Conservation Benefits 

of Rangeland Practices 

Assessment, Recommendations, and Knowledge Gaps

Davidd. Briske, Editor

Conservation Benefits of Rangeland Practices David d. Briske Assessment, Recommendations, and Knowledge Gaps
The expressed goal of this synthesis is to inform deliberations of managers and 

policymakers regarding the current effectiveness and potential improvements to 

rangeland conservation practices and programs. As such, this synthesis represents a 

scientific assessment that was reached independently of the current position or policy 

of the US Department of Agriculture or the United States government.

USDA is an equal opportunity provider and employer.

ISBN 978-0-9849499-0-8

On the cover: “Zion Sunset” highlights several famous 

land forms, including West Temple, Altar of Sacrifice 

and the Sentinel in Zion National Park in southern 

Utah. Photographer: Steve Guymon, Providence, 

Utah; www.steveguymonphotograpy.com.

About NRCS: Helping People Help the Land 

The Natural Resources Conservation Service (NRCS) is a USDA agency that works with 

landowners through conservation planning and assistance to benefit the soil, water, air, 

plants, and animals for productive lands and healthy ecosystems. Working in nearly 

every county in the Nation, NRCS employees understand local resource concerns and 

challenges to assist with conservation solutions that last. Land stewardship by 

private landowners is critical to the health of our Nation’s environment.

Science and technology are critical to good conservation. NRCS experts from many 

disciplines come together to help landowners conserve natural resources in efficient, 

smart and sustainable ways. NRCS succeeds through partnerships, working closely with 

individual farmers and ranchers, landowners, local conservation districts, government 

agencies, Tribes, Earth Team volunteers and many other people and groups that care 

about the quality of America’s natural resources. More information about 

the NRCS is at www.nrcs.usda.gov.

About ARS: Leading America towards a Better Future 

Through Agricultural Research and Information

The Agricultural Research Service (ARS) is the U.S. Department of Agriculture’s chief 

scientific research agency. ARS conducts research to develop and transfer solutions 

to agricultural problems of high national priority and provide information access and 

dissemination to ensure high-quality, safe food, and other agricultural products; assess 

the nutritional needs of Americans; sustain a competitive agricultural economy; enhance 

the natural resource base and the environment; and provide economic opportunities for 

rural citizens, communities, and society as a whole. More information about 

the ARS is at www.ars.usda.gov.

About SRM: Society for Range Management

The Society for Range Management is the professional society dedicated to supporting 

persons who work with rangelands and have a commitment to their sustainable 

use. Rangelands comprise almost one-half of all the lands in the world. They are 

extremely important to society for the goods and services they produce and for the 

ecological services they provide. SRM is dedicated to the conservation and sustainable 

management of rangelands for the benefit of current societies and for future generations. 

More information about the SRM is at www.rangelands.org.
	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	

	 	

Conservation Benefits 

of Rangeland Practices 

Assessment, Recommendations, and Knowledge Gaps 

This document was written to provide an unprecedented 

source of evidence-based information to guide the 

development and assessment of management practices 

and conservation programs on the nation's rangelands.” 

Leonard W. Jolley 

David D. Briske 

i 
	 	 	 	

Conservation Benefits of Rangeland Practices: Assessment, 

Recommendations, and Knowledge Gaps 

United States Department of Agriculture, Natural Resources 

Conservation Service 

David D. Briske, editor 

October 2011 

Second printing with revisions January 2012 

ISBN 978-0-9849499-0-8 

The USDA-Agricultural Research Service and the Society for Range 

Management supported the development of this publication. 

Printed by Allen Press, Inc., Lawrence, Kansas 

Suggested citation: Briske, D.D. [ed]. 2011. Conservation 

Benefits of Rangeland Practices: Assessment, Recommendations, 

and Knowledge Gaps. United States Department of Agriculture, 

Natural Resources Conservation Service. 429 p. 

ii Conservation Benefits of Rangeland Practices 
	 	

	 	

 

	 	 	 	 	 	 	 	 	 	 	 	

 

	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 		 	

	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

table of Contents 

Preface 

David D. Briske	 

iv 

Acknowledgments 

Leonard W. Jolley	 

Introduction 

David D. Briske, Leonard W. Jolley, Lisa F. Duriancik, and James P. Dobrowolski	 

Executive Summary 

David D. Briske, and Thomas L. Thurow	 

v 

1 

9 

ChAPtER 1: An Evidence-Based Assessment of Prescribed Grazing Practices 

David D. Briske, Justin D. Derner, Daniel G. Milchunas, and Ken W. Tate	 

21 

ChAPtER 2: Assessment of Prescribed Fire as a Conservation Practice 

Samuel D. Fuhlendorf, Ryan F. Limb, David M. Engle, and Richard F. Miller	 

75 

ChAPtER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 

Steven R. Archer, Kirk W. Davies, Timothy E. Fulbright, Kirk C. McDaniel, 

Bradford P. Wilcox, and Katharine I. Predick	 

105 

ChAPtER 4: Assessment of Range Planting as a Conservation Practice 

Stuart P. Hardegree, Thomas A. Jones, Bruce A. Roundy, Nancy L. Shaw, and Thomas A. Monaco	 

171 

ChAPtER 5: A Scientific Assessment of the Effectiveness of Riparian Management Practices 

Mel R. George, Randy D. Jackson, Chad S. Boyd, and Ken W. Tate	 

213 

ChAPtER 6: An Assessment of Rangeland Activities on Wildlife Populations and habitats 

Paul R. Krausman, Vernon C. Bleich, William M. Block, David E. Naugle, and Mark C. Wallace	 

253 

ChAPtER 7: Invasive Plant Management on Anticipated Conservation Benefits: A Scientific Assessment 

Roger L. Sheley, Jeremy J. James, Mathew J. Rinella, Dana Blumenthal, and Joseph M. DiTomaso	 

291 

ChAPtER 8: A Landscape Approach to Rangeland Conservation Practices 

Brandon T. Bestelmeyer, Joel R. Brown, Samuel D. Fuhlendorf, Gene A. Fults, and X. Ben Wu	 

337 

ChAPtER 9: A Social and Economic Assessment of Rangeland Conservation Practices 

John A. Tanaka, Mark Brunson, and L. Allen Torell	 

371 

Contributors 

NRCS Partners	 

External Reviewers	 

423 

Index 426 

iii 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	

PREFACE
�

Conservation is a major tenet of American society that is symbolized by Aldo Leopold’s iconic work Sand County 

Almanac. A conservation ethic emphasizes the protection, management, and restoration of natural resources for 

the public benefit, including sustainable social and economic utilization. The importance of conservation was 

imprinted on the national psyche by several episodes of environmental degradation, including severe overgrazing 

of western rangelands in the late 19th century and the Dust Bowl of the 1930s—the latter of which provided 

the impetus for organization of the current-day Natural Resource Conservation Service. Black Sunday—a severe 

dust storm that occurred in the southern Great Plains on April 14, 1935—serves as a symbol of the devastating 

consequences of unsustainable land use on both natural resources and human well-being that is dependent on 

them. The rangeland profession similarly emerged from the actions of early government researchers and managers, 

focused largely on grazing management and restoration, to halt and reverse degradation of western rangelands in 

the late 19th century. These episodes of natural resource degradation have contributed to the growing awareness 

that conservation of the nation’s natural resources is as much about managing human actions and values as it is 

about managing natural resources themselves. 

In spite of broad recognition of the importance of natural resource conservation to the nation, it is necessary to 

substantiate the outcomes of conservation programs in an era of increasing fiscal responsibility and accountability. 

The Conservation Effects Assessment Project (CEAP) was created to assemble the baseline knowledge of rangeland 

conservation programs, inspire innovation in the design and implementation of future programs, and provide a 

blueprint for the delivery of science-based and cost-effective conservation programs. CEAP expressly emphasizes 

that conservation programs address the environmental quality of lands in addition to the sustainable production 

of agricultural goods. Future conservation programs will be increasingly called on to evaluate the benefits of local 

agricultural production relative to the maintenance of regional ecosystem services. 

The academic community has embraced the vision of CEAP and has committed to this synthesis by retrieving 

and evaluating thousands of published research papers and compiling the most relevant information into readily 

accessible forms. The evidence-based recommendations originating from this synthesis can guide the development 

and assessment of future management practices and conservation programs. The knowledge gaps that have been 

identified can inform funding programs of areas in need of further research. Success of the Rangeland CEAP 

Synthesis will partially be determined by 1) the extent to which it can strengthen the linkage between scientific and 

management knowledge, 2) advance conservation science and policy, and 3) promote assessment of societal benefits, 

including both agricultural goods and ecosystems services, emerging from conservation programs. 

Even though the Rangeland CEAP Synthesis was explicitly designed and implemented to assess conservation 

programs of the US Department of Agriculture–Natural Resources Conservation Service, it has broad and significant 

implications to the entire rangeland profession. This synthesis provides a compelling argument for the development 

of an expanded rangeland research agenda that can more effectively articulate and embrace the scope and complexity 

of the conservation challenges that are most pressing to the nation. 

Rangelands are complex adaptive systems that encompass both ecological and social components as well as the 

intricate and poorly understood interactions among these components. This requires that social knowledge of 

rangeland systems, including management, socioeconomics, and policy, merit equal priority to that of ecological 

knowledge because they collectively establish conservation success. Therefore, conservation programs and practices 

within rangeland systems should be designed, implemented, and modified on the basis of multiple knowledge 

sources acquired from both anticipated and unanticipated conservation outcomes. Partnerships among natural 

resource managers, researchers, and policymakers are likely to generate the most relevant knowledge to address the 

emerging conservation challenges confronting the nation. 

David D. Briske	 

Editor and Academic Coordinator	 

Rangeland CEAP Synthesis 

iv Conservation Benefits of Rangeland Practices 
	 	 	

	 	 	

	 	 	

	 	 	 	 	 	 	

ACKNOWLEDGMENtS 

I wish to acknowledge the commitment, skill, and persistent diplomacy on the part of our academic coordinator 

and document editor Dr. David D. Briske, Texas A&M University. The success of this publication would be very 

much in doubt without his capable leadership and consistent vision for the potential of Rangeland CEAP. Dr. 

Tom Thurow, University of Wyoming, conducted a thorough review of the entire draft document to assist with 

matters of tone, content, consistency, and context. His insight and passion for the project are greatly appreciated. 

Special recognition must be afforded the authors since much was asked of these subject matter experts, and most 

of them labor in an environment where synthesis documents are not very helpful on matters of promotion and 

advancement in the research world. They are a very talented group, and their labors on this milestone document 

are greatly appreciated. Nearly 30 external reviewers were involved in the evaluation of this document, and their 

significant inputs enhanced the quality of this work. 

The US Department of Agriculture–Natural Resources Conservation Service (USDA-NRCS) and its leadership 

are to be commended for initiating and consistently supporting this project to scrutinize the science supporting 

rangeland conservation practices. The CEAP Steering Committee was a consistent advocate for the pursuit 

and publication of this literature synthesis. I extent a special thanks to the NRCS specialists who functioned 

interactively with the science writing teams to ensure that the use of conservation practices in the NRCS 

planning environment was fully appreciated. 

The Agricultural Research Service (ARS) allowed several of their scientists to function as authors on this project. 

The long-term nature of the ARS research vision and their selfless research and service contributions are truly 

impressive. Dr. Mark Weltz (USDA-ARS) was crucial to the administration of key agreements in support of this 

project. The Society for Range Management supported Rangeland CEAP by hosting an initial organizational 

meeting at their national headquarters and by managing USDA funds to conduct this project. 

Leonard W. Jolley, PhD 

Leader, CEAP Grazing Lands 

USDA Natural Resources Conservation Service 

Davis, California 

(Photo: Chris Call, Invasive species management; p. 321) 

v 
	 	 	 	

	 	 	

	 	 	

	 	

	 	 	 	

	 	

	 	 	 	

The recent success 

of CEAP provides “ numerous opportunities 

and challenges 

to achieve its full 

potential within 

the USDA and the 

broader conservation 

community” 

Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	 	 	 	 	 	 	

	 	

	 	 	

	 	

	 	

	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	

introduction to the Conservation 

effects assessment Project and the 

rangeland Literature synthesis 

David D. Briske,1 Leonard W. Jolley,2 Lisa F. Duriancik,3 and James P. Dobrowolski4 

Authors are 1Professor, Ecosystem Science & Management, 2138 TAMU, Texas A&M University, College Station, 

TX 77843-2138. 2Leader, Conservation Effects Assessment Project for Grazing Lands, Rangeland and Pastureland 

Ecologist, Resource Assessment Division, USDA-NRCS, Department of Plant Sciences, 2233 PES, Mail Stop 1, One 

Shields Ave, University of California–Davis, Davis, CA 95616-7323. 3Coordinator, Conservation Effects Assessment 

Project, Resources Assessment Division, USDA-NRCS, George Washington Carver Center, Beltsville, MD 20705. 

4National Program Leader, Rangeland and Grassland Ecosystems, Agricultural Water Security, USDA-NIFA, 1400 

Independence Ave SW, Washington DC 20250-2210. 

The ConservaTion effeCTs 

assessmenT ProjeCT 

conservation programming that emphasized 

environmental quality of these lands in 

addition to sustainable agricultural production 

The Conservation Effects Assessment (Mausbach and Dedrick 2004). 

Project (CEAP) is a unique, multiagency 

effort designed to quantify conservation 

effects and to determine how conservation 

practices can be most effectively designed 

and implemented to protect and enhance 

environmental quality (Duriancik et al. 

CeaP Goals 

The primary goal of CEAP is to strengthen the 

scientific foundation underpinning conservation 

programs to protect and enhance environmental 

Rangelands represent noncultivated, non-forested land 

that is extensively managed 

with ecological principles. 

(Photo: David Briske) 

2008). CEAP was jointly initiated in 2003 

by the Natural Resources Conservation 

Service (NRCS) in partnership with the 

Agricultural Research Service (ARS) and the 

National Institute of Food and Agriculture 

(NIFA) in response to requests from Congress 

and the Office of Management and Budget 

for greater accountability to US taxpayers 

following a near doubling of US Department 

of Agriculture (USDA) conservation program 

funding in the 2002 Farm Bill. These funds 

are allocated to multiple conservation 

practices through several USDA-sponsored 

conservation programs, including the 

Environmental Quality Incentives Program, 

Wetlands Reserve Program, Wildlife Habitat 

Incentives Program, Conservation Reserve 

Program, and NRCS Conservation Technical 

Assistance Program. This funding increase was 

concomitant with substantial modifications to 

Introduction to CEAP and the Rangeland Literature Synthesis 1 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	

Rangelands are comprised 

of diverse and extensive ecosystems that provide multiple 

goods and services to society. 

(Photo: Rick Miller) 

quality of managed lands. CEAP is focused on 

establishing principles to guide cost-effective 

conservation practices at landscape scales and 

to achieve multiple environmental quality goals 

by placing specified conservation practices 

or combinations of complementary practices 

at appropriate locations on the landscape to 

maximize their effectiveness. CEAP is also 

developing science-based guidance, information, 

and decision support tools to determine the 

appropriate practices to be implemented at 

various locations on the landscape and to 

provide conservation program managers with a 

blueprint for delivery of science-based and costeffective conservation programs (Duriancik et 

al. 2008). 

A secondary goal of CEAP is to establish a 

framework for assessing and reporting the 

full suite of ecosystem services impacted by 

various conservation practices. Ecosystem 

services represent the benefits that ecological 

processes convey to human societies and the 

natural environment. For example, agricultural 

lands provide flood and drought mitigation, 

water and air purification, biodiversity, 

carbon sequestration, nutrient cycling, and 

aesthetics and recreation, in addition to the 

primary agricultural commodities produced. 

These ecosystem services are often taken for 

granted and unpriced or underpriced by the 

marketplace. Research and assessment activities 

will be integrated within CEAP to provide a 

scientific foundation for assessing the extent 

to which ecosystem services are enhanced by 

conservation practices and programs. 

organization and approach 

The USDA engaged the Soil and Water 

Conservation Society in 2005 to assemble a 

panel of university scientists and conservation 

community leaders to recommend the most 

effective, proactive, and scientifically credible 

CEAP activities—thereby ensuring that 

2 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

D. D. Briske, L. W. Jolley, L. F. Duriancik, and J. P. Dobrowolski 

CEAP products would have wide utility for 

diverse stakeholders within the conservation 

community. CEAP has evolved into an 

assessment and research initiative directed 

at determining not only the impacts of 

conservation practices, but also evaluating 

procedures to more effectively manage 

agricultural landscapes in order to address 

environmental quality goals at local, regional, 

and national scales (Maresch et al. 2008). 

Three principal themes will guide CEAP 

investments and activities in the future 

(Maresch et al. 2008): 

1.	� Research addressing effective and efficient 

implementation of conservation practices 

and programs to meet environmental goals 

and enhance environmental quality. 

•	� Continue and expand CEAP research 

projects on the effects and benefits 

of conservation practices for soil and 

water quality at the watershed and 

landscape scales. 

•	� Implement a new research and 

assessment initiative for grazing lands 

designed to provide scientific evidence 

for implementation of conservation 

practices at the landscape scale. 

•	� Determine the critical processes 

and attributes to be measured at the 

appropriate landscape position for 

evaluation of environmental benefits. 

•	� Expand the scope of assessment to 

include evaluation of a full suite of 

ecosystem services influenced by 

conservation practices and programs. 

2.	� Assessment of the environmental impacts 

of conservation practices for reporting at 

the regional and national scales. 

•	� Continue CEAP activities designed 

to estimate environmental benefits of 

conservation practices and programs. 

•	� Develop a framework for reporting 

impacts of conservation practices 

and programs in terms of ecosystem 

services. 

•	� Identify future conservation 

requirements and provide information 

for setting national and regional 

priorities. 

•	� Expand assessment capabilities to 

address potential impacts of changes 

in agricultural land use and policy and 

define necessary conservation programs 

to meet new environmental challenges 

brought about by alternative land use 

or policy changes. 

3. Translation of science into practice by 

developing a blueprint for integrating 

scientific knowledge into the 

conservation planning and protocols for 

implementation. 

•	� Communicate research findings 

and lessons learned about managing 

agricultural landscapes to a broad 

audience. 

•	� Develop strategies for 

communicating scientific findings 

and recommendations to farmers, 

ranchers, and NRCS field office staff 

describing opportunities to enhance 

environmental quality. 

•	� Conduct studies to determine the 

types of tools and resources field offices 

require to evaluate and implement 

conservation practices within 

landscapes. 

•	� Conduct studies to demonstrate 

effective implementation of 

landscape management and adaptive 

management to conservation planning, 

implementation, and monitoring. 

•	� Develop tools that can be used by 

NRCS field offices to identify the most 

appropriate practices to be applied 

at the most appropriate landscape 

positions to effectively and efficiently 

meet local and regional environmental 

goals. 

CEAP has been organized into four national 

assessments addressing croplands, wetlands, 

wildlife, and grazing lands—grazing lands are 

subdivided into rangelands and pasturelands 

based on distinct climate and management 

considerations. CEAP also includes several 

watershed investigations, which are sponsored 

by ARS, NIFA, and NRCS. These studies 

provide quantitative science-based outcomes 

and methodological innovations to support 

the national assessments. Bibliographies and 

literature syntheses have been developed for 

CEAP has been 

organized into 

four National 

Assessments 

addressing 

croplands, 

wetlands, 

wildlife, and 

grazing lands— 

grazing lands 

are subdivided 

into rangelands 

and pasturelands 

based on distinct 

climate and 

management 

considerations”

Introduction to CEAP and the Rangeland Literature Synthesis 3 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

TabLe 1. Areas treated by each of the seven major conservation practices addressed 

in the rangeland synthesis in the west and central regions of the United States during 

2004–2008. 

Conservation practice Usda code area (ha) 

Prescribed grazing 528 31 359 980 

Prescribed burning 338 370 821 

brush management 314 1 456 837 

range planting 550 517 301 

riparian herbaceous cover 390 12 352 

Upland wildlife habitat management 645 19 165 668 

Pest management 595 7 603 070 

each of the four national assessments and 

the watershed assessment; these document 

the current state of knowledge regarding the 

effectiveness of conservations practices, provide 

recommendations to enhance conservation 

programs, and identify critical knowledge gaps 

that require further research (Duriancik et al. 

2008). 

ranGeLand CeaP 

Rangeland CEAP was formally initiated in 

2006 to evaluate conservation effectiveness on 

rangelands (166 million hectares) and grazed 

forest (23 million hectares) that comprise 188 

million hectares of the nation’s nonfederal 

rural land. It emphasizes conservation practices 

that are routinely applied on rangelands 

west of the 100th meridian to accomplish 

multiple management and environmental 

goals, including maintenance of plant 

community health, protection of water quality 

and quantity, reduction of accelerated soil 

erosion, and promotion of economic stability 

through rangeland sustainability. Conservation 

practices are usually not implemented in 

isolation, but as part of a broader conservation 

plan that may potentially recommend 

implementation of multiple practices. 

Resource management systems represent a 

combination of conservation practices and 

resource management actions prescribed to 

address multiple natural resource concerns that 

meet or exceed the quality criteria for resource 

sustainability. It is fully anticipated that some 

combination of agricultural and environmental 

benefits arise from implementation of these 

conservation practices on rangelands, but 

quantitative measures of their specific effects on 

soil, water, animals, plants, and air are required 

to document the efficacy of these practices and 

systems. 

organization and approach 

Rangeland CEAP encompasses four interrelated 

components: 

1.	� National Assessment. Evaluation of the 

effects of conservation management 

on rangelands across the United States 

accomplished with a combination of 

ground-based measurements, remotely 

sensed data, and hydro-ecological and 

economic simulation models. This effort is 

coordinated by the ARS with emphasis on 

watershed modeling in the Intermountain 

West (Weltz et al. 2008). 

2.	� Watershed Assessment Studies. Quantification 

of the measureable cumulative effects of 

conservation practices and enhancement of 

understanding of the interactions among 

practices in experimental watersheds. 

These watersheds occur in both croplands 

and grazing lands and are intended to 

provide in-depth assessments that are not 

possible at the regional scale to evaluate 

and enhance performance of the national 

assessment models. 

3.	� Bibliographies. Compilation of published 

literature citations addressing the 

environmental benefits of conservation 

practices and programs for grazing lands 

was completed by the USDA National 

Agricultural Library in 2006. Dynamic 

bibliographies using real-time searches in 

the National Agricultural Library catalog 

(AGRICOLA) have been assembled 

(USDA National Agricultural Library 

2007). 

4. Literature Synthesis. Compilation of the 

current status of knowledge concerning 

the ecological effectiveness of major 

conservation practices applied on 

rangelands by systematically mining the 

published scientific literature. 

rangeland synthesis 

Rangelands synthesis CEAP has been 

developed to provide an in-depth 

assessment of the published experimental 

information concerning the effectiveness 

of previously implemented conservation 

4 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	 	 	 	 	

	 	

	 	

	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

D. D. Briske, L. W. Jolley, L. F. Duriancik, and J. P. Dobrowolski 

practices on rangelands. The primary goal 

is to provide the most definitive assessment 

of conservation impacts ever conducted 

within the rangeland profession to serve 

as an evidence-based benchmark for the 

efficacy of current conservation practices. 

This is a necessary and essential step for 

assessing the benefits of existing conservation 

practices and determining whether or not 

current practices require modification in 

either design or implementation to enhance 

their effectiveness in future programs. This 

information, coupled with evidence-based 

recommendations to enhance conservation 

programs, and identification of key knowledge 

gaps in existing information will promote 

development of novel evidence-based 

conservation systems that possess the capacity 

to assess environmental quality and ecosystem 

services in addition to traditional agricultural 

production metrics. 

The rangeland literature synthesis was 

specifically organized around a series of testable 

questions derived from the stated purposes 

or outcomes of seven major conservation 

practices as identified in the NRCS National 

Conservation Practice Standards. These 

conservation practices were selected for 

assessment based on their prominence in 

the conservation planning environment, 

the extent and frequency with which they 

are applied, and the amount of incentive 

payments allocated to them (Tables 1 and 2). 

Rigorous literature syntheses established the 

portion of experimental studies that supported, 

refuted, or were insufficient to assess the 

benefits of these conservation practices. Two 

Table 2. Environmental Quality Incentive Program funds (US dollars) expended on the seven major conservation practices address in the 

rangeland synthesis by region and state during 1997–2003. 

region state 

brush 

management 

Prescribed 

burning 

Prescribed 

grazing 

range 

planting 

riparian 

herbaceous 

cover 

Upland 

wildlife habitat 

management 

West rangeland state 

Arizona 672 345 1 090 536 134 842 1 650 

California 2 192 285 2 625 110 813 592 108 12 116 20 394 

Colorado 185 295 3 785 3 496 391 70 985 8 936 

Idaho 7 250 29 411 18 868 1 126 527 

Montana 923 457 135 236 148 15 895 

Nevada 35 756 90 770 77 883 214 

New Mexico 3 259 774 21 542 421 262 674 895 188 

Oregon 164 759 83 377 188 088 66 35 790 

Utah 415 038 13 402 18 199 241 036 5 629 1 113 

Washington 563 41 862 96 686 75 547 

Wyoming 145 829 5 145 3 136 108 

West total 7 078 894 46 589 6 219 214 2 230 627 19 085 160 362 

Central rangeland state 

Kansas 551 470 33 595 1 321 533 142 215 12 

Nebraska 124 609 2 218 197 443 133 641 21 757 

North Dakota 8 807 1 091 930 203 208 81 976 

Oklahoma 2 215 107 200 046 298 249 174 530 430 

South Dakota 6 376 59 330 100 057 41 820 

Texas 9 297 443 72 136 122 491 1 535 268 25 682 5 290 

Central total 12 203 812 307 995 3 090 976 2 288 919 25 682 151 285 

Grand total 19 282 706 354 584 9 310 190 4 519 546 44 767 311 647 

Introduction to CEAP and the Rangeland Literature Synthesis 5 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	 	 	

	 	

	 	 	

	 	

The recent 

success of 

CEAP provides 

numerous 

opportunities 

and challenges 

to achieve its full 

potential within 

the USDA and 

the broader 

conservation 

community.”

additional chapters emphasizing landscape 

ecology and socioeconomic issues, including 

ecosystem services, were developed on the 

basis of their anticipated importance to future 

conservation programs and planning. These 

final two chapters were organized to be crosscutting among all seven conservation practice 

standards. 

Seven major conservation practices and 

two cross-cutting issues are addressed in the 

Rangeland CEAP literature synthesis. 

•	� Prescribed Grazing 

•	� Prescribed Burning 

•	� Brush Management 

•	� Range Planting 

•	� Riparian Herbaceous Cover 

•	� Upland Wildlife Habitat Management 

•	� Herbaceous Weed Control 

•	� Landscape Analysis (cross-cutting chapter) 

•	� Socioeconomics and Ecosystem Services 

(cross-cutting chapter) 

A writing team was formed for each of these 

nine chapters by recruiting team leaders 

with recognized experience and expertise 

in the respective subject matter areas and 

encouraging them to select two to four 

subject matter specialists with sufficient 

diversity to address the entire scope of 

ecological topics under consideration—soils, 

water, air, plants, and animals—as they relate 

to the seven major conservation practices. 

Geographic representation of team members 

across US rangelands was considered in 

the selection process to the extent possible. 

Teams focused on the development of 

tabular databases comprising quantitative 

information addressing multiple ecological 

responses to conservation practices to 

provide an unprecedented compilation of 

evidence-based information. Databases 

were primarily derived from the refereed 

literature with some quality “grey” literature 

included at the discretion of the writing 

teams. Individual chapters underwent 

rigorous peer review by three recognized 

experts that were not affiliated with CEAP; 

reviewer recommendations were provided 

to the chapter authors for incorporation, 

and the revised chapters were evaluated by 

the academic coordinator of Rangeland 

CEAP. The entire document was evaluated 

for relevance and impact by one nonfederal 

reviewer and one NRCS reviewer prior to 

publication. 

Major sections addressed within each of the 

synthesis chapters include the following: 

•	� Description of conservation practices and 

their purported benefits. 

•	� Evidence-based assessment of conservation 

benefits, including potential tradeoffs and 

risks of not implementing the practice, and 

of unintended negative outcomes. 

•	� Recommendations to modify or develop 

alternative conservation practices to 

more effectively accomplish the intended 

purposes. 

•	� Identification of critical knowledge gaps in 

current information. 

•	� Succinct summary and conclusion of 

findings for each conservation practice. 

•	� Literature-cited section containing 

citations within the text, but not those 

used to support the extensive tabular data. 

These supporting citations will be made 

available in a searchable electronic version 

of this document. 

The rangeland literature synthesis is available 

in both hardcopy and electronic formats. 

The electronic version will be posted on the 

NRCS-CEAP, National Agricultural Library, 

and Society for Range Management Web sites 

and it will be searchable for both citations 

and appendices of tabular data specific to 

each chapter. This document is designed 

to target multiple audiences, including 1) 

policy makers (e.g., executive summary), 

2) practitioners and students (e.g., general 

synthesis), and 3) researchers and modelers 

(e.g., tabular databases and supporting 

references). 

CeaP imPLemenTaTion and The 

road ahead 

Design and implementation of conservation 

practices through use of the best available 

information and technology is a hallmark of 

NRCS. The knowledge generated through 

CEAP-sponsored assessments is critical to 

continuation of this mission by optimizing 

the cost-effectiveness of conservation 

practices and the environmental outcomes 

6 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	

		

	 	

D. D. Briske, L. W. Jolley, L. F. Duriancik, and J. P. Dobrowolski 

that they support. CEAP has generated 

new conservation opportunities to manage 

agricultural landscapes for environmental 

quality, created diverse and valuable 

conservation partnerships, and emphasized 

conservation assessment and planning at the 

watershed and landscape scales. 

Anticipated applications of the information 

created by CEAP include the following: 

•	� Support further development of grazing 

lands management and conservation 

practices within the Soil and Water 

Resources Conservation Act and National 

Conservation Program. 

•	� Informing grazing land initiatives in 

subsequent Farm Bills. 

•	� Advancement of conservation planning 

tools and program delivery mechanisms 

for targeted implementation and enhanced 

adoption. 

•	� Evaluation of mitigation and adaptation 

strategies associated with climate change, 

water security challenges, or changes in 

land use or management. 

•	� Devising inventory and monitoring 

protocols to better document conservation 

benefits for both agricultural production 

and environmental quality. 

CEAP has created new 

opportunities to promote 

rangeland stewardship. 

(Photo: Rick Miller) 

The recent success of CEAP provides 

numerous opportunities and challenges 

to achieve its full potential within the 

USDA and the broader conservation 

community. Implementation of CEAP 

will require reevaluation of procedures 

concerning conservation planning, greater 

knowledge transfer among USDA programs, 

modification of select conservation practices, 

and additional technology development 

and transfer. An expanded culture of 

collaboration among USDA programs and 

agencies, and several nonfederal partners, has 

contributed greatly to the transformational 

influence of CEAP. Continued collaboration 

is necessary both within USDA programs as 

well as with the broader conservation and 

agricultural communities to further capitalize 

on the knowledge and unprecedented 

capacity associated with rapidly emerging 

conservation science to produce the next 

generation of conservation programs for the 

21st century. 

Literature Cited 

Duriancik, L. F., D. Bucks, J. P. Dobrowolski, T. 

Drewes, S. D. Eckles, L. Jolley, R. L. Kellogg, 

D. Lund, J. R. Makuch, M. P. O’Neill, C. 

A. Rewa, M. R. Walbridge, R. Parry, and 

M. A. Weltz. 2008. The first five years of the 

Introduction to CEAP and the Rangeland Literature Synthesis 7 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Conservation Effects Assessment Project. Journal of 

Soil and Water Conservation 63:185–197. 

Maresch, W., M. R. Walbridge, and D. Kugler. 

2008. Enhancing conservation on agricultural 

landscapes: a new direction for the Conservation 

Effects Assessment Project. Journal of Soil and 

Water Conservation 63:198–203. 

Mausbach. M. J., and A. R. Dedrick. 2004. The 

length we go: measuring environmental benefits 

of conservation practices. Journal of Soil and 

Water Conservation 59:96–103. 

[USDA National Agricultural Library] US 

Department of Agriculture National 

Agricultural Library. 2007. Environmental 

effects of Conservation practices on grazing 

lands: a Conservation Effects Assessment 

Project (CEAP) bibliography. Special Reference 

Brief 2006-02. Available at: http://wqic. 

nal.usda.gov/nal_display/index.php?info_ 

center=7&tax_level=4&tax_subject=595&topic_ 

id=2422&level3_id=6990&level4_id=11307. 

Access date: May 2009 

Weltz, M. A., L. Jolley, M. Nearing, J. Stone, 

D. Goodrich, K. Spaeth, J. Kiniry, J. 

Arnold, D. Bubenheim, M. Hernandez, and 

H. Wei. 2008. Assessing the benefits of grazing 

land conservation practices. Journal of Soil and 

Water Conservation 63:214–217. 

8 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

CHAPTER 

1 

An Evidence-Based Assessment of 

Prescribed Grazing Practices 

David D. Briske,1 Justin D. Derner,2 Daniel G. Milchunas,3 and Ken W. Tate4 

Authors are 1Professor, Ecosystem Science & Management, 2138 TAMU, Texas A&M
 

University, College Station, TX 77843-2138; 2 Rangeland Scientist and Leader, USDA-ARS
 

High Plains Grasslands Research Station, Cheyenne, WY 82009; 3 Research Scientist,
 

Forest, Rangeland, and Watershed Stewardship Department, Colorado State University,
 

Fort Collins, CO 80523-1472; and 4 Rangeland Watershed Specialist,
 

Department of Plant Sciences, 2233 PES, Mail Stop 1,
 

University of California–Davis, Davis, CA 95616-7323.
 

Correspondence: Email: dbriske@tamu.edu 

Reference to any commercial product or service is made with the understanding 

that no discrimination is intended and no endorsement by USDA is implied 

21
 
	 	 	 	

	 	

	 	 	

	 	 	

	 	

	 	

	 	

Experimental grazing 

research has produced 

“

consistent relationships 

between stocking rate 

and plant production, 

animal production, and 

species composition 

of herbaceous plant 

communities.”

22 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	

	 	 	

1 An Evidence-Based Assessment of 

Prescribed Grazing Practices 

David D. Briske, Justin D. Derner, Daniel G. Milchunas, and Ken W. Tate 

IntroductIon 

Prescribed grazing is inclusive of many 

interrelated management and conservation 

activities implemented for purposes of 

managing grazed ecosystems. It is supported 

by a loosely organized information base that 

contains management experience, agency policy 

and procedures, and scientific information that 

has been developed throughout the history of 

the rangeland profession. The components of 

prescribed grazing are implemented in various 

combinations to achieve multiple management 

goals and outcomes under a wide variety of 

ecological conditions in diverse rangeland 

ecosystems. A fundamental premise of effective 

grazing management is that it supports 

ecosystem sustainability and restoration of 

degraded ecosystems. Management actions 

have traditionally emphasized livestock species 

and number and their temporal and spatial 

distribution on the landscape (Stoddart et 

al. 1975; Heitschmidt and Taylor 1991). 

The management of grazed ecosystems 

involves multiple human dimensions as well 

as complex ecological processes, making it 

difficult and impractical to attempt to separate 

grazing management from overall enterprise 

management (Stuth 1991). Therefore, 

management practices are commonly designed 

and applied within the context of specific 

landowner operations, management needs, 

and natural resource conservation goals. 

Consequently, prescribed grazing involves a 

continuum of management activities ranging 

from extensive management to those that are 

much more labor and infrastructure intensive. 

Context for the initial development of 

prescribed grazing in the United States 

originated with management recommendations 

to promote sustainable use and recovery of 

rangelands damaged by excessive livestock 

grazing early in the 20th century (Smith 1896; 

Wooten 1916; Sampson 1923, 1951; Hart 

and Norton 1988). Excessively high stocking 

rates (animal units area−1 time−1) common to 

the late 19th and early 20th centuries were 

unsustainable, and the negative consequences 

of those extreme stocking rates adversely 

affected numerous ecosystems throughout the 

Great Plains and West. Early rangeland research 

advocated use of reduced stocking rates and 

simple grazing systems to impose early season 

deferment and season-long rest to halt and 

potentially reverse ecological damage created 

by severe overgrazing. Increased efficiency of 

livestock production became an important 

objective during the 1980s and was associated 

with the introduction of short-duration grazing 

to the United States (Savory and Parsons 1980; 

Savory 1983). These management systems 

were designed to improve the efficiency of 

forage harvest, enhance forage quality, and 

promote livestock production. More recently, 

prescribed grazing has emphasized broader 

conservation goals and ecosystem services. 

Biodiversity conservation, water quality and 

quantity, woodland encroachment, invasive 

species, and carbon sequestration are but a 

few of the current high-profile conservation 

issues considered within grazed ecosystems. 

However, this emphasis is rather recent, and 

the amount of experimental information to 

date is insufficient to draw valid generalizations 

regarding these conservation issues. 

Even though the primary objective of this body 

of information is to promote effective grazing 

management, this is in itself not a sufficient 

foundation on which to evaluate this important 

land use. It is essential that the underlying 

components and processes of effective grazing 

management be recognized, understood, and 

documented to ensure that this information 

base is carefully scrutinized, accurate, and 

Prescribed grazing is the 

management of vegetation 

harvest by grazing or browsing 

animals to achieve desired objectives. (Photo: Derek Bailey) 

CHAPTER 1: An Assessment of Grazing Practices 23 
	 	 	 	

effectively promoted. The primary objective of 

Rangeland CEAP is to organize and evaluate 

the current body of scientific information 

supporting the anticipated benefits of 

rangeland conservation practices implemented 

by the US Department of Agriculture, Natural 

Resources Conservation Service (USDA-

NRCS). This assessment is intended to provide 

the foundation for the next generation of 

planning and assessment procedures that are 

to emphasize environmental quality and the 

assessment of multiple ecosystem services in 

addition to the traditional outcomes of farm 

and ranch productivity (Maresch et al. 2008). 

This chapter evaluates the ecological 

effectiveness of the major purposes and 

purported benefits for prescribed grazing 

as described in the USDA-NRCS National 

Conservation Practice Guidelines. This 

standard defines prescribed grazing as managing 

the harvest of vegetation with grazing and/ 

or browsing animals that is often applied as 

one component of a broader conservation 

management system to achieve one or more of 

the following purposes: •	� Improve or maintain desired species 

composition and vigor of plant 

communities. 

•	� Improve or maintain quantity and quality 

of forage for grazing and browsing animals’ 

health and productivity. 

•	� Improve or maintain surface and/or 

subsurface water quality and quantity. 

•	� Improve or maintain riparian and 

watershed function. 

•	� Reduce accelerated soil erosion and 

maintain or improve soil condition. 

•	� Improve or maintain the quantity and 

quality of food and/or cover available for 

wildlife. 

•	� Manage fine fuel loads to achieve desired 

conditions. 

This definition is very similar to that provided 

in the Society for Range Management 

(SRM) Glossary of Terms (1998)—“the 

manipulation of animal grazing in pursuit of 

a defined objective”—and to that of targeted 

grazing—“the application of a specific kind of 

livestock at a determined season, duration and 

intensity to accomplish a defined objective” 

(Launchbaugh and Walker 2006). Targeted 

grazing emphasizes objectives associated with 

landscape dynamics in addition to livestock 

production. It is important to note that 

prescribed grazing, as defined above, is a much 

broader concept than grazing system, which 

describes a specialized application of grazing 

management based on recurring periods of 

grazing, rest, and deferment for two or more 

pastures (Heitschmidt and Taylor 1991). The 

NRCS National Range and Pasture Handbook 

describes prescribed grazing schedules to 

recommend appropriate periods of grazing, 

rest, and deferment (USDA-NRCS 2003). 

The experimental data addressing these 

purposes were extracted primarily from 

the peer-reviewed literature, summarized 

and incorporated into tabular forms to 

provide an evidence-based assessment of 

how well prescribed grazing achieves these 

stated purposes. In some instances, direct 

comparisons could be made between intended 

conservation outcomes and the experimental 

evidence, but in many others, inferences had to 

be drawn from the most relevant experimental 

data to assess the effectiveness of conservation 

outcomes. Constraints of experimental research 

have influenced both the type of information 

available and the investigations selected for 

inclusion in this assessment. For example, 

spatial heterogeneity may produce conditions 

where most pastures under consideration 

possess generally similar topoedaphic 

characteristics, but in other cases one or 

more pastures may possess distinctly different 

characteristics. Only the first condition 

characterized by relatively homogeneous site 

conditions meets the traditional experimental 

requirements of replication and comparison 

with experimental controls, while decisions 

regarding heterogeneous site conditions can 

be assessed only on a case-by-case basis. Given 

that the goal of this chapter was to evaluate the 

preponderance of evidence supporting major 

grazing management practices, investigations 

that were unreplicated or that did not have 

experimental controls, that applied unequal 

treatments, or that contained minimal data 

were not included. These requirements were 

relaxed to some extent for the evaluation of 

wildlife because investigations addressing 

responses of specific wildlife species or groups 

to unique management practices were often 

limited. Similarly, minor wildlife groups were 

24 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

not addressed because of limited research 

evidence and space limitations. 

This chapter is organized into six major 

headings: introduction, evaluation of prescribed 

grazing purposes, associated considerations, 

recommendations, knowledge gaps, and 

conclusions. The evaluation of prescribed 

grazing purposes is the largest section, and it 

contains seven secondary headings addressing 

each of the conservation purposes previously 

described. Several of these purposes are 

further subdivided into tertiary headings of 

stocking rate and grazing system because these 

two research themes contain a large portion 

of the experimental information associated 

with grazing management. In addition to 

summarizing the experimental evidence 

relevant to prescribed grazing management, 

this chapter emphasizes the strengths and 

weaknesses of the experimental data, provides 

recommendations for improvement of this 

conservation practice, and identifies major 

knowledge gaps in the experimental literature. 

The overarching goal is to describe the current 

status of grazing management information 

to provide a foundation for development 

of the next generation of prescribed grazing 

practices. This chapter was commissioned by 

and is directed toward the NRCS, but it also 

contains important implications to the broader 

rangeland profession. 

EvAluAtIon of PrEscrIBEd GrAzInG 

PurPosEs 

Improve or Maintain desired species 

composition and vigor of Plant 

communities 

Stocking Rate. Stocking rate has long 

been recognized as a fundamental variable 

determining the sustainability and profitability 

of grazed rangeland ecosystems (Smith 1896; 

Wooton 1916; Sampson 1923). The objective 

of stocking rate is to balance the forage 

demand of grazing animals with that of forage 

production over an annual forage production 

cycle. The difficulty encountered when 

setting and maintaining appropriate stocking 

rate on rangelands is the high variability of 

forage production associated with annual 

and interannual precipitation variation. It 

is often recommended that stocking rates 

should be conservatively applied to minimize 

the detrimental consequences of overstocking 

during drought on the economic and ecological 

sustainability of grazed ecosystems. 

The importance of stocking rate to the 

management of grazed ecosystems has attracted 

considerable research attention over the past 

several decades. This research has produced 

consistent relationships between stocking rate 

and plant production, animal production, 

and species composition of herbaceous plant 

communities. Plant production decreases with 

increasing stocking rate, as does individual 

animal production (Bement 1969; Manley et 

al. 1997; Derner and Hart 2007; Derner et 

al. 2008). In contrast, animal production per 

land area increases with increasing stocking rate 

within the limits of ecosystem sustainability. 

These ecosystem responses to stocking rate have 

clear production and conservation implications. 

The response of several ecosystem variables 

indicates that stocking rate is at least indirectly 

correlated with ecosystem function and 

sustainability. High grazing intensities generally 

appear to minimize ecosystem function, 

which often has negative consequences for 

conservation goals and the provisioning of 

ecosystem services. Plant production is the 

most consistent response with 69% (25 of 

36) of the investigations reporting greater 

plant production at lower compared to higher 

stocking rates (Fig. 1). Twenty-eight percent 

(10 of 36) showed no difference in plant 

production with stocking rate. Only four of 

fIGurE 1. Number of 

investigations reporting 

significant effects of 

stocking rate on plant 

production and cover and 

livestock production per 

head and per unit land 

area. 

CHAPTER 1: An Assessment of Grazing Practices 25 
	 	 	 	

fIGurE 2. Number of 

investigations reporting 

significant effects of grazing 

system, categorized as 

short-duration and non– 

short-duration systems, on 

favorable changes in species composition of plant 

communities. 

these investigations considered plant species 

diversity or richness in relation to stocking 

rate, but the trend is for increasing diversity 

and richness with increasing stocking rate, 

which is a consistently observed community 

response to grazing (Milchunas et al. 1988). 

This response is interpreted as a function of the 

suppression of grass dominants at high stocking 

rates, which increases resource availability for 

subordinate species within the community 

(Collins 1987; Anderson and Briske 1995; 

Knapp et al. 1999). However, cases do exist 

where intensively grazed ecosystems are 

required to provide specific habitat for flora and 

fauna (Milchunas and Lauenroth 2008; Derner 

et al. 2009). 

Stocking rate has tremendous potential to 

modify the species composition of herbaceous 

vegetation. Significant change in species 

composition was documented to occur in 

71% (17 of 24) of the stocking rate studies 

evaluated. Plant cover showed a much less 

consistent response than did either production 

or species composition with 67% (14 of 21) of 

the investigations showing no difference with 

stocking rate compared to 24% (5 of 21) that 

did show a positive response. Compositional 

changes largely follow the classical increaser– 

decreaser patterns outlined by Dyksterhuis 

(1949) and more recently verified in a global 

vegetation analysis (Diaz et al. 2007) in which 

tallgrasses are replaced by midgrasses and 

midgrasses by shortgrasses. Eight of these 

studies recorded vegetation responses for ≥ 20 

yr and 14 studies for ≥ 10 yr, but significant 

vegetation change was also recorded in shorter 

time periods. These vegetation responses also 

document the occurrence of equilibrium 

dynamics in which grazing modifies the species 

composition of plant communities in addition 

to weather conditions (Fuhlendorf et al. 2001; 

Briske et al. 2003). The potential for recovery 

of species composition in response to reduced 

stocking rates also documents the high degree 

of resilience associated with many rangeland 

ecosystems (Milchunas et al. 1988). 

Grazing Systems. Although changes in plant 

species composition are often more qualitatively 

assessed than the plant and animal production 

values presented previously, the majority of 

investigations have not shown a clear benefit of 

rotational grazing over continuous grazing in 

promoting secondary succession or improving 

community composition on rangelands 

(Holechek et al. 1999, 2006). In our survey of 

25 grazing experiments, 86% (18) indicated 

no difference in species composition for 

continuous compared to rotational grazing 

at comparable stocking rates. Only 3 of 25 

experiments recorded improvements in species 

composition, and these were all deferredrotation rather than short-duration systems 

(Fig. 2). 

Experimental data referencing biotic diversity 

in grazing systems are limited, especially at 

regional scales, so definitive conclusions are 

unattainable at this point. However, the limited 

experiments addressing plant species diversity 

do not show that grazing systems enhance 

plant species diversity (Holechek et al. 2006). 

In tallgrass prairie, grazing system did not 

influence plant richness or diversity, but both 

variables increased with increasing stocking rate 

(Hickman et al. 2004). Increasing stocking rate 

reduced the abundance of the several dominant 

C4

 grass species and increased the expression 

of several subordinate species. Plant diversity 

responses to grazing are dependent on the 

direct response of various species to grazing 

and the indirect response of other species to 

grazing-induced release from competition 

(Milchunas et al. 1988; Anderson and Briske 

1995). 

Grazing Season and Deferment. Research 

addressing the season and length of grazing 

deferment is surprisingly limited given its 

26 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

	 	

	 	

	 	 	

	 	

	 	

	 	

	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

importance to grazing management. It is 

difficult to draw inferences from the few 

investigations specifically addressing season 

of grazing, especially given the variability 

in production and defoliation responses 

associated with precipitation variation within 

and among years (Zhang and Romo 1994). 

These authors were unable to make conclusive 

recommendations regarding production 

responses of northern mixed prairie to the 

seasonality and frequency of defoliation 

because of weather variation between years. 

Plant species with unique growth periods and 

production potentials contribute additional 

complexity to this assessment (Volesky et 

al. 2004). This underscores the difficulty 

of making generalizations regarding the 

appropriate season of grazing and deferment. 

Inferences regarding the appropriate length 

of grazing deferment can be derived from 

grazing systems research previously evaluated. 

Short deferment periods do not yield benefits 

in those variables measured, that is, plant 

and animal production, species composition, 

and soil characteristics. It can be inferred 

from this extensive data set that successive 

short deferments of 30–45 d are ineffective 

in offsetting short, intensive grazing periods 

of 2–11 d. Conclusions regarding length of 

deferment have been drawn from comparisons 

of short-duration and high-intensity, lowfrequency systems using 42- and 84-d 

deferment periods, respectively (Taylor et al. 

1993). These authors concluded that 80–90-d 

deferment periods were required to maintain 

desired species composition on semiarid 

rangelands. This interpretation has been 

corroborated by research conducted in mesic 

tallgrass ecosystems (Reece et al. 1996). Specific 

ecological mechanisms limiting increased plant 

production and improved species composition 

in response to short-term periodic deferment 

in rotational systems are not entirely clear, 

but they are very likely influenced by the 

time required for plant recovery, especially on 

semiarid rangelands, and the coincidence of 

favorable growth conditions with periods of 

grazing deferment (Briske et al. 2008). 

Grazing deferment relative to the onset and 

recovery from drought has also received 

minimal attention given its significance 

to grazing management. However, several 

conclusions can be drawn from a valuable, but 

limited data set. First, grazing deferment during 

drought has minimal potential to enhance 

plant production or species composition, even 

though it is often necessary to destock because 

of insufficient forage availability (Eneboe et al. 

2002; Heitschmidt et al. 2005; Gillen and Sims 

2006). However, deferment is important to 

maintain sufficient plant cover and density to 

protect soil quality and promote plant recovery 

once rainfall resumes (Wood and Blackburn 

1981a&b; Thurow 1991; Dalgleish and 

Hartnett 2006). Second, grazing deferment is 

not necessarily required for rapid and effective 

vegetation recovery from moderate drought 

conditions (Eneboe et al. 2002; Heitschmidt 

et al. 2005). Investigations demonstrating 

the ability of rainfall to override the effects 

of stocking rate on forage production and 

species composition indirectly support this 

interpretation (Milchunas et al. 1994; Biondini 

et al. 1998; Gillen et al. 2000; Vermeire et al. 

2008). Third, in the cases involving severe, 

prolonged drought, 2 yr or more may be 

required for recovery of species composition 

and productivity. Severe, multiyear drought 

can induce mortality of plants and tillers to 

retard plant growth following the resumption 

of rainfall (Briske and Hendrickson 1998; 

Dalgleish and Hartnett 2006; Yahdjian et al. 

2006). Consequently, several growing seasons 

may to be required for tiller and plant densities 

to recover to predrought values. Plant mortality 

was found to be approximately twice as great 

in heavily compared to more lightly grazed 

Great Plains rangelands following the multiyear 

drought of the 1950s (Albertson et al. 1957). 

Greater plant mortality is likely a consequence 

of the suppressed root growth and function 

that is known to occur with severe grazing of 

individual plants (Crider 1955). 

Improve or Maintain Quantity and 

Quality of forage for Grazing and 

Browsing Animals’ Health and 

Productivity 

Stocking Rate. Experimental data confirm the 

occurrence of a consistent trade-off between 

animal production per head and per land area 

with increasing stocking rate. Eighty percent 

(16 of 20) of investigations reported greater 

animal production per head at low compared 

to high stocking rates, while 82% (14 of 17) 

Research 

addressing the 

season and 

length of grazing 

deferment is 

surprisingly 

limited given 

its importance 

to grazing 

management.”

CHAPTER 1: An Assessment of Grazing Practices 27 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	

	 	 	 	

	 	

	 	

	 	

	 	 	

	 	 	

Individual plant 

production is 

most greatly 

suppressed by 

defoliation during 

the middle of the 

growing season, 

which coincides 

with culm elongation and the 

early boot stage 

of inflorescence 

development...”

showed greater animal production per land area 

at high compared to low stocking rates (Fig. 1). 

This trade-off in animal performance is readily 

explained by the greater availability of plant 

biomass per individual animal with decreasing 

stocking rate and greater forage harvest per 

unit land area with increasing stocking rate. 

Experimental evidence indicates that both 

forage quantity and quality decrease with 

increasing high compared to low stocking rates. 

Forage quality is most likely to decrease during 

grazing periods insufficient for appreciable 

regrowth, where animals initially select the 

highest-quality forage. 

Pattern of tiller defoliation research shows that 

80% or more of all tillers can be defoliated 

in a single grazing period with high stocking 

rates or grazing intensities (8 of 11 studies). 

This indicates that high harvest efficiencies 

and uniform grazing patterns can be obtained 

with large livestock numbers in certain cases. 

However, these data also indicate that multiple 

defoliations occur early within a grazing cycle 

(four of six studies). It has been suggested that 

repeat defoliations begin to occur at about 

the time that 60% of the tillers are initially 

defoliated and that very high grazing pressures 

and paddock numbers would be required to 

minimize the occurrence of multiple grazing 

events within individual grazing periods 

(Jensen et al. 1990a). These data challenge 

the widely held assumption that rotational 

grazing restricts grazing to a single event per 

plant during short (5–10-d) grazing periods 

while simultaneously promoting high plant 

utilization. This may indicate why minimal 

differences in plant defoliation patterns have 

been found between rotational and continuous 

grazing (Hart et al. 1993b). These detailed 

investigations were conducted on very 

small pasture sizes (0.2–24 ha), so caution 

is warranted when scaling these responses 

to larger areas. In addition, frequency of 

defoliation may be more detrimental to plants 

when defoliation events are separated by 

periods of regrowth, as indicated below. 

Frequency of tiller defoliation consistently 

increases with increasing stocking rate (9 of 

10 studies). Defoliation intensity of individual 

tillers also increases with increasing stocking 

rate, but not as rapidly as defoliation frequency 

(four of five studies). 

Forage quality decreased with increasing 

stocking rate within an individual grazing 

period in all four studies evaluated and with 

increasing time of grazing for all three studies 

that carefully evaluated this relationship. This 

clearly indicates that animals compete for 

quality forage, and this process establishes the 

basis for the negative response of individual 

animal performance with increasing stocking 

rate. 

Season of plant defoliation has unique and 

consistent effects on plant production. 

Individual plant production is most greatly 

suppressed by defoliation during the middle 

of the growing season, which coincides with 

culm elongation and the early boot stage 

of inflorescence development, especially in 

bunchgrasses (Olson and Richards 1988). This 

was documented in six of nine investigations 

and in all three studies specifically evaluating 

growth stage responses to defoliation. Early 

season defoliation had the least detrimental 

effect on subsequent plant production, and 

late season defoliation had an intermediate 

effect. However, plant production is 

increasingly suppressed with increasing 

frequency and intensity of defoliation at 

any stage of growth (five of six studies), 

confirming the interpretation that multiple 

defoliations within a growing season are 

detrimental to plant growth and function 

(Reece et al. 1996; Volesky et al. 2004). 

These patterns of grass production responses 

to defoliation at various phenological stages 

substantiate the criticism that has been 

directed toward early season deferment (i.e., 

range readiness) as a valid conservation 

practice. 

The occurrence of patch grazing has been well 

documented in several investigations, and it 

appears to directly relate to the nutritional 

intake of animals when other constraints on 

animal distribution are absent (e.g., distance 

to water and topography). Previously grazed 

patches support forage of higher nutritional 

quality, including crude protein, fiber, and 

digestibility, even though forage quantity 

may be less than on previously ungrazed 

patches (Cid and Brizuela 1998; Ganskopp 

and Bohnert 2006). The primary mechanism 

contributing to patch grazing is animal 

aversion to consumption of senescent plant 

28 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

material, especially current and previous year’s 

culms or stems (Ganskopp et al. 1992, 1993). 

Consequently, patch grazing may provide 

a nutritional benefit to animals at low and 

moderate stocking rates (Cid and Brizuela 

1998). 

Patch structure is relatively consistent within 

season and among years, but it is less stable at 

higher than at lower stocking rates (Willms et 

al. 1988; Cid and Brizuela 1998). At higher 

stocking rates, animals begin to selectively 

graze previously ungrazed patches to maintain 

sufficient forage intake, and they forage greater 

distances to achieve this goal (Ring et al. 1985; 

Ganskopp and Bohnert 2006). Patch grazing 

can be minimized by the removal of senescent 

biomass, especially previous year’s biomass 

with fire, mowing, or periodic heavy stocking 

(Ganskopp and Bohnert 2006). However, 

the implications of patch grazing have been 

shifting from that of an inefficient use of 

forage by livestock to a desirable component of 

vegetation heterogeneity capable of promoting 

biodiversity in the Great Plains (Fuhlendorf 

and Engle 2001, 2004). This is an especially 

relevant consideration, both within and among 

pastures, in light of the CEAP initiative, which 

emphasizes management for environmental 

quality and multiple ecosystem services as well 

as production goals. 

Grazing Systems. Grazing systems represent 

a specialization of grazing management that 

defines the periods of grazing and nongrazing 

(Heitschmidt and Taylor 1991; SRM 1998), 

and they have been given tremendous 

emphasis by both managers and researchers. 

It is important to recognize that constraints 

of experimental research, including the need 

for relatively homogeneous site conditions 

necessary for replication and comparison with 

experimental controls, has emphasized the 

potential for various periods of grazing and 

rest to alter the ecological processes controlling 

plant and animal production. They are unable 

to—and therefore do not—address livestock 

distribution in heterogeneous landscapes 

or livestock movement in response to site 

readiness along elevation gradients. However, 

these latter considerations are also important 

and have been addressed with experimental 

data collected with more appropriate 

experimental approaches. 

The major experimental investigations of 

grazing systems have been categorized by 

geographic location, ecosystem type, relative 

stocking rate, and number and size of pastures 

for each of the respective investigations (Briske 

et al. 2008). Variables were indicated to differ 

between continuous and rotational grazing only 

when they were reported as being statistically 

significant by the authors. For each experiment, 

plant and/or animal production (the most 

quantitative data collected) was characterized as 

1) greater for continuous grazing (CG > RG), 

2) greater for rotational grazing (RG > CG), 

or 3) equal if differences did not exist between 

continuous and rotational grazing (ND). These 

comparative responses were summarized and 

presented as separate histograms for those 

investigations that used similar stocking rates 

between grazing treatments (Fig. 3A), those 

that used greater stocking rates for rotational 

than for continuous grazing (Fig. 3B), and for 

all stocking rates combined (Fig. 3C). These 

experimental comparisons of rotational systems 

included five studies conducted for 9 yr or 

more, and four had pasture sizes greater than 

300 ha, but only two had greater than eight 

pastures per grazing system. 

Eighty-nine percent of the experiments (17 

of 19; Appendix I) reported no differences 

for plant production/standing crop between 

rotational and continuous grazing with similar 

stocking rates (Fig. 3A). When stocking 

rate was less for continuous than rotational 

grazing, 75% of the experiments (three of 

four) reported either no differences or greater 

plant production for continuous grazing (Fig. 

3B). Across all stocking rates, 83% of the 

experiments (19 of 23; Appendix I) reported 

no differences for plant production between 

rotational and continuous grazing, 13% (three) 

reported greater plant production for rotational 

compared to continuous grazing, and 4% (one) 

reported greater production for continuous 

grazing (Fig. 3C; Briske et al. 2008). 

Fifty-seven percent of the experiments (16 

of 28; Appendix I) reported no differences 

for animal production per head between 

rotational and continuous grazing with 

similar stocking rates, and 36% (10) reported 

greater per head production for continuous 

grazing (Fig. 3A). When stocking rate was 

less for continuous than rotational grazing, 

Grazing is a major land use 

on 188 million hectares of 

non-federal lands in the Great 

Plains and western U.S. (Photo: 

Sonja Smith) 

CHAPTER 1: An Assessment of Grazing Practices 29 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

fIGurE 3. Number of published grazing experiments that reported significantly 

higher, equal, or lower plant and animal production responses for continuous compared to rotational grazing at (A) similar stocking rates, (B) higher stocking rates for 

rotational grazing, and (c) across stocking rates for all experiments. Animal production is presented as both a per head and a per land area response (from Briske et 

al. 2008). 

90% of the experiments (9 of 10) reported 

either similar or greater per head animal 

production for continuous grazing (Fig. 3B). 

Across all stocking rates, 50% (19 of 38; 

Appendix I) of the experiments reported no 

differences for animal production per head 

between rotational and continuous grazing, 

8% (three) reported greater production for 

rotational grazing, and 42% (16) reported 

greater production for continuous grazing 

(Fig. 3C). Fifty-seven percent of the 

experiments (16 of 28; Appendix I) reported 

no differences for animal production per unit 

land area between rotational and continuous 

grazing with similar stocking rates, and 36% 

(10) reported advantages for continuous 

grazing (Fig. 3A). When stocking rate 

was lower for continuous than rotational 

grazing, 75% (three of four; Appendix I) 

of the experiments reported greater animal 

production per area for rotational grazing 

(Fig. 3B). Across all stocking rates, 50% 

(16 of 32; Appendix I) of the experiments 

reported no differences for animal 

production per land area between rotational 

and continuous grazing, 16% (five) reported 

greater production for rotational grazing, 

and 34% (11) reported greater production 

for continuous grazing (Fig. 3C; Briske et 

al. 2008). A recent ranch-scale investigation 

comparing four grazing systems over a 

7-yr period that was not included in this 

numerical assessment also reported minimal 

differences in livestock production among 

grazing systems (Pinchak et al. 2010). 

No evidence was found indicating that grazing 

systems override livestock preference for site 

selectivity. Comparisons of continuous seasonlong and rotational grazing on five range 

sites in northern mixed-grass prairie found 

no differences among grass utilization over a 

2-yr period (Kirby et al. 1986). This occurred 

in spite of the fact that the rotational system 

had both a higher stocking rate and a higher 

stock density than did the continuous system. 

Heitschmidt et al. (1989) corroborated these 

conclusions in mixed-grass prairie in northcentral Texas. Paddocks of 30 and 10 ha were 

used to simulate rotational grazing systems 

with 14 and 42 paddocks. Livestock selectivity 

was not modified by either rotational grazing 

system compared to continuous grazing. These 

authors concluded that forage availability, 

rather than stocking density or grazing system, 

was the primary mechanism that modifies 

animal selectivity. However, none of these 

investigations specifically addressed the presence 

of riparian systems in which livestock frequently 

congregate (George et al., this volume). 

Only four studies were found that directly 

compared forage quality in rotational and 

continuous grazing. Forage quality was 

30 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

	 	 	

	 	 	 		

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

comparable among systems in two of the 

investigations (Jung et al. 1985; Heitschmidt 

et al. 1987b), and one each favored continuous 

(Pfister et al. 1984) and rotational grazing 

(Heitschmidt et al. 1987a). Forage quality 

was greater for a seven-pasture short-duration 

system compared to a seven-pasture highintensity, low-frequency system, but similar to 

that of a Merrill four-pasture, three-herd system 

on the Edwards Plateau of Texas (Taylor et al. 

1980). Tiller defoliation patterns in continuous 

and rotational grazing have received only 

minimal attention, but frequency and intensity 

of tiller defoliation was greater for rotational 

grazing in only one (Senock et al. 1993) of 

four investigations (Hart et al. 1993a; Derner 

et al. 1994; Volesky 1994). Collectively, the 

small number of investigations reporting mixed 

results makes conclusions regarding grazing 

systems effects on forage quality and defoliation 

patterns equivocal compared to conclusions 

addressing plant and animal production, and 

species composition. 

Three categories of evidence exist to explain 

why intensive rotational grazing systems have 

not shown greater quantity and quality of 

forage and animal production in experimental 

research. First, short, periodic deferments 

based on established schedules do not always 

coincide with favorable growth conditions 

in rangeland environments (e.g., Taylor et 

al. 1993; Holechek et al. 2001; Gillen and 

Sims 2006). The amount and variability of 

rainfall and the associated predictability, 

duration, and amount of plant growth appear 

to override the potential benefit derived 

from the redistribution of grazing pressure 

in space and time in rotational grazing 

systems (O’Reagain and Turner 1992; Ash 

and Stafford Smith 1996; Holechek et al. 

2001; Ward et al. 2004). Plant growth and 

improvement in species composition will be 

promoted primarily when deferment coincides 

with environmental conditions favorable for 

plant growth (Heitschmidt et al. 2005; Gillen 

and Sims 2006). 

Improper grazing can 

detrimentally affect soil surface 

characteristics to accelerate 

runoff and erosion. (Photo: 

Ken Tate) 

CHAPTER 1: An Assessment of Grazing Practices 31 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	

	 	

	 	

	 	

These results 

refute prior 

claims that 

animal trampling 

associated with 

high stocking 

rates or grazing 

pressures in 

rotational grazing 

systems enhance 

soil properties 

and promote 

hydrological 

function”

Second, rotational grazing may not control the 

frequency and intensity of plant defoliation 

as effectively as often assumed (Gammon and 

Roberts 1978a, 1978b, 1978c; Hart et al. 

1993a). Investigations of tiller grazing patterns 

indicate that it is difficult to achieve a high 

percentage of tiller defoliation (> 80%) before 

multiple defoliations begin to occur within 

a single grazing period (Jensen et al. 1990a; 

O’Reagain and Grau 1995). These data indicate 

that grazing management strategies only 

marginally modify animal selectivity within the 

range of conditions that have been evaluated. 

Third, forage quality is not consistently or 

substantially increased in intensive systems 

compared to continuous grazing (Denny et 

al. 1977; Walker et al. 1989; Holechek et al. 

2000). The absence of experimental evidence 

supporting these three major underlying 

assumptions associated with rotational systems 

is consistent with the production responses 

generated from experimental comparisons of 

rotational and continuous grazing. However, 

conclusions addressing tiller defoliation patterns 

are derived from a small number of experiments 

conducted in very small pastures (0.2–24 

ha) that may not be entirely representative of 

grazing patterns at larger scales. 

These experimental results collectively indicate 

that rotational grazing does not promote 

primary or secondary production compared 

to continuous grazing within rangeland 

ecosystems. These interpretations are consistent 

with those of previous reviews over the past 50 

yr (Heady 1961; Van Poollen and Lacey 1979; 

Holechek et al. 2001), and they clearly support 

the long-standing conclusion that stocking rate 

and weather variation account for the majority 

of variability associated with plant and animal 

production on rangelands (Van Poollen and 

Lacey 1979; Heitschmidt and Taylor 1991; 

Gillen et al. 1998; Holechek et al. 2001; 

Derner and Hart 2007). 

Improve or Maintain surface and/or 

subsurface Water Quality and Quantity 

Stocking Rate. The response of soil 

hydrological characteristics to grazing largely 

parallel those of other ecological variables 

because stocking rate is the most important 

driver regardless of grazing system (Wood and 

Blackburn 1981a&b; Thurow 1991). This 

occurs because the removal of large amounts of 

plant cover and biomass by intensive grazing 

reduces the potential to dissipate the energy of 

raindrop impact and overland flow. The erosive 

energy of water and the long-term reduction of 

organic matter additions to soil detrimentally 

affect numerous soil properties, including 

the increase of bulk density, disruption of 

biotic crusts, reduced aggregate stability, and 

organic matter content, which collectively 

reduce infiltration rate and increase sediment 

yield and runoff. Animal trampling is 

another source of mechanical energy that 

breaks soil aggregates and is therefore 

negatively correlated with maintenance of 

soil structure necessary for high infiltration 

rates (Warren et al. 1986b; Thurow 1991; 

Holechek et al. 2000). These results refute 

prior claims that animal trampling associated 

with high stocking rates or grazing pressures 

in rotational grazing systems enhance soil 

properties and promote hydrological function 

(Savory and Parsons 1980; Savory 1988). 

These hydrological responses to grazing 

are strongly contingent on community 

composition, with communities that 

provide greater cover and obstruction to 

overland flow, such as midgrass-dominated 

communities having greater hydrological 

function, including infiltration rate, than 

shortgrass-dominated communities (Wood 

and Blackburn 1981b; Thurow 1991). 

Grazing System. Short-duration rotational 

grazing systems decreased soil hydrologic 

function at heavy to very heavy stocking rates, 

compared to continuous and deferred-rotation 

grazing systems at moderate to light stocking 

rates. The negative changes in vegetation 

and soil properties controlling infiltration, 

runoff, and soil loss due to heavy stocking 

rates generally cannot be overcome by grazing 

system. These collective results strongly refute 

claims that animal trampling associated 

with high stocking rates or intensities under 

intensive rotational grazing systems enhance 

hydrological function (Savory and Parsons 

1980; Savory 1988). 

There is evidence that soil hydrological 

functions degraded by heavy stocking rates can 

recover with prolonged rest (i.e., ≥ 1 yr). Thus, 

rotational grazing may maintain higher soil 

hydrologic function than continuous grazing 

32 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

at heavy to very heavy stocking rates if the 

deferment period is sufficient (i.e., ≥ 1 yr). 

Similarly, moderately stocked continuous or 

rotational grazing may maintain a consistently 

higher level of hydrologic function compared to 

periodic heavy stocking followed by prolonged 

deferment for hydrologic recovery. 

A few studies have directly examined grazing 

systems (deferred rotation, rest rotation, and 

rotational deferment) in comparison with 

continuous grazing. At moderate stocking 

rates, at which most extensive rotational 

systems were studied, rotational grazing 

systems lead to similar or improved soil 

hydrologic function compared to moderate 

continuous grazing (Ratliff et al. 1972; 

McGinty et al. 1979; Wood and Blackburn 

1981b, 1984). As evidenced by Wood and 

Blackburn (1981) and Thurow et al. (1986), 

these hydrological responses to grazing 

system appear to be strongly contingent 

on plant community composition, with 

midgrass-dominated communities having 

greater hydrological function than shortgrassdominated communities. Gifford and 

Hawkins (1976) emphasize the importance 

that range condition or plant community 

composition has on the hydrological function 

of a site through time in response to grazing 

system. 

Improve or Maintain riparian and 

Watershed function 

There is clear consensus that livestock grazing 

can degrade riparian plant communities, 

hydrologic function, and associated ecosystem 

services. Considerable management attention 

has been directed toward prescribed grazing 

practices with the intent to restore, enhance, 

or maintain rangeland riparian areas. As 

with upland habitats, it is clear that grazing 

intensity is a major factor determining riparian 

response to grazing management. Increased 

grazing intensity is generally associated 

with detrimental effects on riparian plant 

community composition and productivity 

as well as physical degradation of riparian 

soils and stream channels. These primary 

effects can lead to secondary negative effects 

on stream hydrologic functions, which can 

cascade to loss of services, such as fish habitat, 

flood attenuation, and provisioning of clean 

water. Management of grazing intensity is a 

viable conservation practice for riparian areas. 

Season of grazing also determines livestock 

grazing effects on riparian plant communities, 

particularly woody plants, and can be 

managed to conserve riparian habitats and 

their associated services. Livestock distribution 

practices such as water developments, 

supplement placement, and herding are 

effective means of managing the intensity and 

season of livestock grazing in riparian areas. 

Livestock exclusion is an effective practice 

to stimulate immediate recovery for riparian 

plant communities degraded by heavy grazing. 

While the individual effects of some prescribed 

grazing components (e.g., timing, intensity, and 

rest) on riparian habitats have been examined, 

few studies have rigorously examined the 

effects of different grazing systems on 

riparian habitats. The effectiveness of grazing 

management practices on the conservation 

of riparian habitats is covered in depth in the 

chapter on riparian herbaceous cover (George 

et al., this volume). 

reduce Accelerated soil Erosion and 

Maintain or Improve soil condition 

Soil vegetative cover is widely recognized as 

a critical factor in maintaining soil surface 

hydrologic condition and reducing soil 

erosion (Gifford 1985). High stocking rates, 

regardless of grazing system, that reduce soil 

surface vegetative cover below a site-specific 

threshold will increase detachment and 

mobilization of soil particles due to raindrop 

impact, decrease soil organic matter and 

soil aggregate stability, increase soil surface 

crusting and reduce soil surface porosity, and 

thus decrease infiltration and increase soil 

erosion and sediment transport (Blackburn 

1984). Regardless of grazing system, sufficient 

vegetative cover, critical soil cover, or residual 

biomass must remain during and following 

grazing to protect soil surface condition (e.g., 

porosity, aggregate stability, and organic matter 

content) and dependent hydrologic properties 

(e.g., infiltration). Site-specific vegetation cover 

requirements will vary depending on cover 

type (e.g., vegetation, litter, or rock), soil type, 

rainfall intensities, and water quality goals 

(Gifford 1985). 

The majority of research examining soil surface 

hydrologic response to grazing has focused 

on infiltration or proxies for infiltration, such 

Increased 

grazing intensity 

is generally 

associated with 

detrimental effects 

on riparian 

plant community 

composition and 

productivity as 

well as physical 

degradation of 

riparian soils and 

stream channels.”

CHAPTER 1: An Assessment of Grazing Practices 33 
fIGurE 4. Bird species responses to grazing intensity/management treatments for (A) 

shortgrass steppe, (B) mixed-grass prairie, and (c) tallgrass prairie. Drawn from data 

in Giezentanner (1970), Skinner (1975), Kantrud (1981), Kantrud and Kologiski 

(1982), and Milchunas et al. (1998). 

as dry bulk density and soil penetrability. A 

handful of studies have examined soil loss. 

Increased stocking rates from nongrazed to very 

heavy are associated with increased soil loss. 

As with infiltration results, light and moderate 

stocking rates are generally not different. There 

is no consistent result for the effect of grazing 

system on soil loss; in some cases, continuous 

systems are reported to have less soil loss, and 

in other studies, rotational systems are reported 

to have less soil loss. Most of these studies 

are confounded by comparisons of different 

stocking rates among systems, and several 

report that grazing system effect depended on 

plant community (e.g., shrub understory vs. 

interspace). There is no compelling evidence 

that rotational grazing strategies can reduce 

soil loss. Soil vegetative cover (responding to 

stocking rate) and inherent soil characteristics 

are key variables determining site scale soil loss 

(Pierson et al. 2002). 

Improve or Maintain the Quantity 

and Quality of food and/or cover 

Available for Wildlife 

Stocking Rate. Livestock and wildlife 

may directly compete for plant food 

resources, and livestock grazing can alter 

the composition, productivity, and quality 

of plant food resources. Grazing can alter 

community structure through removal of 

recent production and through longer-term 

effects on plant community composition and 

productivity. Cover represents an important 

component of wildlife habitat for escape and 

concealment from predation as well as for 

thermal regulation. Cover requirements for 

specific wildlife species often vary within a 

season and stage of life cycle (e.g., nesting 

vs. foraging). Bird (MacArthur 1965; Wiens 

1969; Cody 1985), rodent (French et al. 

1976; Grant and Birney 1979; Geier and Best 

1980; Grant et al. 1982; Kerley and Whitford 

2000), lagomorph (Flinders and Hansen 

1975), and lizard (Pianka 1966) community 

composition and diversity are often closely 

correlated with vegetation structure. Direct 

behavioral interactions between livestock and 

wildlife are another potential means by which 

grazing may affect wildlife populations. Social 

avoidance can preclude the use of otherwise 

suitable habitat, and it can be influenced by 

the numbers of livestock present (Roberts and 

Becker 1982; Stewart et al. 2002). Trampling 

of nests represents another possible mechanism 

of negative interaction between livestock 

and ground-nesting birds that increases with 

stocking rate (Jensen et al. 1990b). 

34 Conservation	Benefits	of	Rangeland	Practices
D.	D.	Briske,	J.	D.	Derner,	D.	G.	Milchunas,	and	K.	W.	Tate 

There are fewer studies documenting the 

responses of specific wildlife species or groups 

to stocking rate or grazing intensity than there 

are for plant communities. Therefore, studies 

published in the gray literature, including 

symposia and technical reports, have been 

included, but theses, dissertations, or non– 

data-based publications have not. Limited 

data availability also requires that inferences 

be drawn from individual studies rather than 

groups of studies, as has been done in other 

sections of this chapter. Wildlife responses are 

grouped into reptiles, birds, small mammals, 

and large ungulates to more effectively 

assess their potentially unique responses and 

interactions with livestock grazing. 

Reptiles. Ten studies reported on lizard 

communities in grazed versus ungrazed 

treatments, but only one study assessed lizard 

populations over five grazing intensities in 

Arizona (Jones 1979, 1981). The largest 

negative effect of heavy grazing on lizard 

density was found in Sonoran Desert grassland 

(−63%), followed by mixed scrub–dry wash 

(−54%), chaparral (−41%), and cottonwood– 

willow riparian (−20%), with no difference 

in desert scrub. Greater species richness was 

observed in lightly compared to heavily grazed 

desert grassland and cottonwood–willow 

riparian habitat, with no difference in the other 

three communities. The effects of grazing on 

lizard communities were related to differences 

in the cover of short (< 0.3 m) vegetation 

structure and litter cover, but not necessarily 

total vegetation cover. While lizard responses 

to grazing may be expected to be more 

pronounced than for other groups of organisms 

because of their relatively specific microhabitat 

requirements, there are insufficient studies over 

grazing intensities for generalizations to be 

drawn. 

Birds. Bird responses to stocking rate are well 

recognized as being species dependent and can 

be positive, negative, or neutral within any 

one location and treatment comparison (Bock 

et al. 1993; Saab et al. 1995; Knopf 1996). 

Unfortunately, most passerine bird studies 

have compared only grazed and ungrazed 

communities, and the intensity of grazing is 

often not reported. Derner et al. (unpublished 

data) reviewed 27 bird studies/habitats from 

the literature, and only 10 included more than 

fIGurE 5. Bird community (A) dissimilarity (Whittaker [1952] index of community association), (B) abundance (% difference between grazing intensity differential), and (c) 

diversity (H′) across grazing intensity gradients for North America studies. Dissimilarity 

index values range from 0.0 to 1.0, with a value of 0 indicating both treatments having all species in common and in the same proportions (0% dissimilar) and a value 

of 1.0 indicating no species in common (100% dissimilar). Data from Giezentanner 

(1970), Johnson and Springer (1972), Skinner (1975), Grzybowski (1980, 1982), 

Kantrud (1981), Kantrud and Kologiski (1982), and Milchunas et al. (1998). 

one grazing intensity in addition to the longterm ungrazed community. The abundance of 

individual species within a site can be strongly 

affected by grazing intensity. For example, 

CHAPTER	1:	An	Assessment	of	Grazing	Practices 35 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 · 

tABlE 1. Bird community dissimilarity, abundance (numbers), diversity, richness, and dominance in response to grazing averaged by region, 

evolutionary history of grazing, plant community life form, and plant community type. Forests were not included in region or evolutionary 

history categories. Plant community types are for major groupings or those with more than one comparison (from Derner et al., unpublished 

data). 

Birds 

Precipitation 

(average mm yr−1) 

dissimilarity 

(index) 

Abundance (high 

grazed % low 

grazed) 

diversity (high 

grazed /low 

grazed) 

richness (high 

grazed/low 

grazed) 

dominance (high 

grazed/low 

grazed) n 

By region 

Great Plains 487 0.40 38 1.18 1.02 1.12 38 

southwest1 362 0.54 3 0.91 0.90 1.29 4 

northwest1 154 0.54 −22 1.25 1.14 0.83 6 

other grasslands —2 0.54 −33 1.10 0.93 0.96 10 

By evolutionary history 

short history 617 0.53 −27 1.06 0.90 0.98 21 

long history 487 0.40 38 1.18 1.02 1.13 37 

By life form 

Grassland 483 0.43 30 1.18 1.02 1.09 40 

shrubland 291 0.45 −5 1.10 1.03 1.01 11 

forest 1 182 0.52 −45 0.96 0.68 1.06 7 

By community type 

shortgrass steppe 357 0.29 36 0.94 0.83 1.13 6 

Mixed-grass prairie 416 0.35 −2 0.91 0.90 1.25 23 

tallgrass prairie 988 0.61 217 2.52 1.71 0.59 6 

fescue grassland 383 0.60 −31 1.06 1.00 1.09 2 

coastal prairie —2 0.42 −9 0.96 0.54 0.72 2 

southwest grassland 362 0.54 3 0.91 0.90 1.29 4 

shadscale shrubland 154 0.42 −31 1.49 1.43 0.62 2 

1Northwest includes the Great Basin and all communities west of the Rocky Mountains, except for Arizona, New Mexico, and southern California, which are considered Southwest. 

2Number of sites reporting precipitation too few to provide a reasonable mean. 

horned larks respond positively to increasing 

grazing intensity in shortgrass steppe, while 

lark buntings respond negatively (Fig. 

4A). Chestnut-collared long-spurs respond 

positively to increasing grazing intensity in 

mixed-grass prairie, while savannah sparrows 

respond negatively (Fig. 4B). The greatest 

abundance of bird species in tallgrass prairie 

occurred at intermediate intensities of grazing 

(Fig. 4C). While species within a site respond 

differently to grazing intensity, a particular 

species may also have a varied response among 

sites. Knopf (1996) suggested that birds may 

not be generally classified as increasers or 

decreasers in response to grazing, but that 

individual species responses to grazing may 

vary over gradients of potential vegetation 

structure or aboveground primary production. 

Although there are examples for regional 

differences in bird species response to grazing, 

Derner et al. (unpublished data) concluded 

that data over gradients of grazing intensity 

and regional gradients of primary production 

are too limited to produce good models of bird 

preferences for particular grazing intensities 

at particular levels of primary production. 

Reviews by Bock et al. (1993) and Saab et 

al. (1995) provide tables of bird species by 

region within the western United States that 

show general positive, negative, primary 

productivity–dependent, or neutral/mixed/ 

uncertain responses to grazing. 

36 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

At the community level, the change in bird 

community composition relative to the 

ungrazed or lightly grazed condition usually 

increased with increasing grazing intensity 

(Fig. 5A; Table 1). However, dissimilarity was 

generally greater when the communities were 

ungrazed compared to lightly or moderately 

grazed than when grazing intensity further 

increased to moderate or heavy. Total bird 

community abundance showed both positive 

and negative responses with increasing grazing 

intensity across and within community types as 

anticipated (Fig. 5B). Bird community diversity 

was generally slightly negative with increasing 

grazing intensity (Fig. 5C). Exceptions were 

observed for one tallgrass prairie study and 

some mixed-grass prairie sites where slightly 

greater diversity occurred at intermediate levels 

of grazing intensity. In addition to these general 

diversity patterns, management decisions need 

to explicitly evaluate the specific habitat needs 

of bird species of concern. 

Most studies of grazing effects on upland game 

birds (gallinaceous birds) addressed ungrazed 

versus grazed conditions rather than grazing 

intensity gradients, much like research for all 

other wildlife groups. Based on two studies, 

wild turkeys prefer ungrazed/lightly grazed 

vegetation and avoid moderately/heavily 

grazed areas. Similarly, heavy grazing was 

consistently detrimental to sharp-tailed grouse 

(three subspecies) because of a loss of nesting 

cover and tree and shrub density (based on 10 

studies reviewed in Kessler and Bosch 1982). 

There are contrasting positive and negative 

results from ungrazed/grazed studies for sage 

grouse and prairie chickens, but sage grouse 

appear to prefer light/moderate grazed areas 

over heavy grazed areas, but very high cover in 

some ungrazed habitat may be avoided as well 

(some reviewed in Beck and Mitchell 2000). 

Historical evidence suggests that grazing is 

detrimental to quail species in the southwestern 

United States, but recent studies indicate 

that light to moderate grazing intensities may 

be beneficial to Mearn’s quail by increasing 

availability of food resources. Montezuma 

quail prefer high grass cover and tree density, 

while scaled quail prefer high grass cover and 

low tree density. In contrast, five studies of 

bobwhite quail in Texas (see Bryant et al. 1982) 

suggest that grazing is beneficial if intensities 

are not too high. In summary, heavy grazing 

most often results in loss of cover below some 

optimal level for gallinaceous birds, although 

light grazing may be beneficial under some 

circumstances. 

Small Mammals. Small mammals can be 

sensitive to changes in vegetation structure, 

but they may also be affected by grazing 

fIGurE 6. Rodent species abundance across grazing intensities in (A) shortgrass steppe and (B) mixedgrass prairie. Drawn from data in McCulloch (1959) and Grant et al. (1982). 

CHAPTER 1: An Assessment of Grazing Practices 37 
	 	 	 	

 

 

 

tABlE 2. Rodent community dissimilarity, abundance (numbers), diversity, richness, and dominance in response to grazing averaged by 

region, evolutionary history of grazing, plant community life form, and plant community type. Forests were not included in region or evolutionary history categories. Plant community types are for major groupings or those with more than one comparison (from Derner et al., unpublished data). 

rodents 

dissimilarity 

(index) 

Abundance 

(high grazed 

% low grazed) 

diversity (high 

grazed/low 

grazed) 

richness (high 

grazed/low 

grazed) 

dominance 

(high grazed/ 

low grazed) 

unique species 

(high grazed/low 

grazed) n 

By region 

Great Plains 0.35 −27 0.99 0.89 1.11 −2.0 14 

southwest1 0.34 24 0.89 0.85 1.41 −1.0 6 

northwest1 0.43 8 0.81 0.95 1.60 −0.8 19 

By evolutionary history 

short history 0.41 12 0.83 0.93 1.55 −0.9 25 

long history 0.35 −27 0.99 0.87 1.11 −2.0 14 

By life form 

desert 0.18 −43 0.85 0.73 1.40 −1.5 2 

Grassland 0.34 13 0.73 0.92 1.30 −1.4 4 

shrubland 0.38 −0 0.73 0.76 1.62 −1.7 12 

savanna 0.43 14 0.92 1.51 1.53 1.0 3 

forest 0.30 −55 0.82 0.75 0.96 −1.0 2 

By community 

shortgrass steppe 0.19 −9 1.24 1.0 0.68 0.0 1 

Mixed-grass prairie 0.32 −18 0.81 0.78 1.29 −2.9 9 

Grassland 0.47 −13 0.79 0.62 1.53 −5.8 4 

sand sage shrub 0.19 −23 0.82 0.91 1.10 −0.6 5 

tallgrass prairie 0.48 −50 1.34 1.1 0.80 −0.5 4 

desert grassland 0.41 58 0.91 0.92 1.41 −0.8 4 

shadscale shrubland 0.53 1 0.57 0.6 2.36 −2.0 2 

Atriplex shrubland 0.52 −27 0.63 0.96 1.81 0.0 2 

sagebrush shrubland 0.30 30 0.76 0.86 1.49 −1.5 6 

northwest grassland 2 0.34 13 0.73 0.95 1.30 −0.3 4 

1Northwest includes the Great Basin and all communities west of the Rocky Mountains, except for Arizona, New Mexico, and southern California, which are considered Southwest. 

2See Savanna for Northwest savannas. 

induced modification of seed and arthropod of individual species of small mammals to 

food resources. Derner et al. (unpublished grazing intensities are similar to birds, but 

data) reviewed 24 rodent studies/habitats they differ from birds at the community level. 

from the literature, and only six included Like birds, some rodent species are favored by 

more than one grazing intensity in addition to grazing, some decline, and others are relatively 

long-term ungrazed exclosures. The responses neutral (Fig. 6 A and B). The response of some 

38 Conservation Benefits of Rangeland Practices 
D.	D.	Briske,	J.	D.	Derner,	D.	G.	Milchunas,	and	K.	W.	Tate 

fIGurE 7. Rodent community (A) dissimilarity (Whittaker [1952] index of community association), (B) 

abundance (% difference between grazing intensity 

differential), and (c) diversity (H′) across grazing 

intensity gradients for North America studies. Dissimilarity index values range from 0.0 to 1.0, with 

a value of 0 indicating both treatments having all 

species in common and in the same proportions 

(0% dissimilar) and a value of 1.0 indicating no 

species in common (100% dissimilar). Data from 

Frank (1940), Smith (1940), McCulloch (1959), 

Grant et al. (1982), Rice and Smith (1988), and 

Bich et al. (1995). 

species to grazing intensity can be substantial. 

Generalizations concerning rodent responses 

to livestock grazing intensity are less developed 

than those for birds, in part because of fewer 

studies but also because of less consistent 

population responses. 

Derner et al. (unpublished data) assessed the 

number of rodent species unique to various 

grazing intensities to evaluate the general 

patterns of declining rodent diversity with 

increasing grazing intensity. Greater numbers 

of species were likely to be captured on 

ungrazed or lightly grazed communities than 

on moderately or heavily grazed communities. 

However, 19 of 41 cases also displayed 

species unique to the more intensively grazed 

communities as well, but in only five cases 

was the total number of unique species greater 

on the more intensively grazed community. 

The net effects by region and evolutionary 

history were unique and unexpected. The 

numbers of unique species associated with 

heavy grazing were smaller in the Great 

Plains than in the Southwest or Northwest 

and in ecoregions with long rather than short 

evolutionary histories of grazing (Table 2). 

Deserts, grasslands, and shrublands displayed 

somewhat similar reductions in rodent species 

with increasing grazing intensity, and the 

losses were greater compared to savannas and 

forests. The greatest reductions in rodent 

species with increasing grazing intensities 

occurred in mixed-grass prairie. In general, 

no consistent trends could be discerned for 

changes in rodent species composition with 

grazing intensity relative to ungrazed or 

lightly grazed condition (dissimilarity; Fig. 

7A) or abundance; Fig. 7B). Rodent diversity 

generally declines or is unchanged with 

increasing grazing intensity, with the exception 

of a shortgrass steppe study (Fig. 7C). Declines 

in rodent diversity with grazing intensity 

were only small to moderate when they were 

observed. 

CHAPTER	1:	An	Assessment	of	Grazing	Practices 39 
	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	

Bird responses to grazing are 

highly species specific and 

positive, negative and neutral 

outcomes occur. (Photo: USDA: 

Gary Kramer) 

Other Small Mammals. Heavily grazed 

or “overgrazed” communities are generally 

preferred over ungrazed or lightly grazed 

communities by black-tailed jackrabbits in 

eastern Texas (Taylor and Lay 1944), the 

Mojave Desert in California (Brooks 1999), the 

sand hills of Colorado (Sanderson 1959), and 

southern Arizona (Taylor et al. 1935) and by 

Great Plains jackrabbits in mixed-grass (Smith 

1940) and tallgrass prairie (Phillips 1936) of 

Oklahoma. Schmutz et al. (1992) observed 

that rabbits became more abundant as range 

conditions deteriorated in desert grassland. 

MacMahon and Wagner (1985) suggested that 

many areas of the Chihuahuan and Sonoran 

deserts initially altered by fire suppression and 

livestock grazing do not return to previous 

conditions when large herbivores are excluded 

because lagomorphs and rodents, favored by 

the initial changes, maintain the vegetation 

at early seral stages. In contrast, Flinders and 

Hansen (1975) found that cottontail rabbits 

were more abundant in moderately than in 

either lightly or heavily grazed shortgrass 

steppe, white-tailed jackrabbits showed no 

preference, and black-tailed jackrabbits were 

more abundant in lightly and moderately 

compared to heavily grazed communities. 

Changes beneficial to rabbits with increasing 

grazing intensity include increased rabbit 

mobility and improved forage due to increases 

in annuals. 

Wild Ungulates. There is a large body of 

research addressing dietary and habitat use 

overlap between livestock and deer and elk. 

In general, high dietary overlap is observed 

between cattle or sheep and elk, compared 

with much lower overlap between cattle or 

sheep and deer (Skovlin et al. 1968; Mackie 

1970; MacCracken and Hansen 1981; Berg 

and Hudson 1982; Loft et al. 1991). However, 

dietary overlap between deer and cattle can 

increase with increasing intensity of cattle 

grazing (Mackie 1981; Vavra et al. 1982; 

Severson and Medina 1983). Habitat use is 

often separated in time because of seasonal 

migrations of deer and elk and in space 

because of topography or cover requirements 

(Skovlin et al. 1968; Mackie 1970; Berg and 

Hudson 1982). For example, mule deer, elk, 

and cattle observations on slopes steeper than 

10 degrees averaged 50%, 42%, and 18%, 

respectively. Dietary overlap between domestic 

and native herbivores is generally greatest 

during the period in which the herbivores are 

most nutritionally stressed (Olsen and Hansen 

1977; Mackie 1981), and habitat overlap is 

most likely to occur when wildlife are at lower 

elevations during winter, which often represents 

the period of greatest nutritional stress (Wallmo 

et al. 1981). 

Increasing grazing intensities by livestock are 

likely to create a bottleneck in the quantity 

and quality of forage for wild ungulates during 

nutritionally stressed periods (e.g., winter 

or drought). More generalist, large-rumen 

livestock are better able to utilize dormant grass 

forage than deer under conditions of low forage 

availability in heavily compared to moderately 

or lightly stocked pastures (MacMahan and 

Ramsey 1965). Dietary overlap between cattle 

40 Conservation Benefits of Rangeland Practices 
D.	D.	Briske,	J.	D.	Derner,	D.	G.	Milchunas,	and	K.	W.	Tate 

and pronghorn is low, and Schwartz et al. 

(1977) found that pronghorn were able to 

maintain seasonal diet qualities on long-term 

pastures heavily grazed by cattle similar to what 

they did on lightly grazed pastures in shortgrass 

steppe under nondrought conditions. In 

contrast, a pronghorn die-off was attributed to 

very heavy grazing by domestic animals during 

a drought (Hailey et al. 1966). Other studies 

of livestock grazing effects on pronghorn 

populations also show mixed responses. 

In contrast, habitat overlap is a prerequisite 

to facilitation of one herbivore by another. 

Positive or facilitative effects of livestock 

grazing on associated wildlife species may 

result from a reduction in the amount of 

unpalatable, standing dead material (Short 

and Knight 2003) or increased protein content 

and digestibility of forage available late in 

the season (Clark et al. 2000; see below). 

Both competition and facilitation can act 

simultaneously, and competition can be the 

strongest factor (Hobbs et al. 1995). Longerterm facilitative relationships may be based on 

a dichotomy in diet preference of grass versus 

forbs and shrubs. For example, grazing by deer 

and livestock can potentially shift community 

composition toward a composition favored by 

other species of herbivores. 

Grazing can potentially be used as a tool to 

enhance wildlife populations, and this may 

be particularly true when season of grazing or 

deferment of grazing is used to meet specific 

wildlife goals. In some situations, wildlife 

and livestock may overlap in habitat use 

only during particular times of the year. For 

examples, breeding birds may nest only during 

spring/early summer and require specific 

conditions during that time. Elk and deer 

may move down from forested mountainous 

habitat during the winter to occupy foothills 

and plains more likely to be used for livestock 

grazing. Grazing may be imposed or deferred, 

depending on cover and foraging requirements 

of specific species. Some waterfowl or some 

upland game species require dense nesting 

cover, whereas some birds, such as mountain 

plover or curlews, choose nesting sites with 

very little cover and will not nest in ungrazed 

or lightly grazed habitat. Many of the examples 

of season-of-use studies come from wildlife 

refuges or experimental sites where livestock 

fIGurE 8. Responses of nine wildlife categories to rotational compared to continuous 

livestock grazing systems summarized as neutral, negative, or positive. Each wildlife 

category is subdivided into a population (upper) and a habitat (lower) section to indicate the mechanism of livestock impact. The large dark bars represent studies without 

confounding experimental designs, and the small lightly shaded bars represent studies with design problems. Each study may have one to four response variables, so 

each bar does not represent a single study. 

use is optional and management options are 

flexible. 

Successful use of season of grazing may result 

from a facilitation effect of grazing by livestock 

on forage for other ungulates. Alpe et al. (1999) 

showed that early summer grazing improved 

forage quality for wild ungulates in autumn and 

winter if livestock was removed in time to allow 

sufficient regrowth, but that late season livestock 

grazing decreased forage quality for wild 

CHAPTER	1:	An	Assessment	of	Grazing	Practices 41 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	

ungulates. Successful use of season/deferment of 

grazing may also be possible when pastures that 

are not frequently used by wildlife are available; 

otherwise the removal of livestock from one 

pasture must outweigh the effects of increased 

stocking rate in adjacent pastures. Even in these 

cases, experimental outcomes can be neutral, 

positive, negative, or mixed, depending on 

wildlife species or timing of grazing (Medin 

1986; Alpe et al. 1999; Mathis et al. 2006). 

Grazing Systems. Wildlife responses to 

rotational and continuous grazing at relatively 

similar grazing intensities and within similar 

plant communities are evaluated in this section. 

Studies investigating different pastures within 

the same grazing system are also considered 

separately and clearly identified when used. 

Studies are organized by wildlife taxonomic 

groups and summarized across all groups and 

mechanisms for positive, neutral, or negative 

responses to grazing system. 

Birds. Although passerine birds represent the 

most studied group of wildlife in response to 

grazing intensity (see section above), only two 

published studies of rotational compared to 

Ungulate responses to grazing 

are equivocal so that no broad 

conclusions can be drawn. 

(Photo: USDA: Gary Kramer) 

continuous grazing were located. However, 

there are a number of unpublished theses, 

studies within pastures of an individual 

system or that compare rotational grazing 

with ungrazed communities that were not 

considered here. The dissertation of Kempema 

(2007) was unique because it assessed several 

grazing periods of increasing duration, so it 

is summarized here. These studies report that 

passerine responses to grazing systems compared 

to continuous grazing were most often neutral 

(Fig. 8). The rotational systems had the least 

vegetative heterogeneity at both small and large 

spatial scales because of the reduced capacity for 

selective grazing at the bite and patch scale by 

livestock compared to continuous grazing. This 

was accompanied by a decrease in bird species 

richness with decreasing duration of grazing 

(long-continuous richest). In contrast, the 

short-duration system had the highest densities 

of the most species. For most bird species (11), 

there was no significant grazing system effect on 

density, and for the three species that showed 

significant density effects, the responses were 

both positive and negative. Nest success was also 

similar among the three grazing systems. The 

small number of passerine studies specifically 

conducted in grazing systems precludes the 

development of general conclusions. 

Grazing systems studies on gallinaceous birds 

frequently evaluated nest trampling or nest 

predation, often conducted with artificial 

nests. All studies without confounded designs 

show neutral responses to grazing system (Fig. 

8). Larger densities of livestock in smaller 

pastures of rotational systems do not appear to 

increase trampling losses under the densities 

and in the habitats studied. A higher density 

of livestock in some pastures and longer rests 

in others appears to produce a similar mean 

effect. Trampling of nests is often found 

to increase linearly with stocking density 

(Bientema and Müskens 1987; Paine et al. 

1996; Kempema 2007), and the ecological 

significance of nest trampling is greater in more 

productive ecosystems that support greater 

stocking density. In contrast, Koerth et al. 

(1983) found trampling losses of nests to be 

similar between short-duration and continuous 

grazing, even though the short-duration 

system was stocked at a higher rate (5.3 vs. 

8.0 ha · steer−1). Nest trampling may not be 

linear with stocking density because livestock 

42 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	

	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

may travel less in smaller pastures (Koerth et 

al. 1983 and citations therein). Alternatively, 

a reduction in diet selection may increase 

search time exploring new pastures during 

repeated rotations (Spalinger and Hobbs 1992; 

Wilmshurst et al. 1999), and large herd size 

may result in more temporally constant activity 

levels among livestock (Paine et al. 1996). 

Direct studies of population or habitat 

responses between grazing systems are few, 

but positive responses have been reported 

for rotational grazing systems compared to 

continuous grazing for bobwhite quail in 

response to increased bare ground and greater 

forb densities (Fig. 8). There are too few studies 

for sharptail or turkey to draw any meaningful 

conclusions concerning the effect of grazing 

system. No grazing systems studies were found 

for prairie chickens or sage grouse, but some 

management recommendations have been 

made, including multiyear periods of rest to 

restore vegetative cover (Hagen et al. 2004). 

Vegetative cover is an important habitat 

requirement for waterfowl, although very 

dense vegetation can be detrimental to nestsite selection (Kantrud 1990). Ignatiuk and 

Duncan (2001) observed no difference in duck 

nest success in an extensive study of once-over 

rest-rotation or deferred-rotation systems and 

continuous grazing, while additional studies 

compared only pastures within grazing systems 

or conditions following changes in grazing 

regime (Fig. 8). When rest periods were 

from 1 to 3 yr, Gilbert et al. (1996) observed 

increasing duck nest densities with increasing 

years of rest, and regression analyses suggested 

that a 6- to 7-yr rest would be necessary for 

recovery to that of an ungrazed condition. 

Other waterfowl studies with confounding 

experimental designs also suggest that long rest 

periods may be beneficial, but there are too few 

waterfowl studies of grazing systems to form 

robust conclusions. 

Large Ungulates. Grazing systems research has 

been conducted with elk, deer, and pronghorn 

antelope, but the pronghorn study compared 

rest-rotation only with ungrazed pastures. Eight 

studies that included 18 response variables 

were found comparing grazing systems 

with continuous grazing for deer. The most 

common deer response to grazing system was 

negative, followed by neutral and then positive 

responses (Fig. 8). However, population-level 

responses for deer were equally split between 

positive and neutral for rotational compared to 

continuous grazing. Studies assessing habitat 

characteristics important to deer were most 

often negative in rotational grazing systems 

compared to continuous. Only one grazing 

system study reported on social avoidance by 

deer, and it showed deer–livestock competition 

in the short-duration system compared to 

continuous grazing that was attributed to 

habitat modification rather than deer leaving 

the pasture (Cohen et al. 1989). Responses of 

deer to rotational systems are generally mixed 

so that no clear trends can be established. 

Only one study was located that directly 

compared elk responses in a deferred-rotation 

grazing system compared to season-long 

grazing and found no significant response 

when averaged over grazing intensities (Fig. 8) 

but did find a highly significant interaction of 

grazing system with grazing intensity (Skovlin 

et al. 1968, 1975, 1983). Elk preferred seasonlong to deferred rotation at the light grazing 

intensity, but preferred deferred to season-long 

rotation at the high grazing intensity. Elk 

utilized individual plants that had not been 

grazed by cattle, and cattle use of numbers of 

individual plants at the low grazing intensity 

was greater under the rotation system. Forage 

quantity and preference for areas receiving little 

or no prior current-year use by livestock can 

regulate elk movement across larger landscapes 

as well (Mackie 1970). These results and 

the observation that elk preference strongly 

increased with decreasing grazing intensity 

even from light to ungrazed treatments are 

in accordance with the within-system studies 

showing a high degree of elk sensitivity to 

livestock grazing intensity and selection for 

ungrazed units or treatments, unutilized/ 

little utilized areas within grazed pastures, 

and ungrazed individual plants. However, 

the studies within various pastures of a single 

rotational grazing system are often cited to 

support rotational grazing as benefiting elk 

populations. Three of these studies found no 

social avoidance between elk and livestock for 

selection of ungrazed pastures. Livestock grazing 

facilitated use by elk the year following grazing 

in two studies, and elk avoided currently and 

previously grazed pastures in the other study. 

...the most 

frequent wildlife 

response was 

no differences 

between 

continuous and 

rotational grazing 

systems, with the 

remaining cases 

equally divided 

among positive 

and negative.” 

CHAPTER 1: An Assessment of Grazing Practices 43 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	

	 	 	

	 	

	 	 	

	 	

	 	 	

The response 

of soil organic 

carbon to 

stocking rate is 

equivocal, based 

partially on the 

limited number 

of investigations 

that have been 

conducted.”

Summary of Wildlife and Grazing Systems. 

The limited number of available studies does 

not permit generalizations concerning wildlife 

responses to grazing systems and when or 

where or for which species positive, negative, 

or neutral responses may be predicted. There 

appear to be many false claims and few valid 

studies in the literature (Kirby et al. 1992), 

and this assessment applies to the literature 

addressing wildlife responses to grazing 

systems. Collectively, comparative wildlife 

responses to rotational and continuous grazing 

were that 17 showed no difference, eight 

were negative, and eight were positive (Fig. 

8). These experimental data indicate that 

the most frequent wildlife response was no 

differences between continuous and rotational 

grazing systems, with the remaining cases 

equally divided among positive and negative. 

However, most wildlife groups showed mixed 

responses to grazing system, and it is clear that 

there are conditions where rotational grazing 

systems benefit a wildlife species or group, but 

the opposite response is documented as well. 

Much more is known about wildlife responses 

to grazing intensities than grazing systems, but 

even here the majority of studies assess grazed 

and long-term ungrazed communities, which 

are generally not relevant to prescribed grazing 

management (Krausman et al., this volume). 

Manage fine fuel loads to Achieve 

desired conditions 

Grazing does reduce fine fuel loads, and it 

can therefore modify both fire frequency and 

intensity (Belsky and Blumenthal 1997; Briggs 

et al. 2002; Fuhlendorf and Engle 2004). 

This interpretation is supported by the welldocumented inverse relationship between 

stocking rate and aboveground herbaceous 

standing crop (Bement 1969; Milchunas and 

Lauenroth 1993; Manley et al. 1997; Derner 

and Hart 2007). It is often hypothesized 

that woody plant encroachment is partially a 

consequence of reduced fire regimes associated 

with livestock grazing (Scholes and Archer 

1997; Swetnam and Betancourt 1998; 

Briggs et al. 2005). However, beyond these 

broad generalizations, there are only limited 

experiential data to support grazing as a means 

of fuel management (Belsky and Blumenthal 

1997; Davies et al. 2010). This is perhaps 

not that surprising given that fire–grazing 

interactions are strongly influenced by site, 

year, season, and specific fire conditions (Davies 

et al. 2009). 

Patterns of fire and grazing appear to be 

critically linked on the landscape (Fuhlendorf 

and Engle 2004). Grazing may increase the 

variability on fire occurrence by reducing the 

amount and increasing the heterogeneity of fine 

fuel distribution (Holdo et al. 2009). Grazed 

patches have less fine fuel that is less likely to 

burn than ungrazed patches that contain larger 

amounts of combustible fine fuel (Collins and 

Smith 2006; Kirby et al. 2007). However, 

grazing increased fuel homogeneity in a 

bunchgrass-dominated rangeland by reducing 

biomass of individual plants to a greater extent 

than biomass in the plant interspaces (Davies et 

al. 2010). 

Weather and fuel conditions further increase 

the complexity of the relationship between 

fuel load and fire frequency and intensity. For 

example, fine fuel load is strongly correlated 

with fire intensity when fuel moisture is held 

constant, but when fuel moisture is low, 

intense fires can be carried by much lower 

fuel loads (Twidwell et al. 2009). This will 

be influenced by the season, time of day, and 

specific weather conditions associated with 

individual fires. It is no coincidence that most 

wildfires occur during extreme fire conditions; 

during these extreme conditions, fire can 

be carried by a wide range of fuel loads. 

Therefore, it should not be assumed that fire 

frequency and intensity decrease linearly with 

decreasing fuel loads resulting from greater 

grazing intensities. 

The relative proportions of fine and coarse fuel 

loads can also influence the relationship between 

grazing and fire frequency and intensity. Woody 

plant encroachment is often associated with 

a reduction in the amount of fine fuel, but 

coarse fuel loads often increase substantially 

(Hibbard et al. 2001; Norris et al. 2001; Briggs 

et al. 2002). Although coarse fuels have higher 

ignition temperatures, closed-canopy woodlands 

can be highly flammable during extreme fire 

conditions. Therefore, the role of grazing as a 

tool for fuel management is generally supported, 

but it should be cautiously evaluated on a caseby-case basis because fire potential in influenced 

by interactions among several ecosystem 

variables (Fuhlendorf et al., this volume). 

44 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

AssocIAtEd consIdErAtIons 

livestock distribution 

Animal selectivity and foraging behavior 

within landscapes has received considerable 

attention on rangelands (Bailey et al. 1996; 

Launchbaugh and Howery 2005). Herbivores 

naturally select preferred plants and landscape 

positions over others (Van Soest 1994), 

resulting in differential patterns of species 

use within communities and management 

units when stocking rates are not excessive 

and pastures are of sufficient size (Bailey et 

al. 1996; Launchbaugh and Howery 2005). 

Rangelands have traditionally been managed 

to increase uniformity of vegetation use by 

livestock and maximize livestock gains within 

the limits of individual animal performance 

and long-term ecosystem sustainability 

(Bement 1969). This management approach 

has been effective and sustainable from the 

standpoint of livestock and forage production 

(e.g., Hart and Ashby 1998), but it often 

does not mimic the pattern of historic 

disturbance regimes (Fuhlendorf and Engle 

2001) or create habitat structure required for 

many grassland bird species (Knopf 1996; 

see Deferment and Rest section below). 

Livestock distribution and grazing behavior 

can be modified by adjusting the location of 

supplemental feed and water, implementation 

of patch burns, and herding (Williams 1954; 

Ganskopp 2001; Fuhlendorf and Engle 2004; 

Bailey 2005) in addition to the traditional 

practice of fencing. 

Experimental data evaluating the most critical 

variables associated with livestock distribution 

were evaluated from 51 studies and two 

reviews. Treatment responses were categorized 

into 1) general distribution effects, 2) steepslope use, 3) high-elevation use, 4) distance 

from water, 5) plant preferences, 6) uniformity (e.g., fences, salt, and water placement) to 

of grazing, and 7) riparian use. All 51 studies the modification of animal behavior (e.g., 

were short term (< 5 yr), and the vast majority attractants, genetic selection, breeds, and type 

of them used cattle as the livestock species of animal) over the past two decades. Livestock 

(41). Pasture sizes used in these investigations distribution in response to specific conservation 

were generally large (22 > 200 ha). Recent practices have received relatively little attention 

investigations have incorporated technological with the exception of prescribed burning (see 

advances involving GPS devices (e.g., collars) to Fuhlendorf et al., this volume). 

track individual animal movement to provide 

spatial- and temporal-explicit use patterns. The experimental data verify that many of 

Strategies for modifying patterns of livestock the common assumptions regarding livestock 

distribution have shifted from specific practices distribution and preferences for specific sites 

Rangelands play an important 

role in the global carbon cycle 

because of the large reservoirs 

of organic and inorganic 

carbon they contain. (Photo: 

Brandon Bestelmeyer) 

CHAPTER 1: An Assessment of Grazing Practices 45 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	 	

	 	 	

	 	 	 	 	

	 	 	

Prescribed grazing must 

balance the forage demand of 

animals with the physiological 

requirements of plants to be 

sustainable. (Photo: USDA: 

Lynn Betts) 

and conditions are valid. Water distribution 

(11 of 15 studies), steep slopes, and high 

elevations (13 of 17 studies) unequivocally 

influenced livestock distribution. Livestock by 

and large prefer riparian to upland areas (e.g., 

Bowns 1971; Smith et al. 1992; Howery et 

al. 1996, 1998), burned to nonburned areas 

(Coppedge and Shaw 1998; Biondini et al. 

1999), previously grazed compared to ungrazed 

areas (Ganskoppp and Bohnert 2006), and 

fertilized to nonfertilized areas (Samuel et al. 

1980). Range riding and/or herding of animals 

also effectively modified livestock distribution 

(Skovlin 1957; Bailey et al. 2008). A clear 

exception to these generalizations is that salt 

location has only a minor influence on grazing 

distribution within a growing season (five 

of seven studies; Ganskopp 2001). Standard 

approaches to modifying livestock distribution 

are warranted, but it appears that they can only 

minimize animal selection and preferences 

rather than completely eliminate them (Jensen 

et al. 1990a). 

Grazing and soil organic carbon 

Rangelands play an important role in the global 

C cycle because of 1) an extensive land area, 

2) large reservoir of sequestered C that could 

be released back into the atmosphere with 

improper management, 3) potential for high 

rates of soil organic carbon (SOC) accumulation 

by restoration of degraded rangelands, and 4) 

a vast pool of soil inorganic C as carbonates in 

semiarid and arid rangeland soils that may allow 

sequestration or release of CO2 (Schuman et 

al. 1999; Derner and Schuman 2007; Svejcar 

et al. 2008). SOC sequestration is influenced 

by climate (Derner et al. 2006), biome type 

(Conant et al. 2001), management (grazing, N 

inputs, restoration, and fire; Follett et al. 2001; 

Mortenson et al. 2004; Derner and Schuman 

2007; Bremer and Ham 2010; Pineiro et al. 

2010), and environmental conditions (drought 

and climate change; Jones and Donnelly 

2004; Ingram et al. 2008; Svejcar et al. 2008). 

Rangelands are typically characterized by short 

periods of high C uptake (2–3 mo · yr−1), long 

periods of C balance or small losses (Svejcar 

et al. 2008), and climate-driven interannual 

variability in net ecosystem exchange (Zhang et 

al. 2010). Three main drivers that will control 

the fate of C sequestration in rangelands are 1) 

long-term changes in production and quality 

of above- and belowground biomass; 2) longterm changes in the global environment, such 

as rising temperatures, altered precipitation 

patterns, and rising CO2 concentrations, that 

affect plant community composition and forage 

quality; and 3) effects of short-term weather 

conditions (e.g., droughts) and interannual 

variability in climate on net C exchange (Ciais 

et al. 2005; Soussana and Lüschert 2007; 

Ingram et al. 2008; Svejcar et al. 2008; Zhang 

et al. 2010). 

46 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

Application of appropriate management 

practices, such as proper stocking rates, adaptive 

management, and destocking during drought 

conditions on poorly managed rangelands 

(113 M ha), could result in sequestration of 

11 Tg C · yr−1, and continuation of sustainable 

management practices on the remaining 

rangelands would avoid losses of 43 Tg C · yr−1 

(Schuman et al. 2001). 

SOC sequestration rates decrease with 

longevity of the management practice 

(Derner and Schuman 2007), indicating 

that ecosystems reach a “steady state” and 

that changes in inputs would be required to 

sequester additional C (Conant et al. 2001, 

2003; Swift 2001). The response of SOC to 

stocking rate is equivocal, based partially on 

the limited number of investigations that have 

been conducted. Sixty-two percent (five of 

eight) of the investigations showed no response 

of SOC to stocking rate (Smoliak et al. 1972; 

Wood and Blackburn 1984; Warren et al. 

1986a; Biondini et al. 1998; Schuman et al. 

1999) with one showing a decrease (Ingram 

et al. 2008) and two showing an increase in 

response to increasing stocking rate (Manley 

et al. 1995; Reeder and Schuman 2002). The 

two investigations showing an increase in SOC 

with increasing stocking rate occurred in the 

northern mixed-grass prairie during a relatively 

wet period (Manley et al. 1995; Reeder and 

Schuman 2002). It has been demonstrated 

that increasing SOC in these grasslands may 

partially result from increasing dominance 

of the shallow-rooted, grazing-resistant 

species blue grama (Bouteloua gracilis), which 

incorporates a larger amount of root mass in 

the upper soil profile than do midgrass species 

that it replaces (Derner et al. 2006). In a global 

analysis, Milchunas and Lauenroth (1993) 

found that in 19 of 34 comparisons, SOC was 

less in grazed than ungrazed communities, and 

results were similarly mixed for root biomass. 

contributions of Individual Plant 

research to Grazing Management 

Many of the assumptions on which grazing 

management is founded originated from 

defoliation experiments conducted with 

individual plants. Suppression of plant 

photosynthesis, root growth cessation, support 

of regrowth by carbohydrate reserves, and 

regulation of tillering by apical dominance 

represent several of the major assumptions 

(Briske and Richards 1995). The relevance 

of these individual plant-based assumptions 

to grazing management has recently been 

questioned in an assessment of plant and 

animal production responses to grazing systems 

(Briske et al. 2008). In several instances, these 

plant-based assumptions have shown little 

correspondence with the outcomes observed 

in grazing systems. Since the development of 

these plant-based assumptions in the mid-20th 

century, some have been substantiated, but 

others have been refuted from the vantage point 

of greater scientific understanding derived from 

more sophisticated experimental techniques. 

Several plant-based assumptions that have been 

validated and invalidated are summarized below. 

Unfortunately, these assumptions often prevail 

long after they have been refuted by substantial 

experimental evidence. 

Valid Plant-Based Interpretations. Numerous 

plant-based interpretations were developed 

early in the profession to cope with widespread 

overgrazing and rangeland degradation 

that prevailed in the late 19th and early 

20th centuries. These were often based on 

observation and general inference because 

knowledge of plant physiology was very limited 

during this period and did not substantially 

improve until the mid-20th century. Several of 

the more important plant-based interpretations 

that have been supported by current science are 

summarized below. 

Leaf Removal and Subsequent Growth. 

Photosynthetic leaf area provides the energy 

source for plant growth and reductions in leaf 

area suppress both plant photosynthesis and 

growth (Sampson 1923). This interpretation 

has been well supported with additional 

insights addressing the various contributions 

of leaf canopy position and leaf age (Caldwell 

et al. 1981; Gold and Caldwell 1989). The 

validity and consequences of this wellestablished process are reflected in the adverse 

effects of severe and multiple defoliations on 

plant growth within a growing season. 

An important caveat associated with 

plant defoliation experiments, even when 

conducted with field-grown plants, is that 

the defoliation intensities imposed are often 

very severe compared to actual defoliation 

CHAPTER 1: An Assessment of Grazing Practices 47 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	 	

	 	

	 	

	 	 	

	 	

	 	

	 	

	 	

...grazing 

management 

recommendations 

should not be 

developed 

exclusively 

from processes 

derived at the 

individual plant 

level without 

at least partial 

verification of 

the anticipated 

response within 

communities or 

ecosystems.”

patterns documented in the field. Eight of 12 

defoliation studies evaluated defoliated plants 

at ≤6 cm, and three of these eight defoliation 

intensities were imposed on large tallgrass 

species. This suggests that while this research 

is valuable for understanding mechanisms of 

plant response to defoliation, caution should 

be used in translating these responses to actual 

grazing management applications. 

Root Growth and Function. Root growth 

and function are increasingly suppressed 

with increasing intensity and frequency of 

defoliation because they are entirely dependent 

on energy derived from photosynthesis (Crider 

1955). This interpretation has also been well 

supported by subsequent research investigating 

specific physiological mechanisms, including 

root respiration and nutrient absorption 

kinetics (Ryle and Powell 1975; Macduff et 

al. 1989). However, even though suppression 

of root growth following severe defoliation 

of individual plants is well established, the 

evidence that intensive defoliation suppresses 

root biomass within plant communities 

remains equivocal (Milchunas and Lauenroth 

1993; McNaughton et al. 1998; Johnson 

and Matchett 2001). A specific mechanism 

has not been provided for this inconsistency, 

but it likely has to do with compensating 

root growth by less intensively grazed 

plants within the community or a shift 

in species composition to species that 

allocate a greater proportion of biomass 

belowground. Contrasting grazing responses 

between individual plants and communities 

demonstrates that caution should be used 

when extrapolating individual plant responses 

to communities and ecosystems. 

Defoliation-Induced Competitive 

Interactions. The ability of disproportionate 

defoliation intensity among adjacent plants 

to modify intra- and interspecific competitive 

interactions to favor less severely grazed plants 

was initially proposed by Mueggler (1972). 

This interpretation has been substantiated 

with more recent and sophisticated research 

using isotopes of phosphorous (Caldwell et al. 

1985, 1987) and nitrogen (Hendon and Briske 

2002) demonstrating that both the frequency 

and intensity of defoliation can modify 

belowground competition. This series of 

physiological effects on competitive interactions 

is partially reflected in the widely observed 

patterns of increaser and decreaser plant species 

and grazing-induced changes in the species 

composition of plant communities. 

Invalid Plant-Based Interpretations. Several 

well-established interpretations derived from 

individual plant response to defoliation have 

been invalidated with the advent of more 

sophisticated experimental procedures. This 

brief summary of refuted interpretations is 

intended not to criticize this early work, but 

merely to indicate that the knowledge base 

supporting grazing management has and 

will continue to advance as more research 

information is obtained. 

Apical Dominance and Tillering. Apical 

dominance was promoted as the primary 

mechanism controlling tiller initiation 

following defoliation of perennial grasses. It 

was based on the direct hypothesis of auxin 

action indicating that removal of the apical 

meristem terminated supply of the growth 

inhibitor auxin to the axillary buds near the 

base of the tiller and thereby allowed their 

outgrowth into new tillers (Leopold 1949). 

Physiologists considered this concept invalid 

in the 1950s, Jameson (1963) concluded 

that this interpretation of apical dominance 

was not supported by evidence for rangeland 

grasses, and this conclusion was corroborated 

by a larger data synthesis of perennial grasses 

(Murphy and Briske 1992). The traditional 

concept of apical dominance as applied in 

grazing management was a partial and overly 

restrictive interpretation of tiller initiation in 

perennial grasses. A complete understanding 

of the mechanisms contributing to tiller 

initiation is yet to be developed, but it is 

likely a multivariable processes regulated 

by several interacting physiological and 

environmental variables (Tomlinson and 

O’Connor 2004). 

Carbohydrate Reserves as Indicators 

of Regrowth. Carbohydrate reserves were 

proposed as an index of potential plant 

regrowth, and this concept was frequently 

applied in grazing management during the 

latter half of the 20th century and is still 

applied in limited cases. Since carbohydrate 

reserves decrease following plant defoliation, 

it was widely assumed that they must be 

48 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	

	 	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

a major source of carbon supporting leaf 

regrowth (Briske and Richards 1995). A more 

thorough evaluation of plant carbon balance 

indicated that root carbohydrates were used 

primarily within root systems rather than being 

allocated aboveground to support regrowth 

and that reserve pools of perennial grasses 

contained very small amounts of carbon that 

contributed to regrowth for only 1–3 d before 

leaf photosynthesis once again became the 

primary carbon source (Richards and Caldwell 

1985). Moreover, it appears that a consistent, 

positive relationship between the size of the 

carbon reserve pools and grass regrowth had 

never been established in support of this 

widely used interpretation (Busso et al. 1990). 

In retrospect, the concept of carbohydrate 

reserves was founded on an oversimplified 

interpretation of carbohydrate patterns in 

grasses, and it never had great relevance to 

grazing management. Residual leaf area and 

the availability of meristems, in the presence of 

favorable environmental conditions, are now 

recognized to provide more reliable indicators 

of plant regrowth following defoliation (Briske 

and Richards 1995). Ironically, emphasis on 

the maintenance of carbohydrate reserves in 

perennial grasses inadvertently applied these 

valid indicators of plant growth and thereby 

indirectly contributed to efficient grazing 

management. 

The hierarchical structure of ecological 

systems describes the nested levels of 

ecological organization that coincide with 

increasing complexity and interaction among 

components within systems. This hierarchical 

structure determines why it is possible for 

even well-established processes at the level of 

individual plants to not directly translate to 

communities and ecosystems. For example, 

recall that the well-established reduction in 

root growth following intensive defoliation 

of individual plants is not consistently 

expressed as a reduction of root biomass 

within grazed communities (Milchunas 

and Lauenroth 1993; McNaughton et al. 

1998). This inconsistent response suggests 

that processes and interactions within 

populations or communities are overriding 

or mitigating the negative root response of at 

least some of the plant species. Reductionist 

investigations of individual plants produce 

valuable mechanistic insights, but they may 

be too narrow in scope to identify important 

interactions and trade-offs at higher scales to 

make them relevant for direct management 

application (Briske 1991). Plant-based 

research over the past century indicates that 

grazing management recommendations should 

not be developed exclusively from processes 

derived at the individual plant level without 

at least partial verification of the anticipated 

response within communities or ecosystems. 

This is a rather sobering conclusion after 

nearly a century of individual plant-oriented 

research, but it does provide evidence of 

maturation and progress within the rangeland 

profession. 

rEcoMMEndAtIons 

The following recommendations have 

emerged from our evaluation of the benefits 

of NRCS prescribed grazing practices with 

the relevant experimental literature. They 

are presented to enhance the effectiveness 

of the current conservation planning 

standard and to emphasize the CEAP goals 

addressing environmental quality of managed 

lands, including the assessment of multiple 

ecosystem services. 

Priorities and Approaches to 

conservation Planning 

Conservation planning would benefit from a 

substantial shift in priorities that deemphasize 

the independent development of facilitating 

practices (e.g., fencing, roads, and pipelines) 

and reemphasize integration of these practices 

with adaptive management decisions (e.g., 

stocking rate, drought management, and 

monitoring) to promote environmental quality 

of rangelands as recommended by CEAP. 

With the clear exception of improved livestock 

distribution, there is no indication that 

facilitating practices alone directly promote 

effective environmental conservation. The 

function of grazed ecosystems is similarly 

controlled by several dominant environmental 

variables, albeit over diverse social and 

environmental conditions, that are expressed 

in dynamic forage production patterns within 

and among years establishing that management 

decisions, especially during critical periods, can 

have profound effects on grazed ecosystems. 

The environmental variables and many of the 

social variables cannot be directly managed, but 

Conservation 

planning would 

benefit from a 

substantial shift 

in priorities that 

deemphasize 

the independent 

development 

of facilitating 

practices (e.g., 

fencing, roads, 

and pipelines) 

and reemphasize 

integration of 

these practices 

with adaptive 

management 

decisions...”

CHAPTER 1: An Assessment of Grazing Practices 49 
	 	 	 	

 

	 	

	 	

	 	

	 	

	 	

	 	

	 	

Renewed 

emphasis 

on drought 

contingency 

planning must 

integrate both 

economic and 

ecological 

considerations 

to effectively 

encourage 

managers to 

adopt and 

implement 

destocking 

options...”

recognition and planning for their occurrence 

with effective adaptive management plans at 

both the tactical and the strategic level can 

minimize their detrimental consequences 

to both production and conservation goals. 

Increased development and delivery of 

contingency planning protocols are required to 

effectively cope with these variable conditions 

common to most grazing enterprises. These 

tools should emphasize dynamic stocking 

rate determinations and provisions to support 

flexible management strategies, including 

effective destocking and restocking tactics and 

the potential to develop reserve forage supplies 

(e.g., Sharrow and Seefeldt 2006; Hanselka et 

al. 2009; Torell et al. 2010). 

We recommend that additional decision 

support tools and guidelines be developed to 

inform adaptive grazing management decisions, 

especially during critical events and seasons. 

Current information and technology will 

support development of novel, comprehensive 

approaches for implementing dynamic stocking 

rate determinations that can be effectively 

incorporated into management plans and 

monitored by landowners. An undertaking 

of this magnitude will require investment 

of considerable intellectual and financial 

capital, but the experimental evidence strongly 

confirms that site-appropriate stocking rates 

represent the very foundation of sustainable 

grazing management and associated 

conservation benefits. These tools could target 

specific landowners via conservation planning 

or be more generally accessible through AFGC, 

(American Forage and Grassland Council), 

GLCI (Grazing Lands Conservation Initiative), 

SRM (Society for Range Management), or 

SWCS (Soil and Water Conservation Society) 

publications and venues or made available on 

NRCS websites. Incentives could be variously 

structured to encourage use and adoption of 

these tools and approaches. Conservation plans 

may even require participation in a set number 

of instructional activities to attain and maintain 

program eligibility. 

forage Inventory Assessment and 

Monitoring 

Development and implementation of forage 

inventory and monitoring protocols in grazed 

ecosystems requires greater emphasis. This will 

require that the process of balancing forage 

production with animal demand be placed in 

the broadest possible context to include forage 

inventory, seasonal plant growth dynamics, 

and drought management over both short- and 

long-term periods (e.g., Sharrow and Seefeldt 

2006; Hanselka et al. 2009). Static seasonal or 

annual stocking rates provide a broad reference, 

but they are insufficient to addresses wide 

seasonal and interannual variation in forage 

production common to most rangelands. 

Consequently, emphasis on static stocking rates 

results in systems being over- or understocked 

the majority of the time (Hart and Ashby 

1998). Spatial variability of forage production, 

associated with variation in soils, landscape 

position, and local precipitation patterns, also 

minimizes the value of static, regional stocking 

rates. Use of the grazing pressure index, 

describing animal units per unit of forage mass 

over a period of time, has been recommended 

to standardize stocking rates and improve 

clarity of animal–forage relationships (Smart et 

al. 2010). 

Stocking rates based on residual forage, 

determined as a percentage of site-specific 

annual forage productivity, minimizes the 

probability of over- and undergrazing at both 

spatial and temporal scales (Bement 1969; 

Clary and Leininger 2000). Management based 

on residual forage ensures sufficient vegetative 

cover to protect soils during drought and 

dormant seasons, enhances the capacity for 

plant regrowth, and provides food and cover 

for wildlife during stress periods. Stocking 

rates established to promote environmental 

quality on rangelands may also promote 

heterogeneity in structure and diversity of flora 

and fauna because livestock are less likely to 

graze uniformly across local topographic–plant 

community gradients within pastures. 

Experimental information and available 

technology support development of a 

comprehensive approach for implementing 

dynamic stocking rate determinations that can 

be effectively incorporated into management 

plans with landowner participation. An 

undertaking of this magnitude will require 

investment of considerable intellectual 

and financial capital, but the experimental 

evidence directly confirm that site-appropriate 

stocking rates represent the very foundation 

of sustainable grazing management and 

50 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	

	 	

	 	 	 	

	 	

	 	 	

	 	

	 	

	 	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

associated conservation benefits. Management 

for appropriate stocking rates not only supports 

conservation goals, but it also forms the basis 

for effective drought management strategies 

and sustainable long-term economic returns 

(Manley et al. 1997; Hart and Ashby 1998; 

Torell et al. 2010). 

Alternative approaches are required to more 

directly and effectively incorporate dynamic, 

site-specific stocking rate assessments into 

overall management strategies and conservation 

planning. Landowner incentives could be 

provided to encourage adoption of forage 

inventory and monitoring as well as the grazing 

adjustments suggested by these protocols. 

These tools and guidelines are required to more 

closely estimate actual forage utilization or 

grazing intensity so that this information can 

be integrated into an adaptive management 

framework that emphasizes and supports 

flexible grazing management. Existing 

annual forage production curves emphasizing 

specific reference points that are critical to 

the attainment of various management and 

conservation goals (e.g., midpoint and end of 

growing season, critical wildlife requirements, 

and sensitivity of riparian zones) require greater 

attention and user friendly access. Readily 

accessible monthly and seasonal precipitation 

probabilities derived from long-term regional 

climatic records would also support forage 

inventory decisions (Andales et al. 2006). These 

tools may represent simple, direct measures of 

forage availability as well as more complicated 

procedures to forecast drought and forage 

production that could be implemented in 

various combinations at various temporal and 

spatial scales. Specific recommendations to 

support dynamic stocking rate determinations 

and promote adaptive management are 

summarized below. 

Estimation of Residual Biomass to 

Determine Grazing Intensity. Estimates of 

residual forage could be used as a means to 

determine site- and period-specific stocking 

rates and grazing intensities, especially during 

drought conditions. This is a well-established 

management procedure that has a strong 

ecological basis focused on soil protection, 

continued surface hydrological function, 

and maintenance of sufficient residual plant 

material to provide a source of regrowth when 

rainfall occurs (Bement 1969; Bartolome et al. 

1980; Blackburn 1984; Gifford 1985; Clary 

and Leininger 2000). Recommendations could 

be incorporated within conservation plans 

requesting that land managers periodically 

monitor residual biomass, at intervals and 

locations relevant to management objectives, 

following a prescribed set of procedures. These 

residual biomass records could be maintained 

as part of the ongoing conservation plan to 

support longer-term stocking rate adjustments 

and overall adaptive management (Bement 

1969; Clary and Leininger 2000). 

Forage Production and Drought 

Forecasting. Major technical advances have 

occurred in the forecasting of forage production 

and drought that could be used to support 

both tactical (within the growing season) and 

strategic (multiple growing seasons) grazing 

management decisions at regional levels. 

Forage production models such as GPFARM 

(Great Plains Framework for Agricultural 

Resource Management; Andales et al. 2006) 

could be linked with 6–14-d, 1-mo, and 3-mo 

precipitation and temperature forecasts through 

the NOAA Climate Prediction Center (http:// 

www.cpc.noaa.gov/index.php) to provide 

regional projections of forage availability. 

Drought projections are also provided by US 

Drought Monitor (http://www.drought.unl. 

edu/DM/monitor.html) and the Vegetation 

Drought Response Index (http://drought. 

unl.edu/vegdri/VegDRI_Main.htm). This 

index integrates satellite-based (MODIS) 

observations of vegetation conditions based 

on NDVI, climate data, and other biophysical 

information, such as land cover/land use type 

and soil characteristics. Maps of the Vegetation 

Drought Response Index have been produced 

every 2 wk beginning in 2009 throughout 

the conterminous United States that deliver 

continuous geographic coverage over large 

areas, provide regional to subcounty-scale 

information of drought effects on vegetation, 

and have inherently finer spatial detail (1-km2 

resolution) than other commonly available 

drought indicators, such as the US Drought 

Monitor. Incorporation of soil water forecasts 

(http://www.cpc.ncep.noaa.gov/products/ 

Soilmst_ Monitoring/US/Soilmst/Soilmst. 

shtml) could further promote the accuracy 

of these forage production projections. 

Forage projections could be developed for 

The vast majority 

of experimental 

results indicate 

that there is no 

clear advantage 

of any one 

grazing system 

over another 

in terms of 

ecological 

benefits.”

CHAPTER 1: An Assessment of Grazing Practices 51 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	

	 	

The importance of effective 

tactical and strategic decisions 

to successful grazing management is widely acknowledged, 

but only poorly documented. 

(Photo: Alexander Smart) 

specific periods of management interest and 

provide probabilities for forage responses to 

dry, average, and wet conditions to ascertain 

various levels of management risk. Forecast 

information could interface with existing forage 

production curves previously developed by the 

NRCS to generate various forage inventory 

projections to inform management planning. 

Drought Contingency Planning. It is essential 

that monitoring protocols be linked to drought 

contingency planning and management actions. 

It is widely recognized that the commonly 

employed strategy of “optimistic inaction” 

regarding stocking rate adjustments in response 

to developing drought is a major contributor to 

long-term rangeland degradation (Stafford Smith 

and Foran 1992; Thurow and Taylor 1999; 

Torell et al. 2010). However, it is irresponsible to 

delay or fail to implement drought contingency 

planning based on the unpredictability of 

drought given its frequent occurrence on most 

rangelands (Thurow and Taylor 1999). Renewed 

emphasis on drought contingency planning 

must integrate both economic and ecological 

considerations to effectively encourage managers 

to adopt and implement destocking options in 

relation to drought. 

Conservative stocking rates and the formation 

of reserve forage or grass banks are wellestablished strategies for contending with 

economic and environmental aversion to 

drought risk (Thurow and Taylor 1999). 

During normal or wet years, these grass 

banks could serve as restoration programs to 

support prescribed burning or to promote 

critical ecosystem services (i.e., biodiversity 

and carbon sequestration). Flexible stocking 

is also an effective means to cope with 

variable precipitation and forage production 

(Stafford Smith and Foran 1992; Torell et al. 

2010). Cow-calf herds should represent only 

a conservative component of total livestock 

holdings because of the high cost of adjusting 

cow numbers relative to the potential for 

short-term gain. Equal forage allocation to 

cow-calf and stockers has been recommended 

for ranching operations in the western United 

States (Torell et al. 2010). It is important 

to recognize that flexible stocking conveys 

additional costs and financial risks that 

will require specific decision-making tools 

to expand its adoption, and it may not be 

appropriate for risk-averse managers (Tanaka et 

al., this volume). 

the role of Grazing systems 

It is extremely difficult to experimentally mimic 

livestock movements and defoliation patterns 

associated with various applications of grazing 

strategies used by managers. However, grazing 

systems research has carefully evaluated the 

ecological responses of individual plants and 

communities, including wildlife populations, 

soils and surface soil hydrology, and their 

feedbacks on livestock performance, including 

forage intake and weight gain per animal and 

per unit area. These major ecological variables 

integrate numerous ecosystem processes 

sufficiently well to provide reliable guidance 

for the implementation and evaluation of the 

ecological consequences associated with grazing 

systems. The vast majority of experimental 

results indicate that there is no clear advantage 

of any one grazing system over another in terms 

of ecological benefits. Conclusions derived from 

these experimental data provide a sufficient 

basis to establish ecological guidelines for the 

evaluation and application of grazing systems 

in conservation planning and ecosystem 

assessment. These data directly corroborate 

the long-standing conclusions that weather 

52 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	 	

	 	 	 	

	 	

	 	

	 	

	 	 	

	 	

	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

variability and stocking rate account for the 

majority of variation associated with plant and 

animal production and species composition 

changes on rangelands (Heitschmidt and 

Taylor 1991; Holechek et al. 2001; Derner 

and Hart 2007). This interpretation further 

emphasizes the importance of effective adaptive 

management to the successful operation of 

grazed ecosystems, including the establishment 

of clear goals, monitoring of resource 

conditions, and the ability to make appropriate 

and timely management adjustments. Stated in 

another way, there is no indication that grazing 

systems possess unique properties that enable 

them to compensate for poor management 

(Briske et al. 2008). 

This interpretation also emphasizes that it is 

not sufficient to evaluate only whether grazing 

management is effective; we also need to 

determine why it is effective. This information 

is essential to guide development of effective 

conservation practices by determining whether 

emphasis should be focused on facilitating 

practices or on adaptive management skills. 

Although largely undocumented, the importance 

of effective adaptive management to successful 

grazing management is widely acknowledged, 

and it requires much greater emphasis than 

it has received (Stuth 1991; Brunson and 

Burritt 2009; Hanselka et al. 2009). Both 

research and monitoring are required on 

ranch-scale operations to more clearly evaluate 

the contribution of adaptive management to 

the success of conservation practices and to 

investigate the interaction between adaptive 

management and various grazing systems at the 

ranch level (e.g., Jacobo et al. 2006). 

deferment and rest 

Few evidence-based conclusions can be drawn 

regarding the appropriate season for grazing 

deferment and the benefits of long-term rest. 

This is partially illustrated by the inconsistent 

vegetation responses associated with the 

application of rest-rotation systems (Holechek 

et al. 2001). Minimal advantages may have 

resulted because one season of complete rest 

may not have been sufficient to compensate 

for more intensive use of grazed pastures in 

previous years. Vegetation responses to season 

of grazing and deferment are highly dependent 

on 1) the timing and amount of precipitation 

received during the growing season, 2) the 

intensity of defoliation, and 3) the opportunity 

for regrowth following defoliation. Research is 

required to quantify the benefits of long-term 

rest (> 1 yr) and alternating seasons of pasture 

use in successive years. Limited evidence suggests 

that exclusion of livestock is not necessary 

for recovery from moderate drought on wellconditioned rangeland (Heitschmidt et al. 2005; 

Gillen and Sims 2006), but it may be beneficial 

following severe drought that has induced 

substantial tiller and plant mortality (Dalgleish 

and Hartnett 2006; Yahdjian et al. 2006). Plants 

subject to light and moderate grazing often 

show less drought-induced mortality than plants 

that have been severely grazed prior to drought 

(Albertson et al. 1957). 

Grazing can potentially be used as a tool to 

manage wildlife populations, and this may be 

particularly true when season of grazing or 

deferment of grazing is used to meet specific 

wildlife goals. Seasonal livestock use may 

especially benefit wildlife where only part of 

the range is desirable wildlife habitat and social 

avoidance or seasonal migration are important 

considerations, facilitation through improved 

forage quality has been demonstrated, or 

specific nesting requirements are an issue. In 

these cases livestock grazing may be imposed 

or deferred, depending on cover and foraging 

requirements of specific wildlife species. For 

example, some waterfowl or some upland 

game species require dense nesting cover, 

whereas some birds, such as mountain plover 

or curlews, choose nesting sites with very little 

cover and will not nest in ungrazed or lightly 

grazed habitat. Successful use of seasonal and 

deferred grazing may also be possible when 

pastures with limited wildlife value are available 

to minimize livestock use in adjacent pastures 

that contain critical wildlife habitat. 

stronger linkages between science and 

Management 

NRCS Conservation Practice Standards 

should be routinely informed by both 

scientific and management knowledge 

external to the agency to ensure that the most 

current and vetted information available is 

incorporated into the conservation planning 

process. This represents a formidable challenge 

because science and management are not 

directly comparable endeavors (Provenza 

1991), and this may partially explain why 

The knowledge 

gaps identified 

in this synthesis 

need to be at 

least partial 

addressed to 

promote the 

development 

and adoption of 

more effective 

conservation 

practices 

in grazed 

ecosystems.”

CHAPTER 1: An Assessment of Grazing Practices 53 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	

	 	 	 	

	 	 	 	

The diverse ecosystem services 

originating from rangelands 

require greater recognition and 

valuation. (Photo: USDA: Gary 

Kramer) 

stronger science–management linkages have 

not been forged in the rangeland profession. 

Experimental research has focused on specific 

aspects of grazing management, including 

stocking rate, grazing system, and livestock 

distribution, in a static and independent 

manner, rather than on their dynamic 

interaction within adaptively managed 

ecosystems. The critical but poorly defined 

contribution of adaptive management to 

grazed ecosystems is a major impediment 

to the development of linkages between 

research and management because decision 

making is often excluded from experimental 

research even though it is central to grazing 

management (Briske et al. 2008; Brunson 

and Burritt 2009). Research requires 

systematic collection of information to 

document outcomes of various grazing 

strategies, while the outcomes of conservation 

practices standards are seldom monitored 

and documented. This often results in the 

difficult task of comparing quantitative 

research results with qualitative and often 

anecdotal management information. New 

organizational structures are needed to bridge 

the gap between research and management to 

support and incentivize a more comprehensive 

framework for conservation planning (Boyd 

and Svejcar 2009; Svejcar and Havstad 2009). 

The NRCS may wish to adopt a more formal 

research–management framework to address 

conservation programming that could be 

convened each time a conservation practice 

standard undergoes reevaluation. 

Substantial differences between rangeland 

science and management have presented barriers 

to their integration throughout the history 

of the rangeland profession. The extensive 

54 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	

	 	

	 	

	 	 	

	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

synthesis of experimental information provided 

in this document and the science–management 

partnership forged by this 3.5-yr undertaking 

represents an important initial step in attaining 

this goal. Greater integration and information 

exchange among researchers and managers 

would create a “win–win” situation for the 

profession by facilitating development of 

evidence-based conservation practices. This 

represents a necessary step if Conservation 

Practice Standards are to effectively adopt CEAP 

recommendations to provide regular assessments 

of the societal benefits of taxpayer investments 

in conservation practices. It would also enable 

the management community to play a more 

direct role in establishing the rangeland research 

agenda, as suggested in the following section. 

Effective monitoring of conservation practice 

outcomes will be crucial for enhancement of 

science–management linkages by providing a 

quantitative source of information exchange 

between these two groups. 

KnoWlEdGE GAPs 

The following knowledge gaps were identified 

in the process of summarizing and interpreting 

the experimental literature associated with 

prescribed grazing. It is anticipated that by 

highlighting these poorly understood issues, 

they may receive additional research attention 

and funding to promote greater understanding. 

It is critical that these knowledge gaps be 

at least partially addressed to promote the 

development and adoption of more effective 

conservation practices in grazed ecosystems. 

Ecosystem Processes and services in 

Grazed Ecosystems 

Traditionally, grazing research has focused on 

several ecological variables, including plant 

and animal production and, to a lesser extent, 

patterns of species composition change and 

wildlife responses and habitat. These variables 

provide a valuable, but admittedly narrow 

foundation on which to assess ecosystem 

services and environmental quality in grazed 

ecosystems. Research programs designed to 

increase our understanding of ecosystem 

processes and the provisioning of ecosystem 

services are desperately needed. Relevant 

topics include plant functional groups, soil 

health and sustainability, biodiversity, carbon 

sequestration, greenhouse gas emissions, 

drought and drought recovery, and spatial 

heterogeneity of ecosystem and landscape 

structure. 

Ecosystem restoration and 

conservation strategies 

Even though grazing management was 

initiated to halt and reverse the adverse effects 

of overgrazing on rangeland ecosystems, 

restoration of grazed ecosystems has received 

limited research attention in the past several 

decades. Research has been focused primarily 

on optimization of livestock production 

during the past 30 yr with use of intensified 

grazing systems. Consequently, experimental 

information regarding the season of utilization 

or deferment that is most appropriate to 

restore degraded ecosystems or to promote 

various conservation strategies is limited. 

Research addressing individual bunchgrass 

responses to defoliation in the field indicates 

that mid-growing season is the most sensitive 

period for defoliation. However, we are 

unaware of community-level field studies 

that corroborate this conclusion. Similarly, 

individual plant research has imposed very 

severe defoliation intensities compared to 

observed utilization rates in grazed ecosystems 

so that the direct application of these results 

to management is limited. Plant, community, 

and ecosystem responses to realistic grazing 

patterns would benefit from further 

documentation. 

contributions of Adaptive Management 

Management goals, abilities, and 

opportunities as well as personal goals 

and values (e.g., human dimensions) are 

inextricably integrated within grazing 

management, and they are likely to interact 

with the adoption and operation of grazing 

systems to an equal or greater extent than the 

underlying ecological processes (Briske et al. 

2008). Therefore, research and monitoring 

approaches need to explicitly document 

the contribution of adaptive management 

within ecosystems to promote a more 

comprehensive understanding of successful 

grazing management (Brunson and Burritt 

2009; Budd and Thorpe 2009). The potential 

synergistic effects of grazing systems and 

adaptive management inputs have not been 

examined experimentally at the level of the 

ranch enterprise (Briske et al. 2008; Brunson 

Research and 

monitoring 

approaches 

need to explicitly 

document the 

contribution 

of adaptive 

management 

within 

ecosystems to 

promote a more 

comprehensive 

understanding 

of successful 

grazing 

management.”

CHAPTER 1: An Assessment of Grazing Practices 55 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	

	 	

Current NRCS grazing 

practices are appropriate in 

many respects, but multiple 

opportunities exist to improve 

their effectiveness. (Photo: 

Sonja Smith) 

and Burritt 2009). Successful research in 

this area will require direct involvement of 

social and political scientists addressing these 

critical human dimensions issues and their 

interactions with ecological systems. A novel 

experimental approach used by Jacobo et al. 

(2006) compared adjacent ranches that had 

employed unique grazing systems to achieve 

the optimal production outcome. The strength 

of this approach is that it enables researchers 

to evaluate outcomes reflecting the entire 

ranch enterprise, including the capacity 

to adaptively manage for the best possible 

outcomes, within the context of the respective 

grazing system. This approach simultaneously 

evaluates ecological and managerial responses, 

but it has yet to be determined whether it will 

be possible to distinguish between these two 

responses. Similarly, incentives and barriers 

of various social institutions influencing 

the adoption of conservation practices have 

received minimal research emphasis given 

their importance to the management of 

complex adaptive systems (Stafford Smith et 

al. 2007). 

Evaluation of large-scale Ecosystem 

responses 

Grazing research has not adequately assessed 

the effects of grazing at large scales (Bailey et 

al. 1996; Archibald et al. 2005), which often 

demonstrate the occurrence of patch- and areaspecific grazing. Smaller experimental pastures 

usually result in more uniform distribution of 

grazing intensity, which may not appropriately 

describe how domestic grazing animals utilize 

large landscapes or, in the case of native 

ungulates, how they migrate regionally. The 

direct application of research results obtained 

in small-scale experiments (< 200 ha) to 

large ranch enterprises may not be entirely 

appropriate because the ecological processes of 

interest often do not scale in a linear fashion 

(Fuhlendorf and Smeins 1999; Peters et al. 

2006). Investigations of the potential benefits 

of grazing systems at large scales require further 

evaluation, and the evaluation metrics should 

involve a variety of ecosystem services, such as 

firm-level production, biodiversity concerns, 

watershed function, and wildlife habitat. 

Integration of complex Ecosystem 

components 

The complexity of grazed ecosystems resides in 

the broad array of interacting variables associated 

with both ecological and human systems. A wide 

range of ecological variation is associated with 

rainfall regime (i.e., amount, seasonality, and 

intra- and interannual variability), vegetation 

structure, composition, and productivity 

and soils, prior land use, and livestock 

characteristics (i.e., breeds, prior conditioning, 

and previous experience). This tremendous 

ecological variability is paralleled by large, but 

unappreciated variability associated with the 

commitment, ability, goals, and opportunities of 

managers and associated stakeholders dependent 

on the services of these ecosystems (Briske et al. 

2008; Brunson and Burritt 2009). The success 

and benefits that accrue from conservation 

practices within these complex systems is 

dependent on three unique activities. First, the 

conservation practices must be based on sound 

managerial and ecological principles; second, 

practices must be effectively incorporated into 

the overall conservation plan; and, third, they 

must be appropriately applied, maintained, and 

monitored by ecosystem managers. The third 

component addressing manager or landowner 

commitment and capability is most widely 

overlooked and can be addressed only from a 

human dimensions perspective. 

A robust ecosystem management framework 

capable of accommodating both ecological 

processes and human activities, as well as their 

interactions, is required to conceptualize, 

interpret and manage complex adaptive 

systems characteristic of rangelands. This will 

require the development of an information 

base that consists of local knowledge, 

management and policy experience, and 

science-based information mediated through 

an adaptive institutional framework 

56 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

(Herrick and Sarukhan 2007; Reynolds et 

al. 2007). This framework must be coupled 

with current and emerging technologies to 

provide estimates of remotely sensed data to 

address multiple feedbacks between the social 

and ecological components at several scales. 

Ecological site descriptions may provide the 

platform on which to integrate these sources 

of information, but the rangeland profession is 

lacking an institutional structure to house and 

coordinate relevant ecosystem components 

and processes at landscape and regional scales 

(Bestelmeyer et al., this volume). Approaches 

that involve integration of ecological scales 

and human dimensions, coupled with effective 

monitoring protocols capable of evaluating 

both ecological and social metrics, will likely 

drive the next major advance in effective 

rangeland stewardship. 

conclusIons 

An extensive evaluation of the published 

experimental evidence relevant to grazing 

management broadly supports the overall 

USDA-NRCS approach to prescribed grazing 

and validates the ecological foundations 

of many of the purposes addressed in this 

conservation practice standard. The equivocal 

nature of a portion of these findings is a 

consequence of experimental research and 

conservation planning pursuing different 

objectives, with unique approaches that are 

often conducted at different scales and an 

unfortunate legacy of minimal interaction 

between science and management within 

the rangeland profession. Nevertheless, 

inferences drawn from these experimental 

data indicate that the NRCS conservation 

purposes addressing prescribed grazing 

can potentially be realized, if implemented 

appropriately, as indicated by the ability 

for grazing management practices to affect 

all seven stated conservation purposes. 

The challenge of grazing management is 

establishing the appropriate relationships 

between various management practices 

and the intended purposes or outcomes, in 

diverse environmental and social conditions, 

especially when multiple and often competing 

purposes are involved. 

The experimental data unequivocally 

document that stocking rate, coupled with 

effective livestock distribution, is the single 

most important management variable 

influencing production and conservation 

goals in grazed ecosystems. Therefore, 

guidelines, tools and incentives that promote 

appropriate management decisions have 

the potential to enhance the effectiveness 

of conservation outcomes and increase 

the cost–benefit ratio of conservation 

investments. Guidelines promoting the 

goal of balancing forage production with 

animal demand should be placed in the 

broadest possible context to include forage 

inventory, seasonal plant growth dynamics, 

and drought management over both the 

short and the long term. Existing annual 

forage production curves emphasizing 

specific reference points that are critical to 

the attainment of various management and 

conservation goals, supported by monthly 

and seasonal precipitation probabilities, 

require greater emphasis and user-friendly 

access to support forage inventory decisions. 

The adoption of this approach will require a 

major shift in NRCS programmatic emphasis 

from those promoting facilitating practices 

in the form of infrastructure development to 

those promoting timely and effective adaptive 

management actions. 

Experimental evidence indicates that 

grazing systems, in the absence of adaptive 

management, explain little additional 

variability beyond that of stocking rate and 

weather variation for the variables of plant and 

animal production, species composition of 

plant communities, soil surface hydrological 

function, and wildlife populations. In 

addition, the major assumptions on which 

short-duration rotational grazing is partially 

based, including greater control over grazing 

patterns, minimization of multiple defoliations 

within individual grazing periods, and 

greater forage quality, have received only 

equivocal experimental support. Current 

evidence suggests that implementation of 

grazing systems, without incorporation 

of the additional elements of prescribed 

grazing, is insufficient to address the array of 

complex and dynamic conditions inherent 

to grazed ecosystems. However, the potential 

contributions of grazing systems to broader 

conservation goals and ecosystem services, at 

landscape or regional scales, and their potential 

CEAP provides an 

excellent platform 

to promote 

greater science– 

management 

integration by 

bringing together 

researchers and 

NRCS personnel 

for an evidencebased assessment 

of Conservation 

Practice 

Standards”

CHAPTER 1: An Assessment of Grazing Practices 57 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

interactions with adaptive management have 

yet to be evaluated. 

Several important knowledge gaps have been 

identified in the experimental literature 

associated with prescribed grazing. These 

include 1) grazing effects on ecosystem 

services, 2) ecosystem restoration and 

conservation strategies, 3) contributions of 

adaptive management actions, 4) evaluation 

of larger-scale ecosystem responses, and 

5) integration of information within 

complex ecosystems. The greatest deficiency 

encountered in this evaluation of supporting 

experimental data was the paucity of 

information documenting the impact of 

adaptive management on grazing management 

effectiveness and conservation outcomes. It 

is critical that these knowledge gaps be at 

least partially resolved in the near future to 

promote further advances in the ecology and 

management of grazed ecosystems. 

The overarching conclusion of this assessment 

is that even though the current conservation 

practices for prescribed grazing are 

appropriate in many respects, reorganization 

to implement three major modifications 

would greatly increase their effectiveness. 

First, greater emphasis should be placed 

on programs to support management skills 

and management effectiveness beyond 

that of financial incentives supporting the 

independent development of infrastructure. 

There is no clear indication that installation 

of facilitating practices in the form of water 

developments and fencing directly contribute 

to conservation benefits in the absence 

effective management. Second, a system of 

regular and frequent monitoring needs to be 

incorporated into conservation planning to 

directly assess both the short-term and the 

long-term benefits derived from conservation 

practices. Monitoring information will 

directly support adaptive management to 

optimize conservation outcomes per unit 

investments and document the ecological 

benefits of conservation practices on the 

nation’s rangelands as recommended by 

CEAP. Third, incorporate the intent and 

recommendations of CEAP by focusing on 

environmental quality, ecosystem services, 

and societal benefits associated with 

prescribed grazing in addition to sustainable 

production outcomes (e.g., Dunn et al. 

2010). 

Revisions to this conservation practice 

standard should be informed by both 

scientific and management knowledge 

external to the agency to ensure that the most 

current and vetted information available is 

incorporated into conservation planning 

and assessment procedures. CEAP provides 

an excellent platform to promote greater 

science–management integration by bringing 

together researchers and NRCS personnel, as 

well as other stakeholders, for an evidencebased assessment of Conservation Practice 

Standards and approaches to conservation 

planning. We recommend that this integrated 

science–management team approach should 

be formalized in the agency and used to 

revise Conservation Practice Standards and 

set priorities and goals for conservation 

planning. 

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APPEndIx I A. 

Published experiments used to evaluate plant and animal production to continuous grazing (CG) and rotational grazing (RG) 

at (a) equal stocking rates and (b) higher stocking rates for rotational grazing. See Figure 3 for a graphical presentation of the	 

comparative results (modified from Briske et al. 2008). 

study location Ecosystem 

length 

(yr) Grazing system 

stocking rate equal for rotational and continuous grazing 

Mccollum et al. (1999) Oklahoma Tallgrass prairie 5 Short-duration grazing 

(SDG) 

Gillen et al. (1998) Oklahoma Tallgrass prairie 5 SDG 

cassels et al. (1995) Oklahoma Tallgrass prairie 5 SDG 

owensby et al. (1973) Kansas Tallgrass prairie 17 Deferred rotation 

Wood and Blackburn (1984) Southern mixed-grass prairie 5 

High-intensity/lowfrequency and deferred 

rotation 

Kothmann et al. (1971) Texas Southern mixed-grass prairie 8 Merrill 

Merrill (1954) Texas Southern mixed-grass prairie 4 Merrill 

fisher and Marion (1951) Texas Southern mixed-grass prairie 8 Rotational 

McIlvain and savage (1951) Oklahoma Southern mixed-grass prairie 9 Rotational 

derner and Hart (2007a) Wyoming Northern mixed-grass prairie 25 SDG 

Manley et al. (1997) Wyoming Northern mixed-grass prairie 13 SDG and deferred 

rotation 

Biondini and Manske (1996) North Dakota Northern mixed-grass prairie 6 SDG 

Hart et al. (1993) Wyoming Northern mixed-grass prairie 5 SDG 

Hepworth et al. (1991) Wyoming Northern mixed-grass prairie 4 SDG and deferred rotation 

Hart et al. (1988) Wyoming Northern mixed-grass prairie 6 SDG and deferred rotation 

rogler (1951) North Dakota Northern mixed-grass prairie 25 Deferred rotation 

derner and Hart (2007b) Colorado Shortgrass prairie 9 SDG 

smoliak (1960) Alberta, Canada Shortgrass prairie 9 Deferred rotation 

Hubbard (1951) Alberta, Canada Shortgrass prairie 6 Deferred rotation 

laycock and conrad (1981) Utah Sagebrush–grassland 7 Rest-rotation 

Hyder and sawyer (1951) Oregon Sagebrush–grassland 11 Rotational 

Holechek et al. (1987) Oregon Mountain rangeland 5 Rest-rotation and deferred 

rotation 

Murray and Klemmedson (1968) Idaho Cheatgrass (Bromus tectorum) 

dominated 3 Seasonal rotation 

Martin and severson (1988) Arizona Grass–shrub complex 13 Santa Rita, 1-herd, 

3-pasture, 3-yr rotation 

Martin and Ward (1976) Arizona Desert grassland 7 Alternate-year seasonal rest 

Winder and Beck (1990) New Mexico Semidesert 17 3-pasture rotation 

Gutman et al. (1990) Israel Mediterranean grassland 2 Rotational 

Gutman and seligman (1979) Israel Mediterranean Foothill Range 10 Rotational 

ratliff (1986) California Annual grassland 8 Rotational 

70 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

no. of paddocks in 

rotation 

size of 

paddocks (ha) Plant production/standing crop 

livestock production 

per head 

livestock production 

per land area 

8 1.8–3.3 

8 1.8–3.3 

8 1.8–3.3 

3 24 

4 120 

4 ? 

4 24 

5 4 

3 6.7–10 

8 1–2 

4–8 1–3 

6 32 

8 24 

4–8 1–3 

4–8 1–3 

3 9.4 

7 65 

2 61 

3 27–40 

3 447–777 

3 850 

2 57–67 

4 ? 

3 308–1 979 

24 0.004 

3 ? 

6 25.5–33.0 

3 25.5–33.8 

3 30 

ND
 

ND at peak standing crop
 

RG > CG at end of grazing
 

season
 

RG > CG
 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

CG > RG 

CG > RG 

RG > CG
 

ND
 

ND
 

ND
 

ND 

ND
 

ND
 

ND
 

RG > CG
 

ND
 

CG > RG
 

ND
 

ND
 

CG > RG
 

ND
 

ND
 

ND
 

CG > RG
 

ND
 

CG > RG
 

CG > RG 

CG > RG 

RG > CG
 

ND
 

ND
 

ND
 

ND 

ND
 

ND
 

ND
 

RG > CG
 

ND
 

CG > RG
 

ND
 

ND
 

CG > RG
 

ND
 

ND
 

ND
 

CG > RG
 

ND
 

CG > RG
 

CHAPTER 1: An Assessment of Grazing Practices 71 
	 	 	 	

	 	 

	 	 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

APPEndIx I A. continued

study location Ecosystem 

length 

(yr) Grazing system 

stocking rate equal for rotational and continuous grazing 

Heady (1961) California Annual grassland 5 Deferred rotation 

Barnes and denny (1991) Zimbabwe Veld 6 SDG 

fourie and Engels (1986) South Africa Veld 4 SDG 

fourie et al. (1985) South Africa Veld 4 SDG 

Kreuter et al. (1984) South Africa Veld 3 SDG 

Walker and scott (1968) Tanzania 2 Rotational 

Bogdan and Kidner (1967) Kenya Woodland–grassland 5 Rotational, deferred 

rotational 

Higher stocking rate for rotational grazing 

Jacobo et al. (2000) Argentina Temperate grasslands 3 SDG 

Heitschmidt et al. (1987) Texas Southern mixed-grass prairie 4 SDG 

Heitschmidt et al. (1982a) Texas Southern mixed-grass prairie 2 SDG 

Heitschmidt et al. (1982b) Texas Southern mixed-grass prairie 19 Merrill 

reardon and Merrill (1976) Texas Southern mixed-grass prairie 20 Deferred rotation 

Hirschfeld et al. (1996) North Dakota Northern mixed-grass prairie 2 SDG 

volesky et al. (1990) South Dakota Northern mixed-grass prairie 2 SDG 

Kirby et al. (1986) North Dakota Northern mixed-grass prairie 2 SDG 

White et al. (1991) New Mexico Blue grama 6 SDG 

Pitts and Bryant (1987) Texas Shortgrass prairie 4 SDG 

Anderson (1988) New Mexico Tobosa (Hilaria mutica) dominated 2 SDG 

APPEndIx I B. literature cited in the construction of Appendix I A above. 

Anderson, D. M. 1988. Seasonal 

stocking of tobosa managed under 

continuous and rotation grazing. 

Journal of Range Management 41:78–83. 

Barnes, D. L., and R. P. Denny. 1991. 

A comparison of continuous and 

rotational grazing on veld at two 

stocking rates. Journal of Society of 

South African Grassland 8:168–173. 

Biondini, M. E., and L. L. Manske. 

1996. Grazing frequency and 

ecosystem processes in a northern 

mixed prairie, USA. Ecological 

Applications 5:239–256. 

Bogdan, A. V., and E. M. Kidner. 

1967. Grazing natural grassland 

in western Kenya. East African 

Agricultural and Forestry Journal 

33:31–34. 

Cassels, D. M., R. L. Gillen, F. T. 

McCollum, K. W. Tate, and M. 

E. Hodges. 1995. Effects of grazing 

management on standing crop 

dynamics in tallgrass prairie. Journal of 

Range Management 48:81–84. 

Derner, J. D. and Hart, R. H. 2007a. 

Grazing-induced modifications to 

peak standing crop in northern mixedgrass prairie. Rangeland Ecology and 

Management 60:270–276. 

Derner, J. D., and R. H. Hart. 2007b. 

Livestock and vegetation responses to 

rotational grazing in shortgrass steppe. 

Western North American Naturalist 

67:359-367. 

Fisher, C. E., and P. T. Marion. 1951. 

Continuous and rotation grazing on 

buffalo and tobosa grassland. Journal 

of Range Management 4:48–51. 

Fourie, J. H., and E. A. N. Engels. 

1986. Herbage intake by cattle on 

Tarchonanthus veld in the northern 

Cape as affected by stocking rate 

and grazing system. Journal of the 

Grassland Society of Southern Africa 

3:85–89. 

Fourie, J. H., D. P. J. Opperman, and 

B. R. Roberts. 1985. Influence of 

stocking rate and grazing systems 

on available grazing in the northern 

Cape. Journal of the Grassland Society 

of Southern Africa 2:24–26. 

Gillen, R. L., F. T. McCollum III, 

K. W. Tate, and M. E. Hodges. 1998. 

Tallgrass prairie response to grazing 

system and stocking rate. Journal of 

Range Management 51: 139–146. 

72 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	

	 	

	 	

	 	 	 	

	 	

	 	 	 	

D. D. Briske, J. D. Derner, D. G. Milchunas, and K. W. Tate 

no. of paddocks in 

rotation 

size of 

paddocks (ha) Plant production/standing crop 

livestock production 

per head 

livestock production 

per land area 

3 5.4 ND CG > RG CG > RG 

4–8 ? ND ND 

6 5 CG > RG CG > RG 

6 5 ND 

6 ? CG > RG CG > RG 

3 5 CG > RG CG > RG 

3–4 0.6–0.8 CG = RG CG = RG 

10–12 45 ND 

16 33 CG > RG 

10 4 ND RG > CG 

4 ? ND RG > CG 

4 ? ND 

8 16.25 ND 

16 2.2 CG > RG RG > CG 

8 16 RG > CG 

9 45–210 ND 

16 3 ND ND ND 

10 3.5 ND CG > RG CG > RG 

Gutman, M., and N. G. Seligman. 

1979. Grazing management of 

Mediterranean foothill range in the 

upper Jordan River Valley. Journal of 

Range Management 32:86–92. 

Gutman, M., N. G. Seligman, 

and I. Noy-Meir. 1990. Herbage 

production of Mediterranean 

grassland under seasonal and 

yearlong grazing systems. Journal of 

Range Management 43:64–68. 

Hart, R. H., M. J. Samuel, P. S. 

Test, and M. A. Smith. 1988. 

Cattle, vegetation, and economic 

responses to grazing systems and 

grazing pressure. Journal of Range 

Management 41:282–286. 

Hart, R. H., J. Bissio, M. J. Samuel, 

and J. W. Waggoner, Jr. 1993. 

Grazing systems, pasture size, and 

cattle grazing behavior, distribution 

and gains. Journal of Range 

Management 46: 81–87. 

Heady, H. F. 1961. Continuous vs. 

specialized grazing systems: a review 

and application to California annual 

type. Journal of Range Management 

14: 182–193. 

Heitschmidt, R. K., S. L. 

Dowhower, and J. W. Walker. 

1987. Some effects of a rotational 

grazing treatment on quantity 

and quality of available forage and 

amount of ground litter. Journal of 

Range Management 40:318–321. 

Heitschmidt, R. K., J. R. 

Frasure, D. L. Price, and L. 

R. Rittenhouse. 1982a. Short 

duration grazing at the Texas 

Experimental Ranch: weight gains 

of growing heifers. Journal of Range 

Management 35:375–379. 

Heitschmidt, R. K., M. M. 

Kothmann, and W. J. Rawlins. 

1982b. Cow-calf response to 

stocking rates, grazing systems, and 

winter supplementation at the Texas 

Experimental Ranch. Journal of 

Range Management 35:204–210. 

Hepworth, K. W., P. S. Test, R. H. 

Hart, J. W. Waggoner, Jr., and M. 

A. Smith. 1991. Grazing systems, 

stocking rates, and cattle behavior 

in southeastern Wyoming. Journal of 

Range Management 44:259–262. 

Hirschfeld, D. J., D. R. Kirby, 

J. S. Caton, S. S. Silcox, and 

CHAPTER 1: An Assessment of Grazing Practices 73 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

APPEndIx I B. literature cited in the construction of Appendix I A continued

K. C. Olson. 1996. Influence of 

grazing management on intake and 

composition of cattle diets. Journal of 

Range Management 49:257–263. 

Holechek, J. L., T. J. Berry, and 

M. Vavra. 1987. Grazing system 

influences on cattle performance on 

mountain range. Journal of Range 

Management 40:55–59. 

Hubbard, W. A. 1951. Rotational 

grazing studies in western Canada. 

Journal of Range Management 4:25– 

29. 

Hyder, D. N., and W. A. Sawyer. 

1951. Rotation-deferred grazing as 

compared to season-long grazing 

on sagebrush-bunchgrass ranges in 

Oregon. Journal of Range Management 

4:30–34. 

Jacobo, E. J., A. M. Rodriguez, J. 

L. Rossi, L. P. Salgado, and V. A. 

Deregibus. 2000. Rotational stocking 

and production of Italian ryegrass on 

Argentinean rangelands. Journal of 

Range Management 53:483–488. 

Kirby, D. R., M. F. Pessin, and G. 

K. Clambey. 1986. Disappearance 

of forage under short duration and 

season-long grazing. Journal of Range 

Management 39:496–500. 

Kothmann, M. M., G. W. Mathis, 

and W. J. Waldrip. 1971. Cow-calf 

response to stocking rates and grazing 

systems on native range. Journal of 

Range Management 24:100–105. 

Kreuter, U. P., G. M. Brockett, A. 

D. Lyle, N. M. Tainton, and D. 

I. Bransby. 1984. Evaluation of 

veld potential in east Griqualand 

using beef cattle under two grazing 

management systems. Journal of the 

Grassland Society of Southern Africa 

1:5–10. 

Laycock, W. A., and P. W. Conrad. 

1981. Responses of vegetation and 

cattle to various systems of grazing 

on seeded and native mountain 

rangelands in eastern Utah. Journal of 

Range Management 34:52–58. 

McCollum F. T., III, R. L. Gillen, 

B. R. Karges, and M. E. Hodges. 

1999. Stocker cattle response to 

grazing management in tallgrass 

prairie. Journal of Range Management 

52: 120–126. 

Manley, W. A., R. H. Hart, M. 

J. Samuel, M. A. Smith, J. W. 

Waggoner, Jr., and J. T. Manley. 

1997. Vegetation, cattle, and 

economic responses to grazing 

strategies and pressures. Journal of 

Range Management 50:638–646. 

Martin, S. C., and K. E. Severson. 

1988. Vegetation response to the 

Santa Rita grazing system. Journal of 

Range Management 41:291–295. 

Martin, S. C., and D. E. Ward. 

1976. Perennial grasses respond 

inconsistently to alternate year 

seasonal rest. Journal of Range 

Management 29:346. 

McIlvain, E. H., and D. A. Savage. 

1951. Eight-year comparisons of 

continuous and rotational grazing 

on the Southern Plains Experimental 

Range. Journal of Range Management 

4:42–47. 

Merrill, L. B. 1954. A variation of 

deferred rotation grazing for use under 

southwest range conditions. Journal of 

Range Management 7: 152–154. 

Murray, R. B., and J. O. 

Klemmedson. 1968. Cheatgrass range 

in southern Idaho: seasonal cattle 

gains and grazing capacities. Journal of 

Range Management 21:308–313. 

Owensby, C. E., E. F. Smith, and K. L. 

Anderson. 1973. Deferred-rotation 

grazing with steers in the Kansas Flint 

Hills. Journal of Range Management 

26:393–395. 

Pitts, J. S., and F. C. Bryant. 1987. 

Steer and vegetation response to short 

duration and continuous grazing. 

Journal of Range Management 40:386– 

389. 

Ratliff, R. D. 1986. Cattle responses 

to continuous and seasonal grazing of 

California annual grassland. Journal of 

Range Management 39:482–485. 

Reardon, P. O., and L. B. Merrill. 

1976. Vegetation response under 

various grazing management systems 

in the Edwards Plateau of Texas. 

Journal of Range Management 29:195– 

198. 

Rogler, G. A. 1951. A twenty-five 

year comparison of continuous and 

rotation grazing in the northern 

plains. Journal of Range Management 

4:35–41. 

Smoliak, S. 1960. Effects of deferredrotation and continuous grazing on 

yearling steer gains and shortgrass 

prairie vegetation of southeastern 

Alberta. Journal of Range Management 

3:239–243. 

Volesky, J. D., J. K. Lewis, and C. 

H. Butterfield. 1990. Highperformance short-duration and 

repeated-seasonal grazing systems: 

effect on diets and performance of 

calves and lambs. Journal of Range 

Management 43:310–315. 

Walker, B., and G. D. Scott. 1968. 

Grazing experiments at Ukiriguru, 

Tanzania: I. Comparisons of rotational 

and continuous grazing systems on 

natural pastures of hardpan soils. 

East African Agricultural and Forestry 

Journal 34:224–255. 

White, M. R., R. D. Pieper, G. B. 

Donart, and L. W. Trifaro. 1991. 

Vegetational response to shortduration and continuous grazing in 

southcentral New Mexico. Journal of 

Range Management 44:399–403. 

Winder, J. A., and R. F. Beck. 1990. 

Utilization of linear prediction 

procedures to evaluate animal 

response to grazing systems. Journal of 

Range Management 43:396–400. 

Wood, M. K., and W. H. Blackburn. 

1984. Vegetation and soil responses 

to cattle grazing systems in the 

Texas rolling plains. Journal of Range 

Management 37:303–308. 

74 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	

	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

CHAPTER 

2 

Assessment of Prescribed Fire 

as a Conservation Practice 

Samuel D. Fuhlendorf,1 Ryan F. Limb,2 David M. Engle,3 

and Richard F. Miller4 

Authors are 1Sarkeys Distinguished Professor,
 

Natural Resource Ecology and Management, Oklahoma State University;
 

2 Assistant Professor, Eastern Oregon Agricultural Research Center &
 

Department of Rangeland Ecology & Management, Oregon State University;
 

3 Regents Professor, Natural Resource Ecology and Management,
 

Oklahoma State University; 4 Professor, Rangeland Ecology and
 

Management, Oregon State University
 

Reference to any commercial product or service is made with the understanding 

that no discrimination is intended and no endorsement by USDA is implied 

75
 
	 	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	 	

	 	

	 	

	 	 	

	 	

	 	

	 	

	 	

	 	

A number of studies 

across US rangelands 

“

reported that shrubs 

and trees increase and 

herbaceous vegetation 

decreases with longterm fire removal, 

so maintaining or 

increasing relative 

abundance of 

herbaceous vegetation 

requires periodic fire.”

76 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	

2 Assessment of Prescribed Fire as 

a Conservation Practice 

Samuel D. Fuhlendorf, Ryan F. Limb, David M. Engle, and Richard F. Miller 

IntroduCtIon 

Fire has played a key role in the formation of 

most rangeland ecosystems in North America 

(Axelrod 1985) and the world (Bond et al. 

2003; Keeley and Rundel 2005). Alteration of 

fire regimes on US rangelands since European 

settlement has created cases of severely altered 

ecosystems that can eventually result in noanalog, novel, or emerging ecosystems (House 

et al. 2003; Hobbs et al. 2009). According to 

the Landscape Fire and Resource Management 

Planning Tools Project (LANDFIRE; 

an interagency vegetation, fire, and fuel 

characteristics mapping program sponsored by 

both the US Department of Interior and the 

US Department of Agriculture [USDA]–Forest 

Service), three-fourths of US lands dominated 

by native vegetation show moderate or high 

departure from reference conditions as a result 

of altered fire regimes (The Nature Conservancy 

2009). Because most rangelands are considered 

fire-dependent ecosystems, restoring historical 

fire regimes is fundamentally important 

when the management goal is to restore or 

maintain the potential (or historical) natural 

communities. For most ecological sites, the 

historical plant community was maintained 

by fire, and removing fire will cause the 

community to cross a threshold, often to 

woody plant dominance with reduced livestock 

production and loss of other ecosystem services. 

Rapid and extensive woodland expansion on 

rangelands clearly reflects the essential role of 

fire in the maintenance of historical rangeland 

ecosystems. These recent changes in land 

cover patterns emphasize that restoration of 

historical fire regimes are necessary to maintain 

these historical communities as outlined in 

the Natural Resources Conservation Service 

(NRCS) Ecological Site Descriptions. Yet, the 

implementation of prescribed burning as a 

conservation practice has been overshadowed 

by the implementation of other practices, 

especially prescribed grazing. In Oklahoma 

for example, from 2004 to 2008, NRCS 

implemented prescribed burning on 84 700 ha 

compared to 919 800 ha for prescribed 

grazing during 2004–2008. This 10-fold 

difference in the application of these two 

conservation practices clearly identifies the 

higher priority placed on grazing compared 

to that of burning. Considering that NRCS 

grazing practices operate over multiple years 

and that the practice of prescribed burning is 

a one-time application, the effective difference 

is actually considerably larger than 10-fold. 

Disproportionate implementation of these 

two practices is influenced by the complexity 

of social interactions among agencies, the 

general public, and public policy. Social and 

policy concerns are extremely different across 

various rangeland regions, ranging from 

complete acceptance of fire cultures (e.g., Flint 

Hills of Kansas and Oklahoma) to attempts 

to completely remove fire from the landscape 

(e.g., Great Basin). 

With a few exceptions, fire regimes have been 

altered through intentional fire suppression 

and by grazing that uniformly reduces fuel 

loads. Therefore, invasion of woody plants 

(both native and nonnative) has converted 

many shrublands and grasslands to forests 

or woodlands because of the absence of fire 

for abnormally long periods after European 

settlement. In contrast, other rangelands, 

notably those of the Great Basin, have been 

largely invaded by exotic herbaceous species 

that increase fine-scale fuel homogeneity, which 

greatly alters the fire regime by increasing 

fire frequency. State-and-transition models 

suggest conversions to woody plant dominance 

and exotic annuals can eventually become 

irreversible and result in alternative stable 

states. Although rangeland management 

A prescribed fire in a mixed 

grass prairie that has been 

invaded by woody plants. 

(Photo: John Weir) 

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 77 
	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	

Suppression of fire allows 

invasion of woody plants 

into grassland habitats. Prior 

to European settlement, First 

Nations actively burned the 

landscape for many reasons 

and maintenance of those 

ecosystems requires periodic 

fires. (Photo: Sam Fuhlendorf) 

professionals generally support using fire 

in rangeland ecosystems, a long history of 

exclusion, uncertainty about the effects of 

fire, increased wildland–urban interface, 

socioeconomics, and natural resource policy are 

formidable barriers to reintroducing fire except 

in those ecosystems in which the fire return 

interval has been shortened. However, as long 

as maintaining ecosystem structure within a 

historical context is emphasized, fire regimes 

must be restored across most rangelands. 

The USDA-NRCS Practice Standard for 

Prescribed Burning (CODE 338) describes the 

following purposes: 

1.	� to control undesirable vegetation; 

2.	� prepare sites for harvesting, planting, and 

seeding; 

3.	� to control plant diseases; 

4.	� to reduce wildfire hazards; 

5.	� to improve wildlife habitat; 

6.	� to improve plant production quantity and/ 

or quality; 

7.	� to remove slash and debris; 

8.	� to enhance seed production; 

9.	� to facilitate distribution of grazing and 

browsing animals; and 

10. to restore and maintain ecological sites. 

The Conservation Effects Assessment Program 

(CEAP) was initiated to determine the extent 

to which experimental data present in peerreviewed research literature support these 

purposes. The general and value-laden nature of 

these purposes makes them extremely difficult 

to assess against experimental data; therefore, 

we analyzed the research literature to establish 

the ecological effects of prescribed fire from a 

broader perspective. Specifically, we evaluated 

the research literature available on plants, 

soil, water, wildlife, arthropods, livestock, 

78 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

fire management, fire behavior, smoke 

management, socioeconomics, air quality, fire 

history, and human health. These topics were 

selected based on input from rangeland CEAP 

teams focused on other conservation practices 

and initial evaluation of the literature in terms 

of topics that were covered sufficiently to draw 

meaningful conclusions. We also addressed 

issues related to spatial scale, temporal scale, 

and other general descriptions of the body of 

research and we then related our findings to 

the specific NRCS purposes for the practice of 

prescribed burning. 

deFInIng our LIterAture dAtAbAse 

Evaluation of the peer-reviewed literature 

on prescribed fire first required determining 

methods to query the entire body of scientific 

literature on the topic. We wanted to include 

all relevant papers, but we limited the scope 

of the search to minimize less-relevant papers. 

We searched for papers that focused on fire 

(preferably prescribed), but largely excluded 

fire research from forested systems, which 

dominates the fire research literature. Many 

papers that report fire research on rangelands 

do not include the term prescribed, and many 

relevant papers do not use the term rangeland. 

We used multiple approaches (Table 1) to 

identify the most acceptable body of literature 

to evaluate. The data set built from the search 

with the term prescribed fire omitted numerous 

important papers from the pool, and many 

of the papers included some discussion of fire 

but with minimal or no data related to fire. 

Therefore, our final search used the term fire, 

which also located articles with prescribed fire 

in the title, to broaden the search. Although 

this approach excluded papers that reported 

research from regionally important ecosystem 

types (e.g., shinnery oak or chaparral vs. 

shrubland) and papers in which the title 

contained other key fire-related words (e.g., 

burned, burning, and prescribed burning) 

but not fire, the search located more than 

500 papers (Table 2). Assuming our search 

provided an adequate, unbiased sample of the 

literature, we evaluated the search database to 

determine the nature of information available 

through the peer-reviewed literature. We then 

supplemented this information with papers that 

addressed specific topics within our charge for 

this project. As with the comprehensive search, 

Table 1. Number of papers identified for six topics in a Web of Science 

search of peer-reviewed journals. Each number represents the number of 

papers from the Web of Science for each topic listed. 

topic Fire Prescribed Fire 

rangeland 172 48 

shrubland 265 24 

grassland 931 138 

grazing 831 95 

Wildland 494 83 

Forest 6 648 671 

total with forest 9 341 1 059 

total without forest 2 245 318 

Table 2. Numbers of papers identified in a Web of Science search using 

fire (not prescribed fire) and each of the words in the first column in the title 

of the paper. These papers formed the initial database that was analyzed. 

title search combining fire and one 

of the following words number of papers 

shrubland 24 

savanna 157 

grazing 86 

Woodland 61 

Wildland 150 

rangeland 18 

grassland 107 

total 563 

we used Web of Science to search for papers 

on a particular topic. We justified limiting our 

search to the indexed literature on the basis that 

it is widely accepted as scientifically valid and 

the primary science published in peer-reviewed 

literature. 

evALuAtIon oF the dAtA set 

Of the 563 papers (available in January 2008) 

from our query through Web of Science (Table 

2), 474 papers were accessible and confirmed 

to be peer-reviewed research papers. Of these 

474 papers, less than 10 papers were published 

annually from 1967 through 1989 followed 

by a marked increase with 20 or more papers 

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 79 
	 	 	 	

Figure 1. The number of papers published per year from a total of 

563 papers published on rangeland fire. See text and Table 1 for 

explanation of papers selected. 

Table 3. Number of papers published by peer-reviewed scientific 

journals between1967 and 2007 based on a Web of Science 

search. 

Journal 

number of 

papers 

International Journal of Wildland Fire 25 

JrM/rangeland ecology and Management 23 

ecology 15 

Journal of Applied ecology 15 

Forest ecology and Management 13 

vegetatio /Journal of vegetation science 15 

Journal of ecology 12 

Journal of tropical ecology 12 

African Journal of ecology 11 

Austral ecology 10 

Plant ecology 10 

Journal of Arid environments 8 

remote sensing of environment 8 

Australian Journal of ecology 7 

biotropica 7 

ecological Modelling 7 

Journal of biogeography 7 

published each year from 2000 to 2007 

(Fig. 1). These data suggest that the research 

community may view fire with increasing 

importance but these conclusions could be 

limited by the words included in the search. 

An important outcome of the search was that 

rangeland fire research literature is dispersed 

among numerous ecological journals. 

Furthermore, most continents are well 

represented in the research, and topics include 

those not explicitly addressed in the NRCS 

purposes for prescribed burning. More than 

150 journals, mostly international ecological 

or applied ecological journals, published 

rangeland fire research (Table 3). Most of 

the research was located in North America, 

but substantial research was conducted in 

Africa, Australia, South America, and Europe. 

Research from the United States contributed 

214 of the 474 papers in the data set. The 

majority of papers reported research on 

plants, fire management, soils, fire behavior, 

socioeconomics, and wildlife. Authors 

described their papers as addressing a variety of 

vegetation types, with over half of the papers 

classified as savannas and grasslands (Fig. 2). 

Most of the articles recognized by our search 

terms in the United States reported research 

from the Great Plains, followed by the West 

Coast, Intermountain West, Eastern Forests– 

Grasslands, and Desert Southwest (Table 4). 

Topics in the database focused on plants, 

socioeconomics, fire management, soils, fire 

behavior, and wildlife, in respective order from 

highest to lowest, with all other topics having 

10 or fewer papers (Table 4). 

Perhaps the most revealing outcome of 

our search was that it uncovered critical 

limitations to applying the research literature 

to management applications, which is a 

fundamental barrier to constructing researchinformed purposes for prescribed burning. 

First, we found that most of the research 

was conducted at temporal and spatial scales 

inappropriate to management. Second, the 

fire research literature generally ignores fire as 

a dynamic disturbance process that varies in 

frequency, intensity, and time since fire (most 

studies are less than 5 yr postfire). Finally, most 

research failed to evaluate fire in the context 

of other disturbances, such as grazing and 

drought, on complex landscapes. More than 

80 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

half of the papers reported research conducted 

on experimental units of 1 ha or less and many 

studies were conducted on much smaller plots 

and on individual plants (Fig. 3). Fifteen 

percent of the studies were based on modeling 

and 6% on geographic information systems 

with minimal field evaluations or immediate 

application to management. Many studies 

(27%) simply compared a burn treatment to an 

unburned control, which obviously simplifies 

fire to the point of irrelevance to management. 

Fire regime, the features that characterize fire 

as a disturbance within an ecosystem—fire 

frequency, severity, behavior (i.e., fire intensity), 

predictability, size, seasonality, and spatial 

pattern (Morgan et al. 2001)—was rarely 

evaluated. Only 12.5% of the papers focused 

on fire frequency, and only 4% focused on 

understanding changes that occur over variable 

times since fire. Fire season was evaluated in 7% 

of the studies and fire intensity was evaluated 

in 9% of the studies. Most studies failed to 

specifically discuss the interaction of fire with 

Figure 2. Percentage of 474 rangeland fire studies within each 

of seven major vegetation types as described by the authors. 

Papers were published in peer-reviewed journals and located in 

a Web of Science search. 

Table 4. Number of papers published reporting research on topic areas conducted within geographic regions of the United 

States based on a Web of Science search. 

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 81 

topic area 

not 

specific 

eastern forests 

and grasslands 

great 

Plains 

Intermountain 

West 

southwest 

deserts 

West 

Coast total 

Plants 4 9 22 8 13 12 68 

soil 1 2 8 2 2 2 17 

Water 1 0 0 0 0 0 1 

Wildlife 2 1 3 3 1 2 12 

Arthropods 0 0 8 0 1 0 9 

Livestock 1 0 0 0 0 0 1 

Fire Management 4 4 0 7 1 9 25 

Fire behavior 5 3 2 1 2 1 14 

smoke Management 1 0 0 1 0 1 3 

socio-economics 10 2 2 7 1 6 28 

Air 0 1 2 2 2 3 10 

history 1 2 0 3 0 1 7 

health 3 0 0 4 0 3 10 

other 3 0 2 0 1 3 9 

total 36 24 49 38 24 43 214 
	 	 	 	

 

Figure 3. Number of rangeland fire papers published in each of 

four spatial scales. Papers were published in peer-reviewed journals 

and located in a Web of Science search. 

grazing and only 26% and 19% of the studies 

specifically stated that they included ungrazed 

and grazed sites, respectively. Grazing was a 

part of the experimental design in only 13% 

of the studies. Because the vast majority of 

rangeland is grazed, the failure of research 

to address the interaction of fire and grazing 

severely limits applying the research to support 

NRCS purposes for prescribed burning. 

evALuAtIon oF FIre eFFeCts on 

eCosysteM CoMPonents 

Prescribed fire is currently conducted on 

rangelands for many reasons, but a primary 

purpose is to reduce encroachment of invasive 

woody plants (see Fire and Plants section, 

Composition Changes subsection). Because 

fire can both positively and negatively 

influence ecosystem components, fire 

should be evaluated from the perspective of 

all ecosystem components. Therefore, we 

evaluate the literature available on plants, 

soil, water, wildlife, arthropods, livestock, 

fire management, fire behavior, smoke 

management, socio-economics, air quality, and 

fire history. Although we will evaluate the entire 

dataset when appropriate, on occasion we focus 

on data from specific rangeland regions of 

the United States to illustrate differences and 

similarities between regions. 

Fire and Plants 

The data set included 220 papers focused on 

plants and plant communities. Nearly 25% 

of the papers on plants evaluated community 

composition and structure following 

prescribed fire and wildfires. Ten percent of 

the papers evaluated biomass production, 

6% discussed plant diversity, 5% considered 

mortality to individual plants, 5% focused 

on seed germination and establishment, and 

2% considered plant (forage) quality. The 

remainder of the topics considered (plants as 

fuel, physiology, invasive species, harvesting, 

seed production) and were all below 2% of the 

papers evaluated and no conclusions could be 

drawn on these topics. 

Biomass and Forage Quality. Plant 

productivity, plant nutrient content, plant 

diversity, and plant mortality responses to 

fire are highly variable (Blair 1997; Reich et 

al. 2001). Much of this variability depends 

on how productivity is defined. Forage or 

herbaceous production usually decreases for 

1–2 yr after fire followed by a positive to 

neutral effect. An exception to this occurs in 

the tallgrass prairie where productivity can 

be enhanced from one spring fire following 

several years without fire (Blair 1997). Total 

aboveground biomass appears to be negatively 

associated with fire frequency although this 

depends on the ecosystem. Several studies 

indicate an increase in forage quality in mesic 

grasslands following fire (Hobbs et al. 1991; 

see wildlife and livestock sections), but these 

patterns are uncertain in more xeric regions. 

Grazing animals throughout all regions select 

recently burned areas for foraging sites over 

unburned sites, reflecting improved forage 

quality following fire. At a minimum, fire 

maintains openness of grassland landscapes, 

which sustains herbaceous biomass and forage 

quality/quantity. Because individual species 

and communities differ widely in response 

to fire, species diversity and plant mortality 

following fire cannot be generalized. Some 

evidence supports the NRCS purpose of using 

fire to enhance seed production, germination, 

and establishment, but these patterns are 

also species- and site-specific and cannot be 

generalized over all plants and ecosystems. 

Compositional Changes. The majority 

of plant papers focused on community 

82 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

composition. There were 68 papers in the 

data set that evaluated vegetation dynamics 

and plant responses to fire within the United 

States. We evaluated 81 papers to identify 

regional differences and similarities among 

the Great Plains, Intermountain West, and 

the Desert Southwest. Of the papers reviewed, 

65% reported results of prescribed burns, 

21% reported results of wildfires (Fig. 4), and 

most (> 75%) fire treatments were applied in 

spring and summer. Several studies recognized 

that season of fire mostly had a minimal or 

temporary effect (Engle and Bidwell 2001). A 

major concern from the database is the limited 

number of long-term studies (Fig. 5). Twentytwo percent of the studies extended past 10 

yr, of which a majority substituted space for 

time (comparing plant succession across fires 

of different ages), with the majority of studies 

(64%) not extending beyond 3 yr. 

Plant response to fire was highly variable both 

across and within regions and ecological sites 

(Table 5). A large portion of this variability 

can be attributed to the interaction of multiple 

variables, which include site characteristics, 

fuel characteristics, climate, community 

composition, time since fire, fire season, fire 

intensity, and postfire management. However, 

several patterns are evident in fire related plant 

responses across regions and ecological sites. 

Perennial grasses declined in abundance in the 

first postfire growing season in 76% of the 

studies, but usually recovered within the second 

or third year. Abundance of perennial grasses 

increased in only 11% of the studies in the 

first postfire growing season and in 5% during 

the second or third year following fire, but no 

studies reported long-term declines in perennial 

grasses. Annual grasses were usually more 

abundant in the first, second, and third years 

following fires compared to unburned stands. 

Annual and perennial forbs were inconsistent 

in their response the first year following fire, 

but they were more abundant in four out of six 

studies by the second or third year. Abundance 

of both resprouting and nonsprouting shrubs 

(biomass, cover, or volume) was lower during 

the first 10 yr following fire. However, density 

of sprouting shrubs usually equaled or exceeded 

that of unburned communities within 3 yr 

following fire suggesting little or no mortality. 

Full recovery of sprouting shrubs occurred 

within 3–20 yr; recovery took 25–35 yr for 

Figure 4. Number of rangeland fire studies addressing plant responses 

to prescribed fire and to wildfire in three rangeland regions of the United 

States. 

Figure 5. Time period over which individual research studies 

investigated plant response to fire on rangelands. 

nonsprouting shrubs on relatively wet sites 

(e.g., mountain big sagebrush [Artemisia 

tridentata subsp. vaseyana] in the 300–400-

mm precipitation zone) and greater than 45 yr 

on dry sites (e.g., Wyoming big sagebrush [A. 

t. subsp. wyomingensis] in the 200–300-mm 

precipitation zone). 

The use of fire to increase cover, density, and 

biomass of herbaceous vegetation, particularly 

perennial grasses, is only weakly supported in 

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 83 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Table 5. Numbers of studies indicating negative (−), positive (+), and no change the research literature. However, possibly the 

(=) in response of plant groupings (total herbs, perennial grasses, etc.) to fire strongest argument for the use of prescribed fire across specific regions (Great Plains, Intermountain West, and Desert Southwest). is to maintain or restore a desired successional 

community or pattern of communities in 

different stages of recovery following fire. 

Although herbaceous vegetation rarely increases 

the year of the fire, herbaceous dominance 

over woody plants is favored by shorter fire- 7 4 3 2 2 3 0 1 1 return intervals. A number of studies across 

0 0 1 0 1 0 0 0 0 US rangelands reported that shrubs and 

trees increase and herbaceous vegetation 

decreases with long-term fire removal, so 

0 0 0 0 0 0 0 0 0 

20 3 3 3 1 17 0 1 1 maintaining or increasing relative abundance 5 2 1 1 0 5 0 1 0 of herbaceous vegetation requires periodic fire. 

The length of the fire-return interval can also 

determine the proportion of area occupied by 

10 1 2 1 1 7 0 0 1 

5 0 0 1 0 5 0 0 0 grassland, shrub-steppe, and conifer woodland 1 4 0 0 3 1 0 0 0 communities. 

1 2 0 0 1 0 0 0 0 

Fire can be used to change plant composition 0 2 0 0 1 0 0 0 0 (e.g., the proportion of C3

:C4 plants, 0 0 0 0 1 0 0 0 0 herbaceous:woody plants, and forbs:grasses) 

and reduce excessive litter buildup resulting 

in an increase of light to basal tillers (an 

2 3 0 0 3 1 0 0 0 

1 0 0 0 1 0 0 0 0 

issue restricted to highly productive sites). 0 3 1 0 2 0 0 0 0 The literature is mixed on one of the greatest 

concerns over the use of prescribed fire—the 

potential for increasing invasive species. Fire can 

1 0 0 0 0 1 0 0 0 

2 2 1 0 1 1 0 0 0 

act as a trigger to force a desirable stable-state 1 0 0 0 1 1 0 0 0 that may be at risk of resilience loss across a 

1 2 0 0 0 0 0 0 0 threshold to an undesirable invasive plant state. 

Cheatgrass provides an excellent example of this 0 0 0 0 0 0 0 0 0 

dynamic in the Intermountain West. On the 4 0 0 1 0 0 4 1 1 other hand, fire can be used to control invasive 

0 0 0 0 0 0 0 0 0 species through direct control or by focusing 

herbivory on a relatively small burned area 1 0 0 0 0 0 4 1 1 

within a landscape (Cummings et al. 2007). 3 0 0 1 0 0 0 0 0 

9 0 0 5 0 0 0 0 0 Great Plains. Of the 36 papers reviewed from 

the Great Plains, almost all were prescribed 0 0 0 0 0 0 0 0 0 

fires at the sublandscape level (plots and 9 0 0 5 0 0 0 0 0 stands) evaluating burns during the spring 

0 0 0 0 0 0 0 0 0 and summer. Several studies also evaluated the 

timing of burning in the spring (early, middle, 4 0 1 2 0 0 1 0 0 

and late) and reported a significant effect on 0 0 0 0 0 0 0 0 0 vegetation response. However, a literature 

3 0 0 2 0 0 1 1 0 review on the effect of season of burn on 

herbaceous species in tallgrass prairie suggested 0 0 0 0 0 0 0 0 0 

that the data were not conclusive (Engle and 3 0 0 4 1 0 4 1 2 Bidwell 2001). Only a limited number of 

3 0 0 4 1 0 4 1 2 studies reported the method of burning or 

the prefire and postfire conditions. Nearly 0 0 0 0 0 0 0 0 0 

two-thirds of the studies were less than 3 yr in 0 0 0 0 0 0 0 0 0 duration with only 14% exceeding 10 yr. 

Plant grouping 

1 yr postfire 2–3 yr postfire ≥ 4 yr postfire 

(−) (+) (=) (−) (+) (=) (−) (+) (=) 

total herbs 7 4 4 2 3 3 0 1 1 

great Plains 

Intermountain West 

desert southwest 

Perennial grasses 

great Plains 

Intermountain West 

desert southwest 

Annual grasses 

great Plains 

Intermountain West 

desert southwest 

Perennial forbs 

great Plains 

Intermountain West 

desert southwest 

Annual forbs 

great Plains 

Intermountain West 

desert southwest 

shrubs, sprouting 

great Plains 

Intermountain West 

desert southwest 

shrubs, nonsprouting 

great Plains 

Intermountain West 

desert southwest 

trees, nonsprouting 

great Plains 

Intermountain West 

sW deserts 

Woody plants 

great Plains 

Intermountain West 

desert southwest 

84 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

	 	 	

	 	

	 	

	 	

	 	 	

	 	

	 	

	 	

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

Production and composition of herbaceous 

communities following fire were highly variable 

across studies. In the first growing season 

following fire, total herbaceous vegetation 

was less abundant in seven studies, more 

abundant in four studies, and did not differ 

from unburned in three studies. In one study 

burning increased photosynthesis and nitrogen 

uptake in perennial grasses in the first year, 

but biomass was less on burned plots than 

on unburned plots. In the second and third 

postfire growing seasons (most often the third 

year), herbaceous plant abundance generally 

increased to equal or exceed that of unburned 

plots. Late spring fires (May) often increased 

biomass of tallgrasses whereas early spring 

fires reduced biomass. Spring fires favored late 

flowering and C4

 plants, whereas summer fires 

favored C3

 and early flowering plants. Timing 

of burning also influenced the proportion of 

grasses and forbs. Fire appeared to have an 

extended negative effect on herbaceous biomass 

if drought followed in the first season postfire. 

However, one study reported only a very weak 

relationship between fire, weather, and plant 

response. Perennial grass biomass usually 

increased following fire in productive tallgrass 

sites where excessive accumulation of mulch 

occurs. Forb production was often reported to 

be greater on unburned plots. 

Shrub abundance (biomass, cover, volume) 

consistently declined the first year following 

fire and was generally less than the controls 

3 yr after fire. Density of resprouting shrub 

species recovered or exceeded preburn levels 

within 3 yr following fire. Most shrubs in the 

Great Plains are resprouting and fire return 

intervals of 2–5 yr may be required to maintain 

herbaceous dominance. Nonsprouting 

encroaching trees, primarily Ashe juniper 

(Juniperus ashei) and eastern redcedar (Juniperus 

virginiana), increase without fire and gain 

dominance after about 30 yr. The effects of 

these fires depend on grazing intensity, which 

constrains fuel load and fire intensity. 

Intermountain West. Of a total of 36 studies 

reviewed, 32% investigated individual plant 

species and 73% emphasized plant community 

responses. The majority of studies evaluated 

summer burns (57%, many of which were 

wildfires), but more than half considered fall 

or spring burns (35% fall, 21% spring) and 

only a single study evaluated winter burning. 

Spring and fall burns were prescribed fires. 

The majority of studies were short-term (72%) 

and only 8% exceeded 10 yr. Very few studies 

reported the method of burning or prefire and 

postfire conditions. 

Thirteen studies reported total perennial grass 

response in the first postfire growing season, of 

which 10 reported a decline in cover, biomass, 

or density; one an increase in cover; and two 

no change in cover. The majority of these 

studies (7 of 9) showed that perennial grass 

recovered to that of unburned plots within 2–3 

yr, whereas one study showed a decline in cover 

and one study an increase in cover. Perennial 

forbs generally increased as did annual grasses 

in the first postfire growing season. Bluebunch 

wheatgrass (Pseudoroegneria spicata), Sandberg 

bluegrass (Poa secunda), and squirreltail 

(Elymus elymoides) were the most resistant to 

fire whereas Idaho fescue (Festuca idahoensis 

Elmer), Thurber’s needlegrass (Achnatherum 

thurberianum), and rough fescue (Festuca 

campestris) consistently declined in the first 

year and either remained lower or recovered 

densities or cover equal to that of unburned 

plots. Broadleaved grasses and smaller bunches 

typically were more resistant to fire than fineleaved grasses or large bunches. 

Nonsprouting shrubs, primarily mountain big 

sagebrush and bitterbrush (Purshia tridentate, 

a weak resprouter), consistently decreased with 

fire and did not recover for 25–35 yr. However, 

recovery of Wyoming big sagebrush usually 

took longer with one study reporting only 5% 

sagebrush cover after 23 yr following fire. Cover 

of all shrubs (sprouting and nonsprouting) 

was reduced after fire. Few studies have 

evaluated sprouting shrubs (yellow and rubber 

rabbitbrush [Chrysothamnus viscidiflorus and 

Ericameria nauseosa], horsebrush [Tetradymia 

spp.]). However, limited work indicates 

biomass declines of these species in the first 

several years following fire with density 

typically recovering to preburn levels within 3 

yr. In the Southwest, sprouting shrubs were the 

dominate vegetation 25 yr following fire. 

Juniper (Juniperus spp.) and piñon pine (Pinus 

spp.) densities are reduced following fire, 

but large trees are more difficult to kill than 

small trees. Juniper cover of individual trees 

A number of 

studies across 

US rangelands 

reported that 

shrubs and trees 

increase and 

herbaceous 

vegetation 

decreases with 

long-term fire 

removal...”

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 85 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Figure 6. Response of grassland birds to time since focal disturbance by fire and 

grazing at the Tallgrass Prairie Preserve, Oklahoma, 2001–2003. Some birds 

native to the area require recently burned patches that are heavily grazed whereas 

others require habitats that are undisturbed for several years (Fuhlendorf et al. 2006). 

This research emphasizes that 1) the response of grassland birds to fire is highly 

dependent upon the interaction of fire and grazing and 2) fire management should 

not be considered in isolation from other environmental factors, including grazing. 

Figure courtesy of Jay Kerby and Gary Kerby. 

increases slowly for the first 45 yr followed 

by rapid increase during the next 46–71 yr. 

Closed canopies can develop within 80–120 yr 

(Johnson and Miller 2006). Fire in woodlands 

is typically followed by an increase in perennial 

grasses and a reduction of woody plants. 

Sagebrush (Artemesia spp.) reached preburn 

levels in 36 yr and then often declined if piñon 

pine and/or juniper became established on 

the site. Understory cover declined to 5% of 

the adjacent grassland by 100 yr following fire 

as piñon and juniper woodlands developed 

(Barney and Frischknecht 1974; Wangler 

and Minnich 1996). However, understory 

composition following fire is highly dependent 

on the composition and abundance of the 

understory prior to the burn. 

Fire and Wildlife 

Of the 40 papers we evaluated concerning 

the effect of fire on rangeland wildlife, only 

12 papers addressed US rangeland wildlife. 

These 12 papers focus on avifauna and small 

mammals, reflecting the large influence exerted 

by fire on habitat structure, to which these 

vertebrate assemblages are especially sensitive. 

It is interesting to note that measurements 

in most wildlife studies, including those 

we sampled, focus on wildlife population 

response and relatively few (only 2 of the 12 

US studies) measured vegetation attributes 

(e.g., horizontal and vertical structure). Ten 

of the 12 studies were published since 2000, 

which indicates a recent upswing in research 

interest in wildlife response to fire. However, 

only 2 of the 12 studies included private land. 

The 12 studies were spread more or less evenly 

across geographic regions and vegetation types. 

As might be expected from a small number 

of studies, the studies addressed only a small 

number of questions related to the fire regime 

and the grazing environment. For example, 

only one of the studies (Fuhlendorf et al. 

2006) addressed the ecological interaction of 

fire and grazing. From the Fuhlendorf et al. 

(2006) study and related research, we know 

that the ecological interaction of fire and 

grazing strongly influences habitat selection and 

habitat value for virtually all rangeland wildlife, 

including birds (Churchwell et al. 2008; 

Coppedge et al. 2008) and large ungulates 

(Hobbs and Spowart 1984; Coppedge and 

Shaw 1998; Biondini et al. 1999; Van Dyke and 

Darragh 2007). Large herbivores are attracted 

to nutritious regrowth of herbaceous vegetation, 

sometimes emerging outside the growing season 

(Coppedge et al. 1998; Biondini et al. 1999) 

on recently burned areas (Hobbs and Spowart 

1984; Hobbs et al. 1991; Turner et al. 1994). 

In contrast, many, but not all, rangeland small 

mammal and bird species are more suited to 

areas not recently burned and grazed because 

these areas provide vegetative cover required for 

concealment or nesting. However, this influence 

can be mediated by drought (see Meek et al. 

2008) and other factors. 

The context in which prescribed burning is 

applied on rangeland marks the effect on 

wildlife species in question. Wildlife species in 

a given area have variable habitat requirements, 

so positive response by one species will 

likely cause other species to decline (Fig. 6). 

However, because rangeland and rangeland 

wildlife evolved with periodic fire and 

because periodic fire is required to maintain 

habitat suitable for wildlife species native to a 

particular rangeland region, fire is essential for 

maintaining rangeland wildlife populations. 

86 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

Unnaturally long fire-return intervals often lead 

to tree encroachment and other changes that 

reduce habitat suitability for native wildlife 

species that are habitat specialists (Coppedge et 

al. 2001; Reinkensmeyer et al. 2007), some of 

which are species of conservation concern. In 

contrast, fire-return intervals greater than those 

with which an ecosystem evolved can have 

correspondingly deleterious effects on habitat 

and populations of habitat specialists (Robbins 

et al. 2002; Pedersen et al. 2003; Fuhlendorf et 

al. 2006; Rowland et al. 2006). 

Fire and Water 

We reviewed 25 papers that evaluated fire 

effects on various hydrologic processes in 

rangeland (Table 6). Hydrologic variables 

evaluated were water repellency (six papers), 

water quality (two papers), hydraulic 

conductivity or infiltration (six papers), and 

erosion/runoff (five papers). The majority of 

studies were conducted for 3 yr or less: 1 yr 

(52%), 2 yr (20%), and 3 yr (12%). Three 

studies (12%) were conducted for 5–6 yr, and 

one study was conducted for 9 yr. Variables that 

influenced the effects of fire on hydrology were 

aspect, fire severity, and microsites (coppice 

dunes formed beneath shrubs and trees vs. 

interspace). The largest decrease in infiltration 

rate and increase in erosion following fire 

occurred in coppice dunes beneath shrubs and 

trees. Fire had little effect on these two variables 

in shrub or tree interspaces. Water repellency 

usually occurred on both burned and unburned 

sites but usually increased, particularly beneath 

shrub and tree canopies, following fire. 

Hydrophobicity was reported to decline within 

several months to near preburn levels following 

wetting. Soil erosion on cooler, wetter sites 

in sagebrush-steppe communities (e.g., north 

aspects, or sites occupied by Idaho fescue 

compared to bluebunch wheatgrass) were less 

affected by fire than drier, warmer sites. One 

study reported rill erosion as the primary source 

of sediment and several studies reported rills 

readily formed in the coppice dunes. In general, 

these studies suggest that immediate effects of 

fire are largely negative on watersheds, but that 

the effects are short-lived. 

Fire and Arthropods 

Eighteen studies and one extensive literature 

review were evaluated for the effects of fire on 

arthropods. The majority of studies evaluated 

the response of grasshoppers (six studies) or 

arthropods in general (n = 6) to fire. Other 

species studied were ants (three studies), 

beetles (one study), and ticks (one study). In 

a literature review, Swengel (2001) reported 

few studies were conducted at the species 

level. The response of insects to fire was highly 

variable. Short-term and long-term response 

of insects to fire was influenced by intensity, 

complexity or patchiness of the burn, species 

requirements, and plant recovery. Thirteen of 

the 17 studies were conducted for 3 yr or less, 

three studies ranged from 4 yr to 9 yr, and one 

study extended for 25 yr. Insect abundance 

was usually lower (with the exception of 

grasshoppers) immediately following fire (up 

to 1–2 mo). In a 7-yr study across 21 different 

Great Plains sites, Panzer (2002) reported 93% 

of the species were consistent in their response 

to fire over the period of the study. Immediately 

following fire, 26% of arthropod species 

increased and 40% decreased. Of those that 

declined nearly two-thirds recovered within 

2 yr. Insect orders Homoptera and Hemioptera 

appear to generally be more sensitive to fire 

whereas Orthoptera was little affected by 

fire. Fire effects on grasshopper populations 

generally showed limited response, but a shift 

in species composition frequently occurred. 

Fire history 

Obtaining a clear picture of the complex 

spatial and temporal patterns of historic fire 

regimes across the western United States 

before Euro-American settlement is unlikely. 

This can be attributed to limited sources 

of material (e.g., large charred wood or fire 

scars) available for reconstructing pre-historic 

fire regimes on most rangelands and the vast 

variation in fuel composition and structure, 

Table 6. Effects of fire on several hydrologic variables compared to 

unburned plots and time periods required for recovery to near preburn 

levels. 

variable decreased Increased no change recovery 

soil repellency 1 5 0 2.5–3 mo 

Infiltration 6 0 4 2 yr 

runoff 0 6 1 2 yr, 4–5 yr 

sediment loads 0 9 1 1 yr 

Water quality 2 0 0 3–5 yr 

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 87 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	

In many cases, when woody 

plants reach a substantial 

size and/or density, fires will 

either be ineffective or require 

greater expertise to conduct 

them. Using extreme fires for 

restoration is an emerging topic 

in conservation of rangelands. 

An active prescribed burning 

program could help minimize 

the risk associated with extreme 

fires. (Photo: John Weir) 

landscape heterogeneity, ignition from 

aboriginal and lightning sources, weather, and 

topography across this region. We reviewed 

24 papers that attempted to describe prehistoric fire histories across the western United 

States. We tried to capture all of the papers 

that collected quantitative data to reconstruct 

pre-historic fire regimes related to rangelands. 

We also included several studies in ponderosa 

pine in addition to woodlands that evaluated 

the timing of reduced fire occurrence and 

livestock grazing. Twelve of the papers were 

based on fire scar data, two used charcoal or 

ash samples, and three used the presence of 

old-growth piñon or juniper trees. Of the 

12 papers using fire scar data, samples were 

cross-dated in seven studies. Cross-dating 

is a procedure that verifies the exact year of 

the fire, important for calculating fire-return 

intervals and determining the extent of 

individual fires across larger areas. 

Duration of the fire record based on fire scars 

ranged from 250 yr to 500 yr before present 

(BP). Charcoal studies ranged from 5 500 yr 

to 6 000 yr BP. Pre–Euro-American settlement 

fire regimes reconstructed from cross-dated 

fire scars or large charred wood across western 

rangelands are few and primarily restricted 

to the intermountain region. Fire histories 

based on fire-scarred trees are also spatially 

limited to the rangeland–forest ecotones 

in the Intermountain West, which often 

occur as mosaics of conifers and sagebrushsteppe grasslands. Fire-scar samples are 

usually collected from fire-resistant trees 

88 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	 	

	 	

	 	

	 	

	 	

	 	

	 	

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

(e.g., ponderosa pine (Pinus ponderosa) and 

Douglas-fir (Pseudotsuga menziesii) and 

occasionally less fire-resistant trees (e.g., Utah 

juniper [Juniperus osteosperma], piñon pine). 

Several of these studies also evaluated tree age 

structure in adjacent forest and shrub-steppe 

patches. 

Pre-historic (pre-1900) mean fire-return 

intervals reported along ponderosa or Douglasfir–mountain big sagebrush-steppe ecotones 

varied from less than 10 yr (three studies) 

to 10–30 yr (six studies). Studies reporting 

longer fire-return intervals were associated with 

low sagebrush (Artemisia arbuscula; 90 yr to 

150 yr) and were based on tree age structure 

and charred logs and stumps of juniper. The 

relatively short fire-return intervals (< 30 

yr) would have supported grass-dominated 

communities along the forest ecotones. 

Extrapolation of these fire-return intervals 

away from range–forest ecotones is probably 

speculative and likely becomes longer in more 

arid ecological sites, especially those occupied 

by Wyoming big sagebrush. Macroscopic 

charcoal data collected in central Nevada 

suggested that fire-return intervals in the 

drier Wyoming big sagebrush cover type over 

the past several thousand years were up to a 

century, with fire intervals varying with climatic 

fluctuations. 

Several consistent patterns regarding firereturn intervals emerge from these papers. 

First of all, there is consistent evidence that 

most rangelands in the United States have 

experienced a dramatic increase in fire-return 

intervals over the past 100–200 yr. Six of 

the studies reported sharp declines in fire 

occurrences that coincide with the introduction 

of livestock. Piñon–juniper woodlands that 

have persisted for the past several or more 

centuries did not show evidence of highfrequency, low-intensity surface fires. Five of 

the studies reported probability of sites being 

occupied by old-growth trees to be associated 

with rocky surfaces and limited surface fuels 

but none as fire refugia. Three studies reported 

the probability of large fires increases in years 

preceded by wetter than average years. At 

a longer time scale, Mehringer (1987) and 

Mensing et al. (2006) reported a correlation 

of increased fires during periods of wetter than 

average conditions. 

Quantitative measures of pre-historic fire return 

intervals in the tallgrass prairie are not available 

for the Great Plains. The assumption that 

prehistoric fire regimes in the tallgrass prairie 

were characterized by frequent low-intensity 

fires is primarily based on 1) observations 

from early explorers, trappers, and settlers, 

and 2) research showing that in the absence of 

fire these grasslands shift rapidly from prairie 

to woody species (Bragg and Hulbert 1976). 

Several authors have estimated mean fire-return 

intervals of 3–5 yr (Wright and Bailey 1982; 

Knapp and Seastedt 1996). However, little 

is known about the dynamics of native grass 

and woody species prior to Euro-American 

settlement. It is also likely that the influence of 

bison on fuel loads affected fire-return intervals 

across the Great Plains. 

Fire and soils 

The vast majority (45 of 51; 88%) of the 

papers we evaluated on fire effects on 

rangeland soils were published in ecological 

or soil science journals rather than Journal of 

Range Management, Rangeland Ecology and 

Management, or applied ecology journals. 

Therefore, the overall emphasis within 

the research base leans toward ecological 

understanding rather than to explicitly 

answering management questions. Twentyeight papers (55%) reported effects of fire on 

soil chemistry (pH, nutrients), and 17 (33%) 

reported on the effects of fire on one or more 

variables (infiltration, soil water content, water 

repellency, erosion) related directly or indirectly 

to the water cycle. 

The literature on rangeland soils, including 

the effects of fire on rangeland soils, is quite 

voluminous. For example, one of the sampled 

papers is a recent analysis of the literature on 

water repellency. In it, Debano (2000) employed 

a bibliography of over 700 published papers 

reporting on either various aspects of water 

repellency (500 papers) or published papers 

(200) that contribute information directly 

related to understanding the basic processes that 

underlie soil water repellency. Water repellency, 

a global rangeland issue reported for numerous 

vegetation types following fire, also occurs in 

soils other than rangeland. 

The scope of these studies further limits the 

inferential base for applying the results to 

...there is 

consistent 

evidence that 

most rangelands 

in the United 

States have 

experienced a 

dramatic increase 

in fire-return 

intervals over 

the past 100– 

200 yr.”

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 89 
	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	

Bison at the Tallgrass Prairie 

Preserve in Oklahoma grazing 

on a recently burned patch. 

(Photo: Steve Winter) 

management of US rangelands. Although 

physical processes are not place-bound, 

only 16 of the 51 studies reported research 

from US rangelands. Fortunately, these were 

distributed more-or-less evenly across the 

United States (eight in the Great Plains or 

central prairies) and across vegetation types 

(grasslands, shrublands, etc.). However, 

small plots (0.0003–1 ha) were the general 

rule and studies often reported effects from 

a single fire (22 papers), and only 10 of the 

51 studies encompassed time periods of 10 

yr or more. An encouraging sign is a recent 

increase in published studies enhancing basic 

understanding of soil response to fire; the 

majority of papers published since 1998 (35; 

69%) focus on this. 

The influence of fire on soil depends largely 

on the prefire and postfire environment, 

interaction with other factors including grazing 

and invasive species, and the evolutionary 

history of the ecosystem with regard to 

fire frequency and grazing. However, it is 

notable that on US rangeland that are often 

characterized as lacking a long evolutionary 

history of frequent fire most fire research is 

based on observations following wildfires 

rather than controlled studies with prescribed 

fires. For example, portions of the Great Basin 

shrub-steppe have had substantial increases 

in fire-return interval and burn area over the 

past century (Miller et al. 2011). In contrast, 

prescribed burning and the ecological role of 

fire are the context of studies on rangelands 

characterized by a long evolutionary history of 

frequent fire, specifically the Great Plains. 

Soil organic matter, resistant to change when 

rangeland fire is wind-driven and fueled by 

fine fuel, has long been a subject of interest to 

rangeland fire researchers (e.g., Reynolds and 

Bohning 1956; Owensby and Wyrill 1973). 

Recent research has increasingly reported the 

influence of fire on soil organic carbon (and 

CO2

 ecosystem flux) and carbon sequestration, 

which is tied to atmospheric properties related 

to global climate change. In the single study 

90 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

located in the United States that appeared 

in our sample (Ansley et al. 2006a), carbon 

storage in soil increased with fire, likely 

the result of a shift in species composition. 

Recent research in the Intermountain shrubsteppe suggests plant invasions (i.e., Bromus 

tectorum) can reduce soil carbon (Bradley 

et al. 2006; Prater and DeLucia 2006), 

but this did not occur in a similar shrub 

community when perennial native grasses 

dominated postfire (Davies et al. 2007). 

Burning changed soil carbon in a semiarid 

Great Plains rangeland, but the magnitude 

of change was inconsequential partly because 

of a relatively low CO2

 flux (MacNeil et al. 

2008). In subhumid Great Plains rangeland 

where CO2

 flux is markedly greater, soil carbon 

flux increases with periodic burning over 

ungrazed rangeland because burning removes 

accumulated litter that creates temperature and 

light-limiting conditions for plant growth, but 

annual burning will reduce both soil organic 

matter and nitrogen mineralization (Ojima et 

al. 1994; Blair 1997). Annual burning over 

perhaps 20–100 yr may increase the fraction 

of passive soil organic matter at the expense of 

more active fractions, which might ultimately 

reduce total soil organic carbon (Ojima et al. 

1994). Nevertheless, most ecosystem carbon 

loss in rangeland results from combustion of 

aboveground organic material (i.e., fuel), with 

the time to reach prefire levels dependent on 

primary productivity (MacNeil et al. 2008). 

Soil carbon response to the ecological fire– 

grazing interaction has not been investigated, 

but because nitrogen and carbon are coupled 

in the organic matter pool, soil carbon might 

show similar increases following fire–grazing 

disturbances to soil nitrogen (Anderson et al. 

2006). 

Nitrogen in aboveground biomass is volatilized 

in fire and varies with environmental 

conditions and fire characteristics in that 

drier fuels and soils and hotter fires result in 

more intense combustion and more nitrogen 

volatilization (DeBano et al. 1979). Because 

most prescribed burning objectives call for 

conditions that consume most aboveground 

herbaceous fuel, it is often assumed that fire 

depletes ecosystem nitrogen. Indeed, postburn 

soil inorganic nitrogen (NO3

 and NH4) is 

often less, but greater herbaceous aboveground 

annual production and vegetation cover at 

some point after burning suggests plantavailable nitrogen increases following burning. 

Stocker calves grazing on 

areas recently burned in the 

summer and providing new 

growth forage in the fall at 

the Stillwater Research Range. 

Unburned patch is shown in 

the background. (Photo: Sam 

Fuhlendorf) 

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 91 
	 	 	 	

 

 

	 	 	

	 	

	 	 	

	 	

	 	 	

	 	

	 	

...documentation 

of fire behavior 

on rangelands 

does not provide 

suitable guidance 

to address the 

NRCS’s purposes 

of prescribed 

burning.”

Research in subhumid rangeland (Blair 1997) 

and semiarid rangeland (Davies et al. 2007) 

indicates that burning increases nitrogen 

mineralization and enhances other mechanisms 

that result in increased nitrogen. Therefore, 

burning indirectly enhances plant capability 

to utilize nitrogen. Annual burning of 

subhumid rangeland over a period of 20–100 

yr has been predicted to reduce mineralizable 

nitrogen similar to the effect on soil organic 

carbon (Ojima et al. 1994). Because grazing 

reduces the amount of nitrogen available for 

volatilization by fire and because nitrogen 

loss is proportional to biomass available for 

combustion, grazing weakens the effects of 

fire on soil nitrogen (Hobbs et al. 1991). This 

likely explains why nitrogen fertility was not 

diminished with annual burning coupled with 

long-term moderate grazing (Owensby and 

Anderson 1967). Moreover, this mediating 

effect of grazing is subject to the effect of 

scale and preferential grazing of patches 

(McNaughton 1984; Hobbs et al. 1991). When 

fire and grazing interact spatially (i.e., the fire– 

grazing ecological interaction) in a subhumid 

rangeland, plant-available nitrogen increases 

in recently burned, focally grazed patches 

(Anderson et al. 2006), but unburned patches 

with minimal grazing pressure have low levels 

of available nitrogen. No published research on 

the effects of the fire–grazing interaction on soil 

nitrogen is available for other rangelands. 

Some fire prescriptions, wildfire, and fuel 

situations in rangeland can result in extreme 

soil heating, which can markedly change soil 

chemical and physical properties. Brush piles 

and thinning slash, in particular, create intense 

heat that can change biological, chemical, 

and physical properties of soil and can induce 

undesirable vegetation change including plant 

invasions (Neary et al. 1999; Haskins and 

Gehring 2004). Although the mechanisms and 

impact of soil heating are known, other than 

postfire restoration (for example, see Korb et 

al. 2004), we found no studies that fashion 

rangeland fire prescriptions and vegetation 

management guidelines to reduce the impact of 

soil heating with burning brush piles and slash. 

Fire and Livestock 

Among the 476 papers evaluated pertaining 

to fire on rangelands, over 25% of the studies 

referred to grazing by either livestock or 

wildlife. However, only five papers evaluated 

the influence of fire on livestock performance, 

of which only one paper was from the 

United States. Recognizing the potential for 

bias within our initial literature review, we 

broadened our search criteria to include articles 

with “fire or burn” and “livestock, cattle, sheep, 

or horses” in the subject. Even with this less 

restrictive search, only three additional papers 

focused on fire and livestock and they revealed 

a span of 26 yr between studies. Due to the 

limited number and geographic scope of the 

studies, few detailed inferences can be made. 

Early research from the Great Plains region 

focused on livestock performance and indicated 

that midspring and late spring burning 

increased steer weight gains, particularly early 

in the growing season (Anderson et al. 1970; 

Smith and Owensby 1976). Burning portions 

of unimproved Florida rangeland annually 

improved calf production as much as 75% 

over unburned pastures (Kirk et al. 1976). 

The increased benefit from burning alone 

was comparable to supplemental feeding, 

especially when considering the increased 

cost of supplemental feed. Increased livestock 

performance is mostly attributed to increased 

forage quality from increased plant crude 

protein and decreased fiber content following 

fire (Grelen and Epps 1967; Allen et al. 1976; 

Kirk et al. 1976). Within the fire–grazing 

interaction, cattle are attracted to recently 

burned locations, which provide higherquality forage than the surrounding matrix 

of unburned vegetation, and devote 75% of 

their grazing time within recently burned areas 

(Fuhlendorf and Engle 2004). The disturbance 

created by the interaction of cattle grazing and 

fire mimics historical grazing behavior of large 

ungulates and creates a structurally diverse 

landscape. 

Prescribed Fire and Air Quality/smoke 

Management 

Rangeland fire generates a wide variety of byproducts that fall into two broad categories, 

gasses and particulates. Smoke, the visible 

product of partially combusted fuel material, 

contains an array of organic and inorganic 

airborne particulates. Airborne particles can 

be a nuisance, reducing visibility hundreds of 

kilometers downwind from emission sources 

(Ferguson et al. 2003; McKenzie et al. 2006) 

and degrading air quality (Martin et al. 1977). 

92 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 	 	 	 	 	

	 	 	 	 	

	 	 	 	

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

Prescribed fire produces 15–25% of airborne 

particulates and 7–8% of hydrocarbons emitted 

to the atmosphere annually (Martin et al. 

1977). However, environmental conditions, 

fuel characteristics, and characteristics 

of the fire itself influence the amount of 

noncombusted material produced. Relatively 

low-intensity fires, where more complete 

combustion is expected, were modeled to 

produce about 50% less smoke than higherintensity fires with higher rate of spread and 

less complete combustion (Glitzenstein et al. 

2006). Fuel consumption and fire intensity 

clearly influence emissions from rangeland fire. 

Particle size influences the period of suspension 

in the atmosphere. Relatively large particles, 

between 0.07 μm and 1.0 μm diameter, 

may take days to settle out, whereas small 

particles less than 0.07 μm do not settle 

under natural conditions (Martin et al. 1977). 

Larger particles do not remain suspended for 

long time periods, yet can be problematic 

for individuals with asthma or other chronic 

respiratory conditions (Dockery et al. 1993). 

Although smoke management is important as it 

relates to air quality, our review of the literature 

revealed that only six papers addressed smoke 

management on rangeland and thus conclusive 

evidence is limited. Further investigation is 

needed to provide a complete understanding of 

how prescribed fire influences air quality. 

Fire behavior and Fire Management 

In addition to contributing to an 

understanding of the influence of fuels and 

environmental conditions on fire behavior, 

many of the fire behavior studies that we 

reviewed more directly addressed questions 

related to other sections of this report. Only 

a few studies addressed plant responses (e.g., 

tree mortality; Kupfer and Miller 2005) as a 

function of fire behavior, and fire behavior was 

reported in several studies as one of several 

aspects of environmental conditions under 

which the study was conducted (e.g., Engle and 

Weir 2000; Ansley et al. 2006a). Measuring 

a correlate (e.g., char height on trees; Fule 

and Lauglin 2007) of a primary fire behavior 

characteristic (e.g., fireline intensity) was 

common (8 of 11 US papers). 

We examined our sample of published 

papers to determine the extent to which 

they addressed NRCS’s relevant objectives 

of prescribed burning (i.e., prepare sites for 

harvesting, planting, or seeding; reduce wildfire 

hazards; remove slash and debris). Of the 41 

studies sampled on fire behavior, only 11 were 

located in the United States (and three were 

in southeastern forests), so specific application 

to US rangelands is minimal for at least threequarters of the studies. Of the 11 US studies, 

two studies (Sparks et al. 2002; Glizenstein et 

al. 2006) related to wildfire hazard reduction 

and slash removal in southeastern US 

forests, and no studies were related to fuels 

management on rangeland. One study (Gilless 

and Fried 1999) examined a computer model 

for its utility in planning fire suppression. 

Based on our literature search, it appears 

documentation of fire behavior on rangelands 

does not provide suitable guidance to address 

the NRCS’s purposes of prescribed burning. 

For example, mortality or scorch height of 

invasive woody plants is highly dependent 

on fire intensity, which is rarely measured in 

rangeland studies. 

The refereed literature on fire management 

is insufficient to adequately evaluate the 

NRCS’s purposes of prescribed burning (e.g., 

what types of management will promote 

different purposes). However, a significant 

quantity of literature on fire behavior and 

fire management, addressing both prescribed 

burning and fire danger related to rangeland 

is found in federal government documents, 

especially those published from the US Forest 

Service Fire Sciences Laboratory in Missoula, 

Backfire of a prescribed burn 

at the Tallgrass Prairie Preserve 

of Oklahoma. (Photo: Terry 

Bidwell) 

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 93 
	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	 	 	

	 	

	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	

	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	

	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	

	 	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	 		

	 	 	

	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	

	 	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

	 	 	

	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

	 	

	 	 	 	 	 	 	

	 	 	 	 	

	 	

Table 7. Assessment of the 10 purposes within the Natural Resources Conservation Service (NRCS) prescribed burning conservation practice 

standard relative to the supporting experimental evidence. Observations on the evidence provided by the peer-reviewed scientific literature 

supporting NRCS purposes for prescribed burning. 

Purpose Aspects that agree with the purpose 

Aspects that suggest limited or no 

support for purposes Further needs and considerations 

Control undesirable 

vegetation 

Fire can be effective in reducing the stature 

of resprouting, fire-adapted shrubs and trees, 

some invasive herbaceous plants, and the 

encroachment of fire-sensitive shrubs and trees. 

Most effects are too short-lived to support 

the purpose over meaningful management 

time spans. 

Most species are reduced only in stature 

and without mortality. 

Many species are not negatively affected 

by fire, and some species increase after fire. 

Generalizations concerning season, frequency. 

and intensity of fire are mostly unsupported. 

Prepare sites for 

harvesting, planting 

and seeding 

No evidence in our database—additional 

searches1 generated few additional papers 

with controlled comparisons. 

Control plant 

diseases 

One study from a comprehensive search1 

suggested that fire can reduce the density of 

host species (oak), but the authors recommended 

more study to validate their results. 

No evidence in our database—additional 

searches1 generated few additional papers 

with controlled comparisons. 

reduce wildfire 

hazards 

Models and observational studies suggest 

that fire-induced vegetation changes can alter 

subsequent fire regimes and reduce fine fuel 

loads to decrease the likelihood of high-intensity, 

stand-replacing, destructive wildfire. 

Prescribed fire that reduces woody plants 

and maintains grassland productivity can 

increase the likelihood of fire. 

No evidence in our database—additional 

searches1 generated few additional papers 

with controlled comparisons. 

Improve wildlife 

habitat 

Fire can maintain and restore habitat for native 

wildlife species in some situations. 

Any action that improves habitat for one 

species likely degrades habitat for another. 

Studies limited mostly to birds and small 

mammals. Time since fire, the key element, 

has been minimally studied. 

Improve plant 

production quantity 

and/or quality 

Several studies indicate increased plant 

production and forage quality but these are 

mostly restricted to the Great Plains—several 

additional studies indicate animal preference 

for burned sites. Most studies show an initial 

decrease in quantity and then recover y, with 

limited studies showing an actual increase 

over time. However, majority of studies do not 

evaluate response beyond 5 yr. 

Minimal information on grazing animal 

response following fire. Therefore, if the 

purpose is intended to benefit livestock 

production through increased forage production 

or improved forage quality, support is limited. 

The use of terminology such as “plant 

quality” is overly broad and could suggest 

wildlife habitat is improved following fire, 

but this is not supported well in the research 

literature. 

remove slash and 

debris 

Fire can be used in Southeastern pine forest 

to remove slash and maintain savanna and to 

remove brush piles and brush windrows. 

No evidence in our database—additional 

searches1 generated few additional papers 

with controlled comparisons. Because fire 

effectiveness varies, more study is needed on 

fire effects and fire management after brush 

treatments to restore rangeland. 

enhance seed 

production 

Fire can increase seed production of both native 

species and exotic invasive plants. 

Seed production is rarely measured in fire 

research. This purpose is irrelevant to those 

rangeland plants that reproduce vegetatively. 

Facilitate distribution 

of grazing and 

browsing animals 

Recently burned areas generally attract grazers 

because burning increases herbivore access to 

current year’s forage growth. 

The fire–grazing interaction is an appropriate 

tool that employs attraction of large herbivores 

to recently burned areas, but research is limited 

to experimental studies in Oklahoma and 

observational studies of wildlife. 

restore and maintain 

ecological sites 

Fire regimes that mimic evolutionar y conditions 

of the rangeland in question will maintain 

ecological sites and therefore maintain 

grassland, savanna, and shrubland ecosystem 

structure. 

Reintroducing fire will not always restore 

ecological sites. Prescribed burning that 

creates fire regimes that do not mimic 

historic fire regimes can induce site 

degradation by altering biotic and abiotic 

characteristics (e.g., hydrophobic soil). 

Most ecological sites lack a complete fire 

regime description (especially where fire 

intervals are long). 

1Additional literature searches used a topic search in Web of Science for fire AND rangeland AND Reduce wildfire hazards. 

94 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	 	

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

Montana. The early science on fire behavior 

that culminated in Rothermel’s (1983) seminal 

user-friendly fire behavior prediction model 

was applied to rangeland, and many rangeland 

fire managers have used this version of the 

model that does not require a computer. More 

recently, a suite of dynamic computer models 

have greatly expanded the management value 

of the Rothermel model to applications in 

variable terrain and varied fuels and fuel loads. 

Coupling these models with sophisticated 

fuel models informed by state-of-the-art fire– 

weather observing stations and near real-time 

remote sensing of fuels has greatly enhanced 

fire management on privately owned rangeland 

(Carlson et al. 2002; OK-Fire 2009). 

Fire and human dimensions 

From all of the papers evaluated, 36 addressed 

a wide variety of human aspects related to 

rangeland fire ranging from education and 

public perception of fire to health and policy 

issues. Although all regions of the United 

States were covered, over 80% of the studies 

focused on the West Coast or Intermountain 

regions. Human dimensions on rangelands 

have recently gained attention and are reflected 

in 75% of the studies pertaining to fire dated 

2000 or later. Because of the recent interest in 

human dimensions, CEAP devoted an entire 

section to socioeconomics, so we restricted our 

summarizations to limit duplicate information. 

dIsCussIon oF FIndIngs 

Most research evaluated here was not 

developed with the intent of providing specific 

recommendation for management of rangeland 

landscapes. Moreover, constructing research 

based on NRCS purposes for prescribed 

burning is limited by spatio-temporal 

scale of the research; limited description of 

conditions prefire, postfire, and during the 

fire; failure to account for interaction with 

other disturbance processes occurring on 

rangelands (e.g., grazing, drought); simplifying 

a complex fire regime to a single treatment 

event; and the lack of a focus on fire effects 

that are highly dependent on time since fire. 

Even with this paucity of research evidence, 

our evaluation demonstrates that several of 

the NRCS purposes for prescribed burning 

can be justified but with many caveats (Table 

7). Specifically, management of woody plant 

invasion is supported by a fairly consistent 

response in which prescribed fire limited 

invasion. It is less clear if fire can reverse 

woody plant invasions when thresholds have 

been crossed. The purpose of using prescribed 

burning for short-term control of undesirable 

plants is justifiable based on research that 

shows some herbaceous species respond 

negatively to fire in the first growing season 

following fire, especially when combined 

with focal grazing. However, most herbaceous 

species recover within 2–3 yr postfire 

regardless of season of the burn. Contrary 

to this purpose, some regionally important 

herbaceous species in each rangeland region 

of North America respond negatively to fire 

1 yr, 2 yr, and 3 yr postfire. In general, few 

negative effects and more neutral and positive 

effects have been demonstrated on target 

herbaceous species in response to fire. With the 

exceptions of the expansion of invasive annual 

grasses in the Intermountain West following 

fire (Miller et al. 2011), these conclusions were 

surprisingly consistent across the Great Plains, 

Intermountain West, and Desert Southwest. A 

few studies report increased productivity and 

forage quality the year of the fire, but because 

these are mostly from mesic grasslands, this 

NRCS purpose for prescribed burning is not 

broadly supported by research evidence. 

It is widely recognized that the importance 

of all processes on ecosystem structure and 

function are highly dependent on the scale 

of observation. In fact, it has been suggested 

that studies should be conducted at multiple 

scales and the interpretations of research 

should recognize the limited ability to 

translate conclusions across scales. Several 

rangeland studies have demonstrated that 

vegetation dynamics (Fuhlendorf and Smeins 

1996, 1999; Briske et al. 2003) and wildlife 

populations (Fuhlendorf et al. 2002) are highly 

dependent on spatial and temporal scale and 

that important processes at one scale are not 

necessarily transferable to other scales. Our 

analysis suggests that the vast majority of the 

data available on fire is either conducted at 

scales too small to be relevant to management 

(< 1 ha) or, in some regions, based on wildfire. 

Most studies were limited to short-term 

responses (< 3 yr) and often a single application 

of a fire, so they have minimal application 

to the long-term goal of restoring fire to 

...most 

herbaceous 

species recover 

within 2–3 

yr postfire 

regardless of 

season of the 

burn.”

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 95 
	 	 	 	

	 	

	 	

	 	 	

	 	

	 	 	

	 	 	

	 	 	

	 	

	 	

	 	

	 	 	

Experimental fire 

research rarely 

treats fire as a 

regime in which 

fire recurs and 

response to fire 

is dynamic and 

variable with 

fire intensity, 

fire interval, 

and other fire 

variables.”

rangeland ecosystems and strategic ecosystem 

management. 

Conducted largely as short-term studies on 

small plots, much of the research on prescribed 

burning is unable to describe complex patterns 

in space and time that may be associated with 

interactions with other disturbances, such as 

grazing and drought. Because these studies 

are but a step removed from highly controlled 

greenhouse studies with limited application 

in the real world, much of this research fits 

solidly within a “So what?” category when 

evaluated for specific management application. 

For example, a study conducted on vegetation 

response to burning that is conducted by mere 

coincidence in a drought and on small plots 

that were not grazed (e.g., Engle and Bultsma 

1984) cannot be used to support prescribed 

burning relative to vegetation responses 

across spatially variable, grazed pastures in 

nondrought periods. 

Long-term research at management-relevant 

scales that embrace interactive responses and 

complex patterns is insufficient to provide the 

NRCS with data necessary for constructing 

the relevant purposes for prescribed burning. 

The synergistic effects of fire and grazing on 

large landscapes have largely been uncoupled 

within rangeland research and conservation, 

even though most are aware that grazing by 

native and domestic herbivores is a dominant 

feature on all rangelands. An example of the 

uncoupling of fire and grazing is the tendency 

to recommend removal of grazing following 

fire, which does not seem to be supported by 

research. Grazing was part of the experimental 

design of fire studies in a mere13.3% of the 

studies and most of the studies inadequately 

presented the methods or results to provide 

effective conclusions regarding the presence or 

absence of grazing. Fire and grazing operated 

historically as an interactive disturbance 

process in which neither was independent of 

the other. When allowed to interact in space 

and time, fire and grazing create a shifting 

mosaic pattern that cannot be predicted from 

short-term, small-scale studies (Fuhlendorf and 

Engle 2001, 2004). Understanding the effects 

of a fire on grassland soils is highly dependent 

upon grazing (Hobbs et al. 1991). Moreover, 

nonequilibrial dynamics and resilience theory 

predict that episodic events, such as drought, 

can interact with other processes, such as fire 

and grazing, to promote changes that may 

not be predictable from understanding these 

events independently. Finally, most research 

on burning compares two treatments, fire 

and no fire (or burned and unburned), and 

the fire treatment is a single event rather than 

reoccurring fire couched within a complex fire 

regime. 

Much of the research also lacks relevance to 

prescribed burning as a conservation practice 

because it fails to account for the potential 

for fire effects to be highly dynamic and 

variable with time since the previous fire. Only 

4% of the papers included time since fire as 

an important variable when describing the 

magnitude and persistence of fire effects. This 

synthesis indicates that vegetation response 1–3 

yr since the previous fire differs considerably 

from vegetation response in which the most 

recent fire was 4 yr ago or longer. Moreover, 

recovery rate varies greatly among response 

variables. Recovery of soil and water variables 

can be as little as several months to as much 

as decades depending on factors such as soil 

structure, vegetation condition at the time 

of the fire, and fire intensity. Fire frequency 

also compounds temporal response to fire, 

but it was a primary focus in only 12.5% of 

the studies. This further emphasizes that the 

research largely fails to assess dynamic fire 

regimes and the long-term dynamics of firedependent ecosystems. 

Experimental fire research rarely treats fire as 

a regime in which fire recurs and response to 

fire is dynamic and variable with fire intensity, 

fire interval, and other fire variables. Research 

and management often approach fire as a single 

discrete event, so the impact of fire is highly 

dependent upon the conditions at the initiation 

of fire, conditions following fire, fire season, fire 

intensity, and time since fire. Research on fire 

regime rather than on discrete fire events would 

be more comparable to the study of grazing 

systems or constant stocking rates rather than 

the study of a single plant defoliation by a 

herbivore. Land grant institutions and federal 

agencies have been conducting research on 

grazing management for the past 50–100 

yr and many of the methods have become 

standardized (which is not always positive). 

Fire research, on the other hand, has lagged 

96 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 		

	 	 	 	

	 	 	 	 	

		 	 	

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

greatly, and has increased meaningfully in only 

the past decade. The lag in research is largely 

due to limited recognition of the importance 

of fire in a grazing-centric discipline and the 

concomitant limitation in research funding 

(partially alleviated by the Joint Fire Science 

Program [JFSP]). Our synthesis suggests 

that relevance of fire regime research for 

management goals continues to lag behind 

grazing regime research. 

The use of fire to improve wildlife habitat is 

a complex issue that is not easily evaluated 

because some species respond positively and 

others respond negatively to prescribed fire. 

Therefore, fire-improved habitat for one species 

likely translates into fire-degraded habitat 

for another. Groups of organisms that are 

negatively influenced often recover rapidly 

unless they occur in a highly fragmented 

landscape where dispersal from unburned areas 

is limited. These conclusions have led some 

authors to suggest that heterogeneity should 

be promoted to maintain a shifting mosaic 

landscape so that the entire landscape is not 

burned or unburned at any point in time, 

but research is lacking to support this across 

most rangeland types. Moreover, research 

on wildlife response to prescribed burning 

has preferentially focused on birds and small 

mammals. This likely reflects the fact that these 

species groups are more sensitive to fire-altered 

habitat than most other wildlife, and that these 

species are more easily studied. Consequently, 

specific species and not merely species groups 

must be identified when developing burning 

programs to support wildlife. 

The responses of soils and hydrology to fire 

are highly variable, but water repellency (i.e., 

hydrophobic soil) is a negative effect of extreme 

soil heating that occurs mostly under intense 

wildfire where fire has been absent for many 

years or in ecosystems with less evolutionary 

importance of fire. In general, fire increased 

water repellency, runoff, and sediment loads 

and decreased infiltration and water quality. 

Most of these effects disappeared after 2–4 

Prescribed fire on sand sagebrush rangelands in Oklahoma 

that was intended to create 

heterogeneity for wildlife, variable grazing distribution, and 

control the invasion of Eastern 

Redcedar. (Photo: John Weir) 

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 97 
	 	 	 	

	 	 	 	

	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	

Landscape photo of northern 

British Columbia demonstrating 

a shifting mosaic of fire patterns. Vegetation patterns are 

all due to variable times since 

fire resulting in highly variable 

plant communities. (Photo: Sam 

Fuhlendorf) 

yr. Variables that influenced the effects of 

fire on hydrology were aspect, fire severity, 

and microsite characteristics. Soil chemistry 

is also highly variable and dependent on the 

ecosystem studied, pre- and postfire conditions, 

invasive species, and grazing. Some fire 

prescriptions, wildfires, fire-return intervals, 

and fuel situations in rangeland can result 

in extreme soil heating, which can markedly 

change soil chemical and physical properties. 

Otherwise, fire events corresponding to the 

evolutionary fire regime have short-lived and 

slight influence on soil chemistry. 

2. 

reCoMMendAtIons 

1. The need for historic fire regimes to 

maintain the structure and composition 

of historic plant communities is well 

supported by ample scientific evidence. 

Many of the purposes in this conservation 

standard would benefit from the 

incorporation of more-specific goals and 

outcomes that can be more effectively 

compared to and supported by evidence 

in the peer-reviewed literature. Refer to 

Table 7 for evidence provided by the peer 

reviewed scientific literature supporting 

the current NRCS purposes for prescribed 

burning. 

Conservation outcomes of prescribed 

burning are most likely to be attained 

when the specific purposes for prescribed 

burning are tailored to the unique 

characteristics of the ecosystems being 

managed. Highly generalized purposes 

are necessary for the establishment of 

national guidelines, but they may often 

be misleading because ecological processes 

such as seed production, seed germination, 

plant mortality, etc. are likely to be highly 

variable among ecoregions and even within 

communities and soils within ecoregions. 

Even widely accepted generalizations, such 

as the NRCS purpose that fire can be used 

to control undesirable vegetation, carry 

caveats and exceptions when details are 

considered. 

98 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

3.	� Fire effects on ecosystems are often 

considered to be static over time, even 

though the research literature indicates 

that fire effects vary with time since 

fire and time between successive fires. 

Therefore, conservation purposes need 

to incorporate temporal dynamics 

to the extent that this information is 

available. Rangeland ecosystems evolved 

under specific fire regimes rather than 

in response to individual fires, which 

requires that conservation programs 

incorporate comprehensive fire regimes 

that address both short- and long-term 

outcomes. Ecosystem responses to fire 

and the effects of fire are both strongly 

influenced by temporal scale and must 

be carefully considered in conservation 

planning. Ecological site descriptions 

and rangeland research suggest that the 

prescribed burning standard should elevate 

application (i.e., area, number of plans, 

etc.) of prescribed burning so that it 

receives as much or more emphasis as the 

application of the conservation standard 

for grazing. There is no single practice as 

important to the maintenance of rangeland 

ecosystems. 

4.	� Given the foregoing cautions against 

nationally generalized purposes, the 

following purposes for the practice of 

prescribed burning are supported by the 

research literature. These purposes should 

consider the context of the fire regime, 

rather than a single fire in isolation: 

a.	� to alter plant composition, reduce 

undesirable herbaceous plants, and 

reduce accumulated litter for a short 

period of time (generally 1 yr); 

b.	�to improve forage quality for < 3 yr; 

c.	� to limit encroachment of fire-sensitive 

shrubs and trees; 

d.	�to manage stature of resprouting, fireadapted shrubs and trees; 

e.	� to alter distribution of grazing and 

browsing to either promote uniform 

distribution (by spatially homogenizing 

attractiveness) or to promote 

heterogeneity for biodiversity; 

f.	� to reduce potential for high intensity, 

stand-replacing fire by reducing 

accumulated fine fuel; 

g.	� to maintain grassland, savanna, and 

shrubland ecosystem structure, i.e., to 

prevent transitioning to woodland; and 

h.	�to recognize that mosaics of plant 

communities that vary with time since 

fire are critical for wildlife diversity. 

KnoWLedge gAPs 

1.	� Fire is as critical as climate and soils to the 

maintenance of ecosystem structure and 

function in many systems, but only limited 

experimental evidence exists to support 

the specific NRCS purposes for prescribed 

burning, especially those that involving 

long fire-return intervals. 

2.	� The experimental literature supporting 

prescribed burning is in need of greater 

managerial relevance that can be obtained 

by directly addressing spatial scale, 

temporal scale, and interaction with 

other disturbances, including drought 

and grazing. Reliance on information 

resulting from single fires applied on small 

plots tracked for a relatively short time 

interval greatly constrains inferences and 

application to ecosystem management and 

this information should be applied with 

caution. 

3.	� Knowledge of smoke characteristics and 

smoke management, effects of fire on 

wildlife and insects, and fire behavior and 

fire management exist in some regions, but 

are limited in other regions. The lack of 

sufficiently meaningful data on these topics 

makes it difficult to inform prescribed 

burning practices at the national level. 

ConCLusIon 

The vast majority of scientific evidence 

addressing fire in rangeland ecosystems points 

to the value of and need to continue or restore 

fire regimes with conservation programs. This 

is especially relevant given the accelerating rate 

of woody plant encroachment in grasslands 

and savannas, but it is important to other 

conservation outcomes, including altering 

grazing behavior of ungulates and maintaining 

biodiversity. The incorporation of prescribed 

burning conservation programs must include 

an understanding of the dynamic role that 

fire plays in most rangeland ecosystems. Fire 

regimes are equivalent to soils and climate in 

terms of their influence on plant community 

CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 99 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	

	 	

Fire in a mixed grass prairie. 

(Photo: Steve Winter) 

composition and landscape patterns. 

Recognition of fire as a dynamic regime 

that has variable effects depending on the 

interaction with grazing and climate patterns is 

critical to optimizing conservation effectiveness 

in rangeland ecosystems. 

The complex interaction of scientific 

knowledge, social concerns, and variable 

policies across regions are major limitations 

to the successful and critical restoration of fire 

regimes. Successful grassroots actions that have 

led to increased use of prescribed fire include 

the development of prescribed fire cooperatives 

through many of the Great Plains states. 

These cooperatives build on regional strengths 

to bring landowners together to conduct 

prescribed fires on landscapes that have variable 

ownerships. These cooperatives have enabled 

landowners to overcome issues associated with 

labor, liability, and training and to restore 

fire regimes in regions that have had fire 

removed for more than a century. Landowner 

cooperatives have the potential to transform the 

application of fire and indicate that successful 

conservation practices based on fire are possible 

even in areas that do not have a history of 

prescribed burning. 

To address the shortfalls in research applicable 

to prescribed burning on rangelands and the 

limited application of prescribed burning on 

rangelands, we recommend that the NRCS 

position itself to drive rangeland research and 

fire research agendas. Research to support 

NRCS purposes for prescribed burning on 

rangeland, unlike forests, has been limited 

by insufficient funding. However, a NRCSdriven research agenda also is lacking because 

NRCS has been detached from federal fire 

programs, notably LANDFIRE and the JFSP. 

Involvement in these programs would afford 

NRCS a greater opportunity to engage the 

scientific community and interact with other 

federal agencies that are developing valuable fire 

management tools and promoting fire research. 

This would provide NRCS with a platform 

for developing a research agenda relevant to 

supporting fire management on private lands. 

LANDFIRE is an interagency vegetation, fire, 

and fuel characteristics mapping project (www. 

landfire.gov). Developed through cooperation 

of natural resource agencies other than NRCS 

(Bureau of Land Management, US Forest 

Service, National Park Service, etc.) and 

The Nature Conservancy, LANDFIRE was 

initiated by a request from federal agencies to 

develop maps needed to help land managers 

prioritize areas for hazardous fuel reduction 

and conservation actions. LANDFIRE has 

spatial data layers that include all layers 

required to run fire modeling applications 

such as FARSITE and FlamMap, existing 

vegetation type, canopy height, biophysical 

setting, environmental site potential, and 

fire regime condition class, as well as fire 

effects layers. Since its initiation, LANDFIRE 

has been expanded to address the entire 

United States, including private as well as 

public land, but it also has some data needs. 

Vegetation dynamics models that operate 

with Vegetation Dynamics Development 

Tool software form a major component of 

LANDFIRE. These state-and-transition 

models are similar to those being developed 

by NRCS in that they describe pathways of 

vegetation succession and the frequency and 

effects of disturbances; however, these models 

are quantitative and based on extensive field 

analyses and modeling. Currently, rangelands 

are underrepresented and could be enhanced 

by involvement of NRCS in this national fire 

effort. We recommend that the NRCS engage 

the development and use of decision tools 

like LANDFIRE and OK-Fire (2009) that 

facilitate management and application of fire 

on rangelands. 

100 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

S. D. Fuhlendorf, R. F. Limb, D. M. Engle, and R. F. Miller 

The JFSP was created by Congress in 1998 

as an interagency research, development, 

and applications partnership between the 

US Department of the Interior and the US 

Department of Agriculture–Forest Service 

(http://www.firescience.gov/index.cfm). 

Funding priorities and policies are set by 

the JFSP Governing Board, which includes 

representatives from the Bureau of Land 

Management, National Park Service, US 

Fish and Wildlife Service, Bureau of Indian 

Affairs, and US Geological Survey, as well as 

five representatives from the Forest Service. 

This program funds applied research focused 

on the management of fire for natural resource 

managers. NRCS should work to become a 

JFSP partner to support research applicable 

to privately owned rangelands. This would 

provide NRCS with a platform for developing 

a research agenda relevant to supporting fire 

management on private lands. 

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White-tailed deer distribution in response to 

patch burning on rangeland. Journal of Arid 

Environments 72:2026–2033. 

Mehringer, P. 1987. Late Holocene environments 

on the northern periphery of the Great 

Basin. Portland, OR, USA: Bureau of Land 

Management Final Report. Contract YA551-

CTR-340075. 71p. 

Mensing, S., S. Livingston, and P. Barker. 

2006. Long-term fire history in Great Basin 

sagebrush reconstruction from macroscopic 

charcoal in spring sediments in Newark Valley, 

Nevada. Western North American Naturalist 

66:64–77. 

Miller, R. F., S. T. Knick, D. A. Pyke, C. W. 

Meinke, S. E. Hanser, M. J. Wisdom, and 

A. L. Hild. 2011. Characteristics of sagebrush 

habitats and key limitations to long term 

conservation. Studies in Avian Biology (in press). 

Morgan, P., C. C. Hardy, T. W. Swetnam, R. G. 

Rollins, and D. G. Long. 2001. Mapping fire 

regimes across time and space: understanding 

coarse and fine-scale fire patterns. International 

Journal of Wildland Fire 10:329–342. 

Neary, D. G., C. C. Klopatek, L. F. DeBano, 

and P. F. Ffolliott. 1999. Fire effects on 

belowground sustainability: a review and 

synthesis. Forest Ecology and Management 

122:51–71. 

Ojima, D. S., D. S. Schimel, W. J. Parton, 

and C. E. Owensby. 1994. Long-term and 

short-term effects of fire on nitrogen cycling in 

tallgrass prairie. Biogeochemistry 24:67–84. 

OK-Fire. 2009. Weather-based decision support 

products for wildland fire management in 

Oklahoma. Available at: http://okfire.mesonet. 

org/. Accessed September 2009. 

Owensby, C. E., and K. L. Anderson. 1967. 

Yield responses to time of burning in the Kansas 

Flint Hills. Journal of Range Management 20:12– 

16. 

Owensby, C. E., and J. B. Wyrill III. 1973. 

Effects of range burning on Kansas Flint Hills 

soil. Journal of Range Management 26:185–188. 

Panzer, R. 2002. Importance of in situ survival, 

recolonization, and habitat gaps in the postfire 

recovery of fire-sensitive prairie insect species. 

Natural Areas Journal 23:14–21. 

Pedersen, E. K., J. W. Connelly, J. R. 

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Prater, M. R., and E. H. DeLucia. 2006. Nonnative grasses alter evapotranspiration and energy 

balance in Great Basin sagebrush communities. 

Agricultural and Forest Meteorology 139:154–163. 

Reich, P. B., P. Bakken, D. Carlson, L. E. 

Frelich, S. K. Friedman, and D. F. Grigal. 

2001. Influence of logging, fire and forest type 

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boreal forests. Ecology 82:2731–2748. 

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Anthony, and V. E. Marr. 2007. Avian 

community structure along a mountain big 

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Effects of burning on a desert grass–shrub range 

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ecology: United States and Canada. New York, 

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CHAPTER 

3 

Brush Management as a 

Rangeland Conservation Strategy: 

A Critical Evaluation 

Steven R. Archer,1 Kirk W. Davies,2 Timothy E. Fulbright,3 

Kirk C. McDaniel,4 Bradford P. Wilcox,5 and Katharine I. Predick6 

Authors are 1Professor, School of Natural Resources and the Environment, University 

of Arizona, Tucson, AZ 85721-0043, USA; 2Rangeland Scientist, USDA, Agricultural 

Research Service, Eastern Oregon Agricultural Research Center, Burns, OR 97720, 

USA; 3Regent's Professor and Meadows Professor in Semiarid Land Ecology, Caesar 

Kleberg Wildlife Research Institute, Texas A&M University–Kingsville, Kingsville, TX 

78363, USA; 4Professor, Animal and Range Science, New Mexico State University, 

Las Cruces, NM 88003, USA; 5Professor , Ecosystem Science and Management, 

Texas A&M University, College Station, TX 77843-2126, USA; and 

6Senior Research Specialist, School of Natural Resources and the Environment, 

University of Arizona, Tucson, AZ 85721-0043, USA 

Reference to any commercial product or service is made with the understanding 

that no discrimination is intended and no endorsement by USDA is implied 

105
 
	 	 	 	

	 	

	 	 	

	 	

	 	 	

	 	 	

	 	 	

	 	

Woody plant 

encroachment 

“

represents a threat 

to grassland, shrubsteppe, and savanna 

ecosystems and the 

plants and animals 

endemic to them...”

106 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	

	 	 	 	

	 	 	 		

	 	

3 Brush Management as a 

Rangeland Conservation Strategy: 

A Critical Evaluation 

Steven R. Archer, Kirk W. Davies, Timothy E. Fulbright, Kirk C. McDaniel, 

Bradford P. Wilcox, and Katharine I. Predick 

IntRoduCtIon 

Rangelands support the majority of the world’s 

livestock production (Safriel and Adeel 2005) 

and play an important role in human health 

and global carbon, water, and nitrogen cycles 

(Campbell and Stafford Smith 2000). Their 

extensive airsheds and watersheds provide 

habitat for game and nongame wildlife and 

myriad ecosystem goods and services important 

to rapidly growing settlements and cities that 

may be geographically distant. Rangelands 

thus have considerable, multidimensional 

conservation value. Stewardship of vegetation 

composition, cover, and production is 

the foundation of sustainable rangeland 

management, a key component of which is 

maintaining vegetation within a desirable mix 

of herbaceous and woody plants (WPs). 

One of the most striking land cover changes 

on rangelands worldwide over the past 150 

yr has been the proliferation of trees and 

shrubs at the expense of perennial grasses. 

In some cases, native WPs are increasing 

in stature and density within their historic 

geographic ranges; in other cases, nonnative 

WPs are becoming dominant. These shifts in 

the balance between woody and herbaceous 

vegetation represent a fundamental alteration 

of habitat for animals (microbes, invertebrates, 

and vertebrates) and hence a marked 

alteration of ecosystem trophic structure. In 

arid and semiarid regions, increases in the 

abundance of xerophytic shrubs at the expense 

of mesophytic grasses represent a type of 

desertification (e.g., Schlesinger et al. 1990; 

Havstad et al. 2006) often accompanied by 

reductions in primary production (Knapp 

et al. 2008a) and accelerated rates of wind 

and water erosion (Wainwright et al. 2000; 

Gillette and Pitchford 2004; Breshears 

et al. 2009). In semiarid and subhumid 

areas, encroachment of shrubs and trees 

into grasslands and savannas may have 

neutral to substantially positive effects on 

primary production, nutrient cycling, and 

accumulation of soil organic matter (Archer 

et al. 2001; Knapp et al. 2008a; Barger et al. 

2011). While impacts of WP encroachment 

may vary among bioclimatic zones, there is 

one constant: grass-dominated ecosystems 

are transformed into shrublands, woodlands, 

or forest. As such, WP encroachment 

represents a threat to grassland, shrub-steppe, 

and savanna ecosystems and the plants and 

animals endemic to them, a threat on par 

with those posed by exurban and agricultural 

development (Sampson and Knopf 1994; 

Maestas et al. 2003). 

Efforts to counteract the real and perceived 

threats of WP encroachment fall into the 

broad category of brush management. Brush 

management, defined by the Natural Resource 

Conservation Service (NRCS 2003) as the 

removal, reduction, or manipulation of 

nonherbaceous plants, has been an integral 

component of range management since its 

formal emergence in the 1940s. However, 

brush removal has historically been criticized, 

especially when large-scale programs have failed 

to consider the needs of diverse stakeholders 

and the impact on multiple goods and services 

during planning and implementation stages 

(e.g., Klebenow 1969; Belsky 1996). 

Our goal here is to provide a contemporary, 

critical evaluation of “brush management” as a 

conservation tool. We begin with a brief review 

of potential drivers of WP encroachment. An 

understanding of these drivers will 1) shed 

light on the causes for the changes observed 

to date; 2) help us determine if management 

Woody plant encroachment has 

been widespread in rangelands, including these desert 

grasslands in New Mexico. 

(Photo: Paolo D'Odorico) 

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tABlE 1. Potential causes for increases in woody plant (WP) abundance in rangelands. There is likely no single-factor explanation for this 

widespread phenomenon. Most likely, it reflects drivers that vary locally or regionally or from the interactions of multiple drivers. Changes 

in a given driver may be necessary to tip the balance between woody and herbaceous vegetation but may not be sufficient unless cooccurring with changes in other drivers. For detailed reviews and discussions, see Archer (1994), Archer et al. (1995), Van Auken (2000), 

Briggs et al. (2005), and Naito and Cairns (2011). 

driver Mechanism Potential vegetation response 

Climate Increased precipitation Enhances WP establishment, growth, and density 

Decreased precipitation Promote shifts from mesophytic grasses to xerophytic shrubs 

Shift from winter to summer 

precipitation 

Favors WP over grasses, particularly on relatively deep, well-drained soils 

grazing Utilization of grasses by 

livestock 

Herbaceous production and species composition may shift to a community 

more susceptible to WP encroachment; livestock are effective agents of 

dispersal of some WP species; reductions in fine fuel mass and continuity 

(see “Fire”) Seed dispersal 

Browsing Reduced utilization of WPs by 

native herbivores 

Elimination of browsers promotes WP recruitment and growth; WPs kept 

small in size by browsers more susceptible to fire 

Fire Reduced fire frequency, 

intensity, and extent 

Increased WP recruitment and growth (see “Grazing”) 

Atmospheric Co2 Increased atmospheric CO2 

concentrations 

WPs with C3 photosynthetic pathway may be favored over grasses with C4 

photosynthetic pathway 

nitrogen deposition Increased N availability Correlated with forest expansion into grassland 

intervention is realistic; if so, 3) what 

approaches might be most effective; and 4) 

when, where, and under what conditions to 

apply them. We then discuss the ecological 

role of WPs in rangeland ecosystems and how 

human perspectives on WPs in rangelands 

influence management decisions and 

conservation objectives. The ecological impacts 

of WP proliferation are then reviewed with 

the aim of addressing the question, What 

are the environmental consequences of not 

managing WPs in rangelands? As it turns 

out, there are indeed consequences. Many 

of these have emerged relatively recently 

and hence are not yet reflected in current 

management guidelines. Advances in our 

understanding of the ecological consequences 

of WP proliferation in rangelands have 

paralleled changes in both perspectives on 

and approaches to brush management since 

the mid-1900s and have influenced how the 

NRCS has advised landowners. We therefore 

review the evolution of brush management in 

the spirit of putting current perspectives into 

their historical context. The basis for NRCS 

expectations underlying recommendations in 

the NRCS Brush Management Conservation 

Practice Standard matrix (hereafter described 

as “projected effects”) is then evaluated on 

the basis of a pooling of expectations into 

five overarching areas: herbaceous cover, 

production, and diversity; livestock response; 

watershed function; wildlife response; and fuels 

management. Evaluations are then followed by 

recommendations, an itemization of knowledge 

gaps, and a series of overarching conclusions. 

Why hAS WP ABundAnCE 

InCREASEd on RAngElAndS? 

Understanding the drivers of tree/shrub 

encroachment can help identify when, where, 

how, and under what conditions management 

might most effectively prevent or reverse WP 

proliferation. Traditional explanations center 

around intensification of livestock grazing, 

changes in climate and fire regimes, the 

introduction of nonnative woody species, and 

declines in the abundance of browsing animals 

(Table 1). Historical increases in atmospheric 

nitrogen deposition and atmospheric carbon 

dioxide concentration are also potentially 

important drivers. Exploring this important 

question is beyond the scope of this discussion, 

but detailed reviews and discussion can be 

found in Archer (1994), Archer et al. (1995), 

Van Auken (2000), Briggs et al. (2005), and 

Naito and Cairns (2011). Likely all these 

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S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

factors have interacted to varying degrees, and 

the strength and nature of these interactions 

likely varies from one biogeographic location 

to another. Thus, local knowledge is important 

in developing WP management plans. In many 

respects, WP encroachment is a specific case 

of weed and invasive plant management, and 

the concepts and principles developed for those 

perspectives are widely applicable (Sheley et al. 

this volume). 

It is important to note that once the process 

of WP encroachment is set in motion, grazing 

management per se may do little to prevent 

the conversion of grasslands and savannas to 

shrublands and woodlands (e.g., McClaran 

2003; Browning et al. 2008). In fact, on sites 

with a long history of heavy grazing, removal 

of livestock may actually promote rather than 

deter WP encroachment (Smeins and Merrill 

1988; Browning and Archer 2010). However, 

grazing management influences on WP 

encroachment are indirectly important in terms 

of how they affect the amount and continuity 

of fine fuels available for wildfire or prescribed 

burning (Fuhlendorf et al. 2008; Fuhlendorf et 

al. this volume). Because grazing management 

alone is generally not sufficient to curtail or 

reverse shrub encroachment, progressive brush 

management is a potentially important tool for 

grassland conservation. 

Although WP encroachment has been formally 

documented and qualitatively observed in 

some areas, it should not be assumed that 

this transformation has been uniform or 

ubiquitous. Indeed, repeat ground photography 

in western North America documents areas 

where WPs have dominated landscapes since 

the 1800s (e.g., Humphrey 1987; Turner et al. 

2003; Webb et al. 2007). Thus, many areas may 

have been historically comprised of mixtures 

of woody and herbaceous vegetation (e.g., 

shrub-steppe or shrub or tree savannas), and 

efforts to eradicate WPs from such sites may 

be misguided (McKell 1977) and sometimes 

detrimental to native plants and wildlife (e.g., 

Knick et al. 2003). 

PERSPECtIvES on WPs In 

RAngElAndS 

Brush management practices have historically 

focused on the goal of maximizing livestock 

production and promoting groundwater 

recharge and stream flow. Contemporary 

perspectives have been broadened to include 

impacts on biological diversity, ecosystem 

function (primary production and nutrient 

cycles), and land surface–atmosphere 

interactions (Appendix 2; Fig. 1). These 

broader perspectives are recognized to 

varying levels of specificity in NRCS Brush 

Management Conservation Practice Standards 

(code 314) and its projected effects. The 

current challenge lies with articulating these 

more explicitly in the CPPE worksheet, 

exposing landowners and the public to these 

perspectives, and articulating these perspectives 

in terms that can be quantified and objectively 

monitored and evaluated. 

tABlE 2. Perspectives on woody plants (WPs) in rangelands. In areas subject to 

heavy livestock grazing, palatable species typically give way to less palatable, less 

preferred species, and in rangelands, these less palatable species are often shrubs. 

The fact that unpalatable shrubs dominate many grazed rangelands has led to the 

mistaken generalization that all WPs in rangelands are undesirable. WPs have been 

typically viewed as the problem on grazed rangelands, but in fact they are likely a 

consequence of past mismanagement. Brush management conducted in isolation of 

grazing management is therefore treating symptoms rather than addressing the root 

causes of the problem (excessive grazing and fire suppression). When assessing 

whether to invest in efforts to reduce WP cover or density, the points shown in the 

table should be considered. For further discussion, see McKell (1977), Archer and 

Smeins (1993), and Archer (2009). 

• Palatable WPs may have been displaced along with palatable grasses and 

herbs (Lange and Willcocks 1980; Orodhu et al. 1990; Kay 1997). 

• Shrubs may decrease grazing pressure on grasses and provide protection for 

heavily utilized herbaceous species. 

• WPs may provide important habitat for a variety of vertebrate and invertebrate 

wildlife (nongame as well as game). 

• Shrubs may provide an important and underappreciated source of nutrimental 

stability and reduce supplemental feed requirements during cold or dry periods 

(Le Houérou 1980; Coppock et al. 1986; Stuth and Kamau 1990; Styles and 

Skinner 1997). 

• WPs may be the best adapted for the prevailing environmental conditions (Le 

Houérou 1994). 

• Were it not for the “damn brush,” there might be little or no vegetative cover. 

It may not be realistic to expect brush management to enhance herbaceous 

production, especially where soils have extensively eroded. 

• Will brush management stimulate herbaceous production and increase livestock 

carrying capacity sufficiently to offset treatment costs? If so, how much time will 

be required before a follow-up treatment? Will treating one problem perhaps 

create another (i.e., loss of valuable nontargeted species, invasion by weeds or 

exotic species, induced multiple-stemmed growth habit in shrubs, or replacement 

of nonsprouting species with sprouting species)? 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 109 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

FIguRE 1. Traditional and contemporary perspectives on woody plant (WP) encroachment (from Archer 2009). 

thE CoSt oF doIng nothIng 

Changes in WP cover and density represent 

fundamental changes in vegetation 

composition and structure and animal 

(microbes, invertebrates, and vertebrates) 

habitat. These, in turn, can fundamentally 

alter ecosystem primary production, trophic 

structure, biological diversity, nutrient cycling, 

and land surface–atmosphere interactions 

(Fig. 2). 

herbaceous Cover and Production 

The projected effects of brush management 

typically assume that herbaceous cover and 

production will increase following brush 

management. Implicit in this expectation 

is the assumption that WPs have a negative 

impact on ground cover. Does the literature 

support this perspective? The answer to this 

question is context dependent. Herbaceous 

cover and biomass typically decline as WP 

cover and basal area increase. However, the 

specific nature of the response ranges from 

an immediate linear or exponential decline 

to an initial stimulation, followed by a 

subsequent decline (Fig. 3; Table 3). The 

shape of these curves depends on the site and 

its grazing history, its climate, the physiology 

of the herbaceous vegetation (e.g., coolseason C3 vs. warm-season C4 grass), and 

the species of WP and its growth form (i.e., 

evergreen vs. deciduous), canopy architecture 

(i.e., single vs. multiple stemmed), size, 

density, and spatial arrangement (Jameson 

1967; Mitchell and Battling 1991; Scholes 

and Archer 1997; Scholes 2003; Fuhlendorf 

et al. 2008; Teague et al. 2008a). When 

stocking rates are based on total area rather 

than grazable area, WP encroachment can 

intensify grazing pressure to further depress 

grass production unless stocking rates are 

adjusted to compensate for WP-induced 

losses of forage production. WP impacts 

on herbaceous plants must therefore be 

considered in the context of livestock 

management (Briske et al. this volume) and 

the ecological site(s) being managed. 

It is important to note that relationships 

between WPs and grass are typically described 

at either the plant or the stand scale of spatial 

resolution. This can cause confusion and 

apparent contradictions. For example, velvet 

mesquite (Prosopis velutina) typically has 

neutral to positive effects on grasses at the 

plant scale but negative effects at the stand 

scale (McClaran and Angell 2007). Thus, care 

should be taken when generalizing results 

from a given study. Stand-scale assessments 

are generally most appropriate for pasture and 

landscape management. 

How do declines in herbaceous cover and 

biomass that typically accompany WP 

encroachment impact overall ecosystem 

primary production? A recent comparison 

of sites around North America suggests 

aboveground primary production declines with 

WP encroachment in hot and cold deserts 

but that it increases dramatically as a function 

of annual rainfall in semiarid and subhumid 

regions (Knapp et al. 2008a). Recent estimates 

suggest that for every millimeter increase in 

mean annual precipitation above 330 mm, 

aboveground net primary production (ANPP) 

will increase by ~0.6 g C·m−2·yr−1 with shrub 

encroachment (Barger et al. 2011). Thus, losses 

of grass production can lead to a net decline 

in overall ecosystem production in arid areas, 

whereas increases in production attributable 

to WPs more than compensates for declines 

in herbaceous production in other bioclimatic 

zones. 

110 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

FIguRE 2. Drivers of woody plant encroachment (see Text Box 1) and the potential consequences of ecosystem function and land surface– 

atmosphere interactions (from Archer 2009). 

Soil Condition and Erosion 

Projected effects generally assume that soil 

conditions and soil surface stability will be 

slightly to substantially improved by brush 

management and that soil erosion will be 

reduced by WP removal. Although not 

explicitly stated, these assumptions appear 

predicated on the expectation that WP 

proliferation adversely affects these parameters. 

Does WP encroachment lead to a deterioration 

of soil condition and site stability? 

Changes in grass and WP abundance impact 

soils through alteration of above- and 

belowground productivity, quality of litter 

inputs, rooting depth and distribution, 

hydrology, microclimate, and energy balance 

(Fig. 2). The abundance of soil organic matter 

or, more precisely, soil organic carbon (SOC) 

is a good indicator of soil condition, as it 

integrates a variety of ecosystem processes that 

influence fertility, water-holding capacity, and 

site stability. 

A substantial majority of the carbon in 

rangeland ecosystems resides in the SOC 

pool (Schlesinger 1997), but it is not yet clear 

how grazing, climate, and WP encroachment 

and “infilling” (shifts from relatively low to 

FIguRE 3. Potential responses of herbaceous vegetation to increases in woody plant cover or basal 

area. See Table 1 and reviews by Jameson (1967), 

Mitchell and Battling (1991), Scholes and Archer 

(1997), Scholes (2003), Fuhlendorf et al. (2008), 

and Teague et al. (2008a). 

relatively high WP cover or density) interact 

to affect gains and losses from these large 

carbon pools. Despite consistent increases 

in aboveground carbon storage with woody 

vegetation encroachment (Knapp et al. 2008a) 

and dryland afforestation (e.g., Nosetto et al. 

2006), the trends in SOC are highly variable, 

ranging from substantial losses to large gains to 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 111 
	 	 	 	

 

	 	 	

	 	 

	 	 

	 		 

	 	

	 	 

	 	 

	 	 

	 		 	 	 	 

	 	 	 

	 	 	

	 	

	 	 	 	 

	 	 	 	 	 	 	 	 

tABlE 3. Herbaceous response to shrub encroachment (US studies only). 

herbaceous 

dominant woody plant(s) responsea Soils MAP (mm) 

Big sagebrush (Artemisia tridentata) 

Broom snakeweed (Gutierrezia sarothrae ) 

Creosotebush (Larrea tridentata ) 

honey mesquite (Prosopis glandulosa ) 

huisache (Acacia farnesiana) 

Juniper Redberry juniper (J. pinchottii ) 

(Juniperus spp.) 

Redberry juniper (J. pinchottii ) 

Western juniper (J. occidentalis) 

Pine (Pinus spp.) Ponderosa pine (P. ponderosa ) 

Ponderosa pine (P. ponderosa ) 

Ponderosa pine (P. ponderosa) 

P. taeda, P. echinata 

longleaf pine (P. palustris ) 

Pinyon–juniper (Pinus edulis–Juniperus spp.) 

velvet mesquite (Prosopis velutina) 

D1 Aeolian sandy loams 

and loess (aridisols) 210 

A Loam/sandy loam 225–405 (eight sites) 

D2 Silt loam 406 

C — 279 

A Gravelly loam 323 (Vaughn) 

328 (Roswell) 

B Fine sandy loam 480 

B, D1 Sandy loam 221–430 (4 sites) 

D1 Gravelly/loamy 240 

A/B Shallow sandy 

C — 255 

A Gravelly — 

C Clay loam — 

A Sandy loan 231 

B Silt loam/clay loam 648 

B Loamy sand 231 

B Clay loams 665 

D2 Shallow clay 

C Sandy loam 345 

B (total); C	 

(cool-season grass) Clay 850 

B — — 

A (grazed); 

B (ungrazed) Fine loam and clay ~ 525 

B Loams with variable 

rockiness 410 

D2 Clay loam 430 

C — 248 

A — — 

A Limestone derived 560 

A — 380–510 (31 sites, 

varies with elevation) 

A — — 

B Silty loam ~ 584 

A — — 

D3 — 345 

D2 — ~ 330–432 (varies with 

elevation) 

C (one site); D2 

(three sites) — ~ 197–304 (varies with 

elevation) 

A — — 

a

 As per Figure 3 (A, negative exponential decline; B, linear decline; C, initial positive response, followed by decline; and “D,” “Other,” including no change (D1), increase (D2), 

or decline in C4, increase in C3 (D3). 

*Space-for-time substitution (sampling stands of different shrub abundance at one point in time). 

112 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	 	 		 	 	 	

	 	 	

	 	 	

	 	 	

		

	 	 	

	 	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

Scale (plant/ 

stand) Study duration (yr) Study location Reference 

Plant 28 (1950–1978) Southeastern ID Anderson and Holte (1981) 

Stand 20 Northwestern NM McDaniel et al. (2005) 

Stand 4 (1967–1970) Southern ID Hull and Klomp (1974) 

Stand 4 (1951–1954) OR Hyder and Sneva (1956) 

Stand 10 (1979–1989) Vaughn and Roswell, NM McDaniel et al. (1993) 

Stand 2 (1946–1947) TX Ueckert (1979) 

Stand 2* (1981–1982) Chihuahua, Mexico, and southern AZ Morton et al. (1990) 

Stand 14–17 (1984/1987–2001) Southwest NM Perkins et al. (2006) 

Stand 76 (1915–2001) Central NM Baez and Collins (2008) 

Stand Model South-central NM Bestelmeyer et al. (2009) 

Stand 5 (1970–1975) West TX Dahl et al. (1978) 

Stand 105 (1858–1963) South-central NM Buffington and Herbel (1965) 

Stand 6 (1995–2001) North-central TX Teague et al. (2008a) 

Stand 45 (1935–1980) Southern NM Hennessy et al. (1983) 

Stand 1*1998 North-central TX Hughes et al. (2006) 

Stand 46 (1957–2003) Southern AZ McClaran and Angell (2006) 

stand 2 (1978–1979) South TX Scifres et al. (1982) 

Stand 1*1995 Nolan County, TX Johnson et al. (1999) 

Stand 

3 (1984–1986) Western TX McPherson and Wright (1990) 

Stand 

2* (2005–2006) Northeastern CA Coultrap et al. (2008) 

Stand 7 (1975–1982) CA Evans and Young (1985) 

Stand 13 (1991–2004) Southeastern OR Bates et al. (2005) 

Tree 1* Northern AZ Jameson (1967) 

Stand 1* (1988) Northern AZ Moore and Deiter (1992) 

Stand 

1* (1984) Rocky Mountain front range, CO 

and WY Mitchell and Battling (1991) 

Stand 2* (1960–1961) Eastern TX Halls and Schuster (1965) 

Stand 11 Alexandria, LA Grelen and Lohrey (1978) 

Stand 1* Northern and central AZ Jameson (1967) 

Stand 2 (1984–1985) South-central NM Pieper (1990) 

Stand 10 (nonsequential; 

1954–1967) Southern AZ Cable (1971) 

Stand 5 (1945–1950) Southern AZ Parker and Martin (1952) 

Stand — Southeastern AZ Upson et al. (1937) 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 113 
	 	 	 	

 

FIguRE 4. Changes in soil organic pools following shrub encroachment in 

different precipitation zones (from Asner and Archer 2009). Numbers on 

symbols/legend entries indicate references listed in Appendix III. 

no net change (Wessman et al. 2004; Asner and 

Archer 2009; Fig. 4). 

Variation in SOC response to WP 

encroachment is perhaps not unexpected 

given the myriad factors that influence SOC 

pool and fluxes (Wheeler et al. 2007). These 

include growth characteristics of the WPs 

(e.g., evergreen or deciduous, N fixing or not, 

shallow or deep rooted, etc.), climate (mean 

annual rainfall and temperature), soil properties 

(e.g., texture, pH, carbonate content), initial 

conditions (e.g., amount, type, and distribution 

of SOC present at the time WP encroachment 

begins), and prior land management (e.g., 

whether WPs are establishing in native 

rangeland vs. abandoned cropland). In areas 

where shrub-induced increases in SOC have 

been documented, changes are typically 

restricted to the upper 10–20 cm of the soil 

profile, and accumulation appears to be a 

linear function of time since WP establishment 

(Boutton et al. 2009) with rates ranging from 

8 g C·m−2·yr−1 to 30 g C·m−2·yr−1 (Wheeler 

et al. 2007). Some of the uncertainty in 

SOC response may reflect the fact that WP 

encroachment often occurs in areas with a 

history of livestock grazing, which itself has 

positive, neutral, and negative effects on SOC 

pools (Milchunas and Lauenroth 1993; Piñeiro 

et al. 2010; Briske et al. this volume). Where 

historical grazing and WP encroachment effects 

on SOC have been explicitly accounted for, 

it appears that losses of SOC associated with 

heavy grazing can be recovered subsequent to 

WP encroachment and that SOC in the shrubdominated system can be substantially greater 

than that of the original grasslands (Archer et 

al. 2001; Hibbard et al. 2003). 

SOC and N levels are typically highly 

correlated; hence, increases in SOC are 

typically accompanied by increases in soil N 

(Seastedt 1995; Wheeler et al. 2007). When 

the encroaching woody species is a nitrogen 

fixer, soil N levels can increase substantially 

(Geesing et al. 2000; Hughes et al. 2006), thus 

augmenting a key resource that, along with 

water, typically colimits rangeland productivity. 

The resultant increase in soil fertility and waterholding capacity likely drives the increase in 

herbaceous production that typically follows 

WP removal. 

In arid regions, the loss of grass cover due 

to grazing is accompanied by loss and 

redistribution of soil resources from plant 

interspaces to areas beneath shrubs. Many 

studies have investigated this grass-erosion 

feedback, with the consensus that erosion 

by wind and water is capable of removing 

soil resources required for grass growth and 

propagation while creating semipermanent 

fertile islands beneath shrub canopies (see 

Okin et al. 2009). The net result is a dramatic 

increase in wind and erosion resulting from 

increased bare areas in shrublands compared to 

the grasslands they replaced. Aeolian sediment 

flux in mesquite-dominated shrublands in 

the Chihuahuan Desert are 10-fold greater 

than rates of wind erosion and dust emission 

from grasslands on similar soils (Gillette and 

Pitchford 2004). Flow and erosion plots in 

the Walnut Gulch Experimental Watershed 

in Arizona and the Jornada Long Term 

Ecological Research site in New Mexico have 

demonstrated significant differences in water 

erosion between grasslands and shrublands 

(Wainwright et al. 2000). For example, higher 

splash detachment rates (Parsons et al. 1991, 

1994) and interrill erosion rates (Abrahams et 

al. 1988) are observed in shrublands compared 

to grasslands, and shrubland areas are more 

prone to develop rills, which are responsible 

for significant increases in overall erosion 

rates (Luk et al. 1993). Episodes of water 

114 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

	 	

	 	

	 	

	 	

	 	 	

	 	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	 	 	

	 	

	 	 	 	

	 	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

erosion are often associated with decadal 

drought–interdrought cycles because depressed 

vegetation cover at the end of the drought 

makes the ecosystem vulnerable to increased 

erosion when rains return (McAuliffe et al. 

2006). In hot desert systems where shrub 

encroachment has occurred, reestablishment 

of grass cover would help curtail erosion losses. 

However, the loss of topsoil to date, coupled 

with low and highly variable precipitation, 

make these among the most challenging 

environments in which to reestablish perennial 

grass cover once it has been lost (see the 

section “Herbaceous Vegetation and Native 

Communities”). 

Air Quality and land Surface– 

Atmosphere Interactions 

Brush management impacts on air quality 

are treated as “not applicable” or “neutral” 

with respect to particulate matter, ozone, and 

greenhouse gas production. However, these 

factors and others related to pollen production 

and land surface–atmosphere interactions may 

warrant more attention in the next generation 

of projected effects worksheets. 

A synthesis of aeolian sediment transport 

studies spanning a grassland–forest continuum 

suggests 1) that among relatively undisturbed 

ecosystems, arid shrublands have inherently 

greater aeolian transport because of wake 

interference flow associated with intermediate 

levels of density and spacing of WPs and 2) that 

among disturbed ecosystems, the upper bound 

for aeolian transport decreases as a function of 

increasing amounts of WP cover because of the 

effects of the height and density of the canopy 

on airflow patterns and ground cover associated 

with WP cover (Breshears et al. 2009). 

Pollen from WPs trigger nasal allergies and 

asthma (Chang 1993; Gutman and Bush 

1993). Tree/shrub proliferation thus has the 

potential to influence the onset, duration, 

concentration, and total production of 

pollen allergens both locally and at great 

distances (Levetin 1998). However, the role 

of these allergens on human health is not well 

understood (Al-Frayh et al. 1999). 

Climate and atmospheric chemistry are 

directly and indirectly influenced by land cover 

via biophysical and biogeochemical aspects 

of land surface–atmosphere interactions. 

Shifts from grass to WP domination have 

the potential to influence biophysical aspects 

of land–atmosphere interactions related to 

albedo, evapotranspiration, surface roughness, 

boundary layer conditions, and dust loading 

that affect cloud formation and rainfall (Figs. 

1 and 2). Increases in C and N pools that 

occur when WPs proliferate in grasslands and 

savannas may be accompanied by increases 

in trace gas emissions (e.g., carbon dioxide, 

nitrous oxide, and methane; McCulley et al. 

2004; Sponseller 2007; McLain et al. 2008) 

and nonmethane hydrocarbon emissions 

(Monson et al. 1991; Guenther et al. 1995; 

Klinger et al. 1998; Geron et al. 2006). 

Emissions of such compounds can influence 

atmospheric oxidizing capacity, heat retention 

capacity, greenhouse gas half-life, aerosol 

burdens, and radiative properties. As a result, 

air quality (Monson et al. 1991) and energy 

balance can be affected. 

WP encroachment has been accompanied 

by increased dust production in arid regions 

(Gillette and Pitchford 2004). Dust can 

potentially influence weather and climate by 

scattering and absorbing sunlight and affecting 

cloud properties, though the overall effect 

of mineral dust in the atmosphere is likely 

Woody plant 

encroachment 

has been 

accompanied by 

increased dust 

production in arid 

regions”

Changes in vegetation have 

the potential to influence 

biophysical aspects of land– 

atmosphere interactions via 

effects on trace gas and 

hydrocarbon emissions, 

albedo, evapotranspiration, 

surface roughness, and dust 

loading. (Photo: Rich Reynolds) 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 115 
FIguRE 5. (A) Conceptual model of landscape-scale changes in ecosystem 

biodiversity (species, growth form, or structural) that potentially accompany 

woody plant (WP) proliferation in grasslands and savannas (from Archer 

2009). Plot-scale reductions in herbaceous species richness with increases 

in (B) juniper (Juniperus virginiana) density (10-m2 plots) and within Cornus 

drummondii shrub islands and surrounding grasslands (m2 plots; insert) in Tall 

Grass Prairie (Knapp et al. 2008b) and (C) with creosote bush encroachment 

in Desert Grasslands (Baez and Collins 2008). In A, diversity is predicted 

to increase during early stages of WP encroachment because of the mixture 

of woody and herbaceous floral/faunal elements. Maximum diversity might 

be expected in savanna-like configurations where woody and herbaceous 

plants co-occur. As WP abundance increases, loss of grassland components 

eventually occurs. In subtropical thorn woodland and dry forests with high WP 

species richness, a net increase in diversity may result. In other settings, there 

may be no net change in diversity, only a change in physiognomy. Where 

WPs form virtual monocultures with little or no understory (e.g., panels B and 

C), the loss of diversity may be profound. Regardless of the numerical changes 

in biodiversity, the existence of grassland and open savanna ecosystems 

and the plants and animals endemic to them are jeopardized with WP 

encroachment. 

small compared to other human impacts 

(Intergovernmental Panel on Climate Change 

2007). However, the mineral aerosols in dust 

originating from drylands are thought to 

play a major role in ocean fertilization and 

CO2

 uptake (Blain et al. 2007), terrestrial soil 

formation and nutrient cycling (Chadwick et 

al. 1999; Neff et al. 2008), and public health 

(e.g., Mohamed and El Bassouni 2007). Dust 

deposition also decreases albedo of alpine 

snowpack, thus accelerating melt and reducing 

snow-cover duration (Painter et al. 2007). In 

arid regions, erosion increases sediment delivery 

and changes flow conditions of rivers (Jepsen 

et al. 2003) and impacts water quality, riparian 

vegetation, aquatic fauna (Cowley 2006), and 

soil fertility and ecosystem processes (Valentin 

et al. 2005; Okin et al. 2006). Thus, the 

replacement of grasslands by shrublands in arid 

areas has potentially far-reaching ramifications. 

Modified land cover can affect weather and 

climate (Bryant et al. 1990; Pielke et al. 1998). 

Changes in vegetation height and patchiness 

that occur when WPs replace grasses over large 

areas affect boundary layer conditions and 

aerodynamic roughness; changes in leaf area 

and rooting depth alter inputs of water vapor 

via transpiration; and changes in fractional 

ground cover, phenology, leaf habit (e.g., 

evergreen vs. deciduous), albedo, and soil 

temperature influence evaporation and latent 

and sensible heat exchange (Fig. 2; e.g., Graetz 

1991; Bonan 2002). The extent to which these 

changes in structure influence meteorological 

conditions likely vary with annual rainfall 

(e.g., via leaf area changes accompanying shrub 

encroachment [Knapp et al. 2008a]), soil 

texture, shrub rooting depth, and proximity to 

water tables (Jobbaggy and Jackson 2004). 

Effects of WP encroachment on mesoscale 

climate and local weather have not been 

assessed. However, evidence from tree-clearing 

studies suggest that decreases in WP cover 

can potentially influence evapotranspiration, 

the incidence of convective storms, and 

cloud formation (Jackson et al. 2007). Model 

simulations in tropical savannas indicate that 

clearing of woody vegetation could increase 

mean surface air temperatures and wind speeds, 

decrease precipitation and humidity, and 

increase the frequency of dry periods within the 

wet season (Hoffman and Jackson 2000). Thus, 

116 Conservation	Benefits	of	Rangeland	Practices
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	

	 	

	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	 	

	 	 	

	 	

	 	

	 	 	

	 	

	 	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

by extension, we might expect increases in WP 

abundance to have the reverse effect on local 

weather and climate. 

Biodiversity 

Effects of WP encroachment on biodiversity, 

whether quantified as the genetic diversity of 

populations, species richness, or the number 

of plant functional groups or animal guilds 

represented in an area, have not been widely 

quantified. At the landscape scale, colonization 

of grasslands by WPs initially represents 

new species additions and hence promotes 

biodiversity, and shrub modification of soil 

properties, vertical vegetation structure, and 

microclimate may subsequently promote the 

ingress and establishment of other plant and 

animal species (Fig. 5A). In its early stages, 

WP encroachment may have a multiplier 

effect on animal diversity by adding keystone 

structures and habitat heterogeneity (Tews et al. 

2004b) and providing nesting, perching, and 

foraging sites and shelter against predators and 

extreme climatic conditions (Whitford 1997; 

Cooper and Whiting 2000; Valone and Sauter 

2005; Blaum et al. 2007a). Indeed, numerous 

reptiles, birds, and mammals appear to prefer 

heterogeneous grass-dominated landscapes 

where scattered WPs provide up to 15% cover 

(Solbrig et al. 1996; Meik et al. 2002; Eccard et 

al. 2004; Bock et al. 2006; Thiele et al. 2008). 

In arid savanna rangelands, the diversity of 

small carnivores and their prey peaks at about 

10–15% shrub cover (Blaum et al. 2007d). In 

the Chihuahuan Desert, shrub-invaded sites 

harbor four times the number of ant forager 

species found at a relatively pristine desert 

grassland site, suggesting that ant diversity is 

enhanced by shrub invasion and that several 

taxa benefit from it (Bestelmeyer 2005). The 

effects of WP encroachment vary among animal 

taxa and functional groups (e.g., Kazmaier et al. 

2001), but as WP cover increases and habitat 

characteristics continue to shift, shrubland/ 

woodland-adapted species are expected to 

become favored over grassland-adapted species. 

Grassland-obligate plants and animals may be 

affected immediately and negatively by WP 

encroachment (Table 4). Even so, diversity may 

be maintained or enhanced if new species cooccur with the more broadly adapted original 

species and if the displacement of grasslandobligate species is more than offset by the 

arrival of new species (e.g., Sauer et al. 1999; 

Blaum et al. 2007b, 2007c). As WP cover and 

biomass continue to increase, the end result 

may be an overall gain in diversity, no net 

change in diversity, or a net loss in diversity 

(Fig. 5A). Qualitative observations suggest 

that tropical and subtropical grasslands may 

potentially experience net gains in diversity 

with WP encroachment because of large 

pools of tree and shrub species, large pools 

of herbaceous species capable of coexisting 

with WPs, and large pools of invertebrates 

Grasslandobligate plants 

and animals 

may be affected 

immediately and 

negatively by WP 

encroachment”

tABlE 4. Avifauna and woody plant (WP) encroachment. Grassland passerines are declining at a faster 

rate than any other bird group in North America (Peterjohn and Sauer 1999). Woody plant encroachment associated with livestock grazing is among the contributing factors (Bakker 2003; Brennan and 

Kuvlesky 2005). 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 117 

vegetation change Effects on grassland avifauna Citation(s) 

thresholds of WP cover 

and height exceeded 

Reduction in suitable habitat Lloyd et al. (1998); Grant et al. 

(2004); Gottschalk et al. (2007) 

Proximity to 

woodlands 

Decreased food abundance; 

increased predation and brood 

parasitism 

Johnson (2001); Bakker et al. 

(2002); Fletcher and Koford 

(2002); Thiele et al. (2008) 

grasslands converted 

to shrublands 

Increased overall avian species 

richness but declines in groundnesting passerine and gallinaceous 

species 

Whitford (1997); Pidgeon et al. 

(2001); Rosenstock and van Riper 

(2001); Fuhlendorf et al. (2002) 

Juniper encroachment 

into sagebrush–steppe 

communities 

Eliminates sagebrush-obligate 

species habitat 

Connelly et al. (2000); Miller 

et al. (2000); Crawford et al. 

(2004); Reinkensmeyer et al. 

(2007) 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

FIguRE 6. Undisturbed (left) and mowed (right) mountain big sagebrush (Artemisia 

tridentata vaseyana)–bunchgrass plant community in southeastern Oregon. (Photo: 

K. W. Davies) 

and passerine bird species. In other cases, the 

number of encroaching woody species may 

be very small and their traits detrimental to 

the persistence of other plant species. WP 

encroachment may then result in virtual 

monocultures of vegetation (Figs. 5B and 5C) 

with concomitant impacts on faunal diversity. 

Changes in aboveground biological 

diversity with WP proliferation may also be 

accompanied by changes in the diversity of 

microbial communities belowground. Shifts 

from bacterial to fungal populations may 

accompany shifts from herbaceous to woody 

domination (e.g., Imberger and Chiu 2001; 

Purohit et al. 2002), enabling the microbial 

biomass to effectively deal with lower litter 

quality and thus maintain or even increase soil 

respiration and mineralization. Aanderud et 

al. (2008) found differences in gram-positive 

bacteria, Actinobacteria, and fungi communities 

in soils below and between shrubs. Thus, 

changes in microbial communities would be 

expected to accompany changes in composition 

and abundance of shrubs. 

Parasitic nematodes and nematodes feeding on 

bacteria and fungi in the immediate vicinity 

of plant roots are indicator taxa for changes 

in belowground microbial communities. 

The maximum depth of occurrence of these 

organisms increased from 2.1 m in grasslands 

to 4.0 m in areas where WPs have replaced 

grasses, but the composition of the nematode 

food web at this depth was markedly reduced 

from five trophic groups to two (Jackson et 

al. 2002). Invaded woody sites also had lower 

species richness in soils due primarily to the 

loss of root feeding species. 

The conceptual model in Figure 5A is based 

on numerical assessments of species, functional 

group, and structural diversity. However, from 

the perspective of physiognomic diversity, WP 

encroachment is transformative. Grasslands 

become shrub or tree savannas, and shrub 

and tree savannas become shrublands or 

woodlands. Thus, even in cases where numerical 

diversity may be maintained or enriched by 

WP encroachment, there is a loss of grassland 

and savanna ecosystems and the plants and 

animals endemic to them. Thus, while brush 

management has historically been advocated 

from the perspective of potential benefits for 

livestock production and hydrology, it should 

also be considered from the perspective of 

maintaining the existence of grassland and 

savanna ecosystems. 

BRuSh MAnAgEMEnt: A BRIEF 

hIStoRy 

WP encroachment has long been of concern 

to rangeland managers (Leopold 1924). Thus, 

there is a long history of devising management 

tools for reducing WP cover. The basis for 

concerns over WP proliferation was historically 

centered around the adverse effects of shrubs on 

forage production (Fig. 6) and livestock safety 

(e.g., WPs as cover for predators), health (e.g., 

as habitat for insect and arthropod pests and 

parasites such as ticks and horn flies [Teel et al. 

1997]), and handling (difficulty in gathering 

and moving animals with increasing WP 

stature/cover/density). This traditional focus on 

rangeland value for livestock production was also 

the impetus for other management practices, 

such as efforts to eliminate competitors (e.g., 

certain predators, herbivores, and insects) viewed 

as directly or indirectly reducing ranch profits. 

In some cases, these wildlife may have played an 

important role in keeping WPs in check, and 

118 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	 	

	 	 	

	 	

	 	

	 	

	 	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

their systematic elimination may have opened 

the door for WPs to increase in abundance (e.g., 

prairie dogs; Weltzin et al. 1997). 

the 1940s and 1950s 

During the post–World War II era, heavy 

equipment and chemicals were readily 

available and were used on a broad scale. Our 

understanding of ecosystem processes and 

ecosystem goods and services was in its infancy 

during this period, and few environmental 

regulations were in place. Applied research 

in range science focused on the development 

and application of herbicides and mechanical 

techniques (Scifres 1980; Bovey 2001), often 

with the goal of eradicating shrubs. Brush 

management during this period was typically 

applied indiscriminately. 

the 1960s and 1970s 

Efforts aimed at widespread eradication in the 

1940s and 1950s gave way to efforts aimed 

at selective control and containment in the 

1960s and 1970s. By this time, it was clear 

that there were no “silver bullets” for brush 

management. Unlike many herbicide products 

available today that provide a long treatment 

life (15–50 yr; McDaniel et al. 2005; Perkins et 

al. 2006; Combs 2007), treatments in the past 

were relatively short lived (Jacoby et al. 1990a). 

Following chemical spraying, shrub cover often 

returned to pretreatment levels (or higher) 

within 5–15 yr. The necessity of retreating 

landscapes at relatively high frequencies 

made brush management nonsustainable and 

difficult to justify when cost often exceeded 

revenues generated from subsequent livestock 

production. 

Basic and applied research from the 1940s 

to the 1970s led to the realization that brush 

management practices: 

•	� were treating symptoms (the shrubs) rather 

than root causes of land cover change (e.g., 

disruption of historic grazing and fire 

regimes); 

•	� must be conducted in concert with 

progressive livestock grazing management; 

•	� when applied in an indiscriminant manner 

without careful planning, can 

•	� be detrimental to wildlife populations, 

•	� lead to homogenization and loss of 

biological diversity, 

•	� increase risks for catastrophic soil 

erosion and weed invasion, and 

•	� be too costly for a ranching enterprise, 

and 

•	� can be short lived, with shrubs 

reestablishing dominance in 5–10 yr. 

Collectively, these realizations led to the 

development of integrated brush management 

systems (IBMS) in the 1980s (Scifres et al. 

1985; Brock 1986; Hamilton et al. 2004). 

1980s to Present: IBMS 

IBMS are long-term planning processes that 

move away from a purely livestock production 

perspective and toward management of 

rangelands for multiple uses and values. The 

IBMS planning process begins by identifying 

management goals and objectives for a 

specific site and the surrounding management 

unit. These might include increasing forage 

production; maintaining or promoting 

suitable wildlife habitat; augmenting stream 

flow or groundwater recharge; controlling 

pests, pathogens, or invasive species; 

maintaining scenic value; reducing wildfire 

risk; or preserving grassland and savanna 

ecosystems. Specific objectives are refined 

on a comprehensive inventory of ecosystem 

components (plants, animals, and soils), 

projecting the responses of those components 

to brush treatment alternatives, and considering 

the effects of treatment alternatives on 

management goals on other sites (Hanselka 

et al. 1996). Brush management techniques 

(herbicidal, mechanical, and prescribed 

burning) differ with respect to environmental 

impacts, implementation costs, efficacies, and 

treatment longevities. Thus, the IBMS approach 

advocates consideration of the type and timing 

of a given brush management technology and 

makes explicit allowances for consideration of 

the type and timing of follow-up treatments. 

This, in turn, requires knowledge of how woody 

and herbaceous plants will respond and how 

climate, soils, topography, and livestock and 

wildlife management might mediate plant 

responses. The IBMS approach is therefore 

inherently interdisciplinary and dependent on 

the active collaboration of a diverse group of 

management professionals. 

Examples of the IBMS approach abound 

(Teague et al. 1997; Grant et al. 1999; Paynter 

IBMS are longterm planning 

processes that 

move away from 

a purely livestock 

production 

perspective 

and toward 

management of 

rangelands for 

multiple uses and 

values.”

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 119 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	

	 	

	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	

The type, timing, 

and sequencing 

of brush 

management are 

the keys to longterm success”

In areas where shrubs 

are dense, herbicides or 

mechanical treatments, such as 

this crawler tractor equipped 

with push blades and discs, 

may be initially required to 

open up areas to promote 

herbaceous production 

and enable the subsequent 

use of prescribed fire as a 

management tool. (Photo: Tim 

Fulbright) 

and Flanagan 2004; Ansley and Castellano 

2006; Ansley and Wiedemann 2008). Nearly 

every State Cooperative Extension Service in 

the western region offers a long list of guides 

and publications that address the IBMS 

approach and give specific recommendations 

for management of important WPs. There are 

expert system tools (e.g., PESTMAN 2009) 

available to assist land managers in selecting 

appropriate brush management practices and 

techniques (Hanselka et al. 1996). There are 

also special adaptations of the IBMS approach 

for small landholdings (McGinty and Ueckert 

2001). NRCS guidelines should promote the 

IBMS approach and explicitly utilize it in the 

conservation planning process. 

BRuSh MAnAgEMEnt: CuRREnt 

PERSPECtIvES 

Brush management efforts must be viewed as 

an integral part of the overall system for wise, 

efficient use and conservation of grasslands. 

Available brush management and conservation 

methods are complex tools, the effectiveness 

of which depends primarily upon the resource 

manager’s understanding of their proper 

application approached with consideration of 

all potential uses of rangeland. (Scifres et al. 

1983, p. 11) 

This statement, made over 25 yr ago, is still 

relevant today. This rationale is echoed in 

recent textbooks addressing brush management 

and ecosystem restoration (Valentine 1989; 

Whisenant 1999; Bovey 2001; Hamilton et al. 

2004). The NRCS (2006) currently recognizes 

six reasons for undertaking brush management: 

•	� Restore natural plant community balance 

•	� Create the desired plant community 

•	� Restore desired vegetative cover to protect 

soils, control erosion, reduce sediment, 

improve water quality, and enhance stream 

flow 

•	� Maintain or enhance wildlife habitat, 

including that associated with threatened 

and endangered species 

•	� Improve forage accessibility, quality, and 

quantity for livestock 

•	� Protect life and property from wildfire 

hazards 

A Texas survey found that the two primary 

goals of landowners investing in brush 

management were to increase forage 

production and to conserve water (Kreuter et 

al. 2005). Other reasons included improvement 

of aesthetic values, benefit the next generation, 

improve wildlife habitat, and improve real 

estate value. 

Short- and long-Range Planning Is 

Essential 

A well-thought-out, comprehensive resource 

management plan reflecting short-, medium-, 

and long-term goals and objectives should be 

in place before attempting brush management. 

120 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

Evaluating multiple scenarios is useful 

for explicitly assessing the advantages and 

disadvantages of several alternatives. Given 

that the type, timing, and sequencing of 

brush management are the keys to long-term 

success, management plans should identify 

long-term objectives and work to ensure that 

resources and commitments are in place at the 

right time. Programs should include explicit 

information on what constitutes success and, if 

feasible, should address alternatives to primary 

objectives. Objectives will vary with the specific 

needs of the landowner, the community, and 

context of the action. Well-defined short- and 

long-term objectives are critical to determining 

when, where, why, how, and under what 

conditions brush management should be 

undertaken. 

Considerable momentum is now building for 

landscape-scale IBMS projects. Such projects 

usually require close consultation, coordination, 

and cooperation among multiagencies and 

associated stakeholders. Managing rangeland 

brush and weeds requires adapting management 

methods to match species physiological 

and morphological traits and ecological site 

conditions (Fig. 7). The decision-making 

process associated with landscape-scale IBMS 

projects can be restricted by numerous factors, 

including equipment availability, financial 

constraints, available manpower, time needed to 

complete the task, environmental regulations, 

and agency mission and policy. 

the need for Monitoring 

Monitoring is intrinsically linked to the IBMS 

process (i.e., treatment monitoring, control, 

revegetation (natural or planted), and pre- and 

posttreatment monitoring; Scifres et al. 1983; 

Scifres 1987; DiTomaso 2000). Assessing baseline 

(pretreatment) conditions is essential to determine 

the effects of brush management efforts. 

Monitoring of weather and seasonal growth of 

the plants targeted for manipulation should begin 

well in advance (3–6 mo) of planned treatment. 

Posttreatment monitoring should be conducted 

at least annually to evaluate responses. 

Field variables to be measured for baseline 

inventories and posttreatment monitoring 

should be selected on the basis of the objectives 

of the brush management project. The 

following represent initial field variables that 

should be assessed in baseline inventories and 

to evaluate posttreatment responses: 

• Shrub age (or size as a rough proxy), class 

FIguRE 7. Ecosystem 

response to brush management varies with time since 

treatment (Y1, Y2, etc.) and 

is determined by a variety 

of interacting factors. 

distribution (baseline only), height, and 

stem density and diameter 

•	� Plant composition and frequency of 

occurrence 

•	� Ground cover of grass, forb (total and by 

species), litter, and bare ground 

•	� Biomass (seasonal or peak live standing 

crop; preferably by species) 

•	� Specific metrics will depend on the goals 

and objectives of the brush management 

project. 

treatment options 

Understanding the ecology of WPs and 

herbaceous plants and how they interact 

with each other on a particular site is crucial 

in determining the IBMS strategy. Relevant 

stand characteristics include plant community 

composition, plant phenology, plant density, 

plant size (stem diameter; canopy area, 

volume, and height), canopy architecture, and 

patterns of biomass allocation to leaves and 

stems aboveground and roots belowground. 

Each of these can influence the effectiveness 

and longevity of a given brush management 

practice. Realization of brush management 

objectives often benefits from spatially 

explicit prescriptions that take into account 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 121 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

tABlE 5. Factors influencing the effectiveness and longevity of a given brush management practice. Cookbook or “one-size-fits-all” approaches for brush management seldom succeed. A given project should be tailored to individual goals, objectives, and 

circumstances, and these, in turn, will be mediated by the items shown in the table. 

• Site accessibility and terrain: If the site is difficult to reach or traverse (e.g., 

sandy soils, uneven topography), less labor-intensive methods (e.g., aerial 

spraying) might be more effective. 

• Stand characteristics: Extent, age, biomass, and plant density will be factors in 

selecting the most cost-effective methods. 

• Proximity to endangered species: The presence of a federally listed species 

or environmentally sensitive sites may preclude some types of management 

methods or limit the season of application. 

• Presence of desirable plant species or other important resources (e.g., 

archaeological sites): Locally targeted methods (e.g., individual plant 

treatments) may be warranted to protect other resources in the area. 

• Extent of area to be treated: Suitability and efficiency of a given treatment 

method may vary by the size of the area targeted for treatment. 

the topoedaphic heterogeneity of landscapes 

(i.e., uplands, side slopes, riparian, valley 

bottomland, etc.; Taylor and McDaniel 2004; 

Table 5). 

BRuSh MAnAgEMEnt AS A 

ConSERvAtIon tool: A CRItICAl 

ASSESSMEnt 

The projected effects of brush management 

on ecosystems and ecosystem processes 

boils down to its effects on: 1) herbaceous 

vegetation and native plant communities; 

2) livestock; 3) watershed properties related 

to erosion, soil condition, water quality, and 

water quantity; 4) wildlife; 5) air quality; and 

6) human dimension considerations. The 

following sections summarize the scientific 

literature addressing the first five of these 

and the question, Are the outcomes expected 

from brush management being realized? 

For treatment of economics and the human 

dimensions area, see Tanaka et al. (this 

volume). 

Our assessment began with a series of literature 

searches using the Web of KnowledgeSM . 

Search strings included “brush management” 

and terms referring to specific brush 

management techniques. Search results were 

filtered to include only studies conducted in 

the United States on rangelands and only those 

quantifying responses to brush management. 

Studies quantifying herbaceous responses 

dominate the brush management literature 

in the United States and comprised 48.7% of 

those in the sample (Fig. 8). Treatment efficacy 

and shrub regeneration studies accounted for 

another 28.7%. Studies documenting water 

(4.1%) and soil (2.0%) responses were the 

least common. 

herbaceous vegetation and native 

Communities 

The general expectations associated with 

brush management are that it will mitigate 

soil erosion, improve soil condition, enhance 

water quantity and quality (via improvement 

in infiltration and reductions in runoff, 

which interact to reduce sedimentation), 

and improve livestock production. Each of 

these expectations is based on the assumption 

that herbaceous ground cover will increase 

following brush management. How good is this 

assumption? 

Herbaceous Response. The majority (>80%) 

of studies in our literature sampling reported 

positive herbaceous responses following brush 

management (Appendix I). Herbaceous plant 

growth increases an average of 3- to 5-fold for 

brush management conducted on productive 

range sites, including sites with Wyoming big 

sagebrush (Artemisia tridentata; Hyder and 

Sneva 1956; McDaniel et al. 2005) and broom 

snakeweed (Gutierrezia sarothrae; McDaniel et 

al. 1993). Management of other woody species, 

including mesquite (Prosopis glandulosa) and 

creosote bush, can result in substantial forage 

increases on productive sites with adequate 

rainfall (Ethridge et al. 1984; Perkins et al. 

2006; Combs 2007). In semidesert grasslands 

at the Santa Rita Experimental Range in 

Arizona, herbage yields following velvet 

mesquite removal increased (Parker and Martin 

1952; Paulsen 1975; Cable 1976) or remain 

unchanged in zones <1100 m in elevation 

and when velvet mesquite cover was <25% 

(McClaran and Angell 2006). These patterns 

are consistent with field studies in southern 

New Mexico (Warren et al. 1996; Drewa and 

Havstad 2001). 

No consistent relationship between 

posttreatment changes in herbaceous 

production and annual rainfall were found; 

122 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	

	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

however, a survey of data across a range of 

management contexts suggests an upper limit 

for the herbaceous production responses that 

might be expected for a given rainfall zone (Fig. 

9A). Herbaceous response corrected for annual 

rainfall varies with time since brush treatment 

(Fig. 9B). The median first-year response is 0 

and highly variable, with half the treated sites 

responding positively and half negatively. By 

year 2, the median response is slightly positive 

but also highly variable. After year 2, the 

response becomes more consistent and peaks in 

year 5. The response then drops off in years 7 

and 8, being slightly but consistently positive. 

The longevity of brush management 

treatments varies widely by type of treatment 

applied, shrub species, effectiveness of 

the initial treatment, composition of the 

herbaceous vegetation, and soil properties 

(Fig. 10). Variations in the Figure 10 

conceptual model have been illustrated 

for velvet mesquite (Cable 1976), honey 

mesquite (Heitschmidt et al. 1986; Ansley 

et al. 2004b), big sagebrush (McDaniel et 

al. 2005), and creosote bush (Gibbens et al. 

1987; Morton and Melgoza 1991; Perkins 

et al. 2006). The change in foliage cover and 

herbaceous response to brush management 

ranges from 5 to 20 yr for velvet mesquite 

(Cable 1976), from 10 to 25 yr for honey 

mesquite (Dahl et al. 1978; Jacoby et al. 

1991, 1990a, 1990b; Combs 2007), >25 yr 

for sagebrush (McDaniel et al. 2005), and 

>40 yr for creosote bush (Perkins et al. 2006). 

The general curve shape of the overstory– 

understory relationship for these shrub species 

is similar, but average grass yield associated 

with overstory cover is scaled quite differently: 

from 2- to 3-fold greater for mesquite 

relative to big sagebrush and creosote bush, 

respectively. Mesquite management typically 

provides a greater forage response, but it is 

of shorter duration than for big sagebrush 

and creosote bush removal. Accordingly, 

timing of investments to re-treat communities 

dominated by these shrubs would be on the 

order of about 4–12 yr for mesquite, 20–30 

yr for big sagebrush, and >30 yr for creosote 

bush (Torell and McDaniel 1986; Torell et al. 

2005a). 

Although studies have or currently are being 

conducted across different ecological sites 

in the western United States, adequate data 

to statistically estimate the relationships in 

Figure 10 as a function of rainfall, soils, site 

productivity, and so on are not generally Robust 

available. Instead, qualitative assessments by generalizations experienced range scientists and economists 

are currently the norm for projecting forage regarding brush response to brush management (Fig. 11). management 

Soil Condition. Some of the projected effects effects on soil 

of brush management on soil are associated carbon are with the assumption that soil organic matter 

depletion will occur with shrub encroachment. currently not However, as reviewed earlier (see the section possible.” “Soil Condition and Erosion”), this is not a 

robust assumption. It may be true in certain 

cases, most likely those in arid areas where 

disruption of grass cover by grazing has 

accelerated wind and water erosion. But even 

in those instances, soil resources may undergo 

not a net change in abundance but, rather, a 

change from a homogenous to a heterogeneous 

distribution wherein they are concentrated 

within shrub islands (Schlesinger and Pilmanis 

1998). Thus, were it not for shrubs, soil 

resources may have been lost from the site 

because of grazing rather than being spatially 

rearranged. Brush management, by reducing 

shrub cover on fertile shrub islands, may put 

FIguRE 8. Proportion of brush management studies quantifying various categories of 

treatment effects. Published papers resulting from Web of Knowledge search strings 

that included “brush management” and terms referring to specific brush management 

techniques were distilled to a database of 333 articles that reported quantitative 

responses. These were then classified into the categories shown. Articles reporting 

data for multiple metrics were tallied in multiple categories. Thus, the graph reflects 

the information reported in the literature but not on a per-paper basis. 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 123 
A
 B
 

FIguRE 9. (A) Changes in herbaceous biomass 

production (kg · ha−1) 1, 2, and 3 yr after brush 

management as a function of current years’ annual 

precipitation (PPT, mm). Multiple observations for 

a given PPT value reflect multiple sites or different 

brush management applications. PPT was deter-

mined from nearby weather stations if not reported. 

The number of studies pertaining to a given brush 

treatment are listed parenthetically in the key. (B) 

Change in herbaceous biomass per millimeter of an-

nual precipitation received after brush management. 

Responses are from 13 studies representing brush 

management with fire, herbicides, and mechani-

cal treatments. Tukey box plots show inner-quartile 

range (IQR; rectangle) and the median (bold line). 

Whiskers indicate the maximum and minimum 

values or the values within 1.5*IQR of the third 

and first quartiles, respectively. Values beyond 

1.5*IQR of the first and third quartiles are consid-

ered statistical outliers and are indicated with open 

circles. N = 13, 13, 11, 8, 5, 3, and 2 for years 

1 through 7, respectively. The number of studies 

pertaining to a given brush treatment are listed par-

enthetically in the figure legend. Citations used to 

generate the data points are given in Appendix III. 

these sites at risk for net losses of soil nutrients 

unless ground cover is quickly established. 

Alternatively, nutrient losses from shrub islands 

following brush management may help reinstate 

the homogeneous distribution of resources 

by disrupting the processes that lead to the 

concentration of nutrients in and around shrub 

canopies (e.g., Davies et al. 2009a; Ravi et al. 

2009). In the latter scenario, the likelihood of 

getting grasses reestablished within intershrub 

zones may improve. Site-specific factors may 

dictate which of these competing scenarios is 

most likely on a given landscape. 

WP encroachment can have a moderate to 

strong positive impact on SOC and N pools 

on many sites (Fig. 4). This shrub-induced 

improvement in SOC and N may be an 

important factor underlying the extent to 

which herbaceous vegetation production 

increases following brush management 

(Fig. 9). The degree to which shrubs might 

increase soil resources beneath their canopies 

is a function of how long the shrubs have 

occupied the site (Wheeler et al. 2007; 

Throop and Archer 2008). Thus, stand age 

at the time of brush management will have 

an important bearing on soil conditions. 

Removal of individual shrubs causes depletion 

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S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

FIguRE 10. Generalized conceptual model of herbaceous response to brush management. The lag time 

in response (t1−t2), the magnitude of (p1−p2) and 

time (t1−t3) to peak herbaceous response, the duration of peak elevated production response (t3−t4), 

and the time frame over which herbaceous productions decline as shrubs reestablish (t4 onward) vary 

with numerous factors. Knowledge of the relationships depicted in this conceptual model for a given 

ecological site will help determine the type, timing, 

and appropriate sequencing of brush management 

practices in an Integrated Brush Management Systems (IBMS) approach. 

of the associated resource pool and the 

availability of nutrients over the 10–15 yr 

following treatment, the extent depending 

on whether shrubs regenerate (Klemmedson 

and Tiedemann 1986; Tiedemann and 

Klemmedson 1986, 2004). Losses of SOC and 

N accumulating in soils associated with mature 

shrubs killed by herbicide ranged from 67% 

to 106% at 0–5-cm soil depths and from 78% 

to 93% at 5–10-cm soil depths over a 40-yr 

period (McClaran et al. 2008). Data from 

these individual plant perspectives suggest 

that brush management will cause a decline 

rather than an increase in SOC and N pools 

in hot, semidesert rangelands but that shrub 

regeneration can arrest or reverse such declines 

(Hughes et al. 2006). These findings contrast 

with those of Teague et al. (1999), who 

compared SOC and N on sites 4–22 yr after 

root plowing against untreated controls in the 

southern Great Plains to test the hypothesis 

that removal of honey mesquite would result 

in steady decline in SOC because of a loss of 

mesquite inputs and reductions in shading 

(and therefore higher soil temperatures and 

higher oxidation rates). Overall, they found 

no significant differences between treated and 

control sites. Similarly, Hughes et al. (2006) 

found that while aboveground C and N pools 

increased markedly with mesquite stand 

development following brush management 

(more so on sandy sites than shallow, clayey 

sites), near-surface SOC and N pools were 

unaffected. Thus, as with WP encroachment 

(Fig. 4), robust generalizations regarding brush 

management effects on soil condition are 

currently not possible. 

Timing of 

investments 

to re-treat 

communities is 

about 4–12 yr 

for mesquite, 

20–30 yr for 

big sagebrush, 

and >30 yr for 

creosote bush”

FIguRE 11. Projected changes in livestock stocking rates for aerial spraying and mechanical plus grass 

seeding treatment of a Prosopis–mixed brush system in southern Texas relative to untreated controls 

( J. Conner et al., unpublished data, 1998). 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 125 
	 	 	 	

 A 

B 

Non-native species invading or purposely 

seeded following brush management (see Fig. 

12 and the section “Biodiversity and Nonnative 

Species”) may significantly reduce ecosystem 

C accumulating with WP encroachment. 

Indeed, estimates of aboveground C loss 

with conversion of Great Basin shrublands 

and woodlands to annual grasslands are on 

the order of 8 Tg C, with estimates of 50 Tg 

C release to the atmosphere over the next 

several decades (Bradley et al. 2006). In cold 

desert sagebrush steppe ecosystems, this level 

of C release with annual grass invasion could 

completely offset any increases in C with 

woody encroachment that has occurred over 

FIguRE 12. (A) Mojave Desert scrub near Las 

Vegas, Nevada, USA (foreground) and area 

invaded by the nonnative annual grass red 

brome (Bromus rubra; background) following a 

fire that carried from desert floor upslope into 

pinyon–juniper woodlands. This exotic grass has 

instigated a positive disturbance (fire) feedback 

that reduces ecosystem carbon storage, threatens 

the biodiversity, and constitutes a new ignition 

source for fire in upper-elevation woodlands and 

forests. Photo: T. E. Huxman. (B) Nonnative annual 

grass (medusahead, Taeniatherum caput-medusae) 

invaded area (foreground) transitioning into a 

Wyoming big sagebrush (Artemisia tridentata 

wyomingensis) plant community in southeastern 

Oregon. The presence of this exotic annual grass 

increases the probability that the sagebrush plant 

community above it will burn. (Photo: K. W. Davies) 

the past century. However, the story may be 

quite different in southwestern rangelands 

where highly productive, deeply rooted 

perennial grasses introduced from Africa are 

expanding and sequester substantially more C 

than annual grasses (e.g., Williams and Baruch 

2000; Franklin et al. 2006). 

Specific brush management techniques 

will likely differ in their impact on litter 

decomposition, depending on the type of 

disturbance they cause, treatment efficacy, and 

the extent to which they co-occur with other 

land use practices, such as livestock grazing. 

Brush management treatments that minimally 

disturb soils (e.g., herbicide applications and 

prescribed burning) may be most advisable 

for managers wishing to minimize short-term 

SOC losses. In contrast, brush management 

techniques that cause extensive disturbance to 

the soil surface, such as chaining, root plowing, 

and grubbing, may increase decomposition 

rates due to surface soil disturbances. These 

practices likely superimpose a variety of new 

short- and long-term direct and indirect 

effects on decomposition processes via their 

dramatic alteration of surface roughness, water 

infiltration and runoff, vegetation cover, and 

ANPP and by initiating large, synchronous 

inputs of leaf, stem, and coarse woody debris 

onto the soil surface with widely varying 

degrees of contact and incorporation into 

the soil. Such treatments ostensibly increase 

exposure to direct sunlight and UV radiation 

and may promote soil movement via wind 

126 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

and water, particularly during the immediate 

posttreatment period when vegetation is 

reestablishing. 

Vegetation responses to brush management have 

been widely described, but very little is known 

of its effects on soils and nutrient cycling (Fig. 

8). There have been few attempts to model 

brush management effects on ecosystems (but 

see Carlson and Thurow 1996; Grant et al. 

1999), and the future development of such 

models would likely benefit from field studies 

elucidating how various brush management 

practices might impact C and N cycling. 

Lessons learned from studies of temperate forest 

clear-cutting and tropical deforestation would 

be instructive starting points, but it is likely 

that the shrublands, savannas and woodlands 

of drylands would have novel behaviors. For 

example, predicting brush management effects 

on litter decomposition in semidesert grasslands 

will require information on shrub–grass 

interactions and herbaceous biomass influences 

on soil movement at a decadal time scale 

(Throop and Archer 2007). 

Biodiversity Response. Biodiversity responses 

can be assessed at the species (e.g., genetic 

variation in populations), the organismal 

(species richness), the structural (vegetation 

strata and physiognomy), and the functional 

(plant functional groups and animal guilds) 

levels. Studies at the organismal level are 

typically restricted to a select class of organisms 

(e.g., perennial herbaceous plants or small 

mammals) without regard for other classes 

(annual plants, shrubs, reptiles, avifauna, 

large mammals, microbes, etc.). To further 

complicate things, diversity varies with scale 

(e.g., alpha, beta, and gamma diversity) and 

topoedaphic heterogeneity. Objectives aimed 

at preserving, restoring, and monitoring 

biodiversity should thus be phrased to 

specifically articulate the facets of biological 

diversity being addressed. 

From a Web of Knowledge search generating 

333 studies quantifying responses to brush 

management (Fig. 8), 39 articles reporting 

herbaceous plant diversity emerged; of these, 

13 were conducted on rangelands and were 

amenable to comparative analysis. From the 

90 data points emerging from these studies, 

it appears that brush management treatments 

typically have neutral (30% of data points 

exhibiting <10% change) to positive (60% 

of data points exhibiting >10% increase) 

effects on grass/forb diversity (Fig. 13). Cases 

where brush management had negative 

effects on herbaceous diversity (10% of 

data points exhibiting >10% decline) were 

typically associated with herbicide treatments, 

ostensibly reflecting adverse impacts on forbs. 

The few long-term data available suggest that 

posttreatment stimulation of herbaceous 

diversity is relatively short lived (<15 yr). 

In the subtropical southern Great Plains 

characterized by a diverse flora of encroaching 

WPs, WP communities developing after 

brush management have lower shrub diversity 

and higher densities of less desirable browse 

species than the previously existing community 

(Fulbright and Beasom 1987; Ruthven et al. 

1993). In systems where shrubs aggressively 

regenerate vegetatively, use of low-intensity 

fire and herbicides can promote a savanna 

physiognomy (e.g., Ansley et al. 1997, 2003) 

and ostensibly promote diversity. 

FIguRE 13. Changes in herbaceous vegetation diversity following brush management. Results (90 data points) show ratios of values (richness and various indices) 

reported for treated and control areas. Ratios near 1.0 indicate that diversity metrics 

on treated and control areas were similar, values <1.0 indicate that brush management decreased diversity, and values >1.0 indicate increases in diversity. The number of studies pertaining to a given type of brush treatment are listed parenthetically 

in the key. See Appendix III for citations. 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 127 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

		

	 	 	

	 	

	 	 	

	 	

	 	 	

	 	

Brush 

management 

has the potential 

to create 

conditions 

favorable for the 

establishment and 

growth of weeds 

and invasive 

nonnative 

species.”

Faunal diversity response to brush 

management varies with the organisms of 

interest (see the section “Wildlife”). For 

example, although Jones et al. (2000) reported 

that relative total abundance and species 

richness of herpetofauna was similar among a 

variety of treatments, amphibians were most 

abundant in untreated and herbicide-only 

sites, lizards were most abundant on untreated 

sites, and snakes were most abundant on sites 

receiving herbicide and fire. Rodent and avian 

relative frequency, richness, and diversity 

have been observed to be unaffected by brush 

management (Nolte and Fulbright 1997; 

Peterson 1997). 

The biodiversity response to brush management 

may be strongly influenced by the pattern 

of treatment application (see Bestelmeyer et 

al. this volume). A “wall-to-wall” application 

may yield one result, whereas applying a 

treatment or combination of treatments in 

“strips” may have a quite different outcome 

by creating more habitat edge and creating 

patches of grassland habitat interspersed with 

shrubland habitat (e.g., Scifres et al. 1988). 

For example, diversity of native perennial 

grasses may be promoted by a mixture of 

open areas interspersed with cover of mature 

shrubs (Tiedemann and Klemmedson 2004). 

Effects of brush management on biological 

diversity are poorly understood and need to be 

investigated at larger scales across longer time 

periods. 

Biodiversity and Nonnative Species. Brush 

management has the potential to create 

conditions favorable for the establishment 

and growth of weeds and invasive nonnative 

species (Young et al. 1985; Belsky 1996; Bates 

et al. 2007) that can have adverse affects on 

biodiversity. Brush management is therefore 

often conducted in conjunction with seeding 

operations intended to accelerate establishment 

of ground cover and a forage base (see 

Hardegree et al. this volume). In many cases, 

the grasses used for seeding are nonnative 

perennials (Cox and Ruyle 1986; Ibarra-Flores 

et al. 1995; Martin et al. 1995; Christian and 

Wilson 1999). Seeds from such species may be 

more readily available, and their establishment 

success rates may be higher than that of 

natives (Eiswerth et al. 2009). When seeding 

of nonnative grasses is successful, the result is 

often a persistent, long-lived near-monoculture 

of nonnative vegetation. While this may be 

valued for livestock production and ground 

cover and may make the site more resistant to 

invasion by undesirable exotic annual grasses 

(Davies et al. 2010) by virtue of their superior 

competitive ability (Eissenstat and Caldwell 

1987), these plants may represent threats to 

the biodiversity of native plants and animals 

(McClaran and Anable 1992; Williams and 

Baruch 2000; Schussman et al. 2006). Their 

unintended spread into areas beyond where 

they were planted may make it difficult to 

achieve conservation goals on other lands. 

Thus, there are clear trade-offs that should be 

explicitly considered and evaluated. 

A Tool to Promote Landscape Heterogeneity 

and Biodiversity? Disturbances associated with 

fire and herbivory (granivory, grazing, browsing, 

burrowing, trampling, and dung/urine 

deposition) interact with climate variability 

and extremes to generate patchiness across the 

landscape and contribute to the maintenance or 

enhancement of biological diversity. It is now 

recognized that such disturbances should be 

explicitly included in ecosystem management 

and conservation plans (Pickett et al. 1997; 

Fuhlendorf and Engle 2001). 

In the IBMS approach, brush management 

techniques can be targeted for certain portions 

of a landscape and distributed across landscapes 

in both time and space such that mosaics of 

vegetation structures, patch sizes, shapes, and 

age states are created. This, in turn, would 

promote the co-occurrence of suites of insect, 

reptile, mammalian, and avian species with 

diverse habitat requirements (Jones et al. 

2000). The logistics of planning and applying 

spatially heterogeneous brush management 

practices at appropriate scales is facilitated by 

advances in geomatics (e.g., global positioning 

satellites, geographic information systems, and 

remote sensing imagery) and landscape ecology 

that allow habitat and population data to be 

readily linked over large areas. Thus, a lowdiversity shrubland or woodland developing on 

a grassland site can be transformed to a diverse 

patchwork of grassland–savanna–shrubland 

communities using a spatial placement of 

landscape treatments that promotes biological 

diversity at multiple scales (Scifres et al. 1988; 

Fulbright 1996). 

128 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 	

	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

livestock Response 

Livestock grazing contributes significantly to 

the economy and social fabric of most rural 

communities. Brush management is a tool 

used to restore native ecosystems that have the 

capacity to provide a steady source of forage 

for livestock while facilitating other uses and 

resource values (NRCS 2006; US Department 

of the Interior, Bureau of Land Management 

2007). The decision of whether to apply 

brush management for the betterment of 

domestic livestock is influenced by numerous 

factors, including the extent to which declines 

in carrying capacity (Olson 1999), animal 

performance (Ralphs et al. 2000), animal 

loss from poisoning (Williams 1978; Panter 

et al. 2007), animal handling (Hanselka and 

Falconer 1994), and animal health (Teel et al. 

1998) will be impacted. Even when grazing has 

contributed to shrub increases, simply removing 

livestock or reducing their numbers is unlikely 

to remedy a brush encroachment problem 

(Browning et al. 2008). Passive treatments 

may help, but in many situations aggressive 

intervention is necessary (Olson 1999). 

Livestock can be used as part of the vegetation 

treatment program, especially when goats and 

sheep are used to apply browsing pressure on 

Rocky Mountain elk in a 

sagebrush community where 

encroaching western juniper 

trees have been cut to 

preserve habitat for sagebrush 

associated wildlife. (Photo: K. 

W. Davies) 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 129 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	

	 	 	 	

	 	 	

	 	 	 	

	 	 	 		

	 	 	

unwanted shrubs and weeds (Riggs and Urness 

1989; Frost and Launchbaugh 2003). 

Studies quantifying forage response to reductions 

in brush cover are relatively numerous (Fig. 

8), but few have quantified direct commodity 

(livestock) benefits. Potential changes in 

livestock carrying capacity for contrasting 

brush management × precipitation scenarios 

are illustrated in Figure 11. These projections 

illustrate that a range in livestock returns 

should be anticipated because of differences in 

forage response during favorable, normal, and 

unfavorable rainfall conditions that may occur 

over a 20-yr horizon. In this example, using 

cattle prices equal to the average of the past 20 

yr, current operating costs, and current costs 

of brush management practices, the returns on 

aerial spraying and mechanical practices may be 

relatively high when environmental conditions 

support high levels of herbaceous production. 

However, returns are greatly reduced when 

conditions for plant growth are poor. Subjective 

projections such as these are based on the best 

available information, and actual results are 

known to vary widely, depending on the specific 

situation. As research continues, more accurate 

and reliable projections can be developed. 

Increases in available forage following brush 

management do not necessarily warrant 

At advanced stages of shrub 

encroachment, brush management can improve biological 

diversity while potentially benefitting livestock production by 

increasing grazable land area. 

(Photo: K. W. Davies) 

an increase in livestock numbers. In some 

cases, justification for brush management 

may be to maintain stocking rates nearer its 

true capacity (i.e., recognition that current 

stocking rates cannot be sustained). For 

example, big sagebrush management on 

public land helps avoid potential conflict 

and lawsuits with grazing permittees and 

environmentalists because positive steps are 

taken to reduce grazing pressure without 

forcing major herd reductions (Torell et 

al. 2005a). Similarly, forestalling the need 

for controversial grazing reductions was a 

primary benefit of the 11-yr (1962–1972) 

Vale Rangeland Rehabilitation Program 

initiated in eastern Oregon (Bartlett et al. 

1988). In the case of big sagebrush, brush 

management is not always acceptable 

because of its adverse impact on habitat for 

sagebrush-obligate wildlife species (Rhodes et 

al. 2010). 

Reductions in brush and weeds potentially 

benefit livestock operators by increasing 

grazable land area (McDaniel et al. 1978). 

However, returns based solely on gains 

in animal performance are not always 

economically justified, especially when public 

assistance is not available (McBryde et al. 1984; 

Torell et al. 2005b). Lee et al. (2001) found 

that costs for brush management projects in 

the Edwards Plateau area of Texas exceeded 

livestock returns by 7–31%. Similarly, Torell 

et al. (2005b) found that a 30% cost-share 

agreement was required to justify big sagebrush 

management in New Mexico when the added 

forage from the brush management practice 

was valued at an intermediate level of $7/ 

AUM (in 2003 dollars). A range improvement 

practice that increases forage during critical or 

limiting seasons can be economically feasible 

(Evans and Workman 1994). 

While other resources (soil, water, wildlife, 

etc.) may benefit from IBMS, the economics 

of brush management practices continue to be 

evaluated on the basis of the amount of forage 

and meat products gained by implementing 

the practice (Tanaka and Workman 1988; 

Watts and Wamboldt 1996; Lee et al. 2001). 

The economic component of the holistic 

decision support system PESTMAN (2009) 

is driven by the anticipated forage response 

to brush management. Yet, as noted over 30 

130 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

FIguRE 14. Overland flow 

is an important mechanism 

of runoff generation for 

many semiarid landscapes. 

For woodlands such as the 

pinon–juniper stand in the 

Jemez Mountains, New 

Mexico, pictured here, 

much of the runoff and erosion is generated from the 

intercanopy spaces. (Photo: 

Bradford Wilcox) 

yr ago by Smith and Martin (1972), most 

range improvements show a negative benefit/ 

cost ratio (costs exceed benefits) when based 

only on the value of the added forage for 

livestock production. This is a consistent 

and continuing conclusion that increased 

returns from improved animal performance 

and production are often too low for brush 

management to be economically justified 

(McBryde et al. 1984; Lee et al. 2001; Torell 

et al. 2005a). Landowners recognize this, and 

many if not most recognize other benefits 

to conducting brush management beyond 

livestock production. Additionally, most 

landowners conducting a brush management 

project do so under cost-share arrangements 

with state and federal agencies. When the 

value of ecosystem goods and services beyond 

those associated with livestock production are 

taken into account, a more favorable picture 

of brush management begins to emerge (see 

Tanaka et al. this volume). 

Watershed Function 

The NRCS makes a number of assumptions 

related to the hydrological consequences of 

brush removal. These assumptions fall into 

three broad categories: 1) horizontal fluxes— 

the removal of WPs will reduce overland flow 

(surface runoff) and erosion, primarily by 

improving infiltration rates and increasing 

ground cover; 2) vertical fluxes—the removal 

of WPs will reduce the evapotranspiration 

(ET) and thus increase groundwater recharge; 

and 3) landscape effects—as a result of 

assumptions 1 and 2, the removal of WPs will 

reduce gully erosion and increase stream flow. 

We review the validity of these assumptions 

below on the basis of relevant literature. 

Our review is organized by the primary 

geographic regions in the United States where 

information is available: the Southwest, the 

Northwest, and the southern Great Plains. 

Horizontal Fluxes—Surface Runoff and 

Erosion. The expectation that surface runoff 

and erosion are higher from woodlands or 

shrublands than from grasslands is implicit 

in the assumption that reductions in WP 

cover will reduce overland flow and water 

erosion (Fig. 14). In some cases, this is true, 

but in many cases, it is not. It is most likely 

in the xeric climates that support creosote 

bush shrublands and piñon–juniper (Pinus 

spp.–Juniperus spp.) and juniper woodlands. 

The influence of woody species encroachment 

on surface runoff and erosion depends on 

the impacts of encroachment on herbaceous 

vegetation and subsequently bare ground. 

Surface runoff and erosion increase when WP 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 131 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

FIguRE 15. Reductions in herbaceous ground cover resulting from grazing and 

woody plant encroachment can increase connectivity between bare patches and 

lead to higher runoff and erosion (Davenport et al. 1998). (A) pinyon–juniper 

woodland in the Jemez Mountains, New Mexico, and (B) creosote bush shrubland 

in southern New Mexico. (Photo: Bradford Wilcox) 

encroachment decreases herbaceous vegetation 

and increases bare ground; however, if WP 

encroachment does not decrease herbaceous 

vegetation and increase bare ground, then 

surface runoff and erosion would not increase. 

Brush management does not always reverse 

the impacts of WP encroachment on surface 

runoff and erosion. In some cases, depending 

on the woodland type and the method of shrub 

management, surface runoff and erosion may 

actually increase. 

Southwest. There is clear evidence that as 

desert grasslands transition to creosote bush, 

juniper, or mesquite shrublands or woodlands, 

there is more bare ground and better-connected 

interspaces, resulting in lower net infiltration, 

more surface runoff, and higher erosion 

(Fig. 15) (Parsons et al. 1996; Schlesinger et 

al. 2000; Mueller et al. 2008; Wainwright 

et al. 2000). However, the reverse has not 

been demonstrated. In other words, brush 

management on creosote bush shrublands 

does not necessarily curtail surface runoff and 

erosion (Tromble et al. 1974; Tromble 1978, 

1980; Wood et al. 1991). 

There has been relatively little work 

evaluating the hydrological implications of 

managing mesquite in the Southwest. Longterm watershed studies at the Santa Rita 

Experimental Range suggest that runoff and 

sediment yields may decline with mesquite 

removal (Lane and Kidwell 2003). The 

results are, however, equivocal because no 

pretreatment monitoring took place. 

Surface runoff is a relatively small portion of 

the water budget in piñon–juniper woodlands 

(Gifford 1975), primarily because of internal 

storage within the hillslopes (Reid et al. 1999; 

Wilcox et al. 2003a). Surface runoff is higher 

when snowmelt occurs (Baker 1984; Wilcox 

1994). Infiltration rates are higher under 

tree canopies than in the interspaces spaces 

(Reid et al. 1999), even though the hydraulic 

conductivity of canopy and intercanopy soils 

is similar (Wilcox et al. 2003b), likely because 

of the buildup of duff under that canopy. 

In these woodlands, small-plot infiltration 

studies indicate that shrub management 

has little effect or even a negative effect 

on infiltration rates (Gifford et al. 1970; 

Blackburn and Skau 1974; Roundy et al. 

1978). Runoff and erosion are the highest 

following chaining and windrowing. If debris 

is left in place, there is little difference in 

surface runoff between treated and untreated 

locations (Gifford 1973). Watershed-scale 

experiments in Arizona indicate little effect 

of brush management on surface runoff 

(Clary et al. 1974; Collings and Myrick 

1966). Although runoff may be relatively 

small in general, it may be much higher on 

woodlands occurring on slopes >10% (Wilcox 

et al. 1996a, 1996b). In these landscapes, 

cutting trees and leaving the slash in place has 

proven to dramatically decrease erosion rates, 

primarily because of increases in herbaceous 

cover (Hastings et al. 2003). 

Northwest. Major shrublands of the 

northwestern United States are those 

dominated by sagebrush or western juniper 

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S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

(Juniperus occidentalis). Erosion on sagebrush 

rangelands is generally very low (Coppinger 

et al. 1991). Effects of brush management 

on horizontal water fluxes in sagebrush 

landscapes are equivocal. Increasing surface 

runoff and erosion has been documented in 

some cases, and declines have been observed in 

others (Blackburn 1983; Brown et al. 1985). 

Mechanical treatments that disturb soil can 

increase runoff and erosion. For example, 

plowing reduced infiltration and increased 

runoff and erosion up to 12 yr (Gifford 1982). 

Soil erodibility was initially ~2-fold greater 

in burned compared to unburned sagebrush 

communities, but was comparable 1 yr postfire, 

and infiltration and runoff was comparable 

on burned and unburned hillslopes in the 

first and second years postfire (Pierson et al. 

2001). However, in another study, burning 

had little effect on runoff but resulted in a 

large increase in erosion (Pierson et al. 2008). 

Balliette et al. (1986) found little change 

in infiltration, runoff, or erosion following 

herbicide treatment. In contrast, Blackburn and 

Skau (1974) found that plowing and reseeding 

of big sagebrush increased infiltration rates 

and lowered surface runoff. Effects can also 

vary with season. At the small catchment scale 

(2–4 ha), summer runoff and erosion declined 

by 75% and 80%, respectively, following 

conversion of sagebrush to introduced grasses, 

whereas snowmelt runoff increased 12% 

(Lusby 1979). The contrasting results from this 

population of studies likely reflect differences 

in responses of the herbaceous understory 

to sagebrush clearing and differences in 

disturbance impacts associated with various 

brush management techniques. 

Western juniper has been aggressively 

encroaching into sagebrush communities 

across the intermountain West (Miller et al. 

2005). Subsequent to its establishment, western 

juniper excludes other vegetation and increases 

bare ground (Miller et al. 2000). Although 

Belsky (1996) found little compelling evidence 

that surface runoff and erosion were higher 

following western juniper encroachment, 

other work suggests that runoff and erosion 

can be significantly accelerated and that brush 

management can significantly mitigate these 

effects (Buckhouse and Mattison 1980; Gaither 

and Buckhouse 1983). Indeed, Pierson et al. 

(2007) found that 10 yr after juniper removal, 

treated hillslopes had significantly more 

vegetation cover, higher infiltration rates, and 

15-fold less erosion than nontreated sites. 

Southern Great Plains. In the southern Great 

Plains, the major shrublands of concern are 

those dominated by mesquite or juniper. Most 

of the research related to WPs, and water has 

been conducted in relation to Ashe juniper 

(Juniperus ashei) in the Edwards Plateau with 

some additional work completed on mesquite 

woodlands in the Rolling Plains and the South 

Texas Plains. 

Extensive woodlands dominated by Ashe 

juniper occur on the Edwards Plateau of 

central Texas. As with other juniper woodland 

types, there is a widely held perception that 

encroachment by this WP has promoted 

surface runoff and erosion. However, there is 

little evidence in support of this assumption. 

Infiltration rates within this woodland type 

are relatively high, and erosion is low unless 

the area is heavily grazed (Hester et al. 1997; 

Wilcox et al. 2007, 2008b; Taucer et al. 

2008). 

For mesquite shrublands in the Rolling 

Plains of northern Texas, small-plot rainfall 

simulations indicate that shrub management 

may improve infiltration capacity and reduce 

erosion as a result of increased herbaceous cover 

(Bedunah 1982; Brock et al. 1982). Largerscale plot and catchment studies, however, 

suggest the honey mesquite management 

would not significantly alter surface runoff 

and erosion (Carlson et al. 1990; Wilcox et al. 

2006). Weltz and Blackburn (1995) reached a 

similar conclusion for mesquite–mixed shrub 

rangelands in the Rio Grande Plains. 

Vertical Fluxes, ET, and Groundwater 

Recharge. WPs have the potential to alter the 

fluxes of water moving in a vertical direction, 

ET, and recharge by virtue of the fact that deep 

root systems allow WPs access to water not 

available to more shallow-rooted vegetation. 

The ability of WPs to access deep water is, 

however, modulated by soil depth, texture, 

and the underlying geological structure, the 

latter also being a key determinant of whether 

groundwater recharge events will affect stream 

base flow. In principle, in locations where WPs 

are accessing deeper water, there is the potential 

There are 

few examples 

demonstrating 

that brush 

management 

enhances ground 

water recharge or 

streamflow.”

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Historical stream 

flow records in 

the Edwards 

Plateau indicate 

that base flows 

have actually 

increased 

substantially since 

1960 in spite of 

the fact that WPs 

have increased 

markedly since 

that time.”

Large scale brush management 

programs focusing on mesquite 

rangelands in Texas have 

not resulted in increased 

streamflow, in spite of public 

perceptions to the contrary. 

(Photo: Bradford Wilcox) 

to use vegetation management to enhance 

recharge. In practice, however, there are few 

examples of where this has been demonstrated 

and then only at relatively small scales. 

Southwest. There is little difference in ET 

between creosote bush shrublands and desert 

grasslands (Small and Kurc 2003; Kurc and 

Small 2004). Recent work suggests that 

removal of shrubs could increase groundwater 

recharge but not in amounts that would 

appreciably affect water supplies (Sandvig and 

Phillips 2006). 

Recharge rates in most piñon–juniper 

woodlands are very small, and it is unlikely 

that brush management would lead to higher 

recharge (Newman et al. 1997; Sandvig and 

Phillips 2006). However, decreasing piñon– 

juniper cover by chaining increased soil 

moisture in the upper 60–90 cm of the soil 

profile, with only minor differences at greater 

depths (Gifford and Shaw 1973). We are not 

aware of any work comparing ET between 

piñon–juniper woodlands and comparable 

grassland areas. 

Northwest. Removal of sagebrush can increase 

soil water content and presumably recharge 

(Sturges 1993; Seyfried and Wilcox 2006). 

Sturges (1993) suggested that reductions of 

sagebrush cover can increase water yield if 

sagebrush roots are not confined to the same 

volume of soil as grass roots. Along these 

lines, Darrouzet-Nardi et al. (2006) found 

that sagebrush in herbaceous meadows in the 

Sierra Nevada Mountains was in fact accessing 

deeper water than the herbaceous vegetation. 

Sagebrush management decreases water 

withdrawal from the upper 1 m of soil for 2 yr 

posttreatment (Sonder and Alley 1961; Cook 

and Lewis 1963; Tabler 1968; Shown et al. 

1972; Sturges 1977). However, over longer 

periods of time, water depletion to 0.9-m 

soil depth can be greater where sagebrush is 

removed compared to where it is not because of 

increases in herbaceous vegetation production 

(Sturges 1993). The replacement of sagebrush 

by nonnative annual grasses and forbs (e.g., 

Fig. 12) can alter the timing of ET and patterns 

of soil moisture storage. For example, Prater 

and De Lucia (2006) found that early spring 

ET rates were higher from areas converted to 

cheatgrass (Bromus tectorum), an exotic annual 

grass, than for native sagebrush. 

Interception by western juniper canopies can 

reduce the amount of precipitation reaching 

the ground by 20% at the edge of the canopy, 

50% halfway between the canopy edge and 

the trunk, and 70% at the trunk (Young et al. 

1984). Stem flow is low, and thus the moisture 

captured in western juniper canopies is lost 

through evaporation (Miller et al. 2005). 

Cutting western juniper increases soil water 

throughout the growing season in at least the 

first 2 yr posttreatment (Bates et al. 2000). We 

are not aware of longer-term studies evaluating 

the influence of western juniper management 

on soil moisture. 

Southern Great Plains. Ashe juniper 

intercepts 40–50% of rainfall (Fig. 16; Hester 

et al. 1997; Owens et al. 2006). Transpiration 

from an Ashe juniper community should 

be greater than that from an herbaceous 

community because evergreen Ashe juniper 

canopies can transpire much of the year in the 

subtropical portions of their range, and plants 

can access water to deep depths. Mature Ashe 

juniper trees transpire as much as 150 L·d−1, 

the equivalent of about 400 mm·yr−1 (Owens 

and Ansley 1997). Dugas et al. (1998), using 

the Bowen ratio/energy balance method, 

compared ET between intact and cleared Ashe 

juniper stands. For the 2-yr period following 

treatment, the difference in ET was about 

40 mm·yr−1, but the treatment effects on ET 

disappeared in the third year, by which time 

ET was similar in treated and untreated areas. 

For honey mesquite shrublands in the southern 

Texas plains, water balance studies suggest 

that conversion of mesquite to grasslands will 

increase recharge 15–20 mm·yr−1 (Weltz and 

Blackburn 1995; Moore et al. 2008). In the 

Rolling Plains of Texas, honey mesquite utilizes 

both deep and shallow soil water (Ansley et 

al. 1990, 1992a, 1992b), with individual 

plants using 30–200 L·d−1 and plants in open 

134 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

	 	

	 	

	 	

	 	 	

	 	

	 	 	

	 	

	 	 	

	 	

	 	 	

	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

savanna settings using more water per tree 

than plants in dense stands (Ansley et al. 1991, 

1998). At the stand scale, ET was comparable 

on cleared and uncleared honey mesquite 

rangelands (Dugas and Mayeux 1991); hence, 

the potential for increasing soil recharge or 

water yield by reducing mesquite cover in 

these systems is low (Carlson et al. 1990). 

Honey mesquite stands in the southern Great 

Plains can occur on fine, montmorillonitic clay 

soils with high shrink–swell potential. When 

dry, these soils develop extensive fissures that 

allow rapid and deep-percolation of rainfall. 

Mesquite removal on these soils reduced 

ET and increased soil moisture by about 80 

mm·yr−1 (Richardson et al. 1979). 

Landscape Effects. Streamflow. Brush 

management is commonly presumed to 

increase stream flow because of assumed 

increases in the base flow derived from 

increases in groundwater recharge. This has 

not been widely demonstrated except at the 

small-watershed scale, where stream flows are 

generated from winter precipitation (Huxman 

et al. 2005). A very prominent example of 

enhancement of stream flow subsequent 

to brush management is from chaparral 

shrublands characterized by winter rainfall 

(Rowe 1948; Ingebo 1972; Davis and Pase 

1977; Hibbert 1983). 

Increases in stream flow of ~150% were 

demonstrated on a 147-ha watershed following 

herbicide treatment in northern Arizona. 

Stream flow occurred mainly as a result of 

winter precipitation (Baker 1984). A largerscale watershed treatment, however, failed to 

generate additional stream flow (Collings and 

Myrick 1966). Annual water yield initially 

increased 20% on the herbicide-treated 

sagebrush sites (Sturges 1994), then returned 

to pretreatment levels within 11 yr as sagebrush 

density increased. Small-watershed studies in 

western Colorado indicate that runoff from 

summer thunderstorms was reduced following 

conversion of sagebrush to grass (Lusby 1979). 

A paired watershed study in central Oregon 

indicated that late season spring flow may 

increase as a result of juniper management 

(Deboodt et al. 2009). 

In the Edwards Plateau of Texas, Huang et 

al. (2006) found that spring flow increased 

FIguRE 16. Rainfall simulation is a valuable tool for 

understanding how woody plants alter hydrological properties on rangelands. Large-scale rainfall 

simulation experiments, like the one conducted here 

on Ashe juniper rangelands, have the advantage 

of being able to apply known amounts of water 

above the tree canopies and enable quantification of canopy interception and water and erosion 

dynamics at the hillslope scale. (Photo: Bradford 

Wilcox) 

by about 45 mm·yr−1 following Ashe juniper 

removal. Studies of juniper removal on small 

catchments where no springs were present 

found surface runoff was about 20% (13 

mm·yr−1) lower following root plowing, 

which was attributed to increased surface 

roughness that enhanced shallow surface 

storage (Richardson et al. 1979). In another 

study, Dugas et al. (1998) found that when 

juniper cover was removed by hand cutting, 

the treatment had little influence on surface 

runoff from 4- and 6-ha small catchments. 

Similarly, Wilcox et al. (2005) found no 

change in runoff following juniper removal. 

Paradoxically, historical stream flow records in 

the Edwards Plateau indicate that base flows 

have actually increased substantially since 1960 

in spite of the fact that WPs have increased 

markedly since that time (Wilcox and Huang 

2010). The higher base flows were attributed 

Projects that 

remove saltcedar 

and Russian 

olive with the 

intention of 

reducing ET and 

increasing flow 

in streams have 

produced mixed 

results, with most 

studies failing 

to demonstrate 

significant longterm changes.”

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 135 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

FIguRE 17. Brush management is commonly applied with hopes of improving 

stream flow and groundwater recharge. However, studies indicating that brush 

management may not be achieving desired outcomes with respect to water yield 

are accumulating. Estimates of the economic benefits of shrub control based solely 

on water salvage are therefore questionable. However, it may be desirable to 

manage cover of nonnative shrubs, such as the tamarisk shown here, to enhance 

wildlife habitat, biological diversity, and soil health (Shafroth et al. 2005, 2010). 

(Photo: Charles Hart) 

to the fact that ground cover has improved 

across the Edwards Plateau because of livestock 

destocking in the region. In the Rolling Plains 

of Texas, small-watershed and landscape-scale 

evaluations within the plains found little 

evidence that honey mesquite removal had an 

appreciable affect on stream flow (Wilcox et al. 

2006, 2008a). 

Early studies suggested that transpirational 

water loss from WPs such as saltcedar 

(Tamarix spp.) and Russian olive (Elaeagnus 

angustifolia) was substantially higher than 

that of native riparian vegetation. Expansion 

of these nonnative species along riparian 

corridors in the western United States was 

thus presumed to reduce river flows and 

groundwater supplies, and their removal 

was expected to promote stream flow and 

groundwater recharge (Fig. 17). However, 

recent studies indicate that saltcedar and 

Russian olive transpiration is on par with that 

of native species (Owens and Moore 2007), 

and projects that remove saltcedar and Russian 

olive with the intention of reducing ET and 

increasing flow in streams have produced 

mixed results, with most studies failing to 

demonstrate significant long-term changes 

(Shafroth et al. 2010). 

Sediment Delivery. There are few studies of 

brush management effects on sediment yield 

at the catchment or watershed scale. Hastings 

et al. (2003) found that cutting trees and 

spreading slash in piñon–juniper woodlands 

in New Mexico significantly reduced erosion 

from 1-ha catchments. Lusby (1979) found 

that shrub management reduced erosion by 

80% on two 4-ha sagebrush watersheds. Such 

studies suggest that brush management may 

help curtail erosion, but additional studies and 

studies at larger scales are needed before broad 

generalizations can be made with confidence. 

Wildlife Response 

NRCS goals of brush management for 

wildlife include 1) maintaining or enhancing 

habitat—including threatened and endangered 

species, with enhancements encompassing (a) 

slight to substantial improvement in cover, 

usable space, and habitat fragmentation; 

(b) improvement of imbalances among 

and within populations; and (c) neutral 

effects on endangered species—and 2) 

improving food accessibility, quality, and 

quantity. The challenge in meeting these 

goals lies with the fact that wildlife species 

and functional groups vary widely in their 

habitat requirements (Krausman et al. this 

volume). In addition, and as reviewed 

earlier, the response of vegetation and other 

habitat components to brush management 

varies, depending on a variety of factors. 

Impacts—positive, neutral, or negative—of 

brush management on wildlife therefore 

depend on a variety of factors (Fig. 7). 

Stating that brush management maintains 

or enhances wildlife habitat, consequently, 

is an oversimplification. Goals of brush 

management should be stated with the 

interacting factors that influence impacts on 

specific wildlife species taken into account. 

Habitat is species specific, and habitat for one 

species may not serve as habitat for another 

species or group of species (Hall et al. 1997; 

Krausman 2002). Clearing a large tract of 

sagebrush to create grassland, for example, 

may improve habitat for grassland birds 

(Reinkensmeyer et al. 2007) but destroy habitat 

for sagebrush obligates (Klebenow 1969; 

Martin 1970; Green and Flinders 1980). A 

fundamental concept in wildlife management 

is that wildlife species vary in their response 

136 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	 	

	 	 	

	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

to disturbance. Northern bobwhites (Colinus 

virginianus), for example, are frequently 

considered “early ecological succession stage” 

species, whereas white-tailed deer (Odocoileus 

virginianus) are considered “midsuccession 

species” and grizzly bears (Ursus arctos) “climax” 

species (Bolen and Robinson 2002). This 

implies that bobwhites, for example, should 

respond positively to disturbance, whereas 

climax wildlife species may be negatively 

impacted by human-imposed disturbances, 

such as brush management. 

Brush management may affect sexes of the 

same wildlife species differently (Leslie et al. 

1996; Stewart et al. 2003). For example, male 

and female white-tailed deer selected different 

herbicide and fire treatments in Oklahoma 

(Leslie et al. 1996). Anticipated conservation 

benefits should be stated on the basis of the 

species, functional group (e.g., grassland 

birds, woodland birds, large mammals, small 

mammals, etc.) or the gender that they will 

benefit; broad generalizations that all wildlife 

will be benefited by brush management should 

be avoided. 

Brush management may affect the same species 

differently, depending on seasonal use patterns 

of the habitat being treated. For example, 

thinning dense big sagebrush stands can 

benefit sage-grouse (Centrocercus urophasianus) 

during brood rearing (Dahlgren et al. 2006) 

but decrease its value as winter habitat for 

sage-grouse and other wildlife species (Davies 

et al. 2009b). Mechanical brush clearing 

during active nesting can destroy eggs and 

kill nestlings. The magnitude of the impact of 

brush clearing during active nesting on North 

American bird populations is unknown. 

Wildlife species response to brush 

management can also vary by the species of 

brush. Sagebrush-obligate wildlife species 

are negatively impacted by reductions of 

sagebrush abundance (Klebenow 1969; Martin 

1970; Green and Flinders 1980). However, 

sagebrush-obligate wildlife species benefit from 

control of western juniper encroaching into 

sagebrush communities (Miller et al. 2000; 

Reinkensmeyer et al. 2007). 

Density and canopy cover of brush before 

treatment and amount of brush removed 

strongly influence wildlife responses to brush 

management. Clearing some brush in a 

landscape with a 100% canopy cover of WPs, 

for example, may benefit wildlife such as 

white-tailed deer (Fig. 18), whereas clearing 

brush in a landscape with only 25% canopy 

cover may be detrimental (Fulbright and 

Ortega-Santos 2006). In areas where the two 

species overlap, reducing WP canopy cover to 

< 50% favors mule deer (Odocoileus hemionus) 

over white-tailed deer (Wiggers and Beasom 

1986; Ockenfels et al. 1991; Avey et al. 2003). 

Northern bobwhites use habitat patches where 

woody cover is ≥ 30%; therefore, reducing 

woody canopy cover in landscapes that 

marginally provide sufficient woody cover 

may be detrimental to bobwhites (Kopp et al. 

1998; Ransom et al. 2008). 

Climate and Soils Mediate Outcomes. 

Variation in precipitation and soil fertility 

may override effects of brush management 

on wildlife species abundance and richness 

in certain cases. Nutrition, productivity, 

and distribution of white-tailed deer, for 

example, may be more strongly related to 

variation in precipitation than to alterations in 

vegetation resulting from brush management. 

Seventeen years after root plowing, treated 

sites in the eastern Rio Grande Plains of 

Texas were dominated by huisache (Acacia 

farnesiana [L.] Willd.; Ruthven et al. 1994). 

Browse species important to white-tailed 

deer were either absent from the huisache 

communities that replaced the original honey 

mesquite-mixed brush communities or 

present in greatly reduced numbers compared 

to the mesquite–mixed brush community. 

Nutritional condition and population status 

of white-tailed deer, however, were similar in 

untreated and root plowed sites. Changes in 

body condition, reproduction, and diet were 

associated with variation in precipitation 

rather than with plant community differences. 

Similarly, patch burning and grazing had 

little effect on white-tailed deer distribution 

in southern Texas because drought limited 

vegetation response to the treatments (Meek 

et al. 2008). Lack of a difference in the use 

of aerated and aerated and burned patches 

by white-tailed deer has also been attributed 

to lack of precipitation, which constrained 

forb response to the treatments (Rogers et al. 

2004). 

Anticipated 

impacts should 

consider game 

and non-game 

species; and 

should be tailored 

to specific species 

or functional 

groups.”

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 137 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

FIguRE 18. Wildlife 

response to brush management is species and 

situation specific. Mule 

deer and white-tailed 

deer respond differently to 

changes in shrub cover, 

and the white-tailed deer 

depicted in this photo 

may respond positively 

when shrub cover is high 

but may be adversely 

affected by brush management imposed when shrub 

cover is <25%. (Photo: Tim 

Fulbright) 

Brush management effects on wildlife 

food and cover vary with soil productivity 

(Fulbright et al. 2008). Root plowing may 

result in long-term loss of WPs that are 

important as browse for white-tailed deer 

on upland soils, whereas in ephemeral 

drainages, root-plowed sites supported brush 

communities similar in species composition 

and diversity to sites that had not been 

disturbed (Fulbright and Beasom 1987; Nolte 

et al. 1994). Ephemeral drainages receive 

runoff from uplands and tend to have more 

productive soils (Wu and Archer 2005). A 

possible explanation for the lack of reduction 

in species diversity in ephemeral drainages is 

that growing conditions are more favorable 

for the reestablishment of diverse WP species 

following root plowing than in upland sites. 

Vegetation dynamics following brush 

management on fertile soils in mesic 

environments may follow directional 

change toward climax as predicted by 

traditional models of ecological succession. 

In arid or semiarid environments, however, 

vegetation change following disturbance 

may be nondirectional (Briske et al. 2005). 

Disturbance by brush management may push 

vegetation across a threshold to a different 

plant community than existed before treatment 

and one that is relatively stable. This new plant 

community may or may not provide better138 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

	 	

	 	 	

	 	

	 	

	 	

	 	

	 	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

quality habitat for specific wildlife species 

than the plant community that existed before 

brush management. For example, exotic annual 

grasses can rapidly increase and dominate 

plant communities after brush management 

in the intermountain West (Stewart and 

Hull 1949; Evans and Young 1985; Young 

and Allen 1997). Nonnative annual grass 

invasion in sagebrush communities decreases 

their habitat value for sagebrush-obligate 

and facultative wildlife (Davies and Svejcar 

2008). Buffelgrass may increase following root 

plowing or disking in South Texas (Gonzalez 

and Dodd 1979; Johnson and Fulbright 2008) 

with adverse effects on bobwhite populations 

(Flanders et al. 2006). Thus, the potential 

for undesirable shifts in plant communities 

following brush management must be carefully 

considered before implementing treatments 

(see the sections “Biodiversity and Nonnative 

Species” and “A Tool to Promote Landscape 

Heterogeneity and Biodiversity”). 

Scale and Pattern. Effects of brush 

management on wildlife may vary dramatically, 

depending on scale of application (see 

also Bestelmeyer et al. this volume). Many 

grassland-adapted species may respond in a 

positive fashion to broad-scale conversion 

of woodland to grassland (Fitzgerald and 

Tanner 1992; Smythe and Haukos 2010). 

Conversely, these large-scale conversions 

reduce northern bobwhite and Texas tortoise 

populations (Kazmaier et al. 2001; Ransom et 

al. 2008). Extensive brush removal (>60% of 

the landscape) reduces landscape use by whitetailed deer (Rollins et al. 1988; Reynolds et al. 

1992). Large areas of untreated brush provide 

habitat for many nongame bird species, and 

brush management efforts should be limited 

in scope in areas where conservation of this 

wildlife group is a priority (Fulbright and 

Guthery 1996). 

Range management has traditionally 

promoted vegetation uniformity rather than 

heterogeneity (Fuhlendorf and Engle 2004). 

Promoting uniformity, deemed prudent for 

increasing livestock production, included 

practices such as clearing WPs completely 

from the landscape, planting monotypic 

stands of grasses, and taking steps to promote 

livestock grazing distribution. Wildlife needs 

were relegated to lesser importance in this 

traditional management approach. Wildlife 

response to amount and interspersion of 

brush patches varies among species. Many 

wildlife species reach maximum diversity or 

density in heterogeneous landscapes such 

as those containing a mosaic of brush and 

interspersed tracts dominated by herbaceous 

vegetation (Roth 1976; Tews et al. 2004a; see 

also the sections “Biodiversity,” “Biodiversity 

Response,” and “Biodiversity and Nonnative 

Species”). Diversity and richness of birds is 

greatest in plant communities with structural 

heterogeneity (Reinkensmeyer et al. 2007). 

For example, providing a mosaic of plant 

communities including closed-canopy oak 

forest and open pastures derived from forest 

increased breeding nongame birds richness 

in Oklahoma (Schulz et al. 1992). Brush 

management is commonly done in strips 

or other patterns to create mosaics of WP 

communities interspersed with communities 

dominated by herbaceous plants to benefit 

wildlife (Fulbright and Ortega-Santos 2006). 

Brush sculpting is another approach to brush 

management (Fulbright 1997; McGinty 

and Ueckert 2001). Brush sculpting refers 

to selective removal of brush to accomplish 

multiple-use objectives, such as habitat 

improvement for wildlife and increased forage 

for livestock (Ansley et al. 2003). Anticipated 

effects of brush management should take 

into account the extent to which habitat 

heterogeneity is important for wildlife species 

(Fulbright 1996; Kie et al. 2002; Tews et al. 

2004a). 

Patch size is also an important consideration 

when creating vegetation mosaics (Bestelmeyer 

et al. this volume). Selection of patch size 

depends on management objectives and the 

wildlife species or functional group being 

managed. Mosaics may be created to either 

maximize wildlife species diversity or optimize 

habitat for a particular species. Edge and 

interior species are more prone to be affected 

by patch size than are generalist species (Bender 

et al. 1998). A mosaic consisting of patches 

that are too small essentially functions as 

edge and does not provide habitat for interior 

species. For woodland-adapted birds, patch 

size and shape are important because nest 

parasitism and nest predation may increase 

with increasing edge, although this relationship 

has been questioned in recent literature (Patton 

Vegetation 

mosaics may 

be created to 

either maximize 

wildlife species 

diversity or 

optimize habitat 

for a particular 

species.”

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1994; Lahti 2001). Patches that are large with 

relatively little perimeter support fewer edge 

species. 

Patch size and configuration requirements 

vary among wildlife species. Grassland birds, 

for example, require patches >50 ha (Helzer 

and Jelinski 1999). Ratio of patch perimeter 

to area is also important; bird species richness 

is greatest in patches with larger interiors 

that are free from edge effects. For grassland 

birds, landscape composition may interact 

with patch size in that larger core areas may be 

more important in landscapes with a mixture 

of grassland and woodland than in treeless 

landscapes (Winter et al. 2006). 

Although the idea of creating patchy mosaics 

through brush management has been discussed 

in the literature, using brush management to 

achieve an “optimum” size and configuration of 

patches has received little attention (Fulbright 

1996). Part of the reason for the lack of 

attention to the concept of optimal patch size/ 

configuration may be that many of the game 

species that are often the focus of research on 

brush management effects are edge associates 

that show little response to variation in patch 

size. Northern bobwhites, for example, appear 

to be adapted to an almost infinite set of 

patch configurations; therefore, an “optimum” 

arrangement may not exist (Guthery 1999). 

Brush management may increase connectivity 

and reduce habitat fragmentation for 

grassland-adapted species; conversely, brush 

management may fragment habitat of 

shrubland or woodland adapted species if the 

cleared areas limit wildlife movement between 

tracts of woody vegetation. Patches of habitat 

for a wildlife species should be linked by 

corridors that facilitate movements among 

habitat patches (Bennett 2003). Ensuring 

that connectivity exists among habitat patches 

should be a priority when vegetation is 

manipulated. 

Improving Food. Brush management 

may improve food accessibility, quality, 

and quantity for some wildlife species 

or functional groups (e.g., grazers) but 

reduce it for others (e.g., browsers). A 

review of publications in the Journal of 

Range Management, Rangeland Ecology 

& Management, and Ecology and articles 

emerging from a search of BIOONE, JSTOR, 

Science Direct, and Springer using the 

search strings “brush management,” “brush 

management wildlife,” “herbicides birds,” 

“brush control deer,” “brush control prairie 

chicken,” and “brush control sage grouse” 

yielded 50 articles addressing 59 cases of 

effects of brush management treatments or 

combinations of treatments (e.g., fire and 

herbicides) on wildlife food plants. Effects 

on food plants ranged from positive (53%) 

to neutral (32%) to negative (16%). In most 

cases, negative responses occurred where brush 

management reduced mistletoe (a parasitic 

plant on honey mesquite that is eaten by 

Mechanically clearing juniper 

in strips provides edge and 

brush piles for wildlife, forage 

for livestock and opportunities 

for future use of prescribed fire 

as a management tool. (Photo: 

Kirk McDaniel) 

140 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	 	 	

	 	 	 	 	 	

	 	 		 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

Plant communities with mixtures 

of herbaceous- and shrub-dominated patches provide excellent 

habitat for a diversity of game 

and non-game species. (Photo: 

Tim Fulbright) 

deer), reduced browse plants preferred by 

white-tailed deer, or increased thorns or 

secondary compounds in browse regrowth. 

In the review, we considered treatment effects 

to be neutral when they resulted in only 

temporary (<3 yr) increases in forb seeds or 

insects. Chemical, mechanical, and pyric 

brush management methods vary in their 

impact on woody and herbaceous food for 

wildlife. Chemical treatments, for example, 

tend to cause a temporary reduction in 

forbs, whereas fire may stimulate growth and 

abundance of early successional forbs that 

benefit many species of animals (e.g., Fig. 

13) (Beasom and Scifres 1977; Bozzo et al. 

1992a). Fire may top kill WPs, encouraging 

production of palatable sprouts (Schindler 

et al. 2004b). Anticipated benefits of brush 

management to wildlife should be predicated 

on the brush management approach to be 

used and the wildlife species potentially 

affected. 

Endangered Species. Brush management 

potentially reduces habitat for endangered 

species that depend on WP communities, 

such as ocelots (Felis pardalis), which need 

woodland with >97% canopy cover, or pygmy 

rabbits (Brachylagus idahoensis), which forage 

primarily on big sagebrush (Green and Flinders 

1980; Harveson et al. 2004). Conversely, brush 

management potentially could improve habitat 

for grassland-adapted species, such as the 

Attwater’s greater prairie chicken (Tympanuchus 

cupido attwateri). Documentation of the effects 

of brush management on habitat of species 

listed as endangered in the United States is 

lacking, however. 

Herbicide Toxicity. Herbicides used in 

rangeland brush management are usually not 

used in concentrations harmful to wildlife 

and dissipate from the ecosystem following 

the growing season they are applied (Scifres 

1977; Freemark and Boutin 1995; Guynn et 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 141 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	

	 	 	 	

Wyoming big sagebrush 

mowed in strips creates a 

mosaic of treated and untreated 

sagebrush habitat to increase 

diversity and maintain critical 

habitat for sagebrush-obligate 

wildlife. (Photo: K. W. Davies) 

al. 2004). Herbicides are generally not acutely 

toxic to soil organisms (Freemark and Boutin 

1995). Certain aspects of herbicide toxicity to 

wildlife, such as the role of surfactants and inert 

ingredients, and possible synergistic effects of 

multiple chemicals applied simultaneously are 

unknown (Guynn et al. 2004). Herbicides may 

negatively affect insects directly or indirectly, 

but little is known of the effects of rangeland 

herbicides on these organisms. A better 

understanding is needed since native rangelands 

may serve as a reservoir of pollinator and 

predator insects important to crop production 

in nearby cultivated areas (Freemark and 

Boutin 1995). In addition, invertebrates are 

a critically important food resource for many 

grassland bird species (O’Leske et al. 1997). 

Research on herbicide effects on reptiles and 

amphibians is also lacking (Freemark and 

Boutin 1995; Guynn et al. 2004). 

Although rangeland herbicides are generally 

not highly toxic to wildlife, acute effects of the 

herbicide 2,4-D have been documented. The 

herbicide is toxic to cutthroat trout (Salmon 

clarkia) (Woodward 1982). Spraying 2,4-D 

dramatically reduced pocket gopher (Thamomys 

talpoidis) populations in Colorado (Keith et al. 

1959). 

Predators. Anticipated benefits of brush 

management stated by NRCS focus largely 

on forage production and habitat structure 

for herbivores; however, brush management 

also alters predator habitat and may change 

behavioral responses of prey. Ungulates, for 

example, may use cleared patches within 

woodland or shrubland because of enhanced 

visual detection of predators (Bozzo et al. 

1992b). Florida panthers (Felis concolor 

coryi) are attracted to recent prescribed burns 

where prey species such as white-tailed deer 

congregate (Dees et al. 2001). Landscapelevel reduction of brush may remove 

perching structures important for raptors 

and increase susceptibility to nest predators. 

Prickly pear (Opuntia spp.) control, for 

example, has the potential to reduce nest 

sites and increase nest susceptibility to 

predators for bird species that prefer nesting 

142 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

in prickly pear. Treating prickly pear with 

herbicides, however, did not reduce nesting 

success of bobwhites in central Texas 

(Hernandez et al. 2003). Prey population 

densities may also change in response to 

brush management. Effects of mechanical 

brush management on the mortality of small 

mammals and immobile wildlife species at 

the time of treatment are unknown. Habitat 

changes following treatment may have 

unintended consequences, such as favoring 

increased prey densities. For example, cotton 

rat (Sigmodon hispidus) densities were six 

times greater on root-plowed rangeland in 

Texas than in untreated rangeland (Guthery 

et al. 1979). Rodent populations are strongly 

cyclical. Flushes in rodent abundance may be 

followed by increases in predator abundance; 

but subsequent abrupt declines in rodent 

populations may cause the now-abundant 

predators to shift to a prey base of livestock 

or ungulates such as white-tailed deer. 

Brush management may also affect visual 

cues used by predators to locate prey. Logged 

areas in the boreal forests of Canada have less 

debris on the forest floor than uncut stands. 

Efficiency of predation by martens (Martes 

americana) is greater in uncut timber stands 

because coarse woody debris act as sensory 

cues and enhance hunting success (Andruskiw 

et al. 2008). Brush management may likewise 

affect structure and amounts of woody debris 

in shrubland habitats, potentially affecting 

predator efficiency. Herbicide application may 

have little influence on habitat use by coyotes 

(Canis latrans) and bobcats (Felis rufus) 

possibly because standing woody material 

remains after treatment and herbaceous 

community structure is not drastically altered 

(Bradley and Fagre 1988). 

Treatment Longevity. Brush management 

initially reduces shrub canopy cover, but 

over time, stem and foliage cover returns. In 

Texas, the estimated duration of treatments 

range from 10 yr to 20 yr for root plowing 

and from 3 yr to 9 yr for roller chopping 

(Fulbright and Taylor 2001; Schindler and 

Fulbright 2003). Potential benefits of brush 

management for wildlife, therefore, are 

transient. Brush management, for example, 

may benefit a wildlife species initially, but as 

the WP community reestablishes (e.g., Fig. 

10), benefits may be lost. The temporary 

nature of treatments and the need for followup treatments must therefore be explicitly 

considered in statements of anticipated 

benefits (see the previous sections “Integrated 

Brush Management Systems” and “Treatment 

Options”). 

Single applications of mechanical brush 

management with no follow-up treatments 

may adversely impact wildlife habitat. For 

example, density of WPs palatable to whitetailed deer may be lower in WP communities 

that reestablish following root plowing than 

in undisturbed communities (Fulbright and 

Beasom 1987). Density of woody legumes 

such as honey mesquite and huisache may be 

greater on root plowed areas than on untreated 

areas >17 yr posttreatment (Fulbright and 

Beasom 1987; Ruthven et al. 1994). WPs 

that regenerate following roller chopping may 

have longer and more numerous spines than 

undisturbed plants, which could reduce bite 

rate of browsers (Schindler and Fulbright 2003; 

Schindler et al. 2004a). 

Measuring Habitat Improvement. The 

statement of anticipated benefits of brush 

management to wildlife is based on the 

assumption that improvements in food, 

cover, space, imbalance among populations, 

and fragmentation are evidence of habitat 

improvement. Vegetation characteristics are 

commonly linked with habitat quality in the 

wildlife literature (Guthery 1997; Hall et 

al. 1997; but see Johnson 2007). However, 

increases in a specific habitat characteristic 

do not constitute improvement if that 

characteristic is not limiting to wildlife 

(Guthery 1997). For example, rangeland 

disking may increase abundance of seedproducing forbs. However, seeds may not 

be limiting to northern bobwhites (Guthery 

1997). In this case, the assumption that 

increasing food (e.g., seed-producing forbs) 

resulted in habitat improvement may not be 

valid. Further, numerous confounding factors 

exist in natural ecosystems, and an increase in 

food and cover alone may not result in habitat 

improvement if some other factor, such as 

nesting cover, is limiting. 

Brush management is assumed to have 

improved wildlife habitat quality in an area 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 143 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 		 	 	

Brush management may improve food accessibility, quality, 

and quantity for some wildlife 

species or functional groups but 

reduce it for others. (Photo: Tim 

Fulbright) 

if it results in greater food abundance, better 

interspersion of plant communities, and 

habitat requirements, less fragmentation, or 

better cover characteristics. An underlying 

assumption is that population density in 

an area increases with increasing habitat 

quality (Guthery 1997). However, increased 

densities following brush management 

does not necessarily indicate sustained 

improvement in habitat. Treated areas may 

provide resources needed by an organism 

only during part of the year, and untreated 

areas may be needed to meet needs during 

other times of the year. White-tailed deer, 

for example, do not exhibit preference for a 

particular level of woody canopy cover during 

winter, but during summer, deer densities 

increase with increasing WP cover, with areas 

>80% canopy cover receiving greatest use 

(Steuter and Wright 1980). Improvements in 

habitat quality should be expressed in terms 

of increased survival and reproduction in 

addition to increased population densities and 

availability of key habitat components (Van 

Horne 1983; Hall et al. 1997; Crawford et 

al. 2004). For northern bobwhites, evidence 

that their abundance increases with habitat 

quality variables such as food supplies 

and interspersion is limited and equivocal 

(Guthery 1997). Instead, abundance of 

bobwhites is proportional to the amount 

of usable space (habitat for which a species 

is fully adapted), and only practices that 

increase the abundance of usable space are 

likely to improve bobwhite numbers (Guthery 

1997; Guthery et al. 2005). The usable space 

concept has also been applied to white-tailed 

deer management (Hiller et al. 2009). 

Demographic characteristics of wildlife 

populations and usable space are more 

difficult and time consuming to quantify 

than habitat characteristics such as food 

production. As a result, comparisons of survival 

144 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

and reproduction of wildlife on sites with 

and without brush management are limited 

(Appendix II). Consequently, restricting 

statements of anticipated benefits to treatments 

and species for which increased reproduction, 

survival, and density or increases in usable 

space resulting from brush management have 

been documented is impractical. A better 

approach would be to acknowledge that while 

brush management may improve various 

habitat properties, its impact on habitat quality 

for many species is unclear. 

Fuels Management 

Brush management is increasingly being applied 

in shrubland and woodland settings to reduce 

fire risk or create fuel breaks (Keeley 2002; 

Davies et al. 2009b); however, little information 

is available to evaluate its effectiveness. In forest 

systems, mechanical brush management alters 

fuel characteristics and influences fire behavior 

(Kane et al. 2009); however, current fire models 

have not yet been parameterized to represent 

these modified behaviors. Although the impact 

of brush management on fire characteristics and 

spread are unclear, fire suppression efforts can be 

facilitated simply by reducing fuel height (Keeley 

2002). However, while brush management can 

effectively reduce the mass and continuity of 

canopy fuels, it may promote production and 

continuity of fine surface fuels (e.g., grasses) 

and thus promote fire risk (Keeley 2002; 

Perchemlides et al. 2008; Huffman et al. 2009). 

RECoMMEndAtIonS 

•	� Care is needed when using words 

and phrases such as “vigor,” “health,” 

“biodiversity,” “encouraging growth,” and 

“suitable” when projecting the effects of 

brush management. These terms are vague 

or ill-defined and often value laden and 

should be replaced with words and phrases 

that refer to specific and tractable metrics 

to define more specific and measurable 

conservation outcomes. 

•	� Integrated Brush Management Systems 

have proven effective in WP management 

and are likely to yield the greatest 

conservation benefits. Brush management 

is a long-term commitment. Adaptive 

management, coordination with grazing 

management, a plan and funding for 

follow-up restoration and brush treatments, 

and periodic monitoring are essential. 

Emphasize flexibility and objectivity. 

•	� Customize brush management 

prescriptions according to the stakeholder’s 

vision and management objectives and 

the inherent capability or limitations of 

the ecological site. This perspective on 

human dimensions should be incorporated 

into the list of purposes in practice code 

314: “Work closely and cooperatively 

with clientele to apply brush management 

practices that meet both land and personal 

conservation objectives.” 

•	� Evaluate and define when, where, how, 

and under what circumstances brush 

management should be undertaken and 

what specific outcomes are to be attained. 

Recommendations should be thoroughly 

vetted and justified. Do not assume that 

brush management is needed simply 

because shrubs are present. 

•	� Tailor statements of potential hydrological 

benefits of brush management to specific 

bioclimatic zones. 

•	� Anticipated effects of brush management 

should take into account the extent to 

which habitat heterogeneity is important 

for wildlife species. Do not assume 

that brush management will result in 

improvement of habitat for a wildlife 

species or functional group. Tailor 

statements of anticipated benefits of 

brush management to specific habitat 

variables or characteristics, such as food 

production, and to specific wildlife 

species or functional groups. State which 

wildlife species or functional groups 

may be negatively impacted by brush 

management under specific sets of 

circumstances. 

•	� Develop and maintain a relational database 

to evaluate brush management treatments. 

Important information may include 

treatment approaches and longevities; 

location and spatial pattern(s) of treatment 

in relation to soils and topography; preand posttreatment soil, plant, livestock, 

and wildlife responses; environmental 

conditions; and predicted trade-offs and 

outcomes based on published literature 

(Table 6). This database should be updated 

as new information becomes available and 

used to communicate anticipated benefits 

for specific locations and regions. 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 145 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

	 	

tABlE 6. Example of a matrix approach to communicating anticipated benefits of brush management for wildlife. A similar matrix could be 

developed for plants, soils, and so on. 

Brush management 

approach Scale Climate 

Existing woody 

canopy cover (%) Wildlife species or group 

Anticipated 

impact1 

Mechanical Landscape Humid 60–100 Grassland obligates + 

Woodland obligates − 

Edge-associated species − 

Habitat generalists 0 

25–59 Grassland obligates + 

Woodland obligates − 

Edge-associated species − 

Habitat generalists 0 

<25 Grassland obligates + 

Woodland obligates − 

Edge-associated species − 

Habitat generalists 0 

Chemical, fire Mosaic, patch Subhumid, semiarid, arid <25 Habitat generalists 

1+, positive; 0, neutral; −, negative. 

•	� Seeding of nonnative plants following 

brush management should be avoided, 

but if considered, it should be explicitly 

justified. 

•	� Articulate and critically evaluate 

positive and negative trade-offs in brush 

management impacts on various ecosystem 

goods and services. For example, gains 

in livestock production and herbaceous 

diversity accruing from brush management 

may be at the expense of ecosystem carbon 

sequestration. 

•	� Develop a mechanism to integrate 

conservation planning on individual 

properties into and consistent with 

local/regional conservation plans. 

Specific goals and objectives from brush 

management may vary by ownership 

and agency, but by pooling expertise and 

financial resources, there will be better 

opportunities for treating and restoring 

larger areas. 

KnoWlEdgE gAPS 

•	� The extent to which pre–brush 

treatment management conditions 

drive posttreatment responses is largely 

unknown, as are the effects of follow-up 

treatments. 

•	� Projected effects of brush management 

mention numerous variables related to 

air quality as “not applicable.” However, 

available information, albeit scant, suggests 

that changes from grass to WP dominance 

can significantly increase emissions of 

trace gases and volatile organic carbon 

compounds and the production of dust, 

aerosols and allergens. The extent to which 

brush management might reverse these 

is unknown, as are the implications for 

human health, tropospheric chemistry, and 

land surface–atmosphere interactions. 

•	� ANPP can be dramatically enhanced by 

shrub encroachment (Knapp et al. 2008a; 

Barger et al. 2011), but the effects of 

brush management on ANPP are largely 

unknown. Plant production responses 

to brush management have focused on 

the herbaceous vegetation, and there 

is scant data on WP ANPP during the 

postmanagement period. Thus, we are ill 

equipped to evaluate brush management 

from a carbon-accounting perspective. 

•	� The belowground organic carbon 

pool (roots+soil) typically dwarfs 

the aboveground pool in rangeland 

ecosystems. Robust generalizations as to 

how WP encroachment (Fig. 4) and brush 

management affect this large belowground 

146 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	 	 	

	 	

	 	 	 	

	 	 	 	 	

	 	 	

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

pool are not yet possible. Studies that if we are to advance our ability to 

have quantified soil responses to brush comprehensively evaluate the conservation 

management are few (Fig. 8) and have value of brush management. Brush 

relied on comparing random samples from management has the potential to modify 

a treated site(s) to a nearby, untreated site. the provisioning of numerous ecosystem 

Given the extensive edaphic heterogeneity services at both local and regional scales. 

on shrub-encroached rangelands (e.g., Attempts must be made to monitor and 

Bai et al. 2008; Liu et al. 2010), such value these nontraditional nonmarket 

coarse comparisons are probably not too services. 

reliable. Studies quantifying soil resources • Many of the potential benefits of brush 

in a spatially explicit fashion before management depend on the extent to 

and following brush management are which herbaceous production and ground 

sorely needed, as are studies quantifying cover can be reestablished and the duration 

the response of shrub roots to brush of the herbaceous response. General 

management. Decreases in plant and SOC models of WP effects on herbaceous 

pools that may occur following brush vegetation (Fig. 3) need to be better 

management could have important but quantified to determine when it might 

as yet poorly understood implications for be most effective to implement brush 

• 

ecosystem carbon management. 

Quantification of trade-offs between 

livestock production, hydrology, erosion, 

carbon sequestration, biodiversity, and so 

on and approaches for weighting them is 

a current challenge that must be addressed 

management, and conceptual models 

of posttreatment herbaceous vegetation 

response to brush management (Fig. 10) 

need to be made operational to obtain 

quantitative ecological (Fig. 9) and 

socioeconomic (Fig. 11) assessments of 

Woody plant encroachment represents a threat to 

grassland, shrub-steppe, and 

savanna ecosystems and the 

plants and animals endemic to 

them. (Photo: Tim Fulbright) 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 147 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	

	 	 	 	

	 	 	

	 	 	 	

	 	 	 	

	 	 	

	 	

brush management. Simulation modeling 

has been underutilized (Fig. 8). Given 

the advent of inexpensive, user-friendly 

software for personal computers, this 

tool can now be readily used to integrate 

existing information for assessment, 

scenario development, and forecasting (e.g., 

Grant et al. 1999; Fuhlendorf et al. 2008). 

•	� The major knowledge gap related to 

brush management and water is our 

limited understanding of landscape-level 

implications. With the exception of a few 

studies (e.g., Collings and Myrick 1966; 

Wilcox et al. 2008a), there has been 

little documentation of the large-scale 

impacts of brush management on water 

and erosion processes. As a result, there is 

considerable uncertainty concerning the 

efficacy of extrapolating from fine-scale 

studies to the landscape level (Wilcox and 

Huang 2010). 

•	� Biodiversity responses to shrub 

encroachment are poorly documented, 

and responses to brush management 

have focused largely on herbaceous 

vegetation. Responses of various faunal 

groups, including soil biota, are few and 

scattered. The implications of changes in 

biodiversity for ecosystem function have 

been the topic of much discussion in the 

research community but remain poorly 

understood. 

•	� Brush management effects on wildlife 

have focused mainly on game species, 

particularly white-tailed deer, northern 

bobwhites, and sage-grouse. Nongame 

species, including predators, passerines, 

small mammals, and reptiles, have been 

largely neglected. Habitat requirements 

of many nongame species are not well 

understood, making it challenging to 

even speculate about effects of brush 

management. These gaps must be filled 

for statements of anticipated benefits to 

be made for specific species or functional 

groups. 

• The extent to which brush managementinduced changes in habitat attributes 

translate into improvements in carrying 

capacity and animal birth rates, longevity, 

nutritional status and body mass are largely 

unknown. 

•	� Further research that addresses the 

interrelationship between brush 

management and fire behavior is needed 

to provide robust conclusions on its 

effectiveness for reducing fire risk and 

spread. Trade-offs between reducing 

WP canopy mass and continuity and 

promoting fine fuel production needs 

further study among different WP 

communities. 

•	� A framework for conceptualizing how 

climate change, invasions of nonnative 

species, and increases in atmospheric CO2 

and nitrogen deposition might influence 

future grass–woody states and ecosystem 

responses to brush management is 

needed. 

ConCluSIonS 

Successful long-term management programs 

(typically >5 yr) usually involve an integrated 

Rangeland conservation goes 

beyond traditional concerns 

of livestock production to 

include potential effects on a 

variety of ecosystem services. 

The research community is 

challenged with measuring 

and monitoring these varied 

impacts; and the management 

community with creating or 

maintaining woody-herbaceous 

mixtures in arrangements that 

satisfy competing objectives. 

(Photo: Tim Fulbright) 

148 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

brush management systems and restoration 

approach that includes a suite of mechanical, 

fire, biological, and chemical methods. A 

combination of methods customized for 

local ecological site conditions is particularly 

important when the primary objective is to 

achieve long-term native plant stability that 

supports conservation and resource function. 

Assessing revegetation potential is a critical 

first step before proceeding with brush 

management. Brush management and 

revegetation costs are high, and careful 

selection of areas with a high potential 

for reestablishment is necessary for longterm, sustainable brush management. In 

many situations, herbaceous vegetation 

on treated areas will recover naturally after 

brush management without revegetation. 

In other situations, planting or seeding of 

grasses or forbs may be necessary. Sites with 

particularly dense brush cover, poor hydrologic 

integrity, or related conditions may have 

limited revegetation potential. An in-field 

evaluation and soil survey should always be 

used to evaluate soil and other factors that 

will ultimately influence replacement of the 

vegetation community. With these caveats in 

mind, our synthesis suggests the following 

conclusions regarding the conservation value of 

brush management: 

•	� Conservation of grasslands and savannas 

as ecosystem types and the plants and 

animals endemic to them should be a 

high priority (Fig. 19). Loss of grasslandobligate organisms occurs with shrub 

encroachment, even if overall numerical 

biological diversity is enhanced or 

unaffected. Brush management programs 

are essential to maintain grassland, 

steppe, and savanna ecosystems and 

the biodiversity and services they 

provide. Progressive brush management 

protocols will be required to achieve this 

conservation goal in many instances. 

•	� Herbaceous cover, production, and 

diversity are typically enhanced by brush 

management. However, exceptions 

occur, and the possibility for deleterious 

outcomes should always be anticipated and 

considered when planning. Furthermore, 

treatment longevity will vary, so plans for 

follow-up are required. 

•	� Returns arising from improved livestock 

performance and production are 

important, but benefits beyond livestock 

production are being increasingly 

recognized. When the value of ecosystem 

goods and services beyond those associated 

with livestock production are taken into 

account, a more favorable picture of brush 

management begins to emerge. 

•	� Although frequently justified on the 

basis of benefits to water quality and 

quantity, brush management does not 

necessarily produce the hydrological 

benefits that are commonly attributed 

to it. In most cases, these perceived 

benefits are exaggerated and have not 

been documented, and there is little or 

no evidence that brush management is a 

viable strategy for increasing ground water 

recharge or stream flows at meaningful 

scales. Outcomes depend on the vegetation 

type and geological setting. In some cases, 

depending on the vegetation community 

and the method of shrub management, 

surface runoff and erosion may actually 

increase. Local/regional knowledge should 

therefore guide brush management 

prescriptions with respect to hydrological 

impacts. In settings where winter 

precipitation predominates or where 

WPs are accessing deep stores of water, 

there is the potential to use vegetation 

management to enhance groundwater 

recharge and stream flow. However, 

FIguRE 19. Brush encroachment threatens habitat for 

grassland-obligate species 

such as this savanna sparrow. Brush management 

may be required to generate and maintain shrub 

cover amounts and patterns 

within acceptable limits for 

such species. (Photo: Tim 

Fulbright) 

CHAPTER 3: Brush Management as a Rangeland Conservation Strategy: A Critical Evaluation 149 
	 	 	 	

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	

	 	 	 	

	 	 	

A burned (left) and untreated 

(right) mountain big sagebrush 

plant community on the Hart 

Mountain National Wildlife 

Refuge in southeastern Oregon. 

(Photo: K. W. Davies) 

projections for how this translates to 

watershed- and regional-scale hydrology is 

based more on speculation than data. 

•	� Statements that brush management 

maintains or enhances wildlife habitat are 

oversimplifications. Habitat requirements 

of many nongame species are poorly 

understood, making it challenging to 

even speculate about effects of brush 

management on these organisms. Clearer 

definitions of what constitutes a benefit of 

brush management to wildlife are needed, 

and these should be tailored to species or 

functional groups. Statements should focus 

on the habitat characteristics or attributes 

that are anticipated to be improved. 

Technology and the tools available for 

brush management are dynamic and ever 

changing. Keeping educated and up to date 

on new developments is paramount. There are 

knowledge gaps in brush management, but there 

always will be, and it is important that managers 

strive to use the best available information. In 

some instances, practices applied and approaches 

followed to manage a particular WP species may 

150 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

S. R. Archer, K. W. Davies, T. E. Fulbright, K. C. McDaniel, B. P. Wilcox, and K. I. Predick 

not be known. Thus, it is recognized that land 

managers are often placed in situations where 

they must exercise flexibility, responsibility, 

and their best professional judgment when 

developing a planning strategy and carrying out 

an action program. 

Brush management presents a series of 

dilemmas and challenges as a response to 

WP encroachment. The recognition that 

WP proliferation can substantially promote 

ecosystem primary production and carbon 

stocks may trigger new land use drivers as 

industries seek opportunities to acquire and 

accumulate carbon credits to offset CO2 

emissions. WP proliferation in grasslands 

and savannas may therefore shift from being 

an economic liability in the context of 

livestock production to a source of income 

in a carbon sequestration context. Policy and 

management issues related to grazing land 

conservation thus extend well beyond the 

traditional concerns of livestock production 

and game management (wildlife valued for 

sport hunting) to include potential effects on 

hydrology, carbon sequestration, biological 

diversity, atmospheric chemistry, and the 

climate system. The research community is 

challenged with quantifying and monitoring 

these varied impacts and the management 

community with devising approaches for 

creating or maintaining woody–herbaceous 

mixtures in arrangements that satisfy 

competing conservation objectives. 

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landscapes. Landscape Ecology 20:733–742. 

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APPEndIx I. herbaceous Response to Brush Management 

A search of articles with the key word “brush control” in the Journal of Range Management and 

Rangeland Ecology & Management at http://www.uair.arizona.edu/search?page_set = 51 

yielded 1277 separate articles. Among these, about 80% (1021 articles) either assessed target 

plant mortality (764 articles) or described changes in herbaceous (grass and broadleaf) plant 

abundance (e.g., cover, biomass, and frequency; 257 articles). Of the 257 articles reporting 

on herbaceous responses, 216 (84%) characterized the response as positive, 21 (8%) reported 

no change, and 20 (8%) report a negative response to brush management. 

In another, more directed search, we sampled published accounts of how brush management 

influences herbaceous vegetation. Web of Knowledge searches resulted in 532 unique 

references, 36 of which were field studies conducted on rangelands in the United States and 22 

of which measured the response of herbaceous or grass production. 

Among these 22 studies, herbicide was the most frequently assessed brush management 

technique (15 studies, or 68%). As with our initial, broader survey, most of these (18 studies, 

or 82%) reported increases in herbaceous production. The majority of experiments were 

conducted over short periods of time, with only eight studies (36%) lasting more than 5 yr and 

only five (23%) lasting longer than 10 yr. 

APPEndIx II. Brush Management and Wildlife habitat Quality 

•	 Peer-reviewed publications were surveyed to determine the proportion of studies that 

measured effects of brush management on wildlife density. 

•	 A total of 97 publications emerged in this compilation, which included articles in the 

Journal of Range Management, Rangeland Ecology & Management, and Ecology, along 

with those emerging using the search strings “brush management,” “brush management 

wildlife,” “herbicides birds,” “brush control deer,” “brush control prairie chicken,” “brush 

control sage grouse,” “fire sagebrush,” “sage grouse prescribed fire,” and “prescribed fire” 

in the search engines BIOONE, JSTOR, Science Direct, and Springer. 

•	 Only 45% of these articles reported some measure of organism abundance in response to 

brush management. 

•	 Only about 5% reported the demographic information that Van Horne (1983) and Hall et 

al. (1997) suggest as necessary to assess habitat quality.	 

APPEndIx III. Citations for data Points in Figures 4, 9, and 13 

Data points in Figure 4 are from the following: 1–3 = Schlesinger and Pilmanis (1998); 

4–5 = Asner et al. (2003); 6, 8–12, 17 = Geesing et al. (2000); 7 = Hughes et al. (2006); 

13–15 = Boutton et al. (1998); 16 = Tilman et al. (2000); 18–19 = Mordelet et al. (1993); 

20 = San Jose et al. (1998); 21–24 = Wheeler et al. (2007); and 25–34 = Jackson et al. (2002). 

Data points in Figure 9 are from Ansley et al. (2006), Bedunah and Sosebee (1984), Clary 

(1971), Griffith et al. (1985), McDaniel et al. (1982), Morton et al. (1990; mechanical 

treatments), Augustine and Milchunas (2009), Bates et al. (2005, 2009), Cable (1967), Engle 

et al. 1993, 1998), Teague et al. (2008b; prescribed fire), Bedunah and Sosebee (1984), 

McDaniel et al. (1982), Morton et al. (1990; herbicides), and Engle et al. (1993; multiple 

treatments). 

Data points in Figure 13 are from Baeza and Vallejo (2008), Davies et al. (2007), Edwards et 

al. (2007), Maccherini et al. (2007), Maron and Jefferies (2001), Nolte and Fulbright (1997), 

Nolte et al. (1994), Olson and Whitson (2002), Page et al. (2000), Ponzio et al. (2006), 

Ruthven et al. (1993), Ruthven and Krakauer (2004), and Sheley et al. (2006). 

170 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	

	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

CHAPTER 

4 

Assessment of Range Planting 

as a Conservation Practice 

Stuart P. Hardegree,1 Thomas A. Jones,2 Bruce A. Roundy,3 

Nancy L. Shaw,4 and Thomas A. Monaco5 

Authors are 1Plant Physiologist, USDA Agricultural Research Service, 

Northwest Watershed Research Center, Boise, Idaho 83712, USA; 

2Research Geneticist, USDA Agricultural Research Service, Forage and 

Range Research Laboratory, Logan, UT 84322, USA; 3Professor, Department 

of Plant and Wildlife Sciences, Brigham Young University, Provo, UT 84602, 

USA; 4Research Botanist, USDA Forest Service, Rocky Mountain Research 

Station, Boise, ID 83702, USA; 5Research Ecologist, USDA Agricultural 

Research Service, Forage and Range Research Laboratory, 

Logan, UT 84322, USA 

Reference to any commercial product or service is made with the understanding 

that no discrimination is intended and no endorsement by USDA is implied 

171
 
	 	 	 	

	 	 	 	

	 	 	 	

	 	 	

	 	

	 	 	 	

The goal of a 

conservation 

“

management plan is to 

transition an existing 

undesirable plant 

community to a more 

desirable state.”

172 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	

	 	 	 	

	 	 	

	 	 	

4 Assessment of Range Planting as 

a Conservation Practice 

Stuart P. Hardegree, Thomas A. Jones, Bruce A. Roundy, Nancy L. Shaw, 

and Thomas A. Monaco 

IntRoduCtIon 

The Range Planting Conservation Practice 

Standard is used to inform development 

of Natural Resource Conservation Service 

(NRCS) management recommendations 

for improving vegetation composition and 

productivity of grazed plant communities. 

Range planting recommendations are 

generally implemented within an integrated 

conservation management system in 

conjunction with related conservation 

practices such as brush management, 

prescribed burning, prescribed grazing, 

herbaceous weed control, and upland wildlife 

habitat management. The Range Planting 

Standard is defined as “establishment of 

adapted perennial or self-sustaining vegetation 

such as grasses, forbs, legumes, shrubs and 

trees.” The six specific purposes of this 

standard are to: 

•	� Restore a plant community similar to 

the Ecological Site Description reference 

state for the site or the desired plant 

community. 

•	� Provide or improve forages for livestock. 

•	� Provide or improve forage, browse, or 

cover for wildlife. 

•	� Reduce erosion by wind and/or water. 

•	� Improve water quality and quantity. 

•	� Increase carbon sequestration 

Additional conservation effects associated with 

related conservation practices include reduction 

of negative weed impacts and reduction of 

wildfire hazard. Range planting conservation 

practices apply where desirable vegetation is 

below the acceptable level for natural reseeding 

to occur, or where the potential for vegetation 

enhancement by grazing management is not 

satisfactory. 

Range planting was implemented on 517 000 

ha of grazing land in the 17 conterminous 

western states in the 5-yr period, 2004–2008. 

This is a relatively small area compared to 

implementation of some other conservation 

practices within the region over the same 

period: prescribed grazing, 31 360 000 ha; 

upland wildlife habitat management, 19 166 

000 ha; herbaceous weed control, 7 603 000 

ha; brush management, 1 457 000 ha; and 

prescribed burning, 371 000 ha. Conservation 

cover, a closely related conservation practice for 

reducing erosion on retired cropland as part 

of the Conservation Reserve Program (Young 

and Osborn 1990), was implemented on over 

1 600 000 ha during the same period. 

Site-specific conservation management plans 

are developed for areas where existing plant 

community attributes are insufficient to meet 

management goals for productivity or species 

composition, and where natural recovery 

toward a more desirable state is not expected. 

The goal of a conservation management plan 

is to transition an existing undesirable plant 

community to a more desirable state. It is 

assumed that successful implementation of this 

change in state will be associated with specific 

conservation effects. 

The National Standard for range planting is 

usually modified at the state level with specific 

recommendations of regional or local relevance. 

State standards, however, retain the same 

general guidance, and usually vary only in the 

degree to which they include more detailed 

recommendations extracted from regionspecific NRCS technical notes and seeding 

guides. Recommendations for the following 

management elements are common to both 

the national and state range planting standards: 

selection of appropriate plant materials, 

Rangeland drill used for 

site preparation and seed 

placement. (Photo: USFS, 

2006) 

CHAPTER 4: Assessment of Range Planting as a Conservation Practice 173 
	 	 	 	

	 	 	

	 	 	 	

	 	 	

USDA-ARS Plant material 

selection trials, Beaver, UT 

(Photo: Craig Rigby, 2009) 

seed-bed preparation, planting methods, 

seeding depth, seeding rate, time of seeding, 

postplanting management, and weed control. 

The spatial domain of interest in this evidencebased assessment includes rangeland systems 

in the Great Plains, Intermountain West, 

southwestern desert, and interior-California 

hydroclimatic zones (Barbour and Billings 

2000). These areas are characterized by 

different vegetation types, management 

priorities, and climatic syndromes that also vary 

internally along both latitudinal and elevational 

gradients (Natural Resources Conservation 

Service 2006). Range planting issues common 

to all areas are a generally arid or semiarid 

climatology, and high annual and seasonal 

variability in weather and climate. These areas 

are also commonly under pressure from highly 

competitive annual and perennial weeds or 

expanding populations of native woody plants. 

The success of specific conservation practice 

recommendations and the potential ecological 

outcomes realized are both highly dependent 

upon ecological site characteristics, the 

initial degree of deviation from desired site 

characteristics, and weather, all of which are 

highly variable in both time and space. An 

important implication of this high variability, 

in both initial establishment and laterseral processes, is that virtually no direct 

experimental evidence exists to link specific 

range planting conservation practices to 

conservation effects per se. The linkage is 

instead derived indirectly through evidence of 

174 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	 	

	 	

	 	

	 	

	 	 	 	 	

	 	 	

	 	 	 	 	

	 	 	

S. P. Hardegree, T. A. Jones, B. A. Roundy, N. L. Shaw, and T. A. Monaco 

the degree to which specific planting techniques 

have been shown to produce successful plant 

establishment, and evidence supporting the 

positive conservation effects of alternative 

vegetation states. We have, therefore, separated 

our assessment of the conservation effects 

of rangeland planting practices into two 

components: assessment of the direct benefits 

of specific planting techniques recommended 

in the range planting standard, and assessment 

of specific conservation effects of alternative 

vegetation states. The assessment of rangeland 

planting techniques involved a survey of 189 

range planting studies from the refereed journal 

literature. These studies were classified as to 

bioclimatic zone, initial plant community 

and type of disturbance, plant materials and 

seed-mix characteristics, seeding rate, site 

preparation and weed control methodology, 

planting depth, planting season, experimental 

design, weather, and relative success criteria. 

Summary statistics cited in this synthesis were 

derived from the survey. 

Assessment of the dIReCt BenefIts 

of RAnge PlAntIng PRACtICes 

The range planting standard specifically 

requires selection of plant materials that are 

adapted to both climate and microclimate 

as affected by soil type, landscape position, 

and range site characteristics. Gross climatic 

variability generally determines the historical 

complement of native species at a site, but also 

the suitability of introduced plant materials 

(Shown et al. 1969; Shiflet 1994; Vogel et 

al. 2005; Natural Resources Conservation 

Service 2006). Seedbed preparation and 

planting methods are designed to optimize 

microclimatic conditions for planted species, to 

increase the number of favorable microsites for 

germination and establishment, and to mitigate 

or control competition from undesirable species 

(Roundy and Call 1988; Call and Roundy 

1991; Sheley et al. 1996; Krueger-Mangold et 

al. 2006; Sheley et al. 2006). 

A major problem with synthesizing range 

planting research results is the high variability 

in metrics used to evaluate success. Relatively 

few authors have directly evaluated alternative 

criteria for quantification of success (Ries and 

Svejcar 1991). The majority of range planting 

studies use arbitrary, relative criteria for judging 

success, and only consider planting-year or 

first-year effects. Typical criteria for evaluating 

success generally involve measurements or 

ocular estimates of density, frequency, cover, 

and/or biomass. 

seleCtIon of PlAnt mAteRIAls 

Climatic Considerations 

Weather and climate patterns in western 

North America are highly variable in space 

and time (Rajagopalan and Lall 1998). The 

relationship between climate and both 

vegetation distribution and production 

on western rangelands is well documented 

(Barbour and Billings 2000; Natural Resources 

Conservation Service 2006). The general 

importance of climate is acknowledged in 

seeding guides in the form of tables that list 

species and cultivar suitability as a function 

of mean annual precipitation (Jordan 1981; 

Jensen et al. 2001; Lambert 2005; Ogle et al. 

2008a, 2008b). Seeding guides may also cite 

climatic thresholds below which active seeding 

practices are not recommended (Anderson et 

al. 1957; Jordan, 1981). Unfortunately, the 

microclimatic requirements for germination, 

emergence, and seedling establishment are 

much more restrictive than the longer-term 

climatic requirements for maintenance 

of mature plant communities (Call and 

Roundy 1991; Peters 2000; Hardegree et al. 

2003). Current state-and-transition models 

acknowledge that there are perhaps a limited 

set of potential trajectories for moving between 

undesirable and desirable vegetation states 

(Westoby et al. 1989; Batabyal and Godfrey 

2002; Bestelmeyer et al. 2003; Briske et 

al. 2003, 2005, 2006, 2008; Bashari et al. 

2008). Westoby et al. (1989) noted that many 

transition pathways between alternative states 

require the occurrence of a specific and perhaps 

infrequent series of climatic events. 

A B 

virtually no direct 

experimental 

evidence exists 

to link specific 

range planting 

conservation 

practices to 

conservation 

effects per se.”

A Native sagebrush/ 

bunchgrass plant 

community at Jacks Creek in 

southwestern Idaho (Photo: 

USFS) B Cheatgrass 

dominated site in southern Idaho 

after wildfire (Photo: USFS) 

CHAPTER 4: Assessment of Range Planting as a Conservation Practice 175 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	 	

	 	 	

	 	 	

...successful 

establishment 

was frequently 

associated with 

average or 

above-average 

precipitation for 

either the entire 

year, or during 

the season of 

establishment.”

The range planting literature is somewhat 

biased relative to inferences that can be drawn 

from plant-material/climate interactions. Less 

than 60% of the studies reviewed for this 

synthesis reported weather conditions during 

the study, and less than half of these studies 

were replicated for year effects. In studies 

that reported weather conditions, successful 

establishment was almost always associated 

with average or above-average precipitation 

for either the entire year, or during the season 

of establishment. This implies that climatic 

thresholds exist below which management 

actions have little effect on establishment 

success. These thresholds may vary for species 

with different establishment requirements, but 

previous studies have not been designed to test 

this hypothesis specifically. 

The strongest evidence for plant-material 

suitability for a given climatic region is derived 

from observation of historical relationships 

between species and climate, experiencebased observation, and long-term assessment 

of persistence of planted species (Harris and 

Dobrowolski 1986; Shiflet 1994; Barbour and 

Billings 2000; Natural Resources Conservation 

Service 2006). 

Plant-material development 

NRCS has developed relatively specific and 

detailed recommendations for suitability of 

plant materials for different site conditions, 

climatic zones, and management objectives 

(Ogle et al. 2008a, 2008b). Plant-material 

recommendations for both native and 

introduced species are based primarily on 

plant-materials discovery, screening, and 

breeding programs by NRCS Plant Materials 

Centers, and other government research and 

agricultural experiment station programs 

(Hafenrichter 1948; Stewart 1950; Harlan 

1951; Schwendiman 1956; Anderson et al. 

1957; Schwendiman 1958; Harlan 1960; 

Roundy and Call 1988; Alderson and 

Sharp 1994; Asay et al. 2003; Erickson et 

al. 2004). Selected or bred plant materials 

deemed to have superior productivity, 

vigor, establishment, disease resistance and/ 

or seed-production characteristics are then 

cultivated and released for development as 

commercial varieties (Schwendiman 1958; 

Johnson and Asay 1995; Asay et al. 2003). 

The more recent efforts in plant-material 

development and evaluation focus on selection 

for, or comparison of, specific ecological and 

physiological traits (Aguirre and Johnson 

1991b; Johnson and Asay 1995; Arredondo 

et al. 1998; Jensen et al. 2005). These 

efforts incorporate and report more detailed 

experimental design information, but are often 

based on relatively controlled experimental 

conditions in the laboratory, greenhouse, or 

an agricultural field environment (Arredondo 

et al. 1998; Jones et al. 2003). The majority 

of current plant-material recommendations 

are based on evaluations of field performance 

that are not accessible through refereed journal 

publications (Stewart 1950; Schwendiman 

1956; Great Plains Council 1966; Jensen et 

al. 2001; Lambert 2005; Ogle et al. 2008a, 

2008b). 

The literature documenting managementscale range planting is dominated by studies 

in which few inferences can be made about 

relative performance of different species and 

seed sources (Casler 1999). Very few studies 

are replicated in such a way that within- or 

between-species variability can be assessed 

(Kneebone and Cremer 1956; Pitman and 

Jaymes 1980; Asay and Johnson 1983b; 

Rumbaugh and Johnson 1986; Burner et al. 

1988; Asay and Johnson 1990; Kitchen and 

Monsen 1994; Asay et al. 1996; Casler 1999; 

Asay et al. 2001; Vogel and Jensen 2001; Jones 

et al. 2003; Robins et al. 2007). About 60% 

of the rangeland planting studies surveyed for 

this synthesis evaluated performance of either a 

single seed lot, or a unique seed mix. In studies 

that evaluated more than one seed lot of a given 

species, only 6% were replicated at the seed-lot 

level. 

seed Quality 

Seed-quality recommendations for range 

planting conservation practices are generally 

limited to those concerning germination 

testing, and the calculation of seeding rates 

based on estimates of pure live seed (PLS). 

Seed quality, however, has been evaluated in a 

number of studies that have correlated seed size 

and other morphological attributes to seedling 

emergence, growth rate, nutrient utilization, 

and seedling morphology and yield (Trupp 

and Carlson 1971; Carren et al. 1987a, 1987b; 

Limbach and Call 1995a, 1996; Smith et al. 

2003). 

176 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	

	 	 	 	 	

	 	 	

	 	 	 	

	 	 	 	

S. P. Hardegree, T. A. Jones, B. A. Roundy, N. L. Shaw, and T. A. Monaco 

Recurrent selection for increased seed size, 

deep-seeding emergence, or rapid seedling 

growth can result in plant materials with 

improved stand establishment. McKell 

(1972) and Kneebone (1972) recommended 

selection for seed mass to improve 

seedling vigor and establishment. McKell 

(1972) also emphasized the importance 

of rapid germination and pointed out 

that this is often a characteristic of weedy 

opportunistic grasses. Kneebone (1972) 

asserted that seed mass is highly heritable, 

and thus is a trait that will often be 

responsive to selection. 

For cross-pollinated species, genetic 

manipulation through artificial selection 

or hybridization may be used to develop 

plant materials that increase the likelihood 

of seeding success. Kneebone (1972) 

suggested selection for high seed mass with 

hand screens, air columns, or gravity tables. 

This may be combined with selection for 

rapid germination under stress conditions, 

coleoptile length, and deep-seeding 

emergence. Large numbers of seeds and 

seedlings can be easily screened, a feature that 

lends itself to genetic improvement. This 

type of genetic manipulation is unsuitable 

for self-pollinated species where natural 

outcrossing and within-population genetic 

diversity is limited. 

seedbed Preparation and 

Planting methods 

The following conservation practice 

recommendations are directly or indirectly 

related to microclimatic management of the 

seed bed: surface soil modification, microsite 

improvement, seeding depth, seeding rate, 

timing of seeding, and weed control. Seedbed 

microsite improvement can consist of 

operations designed to reduce water loss and/or 

adverse thermal conditions in the seed zone by 

improving infiltration into the soil, improving 

water availability to the seed, reducing water 

loss to the atmosphere or reducing plant 

competition for water. This is accomplished 

through initial mechanical disturbance, soil 

firming and surface modification, control 

of seeding depth, application of soil surface 

amendments, and weed control (Roundy and 

Call 1988; Sheley et al. 1996). 

surface modification 

Soil surface modification is often justified by 

expectations of increased water availability to 

the seed, either by improving seed–soil contact, 

reducing the amount of surface area subject 

to evaporation, increasing infiltration and 

water-holding capacity, or by creating specific 

microsites that either receive or retain water 

more effectively (McGinnies 1959; Roundy 

et al. 1992). In some situations, cultivation 

without surface firming can increase the 

A Seedbed disturbance 

caused by a rangeland drill 

(USFS, 2008) B Seedbed 

disturbance caused by a low-till 

drill (USFS, 2008) C Seedbed disturbance caused by a 

brillion cultipacker (Photo: Alex 

Boehm, 2009) 

A B C 

CHAPTER 4: Assessment of Range Planting as a Conservation Practice 177 
	 	 	 	

	 	

	 	

	 	

	 	 	

	 	 	

	 	 	

	 	 	 	

Specific seedbed 

treatments to 

conserve water 

may not have 

much effect on 

establishment 

success in very 

wet or very dry 

years”

surface area subject to evaporation, reduce 

effective seed–soil contact, reduce seeding 

depth control, decrease hydraulic conductivity 

from deeper soil layers, and stimulate weed 

establishment if seeds are not effectively 

buried (McGinnies 1962; Kyle et al. 2007). 

Subsequent soil firming from press wheels or 

cultipackers improves hydraulic conductivity 

to the seed by reducing soil surface area and 

soil macroporosity (Hyder and Sneva 1956; 

McGinnies 1962). The bulk of range planting 

literature does not separate out treatment 

effects of soil-firming procedures, which are 

usually performed in conjunction with specific 

cultivation and planting procedures (Bement 

et al. 1965; McGinnies 1972; Slayback and 

Renney 1972). Studies that compare multiple 

seed-bed preparation methodologies often 

find differences in relative seeding success 

with different equipment and techniques, 

but specific inferences can only be made at 

the treatment level for a given site and year 

(Hubbard and Smoliak 1953; Hyder et al. 

1955). Few studies of this type have been 

replicated adequately in multiple years or on 

multiple sites (Bement et al. 1965; Eckert and 

Evans 1967; Klomp and Hull 1972; Wood 

et al. 1982; Young et al. 1990; Bakker et al. 

2003). 

Animal trampling, land imprinting, pitting, 

furrowing, and rolling treatments have all 

been used in conjunction with broadcasting 

to capture or preserve moisture, and to press 

surface-applied seed into the soil (Hyder et 

al. 1955; Hyder and Sneva 1956; McGinnies 

1959, 1962; Houston 1965; Haferkamp 

et al. 1987; Roundy et al. 1990; Winkel 

and Roundy 1991; Winkel et al. 1991a; 

Roundy et al. 1992; Ethridge et al. 1997). 

Animal ingestion and subsequent deposition 

of seeds in dung has also been used as a 

mechanism to disperse seeds into favorable 

microsites (Akbar et al. 1995; Andrews 

1995; Ocumpaugh et al. 1996; Auman et al. 

1998; Traba et al. 2003; Gokbulak and Call 

2004). Differential establishment success 

relative to position of soil surface features 

has been reported, and is generally attributed 

to differences in microclimatic conditions 

(Anderson and Swanson 1949; Hyder and 

Sneva 1956; McGinnies 1959; Hull 1970; 

Bragg and Stephens 1979; Hauser 1982; Eckert 

et al. 1986; Roundy et al. 1992). Surface 

modification treatments have also been noted 

to push small seeds too far into the soil or to 

cause surface features to fill with soil from 

wind and water erosion, resulting in seed burial 

beyond establishment depth (Hyder and Sneva 

1956; Kincaid and Williams 1966; McGinnies 

1972; Slayback and Renney 1972; Winkel et 

al. 1991a). Positive effects of these surface 

features may be less relevant in very wet years 

when water is generally available, regardless of 

surface treatment, or in very dry years when 

plantings are unsuccessful regardless of seedbed preparation technique (McGinnies 1968; 

Stuth and Dahl 1974; Wood et al. 1982; Eckert 

et al. 1986; Roundy et al. 1990; Winkel and 

Roundy 1991; Roundy et al. 1992; Romo and 

Grilz 2002). 

mulch Application 

Application of mulch to improve range seeding 

success is frequently advocated as a mechanism 

to reduce water loss and moderate soil surface 

temperatures, although with the caveat that it is 

probably not cost effective for most rangeland 

applications (Lavin et al. 1981; McGinnies 

1987; Ethridge et al. 1997). Relatively 

expensive soil surface amendments such as 

mulch are generally applied only after highimpact disturbance such as mine reclamation, 

or for mitigation of erosion after wildfire on 

topographically complex terrain (Jacoby 1969; 

Meyer et al. 1970; Lavin et al. 1981; Pinchak 

et al. 1985; Schuman et al. 1985; McGinnies 

1987; Schuman et al. 1998; Whisenant 1999; 

Kruse et al. 2004; Groen and Woods 2008). 

An exception may be mulch production as a 

byproduct of mechanical shredding for control 

of juniper and other woody species (Brockway 

et al. 2002). Establishment of a cover crop to 

create standing-stubble mulch is usually limited 

to relatively small areas of major disturbance, 

or higher precipitation zones where grazing 

lands are being reclaimed from cultivation 

(Stroh and Sundberg 1971; Stubbendieck et 

al. 1973; Pinchak et al. 1985; Schuman et al. 

1985; Hart and Dean 1986). Justification for 

mulching practices on rangelands is derived 

from greenhouse, laboratory, and modeling 

studies, all of which confirm general benefits 

of water conservation and mitigation of high 

temperature near the soil surface as a function 

of relative coverage (Hopkins 1954; Bond 

and Willis 1970; Chung and Horton 1987; 

Bristow and Abrecht 1989; Jalota 1993; Brar 

178 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	 	

	 	

S. P. Hardegree, T. A. Jones, B. A. Roundy, N. L. Shaw, and T. A. Monaco 

and Unger 1994; Bussiere and Cellier 1994; 

Gill and Jalota 1996; Novak et al. 2000a, 

2000b, 2000c; Giminez and Govers 2008), 

and field studies, most of which have been 

conducted after tillage or on severely disturbed, 

or otherwise extreme, sites (Dudeck et al. 

1970; Meyer et al. 1970; Stubbendieck et al. 

1973; Schuman et al. 1985; Hart and Dean 

1986; Ethridge et al. 1997; Ji and Unger 2001; 

Dahiya et al. 2007; Groen and Woods 2008). 

Water conservation associated with mulch 

application on range seeding success may not 

be ecologically significant in very high or very 

low precipitation years or on some extreme 

rangeland sites (Gates 1962; Ludwig and 

McGinnies 1978; Lavin et al. 1981; Berg and 

Sims 1984; McGinnies 1987; Bristow 1988; 

Cione et al. 2002; Fulbright et al. 2006). 

For the 21 studies surveyed for this review 

that specifically evaluated mulch treatments, 

62% concluded that mulch application 

improved establishment success. Regardless 

of the variable effects of mulch on seeding 

success, application of mulch for effective 

erosion control and soil stabilization is well 

documented (Meyer et al. 1970; Bautista et al. 

1996; Fulbright et al. 2006; Groen and Woods 

2008). 

seeding depth 

Successful germination and establishment 

is dependent upon placement of seeds in 

favorable soil microsites (Hyder et al. 1955; 

Harper et al. 1965; Young et al. 1990; Call 

and Roundy 1991; Winkel and Roundy 

1991, Winkel et al. 1991b; Roundy et al. 

1992; Chambers and MacMahon 1994; 

Sheley et al. 1996; Ott et al. 2003). A 

major assumption of many site-preparation 

treatments is that they increase the number 

of potential safe sites for germination and 

establishment either by covering the seed, 

by reducing soil water loss from around the 

seed, or by redistributing and concentrating 

resources (Anderson and Swanson 1949; 

Hubbard and Smoliak 1953). 

Mechanical disturbance is generally necessary 

to incorporate seeds into the soil, thus reducing 

the risk of either desiccation or adverse 

thermal effects near the surface. Seeding 

depth recommendations from commonly 

cited seeding guides and technical references 

are relatively specific, but are based on rules of 

thumb regarding seeding depth as a function of 

seed size (Hull and Holmgren 1964; Plummer 

et al. 1968; Jordan 1981; Roundy and Call 

1988; Jensen et al. 2001; Monsen and Stevens 

2004; Lambert 2005; Ogle et al. 2008a, 

2008b). The physical rationale for depth 

recommendations usually assumes a tradeoff between increased water availability and 

increased energy requirements for emergence 

as a function of depth (Roundy and Call 1988; 

Call and Roundy 1991). In some cases, light 

or diurnal temperature fluctuation may regulate 

dormancy to ensure that the seeds germinate at 

an appropriate depth for a given species (Call 

and Roundy 1991; Ghersa et al. 1992; Traba 

et al. 2004). Seed predation has also been 

documented as a potential problem for surfacesown seeds (Nelson et al. 1970). 

Evidence for depth effects is generally 

limited to studies conducted in a controlled 

environment, or over very small spatial scales 

in the field (Kinsinger 1962; Vogel 1963; 

Hull 1964). A major exception is for studies 

comparing the relative establishment of 

broadcast versus planted seeds. Of the 23 field 

studies surveyed for this review that specifically 

compared broadcast versus drill seeding, 73% 

concluded that drill seeding outperformed 

broadcast seeding. These studies, however, 

did not generally include quantification of 

Brillion wheel for pressing 

broadcast seeds into the soil 

(Photo: USFS, 2006) 

CHAPTER 4: Assessment of Range Planting as a Conservation Practice 179 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 		 	

	 	 	 	

	 	 	 	 		

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	

Standard 1X seeding rate of 

247 seeds/m2 or 23 seeds/ft2 

in a 0.25-m2 frame (left). 

Middle frame shows a 2X seed 

rate and right frame shows 

a 5X seed rate (Photo: Alex 

Boehm, 2011) 

the specific depth distribution after planting 

(Stewart 1950; Hyder et al. 1955; Douglas et 

al. 1960; Gomm 1964; Bement et al. 1965; 

Statler 1967; Shown et al. 1969; Nelson et al. 

1970; McGinnies 1972; Drawe et al. 1975; 

Wood et al. 1982; Haferkamp et al. 1987; 

Ott et al. 2003). Relative seeding depth in 

field studies is often reported in the context 

of depth band settings on mechanical seeding 

equipment, but there are very few studies in 

which actual seeding depth has been quantified 

postplanting (Winkel and Roundy 1991; 

Winkel et al. 1991a, 1991b). Laboratory, 

greenhouse, and field comparisons of surfacesown versus planted seeds generally confirm 

that very small seeds establish more frequently 

from near-surface seed placement, larger seeds 

require soil cover for maximal performance, 

and seed performance drops dramatically below 

some threshold depth (Hull 1948; Stewart 

1950; Douglas et al. 1960). Indian ricegrass 

(Achnatherum hymenoides [Roem. & Schulte.] 

Barkworth) has been extensively documented 

for its ability to germinate and emerge from 

relatively deep sowing depths, especially in 

sandy soils (Kinsinger 1962; Jones 1990; Young 

et al. 1994). 

Broadcast and planting recommendations are 

generally not discretionary, as topographic 

complexity and economic considerations 

may preclude the use of planting equipment. 

Broadcast seeding rates are generally 

recommended at two to three times the rates 

for seed that can be incorporated into the soil 

(Stewart 1950; Hyder et al. 1955; Douglas et 

al. 1960; Gomm 1964; Bement et al. 1965; 

Statler 1967; Shown et al. 1969; Nelson et al. 

1970; McGinnies 1972; Drawe et al. 1975; 

Wood et al. 1982; Haferkamp et al. 1987; Ott 

et al. 2003). 

seeding Rate 

General seeding-rate recommendations 

from many technical sources appear to be 

based on a general standard for what could 

be considered a hypothetical dominant 

bunchgrass, planted at optimal depth, in a 

uniform, well-prepared, weed-free seed bed, in 

a favorable establishment year. The standard 

seeding rate for this hypothetical scenario 

seems to be roughly equal to a seed density 

of 1 million seeds per acre or approximately 

23 seeds/ft2 under historical, non-SI units 

of measure (Jordan 1981; Jensen et al. 

2001; Monsen and Stevens 2004; Lambert 

2005; Ogle et al. 2008a, 2008b). The most 

commonly recommended deviation from this 

hypothetical standard is to increase seeding 

rate by a factor of two–five for small seeds 

or for potential location-specific problems 

such as inadequate weed control, lack of site 

preparation, surface application of seeds, 

probability of drought, nonoptimal seeding 

season, or high levels of seed dormancy 

(Jordan 1981; Monsen and Stevens 2004; 

Thompson et al. 2006). Seeding-rate 

recommendations are also generally adjusted 

to reflect the total seed-mix ratio, and ideal 

expectations for composition of the desired 

mature plant community (Pyke and Archer 

1991; Ogle et al. 2008a, 2008b). It is often 

difficult to assess numerical seeding rates, 

as the bulk of the literature reports rate in 

terms of weight of seed planted per unit 

land area. Weight-based recommendations 

in the technical literature, however, are 

generally supplemented by bulk seed density 

180 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

	 	 	

	 	 	

	 	 	

	 	

	 	 	

	 	

	 	 	

	 	

	 	

S. P. Hardegree, T. A. Jones, B. A. Roundy, N. L. Shaw, and T. A. Monaco 

information (Plummer et al. 1968; Jensen et 

al. 2001; Monsen and Stevens 2004; Lambert 

2005; Ogle et al. 2008a, 2008b). 

Seeding-rate recommendations are linked to 

microclimatic considerations, as increased 

seed numbers increase the probability of 

seeds reaching safe microsites, irrespective 

of active depth management (Harper et al. 

1965; Call and Roundy 1991; Roundy et 

al. 1992; Chambers 1995). Relatively few 

studies reporting effects of seeding rate on 

establishment success are replicated in such a 

way to survey annual and seasonal variability 

in seed-bed microclimate (Schultz and Biswell 

1952; Mueggler and Blaisdell 1955; Hull and 

Holmgren 1964; Launchbaugh and Owensby 

1970; Hull 1972a, 1974b; Papanastasis and 

Biswell 1975; Vogel 1987; Masters 1997; 

McMurray et al. 1997; Williams et al. 2002). 

Some studies that include variable seeding 

rates were primarily designed to evaluate 

competition relative to weed-seed numbers, 

but in general, the literature supports the 

concept that higher seeding rates may 

enhance the likelihood of successful initial 

establishment (Vogel 1987; Sheley et al. 1999; 

Wiedemann and Cross 2000; Williams et al. 

2002). Seeding-rate impacts remain highly 

dependent upon threshold requirements 

for water availability in the early stages of 

establishment, and individual seedling growth 

can be negatively impacted by both inter- and 

intraspecific competition later in development. 

The majority of the literature pertaining to 

seeding-rate effects is derived either from 

controlled environment and greenhouse 

studies, or field studies conducted in years 

where reported rainfall conditions were either 

average or above average (Francis and Pyke 

1996; Sheley and Half 2006). Eiswerth and 

Shonkwiler (2006) evaluated a large number 

of range seeding sites and years in Nevada and 

determined that increased seeding rates led to 

higher seedling densities for nonnative grasses 

up to some maximum seeding rate. This study, 

however, did not analyze or report negative 

seeding results, and did not consider weather 

and climate conditions during the years that 

seeding occurred. 

Planting season 

Most studies of planting-season effects on 

establishment success can be linked to climatic 

variability, and often to specific germination 

and dormancy syndromes of various seeded, 

nonseeded, and weedy species (Angevine 

and Chabot 1979). General planting-season 

recommendations require getting the seed 

planted in time to take advantage of the most 

favorable season for plant establishment (Hull 

1948; Stoddart and Smith 1955; Plummer et 

al. 1968; McGinnies 1972; Vallentine 1979; 

Jordan, 1981; Roundy and Call 1988; Ries and 

Hofmann 1996; Monsen and Stevens 2004; 

Stevens 2004). In some cases, dormant-fall 

seeding is recommended well in advance of the 

optimal growing season to take advantage of 

all opportunities for potential establishment 

in a highly variable, and often arid or semiarid 

environment (Hull 1948; Stewart 1950; 

Douglas et al. 1960; Plummer et al. 1968; 

Young et al. 1969b; Nelson et al. 1970; Klomp 

and Hull 1972; Hart and Dean 1986; Young 

et al. 1994; Monsen and Stevens 2004). 

Dormant-fall seeding is also recommended 

when there are logistical concerns for use of 

mechanical equipment during wet-spring 

planting conditions, or to mitigate effects of 

unpredictable spring weather (Stewart 1950; 

Douglas et al. 1960; McGinnies 1973; Hart 

and Dean 1986). Seasonal timing of seeding 

may also be dependent on seasonality of 

weed competition and/or optimal timing of 

weed control measures (Bement et al. 1965; 

Robocker et al. 1965; Hull 1972a; Klomp 

and Hull 1972). The most favorable season 

for establishment varies regionally (Hatfield 

1990): spring in Mediterranean-coastal and 

Intermountain West locations (Douglas et al. 

1960; Nord et al. 1971; Hull 1972a; Harris 

and Dobrowolski 1986), summer monsoon in 

the southwestern desert (Jordan 1981; Abbott 

and Roundy 2003; Hereford et al. 2006), late 

spring through early summer in the Great 

Plains (Robertson and Box 1969; Hyder et 

al. 1971; McGinnies 1973; Hart and Dean 

1986; Ries and Hofmann 1996; Frank et al. 

1998; Romo and Grilz 2002), and late spring 

through early fall in some higher-elevation 

mountain sites (Hull 1966; Currie 1967; Lavin 

et al. 1973; Hull 1974a, 1974b). Postplanting 

microclimate must be favorable for growth, 

but also needs to remain favorable during the 

vulnerable period of seedling establishment 

(Hyder et al. 1971; McGinnies 1973; Frasier et 

al. 1987; Abbott and Roundy 2003). Eiswerth 

and Shonkwiler (2006) confirmed the relative 

Of 52 studies 

surveyed for this 

review, all but 

two concluded 

that weed control 

was either 

necessary, or at 

least beneficial 

to successful 

establishment.”

CHAPTER 4: Assessment of Range Planting as a Conservation Practice 181 
	 	 	 	

 

	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	

	 	 	 	

	 	 	 	 	

	 	 	

	 	 	

A High erosion potential 

for soil surface with low plant 

cover caused by competition 

with western juniper. B Low 

erosion potential for soil 

surface with high plant cover 

after removal of woody-plant 

competition (Photo: Fred 

Pierson, 2003) 

benefits of fall/winter-dormant seeding on 

intermountain rangelands in Nevada with 

the use of meta-analysis of long-term Bureau 

of Land Management fire-rehabilitation 

monitoring data. Very few experimental 

studies of seeding-season effects are replicated 

in more than 1 or 2 yr (Hull 1948; Douglas et 

al. 1960; Robocker et al. 1965; Hull 1974b; 

Ries and Hofmann 1996). Fall-dormant 

planting, however, was found to be superior to 

spring planting in 73% of Great Basin studies 

where planting season was evaluated. 

Weed Control 

Seed-bed preparation and planting method 

recommendations are designed to improve 

microclimatic conditions for desirable 

species, but also to reduce competition 

from undesirable plants (Lavin et al. 1973; 

Gonzalez and Dodd 1979; Ott et al. 2003; 

Mangold et al. 2007). Chemical or mechanical 

weed control, prior to the early stages of 

establishment, are generally required for 

establishment success of both native and 

nonnative plant species (Evans et al. 1970; 

Nelson et al. 1970; Klomp and Hull 1972; 

Stuth and Dahl 1974; Evans and Young 1978; 

Humphrey and Schupp 2002; Mangold et 

al. 2007). Of 52 studies surveyed for this 

review that included mechanical or chemical 

weed control, all but two concluded that 

weed control was either necessary, or at least 

beneficial to successful establishment. Efficacy 

of alternative mechanical and chemical weed 

control treatments is more extensively discussed 

elsewhere in this volume (Sheley et al., this 

volume). 

nRCs teChnICAl ReCommendAtIons 

And the RAnge PlAntIng 

lIteRAtuRe 

General management recommendations, and 

associated NRCS technical references (e.g., 

Ogle et al. 2008a, 2008b), are consistent with 

current rangeland planting technical guidance 

and authorities (Vallentine 1979; Jordan 

1981; Sours 1983; Redente and DePuit 1988; 

Roundy and Call 1988; Roundy 1996; Sheley 

et al. 1996; Whisenant 1999; Jensen et al. 

2001; Monsen and Stevens 2004; Stevens 2004; 

Sheley et al. 2006). These recommendations 

and guidelines do not fundamentally differ 

from earlier-cited works that predate current 

standards for hypothesis testing, statistical 

inference, and experimental design norms 

(Stoddart and Smith 1943; Stewart 1950; 

Stoddart and Smith 1955; Anderson et al. 

1957; Plummer et al. 1968; Vallentine 1979). 

Many of the historical references used to 

justify range planting practices, however, come 

from Agricultural Experiment Station reports, 

internal agency documents, and syntheses 

of unpublished field trials, as opposed to the 

refereed literature (McGinnies et al. 1963; 

Great Plains Council 1966; Plummer et al. 

1968; Gomm, 1974; Cox et al. 1984; Call 

and Roundy 1991). With the exception 

of some specific plant-material selection 

and development programs, the underlying 

principles of these earlier recommendations 

were primarily based on previously established 

agricultural concepts, and a probabilistic 

assessment of best management practices 

derived from the practical experience and 

A B 

182 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

	 	 	

	 	

	 	

	 	

S. P. Hardegree, T. A. Jones, B. A. Roundy, N. L. Shaw, and T. A. Monaco 

personal observations of land management 

professionals. The scientific literature from 

the more recent 40–50 yr has attempted 

to refine and to validate these commonly 

recommended practices experimentally. The 

more recent literature, however, is dominated 

by empirical studies that provide examples of 

field success for specific planting techniques, 

but, individually, are insufficiently replicated 

for general inferences (Call and Roundy 1991). 

We therefore surveyed 189 range planting field 

studies to draw the following general inferences 

regarding assumptions inherent to current 

seeding technical references, and range planting 

conservation practice documents: 

•	� General recommendations supported by 

the aggregate literature must be prefaced 

by an acknowledgement that climatic 

conditions during the establishment year 

must be favorable. Ninety percent of the 

range planting papers surveyed report at 

least one successful treatment. Of the 

57% that reported climatic conditions 

during the study, however, 89% claimed 

average or above-average precipitation in 

the year of establishment for the successful 

treatments. Over half of the studies 

that report successful establishment in a 

below-average precipitation year note that 

the seasonal distribution of precipitation 

was favorable during early seedling 

development. 

•	� Few inferences can be made from the 

range planting literature about the relative 

establishment characteristics of alternative 

plant materials. Very few studies are 

designed and replicated in such a way 

that within- or between-species variability 

can be assessed. Over 35% of studies 

evaluated used only one seed lot or a 

unique seed mix, and 24% compared 

relative establishment among unique seed 

mixes. Of the 86 studies in which more 

than one seed lot of the same species were 

evaluated, only 6% were fully replicated, 

and 19% were partially replicated at the 

species level. Almost half of the studies 

used at least some named varieties, but 

only four specifically evaluated withinspecies variability. The strongest evidence 

for plant materials suitability is derived 

from observation of historical relationships 

between species and climate, experiencebased observation, long-term assessment 

of persistence of planted species, and field 

trials conducted during the process of 

plant-materials selection and development. 

•	� General conservation practice 

recommendations regarding site 

preparation and seeding methodology are 

generally supported from the aggregate 

literature. Drill-seeding treatments 

outperformed broadcast-seeding 

treatments in 73% of the studies that 

included a direct comparison. Application 

of mulch improved establishment success 

in 62% of the studies where there was a 

direct comparison. Increasing seeding 

rate was found to improve establishment 

success in 79% of the 24 studies where 

this was directly tested. Of the 52 studies 

that included mechanical or chemical weed 

control treatments, all but 2 concluded 

that weed control was either necessary or at 

least beneficial to successful establishment. 

Fall-dormant planting was determined 

to be superior to spring planting in 

73% of the Great Basin studies where 

planting season was evaluated. Seedbed preparation, seeding depth, planting 

season, and seeding-rate recommendations 

may be irrelevant in very dry and perhaps 

very wet years. 

•	� The majority of range planting field 

studies are unreplicated in either space 

or time. Only 47% replicate planting 

years and 41% replicate site locations. 

The predominant form of treatment 

replication was within site, with 69% 

of studies having at least two, and 61% 

having at least three, within-site treatment 

replicates. Meta-analysis of studies that 

are individually underreplicated for general 

inferences is hampered by the fact that 

negative results are usually not published, 

and plant-materials selection is often 

based on a priori assumptions about their 

suitability for a given location. 

Assessment of sPeCIfIC 

ConseRvAtIon effeCts 

Evidence supporting positive conservation 

effects of alternative established plant 

communities is generally found in a separate 

body of literature than that examined in 

the Conservation Practices section of this 

There is virtually 

no experimental 

evidence directly 

linking rangeland 

seeding to 

conservation 

effects.”

CHAPTER 4: Assessment of Range Planting as a Conservation Practice 183 
	 	 	 	

 

	 	 	

	 	

	 	

	 	 	

	 	 	

	 	

Application 

of mulch to 

improve range 

planting success 

is probably not 

cost effective for 

most rangeland 

applications ”

chapter. The literature supports the concept 

that seeding, if successful, results in positive 

conservation effects. There is virtually no 

literature, however, directly linking rangeland 

seeding to conservation effects. We have 

limited our review to conservation effects 

related to water quality and erosion, water 

quantity, and soil carbon sequestration, because 

these conservation effects are not specifically 

addressed in the other chapters in this volume. 

Conservation effects related to livestock and 

wildlife needs, weed proliferation, and fire are 

left to the chapters in this volume concerning 

prescribed grazing, upland wildlife habitat 

management, herbaceous weed control, and 

prescribed burning. 

Water Quality and erosion 

Very few studies directly link rangeland seeding 

to conservation benefits from improved 

water quality and reduced erosion (Wright et 

al. 1982; Brown et al. 1985; Beyers 2004). 

There is an extensive literature, however, 

documenting the relationship between 

rangeland soil cover and soil stability (Nearing 

et al. 2005; Bartley et al. 2006; Gimeno-Garcia 

et al. 2007). Removal of plant canopy cover 

by clipping may be insufficient to increase 

sediment loss in the short term when soil is still 

protected by basal vegetation cover and surface 

residues (Giordanengo et al. 2003; Gyssels 

et al. 2005; Nearing et al. 2005; De Baets et 

al. 2006). Range planting, per se, will not 

have a significant effect on soil stability unless 

sufficiently successful to provide adequate 

soil cover (Meeuwig 1965; Gifford 1970, 

1972; Wright et al. 1982; Brown et al. 1985; 

Ziegler and Giambelluca 1998; Aguilera et al. 

2003; Beyers 2004; Groen and Woods 2008). 

Short-term effects of site preparation, fire, or 

other treatments that precede range seeding 

or natural recovery, however, can significantly 

increase potential erosion in the near term 

(Williams et al. 1969; Gifford 1972, 1973; 

Tromble 1976; Roundy et al. 1978; Tromble 

1980; Brown et al. 1985; Benavides-Solorio 

and Macdonald 2001; Gimeno-Garcia et al. 

2007; Grismer 2007; Pierson et al. 2007). 

In most cases, the nature of this cover is less 

relevant than the issue of soil surface protection 

above some threshold level (Mergen et al. 

2001; Aguilera et al. 2003; Descheemaeker et 

al. 2006). Some studies, however, have shown 

differential hydrologic effects under adjacent 

plant communities due to differences in growth 

and litter production, interception, wateruse efficiency, or rooting depth and spread 

(Dunkerley 2002; Bhark and Small 2003; 

Kulmatiski et al. 2006). The relative impact of 

vegetation cover on erosion and runoff is also 

highly dependent on weather, slope and soil 

type (Aguilera et al. 2003; Bartley et al. 2006; 

Nichols 2006). Major soil loss after vegetation 

removal can be exacerbated by intense rainfall 

events (Gifford 1973; 1975; Garza and 

Blackburn 1985; Takar et al. 1990). 

Vegetation affects soil stability and runoff water 

quality by protecting the soil from rainfall 

impact, increasing soil infiltration capacity, 

anchoring the soil mass, and preventing 

the development of rill erosion by slowing 

overland flow rates and increasing surface 

water flow paths (Tromble et al. 1974; Thurow 

et al. 1987; Pimentel and Kounang 1998; 

Aguilera et al. 2003; Imeson and Prinsen 

2004; Puigdefabregas 2005). Invasive woody 

plants have been shown to suppress understory 

species to the point where insufficient soil 

surface cover can result in significant erosion, 

even under relatively low-intensity storm 

events (Davenport et al. 1998; Pierson et 

al. 2007; Petersen and Stringham 2008). 

Annual weed cover can affect seasonal patterns 

of evapotranspiration and soil water use 

(Kulmatiski et al. 2006; Prater and DeLucia 

2006), but there is little evidence that they 

increase site runoff or erodibility when they 

are providing adequate soil cover (Singh 1969; 

Pierson et al. 2002; Wilcox and Thurow 2006; 

Pierson et al. 2007). In the Intermountain 

West, however, invasive annual weeds can 

increase the frequency of periods where 

vegetation cover can be reduced by wildfire 

(Brandt and Rickard 1994; Knapp 1996; 

Young and Longland 1996; Whisenant 1999). 

Indeed, the primary objective of most firerehabilitation seeding practices is to improve 

soil stability and reduce erosion (Richards et 

al. 1998; Bureau of Land Management 1999; 

Beyers 2004; Grismer 2007). 

Application of mulch to improve range 

planting success is probably not cost effective 

for most rangeland applications (Lavin et al. 

1981; McGinnies 1987; Ethridge et al. 1997). 

Mulch application, litter retention after site 

preparation, and mulch production as a result 

184 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	

	 	

	 	

S. P. Hardegree, T. A. Jones, B. A. Roundy, N. L. Shaw, and T. A. Monaco 

of mechanical treatments for woody plant 

control, however, can have a direct conservation 

effect of reducing erosion and runoff on 

severely disturbed or highly erodible range 

sites, or on steep slopes (Meyer et al. 1970; 

Gifford 1975; Knight et al. 1983; Bautista et al. 

1996; Brockway et al. 2002; Benik et al. 2003; 

Grismer and Hogan 2004, 2005; Fulbright et 

al. 2006; Groen and Woods 2008). 

Water Quantity 

Different plant materials and species may have 

different degrees of water-use efficiency and 

biomass production, but in arid rangeland 

systems, plants tend to use all available water 

in the soil profile in most years (Wright 

and Dobrenz 1973; Cable 1980; Trlica and 

Biondini 1990; Dugas and Mayeux 1991; 

Weltz and Blackburn 1995; Hester et al. 1997; 

Wilcox 2002; Huxman et al. 2005; Wilcox and 

Thurow 2006). Seeding grasses after shrub 

removal can result in increased stream flow in 

some Mediterranean-type climates, where the 

principal precipitation season is out of phase 

with the seasonal peak of evapotranspiration, 

or in systems where the woody plant material 

has access to groundwater (Hill and Rice 1963; 

Hibbert 1983; Wilcox 2002; Huxman et al. 

2005; Huang et al. 2006; Wilcox and Thurow 

2006). Shrubland conversion to grassland is 

not generally expected to result in an increase 

in water quantity in arid and semiarid upland 

rangeland systems, except where vegetation 

removal may increase overland flow directly 

to a stream channel (Gifford 1970; Wright 

et al. 1982; Bergkamp 1998; Wilcox 2002; 

Wilcox et al. 2003; Wilcox and Thurow 2006) 

or during extreme rainfall events where runoff 

may be affected by total plant cover (Weltz and 

Blackburn 1995; Quinton et al. 1997). The 

major exception to this would be the relatively 

large potential increase in overland flow after 

major vegetation disturbance, which may also 

generate unacceptable levels of soil erosion 

(Gifford 1973; Osborn and Simanton 1990; 

Takar et al. 1990; Johansen et al. 2001; O’dea 

and Guertin 2003; Pierson et al. 2007). 

Carbon sequestration 

Svejcar et al. (2008) summarized the results 

of a 6-yr regional experiment that monitored 

seasonal carbon flux on western US rangelands 

and found that relatively good condition 

rangeland generally serves as a carbon 

sink, except in the driest areas of the desert 

southwest. The degree of carbon sequestration 

or loss varies primarily in response to seasonal 

and annual weather patterns (Conant et 

al. 2001; Flanagan et al. 2002; Jones and 

Donnelly 2004; Xu and Baldocchi 2004; 

Follet and Schuman 2005; Hastings et al. 

2005; Derner and Schuman 2007; Svejcar et 

al. 2008). A change from tillage and annual 

cropping to perennial grass cover can greatly 

increase soil carbon-sequestration rates, but the 

effect is less on western rangeland soils than 

in more mesic areas (Conant et al. 2001; Guo 

and Gifford 2002; Sperow et al. 2003; Jones 

and Donnelly 2004; Derner and Schuman 

2007). Relatively low sequestration rates, 

however, are offset by the relatively large land 

area occupied by rangelands (Scurlock and 

Hall 1998; Derner and Schuman 2007). Type 

conversion from woody plants to grasses can 

lower carbon-sequestration rates if the initial 

plant community has a higher net ecosystem 

production, but this would probably not be 

the case in most upland arid and semiarid 

rangeland systems (Huxman et al. 2003). 

Restoration of severely disturbed rangeland, 

and activities such as mine reclamation, can 

significantly improve carbon-sequestration 

rates (Follet and Schuman 2005; Derner and 

Schuman 2007). 

Rooting depth, and effective water utilization 

for biomass production, can vary considerably 

among alternative vegetation types (Cline et al. 

1977; Cable 1980; Yoder et al. 1998; Huxman 

et al. 2005; Seyfried and Wilcox 2006). As a 

significant portion of sequestered carbon can 

be deposited below ground by root growth 

(Scurlock and Hall 1998; Jones and Donnelly 

2004; Rees et al. 2005), depth of rooting may 

be a consideration in selection of plant materials 

for rangeland seeding operations. Millions of 

acres of sagebrush–bunchgrass rangeland in 

the Intermountain West have been invaded by 

introduced annual weeds such as cheatgrass 

(Bromus tectorum L.). Cheatgrass-dominated 

systems are characterized by frequent recurrence 

of wildfire and are resistant to management 

actions that are designed to return them to 

a more desirable ecological state (Brandt 

and Rickard 1994; Knapp 1996; Young and 

Longland 1996). Carbon-sequestration rates 

can be nullified when vegetation is periodically 

removed by prescribed fire or wildfire (Suyker 

Current 

conservation 

practice 

standards 

should include 

references to 

more ecologically 

based technical 

literature 

and specific 

guidance for 

both monitoring 

and adaptive 

management.”

CHAPTER 4: Assessment of Range Planting as a Conservation Practice 185 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	 	

	 	

	 	

The most 

useful potential 

technology 

for enhancing 

establishment 

success lies in 

development 

and utilization 

of relatively 

long-range 

weather-forecast 

technology”

and Verma 2001). Management activities, such 

as rangeland planting, that result in an increase 

in forage production can be expected to increase 

carbon-sequestration rates (Ma et al. 2000; 

Conant et al. 2001). 

ReCommendAtIons 

The aggregate literature generally supports 

both the existing conservation practice 

recommendations for rangeland seeding, 

and the inherent assumption that if these 

practices are successful, they will result in 

beneficial conservation effects. Current 

conservation practice recommendations, 

however, are relatively prescriptive in that 

they do not effectively address site- and yearspecific variability, provide no mechanism 

for evaluating or adapting to unsuccessful or 

partially successful treatments, are oriented 

toward short-term management, and are not 

fully integrated with current, ecologically 

based models for management of alternative 

vegetation states. Current conservation 

practice standards should include references 

to more ecologically based technical literature 

and specific guidance for both monitoring 

and adaptive management. Additional 

research needs to be conducted to test the 

ecological underpinnings of existing and 

new ecological models for plant community 

dynamics explicitly, and to develop new 

tools to take advantage of existing and 

emerging knowledge of weather variability 

and anticipated shifts in regional climate. 

Individual seeding studies are seldom 

replicated sufficiently to make valid inferences 

in such a variable field environment. 

Therefore, monitoring protocols that can 

facilitate meta-analysis in support of more 

general inferences need to be developed. 

KnoWledge gAPs 

explicit testing of new Conceptual 

models for dynamic Rangeland 

systems 

Call and Roundy (1991) recommended 

changes to the prevailing research approach to 

address problems inherent in highly variable 

rangeland systems more directly. A more 

general scientific understanding of vegetation 

change may now be achievable with the use of 

more recently developed conceptual models 

for understanding dynamic rangeland systems 

(Westoby et al. 1989; Bestelmeyer et al. 

2003; Sheley et al. 2006). NRCS has already 

adopted some of these paradigms by utilizing 

state-and-transition-model concepts in the 

development of Ecological Site Descriptions, 

but these models do not currently form the 

basis for range planting conservation practice 

recommendations. These models integrate 

multiple processes and acknowledge multiple 

potential trajectories for plant community 

change and will require new and innovative 

approaches for validation and testing in the field. 

development and utilization of 

Weather and forecasting tools 

The stochastic nature of weather variability 

will require adoption of new concepts for 

evaluating revegetation and restoration 

success. Expectations for success need to be 

explicitly linked to the probability of favorable 

conditions for seed germination, emergence, 

and establishment (Krzysztofowicz 2001; 

Bakker et al. 2003). New technologies will 

need to be developed and utilized in order to 

use weather information to inform rangeland 

planting management decisions (Workman and 

Tanaka 1991; Peters 2000; Rayner et al. 2005; 

Andales et al. 2006). 

The most useful potential technology for 

enhancing establishment success lies in 

development and utilization of relatively longrange weather-forecast technology specific to 

rangeland planting applications (Barnston 

et al. 1994, 2005; Garbrecht and Schneider 

2007). Long-term weather forecasts in large 

portions of the Intermountain West are often 

merely synoptic descriptions of historical 

weather patterns and are not based on physical 

or empirical prediction of future weather 

conditions. It may be possible, however, to 

utilize historical weather and seeding data to 

construct models to assess the potential longterm benefits of adopting forecast/modeling 

technology in rangeland restoration planning 

(Batabyal and Godfrey 2002; Schneider and 

Garbrecht 2003, 2006; Bashari et al. 2008). 

Similar technology is in relatively common use 

for more traditional agricultural applications 

and for some rangeland applications (Schneider 

and Garbrecht 2003; Doblas-Reyes et al. 2006; 

Schneider and Garbrecht 2006; Baigorria et al. 

2008; O’Lenic et al. 2008). 

186 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	 	 	 	

	 	 	 	

S. P. Hardegree, T. A. Jones, B. A. Roundy, N. L. Shaw, and T. A. Monaco 

Monitoring native planting at 

the site of the Scooby fire in 

northern Utah (Photo: USFS, 

2010) 

Even low-resolution weather forecasts would 

increase the probability of successful native 

plant establishment if seeding decisions in 

the fall could be based on the anticipation of 

favorable conditions of seedbed microclimate 

in the subsequent winter and spring (Hardegree 

et al. 2003; Hardegree and Van Vactor 2004). 

Weather forecasts could be used to initiate 

contingency plans in areas that have been 

previously identified for restoration, and for 

which premanagement logistics of equipment, 

personnel, and plant materials are in place 

(Bakker et al. 2003; Westoby et al. 1989). 

Separation of restoration planning objectives 

from the wildfire cycle would also simplify 

the problem of predicting management needs 

for native germplasm (Richards et al. 1998). 

Historical climate records could provide a 

relatively stable estimate of the probability 

of favorable establishment years that could 

be used to predict acquisition and storage 

requirements for native seed over the long term. 

Biodiversity and restoration planning objectives 

may require multiple-year strategies for 

replacement of nonnative species only after 

initial site stabilization and suppression of 

annual weed competition (Bakker et al. 2003; 

Cox and Anderson 2004). Weather and climatic 

limitations require definition of realistic goals 

when establishing rehabilitation and restoration 

planning objectives (Call and Roundy 1991; 

Hobbs and Norton 1996; Ehrenfeld 2000; 

Jones 2003). Asay et al. (2001) argue that the 

relatively harsh climatic conditions on many 

rangelands may preclude the realistic use of 

many native plant materials in favor of adapted 

nonnative species. In some years, and on some 

sites, it may be prudent to plant more easily 

established nonnative species, particularly 

after wildfire or other disturbance, when the 

principal objective of rangeland planting may be 

soil stabilization. Biodiversity and restoration 

objectives could then be addressed in years when 

climatic conditions are amenable (Holmgren 

CHAPTER 4: Assessment of Range Planting as a Conservation Practice 187 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	

Adoption 

of minimum 

experimental 

design standards 

would facilitate 

meta-analysis 

of future range 

planting studies.”

and Scheffer 2001; Bakker et al. 2003; 

Hardegree et al. 2003; Cox and Anderson 2004; 

Hardegree and Van Vactor 2004). 

Plant-materials Program development 

and testing 

Previous plant-materials development has 

focused on productivity, vigor, establishment, 

disease resistance, seed production, and specific 

ecological and physiological traits deemed to 

confer superior performance or adaptation. 

Establishment, persistence, and invasion 

resistance of seeded plant communities may 

be enhanced by identification and selection of 

plant materials with functional traits similar to 

the various highly competitive invasive species 

(Arredondo et al. 1998; Pokorny et al. 2005; 

Funk et al. 2008). Functional traits common 

to many weedy invaders include high relative 

growth rate, specific leaf area, leaf nitrogen 

content, and resource-use efficiency (Aguirre 

and Johnson 1991a, 1991b; Grotkopp et al. 

2002; Pokorny et al. 2005; Grotkopp and 

Rejmanek 2007; James and Drenovsky 2007; 

Funk et al. 2008). 

Development of herbicide-resistant native grass 

plant materials may be a useful area of future 

research. In recent years, interest has increased 

in using acetolactate synthase (ALS) inhibitors 

such as imazapic to reduce annual grass 

competition with desirable perennials. Imazapic 

damages fast-growing tissues, especially 

meristems, so weedy annual grasses and crucifers 

are controlled with less damage to perennials 

(Shaner and O’Conner 1991). Some residual 

activity is present, so annual grass control 

is still achieved a year following application 

(Davison and Smith 2007). An advantage of 

this herbicide is that many desirable nongrass 

species, particularly legumes and composites, 

are relatively resistant. Development of ALSinhibitor resistance has been quite successful in 

several crop species (Tranel and Wright 2002); 

thus development of native plant materials with 

such resistance is likely to be successful. When 

such materials are developed by traditional 

plant-breeding methodologies, no special 

Environmental Protection Agency clearance is 

required prior to release. 

For many years, the NRCS and Agricultural 

Research Service have routinely evaluated 

released and experimental materials as part 

of their ongoing plant-material research and 

development programs. For the rangelands 

Wildfire disturbance at the site 

of the Crowbar fire in southwestern Idaho (Photo: USFS, 

2010) 

188 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

S. P. Hardegree, T. A. Jones, B. A. Roundy, N. L. Shaw, and T. A. Monaco 

of the semiarid west, these trials are typically 

dormant planted in late fall for spring 

emergence. Comparisons between plant 

species and among plant materials within 

species are made in order to characterize 

released plant materials and to justify the 

release of new plant materials. The single 

most important trait of these trials is seedling 

establishment. These trials are typically 

replicated complete block design experiments 

subject to statistical analysis, but are usually 

analyzed as individual experiments rather 

than as a collective whole. Because these 

trials are conducted every year, a large volume 

of data have been collected. More robust 

comparisons of species and plant materials 

could be made if these data were compiled 

and subjected to meta-analysis. A Webaccessible program to update and combine 

data sets for analysis would greatly increase the 

statistical power of plant-material evaluations 

and improve decision making. With the use Vargas et al. 2001). Range planting studies 

of additive main and multiplicative interaction also tend to extrapolate results obtained from 

(AMMI) statistical models, NRCS PRISM atypical sites and conditions over larger areas 

climatic data could be included in the data set (Cox et al. 1984), and are seldom replicated 

to facilitate recommendations that take into in multiple seeding years (Casler 1999). 

consideration environmental parameters. Generally, high variability in experimental 

procedures often produces unique individual 

direct Conservation effects of Plant studies from a complex combination of 

materials on nutrient Cycling unique site preparation, plant materials, 

Ecosystem disruption commonly results in seeding rate, soil conditions, and weather 

a shift in nutrient-cycling dynamics from during only 1 or 2 establishment years. An 

systems where nutrients, such as carbon (C) important recommendation is adoption of 

and nitrogen (N), are quickly sequestered by minimum experimental design requirements 

plants and microorganisms to systems that for publication of range planting studies 

contain relatively greater amounts of available relative to specific inferences that are of 

nutrients (Norton et al. 2007). These systems principal interest (Casler 1999; Vargas et al. 

are more susceptible to weed dominance, 2001). 

leach more mineral N, and sequester less 

carbon than effectively functioning systems. Success metrics are highly variable and often 

Although these systems have higher N consist of relative ranking of treatment effects, 

mineralization rates, soil N concentration and there has been very little research to 

is lower (Kulmatiski and Beard 2006). evaluate alternative criteria for quantification 

Nutrient-cycling conservation effects could be of success (Ries and Svejcar 1991). The 

cited as an additional positive purpose for this majority of range planting studies only 

conservation practice standard. consider treatment effects in the first year after 

planting, and studies that are monitored for 

Adoption of standard Protocols for longer periods are generally not replicated for 

evaluating success and development of planting-year effects (Casler 1999). Many 

meta-Analysis of field trials studies that have monitored range planting 

The majority of range planting studies do results in the very long term have noted 

not measure critical environmental factors significant changes from what would have been 

affecting success, but only measure relative measured only 1–3 years postplanting (Bleak 

treatment effects (Call and Roundy 1991; et al. 1965; Hull 1971a, 1973; Lavin and 

Bluebunch wheatgrass 

seedlings (Photo: Lori 

Ziegenhagen, 2007) 

CHAPTER CHAPTER 42: Assessment of Prescribed Fire as a Conservation Practice : Assessment of Range Planting as a Conservation Practice 189 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

Standard rangeland drill 

conducting a native seeding 

following the East Humboldt 

fire, Elko Co., NV. (Photo: 

USFS, 2008) 

Johnson 1977; Eck and Sims 1984; Harris and 

Dobrowolski 1986). 

In general, most individual studies within 

the range planting literature are insufficiently 

replicated to extract valid inferences about 

weather and climate effects, site effects, 

plant-materials effects, and seeding rate. The 

dominant level for validation in the currently 

available literature derives from interspersion 

and within-location replication of seedbed preparation treatments. These studies, 

and a large amount of data contained in 

conference proceedings, technical reports, 

and internal-agency documents, might be 

subject to valuable meta-analysis of treatment 

effects that are difficult or impossible to 

replicate in the context of a stand-alone 

journal publication (Durlak and Lipsay 1991; 

Gurevitch et al. 1992; Adams et al. 1997; 

Michener 1997; Gurevitch and Hedges 1999; 

Osenberg et al. 1999a, 1999b; Gurevitch 

et al. 2001; Johnson 2006). Much of this 

information may only be suitable for low-level 

meta-analysis similar to the summary statistics 

used here to document gross treatment effects. 

It may be possible, however, to develop 

guidelines for establishing some common 

experimental design features for future studies 

that may be amenable to more sophisticated 

meta-analysis. 

Another underutilized research resource is 

the incorporation of extensive managementlevel monitoring information into a 

scientific database format (Pastorok et al. 

1997). Eiswerth and Shonkwiler (2006) 

used a Bureau of Land Management 

data set to evaluate postfire management 

treatment effects on seeded nonnative 

grasses, sagebrush, and annual weeds as a 

function of range site, soil type, and seeding 

prescription. Unfortunately, this data set 

did not evaluate impacts of weather and 

climate variability. Effective utilization of 

these types of data may also require some 

degree of coordination within and between 

management agencies to adopt similar 

monitoring protocols. NRCS Conservation 

Practice Standards could be improved 

by establishing standard monitoring 

requirements to assess both the effectiveness 

of specific management recommendations 

and conservation effects of successful 

practices. Monitoring requirements, 

however, should be based on an explicit 

experimental design that would facilitate 

future meta-analysis. 

ConClusIons 

There is virtually no refereed journal 

literature directly linking NRCS rangeland 

seeding conservation practices to specific 

conservation effects. The aggregate literature, 

however, generally supports general 

conservation practice recommendations 

for rangeland seeding, and the potential 

conservation benefits should these practices 

result in successful establishment of a 

more desirable plant community. A major 

limitation to current conservation practice 

recommendations is that they do not explicitly 

acknowledge or provide management 

guidance to deal with the high variability 

in soil microclimate during germination, 

emergence, and early seedling development. 

Additional guidance is warranted to provide 

recommendations for monitoring and 

adaptive management in these arid and 

semiarid rangelands. Future research efforts 

would also benefit from experimental designs 

that were amenable to meta-analysis, as 

inferences from individual rangeland planting 

trials are generally limited to specific site 

conditions and plant materials. 

190 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

S. P. Hardegree, T. A. Jones, B. A. Roundy, N. L. Shaw, and T. A. Monaco 

literature Cited 

Aase, J. K., and J. R. Wight. 1973. Prairie 

sandreed (Calamovilfa longifolia): water 

infiltration and use. Journal of Range 

Management 26:213–214. 

Abawi, G. Y., R. J. Smith, and D. K. Brady. 

1995. Assessment of the value of long range 

weather forecasts in wheat harvest management. 

Journal of Agricultural Engineering Research 

62:39–48. 

Abbott, L. B., and B. A. Roundy. 2003. 

Available water influences field germination and 

recruitment of seeded grasses. Journal of Range 

Management 56:56–64. 

Abbott, M. L., L. Fraley, and T. D. Reynolds. 

1991. Root profiles of selected cold desert 

shrubs and grasses in disturbed and undisturbed 

soils. Environmental and Experimental Botany 

31:165–178. 

Abrahams, A. D., A. J. Parsons, and J. 

Wainwright. 1993. Resistance to overland 

flow on semiarid grassland and shrubland 

hillslopes, Walnut Gulch, southern Arizona. 

Journal of Hydrology 156:431–436. 

Adams, D. C., J. Gurevitch, and M. S. 

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212 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	

CHAPTER 

5 

A Scientific Assessment of 

the Effectiveness of Riparian 

Management Practices 

Mel R. George,1 Randy D. Jackson,2 Chad S. Boyd,3 and Ken W. Tate1 

Authors are 1Cooperative Extension Specialist, California Rangeland Research and
 

Information Center, Plant Sciences Department, University of California,
 

Davis, CA 95616, USA; 2Associate Professor, Department of Agronomy,
 

University of Wisconsin, Madison, WI 53706, USA; and
 

3Rangeland Scientist, USDA-ARS, Burns, OR 97720, USA.
 

Correspondence: Melvin R. George, Plant Sciences Dept,
 

University of California, One Shields Avenue, Davis, CA 95616, USA. Email:
 

mrgeorge@ucdavis.edu
 

Reference to any commercial product or service is made with the understanding 

that no discrimination is intended and no endorsement by USDA is implied 

213
 
	 	 	 	

	 	 	

	 	 	

	 	

	 	 	 	

	 	

	 	 	

	 	 	 	

	 	 	

The major objective 

of riparian habitat 

“

conservation practices 

is to effectively manage 

riparian vegetation, 

stream channel, and 

soil resources to protect 

or enhance these 

ecosystem services”

214 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

 

 

 

 

 

 

 

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	

	 	 	

A Scientific Assessment of 

the Effectiveness of Riparian 

Management Practices 5 

Mel R. George, Randy D. Jackson, Chad S. Boyd, and Ken W. Tate 

IntRoductIon 

This chapter evaluates the ecological 

effectiveness of the major purposes and 

expected benefits of 21 riparian management 

practices as described in the US Department of 

Agriculture, Natural Resources Conservation 

Service (USDA NRCS), National Conservation 

Practice Guidelines (Table 1). The ecological 

benefits described in the standards for these 

practices include the following: 

• Wildlife habitat 

• Water quantity and quality 

• Stream bank and soil stability 

• Carbon storage 

• Plant and animal diversity 

Riparian management encompasses many 

activities and practices that are applied directly 

to the riparian zone or that are applied in 

the uplands to influence the riparian zone. 

To meet numerous riparian management 

goals, conservation practices are often applied 

as a suite of practices called a resource 

management system. A resource management 

system may include several practices (e.g., 

prescribed grazing, stream crossing, riparian 

herbaceous cover) selected to meet sitespecific conditions and objectives. Riparian 

areas occur along watercourses or near water 

bodies. They are different from surrounding 

lands because of unique hydrologic, soil, and 

plant characteristics that support important 

ecosystem functions and services. 

Riparian areas occupy the transitional area 

between the terrestrial (dry) and aquatic (wet) 

ecosystems. Rangeland riparian areas include 

the stream, stream channel, and adjacent 

riparian vegetation. These areas also include 

seeps, springs, and small wetlands that have 

greater soil water relative to surrounding 

uplands. This does not include marshes, 

impoundments, estuaries, and other wetland 

habitats. Although riparian areas constitute 

only a fraction of the total land area on western 

rangelands, they generally support greater 

overall plant and animal species diversity, 

richness, and productivity than adjacent 

uplands. Access to riparian areas in rangeland 

systems is usually critical to sustaining the 

productive potential of the surrounding 

landscape. Riparian areas are often relatively 

long and narrow in relation to other landscape 

features. This characteristic creates significant 

interaction with other ecological sites within 

the landscape, supporting the exchange of 

materials and energy within the landscape. 

Numerous studies in the western United 

States have shown that riparian areas have 

been negatively impacted by timber harvest, 

road building, irrigation, grazing, and other 

human activities (Kauffman and Krueger 1984; 

Fleischner 1994; Magilligan and McDowell 

1997; Belsky et al. 1999). In many cases, these 

systems have been altered (e.g., down-cutting, 

head-cutting, and stream bank alteration) to 

the point that past geomorphic structure and 

function cannot be restored and returned to 

former conditions. Additionally, installation 

of dams and diversion of water have altered 

runoff timing and amounts, often resulting in 

irreversible changes in riparian characteristics. 

Where irreversible changes have occurred, some 

new desired condition becomes the objective of 

restoration. 

Because grazing is such a widespread practice 

on public and privately owned rangelands, 

assessment of grazing management practices 

is a significant part of this review. Platts 

(1978, 1990) rated the effect of several grazing 

strategies for stream–riparian habitat values 

based on his observations and professional 

Riparian corridor in eastern
 

Oregon. (Photo: Chad Boyd)
 

CHAPTER 5: Effectiveness of Riparian Management Practices 215 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

                   tABLE 1. List of 20 riparian practices and their expected ecosystem services. 

Practice name code 

Ecosystem services 

Wildlife 

habitat 

Water quality and 

quantity 

Stable stream 

banks and soils 

carbon 

storage 

diverse plant 

and animal 

communities 

Animal trails and walkways (feet) 575 X X 

Brush management (acres) 314 X X X X 

channel bank vegetation (acres) 322 X X X 

conservation cover (acres) 327 X X X 

critical area planting (acres) 342 X 

Fence (feet) 382 X X X X 

Filter strip (acres) 393 X 

Pest management (acres) 595 X X 

Prescribed burning (acres) 338 X X 

Prescribed grazing (acres) 528 X X X X 

Range planting (acres) 550 X X X X X 

Riparian forest buffer (acres) 391 X X X X X 

Riparian herbaceous cover (acres) 390 X X X X X 

Stream crossing 578 X X 

Stream habitat improvement and 

management (acres) 395 X X 

Stream bank and shoreline protection (feet) 580 X X 

tree/shrub establishment (acres) 612 X X X X X 

upland wildlife habitat management (acres) 645 X 

use exclusion (acres) 472 X X X 

Watering facility (no.) 614 X X 

experience (Table 2). Similarly, Kovalchik 

and Elmore (1991) rated the compatibility 

of grazing systems with willow-dominated 

communities (Table 3). While the effects of 

many of these grazing systems on riparian 

areas have been documented in case histories, 

rarely have they been tested with rigorous 

experimental designs and appropriate 

statistical analyses (Larsen et al. 1998). 

Both of these evaluations indicate that 

continuous grazing is not compatible with 

riparian areas and that rest or deferment 

from grazing, inherent in various forms 

of rotational grazing, tend to improve the 

riparian habitat values addressed in Tables 2 

and 3. Continuous grazing often results in 

heavy grazing use of the riparian area because 

livestock are attracted to riparian areas from 

the adjacent uplands. Even if the pasture is 

lightly stocked, grazing may be heavy because 

livestock preferentially use the riparian zone. 

Improperly applied rotational grazing systems 

can also result in heavy grazing and damage to 

riparian habitat. 

The objective of livestock grazing strategies 

and practices has been to increase plant 

and litter cover, encourage growth of 

desirable plant species, improve plant species 

composition, increase plant vigor, and protect 

riparian soil and stream banks from erosion. 

Grazing tactics or practices for maintaining 

216 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

                     

                   

                           

                             

                       

                         

                           

                         

                       

                     

                       

                 

     

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

or rehabilitating riparian areas include 1) 

controlling the timing and duration of 

riparian grazing by fencing riparian pastures 

within existing pastures, 2) fencing riparian 

areas to exclude livestock from riparian areas, 

3) changing the kind and class of livestock, 

4) reducing duration of grazing, 5) reducing 

grazing intensity, and 6) controlling season 

of use (Clary and Webster 1989; Platts and 

Nelson 1989). Annual management objectives 

for vegetation attributes (e.g., herbaceous 

plant stubble height, woody plant utilization, 

and vegetative ground cover) are frequently 

recommended or required to guide year-toyear grazing management decisions (Bauer 

and Burton 1993; Hall and Bryant 1995; 

Clary and Leininger 2000). The assumption is 

that meeting annual management objectives 

will be compatible with long-term resource 

objectives (e.g., stream bank stability, 

recruitment of woody plants, clean water; 

Clary and Leininger 2000). 

Scientific documentation that livestock 

grazing could damage riparian areas began 

in the 1980s (Skovlin 1984) and has been 

documented in numerous symposia (e.g., 

Warner and Hendrix 1984; Johnson et al. 

1985; Gresswell et al. 1989; Meehan 1991; 

Clary et al. 1992), literature reviews (Platts 

1981, 1982, 1991; Kauffman and Krueger 

1984; Skovlin 1984; Chaney et al. 1990, 

1993; Armour et al. 1994; Fleischner 1994; 

Rhodes et al. 1994; Kattelmann and Embury 

1996; Ohmart 1996), and government reports 

(US Department of the Interior, Bureau of 

Land Management [USDI BLM] 1994; US 

General Accounting Office [US GAO] 1988). 

Recognizing that riparian ecosystem services 

need to be protected, USDA NRCS, along 

with many other federal and state resources 

agencies, began to apply existing conservation 

practices and to implement new practices with 

the goal of protecting and improving riparian 

habitats. 

dEScRIPtIon oF conSERvAtIon 

PRActIcES And BEnEFItS 

More than 40 practices in the USDA NRCS 

National Conservation Practice Guidelines 

(USDA NRCS 2003) were identified as having 

potential for application to riparian ecosystems. 

For this review, we narrowed the practices in 

Appendix I to a shorter list of 20 that are often 

associated with rangeland or pasture systems 

(Table 1). The purposes or anticipated benefits 

stated in the practice standards for these 20 

practices can be summarized into five main 

ecosystem services: 1) high-quality and abundant 

fish and wildlife habitat, 2) clean and plentiful 

water supply, 3) stable stream banks and riparian 

soils supporting hydrologic functions such as 

flood and pollutant attenuation, 4) carbon 

sequestration, and 5) diverse, rich, productive 

plant and animal communities (Table 1). 

However, the major objective of riparian habitat 

conservation practices is to effectively manage 

riparian vegetation, stream channel, and soil 

resources to protect or enhance these ecosystem 

services (Fig. 1). 

objective and Approach 

Recognizing that anticipated benefits of 

management practices applied to riparian 

habitats are mediated by resource availability, 

especially water, we developed a conceptual 

model that links management practices to 

vegetation attributes and resource constraints 

(Fig. 1). The model acknowledges the 

overriding importance of state factors 

such as climate, parent material, relief, 

geomorphology, past and contemporary land 

uses, and disturbances at the watershed and 

FIguRE 1. Conceptual model of the effect of riparian conservation practices on vegetation, soils, stream banks, and ecosystem services. The outcome of conservation 

practices are contingent on the biophysical context, which is set at a coarse level 

by state factors such as climate, relief, parent material, and the age of the riparian 

area. These state factors interact to shape the geomorphology and current condition 

of the riparian zone. Land use at the landscape level of organization, including conservation practices in uplands, is considered a state factor because it is controlled by 

humans. Within the bounds set by a particular combination of state factors, riparian 

management can affect vegetation and soils. In addition to these direct effects, feedbacks exist between vegetation and soils that condition their individual and collective response to management. Ecosystem services in the form of forage production, 

carbon sequestration, wildlife habitat, and flood and pollutant attenuation emerge 

from these dynamics processes. 

CHAPTER 5: Effectiveness of Riparian Management Practices 217 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	

	 	 	

Practices such as fencing to 

manage grazing pressure on 

riparian vegetation and soils 

are anticipated to enhance 

riparian-based ecosystem 

services. (Photo: Ken Tate) 

larger spatial scales on riparian vegetation 

and soils. It is widely documented that single 

or cumulative watershed-scale management 

practices (e.g., upland brush management, 

upland grazing management) and disturbances 

(e.g., fire, road construction) can affect 

riparian area functions, services, and response 

to site-specific management practices. For the 

purposes of this review, we focused on the 

interaction of various management practices 

with riparian soil and stream bank resource 

availability (i.e., water, nutrients, oxygen), and 

vegetation. We used the model illustrated in 

Figure 1 to generate 21 hypotheses that could 

be evaluated using published experimental 

data. The experimental data associated with 

these selected practices was identified by 

reviewing primarily peer-reviewed literature. 

Support for most hypotheses is summarized 

and incorporated into appendices to 

provide an evidence-based assessment of the 

effectiveness of these riparian management 

practices. 

EvALuAtIon oF RIPARIAn 

MAnAgEMEnt PRActIcES 

We classified 21 hypotheses into three 

riparian management purposes: 1) protection 

or restoration of vegetation attributes, 2) 

protection or restoration of stream channel and 

riparian soil stability, and 3) direct or indirect 

protection or improvement of ecosystem 

services (Fig. 1). In this section, each of the 21 

hypotheses is evaluated against the supporting 

experimental data. 

218 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	 	

	 	 	

	 	

	 	 	

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

Practices that Protect or Restore 

vegetation Attributes 

Hypothesis 1: Management of Time, 

Intensity, Season, and Duration of Grazing 

Affect Herbaceous Species Composition. 

Grazing systems facilitate control of season 

of use, frequency of use, duration of use, 

grazing intensity, and livestock distribution. 

However, herbaceous plant community 

response to grazing management is often 

difficult to predict because the responses are 

contingent on resource availability, namely, 

water (Stringham et al. 2001; Poole et al. 

2006). Resource availability is moderated by 

the biophysical characteristics of the riparian 

area and its watershed (Fig. 1; Goodwin et al. 

2008). In general, increasing grazing intensity 

results in a reduction of slower-growing, largerseeded plant species (i.e., competitive species 

sensu Grime 1979) that are often considered 

desirable, depending on management 

objectives. As grazing intensity increases, the 

abundance of faster-growing, small-seeded 

species (i.e., ruderals) increases; however, this 

response may be less prominent where water is 

in abundant supply. Moreover, where water is 

limiting or the supply is erratic, ruderals may 

dominate even with little or no grazing. This 

model is an oversimplification because in many 

riparian systems, resource limitations (e.g., 

moisture, nutrients, and temperature) occur in 

transient pulses (Seastedt and Knapp 1993). 

Flooding events may create microsites where 

only species tolerant of anoxia can persist. 

Lucas et al. (2004), working in New Mexico, 

found little effect of grazing intensity (no, low, 

and moderate) on herbaceous structure (cover, 

biomass) and composition (diversity). However, 

cool-season grazing promoted herbaceous 

diversity over warm-season grazing. The authors 

were adamant that grazing management 

affects streams in site-specific ways; hence, no 

single prescription is warranted for riparian 

management. This echoes findings of Jackson 

and Allen-Diaz (2006), who found highly 

variable interannual community characteristics 

in spring-fed wetlands that appeared unrelated 

to grazing intensity, while subsequent firstorder streamside vegetation appeared directly 

linked to grazing treatments. 

Lunt et al. (2007), working in a southeastern 

Australian riparian forest, showed that grazing 

exclusion had minimal impacts on understory 

composition and structure over a 12-yr period, 

attributing this to the fact that their system was 

nonequilibrial and responded more to abiotic 

factors than to biotic factors, such as grazing 

management. Clary (1999) found that all 

grazing treatments (0, 20–25%, and 35–50% 

utilization) resulted in increased plant species 

richness on streamsides and meadows as the 

systems recovered from historic heavy grazing. 

This indicates that the ecological condition of 

the riparian habitat at the onset of the study 

has important implications for the potential 

outcomes that may result from various 

management practices. 

Kauffman et al. (1983a) observed a 

phenological shift in the herbaceous plant 

community of mesic and hydric riparian zones 

in eastern Oregon that they ascribed to quicker 

drying of grazed soils resulting from greater 

solar insolation incident on the soil surface. 

Their data showed an increase in undesirable 

plant species with grazing compared to 

exclosures, though the experimental design 

was weak and no estimate of uncertainty was 

reported. The grazing prescription during this 

study was 75% utilization of bluegrass (Poa 

spp.) meadows. 

Lyons et al. (2000a) focused on the effects of 

different types of riparian vegetation on small 

streams in central North America and indicated 

that without grazing, these zones will become 

dominated by woody species that reduce stream 

bank stability. Paine and Ribic (2002) found 

more diverse plant communities and wildlife 

habitat when grassy buffer strips were present 

along riparian zones compared to woodydominated riparian zones. In contrast, Carline 

and Walsh (2007) show that in Pennsylvania, 

exclusion of grazing for 3–5 yr, from formerly 

heavily stocked pastures, resulted in vegetation 

cover increases from 50% to 100%. 

Our review of 11 reports found substantial 

support for the hypothesis that grazing 

intensity influences herbaceous species 

composition. However, managers should 

be aware that grazing effects on species 

composition may be influenced by the 

availability of resources, such as water and 

nutrients. Three of these studies (Lucas et al. 

2004; Jackson and Allen-Diaz 2006; Lunt et 

Eleven studies 

supported the 

hypothesis that 

grazing intensity 

influences 

herbaceous 

species 

composition 

however, grazing 

effects on species 

composition may 

be influenced by 

the availability of 

resources, such 

as water and 

nutrients”

CHAPTER 5: Effectiveness of Riparian Management Practices 219 
	 	 	 	

 

 

 

 

 

                               

	 	

	 	 	 	 	 	

	 	

	 	

	 	

	 	 	 	

	 	 	 	 	 	

	 	

	 	 	 	

	 	 	 	

                         

tABLE 2. Evaluation and rating of grazing strategies for stream–riparian-related fisheries values based on observations of Platts (1990). 

Strategy 

Level to which 

riparian 

vegetation is 

commonly used 

control 

of animal 

distribution 

(allotment) 

Stream bank 

stability 

Brushy species 

condition 

Seasonal plant 

regrowth 

Stream–riparian 

rehabitative 

potential Rating 

continuous season-long (cattle) Heavy Poor Poor Poor Poor Poor 11 

Holding (sheep or cattle) Heavy Excellent Poor Poor Fair Poor 1 

Short duration–high intensity 

(cattle) Heavy Excellent Poor Poor Poor Poor 1 

three herd–four pasture (cattle) Heavy to 

moderate Good Poor Poor Poor Poor 2 

Holistic (cattle or sheep) Heavy to light Good Poor to good Poor Good Poor to excellent 2–9 

deferred (cattle) Moderate to 

heavy Fair Poor Poor Fair Fair 3 

Seasonal suitability (cattle) Heavy Good Poor Poor Fair Fair 3 

deferred rotation (cattle) Heavy to 

moderate Good Fair Fair Fair Fair 4 

Stuttered deferred rotation 

(cattle) 

Heavy to 

moderate Good Fair Fair Fair Fair 4 

Winter (sheep or cattle) Moderate to 

heavy Fair Good Fair Fair to good Good 5 

Rest–rotation (cattle) Heavy to 

moderate Good Fair to good Fair Fair to good Fair 5 

double rest–rotation (cattle) Moderate Good Good Fair Good Good 6 

Seasonal riparian preference 

(cattle or sheep) 

Moderate to 

light Good Good Good Fair Fair 6 

Riparian pasture (cattle or 

sheep) As prescribed Good Good Good Good Good 8 

corridor fencing (cattle or 

sheep) None Excellent Good to 

excellent Excellent Good to 

excellent Excellent 9 

Rest–rotation with seasonal 

preference (sheep) Light Good Good to 

excellent 

Good to 

excellent Good Excellent 9 

Rest or closure (cattle or sheep) None Excellent Excellent Excellent Excellent Excellent 10 

1Rating scale based on 1 (poorly compatible) to 10 (highly compatible) with fishery needs. 

al. 2007), which were conducted in semiarid 

rangeland, concluded that grazing intensity 

effects were either nonexistent or overwhelmed 

by abiotic drivers. There was some support 

for a season, duration, or frequency of grazing 

effect on herbaceous species composition, but 

these studies were conducted mainly in the 

mesic upper Midwest or eastern grasslands 

(Lyons 2000b; Carline and Walsh 2007), 

where resources such as water and nutrients 

are typically in greater and more consistent 

supply. Studies that focused on mesic systems 

supported an increase in woody species with 

grazing exclusion. 

Hypothesis 2: Management of Time, 

Intensity, Season, and Duration of Grazing 

Can Influence Aboveground Herbaceous 

Productivity. Compensatory growth is the 

stimulation of net primary productivity (NPP) 

220 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	 	

	 	 	

	 	

	 	

	 	

	 	 	

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

by defoliation such that regrowth compensates 

for the biomass removed by the defoliation 

process (Bartolome 1993). Overcompensation 

occurs when defoliation results in production 

that exceeds that of undefoliated plants. 

Riparian herbaceous vegetation is often very 

productive since riparian areas usually possess 

abundant nutrients and water. Hence, it 

is plausible that grazing management that 

does not degrade resource availability may 

also result in compensatory growth, as was 

found by Boyd and Svejcar (2004) in eastern 

Oregon. This mechanism was implicated by 

Jackson et al. (2006) as the reason for greater 

nitrate loss from spring-fed wetlands of the 

Sierra Nevada foothill oak woodlands. In 

this study, grazing stimulated production 

compared to no grazing, which promoted 

uptake of nitrate entering the wetlands from 

the surrounding landscape. Alternatively, 

removal of grazing resulted in an immediate 

increase in total standing biomass, but this 

biomass accumulated, as dead plant material, 

on the surface and suppressed subsequent 

productivity, a phenomenon also observed by 

Popolizio et al. (1994) in a Colorado riparian 

zone. 

Huber et al. (1995) observed lower standing 

biomass under moderate grazing intensity 

compared to low and no grazing treatments, 

which were not different from each other. 

Caution must be used when interpreting peak 

standing biomass data because it is difficult 

to know whether a response to the treatment 

or the treatment itself is being measured. 

Productivity could be equal to or greater than 

the control, but standing biomass may be 

lower because of livestock utilization. 

Studying grazing effects of pack stock in 

Sierra Nevada mountain meadows, Cole et al. 

(2004) found reduced vegetation productivity 

over 5 yr in three meadow communities. 

Stohlgren et al. (1989) conducted a clipping 

experiment on high-elevation subalpine 

meadows of the Sierra Nevada. They found 

that clipping for 5 yr to simulate heavy 

grazing negatively affected productivity in 

wet and mesic meadows but not in dry Carex 

exserta meadows. The authors caution that 

these results cannot be extrapolated to address 

grazing at light or moderate levels. However, 

these results support the general notion that 

grazing effects on productivity are likely 

to be more pronounced in systems where 

resource availability is relatively high, such 

as mesic compared to dry meadows, where 

environmentally driven resource limitation 

has a stronger influence. This is to say not that 

grazing management has no effect in resourcepoor systems but rather that productivity is 

inherently low or variable and therefore less 

coupled to management. 

Late-season clipping in a Sierra Nevada 

mountain meadow had no consistent effects 

on above- and belowground response variables, 

such as root growth and photosynthetic 

rates (Martin and Chambers 2002), similar 

to the late-season clipping results of Clary 

(1995, 1999). Kluse and Allen-Diaz (2005) 

clipped Sierra Nevada meadows dominated 

by Deschampsia caespitosa and Poa pratensis 

early in the growing season and found reduced 

productivity in both species but no shift in 

relative species abundance. Huber et al. (1995) 

found that light grazing of a Sierra Nevada 

meadow resulted in vegetation biomass similar 

to ungrazed meadows but encouraged cattle to 

graze away from streamside edge compared to 

heavy grazing. Allen and Marlow (1994) found 

that beaked sedge (Carex rostrata) tolerated 

light to moderate grazing in early summer and 

fall if there was at least 60 d of rest between 

grazing periods to allow production of new 

photosynthetic tissue. 

Nine peer-reviewed reports support that 

grazing intensity can influence herbaceous 

productivity. Two reports (Boyd and Svejcar 

2004; Jackson et al. 2006) support a 

compensatory grazing effect on productivity, 

and two reports (Popolizio et al. 1994; 

Jackson et al. 2006) found that exclusion 

resulted in an accumulation of standing 

biomass that subsequently suppressed 

productivity. One report (Kluse et al. 2005) 

concluded that early-season clipping reduced 

productivity of two grasses. Three of these 

studies (Stohlgren et al. 1989; Huber et 

al. 1995; Cole et al. 2004) concluded that 

resource availability mediated the effect of 

grazing intensity on herbaceous productivity. 

Three studies in Rangelands found that lateseason clipping had no consistent effect on 

above- and belowground productivity (Clary 

1995, 1999; Martin and Chambers 2002). 

The peer-reviewed literature 

generally supports the effectiveness of water 

developments, 

supplement placement and herding 

for reducing riparian vegetation 

utilization, or 

time spent in 

riparian areas”

CHAPTER 5: Effectiveness of Riparian Management Practices 221 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 	 	

	 	 	

Cattle grazing a small riparian 

patch within a sagebrush 

community. (Photo: Chad Boyd) 

We can conclude that there is no universal 

riparian herbaceous production response 

to the complex elements (time, intensity, 

season, or duration) of grazing management. 

Consequently, what a manager learns on one 

site may not be transferable to another site. 

Hypothesis 3: Livestock Distribution 

Practices Reduce Time Spent in Riparian 

Zones or Riparian Vegetation Utilization. 

The peer-reviewed literature generally supports 

the effectiveness of water developments, 

supplement placement and herding for 

reducing riparian vegetation utilization, 

or time spent in riparian areas. Bailey 

(2004, 2005) and George et al. (2007) have 

reviewed practices that attract livestock to 

underused areas and away from riparian 

habitats. Abiotic and biotic characteristics 

of landscapes and pastures influence the 

effectiveness of these practices. A few studies 

document the effectiveness of drinkingwater developments, herding, and strategic 

placement of supplemental feeds for reducing 

grazing use and the time spent in riparian 

areas. Nine out of 10 studies (seven peer 

reviewed, one thesis, and two in rangelands) 

report that development of off-stream stock 

water reduces grazing use or time spent 

in riparian areas. Six of these studies were 

conducted in eastern Oregon. McGinnis and 

McIver (2001) reported that the extent to 

which livestock can be enticed away from 

riparian areas depends on season, topography, 

vegetation, weather, and behavioral differences 

among animals. Ehrhart and Hansen (1998) 

evaluated ecological function on 71 streams 

in Montana and found that off-stream 

water developments resulted in improved 

ecosystem health. A few studies have shown 

that most grazing use occurs within 400 m 

of stock water sources (Pinchak et al. 1991). 

Thus, water developments placed at this 

distance or beyond may be more effective at 

reducing livestock use in riparian areas than 

closer installations. Two studies in California 

(McDougald et al. 1989; George et al. 2008) 

and one in Montana (Bailey et al. 2008a) have 

demonstrated the effectiveness of strategic 

supplement placement for attracting livestock 

away from riparian areas, and one study in 

Montana documented the effectiveness of 

herding with or without supplementation 

for reducing grazing use in the riparian area. 

Additional studies in Montana have shown 

the effectiveness of supplement as a cattle 

attractant. One study in Nevada documented 

the effectiveness of shade structures for 

reducing riparian use. The results of these 

studies are reinforced by studies in California 

and Montana that found that riparian health 

was related to time invested in management 

by the landowner or manager (Erhart and 

Hansen 1998; Ward 2002; Ward et al. 2003). 

Most of the data supporting these findings 

come from Oregon (Great Basin or forest), 

California (oak-woodland and annual 

grassland), or Montana (plains). We conclude 

from these studies that water developments, 

strategic supplement placement, and herding 

can effectively reduce time spent in riparian 

zones and riparian vegetation use by livestock. 

Because the effectiveness of these practices is 

often controlled more by abiotic (topography 

and distance from water) landscape 

characteristics than by biotic characteristics, 

we believe that they can be generalized 

to other rangeland ecosystems. Livestock 

attraction practices work best on gentle slopes 

and become less effective as slope increases. 

Narrow riparian corridors that are bound 

by steep slopes with limited available highquality forage or water are generally not good 

candidates for these practices. 

Hypothesis 4: Under Initially Degraded 

Conditions, Grazing Exclusion Can 

Promote Recovery of Riparian Plant 

Community Composition. The peer-reviewed 

literature generally supports the hypothesis 

that grazing exclusion can promote recovery 

of riparian plant community composition 

in degraded riparian systems. Fencing and 

use exclusion are commonly used to remove 

grazing from riparian areas permanently or 

during recovery periods. Many reports of 

the impacts of grazing on riparian areas and 

associated aquatic ecosystems come from 

comparisons of grazed and ungrazed areas 

(Larsen et al. 1998; Sarr 2002). Working in 

north-central Colorado on montane riparian 

areas, Popolizio et al. (1994) showed that 

long-term grazing altered plant community 

composition and cover characterized by more 

bare ground, dandelion (Taraxacum officinale), 

and clover (Trifolium repens) compared to 

ungrazed areas. Similar findings were reported 

222 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	 	

	 	

	 	

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

by Schulz and Leininger (1990) within the 

riparian zone bordering Sheep Creek in northcentral Colorado. Compositional changes 

from forb- or nonnative grass–dominated 

communities toward native grass– and sedgedominated communities have been widely 

documented in montane riparian meadow with 

grazing exclusion (Leege et al. 1981; Kauffman 

1983b; Schulz and Leininger 1990; Green and 

Kauffman 1995). 

We can conclude from these studies that 

grazing exclusion can promote recovery of 

initially degraded riparian plant community 

composition. However, plant species richness 

has not shown a clear response to grazing 

exclusion, though a few experimental results 

have been reported in the peer-reviewed 

literature (Bowns and Bagley 1986; Green and 

Kauffman 1995). 

Hypothesis 5: Livestock and Other 

Large Herbivores Modify Structure and 

Composition of Woody Plant Communities. 

The literature clearly indicates that livestock 

and native ungulates can modify the structure 

and composition of woody plant communities 

in riparian habitats. The vast majority of 

papers dealing with woody plants were from 

the western and northwestern United States; 

work from the southwestern United States 

was limited, and southern Plains publications 

were lacking. Fourteen of 16 papers (Appendix 

V) indicated structural or compositional 

modification of woody plant communities 

as a result of livestock grazing (Green and 

Kauffman 1995; Samuelson and Rood 2004; 

Holland et al. 2005). Papers by Sedgwick and 

Knopf (1991) and Lucas et al. (2004) did 

not clearly show structural or compositional 

effects of grazing on woody plant communities. 

Two papers indicated negative effects of deer 

(Odocoileus hemionus) browsing (Opperman 

and Merenlender 2000; Matney et al. 2005), 

two papers indicated negative impacts of 

elk (Cervus canadensis) or moose (Alces alces) 

herbivory (Kay 1994; Zeigenfuss et al. 2002) 

on woody plants, and Case and Kauffman 

(1997) reported reductions in woody plant 

abundance as a result of combined deer and elk 

herbivory. 

Establishment and maintenance of woody 

plants can be associated with episodic 

disturbance events (Rood et al. 2007; Bay and 

Sher 2008); therefore, evaluation of the effects 

of grazing on establishment and maintenance 

of woody plants should ideally occur over a 

sufficient time interval to encompass critical 

disturbance events. Auble and Scott (1998) 

reported that recruitment of cottonwood 

decreased with cattle grazing but that 

recruitment was highly dependent on infrequent 

high flow conditions that created suitable 

habitat for seedlings. Conversely, Sedgwick and 

Knopf (1991) thought grazing to be a relatively 

minor impact on willows (Salix spp.) and 

cottonwoods (Populus spp.) in comparison to 

periodic catastrophic flooding (which washed 

out woody plant habitat). Manoukian and 

Marlow (2002) concluded that willow canopy 

cover fluctuated along streams from 1942 

to 1985 but that the trend was upward in a 

USDA Forest Service grazing allotment. They 

concluded that extended periods (>3 yr) of 

rest were not necessary for willow recovery if 

livestock or wildlife use was closely controlled. 

In many cases, livestock use of woody plants 

may constitute only a portion of total use when 

native ungulates are considered. For example, 

Kay (1994) reported that tall willows had 

disappeared from 41 of 44 historical photo sets 

in Yellowstone National Park in association 

with elk and moose herbivory. Grazing can 

also affect woody plants through alterations 

in site hydrology. Such alterations may take 

the form of direct alterations in physical 

characteristics of the stream channel associated 

with changes from high- to low-root-density 

vegetation as discussed under hypothesis 8. 

These modifications could indirectly decrease 

site availability for riparian woody plants by 

decreasing available water. 

The influence of livestock on woody plant 

structure is complex and dependent on a 

variety of management and environmental 

site factors. Hypothesis 3 makes clear that 

livestock usage of riparian areas is variable 

and predicated on a variety of management 

and environmental factors. From hypothesis 

5, we can conclude that livestock and other 

large herbivores can modify the structure and 

composition of woody plant communities, 

but the impacts of livestock on woody plant 

resources are likely to be highly variable 

from location to location and within a given 

location over time. 

Based on these 

studies, there 

is sufficient 

evidence to 

conclude that 

fall livestock 

use of riparian 

areas can lead 

to increased 

utilization of 

woody plants”

CHAPTER 5: Effectiveness of Riparian Management Practices 223 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	

For streams with the site potential to support riparian woody 

plants, consistent late season 

grazing can reduce woody 

plant recruitment and extent. 

(Photo: Ken Tate) 

Hypothesis 6: Late-Growing-Season 

Livestock Use Increases Utilization of 

Woody Plants. Eleven of the 17 papers 

associated with livestock impacts on woody 

plants reported on the effects of late-season 

use. Of those 11 papers, nine reported 

negative structural or compositional 

modification associated with late-season 

livestock utilization of woody plants (Schulz 

and Leininger 1990; Clary et al. 1996; 

Holland et al. 2005), and four papers 

specifically noted increased use during the 

late-season period (Roath and Krueger 1982; 

Kauffman et al. 1983a; Conroy and Svejcar 

1991; Green and Kauffman 1995). One paper 

found that dormant-season clipping had less 

negative impact on woody plant abundance 

than continuous elk use (Zeigenfuss 2002), 

and another paper reported decreased willow 

abundance associated with late-season deer 

use (Matney et al. 2005). Roath and Krueger 

(1982) reported an inverse relationship 

between degree of woody plant utilization and 

phenological maturity of herbaceous cover. 

Kauffman et al. (1983a) and Matney et al. 

(2005) noted that woody plant utilization 

by mule deer did not begin until herbaceous 

availability became limiting. Clary et al. 

(1996) concluded that spring grazing was less 

detrimental to woody plants than fall grazing. 

Based on these studies, there is sufficient 

evidence to conclude that fall livestock use of 

riparian areas can lead to increased utilization 

of woody plants. This temporal pattern of 

woody plant utilization is generally associated 

with reduced herbaceous plant availability or 

forage quality. 

224 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

         

           

     

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

Hypothesis 7: Riparian Burning Can 

Reduce Undesirable Woody Species and 

Restore Desired Herbaceous or Woody 

Vegetation. Few studies have addressed the 

effects of fire on riparian ecosystems (Dwire 

and Kauffman 2003). However, riparian 

species exhibit adaptations that facilitate rapid 

recovery following fire. Several species resprout 

following fire, including quaking aspen (Populus tremuloides), cottonwood, and willows. 

To the extent that fire can remove competition 

from undesirable species, desirable resprouting 

species may be restored. 

Reviews by Dwire and Kaufman (2003) and 

Pettit and Naiman (2007) point out that 

the effectiveness of riparian burning may be 

mediated by resource availability and grazing 

management. Riparian burning is not well 

studied, but these reviews offer several 

hypotheses related to interactions among 

climate, disturbance regime, landscape 

position, and fire frequency and intensity. 

They point out that even in fire-driven 

landscapes (e.g., savannas), riparian plant 

community composition and productivity 

is more likely to be controlled by water 

and nutrient availability afforded by the 

lower landscape position. The effects of 

burning will likely interact with grazing 

management with higher grazing intensities, 

reducing the effects and the likelihood of 

fire in riparian zones (Dwire et al. 2006). 

That said, if sufficient fuel is available, the 

effects of burning may depend on depth to 

the water table. Blank et al. (2003) burned 

riparian sites dominated by big sagebrush (Artemisia tridentata) to reduce its cover 

and favor herbaceous species in areas with 

shallow and deep water tables. Herbs 

that were present at the time of burning 

resprouted and contained higher nutrient 

concentrations following burning. However, 

herbs were less abundant where the water 

table was deeper prior to burning, so the 

postburn response was more favorable with 

shallower water tables. Proportionally more 

of the surface soil nutrients were lost from 

the riparian zones with deeper water tables, 

which does not bode well for the growth of 

herbs in these habitats. Thus, recovery from 

fire depends on the presence of residual 

herbaceous species and an adequate water 

table to support these species. 

tABLE 3. Grazing system compatibility with willowdominated plant communities in riparian habitats 

(Kovalchik and Elmore 1991). 

grazing practice 

compatibility 

with willows 

corridor fencing Highly 

Riparian pasture Highly 

Spring (early-season) 

grazing Highly 

Winter grazing Highly 

two-pasture rotation Moderately 

three-pasture rest 

rotation Moderately 

three-pasture deferred 

rotation Moderately 

Spring–fall pastures Incompatible 

deferred grazing Incompatible 

Late-season grazing Incompatible 

Season-long grazing Incompatible 

Literature concerning use of prescribed fire to 

control undesirable woody species in riparian 

zones relates mainly to the genus Tamarix. 

Three of four studies that incorporated fire 

as a treatment reported successful control 

of Tamarix with mortality rates up to 95% 

(McDaniel and Taylor 2003; Harms and 

Hiebert 2006; Bateman et al. 2008; Appendix 

VII). One study found that control of Tamarix 

was not related to burning or mechanical 

removal but instead was most closely associated 

with site and year factors, the most important 

of which was precipitation, with no Tamarix 

regrowth occurring on sites receiving less than 

20.8 cm of annual precipitation (Bay and Sher 

2008). Busch and Smith (1993) urged caution 

in the use of fire to control both Tamarix and 

Tessaria, as these genera possess ecophysiological 

adaptations that may favor their abundance over 

historically dominant willow and cottonwood 

in the postfire environment. A study in the 

Great Basin on riparian areas affected by stream 

incision and decreased water tables examined 

the use of prescribed fire to remove sagebrush 

and restore riparian obligate herbaceous species. 

Desired plant species increased and herbaceous 

biomass tripled on sites with residual riparian 

species and adequate water tables, but sites that 

CHAPTER5: Effectiveness of Riparian Management Practices 225 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

lacked residual species and had significantly 

lowered water tables were dominated by annual 

weeds postfire (Chambers and Linnerooth 

2001; Wright and Chambers 2002; Blank et al. 

2003). Fire has been used as an effective form 

of control for nonsprouting conifer species on 

upland sites (e.g., Bryant et al. 1983; Engle and 

Stritzke 1995; Miller et al. 2005) and may play 

an important role in controlling encroachment 

of those species into riparian areas. 

While there are insufficient experimental 

data to thoroughly evaluate this hypothesis, 

current literature suggests that fire can be 

used to control some species of woody plants; 

however, the success of fire-based restoration 

may relate strongly to the availability of 

propagules of desired species, which can be 

depleted if riparian degradation has led to 

decreased soil water availability. 

Scientific uncertainty and Livestock 

Exclosure Studies 

Comparison of grazed areas with ungrazed 

areas (exclosures) is a common practice that 

has the potential for erroneous interpretations. 

Sarr (2002) reviewed exclosure studies and 

reported that exclosure-based research has left 

considerable scientific uncertainty because of 

the popularization of relatively few studies, 

weak study designs, a poor understanding 

of the scales and mechanisms of ecosystem 

recovery, and selective, agenda-laden literature 

reviews advocating for or against public 

lands livestock grazing. Exclosures are often 

too small (<50 ha) and improperly placed 

to accurately measure the responses of 

aquatic organisms or geomorphic processes 

to livestock removal. Depending on the 

site conditions when and where livestock 

exclosures are established, postexclusion 

dynamics may vary considerably. Systems can 

recover quickly and predictably with livestock 

removal, fail to recover because of changes in 

system structure or function, or recover slowly 

and remain more sensitive to livestock impacts 

than they were before grazing was initiated. 

Sarr presents suggestions for strengthening the 

scientific basis for livestock exclosure research, 

including 1) incorporation of meta-analyses 

and critical reviews, 2) use of restoration 

ecology as a unifying conceptual framework, 

3) development of long-term research 

programs, 4) improved exclosure placement 

and design, and 5) a stronger commitment 

to collection of pretreatment data. Properly 

designed exclosure studies could provide 

useful insights into grazing effects, but few 

meet these criteria. 

Practices that Protect or Restore Stream 

Bank and Riparian Soil Stability 

Hypothesis 8: Riparian Management That 

Affects Plant Species Composition, Plant 

Vigor, Rooting Densities and Depth, and 

Ground Cover and Influences the Stability 

of Stream Channel and Riparian Soils 

That Derive Their Stability from Riparian 

Vegetation. The linkage between riparian 

management, riparian vegetation, and stream 

channel and riparian soil stability is complex 

(Fig. 1). Stream systems themselves are 

complex; Rosgen (1994) describes almost 100 

stream channel categories. State factors such 

as watershed size, geomorphology, parent 

material, climate, and site-specific riparian 

vegetation attributes interact to define the 

structure and function of each stream segment 

(reach). Each stream reach may support 

and be supported by different riparian plant 

communities. Each reach may respond 

differently to watershed- or landscape-scale 

disturbances, and each may exhibit differing 

response to riparian management practices. 

Gordon et al. (1992), Leopold (1994), and 

Rosgen (1996) are excellent applications of our 

basic understanding of stream hydrology and 

applied river morphology from a watershed 

perspective. 

Eight studies and six reviews provide 

evidence of the importance of riparian plant 

communities and grazing management 

to stream bank and soil stability. Thorne 

(1982, 1990), Gregory (1992), and Trimble 

and Mendel (1995) discuss and document 

the general importance of riparian plant 

communities on stream channel and riparian 

soil stability. In the Sierra Nevada, Michelli 

and Kirchner (2002a) found that the 50-yr 

rate of stream channel migration and erosion 

was 6 and 10 times lower on stream banks 

and associated “wet” riparian areas covered 

with sedge (Carex spp.) and rush (Juncus 

spp.) compared to grass-dominated “dry” 

stream banks and associated riparian areas. 

In a companion study, Michelli and Kirchner 

(2002b) found that the tensile strength of wet 

226 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

riparian soils supporting sedge (Carex spp.) 

and rush (Juncus spp.) plant communities 

were five times stronger than dry riparian soils 

dominated by grass and shrub species. Soil 

tensile strength was positively correlated to 

plant density, biomass, and the ratio of root to 

soil mass. In northwestern Nevada, Manning 

et al. (1989) compared root mass and root 

length density across a soil moisture gradient 

represented by four herbaceous riparian plant 

communities. They found both root metrics 

to increase with soil moisture availability, 

indicating superior site-stabilizing capacity 

in the wetter plant communities. Klienfelder 

et al. (1992) report similar findings for 

riparian areas from central Nevada and eastern 

California. 

Research introduced to test hypotheses 1, 

2, 4, and 5 in this document establishes 

the capacity for riparian grazing, along 

a gradient of heavy to minimal grazing, 

to have primary effects on riparian plant 

composition, biomass, and cover. Combined 

with the discussion above, the capacity for 

riparian grazing management practices to 

have secondary effects on stream channel and 

riparian soil stability can be established. In 

general, it has been well documented (e.g., 

Kauffman and Krueger 1984; Sarr 2002) 

that incorrectly managed livestock grazing 

in riparian areas can 1) reduce plant root 

mass and rooting depth, which is critical 

for stabilizing riparian soils and stream 

banks against stream flow, and 2) shift plant 

community composition from high-rootdensity species to low-root-density species. 

The subsequent potential effects include: 1) 

stream banks and riparian soils becoming 

unstable; 2) stream channels with “hard” 

bottoms eroding laterally and widening; 

and 3) stream channels with “soft” bottoms 

eroding vertically to down-cut the channel 

and lowering the riparian water table. A 

positive feedback loop exists between lowered 

water tables and stream bank stability as it 

becomes increasingly difficult for high-rootdensity species that require wet habitats to 

reestablish (Toledo and Kauffman 2001). 

There are significant limitations with the 

literature addressing grazing impacts on 

stream channel and riparian soil stability 

and associated ecosystem services (e.g., 

aquatic habitat, flood attenuation). Several 

comprehensive reviews of essentially the 

same literature base substantiate the generally 

negative effects of “heavy” grazing, the 

generally positive response of riparian areas 

to complete removal of heavy grazing, the 

need for further research on “proper” grazing 

management strategies for riparian areas, and 

the need for increased rigor and consistency 

in case studies and experiments examining 

these riparian grazing strategies (Rinne 1988; 

Platts 1991; Ohmart 1996; Larsen et al. 

1998; Allen-Diaz et al. 1999; Sarr 2002). 

Based on these reviews and studies, there 

is sufficient evidence that riparian grazing 

management that maintains or enhances key 

riparian vegetation attributes (i.e., species 

composition, root mass and root density, 

cover, and biomass) will enhance stream 

channel and riparian soil stability, which 

will in turn support ecosystem services, 

such as flood and pollutant attenuation and 

high-quality riparian habitat. Lacking in the 

literature are watershed-level, statistically 

robust examinations of how stream channel 

and riparian soil stability are correlated with 

grazing management components, such as 

intensity, frequency, season, and duration of 

grazing across a set of riparian conditions. 

These should be compared for a variety 

to conditions, including degraded and 

undegraded riparian systems, herbaceousdominated and woody-dominated systems, 

and alluvial channel substrates versus bedrockdominated substrates. It is difficult to predict 

the specific impacts of riparian grazing 

management practices under differing levels 

of state variables (Fig. 1) as indicated in the 

results of Lucas et al. (2004) and Jackson 

and Allen-Diaz (2006) for hypothesis 1. 

However, it is clear that riparian grazing can 

be managed to enhance and protect primary 

riparian vegetation attributes that are strongly 

correlated to stream channel and riparian soil 

stability, which support ecosystem services 

provided by riparian areas (e.g., hypotheses 3 

and 4). There may be highly degraded riparian 

conditions, such as down-cut channels, under 

which riparian grazing management practices 

alone cannot restore the site’s former soil 

moisture regime that supported riparian plant 

communities with high rooting densities and 

their associated ecosystem services (Chambers 

and Miller 2004). 

Riparian corridors are
 

important for recreation. (Photo:
 

Chad Boyd)
 

CHAPTER 5: Effectiveness of Riparian Management Practices 227 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

	 	 	 	

	 	 	

Healthy riparian plant 

communities are effective at 

attenuating pollutants carried in 

runoff. (Photo: Ken Tate) 

Practices that Protect or Enhance 

Ecosystem Services 

Hypothesis 9: Riparian Vegetation Can 

Attenuate Pollutants Transported in 

Runoff, and Buffer Strip Effectiveness Is 

Dependent on Site-Specific Factors. The 

management of riparian vegetation to trap 

waterborne pollutants is commonly referred to 

as a vegetative “buffer” or “filter” strip. Small 

wetlands, either natural or constructed, can 

also provide this service. Reviews of research 

relevant to the implementation of vegetative 

buffers in riparian habitats of rangeland 

ecosystems and pastures can be found in 

Castelle et al. (1994), Schmitt et al. (1999), 

Dosskey (2002), Dorioz et al. (2006), and 

Mayer et al. (2007), among others. 

Attenuation efficiencies ranging from ~0 

to greater than 99% have been reported for 

pollutants common to rangelands and livestockgrazed systems, primarily nutrients, sediment, 

and indicator bacteria and pathogens (Dillaha 

et al. 1989; Pearce et al. 1998b; Atwill et al. 

2002, 2005; Bedard Haughn et al. 2004; Tate 

et al. 2004a, 2005; Dosskey et al. 2007; Knox 

et al. 2007, 2008). The variation observed 

across these studies can partially be attributed 

to site-specific differences in biophysical 

factors, such as buffer width, slope, vegetation 

attributes within the buffer, pollutant type and 

attributes, pollutant load entering the buffer, 

overland or flood flow rate entering the buffer, 

hydrologic residence time within the buffer, 

riparian soil attributes within the buffer, and 

228 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	

	 	

	 	

	 	 	

	 	

	 	 	

	 	

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

buffer vegetation management (Castelle et al. 

1994; Schmitt et al. 1999; Mayer et al. 2007). 

A significant number of studies have focused 

on determination of optimal buffer widths. 

No single buffer width can be prescribed for 

all scenarios, and there is increasing demand 

for decision support tools that develop first 

approximations of required buffer widths based 

on site factors (e.g., Dosskey et al. 2005, 2006, 

2008; Parajuli et al. 2008). While biophysical 

site factors determine buffer efficiency, the 

manager must also decide on an acceptable 

level of water quality degradation risk in the 

determination of buffer width. As risk tolerance 

decreases, buffer width must increase (Castelle 

et al. 1994; Atwill et al. 2005; Tate et al. 2005). 

Varying results have been reported for the 

effect of stubble height of herbaceous riparian 

vegetation on sediment and nutrient deposition 

and retention, indicating that this metric may 

not consistently impact, or predict, buffer 

efficiency (Abt et al. 1994; Clary et al. 1996; 

Pearce et al. 1997, 1998a, 1998b; Thorton et al. 

1997; Fraiser et al. 1998; Skinner 1998; Clary 

and Leininger 2000; Marlow et al. 2006). There 

is a consistently positive correlation between 

vegetative ground cover, plant stem density, and 

buffer filtration efficiency for several pollutants 

(e.g., Larsen et al. 1993; Corley et al. 1999; 

McEldowney et al. 2002; Davies et al. 2004; 

Tate et al. 2005). It is important to note that 

these same plant attributes are important for 

determining stream channel and riparian soil 

stability (H 8). Defoliation to manage buffer 

vegetation biomass accumulation, growth 

stage, and nutrient demand affects the nitrogen 

attenuation efficiencies of buffers (e.g., Mendez 

et al. 1999; Matheson et al. 2002; BedardHaughn et al. 2005; Jackson et al. 2006). 

As overland and flood flow rates entering a 

buffer increase and hydrologic residence times 

decrease, buffer attenuation and retention 

capacities can be reduced, if not completely 

eliminated (e.g., Bedard-Haughn et al. 2004; 

Tate et al. 2004a, 2005; Knox et al. 2007, 

2008). Biomass accumulation in buffers can 

create human health concerns. Excessive organic 

carbon near surface drinking water sources may 

lead to formation of carcinogenic–mutagenic 

by-products during chlorination (Krasner et al. 

1989; Jassby and Cloern 2000; Bull 2001). 

Based on 41 peer-reviewed reports, the 

overriding message is that 1) vegetative buffer 

strips can attenuate some portion of most 

waterborne pollutants transported by overland 

and flood flow events, and 2) there is significant 

variation in buffer attenuation efficiency 

attributable to site-specific factors. Supporting 

research ranges across a wide range of systems 

(e.g., urban, agricultural, rangeland), regions 

of the United States and the world and for a 

wide suite of pollutants, including sediment, 

nutrients, microorganisms, and pesticides. 

There is strong evidence supporting the overall 

assertion that riparian vegetation can function 

to attenuate waterborne pollutants in overland 

and flood flow events. 

Hypothesis 10: Practices That Reduce 

Livestock Densities, Residence Time, and 

Fecal and Urine Deposition in Riparian 

Areas and Stream Flow Generation 

Areas Can Reduce Nutrient and Pathogen 

Loading of Surface Water. In conjunction 

with implementation and management of 

vegetative buffers in riparian areas, additional 

water quality protection can logically be 

derived from implementation of livestock 

management strategies that distribute livestock 

fecal material and urine away from riparian 

areas, stream flow generation areas, and 

surface waters. In essence, this will create 

additional buffering length and capacity. 

Recent research on rangelands supports that 

livestock distribution practices can be applied 

to modify the spatial distribution of feces and 

urine deposition, creating buffering distances 

between feces and water bodies with minimal 

establishment of fences (Miner et al. 1992; 

Clawson 1993; Bailey et al. 1996; Bailey and 

Welling 1999; Bailey et al. 2001; Tate et al. 

2003; Blank et al. 2006; Bailey et al. 2008a, 

2008b). Cattle feces and urine distribution 

patterns on rangelands are significantly 

associated with location of livestock 

attractants, aspect, topographic position, and 

season (Tate et al. 2003; Bailey et al. 2008a). 

Strategic location of livestock attractants, 

including stock water, mineral supplements, 

and protein supplements, can have strong 

influences on patterns of cattle fecal and urine 

loads on watersheds. 

There is evidence to support the assertion 

that practices that reduce livestock densities, 

residence time, and fecal and urine deposition 

in riparian areas and stream flow generation 

There is strong 

evidence supporting the overall 

assertion that riparian vegetation 

can function to 

attenuate waterborne pollutants 

in overland and 

flood flow events”

CHAPTER 5: Effectiveness of Riparian Management Practices 229 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	

We conclude 

that grazing 

can decrease 

populations of 

riparian obligate 

avifauna but has 

variable effects 

on generalist 

species”

areas can reduce nutrient and pathogen 

loading of surface water. In addition, 

reducing livestock densities and residence 

time also reduces the negative effects of 

livestock on riparian vegetation attributes 

(e.g., stem density, cover) and soil hydrologic 

attributes (Gifford and Hawkins 1978; Tate 

et al. 2004b; e.g., bulk density, infiltration 

capacity). Thus, reducing livestock densities 

and residence time can benefit buffer 

efficiency (H 9) and site stability (H 8). 

To reduce livestock impacts in the riparian 

zone, stocking rate reductions are not the 

universal solution. The key is implementing 

practices that reduce livestock density in 

the riparian zone. These can be distribution 

practices, fencing to increase control of time 

and duration of grazing, and, in some cases, 

stocking rate reductions. 

Hypothesis 11: Riparian Grazing 

Decreases Habitat Quality for Prairie 

Wetland Avian Species. Nine of 28 avian 

references provided information on grazing 

management of wetland or prairie wetland 

habitat and associated avian species. In seven 

studies focusing on waterfowl (Duebbert et al. 

1986; Ignatiuk and Duncan 2001; Murphy 

et al. 2004), only four had ungrazed controls. 

Habitat quality was unchanged in two studies 

(Barker et al. 1990; West and Messmer 2006), 

decreased in a third (Kruse and Bowen 1996), 

and was unreported in a fourth (Littlefield and 

Paulin 1990); nesting success was unchanged, 

decreased, increased, or was not measured. One 

of three studies without controls indicated that 

heavy stocking rates did not provide adequate 

nesting cover (Duebbert et al. 1986), and two 

studies found no difference in nest success 

or habitat quality between season-long and 

rotational grazing strategies (Ignatiuk and 

Duncan 2001; Murphy et al. 2004). One 

of two passerine studies indicated decreased 

habitat quality and bird diversity with grazing 

(Taylor 1986), and a second study found that 

avian abundance and diversity were unaffected 

by grazing (May et al. 2002). 

The references reviewed here suggest that with 

the exception of heavily grazed areas, grazing 

in wetland habitat does not decrease habitat 

quality for waterfowl. Insufficient data exist to 

determine the influence of grazing on habitat 

quality for wetland passerine species. 

Hypothesis 12: Riparian Grazing Decreases 

Populations of Riparian Avifauna. Effects 

of livestock grazing on riparian avian habitat 

have been reviewed or summarized (Szaro 

1980; Bock et al. 1993; Fleischner 1994; 

Belsky et al. 1999). The importance of 

riparian vegetation as avian habitat has been 

described by numerous authors (Bull and 

Skovlin 1982; Douglas et al. 1992; Sanders 

and Edge 1998; Deschenes et al. 2003). 

Knopf et al. (1988a) reported that riparian 

vegetation attracts over 10 times the number 

of spring migrant birds found in upland 

sites and has 14 times more species during 

fall migration. References were fairly well 

distributed geographically except literature 

for the southern Plains, which was generally 

lacking. Evaluating the influence of grazing 

management practices on riparian wildlife was 

limited by insufficient details in many of the 

studies reviewed. These limitations relegated 

our assessment of grazing responses to a 

presence-and-absence standpoint. 

Nineteen studies report dynamics of riparian 

avifauna as a function of grazing. Of those, 

eight found no change in abundance 

(Kauffman et al. 1982; Sedgewick and 

Knopf 1987; Knopf et al. 1988b; Schulz and 

Leininger 1991; Warkentin and Reed 1999; 

Stanley and Knopf 2002; Scott et al. 2003; 

Martin and McIntyre 2007), five did not 

report or did not clearly report abundance 

(Neel 1980; Crawford et al. 2004; Martin 

et al. 2006; Brodhead et al. 2007; Hall et al. 

2007), and five found decreased abundance 

(Popotnik and Giuliano 2000; Tewksbury 

et al. 2002; Krueper et al. 2003; Earnst et 

al. 2005; Fletcher and Hutto 2008). Four 

studies reported a decrease in species diversity 

or richness of riparian avifauna (Popotnik 

and Giulano 2000; Stanley and Knopf 2002; 

Scott et al. 2003; Hall et al. 2007), and four 

reported static values (Kauffman et al. 1982; 

Schulz and Leininger 1991; Warkentin and 

Reed 1999; Earnst et al. 2005). Bock et 

al. (1993) reviewed abundance data for 63 

neotropical migrant bird species in grazed 

and ungrazed environments. Of these species, 

three declined in abundance in grazed areas, 

and seven additional species were thought to 

be negatively influenced by grazing. These 

species were either shrub, ground, or nearground nesters. 

230 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

Overall, where grazing induced changes in 

habitat structure and composition, avian 

populations tended to change from dominance 

by riparian obligate species to dominance 

by riparian generalists (e.g., Schulz and 

Leininger 1991; Martin and McIntyre 2007). 

Changes in avian abundance were often 

positively associated with habitat quality; 

however, assigning a habitat quality measure 

is somewhat subjective when dealing with 

avian species assemblages; some species may 

benefit from altered habitat, and some may 

be negatively impacted, depending on specific 

habitat requirements (Farley et al. 1994). The 

work of Martin and McIntyre (2007) suggests 

that species diversity may be maximized with 

heterogeneous grazing intensities over space. 

Tewksbury et al. (2002) suggested that avian 

species nesting below 2.5 m would be most 

negatively impacted by livestock grazing. 

We conclude that grazing can decrease 

populations of riparian obligate avifauna 

but has variable effects on generalist 

species. Diversity of species may decrease 

in proportion to grazing-induced decreases 

in habitat diversity (Scott et al. 2003). One 

caveat to this conclusion is that determining 

the specific influence of grazing on riparian 

avian assemblages is challenging and must 

take into account uses and changes in use 

within the surrounding landscape. Avian 

species are highly mobile, and some “riparian” 

species may depend on spatially distant 

habitat types and landscape attributes. In 

an extreme example, assessing the influence 

of management practices on abundance of 

riparian neotropical migrant avifauna should 

involve determination of vital rates (e.g., 

nesting success and juvenile survival) to 

help factor out the proportion of population 

change associated with nonbreeding habitat. 

Management of local-scale riparian issues 

(such as grazing) should be undertaken in 

conjunction with larger-scale efforts to create 

landscapes suitable for attaining conservation 

objectives for riparian avifauna (Martin et al. 

2006; Fletcher and Hutto 2008). 

Hypothesis 13: Riparian Grazing Decreases 

Populations of Macroinvertebrates, 

Herpetofauna, and Salmonids. Limited 

data suggest that grazing does not decrease 

the abundance or overall diversity of 

macroinvertebrates. However, some habitat 

specialists may decrease and be replaced with 

Grazing should be managed 

to allow a site to meet its 

potential to provide in-stream 

aquatic habitat features such as 

over-hanging banks and clean 

gravel beds. (Photo: Ken Tate) 

CHAPTER 5: Effectiveness of Riparian Management Practices 231 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

habitat generalists (Weigel et al. 2000; Bates 

et al. 2007). Data are insufficient to make 

general conclusions regarding the influence 

grazing on herpetofauna populations. Diversity 

of macroinvertebrates remained unchanged 

or increased with grazing in six of nine 

studies (Fritz et al. 1999; Weigel et al. 2000; 

Homyack and Giuliano 2002; Scrimgeour and 

Kendall 2003; Sada et al. 2005; Bates et al. 

2007), decreased in one study (Foote and Rice 

Hornung 2005), and was not reported in two 

studies (Tait et al. 1994; Saunders and Fausch 

2007). Macroinvertebrate abundance remained 

unchanged with grazing in seven studies (Tait 

et al. 1994; Fritz et al. 1999; Homyack and 

Giuliano 2002; Scrimgeour and Kendall 2003; 

Sada et al. 2005; Bates et al. 2007; Saunders 

and Fausch 2007), decreased in one study 

(Foote and Rice Hornung 2005), and was not 

reported in one study (Weigel et al. 2000). Two 

studies reported no effect of grazing on riparian 

herpetofauna (Bull and Hayes 2000; Homyack 

and Giuliano 2002), but a review by Brodie 

(2001) suggests that turtle populations may 

be negatively impacted by increased siltation 

associated with disturbances, such as livestock 

grazing. 

Limited data suggest that livestock grazing can 

decrease salmonid populations, and the bulk 

of papers we examined suggested decreasing 

quality of habitat with livestock use. The 

specific grazing management scenarios under 

which salmonid populations may be negatively 

impacted by grazing are largely unknown given 

that most of the salmonid studies we reviewed 

did not report stocking rate or utilization 

information. Impacts of livestock grazing on 

salmonid habitat and populations have been 

summarized (Meehan and Platts 1978; Platts 

1981, 1991; Armour et al. 1994; Fleischner 

1994; Belsky et al. 1999). Three of six studies 

reported decreased salmonid abundance 

associated with livestock grazing (Keller and 

Burnham 1982; Tait et al. 1994; Knapp and 

Matthews 1996), one study reported no impact 

(Chapman and Knudsen 1980), and one did 

not report abundance as a function of grazing 

treatment (Platts and Nelson 1989). One study 

indicated that salmonid abundance was higher 

for areas grazed with a high-density, shortduration grazing system compared to seasonlong grazing (Saunders and Fausch 2007). 

Three of six studies reported decreased quality 

of salmonid habitat with grazing (Chapman 

and Knudsen 1980; Platts and Nelson 1989; 

Knapp and Matthews 1996), one reported no 

effect (Tait et al. 1994), one did not report 

habitat effects (Keller and Burnham 1982), 

and one reported increased habitat quality with 

short-duration grazing compared to seasonlong grazing (Saunders and Fausch 2007). 

We recognize that additional published 

references are available correlating fish 

abundance with grazing practices. However, 

much of this work has not undergone the 

scrutiny of peer review, suffers from major 

experimental design inadequacies (e.g., lack of 

replication, nonrandom treatment assignment, 

lack of pretreatment data), or has insufficient 

methodological description to determine 

the adequacy of experimental design (Platts 

1982; Rinne 1985; Larsen et al. 1998). These 

problems render affected references useless for 

our purposes in determining the validity of 

hypotheses regarding management practices. 

That said, it should also be pointed out that 

ill-advised grazing practices can lead to loss 

of bank-stabilizing vegetation, resulting in 

altered channel morphology (see discussion 

for hypothesis 8) and that such alterations 

may have strong negative consequences for 

habitat of affected aquatic fauna (Fitch and 

Adams 1998). 

Hypothesis 14: Riparian Grazing Decreases 

Habitat Quality for Riparian Mammals. 

Data are insufficient to determine the impact 

of grazing on large mammal riparian wildlife 

species with two studies reporting either 

decreased quality of fawning habitat (Loft et al. 

1987) or livestock-induced habitat avoidance 

(Loft et al. 1991). Three studies addressed the 

influence of grazing on riparian small mammal 

communities. Two studies found no change 

in diversity of species (Kauffman et al. 1982; 

Schulz and Leininger 1991), and one reported 

decreased diversity (Giuliano and Homyak 

2004). Two of three studies reported decreased 

small mammal abundance with grazing 

(Kauffman et al. 1982; Giuliano and Homyak 

2004), and one study was inconclusive (Schulz 

and Leininger 1991). A fourth study reported 

decreased small mammal biomass with heavy 

compared to light grazing, but responses varied 

by species, and ungrazed comparisons were not 

included (Johnston and Anthony 2008). 

232 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

Available data are insufficient to draw general 

conclusions regarding the impacts of livestock 

on riparian small mammal communities. 

Realistically, a general conclusion for small 

mammals as a group may not be possible 

because of the inherent variability in habitat 

requirements between species and variability 

among years. Species that depend on 

herbaceous cover may decrease with heavy 

livestock use, while the same disturbance may 

increase habitat quality for species requiring 

reduced amounts of herbaceous ground 

cover (Hanley and Page 1982; Johnston and 

Anthony 2008). 

Hypothesis 15: Grazing Removal 

Will Increase Quality of Sage Grouse 

Brood-Rearing Habitat. Because of a 

lack of experimental work on the subject, 

generalizations regarding the influence of 

grazing on sage grouse brood rearing habitat 

cannot be made at this time. Only one 

study, Neel (1980; see also Appendix VI) has 

addressed the influence of grazing on riparian 

brood-rearing habitat for sage grouse. This 

study found that 1 yr of rest from grazing 

increased abundance of forbs important in the 

diet of sage grouse. The author also reported 

that sage grouse selected lightly grazed riparian 

habitat for brood rearing as compared to 

nongrazed habitat. 

Hypothesis 16: Invasive Woody Plant 

Management Can Control Abundance 

of Undesirable Plant Species. Literature 

relating to invasive species management and 

riparian woody plants deals mainly with 

populations of the invasive plant genera 

Tamarix and in some cases Russian olive 

(Elaeagnus angustifolia) in the southwestern 

United States. In three studies, various 

combinations of cutting, plowing, and 

burning were highly effective at removing 

both Tamarix and Russian olive (McDaniel 

and Taylor 2003; Harms and Hieber 2006; 

Bay and Sher 2008; Appendix VII). A fourth 

study reported that while Tamarix seedling 

density was initially higher than that of native 

woody plants, Tamarix seedlings were much 

more susceptible to mortality associated 

with overbank flooding in unregulated river 

systems (Sher et al. 2002). Much of the effort 

to control Tamarix remains unevaluated and 

unpublished. Bay and Sher (2008) reviewed 

control projects ranging from 1 to 18 yr 

posttreatment. They reported that the degree 

of control was not related to time since 

restoration began or specific management 

treatments and that areas with less than 21 cm 

of annual precipitation had only limited longterm Tamarix control. Site factors played a 

strong role in influencing project success, and 

the degree of control was associated positively 

with proximity to perennial water, sufficient 

precipitation, recent flooding, and coarse 

soil texture. Shafroth et al. (2008) noted 

that success of Tamarix control projects was 

highly variable and proposed a framework for 

planning control efforts that focuses on using 

principles of adaptive management. These 

authors stressed that site conditions, including 

soil salinity and texture, current vegetation, 

and availability of desired propagules, have 

a strong influence on restoration success and 

highlighted the importance of considering 

both passive (e.g., flooding) and active (e.g., 

cutting and seeding) management options. 

Because of the spatial and temporal variability 

associated with the success of Tamarix 

control projects, it is not possible to make 

general statements regarding the effectiveness 

of control programs. Future success in 

Tamarix management will likely hinge on 

effective application of adaptive management 

techniques (Reever Morghan et al. 2006). 

Hypothesis 17: Control of Invasive Woody 

Plant Species Increases the Abundance of 

Terrestrial Wildlife. Most of the literature 

regarding invasive riparian woody plant control 

and wildlife abundance relates to the control of 

Tamarix. Four papers relating invasive woody 

plants to wildlife assemblages failed to uncover 

substantive benefit to abundance or diversity of 

avian, butterfly, or lizard assemblages (Knopf 

and Olson 1984; Bateman et al. 2008; Nelson 

and Wydoski 2008; Sogge et al. 2008). In a 

2008 review, Sogge et al. found that not all 

avian species benefit from control of Tamarix, 

particularly when native vegetation does not 

reestablish in the postrestoration environment. 

These authors concluded that 49 avian species, 

including the endangered southwestern willow 

flycatcher, use Tamarix as breeding habitat. 

Van Ripper et al. (2008) reported that for most 

avian species, abundance was highest with 

a mix of native woody plants and Tamarix. 

CHAPTER 5: Effectiveness of Riparian Management Practices 233 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	 	

Seasonal riparian corridor 

in an oak-woodland in 

California’s central Sierra 

Nevada foothills. (Photo: 

Melvin George) 

Knopf and Olson (1984) reported increased 

avian diversity in native riparian communities 

compared to sites dominated by Russian olive 

but noted that Russian olive was used by avian 

species favoring tall shrub habitat and that the 

occurrence of Russian olive near the periphery 

of riparian areas could increase diversity of 

riparian habitats available to avian species. 

At this time, the complexity of wildlife 

responses to Tamarix and Russian olive 

control varies strongly across species and 

geographic location, making generalizations 

regarding the impact of these invasive species 

on terrestrial wildlife difficult (Shafroth et al. 

2005; Sogge et al. 2008). 

Hypothesis 18: Upland Brush Management 

Can Decrease Riparian Erosion and 

Increase Stream Flow. Nine studies addressed 

the influence of woody plant removal on 

watershed hydrology (Appendix VII). Three 

studies reported increased water yield (actual 

or modeled) or stream flow in pinyon-juniper 

(Baker 1984), chaparral (Davis 1993), or 

sagebrush (Sturges 1994) vegetation, and two 

reported no change in stream flow (Wilcox et 

al. 2005) or runoff (Dugas et al. 1998) for Ashe 

juniper (Juniperus ashei). One study found no 

change in basin-level water yield with removal 

of western juniper (Juniperus occidentalis; Kuhn 

et al. 2007). One study reported increased 

potential for deep drainage with burning in 

sagebrush-steppe (Seyfried and Wilcox 2006), 

a second found dramatically decreased runoff 

and erosion in sagebrush-steppe following 

juniper removal and (Pierson et al. 2007), and 

a third found increased runoff with chaining 

and windrowing of pinyon-juniper, but runoff 

was invariant when trees were left in place 

(Gifford 1975). Wilcox (2002) proposed that 

the influence of woody plants on stream flow 

will be a product of interactions between shrub 

characteristics, precipitation, soils, and geology. 

Under this conceptual framework, woody plant 

removal generally will not affect stream flow 

in areas receiving less than 500 mm of annual 

precipitation, and runoff will occur as overland 

flow in the absence of a subsurface connection 

between stream and hillslope. Without 

subsurface flow, water use by woody plants may 

have little impact on stream flow. Huxman et 

al. (2005) echoed the importance of subsurface 

flow for linking woody plants and stream flow. 

Pierson et al. (2007) stressed that increases 

in herbaceous production associated with 

woody plant control can dramatically increase 

infiltration and decrease runoff. 

Given that site characteristics strongly 

influence the relationship between woody 

plant cover and hydrology, definitive 

statements regarding outcomes of this 

interaction are not possible. Newman et al. 

(2006) argued that because of the variability 

and complexity of the relationship between 

woody plants and rangeland hydrology, 

efforts to manage woody plant issues will 

benefit from “place-based science” and an 

interdisciplinary focus on hypothesis testing. 

Hypothesis 19: Shading of the Stream 

Channel by Riparian Woody Vegetation 

Cover Influences Aquatic Ecology 

by Reducing Stream Temperature. 

Macroinvertebrates and fish are sensitive 

to dissolved oxygen content of streams, 

which is influenced by stream temperature. 

Thus, stream temperature is an important 

factor affecting the distribution of aquatic 

vertebrate and invertebrate species (Baltz et 

al. 1987; Lyons 1996; Hawkins et al. 1997; 

Jacobsen et al. 1997; Isaak and Hubert 2001). 

The distribution and abundance of native 

coldwater fisheries in the western United 

States has been reduced since European 

settlement (Nehlsen et al. 1991; Hunnington 

et al. 1996; Thurow et al. 1997), and land 

and water management practices that impact 

stream temperature are considered to be 

partly responsible for these reductions (Isaak 

and Hubert 2001; Poole and Berman 2001; 

Zoellick 2004). Water temperature is a 

particularly important habitat determinant for 

aquatic species in arid rangeland basins of the 

western United States. 

A significant international research base 

indicates that water temperature is a spatially 

and temporally dynamic stream property 

controlled by a complex and interacting set 

of environmental factors, such as local air 

mass characteristics, solar radiation, vegetative 

and topographic shading, channel elevation 

and aspect, channel gradient, adiabatic 

rate, channel width and depth, hydrologic 

residence time, stream flow volume, and 

deep and shallow groundwater inputs (e.g., 

234 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

Constantz 1998; Ebersole et al. 2001; Liquori 

and Jackson 2001; Poole and Berman 2001; 

Ebersole et al. 2003; Johnson 2004; Malcolm 

et al. 2004; Moore et al. 2005; Tate et al. 

2007). Management practices which affect 

these factors have the potential to secondarily 

affect stream temperature dynamics. There 

is clear evidence that shading provided by 

woody plant cover will have some effect on 

stream temperature dynamics. 

There is also clear evidence that the relative 

importance of woody plant canopy cover, 

among the many factors, in determining 

stream temperature is variable across 

ecosystems, watersheds, streams, and 

stream reaches. Larson and Larson (1996) 

hypothesized that when air temperature 

is warmer than water temperature, water 

temperature will increase to approach 

thermal equilibrium with the surrounding 

air mass and that this basic relationship 

is unchanged by the presence of shade 

from woody plants. However, in a field 

study in arid northeastern California, Tate 

et al. (2005) found that daily maximum 

stream temperature was associated with 

air temperature, instantaneous stream 

flow volume, stream order and watershed 

position, and woody plant canopy cover. 

Increased woody plant cover was associated 

with decreased maximum daily stream 

temperature, and a significant interaction 

between canopy and air temperature 

indicated that the cooling effect of woody 

plant cover increased with increased 

maximum daily air temperature. One study 

(Meays et al. 2005) reported that a thermal 

gradient associated with variable elevation 

was the dominant factor controlling stream 

temperature and that exposure time (velocity 

and distance), discharge volume, rate of 

flow, and cool-water inputs had a greater 

influence on stream temperature than woody 

canopy cover. Poole and Berman (2001) 

found that the influence of shade on stream 

temperature was greatest in smaller (first 

and second order) streams and decreased 

with stream size. These authors hypothesized 

that reduced stream shading may lower 

the quantity of air trapped by vegetation, 

which can increase convective and advective 

transfer of heat to the stream surface. Both 

Liquori and Jackson (2001) and Malcolm 

et al. (2004) determined that the type of 

riparian woody plant community affected 

the relationship between canopy cover 

and stream temperature. Stream channel 

protection from incoming radiation is one 

mechanism by which woody plant cover may 

influence stream temperature. In certain 

riparian areas, woody plants may play a 

role in maintaining channel structure (e.g., 

width:depth ratio) in the face of destabilizing 

flood flow events (Winward 2000). To the 

extent that maintenance of channel structure 

is related to stream temperature dynamics, 

woody plants may play an important role 

in moderating in-stream temperature 

fluctuations (Liquori and Jackson 2001). 

There is strong evidence to support the 

assertion that riparian management to 

enhance and sustain riparian woody plants 

can moderate stream temperatures. The 

validity of this conclusion is conditional and 

dependent on 1) site conditions and potential 

and that management of factors such as 

stream flow volume may have a greater effect 

on stream temperature than management 

of woody plant cover; 2) the natural, or 

existing, potential of the riparian site to 

support woody plant communities; and 3) 

the likelihood that management to increase 

woody canopy above natural site potential 

can lead to overall reductions of in-stream 

primary production, diversity, and richness 

of aquatic species (Liquori and Jackson 2001; 

Broadmeadow and Nisbet 2004; Malcolm et 

al. 2004). 

Hypothesis 20: Prescribed Fire Can Increase 

Richness, Diversity, and Abundance of 

Native Riparian Plant and Animal Species. 

Because of the low number of published reports 

concerning prescribed fire in riparian habitat, 

we included wildfire-based publications in this 

discussion. Of seven papers documenting the 

effects of fire on native riparian vegetation, 

four reported little to no effect (Busch and 

Smith 1993; Gom and Rood 1999; Blank et al. 

2003; Smith et al. 2007), and three reported 

an increase in desired species (Stein et al. 1992; 

Kay 1993; Rood et al. 2007; Appendix VII). 

In some cases, fire has been used as a tool to 

rejuvenate dense stands of mature woody plants, 

such as cottonwood, when reproduction became 

limited (Rood et al. 2007). 

There is strong 

evidence to 

support the 

assertion 

that riparian 

management 

to enhance 

and sustain 

riparian woody 

plant cover or 

canopies can 

moderate stream 

temperatures”

CHAPTER 5: Effectiveness of Riparian Management Practices 235 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	

Management which supports 

woody riparian plants and 

deep-rooted herbaceous 

vegetation may increase 

carbon sequestration in riparian 

soils. (Photo: Mel George) 

Existing literature is not sufficient to 

generalize the effects of fire on riparian plant 

species richness, diversity, and abundance. The 

impact of fire on diversity of riparian animal 

species in native riparian habitat is practically 

unexplored. Literature documenting the 

response of riparian birds to fire is lacking 

(Bock and Block 2005; Smith et al. 2007). 

Hypothesis 21: Carbon Storage Is 

Enhanced by Establishment and 

Maintenance of Woody Species, 

Herbaceous Species with High Root Mass, 

and Dominance of Deep-Rooted Perennials. 

Carbon accumulation in soils occurs when 

C inputs to the ecosystem as NPP exceed 

C outputs from the ecosystem as microbial 

respiration of soil organic matter (Post and 

Kwon 2000). Hence, any management that 

increases production and/or decreases microbial 

respiration on an annual basis should promote 

soil C storage. Root detritus is a significant 

contributor to recalcitrant soil C pools (Rees et 

al. 2005), so promotion of belowground NPP 

is believed to be particularly important for C 

sequestration. 

Nine to 18 yr of grazing exclusion from 

herb-dominated wet and dry meadows in 

eastern Oregon resulted in clear increases in 

belowground standing biomass (Kauffman et 

al. 2004). While these authors did not detect 

significant increases in soil organic matter, 

bulk densities decreased significantly, and one 

can infer that soil organic C would increase 

over a longer period of observation. One 

year of late-season clipping of mesic meadow 

species in central Nevada resulted in higher 

rooting activity in the surface 5 cm (Martin 

and Chambers 2002). These authors cite the 

overriding influence of water table depth as 

the reason that larger defoliation effects were 

236 Conservation Benefits of Rangeland Practices Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

not observed. No significant effects on soil 

physical or chemical properties were found 

by Wheeler et al. (2002) as the result of a 

one-time intense grazing event. Clary and 

Kinney (2002) found that simulated seasonlong, heavy grazing significantly reduced 

root production, while simulated moderate 

seasonal grazing had no effect relative to 

unclipped control plots. 

We cannot support or reject this hypothesis 

because few studies have investigated carbon 

storage in riparian zones. Two studies (grazing 

exclusion and late-season clipping) support 

increases in belowground biomass, one study 

found no grazing effect on soil chemical and 

physical properties, and one study found 

that season-long heavy grazing reduced root 

production in riparian systems. 

Incorporating Science into nRcS 

conservation Practices and Systems 

The evidence supporting conservation 

practice effectiveness is mixed with some 

practices being well documented and 

others poorly supported in peer-reviewed 

scientific publications. However, this is not 

the only source of evidence for practice 

effectiveness. Professional experience is also 

an important source of knowledge regarding 

practice effectiveness. Within NRCS and 

other agencies are conservationists who 

have learned to apply practices effectively by 

learning from others and by trial and error, 

much the way agricultural producers learn to 

adapt practices in their farming operations. 

Management based on trial and error is 

often called adaptive management. Adaptive 

management allows managers to monitor and 

evaluate management practices in the field 

as they go along. The nine steps of planning 

used by NRCS make up a form of adaptive 

management that allows conservationists and 

landowners to identify resource concerns 

and alternative practices. Following selection 

and implementation of practices, monitoring 

and evaluation provide feedback regarding 

progress toward objectives and practice 

effectiveness. The knowledge gained during 

planning, implementation, and evaluation is 

seldom reported in peer-reviewed journals. 

Occasionally, it appears in case study reports, 

but more often it goes unpublished (e.g., 

Wyman et al. 2006). 

Early in the planning process, NRCS 

conservationists document and analyze 

resource concerns, including those related to 

riparian systems and associated watersheds. 

This is followed by development of alternative 

practices that may address concerns. Based on 

this analysis, the landowner selects a mix of 

practices. For riparian areas, prescribed grazing 

(528), off-site water (614), fencing (382), and 

riparian herbaceous cover (390) are common 

conservation practices that are often applied 

together because they facilitate control of 

riparian use while enabling use of the broader 

landscape by livestock and wildlife. 

While the effectiveness of these practices may 

not have been documented in the ecosystem 

or site being managed, there is often support 

for their effectiveness from other riparian 

ecosystems in the scientific literature. NRCS 

training programs expedite integration of 

results from other ecosystems into the planning 

process. Conservationists and landowners learn 

what works from these applications, and it 

becomes part of the individual’s experience and 

the agency’s institutional memory in the form 

of state practice standards and specifications. 

REcoMMEndAtIonS 

With more than 40 management practices 

(USDA NRCS 2003) available for application 

to riparian habitats, considerable overlap exists 

among the purposes and benefits stated in 

practice standards. Better riparian practice 

standards could be developed by the following: 

•	� Initiating review teams of NRCS 

conservationists, biologists, and engineers 

to complete practice revisions 

•	� Grouping practices that protect or 

restore vegetation to remove overlapping 

purposes (e.g., channel bank vegetation, 

conservation cover, critical area planting, 

riparian herbaceous cover, stream bank and 

shoreline protection, and tree and shrub 

establishment) 

•	� Grouping and revising buffer and filter 

strip practices into those that apply to 

rangeland, forestland, or cropland 

•	� Updating practice definitions, purposes 

and benefits, criteria and other practice 

standard sections to reflect current 

knowledge 

CHAPTER 5: Effectiveness of Riparian Management Practices 237 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	

	 	 	

	 	

	 	

	 	

	 	

Rigorous monitoring to document 

the effect of resource management systems can 

help resource professionals learn 

more about riparian processes 

and management 

interactions”

•	� Incorporating ecosystem services into 

revised practice purposes and benefits 

•	� Separating structural practices from 

vegetation management practices 

•	� Incorporating riparian purposes and 

benefits into upland practices, such as 

brush management, prescribed burning, 

and prescribed grazing 

While there are opportunities to combine 

and clarify practices and there is evidence 

supporting the effectiveness of many riparian 

management practices, we can provide little 

evidence-based support to USDA NRCS for 

modifications of existing practice specifications 

(practice application) or initiation of alternative 

practices, with one exception in the following 

paragraph. We also recommend addition 

of a collaborative research and monitoring 

component to selected practice implementation 

plans so that the body of evidence supporting 

conservation practices and systems of practices 

can be strengthened. 

We found sufficient evidence to recommend 

that NRCS increase the role of herding and 

supplement placement along with water 

development and fences for manipulating 

livestock distribution. These practices have 

a role where topography does not limit 

their effectiveness and total exclusion is 

not required. While it has not been the 

policy of USDA conservation cost-share 

programs to fund feed purchase or herding, 

placement of supplement and herding 

practices should be included in the overall 

ranch conservation plan. The USDA might 

consider allowing these “feeding” practices to 

be part of the rancher’s share of the cost in the 

Environmental Quality Incentives Program 

and other cost-share programs. 

The need for more effective selection and 

application of management practices on 

a site-specific basis requires much greater 

attention. Recognizing that one practice or 

set of practices cannot meet the conservation 

requirements of biophysically diverse riparian 

habitats and stream systems, USDA NRCS 

applies resource management systems that 

are a flexible mix of practices selected for a 

specific set of site conditions and landowner 

management objectives during conservation 

planning. Rigorous monitoring to document 

the effect of resource management systems 

can help resource professionals learn more 

about riparian processes and management 

interactions while maintaining feedback 

information to both land managers and 

conservation planners. The portfolio of 

research and case studies supporting the 

effectiveness of these practices is limited and 

commonly cannot be extended to other sites. 

Time and funds limit the ability of research 

institutions to investigate the seemingly 

infinite combination of site conditions 

that exist across US rangeland riparian 

zones. To accelerate these investigations, we 

recommend that a partnership of researchers 

and NRCS conservationists implement two 

complementary lines of investigation. In the 

first line of investigation, this team should 1) 

develop, implement, and maintain rigorous 

monitoring of selected practices in selected 

ecosystems and 2) implement monitoring 

systems that can be analyzed and meet 

standards of research peer review. The team 

would manage monitoring data collection and 

analyze the data at appropriate time intervals. 

In short, a research study design (monitoring 

plan) needs to accompany and be funded 

along with the conservation plan. 

The second line of investigation is to develop 

testable hypotheses based on observations 

and findings resulting from team monitoring 

projects. The team can test these hypotheses 

in more controlled studies. These hypotheses 

should attempt to elucidate the intervening 

ecological processes between practice 

implementation and practice effect. Only 

then can we begin to understand relationships 

between grazing, riparian management 

practices, and riparian ecology at relevant scales 

and extrapolate results from one location to 

another. It is crucial that USDA NRCS and 

other agencies support such a monitoring 

partnership between researchers and 

conservationists. 

Finally, we support the completion of riparian 

ecological site descriptions by USDA NRCS. It 

is important to recognize that not all riparian 

areas have the same potential or react to 

management in the same way. Therefore, they 

should be managed according to their unique 

characteristics as described in ecological site 

descriptions. State variables (soils, climate, 

238 Conservation Benefits of Rangeland Practices 
	 	 	 	 		

	 	 	 	 	 	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	 	

	 	 	

	 	 	

	 	

M. R. George, R. D. Jackson, C. S. Boyd, and K. W. Tate 

geomorphology, topography, vegetation, and 

wildlife), vegetation dynamics, and practices 

that have been effective on the site in the past 

are described in the ecological site description. 

We recommend completion of riparian 

ecological site descriptions as a means of 

documenting and communicating proven sitespecific management practices to the NRCS 

planning process. 

KnoWLEdgE gAPS 

Our assessment reveals limited controlled 

experimentation in support of many of our 

hypotheses, resulting in critical knowledge 

gaps across all riparian management practices 

and riparian ecosystems. Linking conservation 

planning and management to research in a 

collaborative program is crucial to filling these 

knowledge gaps. 

Two substantial knowledge gaps exist in the 

riparian literature related to grazing and 

rangeland management. While there are many 

case studies comparing species (animal and 

plant) abundance within and outside riparian 

exclosures, they are often deficient in more 

rigorously designed experiments. For example, 

much of the case study literature concerning 

impacts of livestock on riparian wildlife 

suffers from experimental design inadequacies, 

including lack of pretreatment data, low 

sample size, and lack of randomization of 

treatments (Rinne 1985; Larsen et al. 1996). 

Additional research based on replicated 

experimental designs is needed to better 

understand the relationship between grazing 

and riparian ecology at scales relevant to 

determining the ecological consequences of 

grazing practices. 

The second knowledge gap emphasizes ecosystem services will involve an expanded, 

ecological processes that mediate the effect of research-based focus on the interaction 

management actions on riparian ecosystem between management activities and biophysical 

products and services. Without mechanistic mechanisms responsible for provisioning 

understanding of the intervening ecological ecosystem services. 

processes that mediate cause-and-effect 

relationships, we cannot generalize study concLuSIonS 

results to other sites. This point is particularly 

important given the dependence of riparian While the scientific evidence for many 

plant species on groundwater resources that riparian management practices is inconclusive, 

vary over both space and time (Stringham et there are several practice benefits that are 

al. 2001; Poole et al. 2006). Improving our well documented. There is strong evidence 

knowledge of the effects of management on supporting the influence of management 

Several riparian practices 

have been shown to enhance 

riparian functions and 

dependent ecosystem services. 

(Photo: Ken Tate) 

CHAPTER 5: Effectiveness of Riparian Management Practices 239 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

practices on vegetation in riparian habitats, 

including the following: 

1.	� Grazing intensity influences herbaceous 

species composition and productivity 

(H1, 2). 

2.	� Livestock distribution practices, such 

as water developments, supplement 

placement, and herding, are effective 

means of reducing livestock residence time 

and utilization in the riparian zone (H3). 

3.	� Grazing exclusion can promote recovery of 

riparian plant community composition in 

degraded riparian systems (H4). 

4.	� Livestock and other large herbivores can 

modify the structure and composition of 

woody plant communities (H5). 

5.	� Late-season (usually late summer and fall) 

livestock use of riparian areas can lead 

to increased utilization of woody plants, 

especially when herbaceous plants are 

limited in availability or forage quality 

(H6). 

There is also evidence supporting the influence 

of riparian management practices on riparian 

vegetation and soils (Fig. 1). Riparian grazing 

management that maintains or enhances key 

riparian vegetation attributes (i.e., species 

composition, root mass and root density, cover, 

and biomass) will enhance stream channel 

and riparian soil stability, and this in turn will 

support ecosystem services, such as flood and 

pollutant attenuation and quality of riparian 

habitats (H8). 

Finally, limited evidence indicates that riparian 

habitat management can promote ecosystem 

services by enhancing vegetation and soil 

attributes (Fig. 1): 

1.	� Riparian vegetation can function to 

attenuate waterborne pollutants in 

overland and flood flow events (H9). 

2.	� The design and implementation of 

optimally efficient riparian buffers must 

incorporate site-specific biophysical 

factors, including buffer width, vegetation 

attributes and management, pollutant 

type, pollutant load and concentration, 

flow rate, hydrologic residence time, and 

soil attributes (H9). 

3.	� Practices that reduce livestock densities, 

residence time, and fecal and urine 

deposition in riparian areas and stream 

flow generation areas can reduce nutrient 

and pathogen loading of surface water 

(H10). 

4.	� Grazing in wetland habitat does not 

decrease habitat quality for waterfowl 

except in instances of heavy grazing (H11). 

5.	� Grazing can decrease populations of 

riparian obligate avifauna but may increase 

or have no effect on generalist species (12). 

6.	� Shading provided by woody plant cover 

along with other factors (e.g., elevation, 

topography, and subsurface flow) will 

have some effect on stream temperature 

dynamics (19). 

7.	� Fire can be used to control some species 

of woody plants, but success of fire-based 

restoration may be related to availability of 

water and the availability of propagules of 

desired species following years of limited 

water availability (H7). 

8.	� Grazing does not appear to decrease 

the abundance or overall diversity of 

macroinvertebrates, but these data are 

limited (H13). 

9.	� Limited data suggest that livestock grazing 

practices that are too long in duration and 

poorly timed can decrease salmonid habitat 

quality or populations (13). 

For several hypotheses, the evidence supporting 

or refuting beneficial effects on ecosystem 

services was weak or inconclusive. These include 

1) riparian grazing decreases habitat quality 

for riparian mammals (H14) and sage grouse 

(H15); 2) woody plant control can reduce 

undesirable plant species (H16) or increase 

the abundance of terrestrial wildlife (H17); 

3) the influence of riparian burning (H7) on 

vegetation and animals; 4) prescribed fire can 

increase the richness, diversity, and abundance 

of native riparian plants and animals (H 20); 5) 

upland brush management can decrease erosion 

and increase stream flow (H 18); and 6) carbon 

storage can be enhanced by the establishment 

and maintenance of woody species, herbaceous 

species with high root mass, and dominance of 

deep-rooted perennials. 

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A cross-sectional survey of California’s grazed 

rangeland riparian areas. Proceedings of the 

Riparian Habitat and Floodplain Conference; 

12–14 March 2001; Sacramento, CA, USA. 

Sacramento, CA, USA: California Riparian 

Habitat Joint Venture. p. 479–583. 

Warkentin, I. G., and J. M. Reed. 1999. Effects 

of habitat type and degradation on avian species 

richness in Great Basin riparian habitats. The 

Great Basin Naturalist 59:205–212. 

Warner, R. E., and K. M. Hendrix [eds.]. 1984. 

California riparian systems, ecology, conservation, 

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University of California Press. 997 p. 

Weigel, B. M., J. Lyons, L. K. Paine, S. I. 

Dodson, and D. J. Undersander. 2000. 

Using stream macroinvertebrates to compare 

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southwestern Wisconsin. Journal of Freshwater 

Ecology 15:93–106. 

West, B. C., and T. A. Messmer. 2006. Effects 

of livestock grazing on duck nesting habitat 

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Wheeler, M. A., M. J. Trlica, G. W. Frasier, 

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Rocky Mountain Research Station. RMRSGTR-47. 49 p. 

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CHAPTER 

6 

An Assessment of Rangeland 

Activities on Wildlife 

Populations and Habitats 

Paul R. Krausman,1 Vernon C. Bleich,2 William M. Block,3 

David E. Naugle,4 and Mark C. Wallace5 

Authors are 1Boone and Crockett Professor of Wildlife Conservation, University of Montana, Missoula,
 

MT 59812, USA; 2Affiliate Professor, Idaho State University, Pocatello, ID 83209, USA;
 

3Wildlife and Terrestrial Ecosystems Program Leader, US Forest Service, Rocky Mountain Research
 

Station, 2500 S. Pine Knoll Dr, Flagstaff, AZ 86001, USA; 4Professor, University of Montana,
 

Missoula, MT 59812, USA; and 5Associate Professor, Department Natural Resources Management,
 

Texas Tech University, Lubbock, TX 79409-2125, USA.
 

Correspondence: Paul R. Krausman, Boone and Crockett Program in Wildlife Conservation, Wildlife 

Biology, University of Montana, Missoula, MT 59812, USA. Email: paul.krausman@umontana.edu 

Research was funded in part by the USDA Natural Resources Conservation Service, 

the Boone and Crockett Program in Wildlife and the Wildlife Biology Program, 

the University of Montana, the USDA Forest Service, and Texas Tech University. 

Reference to any commercial product or service is made with the understanding 

that no discrimination is intended and no endorsement by USDA is implied 

253
 
	 	 	 	

	 	 	

	 	

	 	 	 	

	 	 	

	 	 	 	

	 	 	

	 	 	 	

	 	 	 	 	

	 	 	

	 	 	 	

	 	 	

Wildlife responses to 

conservation practices 

“

are usually species and 

even species-habitat 

specific, meaning not 

only that each species 

may respond differently 

to any specific practice 

but also that a single 

species may respond 

differently to the same 

practice in different 

conditions.”

254 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	 	 	 	

	 	 	

	 	 	 	 	

	 	 	

	 	 	 	

	 	 	 	

6 An Assessment of Rangeland 

Activities on Wildlife Populations 

and Habitats 

Paul R. Krausman, Vernon C. Bleich,William M. Block, David E. Naugle, 

and Mark C. Wallace 

IntRoductIon 

Numerous management practices are applied 

to rangelands in the western United States 

to enhance wildlife, including prescribed 

grazing, burning, brush management, 

mowing, fencing, land clearing, planting, 

and restoration to benefit soil and water. 

Indeed, the Natural Resources Conservation 

Service (NRCS) lists 167 conservation 

practices (http://www.NRCS.USDA.gov/ 

technical/standards/nhcp.html). However, 

wildlife responses to conservation practices 

are usually species and even species-habitat 

specific, meaning not only that each species 

may respond differently to any specific 

practice but also that a single species may 

respond differently to the same practice in 

different vegetation associations or conditions. 

When managers apply conservation practices 

to the landscape, habitat is often altered, 

and managers should understand that the 

management will benefit some of the wildlife 

present but may be detrimental to others. 

Conservation practices were designed to help 

ecosystem managers think about the variables 

that accompany any action on the landscape. 

Each conservation practice has specific 

purposes that may influence related resource 

issues. For example, prescribed grazing by 

large herbivores can alter the structure and 

function of ecosystems that have direct and 

indirect effects on wildlife. Primary effects 

are often described in the literature (Mackie 

2000), but there has not been an evaluation 

of how conservation practices affect wildlife 

on rangelands. However, practices like 

prescribed grazing are not a simple treatment 

but have widely divergent effects, depending 

on locale, timing, intensity, and species or 

combination of grazing animals (Briske et al. 

2008). Similarly, small mammals, reptiles, 

amphibians, and bats represent very broad 

wildlife categories that may have diverse 

responses to various conservation practices. 

For example, focusing on Rodentia includes 

species with such widely different habitat 

and life history strategies that responses 

within the group may differ diametrically 

when exposed to the same management 

practice. Furthermore, most of the studies 

that have examined how anthropogenic 

activities on rangelands influence wildlife 

have not classified the management activities 

involved according to the NRCS conservation 

practices. Thus, we refer to related 

conservation practices on rangelands that 

influence wildlife as rangeland activities. 

Wildlife in America has been strongly 

influenced by agriculture; livestock grazing is 

the most widespread land management practice 

in the world (Holechek et al. 2003) and affects 

70% of the land surface in the western United 

States (Fleischner 1994). Traditional practices 

in rangeland management often homogenize 

grazing lands to increase forage production 

and maximize sustainable yield for domestic 

livestock. New management approaches 

that promote the spatial and temporal scale 

of heterogeneity in vegetation structure, 

composition, and biomass so that sufficient 

tracks of particular vegetation associations can 

accommodate desired wildlife populations 

are needed to improve habitat for wildlife 

(Fuhlendorf and Engle 2001; Bruno and 

Cardinale 2008). 

The dynamics of native and domestic 

ungulates, combined with various management 

practices, create a complex interaction that 

influences plant and animal communities 

by altering ecosystem structure, nutrient 

cycling, productivity, recruitment, predator– 

Deer fawn on the Theodore 

Roosevelt Memorial Ranch, 

Dupuyer, Montana along the 

Rocky Mountain Front. (Photo: 

Sonja Smith) 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 255 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	

	 	

Pyrruhuloxias (Cardinalis 

sinuatus) occupy desert scrub 

and mesquite-dominated rangelands in southwestern United 

States. (Photo: Tim Fulbright) 

prey relationships, urination and defecation, 

trampling, and competition. Additional 

modifications to landscapes, including 

roads, fences, anthropogenic water sources, 

agricultural structures, and other developments 

related to livestock production on western 

rangelands, modify wildlife behavior and 

complicate wildlife management. This is 

especially important for wildlife, as domestic 

stock and the related anthropogenic 

developments alter forage availability and 

cover and contribute to habitat alteration and 

fragmentation. Large herbivores may potentially 

modify landscapes in numerous ways (Senft et 

al. 1987; Ohmart 1996; Fuhlendorf and Engle 

2001), but describing them is beyond the scope 

of this chapter. However, it is not surprising 

that the effects of prescribed grazing on wildlife 

have received more attention in the literature 

than other conservation practices. Many of 

the early studies of wildlife parallel livestock 

husbandry and range management theory in 

that grazing and browsing are the primary 

factors affecting the kinds, amounts, and 

quality of forage available (Mackie 2000). 

Our objective was to review peer-reviewed 

literature to examine how conservation 

practices influence wildlife and wildlife habitats 

on rangelands in the United States, with 

specific reference to the NRCS Conservation 

Practice Standard for Upland Wildlife 

Habitat Management. The main purpose of 

this conservation standard is to treat upland 

wildlife habitat concerns identified during 

the conservation planning process that enable 

movement or provide shelter, cover, and food in 

proper amounts, locations, and times to sustain 

wild animals that inhabit uplands during a 

portion of their life cycle. We emphasized the 

literature compiled in the bibliography by 

Maderik et al. (2006) but also considered other 

articles to provide a more complete review. 

We documented rangeland activities that 

influenced (i.e., positive and negative) game 

birds, nongame birds, carnivores, ungulates, 

small mammals, reptiles, and amphibians 

on western rangelands. Carnivores are rarely 

considered by NRCS, but we include them 

in our review because of their importance to 

functioning ecosystems. We also identified 

gaps in scientific knowledge and recommended 

future research to enhance management of 

wildlife on western rangelands in the United 

States. We supplemented the synthesis with 

literature outside the United States when 

similar knowledge within the United States was 

not available. 

Results of tHe lIteRAtuRe 

Assessment 

Very few of the NRCS conservation practices 

that directly affect upland wildlife habitat are 

addressed or evaluated in the peer-reviewed 

literature. We identified specific activities 

when appropriate; however, this review is 

dominated by grazing because of the high 

profile that grazing has received by the scientific 

community. Prescribed grazing, when carefully 

controlled, can be useful in improving habitat 

for specific species, but the frequency, timing, 

256 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

and intensity of livestock grazing for maximum 

wildlife benefits are different than those used 

for maximum livestock benefits (Holechek et 

al. 1982). For wildlife, the amount of critical 

residues left after prescribed grazing is more 

important than the amount removed; the 

condition of most ranges will deteriorate when 

greater than 50% of grazable vegetation is used 

annually (Hyder 1953; Holechek et al. 1982). 

More than 25 yr ago, Holechek et al. (1982) 

reviewed how prescribed grazing could improve 

wildlife habitat and concluded that the 

database was limited. They argued that research 

into how grazing strategies influence wildlife 

should receive high priority. Unfortunately, 

peer-reviewed literature evaluating conservation 

practices for upland wildlife habitat 

management, including prescribed grazing, has 

not received high priority, and the complex 

influences on wildlife and their habitat remain 

largely unknown. 

Rangeland Activities and Habitat for 

Game Birds 

Conservation practices that improve habitat, 

if identified and implemented, may halt the 

decline or, in many cases, enhance the viability 

of game bird populations. Distribution 

and abundance of native grouse (subfamily 

Tetraoninae) that symbolize the biological 

diversity of western grazing lands are in decline 

(Knick et al. 2003; Hagen et al. 2004) or are 

already threatened or endangered (Storch 

2007). Exceptions include spruce grouse 

(Canachites canadensis L. 1758) and blue grouse 

(Dendragapus obscurus Say 1823) populations 

and most white-tailed ptarmigan (Lagopus 

leucurus Richardson 1831) populations. 

Indigenous quail (subfamily Phasianinae) 

populations, though stable locally, are largely 

in decline in the desert Southwest (Saiwana 

et al. 1998; Western Quail Management Plan 

2008) and in the southern Great Plains (Veech 

2006). Species considered here that are native 

to western grazing lands include Gunnison 

(Centrocercus minimus Young et al. 2000) and 

greater sage-grouse (C. urophasianus Bonapart 

1827); lesser (Tympanuchus pallidicinctus 

Ridgeway 1873), greater (T. cupido L. 1758), 

and Attwater’s prairie-chicken (T. cupido 

attwateri); plains (T. phasianellus jamesi L. 

1758) and Columbian sharp-tailed grouse 

(T. p. Columbians L. 1758); wild turkey 

(Meleagris gallopavo L. 1758); bobwhite 

(Colinus virginianus L. 1758); and scaled quail 

(Callipepla squamata Vigors 1830). 

Our synthesis includes U.S. Department of 

Agriculture (USDA) Conservation Practices 

Standards that benefit native grouse and quail 

and is supplemented with information on exotic 

species (e.g., ring-necked pheasant [Phasianus 

colchicus L. 1758] and grey or “Hungarian” 

partridge [Perdix perdix L. 1758]) that are 

abundant regionally and provide recreational 

and economic benefits (Bangsund et al. 2004). 

We do not synthesize the rich literature for ringnecked pheasant because in-depth reviews for 

this species response to Farm Bill conservation 

practices (Haufler 2007) and other management 

are readily available (Trautman 1982; Berner 

1988; Kimmel and Berner 1998). 

We present findings regionally because 

variation in climatic gradients (Fulbright and 

Ortega-Santos 2006), disturbance regimes 

(Coppedge et al. 2008), and contemporary 

land use change (Foley et al. 2005) influence 

vegetation response to management. We 

critically reviewed strength of evidence because 

variation in study design (Guthery 2007) and 

ecological scale of investigation (Manzer and 

Hannon 2005) further influence applicability 

of research outcomes to management. We 

placed recommendations within the context of 

landscape conservation, a well-known ecological 

principle (Lindenmayer et al. 2008) that is being 

used in management of game birds at large scales 

(Hagen et al. 2004; Manzer and Hannon 2005). 

Landscape Conservation. Public land 

managers use holistic strategies that conserve 

entire landscapes because to be effective 

the scale at which conservation practices 

are implemented must match the scale 

of anthropogenic change that threatens 

populations. Tillage agriculture (Walker et 

al. 2007), urban sprawl (Knick et al. 2003; 

Krausman et al. 2008), tree and shrub 

invasion (Fuhlendorf et al. 2002), and energy 

development (Naugle et al. 2011) result in 

broad-scale loss and degradation of habitat 

that overwhelms management of remaining 

fragments (Fuhlendorf et al. 2002; Veech 

2006). Wholesale fragmentation increases 

predation rates (Manzer and Hannon 2005), 

alters historic disturbance regimes (Baker 

Very few of 

the NRCS 

conservation 

practices that 

directly affect 

upland wildlife 

habitat are 

addressed or 

evaluated in the 

peer-reviewed 

literature.”

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 257 
	 	 	 	

	 	 	 	

	 	 	 	 	

		 	 	

Sage-grouse (Centrocercus 

urophasianus) are an important 

species of concern in western 

rangelands. (Photo: Brett 

Billings) 

2006), promotes the spread of invasive 

plants (Bergquist et al. 2007), and facilitates 

disease (Walker et al. 2007). The concept 

of conserving the remaining “usable space” 

is a primary underpinning for quail habitat 

management in the south-central region of 

the Great Plains (Guthery 1997). Under this 

paradigm, managers should strive to increase 

the quantity of quail habitat. Reversing declines 

in game bird populations will require regional 

management of remaining usable space 

(Williams et al. 2004). 

Rangeland Activities. Most literature 

documents the decline or extirpation of 

wildlife populations that result from chronic 

overgrazing. Overgrazing is defined here as 

the combination of stocking rates and timing 

of grazing that reduces wildlife reproduction 

and survival by altering the short- and longterm structure and composition of grassland 

and shrubland vegetation. This chapter may 

be frustrating for some readers looking for 

precise guidance because little experimental 

research has been conducted to know which 

conservation practices benefit game birds. 

Most contemporary studies lack experimental 

controls, are too short in duration, and fail to 

collect pretreatment data. Moreover, findings 

cannot be readily translated in conservation 

practices (e.g., prescribed grazing) because 

existing studies typically compare wildlife 

response to grazed and ungrazed pastures 

without reference to grazing strategy, regime, or 

system. Implications should not be extrapolated 

too broadly because they are most often derived 

from studies of specific species and local-scale 

management actions. 

Grazing. Livestock grazing is a controversial 

practice because indirect evidence 

overwhelmingly suggests that overgrazing 

reduces nest success (e.g., scaled quail [Pleasant 

et al. 2006], ring-necked pheasant [Clark and 

Bogenschutz 1999] and greater sage-grouse 

[Beck and Mitchell 2000]) and brood survival 

(lesser prairie-chicken [Hagen et al. 2005] and 

wild turkey [Spears et al. 2007]) by decreasing 

height and density of herbaceous cover. 

Livestock grazing can have negative or positive 

impacts on game bird habitat, depending on 

timing and intensity of grazing and which 

habitat component is being influenced (Beck 

and Mitchell 2000). Light to moderate grazing 

can promote forb abundance (e.g., food), 

but heavy grazing reduces herbaceous cover 

and promotes invasive species (Crawford et 

al. 2004). Guidelines describing height and 

density of herbaceous cover necessary to 

maintain productive habitats are available 

for many game bird species (Connelly et al. 

2000; Hagen et al. 2004). These guidelines 

provide the “biological sideboards” necessary 

to guide grazing strategies for maintaining 

and enhancing populations; unfortunately, 

the grazing strategies necessary to achieve the 

necessary cover requirements for game birds are 

poorly understood. 

The only empirical evidence of the influence 

of prescribed grazing on game birds we found 

in the literature was an unpublished report 

(Rice and Carter 1982) from a 5-yr study of 

game birds at Fort Pierre National Grassland 

in central South Dakota. Authors compared 

deferred rotation, rest–rotation, and winteronly grazing. Pastures (404 ha) that were 

deferred from grazing until winter provided 

the highest number of plains sharp-tailed 

grouse and greater prairie chicken nests and 

broods. Rest–rotation grazing accommodated 

the second-highest density of nests and broods 

for both species. Deferred rotation did not 

provide blocks of undisturbed cover available 

in the spring for nesting, which was reflected 

in the lowest density of nests and broods. 

Pastures managed under rest–rotation grazing, 

which had the highest cattle stocking rate of 

any system, produced approximately 10 times 

more nest-broods than did pastures managed 

in a deferred rotation system. During the 5-yr 

study, grouse followed the grazing rotation 

seeking the best herbaceous cover for nesting 

and rearing broods. Grouse preferred rested 

pastures for nesting that were at times 4.0 km 

from breeding sites. 

In the south-central United States (e.g., Texas 

and Oklahoma), grazing management can 

be prescribed to benefit bobwhite habitat, 

but a large part of potential quail range in 

the Rolling Plains has been overgrazed and 

excessively treated for brush control (Rollins 

2007). Today, more landowners are tempering 

traditional land management goals with more 

quail-friendly practices, including reduced 

stocking rates (Rollins 2007). Adequate nesting 

cover is a key consideration for quail managers 

258 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

	 	 	

	 	

	 	 	 	 	

	 	 	 	 	

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

(Slater et al. 2001) because food is rarely the 

limiting factor for bobwhites in Texas (Guthery 

2000). Livestock grazing can be an effective 

tool for managing quail habitat, especially in 

manipulating plant succession (Guthery 1986). 

But across most of Texas, bobwhite abundance 

declines as cattle density increases (Lusk et al. 

2002). Light to moderate stocking rates that 

provide 50% grass and 20% to 30% woody 

vegetation result in adequate bobwhite nesting 

habitat in western Oklahoma (Townsend et al. 

2001). Guthery (1986) emphasizes flexibility in 

grazing prescriptions to allow “slack” (Guthery 

1999) in the system to account for variability 

in brush cover and short- and long-term 

precipitation patterns. 

Other than the examples mentioned above, 

little experimental data are available to 

identify beneficial grazing practices that 

increase bird populations levels (e.g., greater 

sage-grouse [Connelly et al. 2000] and lesser 

prairie-chicken [Pitman et al. 2005]) because 

mechanisms are poorly understood (Beck and 

Mitchell 2000; Hagen et al. 2004). Effects 

of livestock grazing vary regionally because, 

unlike the Great Plains where bison (Bos bison 

H. Smith 1827) once flourished (Sanderson et 

al. 2008), many semiarid sagebrush and arid 

desert ecosystems evolved with substantially game birds include agricultural tillage, 

less grazing (Connelly et al. 2000; Knick et herbicide application, mechanical sagebrush 

al. 2003). Wildlife managers in the Great removal, and overprescription of fire in xeric 

Plains readily acknowledge the importance landscapes. Tillage agriculture directly reduces 

of livestock grazing to conservation because the amount of habitat available and fragments 

ranchers whose operations remain profitable are remaining grasslands to the detriment of 

less likely to convert native prairie to cropland wildlife populations (Swenson et al. 1987). 

(Licht 1997; Higgins et al. 2002). Conversely, Various means of mechanical and herbicidal 

wildlife managers in sagebrush and desert removal of sagebrush (Artemisia spp.) directly 

grasslands see grazing as detrimental because reduce the abundance of shrub and herbaceous 

excessive stocking rates often results in severe vegetation that sage-grouse rely on for food 

habitat degradation (Mack and Thompson and cover (Wallestad 1975; Braun and Beck 

1982; Knick et al. 2003). We need more 1996). Periodic fire may rejuvenate grasslands 

experimental studies like those in Europe in the Great Plains (Reinking 2005; but see 

showing how managed grazing was used to Patten et al. 2007), but widespread burning of 

recover a declining population of black grouse sagebrush landscapes is not warranted in xeric 

(Tetrao tetrix L. 1758) in northern England environments farther west (Beck et al. 2008). 

(Calladine et al. 2002). Black grouse numbers Similarly, lesser prairie-chickens in southeastern 

averaged 6.3% higher per year, and brood New Mexico shrublands selected sand shinnery 

survival was 22% higher at sites with reduced oak (Quercus harardii Rydb.) landscapes 

grazing than in overgrazed reference sites. for thermal refugia and protective overhead 

cover; selection for these landscapes suggests 

Vegetation Manipulations Detrimental no justification for shrub control for prairieto Populations. A host of vegetation chicken conservation in these landscapes (Bell 

manipulations that detrimentally impact et al. 2010). 

Western populations of painted 

buntings (Passerina ciris) breed 

in the shrublands of northern 

Mexico and Texas. (Photo: Tim 

Fulbright) 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 259 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	 	

	 	

	 	

	 	 	

	 	 	

	 	

A sciencebased approach 

is the key to 

implementing the 

right practices 

in the right 

places and then 

documenting 

outcomes 

to wildlife 

populations.”

Exotic and Woody Plant Invasions. 

Activities that enable proliferation of exotic 

herbaceous and woody plants (e.g., tree/shrub 

establishment) in rangelands should be avoided, 

but those that reduce or remove unwanted 

invasive species are encouraged (Flanders et al. 

2006). Game bird populations have suffered 

from human fire suppression that promotes tree 

and shrub invasions and establishment of exotic 

plants that eventually results in catastrophic 

wildfire. An increase in tree abundance is 

associated with lower persistence of lesser 

prairie-chicken populations in Oklahoma and 

Texas (Fuhlendorf et al. 2002). Sage-grouse 

do not use mountain big sagebrush (Artemisia 

tridentata Nutt. vaseyama [Rybd.] Bettle) 

landscapes that are invaded by pinyon (Pinus 

spp.)–juniper (Juniperus spp.) woodlands at 

higher elevations in the intermountain West 

(Miller et al. 2000; Crawford et al. 2004); 

the exact mechanism is unknown, but birds 

either experience higher predation rates or 

avoid tall structures in otherwise suitable 

habitats. Similary, scaled quail avoid grasslands 

invaded by trees in the desert Southwest (Van 

Auken 2000; Bristow and Ockenfels 2006). 

Another major problem throughout much of 

the West is proliferation of cheatgrass (Bromus 

tectorum L.), which reduces viability of game 

bird populations. Invasion of rangelands by 

cheatgrass has led to a cycle in which increasing 

abundance of this annual grass promotes large 

fires that allow cheatgrass to increase further, 

causing the loss of perennial bunchgrasses and 

low-elevation communities of Wyoming big 

sagebrush (Artemisia tridentata wyomingensis 

Beetle and Young) (Knick 1999; Baker 2006). 

This phenomenon is particularly troubling 

because no large-scale restoration techniques 

are currently available to restore the millions 

of ha of sagebrush-dominated rangelands that 

have been lost to wildfire. 

Brush Management. South-central Great 

Plains rangelands have changed greatly 

over the past century as mesquite (Prosopis 

glandulosa Torr.) savannas become increasingly 

dense because of a lack of prescribed fire and 

regrowth from chemical and mechanical brush 

management. Light to moderate stocking rates 

usually provide the proper proportions of bare 

ground, herbaceous quail foods, and woody 

cover required to sustain bobwhite populations 

in Oklahoma (Townsend et al. 2001). Grazing 

intensity will range relative to how much brush 

is present; lighter stocking rates are required to 

maintain more herbaceous cover if little brush 

is present, but heavier stocking rates are possible 

if more brush canopy is present (Guthery 

2002). In the Rolling Plains of Texas, bobwhites 

selected rangelands containing higher brush 

canopy cover and overall visual obstruction over 

those with more bare ground (Ransom et al. 

2008). Weather has a tremendous influence on 

the amount of cattle forage available, leading 

Lusk et al. (2007) to conclude that reducing 

livestock stocking rates during dry periods likely 

will foster ground cover more similar to that 

available during wet periods. The main factors 

influencing bobwhite numbers in southern 

Texas were rainfall during the previous growing 

season and type of range, with treatments to 

reduce brush only nominally affecting bird 

abundance (Cooper et al. 2009). In the same 

areas of Texas, application of prescribed fire at 

large spatial scales was deemed a neutral practice 

for managing bobwhite habitat in semiarid 

rangelands (Ransom and Schulz 2007). 

Strategic Approach to Implementing 

Beneficial Practices. Implementing 

practices that are beneficial to game birds 

is often challenging because many of the 

critical experiments have not been done to 

document positive population responses to 

management. A science-based approach is the 

key to implementing the right practices in the 

right places and then documenting outcomes 

to populations to identify and replicate our 

successes, manage adaptively to improve 

delivery, and provide accountability to all our 

audiences. Implementation of conservation 

practices should be linked with field-based 

experimental research to identify the most 

effective and least expensive ways to benefit 

wildlife populations. Many birds use habitats 

at a spatial scale that is larger than that of 

an individual pasture or ranch. Therefore, 

our scientific assessments should reflect 

appropriately large scales at which game bird 

populations use habitat resources year-round 

and transcend that of an individual ranch 

to encompass multiple and nearby ranches 

enrolled in conservation programs. 

The USDA is trying new and innovative 

ways to link science with implementation to 

document the benefits of NRCS conservation 

260 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 		 	

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

practices. For example, the USDA launched 

its new and exciting Sage-grouse Initiative 

(SGI) in March 2010 to provide a holistic 

approach to conserving sage-grouse and 

sustaining working ranches in the West. In its 

inaugural year, the SGI has quickly become 

one of the largest and most recent conservation 

success stories in the West. The SGI’s success 

is in capitalizing on the strong link between 

conditions required to support sustainable 

ranching operations and habitats that support 

healthy sage-grouse populations. The SGI is a 

science-based initiative with evaluations carried 

out by reputable, independent scientists to 

measure the biological response of sage-grouse 

populations to conservation practices, to assess 

SGI effectiveness, and to adaptively improve 

program delivery. 

The SGI follows three primary steps in 

evaluating the benefits of conservation practices 

that may serve as a model for others dealing 

with uncertainty in their implementation 

effectiveness. First, the NRCS worked with 

the Bureau of Land Management to map 

rangewide sage-grouse population centers, or 

“core areas,” to refine SGI delivery ensuring 

that practices benefit large numbers of birds 

(Doherty et al. 2010). Targeting practices 

within core areas ensures that enough of the 

right conservation practices are implemented 

in the right locations to anticipate a positive 

population response. Similar guidance is 

emerging for targeting conservation practices to 

benefit sustainable bobwhite quail populations 

in the West Gulf Coastal Plain (Twedt et al. 

2007). Second, SGI-sponsored studies are 

under way in six states across the West to assess 

benefits of grazing systems and removal of 

encroached conifer. Assessments incorporate 

before–after control–impact designs using 

radio-marked birds across appropriately 

large time and space scales to quantify the 

biological and population-level response of 

birds to conservation practices. Third, the 

NRCS completed a conference report with the 

U.S. Fish and Wildlife Service (USFWS) that 

proactively amends a suite of 40 conservation 

practices to ensure they are either benign or 

beneficial to sage-grouse, including upland 

habitat management, prescribed grazing, and 

brush management for juniper removal. By 

conditioning NRCS conservation practices, 

private landowners enrolled in SGI can rest 

assured that they can continue normal ranching 

operations even if USFWS lists sage-grouse as 

a federally threatened or endangered species. 

Collectively, these three steps offer an approach 

for implementing conservation practices while 

documenting their success and adaptively 

improving them when necessary. 

Rangeland Activities that Improve 

Habitat for nongame Birds 

Rangeland management has great potential to 

improve nongame bird habitat (Haufler and 

Ganguli 2007). To date, most studies address 

management effects, not necessarily benefits, 

on focal species or avian communities. This is 

logical because biologists must first understand 

the nature of the effects (e.g., positive, negative, 

or neutral) to effectively use a given management 

practice as a tool. However, the science has not 

progressed much beyond this preliminary phase, 

and experimental studies designed specifically to 

evaluate management actions to benefit wildlife 

are rare. We approach the review of conservation 

practices to improve habitat for nongame birds 

with a brief mention of key effects papers and 

then review papers that evaluate the efficacy 

of management with the primary objective to 

improve nongame bird habitat. 

Effects Papers. By far, the focus of most 

research has been to address effects of livestock 

grazing on nongame birds (Fleischner 1994; 

Saab et al. 1995; Zimmerman 1997). Research 

Young white-tailed deer 

(Odocoileus virginianus) in 

eastern Wyoming. (Photo: 

David Briske) 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 261 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	 	 	

	 	 		

	 	

has been conducted also to understand effects 

of fire, mowing, and exotic flora and fauna 

(Herkert et al. 1996; Zimmerman 1997; Askins 

et al. 2007). Effects are attributed primarily to 

changes in habitat structure and composition 

(Bock and Webb 1984), although trampling 

of ground nests occur. Indirect effects are 

ascribed to changes in ecosystem structure that 

can influence ecological relationships among 

species. The focus of much attention here 

concerns brown-headed cowbirds (Molothrus 

ater Boddaert 1783) that parasitize nests of 

many cup-nesting species. 

Given that the focus of this chapter is not to 

review these studies, suffice it to say that effects 

on species vary from positive to negative. 

Perhaps the relevance of these effects studies is 

that they indicate management activities that 

are benign, beneficial, or detrimental to species, 

which is a critical first step in developing 

proactive management prescriptions. 

Grazing. Various studies evaluate grazing as a 

tool to enhance nongame bird habitat. Grazing 

is not restricted to exotic domestic herbivores 

but also includes native species, such as bison 

and elk (Cervus canadensis L. 1758). Indeed, 

many, if not most, ecosystems rely on grazing 

by native ungulates to influence vegetation 

structure and composition (Stebbins 1981); 

Pronghorn antelope 

(Antilocapra americana) on the 

Charles M. Russell National 

Wildlife Refuge, Montana. 

(Photo: Jeffrey Wright) 

thus, some form and level of grazing may be 

compatible with natural ecosystems processes. 

Grazing variables that can be manipulated to 

achieve nongame bird goals include stocking 

rates, seasonality, duration, and livestock 

species. A premise of prescribed grazing is that, 

if done correctly, it will enhance horizontal 

heterogeneity and provide a mosaic of 

landscape conditions to meet a wide range of 

bird preferences (Herkert et al. 1996; Derner et 

al. 2009). 

Wetland Birds. Grazing improved habitat 

for wading birds in Austria (Kohler and Rauer 

1991). Two factors led to degraded habitat: 

conversion of pastureland to agriculture 

and the cessation of grazing that allowed for 

encroachment of common reeds (Phragmites 

spp.) and rushes (Juncus spp.) into pastureland. 

Cattle were introduced to control the 

encroachment of reeds and rushes, but Kohler 

and Rauer (1991) noted no tangible increases 

of wading bird populations. Tichet et al. (2005) 

evaluated grazing regimes (stocking levels and 

seasonality) on use by wading birds in French 

wetlands. They found that grazing intensity 

affected species responses differently, depending 

on their habitat requirements. Curlews 

(Numenius arquarta L. 1758) used areas 

with greater spring grazing intensity, whereas 

redshank (Tringa tetanus L. 1758) occupancy 

declined. In autumn, lapwings (Vanellus 

vanellus L. 1758) showed a positive relation 

to grazing, whereas responses by black-tailed 

godwits (Limosa limosa L. 1758) were negative. 

Grassland Birds. Paine et al. (1997) compared 

three grazing regimes in Wisconsin: grass farms, 

continuously grazed pastures, and “bird-friendly” 

rotational systems whereby grazing was deferred 

to create nesting refuges during the breeding 

season. They reported that refuges attracted 11% 

more nesting birds than grass farms and that 

grass farms attracted 65% more nesting birds 

than continuously grazed pastures. Nest success 

for grass farms ranged from 6% to 24% and 

from 30% to 39% for refuges and was 25% for 

continuous grazing during both years of study. 

Most nest mortalities for grass farms and refuges 

were from mowing. Overall avian productivity 

within refuges was greater than that for grass 

farms, which were greater than continuously 

grazed pastures. Productivity is defined as the 

number of birds fledged from nests. Temple 

262 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	 	 	

	 	 	

	 		 	

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

et al. (1999) also compared grazing regimes in 

Wisconsin and reported that diversity, density, 

nest success, and productivity of grassland birds 

was greatest on ungrazed lands. Continuously 

grazed pastures had the lowest diversity and 

densities but were intermediate for nest success 

and productivity. Rotationally grazed pastures 

had intermediate diversity and densities but 

the lowest nest success and productivity. They 

recommended a mosaic of ungrazed and 

rotationally grazed areas to increase productivity 

of grassland birds above that found on a mosaic 

of continuously and rotationally grazed pastures 

(Temple et al. 1999). 

Derner et al. (2009) suggested that livestock 

could be used as “ecosystem engineers” to 

modify vegetation structure within and among 

pastures and provide for habitat needs of 

grassland birds of the Great Plains. Grazing is 

often used in combination with patch burning 

to provide the desired vegetation structure. For 

example, localized grazing and fire could be 

used to reduce vegetation cover and provide 

feeding sites for mountain plover (Charadrius 

montanus Townsend 1853) or nest sites for 

the long-billed curlew (Numenius americanus 

Bechstein 1812). For the upland sandpiper 

(Bartramia longicauda Bechstein 1812), 

reduced grazing could be used to provide 

tall vegetation required for nesting, whereas 

more intensive grazing could increase food 

availability and enhance foraging habitat. 

Riparian Birds. Livestock grazing can have 

positive and negative effects on habitats for 

different species of birds riparian systems. 

Although grazing removes lower vegetation 

layers, it also influences seedling establishment 

and regeneration of shrubs and trees. Indeed, 

dramatic changes in vegetation structure can 

be seen shortly after livestock are removed 

from riparian areas (Krueper et al. 2003). In 

the Northwest, vegetation recovery following 

livestock removal in a riparian meadow was 

complex, given interactions with precipitation 

(Dobkin et al. 1998). Cattle removal resulted in 

a more diverse and abundant avian community 

that was even greater in wet years than in dry 

years. The northern harrier (Circus cyaneus L. 

1766), common snipe (Gallinago gallinago 

L. 1758), short-eared owl (Asio flammeus 

Pantoppidan 1763), song sparrow (Melospiza 

melodia Wilson 1810), and yellow-headed 

blackbird (Xanthocephalus xanthocephalus 

Bonaparte 1826) were found only within the 

cattle-excluded area. In southeastern Arizona, 

density of herbaceous vegetation increased fourto sixfold following removal of cattle (Krueper 

et al. 2003). Mean numbers of detections 

during bird surveys increased for 42 species 

(26 significantly) and decreased for 19 species 

(8 significantly) 3 yr following the removal 

of cattle. Number of individuals detected per 

kilometer more than doubled. Detections of 

open cup-nesting species increased the most 

and Neotropical migratory birds more than 

others. 

Brown-Headed Cowbird Control. 

Reductions in cattle stocking by 86% (752 

animal units [AUM] to 103) were made to 

decrease nest parasitism on the endangered 

black-capped vireo (Vireo atricapilla 

Woodhouse 1852) in Texas (Kostecke et al. 

2003). Rates of cowbird parasitism decreased 

by 13 times after cattle were removed. 

Further, cowbirds needed to travel further 

to breed, resulting in greater energetic costs 

and reductions in numbers of eggs laid. There 

was no evidence of cowbird nest parasitism 

following removal of cattle from a riparian 

area in southeastern Oregon, even though nest 

parasitism was prevalent in nearby riparian 

habitats where cattle remained (Dobkin et al. 

1998). 

Multiple Range Activities. Walk and Warner 

(2000) compared burned, mowed, hayed, 

grazed, and undisturbed management regimes 

on areas of introduced cool-season grasses, 

native warm-season grasses, and annual 

forbs. Eastern meadowlark (Sturnella magna 

L. 1758) and dickcissels (Spiza americana 

Gmelin 1789) were detected most often 

among grazed warm-season grasses. Henlow’s 

sparrows (Ammodramus henslowii Audubon 

1829) and field sparrows (Spizella pusilla 

Wilson 1810) were detected more often 

among undisturbed warm-season grasses where 

eastern meadowlarks and grasshopper sparrows 

(Ammodramus savannarum Gmelin 1789) 

were least abundant. Grasshopper sparrows 

were most abundant among annual weeds 

where Henlow’s sparrows and field sparrows 

were not observed. Overall abundance was 

least among recently burned cool-season 

grasses. Low-intensity late-season grazing was 

Mule deer (Odocoileus hemionus) on the Theodore Roosevelt 

Memorial Ranch, Dupuyer, 

Montana. (Photo: Sonja Smith) 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 263 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	 	 	

	 	 	

	 	

	 	 	

	 	

	 	

...cattle removal 

or reduction 

seems to be an 

effective tool to 

reduce brownheaded cowbird 

numbers and nest 

parasitism on 

open cup-nesting 

birds.”

important for creating a heterogeneous mosaic 

to accommodate many of the grassland birds 

studied. 

Griebel et al. (1998) evaluated bird use of two 

different grazing treatments: 1) bison grazing 

(year-round; 1.2 AUM · ha−1 · yr−1) combined 

with prescribed fire and 2) cattle grazing (15 

May–15 November; 1.0 AUM · ha−1 · yr−1). 

Few differences were reported in bird species 

richness or relative abundance of species 

between grazing treatments, vegetation density, 

and height. During 1 of the 2 yr of study, 

bird species richness was greater in the bisonfire enclosure than in the cattle enclosure; 

abundances of lark sparrows (Chondestes 

grammacus Say 1823) and mourning doves 

(Zenaida macroura L. 1758) were higher 

and grasshopper sparrow lower in bison-fire 

enclosures. Within the bison-fire enclosures, 

differences existed between burned and 

unburned transects, with grasshopper sparrow 

abundance higher in unburned areas and 

mourning dove and lark sparrow abundances 

higher in burned areas. 

Danley et al. (2004) reported few differences 

in bird species diversity or abundance between 

areas that were burned and grazed versus areas 

only burned in North Dakota. The notable 

exception was the brown-headed cowbird, 

which occurred 2.4 times more frequently on 

burned and grazed plots. 

LaPointe et al. (2003) evaluated use of a 

rest–rotation grazing system targeted to 

improve plant cover for nesting ducks and 

grassland birds along the St. Lawrence River, 

Quebec. They evaluated four methods: cattle 

removal, grazing augmented with seeding of 

forage plants, seeding with no grazing, and 

seasonal grazing after duck nesting. Overall 

abundance of birds exhibited no change 2 yr 

posttreatment. However, bobolink (Dolichonyx 

oryzivorus L. 1758) were more abundant in 

areas that were seeded with no grazing and 

where cattle grazed after ducks had nested, 

and red-winged blackbirds (Agelaius phoeniceus 

L. 1756) were more abundant in the two 

treatments with no grazing. 

Rangeland Restoration. Ecological restoration 

is a management paradigm whose objective is 

to return conditions to those that existed in 

the past, typically those that occurred prior 

to European settlement of North America. 

Implicit is that, in doing so, avian community 

structure and composition will be restored also. 

At this point, results of the few studies that 

have evaluated effects of restoration of birds are 

equivocal. 

Fletcher and Koford (2003) evaluated effects 

of restoring native grasslands from former 

agricultural (e.g., hay land and row crops) 

land and reported that 16 of 54 species 

detected increased with restoration. Only 

killdeer (Charadrius vociferus L. 1758 ) and 

cowbird responded negatively to restoration. 

Five of the species that increased are of broad 

regional concern because populations are 

declining. In contrast, Van Dyke et al. (2004) 

found no bird responses, positive or negative, 

to the use of fire and mowing to restore 

tallgrass prairie in Iowa. Results may have 

been influenced by the small scale (< 10 ha) of 

treatments. 

In southeastern Arizona, Malcolm and Radke 

(2008) evaluated effects of active wetland 

and riparian restoration following passive 

restoration (e.g., cattle removal) on bird density 

and diversity. Cattle removal occurred in 1980 

and was followed by active restoration in 2005. 

Active restoration consisted of installation of 

erosion control gabions to create two wetlands 

that were then used to irrigate a desert scrub 

plot. Bird densities increased by 2.3 birds · ha−1 

in 2006 and 8.4 · ha−1 in 2007 following active 

restoration treatments. Species richness showed 

a marginal difference. 

Kennedy et al. (2008) compared cover by 

native versus nonnative plants and the resulting 

influence on nest productivity of passerine 

birds. They reported no association between the 

percentage of nonnative plant cover and nest 

densities, clutch size, productivity, nest survival, 

and nestling size. 

Overall, studies evaluating effects of range 

management directed at improving nongame 

bird habitat are rare. Many studies are 

essentially case studies whose results apply 

largely to the place and time of study. As a 

result, generalizations are difficult at best. 

Some trends that emerged from the papers 

reviewed are that continuously grazed pastures 

264 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	

	 	 	

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

appear to have fewer birds and fewer species 

than areas grazed using a rotational system, 

grazed after the breeding season, or where 

cattle were removed entirely. Additionally, 

cattle removal or reduction seems to be an 

effective tool to reduce brown-headed cowbird 

numbers and nest parasitism on open cupnesting birds. 

Rangeland Activities and Habitat for 

carnivores 

References regarding influences of rangeland 

activities on carnivores are notably sparse and 

are rarely considered by the NRCS. However, 

we include them in this review because of 

their importance to functioning ecosystems. 

We considered 14 taxa to be representative 

of western rangeland habitats: coyote (Canis 

latrans Say 1823), wolf (Canis lupus L. 1758), 

kit fox (Vulpes macrotis Merriam 1888), 

swift fox (Vulpes velox Say 1823), red fox 

(Vulpes vulpes L. 1758), grey fox (Urocyon 

cinereoargenteus Schreber 1775), black bear 

(Ursus americanus Pallos 1780), grizzly bear 

(Ursus arctos horribilis L. 1758), mountain 

lion (Puma concolor L. 1771), bobcat (Lynx 

rufus Schreber 1777), raccoon (Procyon lotor 

L. 1758), striped skunk (Mephitis mephitis 

Schreber 1776), spotted skunk (Spilogale 

spp.), and black-footed ferret (Mustela negripes 

Audubon and Bachman 1851). 

Based on references in the bibliography of 

Maderik et al. (2006), rangeland activities 

appear to influence habitat for spotted skunks 

and striped skunks (Neiswenter and Dowler 

2007). Spotted skunks use areas with more 

large mesquites than striped skunks, and striped 

skunks did not select any habitat relative to its 

availability, but both species appeared to avoid 

agricultural areas. Conservation of western 

spotted skunks may be enhanced by limiting 

brush control for management of livestock on 

mesquite dominated rangelands (Neiswenter 

and Dowler 2007). 

Others reported that the distribution and 

shape of grassland patches, woodland patches, 

pastureland, and farmsteads influenced 

detections of striped skunks, raccoons, and 

red fox. Kuehl and Clark (2002) determined 

that evidence of striped skunks decreased as 

distance from grassland patches increased but, 

in contrast to Neiswenter and Dowler (2007), 

was positively associated with the number 

of farmsteads in their study area. Raccoon 

presence was positively related to presence of 

woody cover, and red fox presence increased 

with greater area of pastureland and greater 

isolation from farmsteads but decreased with 

increasing amounts of habitat arranged in 

strips across the landscape. Ivan et al. (2002) 

reported that alteration of prairie landscapes 

through increases in planted trees, woody 

cover, rock piles, and junk piles enhanced 

conditions for striped skunks and raccoons 

by providing denning habitat. Maestas et al. 

(2003) concluded that ranchlands supported 

relatively more coyotes than exurban 

developments and that ranches are important 

for protecting biodiversity, suggesting that 

future conservation efforts may require less 

reliance on reserves and a greater focus on 

private lands. In ecologically similar areas of 

Arizona, Horejsi (1982) reported that coyotes 

were relatively more abundant on ungrazed 

than on grazed rangelands; however, the 

ungrazed area had been closed to predator 

control for an extended period of time prior 

to the initiation of his research, and the other 

had not. These results have implications 

for predicting the influence of rangeland 

management practices on specific species 

of carnivores and, through their affects on 

Rio Grande wild turkeys 

(Meleagris gallopavo 

intermedia) are found along 

riparian areas and in 

shrublands. (Photo: Tim 

Fulbright) 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 265 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	

	 	 	

...influences 

of rangeland 

activities on large 

native carnivores 

have nearly all 

been negative.”

landscape configuration, on conservation of 

biodiversity in general. 

Hilty and Merenlender (2004) emphasized 

that wide, well-vegetated riparian corridors 

are important in maintaining the connectivity 

of native predator populations to ensure 

their long-term survival. In a similar riparian 

system, Ammon and Stacey (1997) concluded 

that livestock grazing reduced streamside 

vegetation and that grazing could influence 

predator assemblages and, thereby, affect bird 

populations directly and indirectly. Cattle 

grazing did not affect vegetation height or 

density along edges of pasturelands compared 

to the interior of pasturelands. Raccoons and 

other predators may move more freely in 

pasturelands when compared with edges of 

pasturelands, thereby explaining an absence 

of differences in predation risk for nesting 

grassland birds in those habitats (Renfrew et al. 

2005). Conversion of rangelands to irrigated 

agriculture (i.e., alfalfa, mint, and sugar beets) 

may have a positive effect on burrowing owls 

(Athene cunicularia Molina 1782) where those 

small strigids use burrows abandoned by 

badgers (Taxidea taxus Schreker 1777; Belthoff 

and King 2002); presumably, such practices 

have a negative affect on badgers although not 

explicitly stated. 

Numerous references included in the 

bibliography by Maderik et al. (2006) (Beck 

and Mitchell 2000; Townsend et al. 2001; 

Herkert et al. 2003; Cox et al. 2005; Miller and 

Guthery 2005; Renfrew et al. 2005; Shochat et 

al. 2005; Sutter and Ritchison 2005; Grant et 

al. 2006) make inferences about onerous affects 

of grazing on predator assemblages or the 

ability of predators in general to better detect 

and prey on the nests of ground-nesting birds. 

Results of these investigations address primarily 

changes in predation risk to ground-nesting 

birds as a result of modifications to habitat 

structure or composition rather than changes in 

carnivore populations themselves. 

Generalizations About overall effects 

of management on carnivores 

Habitat alteration and loss and harvesting for 

sustenance, sport, and profit have resulted 

in substantial declines in top predators in a 

wide variety of habitats (Bruno and Cardinale 

2008), including rangelands of western 

North America (Laliberte and Ripple 2004). 

Overgrazing of rangelands by domestic 

livestock, sometimes combined with other 

practices, has influenced the structure and 

composition of rangeland habitats, with 

resultant impacts to biodiversity and ecosystem 

function (Blaum et al. 2007). Additionally, 

efforts to enhance livestock production on 

those rangelands have included attempts to 

eliminate carnivores viewed largely as predators 

of livestock. As a result, influences of rangeland 

activities on large native carnivores have 

nearly all been negative. Nevertheless, some 

medium-sized carnivores (e.g., coyotes, skunks, 

and raccoons) have experienced increases in 

populations and distribution, in part resulting 

from an enhanced food base associated with 

human presence or the absence of predators 

that no longer compete with or prey on those 

carnivores. 

Four of the taxa (i.e., wolf, grizzly bear, blackfooted ferret, and San Joaquin kit fox [Vulpes 

macrotis mutica Merriam 1902]) have been 

impacted by activities associated with rangeland 

management (i.e., predator control activities, 

habitat modification, and conversion) to the 

extent that they have been afforded federal 

protection under the Endangered Species Act. 

Two others (i.e., mountain lion and swift 

fox) have suffered substantial reductions in 

distribution and numbers. 

Throughout much of the history of western 

North America, ranchers and other livestock 

producers have viewed large carnivores as 

incompatible with production objectives. 

Ranchers and other rangeland managers viewed 

predator management as an augmentation of 

the efficacy of other practices, and, as such, 

it has become a widespread and accepted 

practice throughout much of the United States. 

Although predator control is not explicitly one 

of the NRCS rangeland management practices 

currently in place, it has been (and in some cases 

likely will continue to be) an activity that occurs 

in conjunction with current NRCS practices 

that place an emphasis on habitat quality 

and enhancement. As such, a brief history of 

predator management and its impacts on species 

and ecosystems is warranted in this chapter. 

Moreover, some carnivores have benefited from 

implementation of selected NRCS management 

practices and warrant recognition. 

266 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

Widespread efforts to eliminate wolves and 

grizzly bears from rangelands in the 48 

contiguous states were largely successful 

(Young 1944; Storer and Tevis 1955; Mech 

1970; Brown 1985), and the use of a variety of 

techniques, including widespread campaigns 

of poisoning, trapping, and shooting, 

ultimately resulted in the previously mentioned 

classification of those large carnivores as 

endangered taxa. Another large carnivore, 

the mountain lion, also was the object of less 

successful but still intensive (Bruce 1953; Hert 

and McMillin 1955) efforts to reduce impacts 

to livestock operations. 

Gray wolves once ranged throughout much 

of North America but were systematically 

eliminated from the majority of historical 

habitats in part because of the threat to 

livestock (Musiani and Paquet 2004). 

Indeed, it is estimated that wolves had been 

eliminated from greater than 85% of their 

former range in rangeland habitats prior 

to restoration efforts (Laliberte and Ripple 

2004). Nevertheless, federal protection, 

combined with efforts to manage wolves in 

the north-central United States (Mech 1970) 

and efforts to restore them within historical 

ranges in the northern Rocky Mountains 

(USFWS 1987), has been successful. 

Wolves remain important predators of 

livestock, but current management strategies 

include provisions for removal of offending 

individuals. 

Grizzly bears once occupied suitable habitat 

across a wide expanse of the continental 

United States, but their geographic range 

has been reduced by 91% in temperate 

grasslands, savannas, and shrublands and 

by 100% in desert and xeric shrublands 

(Laliberte and Ripple 2004), largely a result 

of efforts to eliminate historic conflicts with 

livestock grazing and other human activities. 

Grizzly bears were afforded protection under 

the Endangered Species Act in 1975, and an 

initial recovery plan was completed in 1982 

and revised in 1993 (USFWS 1982, 1993). 

Currently, grizzly bears are categorized as 1) 

an experimental, nonessential population 

segment in parts of Idaho and Montana and 

2) a recovered distinct population segment 

in the Greater Yellowstone Ecosystem of 

Idaho, Montana, and Wyoming. Elsewhere 

in the continental United States, grizzly 

bears remain listed as threatened, but the 

status of populations inhabiting the CabinetYaak Recovery Zone, the Selkirk Recovery 

Zone, and the North Cascades Ecosystem 

Recovery Zone are under review (USFWS 

2008). Recovery of grizzly bears is dependent 

on the maintenance of suitable habitat in 

occupied areas and judicious management of 

individuals that prey on livestock. 

Populations of swift foxes and kit foxes declined 

substantially as a result of rangeland activities, 

and their influences, including habitat loss 

through conversion of native prairies, trapping, 

predator control, shooting, collisions, and use 

of rodenticides to control prey populations, 

likely contributed to the decline of swift 

foxes (Carbyn 1995; Meaney et al. 2006). 

Further, unanticipated trophic cascades due to 

widespread removal of wolves and subsequent 

increases in coyotes and, potentially, red foxes, 

which prey on or compete with these small 

canids, likely have contributed to the decline 

of swift and kit foxes (Carbyn 1995; Cypher 

et al. 2001; Meaney et al. 2006). Alteration of 

native prairies due to grazing and agricultural 

practices has been especially problematic 

for these foxes, and losses were exacerbated 

by poisoning, trapping, and other efforts to 

manage larger predators, including coyotes and 

wolves (USFWS 1983, 1995). 

Mountain lions can be important predators of 

livestock, particularly domestic sheep, which 

are grazed widely on western rangelands. Efforts 

to reduce mountain lion populations were 

intense during the early 20th century (Bruce 

1953; Hert and McMillin 1955), but those 

activities declined substantially in most of 

the western states by the 1970s. Nevertheless, 

it is estimated that the geographic range of 

mountain lions occupying western rangelands 

has been reduced by 49%; distribution of 

those large felids in desert and xeric shrublands 

has, however, remained unchanged (Laliberte 

and Ripple 2004). Although mountain lions 

were successfully eliminated from a substantial 

proportion of their historical distribution, 

they remain the most widely distributed large 

carnivore in North America (Pierce and Bleich 

2003). In some areas of the southwestern United 

States, mountain lion populations have been 

subsidized by increased food supplies in the form 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 267 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

of domestic livestock that allow mountain lions 

to persist at higher densities, and, as a result, 

the effects of predation on native ungulates 

have been exacerbated (Rominger et al. 2004). 

Increased shrub cover on rangelands often is 

associated with overgrazing (Blaum et al. 2004), 

with resultant influences on biodiversity of 

mammalian carnivores (Blaum et al. 2007) that 

may enhance hunting efficiency of mountain 

lions. Reduction of shrub cover on rangelands 

may decrease hunting efficiency of mountain 

lions, and conversion of cow–calf operations 

to steer operations may decrease the benefits 

of livestock operations to mountain lions and, 

thereby, reduce their impacts on native ungulates 

(Rominger et al. 2004). Currently, mountain 

lions are managed as a game species in the 

majority of western states, but exceptions occur 

(Pierce and Bleich 2003; Bleich and Pierce 

2005). 

Black-footed ferrets have declined substantially 

in distribution and once were thought to be 

extinct in the wild. Widespread poisoning 

campaigns to eliminate prairie dogs (Cynomys 

spp., a principal prey of these endangered 

mustelids) from rangelands were implicated 

in the near extinction of that species, as has 

conversion of rangeland to cropland (USFWS 

1988). As a result of a captive breeding 

program, black-footed ferrets have been 

translocated to appropriate habitats in several 

states but remain one of the most critically 

endangered mammals in North America. 

Coyotes have been an unanticipated 

beneficiary of widespread efforts to reduce 

wolves, and the distribution and range of 

coyotes have increased substantially as a result. 

Although direct mortality of coyotes due 

to wolf predation was low, results of recent 

research are consistent with the hypothesis that 

coyote abundance is limited by competition 

with wolves (Berger and Gese 2007). Trophic 

cascades involving wolf removal and resultant 

expansion of the distribution of coyotes, a 

generalist predator, have resulted in further 

impacts to smaller canids, including swift 

fox and kit fox (Cypher et al. 2001). Coyote 

control is an important rangeland activity, and 

substantial research on control efficacy and 

methodology has been conducted (Knowlton 

et al. 1985, 1999; Shivik 2006). Coyotes 

remain an important predator of livestock, 

particularly domestic sheep, but governmentsubsidized predator control alone has failed to 

prevent a decline of the sheep industry (Berger 

2006). Coyote control to benefit livestock 

production can, however, have a positive effect 

on native ungulates, including mule deer (Odocoileus hemionus Rafinesque 1817) and 

pronghorn (Antilocapra americana Ord 1815; 

Harrington and Conover 2007). Similarly, 

interference competition by wolves with 

coyotes has a positive influence on survival of 

pronghorn fawns (Berger et al. 2008). 

In general, mammalian carnivores have 

benefited little from rangeland management 

activities. An exception is the coyote, a 

generalist predator that has expanded its 

distribution substantially as a result of the 

extirpation of the wolf from the majority of its 

historical range. Such shifts have, however, had 

detrimental affects on other native carnivores. 

It is well established that predators play a vital 

role in maintaining structure and stability of 

communities and that removal of predators 

can have a variety of cascading, indirect effects 

(Terborgh et al. 2001; Duffy 2003). Indeed, 

impacts of rangeland activities that have 

targeted predators for reduction to enhance 

livestock productivity extend far beyond 

the anticipated outcomes. Further, current 

investigations of trophic cascades resulting 

from the elimination of top predators can have 

implications beyond the immediate ecosystems 

occupied by those carnivores (Berger et al. 

2001). Moreover, reduction of top carnivores 

can lead to unanticipated detrimental impacts 

to species that may otherwise not have been 

preyed on as a result of mesocarnivore release, 

whereby midsized carnivores benefit from 

a reduction in the numbers or densities of 

top carnivores (Berger et al. 2008). Thus, 

a consequence of the elimination of many 

carnivores from rangelands in North America 

has resulted in indirect impacts to other species 

and other than the rangeland ecosystems 

from which the carnivores in question were 

eliminated. 

Rangeland Activities and Habitat for 

native ungulates 

Because livestock and wild ungulates share 

rangelands, managers have examined the 

influence of cattle and domestic sheep on the 

vegetation used by white-tailed deer (Odocoileus 

268 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

virginianus Zimmerman 1780), mule deer, elk, 

bighorn sheep (Ovis canadensis Shaw 1804), 

and pronghorn. In general, livestock using 

ranges shared with wildlife have historically 

had more negative than positive influences on 

ungulates, and grazing is not always considered 

an important conservation practice with 

beneficial outcomes. However, some studies 

examined how livestock influenced vegetation 

but did not present data related to how those 

influences altered productivity and recruitment 

of ungulates. Below are examples of studies 

that examined the use of prescribed grazing 

as a conservation practice for several ungulate 

species. 

Pronghorn. Pronghorn populations have 

declined on the Anderson Mesa, Arizona, and 

cattle were considered a key factor in altering 

habitat. Five years after cattle were removed 

from Anderson Mesa, hiding cover (for fawns) 

increased by 8% at a distance of 5 m, but 

no differences were reported at 10 or 25 m 

(Loeser et al. 2005). Forb richness decreased 

in the fifth year after cattle removal by 16% 

but not in the following year, and canopy 

cover was unaffected. It will likely take longer 

than 5 yr of cattle absence to reverse damage 

that has occurred to this fragile environment, 

or some mechanism other than grazing was 

involved. However, pastures grazed by livestock 

conservatively or moderately were not used by 

pronghorn in New Mexico (Jamus et al. 2003). 

In the Desert Experimental Range, Utah, 

pronghorn distribution was related to domestic 

sheep grazing, black sagebrush (Artemisia 

nova Beetle and Young), and topographic 

characteristics. Pronghorn-selected areas 

ungrazed by cattle and areas used moderately 

by sheep during dormant periods were not 

favorable for pronghorn (Clary and Beale 

1983). Nevertheless, Mosley (1994) suggests 

that grazing rangelands by domestic sheep 

can be beneficial to wildlife habitat. However, 

Schwartz et al. (1977) suggest that pronghorn 

coexist on rangelands more successfully with 

cattle than with sheep. 

White-Tailed Deer. Most of the studies 

examining livestock interactions with whitetailed deer have documented how deer respond 

to livestock under different grazing systems. 

From these data, conservation practices have 

been recommended. In general, white-tailed 

deer avoid livestock, and livestock operations 

are more profitable when deer are not 

considered in the operation (Bernardo et al. 

1994). Conversely, returns from livestock were 

maximized when wildlife was not considered; 

however, small reductions in net gains (from 

livestock) can improve wildlife habitat 

(Bernardo et al. 1994). 

The diets of white-tailed deer and cattle are 

different (i.e., deer consume forbs, and cattle 

consume more grass), and deer are more 

sensitive to grazing treatments than cattle. To 

enhance forage for white-tailed, cattle should 

be stocked at moderate rates with continuous 

grazing (or even less intensive grazing) to create 

environments where deer can select more 

forbs (Ortega et al. 1997a, 1997b). Dietary 

protein for growth and lactation of whitetailed deer was not met with short-duration or 

continuous grazing. However, the latter system 

may provide deer with more diversity and 

greater nutrition (Ortega et al. 1997b). Deer 

avoided concentrations of cattle and travel 

farther under short-duration than continuous 

grazing systems (Cohen et al. 1989). However, 

home ranges of white-tailed deer were not 

significantly different under short-duration or 

continuous grazing systems (Kohl et al. 1987). 

They also avoid anthropogenic water sources 

in short-duration grazing systems because of 

disturbance from humans, fences, and livestock 

(Kie 1991). Anthropogenic water sources for 

white-tailed deer should be on the periphery of 

short-duration grazing systems if it needs to be 

supplied (Prasad and Guthery 1986; Kie et al. 

1991). 

There are fewer studies examining how 

prescribed grazing by domestic sheep 

influenced white-tailed deer (Ekblad et al. 

1993). In Texas, Darr and Kelebenow (1975) 

reported a negative relationship between 

domestic sheep and white-tailed deer due to 

removal of cover by the former. 

Mule Deer. Overall, the best practice related 

to grazing for mule deer is to minimize cattle 

numbers on deer ranges. Moderate to heavy use 

of deer ranges by cattle reduced hiding cover 

(Loft et al. 1987), caused shifts in habitat (Loft 

et al. 1991, 1993), increased competition for 

forage (especially at high stocking rates and in 

livestock using 

ranges shared 

with wildlife have 

historically had 

more negative 

than positive 

influences on 

ungulates, 

and grazing 

is not always 

considered 

an important 

conservation 

practice with 

beneficial 

outcomes.”

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 269 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	 	

	 	 	

Elk (Cervus canadensis) in 

Rocky Mountain National Park, 

CO. (Photo: David Briske) 

dry years; Smith 1949; Kie et al. 1991; Yeo et 

al. 1993), and influenced foraging behavior 

(Loft et al. 1993; Kie 1996). Mule deer avoided 

pastures occupied by cattle (Wallace and 

Krausman 1987; Austin et al. 1983; Austin and 

Urness 1986; Bailey and Rogotzkie 1991). 

However, several investigators examined how 

forage removal influenced mule deer and 

reported that mowing at 50% removal can 

increase grass and total biomass the following 

spring but that fall cattle grazing leaves more 

nutritious plants available in summer (Taylor 

et al. 2004). According to some, spring and 

summer deer ranges can be grazed by cattle an 

average of 70% (relative utilization) to enhance 

the ranges the following year (Short and Knight 

2003). Burning can also enhance mule deer 

habitat (Williams et al. 1980). 

Domestic sheep grazing deer ranges often 

benefit deer by improving forage quality in 

fall and increasing quantity in spring (Rhodes 

and Sharrow 1990). The degree of range 

improvement due to grazing by domestic 

sheep depends on the intensity of grazing and 

weather. Browse quality will improve with 

moderate grazing (40% to 55%) that ends by 

June (Alpe et al. 1999). 

Guidelines to improve the quality of winter 

range for mule deer in the Great Basin were 

developed by Austin (2000) based on a review 

of grazing studies. The following guidelines 

were established to maintain or increase browse 

production on winter range. 

1.	� Graze livestock between 1 May and 30 

June. 

2.	� Alternate grazing by class of livestock. 

3.	� Use rest–rotation with yearly grazing 66% 

of the total rangeland. 

4.	� Graze livestock to remove 50% of 

understory grasses and forbs. 

5.	� Balance deer browsing in winter and 

livestock grazing in spring. 

6.	� Monitor utilization using permanent plots. 

Elk. Studies examining how livestock influence 

elk were similar to other ungulates examined; 

most work concentrated on the influence of 

livestock on forage and did not directly examine 

population effects. Overall, cattle use of elk 

ranges had little influence on forage quality 

when stocked at 3.7 ha · AUM−1, but it did 

influence the quantity of forage available for elk 

(Dragt and Havstad 1987). Others (Wambolt 

et al. 1997) reported similar results when the 

nutritional values of forage were measured. 

Understanding forage use by wildlife and 

livestock is important for wildlife and livestock 

management. Most studies of elk and cattle 

interactions examined use of pastures under 

different conditions. Because of the varied 

management plans for livestock, managers 

should address multiple herbivore species 

in relation to environmental and climatic 

variation (Werner and Urness 1998). For 

example, in Utah, elk did not influence 

available forage for cattle in June and August 

1994, but use by cattle was greater in areas 

not used by elk in two of three rested pastures 

in June–August 1995. Cattle grazing reduced 

preferred winter elk forage in the initial 

growing season in Montana, but by the second 

season, the standing crop was similar to the 

ungrazed control (Jourdonnais and Bedunah 

1990). Intensive cattle grazing in Washington 

decreased elk use of ranges in 1 of 3 yr by 28% 

(Skovlin et al. 1983). 

Limited research has demonstrated how 

livestock grazing can improve elk forage 

and increase elk numbers. The Bridge Creek 

Wildlife Management area in northwestern 

Oregon was grazed by cattle without a 

270 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

prescribed grazing system and supported 120 

elk during winter over 13 yr. When a livestock 

grazing plan was initiated that incorporated 

rotational grazing, water distribution, properly 

located fences, salt placement, creation of a 

wildlife sanctuary, and closing roads, forage 

quality improved for elk and cattle, the elk 

population increased to nearly 1 200 animals, 

and AUM months for cattle grazing increased 

by 2.6 times (Anderson and Scherzinger 1975). 

In other studies, elk shifted habitats when 

cattle were introduced (Wallace and Krausman 

1987) and selected rested pastures over those 

used by cattle temporally (Frisina 1986; Yeo et 

al. 1993), even though fall cattle grazing and 

mowing (70% and 50% removal, respectively) 

can increase green vegetation the following 

spring (Frisina 1986; Short and Knight 2003; 

Taylor et al. 2004). 

Impacts to elk range from domestic sheep 

depend on climatic conditions and grazing 

intensity. The quality of browse may improve 

with moderate grazing of sheep (40% to 55%) 

that ends by June (Alpe et al. 1999). Others 

(Rhodes and Sharrow 1990) suggest that at a 

stocking rate of 125 to 143 female-days · ha−1, 

domestic sheep can improve forage quality in 

fall and forage quantity in spring. Carefully 

managed late-spring sheep grazing can improve 

winter forage quality on elk winter range (Clark 

et al. 2000). 

Bighorn Sheep. Ranges used by bighorn sheep 

and cattle usually do not overlap spatially, but 

interactions have been documented (Halloran 

and Blanchard 1950; King and Workman 

1984; Dodd and Brady 1986; Steinkamp 

1990). Early reports (Halloran and Blanchard 

1950) simply documented the occurrence of 

both animals, but later reports evaluated the 

relationships between them. Earlier studies 

of cattle and bighorn sheep (Spencer 1943; 

Halloran 1949; Matthews 1960; Arellano 

1961) did not demonstrate competition. 

Habitat preferences for steeper slopes by 

bighorn sheep and gentler slopes by cattle 

precluded competition because there was 

no range overlap. However, Barmore (1962) 

argued that cattle grazing on gentle slopes has 

precluded the use of those areas by bighorn 

sheep, and Bleich et al. (1997) cautioned 

that extensive use of such areas could affect 

forage availability for male bighorn sheep in 

particular. Blood (1961) examined competition 

between cattle and bighorn sheep in Canada, 

where 70% of bighorn sheep winter range was 

used by cattle. He concluded that cattle grazing 

prevented increases in the bighorn sheep 

population. 

King and Workman (1984) reported different 

associations between cattle and bighorn sheep 

in southeastern Utah. They reported bighorn 

sheep in higher, steeper, and more rugged 

talus slopes than cattle, which selected lower, 

gentler slopes and valleys close to roads and 

developed water sources. In addition, diets of 

the ungulates were different; cattle diets were 

dominated by grass, but bighorn sheep were 

browsers. King and Workman (1984) did not 

demonstrate that cattle and bighorn sheep 

competed for space or resources; however, they 

argued that the spatial separation they observed 

may result from a “social intolerance— 

avoidance factor.” McCann (1956), Barmore 

(1962), McCullough and Schneegas (1966), 

Follows (1969), Ferrier and Bradley (1970), 

Dean (1975), Wilson (1975), Gallizioli 

(1977), and Albrechtsen and Reese (1979) 

argued that bighorn sheep avoid areas used by 

cattle. Steinkamp (1990) demonstrated that 

a translocated population of bighorn sheep 

clearly avoided cattle. As cattle moved into 

core areas used by bighorn sheep, sheep moved 

away. Additionally, the closer cattle grazed to 

sheep, the closer sheep remained near escape 

cover. 

Social intolerance (Geist 1971) can have 

serious implications because cattle now graze 

most rangelands that historically supported 

bighorn sheep (Mackie 1978); 70% of public 

lands in the 11 most contiguous western states 

are grazed at least seasonally (US Department 

of the Interior 1986). Livestock grazing, 

even seasonally, appears to result in habitat 

fragmentation (Temple 1984), resulting in 

the exclusion of sheep from what is otherwise 

acceptable habitat. Bissonette and Steinkamp 

(1996) demonstrated that social intolerance 

can be a potent force influencing habitat use 

by sheep. Steinkamp’s (1990) and Bissonette 

and Steinkamp’s (1996) results pertain, 

however, to groups newly translocated into 

unoccupied habitat. Whether social intolerance 

between cattle and bighorn sheep is universal 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 271 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

remains equivocal. Resolution of the dispute 

is clouded by the almost universal disregard 

for spatial scale. Lack of consideration of scale 

effects can have profound implications for 

management. For example, in 1988–1989, the 

bighorn sheep population in Aravaipa Canyon, 

Arizona, was reduced by 52%. Mouton et al. 

(1991) examined the causes of mortality and 

concluded they were “probably the result of 

livestock related viral diseases compounded by 

nutritional stress.” Because range overlap has 

been documented to result in sheep mortality 

by disease transmission, determination of 

overlap and the scale at which it occurs is most 

important. Overlap at the level of home ranges 

may have very different consequences from 

overlap on specific slopes or valley floor areas. 

Additionally, temporal overlap at different scales 

(e.g., seasonal and annual) would appear to have 

important ramifications for management. 

In other areas of the Southwest, grazing by 

cattle has damaged bighorn sheep habitats 

(Gordon 1957; McColm 1963; Riegelhuth 

1965; Gallizioli 1977). Low precipitation levels 

ensure that recovery of ranges will take many 

years, and in some areas damage from livestock 

grazing may be irreversible. Grazing by cattle 

has also influenced bighorn sheep habitat in less 

arid areas (Buechner 1960; Crump 1971; Geist 

1971; Brown 1974) by converting grasslands to 

shrublands (Demarchi 1970). 

Bighorn sheep do not do well when they share 

ranges with cattle. Following the population 

declines of bighorn sheep of the late 1800s 

and early 1900s, they did not recover as well 

as other native ungulates (e.g., mule deer). 

Bighorn sheep are not as tolerant as other 

native North American ungulates to poor 

range conditions, intraspecific competition, 

overhunting, and habitat alteration. In addition 

they are much more susceptible to diseases 

of livestock than other rangeland wildlife, 

especially diseases of domestic sheep. 

Diseases of cattle that influence bighorn sheep 

are poorly documented, but diseases contacted 

from domestic sheep have played an important 

role in bighorn sheep mortality. Throughout 

the western United States, die-offs of bighorn 

sheep and population declines have occurred 

following the introduction of domestic sheep. 

Mortality was the result of competition for 

forage and space and shared diseases (Goodson 

1982). According to Goodson (1982), “Co-use 

of ranges by domestic and bighorn sheep has 

been consistently linked with declines, dieoffs, 

and extinctions of bighorn populations from 

historic to recent times. While much of the 

evidence for competition between domestic 

sheep and bighorn sheep is circumstantial, 

it is sufficiently strong to have prompted 

management decisions against co-use of 

ranges by bighorn and domestic sheep by 

federal land management agencies and state 

wildlife departments.” The Technical Staff of 

the Desert Bighorn Council (1990) reviewed 

24 interactions between bighorn sheep and 

domestic sheep and found that bighorn 

sheep died as a result of all interactions. 

Recent experimental studies confirmed 

field observations; when bighorn sheep are 

exposed to domestic sheep, bighorns die from 

Pasteurella haemolytica (Foreyt 1989, 1990, 

1992; Silflow et al. 1993; Foreyt et al. 1994). 

The actual mechanisms that kill bighorn sheep 

after they come in contact with domestic 

sheep are poorly documented (Jessup 1985), 

but two trends appear clear (Technical Staff of 

the Desert Bighorn Council 1990): 1) a large 

portion of the bighorn sheep population dies, 

and (2) domestic sheep do not suffer ill effects 

because of their contact with bighorn sheep. 

Bighorn sheep are more susceptible to diseases 

they share with livestock. Domestic animals 

have been selectively bred for disease resistance, 

but bighorn sheep have not evolved with 

resistance to the complement of diseases they 

are now exposed in the presence of domestic 

stock. As a result, they have not developed 

effective immunity against livestock diseases. 

Silflow et al. (1991) examined domestic 

and bighorn sheep and concluded that they 

had different control mechanisms for lung 

metabolism, and differences in the metabolites 

released led to different regulation of lung 

defense mechanisms. 

Disease. Biologists are not aware of all the 

factors creating negative interactions between 

domestic stock and bighorn sheep, but scabies, 

chronic frontal sinusitis, nematode parasites, 

pneumophilic bacteria, foot rot, parainfluenza 

III, bluetongue, sore mouth, paratuberculosis, 

and pinkeye are documented decimating 

factors to bighorn sheep (Jessup 1985). 

272 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	 	 	

	 	 	 	

	 	 	

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

Bighorn sheep have coexisted with humans 

for ≥ 30 000 years but now face a precarious 

future. They are an ecologically fragile species, 

adapted to habitats that are increasingly 

fragmented. Fragmentation of habitats 

increases when cattle share the same rangelands 

as bighorn sheep. Domestic sheep pose an even 

greater threat to bighorn sheep. 

small mammals, Reptiles, and 

Amphibians 

There are few studies that address specific 

conservation practices for small mammals, 

reptiles, and amphibians (i.e., access control, 

fences, closing mine shafts, and ponds). While 

limiting human access has positive effects on 

big game survival (Rowland et al. 2000), direct 

data are lacking for effects on small mammals, 

reptiles, and amphibians that are not directly 

harvested by humans. There is evidence that 

trampling caused by high amounts of human 

access (e.g., hiking and off-road vehicles) does 

affect the occurrence of small mammal species 

in montane (Liddle 1975) and urban habitats 

(Dickman and Doncaster 1987). Off-road 

vehicle use has been directly attributed to 

desert tortoise (Gopherus agassizi Cooper 1863) 

and Couch’s spadefoot (Scaphiopus couchi Baird 

1854) declines in California (Berry 1986). 

Such data suggest that controlling human use 

by limiting access may be effective in enhancing 

habitats for small mammals, reptiles, and 

amphibians. 

In rangelands, fences often provide added 

vegetative cover resulting from different 

microclimates and seed deposition by birds 

(Holthuijzen and Sharik 1985). There is some 

research examining the effects of fences on 

small mammals. Merriam and Lanoue (1990) 

used radiotelemetry and showed that whitefooted mice (Peromyscus leucopus Rafinesque 

1818) in farmlands preferentially traveled 

along fencelines, even when associated 

vegetative structure was less than 

1 m wide. 

A primary concern should be for bat maternity 

roosts and hibernacula. Mohr (1972) provided 

data on the importance of these cave resources 

for bats, and Jagnow (1998) reviewed an 

example of effective closure to restrict human 

access but to leave openings for ingress and 

egress by bats. 

Decreased vegetative cover at the edge of 

stock ponds resulting from cattle grazing 

was correlated with decreased abundance of 

Columbia spotted frogs (Rana luteiventris 

Thompson 1913) in Oregon (Bull and Hayes 

2000). Livestock effects on water quality 

were correlated with decreased larval diversity 

and abundance of amphibians in Tennessee 

(Schmutzer et al. 2008). However, the effect 

of cattle on terrestrial habitat quality and 

postmetamorphic survival of amphibians is yet 

to be quantified. 

Prescribed Grazing and upland 

Wildlife Habitat management 

Small Mammals. The greatest proportion 

of literature documenting effects of grazing 

on small mammals has focused on rangeland 

and riparian areas in the western United 

States. The density of aboveground biomass 

is important in structuring small mammal 

communities. Grant et al. (1982) indicated 

that small mammal communities and their 

response to grazing varied widely. Tallgrass 

communities tend to occur in areas of 

reliably high soil moisture and provide a 

high ratio of vegetation to seed with large 

accumulations of litter. These communities 

support highly variable populations of 

herbivorous litter-dwelling small mammals 

with high reproductive rates that can consume 

large amounts of vegetation. Grasslands of 

intermediate productivity have low biomass 

and low diversity of omnivorous and primarily 

surface-living small mammals, but both 

forage consumption and reproductive output 

are somewhat lower than in tallgrass prairie. 

Shortgrass prairie supports high biomass 

and high diversity of relatively long-lived 

omnivorous or granivorous species that 

reproduce opportunistically with precipitation 

and that use available resources (seeds and 

insects) intensively. Communities that differ 

in species composition, niches, and trophic 

dynamics are expected to differ in their 

responses to grazing. Land managers should 

anticipate that small mammals associated 

with herbaceous or shrub cover will decline 

when cattle remove this cover (Moulton 

et al. 1981; Giuliano and Homyack 2004; 

Johnston and Anthony 2008). Livestock 

grazing removes standing plant biomass but 

also prevents accumulation of ground litter 

that may influence small mammal community 

Black-tailed prairie dogs 

(Cynomys ludovicianus) are 

common throughout the Great 

Plains and grazing facilitates 

colony expansion. (Photo: 

David Briske) 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 273 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	

Rocky Mountain bighorn sheep 

(Ovis canadensis), Yellowstone 

National Park. (Photo: Jerod 

Merkle) 

composition, plant species growth, and 

seedling establishment via shading and 

changes in soil temperature and moisture 

(Fowler 1988). In southwestern grasslands and 

shrublands, grazing and fire result in rodent 

communities dominated by heteromyids 

(family Heteromyidae; pocket mice, kangaroo 

rats, and kangaroo mice) instead of murids 

(family Muridae; rats, mice, hamsters, voles, 

lemmings, and gerbils) on mesic sites. In more 

arid sites, grazing and fire favor kangaroo 

rats (Dipodomys spp.) over pocket mice 

(Perognathus spp.; Jones et al. 2003). 

Most studies demonstrating negative impacts 

on small mammal populations have attributed 

those effects to changes in vegetation cover 

and perceived predation risk (Grant et al. 

1982; Uresk and Bjugstad 1983; Heske and 

Campbell 1991; Hayward et al. 1997) or to 

long-term changes in plant species diversity 

(Jones and Longland 1999). However, Steen 

et al. (2005) provided evidence that forage 

competition occurs between livestock and 

voles, herbivores of greatly differing size. 

Grazing can either increase or decrease plant 

community heterogeneity (Adler et al. 2001). 

Detling (2006) provided the most extensive 

review of our state of knowledge concerning 

livestock and prairie dog interactions and 

concluded that we still cannot accurately 

determine the effect of prairie dogs on 

domestic livestock production. However, 

there is evidence that heavy livestock grazing 

can facilitate prairie dog colony expansion. 

Lomolino and Smith (2004) reported that 

prairie dog colonies had similar species 

richness of nonvolant mammals, reptiles, 

and amphibians as adjacent landscapes in 

Oklahoma but harbored different and more 

rare and imperiled species. Milchunas et al. 

(1998) suggested that livestock grazing impacts 

on other grassland herbivores may depend, in 

part, on temporally variable short-term tradeoffs between plant quantity and plant nutrient 

quality. Habitat productivity and herbivore 

densities may mediate shifts from facilitative 

to competitive interactions between differentsized herbivores (Krueger 1986; Cheng and 

Ritchie 2006). Field voles (Microtus agrestis L. 

1761) in Denmark showed a skewed quadratic 

response to grazing intensity (Schmidt et 

al. 2005) with population biomass and 

productivity at light to intermediate grazing 

intensity slightly greater than ungrazed and 

much greater than heavily grazed sites. Grazing 

274 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

on these sites reduced thick vegetative cover 

and promoted more nutritional regrowth, 

and this species of vole responded much the 

way livestock do. Steen et al. (2005) reported 

that field voles in Norway responded similarly 

but that bank voles (Clethrionomys glareolus 

Schreber 1780), whose diet differs, did not 

respond to sheep grazing. 

Reptiles and Amphibians. Kazmaier 

et al. (2001) detected no differences in 

survival or demography of Texas tortoises 

(Gopherus berlandieri Agassoz 1857) between 

moderately grazed (short-duration, winter– 

spring rotational grazing regime; 6–28 AUM 

d · ha−1 · yr−1) and ungrazed sites in the Western 

Rio Grande Plains, Texas. Brodie (2001) 

examined freshwater turtles across North 

America and suggested that increased siltation 

and soil compaction resulting from overgrazing 

in riparian areas could impact reproduction of 

freshwater turtles. 

Smith and Ballinger’s (2001) review indicated 

that lizards that sit and wait in open habitats 

(e.g., collared lizard [Crotaphytus collaris Say 

1823], lesser earless lizard [Holbrookia maculate 

Girad 1851], and side-blotched lizard [Uta 

stansburiana Baird and Girard 1852]) tend to 

be positively affected at the population level by 

livestock grazing, whereas active foragers that 

need vegetative cover (e.g., western whiptail 

[Cnemidophorus tigris Baird and Girard 1852], 

western stone gecko [Diplodactylus granariensis 

Starr 1879], fine faced gecko [Diplodactylus 

pulcher Steindachner 1870], desert spiny lizard 

[Sceloporus magister Hallowell 1854], bunch 

grass lizard [Sceloporus scalaris Weigmann 

1828], and Baja California bush lizard 

[Urosaurus nigricaudus Cops 1854]) tend to 

be negatively affected. Fair and Henke (1997) 

indicated that Texas horned lizards (Phrynosoma 

cornutum Harlan 1825) selected for burned 

plots and did not select for grazed plots in 

southern Texas. Lizard community composition 

in Arizona and desertified arid grasslands 

(Castellano and Valone 2006) was significantly 

different between inside and outside a grazing 

exclosure. Analysis of tail-break frequencies 

suggested that higher predation rates outside 

the exclosure may have contributed to increased 

abundance of eastern fence lizard (Sceloporus 

undulutes Bosc and Daudin 1801) and sideblotched lizards following livestock removal. 

In contrast, the round-tailed horned lizard 

(Phrynosoma modestum Girard 1852) was 

significantly less abundant inside the exclosure. 

Knutson et al. (2004) reported that small 

agricultural ponds in southeastern Minnesota 

provided breeding habitat for at least 10 species 

of amphibians. Gray et al. (2004) reported that 

relative abundance (i.e., average daily capture) 

of New Mexico and plains spadefoot toads 

(Spea multiplicata Cope 1863 and S. bombifrons 

Cope 1863) was greater at cropland than at 

grassland playas but that the abundance of 

other species and diversity of the amphibian 

assemblage was not affected by surrounding 

land use. However, Gray and Smith (2005) 

reported that mass and length of amphibians 

from playas surrounded by grasslands were 

greater than those from agricultural playas. 

They attributed this to altered hydroperiod 

in playas surrounded by agriculture. Body 

size is positively related to the probability of 

survival, reproduction, and evolutionary fitness 

in amphibians (Gray et al. 2004). Thus, if 

cultivation of landscapes surrounding wetlands 

negatively influences postmetamorphic body 

size of amphibians, restoration of native 

grasslands surrounding playa wetlands may 

help prevent local amphibian declines. 

Restoration and management of Rare 

or declining Habitats 

Manipulating riparian herbaceous cover and 

stream habitats are conservation practices that 

have influenced small mammals, reptiles, and 

amphibians. Endangered Columbia Basin 

pygmy rabbits (Brachylagus idahoensis Merriam 

1891) avoided grazed areas with fewer burrows 

than ungrazed areas (Thines et al. 2004). 

Grazing and mowing have been used effectively 

in specific cases to improve habitat for small 

mammal and reptile species that prefer reduced 

vegetative cover. Grazing reduced herbaceous 

and woody cover for the endangered Stephen’s 

kangaroo rat (Dipodomys stephensi Merriam 

1907) in California (Kelt et al. 2005) and 

reduced rhizomatous plant growth to facilitate 

burrowing while increasing sunning spots for 

threatened bog turtles (Clemmys muhlenbergii 

Schoepff 1801) in New Jersey (Tesauro 2007). 

Riparian Herbaceous Cover. Medin and 

Clary (1989) reported that, after 11 yr of 

grazing exclusion, small mammal biomass 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 275 
	 	 	 	

	 	

	 	

	 	

	 	 	 	

	 	

	 	 	

	 	 	

	 	

	 	

	 	 	

	 	

	 	

	 	 	

there are 

common research 

needs and 

recommendations 

that need to be 

considered if 

administrators, 

managers, 

biologists, 

and the public 

are to better 

understand how 

the conservation 

practices of 

NRCS apply to 

upland wildlife 

on western 

rangelands in the 

United States”

was three times greater on an ungrazed aspen 

(Populus tremuloides Mrchx.) and willow 

(Salix spp.) riparian site. Chapman and Ribic 

(2002) reported that ungrazed stream bank 

buffer strips supported more small mammals 

and species than similar grazed areas and that 

rotational grazing was not different from 

continuous grazing as applied to small mammal 

responses. Grazing in wet meadows can 

have indirect effects on small mammals also. 

Whitaker et al. (1983) reported that grounddwelling and fossorial invertebrates in diets of 

vagrant shrews (Sorex vagrans Baird 1857) were 

replaced primarily by volant species on grazed 

sites. Klaus et al. (1999) reported that grazing 

in Wyoming and Montana did not affect 

reproductive activity but did affect survival of 

young water voles (Microtus richardsonii DeKay 

1842), a species of management concern in 

alpine riparian habitats. Land managers should 

anticipate that small mammals associated with 

herbaceous or shrub cover in riparian areas will 

decline when cattle remove this cover (Moulton 

et al. 1981; Giuliano and Homyack 2004; 

Johnston and Anthony 2008). Frog community 

response to grazing intensity was positively 

correlated with grazing reduction of palustrine 

vegetation in an Australian floodplain (Jansen 

and Healey 2003). 

Stream Habitat Improvement and 

Management. Homyack and Giuliano (2002) 

reported that northern queen snakes (Regina 

septemvittata Say 1825) and eastern garter 

snakes (Thamnophis sirtalis L. 1758) were 

more abundant on fenced than unfenced 

stream banks but that most herptofauna may 

require longer than their 4-yr study to respond 

to exclusion from grazing. As with other 

conservation practices, little is available on how 

they influence wildlife populations. 

RecommendAtIons 

Little peer-reviewed research exists that 

examines the effects of conservation practices 

on habitat heterogeneity and diversity of 

wildlife. Most studies that we reviewed failed to 

collect pretreatment data, lacked experimental 

controls, had limited or no replication, or were 

too short in duration. Implications may often 

be extrapolated too broadly because results 

are frequently derived from studies of local 

management actions. 

Research needs and recommendations for 

the different groups of fauna vary. However, 

there are common research needs and 

recommendations that apply to all categories 

that need to be considered if administrators, 

land use planners and managers, biologists, 

and the public are to better understand how 

the conservation practices of NRCS apply to 

upland wildlife on western rangelands in the 

United States. 

1.	� Experimentally designed studies with 

replicates and controls are necessary. These 

studies need to be conducted so that 

scientifically reliable data can be collected. 

2.	� Studies have not been designed to 

understand how NRCS conservation 

practices apply to wildlife. This can be 

acquired only through targeted research. 

Specific studies should be designed 

to determine how specific NRCS 

conservation practices influence wildlife 

and the habitat they depend on, including 

(but not limited to) access control, access 

road, brush management, clearing and 

snagging, conservation cover, diversion, 

early successional habitat development/ 

management, fence, hedgerow planting, 

herbaceous weed control, land clearing, 

reclamation, mine shaft and adit closing, 

pond, range planting, restoration activities, 

spring development, tree and shrub 

establishment, and upland wildlife habitat 

management. 

3.	� Carnivore management is not an aspect of 

NRCS conservation practices for upland 

wildlife, but because of their role in the 

ecosystem, they need to be considered and 

managed. 

4.	� One common theme that is constantly 

emphasized in management theory is the 

importance of monitoring. Unfortunately, 

funds are not provided for these important 

activities. As a result, projects and practices 

are put in place, management plans are 

developed, and short-term research is 

conducted with little or no follow-up. 

This lack of efficient monitoring creates 

numerous information gaps that otherwise 

may have been filled. It is critical that 

monitoring be included in local, regional, 

and national management efforts so 

that the results of those efforts can be 

determined. 

276 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

	 	 	 	

	 	 	

	 	 	 	

	 	 	 	

	 	

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

KnoWledGe GAPs 

Game Birds 

1.	� Experimental evidence of grazing practices 

beneficial to game birds is largely lacking. 

Before–after control–impact field 

experiments are needed to determine 

widespread, relative effects of grazing 

treatments and stocking intensities on 

nesting success and female and chick 

survival (Beck et al. 2000). Investigations 

also are needed to evaluate effects of 

grazing, use levels, and stocking rates 

on abundances of important forbs and 

insects in brood-rearing habitat because 

these responses are poorly understood. 

Experiments should be well replicated and 

of a sufficient time to understand shortand long-term effects on populations. 

2.	� Similarly, investigations are needed to 

understand how to reduce and mitigate 

impacts of energy development and other 

significant sources of human disturbance 

over large landscapes as they relate to 

conservation practices. Recent studies 

show the large-scale and population-level 

impacts of oil and gas development on 

wildlife, including mule deer (Sawyer 

et al. 2009), sage-grouse (e.g., Walker 

et al. 2007), and songbirds (Ingelfinger 

and Anderson 2004). Wind energy will 

reduce our carbon footprint, but impacts 

to wildlife resulting from roads, noise, 

tall turbines, and additional power lines 

are poorly understood (e.g., lesser prairiechicken; Pruett et al. 2009). These studies 

also will require strong statistical designs 

that include treatments and controls at 

spatial and temporal scales relevant to 

landscape-scale impacts (Johnson and StLaurent 2011). 

3.	� A multitude of local-scale questions 

should be addressed as part of larger 

investigations. For example, we should 

determine whether the addition of 

anthropogenic water sources benefits 

quail (and other wildlife) populations 

in the desert Southwest (Western Quail 

Management Plan 2008) and whether 

mortality from fence collisions places a 

role in population dynamics, and, if so, 

we should develop recommendations on 

type and placement of fencing to reduce 

mortality (Wolfe et al. 2007). Studies 

should be conducted long enough to 

capture the short- and long-term influences 

that impact the practice being examined. 

Researchers should collaborate with 

management agencies to develop large and 

experimental projects as part of treatment 

projects planned by state and federal partners. 

In response, researchers and agencies can 

commit to monitoring at appropriate scales to 

evaluate treatment effects and to provide a basis 

for adaptive management. 

nongame Birds 

As noted above, few experimental studies have 

specifically evaluated the use of rangeland 

management to benefit nongame birds. Most 

efforts have been from the midwestern United 

States in the series of studies conducted by 

Herkert et al. (1996, 2003). The degree to 

which their results apply to western ecosystems 

is unknown. To date, most studies conducted 

in the West have consisted of “fence-line” 

observational studies whereby investigators 

compare adjoining pastures with and without 

cattle grazing. Questions concerning grazing 

regime, timing (both longevity and season of 

grazing), stocking levels, and related variables 

have yet to be addressed. To do so will require 

well-designed, replicated studies that can 

determine various sources of variation to 

understand cause–effect relationships. 

Bobcats (Lynx rufus) are a 

common rangeland predator 

that subsist primarily on 

rodents, rabbits, and birds. 

(Photo: Tim Fulbright) 

CHAPTER 6: An Assessment of Rangeland Activities on Wildlife Populations and Habitats 277 
	 	 	 	

 

carnivores 

Fruitful areas of research will include further 

evaluations of the role that top carnivores 

play in ecosystem structure and function 

(Hebbelwhite et al. 2005) and understanding 

the benefits or consequences of restoring those 

predators to historically occupied distributions. 

Additionally, better understanding of 

conditions that result in conflicts between 

humans and large carnivores (Wilson et al. 

2006) may provide opportunities to lessen 

conflicts in the future. Continued efforts to 

improve methods of reducing human–carnivore 

impacts and the implementation of those 

methodologies on rangelands is desirable and 

necessary to conserve large carnivores (Shivik 

2006). Further, responses of small carnivores 

to conservation practices should be explored 

more explicitly because of their importance 

as predators of ground-nesting birds; 

currently, much of the literature addresses 

risk of predation to avian species associated 

with rangeland management practices rather 

than demographic or habitat shifts in small 

carnivores that may result in those shifts in 

predation risk. 

ungulates 

Much of the peer-reviewed literature 

documents the influence of livestock and 

wildlife on range flora, but the studies are 

usually not replicated, are conducted on a small 

scale, and do not indicate how associations 

with livestock influence productivity and 

recruitment of wildlife. 

Additional research is needed to address 

these issues. In addition, because of the 

fragmentation of bighorn sheep habitat by 

livestock (Steinkamp 1990; Bissonette and 

Steinkamp 1996), social intolerance (Geist 

1971), and disease transmission (Jessup 1985), 

most researchers argue that prescribed livestock 

grazing should not occur in bighorn sheep 

habitat. To minimize avoidance of livestock by 

bighorn sheep and, hence, avoidance of habitat, 

livestock and bighorn sheep should not be close 

to each other (Steinkamp 1990; Bissonette 

and Steinkamp 1996). When separation is not 

possible, efforts should be made to minimize 

contact (e.g., placement of anthropogenic 

water sources or fencing critical areas), monitor 

distribution, monitor range conditions, and 

carefully watch for incidences of disease 

outbreaks (Goodson 1982; McCullough et al. 

1980; Technical Staff of the Desert Bighorn 

Sheep Council 1990). 

small mammals, Reptiles, and 

Amphibians 

Responses of small mammals, reptiles, and 

amphibians to grazing and other range 

management practices is species and often 

species-habitat specific. Few general trends 

have been identified, as studies have not 

been adequately designed to understand the 

underlying processes responsible because of the 

highly variable population dynamics of these 

groups of organisms and poor experimental 

designs (Johnson 1982). Experiments need to 

be of sufficient duration (perhaps on the order 

of decades in some ecosystems) and sufficient 

replication (over broad regional ranges) to 

isolate effects of interacting environmental 

factors that are usually not subject to 

experimental control from the effects of 

rangeland “treatments” (Rosenstock 1996). At 

least four avenues would assist in better data: 

1.	� Experimental evidence of conservation 

practices beneficial to small mammals, 

reptiles, and amphibians is largely lacking. 

Experiments designed with pre- and 

posttreatment data and controls are needed 

to determine relative effects of treatments 

on abundance and reproductive success of 

wildlife species. Experiments must include 

regional replications and be of sufficient 

duration to account for the variable nature 

of small animal populations to enable 

managers to understand short- and longterm population effects attributable to 

conservation practices at regional levels. 

2.	� Monitoring the distribution of various 

land uses in different landscapes (e.g., 

clumped or dispersed) and at what scale 

they occur are crucial for assessing longterm population persistence of small 

mammals, reptiles, and amphibians in 

fragmented landscapes. 

3.	� When examining the effects of a 

management practice, comprehensive 

analyses, including the impacts of 

type, frequency, timing, and extent of 

disturbances (e.g., mowing, burning, or 

grazing) of vegetation, are necessary to 

understand the species and species-site-

specific effects of such practices on species 

278 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

P. R. Krausman, V. C. Bleich, W. M. Block, D. E. Naugle, and M. C. Wallace 

abundance and reproductive success. 

4.	� Researchers should collaborate with 

management agencies to develop largescale, cost-effective experimental projects in 

an adaptive resource management strategy 

as part of conservation projects planned by 

state and federal partners. Commitments 

need to be made for monitoring at 

appropriate scales to evaluate treatment 

effects and to provide sound scientific data 

of sufficient scope and scale for assessing 

the true effects of conservation practices. 

conclusIons 

Very few of the 167 conservation practices 

listed by the NRCS have been evaluated in 

the peer-reviewed literature to determine 

their influence on upland wildlife. Activities 

associated with those conservation practices, 

particularly those efforts to enhance livestock 

production by limiting predation, have not 

been adequately investigated with respect to 

their overall impacts to rangeland ecosystems. 

Nevertheless, rangelands are important for 

protecting biodiversity, suggesting that future 

conservation efforts may require less reliance 

on reserves and a greater focus on private lands 

(Maestas et al. 2003). Grazing by livestock 

has received more attention in the literature 

than other conservation practices, but even 

then, studies often fail to distinguish between 

the different types, seasons, and intensities of 

grazing. Peer-reviewed literature evaluating 

how conservation practices influence upland 

wildlife habitat management has not received 

high priority, and their complex influences on 

wildlife and its habitat are largely unknown. 

Furthermore, other uses of rangelands (e.g., 

energy development) result in broad-scale loss 

and degradation of habitat that overwhelms 

other types of management (e.g., conservation 

practices) by increasing predation rates, 

promoting the spread of invasive plants, and 

facilitating disease transmission. However, 

the use of rangelands for sustainable livestock 

production has the potential to ensure the 

maintenance of wildlife habitat, especially 

when compared to energy development and 

urbanization, which will ensure that wildlife 

habitat will persist into the future. 

Studies will need to be designed as targeted 

research, with adequate replicates and controls, 

for outcome-based science if managers and 

scientists are to better understand how NRCS 

conservation practices influence wildlife on 

western rangelands. Future studies should also 

follow rigorous before–after control–impact 

designs, be implemented at the landscape level, 

and be conducted for a sufficient amount of 

time to understand how NRCS conservation 

practices influence ecosystem dynamics. 

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Zimmerman, J. L. 1997. Avian community 

responses to fire, grazing, and drought in 

the tallgrass prairie. In: F. L. Knopf and F. B. 

Samson [eds.]. Ecology and conservation of 

Great Plains vertebrates. New York, NY, USA: 

Springer-Verlag. p. 167–180. 

Zornes, M., and R. A. Bishop. 2008. Western 

Quail Management Plan. Washington, DC, 

USA. Wildlife Management Institute. http:// 

westernquail.org/sites/default/files/quailPlan.pdf. 

Accessed 2 December 2008. 

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CHAPTER 

7 

Invasive Plant Management 

on Anticipated Conservation Benefits: 

A Scientific Assessment 

Roger L. Sheley,1 Jeremy J. James,2 Mathew J. Rinella,3 

Dana Blumenthal,4and Joseph M. DiTomaso5 

Authors are 1Ecologist and 2Plant Physiologist, US Department of Agriculture–
 

Agricultural Research Service, Burns, OR 97720, USA; 3Rangeland Management
 

Specialist, US Department of Agriculture–Agricultural Research Service,
 

Miles City, MT 59301, USA; 4Ecologist, US Department of Agriculture–Agricultural
 

Research Service, Fort Collins, CO 80526, USA; and 5Weed Specialist,
 

University of California-Davis, Davis, CA 95616, USA.
 

Correspondence: Roger L. Sheley 67826-A Hwy 205, Burns, OR 97720; 

roger.sheley@oregonstate.edu. 

Reference to any commercial product or service is made with the understanding 

that no discrimination is intended and no endorsement by USDA is implied 

291
 
	 	 	 	

	 	 	

	 	 	

	 	 	 	

	 	 	

	 	 	 	

A major weakness 

in invasive plant 

“

management is our lack 

of knowledge about 

the efficacy of various 

prevention strategies.”

292 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	

 

 

 

	 	 	 	 	 	 	 	 	 	 		

	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

7 Invasive Plant Management on 

Anticipated Conservation Benefits: 

A Scientific Assessment 

Roger L. Sheley, Jeremy J. James, Mathew J. Rinella, Dana Blumenthal, 

and Joseph M. DiTomaso 

IntroduCtIon 

Invasive plant species have many negative 

impacts on rangelands throughout the world. 

Invasive plants can displace desirable species, 

alter ecological processes, reduce wildlife 

habitat, degrade riparian systems, and decrease 

productivity (DiTomaso 2000; Masters and 

Sheley 2001). Invasive plants are estimated to 

infest about 100 million ha in the United States 

(National Invasive Species Council 2001). 

Experts recognize invasive species are the 

second most important threat to biodiversity 

after habitat destruction (Pimm and Gilpin 

1989; Randall 1996; Wittenberg and Cock 

2001). Furthermore, Wilcove et al. (1998) 

estimate invasive species have contributed to 

the placement of 35% to 46% of the plants 

and animals on the federal endangered species 

list. In 1994, the impacts of invasive plant 

species in United States were estimated to be 

$13 billion per year (Westbrooks 1998). The 

amount of land infested by invasive plants 

is rapidly increasing (Westbrooks 1998) and 

subsequently the negative impacts of invasive 

plants are escalating. To address this issue, 

federal agencies and private land managers have 

developed and implemented integrated pest 

management (IPM) programs on rangeland. 

description of IPM on rangeland 

IPM is a long-standing, science-based, 

decision-making process that identifies and 

reduces risks from pests and pest management– 

related strategies (USDA Regional IPM 

Centers 2004). It was initially used to describe 

agricultural systems, but has since expanded 

to include wildlands and rangelands. IPM 

processes involve the coordinated use of pest 

biology, environmental information, and 

management technologies to prevent significant 

pest damage through economical means, while 

at the same time posing minimal risk to people, 

property, resources, and the environment. 

In recent years, invasive plant management 

has evolved to more frequently incorporate an 

IPM philosophy, as opposed to focusing on a 

single control option with little consideration 

of the ecosystem or the side effects of particular 

control methods. Although IPM approaches 

are not currently used in many regions, 

research has shown that integrating various 

combinations of control options can provide 

more effective control compared to a single 

option. 

IPM strategies have been used in rangelands 

for at least 20 yr, with interest in this approach 

greatly increasing over the last decade. 

Examples of more current IPM approaches 

include the combination of biological control 

agents (Lym and Nelson 2002; Nelson and 

Lym 2003; Wilson et al. 2004; DiTomaso 

2008; Joshi 2008), prescribed burning 

(DiTomaso et al. 2006a), grazing (Sheley et al. 

2004), mowing (Sheley et al. 2003; Renz and 

DiTomaso 2006), and revegetation (Enloe et 

al. 2005). Many of these integrated approaches 

combine nonchemical strategies with judicious 

use of herbicides. In most of these cases, the 

goal of the IPM approach was to establish a 

more desirable plant community that not only 

provides necessary ecosystem functions, but 

also provides some resistance to reinvasion, and 

thus, more effective long-term management of 

invasive plant species. 

description of Assumed Conservation 

Benefits of IPM on rangeland 

Specific plant species have been perceived as 

weeds since agriculture began about 10 000 

yr ago. Early agriculturalists used hoes and 

grubbing implements to control weeds for 

Invasive species have the 

capacity to create complete 

monocultures like the field of 

medusahead near Boise, Idaho 

pictured above. (Photo: Alex 

Boehm) 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 293 
	 	 	 	

 

 

the specific benefit of increased commodity 

production (Radosevich et al. 1997). During 

the 20th century, synthetic organic chemicals 

were extensively used to control invasive weeds 

in rangeland production systems, and natural 

enemies of invasive weeds have been used to 

reduce some undesirable plant populations 

below an economic threshold. The primary 

objective was to enhance grass production, 

while minimizing adverse ecological and 

human impacts of the control effort. 

Currently, ecologists and land managers 

recognize substantial adverse ecological, 

environmental, and economic impacts 

associated with invasive plants (Pimentel 

et al. 1999; DiTomaso 2000; Levine et al. 

2003). In response, they have designed morecomprehensive invasive plant management 

approaches in an attempt to achieve an array of 

benefits in addition to enhanced control and 

grass production. Invasive weed prevention 

strategies and programs are aimed at protecting 

noninfested rangeland. A major focus has 

been to manage invasive plants to establish 

and/or maintain a desired plant community, 

especially to promote restoration of natural 

plant communities. The assumed benefit of 

restoring desired vegetative cover is to create 

and maintain healthy functioning ecosystems 

that reduce reinvasion, protect soils, control 

erosion, reduce sediment, improve water 

quality and quantity, and enhance stream flow. 

Invasive plant management also aims to benefit 

biological diversity and wildlife through habitat 

improvement. 

oBjeCtIveS And APProACh 

The objective of this chapter is to provide a 

comprehensive assessment of the degree to 

which invasive plant management is achieving 

several commonly anticipated and desired 

benefits. We used a comprehensive review of 

peer-reviewed literature to assess the efficacy of 

various invasive plant management practices for 

each of nine conservation purposes developed 

for the Natural Resources Conservation Service 

(NRCS) conservation practice standard of 

herbaceous weed control. In contrast to the 

other conservation practice standards, this 

one was developed simultaneously with the 

Conservation Effects Assessment Program, so 

the stated purposes do not directly match those 

in the new standard. This new conservation 

practice standard is defined as the removal 

or control of herbaceous weeds including 

invasive, noxious, and prohibited plants. 

The writing team developed the following 

conservation purposes at the request of the 

NRCS for this chapter: 1) protect noninfested 

rangeland; 2) enhance quantity and quality 

of commodities; improve forage accessibility, 

quality and quantity for livestock; 3) control 

undesirable vegetation; 4) create a desired plant 

community; 5) change underlying causes of 

weed invasion; 6) restore desired vegetative 

cover to protect soils, control erosion, reduce 

sediment, improve water quality and quantity, 

and enhance stream flow; 7) maintain or 

enhance wildlife habitat including that 

associated with threatened and endangered 

species; 8) protect life and property from 

wildfire hazards; and 9) minimize negative 

impacts of pest control on soil resources, water 

resources, air resources, plant resources, and 

animal resources. The chapter also contains 

a section detailing recommendation and 

knowledge gaps, and conclusions addressing 

this conservation practice. 

ASSeSSMent of IntegrAted PeSt 

MAnAgeMent ConServAtIon 

PrACtICeS 

Protecting noninfested rangeland 

Invasive plant management has traditionally 

focused on controlling invasive plants on 

already-infested rangelands, with less emphasis 

placed on protecting noninfested rangeland 

by preventing invasions (Zalvaleta 2000; 

Peterson and Vieglasis 2001; Simberloff 2003). 

A proactive approach focused on systematic 

prevention and early control provides solid 

economic returns where, on average, every 

dollar spent on early intervention prevented 

$17 in later expenses (OTA 1993). The major 

components of invasive plant prevention 

programs include minimizing invasive plant 

introduction into noninfested areas (often 

through vector management), early detection 

and eradication of satellite patches, and 

increasing the resistance of desirable plant 

communities and soil systems to invasion 

(Davies and Sheley 2007). 

Minimizing Invasive Plant Introductions. A 

substantial amount of literature documents the 

294 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

fIgure 1. The potential fates and pathways of seed. 

modes of plant dispersal throughout the world 

(Riley 1930; Janzen 1982). National, regional, 

and local introductions of invasive plants can 

occur many different ways (Plummer and 

Keever 1963). The most successful methods for 

reducing introductions of invasive plants are to 

create a break or diversion, especially in shortand long-distance dispersal (Fig. 1; Davies and 

Sheley 2007). Identifying vectors that are major 

dispersers of an invasive plant species provides 

vital information necessary for interrupting 

dispersal to new areas (Wittenberg and Cock 

2001; Ruiz and Carlton 2003). 

Dispersal vectors for some invasive species are 

known (Selleck et al. 1962; Brown and Archer 

1987; Miller 1996; Kindschy 1998), but there 

is an obvious paucity of information about 

dispersal vectors. Regardless, Davies and Sheley 

(2007) provide a conceptual framework for 

preventing spatial dispersal of invasive plants. 

The framework identifies major potential 

vectors by incorporating invasive plant seed 

adaptations for dispersal through space and 

infestation locations relative to vector pathways 

(Fig. 2). Land managers can use the framework 

to guide efforts to limit dispersal of invasive 

plant seeds where it is possible. 

A major weakness in invasive plant 

management is our lack of knowledge about 

the efficacy of various prevention strategies. 

Tests of methods of preventing dispersal are 

extremely rare in the literature; however, most 

studies identifying dispersal vectors intuitively 

suggest a method to minimize these vectors. 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 295 
fIgure 2. Conceptual 

framework that identifies the 

major dispersal vectors of 

invasive plant species and 

the corresponding dispersal 

management strategies. 

For example, avoiding human or livestock 

contact with invasive species possessing hooks, 

barbs, and awns during seed production will 

likely help minimize dispersal (Agnew and Flux 

1970; Sorensen 1986). We could only find 

a single study that directly tested a potential 

prevention strategy. In that study, wind 

dispersal was limited by increasing neighboring 

vegetation height for species having large 

plumes (Davies and Sheley 2007). 

Early Detection and Rapid Control 

Response. One key to preventing new 

infestations is early detection of small patches 

that have a high probability of expanding into 

large infestations (Moody and Mack 1988). 

Early detection occurs at multiple levels of 

organization. The United States implements 

a national pest survey and detection program 

through the US Department of Agriculture– 

Animal and Plant Health Inspection Service. 

On the local level, early detection is difficult 

and requires educated and well-informed land 

managers, pest management specialists, and 

private land owners (Navaratnam and Catley 

1986). Systematic weeds surveys (Johnson 

1999), mapping based on sampling (Roberts 

et al. 2004), global positioning systems (Lass 

and Callihan 1993), and remote sensing 

(Steven 1993) have all been used to detect 

new infestations of invasive plants. In spite of 

the importance of detecting small infestations 

of invasive weeds, the cost, difficulty of 

implementation, and lack of reliable technology 

limit effective local early detection programs. 

It is critical that small patches be effectively 

eradicated quickly after they have been located 

(Zamora et al. 1989; Simberloff 2003). 

Eradication involves the destruction of every 

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R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

individual from an area (Newsome and Noble 

1986). Most eradication strategies include 

aggressively repeated monitoring and control 

procedures (Weiss 1999a). There is a paucity 

of successful plant eradications found in the 

scientific literature and most descriptions of 

eradication programs are published in non– 

peer-reviewed formats (Simberloff 2003). Very 

few successful examples exist in the literature. 

The first international conference on eradication 

reported about 15 plant species that were 

eradicated from various areas around the world 

(Simberloff 2001). Pokorny and KruegerMangold (2007) provide evidence that smallscale eradication is achievable by documenting 

the successful removal of dyer’s woad (Isatis 

tinctoria L.) from various counties in Montana. 

One biological tenet of successful eradication 

is that the infestation must be in the initial 

phases of invasion and only dominate a small 

area. Removing an invasive species is possible, 

especially for small infestations, but only under 

some circumstances and with potentially 

unpredictable results (Myers et al. 2000). 

Invasion-Resistant Plant Communities 

and Soil Systems. Promoting desired 

species is a critical component of invasive 

plant management, especially in an attempt 

to prevent invasions (Sheley et al. 1996). 

Researchers have shown that functionally 

diverse plant assemblages resist invasion better 

than less-diverse assemblages (Burke and Grime 

1996; Levine and D’Antonio 1999; Kennedy 

et al. 2002; Pokorny et al. 2005). Invasionresistant plant communities can be achieved 

by maximizing niche complementarity among 

desired species (Tilman et al. 1997; Brown 

1998; Carpinelli 2001; Fargione and Tilman 

2005; Funk et al. 2008). Furthermore, those 

plant communities that maximize biomass 

production also minimize invasion (Hooper 

and Vitsousek 1997; Anderson and Inouye 

2001). 

Strategies aimed at maintaining desired 

plant communities help protect noninfested 

rangeland. For example, prescribed fall burning 

of late-seral big sagebrush–bunchgrass plant 

communities stimulated the herbaceous 

component and increased the resistance of 

the communities to cheatgrass invasion 4 yr 

postburn (Davies et al. 2008). In another 

example, clipping to simulate grazing greatly 

reduced medusahead (Taeniatherum caputmedusae L.) by removing decadent material 

for desired plant species, which stimulated 

regrowth and enhanced competitive ability 

(Sheley et al. 2008). 

It has been proposed that invasion-resistant 

soils can be created by lowering plant-available 

nitrogen (Vasquez et al. 2008). Managing soilavailable nitrogen can be achieved by light to 

moderate levels of grazing. Grazing animals can 

remove nitrogen in plant material, making it 

unavailable to plants (Neff et al. 2005; Steffens 

et al. 2008). Mowing can remove nitrogen if 

the plant material is removed from the site 

(Oomes 1990; Moog et al. 2002). 

Conclusions and Management Implications. 

Protection of noninfested rangeland is 

central to the successful implementation of 

any integrated weed management program. 

Achieving the actual protection benefit is 

possible, but difficult, primarily because of 

a lack of effective techniques to interrupt 

dispersal vectors and our inability to detect 

new infestations before they become large 

infestations. Once found, small patches 

of invasive plants can be eradicated, but 

a comprehensive and intensive long-term 

eradication program must be employed. Many 

social, technological, and economic barriers 

exist that minimize the success of eradicating 

large infestations. The most scientifically 

developed strategy for protecting noninfested 

rangeland from invasion are those that convey 

some degree of invasion resistance to the 

plant community and possibly soils Managing 

desired plant communities to enhance the 

success of late-seral species, enhance diversity, 

and maximize productivity should help to 

minimize invasion and protect noninfested 

rangeland. 

enhance Quantity and Quality 

of Commodities; Improve forage 

Accessibility, Quality, and Quantity for 

Livestock 

The primary marketable commodities garnered 

from rangeland ecosystems are cattle and 

sheep, and, to a lesser extent, goats. Invasive 

plant management can influence quality and 

quantity of forage, as well as its accessibility. 

Consequently, the quantity and quality of 

livestock products can be impacted, but the 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 297 
	 	 	 	

	 	

	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	

Annual grasses significantly 

decrease forage capacity 

of rangeland because cattle 

typically avoid them once they 

begin to develop seedheads, 

which is usually late spring to 

early summer for cheatgrass. 

(Photo: Ryan Steineckert) 

direction and degree of impact varies for 

the three classes of livestock based on forage 

preferences and the plant functional groups 

being managed. In addition, the longevity of 

control impacts varies dramatically among 

invasive weed management strategies and their 

efficacy. 

Cattle. Most invasive weeds decrease forage 

production for cattle (Olson 1999). A 

substantial amount of literature shows an 

increase in forage as a response to invasive plant 

management over an untreated control, but 

experimental evidence showing a reduction 

in forage production with weed invasion is 

limited (except see Maron and Marler 2008). 

In addition to loss of forage, cattle tend to 

avoid areas with heavy infestations of weeds 

(Lym and Kirby 1987; Hein and Miller 1992). 

For example, leafy spurge (Euphorbia esula 

L.) reduces the carrying capacity of infested 

rangeland to near zero because cattle will not 

graze in areas with 10% to 20% cover of this 

weed. Few examples of overall economic costs 

of invasion have been published, but losses of 

forage for cattle on private land in California 

alone are estimated to be $7.65 million per 

year because of yellow starthistle (Centaurea 

solstitialis L.; Alison et al. 2007). 

Among 60 articles addressing invasive weed 

management and forage, 17 indicated 

an increase in forage quantity, quality, or 

accessibility. Increases occur where desired 

species are sufficiently abundant to respond to 

control procedures (Kedzie-Webb et al. 2002). 

Increases in perennial grass biomass ranged 

from 10% (Lym and Messersmith 1990) to 

1 935% (Masters et al. 1996) in response to 

weed control. Most commonly, weed control 

298 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

using herbicides increased forage for cattle 

about two- to threefold after 3 yr (Sheley 

et al. 2000). For example, picloram used to 

control spotted knapweed (Centaurea stoebe L.) 

increased grass yield by 1 513 kg · ha−1 for 2 yr. 

Nearly all studies were 1 yr to 3 yr in duration, 

with the period of invasive plant control being 

about 2 yr or 3 yr. Very little is known about 

the long-term forage production after a single 

herbicide application or a sustained control 

program. However, Rinella et al. (2009) found 

that leafy spurge had increased and grasses 

decreased in comparison to nontreated areas 17 

yr after picloram treatment. 

Effective biological control only exists for 

a small portion of the total invasive weed 

species. However, biological controls have 

increased forage quality and quantity as well 

as accessibility for cattle where they do exist. 

Huffaker and Kennett (1959) reported large 

increases in availability of grasses and forbs 

as cattle forage 10 yr after the release of 

natural enemies of St. Johnswort (Hypericum 

perforatum L.). Increased grass production has 

been reported after release of biological control 

agents (Rees et al. 1996). Longitarsus jacobaeae 

(Coleoptera: Chrysomelidae) has reduced 

the density of tansy ragwort (Senecio jacobaea 

L.) and reduced cattle losses to pyrrolizidine 

poisoning to near zero (McEvoy et al. 1993; 

Coombs et al. 1996). 

Sheep prefer grazing broadleaved plants and 

can be used to shift plant communities toward 

grasses that are preferred by cattle. Johnston 

and Peake (1960) used sheep to reduce leafy 

spurge basal area and increase the basal area 

of crested wheatgrass (Agropyron desertorum 

[Fischer ex Link] Shultes). Similarly, sheep 

grazing increased Idaho fescue (Festuca 

idahoensis Elmer) density and the frequency 

of Kentucky bluegrass (Poa pratensis L.), while 

reducing spotted knapweed (Olson et al. 1997). 

Livestock grazing has also been successfully 

used to reduce annual grasses growing among 

perennial grasses (Havstad 1994). 

Science and technology have not advanced 

to the point that reseeding desired species is 

consistently successful, but seeding desirable 

plants into invasive plant–infested rangeland 

can increase the quantity and quality of forage 

for cattle (Enloe et al. 2005; Sheley et al. 2005). 

In one successful case, applying clopyralid 

plus 2,4-D in combination with streambank 

wheatgrass (Elymus lanceolatus Scribn & Sm.) 

was used to reclaim a rangeland heavily infested 

by Russian knapweed (Acroptilon repens [L.] 

DC.) to a stand dominated by the sod-forming 

grass (Benz et al. 1999). More recently, methods 

for repairing damaged ecological processes 

have increased the success of revegetation 

across highly variable landscapes (Sheley et 

al. 2006, 2009). In these studies, specific 

processes in need of repair were identified 

and modified to foster vegetation dynamics 

toward favorable speciesIntegrated invasive 

plant management strategies also have the 

potential to improve the quantity and quality 

of rangeland for cattle. Among 100 randomly 

selected studies investigating integrated invasive 

weed management, 65 indicated short-term 

positive responses in the quality and quantity 

of forage for cattle. In a few cases, the response 

was synergistic in favoring desired vegetation 

composition (Sheley et al. 2004; Jacobs et al. 

2006). Additive and single main treatment 

effects were the dominant response, and in 

most cases nonnative grasses increased over 

nonnative target invasive weeds (Lym 1998; 

Endress et al. 2008). 

Sheep and Goats. Because sheep and goats 

consume comparatively more forbs than grasses 

in their diets, invasive plant management does 

not benefit these small ruminants (Lym and 

Kirby 1987; Kronberg and Walker 1993). 

Most broadleaved weeds contribute to the 

forage quantity and quality of sheep and goats 

(Olson and Lacey 1994). Although the nutrient 

content of broadleaved invasive weeds varies 

with phenology, most are highly nutritious 

(Bosworth et al. 1980, 1985). 

Conclusions and Management Implications. 

In general, the quantity and quality of cattle 

forage, and thus, cattle, are favored by weed 

management for a short period. Because many 

invasive weed management procedures increase 

forage yield for 2–4 yr, the benefits decrease 

with time following treatment. Sheep and goats 

prefer forbs as a major dietary component, 

and consume many weeds as quality forage. 

Broadleaved weed management has few 

positive benefits for sheep and goats. Invasive 

plant managers may maximize commodity 

production using multispecies grazing. 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 299 
	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	

Control undesirable vegetation 

Controlling undesirable vegetation on 

rangelands is difficult and rarely cost-effective. 

Compared to other land types, rangelands 

generate relatively low revenues per unit area. 

Typically, rangeland managers face expansive 

invasive plant infestations with few dollars 

for management. Additionally, invasive weeds 

tend to have high intrinsic growth rates and 

abundant seed production (Hobbs 1991; 

Rejmanek and Richardson 1996), which allow 

for rapid reinvasion of sites following use of 

herbicides, prescribed fire, and other invasive 

plant control practices (Lym and Messersmith 

1985b; DiTomaso et al. 2006b). Therefore, 

when invasive plants are successfully controlled, 

they often reoccupy the area very quickly. The 

keys to sustainably controlling large rangeland 

weed infestations are frequent use of strategies 

that provide inexpensive, short-term control, 

such as prescribed grazing or infrequent use 

of more expensive strategies that provide 

longer-term control, such as restoration. In this 

section, we review widely used weed control 

strategies with an emphasis on their short- and 

long-term effectiveness, as well as their costs. 

Prescribed Fire. Fire consumes weed standing 

crop, and in this sense fire consistently reduces 

undesirable vegetation. However, to have any 

lasting effect, prescribed fire must reduce the 

Prescribed fire in areas of 

invasion has varying effects 

on different species and can 

often lead to an increase in 

production of invasive species. 

(Photo: Clare Poulsen) 

production of biomass in subsequent growing 

seasons, reduce the existing year’s standing 

crop, and have a neutral or positive effect on 

desirable species. Some studies report increases 

in invasive weed biomass production due to fire 

(Young et al. 1972; Jacobs and Sheley 2003; 

Travnicek et al. 2005; Thacker et al. 2008), 

whereas others indicate decreases in biomass 

(Whisenant et al. 1984; DiTomaso et al. 1999). 

Whisenant et al. (1984) reported an extreme 

reduction of Japanese brome (Bromus arvensis 

L.) due to fire, which temporarily reduced 

this invasive annual grass by 85%. Conversely, 

Jacobs and Sheley (2003) found that fire more 

than doubled production of Dalmatian toadflax 

(Linaria dalmatica [L.] Mill. subsp. dalmatica). 

DiTomaso et al. (2006) concluded that the 

effects of fire depend, in part, on the weed’s 

life history strategy (i.e., annual, biennial, 

perennial) and characteristics of the fire 

regime. Fire-based invasive plant management 

is complex and often not predictable, and 

detailed studies are needed to identify effective 

fire regimes for particular species or similar 

groups of species. 

Applying prescribed fire is costly, so it is 

important for managers to carefully consider 

the longevity of control. Unfortunately, effects 

of fire on invasive plants are usually measured 

for only a year or two postburn. Cheatgrass 

(Bromus tectorum L.) has been measured for 

longer periods, but this weed does not appear 

amenable to fire-based control. On Western 

rangeland, cheatgrass tends to increase with 

fire because fire suppresses the less fire-tolerant 

competitors and available resources are rapidly 

acquired by cheatgrass (Young and Evans 1978; 

Vasquez et al. 2008). Cheatgrass invasion 

often increases fire frequency by increasing fine 

fuel loads, so burning cheatgrass can trigger a 

frequently repeated cycle whereby cheatgrass 

increases fire and fire increases cheatgrass 

(Knapp 1996). Conversely, prescribed fire 

that reduces invasive plants mainly destroys 

propagules, rather than by altering the 

environment in a manner that disfavors 

invasive weeds (DiTomaso et al. 2006a). When 

this is the case, weeds have only to replenish 

their propagule supplies to regain preburn 

abundances. Invasive plants tend to have high 

intrinsic growth rates which allow them to 

regain lost propagules quickly (Rejmanek and 

Richardson 1996; Grotkopp et al. 2002; Pyšek 

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R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

and Richardson 2007). Therefore, prescribed 

fire will tend to provide only short-term 

control, and managers would need to burn 

regularly enough to maintain control for the 

long term. Finally, as with cheatgrass, there 

are many herbaceous perennials that cannot 

be controlled by fire alone, but fire can kill 

large quantities of surface-deposited seeds 

(Vermeire and Rinella 2009), so integrating 

fire with strategies that kill established plants 

may enhance control of some perennial invasive 

plants. 

Herbicides. Herbicides are very useful for 

preventing small invasive plant infestations 

from producing seeds and spreading. They 

are also effective for controlling weeds during 

restoration projects so that seeded species have 

a better chance of establishing (Cione et al. 

2002; Huddleston and Young 2005). These 

uses aside, it is generally not cost-effective 

to control large invasive plant infestations 

with herbicides alone because the repeated 

applications required to maintain control 

(every 1–3 yr) are too expensive (Lym and 

Messersmith 1985b; Sheley et al. 1998; 

Young et al. 1998). Controlling rangeland 

invasive plants rarely increases forage 

production enough to offset the herbicide 

costs (Griffith and Lacey 1991; Bangsund 

et al. 1996). Furthermore, invasive annual 

grasses often proliferate after herbicides 

kill associated invasive forbs, so controlling 

invasive broadleaved forbs often just replaces 

undesirable forbs with undesirable annual 

grasses (Shinn and Thill 2003). 

Another problem with large-scale herbicide 

treatments is that they often kill associated 

native forbs and shrubs (Erickson et al. 2006; 

Sheley and Denny 2006; but see, Rice et 

al. 1997a). Whereas invasive weeds usually 

recover from herbicide quickly, a recent study 

of leafy spurge shows desired native plants 

can fail to recover from herbicides regardless 

of the length of the recovery period (Rinella 

et al. 2009). In that study, leafy spurge filled 

niches left vacant after herbicides removed 

native plants. Paradoxically, when herbicides 

damage native plants, invasive plants may 

ultimately become more abundant in response. 

Although herbicides play a critical role in 

weed prevention and restoration, the scientific 

literature causes us to question their use as 

stand-alone tools for controlling expansive 

invasive plant infestations. 

Prescribed Grazing. Prescribed grazing 

encourages the targeted use of invasive plants 

by manipulating timing, intensity, and 

frequency of herbivory and selecting animal 

classes based on their dietary preferences. For 

example, goats prefer trees and shrubs and forbs 

compared to grasses, so they are sometimes 

stocked on grasslands invaded by pines and 

junipers or invasive forbs, such as knapweed 

(Campbell et al. 2007). 

The key difference between prescribed grazing 

and other invasive plant management strategies 

is that it can be affordably used on an annual 

basis to reduce invasive plant standing crop and 

biomass production in many situations. Sheep 

and goats can be economically profitable in 

well-managed operations or they can serve as 

additional revenue sources in cattle operations 

(Williams et al. 1996; Bangsund et al. 2001). 

Additionally, using sheep or goats to graze and 

reduce exotic forb standing crop can increase 

the amount of forage accessible to cattle and 

increase overall forage utilization (Lym and 

Kirby 1987). 

Desirable rangeland species generally increase 

after land management practices reduce 

invasive plant biomass (Lym and Messersmith 

1985a; Belcher and Wilson 1989; Sheley et al. 

2000). As a consequence, prescribed grazing is 

more worthwhile when it reduces subsequent 

invasive plant biomass in addition to reducing 

standing vegetation. Invasive plant biomass 

responses to grazing depend on the timing, 

intensity, and frequency of grazing, as well 

as the class of livestock. For example, three 

studies reported no effect of sheep grazing on 

leafy spurge production because most grazing 

occurred during and after leafy spurge seed 

production, and at only a single time during 

the season (Lacey and Sheley 1996; Olson 

and Wallander 1998; Seefeldt et al. 2007). 

Alternatively, four other studies reported fairly 

consistent declines in leafy spurge over time, 

and in these studies, grazing occurred multiple 

times prior to seed production (Johnston and 

Peake 1960; Lym et al. 1997; Jacobs et al. 

2006; Rinella and Hileman 2009). Controlling 

invasive plants using livestock requires the 

development of relatively complicated, 

Although herbicides play a critical role in weed 

prevention and 

restoration, the 

scientific literature 

causes us to question their use as 

stand-alone tools 

for controlling 

expansive invasive plant infestations.”

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 301 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

	 	 	

	 	 	 	

	 	 	

	 	

Prescribed grazing and 

biological control can be 

effective treatments when 

implemented into an integrated 

management program. (Photo: 

Sharon Bingham) 

strategically designed strategies for each species 

based on their tolerance and/or resistance to 

herbivory. Detailed investigations are needed 

to identify prescribed grazing strategies that 

are effective for specific invasive species in 

particular environments. 

Biological Control. Biological control agents 

can clearly damage individual plants (Pecinar et 

al. 2007; Thomas and Reid 2007; Zalucki et al. 

2007), but unfortunately, these effects often fail 

to cause appreciable reductions in undesirable 

vegetation (DeLoach 1991). Many agents are 

released in hopes that one or a combination of 

them will prove effective (McEvoy and Coombs 

1999). Unfortunately, this lottery approach is 

rarely effective in controlling large populations. 

Many invasive plants remain highly problematic 

despite being the target of many releases of 

biological control agents for decades (Zalucki 

et al. 2007; Story et al. 2008). Furthermore, 

the risks of deleterious off-target effects increase 

with the number of releases (Louda et al. 2005; 

Pearson and Callaway 2005). 

Risks and failures aside, biological control is 

occasionally extraordinarily successful against 

invasive weeds. The most-cited examples 

include two introduced beetles that reduced 

St. Johnswort density by greater than 99% 

in much of its introduced range (Harris and 

Maw 1984), and three insects that substantially 

reduced ragwort in western Oregon (McEvoy 

et al. 1991; Denslow and D’Antonio 2005). 

Although these are the best-studied examples, 

other rangeland weed species have been 

targeted, with quite varied results, including 

recent introductions that show promise for 

302 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

controlling salt cedar (Tamarix ramosissima 

Ledeb.; Hudgeons et al. 2007). In those few 

special cases, biological control has the unique 

benefit of providing relatively inexpensive 

partial weed control over expansive areas for 

indefinite periods of time. 

Mechanical Control and Seeding. 

Mechanical methods of herbaceous weed 

control include tillage and mowing. Mechanical 

control treatments of tillage and mowing can 

cause substantial reductions in invasive plant 

standing vegetation, but they are only truly 

effective when future biomass production is 

reduced. Invasive plant responses to mowing 

have been mixed, with some studies reporting 

appreciable decreases in weed biomass 

production (Benefield et al. 1999; Rinella et al. 

2001) and other studies reporting no detectable 

change (Benz et al. 1999; Renz and DiTomaso 

1999). Collectively, studies indicate the 

responses of undesirable vegetation to mowing 

depend on species, timing of mowing, and 

other factors. Finally, because invasive weeds 

quickly recover after mowing is discontinued, 

mowing must occur frequently to provide 

continuous control. 

Tillage alone can be used to control invasive 

plants on rangeland under some circumstances. 

In one study, tillage alone provided no 

sustained control of perennial pepperweed 

(Lepidium latifolium L.; Young et al. 1998), 

and in another study, repeated tillage prior to 

a serious frost reduced leafy spurge well (Lym 

and Messersmith 1993). Invasive weeds tend to 

recover quickly when tillage is discontinued. 

In addition to controlling invasive plants, 

tillage provides safe sites for seeded species, 

and some studies report that competition 

from seeded species has partially controlled 

undesirable vegetation (Lym and Tober 1997; 

Bottoms and Whitson 1998; Sheley et al. 

2001; Thompson et al. 2006). However, other 

studies have reported that seeded species 

provided no weed control (Sheley et al. 1999; 

Mangold et al. 2007). In the latter case, it is 

possible data were collected before the seeded 

species grew large enough to compete with 

the reemerging invasive plants. Theoretically, 

when seeded species develop self-sustaining 

populations, these populations should suppress 

undesirable vegetation indefinitely through 

resource competition. Therefore, despite the 

lack of evidence, there are likely to be distinct 

advantages to integrating seeding with other 

practices that provide only short-term control, 

such as herbicides and tillage. 

Conclusions and Management Implications. 

Many questions remain regarding control of 

invasive plants, and many of these questions 

pertain to inconsistencies in the responses of 

undesirable vegetation to various controls. For 

example, there are cases in which individual 

treatments, such as herbicides, grazing, or 

fire have reduced invasive weeds. There are 

also cases in which treatments have failed to 

alter the abundance of invasive plants, and 

there is even some evidence that invasive weed 

control occasionally increases weed species. 

Furthermore, except for a few situations 

using biological control or annual repeated 

grazing, it is often questionable whether or not 

individual invasive weed control strategies are 

worthwhile because the control they provide is 

so ephemeral and expensive. 

Presumably, integrating multiple control 

strategies should lead to more consistent, 

longer-lasting suppression of invasive plants. 

But integrated strategies are more costly, and 

there is still no guarantee that the level and 

longevity of invasive plant control will be 

satisfactory (Sheley et al. 2001; Lym 2005). 

Much research is needed to identify affordable, 

consistently effective strategies for controlling 

undesirable vegetation. 

Create a desired Plant Community 

On most sites, the species that invasive 

plants suppress or displace comprise both 

nonweedy exotic species and natives (Enloe 

et al. 2007). Some of the suppressed natives 

and nonnatives are often valuable forage 

plants, and increased forage production often 

provides the impetus for controlling invaders 

(Lym and Messersmith 1985a). In many cases, 

controlling undesired species does not lead to 

a desired plant community. A variety of other 

objectives may also be met by creating desired 

plant communities including increasing native 

species diversity, increasing habitat for wildlife, 

improving soil and water quality, and reducing 

reinvasion. In this section, we investigate how 

desired, especially native, species respond to 

invasive weed control and examine efforts 

A desired plant community 

should include a variety of 

species that fill multiple niches 

which reduces the likelihood 

of reinvasion. (Photo: Brett 

Bingham) 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 303 
	 	 	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	 	

	 	

	 	

	 	

These studies 

suggest that 

prescribed 

grazing may 

have potential 

for restoring 

desired species, 

but invaded 

communities are 

likely to quickly 

regress to their 

pregrazing 

weedy state 

when prescribed 

grazing is 

discontinued.”

to reestablish desired species from seeds. 

The choice of desired species depends upon 

management goals. In this document, the goal 

is considered to be to establish and maintain a 

healthy, functioning plant community that is 

resistant to invasion and meets other land use 

objectives (Sheley et al. 1996). 

Prescribed Grazing. To our knowledge, only 

two studies have provided detailed assessments 

of plant community responses to prescribed 

grazing of invasive plant–infested rangeland. 

Olson and Wallander (1998) studied responses 

of a leafy spurge–infested plant community 

to prescribed sheep grazing. The authors 

concluded that grazing reduced leafy spurge 

stem height without affecting stem density, 

so grazing presumably lowered leafy spurge 

biomass production. Grazing increased the 

density and frequency of several native (Idaho 

fescue, western wheatgrass [Pascopyrum smithii 

{Rydb.} A. Löve], Sandberg bluegrass [Poa 

secunda J. Presl]) and nonnative (Kentucky 

bluegrass, annual bromes) grasses, and 

decreased density of a nonnative dandelion 

(Taraxacum officinale Weber). In a similar 

study, on spotted knapweed-infested rangeland, 

Olson and Wallander (1997) found that sheep 

grazing reduced spotted knapweed rosette and 

adult plant density. As in the leafy spurge study, 

grazing increased density and frequency of 

native Idaho fescue and nonnative Kentucky 

bluegrass. The native forb arrowleaf balsamroot 

(Balsamorhiza sagittata [Pursh] Nutt.) was not 

influenced by grazing. These studies suggest 

that prescribed grazing may have potential 

for restoring desired species, but invaded 

communities are likely to quickly regress to 

their pregrazing weedy state when prescribed 

grazing is discontinued. To be successful, 

prescribed grazing will likely need to be 

carried out indefinitely. It is unfortunate that 

so few studies have evaluated native species 

responses to prescribed grazing of weed-infested 

rangeland. 

Biological Control. Denslow and D’Antonio’s 

(2005) review of the literature clearly 

demonstrates that successful biological control 

of rangeland invaders can, but does not 

always, have positive effects on suppressed 

desired species. A classic successful example 

is the control of St. Johnswort by the leafbeetles Chrysolina quadrigemina Suffrian 

and Chrysolina hyperici Forster in California 

rangeland. Within 5–10 yr of leaf-beetle 

introduction, St. Johnswort had virtually 

disappeared from sites in several California 

counties (Huffaker and Kennett 1959). It 

was reduced to less than 1% of its prerelease 

cover, and replaced by a combination of native 

and exotic grasses, which greatly increased 

available forage. Similarly, a combination of 

several insects reduced ragwort to a fraction 

of its former abundance in several California 

sites (Pemberton and Turner 1990), and 

throughout western Oregon (McEvoy et al. 

1991). In both situations, control agents 

persisted, ragwort remained under control 

for more than a decade, and desired species 

responded favorably. More recently, a suite 

of biological control agents led to successful 

control of diffuse knapweed (Centaurea diffusa 

Lam.) at sites in Colorado, Montana, Oregon, 

Washington, and British Colombia (Myers 

2004; Smith 2004; Seastedt et al. 2007), 

and successful or partial control of spotted 

knapweed in sites in Colorado and Montana 

with some responses by desired species (Story et 

al. 2006; Seastedt et al. 2007). Similarly, several 

flea beetles (Apthona spp.) have displayed 

variable, but sometimes quite successful 

control of leafy spurge and subsequent release 

of desired species in South Dakota, North 

Dakota, and Montana (Larson and Grace 

2004; Butler et al. 2006; Cornett et al. 2006). 

Additionally, Lesica and Hanna (2004) provide 

an example of positive native plant community 

responses to biological control. 

Herbicides. Rangeland herbicides tend to 

selectively kill either grasses or forbs. Therefore, 

native grasses are typically not damaged by 

the herbicide used to control invasive forbs. 

In fact, many native grasses increase following 

herbicide control of invasive forbs (Sheley et 

al. 2000; Laufenberg et al. 2005; Sheley and 

Denny 2006). Similarly, native forbs often 

increase after herbicides kill invasive grasses 

(Cione et al. 2002; Wilcox et al. 2007). 

Herbicides are sometimes used to control 

invasive grasses even though herbicide-sensitive 

native grasses are present (Kyser et al. 2007). 

Likewise, herbicides are used to control invasive 

forbs growing with native forbs and shrubs 

(Fuhlendorf et al. 2002). Sometimes native 

species escape extensive damage by herbicides 

304 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	

	 	

	 	

	 	

	 	

	 	 	 	

	 	 	

	 	 	

	 	

	 	

	 	

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

or are able to quickly recover from damage. 

For example, when the broadleaf herbicide 

picloram was applied to spotted knapweed– 

infested rangeland during the summer-dormant 

period of most native forbs, Rice et al. (1997b) 

found the herbicide had only mild transient 

effects on native forbs. Similarly, Erickson et 

al. (2006) found that herbicide control of leafy 

spurge with quinclorac did not damage the 

threatened prairie fringed orchid, although 

imazapic damaged the orchid. Also, Simmons 

et al. (2007) found the “nonselective” herbicide 

glyphosate provided substantial short-term 

control of an invasive grass while damaging 

native grasses to a lesser extent or not at all. 

Finally, Barnes (2007) found that grass-specific 

herbicides promoted native warm-season 

grasses by reducing abundance of the exotic 

grass, tall fescue. 

In contrast to these examples, there are other 

cases demonstrating extensive herbicide damage 

to natives. For example, Sheley and Denny 

(2006) concluded that any of three herbicides 

used to control sulfur cinquefoil (Potentilla 

recta L.) continued to suppress native forbs 

2 yr after application. Similarly, Shinn and 

Thill (2004) found that imazapic, a herbicide 

that is active against invasive annual grasses, 

substantially injured native perennial grasses as 

well. Herbicide damage to native species would 

not be a problem if the natives consistently 

recovered, but a recent study showed that 

herbicide control of rangeland weeds can 

pose very serious long-term threats to native 

forb populations (Rinella et al. 2009). In 

addition to sometimes extensively damaging 

native species, herbicides are expensive and 

they generally provide only short-term weed 

control. Therefore, herbicides alone are unlikely 

to create desired native plant communities. 

However, herbicides are critical in preventing 

spread of small weed infestations and in 

integrated weed management. 

Prescribed Fire. Prescribed fire can boost 

native species and reduce populations of annual 

invaders by consuming their seeds (DiTomaso 

et al. 2006b). For example, Harmoney (2007) 

found that fire greatly reduced Japanese brome 

and increased two native grasses above an 

unburned control. Also, fire reduced invasive 

yellow starthistle and three invasive annual 

grasses while greatly increasing diversity and 

species richness of native forbs (Hastings and 

DiTomaso 1996; DiTomaso et al. 1999). 

However, depending on fire timing, species 

identity, and other factors, fire can also boost 

invaders (Young et al. 1972; Jacobs and Sheley 

2003; Travnicek et al. 2005; Thacker et al. 

2008) and cause severe damage to native 

populations, such as big sagebrush (Artemisia 

tridentata Nutt.) growing with cheatgrass 

(Young and Evans 1978; Knapp 1996). 

To our knowledge, no studies have measured 

long-term invasive weed and native species 

responses to prescribed burning. Unfortunately, 

favorable responses likely will be short-lived. 

Weeds generally reinvade very quickly following 

fire (Young and Evans 1978), and native plants 

are likely to revert to their suppressed preburn 

state following reinvasion. It is unlikely that 

managers can burn vegetation regularly enough 

to maintain native populations, but is possible 

to integrate fire with other strategies in hopes 

of providing longer-term restoration of native 

species (MacDonald et al. 2007). 

Mowing. A small number of studies have 

evaluated desired plant responses to mowing 

of invasive weed–infested rangeland. Wilson 

and Clark (2001) found mowing invasive grassinfested rangeland for 4 yr greatly restored 

native prairie grasses. Similarly, MacDougall 

and Turkington (2007) found that 5 yr of 

mowing at the time of invasive grass flowering 

shifted the plant community toward desired 

forbs and grasses. In contrast to these successes, 

Simmons et al. (2007) found mowing had little 

or no detectible effect on an invasive perennial 

grass and native species, and Brandon et al. 

(2004) reported mowing increased invasive 

forb abundances. Collectively, these studies 

indicate native species responses to mowing 

depend on the relative susceptibility of desired 

species and invaders to different timings, 

heights, and frequencies of defoliation as well as 

other factors. Finally, when beneficial mowing 

regimes are identified, they will likely have to 

be carried out indefinitely or combined in an 

integrated management strategy to maintain 

native species because invasive weeds tend to 

recover quickly when mowing is discontinued. 

Seeding. In revegetation projects, the 

desired plant community is in large part 

dictated by the species in the seed mix. A few 

It is unlikely that 

managers can 

burn vegetation 

regularly enough 

to maintain native populations, 

but is possible to 

integrate fire with 

other strategies in 

hopes of providing longer-term 

restoration of native species.”

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 305 
	 	 	 	

	 	 	

	 	

	 	

	 	

	 	

	 	

It has become 

clear that 

nonnative grasses 

have tended 

to outperform 

native grasses 

in revegetation 

studies.”

revegetation efforts have resulted in fairly 

diverse assemblages with impressive stability 

(Blumenthal et al. 2003; Martin et al. 2005; 

Martin and Wilsey 2006; Foster et al. 2007). 

However, these successes were obtained from 

areas with relatively innocuous weed species, 

not highly tenacious rangeland invaders such as 

leafy spurge and spotted knapweed. 

Where invaders dominate rangelands, managers 

often strive for the humble goal of introducing 

one or a few desired or native species in hopes 

they will provide forage, wildlife habitat, and 

other values. Seeding is very expensive, so 

successful seedling establishment is critical. 

Seedling establishment is most effective when 

seeding has been integrated with herbicides 

and/or tillage (Lym and Tober 1997; Ferrell 

et al. 1998; Whitson and Koch 1998; Masters 

et al. 2001; Huddleston and Young 2005; 

Simmons 2005), but establishment failures 

are common (Lym and Tober 1997; Masters 

et al. 2001; Sheley et al. 2001; Wilson et al. 

2004). Seedling establishment depends on 

a myriad of factors, not the least important 

of which is interannual climatic variation 

(MacDougall et al. 2008), which cannot be 

controlled or effectively predicted. Beyond 

establishment, seeded species must maintain 

viable populations over the long term for 

seeding to be worthwhile. Unfortunately, 

desired native species are rarely measured 

for more than a few years after sowing into 

invaded rangelands. Among the longer-term 

investigations, Bottoms and Whitson (1998) 

found that seeded thickspike wheatgrass 

(Elymus lanceolatus [Scribn. & J.G. Sm.] Gould) 

and western wheatgrass maintained healthy 

stands that suppressed Russian knapweed 5 

yr after seeding. Similarly, Ferrell et al. (1998) 

found that seeded wheatgrasses produced 

large quantities of biomass and continued 

suppressing leafy spurge 6 yr after seeding. 

Also, a mixture of four native warm-season 

grasses contributed substantially to biomass 

production 6 yr after being sown into spotted 

knapweed–infested plots (MacDonald et al. 

2007). To our knowledge, Ferrell et al. (1998) 

document the longest-term measurements 

of seeded species in invasive plant–infested 

rangeland. The authors found that two grass 

species remained fairly abundant and partially 

suppressed leafy spurge 10 yr after seeding. 

It has become clear that nonnative grasses 

have tended to outperform native grasses in 

revegetation studies (Ferrell et al. 1998; Asay et 

al. 2001; Sheley et al. 2001). 

Given the expense of seeding rangelands, the 

lack of long-term measurements is troubling. 

Future restoration research should focus on 

determining whether or not native species can 

persist with invasive species beyond a few years 

after seeding. To some extent, restoration efforts 

are predicated on the assumption that improper 

land management causes weed invasions. If 

proper management does not prevent invaders 

from dominating the original community, 

then we should not expect proper management 

to prevent invaders from dominating the 

restored community. It is sobering to consider 

that rangeland restoration may be doomed 

to fail over the long term wherever invaders 

have displaced natives despite good range 

management. 

Conclusions and Management Implications. 

Many studies provide no information on 

desired or native species responses to weed 

management, and a few others provide only 

cursory information. Biological control is 

relatively inexpensive to implement once 

developed, and a few biological control 

programs have restored natives to an impressive 

extent. However, biological control sometimes 

fails completely, and it has many risks. More 

research is needed to elucidate the risks and 

benefits of biocontrol. Desired and native 

species responses to prescribed grazing have 

been limited, but two studies indicate that 

annually applied prescribed grazing can shift 

plant communities toward a desired state. 

Herbicides often control target invaders 

while allowing associated species to increase 

in abundance. However, herbicides are quite 

expensive and herbicides alone provide only 

short-term weed control. Furthermore, 

herbicides pose considerable risks to some 

desired species. Herbicides are useful for 

preventing spread of weed infestation, and 

thus helpful in maintaining a desired plant 

community that has not been invaded. 

Prescribed fire can be beneficial to desired 

species, but it can also harm natives and 

increase invasion. Effective prescribed fire 

regimes will likely need to be repeated regularly, 

and can be used occasionally to restore desired 

plant communities prior to invasion to help 

306 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

keep them resistant. Effective mowing will 

probably need to be carried out often to be 

successful, and in some cases, the plant material 

may need to be removed. A few studies have 

reported that seeded native species remained 

abundant and suppressed tenacious rangeland 

invaders 5 yr or 6 yr after seeding. This is 

promising, but longer-term measurements are 

desperately needed to determine if the benefits 

of seeding warrant its high costs. 

Change underlying Causes of Weed 

Invasion 

The benefits of invasive plant control depend in 

large part on the longevity of control and the 

resulting desired plant community. In turn, this 

depends on the ecological causes underlying 

the original invasion and our ability to alter 

those causes in favor of desired species. Because 

invasive species are rarely eradicated, if the 

original causes of invasion are not repaired, 

reinvasion is likely (Sheley and KruegerMangold 2003). Furthermore, given the large 

areas and low economic returns per unit area 

typical of rangeland, temporary invasive weed 

control is rarely economically sustainable. 

In this section, we consider the durability of 

invasive species control and desired community 

restoration. 

Durability Depends on Original Cause of 

Invasion. Invasion is often the result of changes 

to an ecosystem that inhibit native species, 

and thereby reduce the competition faced by 

invasive species (Facon et al. 2006). Perhaps 

the most important barrier to invasion is the 

presence of desired species. Desired species 

garner much of the water, nitrogen, light, 

and other resources that would otherwise be 

available to invaders. This resistance to invasion 

is often described as “biotic resistance” (Maron 

and Vila 2001; D’Antonio and Thomsen 

2004; Levine et al. 2004), and depends a great 

deal on environmental conditions (Shea and 

Chesson 2002). Because native species are, by 

definition, adapted to historical environmental 

conditions (Landres et al. 1999), changes to 

these conditions are likely to make them less 

well-adapted, and less able to resist invasion. 

Consequently, changes in environmental 

conditions appear to be a common cause of 

invasion (Daehler 2003; Facon et al. 2006). 

These changes can be dramatic, such as soil 

tillage or improper grazing, or subtle, such 

as nitrogen deposition or loss (Vasquez et al. 

2008). Where such changes underlie invasion, 

the key question facing managers is whether 

the change can be reversed. Among the many 

underlying causes of invasion, managers have 

had the most success reversing the following 

three: past disturbances, reversed via successful 

restoration; returning grazing to systems, and 

enemy release, through biological control. 

Restoration as a Long-Term Solution to 

Previous Novel Disturbances. A key question 

for determining the likelihood of long-term 

invasive plant control is the degree to which 

a particular invasion is caused by a novel 

disturbance—a disturbance to which native 

species are not well adapted. Where such 

disturbances underlie the invasion, and can 

be prevented in the future, it is much more 

likely that a stable, desired plant community 

fIgure 3. a Exotic annual 

species (●) are less effective than native perennial 

grasses ( ) at extracting 

soil nitrate at all depths in 

May. b, Monthly sampling 

was conducted at the 0- to 

15-cm depth. Error bars = 1 

SEM; n = 32. 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 307 
	 	 	 	

	 	

	 	

	 	

	 	

	 	

	 	

...reestablishing 

proper grazing 

regimes that 

maintain vigorous 

plants and 

healthy plant 

communities can 

limit invasion.”

can be restored. A combination of invasive 

weed control (typically with herbicides) and 

restoration can potentially reestablish native 

species. Long-lived perennial species often 

garner much of the water, nitrogen, and other 

resources available in grassland ecosystems 

(Wedin and Tilman 1990; Tilman and Wedin 

1991; Baer et al. 2002; Seabloom et al. 

2003; Fig. 3). Consequently, the presence of 

perennial species can provide substantial biotic 

resistance against invasion (Blumenthal et al. 

2003, 2005; Seabloom et al. 2003; Bakker 

and Wilson 2004; Levine et al. 2004). Where 

disturbance damages or removes resident 

species, it can increase resource availability and 

therefore provide opportunities for invasive 

species (Hobbs and Huenneke 1992; Davis et 

al. 2000). However, where desired or native 

species can be restored, their presence may be 

sufficient to reduce resource availability and 

keep invasions from recurring (Blumenthal et 

al. 2003; Seabloom et al. 2003). 

Few studies have actually tested whether 

restoration leads to persistent invasive 

plant control. Ferrell et al. (1998) studied 

the response of leafy spurge to herbicides, 

tillage, and then seeding of several native and 

nonnative grass monocultures. Five years 

postseeding, one grass was very rare, but the 

others substantially suppressed leafy spurge. 

The two most effective species (which were 

nonnative) were reassessed 10 yr postseeding, 

and these species continued to suppress leafy 

spurge. Similarly, Bottoms and Whitson (1998) 

found that herbicides and tillage followed by 

seeding of several native and nonnative grasses 

greatly suppressed Russian knapweed 5 yr 

postseeding. Herbicide, tillage, and seeding of 

native tallgrass prairie species greatly suppressed 

weeds in an old field 7 yr after seeding 

(Blumenthal et al. 2003, 2005). Although 

most of the weed species in this study are not 

considered aggressive invaders of rangeland, 

two of the species inhibited by restoration 

can be desired species or invaders: Kentucky 

bluegrass and smooth brome (Bromus inermis 

Leyss.). Finally, Seabloom et al. (2003) found 

that 5 yr after restoration, native perennials 

comprised the majority of the plant biomass in 

an area otherwise dominated by exotic annuals. 

These studies suggest seeding may sometimes 

provide cost-effective, long-term weed control. 

However, additional longer-term measurements 

are desperately needed to better evaluate the 

long-term benefits of seeding. 

Altering Disturbance Regimes as a Solution 

to Changing Environmental Conditions. 

Where invasion is caused by past disturbance, 

long-term weed control and restoration may 

often be achieved with a combination of 

invasive weed control and seeding. Where 

invasion is caused by ongoing disturbance, 

however, or where past changes have led to new 

stable states, it may also be necessary to change 

the disturbance regime (Suding et al. 2004). 

The most important example of this problem 

in North American rangeland ecosystems is the 

increase in fire frequency that is both caused 

by and helps to perpetuate cheatgrass invasion 

(D’Antonio and Vitousek 1992). Because 

shrub-steppe ecosystems are not well adapted 

to frequent fire, the new disturbance regime 

appears to preclude successful restoration. 

Consequently, fire suppression is required 

to allow native species to compete against 

cheatgrass and other fire-tolerant invaders 

(Brooks et al. 2004). Fire suppression may or 

may not be sufficient to allow native species 

to compete against cheatgrass. If cheatgrass 

invasion is caused by a combination of fire 

and past disturbance, then fire suppression 

and restoration may be sufficient. If cheatgrass 

invasion is also driven by ongoing changes, 

such as nitrogen deposition or the amount/ 

timing of grazing, then fire suppression and 

restoration may not be sufficient. In contrast 

to shrub-steppe ecosystems, tallgrass prairie 

evolved with frequent fire. There, the absence 

of fire can lead to invasion, and prescribed 

fire can be a long-term solution to invasion 

(Smith and Knapp 1999; Copeland et al. 

2002). Similarly, grazing can be a disturbance 

that either favors or inhibits invasive plants, 

depending on the grazing history of the site 

(Mack and Thompson 1982; Milchunas et al. 

1988, 1992; Bock et al. 2007). 

Grazing as a Cause and Solution to 

Invasion. Invasive species have evolved 

with grazing animals, creating a complex 

relationship among grazing preferences and 

plants’ abilities to resist and tolerate grazing 

(Heitschmidt and Stuth 1991). Invasion can 

increase when competitive, desired species are 

defoliated beyond their ability to recover by the 

following growing season (Sheley et al. 1997). 

308 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

Thus, reestablishing proper grazing regimes 

that maintain vigorous plants and healthy plant 

communities can limit invasion (Sheley et al. 

2008). Similarly, lack of grazing of invaders 

can stimulate invasion. Since goats and sheep 

dominate most grazing in areas where many 

invasive weeds evolved, introducing them into 

areas with serious infestations of these species 

can reduce invasive weeds (Olson 1999). 

Biological Control as a Solution for Invaders 

that have Escaped Natural Enemies. It is 

also possible that invaders succeed, overcoming 

biotic resistance, without any change in 

environmental conditions. Invaders may 

have an advantage over native species for 

other reasons. For example, the enemy release 

hypothesis proposes that invasive plants have 

an advantage over native species because they 

have escaped natural enemies when introduced 

to a new range (Maron and Vila 2001; Keane 

and Crawley 2002; Mitchell and Power 2003; 

Colautti et al. 2004; Blumenthal 2006). When 

enemy release is driving invasion, biological 

control, reuniting invasive species with 

specialized enemies, may be the most durable 

control method (Fig. 4). The reversal of enemy of plant species from outside the local plant 

release by biological control is inexact. Only a community (Seastedt et al. 2008). There may 

subset of the enemies from the original range is also be situations in which long-term solutions 

introduced. The introduced enemies, however, simply do not exist. If invasive species are better 

are often missing their own predators, and may adapted to local conditions than are desired 

influence the invasive weed more strongly than species and the local conditions cannot be 

they would have in their native range (Keane altered, management choices may be limited 

and Crawley 2002). to relatively expensive ongoing control or 

learning to manage the invasive species as 

Long-Term Solutions When Causes of novel ecosystems. Given the frequency of 

Invasion Cannot Be Reversed. All of the novel disturbances, opportunities for grazing 

above sections discuss situations in which management, and enemy release, it appears 

reversing the cause of invasion should help possible that the causes of invasion could be 

native species compete effectively against reversed in many cases. Where possible these 

invasive species. However, not all causes of approaches are likely to lead to more persistent 

invasion can be reversed. For example, if invasive weed control and more stable desired 

atmospheric CO2

 enrichment causes invasion plant communities. 

of otherwise healthy native plant communities 

(Smith et al. 2000), there may be no way to restore desired vegetative Cover to 

manage the environment to enable the native Protect Soils, Control erosion, reduce 

plant community to resist invasion. Other Sediment, Improve Water Quality and 

types of global change, such as nitrogen Quantity, and enhance Stream flow 

deposition and altered precipitation, may lead Plant community structure and composition 

to similar problems (Dukes and Mooney 1999; are important drivers of ecosystem function 

Brooks 2003; Vila et al. 2007; Blumenthal and services, including protection and 

et al. 2008). In such situations, efforts to conservation of soil and water resources 

increase biotic resistance, and therefore achieve (Chapin et al. 2000). Our ability to reestablish 

long-term weed control, may require the use desired vegetation on invasive plant–infested 

fIgure 4. Flea beetle (Apthona spp.) (right photo) impact on leafy spurge in Montana 

(bottom photo; 1996) 3 years following their introduction in 1994 (top photo). 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 309 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	

Invasive species can have 

a significant impact on soil 

resources by altering plant 

cover, litter inputs, and the 

amount and distribution of bare 

ground. (Photo: Alex Boehm) 

rangeland is extremely limited, particularly in 

areas with low precipitation. The proportions 

of plant cover and bare ground are generally 

the most important factors that determine the 

degree to which soil and water resources are 

protected and conserved by a plant community. 

To this end, when desired vegetation has been 

reestablished on degraded rangeland with low 

total vegetative cover and a high proportion 

of bare ground, protection and conservation 

of soil and water resources generally increase 

(Pyke et al. 2002; Pierson et al. 2007a). The 

bulk of restoration programs, however, have 

not evaluated if reestablishment of desired 

vegetation has effectively impacted soil or 

water resources. Of the few synthetic efforts 

to date on rangeland, there is little evidence to 

suggest that reseeding efforts are successful in 

establishing enough plant cover to significantly 

improve protection of soil and water resources 

beyond what is attained through natural site 

recovery processes (Pyke et al. 2003; Byers 

2004). Of equal importance is the notion that 

weedy plant communities and desired plant 

communities may not necessarily differ in 

their ability to protect and conserve soil and 

water resources. In some instances, weedy plant 

communities may be important in rapidly 

stabilizing heavily disturbed communities on 

steep slopes, preventing loss and damage to 

soil and water resources (Pierson et al. 2007b). 

Nevertheless, there are a number of instances 

where restoring desired vegetation cover in 

weed-infested communities may benefit soil and 

water resources. Although empirical data are 

limited, some general principles have emerged 

that should allow reasonable prediction of when 

restoring desired vegetative cover on weedinfested rangeland may achieve these benefits. 

Soil Resources. Although a few studies show 

that establishing a desired vegetation cover on 

weed-infested rangeland can increase protection 

of soil resources (Lacey et al. 1989), the bulk 

of evidence is largely observational (Sperber 

et al. 2003), limiting our ability to develop 

generalities. However, processes and factors 

associated with effective soil conservation are 

fairly well defined, which may allow relatively 

accurate predictions of when restoration 

of desired vegetative cover will provide soil 

conservation benefits. Plant cover as well as the 

proportion and connectivity of bare ground are 

central factors determining erosion and sediment 

yield. More canopy cover lowers the effective 

energy of raindrops as well as the amount of 

soil exposed to rainfall impact (Blackburn et al. 

1994). Large, interconnected patches of bare 

ground concentrate runoff and increase flow 

velocities and erosion (Schlesinger et al. 1990; 

Pierson et al. 2007a). The evidence suggesting 

that weeds alter soil physical properties is mixed, 

and appears species-specific (Sperber et al. 

2003; Norton et al. 2004). The most significant 

impacts weeds have on soil resources are related 

to the degree to which weeds affect plant cover, 

litter inputs, and the amount and distribution 

of bare ground (Lacey et al. 1989; Pierson et 

al. 2007a). Therefore, soil conservation benefits 

may be achieved if restoring desired vegetation 

on weed-infested rangeland increases plant cover 

and/or decreases connectivity of bare patches 

and plant interspaces. On the other hand, if 

restoration efforts do not significantly alter these 

parameters, then establishing desired vegetation 

may not significantly improve conservation of 

soil resources. 

Water Resources. The large negative effect of 

invasive plant species on water resources and 

the conservation benefit achieved by restoring 

desirable species is partially documented in 

rangeland riparian systems (Zavaleta 2000; 

Shafroth et al. 2005). However, on upland 

systems the benefits of restoring desired 

species on weed-infested rangeland are less 

well studied and the effects more nuanced. 

Several case studies examine weed effects on 

water resources on uplands and there have been 

extensive studies on individual plant water use 

patterns (Lambers et al. 2000; Enloe et al. 2004; 

Kulmatiski et al. 2006). Therefore, although 

the data on hand are limited, this information 

can be used to develop some general predictions 

as to the effects of weeds on water resources 

as well as the conservation benefits that may 

be obtained by establishing desired species on 

weed-infested rangeland. Patterns and rates of 

plant water use are determined by plant size, 

phenology, rooting depth, and root densities 

(Lambers et al. 2000). Weeds that differ 

significantly from desired vegetation in these 

traits have the potential to alter the pattern 

and amount of water available on rangeland. 

For example, annual grasses that have invaded 

sagebrush steppe systems initiate growth and use 

water earlier in the growing season compared to 

the native perennial bunchgrasses (Kulmatiski 

310 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

et al. 2006). As a consequence, these weedy 

plant communities extract water at a faster rate 

earlier in the growing season than the desired 

plant community. In this case, eliminating the 

weeds and restoring desired perennial plants 

may allow water to remain available to plants 

for a longer duration during the growing season. 

As a contrasting example, deep-rooted forbs 

that invade native bunchgrass communities 

can deplete soil water at greater depths later in 

the growing season compared to bunchgrasses 

(Enloe et al. 2004; Fig. 5). In this case, restoring 

desired bunchgrasses may help conserve deep 

soil water. 

Conclusions and Management Implications. 

Our ability to restore desired vegetation on arid 

and semiarid rangeland is limited. Restoring 

desired species may not always result in a 

conservation benefit in terms of soil and water 

resources. Although few studies have examined 

the conservation benefits of establishing desired 

species, basic knowledge about soil stability 

and hydrological processes allows reasonable 

prediction of scenarios where restoring desired 

species will benefit soil and water resources. 

Namely, if restoring desired species increases 

cover or litter inputs and/or decreases the 

amount or continuity of bare ground, a soil 

and water conservation benefit will likely be 

achieved. The impact of weedy plants on water 

resources and the benefits achieved by restoring 

desired species will mainly depend on the degree 

to which these species groups differ in size, 

phenology, rooting depth, and root densities. 

When these species groups exhibit large 

differences in one or more traits, substantial 

conservation benefits may be achieved. On the 

other hand, when these differences are small the 

conservation benefits may be negligible. 

Maintain or enhance Wildlife habitat 

Including that Associated with 

threatened and endangered Species 

Invasive plant species often change ecosystem 

structure and function, directly impacting 

wildlife habitat (DiTomaso 2000; Masters 

and Sheley 2001). It is not surprising, 

therefore, that invasive species removal has 

been shown to benefit wildlife in a number 

of systems. Relationships between invasive 

plant species and wildlife, however, are often 

more complicated, involving both positive 

and negative effects. For example, saltcedar 

fIgure 5. Volumetric soil water content (%; mean ± SE) by depth averaged 

across time of sampling and year. Plant community by soil depth comparisons: 

yellow starthistle vs. annual grasses (F = 3.33, P = 0.0262), yellow starthistle vs. 

pubescent wheatgrass (F = 0.27, P = 0.8738). 

(Tamarix spp.) can provide suitable habitat 

for the endangered southwestern willow 

flycatcher (Empidonax traillii extimus Nelson) 

and other avian species that nest in midcanopy 

vegetation, but poor habitat for many other 

avian species (Dudley and DeLoach 2004; 

Shafroth et al. 2005; Brand et al. 2008; Fig. 

6). The net benefit of invasive species control 

on wildlife habitat not only depends on the 

balance between negative and positive effects 

of invasive plants on wildlife habitat, but also 

upon the likelihood and time required for 

successful restoration, as well as the direct 

impact of invasive species control measures 

on wildlife and their habitat (Bateman et al. 

2008a). 

The impact of invasive species on wildlife 

habitat, and therefore the benefits gained by 

restoring these habitats, may be relatively 

predictable based on what is known about the 

habitat requirements of particular species. For 

example, deer, elk, and bison rely heavily on 

grasses. When grasslands are invaded by weedy 

invasive forbs, grass production declines and 

animal use of these habitats can decline by up 

to 80% (Thompson 1996; Rice et al. 1997b; 

Duncan 2005). 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 311 
	 	 	 	

 

 

 

 

 

fIgure 6. Avian species richness (mean ± SE) as a function of a, four vegetation types, b, three hydrologic regimes within riparian location 

(floodplain and terrace), and c, nine combined vegetation–hydrologic regime classes on the San Pedro River, Arizona, 1998–2001. EPH 

indicates ephemeral surface water flow; INT, intermittent surface water flow; and PER, perennial water flow. 

Similar effects have been observed on bird 

populations that prefer open grasslands 

(Scheiman et al. 2003). In cases where invasive 

plants alter the preferred forage base or 

structural characteristics of the native plant 

community, restoring these systems likely will 

have a large positive effect on wildlife. 

Even if community structure is not altered, 

restoring natural patterns of plant species 

abundance may greatly improve habitat. For 

example, in western Oregon grassland, tall 

oat grass (Arrhenatherum elatius [L.] P. Beauv. 

ex J. Presl & C. Presl) reduces grassland use 

and egg laying by the endangered Fender’s 

blue butterfly (Icaricia icarioides fenderi 

Macy) (Severns 2008). Oat grass appears to 

reduce use largely by obscuring the butterfly’s 

preferred host plant, Kincaid’s lupine (Lupinus 

oreganus A. Heller var. kincaidii C.P. Sm.), even 

when the lupine is present for butterfly use. 

Plant invasions are likely to have the greatest 

influence on wildlife when their presence leads 

to feedbacks that not only change the plant 

community structure, but also alter ecosystem 

properties. 

For example, in North American rangeland, 

cheatgrass probably has the most widespread 

and severe effects on wildlife of any invasive 

plant. By changing the fire regime, cheatgrass 

can displace shrub-steppe vegetation and 

312 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

associated wildlife species. Furthermore, 

cheatgrass can lead to such changes over 

extremely large areas (Fig. 7). In the 

Intermountain West, it has been estimated 

to occupy 40 million ha (DiTomaso 2000). 

Of primary concern are rare species such as 

greater sage grouse (Centrocercus urophasianus 

Bonaparte), Gunnison sage grouse 

(Centrocercus minimus), Brewer’s sparrows 

(Spizella breweri), sage sparrows (Amphispiza 

belli), sage thrashers (Oreoscoptes montanus), 

and pygmy rabbits (Brachylagus idahoensis). 

Sage grouse are considered to be sagebrush 

obligates (Schroeder et al. 2004), and are 

most likely to persist in large areas with at 

least 25% sagebrush cover (Aldridge et al. 

2008). Both greater sage grouse and Gunnison 

sage grouse have been proposed for listing 

under the Endangered Species Act. Although 

prescribed fire has in the past been suggested 

as a tool to improve sage grouse habitat, recent 

studies suggest fire is most often harmful 

to sage grouse, particularly the frequent fire 

caused by cheatgrass invasion (Connelly et 

al. 2000; Baker 2006). Invasion of sagebrush 

by cheatgrass, and associated increases in fire 

frequency, appear to be primary causes of sage 

grouse decline (Knick et al. 2003; Schroeder 

et al. 2004; Baker 2006). 

Other sagebrush-obligate wildlife species are 

likely to be similarly influenced by cheatgrass 

invasion and loss of sagebrush. Quantification 

of habitat requirements shows that many 

species considered to rely on sagebrush, such 

as pygmy rabbits, sage thrashers, and sage 

sparrows, do in fact have habitats that overlap 

strongly with those of sage grouse (Rowland 

et al. 2006). In eastern Washington, not only 

shrub-nesting sage sparrows, but also a variety 

of ground-nesting birds, were found to be less 

abundant in areas dominated by cheatgrass 

than in shrub–grass plant communities (Brandt 

and Rickard 1994). A variety of avian species 

have also been shown to prefer native perennial 

grass seed to cheatgrass seed (Goebel and Berry 

1976). Small mammals can also be strongly 

influenced by cheatgrass invasion. For example 

bitterbrush-dominated communities have been 

found to support 3–13 times the densities 

of small mammals of cheatgrass-dominated 

communities in central Washington (Gano and 

Rickard 1982; Gitzen et al. 2001). Similarly, 

small-mammal densities and richness were 

higher in intact sagebrush steppe than in areas 

with cheatgrass in Idaho’s Snake River Plain 

(Hanser and Huntly 2006). Finally, a study 

of Townsends ground squirrels (Spermophilus 

townsendii idahoensis Merriam), found higher 

variation in squirrel burrows in cheatgrass 

habitats than in shrub–bunchgrass habitats 

(despite similar mean burrow numbers), 

suggesting cheatgrass provides an adequate but 

unstable food resource for this species (Yensen 

et al. 1992). 

Beneficial Effects of Invasive Plant 

Management on Wildlife. Only a handful 

of studies have actually measured effects of 

rangeland weed control on wildlife or wildlife 

habitat. Chemical control of spotted knapweed 

control was found to release grasses from 

competition and increase winter forage for 

elk by 47% in western Montana (Rice et al. 

1997b). Other wildlife species rely heavily 

on native forbs for food. Consequently, 

restoration of native forbs can be an important 

objective of invasive species control. Fall 

wick application of glyphosate controlled 

Canada thistle (Cirsium arvense [L.] Scop.) 

while increasing shrub biomass, forb biomass, 

and species richness in a Montana waterfowl 

production area (Krueger-Mangold et al. 

2002). Burning can also favor native forbs if 

conducted at the right time. In California, 

burning in the late spring and early summer 

fIgure 7. Bromus tectorum 

invasion in shortgrass 

steppe vegetation near 

Lander, Wyoming. (Photo: 

D. Blumenthal) 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 313 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	 	

There are too 

few direct 

measurements of 

wildlife responses 

to invasive 

species control 

to gauge how 

often control 

yields benefits for 

wildlife.”

after many forbs have set seed can favor native 

forbs over invasive annual grasses (DiTomaso 

et al. 2006b; Meyer and Schiffman 2008). 

Similarly, winter–spring cattle grazing 

appears to reduce annual grass invasion and 

favor native forbs that provide key habitat 

for several butterfly species on California 

serpentine grasslands (Weiss 1999b). A few 

studies have directly documented increased 

wildlife use following invasive species control. 

For example, successful control of spotted 

knapweed with picloram increased elk foraging 

in a Montana old field (Thompson 1996). 

Similarly, mechanical and chemical control 

of saltcedar and Russian olive (Elaegnus 

angustifolia L.) in New Mexico floodplain 

forest increased the abundance of lizards and 

bats, but decreased the abundance of birds that 

nest in midstory vegetation (Bateman et al. 

2008a, 2008b). 

Detrimental Effects of Invasive Species 

Management on Wildlife. Although reduced 

abundance of invasive species is likely to benefit 

wildlife, the methods used to reduce invasive 

species abundance can sometimes harm 

wildlife. A potentially important example of 

direct effects of herbicides on wildlife can be 

found in recent work suggesting that atrazine 

plays a role in global declines in amphibian 

populations (Rohr et al. 2008). Both field 

surveys in Minnesota wetlands and mesocosm 

experiments showed atrazine to be associated 

with increased infection by trematodes (a likely 

proximate cause of amphibian declines) in 

northern leopard frogs (Rana pipiens Schreber). 

Increased infection, in turn, appears to be 

caused by both an increase in the abundance 

of gastropods, which are intermediate hosts for 

trematodes, and decreased immune responses 

on the part of the frogs (Rohr et al. 2008). 

Invasive species control can also harm wildlife 

indirectly, through its effects on the plant 

community. In particular, chemicals that target 

dicots can decrease plant community diversity, 

thereby reducing the food available for some 

wildlife species (Johnson et al. 1996a; Sheley 

et al. 2007). For example, 2,4-D applied to 

western Colorado rangeland to favor grasses 

over forbs and shrubs reduced densities of 

northern pocket gophers (Thomomys talpoides 

Richardson) and least chipmunks (Eutamias 

minimus Bachman), while increasing densities 

of montane voles (Microtus montanus Peale; 

Johnson and Hansen 1969). These effects 

appear to have been caused by a combination 

of reduced food availability and changes 

in vegetation cover. A proposed alternative 

for controlling woody species without 

harming wildlife is tebuthiuron, which can 

reduce woody species without reducing forb 

abundance and diversity (Johnson et al. 

1996b). 

Conclusions and Management Implications. 

There is considerable evidence that invasive 

plants influence wildlife in rangelands. Most 

often this influence is negative, reducing food 

and habitat availability for a wide array of 

wildlife species. Evidence for negative effects 

on wildlife is particularly strong for invaders 

that alter ecosystem structure and function, 

such as weedy forbs invading grasslands and 

annual grasses invading systems historically 

dominated by perennial plants. There are too 

few direct measurements of wildlife responses 

to invasive species control to gauge how often 

control yields benefits for wildlife. Rather, 

wildlife benefits must be inferred from what is 

known about both the relative value of invasive 

and native plant species as wildlife habitat, and 

the effectiveness of management in replacing 

invasive species with native species. 

Protect Life and Property from 

Wildfire hazards 

Wildfires are a regular and natural occurrence 

in many areas of the arid western United 

States and most of these ecosystems are well 

adapted to fires (Brooks et al. 2004). These 

natural ecosystems will return to their preburn 

state within a few years of a fire under normal 

conditions. However, other habitats such as 

riparian corridors, sagebrush scrub, and deserts 

have longer fire-return intervals because of 

sparse and discontinuous vegetation. In these 

areas, the native species are less adapted to 

fire and are susceptible to a short-duration 

fire interval (Brooks et al. 2004). Invasion by 

annual grasses, particularly cheatgrass, red 

brome (Bromus rubens L.), and medusahead, 

have dramatically shortened the intervals 

between fires by providing more continuous 

fuels that are easier to ignite (Brooks et al. 

2004). In addition, invasive annual grasses 

typically reestablish more rapidly than native 

plants after fires. This can further suppress the 

314 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

recovery of the natives and allow the weeds to 

expand their range (Pellant 1990). 

More importantly, if fires occur too frequently, 

some of the native vegetation becomes so 

severely damaged that recovery is no longer 

possible (Pellant 1990; Whisenant 1990). 

This can result in loss of woody species such 

as sagebrush and other important plants and 

wildlife species, and effectively convert highdiversity native plant communities into lowdiversity nonnative communities (Knick 1999). 

In some cases, fire exclusion over a period 

of time can create undesirable conditions 

for both forest sustainability and human fire 

hazard (Keeley 2006). This is the situation 

with some woody species, such as western 

juniper (Juniperus occidentalis Hook; Coultrap 

et al. 2008), which has expanded its range 

dramatically in the northwestern United States. 

Land management agencies, such as the 

USDA Forest Service, Bureau of Land 

Management (BLM), and National Park 

Service, are required to assess site conditions 

following wildfire. Where necessary, they can 

prescribe emergency watershed-rehabilitation 

measures that can 1) help stabilize soil; 

2) control water, sediment, and debris 

movement; 3) prevent permanent impairment 

of ecosystem structure and function; and 4) 

mitigate significant threats to human health, 

safety, life, property, or downstream values 

(USDA Forest Service 2010). Each year 

millions of dollars are spent on emergency 

post-fire rehabilitation treatments (Robichaud 

et al. 2000). 

In southern California, where chaparral 

communities are prone to fire at the wildland– 

urban interface and the societal impacts of 

accelerated postfire erosion are enormous, 

there are pressures to treat burned hill slopes 

with grass seed to protect life and property 

(Gibbons 1995). It was common to seed 

such areas with quick growing annual plants, 

typically nonnative annual ryegrass or 

collections of native and nonnative forbs. This 

practice, however, is no longer recommended 

because the results are often unsuccessful. 

In some cases, heavy rains can wash away 

seeds, or inadequate rainfall prevents good 

seed germination. In addition, some of plants 

used for reseeding can persist and add to 

fIgure 8. Rhinocyllus conicus egg load (untransformed means ± 1 SE) on two native 

thistle species—a, Cirsium flodmanii and b, Cirsium undulate—in grassland patches 

within two landscape types in 2001, and three landscape types in 2002. Results of 

planned contrasts comparing landscape pairs are presented above bars for 2002. 

NS indicates not significant; P < 0.05. 

the invasive plant problem (Bell et al. 2007) 

by competing with the native vegetation 

and preventing recovery. Long-term slope 

stabilization is better achieved by promoting 

the recovery of deep-rooted perennial shrubs 

compared to shallow-rooted annuals. This can 

be accomplished by transplanting shrubs or by 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 315 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

protecting establishing shrubs from herbivory 

or competition from nonnative species. 

Shrub recovery can lead to reduced threat of 

subsequent fires (Bell et al. 2007). 

Conclusions and Management Implications. 

There is a considerable amount of evidence 

to demonstrate the impact of invasive plants, 

particularly annual grasses, on the frequency 

of fires in rangeland systems. In addition, it is 

well recognized that rangeland fires spread by 

invasive plants can cause significant damage 

to property and human health. Although 

few studies have been conducted on the 

interaction between invasive plants, wildfire, 

and impacts to wildlife, it stands to reason 

that these impacts are significant and in most 

cases detrimental to wildlife. With increased 

research on methods to control vegetation and 

protect areas from large catastrophic fires, the 

economic and ecological damage caused by 

invasive plants can be substantially reduced in 

the future. 

Minimize negative Impacts of Pest 

Control on Soil, Water, Air, Plant, and 

Animal resources 

Minimizing negative impacts of pest 

control on biotic and abiotic resources is an 

important step in designing economically 

and ecologically sustainable invasive plant 

management practices (Sheley et al. 2010). 

The most commonly applied control strategies 

for invasive plants on rangeland include 

herbicides, biocontrol, grazing, fire, or 

mechanical control such as tilling (Jacobs et 

al. 1999). Impacts of these control strategies 

on abiotic and biotic resources have been 

assessed to varying degrees and in some cases, 

general ecological patterns and principles are 

beginning to emerge. For example, the fate 

and ecological impact of various herbicides 

on rangelands has been documented and 

there is much evidence suggesting that as 

disturbance (e.g., herbicide use, tilling, 

grazing) intensity increases, invasibility of a 

system also increases (Hobbs and Huenneke 

1992; Davis et al. 2000). Nevertheless, large 

gaps in our understanding of pest control 

impacts on abiotic and biotic resources 

remain. For example, a key component of 

ecologically based invasive plant management 

is to apply pest control strategies that reduce 

the performance of invasive species more than 

the performance of desirable species (Sheley 

et al. 2006). However, a Web of Science query 

that included the search terms “herbicide” and 

“rangeland” demonstrated that only 28% (20 

of 70) of field studies published between 1976 

and 2008 examined herbicide effects on both 

desirable and weedy vegetation. 

Impacts of Control on Soil, Water, and Air 

Resources. Impacts of pest control on soil, 

water, and air resources vary depending on 

pest control strategy, but in general, effects are 

relatively predictable. For example, intense soil 

disturbances contribute to erosion, decreased 

water quality, and dust production and also 

release nutrients, which favors the growth 

of weeds compared to natives (Greene et al. 

1994; Davis et al. 2000; McEldowney et al. 

2002; Zhao et al. 2005). Because of this, 

current management frameworks for weedinfested rangeland focus on using tools that 

will minimize disturbance such as no-till drills 

and moderate grazing in efforts to direct a 

plant community toward a more desirable state 

(Mangold et al. 2006). 

Concerns over the effects of herbicides on 

soil, water, and air resources have been raised 

due to the potential of herbicides to affect soil 

processes, to contaminate groundwater, or 

to be transported on wind-eroded sediment 

and potentially inhaled by humans (Larney 

et al. 1999; Liphadzi et al. 2005; Borggaard 

and Gimsing 2008). The impact of these 

herbicides on these resources is dependent on 

type of herbicide used, application rate, and 

soil characteristics, among other factors. For 

example, glyphosate tightly adheres to soil, 

which makes it difficult for this compound to 

leach into groundwater or affect soil biological 

processes (Borggaard and Gimsing 2008). On 

the other hand, compounds such as dicamba 

and picloram are highly mobile in the soil 

(Krzyszowska et al. 1994). High application 

rates, high rainfall following application, or 

direct application of these compounds to 

water bodies can pose a significant threat to 

water resources. Overall, careful application 

of herbicide following recommended 

procedures coupled with the relatively low 

application rate of herbicides commonly used 

on rangeland tends to minimize the negative 

effects of herbicides on rangeland soil, water, 

and air resources. 

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R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

Impacts of Pest Control on Plant and 

Animal resources 

Herbicides. Only a subset of studies (20 

of 70) has examined herbicide effects on 

both invasive and desirable plant species 

in the field. Although the responses are 

dependent on a number of factors, such as 

mode of herbicide action and site-specific 

environmental conditions, two important 

trends have emerged. First, desirable species 

functionally or taxonomically similar to the 

invasive plant species targeted for control tend 

to be more negatively impacted by herbicide 

application. For example, herbicides such as 

2,4-D, clopyralid, or picloram are commonly 

applied to control broadleaf weeds such as 

knapweed, leafy spurge, and sulfur cinquefoil 

on rangeland. Because grasses are capable 

of metabolizing these compounds, desirable 

rangeland grasses are generally unaffected by 

these herbicides (Sheley and Jacobs 1997; 

Sheley et al. 2002; Laufenberg et al. 2005). 

However, these herbicides can greatly decrease 

native forb density and cover (Sheley and Effects of herbicides on mammals, birds, and 

Denny 2006). There is evidence suggesting invertebrates are generally identified during the 

that herbicide effects on native forbs are long- ecological risk assessment prepared with each 

lasting and can drive a local decline in species herbicide and for public land management 

richness (Fuhlendorf et al. 2002; Rinella activities during environmental impact 

et al. 2009). As another example, desirable reporting. Although a number of herbicides are 

rangeland grasses have shown varying degrees available to control weeds on rangeland, 70% 

of susceptibility to imazapic, a herbicide of the land treated with herbicides by the BLM 

used to control invasive annual grasses, uses 2,4-D, glyphosate, picloram, tebuthiuron, 

with evidence suggesting grasses within or imazapic. Of these, glyphosate, picloram, 

the Hordeae tribe may be more tolerant to and imazapic show low toxicity to terrestrial 

imazapic than other grass species (Kyser et al. animals whereas tebuthiuron and 2,4-D 

2007). A second trend is that the impact of demonstrate moderate toxicity. The low rates 

herbicides on desirable vegetation depends on of herbicide applied on rangeland combined 

the rate and timing of herbicide application. with relatively low toxicity and lack of chronic 

In general, when herbicides are applied exposure suggest herbicides have minimal effect 

several weeks prior to seeding or during a on terrestrial animal species on rangeland. 

dormant seeding, herbicides have a greater 

selectively for weeds compared to seeded Biocontrol. Development and release of 

species (Jacobs et al. 1999; Kyser et al. 2007; biocontrols follows international and national 

Sheley 2007). Higher herbicide application guidelines designed to minimize the possibility 

rates can have negative impacts on seeded that biocontrol releases will negatively impact 

species, even during fall dormant plantings desirable vegetation (FAO 1996; Wilson 

(Monaco et al. 2005). Even with a given rate and McCaffrey 1999). Biological control has 

and timing of herbicide application, desirable been implemented successfully in a number 

species response can vary substantially across of systems (e.g., Huffaker and Kennett 1959; 

sites in a given year and across years in a McEvoy et al. 1991; Lym 2005) and when 

given site (Monaco et al. 2005; Sheley et al. operating under current protocols there are 

2007). Beyond a few generalities, the effect relatively few documented direct effects of 

of herbicide on desirable vegetation remains biological control on desirable vegetation 

difficult to predict. given the number of biocontrol releases 

The impacts of herbicides on 

desirable and undesirable 

species depend on the rate and 

timing of the herbicide application. (Photo: Rob Wilson) 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 317 
	 	 	 	

(Funasaki et al. 1988; Center 1995). There 

is, however, mounting evidence suggesting 

that poor monitoring efforts, difficulty in 

predicting biocontrol effects, and the largely 

unrecognized indirect effects biocontrols 

can have on ecosystems contributes to an 

underestimation of the detrimental effects of 

biocontrols on desirable vegetation (Simberloff 

and Stiling 1996; Thomas and Willis 1998; 

Pearson and Callaway 2008). For example, 

the bulk of biocontrol monitoring focuses 

on release sites with little attention paid to 

offsite biocontrol effects even though there is 

strong evidence demonstrating landscape-scale 

variation in biocontrol effects on desirable 

vegetation (Simberloff and Stiling 1996; Rand 

and Louda 2004). In addition, it is estimated 

that less than half of the biological control 

efforts targeting invasive plants in the United 

States demonstrated any evidence of control 

(OTA 1995). Given that our ability to predict 

biocontrol effects on well-studied target 

vegetation is so low, some researchers have 

questioned the ability to predict biocontrol 

effects on desirable vegetation (Thomas and 

Willis 1998). Although examples of direct 

effects of biological control on desirable 

vegetation in a number of systems supports 

these concerns (Simberloff 1992), of equal 

importance is the recent literature showing 

complex indirect effects of biocontrol on 

desirable vegetation. For example, following 

the collapse of the target pest population, 

intense competition among biocontrol agents 

can cause a transient increase in host plant 

range, which results in the biocontrol agents 

attacking desirable vegetation (Lynch et al. 

2002). Alternatively, when biocontrol agents 

only moderately damage invasive plants they 

may increase invasive plant competitive ability 

by stimulating compensatory growth (Callaway 

et al. 1999). In this situation, moderately 

damaged invasive plants may serve to maintain 

biocontrol densities at high levels, increasing 

biocontrol impacts on desirable vegetation 

(Rand and Louda 2004; Figs. 8a and 8b). These 

patterns of responses suggest, at a minimum, 

that current procedures do not adequately 

prevent biocontrol efforts from having 

significant impacts on desirable vegetation. 

Concerns over the effects of biocontrol on 

animal resources largely have been centered 

on within-guild (e.g., insect) interactions. For 

example, if an introduced biocontrol insect 

shares food sources or parasites with a native 

insect, then biocontrol can have direct and 

indirect effects on native insect populations 

(Louda et al. 1997; Willis and Memmott 

2005). An analysis of 17 food webs in Australia 

showed that a weed biocontrol agent with high 

weed host specificity was associated with a 

decline in native insect diversity (Carvalheiro 

et al. 2008). Although the magnitude of 

these direct and indirect effects are difficult to 

quantify and are generally underreported in 

the literature, basic community ecology theory 

predicts that such affects may be common 

(Holt 1977). In some cases, however, effects of 

biocontrol on a desirable plant community can 

be complex and difficult to predict, involving 

multiple interactions within a food chain. For 

example, introduction of gall flies to control 

spotted knapweed dramatically increased 

deer mouse populations that used gall flies as 

a food source (Ortega et al. 2004). Because 

deer mice also use native plant seed as a food 

source, introducing gall flies increased deer 

mouse populations which resulted in increased 

predation on native seeds and overall decrease 

in native plant density (Pearson and Callaway 

2008). Although theory and empirical evidence 

suggest biocontrols likely will have some 

negative effect on native animal populations, 

biocontrol may still be an appropriate option 

if benefits outweigh the costs. Namely, if 

biocontrols have a large negative effect on 

weed populations, this benefit may outweigh 

moderate negative impacts of biocontrol on 

native plant and animal populations. 

Grazing. Prescribed grazing effects on 

nontarget vegetation depend on a number of 

factors, including animal species used, timing 

of grazing relative to the phenology of desirable 

vegetation, and forage quality and quantity 

of weedy vegetation relative to desirable 

vegetation, as well as grazing tolerance of weedy 

and desirable species. Moderate grazing using 

animals or mixtures of animals (e.g., sheep 

and cattle) that demonstrate certain dietary 

preferences for a particular weed can be used 

to decrease weed density and increase density 

of desirable plants (Bowns and Bagley 1986; 

Sheley et al. 1998). In general, when grazing is 

limited to periods when weedy species are most 

susceptible to defoliation and desirable plants 

are largely dormant, the impact of grazing on 

318 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

Prescribed grazing can be an 

effective tool in reducing the 

vegetative growth of invasive 

species. (Photo: Brenda Smith) 

desirable vegetation can be minimized and 

benefit of grazing for weed control maximized 

(Kennett et al. 1992). For example, utilization 

of grasses by sheep in areas infested with 

knapweed was decreased by timing grazing 

to occur when knapweed was still growing 

and vegetative growth of desirable grasses had 

largely stopped for the season (Thrift et al. 

2008). If weedy and desirable vegetation have 

comparable forage quality, grazing animals 

largely will consume plants in proportion to 

their abundance. For example, diets of sheep 

used to graze spotted knapweed were over 50% 

grasses in areas with low spotted knapweed 

density, but were less than 20% grasses in areas 

with high spotted knapweed density (Thrift 

et al. 2008). When weeds have much lower 

forage quality compared to desirable vegetation, 

grazing animals can have a larger preference 

for and a much greater negative impact on the 

desired vegetation (Ralphs et al. 2007). Despite 

these general guidelines it is difficult to predict 

the effect of grazing on desirable vegetation. 

For example, grazing leafy spurge infestations 

has been found to decrease (Jacobs et al. 2006), 

increase (Seefeldt et al. 2007), or have no 

effect on (Lacey and Sheley 1996) the cover 

of desirable grasses. Although a portion of 

this variation may be explained by differences 

in grazing systems, differences in grazing 

tolerance between weedy and desirable species 

at a particular site also may be important 

(Kennett et al. 1992; Kirby et al. 1997; 

Olson and Wallander 1997). If weedy species 

demonstrate a greater tolerance to grazing than 

desirable species do, then prescribed grazing 

may be a counterproductive control strategy 

even if grazing animals demonstrate greater or 

equal preference for weedy species compared 

to desirable species (Kimball and Schiffman 

2003). Although the value of prescribed grazing 

for weed control has been demonstrated in 

a number of systems, negative impacts on 

desirable vegetation have been demonstrated, 

highlighting the need to closely monitor 

prescribed grazing efforts. 

Conclusions and Management Implications. 

A key step in designing economically and 

ecologically sustainable invasive plant 

management practices is to apply management 

techniques that minimize negative impacts 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 319 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

on biotic and abiotic resources. Some general 

principles are beginning to emerge allowing 

progress to be made toward this goal, such 

as our understanding of the relationship 

between disturbance intensity and invasibility. 

Although some negative effects of pest control 

strategies on native plant and animal resources 

are likely, herbicides, biocontrol, and grazing 

can be applied in ways that greatly minimize 

these impacts if ecological processes and 

mechanisms are considered beforehand and 

control strategies are adjusted to address 

these factors. Identifying these processes 

and mechanisms and making necessary 

adjustments in management, however, is far 

from straight forward. Complex direct and 

indirect effects of control efforts on desirable 

plant and animal resources occur, requiring 

careful implementation of control efforts, 

comprehensive monitoring, and a broad 

determination of costs and benefits achieved by 

control efforts that include multiple ecosystem 

components. 

reCoMMendAtIonS And 

KnoWLedge gAPS 

Our recommendations are centered on three 

general aspects of invasive plant management. 

The first revolves around improving and 

standardizing data collection and risk analysis 

needed to better inform management 

decisions. Second, progress toward sciencebased management of rangeland threatened 

and/or dominated by invasive species must 

be greatly accelerated. Third, invasive plant 

management would greatly benefit from 

the development and implementation of a 

comprehensive education and technology 

transfer program. The objective of this 

portion of the document is to provide critical 

recommendations to guide future development 

of invasive weed management and to identify 

important knowledge gaps. A brief rationale 

and justification for each recommendation and 

knowledge gap are provided as well. 

Standardized data Collection, risk 

Analysis, and Prioritization Procedures 

The magnitude and complexity of invasive 

plant management requires that ecologists 

garner maximum information from all 

datasets. Data collection for both invasive 

plants and desired species is central to 

developing appropriate management 

programs in the future. Standardized data 

collection will be required in order to allow 

data comparisons among years and data 

combinations to conduct meta-analysis needed 

for development of robust principles for 

management. Managers need standardized 

data collection procedures to create accurate 

vegetation assessments that allow periodic 

evaluations of their management. Inventory 

data must be summarized and analyzed to 

forecast likely future vegetation patterns so 

ecological and economic risk/benefit analysis 

can be accurately conducted. Standardized 

ecological and economic data collection would 

be critically valuable to determine land areas 

with characteristics that favor the likelihood 

of success in response to a particular control 

strategy. 

Science-Based Solutions to Invasive 

Plant Management 

Just as physics provides the scientific 

principles for engineering, ecology must 

provide the scientific principles for invasive 

plant management. We strongly recommend 

further development of ecologically based 

management frameworks that can be used 

to guide the incorporation and application 

of ecological principles for invasive plant 

management. Frameworks must be useful to 

researchers and managers, so the connection 

between these complementary endeavors 

is natural and direct. State-and-transition 

models that utilize ecological processes 

and the influence of management on these 

processes to predict vegetation dynamics 

represent a viable framework for various 

ecological site descriptions. A process- and 

evidence-based approach is central to 

advancing invasive plant management from 

misapplied treatments that address only 

symptoms to management programs that 

emphasize the underlying cause of invasion, 

retrogression, and succession. 

Complex interrelationships among various 

components within ecosystems create multiple 

indirect responses to vegetation management 

that are very difficult to predict. This creates 

a strong need to manage invasive plants 

within the context of the entire ecosystem. 

Invasive plant management must become 

more integrated within a systems approach to 

320 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	 	

	 	 	 	

	 	 	

	 	

	 	 	 	

	 	 	 	

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

facilitate problem solving and the attainment 

of well-defined goals, rather than only 

practice-based outcomes. Management must 

assess the complex interrelationship among 

ecosystem components and processes and 

design management strategies that influence 

the underlying ecological cause of invasion 

and dominance by invaders with predictable 

outcomes. 

Imposing management that addresses the 

actual cause of invasion is clear in some 

cases. For example, the increase in invasive 

wetland species in flooded waterfowl habitat 

on the Malheur National Wildlife Refuge 

requires flooding regimes to be less frequent, 

allowing substantial dry periods to shift the 

balance in favor of diverse vegetation. In 

many cases, the actual causes of invasion 

are less obvious and may actually be a result 

of multiple direct and indirect interactions 

that determine successional dynamics. Thus, 

weed ecologists and scientists must develop 

guidelines to evaluate causes of invasion, 

succession, and retrogression. Once these 

guidelines are developed, ecological principles 

must be developed that provide guidelines 

for managers to impose tools and strategies 

to influence conditions, mechanisms, and 

processes in favor of desired vegetation. As 

multiple interactive ecological processes 

require amendment, integrated plant 

management strategies can be developed 

and employed much more effectively. In this 

way, various plant management strategies 

can be designed based on how the treatments 

influence the ecological processes that direct 

ecosystem change. Tools and strategies that 

are based on sound ecological principles 

could enhance our ability to employ effective 

integrated management. 

Enhancing our ability to prevent invasion 

is critical for successful implementation of 

integrated invasive plant management. Given 

the complexity and persistence of invasive 

plants, a proactive approach focused on 

systematic prevention and early intervention 

Invasive species management 

that addresses the actual 

cause of invasion and 

employs ecological principles 

in management strategies 

increases the likelihood of 

success. (Photo: Chris Call) 

CHAPTER 7: Invasive Plant Management on Anticipated Conservation Benefits 321 
	 	 	 	

 

 

	 	 	

	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	

Practicing prevention of 

invasive species, such as 

medusahead, is more economical and more effective than 

costly restoration. (Photo: Ryan 

Steineckert) 

could be much more effective than the existing 

reactive approach. Most managers recognize 

the importance of prevention, but lack the 

ability to effectively employ it. Sciencebased prevention strategies that are based 

on the ecology of seed dispersal are severely 

needed. Land managers need a conceptual 

framework and associated tools that assist 

them in indentifying which vectors are major 

contributors to invasive species dispersal and 

propose dispersal management strategies to 

minimize or interrupt these major vectors. 

Effective methods for containing existing 

infestations are also needed. 

Invasive plant management is currently 

applied in a somewhat haphazard way based 

on political pressure and funding resources. In 

the future, more emphasis should be focused 

on prioritizing invasive plant management in 

areas that have the highest likelihood of success 

both economically and ecologically. Methods 

for prioritizing invasive plant management 

will continue to be increasingly necessary as a 

means to effectively allocate scarce resources. 

Moreover, the lack of successful control 

of invasive species indicates that we may 

transition toward a management philosophy 

that minimizes the negative impacts of invasive 

species and maximizes the ecological and 

economic benefits garnered from invasive 

weed management programs. Concepts, such 

as economic/ecological injury levels, biomass 

optimization models, and thresholds will 

need to be carefully developed in a manner 

that helps managers prioritize management 

programs to address invasive species. 

Comprehensive education and 

technology transfer Programs 

Although some of the necessary infrastructure 

to conduct educational programs effectively is 

in place, a unified, progressive, and outcomebased educational and technology transfer 

program would have strong synergistic effects 

on invasive plant management. Educational 

programs vary widely in their objectives, 

content, and outcomes. Current programs 

lack continuity of message and the ability to 

progressively advance managers’ understanding 

of science-based management. We propose 

that various ecological societies, managers, 

and researchers develop a comprehensive 

science-based curriculum promoting the most 

state-of-the-art, science-based assessment 

and management strategies. Once developed, 

training modules could be developed 

for various portions of the educational 

infrastructure having responsibility for natural 

resource extension and outreach. 

Restoration of invasive plant–dominated 

rangeland is extraordinarily risky and 

expensive. Based on our assessment, the 

continued application of “farming system” 

seeding methods is unlikely to provide 

sustainable replacement of invasive species. 

Many ecological barriers to seed germination, 

seedling establishment, and population 

development exist in restoration areas where 

322 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	

R. L. Sheley, J. J. James, M. J. Rinella, D. Blumenthal, and J. M. DiTomaso 

invasive plants dominate. Managers must have 

an understanding of these barriers and methods 

for overcoming them if restoration is to become 

a useful strategy to restore previously invaded 

sites and prevent reinvasion in the future. 

Restoration approaches must be founded 

upon ecological principles that can be applied 

to specific sites and varied as environmental 

conditions vary across landscapes. 

Invasive plant problems and solutions are 

complex and management outcomes are rarely 

predictable. Ecologists and managers are often 

uncertain about the best management practices 

to employ, or if management will actually 

repair plant communities and the associated 

ecological processes. In most cases, simple 

answers to complex situations do not exist and 

solutions to invasive plant problems are elusive. 

Managers need scientifically credible methods 

for testing various management strategies that 

can be used when management programs 

are being planned and implemented. These 

adaptive management strategies should include 

controls for comparisons and designs that use 

simple experimental hypothesis testing, in 

addition to monitoring previous effectiveness. 

A major strength of adaptive management is 

that it would allow managers to continuously 

evaluate the effectiveness of current invasive 

plant management programs and assist 

with identification of the most successful 

management programs. 

ConCLuSIonS 

Invasive plants negatively impact rangelands 

throughout the western United States by 

displacing desirable species, altering ecological 

processes, reducing wildlife habitat, degrading 

systems, altering fire regimes, and decreasing 

forage productivity. Assessing the influence of 

conservation practices on the perceived benefits 

to ecosystems is critical to understanding their 

usefulness in maintaining sustainable ecological 

and economic systems. We conducted a 

comprehensive synthesis of peer-reviewed 

literature to determine the efficacy of various 

invasive plant strategies on several anticipated 

benefits. The literature documented only shortterm vegetation responses to invasive plant 

management and rarely addressed long-term 

ecological outcomes associated with invasive 

plant management. Our ability to protect 

noninfested lands is encumbered by the lack 

of early detection techniques and effective 

eradication efforts once new infestations are 

identified. Several strategies for maintaining 

invasion-resistant plant communities are 

beginning to emerge. Herbicides provided 

short-term control of most invasive weeds, but 

without additional management, weeds often 

return rapidly. Documentation of the efficacy 

of biological control on plant development is 

well established, but positive effects on control 

and vegetation dynamics are exceedingly rare. 

Grazing management is emerging as a useful 

method for managing invasive plant species, 

but the timing, intensity, and frequency 

of grazing, as well as the class of livestock 

are only known for a few invasive species. 

Restoration of infested rangeland is difficult 

and only successful about 20% of the time 

when nonnative plant material is seeded and 

the probability is even less when native species 

are used. There are cases in which invasive 

plant management strategies can be effective, 

and in those cases, the management strategies 

appear to favorably affect wildlife and other 

important ecological attributes of ecosystems. 

However, most strategies are associated with 

high ecological risks and high risk of failure in 

the long term. It is clear that more research is 

necessary if the anticipated benefits of invasive 

plant management are to be achieved. This 

synthesis indicates that long-term invasive plant 

management is lacking for most applications 

and that ecologically based invasive plant 

management is desperately needed to meet this 

escalating problem on rangelands. 

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CHAPTER 

8 

A Landscape Approach to 

Rangeland Conservation 

Practices 

Brandon T. Bestelmeyer,1 Joel R. Brown,2 Samuel D. Fuhlendorf,3 

Gene A. Fults,4 and X. Ben Wu5 

Authors are 1 Research Ecologist, US Department of Agriculture Agricultural Research Service Jornada
 

Experimental Range, Las Cruces, NM 88003; 2Rangeland Ecologist, US Department of Agriculture
 

Natural Resources Conservation Service Jornada Experimental Range; 3Sarkeys Distinguished
 

Professor, Department of Natural Resource Ecology and Management, Oklahoma State University,
 

Stillwater, OK 74078; 4Rangeland Management Specialist, US Department of Agriculture
 

Natural Resources Conservation Service West National Technology Support Center,
 

Portland, OR 97232; and 5Professor, Department of Ecosystem Science and Management,
 

Texas A&M University, College Station, Texas 77843.
 

Correspondence: Brandon T. Bestelmeyer, USDA ARS Jornada Experimental Range , MSC 3JER, 

Box 30003, New Mexico State University, Las Cruces, NM, 88003-8003. Email: bbestelm@nmsu.edu. 

Reference to any commercial product or service is made with the understanding 

that no discrimination is intended and no endorsement by USDA is implied 

337
 
	 	 	 	

	 	

	 	

	 	

	 	 	

	 	 	

	 	 	

	 	 	

 

...field-scale (local) 

evaluations of 

“

conservation practices 

that have been 

emphasized in this 

document do not 

enable evaluation of 

the broader-scale, 

cumulative effects of 

practices”

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8 A Landscape Approach to 

Rangeland Conservation Practices 

Brandon T. Bestelmeyer, Joel R. Brown, Sam D. Fuhlendorf, Gene A. Fults, 

and X. Ben Wu 

IntRoduCtIon 

A primary objective of the Conservation 

Effects Assessment Project (CEAP) has 

been to evaluate whether or not rangeland 

conservation practices (hereafter “practices”) 

supported by the US Department of 

Agriculture Natural Resources Conservation 

Service (USDA NRCS) yield the 

environmental benefits that we ascribe to 

them. In doing so, the authors of most CEAP 

chapters have focused on specific conservation 

practices (e.g., grazing management or 

brush control) and sought generalities about 

their effectiveness based on a review of the 

literature. Often, a weight-of-evidence–based 

interpretation is drawn from the percentage of 

studies that support a particular assertion. The 

goal of this approach is to produce general 

recommendations about the implementation 

of practices, sometimes tailored to broad 

ecosystem types. The limitations of this 

approach, however, are revealed when the 

weight of evidence for or against the utility of 

a practice is equivocal. Often, practices both 

succeed and fail in different situations. Thus, 

the evidence suggests that we do not have 

the data to discover under what circumstances 

a practice succeeds or fails (Michener 1997). 

In other words, we cannot yet account for 

spatial heterogeneity, including variations in 

soils, climate, vegetation state, within-site 

patchiness, and landscape position relative 

to dispersal, water movement, and other 

processes that strongly influence the success 

of practices. In the face of this information 

shortage, we inevitably overgeneralize and fail 

to recognize important variations in context. 

A related issue is that many practices are 

believed to have benefits that are manifested 

at spatial scales larger than the treatments 

themselves. A primary example is the 

expectation that prescribed grazing or brush 

management applied to uplands will have 

measurable effects on riparian or watershed 

function (Goodwin et al. 1997). It is also 

expected that practices will have a measurable 

positive impact on rangeland conditions 

at a landscape to regional level and that an 

improvement in one location is not offset 

by degrading processes at another location. 

Thus, field-scale (local) evaluations of practices 

that have been emphasized in this document 

do not enable evaluation of the broaderscale, cumulative effects of practices (e.g., 

Kondolf et al. 2008). Because of both spatial 

heterogeneity and differences in how multiple 

local treatments scale up to affect broad-scale 

attributes, we cannot simply assume that 

more is better in a linear way. We will have to 

measure directly attributes at broad scales and 

relate them to the locations and consequences 

of field-scale practices. Such linkages are 

currently rare because conceptual models of 

cross-scale interactions are only now being 

developed and resource managers are usually 

not certain how to apply them. 

Similarly, from a sociological standpoint, 

there is an expectation that successful 

practices accelerate their adoption by other 

landowners in the immediate area (Kreuter et 

al. 2005). Among the hypothesized benefits 

of conservation programs, including technical 

assistance and cost sharing, is that local 

demonstration of benefits encourages the use 

of practices among neighboring managers. The 

spread of practices among managers provides 

another means for local practices to have effects 

at broader scales. 

In this synthesis, we promote the development 

of a systematic approach by which the NRCS 

Blue grama (Bouteloua gracilis) 

grass (foreground) and the Animas Mountains (background) 

in the Malpai Borderlands of 

southwestern New Mexico. 

(Photo: Brandon Bestelmeyer) 

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...information 

provided by 

this “landscape 

perspective” 

could enable 

planners 

to increase 

successful 

application, use 

federal resources 

more efficiently, 

and assess more 

effectively the 

consequences 

of practices.”

and other agencies can evaluate both the local 

and landscape context of practices—that is, 

where they occur in a landscape and region 

and the varying processes and constraints 

associated with those locations. We further 

emphasize that this approach should include 

increased attention to spatial pattern as an 

attribute that contains valuable information, 

in addition to averages or sums of variables 

(such as plant cover) that are typically 

emphasized. Collectively, the information 

provided by this “landscape perspective” 

could enable planners to increase successful 

application, use federal resources more 

efficiently, and assess more effectively the 

consequences of practices. The empirical 

basis for these assertions within the rangeland 

conservation literature is weaker than for 

other chapters due to limited development of 

landscape perspectives in rangeland ecology 

and the consequent paucity of studies. 

Nonetheless, evidence from the broader 

literature in landscape ecology (e.g., Liu 

and Taylor 2002) and some key examples 

in rangelands supports our contention that 

it is essential for NRCS and its partners to 

develop 1) interpretive tools that facilitate 

a consideration of landscape context and 

spatial pattern in conservation planning and 

assessment and 2) database systems that link 

practice effects to ecological sites, state-andtransition models (STMs), and the mosaic of 

ecological sites and states in a landscape. 

This chapter is organized into seven sections. 

Following this introduction, we review 

the current processes used by NRCS in 

conservation planning at different spatial scales. 

We then review concepts that can be used to 

better place practices in a landscape context 

and introduce a spatial hierarchy to facilitate 

application of landscape concepts. We then 

outline a model-based approach that could be 

used to design and test the effects of practices, 

taking into account landscape context and 

linking tests to ecological site descriptions 

(ESDs) and STMs. We offer some general 

recommendations for incorporating landscape 

perspectives in conservation planning, identify 

knowledge gaps that must be overcome to act 

on some recommendations, and conclude that 

landscape perspectives are useful and feasible. 

Because the language used to describe elements 

of the landscape perspective is not well-known 

or standardized, we encourage readers to refer 

to definitions for terms and phrases used in this 

chapter (Table 1). 

CuRRent StAte of LAndSCAPe 

PeRSPeCtIveS In ConSeRvAtIon 

PRACtICeS 

Conservation Planning at Multiple 

Scales 

Most NRCS staff currently involved in 

conservation planning and implementation 

have been trained as “progressive” planners, 

which means that NRCS planners and clients 

incrementally implement conservation practices 

within a specified area. Progressive planning 

has the advantage of focusing resources on 

immediate concerns, but lacks the spatial 

and temporal perspectives necessary to meet 

landscape-scale goals and, more importantly 

here, to provide a consistent and transparent 

basis to assess conservation effects. The end 

result of the planning process should be a 

conservation plan that identifies specific actions 

to be taken by land managers in order to meet 

objectives for specific land areas. For a variety 

of reasons, progressive planning often does 

not result in a comprehensive strategy that 

addresses the variety of conservation needs at 

different scales. 

Regardless of the planning approach, 

conservation plans rely on the implementation 

of individual or combinations of conservation 

practices to achieve objectives. “Conservation 

practices” are protocols for actions taken 

by land managers to improve or maintain 

the condition of rangelands (USDA NRCS 

2003a). Practices are classified as 1) vegetation 

management practices, 2) facilitating practices, 

or 3) accelerating practices. To a large extent, 

this classification also reflects the amount of 

resources required to implement the practices: 

1.	� Vegetation management practices are 

intended to influence the use and growth 

of the vegetation and are specifically 

evaluated in other chapters. Examples 

include prescribed grazing and prescribed 

burning. 

2.	� Facilitating practices are intended to create 

infrastructure that aids in vegetation 

management and are only indirectly 

evaluated in other chapters. Examples 

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B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

Cross-scale interactions 

Land unit/plant community 

Landscape 

Landscape/spatial context 

Landscape level/scale 

Multiple scale 

Patch 

Patchy/Patchiness 

Scale 

Scale of heterogeneity/spatial pattern 

Scaling up 

Spatial heterogeneity 

Spatial interaction 

Spatial or landscape pattern 

tABLe 1. Glossary of landscape-related terms and phrases used in this chapter. Many definitions adapted from Turner et al. (2001) 

How processes at one spatial or temporal scale interact with processes at other scales 

(e.g., fine-scale plant growth interacting with flows of surface water in a landscape) 

Areas of land that are sufficiently large to be of management interest and are ecologically 

homogeneous with respect to management issues 

An area that is spatially heterogeneous in a property of interest, usually with respect to plant 

communities 

The influence of the location within a landscape or in space, including both underlying 

heterogeneity and spatial interactions with neighboring locations 

A level of biological organization characterized by a mosaic of plant communities that have 

developed in response to common soil and geomorphic processes and that often exhibit spatial 

interactions (e.g., a watershed or multiple watersheds) 

Simultaneous consideration of patterns and processes occurring at different spatial scales 

A relatively homogeneous area that differs from its surroundings; used here in reference to 

distinct fine-scale assemblages of plants or ground cover that make up a plant community 

Being made up of different patches; the degree to which an area is made up of diverse 

patches 

Spatial dimension of a measured attribute or process, characterized by its grain (smallest 

resolved unit) and extent (the area across which measurements are taken) 

The idea that heterogeneity and pattern can be defined differently at different spatial scales 

Using measurements of properties gathered at finer scales to estimate properties at broader 

scales 

Variability or dissimilarity of properties of interest across a defined area 

The flow of matter, disturbance, or information from one location to another 

The arrangement of patches or land units of interest in geographical space and relationships 

between these units 

include water developments, stock trails, individual field and property boundaries. 

and fencing. Although this level of planning may involve 

3.	� Accelerating practices are intended to a high degree of precision regarding the 

supplement vegetation management by placement of practices (i.e., fencing, water 

promoting plant community change more development, roads), there is seldom a clearly 

rapidly than is possible through vegetation defined link to important processes occurring 

management alone, often at great expense. at larger spatial scales. For example, a water 

Examples are brush management, range development may be planned and implemented 

planting, or channel stabilization. within an individual pasture with the 

objective of improving grazing distribution. 

Conservation planners work with individuals However, the improved grazing distribution 

and groups to inventory resources, identify is seldom documented quantitatively 

concerns and objectives, and develop a and the assumed larger-scale effects of 

conservation plan. Once all of the planning improved grazing distribution (water quality, 

inputs have been documented, individual habitat improvement) also lack consistent 

practices are assembled into conservation measurement. Notable exceptions to this 

systems to meet the needs of clients (e.g., a generalization include cases in which planning 

resource management system). This process considered sage grouse movements between 

involves map-based decisions about where nesting and lekking sites (Connelly et al. 2000) 

to implement particular practices within or where planning was designed to prevent 

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 341 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	

	 	 	 	 	

	 	 	

A semiarid grassland 

(foreground) and creosotebush 

shrub savanna (background 

on hill) in southcentral New 

Mexico. (Photo: Brandon 

Bestelmeyer) 

the spread of waterborne bacteria (Escherichia 

coli O157:H7) from rangelands into green 

leafy produce fields (Tian et al. 2002; Jay et al. 

2007). 

One limitation to the broader use of 

landscape perspectives is the current focus 

on the relatively homogeneous “site” as 

the fundamental spatial unit of rangeland 

inventory, planning, and assessment (Brown 

et al. 2002; Washington-Allen 2006). The 

size of a site varies, but it is roughly 5–50 

ha. Following current NRCS protocols, sites 

are classified to an “ecological site” based on 

the relationship of potential vegetation to 

differences in climate, soils, and landscape 

position relative to water movement or solar 

energy inputs. The potential vegetation and 

associated vegetation dynamics described 

for ecological sites (using STMs) have been 

used as a benchmark with which to gauge 

the effects of management on rangelands. 

The current ecological site system defines 

benchmarks on a “piece-by-piece” basis, 

however, and does not directly address how 

the composition and arrangement of multiple 

sites (i.e., landscape context) or variability 

within sites should influence decisions 

(USDA NRCS 2003a). Research in landscape 

ecology and the practical experiences of 

planners suggest that the landscape context 

of a land unit can be critically important 

(explanation below). Although management 

units commonly contain multiple ecological 

sites, there has been little attempt to 

formalize procedures for using information 

about the composition and arrangement of 

sites or variability within sites in planning 

or assessment. Historically, planning and 

implementation of practices did not consider 

342 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	

	 	

	 	

	 	

	 	 	

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

landscape context in the placement of 

water facilities, fences, or other facilitating 

practices. Multiple-site planning most often 

occurred when designing grazing systems to 

accommodate livestock movement, unique 

plant growth patterns, and nutritional needs. 

However, there has been little effort to search 

for principles relating to the assessment of 

practice effects at multiple spatial scales. 

Furthermore, there is little guidance for 

integrating multiple properties into planning 

or assessment, which represents an even 

broader scale of heterogeneity and pattern. 

There is some precedent for planning and 

assessment at multiple scales in the form of 

“area-wide conservation plans” referred to in 

the National Planning Procedures Handbook 

(USDA NRCS 2003b). 

“Area-wide conservation plans are voluntary, 

comprehensive plans for a watershed or 

other large geographic area. Area-wide 

conservation plan development considers 

all natural resources in the planning area as 

well as social and economic considerations. 

Plan development follows the established 

planning process to assist local people, through 

a voluntary locally led effort, to assess their 

natural resource conditions and needs; set 

goals; identify programs and other resources 

to solve those needs; develop proposals and 

recommendations to do so; implement 

solutions; and measure their success.” 

This statement allows for the scale of the 

Conservation Management Unit to be 

determined by the planner and acknowledges 

that goals can be defined for multiple scales, 

but provides little guidance for planning and 

assessment across the scales. 

The Watershed Protection and Flood 

Prevention Act (83-566), referred to as PL 

566, authorizes NRCS to work with clients at 

spatial scales greater than the property level. 

Guidance for this program is found primarily 

in the National Watershed Manual (USDA 

NRCS 2009). The tools for planning, however, 

are primarily those used for individual property 

planning. Assessment tools differ slightly in 

that outcomes are usually expressed at the 

watershed scale, often using economic variables 

(e.g., flood prevention benefits). 

The National Biology Handbook (USDA 

NRCS 2004) also considers multiple scales 

for programs to enhance wildlife habitat. 

In this document, the concepts of core 

reserves (nodes), corridors, and buffer zones 

are introduced and applied by example. In 

addition, qualitative metrics for assessing patch, 

corridor, matrix, and structural attributes are 

introduced. In discussions of these attributes, it 

is clearly stated that planning, implementation, 

and assessment may span several spatial scales. 

In spite of these precedents, NRCS policy, 

guidance documents, and reports of 

accomplishments have focused largely on 

planning and outcomes at the individual 

management unit or ranch scale. In addition, 

the dominance of the Environmental Quality 

Incentives Program (EQIP) in agency 

activities often dictates that planners work 

on Conservation Management Units that 

are defined by management unit boundaries. 

The lack of tools to assist in planning or 

quantitative assessment of ranch-, watershed-, 

or landscape-level outcomes limits the ability of 

planners to address broad-scale problems. 

Measurement of Practice effects 

There is currently no specific NRCS protocol 

or program designed to measure conservation 

effects at multiple spatial scales. The only 

dataset that covers all nonfederal rangelands 

(and all other nonfederal lands) is the National 

Resource Inventory (NRI). The NRI is a 

“longitudinal survey of soil, water, and related 

environmental resources designed to assess 

conditions and trends every five years on nonfederal US lands” (Nusser and Goebel 1997). 

The sampling framework for NRI is based on 

sampling areas (primary sampling units or 

segments) of 160 acres (64.8 ha). The number 

of segments has varied throughout the life 

of the NRI, ranging from 108 000 to over 

300 000 nationwide. Within each segment, 

three sample points are randomly located and 

various attributes are measured in them. This 

structure permits statistical inferences about 

changes in land use and practice application 

at regional, state, and national levels. The 

NRI was designed initially to detect changes 

in broad classes of land cover and use (e.g., 

cropland to pasture, or wheat to corn). Most 

NRI analyses are based on photointerpretation. 

In 2003, a special study was established for 

There is currently 

no specific 

NRCS protocol 

or program 

designed 

to measure 

conservation 

effects at multiple 

spatial scales.”

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 343 
	 	 	 	

 

 

 

	 	

	 	 	

	 	

	 	

	 	 	 	

	 	 	

	 	

	 	 	 	 	

	 	 	 	 	 	

	 	

	 	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	

	 	 	 	 	 	

	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	 	

	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	 	

	 	 	 	

	 	

	 	

	 	 	 	

	 	

	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	

	 	 	 	

tABLe 2. Examples of specific applications of landscape concepts discussed in this chapter to rangeland conservation practices discussed in 

other chapters. 

Conservation practice type Spatial heterogeneity Spatial pattern Scaling 

Brush management Shrub encroachment rates 

vary with ecological site 

Grass fragmentation with 

encroachment determines 

erosion rates 

Mosaics of shrublands and 

grasslands can be conser vation 

goals for wildlife 

Invasive species 

management 

Certain sites are more likely 

to be invaded first and serve 

as early warning 

Formation of gaps in vegetation 

may facilitate initial invasion 

When scale of infestation is 

large, it is difficult to control with 

herbicides 

Prescribed burning Fire prevalence depends on 

climate 

Fuel connectivity determines 

scale of fire spread 

Scale of fire-affected vegetation 

determines habitat values to wildlife 

Prescribed grazing Certain landforms or soils 

attract more use by livestock 

Patch grazing breaks down with 

increased stocking rate 

Increases in scale of pasture results 

in less uniform livestock distribution 

Range planting Species plantings more 

likely to be successful on 

certain soils 

Plantings lead to reduced bare 

ground connectivity and runoff 

Use of key microsites for 

establishment may lead to spread 

of seed to larger areas 

Riparian management Geomorphic valley type 

determines potential riparian 

function 

Spatial pattern of channel 

(meander width ratio) determines 

riparian vegetation states 

Upland vegetation management 

affects sediment deposition in 

riparian areas 

Wildlife habitat 

management 

Multiple plant communities 

may constitute desirable 

habitat 

Fragmentation of forest leads 

to increased predation or 

parasitism rates 

Dispersal limitation due to 

landscape fragmentation may 

override quality of local habitat 

rangelands that included field data collection. 

This will allow detection of more subtle 

changes in ecological state or condition within 

rangeland vegetation in future evaluations, but 

only at very broad scales (http://www.ncgc.nrcs. 

usda.gov/products/nri/range/2006range.html). 

There are two significant limitations to using 

existing NRI data to evaluate conservation 

practices. The first is that it is virtually 

impossible to obtain reliable information 

about the practices being applied at any 

particular random point. Although observable 

conservation practices are recorded, most of 

the more significant practices are not easily 

documented without expensive (and often 

unavailable) landowner interviews. Some 

of these data do exist, but they are housed 

in separate databases, are subject to privacy 

limitations, and generally lack sufficiently 

precise location information to relate them to 

field measurements. 

The second limitation is that the sampling 

design was intended to provide estimates 

of change in major land use and land cover 

classes at regional to national scales. The 

design was not intended to detect differences 

among ecological sites or among variations 

in the arrangement of practices within a 

landscape. Such analyses require sampling that 

is structured with respect to the distribution 

of treatments and ecological sites in particular 

landscapes. Analysis must also consider 

particular models of interactions between 

locations in a landscape. Further, consideration 

of within-site heterogeneity may require 

sampling protocols that are tailored to specific 

processes (e.g., habitat use by birds). Thus, 

we assert that NRI is insufficient to provide 

an understanding of the effects of landscape 

context on conservation outcomes. This 

assertion suggests that additional conservation 

assessment strategies are needed at local to 

regional levels. 

PLACIng ConSeRvAtIon PRACtICeS 

In A LAndSCAPe Context 

Most chapters in this volume review literature 

to provide science-based evaluations of the 

effectiveness of individual practices and several 

of them point to the importance of landscape 

context. In the last 15 yr, the number of 

344 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	

	 	

	 	 	 	

	 	

	 	

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

rangeland studies that mentioned “spatial” or 

“landscape” increased from about 100 to over 

700 per year, suggesting a growing awareness 

of the value of landscape perspectives within 

the rangeland science community. The number 

of studies referencing these terms that were 

associated with conservation practices, however, 

increased only slightly and is still a small 

number (< 40/yr). Consequently, this chapter 

does not on focus on specific practices, but 

instead offers a forward-looking synthesis of 

concepts and approaches that could facilitate 

a shift from the typically site-scale or local 

perspective toward a landscape perspective (Fig. 

1). The local perspective uses information about 

the ecological state or plant community phase 

(plant community variants within states of an 

STM) to select practices to achieve target states 

or phases, sometimes tailored to an ecological 

site. The landscape perspective additionally 

considers how the surrounding landscape— 

including its states, ecological sites, and 

connections among them—influence the effects 

of practices. Attention is also paid to spatial 

heterogeneity or patchiness within the site as a 

regulator or objective of conservation effects. 

Finally, the landscape perspective considers how 

a localized conservation effect influences the 

landscape in which it is embedded. Below, we 

offer a review of several concepts underpinning 

the landscape perspective. Examples of how 

these concepts apply to practices discussed in 

other chapters are presented in Table 2. 

Spatial Heterogeneity 

Spatial heterogeneity refers to the variability of 

properties over space including soils, vegetation, 

and process rates such as water runoff or 

erosion (Turner and Chapin 2005). Spatial 

heterogeneity within rangelands has been 

accounted for at some scales, but not others. For 

example, spatial heterogeneity due to relatively 

static variations in the landscape (reflected 

in the ecological site) and differences in past 

vegetation dynamics or “historical legacies” 

(reflected in ecological states or community 

phases) is known to influence the success 

of practices and is widely used in planning 

(Creque et al. 1999; Bestelmeyer et al. 2009). 

Different ecological sites or states/phases are 

predicted to differ in both the possible changes 

that can be observed and the success of certain 

management interventions. Different soils vary 

in their susceptibility to shrub invasion and the 

loss of perennial grasses, and in the potential for 

recovery of different grass species (Fuhlendorf 

and Smeins 1998; Hamerlynck et al. 2000; 

Wu and Archer 2005). For example, sites with 

well-developed argillic (clay-rich) horizons 

may limit invasion by both Larrea (McAuliffe 

1994) and Prosopis (Archer 1995; Miller et al. 

2001). On the other hand, the phytotoxicity of 

chemicals used to control shrubs may also be 

limited on clay-rich soils (Duncan and Scifres 

1983). Once shrubs are removed, the recovery 

of grasses in response to removal may depend 

on clay content (Bestelmeyer et al. 2006a) and 

the degree of soil interspace erosion that is 

captured in classifications of alternative states 

(Herrick et al. 2006). The linkage of this type 

of information to ESDs and maps is potentially 

very useful. ESDs, however, often do not 

clearly present information on the likelihood of 

practice success or failure among ecological sites 

or states. 

Spatial heterogeneity may also be of value in its 

own right, although considering heterogeneity 

in this way for planning is not common. It has 

long been recognized that habitat heterogeneity 

promotes biodiversity, but conventional 

measurements tend to emphasize within-habitat 

heterogeneity (such as the vertical complexity 

of vegetation at a point in space; Tews et al. 

2004) rather than landscape heterogeneity 

(the composition and arrangement of different 

habitats across a landscape). Landscape 

heterogeneity in rangelands can be produced 

by the simultaneous and dynamic coexistence 

of different plant communities due to patchy 

disturbances and corresponding asynchrony 

in successional stages among patches. Such 

patch dynamics or a “shifting mosaic” 

(Bormann and Likens 1979) can yield desirable 

properties at broad scales. For example, 

shifting mosaics in grasslands caused by grazing 

and fire disturbances can create a mix of 

structurally simple and structurally complex 

plant communities that sustain wildlife 

populations that exploit resources in different 

plant communities over time (Archibald et al. 

2005; Fuhlendorf et al. 2006). Spatiotemporal 

variation in vegetation can also be used as a tool 

to manage livestock distributions (Fuhlendorf 

and Engle 2004). Some valued wildlife 

species may even be associated with elements 

characterizing traditionally undesirable 

states, such as mesquite shrubs in semidesert 

...a forwardlooking synthesis 

of concepts and 

approaches that 

could facilitate 

a shift from the 

local perspective 

toward a 

landscape 

perspective”

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 345 
	 	 	 	

fIguRe 1. A schematic of the landscape perspective. The top panel illustrates patchiness in vegetation 

cover (dark) interspersed with sparsely vegetated areas (light). The middle panel is a map of ecological 

sites, each representing one or more vegetation states. The bottom panel illustrates the landscape within 

which the mosaic of ecological sites is embedded. These are three scales of pattern that promote a 

landscape perspective. 

grasslands (Lloyd et al. 1998; Saiwana et al. 

1998) or bare and structurally simplified areas 

in the Great Plains (Knopf 1996; Derner et al. 

2009). Thus, the presence of even persistently 

altered states in parts of a landscape can have 

management value. In rangelands, however, 

management typically seeks to maximize the 

coverage of a single, desired plant community 

and deemphasizes landscape heterogeneity 

(Fuhlendorf and Engle 2001). 

Spatial heterogeneity may be similarly 

important to the basic functioning of 

ecosystems. Shifting mosaics caused by 

localized fire and succession are seen as 

prerequisites for long-term sustainability of 

desirable forest structures (Allen et al. 2002; 

Bond et al. 2005). At a finer scale, alternating 

areas of bare ground that generate water runoff 

and vegetated patches that intercept runoff 

can sustain productive vegetation in arid 

environments where a more homogeneous 

distribution of vegetation could not be 

sustained (Noy-Meir 1973; Ludwig and 

Tongway 1995; Rietkerk and van de Koppel 

2008). Consequently, conservation and 

restoration approaches to increase cover and 

production of desired species may focus on 

maintaining or creating heterogeneity and 

specific spatial patterns in vegetation (Noble 

et al. 1997; Miller and Urban 1999). In such 

cases, the existence of patches of bare ground 

or “early seral” vegetation may not reflect 

the initiation of ecosystem degradation, as is 

sometimes assumed, but rather are necessary 

components of some ecosystems that sustain 

productivity and biodiversity. 

Spatial heterogeneity of the kind noted above 

is especially important for vegetation sampling. 

Point samples gathered without sufficient 

replication or stratification to different patch or 

community types can lead to misinterpretations 

about vegetation conditions when extrapolated 

346 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

	 	

	 	 	

	 	 	

	 	

	 	

	 	 	

	 	

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

to the management unit or landscape scale. 

Sampling that ignores spatial heterogeneity 

and replication needs is a common source of 

monitoring failure (Herrick et al. 2005). 

We conclude that explicit consideration 

of spatial heterogeneity in conservation 

planning and assessment is valuable because 

it contains information about practice effects. 

Spatial heterogeneity can also be a goal of 

practices. The information provided by 

spatial heterogeneity can only be obtained if 

our treatment of data moves from reporting 

averages or percentages to explaining the 

variability in the data as a function of ecological 

site, state, and patchiness variables. Such 

data can then be used to improve ESDs so 

that they better predict variations in success. 

Our concepts of ecological sites and states 

also need to describe spatial heterogeneity 

as an attribute of interest, rather than solely 

describing spatially averaged attributes and 

implicitly treating management for desired 

plant communities as an “all-or-nothing” 

proposition. 

Spatial Pattern 

Without direct consideration of heterogeneity, 

attributes such as soils, cover, and species 

composition become nonspatial. Spatial pattern 

within and among plant communities in a 

landscape, however, may have predictive value 

(Reitkerk et al. 2004; Barbier et al. 2006). 

Whereas spatial heterogeneity describes the 

condition of possessing dissimilar patch types 

in an area, spatial patterns communicate where 

those patch types are and how they are shaped 

and arranged. Spatial pattern can be measured 

by the density, size, shape, and location/ 

adjacency of patches and plant communities, 

typically via a geographic information system 

(GIS), spatial statistics, and derived maps 

(Gustafson 1998; Turner 2005). Spatial 

patterns are descriptors of the potential for— 

and consequences of—spatial interactions 

involving the flow of matter (e.g., water, seeds, 

or animals), disturbance (fire or erosion), or 

information (cues for habitat selection). They 

thus communicate valuable information about 

specific ecological processes that affect the 

success of practices (Turner 2005). 

Spatial patterns in vegetation give rise to 

variations in the availability of limiting 

resources or the intensity of disturbance. 

These variations subsequently alter the spatial 

patterns—a phenomenon sometimes referred 

to as “self-organization” (Watt 1947; Rietkerk 

et al. 2002). Self-organized spatial patterns 

are the consequences of feedbacks between 

the initial spatial pattern in vegetation and 

processes such as water redistribution. For 

example, a vegetation patch in a matrix of open 

ground will intercept the flow of water and 

increase infiltration and plant-available water 

resulting in increased plant growth. Increased 

plant growth may enable the patch to intercept 

more water and so on until resources from 

the adjacent bare area become limiting. As a 

consequence of feedback effects, rangeland 

landscapes often exhibit characteristic spatial 

patterns. As overall resource availability to 

the site changes (e.g., due to aridity), the 

spatial pattern may change in predictable 

ways (Rietkerk et al. 2004). Disturbances to 

patches that interfere with feedbacks also lead 

to predictable effects on spatial patterns and 

feedbacks (Ares et al. 2003; Kéfi et al. 2007). 

Large patches tend to become fragmented, 

leading to decreased resource capture, 

production, and soil degradation (Wu et al. 

2000; Bestelmeyer et al. 2006b). The decrease 

in large patches also leads to characteristic 

changes in the distribution of patch sizes in 

an area. Thus, changes to spatial patterns have 

been promoted as early warning indicators 

of rangeland degradation (and presumably 

restoration success; Kéfi et al. 2007; Scanlon et 

al. 2007). The value of pattern-based indicators 

relative to standard measures of ground cover is 

being debated (Maestre et al. 2009). 

The spatial pattern of ecological states, soils, 

and topography in a landscape governs the flow 

of resources and disturbances. The impact of 

practices on potential water yield is especially 

sensitive to the locations in a landscape where 

the practices are applied (Ludwig et al. 2005). 

For example, Wu et al. (2001) found that 

policy incentives for brush control on the 

Edwards Plateau need to clearly specify the 

optimal locations for treatment in order to 

influence water yield. Spatial specification is 

important because there are tradeoffs between 

strategies designed to increase potential forage 

productivity vs. water yield potential (Redeker 

et al. 1998). Similarly, the spatial design 

of infrastructure including the locations of 

Spatial pattern 

can be measured 

by the density, 

size, shape, 

and location/ 

adjacency of 

patches and plant 

communities, 

typically via 

a geographic 

information 

system”

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 347 
	 	 	 	

 

	 	

	 	

	 	

	 	 	

	 	

	 	 	

	 	 	

	 	

	 	

...the spatial 

pattern of 

vegetation cover 

can be even 

more important 

than the average 

amount of cover 

in determining 

runoff and 

erosion under 

some conditions.”

fences, watering points, and feeders are used 

to modify patterns of animal movement and 

forage utilization, taking into account livestock 

behavior and the template of topography and 

plant communities to which livestock respond 

(Laca 2009). In this way, the spatial locations 

of rangeland infrastructure can have a large, 

indirect impact on overall vegetation change. 

Similarly, the spread of fire in a landscape 

depends upon where the fire is initiated relative 

to the spatial arrangement of ridges and valley 

bottoms (Swanson et al. 1988) as well as the 

connectivity of fuel loads (Allen 2007). 

The spatial pattern of vegetation patches 

resulting from practices, in turn, affects 

key processes and services. For example, 

Ludwig et al. (2007a) showed that the spatial 

configuration of vegetated and bare patches had 

a significant influence on erosion and sediment 

loss in northeastern Australia. A catchment 

with a coarse-grained patch structure (few 

large patches) and 54% grass cover had 43 

times greater sediment loss than a catchment 

with fine-grained patch structure (many small 

patches) and 43% grass cover. This result 

suggests that the spatial pattern of vegetation 

cover can be even more important than the 

average amount of cover in determining runoff 

and erosion under some conditions. With 

similar amounts of total cover, the pattern of 

vegetation patches comprising this cover can 

be arranged such that they either slow runoff 

and retain sediment or allow it to leave the 

site. Larger bare patches, bare patches that are 

elongated parallel to the direction of flow, and 

bare patches that occur lower on a hill slope are 

less able to slow the movement of water and 

sediment and prevent its transfer into channels. 

Water and sediment loss rates at sites 

with intermediate values of cover close to 

“percolation thresholds” can be very sensitive to 

changes in plant cover (Davenport et al. 1998). 

Percolation thresholds describe how a small 

change in cover over a defined area can result in 

a cover type becoming connected with respect 

to a process, such as the spread of fire (Turner 

et al. 1989) or the dispersal of species through 

certain cover types (King and With 2002). 

The shape and size of patches comprising 

cover affects the critical threshold value, but in 

general the shift from fragmented to connected 

occurs at intermediate cover values (Miller 

and Urban 2000). Areas with very low cover 

are necessarily fragmented and areas with very 

high cover are necessarily connected via a 

single large patch. Spatial pattern matters most 

when cover is intermediate because of the wide 

variety of possible arrangements of patches 

(Gergel 2005). Thus, attention to changes in 

connectivity can help us understand nonlinear 

relationships between cover and ecological 

processes in some conditions (Peters et al. 

2004). 

These examples suggest that various 

measurements of connectivity (and its converse, 

fragmentation) could be used to estimate 

critical ecological processes that mediate the 

effects of practices (Debinski and Holt 2000; 

Tischendorf and Fahrig 2000; Goodwin 2003). 

Some specific measures include the frequency 

of patches of different size or weighted 

mean patch size (Li and Archer 1997), the 

aggregation index (He et al. 2000), and the 

landscape leakiness index (Ludwig et al. 2002, 

2007b) as well as several others (McGarigal 

and Marks 1995; Calabrese and Fagan 

2004). As one example, we would expect that 

decreasing connectivity of vegetation patches 

and increased connectivity of bare patches 

would be reflected in a “landscape leakiness 

index.” Increases in this index are correlated 

with reduced water infiltration and nutrient 

retention. In selecting a metric, it is important 

to link it to a conceptual model of the pattern– 

process relationship, which, in turn, should 

indicate the appropriate spatial scale at which 

the metric is measured. 

Biophysical Scaling effects 

The spatial extent of observation determines 

how we perceive natural resource problems 

as well as the practices we use to solve them. 

For example, woody plant encroachment 

in landscapes of South Texas was shown 

to be strongly scale-dependent. In areas 

encompassing multiple soils, woody plant cover 

was associated with high-clay soils and wetter 

portions of the landscape (Wu and Archer 

2005). At finer scales within upland soils, 

woody plant cover was negatively related to 

soil clay content and was unrelated to surface 

hydrology (Archer 1995; Wu and Archer 

2005). Thus, the correlates of woody plant 

dominance depend upon the spatial scale of 

measurement. Correlations with particular 

348 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

fIguRe 2. Example of a spatial hierarchy in a Chihuahuan Desert rangeland, following Table 2 and 

Bestelmeyer et al. (2009). A. Major Land Resource Area (MLRA) 42 and the Land Resource Unit MLRA 

42.2, shaded. B. A soil–geomorphic system along the Rio Grande River valley comprising relict piedmont, 

ballena, and inset fan landforms associated with the Pleistocene–Holocene entrenchment of the river. 

The inset shows detail on soil map units and watershed boundaries (US Geological Survey hydrological 

unit code 11), illustrating the repeating patterns along the north–south axis. C. Each ecological site is 

represented by a different primary color. Various states occur within these delineations, including shrub 

savanna, shrub-dominated, and shrubland. States within sites are different shades of the primary color. 

d. Patches of Bouteloua eriopoda grass occurring in the interspaces of Larrea tridentata shrubs, a spatial 

pattern in the shrub-dominated state that indicates restoration of grass cover is possible. 

variables (e.g., clay content) can change sign 

across scales due to interactions with factors 

that vary at broad scales (e.g., topographic 

position overrides the negative effect of 

clay on woody plant cover). Management 

models used to predict patterns of woody 

plant encroachment thus need to recognize 

the scale dependence of variables governing 

encroachment. 

Different animal species or groups respond 

to rangeland attributes that are measured at 

different spatial scales. For example, although 

habitat quality for many bird species focuses on 

local vegetation structure, practices designed 

to promote highly mobile wetland bird species 

should focus on the distribution of a spatially 

dispersed mosaic of sites that are used at 

different points in the annual cycle (Haig et 

al. 1998). Considering the response of insect 

communities to grazing and mowing for hay 

in tallgrass prairie, Stoner and Joern (2004) 

showed that the species diversity of generalist 

and herbivore insect guilds in prairie fragments 

was largely controlled by local (withinfragment) plant community composition. 

This suggests that practices should focus on 

plant community attributes at the local scale 

to maintain populations of these insects. The 

predator insect guild, however, responded more 

to broader-scale attributes such as the shape 

of the fragments, thereby producing indirect 

effects on the other generalist/herbivore 

guilds. Attention to both fragment quality 

and fragment shape would be important 

conservation objectives in this case. 

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 349 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

fIguRe 3. An example of ecological site concepts developed for an area near 

Dalinour Lake in the Inner Mongolia autonomous region near Xilinhot, China. The 

ecological site boundaries are defined by varying depths of sandy sediments over 

saline lake sediments that are associated with shifts in vegetation and vegetation 

change/soil processes. Because the three ecological sites are coevolving in 

response to a common process (wind-driven deposition of sand over saline lake 

sediments) they can be considered as part of the same soil–geomorphic system. 

The spatial scale of observation also 

determines our ability to perceive processes 

that link the success of local practices to 

conditions in adjacent parts of a landscape. 

This perception is important when practices 

carried out at a fine scale influence adjacent 

areas due to spatial interactions. Conversely, 

the consequences of a practice within a local 

area may depend on influences from adjacent 

areas; both of these effects are termed “crossscale interactions” (Peters et al. 2004). Clearly, 

some of the primary benefits of erosion 

control occur off-site in the form of decreased 

erosion in adjacent sites, improved water 

quality, and decreased siltation of streams 

and reservoirs (Pimentel et al. 1995; Pringle 

et al. 2006). Cross-scale interactions can also 

thwart local management efforts. Historical 

overstocking and drought initially converted 

extensive areas from perennial grassland to 

eroding shrubland until the 1950s in the 

Jornada Basin of southern New Mexico. Once 

the eroding shrublands became sufficiently 

extensive, many remaining grassland areas 

were converted to coppice-dune shrublands 

even when domestic grazers (and native 

grazers) were excluded via fencing (Peters et 

al. 2006). Grassland-to-shrubland transitions 

after the 1950s had become decoupled from 

the local processes that had previously caused 

them. Instead, they were controlled by broadscale erosion and sediment movement that 

led to local soil instability with abrasion, 

burial, and mortality of grasses, occurring 

even in ungrazed areas (Okin et al. 2009). 

In such cases, the local management of 

vegetation or soils may not be adequate to 

predict the trajectory of vegetation change, 

as is often assumed in the use of assessment 

and monitoring indicators. A characterization 

of the functioning of the broader landscape 

would be required. 

These examples indicate that in addition to 

variation in the properties of a specific land area 

(e.g., its ecological site or state) planners should 

carefully consider the landscape context within 

which an area is embedded. Furthermore, 

carefully chosen intervention points can induce 

nonlinear responses or emergent effects that are 

not predicted based on a simple linear scaling 

of the areas that are treated. One example 

might be to increase grass cover in critical 

portions of a watershed to reduce watershedscale sediment loss. In both cases, there may 

be critical spatial scales at which the effects 

of key spatial interactions can be predicted, 

determined by factors such as geomorphology, 

hydrology, or species behavior (Turner et 

al. 2001). Our ability to predict these scales 

remains limited, but planners can integrate 

expert judgment with GIS to make informed 

(and testable) predictions. 

Societal Heterogeneity and Scaling 

Consideration of biophysical scaling will often 

lead planners to look to scales larger than a 

management unit. In these cases, the identity, 

heterogeneity, and spatial arrangement of 

management units in different ownership or 

tenure must be considered. Activities on one 

management unit may have off-site effects 

on adjacent units that are unrecognized, 

or diffuse effects from multiple units may 

influence attributes of communal interest, 

such as water table depth (Swallow et al. 

2001; Standish et al. 2009). Collaborative 

approaches are thus necessary precursors to 

broad-scale practices such as fire management 

or species conservation (Sayre 2005). In order 

350 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 	 	 	

	 	 	

	 	 	

	 	

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

for multiproperty practices to be successful, 

an atmosphere of trust and cooperation is 

required, calling for careful attention to an 

inclusive process in strategic and tactical 

planning (Duff et al. 2009). Much like 

the processes that lead to patchiness and 

sustainability in vegetation, “self-organized” 

groups of interested property owners working 

with agency representatives and scientists 

(e.g., prescribed fire associations) can lead 

to successful broad-scale conservation 

efforts (Biggs and Rogers 2003). Ultimately, 

however, successful practices must foremost 

be perceived to benefit individual landowners 

(Swallow et al. 2001). As with the use 

of ecological sites and states, it would be 

useful to document the societal contexts 

within which certain practices succeed or 

fail as a means of developing more effective 

approaches to conservation planning (Paulson 

1998). 

A SPAtIAL HIeRARCHy foR 

ConSeRvAtIon PLAnnIng And 

evALuAtIon 

The preceding review makes a compelling 

case for the value of landscape perspectives in 

conservation planning and assessment, but 

how can we most effectively incorporate these 

perspectives? Some approaches, including 

the use of spatial simulation models, are too 

technically complex to be widely implemented 

at the present time. We suggest that informed 

judgment combined with GIS and selected 

use of some existing models (e.g., hydrologic 

models) provide a practical means to develop 

landscape perspectives. The concepts described 

here rely on spatial data. Such data can be 

used to detect patterns at different scales and 

then design and evaluate practices based on 

the patterns. Mapping activities—usually of 

management units, vegetation, and ecological 

Water running in a rill 

after a rainfall event within 

a desertified grassland, 

southcentral New Mexico. 

(Photo: David Toledo) 

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 351 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	 	

	 	

	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	 	

	 	

	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	

	 	 	

	 	 	

	 	

	 	 	 	 	 	 	 	

	 	

	 	

	 	 	

	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

tABLe 3. General levels of the land hierarchy discussed in this chapter, distinguishing characteristics for each level, 

approximate map scales, and analogous levels in the National Hierarchy of Ecological Units and Terrestrial Ecological 

Unit Inventory of US Forest Service. Entries in parentheses are not formal levels but are discussed in literature. This hierarchy 

mixes a spatial hierarchy (Major Land Resource Area to watershed) with elements of a classification hierarchy (ecological 

sites to patch) that can be delineated as nested spatial units. 

general level used in this 

chapter1 distinguishing characteristic Map scale uSfS2 

Major Land Resource Area An area of similar gross physiography and 

continental weather pattern 

1:3 500 000 Section 

Land Resource unit A class or area based on regional climate 

variation or geology within Major Land 

Resource Areas; may or may not be spatially 

explicit 

1:1 000 000 Section/subsection 

Soil–geomorphic system An area of similar geology and linked 

geomorphic/biotic processes that control 

landscape evolution 

1:250 000 Land-type 

association 

Watershed/Airshed An area that is internally connected by a 

dominant spatial interaction (typically water 

flow, but could be eolian soil redistribution, 

fire, or animal movement) 

~ 1:100 000, 

variable 

(hydrologic unit 

code 11) 

(Watershed) 

ecological site A class of land of similar potential vegetation, 

soil, geomorphic setting, topographic 

position, and microclimate 

1:24 000 to 

~ 1:150 000 

Ecological type 

Plant community/state An area of similar plant species composition ~ 1:5 000 to 

1:12 000 

Plant association/ 

structural stage 

Patch A discrete unit of homogeneous vegetation 

and soil surface properties, ca. 1–100 m2 

1:1 (Patch) 

1USDA NRCS (2003a) and Bestelmeyer et al. (2009). 

2Winthers et al. (2005) and Cleland et al. (1997). 

sites—are already part of planning and 

assessment process. Our recommendation is to 

plan and evaluate conservation practices with 

regard to multiple hierarchical levels of pattern 

in rangelands (Fig. 2) and forge more explicit 

connections with existing databases on soils 

and ecological sites. First, we describe a series 

of hierarchical levels in rangelands (from fine to 

broad scales) and the data that can be used to 

represent them (Table 3). 

Patches 

A patch is a relatively homogeneous area, 

often defined by local aggregations of plants 

or the absence of plants (e.g., a bare ground 

patch). Patch is a concept that can be used at 

any scale depending on the process or species 

of interest. The use of patch in this document 

is similar to the concept of the “pedon” used 

to describe a homogeneous unit of soil. Patch 

spatial patterns (the size, arrangement, and 

composition of patches) at fine scales (e.g., 

0.1–1 ha) are used to define patterns within a 

plant community that affect specific processes 

such as erosion or habitat use. Information 

about patchiness serves three functions: 1) it 

changes our predictions when compared to the 

default assumption of uniformity (e.g., patchy 

vs. uniform grazing pressure), 2) it can be an 

objective of management (e.g., using fire to 

increase habitat heterogeneity), and 3) once 

recognized, patchiness can be altered to affect 

processes (e.g., to decrease landscape leakiness). 

For example, patchiness in grazing pressure can 

produce localized changes in vegetation that 

would not be predicted assuming a uniform 

grazing distribution, including changes 

considered to be useful or to be degradation 

(Adler and Lauenroth 2000; Augustine 2003). 

Patchiness produced by grazing, fire, or their 

352 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

interaction has been promoted as a means to 

increase biodiversity (Fuhlendorf and Engle 

2001). Practices can also take advantage of 

existing patchiness; erosion control structures 

can connect vegetation patches to form large 

obstructions to overland flow (see Ludwig et 

al. 1997; Reid et al. 1999). Slash additions 

combined with seeding may initiate the 

development of grass patches in eroded areas 

(Stoddard 2006) and the new grass patches 

may then expand over time.The alteration of 

patch spatial patterns can be measured using 

a number of tools. Ground-based transect 

approaches, including gap intercept, measure 

changes in the frequency distribution of 

fine-scale bare patches (Kuehl et al. 2001; 

Herrick et al. 2005). Aerial photography or 

high-resolution satellite imagery coupled to 

image classification are used to map vegetated 

patches (Bastin et al. 2002; Laliberte et al. 

2004; Bestelmeyer et al. 2006b) and calculate 

a variety of patch metrics (Gergel and Turner 

2001). More easily, Google Earth can be 

used to detect patch patterns across the 

globe and can now be linked to traditional 

GIS shapefiles (http://earth.google.com). 

Finally, the USDA Agricultural Research 

Service (ARS) Jornada Experimental Range 

has produced simple nominal and ordinal 

indicators that capture patch spatial patterns 

and associated soil redistribution processes 

(http://jornada.nmsu.edu/sites/default/files/ 

FieldGuidePedodermPattern.pdf). The success 

of these approaches depends upon the spatial 

grain and the detectability of the patches 

involved. Measurements of patch spatial 

patterns (types, sizes, density, connectivity) 

could be explicitly defined as objectives (or 

preconditions) for state or Major Land Resource 

Area (MLRA)-level practice guidelines. 

Plant Community 

Plant communities, considered as spatial 

units, are assemblages of plant species and 

patches that exist at a particular place and 

time (Vellend 2010). Different communities 

in a landscape can be distinguished based on 

spatiotemporal shifts in the composition and 

abundance of species. Plant communities 

can be classified to states based on their 

responses to natural and management drivers 

and ecological sites may exhibit one or 

more plant communities (phases) or states 

(Bestelmeyer et al. 2009; Fig. 2). Practices 

are expected to produce or limit shifts among 

communities occurring in different states 

(accelerating practices) or produce shifts 

among communities within a state (vegetation 

management or facilitating practices; 

Stringham et al. 2003). Thus the identity of a 

community carries with it explicit predictions 

about its likely response to a practice. 

In addition to the obvious role of plant 

communities as planning tools and 

assessment strata, data on the success of 

practices could be linked to communities 

and states in the Ecological Site Information 

System. These data can be used to refine 

STMs. Classifications of plant communities 

can also be linked to the responses of key 

animal species (Holmes and Miller 2010). 

Plant communities can be mapped using 

vegetation maps based on standardized 

vegetation classifications available through 

some gap analysis programs and other detailed 

mapping efforts (http://www.natureserve.org/ 

prodServices/ecomapping.jsp). It is important 

to recognize, however, that coarse vegetation 

maps may combine several plant communities 

(and states) that are distinguished in STMs. 

Alternatively, plant communities can be 

mapped directly against a background 

ecological site layer using aerial photography 

or satellite imagery and derived spectral 

indices, often resulting in map units featuring 

associations or complexes of communities. 

Watersheds in Kalalau valley, 

Kauai, Hawaii. (Photo: 

Brandon Bestelmeyer) 

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 353 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Currently, maps of plant communities are 

seldom available so they can be produced as 

needed for setting landowner objectives and 

implementing practices. 

ecological Site 

Ecological sites are classes of land that differ 

in potential natural vegetation (Landres et 

al. 1999), historical range of variation, and 

response to disturbance as a function of 

differences in soils, landforms, and climate 

(USDA NRCS 2003a). The Terrestrial 

Ecosystem Survey of the US Forest Service 

provides land units similar to ecological sites 

(Winthers et al. 2005). Ecological sites are 

nested within climate-based classes called Land 

Resource Units (LRUs) or MLRAs (similar to 

ecoregions). In conjunction with its STM, the 

ecological site communicates the breadth of 

possible plant communities known to exist on a 

site. Even when STMs are similar, soil variations 

represented in ecological sites may influence 

the effects of management. Examples include 

the success of herbicide use with varying soil 

clay content, or variation in the success of grass 

seeding with climate variation among LRUs. 

Thus, planning and evaluation should be linked 

directly to the ecological site, and better still, to 

local information on soil/landform variations 

within ecological sites (Bestelmeyer et al. 2009). 

In this way, the classification of ecological sites 

can be updated to better reflect differences in 

ecological resilience or other responses, or the 

effects of important variations within ecological 

sites can be described. 

Ecological sites are correlated to soil map unit 

components and are represented spatially via 

soil map units of the National Cooperative 

Soil Survey (e.g., http://soildatamart.nrcs.usda. 

gov). Given the scale of soils mapping in many 

rangelands, soil map units usually have a one-

to-many relationship with soil components. 

As a result, soil map units describe soil 

complexes, associations or consociations that 

often translate to multiple ecological sites 

per soil map unit. For the purposes of initial 

stratification, many soil map units can usefully 

be classified according to the spatially dominant 

ecological site within them while recognizing 

that they are not necessarily homogeneous. 

Visual cues obtained in the field (e.g., surface 

soil color, gravel, slope) can be used to more 

precisely classify areas. 

Watersheds/Airsheds/firesheds 

Ecological site and state units in many 

landscapes are connected to one another via 

hydrology and eolian transport or potential 

fire spread. Thus, management in one unit 

will impact others in connected landscapes. 

Watershed manipulations are often designed to 

take advantage of these connections; this has 

been reviewed elsewhere (Williams et al. 1997). 

To understand hydrological consequences, 

the appropriate order (or scale) of a watershed 

(i.e., Hydrological Units of the US Geological 

Survey) should be specified, alongside an 

expectation of the response to a hydrological 

manipulation at different scales and in specific 

parts of the watershed. Such predictions are 

more common within riparian zones compared 

with upland areas affecting riparian zones 

(Goodwin et al. 1997). Informal conceptual 

models or distributed hydrologic models can be 

used to develop such predictions. The concept 

of airsheds may be especially important in 

arid zones where wind erosion and sediment 

deposition processes are important (Okin et al. 

2006). Airsheds could also assist prediction of 

the movement of smoke from prescribed fires. 

Similarly, “firesheds” have been conceived as 

the possible or expected area influenced by a 

single fire ignition as constrained by natural 

barriers, fuel, terrain, and weather within a 

given period of time (Stratton 2006). Formal 

units for the latter types of “-sheds” may not 

exist, but can be estimated from models or in a 

GIS. Accounting for physical connections is a 

key element in developing estimates of off-site 

effects of practices. 

Soil–geomorphic Systems(SgSs) 

The SGS is a new concept and refers to a 

discrete land area with a characteristic spatial 

arrangement of ecological sites (and often 

plant communities) that are linked by fluxes 

of materials, organisms and disturbances, soilforming processes, and ecological processes 

(Bestelmeyer et al. 2009; see Figs. 2 and 3). 

They are similar in scale to the landtype 

association of the National Hierarchy of 

Ecological Units (Cleland et al. 1997). Land 

areas within an SGS feature similar landscape 

organization and may encompass multiple 

watersheds or airsheds (depending on their 

scale). The interaction of management and 

soil or landscape attributes should be similar 

across an SGS. In other words, the rules 

354 Conservation Benefits of Rangeland Practices 
B.	T.	Bestelmeyer,	J.	R.	Brown,	S.	D.	Fuhlendorf,	G.	A.	Fults,	and	X.	B.	Wu 

fIguRe 4. An example of a state-and-transition model developed for the Sandy ecological site, Major Land 

Resource Area 42 and Land Resource Unit 42.2. BOER = Bouteloua eriopoda (Torr.), Other PG = other 

perennial grasses (including dropseeds, Sporobolus spp.), and PRGL = Prosopis glandulosa (Torr.). Cover 

values (%) reported are foliar canopy cover from a variety of datasets. Key portions of state narratives 

relating to spatial patterns are highlighted to illustrate their use (right column). Specific predictions for 

transitions (T) and restoration pathways (R) occur below the figure. 

governing spatial interactions and determining 

the success or failure of a management action 

in a land area are similar within an SGS and 

will differ in distinct SGSs. Additionally, 

the spatial scales at which responses will be 

manifested, and should be monitored, can also 

differ among SGSs. SGSs can be used to tailor 

management and monitoring programs to 

landscapes that are structured differently. The 

extent of SGSs can be hand-digitized in a GIS 

using a digital elevation model and geology 

maps alongside a basic knowledge of hydrology 

and geomorphology or created by aggregating 

State Soil Geographic (SSURGO) database soil 

map units (sometimes State Soil Geographic 

(STATSGO) database map units can be used). 

MLRAs and LRus 

Considerations at these scales are similar to 

those at the ecological site scale. It is useful to 

understand the location of an ecological site 

within a MLRA or LRU, given the continuous 

variations in climate that exist across the 

extent of these broad areas. Modeled climate 

products with national coverage, such as the 

PRISM model (Daly et al. 2002) can be used 

to quantify within-MLRA/LRU variations. 

The types of practices used and their outcomes 

vary strongly among MLRAs/LRUs, so it 

would be useful to assemble guidelines at 

these levels. For example, rangeland seeding 

has been recommended in the 10–14-inch 

precipitation zone (LRU) of MLRA 35 

(Colorado Plateau) but not in the 6–10-inch 

precipitation zone. 

A ModeL-BASed, LAndSCAPe 

APPRoACH to ConSeRvAtIon 

PRACtICeS 

Using the spatial data and concepts described 

above, we recommend that the following 

steps be considered to design and test the 

effects of practices and then link what we 

CHAPTER	8:	A	Landscape	Approach	to	Rangeland	Conservation	Practices	 355 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

have learned from these tests to an expanding 

database. Planning starts with collaborative 

development of a conceptual model of the 

intended effects and ends with an update to 

the model, paralleling statistical approaches 

that are advocated for adaptive environmental 

management (Ellison 1996). In this case, 

the “model” includes recognition of spatial 

heterogeneity, spatial pattern, and landscape 

context. 

define Boundaries of the Management 

Area and Critical natural Resource 

Issues 

This is perhaps the most important step to 

incorporate biophysical and societal scaling. 

Knowledge of the primary conservation 

problems, the biophysical and social 

mechanisms involved, and the scales of spatial 

interactions associated with those mechanisms 

are used to delineate the spatial extent of a 

management area. General information about 

MLRAs, LRUs, and SGSs can be used to 

identify the types of mechanisms and critical 

scales that characterize an area (based on 

patterns in soils, geomorphology, hydrology, 

and climate) and therefore, the land area that 

needs to be considered to solve a problem. 

The extent of the management area alongside 

patterns of land ownership is defined in a GIS. 

Patterns of land ownership within the focal 

area will determine what resource concerns 

can be addressed on individual properties and 

therefore, which conditions can be expected to 

improve. 

Collaborating stakeholders and planners 

then identify and prioritize natural resource 

problems and the specific locations of interest. 

The causes of the problems are identified, 

and historical perspectives on ecosystem 

conditions and drivers can help to recognize 

the key issues. Participatory mapping exercises 

(Reed et al. 2008) and workshops structured 

around general conceptual models of land 

change for an area (Reynolds et al. 2007) are 

useful approaches. The goal of this step is to 

focus limited resources on the highest-priority 

problems. 

develop Models of Conservation effects 

Soil or landform mapping is used to identify 

and locate the set of ecological sites present 

within the management area. This activity 

effectively stratifies the management area 

according to different conservation objectives 

and expected responses to practices. Each 

ecological site is then linked to a specific 

STM that describes the plant communities 

that are possible for each site and the drivers 

or interventions needed to achieve them. 

Ecological sites and STMs may already exist for 

the management area. Alternatively, ecological 

sites and STMs may need to be developed by 

project personnel in cooperation with NRCS 

and other agencies. Existing STMs must often 

be expanded to provide explicit predictions 

about practice effects. 

A primary means to develop or expand 

STMs that serves the design of conservation 

practices is to examine historical applications 

and reconstruct their effects. This can be 

accomplished in many areas via comparisons 

of historical aerial photography and sometimes 

via ground-based data or photography. Local 

knowledge on how the practices were applied 

and information on the initial state and 

ecological site are essential. It would be useful to 

store information on past effects of conservation 

practices for each ecological site and state. Such 

a database does not yet exist at the national 

level, but the Land Treatment Digital Library 

provides a model for such a database (http:// 

pubs.usgs.gov/fs/2009/3095) and similar 

databases could be developed locally. Additional 

sources of information include inventory data of 

the properties of plant communities associated 

with the same ecological sites and different 

management histories, recent monitoring data 

including responses to climate change and 

management interventions, and process-based 

studies that test for the mechanisms causing 

or constraining ecosystem responses, often 

associated with long-term research sites. 

ESDs and STMs are used to subdivide the 

landscape according to conservation objectives 

and to specify the target states or plant 

community phases for each ecological site. A 

reasonable target depends partly on ecological 

potential, which depends on soil variations 

reflected in ecological site classification (e.g., the 

depth to saline sediments strongly affects the 

potential composition of plants; Fig. 3). The 

selection of targets and practices also depends 

upon either the risk of degradation or the 

nature of restoration thresholds that must be 

356 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	

	 	 	

	 	

	 	

	 	

	 	 	

	 	 	

	 	

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

fIguRe 5. A schematic of the relationships between state transitions (boxes and arrows) in stateand-transition models (STMs) associated with different ecological sites sharing a spatially interactive 

landscape. Transitions within the hydrologically isolated STM (Gravelly ecological site) have no effect on 

other transitions. A transition within the Draw ecological site has cascading effects downslope in the linked Spatial data STMs. See text for additional explanation (from Bestelmeyer et al. 2011). on states and 

ecological sites 

are the critical 

elements needed 

to connect 

predictions to 

specific sites and 

to assess spatial 

interactions 

across a 

landscape.”

overcome to achieve the target state. STMs can 

be used to formally define predictions about the 

responses of a discrete land area to conservation 

practices (Fig. 4). Most important, the STMs 

then become the repository for information 

learned from monitoring of practice outcomes. 

STMs linked to ecological site classifications 

function as evolving libraries and the interface 

between knowledge and action. 

Although not generally available in existing 

STMs, local and landscape spatial patterns 

may be described as attributes defining atrisk community phases or alternative states 

(e.g., Ludwig and Tongway 1997; Fig. 4). For 

example, the presence of large open ground 

patches may signal an increased risk of invasion. 

Indicators of risk may also occur elsewhere in 

the landscape. For example, a head-cut gully 

several kilometers away might soon affect the 

vegetation and hydrology of an upslope area. 

Identify natural Resource goals Across 

ownership Boundaries 

When conservation objectives suggest that 

cross-boundary coordination will be needed, 

stakeholders may differ in their preferences 

and perceptions of tradeoffs. Goals must 

sometimes be negotiated alongside building of 

trust between coordinating parties. Kitchentable to community-level discussions and 

gradual consideration of the options are 

essential. 

develop Maps of ecological Sites, 

States, and Landscape Models 

Spatial data on states and ecological sites 

are the critical elements needed to connect 

predictions from STMs to specific sites 

and to assess spatial interactions across 

a landscape. Several tools currently exist 

to support this (e.g., Web Soil Survey, 

http://websoilsurvey.nrcs.usda.gov; Soil 

Web, http://casoilresource.lawr.ucdavis. 

edu/drupal/node/902). In relatively small 

areas, the solution is simple: conduct field 

assessments of vegetation and other state 

attributes (e.g., soil surface properties, 

patch patterns) alongside verification of the 

ecological sites. In large landscapes, we have 

used aerial photography and other layers 

(e.g., digital elevation models, soil maps) 

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 357 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

fIguRe 6. An example of a conservation effects assessment monitoring design within 

a brush management area treated with herbicide. The treated area was classified 

as having two ecological sites (green = Clayey, tan = Gravelly). Subsequent field 

assessments identified the Hills ecological site as an inclusion within one Gravelly 

polygon. ArcGIS was used to create a grid of systematically distributed points. 

A random number generator was used to select a subset of points for assessment 

of treatment efficacy and low-intensity monitoring (shrub mortality assessment and 

repeat photography points) in each ecological site (colored circles). One random 

subset point located in a nontarget ecological site (a draw that was misclassified in 

the spatial dataset) was rejected for use. An intensive monitoring point (outlined in 

red) in the spatially -dominant ecological site was randomly selected from the subset 

points at which standard monitoring procedures were employed. 

in GIS to produce maps of ecological states 

and ecological sites (hereafter a “state map”). 

In using this approach in arid ecosystems, 

staff members at the Jornada Experimental 

Range have delineated map units that were 

believed to be internally homogeneous with 

respect to ecological site and state, but the 

identity of the state or ecological site was 

uncertain due to data limitations. Thus, 

they used the map to structure rapid field 

assessments and subsequently attributed the 

polygons. Assessments that were coupled to 

state mapping allowed a trained technician 

to evaluate the ecological site and state of 

13 000–25 000 ha · day−1 . 

The resulting maps can be combined with 

conceptual landscape models of likely spatial 

interactions or with GIS-based landscape 

models (e.g., Soil and Water Assessment 

Tool; Di Luzio et al. 2004) to predict how 

practices will influence broader scales via 

spatial interactions. As a simple example 

of the former, a planner can consider how 

STMs operating on different ecological sites 

may be linked within a watershed via a map 

of soils and a digital elevation model (Fig. 

5; Bestelmeyer et al. 2011). Some models 

associated with hydrologically isolated 

landforms (e.g., Gravelly sites on erosional 

fan remnants, sensu Peterson [1981]) need 

not involve consideration of interactions with 

other models. On other sets of landforms, 

transitions among states are linked among 

ecological sites. A transition from a grassland 

to sparsely vegetated or bare state in a draw 

(inset fan) would result in a shift to shrubs 

and grassland species tolerant of drier 

conditions in a downslope ecological site 

(e.g., Devine et al. 1998). The sparser cover 

and increased runoff from this site, in turn, 

might lead to increased production at the 

edge of a Clayey basin floor and a shift from 

drier- to wetter-adapted grass species (e.g., 

Peters et al. 2006). Thus, in this example, a 

practice to repair the gully in the draw might 

result in both desirable and undesirable 

transitions in downslope ecological sites. 

This general set of interactions would operate 

throughout the SGS. In this way, the linkage 

of state and ecological site maps to landscape 

models can be used to delineate land units 

that require different practices and predict 

how they and adjacent units will respond. 

design Practices for Individual or 

Multiple Combined Land units 

For each land unit or for groups of land 

units that interact spatially (e.g., via 

hydrological connections), practices are 

specified to maximize the likelihood of 

successful maintenance of, or restoration to, 

the target state or phase. Where possible, an 

experimental component can be included 

using matched controls and pre-intervention 

measurements to allow a before–after–controlintervention statistical design (Block et al. 

2001), preferably over suitable periods of 

time. Decisions are made for every land unit, 

including the decision not to intervene. The 

rationale for these decisions, based on the 

STM and stakeholder goals, can be stored 

with the state map database in ArcGIS 

software. 

358 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

Monitor and update eSds 

Monitoring stratified to different land 

units can test both the effectiveness of the 

practice application and the effects of the 

practice given its successful application. 

The monitoring program should be able to 

distinguish between application effectiveness 

and effect given successful application to 

provide fair evaluations of the causes of failure. 

Stratification by ecological state, ecological site, 

and surrounding states and sites allows contextdependent tests of practice effects. There 

should also be careful consideration of the 

hypothesized response attributes and timelines 

for change. Without careful consideration of 

these design elements, monitoring programs 

are often incapable of providing valid tests of 

practice effects. 

In keeping with the collaborative nature 

of this approach, the interpretation of the 

monitoring data should be discussed among 

planners, science specialists, and stakeholders. 

Because the effects of intervention often 

unfold over long time periods and are 

influenced by short-term climate variability 

and other events, the results are sometimes not 

straightforward to interpret. The limitations of 

the data obtained at any given time should be 

recognized and interpretations can evolve with 

additional data. 

The evidence obtained is used to modify or 

revise the appropriate STM, ecological site 

classification, and local landscape or general 

SGS models. As a result, the criteria for states 

and ecological sites may be changed and the 

likelihood of success of a practice within a 

state or ecological site can be quantified. The 

attributes of state maps can be updated and 

subsequent practices modified. 

An example: What Are the Benefits 

of the Model-Based, Landscape 

Approach? 

The sequence of activities discussed above 

are new proposals, therefore we cannot 

provide direct evidence of their effectiveness. 

We can, however, provide an example of 

how they are currently being applied and 

the benefits we anticipate. The USDA ARS 

Jornada Experimental Range has worked with 

the Bureau of Land Management (BLM) 

in New Mexico to evaluate the effects of 

brush management practices that have been 

supported by both BLM and USDA EQIP 

funds as part of the Restore New Mexico 

program. Initial meetings with BLM staff 

were used to create explicit descriptions of 

the expected benefits of brush management. 

These meetings were also used to specify 

explicit hypotheses for vegetation responses in 

different ecological states and ecological sites. 

Digital maps were then used to design the 

brush control treatments. Soils, landform, and 

the pattern and cover of ecological states from 

aerial photography were interpreted according 

to STMs to identify suitable treatment areas 

within selected allotments. This selection 

procedure was based on interactions between 

BLM staff and grazing permittees. For example, 

shrubland states on thin, rocky soils are 

unlikely to yield much herbaceous response 

from brush control and were avoided in favor 

of slightly deeper soils. They used a simple 

spatial pattern indicator—the aggregation of 

perennial grasses under shrub canopies— to 

identify suitable states for treatment during 

field visits. They focused on states where 

remnant perennial grasses were distributed 

throughout shrub interspaces and in which 

vegetative recruitment could lead to rapid 

recovery (Bestelmeyer et al. 2009). 

Once brush control treatments were applied, 

the same spatial data were used to design a 

monitoring program. Spatial data were used 

to stratify plots to target ecological sites and 

states and then randomly select plots within 

target ecological sites to achieve a spatially 

balanced sample (Fig. 6). They also established 

sampling plots in areas outside of treated 

areas on the same leased properties to evaluate 

changes to herbaceous cover that may occur 

when stock numbers are redistributed to other 

pastures due to grazing deferments in treated 

pastures. The size of the monitoring units (50-m 

transects) and monitoring methods were chosen 

considering the size and distribution of remnant 

grass cover patches. Line–point intercept, 

gap intercept, and belt transects (Herrick et 

al. 2005) are being used to monitor trends in 

herbaceous plant recovery and shrub mortality 

and recruitment, tailored to the expected 

responses of the brush control treatment. We 

have planned to obtain repeated readings over 

a 12-yr time horizon; desert grassland recovery 

is slow at best. Across the body of brush control 

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 359 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	

	 	 	 	

	 	

	 	 	

	 	 	

	 	

	 	 	

	 	 	

planning data 

entered by field 

staff need to be 

integrated with 

spatial data and 

followed up by 

monitoring in 

treated areas as 

well as offsite 

areas”

treatments (over 50 sites), herbaceous response 

will be modeled as a function of soil properties, 

landform, landscape position, size of the 

treatment, weather, and posttreatment grazing 

management. Thus, this design integrates 

data on both local and landscape factors to 

design and test the practice. The need for local 

stratification, randomization, and measurements 

with respect to the specific treatments indicates 

the need for a carefully designed monitoring 

and spatial information system. The results of 

the Restore New Mexico monitoring will be 

used to update STMs for the area. 

We anticipate that the model-based approach 

will produce results and benefits that have 

heretofore not been achieved. In spite of a 

long history of brush management activities, 

the lack of an ecological site and STMguided experimentation and monitoring has 

circumvented a quantitative understanding of 

the conditions under which brush management 

succeeds and the characteristics of success. 

We anticipate that the Restore New Mexico 

monitoring will be able to more precisely target 

brush management activities to achieve desired 

results in the future. Model-based conservation 

planning might save millions of dollars that 

would be spent on ineffective treatments in 

southern New Mexico alone. 

ReCoMMendAtIonS 

Incorporate Landscape Perspectives into 

Conservation Planning 

How can agencies and conservation planners 

implement model-based approaches such as 

those discussed above? First, it is unreasonable 

to assume that all conservation planners will 

be able to access, manipulate, and store spatial 

data and process models underpinning the 

model-based approach. Expertise in GIS, 

remote sensing, and model implementation 

needs to be available to planners given a 

general knowledge of the uses of these tools. 

Such support within NRCS could occur 

initially via training and collaborations with 

other action and science agencies and with 

academic partners to make spatial data available 

in the planning process. The production of 

training materials and simple Web-based tools 

should be a priority. There should also be 

institutional support within NRCS to make 

expertise in GIS, remote sensing, and model 

implementation available to planners. Such 

support within NRCS could occur via spatial 

data specialist positions; similar expertise 

already exists in support of soil surveys. 

Second, agencies alongside academic programs 

at universities should invest in longer-term 

training in landscape ecology and related tools 

and concepts, particularly GIS, hydrology, and 

soils/geomorphology. Most programs already 

emphasize elements of this training, but these 

elements are seldom integrated with ideas 

including ecological sites, STMs, monitoring, 

and approaches to specific rangeland 

practices. There is a clear need to develop 

integrative courses and texts that link the more 

disciplinarily specialized bodies of knowledge. 

Third, administrative changes are needed in 

the development of conservation plans to 

include systematic consideration of off-site 

effects. Modifications to the Conservation 

Practice Physical Effects planning document 

to add an “off-site effect” category would 

be one approach. Historically, conservation 

planners incorporated landscape perspectives 

via rules of thumb such as “look across the 

fence to see what is coming at you,” but 

administrative requirements would ensure that 

spatial interactions are taken into account when 

needed. Finally, there must be an institutional 

structure within which STMs can be updated. 

This is a critical step if we are to learn from the 

study of conservation effects. 

develop Landscape Approaches to 

Conservation effects Monitoring 

Structured monitoring should be part of the 

budget for broad-scale conservation practices. 

Clear guidelines for design, institutional 

support for the implementation of the design, 

and a mechanism to incorporate what is 

learned from the monitoring within ESDs 

should be developed (as illustrated above). 

To accomplish this, planning data entered by 

field staff need to be integrated with spatial 

data and followed up by monitoring in 

treated areas as well as off-site areas. Careful 

monitoring design, including stratification and 

sufficient replication, must be supported if the 

monitoring is to be useful. 

There should be a system in place to document 

the cumulative benefits of practices with 

360 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 	 	 	 	

	 	 	 	

	 	 	

	 	 	 	 	 	

	 	 	 	

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

regard to large areas, long timeframes, and a 

broader range of services (Tanaka et al. 2005). 

Although there have been few attempts to 

project cumulative benefits of widely dispersed, 

small-scale restoration projects, Kondolf et al. 

(2008) developed a relatively simple approach 

to prioritizing projects to achieve maximum offsite benefits. This approach allows for an explicit 

statement of the assumptions about the delivery 

of ecosystem services offsite via the application 

of traditional site-specific management practices. 

Similar approaches merit closer consideration by 

NRCS. 

develop a Spatial Information Support 

System and Associated tools 

A spatial information system designed to 

support both the planning process and the 

design of monitoring programs requires a 

spatial database of maps and tools that can be 

used by planners, perhaps with assistance by 

spatial data specialists. The Web Soil Survey 

(http://websoilsurvey.nrcs.usda.gov) already 

provides a remarkable array of data and tools 

that can be used in the planning process and 

mirrors some of the steps described above. 

Soil Web, another recent online soil survey 

tool, links soil maps to Google Earth imagery 

(http://casoilresource.lawr.ucdavis.edu/drupal/ 

node/902). It should be possible to expand 

Web Soil Survey and Soil Web by linking their 

data layers to others via Web-based distributed 

networks (e.g., climate from PRISM, http:// 

www.prism.oregonstate.edu, or fire locations 

from GeoMAC, http://www.geomac.gov) and 

to project-level data housed in local servers. 

The Landscape Toolbox (http:// 

landscapetoolbox.org) provides an example 

of how various tools can be linked to specific 

A restoration treatment using 

woody debris to create 

vegetated bands (patches) 

in eroded soils of Big Bend 

National Park, Texas. (Photo: 

Brandon Bestelmeyer) 

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 361 
	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

Recently (left) and less recently 

(right) burned tallgrass prairie 

in east-central Kansas. (Photo: 

Brandon Bestelmeyer) 

natural resource problems. The Landscape 

Toolbox employs an analytical framework to 

link specific problems to tools that provide 

information at different spatial scales. Projects 

such as the Landscape Toolbox highlight 

the value of linking disparate information 

sources and reveal the need for standards of 

information transfer (e.g., data formats and 

metadata) in the use of spatial data from 

multiple sources. 

Maps that Facilitate Evaluation of 

Landscape Context. Such maps can be 

based on stable physical attributes derived 

from digital elevation models, including 

drainage networks, flow directions, and flow 

accumulation developed using GIS-based 

models. In addition to these more static 

attributes, climate data could be used to 

evaluate patterns of wind direction and velocity 

and fine-scale patterns of precipitation (e.g., 

via Doppler radar maps). Fire extents and 

characteristics can also be mapped. Readily 

available, preprocessed spectral data from the 

MODIS satellite can be used to document 

variations in production at landscape scales 

(e.g., maps of the Normalized Difference 

Vegetation Index with 250-m resolution). 

These data, in conjunction with visual 

interpretation of land surface characteristics 

in aerial imagery, can allow a trained specialist 

to recognize several important physical spatial 

processes. 

Simple Spatial Models. Expanding upon the 

maps developed above, hydrological, fire or 

other models can be used to simulate landscape 

processes using digital elevation models, soil 

maps, vegetation maps, and other data as 

inputs (e.g., Stratton 2006). Such models 

362 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

could be used to target, for example, particular 

portions of a watershed for treatment based 

on the sensitivity of water output to a change 

in plant cover expected in an area (Wu et al. 

2001). More complex process-based models are 

being developed, but they are often designed 

for the purpose of exploring the effects of 

specific processes and it is unlikely that these 

models will be useful for guiding management 

anytime soon. On the other hand, simple 

models with user-friendly interfaces could 

enable the planners to develop alternative 

spatial designs for practices and compare the 

anticipated effectiveness of these options. 

Spatial Pattern Indicators. Such indicators 

would be used to evaluate differences in patch 

or landscape pattern (e.g., patch size, patch 

density, connectivity, landscape leakiness) 

that are important mediators of ecological 

processes including water redistribution, 

erosion, or wildlife movement. Calculation 

of these indicators usually relies on classified 

satellite data or aerial imagery and the type 

of classification used depends on the process 

in question. Consideration of spatial pattern 

need not involve GIS-based calculations, 

however. Simple ground-based indicators 

for field inventory are available for upland 

(Tongway and Hindley 2004; Herrick et 

al. 2005; http://jornada.nmsu.edu/sites/ 

default/files/FieldGuidePedodermPattern. 

pdf) and riparian areas (Prichard 1998). The 

specific indicators required will vary with the 

ecological process involved. As with vegetation 

composition, the measurements will change 

in response to practices and thus will be useful 

for monitoring. Assistance from spatial data 

specialists should be made available to support 

remote-sensing based monitoring. Protocols 

for the selection and use of specific indicators 

could be specified for MLRAs or groups of 

MLRAs that share similar ecological processes. 

Research support will be needed to identify 

useful spatial pattern indicators and to interpret 

their values. In this vein, reference values 

should be associated with descriptions of state 

and plant community phases. 

Link Results to a national ecological 

Site database 

Although the spatial information support 

system discussed above facilitates conservation 

planning, we must also consider where to 

house the monitoring data and the lessons 

learned from them. Similar to existing 

databases such as the National Soil Information 

System and the Ecological Site Information 

System, raw monitoring data at points and 

the interpretations derived from those data, 

respectively, will likely require separate, but 

linked databases. Current revisions to database 

structures planned with the reorganization of 

responsibilities for production of ESDs within 

NRCS provide an opportunity to consider 

how monitoring data on conservation effects 

could be linked to ESD interpretations. In 

any case, a database system and process to link 

CEAP monitoring data to ESDs must be a high 

priority within NRCS. 

Support Research to Better Integrate 

Concepts, tools, and Applications 

Development of a systematic approach to 

conservation planning at the landscape level 

would benefit from research that addresses how 

to integrate information from landscape scales, 

spatial patterns, models, conservation planning 

field data, and NRI and other monitoring 

data. Specifically, such research would illustrate 

how field measurements should be gathered 

so that they can be scaled up or integrated 

with models and spatial data from broader 

spatial scales. Case studies centered on specific 

landscapes or MLRAs, supported by the USDA 

National Institute of Food and Agriculture 

and USDA Conservation Innovation Grants, 

could be used to explore how to bring together 

the variety of tools and approaches. Research 

is also needed to determine how best to scale 

up interpretations of conservation effects 

to state, regional, and national levels. Case 

studies illustrating how real-world conservation 

planning is linked to landscape research could 

provide an effective assessment of the benefits 

of the landscape perspective. 

KnoWLedge gAPS 

We identify the following knowledge 

and administrative gaps that need to be 

overcome in order move forward with our 

recommendations. 

1.	� ESDs and STMs need refinement and 

elaboration so that they contain explicit 

predictions about how plant communities 

and dynamic soil properties are assumed 

CHAPTER 8: A Landscape Approach to Rangeland Conservation Practices 363 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

to change as a function of conservation 

practices. For predictions involving spatial 

interactions, levels that aggregate multiple 

ecological sites (such as SGSs) will need 

to be specified and carry the predictions. 

These predictions become the hypotheses 

for monitoring efforts and tests of them are 

used to update ESDs and refine our use of 

practices. 

2.	� The lack of synthetic models at the level 

of MLRAs or LRUs is a clear limitation 

to developing consistent ESDs and STMs 

across the United States. Such models are 

needed to develop comparisons among 

different ecological sites and regions (e.g., 

the comparative likelihood of success or 

the magnitude of an effect in different land 

areas) and to represent spatial interactions at 

landscape scales (e.g., wildlife populations 

that cross multiple ecological sites). 

3.	� Readily available maps of ecological 

sites, and especially ecological states, are 

generally not available to assist planning. 

Without maps that are connected to ESDs 

and STMs, planners will find it difficult to 

use these tools. 

4.	� Models for spatial interactions in 

landscapes that specify how conservation 

practices in one state/ecological site 

mapping unit should affect the states 

of adjacent mapping units are poorly 

developed. This will require creative 

research and modeling coupled to field 

studies. 

5.	� Spatial pattern indicators that aid in 

predicting the trajectories that states will 

take under different conservation practices 

are seldom available when they could be 

useful. Addressing this gap will require the 

integration of pattern analysis, using fieldor image-based approaches, coupled to 

monitoring of conservation effects. 

6.	� We lack administrative and database 

mechanisms to update ESDs and STMs. 

If the ESDs and STMs are not improved 

as a function of the monitoring tests, then 

learning cannot occur and the efficiency of 

conservation practices will not improve. 

ConCLuSIonS 

Our assessment indicates that landscape 

perspectives and applications are needed to 

promote long-term success and effectiveness 

of conservation practices on rangelands. 

A large body of literature supports the 

utility of a landscape perspective (e.g., 

Naveh and Lieberman 1984; Turner et al. 

2001). We reviewed the implications of 

spatial heterogeneity, spatial pattern, and 

spatial scaling for the design of practices 

and interpretation of conservation effects. 

Spatial heterogeneity is used to understand 

why a practice succeeds or fails in areas of 

differing climate, soil, and spatial context. 

Spatial heterogeneity can also be a primary 

goal of practices, for example, by supporting 

the varying habitat elements used by animal 

species. Spatial patterns are used to indicate 

critical processes that are not reflected in 

other measures, such as connectivity for 

wildlife movement or runoff and erosion 

potential. Patterns too can be a conservation 

objective (e.g., wildlife corridors or areas of 

low landscape leakiness). Spatial scaling is 

used to understand the dimensions of the 

land area over which spatial interactions 

link practices in one place to effects in other 

places, and conversely, how characteristics 

of the landscape affect the local success of a 

practice. 

Landscape perspectives encompassing 

spatial heterogeneity, pattern, and scale are 

increasingly being connected to practical tools 

that can be used by conservation planners. 

Such tools include indicators, classifications 

and maps of ecological sites and states, and 

hydrologic models. These tools can be used 

both to help design practices and to design 

the monitoring programs that evaluate their 

effects. A spatial hierarchy focuses attention on 

the data needed at each spatial scale governing 

ecological processes of interest. In order of 

decreasing scale, MRLAs, SGSs, watersheds, 

ecological sites, plant communities, and 

patches each relate to processes governing the 

management of rangelands. Furthermore, 

consideration of societal information such 

as land ownership is usually needed at broad 

scales. Each of these data sources can be 

consulted in a systematic way, which we 

described in six steps, to design and evaluate 

conservation practices in a landscape. 

We recommend that conservation practitioners 

consider several scales of spatial pattern and related 

spatial processes, including cumulative effects, each 

364 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

B. T. Bestelmeyer, J. R. Brown, S. D. Fuhlendorf, G. A. Fults, and X. B. Wu 

time a practice is applied. This synthesis indicates 

that a systematic approach to planning that 

incorporates landscape perspectives would, in 

many cases, lead to more effective interventions 

by 1) recognizing indicators foretelling the 

likelihood of success; 2) targeting interventions 

to ecological sites, states, and spatial contexts in 

which success is most likely; and 3) maximizing 

(and measuring) the cumulative, positive effects 

of practices over the long term at broad spatial 

scales. 

Although some of the tools and approaches 

supporting landscape perspectives are already 

used by conservation planners, the development 

of others will require a scientific and institutional 

investment by the federal government and 

support by universities and funding agencies. 

Spatial data information systems should 

be developed that link maps, models, and 

pattern-based metrics to support planning and 

monitoring design. Databases are needed to 

house the resulting data. The interpretations of 

these data should be linked to ESDs. Foremost, 

we must invest in training and research to instill 

an understanding of the concepts and a capacity 

for reasoning about landscape processes (e.g., 

Gergel and Turner 2001). We suspect that such 

investments would pay for themselves, and then 

some, by improving conservation effectiveness 

in the millions of acres of rangelands that will be 

treated in years to come. 

Acknowledgments 

Thanks to Jason Karl, Jeff Herrick, Kris 

Havstad, David Briske, Tony Svejcar, and two 

anonymous reviewers for comments on this 

chapter. Laura Burkett, Caiti Steele, and Jeff 

Herrick have been partners in developing many 

of these perspectives. These perspectives are 

based in part on research funded by USDA 

CSREES National Research Initiative grant 

2008-35320-18684 to BTB. 

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CHAPTER 

9 

A Social and Economic 

Assessment of Rangeland 

Conservation Practices 

John A. Tanaka,1 Mark Brunson,2 and L. Allen Torell3 

Authors are 1Department Head and Professor, Department of Renewable Resources,
 

University of Wyoming, Dept 3354, 1000 E. University Ave, Laramie, WY 82071, USA;	
 

2Professor and Department Head, Department of Environment and Society,
 

5215 Old Main Hill, Utah State University, Logan, UT 84322-5215, USA; and
 

3Professor, Department of Agricultural Economics and Agribusiness,
 

New Mexico State University, Las Cruces, NM 88003, USA.
 

Correspondence: John A. Tanaka, Department of Renewable Resources,
 

University of Wyoming, Dept 3354, 1000 E. University Ave, Laramie, WY 82071;
 

jtanaka@uwyo.edu
 

Reference to any commercial product or service is made with the understanding 

that no discrimination is intended and no endorsement by USDA is implied 

371
 
	 	 	 	

	 	 	 	

	 	 	

	 		

	 	 		

	 	

	 	 	 		

	 		

The failure to include 

social and economic 

“

nonmarket values 

in decision-making	 

and analysis will 

likely undervalue the	 

net benefits of our 

nation’s investments 

in conservation.”

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9 A Social and Economic 

Assessment of Rangeland 

Conservation Practices 

John A. Tanaka, Mark Brunson, and L. Allen Torell 

IntRoduCtIon 

Rangelands provide a wide variety of ecosystem 

goods and services, and the conservation 

practices implemented on them produce a 

variety of direct and indirect economic and 

social effects. Basic ecological relationships 

and varying degrees of natural resource 

management determine the magnitude and 

quality of goods and services produced. Society 

determines what the relative values of these 

goods and services are at any particular location 

and time (Fox et al. 2009). 

In this chapter, we examine the literature 

related to the economic and social aspects 

of ecosystem services impacted by the 

conservation practices of the Natural 

Resources Conservation Service (NRCS) of 

prescribed grazing, prescribed burning, brush 

management, upland wildlife habitat, riparian 

management, and range planting. In addition, 

we examine the social and economic aspects 

of invasive species management that cross 

different conservation practices. At the time 

of this synthesis, invasive species management 

was not a specific conservation practice, but 

the NRCS recently created a new conservation 

practice titled Herbaceous Weed Management 

that is evaluated in a separate chapter of this 

document. Valuation of ecosystem goods 

and services potentially impacted by the 

specified conservation practices, particularly 

those services for which markets do not exist, 

is also examined. Understanding valuation 

methodology is important in evaluating 

conservation practice implementation 

and funding decisions. In some cases, the 

nonmarket valued ecosystem goods and services 

are those most valued by society. 

The reason for estimating some measure of 

value for ecosystem goods and services is that 

landowners and managers need to evaluate 

trade-offs for decision making (e.g., Maguire 

and Justus 2008; Nelson et al. 2009). One 

way to make this evaluation workable is to 

put all the resources in the same units, and the 

assignment of monetary value is one way to 

accomplish this. However, the concept that an 

ecosystem good (e.g., an endangered species) or 

service has an intrinsic value that is “priceless” 

or “infinite” does not serve decision makers 

well when choices have to be made. The failure 

to include social and economic nonmarket 

values in decision-making processes will likely 

lead to undervaluing the net benefits and 

lead to inefficient allocations of our nation’s 

investments in conservation. 

The NRCS has recognized that ecosystem 

goods and services are directly and indirectly 

affected by the various conservation practices 

that they implement on rangelands. In the 

description of each conservation practice, 

the purposes describe the expected benefits 

or outcomes of practice implementation. 

Additionally, for each conservation practice, 

a physical effects worksheet is published that 

more specifically describes the benefits and 

outcomes. Both of these are on the electronic 

Field Office Technical Guide (http://www. 

nrcs.usda.gov/technical/efotg) sections of 

the NRCS website. This is shown in the 

descriptions of conservation practices and 

in the economic analysis of benefits and 

costs. In examining the conservation practice 

descriptions, there are a variety of different 

ecosystem goods and services listed as being 

positively or negatively impacted by the 

different practices. Table 1 shows a list of 

potential goods and services that can come 

from rangelands as currently recognized by 

the NRCS. As shown, there are many facets 

of each general good or service, each of 

which can have its own effect on the quality 

Western Juniper expansion, 

Dufur Wildlife Management 

Area, Oregon. (Photo: John 

Tanaka) 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 373 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

tABLE 1. NRCS estimated impacts of different conservation practices on different ecosystem goods and services. 0 = not applicable, 

1 = neutral, 2 = slight impact, 3 = moderate impact, and 4 = substantial impact. Parentheses indicate a negative impact. Adapted from 

NRCS Physical Effects Worksheets for each conservation practice (available at http://nrcs.usda.gov/technical/efotg as of March 2008). 

Brush 

management 

Prescribed 

burning 

Prescribed 

grazing 

Range

planting 

upland wildlife 

habitat Riparian 

Soil—erosion 

Sheet and rill 2–4 2–4 3–4 3–4 3 2–3 

Wind 2–4 2–3 3–4 3–4 3 2–3 

Ephemeral gully 2–4 2–4 3–4 3–4 3 2 

Classic gully 2–4 2 2–3 2–3 2–3 0 

Stream bank 2 2 2–4 2–3 2 3–4 

Shoreline 2 2 2–4 2–3 2 3–4 

Irrigation induced 0 0 2–3 0 0 0 

Mass movement (2–3) 0 1 1 2 2 

Road, roadsides, and construction sites 0 0 0 0 0 0 

Soil—condition 

2–4 2 3–4 3–4 1 3–4 

2–4 2–3 3–4 3–4 0 2–4 

(2) 0 2–4 3–4 0 3–4 

0 0 0 2 0 0 

2–3 (2) 2–3 2 0 2–3 

0 2 2–3 2–3 0 2–4 

0 2 2–3 2–3 0 2–4 

0 2 2–3 2–3 0 2–4 

0 2 2–3 2–3 0 2–4 

0 2 2–3 2–3 0 2–4 

0 2 2–3 2–3 0 2–4 

(2) 2 2–3 2–3 0 2–3 

2–3 2 3 2–4 0 2–3 

Water—quantity 

Rangeland hydrologic cycle 2–4 2–4 2–4 3–4 0 2–4 

(2–3) 0 1 1 0 2–3 

2–3 2 2–3 2–3 (3) (3) 

(2–3) 0 2 1 2–3 2–3 

0 0 0 0 0 0 

2 2 2 2 0 0 

0 0 (2) 0 0 0 

2–3 0 2–3 2–4 0 0 

2–4 2 2–3 2–4 2 2–3 

2–4 2 2 2–3 2 4 

2 0 1 1 0 0 

Insufficient flows in water courses 2–4 0 1 2–3 2 2–3 

Water—quality 

(2) 0 2–3 2–3 0 2–3 

0 2 2 2–3 0 4 

(2) 0 2 2 0 4 

1 0 2 2 0 2 

1 0 2 2 0 2–3 

1 0 0 0 0 0 

(2) 0 2–3 2–3 0 2–3 

1 2–3 2 2 0 4 

2–4 2 2–4 2–4 2–3 3–4 

organic matter depletion 

Rangeland site stability 

Compaction 

Subsidence 

Contaminants 

• Salts and other chemicals 

• Animal waste and other organics—n 

• Animal waste and other organics—P 

• Animal waste and other organics—K 

• Commercial fertilizer—n 

• Commercial fertilizer—P 

• Commercial fertilizer—K 

• Residual pesticides 

damage from sediment deposition 

Excessive seepage 

Excessive runoff, flooding, or ponding 

Excessive subsurface water 

drifted snow 

Inadequate outlets 

Inefficient water use on irrigated land 

Inefficient water use on nonirrigated land 

Reduced capacity of conveyances by sediment

deposition 

Reduced storage of water bodies by sediment

accumulation 

Aquifer overdraft 

In groundwater 

• Harmful levels of pesticides 

• Excessive nutrients and organics 

• Excessive salinity 

• Harmful levels of heavy metals 

• Harmful levels of pathogens 

• Harmful levels of petroleum 

In surface water 

• Harmful levels of pesticides 

• Excessive nutrients and organics 

• Excessive suspended sediment and

turbidity 

374 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

J. A. Tanaka, M. Brunson, and L. A. Torell 

tABLE 1. continued. 

Brush 

management 

Prescribed 

burning 

Prescribed 

grazing 

Range

planting 

upland wildlife 

habitat Riparian 

• Excessive salinity 2 0 2–3 2 0 2 

• Harmful levels of heavy metals 2 2 2 2–3 0 2–3 

• Harmful temperatures 1 1 1 1 1 2–3 

• Harmful levels of pathogens 0 0 2 2 0 3 

• Harmful levels of petroleum 0 0 0 0 0 2 

Air—quality 

Particulate matter less than 10 μm in diameter 

(PM 10) 0 2–3 2–3 2–3 2–3 2 

Particulate matter less than 2.5 μm in diameter 

(PM 2.5) 0 2–3 2–3 2–3 2–3 2 

Excessive ozone Neutral 1 1 1 1 1 

Excessive greenhouse gas 

• Co2 (carbon dioxide) 0 3–4 2–3 2–3 2 2–3 

• n20 (nitrous oxide) 

• CH4 (methane) 

Ammonia (nH3) 

Chemical drift 

objectionable odors 

Reduced visibility 

undesirable air movement 

Adverse air temperature 

0 

0 

0 

(2–3) 

0 

0 

0 

(2–3) 

0 

(3) 

0 

0 

(2) 

1 

0 

(2–3) 

0 

0 

0 

0 

1 

2–3 

0 

0 

0 

0 

0 

0 

0 

2–3 

0 

0 

0 0 

0 0 

0 0 

0 0 

0 0 

2 0 

2–3 0 

2–4 1 

Plants—suitability 

Plants not adapted or suited 3–4 3–4 3–4 4 3–4 3–4 

Plants—condition 

Productivity, health, and vigor 2–4 4 4 4 3–4 4 

threatened or endangered plant species 

• Plant species listed or proposed for listing

under the Endangered Species Act 

• declining species, species of concern 

noxious and invasive plants 

Forage quality and palatability 

Wildfire hazard 

1 1 1 1 1 1 

1 1 1 1 1 1 

3–4 3–4 3–4 3–4 3–4 3–4 

3–4 4 3–4 4 3–4 3–4 

3–4 4 2–4 0 0 0 

Animals—fish and wildlife 

Inadequate food 2–4 2–4 2–4 2–4 4 3–4 

Inadequate cover/shelter 2–4 2–4 2–4 2–4 4 3–4 

Inadequate water 0 0 0 0 0 2–4 

Inadequate space 

Habitat fragmentation 

Imbalance among and within populations 

threatened and endangered fish and wildlife 

species 

• Fish and wildlife species listed or

proposed for listing under the

Endangered Species Act 

• declining species, species of concern 

2–4 3–4 3–4 3–4 4 2–4 

2–4 3–4 3–4 3–4 3–4 2–4 

2–4 2–4 2–4 2–3 4 2–4 

1 1 1 1 3–4 1 

1 1 1 1 3–4 1 

Animals—domestic 

Inadequate quantities and quality of feed and

forage 304 4 4 4 2–3 3–4 

Inadequate shelter (2–3) (2) 2–4 0 0 0 

Inadequate stock water 0 0 0 0 0 0 

Stress and mortality 2–4 2–3 3–4 2–4 0 3–4 

Human—economics 

Land—change in land use 

Land—land in production 

Capital—change in equipment 

Capital—total investment cost 

0 0 0 2–4 0 2–4 

0 3 0 4 0 2–4 

3 2 2 2 2 2

2–4 2 0 3 2 4 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 375 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

tABLE 1. continued. 

Brush 

management 

Prescribed 

burning 

Prescribed 

grazing 

Range

planting 

upland wildlife 

habitat Riparian 

Capital—annual cost 1–3 1 1 0 1 2 

Capital—credit and farm program eligibility Situational Situational Situational Situational Situational Situational 

Labor—labor 1–3 2–3 2–3 2 2–3 2–3 

Labor—change in management level 1–3 2 2 2 Negligible 2 

Risk—yield (3–4) (2) (2–3) 0 (2–3) 0 

Risk—flexibility (3–4) 3 2–3 2 2–3 0 

Risk—timing 4 4 4 4 0 0 

Risk—cash flow 2–4 (2) (2–3) 2 2 2–3 

Profitability—change in profitability 2–3 2–3 2–3 2 (2) Situational 

Human—cultural 

Cultural resources and/or historic properties 

present or suspected to be present 2–4 2–4 0 2–4 0 (2–4) 

Human—energy 

depletion of fossil fuel resources No effect (2–3) 0 0 2 (2–3) 

underutilization of nonfossil energy resources (2) 2 0 0 0 0 

and quantity of the good and service and its 

relative value to society. 

Table 1 shows the expected change in the 

ecological or societal parameters from six 

main conservation practices evaluated in this 

document (invasive plant species management 

is not included). An examination of Table 

1 indicates the ecosystem good or service 

followed by the expected level of impact from 

each of the conservation practices. While 

some goods and services listed are considered 

“bads,” such as soil erosion (services can be 

either positive or negative, and those with 

negative outputs are called “bads” as opposed 

to “goods”; they are the outputs on which 

humans either place positive or negative 

values), the numbers indicate whether 

the conservation practice will minimize 

(positive numbers) or accelerate (numbers in 

parentheses) soil erosion. For example, in their 

critique of the ecological impacts of ranching, 

Freilich et al. (2003) identified some of the 

potential benefits that may arise from proper 

livestock management and also factors that 

need to be mitigated as those practices are 

implemented. 

To put the values from Table 1 in context, 

Table 2 shows the number of hectares treated 

by each of the major conservation practices 

from 2004 to 2008 by state and rangeland 

region. Table 3 shows the conservation 

practice expenditures from all NRCS 

programs by state from 2005 to 2009. From 

an economic point of view, the scale of 

the practices being implemented over this 

period determines the potential size of the 

impact. Different states use some practices 

more than others (Tables 2 and 3) with 

expected differences in goods and services 

produced. The differences in expenditures 

by state and conservation practice (Table 3) 

may be due to a variety of factors, including 

local preferences, acceptability, and needs. 

Expenditures by state will also vary based 

on the amount of private rangeland and 

the willingness of those landowners to 

participate in NRCS programs. Expenditures 

in conservation practices from 2005 to 2009 

(Table 3) in aggregate provide some indication 

of how practices have been implemented in 

different states. Table 4 shows the total annual 

expenditures for the conservation practices 

from 2005 to 2009. Since NRCS conservation 

programs are cost-share programs, these 

expenditures indicate only the government’s 

share of the total investment in conservation. 

These figures also illustrate that federal 

expenditures have generally been increasing 

for these conservation practices over the 5-yr 

period. The question being asked is, “Do the 

net societal benefits, including both market 

and nonmarket values, from these practices 

offset the known costs?” 

In this chapter, we discuss the various ecosystem 

goods and services impacted by these seven 

376 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	

	 	

	 	

	 	

	 	 	 	

	 	

	 	 	 	

	 	

	 	

	 	 	 	

	 	 	 	 	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

tABLE 2. Area of the main conservation practices implemented in rangeland states from 2004 to 2008. 

Location 

Brush 

management (ha) 

Prescribed 

grazing (ha) 

Range planting 

(ha) 

Prescribed 

burning (ha) 

Riparian herbaceous 

cover (ha) 

upland wildlife habitat 

management (ha) 

national 1 590 489 32 682 716 519 881 619 786 15 781 21 860 411 

Rangeland region 

Central 1 046 181 16 780 643 257 038 336 161 8 630 12 185 583 

West 410 656 13 815 577 260 263 34 660 3 722 6 980 085 

West state 

Alaska 384 1 416 984 0 3 0 184 083 

Arizona 34 413 2 607 496 5 228 3 148 59 332 011 

California 28 491 588 229 6 924 6 197 511 152 638 

Colorado 25 196 2 245 726 88 363 4 119 154 668 710 

Hawaii 4 136 23 822 123 1 330 0 57 358 

Idaho 8 996 239 791 3 359 1 790 193 180 574 

Montana 37 1 046 954 94 561 9 406 31 735 365 

nevada 7 196 59 813 1 733 0 29 7 934 

new Mexico 239 599 2 630 730 15 684 6 014 12 3 202 863 

oregon 12 061 290 328 5 006 987 2 619 222 809 

utah 27 561 423 301 37 417 140 3 217 394 

Washington 5 65 912 433 103 74 124 209 

Wyoming 22 581 2 176 491 1 432 1 421 38 894 138 

total 410 656 13 815 577 260 263 34 660 3 722 6 980 085 

Central state 

Kansas 96 700 673 583 85 938 149 027 2 455 409 

nebraska 13 832 1 370 129 58 034 16 558 1 252 186 720 

north dakota 3 306 541 545 4 366 113 3 885 158 965 

oklahoma 170 900 1 247 742 32 143 84 726 645 303 599 

South dakota 1 709 343 747 2 894 302 2 140 158 251 

texas 759 734 12 603 897 73 663 85 436 706 10 922 639 

total 1 046 181 16 780 643 257 038 336 161 8 630 12 185 583 

major conservation practices and determine if 

the peer reviewed literature provides measures 

of the quantity of change in the ecosystem 

good or service or merely considers those 

changes without quantifying them. We then 

evaluate what the literature indicates about the 

social effects associated with implementation 

of the conservation practices, the economic 

consequences of the conservation practices, 

and the economic valuation of the ecosystem 

goods and services. Our intent is not to specify 

what these values are today because they will 

change over time. Rather, we seek to define the 

ecosystem goods and services and their social 

and economic benefits as well as how those 

values may be used in decision making. 

ECoSyStEM SERvICES 

Ecosystem goods and services are defined 

as those things or experiences produced by 

natural systems on which humans place 

value (Alcamo et al. 2005; Fisher et al. 2009; 

Fox et al. 2009; Kremen and Ostfeld 2005; 

Millennium Ecosystem Assessment [MEA] 

2005). In this section, we examine the types 

of ecosystem services that may be increased 

or decreased from the implementation 

of conservation practices, identify major 

relationships among ecosystem goods and 

services, and describe the primary issues 

associated with measurement of those 

relationships. 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 377 
	 	 	 	

 

	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	

	 	 	 	

	 	

	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	

	 	 	 	

	 	 	 	 	 	 	

	 	

	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	

tABLE 3. All NRCS program funds expended on conservation practices by state, 2005–2009. 

Location 

Brush 

management 

Prescribed 

burning 

Prescribed 

grazing 

Range 

planting 

Riparian 

herbaceous 

buffer 

Wildlife 

upland habitat 

management total 

West rangeland state 

Alaska 4 768 — 405 184 — — 33 208 443 160 

Arizona 2 896 930 17 256 4 573 259 326 474 — 65 166 7 879 085 

California 3 311 387 15 270 1 045 000 928 945 13 104 92 985 5 406 692 

Colorado 1 645 065 4 176 2 405 289 575 649 31 140 504 4 770 714 

Hawaii 2 249 800 — 546 932 89 765 — 1 490 2 887 986 

Idaho 270 053 4 372 929 023 349 916 2 013 7 199 1 562 577 

Montana 83 108 — 3 266 740 722 150 588 4 196 4 076 782 

nevada 758 417 — 176 834 171 984 — 307 1 107 542 

new Mexico 18 185 311 68 398 1 333 336 1 084 828 338 351 235 21 023 445 

oregon 1 735 163 97 104 3 195 510 524 326 — 150 912 5 703 015 

utah 1 087 580 2 152 773 954 860 948 525 20 790 2 745 949 

Washington 219 — 149 788 145 726 2 264 191 632 489 629 

Wyoming 717 663 21 132 2 950 802 48 960 — 67 070 3 805 627 

West total 32 945 463 229 860 21 751 651 5 829 671 18 862 1 126 694 61 902 202 

Central rangeland state 

Kansas 4 141 308 417 368 11 378 028 299 756 — 1 058 16 237 518 

nebraska 2 281 417 238 001 1 081 426 681 287 311 626 510 4 908 952 

north dakota 70 628 — 1 350 708 189 920 266 341 624 1 953 145 

oklahoma 10 301 158 766 992 1 067 643 691 442 — 64 273 12 891 508 

South dakota 6 664 — 609 149 351 212 — 71 877 1 038 902 

texas 77 507 317 436 684 9 585 612 5 720 478 4 024 44 434 93 298 549 

Central total 94 308 492 1 859 045 25 072 567 7 934 094 4 601 1 149 775 130 328 573 

Grand total 127 253 955 2 088 905 46 824 218 13 763 765 23 463 2 276 469 192 230 775 

There are a variety of conceptual models used 

to organize and classify various ecosystem 

goods and services for purposes of informing 

management decisions (Ruhl 2008; Swinton 

2008). The Millennium Ecosystem Assessment 

(MEA) is one such model (Carpenter et al. 

2006) that sorted ecosystem services into 

provisioning (e.g., food, freshwater, fuel 

wood, and genetic resources), regulating (e.g., 

climate regulation, disease regulation, flood 

regulation, and erosion regulation), cultural 

(e.g., spiritual/inspirational, recreational, 

aesthetic, and educational), and supporting 

(e.g., soil formation, nutrient cycling, and 

primary production) categories. The conceptual 

model by Fox et al. (2009) provided the 

framework by which the ecological systems 

interact with the social and economic systems 

and defined the ecosystem goods and services as 

extractable goods and tangible and intangible 

services. While the MEA model has gained 

acceptance in defining these ecosystem goods 

and services, there seems to be a large amount 

of double counting that could occur if it were 

implemented on the ground. From a valuation 

viewpoint, double counting creates problems 

in summing up the total effects. The Fox et al. 

(2009) conceptual model did not seek to define 

the different ecosystem goods and services but 

instead made explicit that ecosystem goods 

and services provide the connection between 

human systems and their environment and 

hence their source of value. 

Part of the issue with defining ecosystem 

services associated with implementation of 

378 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

tABLE 4. Total annual expenditures through all NRCS programs for selected conservation practices in 2005–2009. 

Conservation practice 

year 

2005 2006 2007 2008 2009 

West rangeland states 

Brush management 3 747 652 5 634 082 6 052 764 9 013 795 8 497 171 

Prescribed burning 3 165 29 918 61 380 83 048 52 350 

Prescribed grazing 3 205 190 3 731 971 5 842 649 5 116 415 3 855 425 

Range planting 788 901 774 455 1 439 449 1 548 191 1 278 675 

Riparian herbaceous buffer 7 882 3 989 2 635 3 430 927 

Wildlife upland habitat management 30 415 95 562 210 273 328 152 462 292 

West total 7 783 205 10 269 975 13 609 150 16 093 031 14 146 840 

Central rangeland states 

Brush management 14 198 925 17 012 119 17 297 158 24 859 881 20 940 409 

Prescribed burning 232 389 186 525 258 648 643 463 538 020 

Prescribed grazing 1 802 474 3 400 594 8 407 342 5 071 747 6 390 410 

Range planting 1 267 242 1 497 869 1 521 917 2 195 160 1 451 906 

Riparian herbaceous buffer 64 694 — 3 688 155 

Wildlife upland habitat management 123 393 173 081 244 566 365 460 243 274 

Central total 17 624 488 22 270 883 27 729 631 33 139 399 29 564 174 

Grand total 25 407 693 32 540 858 41 338 781 49 232 430 43 711 014 

conservation practices is that there is little 

research on their production functions 

(Kremen and Ostfeld 2005). Production 

functions describe the relationship between 

the quantities and qualities of inputs used 

to produce various quantities and qualities 

of outputs. In addition, there is a need to 

understand the form of the relationship 

between different outputs described as the 

production possibility frontier by economists. 

For example, a given amount of input (e.g., 

land) can produce a variety of outputs 

(e.g., cattle vs. wildlife). Herrick (2000) 

concluded that we need to demonstrate 

causal relationships between soil quality and 

ecosystem functions such as biodiversity and 

biomass as well as the ecosystem’s response to 

disturbance. Lal (2007) similarly presented 

arguments that soil science is the key if 

we are to pursue going to a carbon-based 

economy and meet the diversity of needs 

and wants from ecosystems. In any case, 

the characteristics of many ecosystem goods 

and services that will make them difficult to 

assess in decision making include their public 

good aspects, spatial and temporal dynamics, 

joint production, complexity of ecosystems, 

interdependence benefits, and the interactions 

among these characteristics (Fisher et al. 2009). 

There are many pressures on the production of 

ecosystem services originating from agricultural 

and natural resource management, including 

conservation practices, and societal issues, 

such as urbanization and land fragmentation. 

Converting land use from the production of 

typical agricultural products to the production 

of biomass for energy can also drastically alter 

the quantity and type of ecosystem services 

produced on a given land tract. Cook et al. 

(1991) estimated that the potential exists for 

20 million ha of US rangeland to be converted 

to energy-producing biomass with impacts 

on wildlife habitat, soil erosion, salinization, 

groundwater depletion, and subsidence. The 

study did not, however, quantify the expected 

changes in those ecosystem services. Higgins 

et al. (2002) looked at the potential impacts 

of agricultural practices and development as 

threats to future waterfowl habitat conservation 

over time. They concluded that changing 

economic and policy pressures on farmers 

and ranchers have the potential to modify 

management practices to bring marginal land 

currently in the Conservation Reserve Program 

(CRP) back into crop production. 

In a study to evaluate the effects of western 

juniper (Juniperus occidentalis Hook) control 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 379 
	 	 	 	

on an eastern Oregon ranch, environmental 

services were included in the ranch model 

to evaluate their response to juniper control 

(Aldrich et al. 2005). While this study did 

not estimate the production function for 

environmental services, it did show how 

erosion and wildlife populations may change 

in response to the implementation of different 

management alternatives. Wildlife species 

(quail, deer, and elk) responded differently 

because of unique habitat needs as the percent 

canopy cover of juniper changed from the 

alternative juniper control practices and the 

implementation timing. Erosion potential 

decreased as the trees were removed and cover 

of grasses and shrubs increased. Other models 

have been used in similar situations to evaluate 

tree control to increase forage production 

(Engle et al. 1996). Unfortunately, the manner 

in which various responses of ecosystem 

services might impact landowner decisions was 

not considered in either study. 

A comparison of cattle production with quail 

and deer habitat in Oklahoma indicated that 

ranch returns varied based on the amount 

of wildlife present and the various brush 

management treatments that were used (e.g., 

prescribed burning, herbicide applications, and 

mechanical treatments; Bernardo et al. 1994). 

Lease hunting showed higher net returns per 

hectare compared to cattle production when 

wildlife populations were abundant but with 

no lease hunting income. 

The development of a production possibilities 

frontier (PPF) requires knowing how each 

ecosystem good or service responds to common 

inputs such as vegetation (McCoy 2003). A 

PPF was developed to compare cattle and 

antelope in Wyoming (Bastian et al. 1991); 

however, the PPF did not show a great deal 

of substitutability because dietary overlap 

between the two species was minimal. It was 

concluded that it would take extreme value 

differences between cattle and antelope for the 

economically optimal combination of species to 

be only cattle or only antelope. 

A PPF describing the effect of cattle grazing 

on carbon and nitrogen balance of mixed-grass 

rangelands was developed comparing light, 

heavy, and ungrazed pastures (Schuman et al. 

1999). Aboveground biomass, carbon, and 

nitrogen showed a curvilinear (decreasing at an 

increasing rate) decline with increasing grazing 

intensity. When considering both above- and 

belowground biomass, a U-shaped curve was 

observed where total carbon decreased at an 

increasing rate and total nitrogen was about 

linear because of grazing intensity. In these 

kinds of studies, three treatment levels can 

begin the process of identifying the PPF curve 

in order to assess the trade-offs among unique 

combinations of the two products. More 

treatment levels will lead to better decisionmaking capabilities, allowing for more finely 

defined points to ascertain continuous trade-

offs. 

The quantity and quality of forage produced is 

one of the major ecosystem goods that is valued 

in the market, which makes it relatively easy 

to value compared to other ecosystem goods 

and services (Bartlett et al. 2002; Council for 

Agricultural Science and Technology 1996). 

However, forage production can have other 

values beyond that which is placed on it in 

the marketplace. The market value of forage 

tends to be heavily weighted toward domestic 

livestock production, but it can also have 

other values, such as wildlife feed and habitat, 

erosion control, and quality of life. These 

additional values are likely to be captured 

only through nonmarket valuation methods. 

It would be necessary to estimate a livestock 

production value, a wildlife feed and habitat 

value, an erosion control value, a quality-of-life 

value, and social benefits and costs if the total 

economic value for forage is to be estimated 

(Bartlett et al. 2002). 

Implementing conservation practices can have 

the unintended consequences of reducing some 

ecosystem goods and services. For example, it 

is generally viewed that controlling salt cedar (Tamarix spp.) is a desirable practice when 

the objective is to improve riparian habitat for 

numerous wildlife species. However, as Dudley 

and DeLoach (2004) pointed out, when an 

endangered species such as the southwestern 

willow flycatcher (Empidonax traillii extimus) 

uses salt cedar for habitat, controlling salt cedar 

presents the problem of an incidental taking of 

an endangered species even though the native 

vegetation might eventually provide better 

habitat. In addition, salt cedar or its control 

can have impacts on water availability, other 

380 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

	 	 	 	 	

	 	 	

	 	 	 	

	 	 	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

Small acreage owners in Utah 

County, Utah, maintain livestock that can affect surrounding agricultural land. (Photo: 

Mark Brunson) 

wildlife, aesthetics, and forage availability, 

resulting in multiple trade-off decisions. 

Beyond these trade-offs, one of the main 

reasons to control invasive plant species such 

as salt cedar is to prevent its eventual spread to 

other areas. The bottom line for such decisions 

are both spatial (e.g., water for downstream use 

vs. wildlife habitat at the point of control) and 

temporal (spread over time). 

In a Texas study of CRP lands, a comparison of 

targeting only high potential programs based 

on costs, benefits, or the benefit-to-cost ratio 

found mixed results for environmental benefits 

(Babcock et al. 1996). They found that using 

the proper criterion (benefit to cost) can result 

in greater environmental benefits, including 

reduced wind and water erosion, increased 

surface water quality, or better wildlife habitat, 

compared to other selection criteria for enrolled 

lands. They noted that heterogeneity of 

environmental quality and productivity affects 

the magnitude of changes in the environmental 

effects. They were not able to estimate the 

impact on the production quantity for many 

ecosystem services because of the lack of ability 

to quantify physical trade-offs among the 

ecosystem services and the absence of a social 

value function to evaluate societal trade-offs. 

It may be possible to increase ecosystem 

services through changes in management 

options. Integrating crop and livestock systems 

in Texas was shown to improve nutrient 

cycling, reduce soil erosion, improve water 

management, interrupt pest cycles, and spread 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 381 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	

	 	 	 	

	 	

Salt Cedar (tamarisk) control, 

Bosque del Apache National 

Wildlife Refuge, New Mexico. 

(Photo: John Tanaka) 

economic risk through diversification (Allen 

et al. 2008). Grazing was shown to increase 

soil organic carbon and nitrogen contents 

with light grazing compared to no grazing or 

heavy grazing (Ganjegunte et al. 2005). Brush 

management may be a way to increase water 

yield as well as bird habitat for species that 

require grasslands (Olenick et al. 2004b). 

In Arizona, an estimate was made of the value to 

home owners from riparian habitat restoration, 

and it was concluded that the benefits exceeded 

the costs in this case (Bark-Hodgins and 

Colby 2006). Although the value is not solely 

attributable to riparian restoration, it does 

identify hedonic pricing models (essentially 

a regression model that uses characteristics to 

explain differences in price) that relate numerous 

attributes to the value of the land as one method 

of estimating the value to property owners of 

some ecosystem goods and services. A similar 

hedonic model was used to estimate amenity 

values for agricultural land in Wyoming. Land 

that could offer several economic services, 

including scenic views, elk habitat, sport fishing, 

and distance to a nearby town, was shown to 

command a higher selling price (30%) than 

similar land that did not provide these ecosystem 

services (Bastian et al. 2002). 

The relationship between the type of ecosystem 

being improved and the management uses of 

the riparian zone affect how a conservation 

practice can be implemented to improve 

riparian vegetation. Quinn et al. (2001) 

developed a relationship between a riparian 

zone classification and the potential for 

riparian zone improvement in ecological 

health. Implementing practices such as offstream water development to draw cattle 

out of riparian areas can have beneficial 

effects on both livestock and the riparian 

area (Stillings et al. 2003). Changing the 

status of riparian habitat was shown to have 

differential effects on amphibians, reptiles, 

birds, and mammals (Ekness and Randhir 

2007). They found that the higher the degree 

of disturbance to the riparian area, the greater 

the negative impact on these four wildlife 

groups. They concluded that spatial targeting 

of conservation practices will have the greatest 

positive effect when targeting headwaters and 

lower-order watersheds. They also developed 

a conceptual model to evaluate the role of 

conservation practices that affect watershed 

characteristics important to the wildlife groups 

that they studied. Freemark (1995) developed 

a spatial–temporal hierarchy to illustrate the 

scales at which conservation practices and 

other stressors can affect wildlife in agricultural 

landscapes. The implication is that at these 

different scales, conservation practices may 

have differential impacts on wildlife habitat 

and populations. 

In this section, we have sought to define 

and identify the kinds of ecosystem services 

that can be expected to arise from the 

implementation of conservation practices. The 

basic premise is that these ecosystem services 

arise, either intentionally or unintentionally, 

from the conservation practice and can 

have either a positive or a negative value. 

Understanding the relationships and relative 

values of the ecosystem services is crucial for 

making investment allocation decisions and 

to determine whether an investment in a 

conservation practice is going to be profitable. 

EConoMICS 

Prescribed Grazing 

The NRCS conservation practice standard 

for prescribed grazing (US Department of 

Agriculture [USDA]-NRCS 2007) defines 

“prescribed grazing” to be the controlled 

382 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

	 	 	 	

	 	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

harvest of vegetation with grazing animals, 

managed with the intent of achieving a 

specific objective. Sustainability of forage and 

livestock production are central concerns when 

designing a grazing strategy, but there is a 

body of literature dealing specifically with the 

economics of grazing. Choosing an optimal 

stocking rate is considered one of the most 

important grazing management decisions 

because the stocking rate decision affects 

vegetation, livestock production, wildlife and 

economic returns (Holechek et al. 2004). We 

first describe the economic model typically 

used to define economically optimal stocking 

rates and then review the literature dealing with 

the prescribed grazing conservation practice. 

Optimal Stocking Rates. The stocking 

rate decision is a classic example of the wellknown production economic model of profit 

maximization when defined from the input 

perspective (Debertin 1986; Workman 1986). 

The traditional myopic single-year economic 

model ignores potential interyear grazing 

impacts and equates the added economic value 

of an additional grazing animal to the added 

cost of that animal, a principle commonly 

known as equating value of marginal product 

to marginal factor costs (VMP = MFC). With 

diminishing rates of gain as more animals 

are added to the pasture, each animal added 

contributes less to profit than did the previous 

one, and at the economically optimal stocking 

rate, the last animal adds nothing to profit. 

The conceptual economic model was described 

over 45 yr ago by Hildreth and Riewe (1963) 

and has been applied primarily to yearling 

stocker cattle because of the added complexities 

of cow–calf production (but see Hart et al. 

1988b). Regardless of the animal class, the 

expected production rate for grazing animals 

is related to the number of animals grazing 

a given land area. Based on declining per 

head performance, a production function is 

defined that relates the gain per hectare that 

would be realized at alternative stocking rates. 

The principle of diminishing returns applies 

and animal gains will eventually decrease as 

stocking rate is increased. Total gain per hectare 

will also eventually fall, but this may occur 

at a stocking rate that is well beyond what 

would be detrimental to rangeland condition 

and future forage production. In this case, 

rangeland condition and sustainability over 

time become of key importance, and a dynamic 

economic model is needed. However, given no 

major year-to-year interactions, the single-year 

economically optimal stocking rate will lie 

somewhere between the relatively low stocking 

rate that would yield the biggest calf and the 

relatively high stocking rate that would give the 

most gain per hectare (Torell et al. 1991). 

A search for literature in AGRICOLA with 

screening on rangelands and the search term 

“stocking rate” in the title or subject field, with 

“economics” and “rangelands” in the key words, 

identified 156 papers, of which no more than 

about 40 actually did an economic assessment 

of stocking rate alternatives. Numerous studies 

applied some variation of the single-period 

model of optimal stocking rates as described 

by Workman (1986) with notable examples 

including research conducted in Wyoming 

by Hart and various coauthors (Hart et al. 

1988a, 1988b; Hart 1991; Manley et al. 

1997). Hart’s model applications improved the 

economic assessment of optimal stocking rates 

by modifying the input to be the number of 

animals grazing per unit of forage produced 

(grazing pressure [GP]) and not animals per 

hectare (stocking rate [SR]). The Hart studies 

were also unique in that long-term grazing 

studies were used to define key production 

relationships. Most economic studies about 

stocking rates and rangeland investment 

analysis have typically used biophysically 

simulated data (Huffaker and Cooper 1995; 

Aguilar et al. 2006; Teague et al. 2008). 

Hart’s revised definition of grazing input 

showed that a given cost–price situation results 

in an economically optimal GP for the current 

grazing period, but the optimal number of 

animals stocked per unit area (SR) is also 

determined for the given forage condition. The 

economically optimal stocking rate will depend 

on sale prices and production costs and will 

vary annually. Further, the production function 

captures key input–output relations crucial to 

the economic assessment. A favorable rainfall 

year means more forage, and the production 

function shifts upward. Rates of gain will be 

different for different grazing seasons, and the 

economic assessment is defined for a particular 

grazing season and grazing system. An altered 

grazing season or rotational scheme potentially 

Grazing in Northeast New 

Mexico. (Photo: Victor Espinoza) 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 383 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

	 	 	

	 	 	 	

Low sagebrush grassland, 

Indian paintbrush, Big Horn 

Canyon National Recreation 

Area, Wyoming. (Photo: John 

Tanaka) 

shifts the production function up or down. 

A limitation is that the traditional economic 

model does not include the economic value 

of other ecosystem services, and it considers 

production only during the current period. 

Because of the complexity and lack of response 

data, few economic studies have moved beyond 

the simple single-period economic model of 

stocking rates. Yet, for the cow–calf producers 

most commonly using western rangelands, 

interyear interactions always occur, and the 

stocking rate decision becomes much more 

complex with the additional uncertainty about 

available forage now and in the future. The 

cow herd must be maintained across years, and 

forage availability is highly variable between 

years. Overgrazing during dry years becomes 

problematic and controversial when conflicts 

arise between livestock producers and land 

agency personnel about reducing stocking rates 

during drought periods. 

Early dynamic economic studies of optimal 

stocking rates included Burt (1971), Karp 

and Pope (1984), Pope and McBryde (1984), 

Rodriquez and Taylor (1988), Garoian and 

Mjelde (1990), Torell et al. (1991), and 

Huffaker and Cooper (1995). There is a 

widely held belief that individual shortterm optimization is at odds with long-term 

sustainability of an ecological–economic 

system, suggesting that a dynamic approach 

is needed. Several studies have not found this 

to be the case, however. Torell et al. (1991) 

found the intertemporal grazing impacts on 

forage production were not that important. If 

profit-maximizing livestock producers would 

maximize profit during the current period, 

then nearly identical stocking decisions would 

be made as those obtained from a dynamic 

decision model, and optimal stocking rates 

would be at sustainable levels. Falling animal 

performance was the critical driver for stocking 

rate decisions. The number of stocker animals in 

384 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	 	 	

	 	 	 	

	 	

	 	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

the pasture would optimally fluctuate annually 

with forage conditions, beef prices, and 

production costs. Only occasionally would the 

cost–price situation be such that major interyear 

forage impacts would occur (Torell et al. 1991). 

Similarly, Quaas et al. (2004) concluded that for 

typical semiarid rangelands, under plausible and 

standard assumptions, short-term optimization 

leads to sustainable outcomes. Contrary to this 

finding, Teague et al. (2009), using a simulation 

model of semiarid savanna rangeland, found 

that sustainable stocking rates were 67–75% 

those that would maximize profit to livestock 

producers. They note that earning potential 

was four times higher for range in excellent 

condition and suggest a need to stock rangeland 

more lightly so as to prevent rangeland 

degradation and improve range conditions. The 

risk of potential herd liquidation and the need 

for feed purchase along with other negative 

impacts to the rangeland resource increase as 

stocking rates increase. 

Matching stocking rates with dynamic 

forage conditions has emerged as the most 

consistent management variable influencing 

both plant and animal responses to grazing 

(Briske et al. 2008). It follows that it is also 

the most important factor influencing ranch 

profitability and economic responses to 

grazing. Typical drought management strategies 

include increased supplemental feeding, 

maintaining a conservative stocking rate so 

that destocking is rarely necessary, maintaining 

grazing flexibility by having yearlings as one of 

multiple enterprises on the ranch, and leaving 

a significant amount of herbaceous production 

at the end of the grazing season (Stafford Smith 

1992; Hart and Carpenter 2005). Torell et al. 

(2010) found that interyear forage variability 

decreased net ranch returns by 46% relative to 

what could be obtained without variable forage 

conditions. 

Economics of Grazing Systems. A deferred, 

rotational or other type of grazing system must 

result in one of two production responses for 

the practice to be economically beneficial to a 

livestock producer. First, animal performance 

must be improved with the alternative grazing 

system (shifting the production function up), 

or, second, forage production must improve 

over time (shifting the production function 

up gradually over time). It is also possible 

that these grazing systems may support 

more effective management decisions by 

some managers to induce these production 

responses (see the section “Social Aspects 

of Conservation Practices”). The prescribed 

grazing system could result in lower cost, and 

that would potentially justify the practice. 

From society’s perspective, grazing practices 

may also reduce fire danger, provide other 

habitat improvements that are valued, or allow 

integration and adoption of other management 

practices that add value. 

A recent synthesis paper (Briske et al. 2008) 

summarized key findings from many different 

studies about the benefits of rotational grazing 

systems as compared to a continuous, seasonlong grazing strategy. The main conclusion 

drawn from the review was that “rotational 

grazing as a means to increase vegetation and 

animal production has been subjected to as 

rigorous a testing regime as any hypothesis in 

the rangeland profession, and it has been found 

to convey few if any, consistent [ecological] 

benefits over continuous grazing” (Briske et al. 

2008, p. 11). As noted in the review, there has 

generally not been an economically measurable 

difference in plant production/standing crop 

or animal production between rotational and 

continuous grazing with similar stocking rates. 

The production function does not appear to 

shift up within a given year from improved 

animal performance or over time because of 

increasing forage production. Economically, 

this means that if rotational grazing requires 

more labor, capital, and management inputs, 

the lower-cost continuous grazing alternative 

would be preferred based solely on net ranch 

returns. This preference may be altered by other 

goals, such as nonvalued ecosystem services or 

other societal benefits. 

Considering specifically the economics of 

implementing grazing systems, the CAB 

abstracts populated with 60 citations using the 

key words of “grazing systems” in the title and 

“economics” and “rangelands” in any other 

field. Further screening indicated only 23 of the 

articles were relevant. Nearly half the economic 

studies were conducted in Switzerland, Africa, 

and Australia. Some studies compared primarily 

different stocking rates or grazing intensities 

(Behnke 2000; Rook et al. 2004; Trapnell et 

al. 2006). Two African studies evaluated the 

...economic 

evaluations of 

grazing systems 

have consistently 

found seasonlong continuous 

grazing to be the 

most economical 

or not different in 

production and 

rate of economic 

return.”

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 385 
	 	 	 	

	 	 	

	 	 	

	 	 	

	 	

	 	 	

	 	

The most valuable 

forage is not 

necessarily of the 

highest quality; 

rather, it is 

available when 

few alternatives 

exist.”

economics of multipaddock systems and found 

that few paddocks rather than many paddocks 

were most economical (Beukes et al. 2002; 

Mentis 1991). Hart et al. (1993) noted that 

cross-fencing and water development were 

important to achieve uniform utilization of 

forage and for minimizing grazing energy costs, 

but these goals can be achieved independently 

of the grazing system. 

Not surprising, given the finding of the Briske 

et al. (2008) literature review that few forage 

and livestock benefits accrue from rotational 

grazing, economic evaluations of grazing 

systems have consistently found season-long 

continuous grazing to be the most economical 

or not different in production and rate of 

economic return (Heitschmidt and Kothmann 

1980; Quigley et al. 1984; Van Tassell and 

Conner 1986b; Hart et al. 1988a; Heitschmidt 

et al. 1990). A somewhat different conclusion 

was reached by Owensby et al. (2008) where 

season-long stocking of a tallgrass prairie 

site in Kansas was found to have the lowest 

economic risk (i.e., variability), but returns 

per hectare were higher for an intensive early 

stocking system. Given added capital and 

labor requirements for more intensive grazing 

systems, not only must there be measurable 

production responses from the practice, but 

those responses must be substantial enough to 

justify the added expense. The literature does 

not show this to be the case, and, as noted in 

the prescribed grazing chapter, there is minimal 

information documenting the influence of 

intensive grazing systems on the effectiveness of 

adaptive management. 

Briske and coauthors (this volume) extensively 

explored published literature dealing with 

grazing/wildlife interactions. They note that 

many wildlife species, including birds and 

wild ungulates, demonstrate a relative neutral 

response to the type of grazing system in place. 

Both positive and negative responses were 

noted. Given the general lack of response, it is 

not surprising that we found no studies that 

explored the economics of wildlife interactions 

and grazing systems. 

Another area where grazing systems and 

deferred grazing has been shown to be 

beneficial is as an adjustment mechanism to 

drought and seasonal forage shortages. As 

noted by Tanaka et al. (2007), the seasonality 

of forage use is an important consideration in 

ranch planning because the number of forage 

alternatives is limited during certain months of 

the year, and some forages and harvested feeds 

are considerably more expensive. The most 

valuable forage is not necessarily of the highest 

quality; rather, it is available when few other 

alternatives are. A grazing scheme that leaves 

residual forage for carryover and use during 

a future short-supply period, allows riparian 

areas to be rested, extends the grazing season, 

and/or replaces an expensive feed alternative 

has substantial economic value (Stillings et al. 

2003; Tanaka et al. 2007). 

Greater reliance on livestock grazing compared 

to harvested forages is an effective way to 

reduce feed costs that requires a planned 

grazing strategy. Adams et al. (1994) estimated 

that the weaning weights of calves were 

increased 5 kg by grazing meadows during 

May instead of feeding hay, and feed costs were 

substantially reduced. Extending the grazing 

season in winter and spring increases ranch 

returns over traditional systems that used a 

greater amount of harvested forage. Winter 

feeding costs are the largest expense for many 

livestock operations (Prevatt et al. 2001), and 

innovative grazing schemes have the potential 

to reduce those costs. 

Briske et al. (2008) noted that even if evidence 

for production benefits from rotational 

grazing is inconclusive or nonexistent, many 

livestock producers believe that such benefits 

do exist. The website by Holistic Management 

International (http://www.holisticmanagement. 

com/n7/results_07.html; last accessed March 

27, 2009) documented the stepped-up 

level of management and perceived benefits 

that ranch managers practicing holistic 

management have received. The survey of 

43 ranch managers in the northern Rockies 

indicated that a high percentage of participants 

now do annual ranch planning, set goals, 

and have annual and formally documented 

land monitoring programs in place. It is from 

these activities that the majority of benefits 

from added management likely occur rather 

than a specific grazing practice. Further, they 

largely believed that production benefits do 

in fact exist, in contrast to the experimental 

evidence summarized by Briske et al. (2008). 

386 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 	 	 	 	 	 	 		 	 

 

J. A. Tanaka, M. Brunson, and L. A. Torell 

Management and financial skills have typically 

been taught at holistic management schools, 

and these skills may be the biggest benefit 

that livestock producers have received from 

intensive grazing system training. These 

benefits are discussed in greater detail in 

the section “Social Aspects of Conservation 

Practices.” 

Brush Management 

“The economics of brush control must be 

determined by the amount of forage and 

meat products gained; however, the principal 

objective in brush control should be an upgrade 

in range condition” (Hyder and Sneva 1956, 

p. 34). This statement, made over 50 yr ago, 

clearly articulated what was then and continues 

to be the main reason and economic rationale 

for brush control practices. The NRCS 

now recognizes six broad reasons for brush 

management (USDA-NRCS 2003): 

•	� Added forage for livestock 

•	� Restoration of natural plant community 

balance 

•	� Creating the desired plant community 

•	� Controlling erosion, reducing sediment, 

improving water quality, and enhancing 

stream flows 

•	� Maintaining and enhancing wildlife 

habitat including protection of endangered 

species 

•	� Protection of life and property from 

wildfire hazards 

A Texas landowner survey describing incentive 

for brush control found that increased forage 

production and water conservation were most 

important (Kreuter et al. 2005). Secondary 

incentives were to improve aesthetic values, 

benefit the next generation, improve wildlife 

habitat, and improve real estate values. 

Forage Production Benefits. The traditional 

brush control economic analysis that Hyder 

and Sneva (1956) described uses standard 

net present value (NPV) tools with the cost 

of the brush control treatment compared 

to discounted net future forage production 

benefits expected to be realized over some finite 

treatment life. The key tasks and elements 

of the economic assessment are to define the 

expected forage response (an assessment of 

forage productivity with and without the 

treatment) and estimate the added livestock 

carrying capacity possible over time with brush 

control, select an appropriate discount rate to 

properly account for timing difference between 

benefits and costs, and price and value the 

added grazing capacity (Workman and Tanaka 

1991). A positive NPV or a benefit-to-cost 

ratio greater than one implies an economically 

feasible rangeland management practice 

(Workman 1986). 

The economics of controlling brush for 

enhanced livestock production is variable 

depending on the economic value assigned to 

the forage, the assumed rate of forage response 

and treatment longevity, the assumed proper 

use rate or allowance for how much of the 

additional forage will be harvested to generate 

additional livestock income, and the discount 

rate used. The economic value of the added 

forage is influenced by the quality of the forage 

for grazing, by the forage and feed alternatives 

available, and by livestock prices. Most 

important is whether the added forage would 

be available during periods when other forages 

are scarce and costly (Evans and Workman 

1994). 

The expected longevity of brush control 

treatments varies widely by brush species and 

range site, as does the forage response. Yet a 

consistent overstory–understory relationship 

has been noted for many shrubland 

communities and species. These relationships 

generally show a downward-sloping sigmoid 

or exponential curves when herbaceous yield 

(kg · ha−1) is plotted against brush canopy (%; 

Ffolliott and Clary 1972). This suggests that 

increased brush cover diminishes understory 

forage production but at a decreasing rate. 

Herbaceous production has been shown to 

increase an average of three to five times 

following effective control of many brush 

species located on productive range sites, 

including sites infested with Wyoming big 

sagebrush (Artemisia tridentata wyomingensis; 

Hyder and Sneva 1956; McDaniel et al. 2005), 

broom snakeweed (Gutierrezia sarothrae; 

McDaniel et al. 1993), pinyon-juniper (Clary 

et al. 1974; Pieper 1990), and redberry juniper 

(Juniperus pinchotii; Johnson et al. 1999). 

Successful control of other species like mesquite 

(Prosopis glandulosa), salt cedar (Tamarix 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices CHAPTER 2: Assessment of Prescribed Fire as a Conservation Practice 387 
	 	 	 	

 

 

	 	

	 	 	 	

	 	

	 	 	

	 	

	 	

	 	

...forage 

production, not 

water yield, is the 

primary benefit 

of the brush 

control practice, 

especially on 

more mesic sites.”

spp.), and creosotebush (Larrea tridentata) 

resulted in an increase in grass cover, but with 

minimal changes in harvestable forage except 

on productive sites with adequate rainfall 

(Ethridge et al. 1984; Harms and Hiebert 

2006; Perkins et al. 2006; Combs 2007). 

The economics of brush management practices 

continues to be evaluated on the basis of the 

amount of forage and meat products gained 

by implementing the practice. The economic 

component of PESTMAN, a holistic decision 

support system currently in development at 

Texas A&M University (PESTMAN 2009), 

is driven by the anticipated forage response 

to a selected brush control treatment. Yet, as 

noted over 30 yr ago by Smith and Martin 

(1972), based on livestock production value, 

most rangeland management practices showed 

a negative benefit-to-cost ratio (costs exceed 

benefits) based only on the value of the added 

forage. This is a consistent and continuing 

conclusion from studies dealing with the 

economics of brush control practices. Increased 

returns from improved animal performance 

and production are usually too low for brush 

control to be economically justified (McBryde 

et al. 1984; Lee et al. 2001; Torell et al. 2005a). 

Landowners recognize this, and many brush 

control projects are implemented under costshare arrangements with state and federal land 

management agencies. 

Torell et al. (2005a) found that a cost-share 

payment of about 30% of the treatment cost 

was required to justify control of big sagebrush 

in northwestern New Mexico when the added 

forage from the brush control practice was 

valued at an intermediate level of $7 · animal 

unit month−1 (AUM; in 2003 dollars). The 

NPV of the investment was positive except at 

two relatively unproductive sites, when forage 

was valued at $10 · AUM−1. A rangeland 

management practice that adds forage during 

a critical and limiting season makes forage 

valuable, and many times added forage 

production alone justifies the improvement in 

these cases (Evans and Workman 1994). 

If a brush control project successfully increases 

available forage, increasing livestock numbers 

is not always justified. In some cases, a 

justification for brush control is that the 

stocking rate on the area can be maintained 

nearer its actual capacity by recognizing that 

current stocking rates are not sustainable. 

Sagebrush control near Farmington, 

New Mexico, helped the Bureau of Land 

Management (BLM) avoid potential conflict 

and lawsuits with grazing permittees and other 

parties because positive steps were taken to 

reduce grazing pressure but without forcing 

major herd reductions (Torell et al. 2005a). 

Similarly, forestalling the need for controversial 

grazing reductions was a primary benefit 

of the 11-yr (1962–1972) Vale Rangeland 

Rehabilitation Program initiated on BLM 

lands in eastern Oregon (Bartlett et al. 1988). 

Therefore, the maintenance of sustainable 

livestock carrying capacity should be given 

explicit consideration when evaluating the 

effectiveness of brush management programs. 

Watershed Benefits. Watershed benefits are 

increasingly used as justification for public 

expenditures for brush control. Large tracts 

have been treated at great expense to control 

brush, especially salt cedar, to realize perceived 

watershed benefits, including added stream 

flow, water yield, and aquifer recharge. For 

various reasons, trees and brush are perceived 

to have a higher evapotranspiration (ET) 

rate than herbaceous species within the 

understory (Wilcox and Thurow 2006). 

The argument is made that if ET loss can be 

reduced by managing rangelands for a greater 

grass component and a lesser tree and shrub 

component, more water will be available for 

runoff and/or deep drainage. As noted by 

Wilcox and Thurow (2006), this argument has 

been shown to be true in a variety of humid, 

montane and Mediterranean climates, where 

studies have shown increases in water yields 

tied to removal of trees and shrubs. In semiarid 

rangelands, however, water yield benefits have 

not been demonstrated on scales that would 

greatly alter regional water supplies (Wilcox 

2002). Sturges (1983) noted that the response 

of the soil water regime to a brush control 

treatment is inversely related to the response 

in herbaceous production. This suggests that 

much of the added water from brush control in 

arid areas is used to produce greater herbaceous 

production. Thus, added forage production, 

not water yield, is the primary benefit of the 

brush control practice, especially on more 

mesic sites. It appears that there is no real 

potential for increasing stream flows in addition 

388 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

	 	 	

	 	 		

	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

Prescribed burning mosaic, 

near Lakeview, Oregon. 

(Photo: John Tanaka) 

to the forage benefit unless annual precipitation 

exceeds 450–500 mm (Wilcox 2002). 

Research addressing the economics of brush 

control for enhanced water yield has been 

conducted only on Texas watersheds. As 

noted by Wilcox (2002), the perception is 

widespread that the water supply in Texas can 

be substantially increased through aggressive 

control of mesquite and juniper. Several 

studies have explored the cost implications of 

Texas brush control practices from a rancher 

perspective (Lee et al. 2001; Olenick et al. 

2004a), and surveys have been conducted to 

evaluate ranchers’ willingness to participate 

in brush control projects designed to enhance 

water yields (Thurow et al. 2000; Kreuter et 

al. 2004, 2005). These landowner surveys 

indicated that a subsidized public cost-share 

program would be necessary for widespread 

participation in brush control projects by Texas 

ranchers if public watershed benefits were the 

goal. 

Economic studies have generally evaluated 

the economic value of watershed benefits 

indirectly using the procedure described by 

Lee et al. (2001) and Olenick et al. (2004a). 

The discounted forage production benefits are 

assumed to go to private ranchers, and they 

have an assumed willingness to participate 

at a maximum cost up to this level. Beyond 

this point, the NPV of the investment would 

be negative (benefits < costs) for the private 

landowner. Watershed benefits or the public’s 

benefit is estimated to be the present value of 

the brush treatment cost minus landowner 

forage benefits. This residual value is a surrogate 

measure of the value that society must place 

on watershed benefits if the investment is to 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 389 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

Prescribed fire in pinyonjuniper, Arizona. (Photo: John 

Tanaka) 

be economically efficient and have a positive 

overall NPV. The analysis makes no attempt 

to estimate what watershed benefits actually 

were but instead estimated the level of required 

public benefits required to justify the total cost 

of brush control. The obvious justification and 

assumption was that the cost-share program 

accurately reflected social priorities. Skeptics 

can counter, however, that land management 

agencies’ budgets and spending priorities 

most often reflect political and bureaucratic 

objectives and are not a reflection of social 

value (Skaggs 2008). 

Additional Texas studies have used plant growth, 

hydrologic, and economic models to determine 

costs and benefits (added water) resulting 

from brush control (Bach and Conner 1988; 

Conner and Bach 2000; Lemburg et al. 2002; 

Olenick et al. 2004a). Simulated estimates 

indicated that the public cost of additional 

water ranged from $26 to $129 per 1 000 m3 

depending on area and type of treatment. This 

compared with an estimated $65 per 1 000 m3 

for leasing water pumped from the Edwards 

Aquifer (Olenick et al. 2004a). 

Tamarix, or salt cedar, is often controlled 

on the basis of an economic justification 

associated primarily with water conservation. 

It is an expensive species to control ($4 000– 

$12 000 · ha−1), requiring repeated mechanical, 

fire, chemical, and revegetation treatments 

as summarized at http://saltcedar.nmsu.edu. 

Economic feasibility studies assessing the costs 

and benefits of salt cedar control have been 

conducted for some western US waterways 

(Great Western Research Inc. 1989). Horton 

and Campbell (1974) estimated that water 

savings, based on the difference in water use 

between salt cedar and native vegetation, were 

as high as 3 000 acre-feet · yr−1 following salt 

cedar control on the Colorado River. It has 

been estimated that 568 000 acre-feet · yr−1 of 

water are lost to salt cedar from the Bonneville 

Unit of the Central Utah Water Project on the 

Colorado River at an estimated cost of $27 

million annually (Brotherson and Field 1987). 

Zavaleta (2000) estimated that marginal water 

losses to salt cedar are comparable to annual 

precipitation totals for the arid western states 

where salt cedar has been a problem. She 

estimated that Tamarix stands consume 3 000– 

390 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

J. A. Tanaka, M. Brunson, and L. A. Torell 

4 500 m3 · ha−1 · yr−1 (8 219–12 329 L · d−1) 

of water, more than the native vegetation 

it replaced. Lost economic value in 1998 

dollars was estimated to be $284–$447 · ha−1 

of land infested by the invasive species. 

Using a multiyear average treatment cost of 

$5 000 · ha−1 and with a 6% discount rate, 

Zavaleta (2000) estimated that it would take 16 

to 50 yr to break even on the initial and followup treatment costs, depending on the value 

assigned to the water saved. 

The economic assessment by Zavaleta 

(2000) likely used the popular Tamarix 

water consumption value of about 757 

L · d−1 · plant−1. Owens and Moore (2007) 

reviewed the literature on water use by Tamarix 

and concluded that a more realistic estimate 

of the maximum daily water use was less than 

122 L · d−1 · plant−1. As Owens and Moore 

(2007) noted, at this reduced estimate of water 

savings, the economics of spending $5 000– 

$8 000 · ha−1 to control salt cedar would be 

dismal, based only on water conservation 

benefits. 

Wildlife Benefits. Wildlife habitat and, 

therefore, wildlife populations are variously 

influenced by the relative cover of brush and 

tree species, and many wildlife species prefer a 

much denser overstory then would be optimal 

for livestock forage production and water 

yield alone (< 5% brush canopy). Surveys of 

Texas landowners indicated an average brush 

canopy cover of 41% on their ranches as 

compared to a preference of 27%, a level that 

would maximize the value of lease hunting for 

white-tailed deer (Thurow et al. 2000). Many 

Texas ranchers are more interested in brush 

thinning than brush eradication because of 

the high revenues derived from lease hunting 

(Kreuter et al. 2004). 

Few studies have evaluated the economics of 

brush management for enhancing wildlife 

values and especially with consideration of 

the trade-offs with other resource values. 

Aldrich et al. (2005) developed a multiperiod 

linear programming model to evaluate the 

economics of western juniper control in 

central Oregon. Profit from livestock income 

was maximized by choosing economically 

optimal juniper management strategies so as 

to manipulate available forage resources for 

livestock production. Wildlife income was 

not considered a source of ranch revenue, 

but equations were included to evaluate how 

optimal juniper control strategies for livestock 

production would impact quail, deer, and elk 

numbers. Given the conflicting level of desired 

juniper for cattle versus wildlife production and 

with maximization of income from cattle only, 

wildlife numbers were projected to decline 

following livestock-revenue profit-maximizing 

strategies. 

Standiford and Howitt (1993) developed 

an optimal control model that considered 

multiple resource values from management of 

California’s hardwood rangelands, including 

forage production, oak wood sales, and hunting 

revenue. Key relationships were identified, 

including expected oak tree growth rates, the 

interaction of tree overstory versus livestock 

forage production, and hunting revenue 

potential under different tree canopies. 

Revenue from hunting was defined to increase 

with oak crown cover, whereas revenue from 

cattle decreased with an increasing canopy 

of oak trees. Economically, optimum oak 

canopy was found to vary depending on the 

resource values considered, but the optimum 

was consistent with expectations. When only 

livestock production value was considered, 

the optimal control model indicated that oak 

trees would be gradually cleared because of 

the resulting additional forage for livestock. 

Over the assumed 13-yr planning horizon, 

oak canopy would be reduced from 55% to 

less than 5%, but immediate tree clearing was 

not feasible given imposed realistic budget 

constraints. Adding firewood harvest to 

livestock income resulted in a light tree harvest 

for firewood (2–3% of oak canopy). The NPV 

addition from firewood revenue was 3%. 

Managing for cattle and wildlife increased NPV 

by 40% and greatly changed management 

of the oak canopy. The oak canopy would be 

maintained at 55%, no firewood would be 

harvested, and reduced cattle numbers would 

be optimal. The marginal economic value of 

the oak canopy for wildlife habitat exceeded the 

marginal value of greater livestock forage and 

firewood volume. 

Bernardo et al. (1994) highlighted that wildlife 

is an increasingly important source of ranch 

income, and hedonic ranchland valuation 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 391 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	

Spraying salt cedar along 

Pecos River. (Photo: Kirk 

McDaniel) 

models show market preferences for ranches 

with wildlife revenue and hunting potential. 

Ranches with quality wildlife habitat and 

scenic appeal bring premium prices in the 

ranch real estate market. Torell et al. (2005b) 

estimated that 25% of New Mexico ranches 

have wildlife income-earning potential and that 

these ranches sell for premium prices, especially 

scenic mountain ranches with elk herds present. 

As discussed above, Standiford and Howitt 

(1993) have shown optimal brush management 

strategies for wildlife to be different from 

optimal production strategies for livestock on 

California’s hardwood rangelands. Bernardo 

et al. (1994) similarly concluded that some 

reduction in cattle grazing was necessary to 

maintain deer and quail habitat at desired 

levels. As noted by the Texas Parks and Wildlife 

Management Department (2008), most wildlife 

species are selective foragers, preferring to feed 

on a wide variety of plants rather than a few 

specific ones. Therefore, habitat improvement 

recommendations should emphasize the need 

for an even distribution and high availability 

of potential forage plants from season to 

season. Brush management can allow solid 

stands of woody vegetation to be interspersed 

with cleared areas over the landscape. Cleared 

strips or blocks can produce desirable forb and 

browse production while retaining an adequate 

mosaic of woody cover for escape, nesting, or 

protection from the elements. Properly utilized 

brush management practices can improve the 

availability of escape cover and food plants for 

both wildlife and livestock, though livestock 

production profitability will not be maximized. 

Prescribed Burning 

Various search engines, including 

AGRICOLA (Agriculture abstracts), CAB 

Abstracts, and wildlife and ecology studies 

worldwide, were used to evaluate the peerreviewed literature that is available on the 

economics of prescribed fire as a conservation 

practice. “Fire” was required in the title 

and key words, or the title had to include 

“economics.” After some screening and 

elimination of irrelevant papers, 46 citations 

dealt with fire and economics. Seven of these 

papers also used the term “social” in either 

the title or the key words. Six of the papers 

dealt primarily with fuel reduction on forested 

lands using prescribed fire. 

Fire Hazard Reduction, Liability, and Risk 

Concerns. Of the papers dealing with human 

and economic aspects related to rangeland fires, 

nearly half the papers dealt with wildfires and 

using prescribed burning as a way to reduce 

the risk and economic damage from wildfires 

(Kaval et al. 2007; Mercer et al. 2007; Yoder 

and Blatner 2004). As noted by Kaval et al. 

(2007), residents within wildland–urban 

interface zones recognized the positive role 

that prescribed burning can have in reducing 

the dangers of wildfire and are willing to pay 

for positive fire risk mitigation measures. In 

this Colorado study, residents responding to 

the willingness-to-pay survey were willing to 

pay an average of nearly $800 year-1 for fuel 

management treatments to reduce fire risk. 

At a low $5 · household−1 annual bid price, 

all survey respondents were willing to pay for 

prescribed fire so as to reduce fire danger, while 

at a relatively high level of $1 500 · household−1 

annually, only 14% were willing to pay for 

prescribed fire treatments. 

Cost studies have shown prescribed fire to 

be a cost-effective fuel reduction method. 

Yet Hartsough et al. (2008) found that using 

prescribed fire to reduce fire danger was 

relatively high cost in the western United 

States because of terrain and stand conditions, 

high fuel loads, and the need to ensure that 

prescribed fires do not escape. Evaluating data 

392 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

	 	

	 	

	 	 	

	 	 	

	 	

	 	

	 	 	 	

	 	 	 	

	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

at seven sites in the western United States 

indicated that gross costs of mechanical fuel 

reduction treatments were more expensive 

than those of prescribed fire, but net costs were 

similar or less after the market value of the 

harvested wood products were deducted. The 

relative merit of using prescribed fire versus 

mechanical canopy removal was highly sensitive 

to the market value of forest products generated 

by the mechanical operation (Hartsough 

et al. 2008). As noted by Omi (2008), fire 

hazard reduction through fuels management 

is controversial, and the literature on fuel 

treatment effectiveness in reducing fire danger 

is nearly nonexistent, as are the data required 

to assess the trade-offs between alternative fuel 

reduction treatments. 

Seven papers addressed fire escape and the 

liability risk associated with prescribed fires, 

all published since 2000. As noted by Yoder 

(2008), prescribed fire is considered a useful 

but risky method of reducing wildfire risk, 

increasing forage production, and improving 

wildlife habitat. This risk has resulted in new 

laws and reforms (Sun 2006), and Yoder found 

that these new liability laws and regulations 

have effectively reduced the incidence and 

severity of escaped fires. 

Improve Forage Production. Some type 

of intervention may be needed to improve 

range condition and to redirect what may be 

a continued decline in rangeland productivity. 

However, Fuhlendorf and coauthors found a 

very weak argument for using fire to increase 

herbaceous vegetation production, particularly 

perennial grasses (this volume). They note 

that perennial grasses generally declined in the 

years immediately following prescribed fire but 

that most perennial species recovered within 

2–3 yr. The delay in realized grazing benefit 

is problematic for realizing positive economic 

returns from the prescribed fire treatment given 

the time value of money. Further, conducting 

a prescribed burn may be limited or delayed by 

air temperature conditions, relative humidity, 

wind speeds, and the availability of fine fuels 

to carry the fire so as to minimize the risk of 

fire escape, to minimize damage to perennial 

grasses, and to carry the fire over the desired 

area. Fire is not as easy or as convenient to 

use as chemical treatments for brush control. 

McDaniel et al. (1997) noted that over a 

6-yr study period on the shortgrass prairie 

of New Mexico, the desired fire conditions 

recommended by Wright and Bailey (1980) 

were rarely observed. For many arid rangelands, 

accumulating fine fuels under a dense brush 

canopy can be particularly problematic for 

implementing prescribed burning treatment 

(Bastian et al. 1995; McDaniel et al. 1997; 

Teague et al. 2001). 

Eight papers studied the economics of 

prescribed fire as a strategy for reducing brush 

overstory and increasing the production of 

understory forage species. Two of these studies 

were outside the United States (Trollope 1978; 

Henkin et al. 1998). Most of the US studies 

evaluated the economics of prescribed burning 

for control of honey mesquite and cactus in 

the Texas Rolling Plains (Teague et al. 2001, 

2008). Other economic studies addressed oak– 

hickory forests (Bernardo et al. 1992), eastern 

red cedar (Juniperus virginiana; Bernardo et al. 

1988), Macartney rose (Rosa bracterata Wendl.; 

Garoian et al. 1984), and big sagebrush 

(Bastian et al. 1995). 

Of the limited economic studies about 

prescribed fire to enhance forage production, a 

common prescription was an initial chemical 

treatment to reduce the brush overstory 

followed by prescribed fire treatments at 5–7-

yr intervals as a maintenance treatment (Van 

Tassell and Conner 1986a; Teague et al. 2001). 

The economics of prescribed fire treatments 

were estimated to be better than chemical 

treatments even if the burn treatment was 

considered to be considerably less effective 

in overstory reduction and longevity (Teague 

et al. 2001). This is because, ignoring fire 

risk, the cost of prescribed burning was 

assumed to be much cheaper than chemical 

treatments in all the studies. Bastian et al. 

(1995) estimated that the cost of prescribed 

fire was only half that of chemical treatment, 

and Teague et al. (2001) estimated the cost of 

follow-up burn treatments to be only 10% of 

the initial $56.81 · ha−1 chemical treatment. 

Perhaps the biggest limitation of the economic 

studies of brush control, including both 

chemical and fire, was the estimation of the 

forage production response curves. Forage 

response can be expected to be highly variable 

depending on soils, pretreatment brush cover, 

climatic conditions, and forage and brush 

Perhaps the 

biggest limitation 

of the economic 

studies of brush 

control, including 

both chemical 

and fire, was the 

lack of data to 

estimate forage 

production 

response curves.”

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 393 
	 	 	 	

 

	 	 	 	 	

	 	 	

		 	 		

	 	

Snow geese at Bosque del 

Apache National Wildlife 

Refuge, New Mexico. 

(Photo: John Tanaka) 

species. The economic studies were largely 

based on relatively short response studies 

(< 10 yr) and various unmeasured assumptions 

about the rate of brush reinvasion. Given the 

general nonlinearity of overstory–understory 

relationship, the economic studies concluded 

that a relatively dense stand of brush must be 

present initially to be economically feasible 

for control by either fire or chemical methods. 

None of the studies mentioned or considered 

the reduction in perennial grass cover in the 

immediate years following prescribed burning 

treatments as compared to chemical treatments. 

Wildlife Benefits. We found only two studies 

investigating the economics of prescribed 

fire to improve big game habitat and increase 

wildlife income. A study by Gonzalez-Caban 

et al. (2003) (also published as Loomis et al. 

2002) developed production functions relating 

deer harvest response to prescribed burning. 

Diminishing marginal benefits were noted. An 

additional 445 ha of prescribed burn increased 

deer harvest by 33 head, whereas the next 

1 502 ha of prescribed burn increased deer 

harvest by eight head. When compared to the 

estimated $519–$593 · ha−1 cost of conducting 

prescribed burns, the economic value of the 

added deer harvest was only 3.4% of the total 

cost for the first 445 ha burned. 

Teague et al. (2001) studied the economic 

response of honey mesquite control in the 

Rolling Plains of Texas from both herbicide 

and prescribed fire treatments. They noted that 

burning at a 5–7-yr interval improved wildlife 

habitat. Their economic evaluation of herbicide 

and prescribed fire treatments was most 

sensitive to realizing a wildlife income response. 

If treatment on any part of the ranch increased 

wildlife income, then the NPV of the brush 

control treatment was substantially increased. 

Rangeland Planting 

Rehabilitation of rangeland by seeding and 

planting began in the western United States 

in the late 1800s, and, according to Heady 

and Child (1999), more literature exists 

on range seeding than any other practice 

in range management. They also note that 

the environmental movement after 1970 

demanded less seeding of rangeland with 

monoculture species and more with native 

species. Rehabilitation and prevention 

of erosion have replaced increased forage 

production as the primary objective for 

394 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

J. A. Tanaka, M. Brunson, and L. A. Torell 

seeding public lands. Rangeland seeding is also 

considered advantageous for managing weeds 

and cheatgrass (Bromus tectorum; Young and 

Clements 2009). 

The Heady and Child (1999) textbook chapter 

(chapter 24) provides a wealth of information 

about ecological considerations about 

rangeland planting. Seeding guidelines they 

identify include the following: 

•	� Is the seeding needed? Removal of the 

competitive brush overstory may be 

adequate. Seeding has the greatest 

potential for profitable returns when native 

vegetation does not exist. 

•	� Is the climate favorable? Successful plantings 

are infrequent in areas receiving less than 

250 mm of precipitation per year, but 

failure for areas with greater than 600 mm 

of rainfall are less frequent. 

•	� Is the habitat favorable? Select seeding 

sites with the most herbaceous response 

potential. 

•	� What species should be planted? Consider 

seasonal forage demands and the potential 

to replace expensive feeding alternatives. 

Recognize that a mixed diet is generally 

more desirable and often will produce 

greater livestock gains than a monoculture. 

•	� Manage the seeded area. Provide grazing 

deferment after establishment and do not 

overgraze. 

The steps required to analyze the economics 

of revegetation projects include incorporation 

of all these considerations as they relate to 

potential costs, benefits, and risks (Workman 

and Tanaka 1991). 

Economic literature evaluating rangeland 

seeding are generally at least 30 yr old, and 

seeding success is variable and depends largely 

on selecting a desirable site in an adequate 

rainfall area. Much of the literature exists in 

the form of extension guides and bulletins. 

These bulletins generally provide guidelines for 

designing an economically successful project 

instead of evaluating the economics of specific 

improvement projects or case studies (Lloyd 

and Cook 1960; Wiens et al. 1969; Kearl and 

Cordingly 1975; Wambolt 1980; Kearl 1986; 

Workman and Tanaka 1991; Heady and Child 

1999). 

The probability of successful seeding 

establishment was highlighted in the only 

study found concerning the economics of range 

reseeding in the desert Southwest (Ethridge 

et al. 1997). This 6-yr study of seeding trials 

on the Jornada Experimental Range near Las 

Cruces, New Mexico, considered 14 different 

plant varieties, including introduced and native 

species. The study indicated that reseeding was 

not an advisable financial investment for the 

Chihuahuan deserts of southern New Mexico 

because of the high probability of stand failure. 

Estimated NPV was negative for all species 

planted and seedbed preparation strategies. 

Most of the economic studies on rangeland 

planting—or rangeland seeding or reseeding, 

as it is categorized in the literature—are about 

the economics of seeding crested wheatgrass 

(Agropyron desertorum and A. cristatum). From 

1945 until 1965, several million hectares of 

sagebrush rangeland were seeded to crested 

wheatgrass in the Intermountain West (Young 

1994). It was estimated that in Nevada, 0.4 

million of the 11 million ha of sagebrush 

rangeland were seeded to wheatgrass. The 

seeded area constitutes only 2% of the total 

rangeland in Nevada but produces 10% of the 

harvestable rangeland forage (Young and Evans 

1986). 

Seeding sagebrush rangelands to crested 

wheatgrass was generally found to be a very 

economical practice because forage production 

was 3–20 times greater than that of the native 

plants it replaced, calf crop and average 

weaning weights increased, and early spring 

use replaced expensive hay as an alternative 

lower-cost feed. In some studies, rates of 

return were estimated to be in the range of 

10–22% with an anticipated stand life of 

25 yr or more (Kearl and Cordingly 1975; 

Shane et al. 1983). Not all economic studies 

estimated positive economic returns, however. 

Godfrey (1986) reviewed 24 economic studies 

conducted between 1943 and 1979 that dealt 

with seeding crested wheatgrass and found net 

economic returns to be positive in nine of the 

studies and variable or unknown in the other 

studies. He attributed the variability in NPV 

estimates to four main reasons: 1) some of the 

plantings were failure, 2) low-production areas 

were seeded instead of high-potential areas 

(they used a worst first selection criteria), 3) 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 395 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 	

	 	

Bison at the Wichita Mountains 

Wildlife Refuge, Oklahoma. 

(Photo: John Tanaka) 

too much was spent for brush removal, and 

4) other, less productive improvements were 

included along with the seeding project. 

Widespread planting of crested wheatgrass is no 

longer common, as it is an exotic monoculture, 

and successful widespread planting altered 

sagebrush habitat required for sage-grouse 

(Centrocercus urophasianus; Connelly and 

Schroeder 2000). As noted by Young and Evans 

(1986), the golden age of planting crested 

wheatgrass lasted for barely a decade, from 

the mid-1950s until the mid-1960s. Its role 

is now considered to be in the reclamation of 

drastically disturbed lands (Depuit 1986), and 

it has potential when seeded with forage kochia 

(Kochia prostrata ssp. virescens) to outcompete 

cheatgrass (Bromus tectorum; Harrison et al. 

2000). Widespread use and plantings indicates 

that this forage species was one of the most 

economically successful rangeland plantings in 

its day, when only livestock forage production 

was valued. 

Riparian Herbaceous Cover 

Literature dealing with the economics of 

improving riparian herbaceous cover focuses 

primarily on nonrangeland (e.g., forested 

wetlands) areas and on tree cover rather than 

herbaceous cover. An EBSCO search with the 

terms “contingent valuation” and “riparian” 

returned 25 citations, but there were no studies 

found that specifically dealt with the economic 

value of establishing riparian herbaceous cover 

in rangeland areas as a way to improve riparian 

areas. 

One of the few economic studies, a contingent 

valuation to estimate the benefits and costs of 

riparian restoration projects along the Little 

Tennessee River in North Carolina (Holmes 

et al. 2004), found net benefits from riparian 

ecosystem restoration to be strongly positive 

but much larger for large-scale projects. 

Restoration benefits were described in terms of 

five indicators of ecosystem services: abundance 

of game fish, water clarity, wildlife habitat, 

396 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	

	 	

	 	 	

	 	

	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

allowable water uses, and ecosystem naturalness. 

Other economic studies explored the willingness 

to pay and willingness to accept payment for 

provision of a riparian strip or corridor for 

habitat preservation (Amigues et al. 2002; 

Qiu et al. 2006). The equivalent of a positive 

benefit-to-cost ratio was indicated with results 

consistent with housing price differentials 

(house prices with or without riparian strip or 

corridor for habitat preservation) for stream 

access in the area (Qiu et al. 2006). The 

willingness of recreational visitors to pay for 

riparian area preservation that maintained 

bird diversity in the San Pedro River basin of 

Arizona was explored by Colby and Orr (2005). 

They estimated that a one-time aggregate 

monetary willingness to pay by nonlocal visitors 

for riparian area preservation was $2.77 million. 

upland Wildlife Habitat Management 

As previously noted, wildlife income is an 

increasingly important part of total ranch 

income, and ranchland market values are 

greatly influenced by scenic views, recreation, 

and hunting opportunities. Torell et al. (2005b) 

found that adding wildlife income to a New 

Mexico ranch contributes 2.5 times more to 

ranchland market value then does a similar 

amount of livestock income. These ranchland 

real estate market effects have been noted for 

many years (Pope 1985; Torell et al. 2005b). 

The obvious implication is that significant 

opportunities exist to increase economic values 

through upland wildlife habitat management 

if increased hunting and wildlife viewing 

opportunities can be created. 

A great deal of research has been conducted 

to estimate the demand and economic value 

of various wildlife species, with obvious value 

implications for wildlife habitat improvement. 

One notable author, John B. Loomis at 

Colorado State University, has contributed 

greatly to the development and application of 

nonmarket valuation procedures. However, 

much of his research and other related research, 

as noted by Daniels and Riggs (1988), has 

concentrated on estimating the value of the 

wildlife and not the habitat sustaining wildlife. 

As noted by Bernardo et al. (1994), the weak 

link is the lack of data required to translate 

physical effects of habitat improvement 

practices into altered wildlife numbers and 

economic benefits. The Bernardo et al. paper 

provides one of the limited cases where the 

production trade-offs between cattle grazing 

and wildlife habitat were estimated. A second 

study was a California study on oak rangeland 

described earlier by Standiford and Howitt 

(1993). A third study involved estimation of 

the production possibility frontier between 

stocker cattle and antelope by Bastian et al. 

(1991). Glover and Conner (1988) developed 

a linear programming model to evaluate the 

optimal (profit-maximizing) mix of cattle, 

sheep, goats, and deer on a representative 

ranch in the Edwards Plateau region of Texas 

and found that active management for wildlife 

added to net ranch income. Another study by 

Loomis et al. (1991) evaluated livestock grazing 

strategies that would potentially improve 

deer habitat in California and concluded that 

implementing a rest–rotation livestock grazing 

system with 1 yr or more of nonuse in a 3-yr 

cycle would increase hunting value far beyond 

the value lost from reduced livestock grazing. 

Others have estimated forage values for cattle 

versus wildlife (Martin et al. 1978; Cory and 

Martin 1985; Loomis et al. 1989), but they 

did not provide estimates within a multipleenterprise context that estimated production 

possibilities and trade-offs. 

Wildlife valuation procedures use various 

techniques to estimate a consumer’s willingness 

to pay where no established market exists, 

relying on demand analysis and consumer 

surplus estimation (Sorg and Loomis 1985; 

Champ et al. 2003). Valuation of wildlife 

habitat uses these value estimates for wildlife 

and expands to a benefit-to-cost assessment 

where the economic value of increased wildlife 

numbers is compared to the cost of practices 

that improve habitat and ultimately wildlife 

numbers. Studies that have attempted to 

estimate the linkage between altered habitat 

and wildlife numbers include the prescribed 

burning assessment described above (Loomis 

et al. 2002; Gonzalez-Caban et al. 2003) 

where benefits from additional deer harvest 

was estimated to be no more than 3.4% of 

prescribed burning treatment cost, suggesting 

that deer hunting benefits represent only 

a small part of the multiple-use benefits of 

prescribed fire. 

Lenarz (1987) concluded that treatments 

to increase forest openings were never cost 

Very little if any 

research exists 

showing the direct noneconomic 

effects of NRCS 

rangeland conservation practices 

on individuals, 

households, or 

social systems.”

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 397 
	 	 	 	

	 	 	 	 	

	 	 	 	 	

Planning in the tallgrass prairie, 

near Bowie, Texas. (Photo: John 

Tanaka) 

effective based on the value of hunting licenses, 

but it was cost effective based on total gross 

hunting-related expenditures. Daniels and 

Riggs (1988) recognized and corrected the low 

economic value assigned by Lenarz (1987) 

by considering only the value of the hunting 

leases. They based deer values on standard 

willingness-to-pay measures and concluded 

that a positive NPV would be realized from 

investments to create forest openings whenever 

cleared areas were less than the 3% level and a 

reasonable discount rate was used. 

Garrett et al. (1970) estimated the demand for 

deer hunting in Nevada and valued the habitat 

that supported hunting activity. They compared 

the habitat value to selected rehabilitation 

projects expected to alter deer numbers. The 

first improvement considered was crested 

wheatgrass (Agropyron cristatum) seeding 

with the estimate that the practice would be 

detrimental to deer numbers and wildlife value 

because of the decrease in forage species most 

desirable to deer. Chaining of pinyon-juniper 

at two sites was estimated to result in a positive 

benefit-to-cost ratio, ranging from 1.65 to 

2.09 for the two sites, based only on increased 

economic value from deer utilization supported 

by a more open woodland canopy. 

Similar to the hedonic modeling approach 

used by Torell et al. (2005b) to evaluate the 

contribution of wildlife to western ranchland 

values, Netusil (2006) used urban housing 

sales within the Fanno Creek Watershed within 

the city of Portland, Oregon, to evaluate how 

real estate prices varied with different amounts 

of upland wildlife habitat. Close proximity 

to a stream increased property values. A 

property’s sale price was found to increase as 

the percentage of regionally significant habitat 

on the lot increased but at a decreasing rate. 

Property owners placed a premium price 

on lots with habitat providing the highest 

ecological values (large forest patches, wetland 

areas, and large contiguous patches) and 

398 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	 	 	

	 	 	 	

discounted lots with lower-valued habitat. The 

maximum impact on house lot sale price was 

when upland wildlife habitat coverage on the 

property was about 38%. 

SoCIAL ASPECtS oF ConSERvAtIon 

PRACtICES 

Very little if any research exists showing the 

direct noneconomic effects of NRCS rangeland 

conservation practices on individuals, 

households, or social systems. It is likely that 

many producers do realize psychological 

benefits from conservation, as stewardship 

outcomes typically rank high among the 

management goals of livestock producers 

(Huntsinger and Fortmann 1990; Sayre 2004). 

Moreover, livestock producers who believe 

strongly in a responsibility to society are more 

likely to engage in environmentally desirable 

management practices, such as invasive weed 

control and riparian protection (Kreuter et 

al. 2006). Thus, non–peer-reviewed feature 

articles often refer to the psychological 

rewards ranchers enjoy because they employ 

conservation practices on land they hope 

to preserve for posterity (e.g., Little 2005; 

Smith 2008). Such rewards are often hard to 

document scientifically, however. 

Indirect evidence of psychological benefit from 

implementing conservation practices comes 

J. A. Tanaka, M. Brunson, and L. A. Torell 

from Holistic Management (HM), a program 

that typically advocates rotational grazing as part 

of an overall ranch management plan. Montagne 

and Orchard (2000) found that participating 

ranchers in the northern Rockies reported 

increased personal satisfaction after having 

adopted an HM approach. Stinner et al. (1997) 

found after interviewing HM participants 

nationwide that 91% reported improvements 

in quality of life after HM training. However, 

HM is a whole-ranch program that focuses on 

time management, goal setting, and monitoring 

as well as prescribed grazing, and neither study 

separated the effects of different aspects of the 

program. Moreover, such studies can speak 

only to the perceived effects of conservation; we 

have found no evidence that land managers, 

farm/ranch households, or group members that 

engage in conservation practices actually score 

higher on measurements of psychological or 

social well-being than producers who do not use 

such practices. 

Much more is known about why people choose 

to adopt conservation practices than the relative 

effectiveness of their implementation. Studies 

of innovation adoption offer insight as to 

which outcomes are anticipated by landowners 

and managers and the circumstances under 

which those outcomes are likely to be 

sufficiently valued to produce a change in 

management practice. Several thorough reviews 

New residential subdivisions 

in former ranchland, Eagle 

Mountain, Utah. (Photo: Mark 

Brunson) 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 399 
	 	 	 	

 

 

have been produced over the years (e.g., Nowak 

and Korsching 1983; Clearfield and Osgood 

1986). Most recently, Prokopy et al. (2008) 

reviewed 55 separate studies over a 25-yr period 

that explored adoption of agricultural best 

management practices (BMPs) in the United 

States. Their goal was to identify general trends 

in how adoption of conservation practices 

is related to social-psychological, enterprisebased, and social and economic factors. Most 

of the reviewed studies focus on soil, nutrients, 

and pest management; very few focused on 

the water or livestock management practices 

pertinent to grazing lands. Nonetheless, their 

findings offer general guidance about the role 

of anticipated benefits in the implementation 

of practices. 

The variables most strongly associated with 

adoption of BMPs were attributes of the 

decision maker or of the farm/ranch operation: 

demographic factors, such as income and 

education; access to information, capital, and 

social support; and farm size (Prokopy et al. 

2008). Producers’ awareness of environmental 

problems and their overall environmental 

attitudes were positively associated with BMP 

adoption, not surprisingly suggesting that 

farmers and ranchers are more likely to adopt 

conservation practices if they believe that 

conservation is important. 

Rogers (2003), whose theories on innovation 

adoption and diffusion have been highly 

influential in many fields including agriculture, 

identified three categories of factors that affect 

adoption rates: attributes of the potential 

adopter, the adopter’s social system, and 

the innovation itself. In this chapter, we are 

most interested in the latter category, as it 

encompasses the perceived personal, family, or 

social benefits that potential adopters ascribe 

to conservation activities on grazing lands. 

Wejnert (2002) further divides the pertinent 

innovation attributes into two decision 

metrics: the degree to which the benefits 

of a practice are thought to outweigh the 

costs and the degree to which the positive or 

negative consequences of adoption accrue to 

the private individual versus the public good. 

Both criteria relate to producers’ beliefs about 

personal, social, and economic factors as well 

as environmental benefits of a practice under 

consideration for adoption. Perceived costs can 

be psychological or social, just as are benefits; 

for example, Grigsby (1980) argued that one 

of the most significant barriers to innovation 

among ranchers is a belief that the innovation 

somehow threatens their ranching lifestyle. 

InnovAtIon–AdoPtIon StudIES 

Innovation–adoption studies abound in 

agriculture. Most of these focus on crop 

producers, but a number of researchers have 

explored practices recommended by the 

USDA-NRCS for grazing lands conservation. 

In this section, we describe more general 

studies that may include multiple practices; 

practice-specific research is described in 

subsequent sections. 

In southwestern Oregon, Habron (2004) 

found that landowners implemented upland 

conservation practices such as off-stream 

livestock water developments and rotational 

grazing more often than fencing or tree 

planting in riparian areas. Key influences on 

whether producers adopted any practice at 

all were whether they used irrigation, shared 

management decisions with a spouse, believed 

in scientific experimentation, and discussed 

conservation with others. The key factors 

predicting adoption of specific BMPs depended 

on the kind of practice implemented. 

A Utah study asked ranchers who have 

reputations as innovators what outcomes had 

led to their adoption decisions (Didier and 

Brunson 2004). One of the most influential 

outcomes was social: interviewees often 

reported that they were motivated by a desire 

to demonstrate stewardship to federal land 

managers and/or the public. The authors did 

find that innovation attributes were important 

to ranchers, especially in the negative. For 

example, interviewees reported that they or 

neighbors had rejected conservation practices 

because of perceived poor cost-to-benefit ratios 

or difficulty in pilot-testing a practice before 

full adoption. Brush management was cited 

as an example of the former barrier, while 

prescribed grazing—especially in the form of a 

short-duration rotational grazing system—was 

typically noted as an example of the latter. 

Barao (1992) surveyed Maryland producers 

who had attended an extension field day 

400 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	

	 	 	 	

	 	 	

	 	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

Measuring CO2 and 

temperature effects, High Plains 

Research Station, Cheyenne, 

Wyoming. (Photo: John Tanaka) 

to learn which demonstrated practices, if 

any, were subsequently adopted. Grazing 

management was the most common 

practice adopted (34%); livestock nutrition 

improvements, such as changing pasture 

species composition or analyzing forage/ 

feed, were less commonly adopted. Perceived 

outcomes of the practices were not found to 

be as important to these decisions, however, 

as the ease with which the practice can 

be learned and the results of a change can 

be observed. In another assessment of an 

extension program, a Livestock Systems 

Environmental Assessment tool, Koelsch et 

al. (2000) found that producers who used 

the tool were most likely to cite a desire for 

improved environmental stewardship as the 

most important reason for doing so. 

nonadoption 

A few studies have taken the opposite approach 

to understanding adoption decisions, and 

these also may prove useful—if people are 

not adopting conservation practices because 

they do not consider them worthwhile, that 

would suggest that such practices are not 

thought to provide personal benefits. However, 

Gillespie et al. (2007) found in a survey of 

1 700 Louisiana beef producers that the 

most influential reasons for nonadoption of 

16 BMPs were because they felt the practice 

would not work on their property or were 

unaware of the practice. Similarly, Prokopy et 

al. (2008) found that access to information 

influenced likelihood of adoption. Thus, 

knowledge of USDA conservation programs 

can also be an impediment to adoption. In 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 401 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	 	 	

	 		 	 	

Elk in the Pequop Range in Elko 

County, Nevada. (Photo: Tim 

Torell) 

Kansas, Smith et al. (2007) found that 80% of 

survey respondents knew about EQIP; 31% 

participated in the program. Those figures are 

considerably higher than in an earlier study 

by Cable et al. (1999), who found that only 

54% of survey respondents were aware of state 

or federal cost-share programs for private land 

conservation practices. For those who are aware 

of these NRCS cost-sharing programs but do 

not participate in them, lack of interest in 

conservation was not a significant influence on 

nonadoption; instead, ranchers tended to cite 

perceived regulatory impacts and paperwork 

as important reasons not to enroll (e.g., Didier 

and Brunson 2004; Smith et al. 2007). 

Geographical Regions 

Most of the studies described here pertain to 

a specific region, such as brush management 

in Texas (Taylor 2005; Kreuter et al. 2008) or 

prescribed burning in the desert Southwest 

(Sayre 2005). Some results of these studies 

are likely to be applicable to grazing lands 

nationwide, especially those relating to 

characteristics of land managers themselves, 

such as the importance placed on conservation 

as a management goal or sociodemographic 

and information access influences on 

adoption. 

However, geographic factors are likely to 

influence landowners’ beliefs about the 

perceived outcomes of conservation practices 

and thus the likelihood that those practices 

will be implemented. For example, Regen 

et al. (2008) found that protecting wildlife 

habitat and prairie restoration were important 

issues for a majority of landowners in a region 

straddling the Iowa–Missouri border. Most 

respondents reported using brush management 

practices to control eastern red cedar, but 

only 25% used prescribed burning, a practice 

frequently recommended for control of this 

species. In contrast, 38% of respondents to 

a survey by Kreuter et al. (2008) had used 

prescribed burning, mainly to control brush. 

Similarly, Liffman et al. (2000) found that 

25% of landowners in Alameda and Contra 

Costa counties, California, had used prescribed 

burning in the previous 5 yr, while 34% of 

those in Tehama County had done so. In both 

the Midwest and California situations, a likely 

explanation for lower burning rates is likely 

to be the juxtaposition of grazing lands with 

other land uses (cropland in Iowa and Missouri 

and rapid exurban development in Alameda 

and Contra Costa counties), whereas burning 

is less likely to pose liability and permitting 

difficulties in areas where grazing lands 

dominate. 

Prescribed Grazing 

As was noted previously, researchers 

studying HM have found that practitioners 

report improved quality of life as a result 

of participation in their decision-making 

program (Stinner et al. 1997; Montagne and 

Orchard 2000). A similar conclusion was 

derived from the Sustainable Grazing Systems 

program in southern Australia. This program 

was established in 1996 to address declining 

pasture productivity. Nearly 10 000 Australian 

livestock producers received training and new 

skills, participated in demonstrations, and 

integrated management and goal setting into 

their ranching operations, similar to those 

participating in the HM program. 

It is not known whether perceived 

improvements from program participation was 

a result of having employed a prescribed grazing 

system or some other factor associated with the 

learning experience but in the Montagne and 

Orchard (2000) study, interest in alternative 

grazing systems was one of the most frequently 

cited reasons for change. Overall, ranchers 

reveal positive ecological changes on the land 

and increased economic as well as personal 

satisfaction. In Minnesota, a psychological 

benefit was reported in collaborative research 

project by farmers and scientists that found 

that rotational grazing not only improved soil, 

pasture, and stream quality but also boosted 

the confidence of the farmers in their ability 

to employ more sustainable grazing practices 

(Badgley 2003). Again, it is not clear whether 

it was the grazing system or the collaborative 

process that improved farmers’ confidence levels. 

Collective interests of groups can also benefit 

from conservation practices beyond the 

individual benefits. Armstrong and Warner 

(1992) reported that adoption of a rotational 

grazing strategy by the Walker River Tribe 

in Nevada promoted tribal interests by 

benefiting all natural resources. This assumes, 

however, that resource benefits truly exist. As 

noted above, Briske et al. (2008) states that 

402 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 		

J. A. Tanaka, M. Brunson, and L. A. Torell 

even if evidence for benefits from rotational 

grazing is inconclusive, many livestock 

producers believe that such benefit exists. 

They suggest that “personal goals and values 

. . . are inextricably integrated within grazing 

systems, and they are likely to interact with the 

adoption and operation of grazing systems to 

an equal or greater extent than the underlying 

ecological processes” (p. 10). A basic theory 

of psychology holds that people are motivated 

to behave in ways that are consistent with 

their beliefs, and when evidence suggests that 

those behaviors are unhelpful, they may tend 

to reject the evidence rather than reject the 

belief (Festinger 1957). Consumer researchers 

(e.g., Mano and Oliver 1993) have applied 

this theory to explain postadoption satisfaction 

levels, suggesting that the act of having 

adopted a new product or behavior predisposes 

one to evaluate it positively. If we apply this 

to rotational grazing, one explanation for the 

continued perception of realized benefits is 

that there are psychological rewards associated predictable method for controlling brush.” 

with doing so. Interestingly, the authors recommended that 

to increase adoption rates, technology transfer 

Brush Management professionals should emphasize the short-term 

While brush management is a recognized economic benefits of Brush Busters rather 

NRCS conservation practice, it may or may than the long-term environmental benefits. 

not be considered “conservation,” depending This suggests that ranchers who implement 

on the purpose of the practice and the historic this practice may not obtain personal benefits 

and current conditions where it is implemented. from implementing brush management as 

For example, removing encroaching junipers to a conservation practice; rather, the benefits 

improve wildlife habitat and water availability in may accrue to society. Similarly, Thurow et 

central Texas might be considered conservation, al. (2001) found that economic factors were 

whereas removing all sagebrush from a native associated with ranchers’ willingness to enter 

shrub–steppe community and planting a into a brush control cost-share contract but that 

nonnative forage grass would not. Even in the conservation motives were not. Further evidence 

former case, brush removal might constitute that conservation alone cannot motivate 

“conservation” if intended to benefit black- brush management comes from Olenick et 

capped vireo (Vireo atricapilla) yet detrimental if al. (2005), who found in a 2003 Texas survey 

the site is occupied by golden-cheeked warblers that landowners generally held favorable views 

(Dendroica chrysoparia). Unfortunately, brush toward programs that would reduce brush cover 

management implementation studies do not to increase water yields or to improve wildlife 

necessarily distinguish between management habitat, but they disapproved of programs that 

for conservation and management for other would encourage the proliferation of woody 

purposes. plants in an attempt to increase atmospheric 

carbon sequestration. Landowner attitudes 

Kreuter et al. (2001) surveyed Texas county were also associated with the voluntariness and 

extension agents to assess landowner interest in flexibility associated with any proposed program 

and adoption of Brush Busters, a collaborative to enhance ecosystem services. 

extension/research program. Respondents 

reported that the landowners with whom Prescribed Burning 

they work perceive the program to be an Prescribed burning is a practice where the 

“inexpensive, convenient, safe, effective, and conservation benefits are offset by potential 

Moving cattle in the Blue 

Mountains, Oregon. (Photo: 

John Tanaka) 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 403 
	 	 	 	

risks, including a loss of forage, regulatory 

difficulties associated with smoke and burning 

permits, weak legal protections against 

liability, and potential escape of the fire onto 

a neighboring property (Liffman et al. 2000; 

Brunson and Evans 2005). Ranchers may 

believe that prescribed fire would benefit 

their land but are reluctant to implement. 

For example, Sayre (2005) studied eight 

locations in southern Arizona and New Mexico 

where prescribed burning was part of wildlife 

conservation efforts on grazing lands. He 

found that while interest in restoring fire to the 

landscape was high, use of prescribed burning 

was limited by trade-offs between conservation 

goals and forage availability and by real or 

perceived regulatory scrutiny. Burning was 

most feasible where there were institutional 

structures that allowed for collaborative 

management across ownership boundaries. 

In the Great Plains, an institution has arisen 

for just that purpose. Prescribed burning 

cooperatives offer landowners a chance to 

learn from peers how to apply fire safely and 

effectively and reduce liability concerns (Taylor 

2005). In Texas, Kreuter et al. (2008) found 

that members of a large cooperative had more 

positive attitudes than nonmembers about the 

ecological role of fire and the use of prescribed 

fire. The authors suggest that the group not 

only offers opportunities to learn and reduce 

liability but also promotes cooperative behavior 

that can benefit a ranching community. 

Thus, in the context of a prescribed burning 

association, there may be a social benefit to 

implementing this conservation practice. 

Rangeland Planting 

Few authors have specifically addressed 

the social aspects of rangeland planting as 

a conservation practice, but these provide 

interesting insights as to the influence of 

changing societal norms. As noted above, 

seeding of crested wheatgrass was one of 

the most common vegetation management 

practices in the western United States during 

the 1950s and 1960s. By many estimates, it 

was also one of the most economical practices 

because of the species’ competitiveness, early 

grazing use, and productivity. However, as the 

use and restoration of native plants has grown 

more popular, societal opinion has turned 

against the idea of replacing native rangeland 

with a monoculture of an exotic species that 

can persist for decades (Johnson 1986; Conner 

and Bach 2008). Negative characterizations 

of this highly adaptable and productive forage 

species have some critics who see planting 

crested wheatgrass as an ill-advised subsidy 

of ranching on public lands (Abbey 1988; 

Hess 1992) and those who complain that it 

has reduced sagebrush habitat for Greater 

sage-grouse (Connelly and Schroeder 2000). 

While it is still used on private ranches, crested 

wheatgrass seeding in public land grazing 

allotments has declined (Conner and Bach 

2008), often restricted to areas vulnerable to 

invasion by exotic annual grasses where rapid 

revegetation is needed for site stabilization after 

wildfire. Even then, plans may call for the use 

of “assisted succession” (Cox and Anderson 

2004) to replace crested wheatgrass with native 

perennial species as soon as is practical. 

Young and Clements (2009, p. 179) described 

these largely social pressures affected rangeland 

management: 

The politics of bureaucratic survival called for 

saying and doing as little as possible. Public 

land managers learned never to propose a 

seeding to increase forage supplies because 

government agencies, environmentalists, 

and archaeologists would descend en masse 

demanding a full environmental impact 

statement. Even wildfire burns were not seeded. 

When public pressure dictated seeding some 

very important habitat, the seed mixture 

was composed of species that had no chance 

of establishment and was seeded by aerial 

broadcasting on unprepared seedbeds. Young 

managers who tried seeding and failed were 

careful never to try again. 

Riparian Herbaceous Cover 

Considerable research has explored attitudes 

toward riparian restoration and protection, but 

almost all has focused on tree cover rather than 

herbaceous cover. For example, Lucht (2007) 

analyzed interest in adoption of agroforestry 

and conservation practices, including riparian 

planting, among agricultural producers, 

resident nonfarm landowners, and absentee 

nonfarm landowners. She found comparatively 

high interest but low knowledge levels 

compared to other practices. Absentee nonfarm 

landowners had the highest level of interest. 

404 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

	 	 	 	

	 	 	 	

	 	 	 	

	 	

	 	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

Public information campaign in 

an area undergoing transition 

from farm/ranch land to smallacreage subdivisions, Arimo, 

Idaho. (Photo: Mark Brunson) 

Ryan et al. (2003) found that Michigan farmers 

were motivated primarily to adopt conservation 

practices along riparian zones, not for the 

economic returns it would provide but because 

of their strong attachment to the land and 

their desire to convey the message that they 

are good stewards of the land. They noted that 

strategies for conservation must respect farmers’ 

attachment to the land, the desire to practice 

good stewardship while deriving income from 

the land. 

One study that did focus on nonwoodland 

planting was by Smith et al. (2007), and 

they reached conclusions similar to Ryan et 

al. (2003) about developing strategies for 

conservation. They found that Kansas farmers 

and ranchers were less likely to plant riparian 

filter (buffer) strips than to employ other forms 

of best management practices because filter 

strips must be enrolled into the Continuous 

Conservation Reserve Program, thereby 

incurring restrictions on haying and grazing 

use. 

upland Wildlife Habitat Management 

Upland wildlife habitat management, more so 

than other practices described here, is likely to 

enhance the use value for landowners as well 

as the bequest or existence values. The benefits 

that can be realized from wildlife are many and 

well known (Manfredo 2008). Because habitat 

enhancement also enhances the likelihood 

of being able to successfully view or hunt 

wildlife, many landowners will improve their 

land for wildlife. For example, when Cable et 

al. (1999) surveyed 900 Kansas agricultural 

producers about wildlife and riparian 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 405 
	 	 	 	

 

 

areas, they found that more than a third of 

respondents reported that they had idled land 

or changed management practices specifically 

to help wildlife. The most common “extremely 

important” motivations for doing so were to 

preserve wildlife for future generations (55.6%) 

and because the landowner enjoyed watching 

wildlife (51.7%). A nationwide survey of 

farmers and ranchers by Conover (1998) found 

that about half of respondents manage their 

properties to enhance attractiveness to wildlife. 

Both personal and monetary benefits 

influence decisions to implement wildlife 

habitat improvements, and it can be 

difficult to separate the two. Van Kooten 

and Schmitz (1992) found that agricultural 

producers participating in a waterfowl habitat 

enhancement project in western Canada held 

more positive attitudes toward wildlife than 

nonparticipants and therefore can be assumed 

to obtain personal benefits from participation, 

but positive attitudes alone were not sufficient 

to motivate habitat improvements in the 

absence of economic incentives. While this 

study was not completed in the United States, 

other studies conducted in this country (e.g., 

Troy et al. 2005) support the idea that the 

social or psychological benefits of wildlife 

habitat enhancement typically do not offset 

costs of doing so without some sort of 

economic incentive. 

Incentive to entice landowners to adopt 

wildlife management programs may take 

several forms. Conover (1998) reported that 

nearly 80% of farmers and ranchers had 

encountered some sort of wildlife damage 

on their properties. This is one of several 

impediments to implementation of wildlife 

habitat improvements. Cable (2002) reported 

that while Kansas agricultural producers 

believe that it is important to protect wildlife 

habitat, fewer than half had set aside any of 

their property for wildlife. The primary reasons 

were fear of increased trespassing by hunters 

if their land became known as especially 

attractive to wildlife and the cost of idling any 

of their land. Costs are even greater when the 

wildlife may be protected under the federal 

Threatened and Endangered Species list (Brook 

et al. 2003; Elmore et al. 2007). Thus, any 

positive personal outcomes associated with 

implementing the upland wildlife habitat 

management practice must be great enough 

that a landowner is willing to take on the 

risk of other consequences, such as damage, 

trespass, or increased regulatory scrutiny and 

reduced management flexibility. 

Incentives for wildlife habitat management 

on private land can be nonmonetary as well as 

monetary. Programs such as the Safe Harbor 

and Candidate Conservation Agreement with 

Assurances programs of the U.S. Fish and 

Wildlife Service as well as other collaborative 

land management efforts that exist across 

the West seek to protect landowners against 

regulatory risks in exchange for taking actions 

on behalf of wildlife (Belton 2008; Womack 

2008). Belton (2008) surveyed members of 

local working groups that attempt to maintain 

habitat for the greater sage-grouse. Participation 

in these efforts can promote cohesion among 

landowner neighbors and enhance cooperation 

with government agencies, but the “peace 

of mind” that such efforts are intended to 

provide are dependent on trust levels in the 

agencies responsible for wildlife management 

and protection. As before, Belton’s (2008) 

research suggested that people are more 

likely to participate if they receive monetary 

compensation for providing habitat for grouse. 

vALuE oF ECoSyStEM SERvICES 

types of values 

Ecosystem services benefit society in numerous 

and diverse ways. We can differentiate between 

those goods and services that are place bound 

(in situ) and those that can be derived from 

multiple locations (ex situ). There are a variety 

of classifications, including consumptive and 

nonconsumptive, market and nonmarket, 

primary and secondary, and in situ and ex 

situ (Brown et al. 2007; Cooper and Dobson 

2007; Breckenridge et al. 2008). The basic 

issue is how to account for all the benefits and 

costs associated with the services derived from 

rangeland ecosystems. As noted earlier, each 

of the conservation practices can potentially 

produce different kinds, qualities, and amounts 

of these goods and services, depending on 

location, natural potentials, current states, and 

other factors. 

Brown et al. (2007) used a traditional 

approach by dividing ecosystem goods into 

406 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	

	 	

	 	 	 	 	

	 	 	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

nonrenewable and renewable. Nonrenewable 

goods included rocks, minerals, and fossil 

fuels, while renewable goods include wildlife 

and fish, plants, water, air, soils, recreation, 

and aesthetics. Ecosystem services include 

purification of air and water, nutrient cycling, 

maintenance and renewal of soil and soil 

fertility, pollination of crops and natural 

vegetation, dispersal of seeds, maintenance of 

regional precipitation patterns, erosion control, 

biodiversity maintenance, control of pests 

affecting plants or animals, protection from the 

sun’s harmful UV rays, partial stabilization of 

climate, moderation of temperature extremes 

and the force of winds and waves, and 

mitigation of floods and droughts. In a listing 

of primary and secondary benefits from using 

pesticides, Cooper and Dobson (2007) divided 

primary benefits into agricultural production, 

energy needs, and preventing problems and 

secondary benefits into farming communities, 

national issues, and global issues. The 

interaction among all these benefit categories 

is complex and makes such separation difficult 

(see the section “Ecosystem Services”). 

Economic valuation 

In valuation of nonmarket ecosystem goods 

and services, there are few acceptable methods 

used in the literature. It is important to note 

that all these seek to estimate what a person 

or household would willingly pay to have that 

good or service or to put a value on damages 

from losses or costs avoided (Olewiler 2004). 

The comparability of these values with goods 

and services that are actually paid for out of the 

individual’s income remains a question. In other 

cases, for the experiments to have validity, the 

consumer being questioned needs to have a very 

clear idea of what goods or services are at stake. 

Before delving into the valuation question, it is 

important to note that the nature of ecosystem 

goods and services that are not traded in private 

markets lead to what is known as market 

failure and results in goods and services being 

either undersupplied or overused (Lant et al. 

2008). Goods and services that are provided 

through private markets involve feedbacks 

that can potentially provide efficient levels of 

production. Rules or incentives put in place to 

deal with market failures can lead to inefficient 

levels of production and essentially require the 

regulating agency to guess at market-clearing 

prices or quantities. It is important that 

economists work closely with ecologists and 

other specialists in a truly collaborative process 

in their efforts to estimate values (Heal and 

Barbier 2006). 

In a study that estimated the value of “diversity 

in biodiversity,” Christie et al. (2006) found that 

the public did not generally understand different 

attributes of biodiversity even as they valued 

biodiversity itself. The public was also found to 

be relatively indifferent as to how biodiversity 

was achieved, but most attributes of biodiversity 

examined had positive values. Of course, 

biodiversity as an attribute is a multifaceted 

concept that occurs at many scales (West 1993), 

and one has to carefully define what is meant by 

the term before it can be valued. 

It is also important to recognize who is likely to 

be the recipient of the benefits or who is setting 

the value (Burger et al. 2008). As they note, 

when valuing an individual species that does 

not have immediate or direct value, it is usually 

conservationists or regulators that set the value. 

On the other hand, when the individual species 

has a direct value to individuals, values are set 

through businesses, social scientists, or others 

with a direct connection to that species. The 

other types of resources they examined were 

the value of ecosystems to human communities 

and intact ecosystems with ecological, aesthetic, 

and existence values to people. Burger et al. 

(2008) also suggests specific economic value 

estimation methods that are appropriate for 

each type of environmental good or service 

(Table 5). The valuation methods that may 

be used in different situations are travel cost, 

tABLE 5. Economic estimation methods suggested by Burger et al. (2008). 

type of ecosystem good Methods 

Resources themselves Use survey of selected businesses 

Specific resources for 

individuals 

Use sample surveys to estimate direct 

values; estimate direct, indirect, and 

induced values using regional economic 

models 

Resources for 

communities 

Estimate replacement value, insurance 

costs, regional economics 

Intact ecosystems Use contingent valuation to estimate 

existence values; estimate regional 

economics 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 407 
	 	 	 	

 

 

 

 

 

 

 

 

 

FIGuRE 1. Water erosion benefit estimates (adapted from Hansen and Ribaudo 2008). 

FIGuRE 2. Wind erosion benefit estimates (adapted from Hansen and Ribaudo 2008). 

contingent valuation, hedonics, cost-based 

approaches, and factor-income approaches 

(Swinton et al. 2007). It is beyond this chapter 

to discuss these methods, but suffice it to say 

that each method is applicable to different 

situations and that the estimates derived have 

varying levels of confidence and comparability 

(Randall 2007). 

In the only national-level estimate of 

conservation values with regional applications, 

Hansen and Ribaudo (2008) estimated the 

values of various soil conservation benefits. 

Although their focus was primarily on the 

impacts on end uses of water, they do provide a 

value for reducing soil erosion. Different values 

were derived for water and wind erosion, and 

these values varied by region (Figs. 1 and 2). 

Huszar (1989) estimated that off-site economic 

costs exceeded on-site costs from wind erosion 

and concluded that if public action were to be 

warranted, it should be aimed at reducing offsite impacts. 

One of the issues in many valuation studies 

using willingness-to-pay measures is that while 

values are estimated, they do not address how 

values will change as supply and demand for 

that ecosystem good and service change. As 

an example, Loomis (2005) estimated values 

for outdoor recreation on public lands based 

on numerous studies. While these values may 

be valid, local conditions and the quantity of 

these goods and services nearby will affect these 

values. 

In a study to compare the value of ecosystem 

services from restored versus native land, 

Dodds et al. (2008) estimated values based on 

a broad literature search. Estimated values for 

rangeland regions are shown in Table 6. It is 

important to note that they generally estimated 

lower values for restored lands. The implication 

of these values may be that society values the 

maintenance of native rangelands more than 

restored or that restored lands have not been 

shown to be as productive in producing these 

ecosystem goods and services as intact native 

rangelands. 

Aesthetics are often cited as one of the 

important ecosystem services derived from 

rangelands. Most studies dealing with aesthetics 

have used contingent valuation. In one study 

that sought to actually quantify what ranch 

buyers would pay for a “quality-of-life” amenity 

that comes with owning the ranch, Torell 

et al. (2005b) found that ranch location, its 

scenic view, and the desirable lifestyle had 

more of an influence on ranch price than 

its potential to produce income. One of the 

implications of this is that while ranch owners 

may not respond to conservation practice 

408 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

J. A. Tanaka, M. Brunson, and L. A. Torell 

implementation to improve ranch income, they 

may respond to practices that enhance these 

factors. 

Social valuation 

There is a small but growing literature on the 

valuation of ecosystem goods and services 

with metrics other than monetary valuation. 

These are presented here under social valuation 

since they seek to find alternative metrics for 

evaluating trade-offs among different products 

or outcomes. 

One such approach has been to define a 

social-ecological system and solving a system 

of structural equations (Asah 2008). Finding 

the relationships between social systems and 

ecological systems remains a challenge. This 

method sought to relate a management goal 

with social knowledge of ecological responses. 

The results, however, are difficult to extrapolate 

because of the “place-specific nature of humanenvironment interactions” (Asah 2008). 

Another system uses what are termed “holistic 

ecosystem health indicators” to integrate 

ecological, social, and interactive indicators 

(Munoz-Erickson et al. 2007). The ecological 

indicators use biophysical measurements; the 

social indicators use demographics, economics, 

and quality-of-life metrics; and the interactive 

indicators use land use practices, policy, and 

collaboration measurements. Each of the 

measurements is weighted to derive a measure 

of overall holistic ecosystem health. 

RECoMMEndAtIonS 

Use of social and economic information can 

be incorporated into the NRCS conservation 

planning processes in a variety of ways. NRCS 

has done a commendable job of considering 

the ecosystem services in their planning and 

management processes. Acknowledgment 

of potential ecosystem services in the 

Conservation Practice Physical Effects 

Worksheets is an important step forward. We 

encourage the NRCS to continue to develop 

these worksheets, to refine their physical effects 

and rationale, and to define them on a more 

site-specific basis (i.e., by major land resource 

areas or ecological sites, as appropriate). We 

further recommend that the economic and 

cultural categories be expanded values of 

tABLE 6. Estimated values of ecosystem services per native hectare per year (in 2005 

dollars) (Dodds et al. 2008). 

Ecosystem service 

Great Plains north American deserts 

native Restored native Restored 

Gas regulation 7 6 — — 

disturbance regulation 7 7 2 1 

Water supply 28 19 85 25 

nutrient cycling 22 15 60 18 

Soil erosion control 241 175 237 65 

Commodities 3 853 2 490 — — 

Biodiversity 46 50 — — 

Recreation 1 003 1 003 16 16 

ecosystem services and social impacts at the 

individual, ranch or farm, and community 

levels. 

The method currently being used by 

NRCS to evaluate the benefits and costs 

contains the main features present in all 

the standard economic analyses (NPV) 

with the addition of values for selected 

ecosystem goods and services identified in 

each practice’s description (H. Gordon, 

personal communication, 2008). The 

NRCS should continue to refine how it 

incorporates ecosystem goods and services 

into its conservation practice analysis. While 

the current economic analysis spreadsheet 

incorporates factors from the Physical Effects 

Worksheets and attempts to place monetary 

values on each item, justification for those 

monetary values needs to be developed and 

standardized. While the approach is sound, 

without a sound basis, values for the various 

ecosystem goods and services can easily be 

manipulated to justify any project. It should 

also be recognized that a complete benefitto-cost assessment of selected conservation 

projects is not possible until valid estimates 

of economic value is assigned to presently 

unvalued ecosystem services. 

The NRCS should seriously review its costshare policies and requirements. If they are 

truly designed to pay for that portion of a 

conservation practice that benefits society, then 

the percentages should reflect that split. There 

are cost-share options that could be examined. 

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 409 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	

	 	 	

Bighorn sheep near Gabbs, 

Nevada. (Photo: Tim Torell) 

We recommend serious consideration of the 

first interpretation of cost share below: 

1.	� Set cost share based on the split between 

expected private and public benefits. In 

some instances, this could range from 

0% to 100%. Thus, rather than trying 

to estimate ecosystem goods and services 

values for each project, they could be 

determined for each practice through 

setting the appropriate cost share. Private 

landowners could then determine if their 

share of the total project cost would be 

covered by changes in their private benefits 

over the life of the project. 

2.	� Determine whether conservation is a 

priority and use cost-share amounts to 

promote adoption of those practices with 

the highest public benefit. If the only 

purpose behind cost share is to get the 

private entity to become invested in the 

practice, the level of investment should be 

examined with a view toward the impacts 

that higher, the same, or lower cost shares 

would have on adoption rates. 

KnoWLEdGE GAPS 

Knowledge gaps in the social and economic 

realms of conservation practices are numerous. 

Here we cover a few specific research needs 

for economics, ecosystem services, and social 

science. 

Economics of Conservation Practices 

Since most rangeland conservation practices will 

be implemented by private ranchers throughout 

the western United States, it is necessary to 

understand how changes affect the overall 

economics of the ranching operation. Economic 

analysis of conservation practices can start with 

a basic efficiency estimate, such as present net 

worth, benefit-to-cost analysis, or internal rate of 

return. While that will provide a basic estimate 

of profitability, ranchers also need to understand 

how the change will affect their entire operation, 

and society needs to understand the larger-scale 

social benefits and costs. 

In terms of the actual conservation practice, the 

methodology for economic analysis as related 

to livestock production is well understood. 

Research that is needed at this level is 

knowledge about the physical responses (e.g., 

additional production and seasonality), the 

costs of inputs and outputs, and the timing of 

benefits and costs. Caton et al. (1960) noted 

that early attempts by agricultural economists 

participating in a West-wide regional research 

project to quantify the economics of rangeland 

management practices were hampered by a lack 

of response data. Potential benefits of many 

improvements could not be assessed because 

long-term studies had not been undertaken to 

quantify the forage and livestock responses that 

were realized from the various practices. This 

limitation continues. 

The economic impacts of the conservation 

practice on the entire ranching operation 

require additional research. Most of the 

relationships are biological in nature (e.g., 

livestock production and forage production) 

and require knowledge about annual cycles 

and longer-term responses. Each livestock 

production cycle will have its own unique 

attributes, depending on geographic location, 

type of animal species, and production goals. 

In order to model within a year, these factors 

must be understood. When management 

changes with resulting changes in herd size, it 

may take several years for the ranch to come 

to a new equilibrium herd size. On the forage 

side, within-year variation of production affects 

the amount of feed available for the herd. 

Within the year and across years, temperature 

and precipitation will also affect the amount 

of feed available for livestock. Many of the 

conservation practices will affect the potential 

amount of forage production. 

Social Aspects of Conservation 

Practices 

There are many knowledge gaps related 

to social aspects of conservation practices. 

410 Conservation Benefits of Rangeland Practices 
	 	 	 	 	 	 	 	

	 	 	 	 	 	 	 	 	 		

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	

	 	 	

J. A. Tanaka, M. Brunson, and L. A. Torell 

The problem with many of these is that the 

knowledge that is needed is place based. What 

might be socially acceptable for a national 

program may not be at the local level. While 

there have been a few studies looking at 

adoption of conservation practices, these need 

to be done at more locations and specific to 

NRCS conservation practices on rangelands. 

In addition to location, such information for 

individuals, operations, and social systems will 

be useful in designing programs. 

Ecosystem Goods and Services 

valuation 

Quantifying societal benefits and the economic 

value of previously nonquantified ecosystem 

services are the areas where economic 

evaluation of conservation practices is most 

lacking. As noted earlier, livestock production 

benefits do not justify the total cost of many 

conservation practices (McBryde et al. 1984; 

Lee et al. 2001; Torell et al. 2005a; Skaggs 

2008). Economic evaluations are incomplete, 

and the assumption is made that ecosystem 

services not quantified in the analysis offset the 

excess cost of the practice not justified from 

marketable goods and services. This may not 

be the case. Only by quantifying and assigning 

economic value to selected ecosystem services 

that are currently only qualitatively noted can 

a complete economic assessment be made. 

Quantifying benefits accruing to the public at 

large could also be used to justify cost-share 

percentages on a case-by-case (or location-bylocation) basis. This would mean an expanded 

use of nonmarket valuation techniques (see 

Champ et al. 2003). We note that while 

it would be advantageous to know how 

people value each of the goods and services 

produced by different conservation practices, 

it is probably neither likely nor feasible. Yet 

expanding the economic analysis to quantify 

all potential benefits has several important 

implications. First, the quantification may 

show a very positive benefit-to-cost ratio from 

society’s point of view, suggesting that even 

more should be done. In other cases, it may 

demonstrate that the conservation practice is 

not justified. If not valued by society, based on 

economics, transfer payments and willingness 

to pay, inaction, and a deteriorated ecosystem 

may be the preferred state. Dismal federal and 

state budget situations highlight that tradeoffs exist. 

Most of the knowledge base on the values 

of nonmarket goods and services is time 

and space specific given the methodologies 

currently in use. Aggregation of numerous 

studies can provide value ranges that various 

researchers have estimated using methods such 

as travel cost, contingent valuation, or hedonic 

models. Whether these are appropriate for the 

NRCS to use in evaluating their conservation 

practices through extrapolation is a question 

for investigation. The basic knowledge gaps are 

the values for each ecosystem good or service in 

each location and at the specific time. 

ConCLuSIonS 

use of Economic Information in 

Resource Planning and Management 

Economic values for market and nonmarket 

ecosystem goods and services will vary by 

location and time. What might be available at 

the scale necessary for ranch-level planning are 

indicators of relative values. There have been 

many studies of individual benefits for specific 

goods and services. Those that exist in markets 

provide what people are truly willing and 

able to pay. Some methods, such as hedonic 

pricing and willingness-to-pay studies, can help 

determine values of specific characteristics. 

Relative values of the various goods and services 

can provide information to planners and 

managers if they are collected appropriately 

and in a consistent manner. There are 

numerous peer-reviewed articles for many 

of the conservation practices that have done 

an economic analysis of the specific practice 

comparing the cost of the practice with the 

estimated market benefits. We have not 

reviewed those studies because they are not 

particularly enlightening for this project. The 

method currently being used by the NRCS 

to evaluate the benefits and costs contains 

the main features present in all the standard 

analyses with the addition of values for the 

various ecosystem goods and services identified 

in each practice’s description (H. Gordon, 

personal communication, 2008). 

If economic feasibility of various conservation 

practices is important, it would be prudent for 

NRCS to become seriously involved in finding 

ways to assess the economic value of ecosystem 

goods and services beyond those found in 

Quantifying 

societal benefits 

and the 

economic value 

of previously 

nonquantified 

ecosystem 

services are the 

areas where 

economic 

evaluation of 

conservation 

practices is most 

lacking.”

CHAPTER 9: A Social and Economic Assessment of Rangeland Conservation Practices 411 
	 	 	 	

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

	 	 	 	 	

	 	 		

	 	

Old cabin and scenic view, 

Owyhee County, Idaho. 

(Photo: John Tanaka) 

the marketplace. As noted, when considering 

only forage value or livestock production 

from rangelands as the primary benefit, 

many conservation practices will not show a 

favorable benefit-to-cost ratio for conservation 

programs. As the ecosystem goods and services 

are considered and valued, the NRCS can 

allocate the conservation practice costs to the 

private landowner and to taxpayers through its 

cost-share mechanism based on the expected 

proportion of benefits going to the different 

parties. For example, if a prescribed grazing 

practice does not increase livestock production 

and net returns but produces significant social 

benefits, the taxpayers may be allocated a larger 

proportion of the costs of implementation than 

if livestock benefits were higher. Some of the 

social benefits may go to the private landowner, 

and this should also be taken into account (e.g., 

maintaining a way of life). At present, while 

cost-share mechanisms would seem to implicitly 

recognize these social benefits, there is little 

information to justify the cost-share percentages 

on a case-by-case (or location-by-location) basis. 

It is not reasonable to assume that the social 

benefits from a given conservation practice 

is the same everywhere or that they offset 

substantial practice costs in many cases. 

Quantitative estimates of the value of 

ecosystem service are largely nonexistent, but 

quantification of traditional market values 

are lacking as well. Efforts to quantify the 

economics of rangeland management practices 

have been hampered by a lack of response 

data. Long-term range and grazing studies 

that monitored the production response of 

management practices were found to be a 

major shortcoming of economic assessments 

of rangeland management practices that have 

continued since the 1960s. These economic 

assessments of rangeland management 

practices are based on very limited data and 

usually use simulated biophysical data. Longterm range and grazing studies are becoming 

even less common. Response relationships 

among conservation practices and other 

ecosystem goods and services are even rarer or 

nonexistent. 

use of Social Information in Resource 

Planning and Management 

Social values and attitudes have clearly 

impacted adoption of rangeland management 

practices and how rangeland policy and 

management has progressed. As an example, 

Young and Clements (2009, p. 178.) 

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J. A. Tanaka, M. Brunson, and L. A. Torell 

concluded, “Public rangeland management 

agencies did not drop the use of herbicides 

because they were afraid of the environmental 

consequences of using pesticides; they dropped 

them because they were afraid of comments 

from a highly vocal but not necessarily 

knowledgeable portion of the general public. 

Congress stopped appropriating money for the 

improvement of publicly owned rangelands to 

avoid criticism from environmental groups.” 

The numerous other examples described above 

clearly show that social factors have motivated 

land managers to behave in ways beyond profitmaximizing behavior. 

If social information is going to be used in 

resource planning and management, social 

indicators need to be added to the list of 

benefits along with a description of how 

to use and interpret the indicators. Our 

assumption is that most resource managers 

do not know what indicators would be 

appropriate or how to use them in decision 

making. The significant lack of social research 

on the effects of the conservation practices 

on landowners will make implementation of 

this recommendation difficult in any sort of 

quantitative manner. 

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422 Conservation Benefits of Rangeland Practices 
	 	 	

 

 

	 	

	 	 	

 

	 	

	 	 	

	 	 	

	 	

	 	

 

	 	

	 	

	 	

NRCS AdviSeRS to CeAP RANgelANd liteRAtuRe SyNtheSiS teAmS 

CoRe Review teAm 

Leonard Jolley (Leader) 

George Peacock 

Dennis Thompson 

Kenneth Spaeth 

Joel Brown 

Curtis Talbot 

PReSCRibed gRAziNg 

Patrick L. Shaver (Leader) 

Arnold Norman 

Steve Barker 

Stan Boltz 

Shane Green 

heRbACeouS weed CoNtRol 

Jeff Repp (Leader) 

Loretta Metz 

Jeff Printz 

PReSCRibed buRNiNg 

Patrick L. Shaver (Leader) 

Homer Sanchez 

Chuck Stanley 

Gene Fults 

bRuSh mANAgemeNt 

Patrick L. Shaver (Leader) 

Homer Sanchez 

Chuck Stanley 

George Chavez 

RANge PlANtiNg 

Arnold Norman (Leader) 

Chuck Stanley 

Jon Gustafson 

Jeff Burwell 

RiPARiAN heRbACeouS CoveR 

Jeff Repp (Leader) 

Gene Fults 

Loretta Metz 

Marcus Miller 

wildlife uPlANd hAbitAt 

mANAgemeNt 

Wendell Gilgert (Leader) 

Gene Fults 

George Chavez 

Gerald Jasmer 

Reggie Blackwell 

lANdSCAPe ANAlySiS 

Gene Fults (Leader) 

Jon Gustafson 

SoCioeCoNomiCS ANd 

eCoSyStem SeRviCeS 

Hal Gordon (Leader) 

Reggie Blackwell 

Jeff Repp 

Advisors and External Reviewers 423 
	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	

	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	

	 	 	 	

	 	

	 	 	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	

	 	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	

	 	 	

	 	 	 	 	 	 	

	 	 	 	

	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

eXteRNAl RevieweRS 

Jim Ansley, Professor, Texas AgriLife
 

Research, PO Box 1658, 11708
 

Highway 70 S, Vernon, TX 76384-1658
 

Mike Borman, Professor and Head,
 

Department of Rangeland Ecology and
 

Management, 200A Strand Agriculture
 

Hall, Oregon State University, Corvallis,
 

OR 97331-2218
 

John Briggs, Professor and Director,
 

Konza Prairie Biological Station, 116
 

Ackert Hall, Division of Biology, Kansas
 

State University, Manhattan, KS 66506
 

John Brock, Professor Emeritus,
 

Environmental Resources, School of
 

Planning and Landscape Architecture,
 

Arizona State University, Tempe, AZ
 

85287-2005
 

Steve Bunting, Professor, Department of
 

Rangeland Ecology and Management,
 

University of Idaho, Moscow, ID 83844-

1135
 

Chris Call, Professor, Department of
 

Wildland Resources, 5230 Old Main
 

Hill, Utah State University, Logan, UT
 

84322-5230
 

Ray Callaway, Professor, Division
 

of Biological Sciences, 32 Campus
 

Dr, HS104, University of Montana,
 

Missoula, MT 59812
 

Jean Chambers, Research Ecologist,
 

US Forest Service, Reno Great Basin
 

Ecology Laboratory, 920 Valley Rd,
 

Reno, NV 89512
 

Richard Conner, Professor, Department
 

of Agricultural Economics, 2124 TAMU,
 

Texas A&M University, College Station,
 

TX 77483-2124
 

Robert Cox, Assistant Professor,
 

Department of Natural Resources
 

Management, Goddard Building, Texas
 

Tech University, Box 42125, Lubbock,
 

TX 79409
 

Randy Gray (retired), National Wildlife 

Biologist, USDA Natural Resources 

Conservation Service, Washington, DC 

Jonathan Haufler, Executive Director,
 

Ecosystem Management Research
 

Institute, 210 Borderlands, PO Box 717,
 

Seeley Lake, MT 59868
 

Anne Hild, Professor, Department of
 

Renewable Resources, University of
 

Wyoming, 1000 E University Ave,
 

Laramie, WY 82071
 

Jerry Holechek, Professor, Animal and
 

Range Sciences, Knox Hall, Box 30003,
 

MSC 3-I, New Mexico State University,
 

Las Cruces, NM 88003-8003
 

Carolyn Hunsaker, Research Ecologist,
 

US Forest Service, Pacific Southwest
 

Research Station, 800 Buchanan St,
 

West Annex Building, Albany, CA
 

94710-0011
 

Mort Kothmann, Professor, Ecosystem
 

Science and Management, 2138
 

TAMU, Texas A&M University, College
 

Station, TX 77843-2138
 

Clayton Marlow, Professor, Department
 

of Animal and Range Sciences, PO Box
 

172900, Montana State University,
 

Bozeman, MT 59717-2900
 

Mitch McClaran, Professor, School of
 

Natural Resources, 325 Biosciences
 

East, University of Arizona, Tucson, AZ
 

85721
 

424 Conservation Benefits of Rangeland Practices 
	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	

	 	 	 	

	 	 	 	

	 	 	 	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	

	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	

	 	 	 	

	 	 	

	 	 	 	 	

	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

	 	

	 	 	 	

	 	 	 	 	

	 	 	 	

Don McLeod, Associate Professor, 

Department of Agricultural and Applied 

Economics, University of Wyoming, 

1000 E University Ave, Laramie, WY 

82071-3354 

Keith Owens, Professor and Head, 

Department of Natural Resources 

Ecology and Management, 008C 

Ag Hall, Oklahoma State University, 

Stillwater, OK 74077 

Fred Pierson, Research Leader, 

Northwest Watershed Research Center, 

800 Park Blvd, Suite 105, Boise, ID 

83712-7716 

Alan Rasmussen, Dean, College of 

Agriculture, Natural Resources and 

Human Sciences, Support Services 

Building, MSC 156, Texas A&M 

University–Kingsville, Kingsville, TX 

78363-8202 

Neil Rimbey, Professor, Caldwell 

Research and Extension Center, 

University of Idaho, 1904 E Chicago 

A-B, Caldwell, ID 83605 

Bill Rogers, Associate Professor, 

Ecosystem Science and Management, 

2138 TAMU, Texas A&M University, 

College Station, TX 77843-2138 

Tony Svejcar, Research Leader, Eastern 

Oregon Agricultural Research Center, 

67826-A, Highway 205, Burns, OR 

97720-9399 

Amy Symstad, Research Ecologist, 

USGS Black Hills Field Station, 26611 

US Highway 385, Hot Springs, SD 

57747 

Robert Washington-Allen, Assistant 

Professor, Ecosystem Science and 

Management, 2138 TAMU, Texas A&M 

University, College Station, TX 77843-

2138 

Advisors and External Reviewers 425 
	 	 	 	

	 	 	 	

	 	 	

	 	

	 	 	 	 	

	 	 	

	 	 	 	 	

	 	

	 	 	

	 	 	 	

	 	

	 	 	

	 	

	 	

	 	

	 	 	

	 	 	

	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	 	

	 	

	 	

	 	 	 	

	 	

	 	

	 	 	

	 	

	 	 	 	 	 	

	 	 	 		

	 	

	 	 	

	 	

	 	

	 	 	

	 	

	 	

	 	

	 	

	 	 	 	 	

	 	 	 	

	 	 	 	

	 	 	

	 	

	 	 	

	 	 	

	 	 	

	 	 	 	 	

	 	 	 	 	

	 	

	 	 	 	 	

	 	 	

	 	

	 	 	

	 	

	 	 	

	 	 	

	 	 	 	 	

InDEX 

A 

amphibians, conservation practices &, 273, 

275–276, 278–279 

arthropods, fire &, 87 

atmosphere 

fire &, 92–93; 

woody plant encroachment &, 115–116, 

146 

B 

biodiversity 

economic value of, 407; 

woody plant encroachment &, 117–118, 

127–128, 148 

brush management 

assessment of, 107–170; 

economics of, 387–392, 403; 

ground water recharge &, 133–135 

history of, 118–120; 

knowledge gaps of, 146–148 

livestock &, 129–131; 

planning of, 120–122; 

recommendations on, 145–146; 

soils &, 111–115, 123–125 

treatment longevity of, 143–145 

wildlife &, 136–145 

burning 

brush management &, 145, 148; 

conservation practices &, 77–104; 

economics of, 392–394, 403–404; 

knowledge gaps of, 99 

riparian management &, 225–226, 240; 

recommendations for, 98–99 

smoke management &, 92–93 

weed control &, 300–301, 305 

C 

carbon sequestration 

fire &, 90–92; 

grazing &, 46–47; 

range planting &, 185; 

riparian management &, 236–237; 

woody plant encroachment &, 111, 114, 

124–127 

carnivores 

conservation practices &, 265–268, 

276, 278; 

livestock &, 265–268 

conservation effects assessment project 

goals of, 1–2 

overview of, 3–4 

practices evaluated &, 5–6 

rangelands &, 4–5 

climate 

brush management &, 137; 

carbon sequestration &, 46; 

fire &, 99–100; 

range planting &, 175–176, 181–183, 

186–188; 

woody plant encroachment &, 116–117 

cultivation, range planting &, 177–178 

current status of knowledge 

summary of, 12–19 

knowledge gaps of, 19–20 

D 

defoliation, 28, 31–32, 47–49 

drought, grazing &, 26–27 

E 

economic value, conservation practices &, 

129–131, 373–422 

ecosystem services, value of, 228–229, 

377–382, 403–409 

ecological site, definition of, 354, 359– 

360 

erosion 

brush management &, 131–136, 149; 

control of, 309–310; 

fire &, 87, 97–98; 

grazing &, 33–34; 

range planting &, 184; 

riparian systems &, 233 

woody plant encroachment &, 111, 114 

426 Conservation Benefits of Rangeland Practices 
	 	 	 	

	 	 	

	 	

	 	 	

	 	 	

	 	

	 	

	 	 	 	 	

	 	

	 	 	 	

	 	

	 	 		

	 	 	

	 	

	 	

	 	 	 	 	

	 	 	 	

	 	 	 	 		

	 	 	

	 	 	 	

	 	

	 	 	

	 	 	 	 	

	 	 	

	 	 	 	 	

	 	 	

	 	 	

	 	 	

	 	

	 	 	

	 	 		

	 	

	 	 	 	 	

	 	 	

	 	 	

	 	 	 	 	 	

	 	 	 	 	 	

	 	 	

	 	 	 	

	 	 	 		

	 	

	 	 	

	 	 	 		

	 	 	

	 	 	 	 		

	 	

	 	

	 	 		

	 	

	 	

	 	

	 	

	 	 	 	

	 	

	 	 	

	 	

	 	

	 	

	 	 	

	 	 	

	 	 	

	 	 	 	

	 	 	 	

	 	 	

	 	

F 

fencing 

livestock &, 45, 49, 58; 

riparian areas &, 217 

fire 

behavior of, 93–95; 

brush management &, 145, 148; 

conservation practices &, 75–104; 

grazing &, 44; 

historical perspective of, 87–89 

invasive plant management &, 300–301, 

308; 

management of, 93–95; 

native species &, 305, 308; 

research, 95–101; 

riparian management &, 225–226, 

235–236, 240 

forage quantity and quality 

brush management &, 129–131; 

fire &, 82; 

grazing &, 25, 27–29, 49–53, 269–274; 

invasive plant management &, 297–303 

G 

game birds, conservation practices &, 

256–261, 277 

grazing 

amphibians &, 273, 275–276; 

continuous, 26, 28–30, 32–33, 42–44; 

drought &, 26–27; 

economics of, 382–387, 402–403; 

fire &, 44, 82, 86, 96; 

game birds &, 258–259; 

invasive plant management &, 301–302, 

304; 

knowledge gaps of, 55–57 

livestock &, 45–46, 258–265; 

nongame birds &, 262–265; 

recommendations for, 49–55 

reptiles &, 273, 275–276; 

riparian management &, 215–224, 

226–227, 229–233, 240; 

rotational, 26, 28, 29–34, 42–44, 57 

small mammals &, 273–276; 

soil carbon &, 46 

stocking rate &, 25, 27, 32, 34 

systems of, 26, 29, 32, 42, 52
 

wildlife &, 256–259, 268–273;
 

woody plant encroachment &, 109
 

H 

habitat 

conservation practices & wildlife, 

255–290; 

fire &, 86–87; 

invasive species &, 311–314 

herbicides 

brush management &, 119, 124, 

126–128, 133, 135, 141–142; 

invasive plant management &, 301, 

304–305, 317–319 

herbivores 

brush management &, 129–131; 

riparian management &, 215–224, 

226–227, 229–233, 240 

heterogeneity 

societal, 350–351; 

spatial, 345–347 

I 

invasive plants 

biocontrol &, 317–319 

encroachment of, 107–170; 

fire &, 84; 

integrated pest management &, 293–294 

invasion mechanisms of, 307–309 

knowledge gaps of, 319–323 

management of, 291–336; 

prevention of, 294–297 

recommendations for, 319–323 

riparian management &, 233 

wildfire prevention &, 314–316 

L 

landscape approach 

conservation value of, 344–351 

conservation practices &, 257–258, 

337–370; 

definition and concepts of, 340–343 

knowledge gaps of, 363–364 

model-based, 355–360 

Index 427 
	 	 	 	

	 	 	 	

	 	

	 	

	 	

	 	

	 	

	 	 	

	 	

	 	 	 	 	

	 	 		

	 	

	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	

	 	 	 	

	 	 	 		

	 	 	 		

	 	

	 	 	

	 	 	

	 	

	 	 	 	

	 	 	 	

	 	 	 		

	 	 	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	

	 	 	 	 	

	 	

	 	 	

	 	 	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	

	 	

	 	 	

	 	

	 	 	

	 	 	 	

	 	 	

	 	

	 	

	 	 	 	

	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	 	 	

	 	 	

	 	 	 	 	 		

	 	 	 	

spatial hierarchy &, 351–355
 

recommendations for, 360–363
 

livestock 

brush management &, 129–131; 

distribution of, 45–46; 

fire &, 92; 

grazing &, 258–265; 

invasive plants &, 297–299 

predators &, 265–268; 

production of, 23–24, 28, 30, 55; 

riparian management &, 215–224, 

226–227, 229–233, 240; 

wildlife &, 256–259, 262–265, 268– 

276 

M 

mowing, invasive plant management &, 

303, 305 

mulching, range planting &, 178, 184–185 

N 

nongame birds 

brush management &, 139, 260; 

grazing &, 35–37, 261–265, 277 

P 

plant-materials, range planting &, 

176–177, 188–189 

pollution, riparian management &, 

228–229, 240 

precipitation 

brush management &, 137; 

range planting &, 175–176 

predators, livestock &, 265–268 

R 

range planting 

brush management &, 128 

depth and rate of, 179–181 

economics of, 394–396, 404; 

knowledge gaps of, 185–190 

integrated pest management &, 303, 

305–306 

materials selection &, 175–177
 

recommendations for, 185
 

seedbed preparation &, 177–178
 

weather forecasting &, 186–188
 

reptiles, conservation practices &, 35, 273, 

275–276, 278–279 

riparian management 

assessment of, 215–252; 

economics of, 396–397, 404–405; 

grazing &, 33, 219–224, 263–266, 

273–276; 

invasive plant management &, 310–311; 

knowledge gaps of, 239 

recommendations on, 237–239 

restoration of, 226–227 

stream temperature &, 234–235 

water quality &, 229–230 

S 

socioeconomic assessment 

conservation expenditures &, 378–382 

conservation impacts &, 373–376 

innovation and adoption &, 400–406 

knowledge gaps of, 409–411 

social considerations of, 399–406 

recommendations for, 409–410 

small mammals 

conservation practices &, 273–276, 

278–279; 

fire &, 86, 97; 

grazing &, 35, 37–38, 40; 

riparian systems &, 232–233 

smoke management, fire &, 92–93 

social value, conservation practices &, 

371–422 

soils, fire &, 89–92, 97–98, 

spatial pattern, landscape approach to, 

347–348 

stocking rate 

economics of, 383–386, 388; 

grazing &, 23, 25–30, 32–35, 42, 

44–45, 47, 49–54, 57, 258–260 

428 Conservation Benefits of Rangeland Practices 
	 	 	 	

	 	 	

	 	

	 	 	

	 	 	 	 		

	 	

	 	 	

	 	

	 	 	 	 	

	 	 	

	 	

	 	

	 	

	 	 	 	 	

	 	 	 	

	 	

	 	 	

	 	

	 	 	

	 	 	 	

	 	

	 	 	

	 	

	 	 	

	 	 	 	

	 	 	

	 	 	

	 	 	 	 	 	 	

	 	

	 	

	 	 	 	

	 	

	 	 	

	 	

	 	 	

U 

ungulates 

conservation practices &, 268–273, 278; 

fire &, 86, 99; 

grazing &, 40–43; 

riparian systems &, 223 

upland wildlife habitat, economics of, 

397–399, 405–406 

W 

water 

fire &, 87, 97–98; 

grazing &, 32–33; 

invasive plant management &, 309–311, 

316; 

range planting &, 184–185 

watershed 

brush management &, 131–136; 

grazing &, 33 

weather 

brush management &, 137; 

range planting &, 175–176, 181–183, 

186–188; 

woody plant encroachment &, 116–117 

weed control 

biological, 302–303, 309; 

grazing &, 301–302, 308–309; 

herbicides &, 301; 

mechanical, 303; 

range planting &, 182 

wildfire 

brush management &, 136–145, 148, 

150; 

grazing &, 44 

wildlife 

conservation practices &, 253–290; 

economics of, 397–399 

fire &, 86–87, 97; 

grazing &, 34–44, 262, 273–275 

invasive species &, 311–314 

knowledge gaps of, 277–279 

stocking rate &, 24, 34–35, 37, 40–42, 

44, 50–53, 55–57 

recommendations on, 276 

reptiles and amphibians &, 275–276 

riparian management &, 230–233 

woody plants 

encroachment of, 107–170, 260; 

fire &, 95–96; 

riparian management &, 234–235, 240 

Index 429 
 

 

 

 

Contributors AnD ExtErnAl rEviEwErs 

Academic Coordinator 

David D. Briske 

nrCs Coordinator 

Leonard Jolley 

External book reviewer 

Thomas L. Thurow 

ExtErnAl rEviEwErs 

Jim Ansley 

Mike Borman 

John Briggs 

John Brock 

Steve Bunting 

Chris Call 

Ray Callaway 

Jean Chambers 

Richard Conner 

Robert Cox 

Randy Gray 

Jonathan Haufler 

Anne Hild 

Jerry Holechek 

Carolyn Hunsaker 

Mort Kothmann 

Mitch McClaran 

Don McLeod 

Clayton Marlow 

Keith Owens 

Fred Pierson 

Alan Rasmussen 

Neil Rimbey 

Bill Rogers 

Tony Svejcar 

Amy Symstad 

Robert Washington-Allen 

PrEsCribED GrAzinG 

team leader—David Briske 

Members—Justin Derner, Dan Milchunas, Ken Tate 

nrCs team leader—Pat Shaver 

Members—Arnold Norman, Steve Barker, Stan Boltz, Shane Green, Jeff Printz 

PrEsCribED burninG 

team leader—Sam Fuhlendorf 

Members—Ryan Limb, Dave Engle, Rick Miller 

nrCs team leader—Pat Shaver 

Members—Homer Sanchez, Chuck Stanley, Gene Fults 

brush MAnAGEMEnt 

team leader—Steve Archer
�

Members—Kirk Davies, Tim Fulbright, Kirk McDaniel, Brad Wilcox, 

Katharine Predick 

nrCs team leader—Pat Shaver 

Members—Homer Sanchez, Chuck Stanley, George Chavez 

rAnGE PlAntinG 

team leader—Stuart Hardegree 

Members—Tom Jones, Bruce Roundy, Nancy Shaw, Tom Monaco 

nrCs team leader—Chuck Stanley 

Members—Arnold Norman, Jon Gustafson, Jeff Burwell 

riPAriAn hErbACEous CovEr 

team leader—Mel George 

Members—Randy Jackson, Chad Boyd, Ken Tate 

nrCs team leader—Jeff Repp 

Members—Gene Fults, Lori Metz, Marcus Miller 

uPlAnD wilDlifE hAbitAt MAnAGEMEnt 

team leader—Paul Krausman 

Members—Vern Bleich, Bill Block, Dave Naugle, Mark Wallace 

nrCs team leader—Wendell Gilgert 

Members—Gene Fults, George Chavez, Jerry Jasmer, Reggie Blackwell 

hErbACEous wEED Control 

team leader—Roger Sheley 

Members—Jeremy James, Matt Rinella, Dana Blumenthal, Joe DiTomaso 

nrCs team leader—Jeff Repp 

Members—Lori Metz, Jeff Printz 

lAnDsCAPE AnAlysis 

Chapter leader—Brandon Bestelmeyer 

Members—Ben Wu, Joel Brown, Sam Fuhlendorf 

nrCs team leader—Gene Fults 

Members—John Gustafson 

soCioEConoMiCs AnD ECosystEM sErviCEs 

team leader—John Tanaka 

Members—Allen Torell, Mark Brunson 

nrCs team leader—Hal Gordon 

Members—Reggie Blackwell, Jeff Repp