others 

2003). The lopped and scattered treatment did improve 

total soil nitrogen, plant available phosphorus, and 

soil pH relative to the controls.

A forestry technician downloads hydrologic data to a 

laptop computer from a Parshall flume installation at 

the Cascabel Watersheds, New Mexico. The small 

flume measures typical low streamflow events and the 

larger flume measures higher events. The 12 small 

instrumented watersheds will be used to evaluate the 

effects of warm and cool season prescribed burning 

and unburned conditions on the physical, chemical, 

and biological compenents of the oak savanna ecosystem. (Photo by John Yazzie)

Prefabricated steel flumes, 

which can be assembled in the 

field, have been used to measure 

streamflow in remote, relatively 

inaccessible areas. (Photo by 

Gerald Gottfried)

Holistic ResourceManagement (HRM)—Savory 

(1988) introduced the idea of HRM into the United 

States from Zimbabwe. HRM is a grassland system or 

planning model that considers social, economic, and 

biological needs. However, it basically uses livestock 

to accomplish its goals. It involves a high-intensity, 

short-duration, time-controlled grazing system based 

on the phenological and physiological needs of the 

plant and animal species being managed, followed 

by periods of nonuse. The number of paddocks in the 

system is a function of the length of the deferment 

period that the plants require to maintain good 

health and high vigor; this usually is between 45 
152 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

feeding also stimulates new growth on existing plants, 

and microorganisms associated with the hay and cattle 

droppings improve soil fertility and help buffer the soil 

pH of the tailings. Irrigation systems are available on 

many tailing sites. Short-duration/high-intensity grazing has been tried on a severely eroded site in the San 

Bernardino Valley of southeastern Arizona (Gottfried 

and others 1999). Mechanical restorations could not 

be used on the site because of valuable archeological 

resources. Native hay was spread on the fenced site 

before cattle were introduced for less than 3 days, 

and the site was seeded with a mixture of native 

grasses. Most of the resulting grass seedlings—cane 

beardgrass (Bothriochloa barbinoides) and, Arizona 

cottontop (Digitorial californica)—were from the hay. 

Initial results have not been as good as expected and 

were similar to the control area because of summer 

droughts and insect herbivory, both of which are common problems in the Southwest.

Watershed Management: Best Management 

Practices

Author: Penny Luehring, Southwestern Region

The Clean Water Act of 1972 (Federal Water Pollution 

Control Act, amendments of 1972; PL 92-500, 33U.S.C. 

1311-1313, 1315-1317) defines a Best Management 

Practice (BMP) as: “a practice or a combination of 

practices, that is determined by a State (or designated 

area-wide planning agency) after problem assessment, 

examination of alternative practices and appropriate 

public participation, to be the most effective, practicable 

(including technological, economic, and institutional 

considerations) means of preventing or reducing the 

amount of pollution generated by non-point sources to 

a level compatible with water quality goals.”

Nonpoint source pollution is water pollution that 

originates from many indefinable sources. Nonpoint 

source pollutants are generally carried over, or through, 

the soil and ground cover via streamflow processes. 

Resource management activities likely to occur in 

grassland ecosystems that may be considered nonpoint 

sources include runoff from grazing, construction, 

revegetation, restoration, prescribed burning, wildfire 

suppression, pest or invasive-plant control, developed 

recreation sites, mining, road construction, and road 

maintenance.

Best Management Practices must be designed 

on site-specific basis for planned activities, taking 

into consideration the degree of surface disturbance 

anticipated, drainage patterns, climate, slope, soil 

erodibility, and proximity to stream channels. BMPs 

may be operational or administrative. Operational 

BMP examples include putting drainage dips in roads, 

leaving untreated buffer strips between an activity 

and the stream channel, or applying mulch. Examples 

to 90 days, depending on the timing and amount of 

precipitation. Paddock use is scheduled to utilize both 

warm and cool season plants. Livestock are moved 

on a fast rotation during the growing season and are 

only allowed to remove 5 percent of the forage crop 

on each visit to a paddock before it is deferred. HRM 

requires a high degree of balance between annual 

forage consumption and annual forage production, the 

latter of which depends on annual climatic conditions, 

and HRM requires intensive herd management to 

achieve the desired effects.

High stocking densities are projected to improve 

water infiltration into the soil, increase mineral cycling, 

increase the number of plants consumed, improve the 

leaf area index, improve the distribution of grazing, 

increase the period when green forage is available, and 

reduce the percentage of ungrazed plants. However, 

Savory’s system is controversial (Holechek and others 

2000, Sayre 2001), and some of the claimed benefits 

have not been proven—for example, increased water 

infiltration has not been found, and sedimentation is 

higher under this system than under more moderate 

grazing systems. Holechek and his associates (1998, 

2000) and Sayre (2001) provide a more complete 

analysis of HRM. They indicate that it may not be appropriate for arid grassland ecosystems because short 

growing seasons minimize the value of repeated periods 

of defoliation and nonuse. HRM may be best suited for 

monocultures of extremely grazing-tolerant grasses in 

subhumid environments, but it may not be suitable 

for lands where the goal is to manage for plant species 

diversity and variation in residual covers (C.H. Sieg 

2002, personal correspondence). Nevertheless, some 

ranchers have adopted the short-duration grazing and 

have been satisfied with the results (Sayre 2001).

It is usually accepted that short-duration grazing is 

not the same as Savory’s HRM, but many researchers 

have related their findings to HRM (Holechek and 

others 2000). Several studies have compared shortduration grazing to continuous grazing systems. 

A review of results indicates that the two systems 

produce similar results when stocking rates are 

equivalent with respect to forage production, plant 

succession and range condition, livestock productivity, 

and harvest efficiency (Holechek and others 2000). 

Short-duration/high-intensity grazing without the 

pasture rotation system has been used on some difficult restoration sites. For example, cattle have been 

grazed to rehabilitate mine tailing sites in central 

Arizona. Hay is often spread to encourage the cattle 

movements around the tailing site where they break 

up the tailing material by hoof action, fertilize and mix 

organic material into the substrate, and create small 

depressions to catch precipitation (Wheeler 1998). Hay 

also serves as mulch, reducing soil evaporation and 

soil surface temperatures. Wheeler (1998) claims that 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 153

of administrative BMPs include controlling livestock 

numbers and season of use, timing of construction 

activities, and designing roads to minimize stream 

crossings. Most State water quality regulatory agencies and Federal land management agencies maintain 

handbooks of water quality protection techniques and 

tools that can be used to prescribe BMPs.

Monitoring of BMPs is a critical part of the process. 

Monitoring should be done first to see if the BMPs were 

implemented as prescribed. Then, BMPs should be 

evaluated as to their effectiveness so that prescribed 

protection measures can be continually improved 

and BMP knowledge can be recycled to benefit future 

projects.

Road Management

Author: Bill Woodward, Southwestern Region

Grassland roads provide legal access necessary for 

administration and use of Federal, State, and private 

lands. Federal grasslands are often fragmented, and 

as a result, grassland roads are generally branches 

of State, County, and other road systems that serve 

as primary access roads to these areas. Land fragmentation usually results in significant increases in 

road densities across previously undivided landscapes 

(Mitchell and others 2002).

Most grassland roads are for use by high clearance 

vehicles. Road densities on Federal lands are often less 

than 2 miles per square mile (equivalent to having a 

road around each section of land). For most grassland 

activities, such as cattle grazing and hunting, this is 

adequate to meet management objectives. However, 

in areas where oil and gas operations are a primary 

activity, road densities as high as 6 miles per square 

mile may be necessary to maintain wells and related 

equipment. One study in eastern Colorado found road 

densities for subdivided grasslands were between these 

two values. Road densities on two subdivided ranches 

were 3.4 to 5.7 miles per section compared to between 

0.6 and 1.5 miles per section on two neighboring intact 

ranches (Mitchell and others 2002).

Existing roads and road construction and maintenance traditionally have been a concern in forest and 

grassland management. Older roads were often constructed along riparian corridors and across sensitive 

meadows and wetlands. Most of these roads were poorly 

designed, if designed at all, and had inadequate drainage. Poorly designed or constructed roads, properly 

designed roads that are not maintained, and roads with 

inadequate drainage and poor culvert design contribute 

significantly to erosion and sedimentation. Efforts 

to correct these problems through construction and 

maintenance have been hindered by budgets, priority 

safety items, and environmental considerations.

Today’s emphasis for grassland roads on Federal 

lands is to locate, design, construct, maintain, and 

manage to minimize erosion and sedimentation effects 

and to reestablish wetland and riparian areas. Offroad 

use is often prohibited or restricted to minimize creation 

of unneeded travelways and the spread of noxious and 

invasive vegetation species. Unneeded roads are to be 

decommissioned or converted to other nonroad uses. 

Road conditions and management emphasis on private 

lands, which often have heavy use, are variable.

Standard USDA Forest Service manuals are 

available to ensure proper design, construction, and 

maintenance of roads and for implementation of soil 

An illegal four-wheel drive vehicle 

road on steep slopes in the SP Crater 

area north of Flagstaff, Arizona. Such 

activities degrade vegetation, soil, 

water, and esthetic resources. (Photo 

by John Yazzie)
154 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

and water conservation practices within grassland 

communities. Similar manuals also are available 

from other government agencies and from educational 

institutions.

Recreation Management

Author: Rick Atwell, Southwestern Region

Pressures for increased recreational opportunities 

are increasing throughout the West as the region’s 

primarily urban populations continue to grow. While 

forests and associated lakes and streams are the primary recreational focus for many in the arid Southwest, 

grasslands also are being impacted by increased road 

traffic and offroad travel. Certain grasslands, especially 

mountain meadows, receive heavy use by the recreating 

public, and sometimes these sensitive areas must be 

protected from overuse.

One example: the measures used to protect the 

Kiwanis Meadow on the Sandia Ranger District of 

the Cibola National Forest. The meadow is used by 

hundreds of thousands of visitors who are traveling 

along the 1.5 mile trail between the Sandia Mountain 

Aerial Tramway and the Sandia Crest observation 

point. These two destinations receive the highest 

visitor counts for National Forests in northern New 

Mexico. The District employed a three-part plan. The 

first part was to reroute the popular Crest Trail from 

the meadow to its east side, and a second step was to 

build a buck and pole fence around the meadow. This 

provided a physical barrier to encourage visitors to 

stay on the established trails and not use the meadow 

for lounging or picnicking. The materials for the fence 

came from other areas of the District, often resulting in 

the creation of beneficial small wildlife openings. The 

third step was to place directional and informational 

signs around the area.

Integrated Weed Management

Author: Gene Onken, Southwestern Region

The introduction, adaptation, and spread of nonnative invasive plant species have become a serious threat 

to native grassland ecosystems of the Southwestern 

United States. Significant spread of invasive weeds 

in the Southwest has occurred relatively recently, 

especially during the past 15 years. Because of more 

xeric climate, harsh sites, and marginal soil productivity, the invasive weed problem reached serious 

dimension in the Southwest later than in other parts 

of the country. In the Southwestern Region of the 

USDA Forest Service, remoteness may also be a factor because the Region has a less dense road system 

than more highly developed agricultural production 

areas in other parts of the Southwest. These factors 

translate to a somewhat slower initial rate of weed 

introduction and spread.

Some of the most invasive of species evolved in 

Eurasia where the climate is similar to that of the 

Southwestern United States. Once these species were 

introduced here, the infestations spread rapidly. The 

weeds were and are often able to outcompete the native species for available moisture and space on the 

landscape. The overall problem is now highly significant 

both from an ecological and an economic perspective. 

To sustain the native grassland species, both integrated

and adaptive management actions are now required 

for managing the invasive species threat.

Integrated weed management (IWM) is a systems 

approach to managing undesirable plants. It is defined 

in the Federal Noxious Weed Act of 1974 (PL 93-626, 

January 1975, U.S.C. 2801-2814) as:

[A] system for the planning and implementation 

of a program, using an interdisciplinary approach, 

to select a method for containing or controlling an 

undesirable plant species or group of species using all 

available methods, including...education; prevention; 

physical or mechanical methods; biological control 

agents; herbicide methods; cultural methods; and 

general land management practices.

The concept of integrated weed management has 

already been adopted by most of the Federal and State 

agencies in the Southwest.

IWM involves using the best control techniques 

available for the target weed species. IWM requires a 

planned and coordinated program to limit the impact 

and spread of the weed. Control methods should be 

determined by: management objectives for the land, 

effectiveness of the control technique on the target 

species, environmental factors, land use, economics, 

policy and legal restrictions, safety to humans and 

the environment, and the extent and nature of the 

infestation.

An IWM approach for addressing invasive plant 

problems is particularly suited for an adaptive management strategy, which provides a way to describe 

and evaluate the consequences of dynamic and rapidly 

changing invasive plant populations on the landscape. 

Within a rather short time, invasive weeds spread 

rapidly from existing infestations into new locations. 

New species also may become established in any 

given locality. New technologies for weed control are 

continually being developed as new biological agents 

and herbicides become available.

An adaptive management strategy requires sitespecific explanations of what actions the land managing 

agency or landowner will take under various conditions and what the environmental effects will be for 

those weed control actions. Weed management actions 

are continually reevaluated by land managers as 

monitoring indicates changes on the landscape and 

as new control options are developed. This reevaluation is done within the framework of the original 

integrated weed management plan. Then the next 

control actions are appropriately adjusted or adapted 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 155

to the new and changing conditions and technologies. 

Thus, an adaptive management strategy is a continuing process of:

• Management action

• Monitor the results

• Evaluate the changes

• Adapt the treatment plan

• Implement the next (adapted) management 

action

• Monitor the results

An adaptive management approach shows how 

weeds will be treated without listing individual species or individual sites. Managers should specify an 

approximate number and/or percentage of treatment 

acres across a geographic area. This estimate forms the 

basis for predicting the environmental consequences 

and for defining the various weed management strategies.

Now that nonnative invasive species have become 

firmly established within the native grassland ecosystems of the Southwestern Region, aggressive and 

continuing eradication or control of new or existing 

nonnative invasive populations or infestations must 

be part of management to prevent spread or to restrict 

unacceptable impacts on existing native plant communities. Economic, environmental, or legal constraints 

may prevent eradication or the desired degree of 

invasive weed control. Therefore, implementing integrated weed management practices under an adaptive 

management strategy provides the most practical 

approach toward addressing this problem.

In summary, invasive plant species will continue to 

impact our landscapes and place our natural grassland ecosystems at risk. A perpetual expenditure of 

effort and resources will be required to sustain the 

grasslands.

Remote Sensing, GIS Applications, 

and Database Management_________

Remote Sensing

Author: Bill Krausmann, Southwestern Region

Remote sensing has a demonstrated potential to aid 

grassland managers in maintaining long-term viability 

of the resource. Remote sensing can be defined, in 

this context, as the analysis of grassland responses to 

electromagnetic radiation as collected and recorded in 

an image format. Examples of applications in remote 

sensing that deal with range management aspects of 

grassland ecosystems date back well into the 1930s 

(Tueller 1989).

The types of imagery and imagery-based analytical 

processes that can be applied to grassland management have grown exponentially over the past 25 

years. Several forms of satellite imagery are currently 

available for analysis, in addition to various forms of 

aerial photography—a standard tool for managers. 

New satellite imaging systems are being brought on 

line each year, further increasing the potential value 

of remote sensing as a management tool. The marked 

increase in available imagery and the development 

of new techniques are fortuitous because of a need to 

augment standard inventory and monitoring methods 

to meet current requirements placed on resource 

management agencies.

Table 8-1. Map scale, vegetation, and appropriate imaging technology.

Vegetation Example

Map scale Potential Existing Potential Existing Image data type

1:1,000,0001 Class Lifeform Close forest Coniferous forest AVHRR

Subclass Lifeform Mainly evergreen Coniferous forest

1:500,000 Group Lifeform Temp. evergreen Coniferous forest

 needleleaf

1:250,000 Formation Structural stage Pine forest Seedling/sapling Landsat MSS

1:126,720

1:100,000 Series Cover type Ponderosa pine Landsat TM

1:50,000 Subseries Domance type Ponderosa pine/ SPOT XS

 gambel oak SPOT Pan

Association Community type Ponderosa pine/ Ponderosa pine/ Digital camera

 Arizona fescue/ mutton bluegrass videography

 gambel oak 

 phase

Air photos
156 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

Vegetation Mapping—Managers are interested 

in the type, distribution, and condition of grassland 

vegetation, and for grazing analysis, the forage base as 

it occurs across space and time. While remote sensing 

techniques have been utilized since the 1960s to map 

parameters of vegetation, the number of useful and 

fully applied procedures or techniques is quite low 

(Tueller 1989).

The level of detail that can be mapped within vegetation communities is directly related to the scale of 

a given project. Selecting an image data source that 

is appropriate for the scale of the mapping project is a 

critical element in the successful application of remote 

sensing in grassland management (table 8-1).

Aerial photography has long been a proven tool for 

mapping vegetation. Research has shown that photography acquired at a scale of 1:10,000 is optimum for 

mapping vegetation at the ecological site level (Tueller 

1979). As photo scale becomes smaller, the range of 

vegetation units that can be detected becomes more 

general, as indicated in table 8-1. It is inappropriate, 

for example, to use Landsat Thematic Mapper imagery 

for a project where mapping at a species association 

level is of interest. On the other hand, mapping series 

level vegetation across an entire National Forest using aerial photography could be done but would be 

extremely laborious and expensive.

Computer mapping of vegetation types revolves 

around the processes of supervised and/or unsupervised 

image classification. In the supervised approach, areas 

of homogeneous training sites of known vegetation 

type are identified on the imagery. There are typically 

several training sites per class of vegetation. Statistics 

derived from the spectral response of the pixels within 

the training sites are used by one of several classification algorithms to separate the image into specified 

vegetation classes.

Generally, unsupervised classification seems to 

work best on grasslands (Tueller 1989). The analyst 

provides the computer with the number of clusters to 

be derived and parameters that drive cluster merging 

and splitting in the unsupervised process. Pixels in 

the image are then clustered into groups with similar 

spectral response by the computer. The groupings that 

are generated in this process are called spectral classes. 

They represent areas with similar spectral response. 

It is the analyst’s responsibility to develop information classes from the spectral classes. The classes 

are developed by building relationships between the 

spectral classes and areas of the surface with known 

vegetation cover. Information classes can represent, 

among other things, plant communities, grazing allotments, or range improvements.

Mapping vegetation in rangelands, including 

grasslands, using satellite data is a difficult process 

in the Southwest. Several problems combine to reduce 

classification accuracies. Some of these problems 

include high soil background response, spatially 

heterogeneous precipitation patterns, spatially heterogeneous grazing patterns, and timber-covered range 

allotments. On average, for classifications produced 

at the vegetation series/sub-series level (table 8-1), 

overall classification accuracies between 65 and 75 

percent could be expected.

Change Detection—The detection of change over 

time in grasslands and other rangelands is perhaps 

the most significant application of remote sensing to 

range management. Remote sensing instruments can 

produce imagery at scales from the site level to subcontinental areas that can be used to detect fluctuations 

in productivity.

Digital change detection involves using a computer 

to compare the spectral response of two or more 

images acquired on different dates. The comparison 

is performed at the pixel level and requires that the 

imagery data sets be accurately coregistered. There are 

several examples of change detection methods applied 

to range management issues (Chavez and MacKinnon 

1994, Knight 1995, Pickup and others 1993, Ringrose 

and others 1999, Wallace and others 2003). Several 

algorithms have been developed for change detection, 

and a review of methodologies exists (Singh 1989), 

as does an assessment of Landsat Thematic Mapper 

imagery for change detection (Fung 1990). The most 

common change detection methodology is image differencing. In image differencing, change between two 

images is highlighted by subtracting one image’s pixel 

values from the other.

Other Applications—Remote sensing has been 

used to monitor and evaluate grazing management. 

Pickup and Chewings (1988) used Landsat imagery 

and animal distribution models to estimate the distribution and grazing pattern of cattle on large pastures 

in Australia. Utilization levels of 25 to 40 percent are 

recommended in the semiarid Southwest (under 300 

mm of precipitation annually) (Holechek and others 

1998). Acquiring information on utilization levels 

requires considerable fieldwork. Repetitive aerial photography, videography, and digital camera imagery can 

provide baseline data on range readiness, utilization, 

livestock distribution, or other parameters that may 

reduce the amount of field work required to support 

management decisions.

Soil loss from gullying, overland flow, or eolian 

processes is a significant concern of grassland managers. Researchers have mapped soils, assessed soil 

loss, and measured gully erosion (Pickup and Nelson 

1984, Westin and Lemme 1978) primarily using various 

forms of aerial photography.

Conclusion—In 1989, Tueller described the future 

of remote sensing applications in range management 

as “hazy” (Tueller 1989). The same could probably be 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 157

said today. While research continues on applications of 

remote sensing to rangeland management, much of it 

relates to satellite monitoring of large, semiarid regions, 

and the questions being addressed are subcontinental 

in nature. This work does not address the needs of 

range managers working with several allotments on 

a USDA Forest Service Ranger District. If remote 

sensing is to become a standard management tool, 

procedures that directly address range readiness and 

utilization in a cost-effective manner at large scales 

must be developed.

New tools such as MODIS (a 36 band imaging radiometer) and the digital camera, combined with Landsat 

Thematic Mapper imagery and aerial photography, may 

provide opportunities to better monitor rangelands, 

including grasslands, at both the macro and micro 

scales. It is hoped that within 20 years much of the data 

required by managers may be derived from imagery.

Geographic Information System as a Tool 

for Managing Grasslands

Author: Pat Frieberg, Southwestern Region

The emergence of Geographic Information System 

(GIS) technology in the past few years has provided 

the grasslands manager with a decision support 

system that facilitates the decisionmaking process 

when addressing management issues. GIS provides 

the analytical capabilities for spatial data mapping, 

management, and modeling.

GIS is defined as a collection of computer hardware, 

software, geographic data, and personnel that are 

designed to hold, manipulate, analyze, and display 

all forms of geographically referenced information 

(Environmental Systems Research Institute 1995). 

GIS can address the impacts of multiple variables 

simultaneously—for example, to determine the capability of a grassland to sustain livestock grazing. Using 

basic layers of slope, aspect, vegetation, and water 

sources, along with certain known habits of livestock, 

one can model the grazing pattern and subsequent 

ability of the grassland to sustain grazing. Once 

the forage production and utilization information is 

measured and collected in the field, it can be entered 

into the model. The model then calculates the carrying capacity based upon actual use, and the criteria 

are entered into the model. The Apache-Sitgreaves 

National Forest in east-central Arizona is an example 

of a Forest that has used GIS for this purpose. The 

Range Management Staff developed a model using 

forage production, soil stability, distance from water, 

and steepness of slope to determine grazing capacities 

for allotments scheduled for new or revised allotment 

management plans.

Using the relationship between forage production 

and forest overstory density, a GIS map depicting 

herbaceous forage production classes for each allotment 

was developed and field checked. The result was an 

estimate of forage production in pounds per acre for 

each class. A percentage of allowable use for each of 

these forage production classes was established based 

upon range condition and management strategy. GIS 

maps showing limitations for soil stability, distance 

from water, and slope gradient were generated and 

used to further refine the percent allowable use in each 

forage production class. When the refined allowable 

use percentage was multiplied by the forage production in pounds per acre, the result was the pounds of 

forage available for consumption. This was multiplied 

by the GIS acres in each forage production class, then 

the total pounds of forage available for consumption by 

wildlife and livestock were computed for each pasture, 

allowing the range manager to estimate the livestock 

capacity for each pasture. The capacity of all pastures 

in the allotment was added to establish a capacity for 

the allotment.

GIS can be used for virtually every facet of resource management. Providing a spatial inventory 

of different aspects of the grassland resource, such 

as the location of noxious weeds, vegetation types, 

roads, and so forth, is a common use. Combining the 

database and mapping capabilities within the GIS 

system permits a variety of analyses to be performed, 

for example, analyzing counties and communities by 

economic characteristics. A study in the Sevilleta 

Long Term Ecological Research Program (LTER) 

is examining prehistoric and historic land use and 

exotic plant invasion by overlaying vegetation maps 

with archaeological site maps. Typically, GIS data are 

linked to large databases for decision support, such 

as the Forest Service’s NRIS and INFRA databases 

and the Bureau of Land Management’s Rangeland 

Information System (RIS). Remote sensing data can 

also be incorporated into GIS layers for analysis and 

for change detection purposes.

GIS data are increasingly being linked with modeling 

programs to create management applications. Some 

examples of the variety of applications:

• FRAGSTATS, a computer software program designed to compute a wide variety of landscape 

metrics for map patterns.

• The Agricultural Research Service’s Arid Basin 

(ARDBSN) model to predict the amount of runoff 

resulting from rainstorm events.

• Predictive wildlife habitat models for current 

management or for potential reintroduction—

for example, mountain lion (Felis concolor), 

black-tailed prairie dog (Cynomys ludovicainus 

arizonensis), and aplomado falcon (Falco femoralis septentrionalis).

• Texas A&M’s PHYGROW, a hydrologic based 

plant growth simulation model using soil 
158 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

characteristics, plant community characteristics, 

and weather data for a particular location to 

predict forage production.

GIS can also be an important tool in monitoring, 

not only to provide a spatial context to the monitoring effort but also to select sampling sites based on 

specific criteria, analyze changes in parameters over 

time, and for additional modeling to assess the temporal and spatial variability of ecosystem processes 

in an area.

Forest Service Corporate Databases That 

Cover Grasslands

Author: Reuben Weisz, Southwestern Region

The USDA Forest Service is currently investing in 

four major database projects that may be useful for 

grassland assessments:

• The Geographic Information System (GIS) Core 

Data Project

• The Automated Lands Project (ALP)

• The INFRASTRUCTURE (INFRA) Project

• The Natural Resources Information System 

(NRIS) Project

These database projects provide a set of interrelated 

databases and computer applications containing spatial (map) and tabular (numbers) data, collected and 

recorded in a consistent manner. The GIS Coordinator 

for a particular grassland area will be able to provide 

more details about applications of these databases.

Briefly described, the GIS Core Data Project is putting into place, “wall to wall” across all Forest and 

Grasslands boundaries, 15 standard GIS layers. These 

contain the minimum data required to do business 

everywhere in the Forest Service. Broadly speaking 

there are three categories of GIS layers:

1. Land Information Layer—The ALP database 

manages the spatial and tabular data associated 

with information about ownership, jurisdiction, land surveys, and restrictions and rights. 

Examples of ALP data include information about 

who owns a grassland area and what restrictions 

and rights apply to it.

2. Constructed Features Layers—The INFRA 

database manages spatial and tabular data 

describing those things in the ecosystem that 

are constructed or created by people. Examples 

of INFRA data include information about roads, 

trails, allotment and pasture boundaries, allotment and permit management information, and 

range improvements.

3. Natural Resource Layers—The NRIS database manages the spatial and tabular data about 

those parts of ecosystems that occur naturally 

such as air, water, terrestrial ecologic units, 

existing vegetation, threatened and endangered 

species occurrences, topography, water and 

watersheds, and the human dimension. In a 

given grassland area, this might contain useful 

information about water quality, water uses and 

water rights, threatened and endangered species, 

and watershed condition.

GIS Applications for Wildlife Management

Author: Bryce Rickel, Southwestern Region

GIS applications have been expanded to provide land 

managers with tools specific to wildlife management. 

Two examples are the Southwest Wildlife Information 

System and the Habitat Quality Index.

Southwest Wildlife Information System 

(SWIS)—This is an ArcView application that has 

been developed to provide field personnel with an 

easy way to search, query, and analyze basic wildlife 

species habitat relationships across landscapes. SWIS 

allows the user to perform species/habitat tabular and/

or spatial searches and queries using ArcView. The 

focus of the current application is providing information at the Forest and project levels. SWIS also allows 

the user to display Arizona and New Mexico Natural 

Heritage data.

Habitat Quality Index (HQI)—This model is an 

ArcView application that allows Forest and District 

biologists to develop their own species habitat quality 

models. The primary modeling approach employed 

is simple and has been used for almost two decades. 

The idea behind the model is that a particular habitat 

type, per season, has certain cover and forage values for a particular wildlife species. An HQI model 

for a species will produce a GIS map with habitat 

qualities for each habitat type (or polygon) across a 

landscape.

Database Use to Assess Effects of Grazing 

on Southwestern Biodiversity: An Example

Authors: Curtis Flather and Patrick Zwartjes, 

Rocky Mountain Research Station

The decisions the Forest Service has made regarding the use of public lands for livestock grazing have 

become a contentious issue in the Southwest, resulting 

in many of these decisions being challenged (particularly through litigation) by a variety of parties with 

competing interests (for example, environmentalists 

and cattle ranchers). The Forest Service has recognized 

that National Forest managers and biologists have 

little information on the impact of grazing (both by 

permitted livestock and by native ungulates) on the 

various animal and plant species, as well as the overall 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 159

biodiversity, found on National Forest System lands. 

To improve management decisions regarding species 

that may be at risk from permitted and/or native 

ungulate grazing, the Forest Service’s Washington 

Office entered into a cooperative project with the 

Forest Service Southwestern Region (Region 3) and 

the Rocky Mountain Research Station (RMRS). The 

objectives of this project are to:

• Describe and classify habitat types on grazed 

lands in the Southwest.

• Determine species that are sensitive to grazing.

• Determine source habitats for sensitive species.

• Develop distributional maps and ecological/life 

history summaries for each species.

• Describe the effects of specific grazing management regimes on these species and their 

habitats.

Scope, Organization, Goals—Data acquisition 

for this project is focused on the compilation of existing information, both at a landscape level (including 

species distributions and general habitat associations), 

and at the individual species level (which includes data 

on specific habitat requirements and other pertinent 

ecological information). The geographic scope includes 

all land areas (not just National Forest System lands) 

in Arizona and New Mexico, with a taxonomic scope 

that includes select species of terrestrial vertebrates, 

aquatic vertebrates, and plants.

The responsibilities for the different objectives have 

been assigned to three units within the RMRS:

• Fort Collins: Characterize regional biodiversity 

patterns; develop models of species occupancy.

• Albuquerque: Identify species sensitive to grazing; develop species accounts for each with detailed 

habitat, ecology, and life history information.

• Flagstaff: Identify broad vegetative zones in the 

region; assess impacts of grazing on these zones 

as well as on select individual plant species.

In general, the goals of this project are to provide 

fundamental information on species distribution, occurrence, ecology, and life history for use by range and 

forest managers. It is important to note that this study 

is not intended to be a regional viability assessment, 

nor is it an assessment of Region 3’s grazing program; 

rather, it is a tool to assess the habitat needs and 

vulnerabilities of individual species.

Geospatial Analyses—The objectives of this 

segment of the project are based on a geographic approach, one that will develop a geospatial database of 

species occurrence information as well as geographic 

information on land cover and land use throughout 

the Southwestern Region.

The format for this database is an ArcInfo 8/Microsoft 

SQL relational database, distributed as ArcInfo coverages. The content is divided into three map types:

• Species Occurrence Maps—Two general categories: (1) maps using species point observation 

data, collected from museum collection records, 

survey data (for example, USFWS Breeding Bird 

Survey), and biological atlases; and (2) predicted 

occurrence range maps, based on habitat associations, and produced by the National Gap Analysis 

Program. Of particular importance will be the 

capacity to generate species lists by spatial queries, such as within a particular National Forest 

or grazing allotment.

• Biodiversity Maps—Overlays geographic 

information for different species within a broad 

grouping (for example, birds, reptiles) to estimate 

species richness and relative biodiversity among 

geographical areas. Potential applications include 

identifying areas with the greatest number of 

species, and ranking different areas with respect 

to species richness.

• Base Layer Maps—Includes maps of topography/

digital terrain, land use, land cover/vegetation, 

administrative boundaries, and so forth.

The main product for this project will be a Rocky 

Mountain Research Station General Technical Report 

(GTR) that will detail the methods and procedures 

utilized for this analysis, as well as the results of an 

analysis of biodiversity patterns in the Southwestern 

Region. The GTR will include a CD-ROM containing 

software and map files for analysis by the users. This 

CD-ROM will contain more than 30 GIS coverages, 

range maps for more than 800 species, more than 

500,000 point locations for individual species, analysis 

algorithms, and a simple user interface.

Vertebrate Species Accounts—The project was 

designed to examine all terrestrial and aquatic vertebrate species in the Southwestern Region, and (1) 

identify those species with the greatest potential to be 

negatively impacted by grazing, and (2) collect detailed 

habitat and life history information into individual 

accounts based on information from both the published 

literature and the expertise of vertebrate zoologists 

working in Arizona and New Mexico.

Panels of zoological experts were assembled according to several broad categories of vertebrate taxa:

• reptiles and amphibians

• grassland-desert scrub birds

• woodland birds

• riparian birds

• small mammals

• carnivorous and ungulate mammals
160 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

The complete list of participants will include more 

than 40 zoological experts from Arizona and New 

Mexico, including university scientists, Federal and 

State scientists, wildlife experts, and private experts 

and consultants.

The panel process was organized into two stages, with 

an intervening period during which the Albuquerque 

RMRS developed draft species accounts based on the 

published literature:

• Stage I—The panel of experts considers and 

discusses the ecological and habitat needs of each 

species, followed by each panelist recording their 

assessment of the direction, magnitude, and likelihood of grazing impacts on this species. Species 

were selected for further consideration if at least 

one of the panelists determined the species had 

at least the potential to be severely negatively 

impacted (regardless of likelihood), or that the 

species is known to be negatively impacted (but 

of unknown magnitude).

• Stage II—Draft species accounts, developed for 

all species selected in stage I, are reviewed by 

panel members, who fill in gaps in information 

from the published literature, suggest additional 

sources of information, and make selections from 

a series of menus that contain habitat and life 

history descriptors that serve to characterize the 

ecological needs of each species. In addition, the 

stage II panel has the authority to either remove 

species from the database that were selected in 

stage I, or to add species, which they consider to 

be erroneously excluded.

The final version of the species accounts will be 

based on the inputs of the panelists in combination 

with the information from the published literature. The 

database of accounts will be contained on a CD-ROM 

in Microsoft Access format, with a user interface that 

allows for searches based on species names, or queries 

that will generate a species list based on shared habitat 

descriptors in the menu selections, such as general 

habitat association, specific attributes of habitat (tree 

heights, grass densities, and so forth), season of use, 

grazing effects, and others. Individual accounts will be 

able to be viewed on screen, or printed out individually 

according to the needs of each user. In addition, the 

CD-ROM will contain a ProCite file of all literature 

cited, and the Albuquerque RMRS will deliver to Region 

3 the complete collection of literature used to develop 

all species accounts. We anticipate the final product to 

contain more than 300 terrestrial vertebrate and more 

than 50 aquatic vertebrate species, with a literature 

database of more than 2,000 entries. As of publication of 

this Assessment (2004), the species accounts database 

is well under construction, with a large percentage of 

accounts completed.

Progress and Schedules— The geospatial analyses and the panel process were largely completed by 

the start of Federal government fiscal year 2003. 

Development of the GTR, user interfaces for the CDROMs, and editing and completion of the individual 

species accounts are continuing as of the publication 

of this Assessment. The complete package should be 

available to forest managers and biologists as a new 

management tool by 2006.

Grasslands Education and Communication 

in the Southwestern Region

Author: Jean Szymanski, Southwestern Region

Education for public and private land managers and 

for the interested public is vital for sound grassland 

management in the Southwest. Managers need the 

knowledge to be good stewards of their lands and 

to conduct ecologically and, especially for private 

landowners, financially sound treatments. Educated 

ranchers and land managers can help with many 

activities on their own or on leased grasslands. One 

example would be public agency managers who understand how to monitor grassland conditions, thus 

allowing educated decisions on livestock stocking 

levels and rotations and providing a basis for discussions on permits with the public, who may question 

their management decisions. Education of teachers, 

especially at the high school and university level, 

often pays large dividends.

The Coconino National Forest has a session on 

rangelands as part of the Arizona Natural Resource 

Conservation Workshop for Educators program. In 

this session, participants lay out transects to measure 

plant composition and amount of ground cover. The 

emphasis in this program is on rangelands as a type 

of landscape with many uses, not just grazing.

The Cuba Soil and Water Conservation District is 

the lead sponsor of the New Mexico Forestry Camp; 

the Forest Service is one of the many co-sponsors, with 

involvement from the Regional Office and the Santa 

Fe, Cibola, and Carson National Forests. Two sessions 

on grasslands are conducted at this weeklong camp 

for New Mexico youth (13 through 18 years old). One 

session focuses on identification of forbs and grasses, 

and the other focuses on grazing studies, soils, noxious 

weeds, and weed control methods.

The Malpai Borderlands Group conducts three 

“Ranching Today” workshops on the Borderlands near 

Douglas, AZ. These workshops are 3 to 5 days long 

and attended by ranchers, the public, and Nature 

Conservancy members. People from throughout the 

Western United States often attend these workshops to 

learn if the Malpai model would work in their areas. The 

workshops discuss examples of grassbanking, private 

land conservation easements, ongoing research, and 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 161

endangered species. The Coronado National Forest 

participates in these activities. The Malpai Borderlands 

Group also supports activities at Douglas High School 

related to the supervised breeding of endangered 

Chiricahua leopard frogs (Rana chiricahuensis), and 

supports local students who serve as summer interns 

on sponsored research studies.

Most of the Western State sections of the Society for 

Range Management hold weeklong summer camps for 

high school students to learn about rangeland management. The Arizona Section has conducted camps in 

the State since the 1960s. Camps have been held at 

the Sierra Ancha Experimental Forest or at Mormon 

Lake.

The Cibola National Forest, as a partner with the 

Playa Lakes Joint Venture, has access to several 

Grassland Education “trunks” to use as a teaching 

tool. The trunks contain books, tools, and videos to aid 

in educating about grasslands. The Kiowa National 

Grasslands has been involved in outdoor education 

days, where countywide third graders are brought out 

to the National Grassland and given environmental 

instruction about vegetation. The National Grassland 

also serves as a site for workshops for landowners on 

grazing management and riparian enhancement.

The Washington Office of the USDA Forest Service 

is producing a series of posters, similar to the Smokey 

Bear poster series, that relate to grassland ecology 

and management. These posters will be distributed 

throughout the country. Some topics are:

• Grassland fire ecologies

• Rainfall in grasslands

• Noxious weeds

• What keeps a prairie a prairie?

• State grasses

• The cycle of life on a grassland

• The anatomy of grass

• The function of grass roots

Other Federal agencies have education programs 

related to grassland and general range management. 

The USDA Agricultural Research Service in Boise, 

ID, incorporates students in its watershed program 

to provide an educational experience and to broaden 

their awareness of the biological and physical processes, 

landscape attributes, and social and economic factors 

that affect viability of range management (Northwest 

Watershed Research Center, USDA ARS, undated).

Educational activities also include measures designed to inform the public about grassland ecology 

and management decisions. The Kiwanis Meadow 

near the Sandia Crest, north of Albuquerque on the 

Cibola National Forest, receives heavy recreational use 

because of its proximity to the Sandia Tram. Recent 

efforts to protect the alpine meadow by rerouting the 

trail have included the placement of directional and 

informational signs. The directional signs show visitors 

the new routes and where traffic is prohibited, and the 

informational signs explain why the Forest undertook 

the project. Signs that explained that meadow protection benefited wildlife were especially effective.

Acknowledgments_________________

I greatly appreciate the technical reviews by E. 

J. Alford, Tonto National Forest, Phoenix, AZ; R. J. 

Bemis, USDA Natural Resources Conservation Service, 

Douglas, AZ; B. McDonald, Malpai Borderlands 

Group, Douglas, AZ; and C.H. Sieg, Rocky Mountain 

Research Station, Flagstaff, AZ. Their comments 

and suggestions greatly improved the manuscript. 

The reviews by Rex Pieper and Wil Moir were helpful 

and appreciated.

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164 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

As part of the assessment effort, the team hosted a 

series of conference calls to supplement the information 

that had been collected and to identify any gaps in the 

assessment report. Persons participating in the conference calls included USDA Forest Service Southwestern 

Regional Office program managers, National Forest 

planners and resource staff officers, District Rangers 

and resource specialists, researchers, and academia. 

Participants were asked nine discussion questions, 

and they provided their answers either during the 

calls or through written responses. This appendix is 

a summary compilation of all the responses for each 

question and includes only the key points and not the 

supplementary detail.

1. What are the existing issues that you are 

dealing with in Southwestern grasslands?

• Fragmentation of grassland ecosystems, both biologically and politically (fragmented ownership), 

which increases the complexity of their management. We are losing grasslands to development as 

human communities expand into open space.

• Identifying desired composition and disturbances 

for grasslands (integrating social and economic 

factors into a landscape-level desired condition). 

There is the perception that grasslands have less 

value than forests.

• A lack of understanding of grassland ecology, the 

role of grazing in grassland ecosystems, and how 

grasslands respond to climatic variation.

• Monitoring is an issue due to a lack of consistency, 

a lack of resources (both people and funding), 

training challenges, and not having permanent 

photo monitoring points. The lack of personnel 

affects the level of management as well.

• A reduction in acreage and vegetation diversity 

of semidesert grasslands as a result of woody 

species encroachment and a shift from perennials to annuals. The increase in woody species 

is throughout all grassland types, and most 

respondents attributed the increase to overgrazing and lack of fire.

• A decrease in long-term soil productivity (assessed 

by Region 3 soil condition protocol) in regards 

to porosity, nutrient cycling, compaction, and 

erosion. Some unsatisfactory soils may never be 

able to be restored.

• A decrease in productivity in montane meadows, 

loss of meadows from forest encroachment, change 

of species composition (increase in Kentucky 

bluegrass), increase in elk populations causing 

damage (particularly when using winter range), 

lowering of the water table, and impaired or unsatisfactory soil condition. Some of these changes 

are due to livestock and elk grazing and recreation 

impacts such as RV camping. Soil recovery takes 

longer than vegetation recovery. Higher elevation 

meadows typically have better ground cover and 

are more resilient than lower elevation meadows, 

where the vegetation is more brittle and the soil 

quicker to erode.

• Maintaining the diversity of plants and animals 

in grassland ecosystems. Many grassland wildlife 

species appear to be declining in numbers and 

distribution. Some key species of concern include 

the Sonoran pronghorn, Mountain plovers, 

Northern Aplomado falcon, lesser-prairie chicken, 

black-tailed prairie dogs, and native nongame 

fish. Southwestern grassland ecosystems have 

received less attention than other ecosystems so 

it is probable that many other species warrant 

concern. Single-species management often creates 

conflicts in management activities and hampers 

our efforts to manage in an ecosystem context.

• Controversial management affecting grasslands 

include grazing, roads, and offroad vehicle use.

• Reintroduction of natural fire regimes. Concerns 

include smoke, public fear of fire, differences in 

landowners’ views of fire, and difficulty in controlling grassland fires.

Appendix:

Discussion Question Summary
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 165

• Effects of drought, particularly extended droughts 

of several years.

• Spread of exotic or weed species into Southwestern 

grasslands.

• Land grant and Native American claims to land 

in northern New Mexico have resulted in difficult 

social and political issues, particularly grazing 

issues.

• Some conversions from shrub to grasslands in 

northern New Mexico done in the 1960s and1970s 

have since reverted.

• Lack of information, particularly habitat needs for 

designated threatened, endangered, and sensitive 

(TE&S) species.

2. What do you foresee as emerging issues in 

Southwestern grasslands?

• Increased need and pressure for regional planning 

to address issues.

• Urban encroachment, with ranchers selling their 

properties for development. Increasing wildland/

urban interface conflicts will take many forms 

(fire concerns, an increase in nonnative plants, 

livestock conflicts, free-ranging dogs).

• Wildlife viability. The National Grasslands are 

becoming a repository for species declining on 

private lands. There is the potential for reintroductions of some species. White-tail prairie dogs 

may be in a more serious decline than black-tailed 

prairie dogs.

• Ecological viability is being threatened due to 

accelerated loss of soil productivity.

• Increasing populations of exotic or weed species.

• Greater public interest and demand for access 

to grassland areas, while being faced with limited recreation and public access facilities and 

infrastructure.

• Managing grasslands to reduce the risk of 

catastrophic fire. Using prescribed fire as a management tool is a challenge because of air quality 

concerns and the difficulty of fire planning with 

a variety of landowners.

• The drying of montane meadows from lowered 

water tables and loss of ground cover.

• Management tools: Best Management Practices 

(BMPs) for grazing, grassbanking, fencing, information management, and modeling.

• Implementing the Clean Water Act to address 

impaired streams.

• Recreation impact in grasslands, particularly 

in montane meadows. Offhighway vehicles are 

causing compaction of the soil.

3. What are the highest priority research 

needs for Southwestern grasslands?

• Establishment of reference areas and determination of reference conditions, not only flora but 

also fauna (such as Mirriam’s elk). Identification 

of historic grassland distribution and conditions 

including disturbance regimes (fire frequency, 

ungulates, prairie dogs).

• Increasing depth of understanding of grassland 

ecology—disturbance patterns, the mechanism of 

recovery, effects of fragmentation on ecosystem 

function and plant species distribution, effects 

of changing climate (particularly drought), succession, linkages between factors, timing and 

density of utilization, the interrelationships 

between invasive and native plants, long-term 

implications of diversity, key indicators of 

grassland health, potential natural plant communities, the role of pollinators today relative to 

their historic role—and how they vary between 

ecotones. A study on what is causing the drying 

out of montane meadows.

• Understanding the tradeoffs between different 

composition mixes to help in the determination of 

desired conditions, and how/when to use disturbance events to move toward these conditions.

• Specific quantitative attributes for individual 

TE&S species habitat requirements. Identifying 

and understanding the role of small mammals in 

grassland ecosystems. Understanding of prairie 

dog habitat needs and reintroduction factors 

such as minimum numbers and distributional 

patterns. Social science research on dividing 

fisheries between sport fish and native nonsport 

fish, and fish ecology in general.

• A greater depth of understanding of the effects of 

management (grazing, timber, fire) on grassland 

ecology and priority species of concern (prairie 

dogs, mountain plovers). Determinations of the 

density of surface activities, such as the density 

of oil wells and their infrastructure. A study of 

the effectiveness of techniques to keep elk out 

of meadows.

• A values survey to capture the social aspects of 

why people think grasslands are special.

• Development of grassland ecosystem models and 

the understanding of the vegetation dynamics 

that feed the models.

• Exploring research opportunities on Tribal 

lands.
166 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004

4. How can we achieve sustainability of the 

grasslands resource while addressing the 

wants and needs of people/communities?

Comments ranged from the general (understand 

and apply broad-scale ecologically sound approaches) 

to the specific (exclude cattle for 10 years, calculate 

stocking rate under the correct carrying capacity, then 

reintroduce fire and cattle). Conserving larger blocks 

of grasslands would make landscape-scale management possible, which is particularly important for 

wildlife habitat needs. Understanding the historic 

and present-day ecology of grasslands is necessary to 

develop desirable ecological structure, function, and 

processes that would sustain grassland resources and 

the needs of people. A public education program would 

help people understand how their needs and wants 

affect sustainability. Collaboration between academia, 

managers, and landowners is crucial. One mechanism 

to bring diverse interests together is through monitoring training.

5. What are the changes, both short term and 

long term (over 100 years), to Southwestern 

grasslands? Which changes are attributable to 

human influences and which are not?

• Fragmentation of larger contiguous grassland 

areas. Decrease in the amount of grasslands, as 

can be seen from historic photos. Development 

occurring at the landscape scale. Habitat fragmentation (for example, pronghorn antelope).

• Simplification of grassland types. Loss in quality 

of the grasses.

• Interruption of natural processes such as fire 

and water (flow patterns and the availability of 

surface water and groundwater), and a reduction 

in the nutrient and energy cycles from herbage 

removal. Drying out of montane meadows.

• Unsatisfactory soil conditions in many grassland areas and continued loss of soil and soil 

productivity.

• Increasing exotic species, particularly plants (accelerating). Some of these can not be eliminated, 

yet they change fire frequencies, wildlife needs, 

and so forth.

• Increasing juniper, pine, and shrubs in some 

grassland areas (rate of increase is slowing).

• Declining water tables and down-cut channels.

Where identified, changes were attributed to human 

influences with the exception of the increase in woody 

vegetation, which was viewed by some participants as 

a combination of human influence and climate.

6A. What do we need to be doing differently in 

the management of Southwestern grasslands?

• Implement larger scale management across 

regions. Continue broad-scale application of 

management based on ecological understanding 

at both local and regional scale. Having a separate 

grassland plan for each forest recognizes the 

uniqueness of grassland ecosystems.

• Take the time to develop relationships. 

Memoranda of understanding (MOU) are occurring at a larger scale. This takes dollars and 

people’s time, and the ability to travel. Pursue 

cooperative and coordinated funding sources.

• Take an ecological approach rather than managing species by species as they are listed. Revise 

Forest Land Management Plans to reflect both 

an ecological approach while addressing individual species of concern. Consider revision of 

the Mexican Spotted Owl Recovery Plan as it 

applies to marginal forested habitats and historical savannah ecosystems.

• Develop staff working on the ground so they can 

effectively manage and monitor the resource. 

Find better ways to monitor grasslands given 

our dollar constraints.

• Improve management: Ensure plant recovery 

takes place each year rather than focus exclusively 

on grazing regimes. Manage offroad vehicle use. 

Monitor prescribed burning activities more closely. Address smoke management issues. Protect 

stable soils. Pursue more active restoration—not 

just a reduction in livestock numbers. Use more 

of an adaptive management approach rather than 

a “punitive” management approach.

• Recognize and incorporate the increasing emphasis and dependence on ecologically sound nature 

tourism to replace some of the failed economic 

efforts of the past.

• Increase public education programs on the value 

and role of grasslands.

• Map existing grasslands and compare to historic 

photos. Identify the desired pattern of grasslands 

within the Southwestern landscape. Establish 

photo points and frequency transects for monitoring.

• Work with the State Game and Fish Departments 

to develop realistic elk numbers, particularly in 

drought years.

6B. What do we need to continue doing?

• Take an ecological approach to management.

• Monitor and study prescribed and natural fire 

and woody plant control.

• Grassland restoration efforts.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 1. 2004 167

• Use the Forest Interdisciplinary Team approach 

to handle NEPA work, thus freeing the Districts 

to continue their work.

7. What do you feel are the current trends in 

public perception and patterns of use related 

to Southwestern grasslands?

• Greatly increasing public interest and demand 

for a wide array of use opportunities, as well as 

enhancement of declining species and habitats. At 

the same time, there is a backlash toward State 

or Federal listed species. Even with increasing 

interest, the general public do not assign as 

great a value to grasslands as they do to other 

ecosystems and are poorly informed on grassland 

issues. However, the public is rapidly gaining a 

greater appreciation for open space and is generally becoming more aware of the environment.

• Increasing recreational use including offhighway 

vehicles driving and hunting. Most recreational 

use is concentrated in riparian areas. Hiking 

is likely to increase at a slower rate due to the 

constancy of vistas.

• Fear of loss of grazing use among permittees. 

Local support for subsistence grazing (northern 

New Mexico) with effective political lobbying.

• Backlash against prescribed burning.

• Further polarization between environmental and 

resource use nongovernmental organizations.

8. Who are your key stakeholders and why are 

they considered key stakeholders? How do you 

work with other grassland managers, stakeholders, and researchers, and what barriers 

do you face in initializing and maintaining 

cooperative efforts?

• Key stakeholders include: permittees, Ducks 

Unlimited, Playa Lake Joint Venture, N. Great 

Plains Initiative, Quivera Coalition. Barriers 

include lack of funding, current workload including litigation, lack of time needed to develop 

partnerships.

• Most significant stakeholder is private landowner 

as most grassland regions are predominately privately owned. Other big players are State wildlife 

agencies, State extension, agricultural research 

universities, NRCS, Fish and Wildlife Service 

(High Plains Partnership), Western Governors 

Association, rural development agencies, Nature 

Conservancy. Numerous efforts that often bring 

these folks together, such as Playa Lakes Joint 

Venture, Lesser Prairie Chicken Interstate 

Working Group, BTPD working groups. As a 

whole the key players and agencies seem much 

more united and working from the same viewpoint 

than occur with most regional applications in 

other parts of the Western United States, but 

sometimes it is hard to tie all the various group 

efforts together. One possibility would be to have 

an area serve as a “showcase.”

9. If you had unlimited funding, what would 

you do differently to change or improve management of the Southwestern grasslands?

• Have more public education opportunities; do 

more monitoring work to put a grazing association in place; encourage more cooperative efforts 

across boundaries involving users and more of 

the public; make decisions that are ecologically 

based rather than economic based; implement 

more community based projects; focus more of 

our efforts on bringing about changes to improve 

grassland health.

• (1) More verified ecological understanding of 

effects of applied and natural processes on an 

ecoregion basis through research; (2) more public 

use/access facilities and programs/emphasis; (3) 

more funding to return landscapes to more natural 

ecological states (fire, ecologically based grazing, 

elimination of noxious/invasive plants such as the 

black locust, reintroduction of lost species, such 

as BTPD and lesser PC, where appropriate); (4) 

rural economic development emphasis on nature 

tourism; (5) financial incentives to private landowners to maintain or develop ecologically sound 

lands supported by income opportunities and/or 

public subsidy, such as the Farm Bill currently 

provides in CRP, WRP, WHIP, and development 

cost-sharing efforts, such as F&WS Partners for 

Wildlife, State programs, and so forth.

• Collection of baseline vegetation information 

NOW, rather than as current funding allows. 

Funding to support consolidation of grasslands, 

funding, and resources to support interagency/

interorganization plans on landscape scale. This 

requires not only additional funding but also 

human resources and flexibility by all agencies/organizations to get past agency-specific 

barriers that make interagency implementation 

difficult. On our grasslands that are within our 

mountain districts, the biggest issue and area of 

need is pinyon-juniper encroachment in the small 

openings. Maintaining and restoring meadows 

through burning and/or mechanical thinning is 

needed.
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