United States

Department

of Agriculture

Forest Service

Rocky Mountain

Research Station

General Technical

Report RMRS-GTR-169

April 2006

Responses of Plant Communities 

to Grazing in the Southwestern 

United States

Daniel G. Milchunas
Rocky Mountain Research Station

Natural Resources Research Center

2150 Centre Avenue, Building A

Fort Collins, CO 80526

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Milchunas, Daniel G. 2006. Responses of plant communities to grazing in the southwestern 

United States. Gen. Tech. Rep. RMRS-GTR-169. Fort Collins, CO: U.S. Department of Agriculture, 

Forest Service, Rocky Mountain Research Station. 126 p.

Abstract

Grazing by wild and domestic mammals can have small to large effects on plant communities, depending on characteristics of the particular community and of the type and intensity of grazing. The broad 

objective of this report was to extensively review literature on the effects of grazing on 25 plant communities of the southwestern U.S. in terms of plant species composition, aboveground primary productivity, and root and soil attributes. Livestock grazing management and grazing systems are assessed, as 

are effects of small and large native mammals and feral species, when data are available. Emphasis 

is placed on the evolutionary history of grazing and productivity of the particular communities as determinants of response. After reviewing available studies for each community type, we compare changes 

in species composition with grazing among community types. Comparisons are also made between 

southwestern communities with a relatively short history of grazing and communities of the adjacent 

Great Plains with a long evolutionary history of grazing. Evidence for grazing as a factor in shifts from 

grasslands to shrublands is considered. An appendix outlines a new community classification system, 

which is followed in describing grazing impacts in prior sections.

The Author

Daniel G. Milchunas is Research Scientist, Forest, Range, and Watershed Stewardship Department, 

Colorado State University, Fort Collins.

Acknowledgments

Primary funding was provided by the USFS Rocky Mountain Research Station, with additional support 

from the Shortgrass Steppe Long-Term Ecological Research project. Curt Flather of the USDA Forest 

Service coordinated the overall project. Curt Flather, Pat Zwartjes, Charlie McDonald, and Deborah 

Finch of the Forest Service provided reviews of an early draft. Jane Bock and Rex Pieper gave constructive and useful review comments that improved the final version of the manuscript. I thank Lane Eskew 

for technical editing and Joyce Vandewater for help with graphics. Norman Ambos of the Forest Service 

provided unpublished data for Dutchwoman butte.

Cover photo: On the 300-square-mile Jornada Experimental Range near

Las Cruces, New Mexico, technicians Rob Dunlap (left) and

John Smith round up cattle. Photo by Scott Bauer.
Executive Summary

Two primary factors determine a plant community’s relative response to livestock grazing: 

the evolutionary history of large-herbivore grazing and the plant community’s aboveground 

primary productivity. Bison were generally not part of the evolutionary history in the southwest, 

except for the edge of their range (less intense grazing) in far eastern New Mexico where a 

smaller bodied (more selective) grazer exerted possibly different pressure than that occurring 

in the central Great Plains. Because aridity (low productivity) and grazing are convergent 

selection pressures, communities of the southwestern United States may be intermediate in 

their response to livestock grazing. However, grazing by domestic livestock is controversial in 

the southwest, and grazing has been considered a reason for shrub encroachment in many 

communities.

This report assesses the effects of grazing on 25 plant communities of the southwestern 

United States, in terms of plant species composition, aboveground primary productivity, and 

root and soil attributes. All possible studies were considered in an attempt to 1) determine the 

state of knowledge for each community and suggest where research is lacking, 2) assess the 

relative responses among communities, 3) present an unbiased, literature-based overview 

of grazing and the shrub encroachment issue, and 4) broadly view grazing responses in the 

southwest relative to responses observed in adjacent semiarid to subhumid Great Plains 

communities that have a long evolutionary history of grazing by bison. Responses of individual communities are summarized at the end of each section. Community response to grazing 

increased, following theoretical predictions, as such: unproductive/long history < unproductive/short history < productive/short history < productive/long history. However, variables such 

as precipitation, grazing intensity, and years of protection from grazing could much better 

predict community response on the Great Plains than the southwestern United States. This 

degree of unpredictability for the southwest, together with a large difference between nevergrazed geologic refuges and human-made exclosure comparisons with grazed treatments in 

the southwest, raises questions as to whether some livestock effects may have been historic, 

or whether southwestern communities require temporally rare and unpredictable environmental events for change to occur. Riparian communities were particularly responsive to grazing. 

The response of a plant species common to both Great Plains and some southwestern communities was not the same in the two locations. In general, southwestern communities were 

more sensitive to grazing than the adjacent shortgrass steppe, but were similar in response 

to other semiarid/arid and subhumid communities with a short evolutionary history of grazing. 

Further, the temporal variability in plant community species composition due to weather appears to be much greater in southwestern than in Great Plains communities.

The majority of evidence from experimental literature indicates that shrub encroachment 

occurs into ungrazed sites as well as grazed sites. There are four possibilities that could still 

invoke livestock grazing as a factor in this conversion of communities: 1) previous very heavy 

grazing during the 1800s resulted in an alternate stable state, 2) grazing interacts with fire, 

3) there are highly erodible areas where long-term studies have not been in place, where 

current grazing could initiate alternate stable states, and 4) grazing interacts with some other 

factor such as climate or rodents. The evidence for these are reviewed and found to be either 

weak or experimentally untestable under current conditions but having factors that weaken a 

supporting argument. There is an increasing amount of evidence implicating changes in the 

seasonal proportions of precipitation as a primary cause of community change in the southwest. However, no one factor alone appears to adequately explain the complex spatial pattern 

of community change through parts of this region.

Intense non-selective grazing is an exogenous disturbance in most southwestern communities. From a conservation standpoint, there are places where grazing by livestock should 

be encouraged in the absence of the native grazer, and other places where livestock grazing 

should be discouraged or at least cautiously managed because similar grazers were not present historically. From a production standpoint, light to moderate grazing can be sustainable in 

the southwest in many situations. For some communities there are little data to base conclusions, and for other communities there are contradictory findings. However, the fate of private 

lands will often closely be coupled to use of public lands. Alternatives to ranching such as 

development and subdivision for ranchettes or agriculture may potentially be a greater threat 

to the integrity of southwestern ecosystems than ranching of native communities.
Contents

Executive Summary ............................................................................................................ i

Introduction......................................................................................................................... 1

History of Grazing by Native and Domestic Large Herbivores ...................................... 3

The Importance of Evolutionary History and Environmental Conditions in 

Grazing Responses.................................................................................................. 3

Evolutionary History of Grazing in Southwestern U.S. Plant Communities ................. 4

History of Grazing by Domestic, Feral, and Exotic Animals in the 

Southwest United States.......................................................................................... 7

Grazing Effects on Plant Communities ............................................................................ 9

Alpine Tundra............................................................................................................... 9

Boreal Forest ............................................................................................................. 10

Mixed Conifer Forest ................................................................................................. 11

Ponderosa Pine and Madrean Pine Forests.............................................................. 12

Species Composition ........................................................................................ 13

Aspen Communities .......................................................................................... 16

Primary Production............................................................................................ 17

Root Responses................................................................................................ 17

Soils .................................................................................................................. 18

Fire-Grazing Interactions................................................................................... 18

Grazing Management........................................................................................ 19

Pinyon-Juniper........................................................................................................... 20

Species Composition ........................................................................................ 20

Primary Production............................................................................................ 25

Soils .................................................................................................................. 26

Fire-Grazing Interactions................................................................................... 27

Grazing Management........................................................................................ 27

Madrean Woodland.................................................................................................... 28

Great Basin Juniper-Sagebrush ................................................................................ 31

Interior Chaparral....................................................................................................... 32

Subalpine Grassland ................................................................................................. 35

Montane Meadow ...................................................................................................... 37

Species Composition ........................................................................................ 37

Primary Production............................................................................................ 39

Rooting Characteristics and Soil Properties...................................................... 40

Temperate Grassland........................................................................................ 40

Species Composition ........................................................................................ 40

Primary Production, Root, and Soil Responses ................................................ 44

Grazing Management........................................................................................ 44

Mesquite Savanna ..................................................................................................... 45

Species Composition ........................................................................................ 45

Primary Production............................................................................................ 60

Root and Soil Responses.................................................................................. 61

Grazing Management........................................................................................ 61

Mesquite Savanna Conversion to Chihuahuan Desert: Shrub Invasion, 

Livestock Grazing, Fire, Small Mammals, and Climate.......................................... 63

Chihuahuan Desert.................................................................................................... 68

Species Composition ........................................................................................ 68

Management ..................................................................................................... 72

ii
Great Basin Sagebrush Shrubsteppe ........................................................................ 72

Species Composition ........................................................................................ 73

Primary Production............................................................................................ 74

Management ..................................................................................................... 74

Shinnery Oak/Sand Sage .......................................................................................... 75

Great Basin Temperate Desert .................................................................................. 75

Species Composition ........................................................................................ 75

Management ..................................................................................................... 76

Sonoran/Mohavian Deserts ....................................................................................... 77

Riparian and Wetland Communities .......................................................................... 80

Narrowleaf Cottonwood..................................................................................... 80

Broadleaf Cottonwood Forest ........................................................................... 81

Riparian Scrub .................................................................................................. 84

Other Riparian................................................................................................... 84

Dry Washes....................................................................................................... 85

Wetlands ........................................................................................................... 85

Grazing Management........................................................................................ 86

Relative Effects of Grazing Among Southwestern Community Types, and 

Comparison of SW with Great Plains Community Responses: 

Does Evolutionary Grazing History Matter?............................................................... 88

Conclusions...................................................................................................................... 96

References ...................................................................................................................... 100

Appendix 1. Synonyms for the plant community classification system of Moir (2000).... 117

Appendix 2. List of scientific and common names of plants used in text. ....................... 119

iii
USDA Forest Service RMRS-GTR-169. 2006.

The objective of this review is to examine the responses 

to grazing in plant communities of the southwestern 

United States. More specifically, the review focuses on 25 plant community types in Arizona and New 

Mexico and the manner in which grazing affects species composition, primary productivity, soil, and other 

belowground attributes. Focusing on each particular 

plant community, rather than a region, allows for a better assessment of the specific states of knowledge and 

research needs. This review also discusses how the evolutionary history of plant communities interacted with 

native ungulate grazing to shape responses to grazing 

by livestock. Grasslands of the Great Plains have a long 

evolutionary history of grazing by bison. The impact of 

livestock grazing on the Great Plains differs from the 

impact of livestock grazing on other lands where native 

ungulates did not graze as much in the past, such as 

southwestern lands (Bock and Bock 1993; Milchunas 

and others 1998; Bock and Bock 2000), the Great Basin, 

and the far west (Mack and Thompson 1982; Milchunas 

and others 1988; Lauenroth and others 1994) that did 

not have similar levels of grazing by large generalist 

herbivores. Evolutionary history of grazing and environmental moisture or primary productivity interact 

in determining species adaptations for tolerance or 

avoidance of herbivory (Milchunas and others 1988) 

and in community responses to grazing (Milchunas 

and Lauenroth 1993). While quantitative global-scale 

analyses have tested the hypotheses concerning the history-productivity interaction (Milchunas and Lauenroth 

1993), no specific attempt has been made to compare 

southwestern with adjacent Great Plains plant communities with respect to relative responses to grazing.

The literature concerning southwestern plant communities often indicate that very large changes in 

compositions and distributions have taken place within the lifetimes of individuals and coincidental with 

the period of large-scale introduction of livestock. 

Controversy over the role of livestock grazing in those 

changes is high. Some reviews and paper discussion 

sections in even the scientific literature appear to pick 

and choose studies or evidence supporting a particular 

viewpoint. Therefore, an additional objective of this review was to do an extensive coverage of the studies for 

each community type whereby all available information 

is presented. The long literature cited section for this 

paper attests to the volume of studies associated with 

this endeavor, but also to the possibility that important 

pieces could have been missed.

Several other procedures/conventions were taken in 

compiling the effects section to help the reader as well 

as this author in interpretation. First, observation versus 

experimentation is differentiated. Observation-based 

information is a large proportion of the knowledge 

available for some community types, and less for others. 

In the case of a plant community type for which a large 

number of experimental studies were available, an even 

lesser amount of observational evidence may have been 

presented. Sections on community types are generally 

organized by observational works first and experimental 

second, and presented in a way whereby it is apparent as 

to which is which. This is not to demean observational 

information, but I know that I have often been surprised 

when my experimentation has not supported my perceptions based on observations of a system that I worked in 

for years.

A second convention was to sometimes comment on 

particularly good experimental designs and point out 

interpretive peculiarities of others. In general, however, 

grazing studies require large pastures and are often inadequately or improperly replicated, leading to an inability 

for proper characterization and inference (Hurlbert 1984; 

Brown and Waller 1986). Pseudoreplication and/or 

complete lack of statistical testing is common especially 

among early research, yet this research is particularly 

important for our understanding grazing effects when 

conducted through long-term weather cycles and when 

it provides a historical record of previous condition of 

Introduction
USDA Forest Service RMRS-GTR-169. 2006.

plant communities. Animals do not graze uniformly 

across the landscape topography even within areas of a 

square kilometer in the same general plant community 

type, and the effects of grazing can differ substantially 

over this heterogeneity (Milchunas and others 1989) as 

well as over large regional differences within a plant 

community type (Sims and others 1978; Milchunas and 

Lauenroth 1993). Therefore, reaching consensus in results from multiple studies, or differentiating factors that 

may lead to divergent results, are particularly important 

potential outcomes of extensive reviews. Alternatively, 

a lack of studies for particular communities points to 

the need for additional study.

Third, an attempt was made to mention the type of 

sampling procedure used by the author. Canopy cover, 

basal cover, density, and biomass each have advantages 

and disadvantages depending on the particular objectives of the research, but some can confound current-year 

removal by the herbivore with longer-term plant population dynamic changes associated with grazing when 

the objective here is the latter. When contrasted with 

an ungrazed treatment for example, canopy cover and 

biomass not sampled from within temporary cages on 

grazed areas measures both current-year removal by the 

animal plus any of the longer-term effects of the grazing 

on mortality and growth capacity, whereas basal cover 

does not include effects of current removal. Since grazing is by definition removing the canopy, studies where 

the objective is long-term effects on plant populations 

would not have wanted to include the confounding effect of utilization inherent in canopy cover, but studies 

where the objective was the effects on other native consumers would have wanted to include the short-term 

effects on structure and species availabilities included 

in current-year removal by the livestock. Density has a 

size dependent bias, but is also an unbiased estimator 

of long-term population dynamics. In terms of aboveground net primary production, use of end-of-season 

peak standing crop from annually-moved temporary 

cages was considered a valid estimate (see Milchunas 

and Lauenroth 1992; McNaughton and others 1996 for 

advantages and disadvantages), but several papers were 

omitted from production portions of this review when 

uncaged grazed biomass estimates were called production. Moving cages multiple times throughout a growing 

season, and clipping caged and uncaged biomass, is a 

method that captures the effects of current-year grazing 

on compensatory regrowth but can also result in biases 

that cause overestimations, and few studies have used 

this method (McNaughton and others 1996).

The effects on plant communities of herbivory by 

domestic, feral or exotic, and native animals are considered in this paper. However, the majority of studies 

available concern domestic livestock. While there may 

be many reasons for the prevalence of livestock studies, 

it should not lessen our consideration of the potential 

importance of the other herbivores. In an introduction 

to a proceedings on research on exotics in National 

Parks, Dennis (1980) notes that 14 of 24 papers concerned ungulates. Feral or exotic species often represent 

a novel perturbation to a system, which can have large 

impacts. The works of Brown and colleagues on rodents 

in the southwest are prominent examples of how native 

small mammals can act as keystone species in their effects on a system (Brown and others 1979; Brown and 

Heske 1990). The classic textbook example of predator control, herbivore population eruption, damage to 

plant communities, and then herbivore population crash 

comes from the Kaibab deer herd in the southwest 

(Rasmussen 1941; Leopold 1943; Mitchell and Freeman 

1993; Young 2002). Deer populations in many areas of 

North America are at higher densities than they have 

been in the past several hundred years, because of their 

opportunistic nature and use of early seral stages created by humans (Rooney 2001). However, there appears 

to be more studies on the effects of livestock grazing on 

wildlife populations than on the effects of wildlife on 

vegetation. Information on these other groups of herbivores/granivores are included when available.

The organization of this paper is to first examine the 

history of grazing by native and domestic herbivores in 

the southwest at a broad geographic scale. Additional 

material specific to particular communities, when available, is included in the section on grazing effects on that 

community. The summary of available studies on the 

effects of grazing on each of the communities follows 

a community classification system recently developed 

by the USFS (Moir 2000, Appendix 1 provides outline 

and synonyms from other classification systems). Some 

departure from the numerical order of communities in 

the Forest Service classification system is made for the 

purpose of conceptual flow. In one case, two community 

types are combined because of difficulty in placement of 

the available studies into separate categories, the small 

amount of information available, and similarity between 

the two communities. Each community type begins with 

a general description or definition, followed by a discussion of grazing effects on species composition, primary 

production, and, where available, root responses, soils 

(including erosion), fire-grazing interactions, and grazing management. A comparison among community 

types, and between southwestern and Great Plains communities, is made with respect to sensitivity to grazing. 

Conclusions are based on the within and between community responses. Appendix 2 provides a list of scientific 

names of plant species referenced in the text.
USDA Forest Service RMRS-GTR-169. 2006. 

The Importance of 

Evolutionary History and 

Environmental Conditions 

in Grazing Responses________

A theoretical, generalized model has been proposed 

to explain the very different sensitivities of plant communities to grazing by large herbivores (Milchunas 

and others 1988) and has subsequently been validated 

by quantitative, meta-analyses of grazing studies from 

around the world (Milchunas and Lauenroth 1993) and 

various other experimental and synthesis efforts (Painter 

and others 1993; Proulx and Mazumder 1998; Huisman 

and Olff 1998; Kotanen and Bergelson 2000; Loreti 

and others 2001; Osem and others 2002). Basically, the 

theoretical model views the primary determinants of 

plant community response to grazing as an interaction 

between evolutionary history of grazing and site productivity. Various avoidance and tolerance mechanisms 

develop in plant populations that are exposed to high 

levels of herbivory (prostrate low-growing morphologies, investment in crown and root storage unavailable 

to grazers, and so forth). However, adaptations for 

successfully competing in water limited semiarid systems are convergent (act in the same manner) to those 

inferring tolerance and/or avoidance to grazing, while 

adaptations for successfully competing in productive 

environments are divergent (act in the opposite manner) 

to those inferring tolerance and/or avoidance to grazing. Relatively greater investment in crown and root 

structures are positively associated with the ability to 

compete for limiting soil-water resources in semiarid 

areas, and is also an avoidance of grazers because a 

greater proportion of the plant is unavailable to large 

ungulates and supplies necessary resources for regrowth 

after a defoliation event. In contrast, light in a space-restricted canopy can become relatively more limiting in 

productive environments where belowground resources 

are relatively not as limiting. Adaptations for competition in the canopy (tall growth forms, relatively more 

proportional aboveground investments) make a plant 

more susceptible to greater injury and lower potential 

regrowth reserves following defoliation. Further, large 

herbivores consume the plant canopy, thereby having a 

direct influence on mechanisms of plant-plant competition in productive communities, while only indirectly 

influencing belowground competition in semiarid environments. The role of small herbivores and granivores 

can also be large in some systems (Brown and Heske 

1990), but less experimental and theoretical work 

has been done on their potential effects on plant 

communities.

The interaction between evolutionary history of 

grazing and precipitation/primary production result in 

differences in the response to grazing that can be described by four extreme cases along the two continuous 

axes (Fig. 1). At one extreme, rapid switching in species 

adapted to competing in the canopy and those adapted 

to grazing tolerance and avoidance in productive communities with a long evolutionary history of intense 

grazing result in large changes in species composition 

and diversity when ungrazed-grazed comparisons are 

made. At the opposite extreme, communities of low 

production and short history of grazing are not as responsive to grazing as the former in terms of changes 

in species composition and diversity, but can be more 

susceptible to invasions by exotics than systems with a 

long history of grazing. Adaptations to semiaridity temper the drastic changes seen in productive communities 

with a short history of grazing, but are greater than responses in semiarid communities with long history of 

grazing because grazing pressure was not similar to 

aridity pressures when acting on the evolution of adaptations. Plant basal cover is more likely to increase in 

semiarid communities with a long grazing history and 

History of Grazing by Native and 

Domestic Large Herbivores
 USDA Forest Service RMRS-GTR-169. 2006.

decrease in semiarid communities with a short grazing 

history because of differences in grazing tolerance and 

avoidance mechanisms (Milchunas et al. 1988, 1989). 

Invasion of exotic or native opportunistic species is 

more likely where unoccupied space increases rather 

than decreases (Milchunas and others 1992).

The importance of evolutionary history of grazing 

and community productivity in determining responses 

to grazing in plant community species composition and 

net primary production was quantitatively assessed 

over 236 ungrazed-grazed comparisons from around 

the world (Milchunas and Lauenroth 1993). At this 

global scale, sensitivity to grazing was more explained 

by ecosystem and environmental variables (evolutionary history of grazing, productivity, and so forth) than 

by current grazing-related variables (intensity, years 

of protection, and so forth). Current grazing practices 

undoubtedly become important within communities 

or regions. Because of the importance of evolutionary 

history of grazing and plant community productivity, 

I first review the evolutionary history of grazing in 

southwestern communities, and in the 

Plant Community response section I 

include information on primary productivity when available.

The Milchunas and others (1988) 

theoretical model is entirely based on 

plant-plant interactions as historically 

and currently influenced by large generalist herbivores over a semiarid to 

subhumid gradient. This may be a partial weakness of the model when arid 

or highly erodible systems are considered. Soil processes mediated by large 

herbivores have also been suggested in 

a model describing responses to grazing 

in arid-to-semiarid regions in the southwestern United States (Schlesinger and 

others 1990; Whitford 2002) and elsewhere (van de Koppel and others 1997). 

Effects of intensive grazing on albedo 

feeding back on local climate change 

has also been suggested as a model for 

community desertification and drought 

cycles in the Sahel (Charney 1975). 

The El Nino Southern Oscillation has 

also been implicated as interacting with 

herbivores in affecting the dynamics of 

plant communities (Holmgren and others 2001). There is evidence supporting 

all of these models, and it is possible 

that all or some operate simultaneously at any particular 

location. This possibility in reference to the southwest 

will be further pursued in the conclusion section of this 

paper, after reviewing literature regarding plant community responses.

Evolutionary History of 

Grazing in Southwestern U.S. 

Plant Communities__________

Bison were the primary large, generalist herbivore 

that was important in structuring plant communities 

of the Great Plains. Although other herbivores such 

as deer and elk also occurred throughout much of the 

region, in the plains they generally did not reach the 

large herd sizes associated with bison. Further, the body 

size and therefore diet of bison and cattle are generally 

similar, although bison are a little more of a generalist, 

non-selective “lawn mower” type grazer than cattle in 

Figure 1—Theoretical model of plant community species diversity response 

to grazing intensity over gradients of precipitation and evolutionary history 

of grazing. The semiarid to subhumid precipitation gradient (approximately 

>250 to <1000 mm/yr average) is considered interchangeable with low to 

high aboveground primary production. The four graphs represent the four 

extreme cases over the environmental gradients, and intermediate gradations occur from one curve to another. Grazing is for large generalist 

herbivores such as cattle, sheep, bison, and so forth. The effects of, or 

interactions with, small specialists herbivores are not included in the model. 

Adapted from Milchunas and others (1988).
USDA Forest Service RMRS-GTR-169. 2006. 

shortgrass steppe communities (Lauenroth and 

Milchunas 1991). There are numerous accounts illustrating the intense grazing pressure that was applied 

by a large herd of bison. For example, Hornaday (1889, 

quoted in Larson [1940]) wrote that bison “at times so 

completely consumed the herbage of the plains that 

detachments of the United States Army found it difficult to find sufficient grass for their mules and horses.” 

Col. Dodge was reported to have estimated one herd at 

over four million individuals (Roe 1951). Seton (1953) 

estimated numbers at 40 million in the “open plains” 

and 30 million in the “prairies”. Lott (2002) estimated 

an average population size of 24 to 27 million bison in 

North America, based on carrying capacity estimates 

and population time-lags in responding to weather cycle 

effects on resource availability. Because of the intensity 

of grazing by bison for 10,000 years and general similarities between the animals, herbivore adapted grasses 

in the Bouteloua (grama grass) Provence east of the 

Rocky Mountains in North America were not adversely affected by the introduction of domestic livestock 

(Mack and Thompson 1982; Milchunas and others 

1988). However, the steppe of the Agropyron (wheatgrass) Provence west of the Rockies, which lacked 

large herds of mammals throughout the Holocene and 

hence dominant caespitose plants not well adapted to 

herbivory, showed declines in herbaceous vegetation 

and increases in opportunistic species with the introduction of livestock. In these grasslands, the dominant 

C3 grasses and the prominent cryptogram layer were 

soon altered by domestic livestock and replaced largely 

by alien winter annuals. Further, in the Great Basin, a 

greater pressure on the shrub component in sagebrush 

shrubsteppe occurred during the Pleistocene by browsers such as camels and ground sloths (Martin 1970, see 

Great Basin sagebrush shrubsteppe section).

With respect to range of bison in the southwestern 

United States, elevation and latitudinal/longitudinal 

factors with the adjacent Great Plains grasslands were 

involved in distributional patterns. Large generalist 

herd animals such as bison generally do not utilize areas of steep topography (see evolutionary history part 

of Alpine Tundra section below). Since bison proliferated in the Great Plains, but existed for a relatively short 

period, or in isolated locations west of the Rockies, the 

potential for an influence of bison on southwestern communities could only come from the east and northeast, 

similar to the locations of the major biomes (Fig. 2). 

However, Gosz and Gosz (1996) delineate the boundary 

for Great Plains grasslands farther west than Lauenroth 

and Milchunas (1991) delineate the boundary for the 

shortgrass steppe (Fig. 2 vs. Fig. 3), because of the influence of blue grama in plant communities further west 

Figure 2—Distribution of four major biomes in the state of 

New Mexico. Adapted from Gosz and Gosz (1996).

Figure 3—Historic estimated distribution of bison in Arizona 

and New Mexico based on several sources and locations 

where archeological remains have been found (adapted 

from Truett 1996) with the boundary of the shortgrass 

steppe superimposed (as mapped in Lauenroth and 

Milchunas 1991).
USDA Forest Service RMRS-GTR-169. 2006.

than what may be considered shortgrass steppe, but also 

because the Great Plains geomorphic province extends 

farther west than the shortgrass steppe biotic community (Sims and others 1978). Although grasslands in 

southeastern Arizona have been designated as “Plains 

and Great Basin Grasslands” and have some species in 

common with the Great Plains, similarities of the flora 

are greater to the southeast (Mexico and southwestern 

Texas) than to the northeast (central U.S.) (McLaughlin 

and others 2001). Madrean and Apachian elements are 

the largest part of the native flora.

Interestingly, the distribution of bison into the southwest has been mapped intermediately between boundary 

of the shortgrass steppe, which is more to the east (Fig. 

3), and the boundary of the Great Plains biome, which is 

more to the west (Fig. 2). Clearly, biomes, communities, 

or animal distributions usually do not have sharp welldefined boundaries, but they gradually merge in broad 

ecotone transitional zones. In reviewing the evidence 

for distributions of both bison and elk, Truett (1996) 

used the Bailey (1931) range map and early explorer accounts that indicated only occasional occurrence west 

of the Pecos River. Bison remains have been found at 

many sites in other areas of New Mexico and Arizona 

(Parmenter and Van Devender 1995; Truett 1996), but 

numbers were generally thought to be very low. Effinger 

and Lucas (1990) also map most bison remains in 

eastern New Mexico. Bailey concluded that bison never 

occupied the Rio Grande Valley in large numbers, because of the barrier provided by arid valleys and due 

to hunting by the Navajo and Pueblo Indians that inhabited the valley. There are no explorer accounts of 

bison in the Rio Grande Valley, but there were numerous reports of bison along the Pecos River, named Rio 

de las Vacas for the abundance of bison. Bailey quotes 

Coronado from the mid 1500s as saying “the plains are 

as full of crooked-back oxen as the Mountain Serene in 

Spain is of sheep,” and Espejo in reference to the Pecos 

Valley in 1584 with “all the way passing through great 

herds of buffaloes.” Truett also maps the historic distribution of elk in New Mexico and Arizona (based on a 

number of early references) as very limited and in only 

some mountain ranges, about only half of today’s range. 

Truett investigated a number of potential reasons for 

the scarcity of these large herbivores in the southwest 

and concluded that the most important causes were the 

restricted distribution of perennial water together with 

hunting by late-prehistoric peoples. Archeological site 

excavations suggest a relative abundance of the smaller, 

more dietary specialist deer and pronghorn antelope 

in the region. These accounts cover the more recent, 

11,000-year history during the Holocene. Bison species 

originally immigrated to North America about 150,000 

years ago in the late Pleistocene and occupied not only 

the Great Plains, but much of the southwest as well (Van 

Devender 1995). About two-thirds of the large mammals of North America went extinct 11,000 years ago, 

and desert grasslands after that time have had the poorest fauna of large herbivores ever in the last 20 million 

years of its history. Lott (2002) also mapped bison distribution in 1500 as extending only into the eastern strip 

of New Mexico, but he suggested dense numbers only in 

the northeastern corner of New Mexico with southeastern New Mexico being only “sparsely occupied during 

periods of typical climate.” Carrera and Ballard (2003) 

reviewed the evidence of elk distributions, and conclude that Arizona, New Mexico, and northern Texas 

were the southernmost edge of the distribution in the 

late Holocene and that elk never occurred naturally in 

Mexico.

It can be concluded that only a small eastern strip of 

New Mexico was subjected to the long-evolutionary history of grazing that was a strong force in shaping recent 

species adaptations and plant communities in the adjacent Great Plains region. More specialist herbivores were 

present, and specialist herbivores alone rarely consume 

a large percentage of plant production in communities, 

except under occasional population outbreaks. Both the 

type and intensity of herbivory differed between the 

adjacent shortgrass steppe and much of the southwestern United States. Further, the Indians recognized two 

types of bison, a smaller southern type and a northern 

larger bodied animal (Bailey 1931). Bailey places the 

larger bison in the Arkansas River area and northward, 

and the smaller animal as ranging through Texas and 

eastern New Mexico. Allen (1874 cited in Bailey 1931) 

also reported that the Indians and Mexicans recognized 

a ‘mountain buffalo’ and a ‘plains buffalo’, but Allen 

did not consider two separate species due to conflicting accounts of characteristics of the mountain buffalo. 

Bailey’s talks with Indians indicate that a cow of the 

northern type was as large as a bull of the southern 

type. A body size that different may also suggest a more 

selective grazer in the south than in the north, possibly representing a somewhat different pressure by bison 

grazing of southern compared to northern shortgrass 

steppe communities. Therefore, not only was it a small 

eastern strip of New Mexico that was subjected to bison 

grazing, but the type of grazing pressure received at this 

limited border of occupation could well have been by 

a relatively more selective grazer than current-day cattle. Responses of even these communities to the novel 

grazing pressure of domestic livestock may, therefore, be 

greater than in the Great Plains where past and current 
USDA Forest Service RMRS-GTR-169. 2006. 

grazing pressure may have been more similar. Further, 

plant populations near the edge of their range may also 

be more susceptible to change.

History of Grazing by 

Domestic, Feral, and Exotic 

Animals in the Southwest 

United States_______________

There have been many detailed reviews of the history 

of livestock grazing in the southwest (Wagoner 1952; 

Schickedanz 1980; Bahre 1991; Elson 1992; Cartledge 

and Propper 1993; Scurlock 1995; Brugge and Gerow 

2000; Eastman and others 2000; Fowler 2000), and it 

is not the objective of this paper to repeat that here. It 

is, however, important in the context of this paper to 

examine the degree and extent of overgrazing in the 

late 1800s to early 1900s, and to place this in a temporal context to evaluate in the discussion of grazing 

impacts and interactions with other stressors that are 

addressed later in this paper. This is because of the coincidence in the period of intense, unregulated grazing 

with the start of profound changes in plant community 

species compositions and distributions. We therefore 

present only a brief outline of the history of grazing 

by domestic livestock to illustrate the time frame and 

extent of the grazing. “Overgrazing” is a qualitative 

term that is difficult to define scientifically, because it 

requires knowledge of a particular threshold. The term 

is used here only in the context of the historic period of 

unregulated, intense grazing of large, unfenced areas 

to emphasize the situation that is generally acknowledged by ranchers, range managers, and scientists as a 

period of overgrazing.

Humans occupied and modified southwestern ecosystems long before domestic livestock were introduced. 

Evidence of humans dates back earlier than 10,000 

B.C., although permanent habitation did not occur until 

300 B.C. (Elson 1992). The Anasazi were the dominant 

Indian communities around 1100 to 1300 and possibly 

had large impacts on plant communities through the 

gathering of fuel wood and food and cultivation activities (Cartledge and Propper 1993). There are several 

lines of evidence that indicate significant deforestation 

due to the activities of these people.

The first cattle were brought into Mexico in 1521 

by the Spanish (Wagoner 1952). They were of the 

Andalusian breed originally from western Spain. 

Coming from a similar environment, they were 

well-adapted to the semiarid southwest. Cattle were first 

brought into what would become southwestern United 

States by Coronado in 1540, and many were abandoned. 

Thousands of their descendants were discovered later 

by frontiersmen. There were 75,000 head by 1586. Wild 

asses or burros were similarly introduced (Douglas and 

Leslie 1996). The numbers of cattle, sheep, and goats 

owned by Spanish and Indians, and that were wild, increased through the 1700s (Wagoner 1952). By 1757 

there were 7,000 horses, 16,000 cattle, and 112,000 

sheep in New Mexico, primarily in the Rio Grande 

Valley (Elson 1992). Some signs of overgrazing began 

to appear locally by 1846: in the plains south and east 

of Santa Fe, the Rio Grande Valley, and the Rio Puerco 

watershed. Bahre’s (1991) review of the historical status of Arizona borderlands concludes that overgrazing 

was insignificant through the mid-1800s and much of 

the descriptions of southeastern Arizona were of pristine conditions. While the history of domestic livestock 

grazing is longest in the southwest than anywhere else 

in the western United States, areas heavily impacted by 

grazing prior to Anglo entry in the mid-1800s were usually near settlements along major drainages or on some 

Indian occupied areas (Bahre 1991).

The end of the Civil War, the establishment of railways, and removal of Indian threats to settlement all 

combined to produce a boom in the cattle industry in 

the 1880s. Based on accounts of locals in southeastern 

Arizona at the time, Bahre (1991) writes that “every 

running stream and permanent spring had been claimed 

and adjacent ranges stocked” and “the entire region must 

have looked like one big cattle ranch.” A major drought 

in 1891 to 1893 killed large proportions of the livestock 

and many areas experienced major topsoil erosion after 

loss of vegetative cover. Overstocking continued after 

the drought, but animal numbers had peaked in 1891. 

Wagoner (1952 appendix I) lists numbers of cattle for 

Pima county, Arizona, as 11,741 in 1880, 121,377 in 

1891, and 49,599 in 1893. Cable (1975 citing Croxen 

1926) describes the stocking of Tonto National Forest 

as peaking in about 1900 at 15 to 20 times the number 

present in 1926. The decline in mining and introduction 

of railroads led also to abandonment of burros (Douglas 

and Leslie 1996).

The 1880s through early 1900s represented a period 

of grazing at extreme intensities greatly over carrying 

capacity of the plant communities. Bahre (1991) considered Forest Service land to have been overgrazed 

through 1906, and Bureau of Land Management land 

continued to be heavily overgrazed until enactment of 

the Taylor Grazing Act in 1934. Homesteading in the 

1910 to ‘20s resulted in decreasing amounts of open, 
 USDA Forest Service RMRS-GTR-169. 2006.

unfenced range, and Forest Service lands were generally fenced by 1930.

Changes in fire regimes also occurred along with settlement and fencing and management of public lands. 

Fire suppression accentuated lower frequencies of fires 

due to lowering of fuel loads by livestock removal of the 

forage. Frequent fires started by Indians were also no 

longer a part of the environment. Grazing, but also grazing as interacting with fire, became a primary suspect in 

the very large changes in plant community species composition and in the distributions of plant communities 

that were also occurring in the late 1880s and continue 

to present. Bahre (1991) provides an overview review of 

fire incidence in the southwest, and more recent reports 

and analyses relevant to grazing are examined below for 

each particular community type. Several free-ranging 

exotic animals are also found in New Mexico, and include Barbary sheep, gemsbok, Persian wild goats, and 

Siberian ibex (Douglas and Leslie 1996).
USDA Forest Service RMRS-GTR-169. 2006.

There are several classification systems that cover 

southwest U.S. plant communities. In this report, we 

follow a recent one developed by W. H. Moir (2000) 

for the Forest Service in New Mexico and Arizona. 

Synonyms for each community that are associated with 

other classification systems are listed in Appendix 1. A 

list of scientific names for the common names of plants 

used in the text appears in Appendix 2.

Alpine Tundra______________

Alpine plant communities are subject to many environmental stresses such as high wind and UV-radiation, 

cold temperatures, and short growing seasons. Plants in 

this environment, therefore, grow very slowly and have 

low annual primary productivity. These characteristics 

make alpine systems particularly susceptible to aboveground disturbances such as grazing. Alternatively, 

alpine plant communities are generally considered to 

have high root to shoot ratios (Paulsen 1975; Billings 

1979; Thilenius 1979), which is a characteristic of systems that imparts avoidance and tolerance to grazing 

(Milchunas and others 1988; Milchunas and Noy-Meir 

2002). Reserves in crowns and roots are inaccessible 

to large aboveground herbivores and serve as a source 

of carbohydrates for regrowth following defoliation. 

Thilenius (1975) and Paulsen (1975) report that 80 to 

95 percent of alpine plant biomass in the Rockies is belowground and proportions are greater for forbs than for 

grasses, and Billings (1979) gives a range in root:shoot 

ratios from 1 to 6. In comparison, however, root:shoot 

ratios for North American grasslands range from 2 to 

24, from 2 to 4 for a desert grassland site in New Mexico 

(Sims and others 1978), and 10 to 15 for the highly grazing tolerant shortgrass steppe in Colorado (Sims and 

others 1978; Milchunas and Lauenroth 2001). Tundra 

plants growing under low temperatures have high carbohydrate contents, which may account for the high 

degree of tolerance to defoliation of some graminoids 

(Tieszen and Archer 1979). Thilenius (1975) speculated that competition among alpine plants is less intense 

than for plants in lower elevations, and that they respond 

more to environmental conditions than to biotic influences. If true, then the effects of grazing would need 

to manifest through direct mortality or modification of 

the environment rather than by altering competitive outcomes. However, Griggs’ (1956) observations suggest 

competition is important in harsh alpine communities, 

and Fowler (1986) reviewed evidence that supports the 

importance of competition in harsh arid environments.

The evolutionary history of grazing in alpine communities is not well understood but may have had only 

a minor role in structuring these systems (Tieszen and 

Archer 1979). Basic similarities of alpine communities in areas where large herbivores were and were not 

present indicates convergence (similarly structured communities) regardless of herbivore presence (Tieszen and 

Archer 1979), and the number and density of herbivores 

in alpine systems is generally much lower than lower-elevation grasslands in the same region (Hoffman 1974). 

Other reports in the literature appear contradictory. 

Tieszen and Archer (1979) cite studies in New Zealand 

and Ethiopia (Mark 1965; Klotzli 1977) as showing tall 

tussock communities changing to short-tussock communities when defoliated or grazed, suggesting that 

these systems that evolved in the absence of large herbivores display unusually dramatic responses. However, 

the responses may be due to the unusual tall structure of 

these particular communities. Bock and others (1995) 

specifically attempted to test the hypothesis that responses to defoliation would be different between alpine 

communities in the Republic of Georgia with a 2000 yr 

history of livestock use and a Rocky Mountain National 

Park, Colorado, alpine community lacking a history of 

settlement and livestock grazing. These authors found 

defoliation decreased productivity in both systems, and 

they found an even greater depression in the Georgian 

Grazing Effects on Plant Communities
10 USDA Forest Service RMRS-GTR-169. 2006.

plants. The clipping studies did not support directional 

selection by a long history of grazing in this system, 

but plant family patterns and morphology may have resulted. Grazing can rapidly change species composition 

and morphology of communities, but 2000 years may 

be insufficient for individual or population level changes. However, if prior, longer-term selection had been 

acting, a genetic pool containing characteristics relating to tolerance or avoidance would have been present 

and influenced community dynamics when domestic 

animals were introduced. Bock and others (1995) did 

observe differences in community species composition 

and plant height structure that indicate selection at this 

level.

There are not large amounts of alpine vegetation in 

Arizona and New Mexico. Two plant associations with 

140 plant species have been described for New Mexico, 

and three associations with 80 species in Arizona 

(Chambers and Holthausen 2000). For alpine areas elsewhere and in general, Thilenius (1975) considered that 

no more than 200 to 300 species usually occupied the 

alpine zone of a particular mountain range. Although 

the alpine areas in Arizona and New Mexico were used 

for sheep grazing in early settlement years, they are no 

longer intensively grazed. Sheep numbers in the United 

States were highest from the 1880s through the 1940s 

and have declined since (Joyce 1989). Cattle are poorly 

adapted to this environment (heart damage - brisket disease) but can be acclimated (Thilenius 1979).

No studies of the effects of grazing on southwestern 

alpine plant communities could be located. Thilenius 

(1975, 1979) and Tieszen and Archer (1979) reviewed 

research concerning grazing of alpine communities in 

other locations. Since grazing study sites are potentially very different from those in the southwest, specific 

changes in structure or function of alpine communities 

cannot be considered here. Some very general conclusions reached by these authors, that may or may not 

be applicable to southwestern alpine communities, include: 1) Some graminoids show good regrowth after 

even chronic grazing pressure, whereas deciduous and 

evergreen shrubs can be damaged more severely. 2) The 

regrowth of graminoids results in reduced root growth 

which can take more than two years to replace. 3) There 

is the potential for changes in species composition, but 

the small amount of work in this area makes grazing 

management of these systems largely empirically derived. 4) A reasonable average production for grazable 

alpine vegetation is ~90 to 100 g/m2 and wetter or drier 

sites should be considered unsuitable range because wet 

sites are susceptible to trampling damage and drier sites 

are susceptible to erosion. Costin (1967) also concluded 

that the low productivity is a management problem, 

and makes alpine areas unsuitable for efficient grazing use. Chapin (1980) commented on the potential for 

defoliation to deplete nutrients in plants growing in nutrient-poor soils such as those found in arctic and alpine 

habitats.

Particular problems of grazing management in these 

systems include proper animal distribution due to shelter from wind, topography, and snow-melt distributions, 

and trampling impacts in saturated soils below snowdrifts (Thilenius 1975, 1979). Light to moderate use is 

generally considered to be 20 to 30 percent of aboveground biomass, but actual grazing is often lighter at 7 to 

10 percent. The 20 to 30 percent utilization recommendation is the lightest of all plant-community types in 

New Mexico (Holechek and Pieper 1992). Past and most 

current grazing uses a sheep herder, although herderless 

grazing may also occur. Paulsen (1975) and Thilenius 

(1975, 1979) outlined USDA-Forest Service herding 

practices that would seldom show excessive use, improve range condition, and show reduced erosion. These 

included slow moving, loosely bunched distributions, 

watered once a day, one-time per season defoliations, 

and one-night beddings where salt is provided in movable containers. Grazing should be deferred until after 

the period of most active plant growth, and rest-rotation 

systems may also improve range condition. Slopes of 

greater than 40 percent are considered ungrazable due 

to the high potential for erosion problems to develop. 

Thilenius (1975, 1979) also stressed that condition 

standards for Rocky Mountain alpine communities are 

poorly understood, and current guidelines are of doubtful value. Recovery of misused alpine systems requires 

long time periods. While extensive proportions of alpine 

areas have been disturbed by grazing at some point in 

time, Johnson and Brown (1979) considered other types 

of disturbances to be currently more intensive.

Boreal Forest_______________

Boreal forest occupies approximately 500,000 ha in 

the southwest, with 80 percent in New Mexico (Chambers 

and Holthausen 2000). Precipitation ranges from 580 to 

880 mm/yr and is the second highest for all Arizona - 

New Mexico community types. At five sites in Arizona, 

understory productivity of grasses and forbs averaged a 

little over 8 g/m2 in conifer forest and 51 g/m2 in associated aspen stands (Reynolds 1969). This is extremely 

low herbaceous production. Reynolds (1962) considered this community type to be of little importance for 

range production, but of high value for water yields and 
USDA Forest Service RMRS-GTR-169. 2006. 11

recreation. Logging or thinning of overstory increases 

forage production (Ffolliott 1979), but the inverse relationship between overstory-understory production 

is not as evident in the spruce-fir type as in other coniferous forests in the region (Ffolliott and Gottfried 

1999). Understory in closed-canopy spruce-fir is mainly lichens, fungi, sedges, mosses, and liverworts, with 

grasses, forbs, and half-shrubs becoming more abundant with opening of the canopy by disturbances. In 

addition to low forage production, the growing season 

is short—often less than 75 days.

There is little actual grazing in the spruce-fir community as animals primarily utilize meadows within 

this type, and grazing capacity of the area is based on 

the meadow communities within. Deer and cattle use 

of associated aspen stands can also be greater than use 

of the conifer forest (Reynolds 1969). As with alpine 

communities, the boreal forest type is considered more 

suitable for sheep than cattle grazing (Reynolds 1962). 

Sheep can, however, be more damaging to vegetation 

than cattle when heavy grazing occurs because they 

are capable of grazing plants closer to the ground. It 

is unlikely that boreal forest types are or were heavily grazed by native large herbivores, although use for 

cover may be high. Plant-herbivore relationships are 

probably poorly developed, except for associated aspen 

stands where utilization can be heavy.

Chambers and Holthausen (2000) commented that 

there is very little literature quantifying the effects of 

grazing in montane communities in the southwest, and 

they list some important research needs. I was unable 

to locate any quantitative data comparing ungrazed and 

grazed plant communities for boreal forest types.

Grazing has been considered by some to be an indirect factor in changes in many southwestern community 

types through its effect on reducing fuel loads and altering natural fire regimes (see other community type 

sections below). This may not be an effect of grazing in 

the boreal forest type, since the natural fire return interval is very infrequent anyway. White and Vankat (1993) 

estimated the natural fire return interval for boreal forest was 70 to 250 years, which is relatively infrequent 

compared to lower elevation communities.

High value for recreational use would conflict with 

use as grazing land. Further, the spruce-fir type is estimated to require twice the area than pine-fir forest 

to support an AUM, and six-times the area for mountain grassland (Reynolds 1962). Aspen stands within 

this type provide much more forage, but heavy grazing of young aspen shoots by wild or domestic large 

herbivores can eliminate reproduction and result in old, 

relatively even-aged stands. Chambers and Holthausen 

(2000) suggested livestock grazing should be avoided 

in communities where animal dietary preference for regenerating trees is high.

Mixed Conifer Forest________

The two community types differentiated according to 

Appendix 1 are combined for the purpose of assessing 

grazing impacts because no data specific to Madrean 

montane forests were located. Data located is applicable 

to the upper montane forest, white fir/douglas-fir, pine/

douglas-fir, and Petran conifer forest type.

Annual precipitation ranges from 450 to 800 mm in 

this and the Madrean mixed conifer type (Ffolliott and 

Gottfried 1999). Grasses and forbs are more abundant 

than in adjoining higher elevation boreal forest, although dense canopies and high litter buildup can limit 

understory production. Management practices closely 

parallel those in adjoining, lower elevation ponderosa 

pine communities. These communities have high value 

for recreation and wildlife habitat, and are valuable for 

summer forage for mule deer and elk. Exotic or noxious 

plants are not important components of this community, 

although Kentucky bluegrass is an introduced species 

that is considered a good livestock forage (Chambers 

and Holthausen 2000) but is primarily associated with 

grassland openings in the forest. Gambel oak may 

replace ponderosa pine and Douglas fir stands after removal by fire or logging in southern Colorado (Engle 

and others 1983).

Humphrey (1960) described “fir” forests as “...so 

dense as to let little sunlight reach the ground and, as 

a result, except for a few shrubs and weeds and an occasional grass, there is little available to serve as feed 

for either game or livestock.” Timber harvesting of 

overstory can increase grass and forb production from 

11 to 17 g/m2/yr and increase deer use of the habitat 

(Patton 1976). Thill and others (1983) reported understory production values from 4 to 16 g/m2

 in unlogged 

mixed conifer forests, depending on tree basal area. 

Comparable values for thinned sites increased up to 74 

g/m2/yr for lowest basal area sites. Small clear cuts can 

increase forage for both domestic and native large herbivores from 8 to 28 g/m2

/yr (Ffolliott and Gottfried 

1989). Utilization by livestock and wildlife can increase 

but varies considerably between cleared patches. The 

very low forage production for this community would 

suggest that it would be easy to overutilize the forage 

base with only slight overstocking. Clearcutting of aspen stands in Colorado, however, resulted in few major 

changes in understory characteristics (Crouch 1983). 
12 USDA Forest Service RMRS-GTR-169. 2006.

Production of grasses, forbs, and shrubs was similar to 

uncut stands, except for some depression the first year 

after cutting. The largest effect of cutting was to increase aspen sprouting. Utilization by cattle was greater 

on older than recently logged or uncut sites. Reynolds 

(1962) recommended stocking rates of 10+ acres/AUM 

for pine-fir forests and 2 to 4 acres/AUM for associated 

grasslands.

Natural fire return intervals are considered to have 

been less frequent than those in the ponderosa pine 

community, but fire has played a role in structuring 

mixed conifer forests (Chambers and Holthausen 2000). 

Fire return intervals have been reported to be from 2 

to 27 years to 6 to 27 years, depending on forest type. 

However, Grissino-Mayer and others (1994) suggest a 

low intensity surface fire return interval of 4 to 6 years 

in the Pinaleno Mountains, based on fire scar data. This 

is similar to that for ponderosa pine communities. Fire 

regimes changed after European settlement. These authors suggest that spreading fires initiated naturally and 

by Native Americans were common until the period 

when livestock grazing intensified and fire suppression was initiated, and that grasses were an important 

component of fine fuel loads. Kaib and others (1999) 

and Swetnam and others (2001) studied fire scar data 

traversing a gradient in community type from desert 

grassland, Madrean pine (canyon pine-oak), to mixed 

conifer (Madrean) and estimated fire return intervals of 

4 to 8, 3 to 7, and 4 to 9 years, respectively, based on 

scars on at least 10 to 25 percent of the trees sampled.

There is only one quantitative grazing intensity comparison for this community type in Arizona and New 

Mexico. One of the 34 grazing exclosure plots assessed 

by Potter and Krenetsky (1967) and Krenetsky (1971) in 

the upper range of the ponderosa pine type was dominated by Douglas-fir. Grazing intensity at this site was 

probably moderate from ~1940 through 1951, when the 

grazing allotment was closed to grazing and use was 

only by recreationists. Grazing was for two months of 

the year by sheep, and the plot was in good condition 

and underutilized when the exclosure was established. 

Re-sampling of the plot was done in 1964, thus the area 

outside the exclosure was also ungrazed by livestock the 

previous 13 years. Unlike some other sites in this study, 

protection from grazing favored tree growth. Krenetsky 

(1971) concluded that this hindered grass recovery, but 

grass cover increased from 1940 to 1964 to a greater 

extent in the exclosure (0.9 to 6.7 percent) than outside 

(0.2 to 3.5 percent). Largest increase in grass was for 

creeping muhly. Forb cover decreased with protection 

from grazing, although it was always below 1 percent 

cover. The large increase in browse was primarily due to 

Douglas fir, which showed a 22 percentage-unit greater 

increase in the ungrazed compared to the grazed plot. 

Total vegetative cover was 4.4 and 36.6 percent in the 

ungrazed area in 1940 and 1964, respectively, compared 

to 0.8 and 8.85 percent in the adjacent grazed area. The 

diversity (richness) of plant species sampled decreased 

from 17 to 9 in the exclosure from 1940 to 1964, respectively, but increased from 9 to 12 in the grazed area. 

The loss of species with protection from grazing was 

due to the increase in tree cover and a loss of forbs and 

grasses that were minor components of the 1940 cover. 

Exclosure from grazing resulted in greater soil infiltration rates and concentrations of Ca and Mg, with other 

characteristics similar to the grazed area. Litter biomass 

was also similar between grazing treatments. The very 

low cover of grasses and forbs compares well with a 

geologic refuge (butte) in southern Utah that has never 

been grazed by domestic livestock and has a cover of 

less than 3 percent for the grass plus forb component 

(Mandany and West 1984).

Browsing of fir, spruce, and pine species by deer is 

considered in the next section (Ponderosa Pine).

No solid conclusions should be attempted concerning 

the effects of grazing on mixed conifer forests of the 

southwest based on only a single-site ungrazed/grazed 

comparison.

Ponderosa Pine and Madrean 

Pine Forests________________

The Madrean pine forests (often dominated by ponderosa pine) and ponderosa pine types are included in 

one section because of a general lack of information for 

the Madraen type. The latter type is differentiated in the 

text when possible.

Grazing of domestic livestock in the ponderosa type 

became intense during the 1880 to ‘90s (Currie 1975). 

Grazing was primarily by cattle until World War I, 

when sheep grazing became more important. Livestock 

numbers declined after the mid-1930s and are now less 

than one-third of peak numbers.

Reynolds (1962) considered the ponderosa pine type to 

be the most important community in the southwest with 

respect to timber production, water yields, trout habitat, 

and recreational use, second most important in big game 

habitat, and third in range production. Ponderosa pine 

communities of the southwest are sometimes separated 

into those occurring in the southern Rocky Mountains 

of Arizona, and those in the central Rocky Mountains 

from northern New Mexico, through Colorado and into 
USDA Forest Service RMRS-GTR-169. 2006. 13

the mountains of southern Wyoming. A large amount 

of the controlled experimental grazing work in the ponderosa pine type has been done at the USFS Manitou 

Experimental Forest in central Colorado. These communities are similar enough to those in New Mexico 

that we here include results and conclusions from the 

Colorado studies.

Precipitation in the Arizona communities ranges from 

356 to 457 mm/yr, and falls in two pronounced periods of 

winter and summer (Paulsen 1975). Precipitation in the 

central Rocky communities falls primarily during the 

April through September growing season, with means 

from 381 mm in New Mexico, 391 mm/yr at Manitou 

in central Colorado, to 635 mm/yr in Wyoming (Smith 

1967; Paulsen 1975).

Mountain meadow grasslands are interspersed 

throughout ponderosa pine stands of varying density. 

These grassland types are considered below in the montane meadow community type. As with other montane 

forest types, cattle often prefer grazing the grassland 

meadows over the forested sites (Smith 1967; Paulsen 

1975), making stocking rate determinations more 

difficult.

Species Composition

Farish (1889, cited in Cooper 1960) reported at the 

height of historic very heavy grazing in Arizona that 

much of the better forage plants had been lost within 12 

years (see History of Grazing by Domestic Animals section). Very high stocking rates coupled with a drought 

caused heavy losses of livestock. Governor Hughes 

(1893, cited in Cooper 1960) stated in his annual report: 

“In nearly all districts, owing to overstocking, many 

weeds have taken the place of the best grasses. In other 

places where ten years ago the end of the wet season 

would find a rich growth of grass, now is of inferior 

quality, or less quantity, or does not exist at all.”

In terms of species composition today, the understory of Arizona ponderosa communities on the Colorado 

Plateau are dominated by Arizona fescue and mountain 

and screwleaf muhlenbergias (Paulsen 1975). Central 

Rocky Mountain communities in the south are Arizona 

fescue and mountain muhly dominated, and Idaho fescue can replace Arizona fescue in the north or upper 

elevation limits, while Thurber and Idaho fescue may 

replace Arizona fescue in western Colorado and northern New Mexico. Forbs and shrubs are generally small 

components, and variable in space and time.

Arnold (1955) and Clary (1975) outlined successional stages of degradation due to grazing for Arizona 

ponderosa pine in forest openings:

Bunchgrass stage (Arizona fescue and mountain 

or screwleaf muhly)

Sod-forming grass stage (black dropseed, 

Kentucky bluegrass, blue grama)

Prostrate perennial forbs stage (pussytoes and 

sandwort)

Short-lived half-shrub stage (broom snakeweed 

and hymenoxys)

Annual stage (annual dropseed)

Denuded soil stage

Species in each progressively lower stage are shorter-lived. Arnold emphasized the use of life-forms in 

assessing condition of ponderosa range due to grazing. 

The replacement of tall and/or midgrasses by shortgrasses with grazing is a common response to grazing 

across plant communities worldwide (Milchunas and 

Lauenroth 1993; Diaz and others 2006). However, the 

presence of “sod-grass” forming species is often related 

to systems with a long evolutionary history of grazing 

(Milchunas and others 1988). While that is not the case 

for these systems, a sod-forming stage contributes a degree of protection from grazing to the system whereby 

erosion potentials are lowered even though species 

composition changes are occurring. Recovery from the 

denuded stage to a bunchgrass stage may require 100 

years even under reduced grazing pressure, based on 

observations of a half-shrub stage after 30 years (Clary 

1975). Successional stages of recovery may not follow 

degradation stages after complete protection from grazing, because bunchgrasses may establish immediately 

along with annuals and other life-forms, including pine. 

The savanna structure of ponderosa pine climax is confirmed by accounts of early explorers (Cooper 1960) as 

well as examination of relict stands on geologic refuges 

(mesas, and so forth) (Mandany and West 1984). Natural 

fire is a reason stands are not closed canopy (see section 

below). Forests are much more dense today than in the 

past, and succession in the absence of fire can proceed 

to a closed canopy structure (Cooper 1960).

The successional scheme described above by Arnold 

(1950) was based on five exclosures and adjacent grazed 

areas that were established in 1912 and resampled 36 

years later. Within the grazing treatments there were 

areas ranging from open to relatively closed stands. 

Plant species were grouped into bunchgrasses, grasses 

that avoid or tolerate grazing (sod grasses), and “weeds” 

(colonizing ruderals). Under protection from grazing 1) 

bunchgrasses increased (mountain muhly, Arizona fescue, mutton grass, pine dropseed, and less abundant spike 

muhly, deergrass, little bluestem, junegrass, and needle 

grasses), 2) sodgrasses decreased from their dominance 

•

•

•

•

•

•
14 USDA Forest Service RMRS-GTR-169. 2006.

in grazed treatments (blue grama, squirreltail, black 

dropseed, Fendler threeawn, and Arizona threeawn 3) 

“weeds” (colonizing ruderals) were a small proportion 

of the total and consisted of mainly perennials (yarrow, pussytoes, sandworts, thistles, daisies, asters, and 

penstemons), compared to mainly annual opportunistic ruderals in the grazed treatments. Blue grama and 

black dropseed dominated grazed communities. Even 

these short prostrate species declined in open-canopied 

sites where grazing pressure was relatively much more 

intense. Thus, greatest herbaceous cover was found in 

ungrazed open-canopied sites, whereas the least was 

found in grazed open canopied sites. Bunchgrasses in 

grazed treatments were often confined to refugia under 

denser pine cover. The survival of the awned species 

under grazing was attributed to the defensive structure of awns and that some were cool-season species 

that grew largely before the grazing season began. 

Arnold concluded that canopy closure under ungrazed 

conditions reduced productivity but had little effect on 

species composition. Grazing reduced productivity and 

had effects on species composition. Further, declines 

in production caused by canopy closure “will in time 

encourage improper grazing use and thus bring about 

discernable changes in plant composition.” Managers 

need to continually adjust stocking rates when overstory 

canopy conditions are in transition.

Cooper (1960) described a site where “livestock have 

been virtually excluded”...“ground cover approximates 

to some degree that of pre-white man days...” as “...as 

dense as is to be found anywhere in the southwestern 

pine zone,” and composed of 90 percent screwleaf muhly with few other grasses and forbs. However, Hanks 

and others (1983) classification of ponderosa pine types 

includes a description of a ponderosa/screwleaf muhly 

type which is the wettest most productive habitat, with 

screwleaf muhly as the dominant.

Potter and Krenetsky (1967) and Krenetsky’s (1971) 

reassessment of plots inside and outside grazing exclosures after 25 years of treatment included six locations 

in ponderosa pine. They report a definite increase in understory cover with protection, primarily due to grasses, 

but depending on elevation (see below). The increase in 

grasses was the greatest for any of the seven community 

types studied. Grass species increasing with protection 

from grazing were blue grama, Arizona fescue, prairie junegrass, and creeping muhly. It is interesting to 

note that blue grama displayed the greatest increase, because 25 years is a considerable period and blue grama 

is a component of Arnold’s (1955) and Clary’s (1975) 

second seral stage. Blue grama is generally considered 

an increaser with grazing throughout all Great Plains 

grasslands (Lauenroth and others 1994), and it decreases with protection from grazing in the shortgrass steppe 

with a long evolutionary history of grazing (Milchunas 

and others 1989). Potter and Krenetsky (1967) and 

Krenetsky (1971) observed no effect of protection on 

forbs, which were minor overall components. Browse 

increased more under protection than under grazing, 

and this was the only community studied where ponderosa pine increased under protection rather than when 

grazed.

An important relationship between response to 

protection and elevation of the site was discussed by 

Krenetsky (1971). Recovery of grasses with protection 

was only found for the two low elevation sites. At the next 

highest elevation site, there were no significant effects 

of protection on grass or shrub cover, although vegetation was sparse at this site. At the next higher elevation 

site, there was greater browse cover and protection favored tree overstory cover. At the highest elevation site, 

grazing significantly favored the overstory, with no 

significant effect on grass cover. Similar relationships 

were observed for sites in the pinyon-juniper type (see 

section below).

In northwestern New Mexico, Gardner and Hubbell 

(1943) studied two paired grazed and ungrazed plots after eight years of protection that were selected for study 

because they were severely overgrazed and invaded by 

the “weed” pingue, an indigenous bare-ground colonizing species. Heavy grazing continued on one of the 

grazed treatment sites and was reduced to light grazing 

on the other site. The sites were in openings in open savanna stands, so it is difficult in this case to classify as a 

ponderosa type or a mountain meadow type. Pingue decreased slightly over the eight year period on the heavily 

grazed treatment, compared to “marked” decreases on 

the ungrazed and lightly grazed plots. Three species 

of forbs greatly increased on the ungrazed and lightly 

grazed plots. Of the grass species sampled, sand dropseed decreased and bottlebrush squirreltail increased. 

New pine trees in the heavily grazed plot were shorter 

than those in the other three, although this was based on 

a limited amount of data. This is contradictory to other 

reports indicating increases in pine cover with fire suppression and grazing (see section below).

Quantitative controlled studies of grazing at the 

Manitou Experimental Forest in Colorado provide the 

most detailed accounts of grazing effects on plant species composition in the ponderosa type. Cattle favored 

open grassland areas within the large pastures (Smith 

1967). Utilization of open and timbered stands increased with increasing intensity of grazing. Use of the 

open ponderosa type was very low under light stocking 
USDA Forest Service RMRS-GTR-169. 2006. 15

levels. Use of dense timber areas was never greater than 

13 to 14 percent even under heavy stocking rates, and 

uniform utilization between grassland and open ponderosa was not achieved even with heavy stocking. Percent 

utilization was more closely related to the number of 

individual plants grazed than to the amount removed 

from an individual plant. Arizona fescue and mountain 

muhly were the most important forage species. Little 

bluestem and sedge were also palatable, but blue grama 

and fringed sagebrush were relatively unpalatable although the former provided a significant amount of 

forage biomass later in the season.

Responses of the plant communities at Manitou 

were studied over 14 years, although the period also 

included a seven year drought in the 1950s. For the 

ponderosa type, total plant canopy cover remained 

similar between 1940 and 1957 on lightly (10 to 20 

percent utilization) and moderately (20 to 40 percent 

utilization) grazed treatments but declined considerably 

on heavily (>50 percent utilization) grazed treatment 

(Smith 1967). These responses were due to grasses and 

sedges, rather than to forbs and other plant groups. 

Mountain muhly was resistant to grazing at light and 

moderate levels, although cover decreased by half on 

heavily grazed treatment. Arizona fescue decreased 

slightly under light and moderate grazing in grassland 

and remained similar in ponderosa, but it was reduced 

even more than muhly with heavy grazing. Blue grama 

increased in grassland but was not found commonly in 

the timbered stands. Cover of the short statured blue 

grama generally increases with grazing throughout its 

range (see references in Smith 1967 and Lauenroth and 

others 1994). Bottlebrush squirreltail increased with 

light grazing and declined with increasing intensity of 

grazing after that. Lupines, bearberry, Arkansas rose, 

and mountain mahogany all declined with grazing. 

Increasers with grazing, especially heavy, included a 

variety of colonizing ruderals species that were either 

annuals, low-growing, or stoloniferous and included 

tumblegrass, pussytoes, groundsels, and goosefoots. 

Sun sedge and fringed sagebrush showed no responses 

to the grazing intensities. A large number of species 

increased on protected sites, but only blue grama, 

pussytoes, and fendler sandwort decreased. The number of Arizona fescue, mountain muhly, or blue grama 

plants bearing flower stalks were generally similar for 

light and moderately grazed treatments, but all declined with heavy grazing. A study conducted to assess 

time for recovery of flower stalk production of heavily 

grazed plants after complete protection indicated three 

years was required for Arizona fescue and mountain 

muhly.

Grazing may also affect ponderosa pine regeneration 

through its effect on fuel loads and fire (see section below), by reducing competition between grasses and pine 

seedlings, or directly through trampling and consumption. Koehler and others (1989) found that mechanical 

removal of other plants around seedlings significantly 

increased their growth in both height and diameter, 

whereas chemical removal increased only height. A 

further study involving grazing cattle in enclosures resulted in enhanced growth of seedlings while causing 

only one percent mortality. A third study used large areas of ungrazed, moderately grazed in alternate years, 

and moderately grazed in consecutive years. Growth of 

pine increased with increasing grazing frequency, but 

density of seedlings was not affected by the grazing 

treatments. Currie and others (1978) found that seedling damage was high on heavily grazed treatments, 

but not on moderately or lightly grazed treatments. 

Rodents and rabbits caused more damage than cattle on 

the lightly and moderately grazed treatments. However, 

average height growth was greater in the lightly than 

either moderately or heavily grazed treatment in this 

study. Similar results for pine seedling mortality from 

other parts of the country are discussed by Koehler and 

others (1989). Although not in the southwestern United 

States, Neils and others (1956) reported that 90 percent 

of ponderosa pine seeds were consumed by small mammals before they had time to germinate, and that 60 

percent seedling mortality due to deer occurred in one 

year, and 99 percent mortality over three years.

In Colorado, Currie and others (1977) noted consumption of ponderosa pine seedlings when deer “happened 

upon them,” and that they bit off the terminal bud and 

upper needles. Newly fallen dead needles and live 

needles on lower branches of mature trees are also readily 

consumed. Fungi and dwarf mistletoe were large parts 

of the diet in heavily forested areas. Although grasses 

are a large part of the diet in more open habitat, there is 

little dietary overlap between deer and cattle, so grazing 

effects on the plant community would be expected to be 

different. Ponderosa pine is a considerable component of 

the diets of mule deer in many areas (Kufeld and others 

1973), and most commonly in winter (Crouch 1981). 

Deer can damage ponderosa pine as well as several fir 

and spruce species, but it is generally not considered as 

economically important in the southwest as it is in the 

northwest or northern Rockies (Crouch 1981). Local 

damage can be intense in some instances but is generally 

not widespread (Heidmann 1972). On the other hand, 

the classic example of herbivore overpopulating a range 

after predator control and drastically affecting plant 
16 USDA Forest Service RMRS-GTR-169. 2006.

communities comes from the Kaibab Plateau herd in 

the 1910s through 1930s (Rasmussen 1941; Leopold 

1943; Young 2002). While this example of deer altering 

plant communities includes ponderosa pine habitat, the 

Kaibab deer eruption is discussed in the Pinyon-Juniper 

section below. Browsing of understory gambel oak may 

be considerable, because it is a large component in the 

diet of mule deer in northern Arizona (Neff 1974 cited 

in Carpenter and Wallmo 1981), but heavy browsing 

stimulates sprouting that is considered desirable for 

wildlife (Sheperd 1971 cited in Carpenter and Wallmo 

1981). Elk were observed to browse an average of 57 

percent of the current-year branches of buckbrush. This 

shrub is an important component of midstory structure in 

ponderosa communities in the southwest and therefore 

important for diversity of other wildlife. These authors 

indicate that Rocky Mountain elk introduced into the area 

to replace the extirpated Merriam’s elk have increased in 

numbers to about 30,000 in Arizona and about 45,000 in 

New Mexico, and that these numbers are greater and the 

elk’s distribution greater than in the past.

There are not a large number of studies to base 

conclusions on the effects of grazing on species composition in ponderosa pine communities. Some results 

differ among studies, which may be due to elevation as 

suggested by one study, or simply to site or treatment 

differences. The studies suggest lower abundance of 

colonizing ruderal, annual, and low growing species in 

exclosures compared to grazed sites. A long-term well 

documented study showed little change through time in 

communities lightly or moderately grazed, but declines 

in understory cover with heavy grazing. There is contradictory evidence concerning the effect of grazing on 

ponderosa pine.

Aspen Communities

Aspen communities within the ponderosa type are 

important habitat for wildlife, and utilization can be high 

(Patton and Jones 1977; DeByle and Winokur 1985). 

This community can also be found interspersed within 

boreal forest and mixed conifer forest types. Understory 

productivity of grasses and forbs can be six times that in 

adjacent conifer forest (Reynolds 1969), making it particularly attractive to wild and domestic grazers. When 

grassland surrounds aspen stands, livestock prefer the 

grasslands, but use of the stands for shade can be heavy 

(DeByle 1985a).

Browsing of new aspen shoots by elk can be heavy under some circumstances and in some locations. Bailey and 

Whitham (2002) found that there was a positive relationship between severity of a burn and aspen sprouting in the 

absence of elk (in exclosures). However, elk selectively 

browsed aspen sprouts in the more severely burned sites, 

averaging 85 percent of shoots compared to only 36 percent in intermediately burned sites. The heavy browsing 

by elk more than negated the positive effect of intense 

burning on regeneration, and intermediately burned sites 

showed much greater regeneration. The differences in 

standing biomasses in the severely burned browsed and 

unbrowsed sites were dramatic, with 8967 g/m2 in the 

unbrowsed and 135 g/m2 in the browsed stands. Each 

combination of burning intensity and browsing intensity 

resulted in distinctly different arthropod communities 

associated with the different plant communities. Lack of 

regeneration of aspen in Rocky Mountain National Park 

in Colorado due to elk grazing has been of concern for 

many years, but survey studies indicate the problem to be 

of local and not regional or landscape concern. Kaye and 

others (2003) found a very patchy distribution of stand 

structure, no difference in regeneration between concentrated-elk-use winter range stands and dispersed-elk-use 

summer range stands, and no evidence of aspen decline. 

Suzuki and others (1999) found that regeneration was 

common within stands except in local areas of high elk 

utilization. It can take 6 to 8 years for sprouts to go to a 

height out of reach of elk (Patton and Avant 1970). Elk 

also chew the bark of mature trees during winter, and 

this can have an impact on stands by providing an entry 

way for pathogens, while complete girdling is not common (Debyle 1985a; DeByle and Winokur 1985). The 

effects of elk grazing on other plant communities has 

been reviewed by Wisdom and Thomas (1996). Deer, 

lagomorphs, and rodents may also remove bark, twigs, 

and sprouts for food, and these herbivores and birds utilize buds (DeByle 1985a,b; DeByle and Winokur 1985). 

Both aspen sprouts and aspen bark are high quality forages in term of nitrogen and digestibility (Mueggler 

1985; DeByle 1985b).

Gophers were estimated to consume from 6 to 39 

percent of the annual belowground primary production in aspen stands in Utah, and from 10 to 31 percent 

in meadow, and 5 to 6 percent in fir communities 

(Andersen and MacMahon 1981). However, means of 

estimating belowground net primary production at that 

time were limited and open to many serious errors and 

biases (Milchunas and Lauenroth 1992). A study by 

Cantor and Whitham (1989) suggests that belowground 

consumption of aspen by gophers in meadow communities in northern Arizona may be substantial, to the point 

that gophers limit the distribution of aspen communities 

to rock outcrop refugia where gophers cannot forage effectively. Mortality of aspen was 90 percent at meadow 

edges where gopher densities were high, compared to 
USDA Forest Service RMRS-GTR-169. 2006. 17

mortality of less than 20 percent where gophers were in 

low abundance. Aspen survival increased 3.5 times over 

control when gophers were removed, and tree growth 

rates increased 3 times.

Sheep compared to cattle grazing can be particularly 

hard on aspen sprouts, weakening a stand’s capacity to 

produce suckers after several years of heavy browsing 

(Patton and Jones 1977, DeByle 1985a). Grazing by 

both animals, if heavy, can have very large impacts on 

the community, as sheep remove forbs and browse and 

cattle remove grasses (DeByle 1985a). Heavy prolonged 

grazing generally results in a loss of plant species diversity and greater composition of less palatable species 

(Mueggler 1985). DeByle reviews grazing in the aspen 

type throughout its range and discusses competition between the various types of herbivores that may utilize 

the habitat.

Primary Production

Clary (1975) reviewed literature for understory 

aboveground primary production in Arizona ponderosa pine and gave typical ranges of 22 to 36 g/m2

/yr. 

Exponential and linear negative relationships with overstory basal area have been reported. Productive sites 

with dense overstory may produce 6 to 8 g/m2/yr and 

ungrazed open productive sites may reach 156 g/m2/

yr, while sites with infertile, shallow soils may be half 

as productive. In Colorado, understory aboveground 

production in ungrazed open timber compared to surrounding grassland was 37 and 181 g/m2

/yr, respectively 

(Smith 1967). Grasses and forbs are a large component 

of the herbaceous standing crop, and most are perennials. Grasses, forbs, and shrubs comprised 63, 29, and 8 

percent, respectively, of the production on cutover stands 

with a moderate canopy (Brown and others 1974), while 

for dense canopy stands values were 86, 17, and trace 

percent, respectively (Pearson 1964). The largest variability in production due to year-to-year fluctuations in 

weather were found where overstory canopy coverage 

was least (Paulsen 1975).

Several investigators have examined the effects of 

grazing on understory aboveground primary production, 

and some of these entailed controlled experiments and 

quantitative measurements. An early account of grazing effects was that of Arnold (1955), where values of 

production were given for range sites in excellent, good, 

fair, and poor condition classes in Arizona. Although 

experimental plots were described for species cover 

measurements (Arnold 1950), it is not clear how they 

were sampled for production, but graphs made from the 

Arnold values for two different community types appear 

similar and curves appear rather artificially smooth (Fig. 

4A). This may lead one to possibly doubt that these sites 

were actually measured, and that values represent qualitative estimates by the author. Sites in excellent condition 

were based on observation of relict areas and represent 

the ungrazed condition. If heavily grazed sites were represented as poor condition, then they produced only 36 

percent of that on the ungrazed excellent condition site.

Data from controlled studies in Colorado do not 

show a continual decline in production with increasing 

grazing pressure (Fig. 4B; Smith 1967). There were no 

differences among treatments two years after initiation. This was possibly because timbered sites were in 

relatively good condition prior to initiating the treatments, as cattle grazed associated grasslands more 

heavily. Associated grassland sites even increased production in response to a reduction in grazing pressure 

imposed two years before starting the grazing treatments. Understory production of ponderosa sites were 

not different among ungrazed, light (10 to 20 percent 

utilization), and moderate (30 to 40 percent utilization) 

grazing after 7 years of treatment, but heavily (>50 

percent utilization) grazed sites produced 58 percent of 

ungrazed sites. After 17 years of treatment, moderately 

grazed sites had greater production than ungrazed or 

lightly grazed sites, and heavily grazed sites produced 

46 percent of ungrazed sites. Heavy grazing affected 

early leaf growth, as indicated by leaf height measurements. Flower stalk production is discussed in the 

above section.

There is very little evidence to base conclusions 

concerning grazing effects on ponderosa pine community 

understory production. Both an apparent observational 

report of condition classes that may confound site with 

grazing effect and a well sampled controlled experiment 

suggest decreases in production with heavy grazing.

Root Responses

Root biomass in response to grazing exclosure, moderate, and heavy use was studied at the Manitou site in 

Colorado after 18 years of treatment (Schuster 1963; 

Smith 1967). Root biomass decreased with increasing 

intensity of grazing, and grazing resulted in a vertical 

distribution of roots (mostly grasses) more concentrated near the surface. Average maximum root depth 

was significantly reduced by heavy grazing, but there 

was no difference between moderately grazed and ungrazed treatments. There was a trend for lateral spread 

of roots to decrease with increasing grazing intensity, 

but this was not statistically significant. Individual roots 
18 USDA Forest Service RMRS-GTR-169. 2006.

of three important grass species were smaller in diameter and branched less under heavy grazing than those 

under lesser grazing intensities. Smaller effects on these 

variables were observed in moderately grazed treatments, and this may have been due to a lack of response 

by blue grama roots. Root biomass of blue grama is also 

very tolerant of grazing in shortgrass steppe (Milchunas 

and Lauenroth 1989).

Soils

Infiltration and erosion in response to light, moderate, and heavy grazing use were studied at the Manitou 

site in Colorado after 12 years of treatment, and in exclosures and adjacent grazed sites after 7 and 14 years 

of treatment (Dortignac and Love 1961; Smith 1967). 

No differences in infiltration were detected among 

light, moderate, and heavy use treatments, but data 

were highly variable. Infiltration inside exclosures in 

open ponderosa were greater than in grazed sites after 

7 years but did not increase further after 14 years of 

protection. Erosion rates in grazing exclosures tended 

to be less than in lightly and moderately grazed treatments but were substantially greater in heavily grazed 

treatment. However, the rates of erosion in the heavily grazed treatment were not considered excessive by 

the authors. Dunford (1954) reported increased runoff 

from these treatments with increasing grazing intensity, 

with no effect of moderate grazing on soil loss. Arnold 

(1950) mentioned that soils were less compacted, and 

organic matter greater, inside exclosures after 36 years 

of protection of northern Arizona ponderosa pine communities, but no data were provided.

Fire-Grazing Interactions

Mean fire return intervals for southern ponderosa 

pine (Madrean, canyon pine-oak) forests were estimated to be 3 to 7 years, based on fire-scar data (Kaib and 

others 1999; Swetnam and others 2001). Grazing by domestic livestock and subsequent fire suppression with 

increased settlement by European humans resulted in 

very evident reductions in fire frequencies throughout 

the southwest starting about 1880 to 1890 and continue 

Figure 4—Understory aboveground primary 

production (g/m2/yr) for A) ponderosa pine 

and mountain meadow communities in excellent, good, fair, or poor range condition 

classes (Data from Arnold 1950, 1955), 

and B) ponderosa pine communities ungrazed, lightly (10 to 20 percent utilization), 

and moderately (30 to 40 percent utilization), and heavily (>50 percent utilization) 

grazed for 2, 7, and 17 years (Data from 

Smith 1967).
USDA Forest Service RMRS-GTR-169. 2006. 19

to decline to present (Swetnam and Betancourt 1998). 

Cooper (1960) believed that fire suppression was the 

most important of the two but that grazing may have 

had an indirect effect on fire frequency through lowering of fuel loads. Prior to settlement, Indians had set 

fires for 10,000 years and lightning fires occurred naturally even prior to that. Removal of competition for 

pine seedlings and greater cover of bare mineral soil 

resulted in dense thickets of pine regeneration. A more 

open nature of the stands was reported by early explorers (Cooper 1960) and is also found on relict areas large 

enough to be exposed to lightning fires (Mandany and 

West 1984). Cooper found that fire thinned stands of 

young trees; only 18 percent of trees less than 1 inch 

dbh survived. Stands not exposed to fire had diameter 

distributions skewed toward smaller trees.

Fire suppression efforts and grazing effects on fuel 

loads were believed by most in the past to be the sole 

reason for the even-aged dense stands of ponderosa 

pine that established in the early 1900s. Cooper (1960) 

concluded that “long term climate changes do not 

appear to be a factor contributing to these changes.” 

However, Savage and Swetnam (1990) found that firescar records from stands in the Chuska Mountains 

showed a reduction in fire frequency four decades 

earlier than other parts of the southwest, but that 

dense even-aged stands established at the same time 

everywhere. The coincidence of the earlier decline in 

fire on this Navajo Reservation with its earlier history 

of heavy grazing by livestock supported the hypothesis 

that livestock grazing has had large effects on fire 

frequencies in the ponderosa type. Further, the decline 

in fire for the southwest in general preceded organized 

fire suppression. However, the regeneration pulses in 

forests both within and outside the reservation could 

be attributed to periods of both good seed supply 

and temperature and precipitation patterns highly 

favorable to seedling establishment. These authors 

concluded that the structure of present-day ponderosa 

stands was a complex interaction of anthropogenic 

disturbance and climatic conditions, and that fire and 

grazing regimes alone were not sufficient to cause the 

regeneration pulses. Swetnam and Betancourt (1998) 

present further evidence indicating a very important 

role of long-term weather cycles in driving recruitment/

mortality cycles. Drought was shown to have large 

impacts on tree populations, and when preceded by 

cycles of high production, also resulted in high fuel 

loads/dry conditions with widespread fires across the 

region. Grazing may play a role in these weather/fire 

cycles, but the primary driver is weather. Based on 

preliminary studies, Curtin (2002a) also stresses that 

“climate is the overriding factor determining the 

outcome of management actions.” Because “climatic 

trends may not be reflected in the age structure of a 

single forest or woodland” (Swetnam and Betancourt 

1998), early researchers studying single stands would 

be likely to miss or underestimate the importance 

of these longer-term cycles. It is interesting to note, 

however, that Ellison (1960) surmised the importance 

of drought in woody-plant mortality, but also suggested 

an interaction with grazing based on Young’s (1956) 

drought observations of greater mortality on lightly 

than heavily grazed ranges due to greater herbaceous 

cover effects on soil water. Young also noted that 

pricklypear cactus and mesquite were not as affected 

by drought.

Grazing Management

Approximately 80 percent of the ponderosa range 

in Arizona is grazed only in summer due to cold temperatures and snow accumulations (Clary 1975). At the 

time of the Clary report, continuous grazing occurred 

on 9 percent of the USFS allotments, rotation 15 percent, deferred 19 percent, rest-rotation 42 percent, and 

deferred-rotation 11 percent. Studies of advantages or 

disadvantages to different grazing systems have not 

been investigated for this range type. Currie (1976) 

reported no significant difference in production of 

montane meadows within ponderosa pine type to several grazing systems (see Montane Meadows section 

below).

Average grazing capacities were estimated at 7, 10, 

18, and 41 acres/AUM for good, fair, poor, and very 

poor range-class conditions, which is approximately 30 

to 40 percent utilization (Clary 1975). Currie (1975) 

based recommendations on slope and condition class 

with lower stocking on steeper slopes. Average acres/

AUM were estimated at 2 to 6, 3 to 12, 4 to no grazing, and 8 to no grazing for excellent, good, fair, and 

poor range-class conditions. Currie suggested that stubble heights of Arizona fescue should be 5 to 6 inches, 

and mountain muhly 1.5 to 2 inches, at the end of the 

grazing season to obtain the 30 to 40 percent utilization. Steepness and length of slope (except in extreme 

cases) and distance to water were factors important in 

utilization of ponderosa ranges based on an early study 

(Glendening 1944), but not in a similar study conducted 

by Clary and others (1978). This was attributed to fencing to smaller pastures and increased development of 

water-site densities. Holecheck and Pieper (1992) also 
20 USDA Forest Service RMRS-GTR-169. 2006.

recommended 30 to 40 percent utilization, but of key 

species. This amount of utilization would agree with the 

studies at Manitou in Colorado (see above) that found 

few negative effects of moderate grazing at 30 to 40 

percent. However, utilization in the ponderosa type was 

much less than that for grassland meadows within the 

ponderosa pastures. Utilization of grasslands in moderately stocked pastures ranged from 24 to 53 percent, 

whereas utilization of open ponderosa stands in the 

same pasture/years ranged from 11 to 40 percent, and 

dense stands from 0 to 4 percent. Management of ponderosa communities cannot be done independently of 

the associated mountain grasslands that are interspersed 

in varying amounts. Possibly a better approach would 

be to manage to keep associated meadow grasslands in 

a productive state, utilizing the forested sites even less, 

since light to moderate grazing may be sustainable in 

this system based on the Manitou studies. Sustainable 

in this case refers to primary production and secondary 

livestock production, not necessarily wildlife populations or desirable plant species composition.

Pinyon-Juniper______________

The pinyon-juniper type is one of the largest communities in the southwest in terms of area occupied 

(Gottfried and Pieper 2000). There is large variability 

in climate, topography, soils, and therefore plant species composition and stand structure (Springfield 1976). 

Precipitation generally ranges from 280 to 560 mm/yr, 

and amounts and seasonal distributions vary with elevation and geographic location. Precipitation is relatively 

more Mediterranean (45 percent April to September) in 

northwestern Arizona and understories have affinities 

to Great Basin plant communities, whereas in eastern 

New Mexico precipitation is continental (75 percent 

growing season) and plant communities have affinities to shortgrass steppe Great Plains grasslands (Lavin 

1953, 1964 cited in Springfield 1976). Mediterranean 

climates favor annual species and deep-rooted (shrub) 

communities over continental climates if annual average precipitation and temperatures are held constant, so 

responses to and thresholds to grazing may be different. 

Further, the long, intense evolutionary history of grazing by native bison in the Great Plains compared to the 

low intensity of historical grazing in Great Basin plant 

communities can also shape responses to current-day 

grazing by domestic herbivores (Mack and Thompson 

1982; Milchunas and others 1988), and this may or may 

not translate through affinities in communities that were 

not actually historically grazed.

The history of livestock grazing in the pinyon-juniper 

type follows the general history for lower elevation sites 

(see history section above), except that Gottfried and 

Pieper (2000) mention that most of the major Spanish 

settlements and Indian Pueblos were located within 

or close to pinyon-juniper communities. Gottfried and 

Pieper (2000) present an excellent summary and conceptual model of historical pinyon-juniper changes on 

the Pajarito Plateau, New Mexico (Fig. 5). Four main 

periods and five general states of pinyon-juniper communities are outlined. A dense understory vegetation and an 

open woodland structure were maintained by periodic 

fires prior to the Anasazi impacts from A.D. 1200 to 

1550. During this period, there was possibly an increase 

in fire frequency due to the use of fire by the Indians and 

a decrease in grazing by wildlife due to hunting. This resulted in a decreased cover of pinyon-juniper due to fire 

and cutting and an increase in understory as a result of the 

lower overstory and reduced grazing, and little overall effect on erosion due to localized farming and settlement. 

The Spanish colonization period from the late 1500s 

through the early 1800s saw some cutting of pinyon-juniper and limited introduction of domestic livestock, 

resulting in only localized changes in vegetation and 

soil erosion. Anglo settlement after 1880 brought similar 

conditions described for other communities above: large 

increases in domestic livestock and reduced fuel loads, 

fire suppression, and several periods of drought, and several changes in wildlife populations some of which may 

be particular to this location. These conditions combined 

to increase pinyon-juniper cover, bare soil and erosion, 

and the difficulty for herbaceous plant re-establishment, 

resulting in unstable “pinyon-juniper rocklands.”

Species Composition

The response of pinyon-juniper communities to grazing may depend in part on the successional stage being 

grazed. Gottfried and Pieper (2000) show three successional scenarios for the pinyon-juniper type based 

on different authors and different parts of the country. Fire is the succession initiating force in all three 

models. Annual, perennial grass - forb - shrub, and 

woodland stages are common to all. There are few 

large-scale problems with exotic species invasions into 

pinyon-juniper communities, although Russian thistle 

is widespread with many less abundant taxa. The effects of grazing on fire regimes is discussed in the Fire 

- Grazing Interactions section below.

Preference for particular species by livestock is one of 

several factors that can influence the directional change 

the species may take with grazing. In northern Arizona, 
USDA Forest Service RMRS-GTR-169. 2006. 21

Clary and Pearson (1969) observed the following preferences (percent utilization): prairie junegrass 43 percent, 

bottlebrush squirreltail 40 percent, mutton bluegrass 33 

percent, sideoats grama 30 percent, black dropseed 15 

percent, blue grama 5 percent, and spike muhly 4 percent. Blue grama is generally the most abundant grass 

in the pinyon-juniper type (Gottfried and Pieper 2000), 

which may account for its low preference.

Qualitatively, Paulsen (1975) suggested that spring 

grazing has been especially detrimental to cool-season 

grasses such as muttongrass, bottlebrush squirrel-tail, 

and western wheatgrass. Humphrey (1960) described 

condition classes for Arizona pinyon-juniper type. 

Excellent condition was considered to have almost 

entirely palatable perennial grasses such as hairy and 

sideoats grama, plains lovegrass, woolly bunchgrass, 

pinyon ricegrass, Texas beardgrass, little bluestem, and 

curly mesquite. This condition was thought to be sustainable if proper utilization levels were maintained, but 

condition could decline during drought. Grazing intensities should be especially watched and adjusted under 

drought conditions. Good condition range has enough 

cover, and of the palatable species, to resist erosion losses. Junipers are widely spaced. Fair condition is marked 

by low cover and density of vegetation, although it is 

still composed of perennial grasses. Production is half 

of potential, and grazing should be deferred during the 

summer at least every third year. Poor condition range 

produces almost no forage, and plants that are there produce little seed or vegetative reproduction. Removal of 

grazers for a couple of growing seasons and re-seeding 

of native forage plants was recommended.

There are several studies of pinyon-juniper communities on isolated, “relict” geologic refuge sites. Large 

geologic refuges such as buttes and mesas have advantages and disadvantages when used to compare to 

mainland conditions (Milchunas and Noy-Meir 2002). 

An advantage is that the refuge communities have never 

been grazed by very large herbivores such as domestic 

animals, whereby potential shifts to alternate states due 

to historical very heavy grazing before scientific study 

can be eliminated as a potential confounding factor in 

present-day condition. On the other hand, these isolated 

islands can differ in abiotic conditions, fire frequency 

may be lower than the historical average because spread 

of natural fires to the isolated area may have also been 

inhibited, and in some cases the small mammal and 

even insect assemblages may be limited by dispersal 

Figure 5—Three successional models for pinyon-juniper communities in different locations. From Gottfried and Pieper 

(2000).
22 USDA Forest Service RMRS-GTR-169. 2006.

capabilities. In other cases, a large size can allow good 

chances for lightning strikes and fire spread, and ease of 

accessibility to all but large herbivores can make these 

communities important for comparative purposes.

Baxter (1977) compared a 10 acre mesa in northcentral Arizona to surrounding grazed sites that were 

either continuously grazed or under a rest-rotation system with an apparently lower grazing intensity. The 

mesa was accessible to rabbits, but fire was probably 

not common on this small, isolated area. Baxter states 

that he used to think blue grama was an important forage species of woodland sites where it often heavily 

dominates, and that many in the area think “blue grama 

has been the salvation of the rancher” in the southwest 

because other species had been eliminated by grazing 

long ago. Observations and measurements taken on the 

mesa show a plant community where the cool-season 

grasses mutton bluegrass and bottlebrush squirreltail 

make up 60 percent of the plant community, and blue 

grama only 7 percent. In contrast, the two cool-season grasses were not encountered on the continuously 

grazed site, and blue grama comprised 77 percent of 

species composition. Cool-season grasses and blue 

grama were equally abundant in the rest-rotation site 

(35 percent each). Forbs did not vary much with grazing 

treatments. Bare ground was 65 percent on the continuously grazed, 25 percent on the rest rotation site, and 8 

percent on the mesa. A lack of juniper on the mesa was 

attributed to the competition from grasses, since fire 

was probably not a factor. Production was estimated to 

be about 100 g/m2 compared to 10 g/m2 in the grazed 

area. Biomass sampling appears very weak, but Baxter 

believed that there were very large differences between 

the three sites. The continuous versus the rest-rotation 

system comparison was confounded by also having differences in grazing intensity, and it was considered a 

“managed” versus non-managed comparison. The responses reported occurred over a 20-year period.

Schmutz and others (1976) studied another relict, 

geologic refuge area in northeastern Arizona. The 12 ha 

(30 acre) butte is protected from grazers by 150 to 350 m 

high sheer cliffs, so small mammal populations and fire 

may also be affected by the butte structure. Comparable 

grazed areas were not sampled, so only qualitative 

visual observations suggested “a marked contrast in 

luxuriance of growth of grasses on ungrazed” compared 

to “the typically overgrazed areas on the mainland.” 

Plant communities were dominated by perennials, with 

almost no annuals. However, the perennials were species of low to moderate palatability similar to those on 

the mainland, which may be attributed to the shallow 

soils of the butte. Thatcher and Hart (1974) reached 

similar conclusions after studying another 40 acre mesa 

in northern Arizona. Some parts of the mesa had pure 

stands of grass and pinyon, juniper, shrubs, and very 

little grass in others. Differences could be attributed to 

soil types and fire frequency. Pinyon-juniper was dominant in the absence of fire regardless of soil type, but 

grass was abundant after fires only on the sandy surface-textured soils. Mandany and West (1984) observed 

very low herbaceous cover in a pinyon-juniper community on an ungrazed mesa in southern Utah. Johnson 

(1983) studied five ungrazed buttes and a grazed area in 

southern Utah and reported floristic similarity between 

grazed and ungrazed sites but that the grazed area had 

only one species of grass while each butte site had at 

least six. However, some of these buttes did not have 

similar small mammal and insect populations as found 

on the grazed mainland site. (See also the relict, geologic refuge studies in the Great Basin Juniper-Sagebrush 

savanna section below because of similarities with this 

community type.)

The overall conclusion from the studies of relict, 

geologic refuges never grazed by domestic livestock 

compared to grazed areas is that there are sometimes 

very large differences in species composition and species richness and other times there are not, and this may 

be attributed to soil type and fire history. In some locations, infertile shallow soils rather than past or current 

grazing abuse may be responsible for low coverage of 

herbaceous understory plants. More highly productive 

sites are more likely to show effects of grazing, which 

is consistent with qualitative and quantitative models at 

a global scale (Milchunas and others 1988; Milchunas 

and Lauenroth 1993).

There are only a few controlled studies on the effects 

of grazing on species composition of the pinyon-juniper 

type. This seems unusual given the extensive distribution and use of this type for grazing. Springfield (1959 

cited in Springfield 1976) indicated that responses to 

grazing protection for 10 to 14 years depended on tree 

canopy cover. The effects of grazing were “negligible” 

when the canopy was greater than 30 percent, whereas 

perennial grasses increased “substantially” with protection when tree cover was less than 20 percent. This 

would suggest poor possibility for improving conditions 

in dense woodland by protecting them from grazing 

(Potter and Krenetsky 1967). Springfield did not suggest 

why this may have occurred, provide grazing intensity 

or distributions at the grazed sites, or indicate the prior 

condition of the exclosure(s). However, the common response to grazing worldwide (Milchunas and Lauenroth 

1993) was observed, whereby short prostrate species 

increase with grazing and midgrasses such as sideoats 
USDA Forest Service RMRS-GTR-169. 2006. 23

grama increase with protection. Springfield also noted 

that annuals increased with grazing and the palatable 

fourwing saltbush decreased. While shrubs are more 

likely to increase worldwide with grazing (Milchunas 

and Lauenroth 1993), the highly nutritious and palatable C3

 shrub fourwing saltbush generally declines with 

grazing throughout its range.

Potter and Krenetsky (1967) and Krenetsky (1971) 

re-evaluated six exclosures at five different sites in the 

pinyon-juniper type 25 years after their establishment. 

These were the same sites examined by Springfield 

(1959 cited in Springfield 1976) after 10 to 14 years 

of protection (paragraph above). Potter and Krenetsky 

(1967) and Krenetsky (1971) analyzed and presented 

the same data differently. Potter and Krenetsky’s (1967) 

analysis did not relate changes to differences in canopy cover as did the earlier Springfield assessment or 

the Krenetsky (1971) analysis. Language in the Potter 

and Krenetsky (1971) paper is confusing and not well 

defined, but apparently states that grass cover was “consistently” higher in ungrazed than grazed sites, with 

prairie junegrass becoming more abundant. Forbs may 

be slightly greater on grazed sites, and shrub cover as 

well as pine and juniper cover increased over time to 

a greater extent in grazed sites. Krenetsky (1971), in 

the more detailed dissertation, specifically differentiates between overstory cover and response to grazing, 

and he relates tree cover to elevation. Development of 

overstory cover was positively related to increasing 

elevation, as was also observed for the ponderosa pine 

type. Differentiating in this respect resulted in similar 

conclusions as those of Springfield; grass recovery was 

favored by protection when canopy cover was lower. 

This, however, was based on two low elevation sites 

where grass was significantly greater with protection, 

compared to three higher elevation sites with greater 

overstory cover where differences between grazing 

were not significant. Krenetsky (1971) placed some 

confidence in these findings because they appeared 

consistent in other vegetation types as well as in pinyonjuniper, but the response was strongest in pinyon-juniper 

vegetation. Krenetsky’s overriding conclusion was that 

recovery of ranges by even complete protection must in 

some cases “be seriously questioned.”

At a southeastern New Mexico pinyon-juniper 

site (precipitation 389 mm/yr), Pieper (1968) studied 

grazed (intensities not provided) and adjacent ungrazed 

sites for 12 years on loamy uplands and lowlands and 

a stony hills site. The loamy sites were open grasslands 

and the stony hills site had an overstory. Plant cover was 

not affected by grazing treatment on either loamy site, 

but it was greater on grazed sides of the fence on the 

stony hills site (possibly due to blue grama increase). 

In terms of compositional percentages, no statistically 

significant differences were observed on the stony hills 

site, and means were generally close. Some compositional changes were observed on the two loamy sites. 

On bottomlands that were probably relatively heavily 

grazed, blue grama and western wheatgrass composition declined with grazing while mat muhly increased. 

On loamy upland sites, mat muhly also significantly 

increased with grazing, while Carruther’s sagewort 

significantly decreased and blue grama tended to decrease. At the same research station, Gamougoun and 

others (1984) estimated vegetative cover in exclosures 

to be 43 percent compared to 50 percent in moderately 

stocked treatment, 40 percent in heavily stocked, 35 

percent in rested rotation pastures, and 36 percent in 

grazed rotation pastures. It is interesting to note that 

blue grama, a species that increases throughout Great 

Plains grasslands (Lauenroth and others 1994), sometimes decreases even in this area possibly influenced by 

bison grazing (see Fig. 3).

Long-term (40 yrs) protection from grazing was 

studied in northeastern Arizona and compared to areas moderately to heavily grazed with sheep and cattle 

(Brotherson and others 1983). This is a relatively dry 

site receiving only 290 mm/yr precipitation. Vascular 

plant and litter cover showed no differences among 

treatments, although bare soil cover was greater in 

grazed areas due to loss of cryptogamic crusts (see 

below). Trees, shrubs, and forbs had similar covers between treatments, but grasses were 2 1⁄2 times greater 

on ungrazed compared to grazed areas. Some species 

increased in response to long-term protection (prickly 

gilia, brittle pricklypear, Indian ricegrass, longtongue 

muttongrass) and some decreased (slenderbush eriogonum, slender gilia, sandhill muhly).

There have been a couple studies of cryptogamic 

crusts of ungrazed and comparable grazed sites in the 

pinyon-juniper type. Beymer and Klopatek (1992) studied five sites in northern Arizona that ranged from a 

never-grazed relict, to sites not grazed since the 1930s, 

to sites currently grazed. Unlike findings from other 

communities, these authors reported that species presence was not related to the amount of grazing. Grazing 

reduced cover of cryptogamic crusts from 23 to 5 percent but did not result in loss of species. Brotherson and 

others (1983) studied cryptogamic crusts in northeastern 

Arizona in heavily grazed and adjacent ungrazed sites 

fenced for 40 years. Vascular plants were less affected 

by grazing than cryptogamic crusts. Mosses and lichens 

declined more than 80 percent with grazing, while algal and grass cover were reduced less than 60 percent. 
24 USDA Forest Service RMRS-GTR-169. 2006.

Twenty-two species were sampled on the ungrazed 

treatment compared to only 8 on the grazed. These 

authors suggested periods of rest from grazing may enhance reestablishment. Ladyman and Muldavin (1996) 

reviewed the effects of grazing on cryptogamic crusts 

in a number of arid western communities and concluded 

that recovery times can be substantial. Grazing during 

periods of the year when crusts are more likely to be 

wet minimizes the impacts, and they are also less susceptible to damage in winter as well. However, climatic 

conditions in some parts of the pinyon-juniper range in 

the southwest are never conducive to mitigating grazing 

impacts on this sensitive part of the community.

Grazing by feral animals is an important factor 

in many parts and in many plant communities of the 

southwest. Feral burros were first observed in Bandelier 

National Monument in north central New Mexico 

in the late 1930s. A canyon that is a barrier to burro 

migration provided a comparison between burro-impacted and relatively ungrazed sections of the monument 

(Koehler1974). Large differences in condition of pinyon-juniper vegetation was observed on opposite sides 

of the canyon. Under very heavy grazing pressure, 

grasses occurred only under protection of Opuntia or 

dense brush cover. Common wolfstail, sideoats grama, 

and New Mexico feathergrass were among the species 

most sensitive to grazing, while blue grama, threeawn 

species, and ring muhly were the least sensitive. Heavily 

grazed sites showed a high degree of pedestalling of 

plant crowns and little reproductive output. Some use 

of ponderosa pine communities was also observed, 

but there was limited use of higher elevations, heavily 

forested areas, or steep slopes. Utilization of forages in 

heavily grazed sites was estimated to be up to 80 percent. Deer and trespass cows also contributed to the 

observed conditions. The effects of burro grazing on 

vegetation translated into impacts on bird populations 

as well, and these effects were most pronounced on 

foliage nesting birds, which showed a proportionately 

greater decline on burro-grazed compared to ungrazed 

areas of the park than other nesting guilds (Wauer and 

Dennis 1980).

The Kaibab Plateau in north-central Arizona is a mix 

of pinyon-juniper at lower elevations, leading to ponderosa pine, and then spruce-fir mixed conifer forest with 

aspen as elevation increases (Rasmussen 1941). For a 

large part of the 1900s the Kaibab deer eruption was cited as the classic example of predator control, herbivore 

population eruption, destruction of forage base, followed 

by population crash and decline to lower than historic 

levels (Rasmussen 1941; Leopold 1943; Kimball and 

Watkins 1951; Mitchell and Freeman 1993; Young 2002). 

Large numbers of mountain lions, wolves, coyotes, and 

bobcats had been killed in predator control efforts that 

were popular wildlife management techniques of that 

time (Rasmussen 1941; Mitchell and Freeman 1993; 

Young 2002). In 1970, Caughley argued that habitat and 

resource supply, not predation, is generally most likely 

to result in eruptions, which lead to much debate over 

the classic story involving predator control. Mitchell 

and Freeman (1993) conducted a review of the evidence 

for the Kaibab situation and concluded that the eruption 

was due to predator control (primarily mountain lion), 

removal of large numbers of livestock prior to the eruption and the release of the forage base, and a prohibition 

of hunting in an area that had traditionally been heavy 

hunted by Native Americans. They concluded that the 

decline of the deer herd was also due to multiple factors 

that included the extreme overutilization of the forage 

by the large numbers of deer, the effects of the new fire 

control efforts of the time on decreasing understory 

production, and a drought that coincided with the peak 

of the eruption. These are basically the original factors 

presented by Rasmussen (1941). Although removal of 

predators was one primary reason for the irruption, the 

historical simplification of the situation for textbook 

purposes and for a philosophical need at the time to emphasize the importance of predators as a component of 

the ecosystem rather than an evil led to the Kaibab herd 

becoming the classic story that it has become. Regardless 

of specific reasons, the Kaibab deer herd represents an 

unusual situation where a number of rapidly changing 

factors coincided to produce very large effects of a wild 

herbivore population on vegetation communities. It was 

estimated that 60,000 deer starved to death within an 

area of 2,980 km2. Reports from the time indicate that 

juniper, pine, fir, spruce, and a variety of shrubs were 

severely utilized and damaged (Kimball and Watkins 

1951). Leopold (1943) reported that the effect on winter 

food plants was very severe. Boone (1938) suggested the 

damage was so severe that there was a loss in carrying 

capacity of as high as 90 percent. Vegetation that was 

first observed to be in a downward trend in 1919 was 

not considered to have generally recovered until 1941 

(Kimball and Watkins 1951).

Krausman and others (1992) conducted a followup 

to an earlier Leopold and others (1947) survey of “overpopulated deer herds” in the United States. The Leopold 

survey identified nearly 100 herds, six of which were in 

Arizona and eight were in New Mexico. Of these herds, 

all were considered to be at or below carrying capacity in 

the 1980s, except for occasional or limited overbrowsing 

at the Kaibab north and Kaibab south ranges. However, 

the original Leopold term of over-populated as well as 
USDA Forest Service RMRS-GTR-169. 2006. 25

current terms concerning carrying capacity are now 

considered value laden and/or difficult to scientifically 

define.

There are several studies of grazing effects on species 

composition in the pinyon-juniper type. However, it is 

a widespread community type that varies greatly in 

species composition. Soil type, tree cover, and proximity 

to areas historically grazed by bison may affect level 

of response to grazing. Cryptograms generally show 

large declines with grazing. There is some evidence 

that increasing overstory cover can override grazing 

effects, but that in more open sites there is loss of 

species palatable to livestock and increases in annuals. 

Sites in proximity to areas historically grazed by bison 

may display less loss of overall cover due to increases 

in species with low growing prostrate growth form. 

However, even at these sites the soil type has influences 

on response to grazing. Blue grama may or may not 

increase with grazing, depending on site, even though it 

is a consistent increaser in Great Plains grasslands.

Primary Production

As with many other woodland types in the southwest, tree canopy cover can vary from dense to open 

savanna; precipitation in open stands is generally at the 

low annual average end and is at the high end in dense 

woodlands (Springfield 1976). Productivity can vary up 

to 50 percent with local soil types in northern Arizona 

(Jameson and Dodd 1969). Thatcher and Hart (1974) 

estimated production on a never-grazed relict in northern Arizona and found a range across soil types of 45 

to 147 g/m2/yr. Pieper (1995) found temporal variation 

in productivity ranging from 19 to 73 and 30 to 100 g/

m2/yr for locations over 12 years of measurement. Clary 

(1987) reviewed seven studies across a geographic range 

through Arizona and New Mexico and reported from 7 

to 93 g/m2/yr. Kruse and Perry (1995) studied mature 

old-growth stands and reported understory production 

of only 3 to 14 g/m2/yr. There are extreme effects of 

overstory canopy cover on productivity; Jameson and 

Dodd (1969) report 58 g/m2/yr when canopy cover was 2 

percent to 10 g/m2/yr when cover was 30 percent. Clary 

(1987) and Pieper (1995) reviewed the large number of 

studies on overstory-understory relationships. Increases 

in production after pinyon-juniper control ranged from 

33 to 2900 percent based on six studies across eight 

locations (Pieper 1995). Springfield (1976) states that 

there are some stands, usually dominated by pinyon 

pine, so dense as to have practically no understory and 

have little or no value for grazing.

Because of the generally low forage productivity, 

increases in juniper cover and invasion into grassland 

and drier areas of ponderosa pine led to large numbers 

of conversion projects after World War II, which were 

and still are controversial (Gottfried and Pieper 2000). 

Springfield (1976) states that there is more information about control of pinyon and juniper than any other 

aspect of management of this community. Clary and 

others (1974) concluded that increased yields were not 

sustained long enough to make conversion economically justifiable. Recently, Ffolliott and Gottfried (2002) 

questioned earlier perceptions of “massive invasion” by 

trees. A site followed from 1938 to 1991 showed an increase of only 1.2 trees/acre, but large increases in tree 

crown diameter and height could possibly be misperceived as invasion. An additional alternative view is 

that invasion may not necessarily be into new areas, 

but just be a re-occupation of previously occupied areas (Gottfried and Pieper 2000). Conversion to improve 

livestock forage production or wildlife habitat is a human decision and action, not a direct or indirect effect 

of the grazing animal, so it will not be discussed in 

detail here. However, Lanner (1993) argues that conversion, which he terms “deforestation,” 1) economically 

only breaks even in terms of forage production for even 

successful projects, 2) does not increase water yields as 

hypothesized, 3) does not reduce erosion, and 4) fragments habitat for wildlife rather than improves wildlife 

habitat because any manipulation of habitat is beneficial 

for some species and detrimental for others.

The effects of grazing on primary production of the 

pinyon-juniper type have been assessed both qualitatively and quantitatively, although data are very limited 

for both. Clary (1987) reviewed production values for 

grazed sites and separate independent sites that were not 

grazed (sites were not experimentally paired, but were 

from different studies), and estimated that ungrazed 

sites often produced more than 135 g/m2/yr compared 

to generally less than 100 g/m2/yr for grazed sites. 

Apparently based on ocular estimates, Baxter (1977) estimated production on a relict mesa to be approximately 

100 g/m2/yr compared to 10 g/m2/yr on nearby grazing 

allotments.

There are only two quantitative studies of production in the pinyon-juniper type, and only one provides 

any statistical tests. Springfield (1959, cited in Potter 

and Krenetsky 1967 from the sites described above in 

Species Composition section) gave estimates for dense 

woodlands of from 20 to 26 g/m2/yr in protected compared to from 13 to 25 g/m2

/yr in grazed plots. These 

values are very similar, as were differences between 

grazing treatments in species composition. However, in 
26 USDA Forest Service RMRS-GTR-169. 2006.

open pinyon-juniper sites production on protected sites 

was twice as much as on grazed sites. Pieper (1968) 

provides production estimates for 12-year-protected 

and grazed stony hill topographic sites of 63 and 53 g/

m2/yr, respectively, for loamy bottomland of 68 and 33 

g/m2/yr and for loamy upland of 73 and 62 g/m2/yr, for 

the respective treatments. All three protected vs. grazed 

comparisons were statistically significant. The large 

difference with grazing in production on bottomlands 

represents the relatively high grazing pressure in this 

topographic position within a particular pasture stocking 

rate. Other locations show small decreases in production 

with grazing. This site is in southeastern New Mexico, in 

an area possibly influenced by bison grazing.

Reproductive effort of plants is often used as an indication of condition or stress. In a review of this concept, 

Obeso (1993) found defoliation caused mainly reductions in the reproductive output of herbaceous perennials. 

Pieper (1968) measured the number of seed stalks per 

unit ground area in protected and grazed pinyon-juniper 

and observed significant reductions on loamy bottomland and upland sites, but not on stony hill sites. Stony 

hill sites were more heavily dominated by blue grama, 

and there were no differences in species composition 

between grazing treatments. Average heights of seed 

stalks were significantly lower in grazed compared to 

protected treatments at all three topographic sites.

No data were located concerning root responses to 

grazing in the pinyon-juniper type.

There are few studies to base conclusions on effects 

of grazing on primary production of the pinyon-juniper 

type. Available evidence suggests large declines in 

production in open-canopy juniper communities distant 

from the geographic range of bison, small declines 

in dense woodland, and relatively small declines in 

communities within the historic distribution of bison.

Soils

Grazing can affect both soil nutrient and water-holding capacity through influences on erosion of topsoil 

with reductions in plant cover and by decreasing infiltration rates through soil compaction. Soil organic 

matter can be lost through erosion as well, and/or inputs 

of carbon to soils can be reduced if primary production 

decreases or decomposition increases due to increased 

soil temperatures. While pinyon-juniper systems are 

generally nitrogen deficient, and sometimes phosphorus 

and potassium deficient, water stress more often limits 

plant growth (Gottfried and Pieper 2000). Very little data 

are available on grazing effects on soil nutrient or carbon 

levels in the pinyon-juniper type. Beymer and Klopatek 

(1992) did not find trends in soil organic carbon, texture, or pH over a five-site gradient from ungrazed to 

increasing grazing pressure. Grazed rather than ungrazed sites tended to have insignificantly higher soil 

nitrogen concentrations. Krenetsky (1971) found no 

consistent directional differences between grazed and 

ungrazed plots at five sites for soil texture, N, P, K, Ca, 

or Mg concentrations. Wilcox and Wood (1988) found 

only slight differences between two ungrazed and two 

lightly grazed sites in southeastern New Mexico; ungrazed sites had lower organic carbon and sand content. 

No differences were apparent between grazing levels 

and soil bulk density or soil depth. Organic carbon was 

also lower in exclosures compared to moderate or heavy 

continuous grazing or rested or grazed short duration 

treatments in central New Mexico at one site, but not at 

another (Weltz and Wood 1986).

Erosion in the pinyon-juniper type is highly variable both within and between geographic location 

(Gottfried and Pieper 2000). Within site, movement of 

water and soil is least under trees where litter buildups 

produce micro-topographic hummocking, and greatest 

between trees where there can be less cover and greater 

compaction by livestock and wildlife. Between sites, 

observations from studies are often inconsistent (see 

review in Gottfried and Pieper 2000). Davenport and 

others (1998) proposed a threshold model that operates 

very differently under different site conditions to explain 

the reason why some pinyon-juniper systems are undergoing accelerated, heavy erosion, others are undergoing 

little or no erosion, and few intermediate cases exist. On 

sites with high soil erosion potential (due to climate, soil 

type, geomorphology), erosion rates are very sensitive 

to even small changes in cover. Conversely, increasing 

site erosion potential has little effect on erosion under 

high cover, but under low cover erosion rate is very 

sensitive to increasing site erosion potential. The tree 

component can be important in contributing to the 

threshold through its influence on understory cover. 

Grazing can lead to threshold conditions directly by 

affecting understory cover, and indirectly by reducing competition for tree establishment or by reducing 

fuel-loads and tree mortality by fire, thereby further decreasing understory cover. Jameson (1987) also proposed 

a cusp-catastrophe model of alternative steady states for 

pinyon-juniper communities, where fire, grazing, and 

climate interact in producing the alternate states.

At locations where erosion rates have been high, reestablishment of understory cover is difficult (Gottfried 

and others 1995). Wildlife may be limiting seed availability through consumption, and eroded soils are more 
USDA Forest Service RMRS-GTR-169. 2006. 27

xeric, nutrient poor, and warmer, thereby limiting seedling establishment. Gottfried and others (1995) indicate 

that simply eliminating livestock that contributed to the 

high erosion situation will not halt or reverse the problem because the system is currently driven by physical 

abiotic processes rather than by biological processes. 

This was evident from areas where livestock were removed in 1932 and 1943 and erosion continued to be 

high.

However, there are only a few actual studies of 

infiltration/erosion in pinyon-juniper communities. At 

a 384 mm/yr precipitation site in south-central New 

Mexico on fine-loamy soils, Gamougoun and others 

(1984) found greater infiltration rates in exclosures 

(ungrazed 30 years) than in continuous moderately or 

heavily grazed treatments, and the latter had greater 

infiltration than rested and grazed pastures in a rotation 

system. There were no differences between moderate 

and heavy grazing treatments. Sediment production was 

highly variable within and between years. Sediment 

production tended to be less in ungrazed exclosures 

than all grazing treatments in all three years, but it was 

significantly less in the ungrazed exclosure in all years 

only when compared to the grazed rotation pasture. 

Significantly less sediment production was observed 

in the exclosure compared to the continuous heavily 

grazed pasture in only one of the three years and was 

never statistically different between the exclosure and 

the moderately grazed pasture. At the same site during 

a fourth year of study, Weltz and Wood (1986) observed 

greater infiltration rates in the exclosure than any of 

the grazed treatments, and greater rates in moderately 

grazed than heavily grazed treatments. At a second 

fine-loamy site where the exclosure had been ungrazed 

since the turn of the century, infiltration rates were not 

different among ungrazed, moderately grazed, and a 

rested short duration pasture. The grazed short duration 

pasture showed much lower infiltration rates than any 

of the other three sites. At a site in southeastern New 

Mexico with 500 mm/yr precipitation and shallow 

gravelly loams to clay loams, Wilcox and Wood 

(1988) found 12 to 17 percent lower infiltration on 

ungrazed compared to lightly grazed slopes averaging 

a 50 percent steepness when soil was initially dry or 

wet, respectively. The difference was significant, 

however, only for the dry soil condition. Sediment 

concentrations were similar between grazing treatments 

for initially wet soil conditions, and only slightly 

higher with grazing for dry soil conditions. Sediment 

production values between treatments were similar to 

results for infiltration.

Fire-Grazing Interactions

Interactions between grazing and fire frequency and 

the effects on community structure in pinyon-juniper 

communities is similar to that reported in the ponderosa 

pine section above and will not be repeated here. Fig. 5 

gives an overview of the history and role of fire in the 

pinyon-juniper community. Estimates of natural firereturn intervals for pinyon-juniper are rare and studies 

were in the upper border where it occurs with ponderosa 

(Gottfried and others 1995). Fire scars on pinyon are 

rare (trees are often killed), and junipers cannot be accurately dated due to numerous false and missing rings. 

Gottfried and others (1995) reviewed the available studies and give estimates of 11 scars over 200 years, a 20 

to 30 year return interval, and a 15 to 20 year return 

interval.

Jameson (1987) considered not only grazing-fire 

interactions in driving alternative steady states in the 

pinyon-juniper type, but also the interaction between 

grazing and weather. Years of precipitation sufficient 

for tree establishment would be less common in ungrazed sites where grasses compete for soil water but 

more common when grazing removed the vegetation.

Grazing Management

As mentioned earlier, there is probably more research concerning control, conversion, and thinning of 

pinyon-juniper than any other aspect of management, 

use, or ecology of the system. While large-scale conversions are presently controversial from both ecological 

and economical standpoints, some modification is still 

considered a management option (Gottfried and others 

1995; Gottfried and Pieper 2000). These authors suggest 

that clearing small, dispersed areas can create a mosaic 

of habitats that would increase diversity and abundance 

of large and small mammalian wildlife and birds, as 

well as increase forage production for livestock. Open 

areas should be designed so that the longest distance an 

animal could be seen across would be 182 m and continuous corridors of adequate width be maintained. In 

some cases, thinning and seeding with native understory 

species has increased ground cover and reduced soil 

erosion losses (Chong 1994, cited in Loftin and others 

1995). Grazing should be deferred two growing seasons 

after either thinning, conversion, or seeding to allow 

plants to establish (Gottfried and others 1995).

Very little work has been done concerning different grazing systems on pinyon-juniper communities. 

However, given the potential in some of the communities for thresholds in erosion mediated by plant cover, 
28 USDA Forest Service RMRS-GTR-169. 2006.

the positive feedback mechanism for continued erosion 

once the threshold is passed regardless of a grazing factor, and the observations that some grazing systems that 

concentrate animals even for short periods can increase 

erosional losses (see soils section above), it would seem 

prudent to base all grazing management on maintaining 

critical cover at all times, regardless of the duration or 

frequency of rest periods following the grazing. Pointin-time grazing intensity maximums should be tightly 

regulated even if long periods of rest or deferment are 

scheduled in the future.

Several grazing intensity guidelines have been proposed for pinyon-juniper communities. Reynolds (1962) 

recommended 3 to 10 acres/AUM, depending on site 

condition and potential. Holechek and Pieper (1992) 

recommended 30 to 40 percent use of key species as a 

moderate intensity, with the high end being for sites in 

good condition or grazed during the dormant season. 

Holechek and Galt (2000) combine the shortgrass and 

the pinyon-juniper type together and use some guidelines developed by Bement (1969) for shortgrass steppe 

in the eastern Colorado Great Plains. Moderate in this 

case is 41 to 50 percent utilization, which is similar to 

moderate for shortgrass steppe (Klipple and Costello 

1960; Bement 1969). A stubble height of 1.5 to 2 inches 

is an indicator of this level of use. However, there can 

be a major difference in the response of the dominant 

species in these two systems. Blue grama increases with 

even heavy grazing in the shortgrass steppe and in other 

grasslands of the Great Plains (Milchunas and others 

1989; Lauenroth and others 1994), resulting in greater 

cover of vegetation and exploitation of soil volumes 

(Milchunas and Lauenroth 1989). The cover of blue 

grama and of total vegetation in some pinyon-juniper 

communities can decrease with grazing (see sections 

above). While Holechek and Galt (2000) state that observations of blue grama residues in New Mexico blue 

grama rangelands match those developed for shortgrass 

steppe by Bement (1969), differences in behavior of the 

dominant species in some pinyon-juniper communities 

should raise caution in applying blue grama stubble 

height as a guide and using similar estimates of what 

“moderate grazing” entails. Holechek and Galt (2000) 

provide guidelines for conservative and light to nonuse grazing intensity categories representing 31 to 40 

percent and 0 to 30 percent utilization. Their “conservative” level would be similar to a lower range moderate 

in shortgrass steppe and may better apply as moderate 

in those particular communities where vegetative cover 

declines with grazing. Paulsen (1975) states that most 

range managers consider use of more than 40 percent 

of forage grasses to be detrimental, and he recommends 

leaving one-third of the culms ungrazed at the end of the 

season. This is similar to Holechek and Galt’s (2000) 

conservative level of grazing, where one-third to onehalf of primary forage species show use on key areas.

Humphrey (1955), Paulsen (1975), and Springfield 

(1976) also recommended resting areas that are in poor 

condition every summer, fair condition range every 

other summer, and good condition range every third 

summer. Springfield (1976) indicates that some system 

of seasonal use in conjunction with deferment is preferable to year-long grazing. The grazing system should 

allow for rest for both cool- and warm-season grasses, 

and guidelines for timing to accomplish this are provided in Springfield (1976). This author reviewed the 

potentials for using controlled burning as a management tool.

Humphrey (1955) provides indicator species for good 

condition range as black, blue, and sideoats gramas, 

Indian ricegrass, junegrass, western wheatgrass, and 

galleta, and those for poor condition as ring muhly, 

blue grama, sand dropseed, snakeweed, pingue, and 

rabbitbrush.

Madrean Woodland_________

Precipitation in these communities exceeds 400 mm/

yr, of which half to two-thirds falls during the growing season (Brown 1982; Ffolliott and Guertin 1987; 

Gottfried and others 1994). Ffolliott (1999a) gives 

general ranges of 300 to 560 mm/yr, and up to 800 

mm/yr in some locations. There are 12 habitat types in 

the southwest, and they generally occur on thin soils, 

rocky slopes, and along drainages (three major types 

are described in the Plant Community Descriptions 

Appendix 1). This characteristic could make them susceptible to erosion when grazed. Because the dominant 

trees are hardwood oaks, important uses of these communities includes fuelwood and fenceposts. However, 

silva-pastoral management is considered of particular 

importance in these woodlands. The history of land-use 

in these communities is similar to that of the pinyonjuniper type, but some areas were settled somewhat 

later due to their more mountainous terrain (Ffolliott 

1999a). McClaran and others (1992) reviewed the history of settlement in this plant community and came to 

a similar conclusion concerning terrain, but they also 

concluded that Indian attacks prevented ranch establishment until chief Geronimo surrendered in 1886. Nearly 

three-quarters of these communities are part of the 

Coronado National Forest and were closed to settlement 

between 1902 and 1910. Most grazing occurred in lower 
USDA Forest Service RMRS-GTR-169. 2006. 29

desert grasslands until the mid to late 1930s, when state 

and federal programs developed stock-watering facilities and fencing.

Extremely little is known of the effects of grazing 

by domestic livestock in Madrean woodland. There are 

few controlled quantitative experiments and even very 

little qualitative descriptive information. Little is even 

known in terms of silviculture practices (Gottfried and 

others 1994). Large-scale range improvement practices 

have generally not been undertaken in these communities (Ffolliott and Guertin 1987). However, grazing 

is an important land-use (McClaran and others 1992; 

Ffolliott 1999b), and Humphrey (1960) considered these 

communities as providing “some of the best grazing in 

the state.” Brown (1982) considered it the “principal biotic community for the white-tailed deer.” An estimate 

of herbaceous production may be obtained from Caprio 

and Zwolinski’s (1994) data of litter plus plant biomass, 

because they sampled 10 months after burning (litter 

removed and plants grew for 10 months). These data indicate levels of herbaceous production of 108, 150, and 

136 g/m2/yr on south, east, and north slopes, respectively. 

Standing biomass of herbaceous understory can be very 

large; Haworth and McPherson (1994) gives estimates 

of from 800 g/m2 within tree canopies, more than 1200 

g/m2 2 m from canopy edge, to over 1500 g/m2 in interstices between trees. Natural Resource Conservation 

Service gives estimates of production from 22 to 34 g/

m2/yr in mountainous sites to 112 to 280 g/m2/yr in deep 

sandy loam soils of only 1 to 15 percent slopes and 20 

percent canopy cover (Soil Conservation Service 1988 

cited in McClaran and others 1992). Humphrey (1960) 

comments that this is a relatively productive range with 

adequate precipitation to support a variety of grass and 

shrub forage plants. Carrying capacity of range in excellent condition is about 10 to 15 animal units per section 

year-long.

While thinning overstory increases understory 

production in many systems, the effects of thinning 

oaks for wild or domestic herbivore forage may be 

small in this system (Haworth and McPherson 1994). 

McPherson (1993) reported increased forage production 

of 10 to 20 g/m2/yr with canopy removal. Increases are 

short-lived due to rapid resprouting of oaks (McClaran 

and others 1992). Regeneration by sprouting after cutting can be highly variable, and reproduction by seed 

is episodic, whereby sustainability of coppicing is not 

known (Gottfried and others 1994). Extremely dense 

stands can hinder animal movement, and thinning of 

stands has been attempted to improve wildlife and cattle habitat (Clary and Tiedemann 1992). Animal weight 

gains increased 60 percent after clearing of oaks in 

southern Colorado (Marquiss 1972). Changes in species 

composition with cutting of Emory oak overstory were 

also small (Haworth and McPherson 1994). Ffolliott 

(2002) recommended that openings to improve livestock forage should be less than 15 to 200 m wide, but 

that these could improve insect abundance for wild turkey populations. Cutting of oaks in these communities 

is not always necessary to open stands, because tree 

densities naturally vary from few, scattered trees to over 

150/ha, and grasses and forbs occur in savannah-like 

mosaics (Ffolliott 1985, cited in Ffolliott and Guertin 

1987; Ffolliott 1999b). Herbaceous species canopy cover at a savannah-like site averaged about 60 percent, and 

grasses contributed a substantial proportion (Caprio and 

Zwolinski 1994).

Ffolliott (1999b) suggested that grazing may favor oak 

seedling establishment if competition with herbaceous 

species is reduced, but Phillips (1912, cited in Ffolliott 

1999b) and Bahre (1977) suggested that poor regeneration from seed was due to livestock grazing. McClaran 

and others (1992) states there is no obvious difference 

between oak recruitment in ungrazed exclosures compared to adjacent grazed areas, and indicates this is 

different from the situation in less productive California 

oak communities. Humphrey (1960) also mentions that 

there was more grass and fewer trees in the past than 

there is today, but this could be due directly to fire thinning of young trees (McPherson 1992). Bahre (1977) 

and McPherson (1992) think tree density has decreased 

over the last century. In general however, fire favors 

oaks, but use for fuelwood, fenceposts and other types 

of construction can have impacts on their abundance 

(McPherson 1992). McPherson (1993) showed that 

seedling mortality was directly linked to belowground 

competition from other plants. Defoliation of plants 

growing in the field was a major cause of mortality, but 

defoliation in growth chambers was not. The majority 

of mortality due to defoliation was from invertebrates. 

Humphrey (1960) suggests that junipers have increased 

on some sites, but not as much as in other communities with junipers present. There does not appear to be 

a consensus in opinion concerning tree densities condition in these communities, let alone the direct and/or 

indirect effect of grazing. Cattle are generally not large 

consumers of the mature oaks themselves, although 

some consumption of Emory and silverleaf may occur 

in winter (McClaran and others 1992).

In some cases the increase in pinyon and junipers 

has been reported to be substantial. Tolisano (1995) 

indicate an increase from 20 to 40 percent, with decreases in Madrean woodland communities over a nearly 

30 year period in Apache National Forest. Some 
30 USDA Forest Service RMRS-GTR-169. 2006.

quantitative data on range condition were collected between 1962 and 1968 and again in 1993 to 94. 

Intermediately, cattle were removed from one grazing 

allotment in 1983. A comparison of the 10 year ungrazed and adjacent grazed allotments indicated some 

recovery of tall bunch grasses and cool-season grasses 

and increased grass and forb diversity for one-third of 

the study plots (Tolisano 1995). One-third of the plots 

showed no trend in recovery, and one-third showed 

some declines in cover. The author indicated that some 

deterioration of the range may have occurred between 

the time of the initial sampling and the removal of cattle 

but concluded that a century of rest from grazing had 

not resulted in large improvement, suggesting a very 

slow recovery that may take centuries.

Humphrey (1960) described range condition classes 

as follows: Excellent - perennial grasses make up 90 

percent of plant cover; plants are “vigorous” and there is 

no erosion. Good - understory is primarily good-forage 

grasses and low-shrubs, but there are more oaks and bare 

ground present, which could be improved by thinning of 

oaks with no reduction in stocking rate. Fair - density of 

perennial grasses is low and plants are small, producing 

few seeds, and there are patches of bare ground. Poor 

- forage production is 25 percent of range in excellent 

condition, perennial grasses are replaced by annual 

grasses and weeds (colonizing ruderals) and unpalatable 

shrubs, and erosion is evident and can be severe. Ranges 

in poor condition should be left ungrazed for 3 to 5 years, 

or grazing should at least be discontinued during July 

through September with grazing intensities reaching no 

more than 50 percent by this time. Carrying capacity of 

cows for range in poor condition is reduced by a quarter. 

Holechek and Pieper (1992) recommend 30 to 40 percent 

utilization of key forage species as a moderate grazing 

intensity. Brown (1982) also commented on bare ground 

as an indicator of very heavy grazing of these communities, although Ffolliott and Guertin (1987) suggest 

similar or lower erosion potentials in these systems as 

for pinyon-juniper communities, based on observation. 

The characteristic steep slopes and thin soils on which 

these communities are commonly found would suggest 

high erosion potentials should soil become unprotected 

due to loss of vegetative cover.

Actual grazing intensities reported in 1991 on the 

150 allotments in Coronado National Forest averaged 

35 to 55 percent utilization of herbaceous understory 

vegetation (McClaran and others 1992). This was 3.8 

acres/cow/month, or 2.5 acres/cow/month if unsuitable 

steep slopes were removed from the grazable area. Yearlong or seasonal grazing was the most common grazing 

practice at this time, although two or three-pasture 

rest-rotation systems were becoming more common. 

There are no data to base conclusions on as to whether 

the grazing systems improve range condition or not. 

Ffolliott (2002) suggested a maximum forage utilization 

of 35 percent in areas managed for Merriam’s turkeys 

and that grazing be deferred until September 1 in high 

density brood habitat.

Grazing can reduce fuel loads in terms of herbaceous 

canopy and litter biomass and can affect site productivity, thereby influencing fire regimes. There is evidence 

that oak woodlands were historically a part of a fire 

chain, whereby fires started in lower elevation grasslands 

spread through oak woodlands and into higher elevation mixed-conifer forests (Bahre 1991). In this case, 

heavy grazing of fuel loads and fire suppression in lower elevation grassland affected fire regimes of neighbor 

communities. However, McPherson (1992) and Bennett 

and Kunzmann (1992) comment that the absence of reliable data make it difficult to estimate fire frequencies. 

However, Swetnam and others (1992) conducted firescar studies in southeastern Arizona oak-pine woodland 

and estimated a mean fire return interval of 3.9 years for 

any one tree, and 13.2 years for fires scarring 25 percent 

of the trees. They concluded that surface fires occurred 

every 1 to 38 years up until 1890 when livestock grazing 

became an important landuse. Caprio and Zwolinski’s 

(1994) study of fire effects in Madrean woodland show 

large effects on species composition. Woody species 

were reduced and grasses and forbs increased after an 

initial depression. Oak species resprouted well after fire 

but had about a 15 percent mortality. The rapid accumulation of fuels in these communities suggest a potential 

short interval between fires of 1 to 2 years prior to fire 

suppression and grazing by livestock.

There are no controlled grazing studies to base any 

conclusions concerning effects of livestock in these 

communities. However, the qualitatively described 

switch from perennial grass dominated vegetation to annuals with heavy grazing pressure, and the potential for 

large areas of bare ground to develop and erode, would 

suggest this is not a system well adapted to heavy grazing. Productive communities with a long evolutionary 

history of grazing can display a rapid switching from 

perennial tall grasses to perennial short “sod grasses” 

with changes in grazing pressure over short periods of 

time, and semiarid communities with a long history of 

grazing can display increases in basal cover of prostrate, 

grazing tolerant, perennial short grasses (Milchunas and 

others 1988). While blue grama is a dominant in the latter, nearby shortgrass steppe communities and can be a 

component of Madrean woodland communities as well, 

it is interesting to note that genotypes of blue grama 
USDA Forest Service RMRS-GTR-169. 2006. 31

here may not respond to grazing as they do in shortgrass steppe (Milchunas and others 1989; Milchunas 

and Lauenroth 1989) or blue grama may not be present 

in sufficient quantities to detect a response at the community level.

Great Basin JuniperSagebrush__________________

Some of the general literature overlaps when differentiating between pinyon-juniper woodlands and Great 

Basin juniper-sagebrush community types, and between 

Great Basin juniper-sagebrush and Great Basin sagebrush shrubsteppe, or between Great Basin sagebrush 

shrubsteppe and Intermountain sagebrush steppe. In the 

case of the former, sagebrush is an important component 

of one but not the other, but pinyon-juniper is sometimes 

applied regardless of understory components. Schmutz 

and others (1967) considered their study site “borderline between the sagebrush northern desert shrub and 

the pinyon-juniper,” although the site is dominated by 

sagebrush. In the case of the middle, relative amounts 

of the overstory can determine the classification, and 

juniper and pinyon can invade Great Basin sagebrush 

shrubsteppe (Ellison 1960; Rowlands and Brian 2001). 

Community analysis of a Great Basin juniper-sagebrush 

and a Great Basin sagebrush shrubsteppe suggested that 

the classification is based primarily on just the absence/

presence of pinyon and juniper. In the case of Great Basin 

sagebrush shrubsteppe and Intermountain sagebrush 

steppe, West (1983a) states that, because of structural 

similarities, the two types have often been combined in 

reviews, but the more arid Great Basin sagebrush shrubsteppe reacts to disturbance more like a desert and the 

Intermountain sagebrush steppe more like a semi-arid 

grassland. However, because of these gradients and 

some similarities between pinyon-juniper, Great Basin 

juniper-sagebrush, and Great Basin sagebrush shrubsteppe, and because of lack of abundant data for all, 

cross referencing between the three types covered in 

this review and outside reviews of Intermountain sagebrush steppe may be helpful to the reader. Specific to 

this review, abiotic conditions, primary productivity, and 

grazing management for this community are covered in 

the general references contained in the pinyon-juniper 

section above or the Great Basin sagebrush shrubsteppe 

section below.

There are only four studies of grazing effects in this 

community type at three locations. Jameson and others 

(1962) studied a large 1,430 acre mesa in northwestern 

Arizona in 1958, and Rowlands and Brian (2001) returned to the same mesa in 1996. The mesa is grazed 

by deer but is inaccessible to livestock. Jameson and 

others noted that fires had occurred long ago, that trees 

would dominate nearly 90 percent of the area were it not 

for fires, and that burned areas were not rapidly reinvaded by trees. Pinyon density was increasing some, as 

evidenced by the presence of seedlings and young trees. 

However, junipers were mature and even-aged. This indicates a direct effect of fire on the tree component and 

suggests that the lack of grazing may slow invasion by 

juniper but not necessarily pinyon. Deer grazing was 

“appreciable” but generally did not cause severe damage 

to the vegetation. Muttongrass was the dominant grass 

of the mesa but was not reported in studies on the mainland. In contrast, blue grama was the dominant on the 

mainland but was in very low abundance on the mesa 

except on hard shallow soils. Rowland and Brian’s intensive sampling revealed that the cover of most species 

had remained stable over the 38 year interval since the 

Jameson and others study. However, a few changes were 

noted. Muttongrass had increased, while Torrey jointfir, prickly pear cactus, and snakeweed had decreased. 

Sagebrush, juniper, and pinyon pine showed no change.

Daddy and others (1988) compared an area protected 

from grazing for 21 years with moderately and heavily 

grazed areas in northwestern New Mexico. The measurements were quantitative, but statistical power was 

poor, whereby differences reported as not significant 

may be due to high variance. Excluding large herbivores 

did not retard woody plant growth. Big sagebrush did 

not differ between treatments but appeared more “luxuriant” in the exclosure. West (1983a) also commented 

that sagebrush dominates Great Basin sagebrush shrubsteppe communities even when they are pristine or 

near-pristine sites. Daddy and others do not give data 

for juniper. Total herbaceous cover and biomass were 

greatest in the moderately grazed treatment, primarily due to large amounts of galleta. Because the sites 

were not grazed during the year of study, end of season biomass values can be used as estimates of primary 

production. Protection for 21 years did not result in 

increases in production, as values were 19, 50, and 25 g/

m2/yr for the exclosure, moderately, and heavily grazed 

treatments, respectively. Precipitation at this site is only 

200 mm/yr. Blue grama tended to increase with increasing grazing intensity, but differences were not significant. 

The heavily grazed treatment had a large proportion 

of opportunistic ruderals (three-awns) and exotics 

(cheatgrass, Bromus tectorum). Species diversity was 

greatest in the moderately grazed treatment and similarly lower in the heavily grazed and ungrazed treatments. 
32 USDA Forest Service RMRS-GTR-169. 2006.

There were no differences among grazing treatments in 

root biomass. However, root distribution was deeper in 

the exclosure, where only 40 percent of the biomass was 

in the 0 to 40 cm depth compared to 80 percent in the 

grazed treatments. Soil water content was lowest in the 

heavily grazed treatment and highest in the moderately 

grazed. The heavily grazed treatment was driest near 

the surface.

Pinyon-juniper communities in Grand Canyon 

National Park were heavily impacted by feral burro 

grazing. Bennett and others (1980) noted reductions in 

palatable perennial grasses, seed production, total plant 

cover and density, and plant diversity in both richness 

and evenness components.

These very limited number of studies suggest large 

effects of protection from grazing on species composition 

but little effect on recovery of production. West (1983a) 

also commented that damage to the somewhat similar 

Great Basin sagebrush shrubsteppe is much less 

reversible, and manipulations less successful, than for 

the more mesic Intermountain sagebrush steppe.

Interior Chaparral___________

Chaparral of the southwest is located almost exclusively in a discontinuous strip across central Arizona and 

occupies only 13,000 km2 (Cable 1975). Mean annual 

precipitation averages about 480 mm/yr in the center, 

with a range from 480 to 635 mm/yr. Approximately 

half the precipitation falls from May through October. 

Chaparral is found on coarse soils, particularly on steep 

slopes, where precipitation is equally divided between 

summer and winter (McGinnies and others 1971, cited 

in Cable 1975). Heavier soils are occupied by grassland, and grazing can favor invasion of chaparral into 

grassland. Oak woodland occurs under similar environmental conditions but where there is more precipitation. 

Shrub live oak is from 45 to 80 percent of the total shrub 

cover, which varies from 25 to 80 percent with increasing elevation and precipitation Cable 1975). Herbaceous 

understory species are sparse where the overstory cover 

is high. As with the pinyon-juniper type, there have 

been some attempts to thin overstories in order to increase forage production (Reynolds 1962, Cable 1975). 

Root-plowing is most effective and fire or herbicides 

are not effective means of shrub control. In the past, 

exotic species such as Lehmann lovegrass or crested 

wheatgrass have sometimes been seeded after shrub 

control efforts. Productivity of the herbaceous understory can increase from near zero when shrub cover is 

30 percent or more to about 200 g/m2/yr where shrub 

cover is 5 percent or less (Pond 1961, cited in Reynolds 

1962). Reynolds (1962) list the more important uses of 

chaparral as wildlife habitat and rates this community 

very low for range production.

The history of grazing by domestic livestock is a little different for the chaparral than for other community 

types because of the rough topography (Cable 1975). 

The first cattle were brought in 1874, and most of the 

chaparral was stocked within 10 years. Early reports 

were that the shrub stands were quite open with productive grass cover between (Croxen 1926 cited in Cable 

1975). Livestock numbers peaked in 1900 at about 15 

to 20 times the number present in 1926 when numbers 

were reduced to the approximate carrying capacity. By 

1926 chaparral was dense shrub stands with only a small 

proportion of the previous grass production. McGinnes 

and others (1941, cited in Cable 1975) also commented 

on the change from open shrub with good grass cover to 

dense shrub with little herbaceous cover over 30 years 

due to fire and excessive grazing. As a result, there has 

been invasion of this community type into adjacent 

types and loss of perennial grasses such as blue, black, 

hairy and sideoats gramas, dropseeds, threeawns, curly 

mesquite, bluestems, and wolftail. Grazing of chaparral 

today is primarily by cattle, but goats were an important 

component as late as 1942 (Cable 1975). Rigden and 

Parker in 1943 commented that as many as 250 goats 

and 10 cows per section (259 ha) grazed many chaparral areas year-long, and that this intensity of grazing 

resulted in mortality of perennial grasses and was not 

sustainable. Goats are browsers while cattle are generalist grazers. Dual grazing by goats with cattle would 

tend to balance grazing pressure, with goats utilizing 

more of the shrubby component. The two-animal system was economically beneficial (Cable 1975).

Grazing of goats in chaparral can, however, have 

some undesirable consequences as well. Severson and 

Debano (1991) studied four stocking levels of goats in a 

short duration grazing system for four and a half years. 

Total shrub cover was lowered by goat grazing at heavy 

(38 percent cover), moderate (39 percent), and light 

(35 percent) stocking densities compared to unstocked 

controls (51 percent cover). A shrub-crushing treatment further increased the effectiveness of the goats. 

However, shrub species not preferred by the goats were 

not affected, and neither were perennial understory species. Annual grasses increased due to soil disturbance 

by either the hoof action of the goats or the mechanical 

crusher. Lower amounts of litter were found between 

shrubs, and this translated to lower soil organic matter levels. Concentrations of N and P were also lowered 
USDA Forest Service RMRS-GTR-169. 2006. 33

under some shrub species and soil bulk densities increased. Shrubs preferred by goats were also those 

preferred by deer, and heavy browsing of nitrogen-fixing shrubs could eliminate these, lower forage diversity, 

and possibly lead to increased nutritional stress for wildlife. The lower litter accumulation and reduced tissue 

quality was thought to have potential long-term effects 

on stability of the system by affecting decomposition 

processes. This and two other goat browsing studies in 

Arizona suggest limited success in using goats to control shrubs in chaparral. Goat browsing of gambel oak 

was found to not be a useful control measure in ponderosa stands in southern Colorado unless combined 

with other mechanical treatments, suggesting that this 

may apply to other communities as well. Shrubs are a 

very important component of deer diets in chaparral, 

but little information exists on their impact on the plant 

community (Urness 1981; Wallmo and others 1981).

Rigden and Parker (1943) emphasized the need to 

balance carefully the numbers of goats and cattle in 

relation to the proportions of shrub and herbaceous 

forage. They also conducted one of the few exclosure 

experiments in chaparral, for five years on two ranches 

grazing both cattle and goats and for two years on an 

additional goat only ranch. Perennial grasses increased 

with time of protection from grazing. Half-shrubs also 

increased with protection, but the ruderal half-shrub 

snakeweed decreased slightly even though it increased 

over time in the grazed treatments. Browse increased 

slightly under protection from grazing only at one site. 

While very heavy grazing can kill shrub species, the 

usual impact under less intense grazing is to gradually 

shift species composition from a mix of palatable/unpalatable species towards a dominance of unpalatable 

species. Less palatable species include scrub live oak 

and manzanita, and palatable shrubs include cliffrose 

and desert ceanothus. Rigden and Parker (1943) also 

mention that heavy grazing and brush fires reduce the 

litter layer and soil organic matter, thereby resulting 

in dryer conditions that will not support a good grass 

cover. Grasses dependent on heavy summer rains that 

come as large events. These events produce more runoff 

if a good litter cover is not present. In contrast, deep 

rooted shrubs are favored by winter precipitation that 

comes as small events but that accumulates through 

time deep into the soil profile.

Rich and Reynolds (1963) compared watersheds 

moderately grazed (40 percent utilization) and heavily 

grazed (80 percent utilization) to sites ungrazed for 9 to 

12 years. Perennial grasses and half-shrubs declined in 

composition with heavy grazing, cacti increased, and 

shrubs did not appear to respond greatly. The moderately 

grazed watershed was generally similar to ungrazed 

sites, with perennial grass and half-shrub composition 

slightly higher than the ungrazed sites. Data were presented in terms of composition as a percent of total, so 

actual quantities are not known from the data presented.

A range condition classification for chaparral was developed by Rigden and Parker (1943) as follows:

“Good. Perennial vegetation cover 1/3 to 1⁄2 of 

ground surface; blue, hairy, sideoats, and black 

gramas compose 20 percent of all vegetation; 

shrubby buckwheat and menodora plentiful; 

shrubs compose about 70 percent of all vegetation; 

most palatable species such as Ceanothus, Garrya, 

and mountain-mahogany in healthy growth; use 

on scrub oak light.

Fair or Unsatisfactory. Perennial vegetation usually covers 1/5 to 1/3 of ground surface. Grama 

grasses compose less than 10 percent of vegetation but increase of 3-awn grasses may raise total 

to 20 percent; half-shrubs such as snakeweed increasing; may comprise 20 percent of vegetation. 

Palatable shrubs disappearing from range.

Poor or deteriorated. Vegetation usually covers 

less than 1/5 of ground surface. Perennial grasses 

including 3-awns absent or confined to protection 

of shrubs and rocks. Soil unstable, erosion active, 

no litter. Increase in snakeweed and pricklypear. 

Hillsides covered with prominent pattern of crisscross livestock trails and trampling excessive. 

Shrubs dying out with many dead and half-dead 

plants.

Spring annuals such as filaree will occur in all 

condition classes but not dependable every year 

for forage.”

Humphrey (1960) also described range condition 

classes for chaparral in the somewhat more productive 

southern ranges. In addition to the generalizations given 

by Rigden and Parker (1943) above, Humphrey suggests 

burning as a means of killing manzanita while leaving 

large oaks for shade. This author also suggests stocking 

steep slopes at lighter rates and very steep slopes left 

ungrazed. Ranges in poor condition should be left ungrazed for a two to three year period, or at least deferred 

during July, August, and September leaving at least 50 

percent of the grass production as residual.

A rotation-deferred grazing system was described by 

Freeman (1961) whereby one or more pastures are rested during the growing season and use of other pastures 

is rotated. Although there were no data, the grazing 

system was reported to have improved range condition 

and grass production. Salting a good distance away from 

•

•

•

•
34 USDA Forest Service RMRS-GTR-169. 2006.

watering locations was an additional important aspect in 

recovery of especially degraded areas near water. Yearlong grazing is often practiced in the chaparral because 

shrubs provide some winter forage and grasses summer 

forage, and because weather is mild throughout the year 

(Cable 1975). Suggested grazing intensities range from 

40 percent utilization of perennial grasses to 30 to 35 

percent of current-year twig growth of shrubs.

Mule deer are the common deer species of chaparral, but Coues white-tailed deer can be locally abundant 

(McCulloch 1972, cited in Wallmo and others 1981). 

Wallmo and others (1981) reviewed literature on densities of deer in chaparral and reported 4 to 5 deer/km2

usual, with local densities that reached 20 to 30 deer/

km2. All shrubs except possibly pringle manzanita were 

found in diets, but effects on the plant community are 

generally not known.

Runoff and erosion on ungrazed (9 to 12 yrs) and 

moderately and heavily grazed watersheds were estimated by Rich and Reynolds (1963). No significant 

effects of grazing were found for water yield, and differences among treatment means were small. Sediment 

trapped was so variable that differences among grazing treatments were not significant. Slopes in the study 

watersheds were moderate. The authors concluded that 

reductions in densities of perennial grasses must be 

severe in order to change runoff from subsurface to surface phenomena. However, Humphrey (1960), Freeman 

(1961), and Rigden and Parker (1943) mention grazinginduced erosion in relation to range condition. Rigden 

and Parker (1943) even use the terms “erosional pavement 

is pronounced” and “grasses are distinctly pedestaled” 

in relation to poor or deteriorated range condition class. 

Water yields generally increase following conversion 

from shrub to grass (Cable 1975). Increases are attributed to reduced transpiration from deep rooted shrubs 

compared to shallow rooted grasses. Erosion following 

fire that removes the shrub and litter cover can be extensive, but sediment movement declines sharply as cover 

recovers and returns to the normal small amounts about 

four years after the fire (Pase and Ingebo 1965, cited in 

Cable 1975).

The role of fire and grazing by fire interactions in 

chaparral may not be as important as in many other 

southwestern communities. Shrub cover in chaparral 

recovers rapidly after fire with only minor changes due 

to prolific seedling establishment of fire-scarified seed 

or by sprouting (Paulsen 1975; Cable 1975). Cable suggests that fire is an advantage to the shrub component 

after a short period when grasses and forbs are favored. 

Knipe (1983) noted that the only shrubs killed by fire 

were manzanita and desert ceanothus, but that fire 

stimulated germination of their seeds. Because other 

means of shrub conversion, such as root plowing, are 

not feasible over much of the rugged terrain and rocky 

soils this community occurs on, Knipe (1983) assessed 

the possibility of using goat browsing following burning and seeding of range to trample seeds into the soil 

and utilize new shrub sprouts until the grasses became 

established. The system of fire and timed grazing was 

more successful than fire alone. Trampling of the seed 

into the soil and the use of a wider range of vegetation 

than that used by livestock and wildlife were considered 

positive aspects of the treatment, although heaviest use 

was of species preferred by cattle, deer, and elk. Desert 

bighorn sheep may act similarly, but numbers may not 

be great enough to have much of an impact. Krausman 

and others (1989) found that diets of desert bighorn 

sheep contained a large browse component, and that 

diet overlap was insignificant among the bighorn sheep, 

desert mule deer, and feral burros.

Rigden and Parker (1943) made a very important point 

concerning the balance between the way a shift in the 

half summer and half winter precipitation pattern could 

favor grasses versus shrubs respectively. Mediterranean 

climates favor deep rooted shrubs due to over-winter 

storage of soil water that can accumulate deep in the 

soil profile in the absence of large surface-evaporative 

and surface transpirational demands. Annuals are also 

favored in this type of climate because they can rapidly 

utilize stored water in spring before the perennials are 

active. Severson and DeBano’s (1991) observation of 

increased annuals with grazing may not have been due 

to hoof action as suggested, but could possibly be due 

to shifts in temporal dynamics of soil water availability. 

Grazing in these systems can shift the balance, or they 

can create an alternate stable point if a successful shrub 

or exotic annual invades. This is evident in the case 

of cheatgrass and the conversion of California perennial grasslands to an annual dominated system. Both 

Rigden and Parker (1943) and Severson and DeBano 

(1991) consider litter and soil organic matter to control 

the balance in chaparral through their effects on soil 

water dynamics. Litter cover can rapidly be recovered 

with protection from or reductions in grazing pressure, 

but soil organic matter has fast, intermediate, and recalcitrant pools (Burke and others 1997). Recalcitrant 

pools can have turnover times in the hundreds of years, 

and loss of this pool through erosional processes could 

lead to an alternate stable point in ecological time.

There are insufficient studies in the chaparral to 

suggest whether a shiftable balance exists or whether 

an alternate state has been reached. Rigden and Parker’s 

(1943) results show perennial grass recovery is possible, 
USDA Forest Service RMRS-GTR-169. 2006. 35

whereas Severson and DeBano (1991) results suggest 

that perennial grass may not respond. The latter is 

unclear since the consumption of some grass by goats 

may have countered the effects of shrub overstory 

reduction.

Subalpine Grassland_________

Subalpine grassland is associated with and interspersed through boreal forest with 500 to 1000 mm/yr 

precipitation, of which 150 to 250 mm/yr falls in the 

summer (Paulsen 1975; Turner and Paulsen 1976). 

Most of this community type occurs in Colorado, small 

amounts in New Mexico, and only a few isolated locations in Arizona. Herbaceous primary production 

ranges from 115 to 225 g/m2/yr and can reach as high 

as 335 g/m2/yr. There is very high variability in production from year to year within a site, and the grass 

component can vary from 35 to 62 percent of total 

vegetation. Along with the patchy distribution of the 

grassland within the forest type (see section above), 

this potentially makes it difficult to adjust stocking 

rates to available forage without grazing the community more heavily than desired.

Paulsen (1975) and Turner and Paulsen (1976) describe range condition changes associated with livestock 

grazing. Thurber fescue is an important component 

of the community, is sensitive to grazing, and may be 

replaced by other small bunchgrasses. Forbs and secondary grasses increase as Thurber fescue and Parry 

oatgrass decline. Unpalatable forbs and shrubs are more 

likely to replace palatable grasses with cattle grazing, 

whereas sheep grazing is more likely to result in a shift 

from palatable forbs to grasses. Total cover may remain similar in this early stage of range deterioration, 

and primary production and litter may decrease only 

slightly. In poor condition range, bare soil and erosion 

become more evident, and species composition becomes 

forb and shrub dominated. Costello (1944) observed 

Thurber fescue becoming the dominant species within 

aspen stands after prolonged heavy grazing eliminated preferred forbs and grasses. Moir (1967) suggested 

Kentucky bluegrass replaced Thurber fescue after heavy 

grazing. After protection from grazing, Paulsen (1970) 

observed increases through time in slender wheatgrass 

and butterweed groundsel and decreases in Letterman 

needlegrass and Fremont geranium. Based on this study, 

however, Turner and Paulsen (1976) suggested that 

protection from grazing for eight years had little effect 

on grass-forb composition.

Several studies of grazing effects in subalpine grassland have been conducted. A study in western Colorado 

conducted by Klemmedson (1956) must be interpreted 

with extreme caution. Study sites were chosen by following a “Range Condition and Site Guide” based on a 

listing of species thought to be increasers, decreasers, 

and invaders with grazing, and then various aspects of 

the vegetation and soil were measured. No indication is 

given of actual previous grazing intensities on the different sites. Studies such as this may be of limited use 

to confirm accuracy of condition guides based on species composition if soil type remained constant among 

sites, but this information is not provided either. Total 

vegetative cover declined from 48 percent for ‘good’ 

condition range to 42 percent for fair and 24 percent 

for poor, possibly confirming the use of Thurber fescue, 

Carex species, nodding brome, and sheep fescue as “desirable” species and a number of forbs and half-shrubs 

as “undesirable.” Soil organic matter, degree of plant 

pedestaling, some erosion indices, and infiltration were 

also measured, all of which followed the predetermined 

condition classification.

In a similar situation again in western Colorado, 

Turner and Dortignac (1954) chose “six common plantcover types which are readily recognizable by those 

directly concerned with the management of mountain 

grasslands.” No indication is given of previous grazing 

histories of the sites or of soil types. The potential for 

different responses to grazing on different soil types 

is recognized. The authors’ observations indicate that 

Thurber fescue is commonly replaced by Idaho fescue 

and needlegrasses with heavy grazing on coarse-textured soils, and by Kentucky bluegrass on fine-textured 

soils. The six community types studied were dense 

Thurber fescue, open Thurber fescue, bluegrass, needlegrass, lush-weed, and poor-weed. Over this gradient of 

community types, grass biomass declined from 228 to 

21 g/m2, while opportunistic ruderals increased from 

48 to 125 g/m2. Litter declined over the community 

gradient from 1148 to 57 g/m2, while bare ground cover 

increased from 6 to 58 percent. Infiltration was slowest in the needlegrass and the poor-weed community 

and highest in the dense Thurber fescue. Erosion losses 

ranged from 9 g/m2 in dense fescue to 490 g/m2 from 

the poor-weed community.

One of the Potter and Krenetsky (1967) and Krenetsky 

(1971) exclosures in this type showed increases in 

Thurber fescue and Columbia needle-grass after 25 

years. Carex species, Arizona fescue, and Kentucky 

bluegrass declined with protection. Forbs declined with 

protection, while increasing through time on the grazed 

site. Browse increased from 1939 to 1963 on both 
36 USDA Forest Service RMRS-GTR-169. 2006.

grazing treatments. Part of an aspen stand contained 

in the original exclosure increased inside but not outside the fence where sheep grazed, and the aspen stand 

in turn was displaced by a mixed conifer association. 

Overall, however, grass cover and total herbaceous 

cover increased through time more in the grazed area 

than the ungrazed, and this led Krenetsky to conclude 

protection does not speed grass recovery. This site was 

in good condition when the exclosure was built in 1939 

and had been grazed by sheep.

A controlled but complex study of interactions between gophers and cattle grazing, and by a change from 

heavy to light grazing, was conducted on a mesa near 

Grand Junction, Colorado (Turner 1969). The area had 

been heavily grazed to the point where dense stands 

of Thurber fescue that were documented in 1889 had 

been replaced or obscured by sagebrush and orange 

sneezeweed. The area had also become heavily infested 

with pocket gophers. Condition of the site at the start 

of the study was described as very poor. Initial treatments included construction of replicated grazing 

exclosures and reduction in grazing intensity by onehalf, and gopher control or no gopher control. Averaged 

over gopher treatments, grazed sites had beginning and 

19-year-later grass biomass of 14 and 34 g/m2 , compared to 15 and 40 g/m2

 beginning and 19-year-later 

biomass, respectively, for the ungrazed. The switch to 

light grazing had nearly as much of an effect as complete protection for grass recovery. For forbs, grazed 

sites had beginning and 19-year-later biomass of 63 and 

49 g/m2

 compared to beginning and 19-year-later biomass of 67 and 49 g/m2

 for the ungrazed. The switch 

from heavy to light grazing in this case had the same 

effect as complete protection. However, sneezeweed 

production was reduced only half as much when heavy 

grazing was reduced to light grazing compared to heavy 

grazing being reduced to no grazing at all. Shrubs increased slightly on grazed sites and did not change from 

initial values on protected sites, but differences were not 

significant. There was an apparent gopher treatment by 

grazing treatment interaction, whereby grass recovery 

was greatest when gophers were present on ungrazed 

sites and grass recovery was least when gophers were 

controlled on grazed sites. Species composition changes 

were large. The 19-year protection resulted in initial and 

final values of 35 and 18 percent for Letterman needlegrass, 8 to 34 percent for slender wheatgrass, and both 

bromegrass and trisetum increased from 3 to 15 percent each. Forb composition did not change as much as 

grass species, but western yarrow was a decreaser and 

agoseris was an increaser. The improved conditions on 

ungrazed and lightly grazed compared with previous 

heavily grazed conditions were also evidenced by increased flower stalk height for many of the later seral 

stage species. The authors concluded that “deteriorated 

mountain grassland range may improve almost as rapidly under light grazing as under non-use.”

A third study, conducted in western Colorado in 

moderately grazed areas with occasional heavy use, indicated that 8 years of protection had very little effect 

on plant community species composition. Deep rooted 

shrubs, and proportions of grasses and forbs, remained 

generally unchanged (Paulsen 1970). Thurber fescue 

was increased under protection.

Little literature exists concerning grazing management in this community. Paulsen (1975) suggests less 

than 40 to 45 percent utilization of palatable grasses. 

Forbs are more abundant earlier in the growing season 

and grasses later, whereby seasonal grazing can be used 

to vary pressure on these components. Palatable perennial grasses should be headed out, and soil sufficiently 

dry to avoid excessive compaction and trampling before animals are released onto the range in the spring. 

Although conducted in a Wyoming mountain grassland, 

a seven-year comparison of continuous grazing versus 

a rotational grazing system did not show conclusive 

differences (Pond and Smith 1971, cited in Turner and 

Paulsen 1976). Little is known of fire effects on this 

community, although fire tends to maintain grassland 

openings from tree encroachment (Wright 1971, cited 

in Turner and Paulsen 1976). However, this grassland 

is interspersed within the boreal forest type and White 

and Vankat (1993) estimated the natural fire return interval for boreal forest was 70 to 250 years. This is a 

relatively very infrequent fire return interval compared 

to lower elevation communities. Primary production 

may be increased when fire removes large buildups of 

litter in productive sites. Litter can regulate soil temperatures and therefore soil water dynamics, tie up 

nutrients, and provide a positive or negative microsite 

for seed germination depending on particular species 

requirements (Facelli and Pickett 1991). However, the 

effects of fire on primary production may vary with climatic conditions after the fire, whereby production may 

be increased when sufficient moisture is available or decreased when conditions are dry after the a fire (Turner 

and Paulsen 1976).

There is very little information to base general 

conclusions of the effects of grazing on subalpine 

grasslands. Based on qualitative, observational evidence, 

there would appear to be very large differences in 

condition classes ranging from poor to good. Actual 

experimental evidence for improving condition classes 
USDA Forest Service RMRS-GTR-169. 2006. 37

through changes in grazing management must be 

based on only three studies. These would suggest that 

a change in management from heavy grazing to either 

light grazing or no grazing gives similar results when 

initial condition is an early seral stage, and ranges in 

later seral stages initially may not change with complete 

protection.

Montane Meadow___________

This community type is closely associated with and 

interspersed among primarily the ponderosa pine and 

some mixed conifer types. Therefore, much of the information reviewed in the ponderosa pine section above 

is relevant to this community and will not be repeated 

here, particularly concerning overstory-understory relationships and primary production, tree encroachment, 

fire, and grazing management. This section will focus 

only on plant community response directly to grazing 

treatments. Small areas of this grassland among ponderosa pine can occur on the same soil types, but larger 

areas can be on different soil types than the adjacent 

forest (Klemmedson and Smith 1979, cited in Chambers 

and Holthausen 2000).

Species Composition

A shift to short, prostrate growth forms has throughout the world been associated with increased grazing, 

especially in systems with a long evolutionary history 

of grazing (McNaughton 1984; Milchunas and others 1988; Lauenroth and others 1994). Arnold (1955) 

observed similar changes in montane meadows of 

northern Arizona. A difference between the montane 

meadow and systems with a long history of grazing 

arises in the increase in weed (colonizing ruderals) or 

exotic species with grazing in the former (Arnold 1955; 

Johnson 1956) and the absence of weed increases in 

systems such as the shortgrass steppe or Serengeti grasslands (Milchunas and others 1988). Arnold considered 

indicators of deteriorated range condition in montane 

meadow to be closely associated with increasing composition of decreasingly desirable growth-forms from 

perennial tall grasses (redtop, timothy), perennial midgrasses (Arizona fescue, mountain muhly), perennial 

short grasses (black and pine dropseed, blue grama), 

perennial prostrate forbs (pussytoes, sandworts), shortlived unpalatable half-shrubs (snakeweed), to annuals 

(annual dropseed).

Quantitative studies in general somewhat support 

Arnold’s classification, but there is not necessarily a 

linear relationship of increasing grazing intensity and 

the indicators described above. In a controlled replicated study in central Colorado (Manitou Experimental 

Forest), Johnson (1956) sampled 3 replicate blocks 

of meadows ungrazed 10 years, and 9 years of lightly, 

moderately, and heavily grazed treatments. Grazing was 

June through October, and the treatments represented 

utilization of approximately 0, 10 to 20, 30 to 40, and 50 

percent or greater for the ungrazed, lightly, moderately, 

and heavily grazed treatments, respectively. The treatments were sampled for four years, the last of which was 

a drought year. Johnson (1956) observed similar species composition between ungrazed and lightly grazed 

treatments, and similar grass and colonizing ruderal 

compositions among ungrazed, lightly, and moderately 

grazed treatments (Table 1). Species composition on 

heavily grazed treatments differed considerably from the 

other three treatments. Species considered weeds (colonizing ruderals) were approximately 50 percent of the 

heavily grazed community and 25 to 30 percent of the 

other three treatments. For grasses, mountain muhly and 

blue grama dominated composition of the heavily grazed 

treatment while mountain muhly and Arizona fescue 

dominated the exclosures. Density was related to grazing intensity for most species. Sedges and rose pussytoes 

were most abundant in the heavily grazed treatment, and 

sedges and mountain muhly were most abundant on the 

ungrazed treatment (Table 2). Total density of all species was greatest in the heavily grazed treatment, but 

plants were smaller and the greater number may have 

been due to disintegration of originally large clumps. In 

many cases, species considered unpalatable were more 

abundant in moderately and heavily than in lightly or 

ungrazed treatments. Heavy grazing resulted in more 

prostrate, short plants with sod-like physiognomy compared to those in the ungrazed treatment. Colonizing 

ruderal seedlings of three of four species were more 

abundant on heavily grazed than ungrazed treatment. 

Smith (1967) reported on the same treatments initiated 

by Johnson (1956) but after 17 years of treatment. Plants 

were categorized into those favored by lighter grazing 

(mountain muhly, Arizona fescue, blue grama, bottlebrush squirreltail lupines, bearberry, and Arkansas rose), 

those favored by heavy grazing (tumblegrass, pussytoes, 

groundsels, and goosefoot), and those not affected (sun 

sedge and fringed sagebrush). However, the only species 

with greater cover after 17 years of heavy grazing compared to initial values were blue grama and goosefoot. 

In terms of composition, perennial forbs increased and 

perennial grasses decreased over the 17 years (Table 1).

At the same research station as the Johnson (1956) 

study, Currie (1976) assessed treatments of protection, 
38 USDA Forest Service RMRS-GTR-169. 2006.

alternate rest, or seasonal spring or fall grazing on montane meadow that previously had been heavily grazed 

season-long (June thru October) for 23 years. The experiment had three replicates and also included fertilizer and 

herbicide treatments in a randomized split-block design. 

The concentration of livestock in the meadow areas of the 

surrounding ponderosa pine range had led to dominance 

by unpalatable species. Grazing intensities remained at 

the previous heavy rate of 55 percent utilization. The 

study included a pre-treatment sampling year followed 

by four years of treatments and measurements. Total 

plant density was not affected by grazing treatments, and 

only two species responded to the grazing treatments. 

Pennsylvania and horse cinquefoil were greater in plots 

grazed in spring or season-long compared to the other 

treatments. Currie concluded that no season or system 

of grazing, including complete protection, was more efficient in promoting recovery of depleted range. Currie 

did not comment on why his results were different from 

those of Johnson (1956) at the same research station. 

Currie’s sites were described as previously depleted, while Smith (1967) described Johnson’s 

sites as at one time heavily grazed but does not 

indicate a depleted condition.

In Arizona montane meadows, Rambo and 

Faeth (1999) studied three paired grazed and 

8 to 9 year ungrazed sites where consumption 

by cattle, elk, and deer was about 70 percent 

of standing crops. No consistent differences 

between grazing treatments were observed 

for the relative abundance of grasses, forbs, 

shrubs, or introduced exotic species. However, 

the method for estimating abundance was not 

presented (cover, density, biomass), and sampling too close to fence lines may have biased 

results. Some individual species and diversity 

responses were seen. Ungrazed sites had lower 

percentages of the abundant western yarrow, 

Table 2—Plant species density (number of individuals/0.93 m2) of four 

grazing treatments in montane meadow communities. Treatments 

were initiated in 1941 and values are an average for sampling from 

1949 and 1950, adapted from Johnson (1956).

Grazing treatment plant density (number/plot)

Species Heavy Moderate Light None

Grass

Blue grama 30.1 24.0 5.9 7.8

Arizona fescue 2.0 8.8 4.0 12.7

Mountain muhly 38.8 37.0 29.0 25.9

Sedge 106.7 58.9 44.0 53.8

Weeds (ruderals)

Rose pussytoes 60.3 32.1 21.4 10.0

Nickleaf milkvetch 3.2 4.6 2.2 6.1

Trailing fleabane 20.7 11.0 14.6 7.0

Fringed sagebrush 19.5 12.1 5.2 14.8

Table 1—Species composition (percent of total weight) of four grazing treatments in montane meadow communities. Treatments 

were initiated in 1941. Data for 1949 to 50 adapted from Johnson (1956), and 1940 and 1957 adapted from Smith (1967)

Grazing treatment species composition (percent)

Heavy Moderate Light None

Species 1940 ’49-’50 1957 1940 ’49-’50 1957 1940 ’49-’50 1957 ’49-’50 1957

Grass

Blue grama 2.8 16.2 7.7 4.6 7.1 9.8 1.3 2.1 5.8 1.3

Arizona fescue 8.0 1.8 1.5 8.7 29.5 8.2 4.5 15.6 4.1 14.9

Mountain muhly 20.8 20.1 14.6 17.2 26.9 15.4 19.1 30.4 19.4 45.0

Other grasses 24.6 1.9 23 8.1

Sedge 4.9 5.1 4.5 3.5 5.0 2.7 3.7 3.2 3.2 5.8

Total grass & sedge 44.0 47.6 34.0 36.0 70.3 51.0 35.0 74.3 51.0 74.9 59.0

Weeds (ruderals)

Rose pussytoes 22.4 6.5 8.6 3.9

Nickleaf milkvetch 1.6 1.3 0.7 2.3

Trailing fleabane 2.3 0.4 1.0 0.5

Fringed sagebrush 14.1 14.2 16.1 13.3 7.0 12.1 10.2 2.6 11.3 12.3

Other ruderals 12.2 14.5 12.9 6.1

Total ruderals 52.4 29.7 25.7 25.1

Perennial forbs 55.0 63.0 62.0 46.0 53.0 47.0 36.0

Total 99.0 100.0 97.0 98.0 100.0 97.0 98.0 100.0 98.0 100.0 95.0
USDA Forest Service RMRS-GTR-169. 2006. 39

thereby increasing evenness. Species richness was 

greater on two of the three grazed sites.

Primary Production

Results from studies of primary production responses to grazing of montane meadows vary. For montane 

meadows in northern Arizona, Arnold (1955) reported 

primary productivity values declined in a nearly linear 

fashion with decreasing range condition from excellent 

to very poor (Fig. 4A). Very poor condition meadows 

supported only 21 percent the productivity of meadows 

in excellent condition. As noted in the ponderosa pine 

section, however, side-by-side plotting of the Arnold 

production values for montane meadows and openings 

between ponderosa pine trees gave similar and smooth 

curves, and it is not described how the sites were chosen 

or the data were collected. As described in the section 

above, early studies sometimes chose sites based on preconceived ideas of condition, and then sampled these 

condition classes. No indication of grazing intensity or 

history is given for the sites.

In the Johnson (1956) study described above, primary 

production of the heavily grazed treatment was 47 and 

79 percent of the ungrazed treatment for the normal and 

drought year, respectively (Table 3). Grazing effects on 

total production were less under drought condition. The 

effect of drought on lessening differences between the 

heavier and lighter grazing treatments was also evident 

in grass and sedge production. The drought reduced 

primary production 28, 22, 43, and 51 percent from the 

previous year’s level on the heavily, moderately, lightly 

grazed, and ungrazed treatments, respectively. Johnson 

suggested additional study to explain causes of this unusual result. Production in the normal year was generally 

heavy < light, moderate < ungrazed, while differences 

among means in the drought year could be attributed 

to variance in the data. Weed (colonizing ruderal) production was always highest on the heavily grazed and 

lowest on the ungrazed treatment. Decreased seed 

stalk production with increasing grazing intensity was 

observed for only two species (mountain muhly and little 

bluestem), and not for any other species sampled. After 

17 years of treatment, Smith (1967) reported nearly five 

times the production of grass and sedge on protected 

versus heavily grazed meadows, while moderately and 

lightly grazed treatments were approximately 25 and 50 

percent of protected areas. Smith concluded that light 

and moderate grazing maintained satisfactory production throughout the study, while heavy grazing resulted 

in increasingly reduced production.

In the Currie (1976) study of protection, alternate 

rest, or seasonal spring or fall grazing (see details 

above) there were no significant effects of treatment by 

year or main effect of treatment on primary production, 

and average range among treatments was low at 125 to 

135 g/m2/yr. Although protection or grazing seasons or 

systems did not promote recovery, fertilizer and herbicide treatments were effective in improving condition of 

Table 3—Primary production (g/m2/yr) of four grazing treatments in montane meadow communities. 

Treatments were initiated in 1941. Adapted from Johnson (1956).

Grazing treatment primary production (g/m2

/yr)

Heavy Moderate Light None

Species 1949 1950 1949 1950 1949 1950 1949 1950

Grass

Blue grama 15.7 14.9 8.3 7.3 1.3 2.2 2.4 1.7

Arizona fescue 2.5 1.0 38.4 27.7 2.5 8.3 28.5 18.5

Mountain muhly 22.9 16.0 33.5 26.1 38.0 22.3 109.7 43.9

Other grasses 6.7 2.2 2.9 1.5 24.3 19.5 17.9 8.5

Total grasses 47.8 34.2 83.2 62.5 88.6 52.3 158.5 72.6

Sedge 5.7 4.1 4.8 5.9 5.7 1.3 8.9 8.2

Total grass & sedge 4.9 38.3 88.0 68.5 94.3 53.7 167.4 80.8

Weeds (ruderals)

Rose pussytoes 19.1 22.9 9.9 4.9 8.5 7.6 10.0 3.6

Nickleaf milkvetch 2.4 0.8 1.6 1.2 0.9 0.6 4.1 2.9

Trailing fleabane 4.0 0.6 0.3 0.6 0.3 1.2 0.2 1.0

Other weeds 14.8 8.7 24.7 9.6 19.5 7.4 18.2 4.4

Total ruderals 40.2 33.0 36.4 16.3 29.3 16.8 32.5 11.9

Fringed sagebrush 14.6 14.9 7.3 8.1 1.5 2.9 22.1 15.9

Total 105.3 86.2 131.7 92.8 125.0 73.4 221.9 108.6
40 USDA Forest Service RMRS-GTR-169. 2006.

depleted range. Again, Currie did not comment on why 

his results were different from those of Johnson (1956) 

at the same research station.

Rooting Characteristics and Soil Properties

Various aspects of root biomass and distributions and 

soil properties were assessed in the Colorado studies 

described above. Johnson (1956) found no significant 

differences among ungrazed, lightly, and moderately 

grazed treatments in aboveground to belowground ratios 

of little bluestem, Arizona fescue, or mountain muhly. 

However, the ratios were about 2.5 for heavily grazed 

treatment compared to 3.5 for the three treatments under lower grazing intensity. Heavy grazing reduced 

root biomass in the upper soil profile. After 17 years 

of treatment, Smith (1967) extracted root monoliths 

from the treatments to a depth of three feet. Root biomass in moderately grazed treatments was 84 percent of 

that in ungrazed, and heavily grazed was 68 percent of 

that in ungrazed treatment. Differences were due to the 

grass component. The vertical distribution of roots was 

shallower under heavy grazing intensity, with 81 percent in the top third of the profile in ungrazed treatment 

compared with 86 percent in the heavily grazed treatment. Lateral spread of grass roots tended to decrease 

with increasing grazing intensity, but differences were 

not significant. Lateral spread of fringed sagebrush and 

pussytoes roots was greatest under moderate grazing 

intensity. The diameter and strength of primary roots 

and the number of branch rootlets of mountain muhly 

and Arizona fescue decreased with increasing grazing 

intensity. In contrast, ungrazed and moderately grazed 

blue grama roots did not differ, but declined only with 

heavy grazing. Roots are the primary form of input of 

organic matter to grassland soils, whereby differences 

with grazing would eventually manifest as differences 

in soil carbon contents.

Infiltration rates on the grazed treatments were 

highly variable and generally did not change over 

the period of study. Results were clear only in the 

ungrazed exclosures, where rates were 116 percent of 

initial after 6 years of protection and 167 percent after 13 years. Erosion rates were measured only after 11 

years of treatment. Erosion rates on lightly and moderately grazed treatments did not differ, but were 66 

percent of those on heavily grazed treatment. Losses of 

root biomass and aboveground plant cover under heavy 

grazing were great enough to affect soil loss, thereby 

affecting potential recovery rates of the system even if 

grazing management changes were undertaken.

Overall, there are too few controlled studies of the 

effects of grazing on montane grassland to base any 

strong conclusions. The two strong studies that are 

available reach different conclusions possibly due 

to different initial conditions or to the length of the 

studies. Heavy grazing appears detrimental to the 

aboveground and belowground plant community and to 

soil properties.

Temperate Grassland________

The temperate grassland group is a very large mix 

of communities in semiarid to subhumid environments 

and with short to long evolutionary histories of grazing 

(see History of Grazing by Native and Domestic Large 

Herbivores section above). With a long history of grazing by bison of the Great Plains, the shortgrass steppe 

which occupies the far-western part of New Mexico 

(Lauenroth and Milchunas 1991; Fig. 3) would be expected to be particularly tolerant of grazing by large 

herbivores, while other communities in this type that 

occur in conjunction with Great Basin communities 

would be expected to be sensitive to grazing (Mack and 

Thompson 1982; Milchunas and others 1988). Gosz and 

Gosz (1996) map of the Great Plains grasslands extends 

farther west than Lauenroth and Milchunas’s (1991) 

map of the shortgrass steppe (Fig. 3), because of the 

influence of blue grama in plant communities farther 

west than what may be considered shortgrass steppe, 

but also because the Great Plains geomorphic province 

extends farther west than the shortgrass steppe biotic 

community (Sims and others 1978). Blue grama is the 

overwhelming dominant of the shortgrass steppe and 

one of the species that imparts tolerance to aboveground disturbances such as grazing or fire (Milchunas 

and others 1990). Blue grama may or may not respond 

similarly to grazing in different locations. The behavior of blue grama in response to grazing in different 

communities and locations is discussed further in the 

Conclusions section below. In this section, we start with 

assessment of the southern shortgrass steppe, followed 

by other communities dominated by short grass species, 

and then examine other plant communities, thereby 

attempting to organize the studies from long to short 

history of grazing.

Species Composition

The northern shortgrass steppe is among the communities most resistant to grazing in the world (see Fig. 
USDA Forest Service RMRS-GTR-169. 2006. 41

13 in Comparison of Southwestern Communities with 

Great Plains Communities section below). However, 

some differences exist between the southern and northern shortgrass steppe, including greater precipitation, 

temperature, and aboveground plant production in the 

south (Sims and others 1978; Sims and Singh 1978a, 

b). In west Texas near Amarillo, the IBP Pantex southern shortgrass steppe site is included here because of 

its proximity to, and similar latitude of, New Mexican 

shortgrass steppe. Precipitation averages 499 mm/yr. A 

five-year-old exclosure was compared to heavily grazed 

communities where utilization of aboveground production was 65 percent. This is very heavy utilization 

compared with that generally given for other communities reviewed in this report but is similar to the 60 percent 

utilization for heavy grazing treatments in the northern 

shortgrass steppe (Milchunas and others 1989). Grazing 

treatments did not affect the composition of warm- or 

cool-season grasses, which together comprised 75 percent or more of the vegetation (Table 4; Sims and others 

1978). Forbs were less than 10 percent of the community and were greater on ungrazed treatment. Cacti were 

important but too patchily distributed to accurately 

assess with the number of quadrats used. In terms of 

species composition, blue grama increased from 40 

percent of the vegetation on ungrazed treatment to 57 

percent on the very heavily grazed treatment (Table 4; 

Grant 1971). Buffalo grass (a very short stoloniferous 

species) and scarlet globemallow (a very deep rooted, 

rhizomatous forb) also increased with heavy grazing. 

Overall, however, there were very small differences 

between the grazed and ungrazed plant communities, 

with a calculated a similarity index of 96 percent (Grant 

1971). That degree of similarity was equal to or greater 

than what was observed for the other nine IBP grassland 

sites in North America, with the northern shortgrass 

steppe site having a value of 88 percent similarity for 

heavily grazed compared to ungrazed and 90 percent 

for moderately grazed compared to ungrazed. Other 

aspects of community structure also showed little difference between grazing treatments. Live plant biomass 

was slightly greater in grazed than ungrazed treatment, whereas old-standing dead material was greater 

on ungrazed. This resulted in no difference between 

standing plant biomass between heavily grazed and 

ungrazed treatments. Both root biomass and crown 

biomass were greater on heavily grazed compared to 

ungrazed treatment, probably due to the increase in blue 

Table 4—Plant biomass, functional group composition, and species composition 

based on biomass at an ungrazed or grazed southern shortgrass steppe site in 

western Texas. Adapted from Sims and others (1978) and Grant (1971).

Ungrazed Grazed

Biomass component Plant biomass (g/m2

)

Live 70 86

Recent dead 33 33

Old dead 96 79

Total above ground standing crop 200 199

Litter 232 215

Crown 254 299

Roots in upper soil layer 530 668

Total roots 620 725

Groups Functional group composition (percent)

Warm-season grasses 75 69

Cool-season grasses 5 5

Warm-season forbs 7 2

Cool-season forbs 5 3

Cacti 10 22

Species Species composition (percent)

Blue grama 40.2 56.7

Buffalograss 1.2 2.4

Hedgehog cactus 0.0 0.0

Plains prickly pear 56.3 40.3

Prairie coneflower 1.9 0.0

Scarlet globemallow 0.0 0.5
42 USDA Forest Service RMRS-GTR-169. 2006.

grama with large crown and root biomass compared to 

aboveground biomass. Similar results have been observed for northern shortgrass steppe (Milchunas and 

Lauenroth 1989; Milchunas and others 1989), but will 

not be reviewed here because of their distance from 

southwestern communities.

The Fort Stanton Experimental Ranch in south-central New Mexico is west of the shortgrass steppe but 

within the Great Plains geomorphic province. The 

grassland at this site is dominated by blue grama, with 

wolftail, sideoats grama, ring muhly, sand dropseed, 

and other important grasses. Although dominant, blue 

grama is roughly half of the grass biomass and a third 

of the total biomass. The open grassland sites are found 

between surrounding pinyon-juniper communities. 

Precipitation is about 400 mm/yr, with 60 percent falling during the June through September growing season. 

Pieper and others (1978, 1991) compared moderate, 

“proper” stocking rates to a 25 percent greater heavy 

stocking rate for seven years. Blue grama, other grasses, and forbs were lower on the heavily compared to 

moderately grazed treatment for all of years 2 through 

7. The reductions in blue grama under the heavy grazing 

treatment were often large, with half as much biomass 

compared to moderately grazed. The differences were 

not due to consumption, since sampling occurred under 

temporary cages that were moved each year. A drought 

near the middle of the study did not greatly affect relationships between the two treatments. Similar treatment 

effects were observed for basal cover of all species and 

blue grama, except that differences did not appear until 

the fourth year of study. Larger grazing treatment effects 

were observed between moderate and heavy grazing 

at this site than were observed between ungrazed and 

heavy grazing at the shortgrass steppe site above.

Five exclosure-grazed paired grasslands were among 

the large number of sites compared between 1940 and 

1964 by Krenetsky (1971). The sites were considered 

the shortgrass association due to the dominance by blue 

grama, but they spanned a wide geographic range in 

three different National Forests, a precipitation range 

from 300 to 500 mm/yr, and were sometimes closely 

associated with pinyon-juniper or ponderosa pine communities. Averaged over the five sites, there was little 

difference between grazing and protection in the percentage change between 1940 and 1964 of grasses, 

forbs, browse, or total vegetation basal cover. The largest 

difference was for grasses, which increased 15 percent 

over time in the ungrazed compared to an increase 

of 11 percent on the grazed sides of the fences. There 

were large differences among sites, but the differences 

among sites in the temporal changes were greater than 

differences due to grazing. Blue grama did not respond 

to grazing treatment at three of the five sites, showed 

small relative increases with protection at one site, and 

showed large relative increases at another. Blue grama 

generally increases with grazing throughout Great 

Plains grasslands (Lauenroth and Milchunas 1991), so 

this is a rather unusual response, but one that is not uncommon in the southwest as shown in this review. The 

latter site was in Gila National Forest in western New 

Mexico. Other unusual dynamics at the Gila-8 site were 

the loss of a number of forb species on both treatments 

between 1940 and 1964 and the appearance of mesquite 

in the exclosure. Krenetski concluded that there was 

little difference between grazed and protected sites except for some differences for the two at lower elevation. 

New species of grasses did not come into the exclosures. 

From Krenetski’s tables, understory species richness averaged 8.4 in ungrazed exclosures compared to 8.0 in 

the adjacent grazed areas.

Orodho and others (1990) studied an Indian ricegrass/fourwing saltbush community in northwestern 

New Mexico. Grass biomass alone at this productive 

site was 187, 261, and 235 g/m2 for upland, slope, and 

swale topographic position, respectively. Communities 

protected from grazing for about 50 years within Chaco 

Culture National Historical Park were compared to 

adjacent heavily grazed sites. The response to grazing 

treatments differed with topographic position. The density of Indian ricegrass was greater on protected slope 

and swale, while the opposite was found on the upland. 

Similar treatment effects were observed for other grasses, except for no difference with grazing on the slope 

community. Forb density was greater only on swales. 

Fourwing saltbush is a valuable forage plant for livestock, especially in winter, and this species increased 

with protection. No differences in grazing treatments 

were observed for biomass of Indian ricegrass, other 

grasses, or forbs when plots on the grazed sites were 

protected during the year of study.

Natural geologic refuges that have never been 

grazed can provide interesting comparisons with adjacent grazed areas, although there are other potential 

confounding effects of their isolation (Milchunas and 

Noy-Meir 2002). A study of a “park” isolated by steep 

surrounding walls in southeastern Utah’s Canyonlands 

National Park is included here because of climate and 

soils that are reasonably similar to nearby northern New 

Mexico and Arizona (Kleiner and Harper 1972, 1977). 

Because of the nature of isolation, abiotic conditions 

are similar at the protected and grazed sites with 250 

mm/yr precipitation. Relative frequencies of plant and 

cryptogram species in galleta/needle-and-thread grass 
USDA Forest Service RMRS-GTR-169. 2006. 43

communities displayed very large differences between 

species composition and richness. Sixteen species were 

significantly more abundant in the refuge and 12 were 

more abundant in the grazed site (Table 5). The average number of vascular species per stand was 15 and 11 

for ungrazed and grazed sites, respectively, and was 6 

and 2 for cryptogram species. Vascular plant cover was 

similar between ungrazed and grazed sites (17 percent), 

but cryptogram cover was much higher with protection (38 vs. 5 percent). Annuals were more abundant 

on the grazed site, primarily due to six-weeks fescue 

grass. The exotic cheatgrass was more abundant in the 

refuge, while the naturalized exotic Russian thistle 

was more abundant in the grazed site. Of the dominant 

grasses, galleta was more abundant in the refuge and 

needle-and-thread more abundant in the grazed site. 

These very large differences existed even though the 

grazed site was utilized lightly to moderately during 

winter due to lack of adequate water, and was ungrazed 

for five years prior to sampling. Kleiner (1983) assessed 

successional trends in the grazed site 10 years later and 

reported continued directional change towards the relict 

ungrazed condition. The recovery of cryptogramic crust 

was more rapid than earlier believed.

Table 5—Species relative abundance index for ungrazed natural geologic refuge and grazed 

communities in Canyonlands National Park. Adapted from Kleiner and Harper (1972).

Relative abundance index *

Species Ungrazed Grazed

Galleta 5,206 3,553

Needle-and-thread 3,626 4,902

Wooly Indian-wheat 3,304 516

Six-weeks fescue 1,876 2,890

Prickly pair 1,377 39

Indian ricegrass 1,141 1,343

Sand dropseed 663 1,530

Desert tansy mustard 588 48

Rosy gilia 538 0

Blueburr stickseed 476 6

Hoary tansyaster 465 296

Nodding wild buckwheat 312 138

Downy chess 183 9

Praire pepperweed 163 64

Aster spp. 36 230

Small wirelettuce 36 4

Plains hiddenflower 27 105

Sego lily 18 0

Desert false-yarrow 17 0

Blue grama 16 522

Green joint-fir 3 360

Winterfat 3 90

Pale evening-primrose 1 15

Russian thistle 0 24

Silvery globe mallow 0 1,463

Floral characteristic

Average number of species/stand (species richness) 22 13

Average number of vascular species/stand 15 11

Average number crytogam species/stand 6 2

Average number of species/quadrat 7 3

Vegetational characteristic

Average total living cover (percent) 55 22

Vascularplant living cover (percent) 17 17

Crytogamic cover (percent) (1⁄2 moss - 1⁄2 lichen) 38 5

Average number of hits/stand on living vascular tissue 23 26

Litter cover (percent) 12 10

Contribution of annual to the total vegetation (percent) 12 16

* constancy times frequency, with a possible range from 0 to 10,000.
44 USDA Forest Service RMRS-GTR-169. 2006.

Primary Production, Root, and Soil Responses

Aboveground primary production for the shortgrass 

steppe site described above was greater in ungrazed 

than grazed treatments in two out of three years of 

study (Sims and others 1978). Average values were 

257 and 225 g/m2/yr for the ungrazed and very heavily 

grazed treatment, respectively. However, these values 

may not be statistically different, and the method used 

for calculating production was the sum of individual 

species peaks through the growing season. Although 

this was chosen as the best out of more than 30 methods of calculation at the time (Singh and others 1975), 

the summation of peaks method can generate large biases whereby peak-standing crop of temporarily caged 

vegetation gives more reliable estimates with less bias 

(Milchunas and Lauenroth 1992). Peak-standing crop 

estimates of aboveground production from the same 

data yield estimates of 103 and 179 g/m2/yr for the ungrazed and grazed treatments, respectively, which is 

opposite treatment direction to the sum of peaks estimate. Similarly, estimates of root and crown production 

based on summation of peaks should be viewed with 

caution, but in this case only an estimate of biomass 

is valid for peak standing crop values since crown and 

root turnover takes many years. Estimates of crown, 

near-surface root, and total root biomass were greater in 

grazed than ungrazed southern shortgrass steppe (Table 

4). Very little difference in crown and root biomass 

between heavily grazed and 47-year ungrazed treatments have been observed in northern shortgrass steppe 

(Milchunas and Lauenroth 1989).

Estimates of total herbage production for the Fort 

Stanton Ranch study described above were greater for 

the moderate compared to heavy grazing treatments in 

all eight years but two, and the two were drought years. 

For the non-drought years near the end of the study, 

aboveground production was greater in the ungrazed 

compared to grazed treatment by as little as 12 g/m2/yr 

to as much as 50 g/m2/yr.

Primary production did not differ with grazing 

treatment at the Indian ricegrass/fourwing saltbush 

community described above (Orodho and others 1990). 

However, soil bulk density was 8 percent greater at the 

grazed compared to long-term protected site. The largest difference between grazing treatments was for an 

upland site, and the least for a swale site. Bulk density 

did not differ at the Canyonlands site in Utah, but soil 

organic matter was 1.33 percent for the refuge compared to 0.88 percent for the previously grazed site. Soil 

potassium and phosphorous were also slightly higher in 

the ungrazed communities, but nitrogen did not differ 

with grazing.

Grazing Management

Two grazing systems were evaluated at the Fort 

Stanton Experimental Ranch. While comparing the 

moderate and heavy continuous grazing treatments 

described above, Pieper and others (1978, 1991) also assessed a four-pasture, one-herd rotation system stocked 

at the same level as the heavy continuous pasture. The 

rotation system resulted in greater primary production 

than the continuously grazed pasture in five out of eight 

years. Grasses other than blue grama showed the greatest increase in response to rotational grazing, and forbs 

showed declines. There was no advantage of the rotation system in terms of cattle performance.

A short-duration seven to nine cell rotation system was compared to a continuously grazed pasture. 

However, the short duration pasture was stocked from 

1.1 to 2 times more heavily than the continuous pasture and received generally less precipitation as well. 

Blue grama increased production and basal cover in 

the short duration pasture. However, no other species 

significantly increased under the short duration system, 

and basal cover of both total grasses and total forbs decreased. Average total production over growing periods 

were only slightly greater on the continuously grazed 

pasture.

As mentioned above, grazing intensities for the 

shortgrass type are generally higher than for other 

southwestern communities. Of nine native communities, Holechek and Pieper (1992) recommended 40 to 

50 percent utilization of shortgrass steppe and southern 

mixed-grass prairie of the Great Plains, compared to 20 

to 30 to 30 to 40 percent utilization for other southwestern communities. This reflects the relative tolerance 

to grazing of these systems that have a long history of 

grazing by native large herbivores.

There are a limited number of studies given the 

diverse nature of plant communities in the broad 

general type. A true shortgrass steppe community 

displayed a high tolerance to relatively very high 

grazing intensity. Magnitude of responses to grazing 

appeared to increase over a gradient of shortgrass 

steppe, short grass communities within pinyon-juniper 

range but within the Great Plains grassland Province 

(see Fig. 2, 3), to grasslands of the Colorado Plateau 

Shrub-Steppe associated with the Great Basin. The 

response of blue grama also changed across the history 

of grazing gradient, from increases with grazing to a 

mix of decreases or no effect. The large and important 
USDA Forest Service RMRS-GTR-169. 2006. 45

belowground component of the southern shortgrass 

system also appears highly tolerant of heavy grazing; 

although based on only one study, the data are supported 

by similar studies in the northern shortgrass steppe. At 

the other extreme of evolutionary history of grazing, a 

never-grazed natural geologic refuge shows very large 

effects of grazing.

Mesquite Savanna___________

Mesquite savanna communities receive from 200 to 

500 mm/yr of precipitation, and there is a major west 

- east gradient in the seasonal distribution (Martin 

1975). The percent of annual precipitation occurring 

April through September is only 45 percent in western Arizona, increases to 65 percent in western New 

Mexico, and peaks at 75 percent in western Texas. As 

mentioned above, Mediterranean climates are conducive for annual and/or shrub dominance, and the degree 

of a Mediterranean climate has important implications 

for the success of these life-forms and their potential 

response to disturbance. Another important aspect of 

the climate that drives vegetation composition is the 

high variability of precipitation and the periodic cycling 

of drought periods. Mesquite savanna, which includes 

what is commonly called desert grassland and semidesert grasslands, has undergone more change over the 

past hundred years than possibly any other non-wetland 

community type in North America. Large-scale shrub 

invasions (mesquite, creosote-bush, tarbush, snakeweed, burroweed, cholla) into grassland coincided with 

the historic period of very heavy grazing by domestic 

livestock, which led to a century of debate concerning 

mechanisms associated with livestock grazing that may 

be responsible for such drastic shifts in communities. 

This will be discussed in the Shrub Invasion, Livestock 

Grazing, Fire, Small Mammals, and Climate section 

below, along with other hypotheses about contributing factors. In many cases, what is now a Chihuahuan 

Desert mesquite dominated shrubland used to be a mesquite savanna grassland. This transition is still ongoing. 

Therefore, there is some ambiguity concerning classification and an overlap between this section and the 

Chihuahuan Desert section below.

There are possibly more studies of grazing in the 

mesquite savanna type than all others combined. This 

is because mesquite savanna occupies large areas of 

several long-term research sites and because livestock 

grazing is an important landuse in the community. 

The sites include the Jornada Experimental Range 

and the nearby Chihuahuan Desert Range Research 

Center College Ranch (231 mm/yr precipitation) near 

Las Cruces, New Mexico, in the Jornada Del Muerte 

Basin, the Santa Rita Experimental Range south of 

Tucson, Arizona (384 mm/yr precipitation), the PageTrowbridge Experimental Ranch in southern Arizona 

(394 mm/yr precipitation), and the Appleton-Whittel 

Research Ranch in southeastern Arizona (430 mm/yr 

precipitation, half in July and August). Some of these 

research sites provide long-term monitoring of controlled grazing treatments through periods of drought 

and favorable growth and spanning shrub invasions. 

These long-term data sets are particularly valuable in 

disentangling the complexity of interacting factors involved in community shifts. The following section on 

species composition is organized by research site and 

plant community, from earliest to latest study for sites 

with a history of research.

Species Composition

Jornada black grama and dropseed communities—Interactions between drought and grazing on 

black grama communities were noted by Campbell 

and Bomberger (1934). Severe drought and some very 

heavy grazing had large effects on the basal cover of 

black grama and snakeweed. Black grama is a primary 

component of the late seral stage and is a valuable forage species. Black grama does not often reproduce by 

seed but spreads by stolons. Recovery of black grama is 

therefore often slow. Snakeweed is a shrub sometimes 

toxic to cattle and was thought to represent deteriorated range condition (Jardine and Forsling 1922). These 

authors had previously shown that very heavy grazing 

leads to loss of black grama and establishment of snakeweed. Drought and heavy grazing were believed to also 

result in movement of the mesquite sand dune shrub 

association into black grama grassland. Campbell and 

Bomberger showed that with above-average precipitation and light grazing, black grama basal cover increased 

nearly five to six times after six years of recovery from 

heavy grazing and drought, while snakeweed was only 

two to three times greater.

Nelson (1934) studied black grama communities over 

13 years that spanned two drought periods. Early grazing studies often used distances from water, together 

with exclosures, to obtain grazing intensity gradients. 

Conservative grazing was considered to be 75 to 80 

percent utilization of black grama, whereas too intense 

grazing pressure was considered complete use of black 

grama and a range that appeared trampled after the 

grazing season (see below for more recent definitions 
46 USDA Forest Service RMRS-GTR-169. 2006.

of conservative grazing intensities). Drought alone was 

found to have extremely large effects on black grama. In 

plots protected from grazing, black grama basal areas 

declined 42 percent during the first drought, nearly recovered after two favorable growing seasons, declined 

to 11 percent of the original cover after an additional two 

years of drought, and then recovered again. Conservative 

grazing resulted in similar average basal covers to that 

of complete protection. However, the recovery from 

drought cycle was faster under conservative grazing 

than it was under no grazing. Increased tillering from 

a greater number of, but smaller more evenly distributed, tufts in conservative compared to ungrazed areas 

was thought to better utilize spatial distribution of soil 

water. Heavier grazing intensities prevented full development of black grama following drought, and allowed 

other species to increase. Very heavy grazing resulted 

in almost complete disappearance of black grama during drought and lead to unstable soil conditions. Height 

growth of black grama under very heavy grazing was 

20 percent of that on ungrazed treatments during dry 

years and 75 percent during favorable years.

An additional early report of responses of black 

grama communities to grazing was provided by Paulsen 

and Ares (1961, 1962) for 1916 through 1952. By this 

time, definitions for grazing intensities had changed. 

Heavily grazed sites were considered to have more than 

55 percent utilization of black grama, intermediate 40 

to 55 percent, and conservative less than 40 percent. 

These authors reinforced Nelson’s (1934) conclusions. 

The prolonged dry period between 1916 and 1926 reduced basal cover of black grama to approximately the 

same level regardless of grazing intensity or initial cover conditions. However, basal cover declined the most 

in protected plots after the severe drought of 1923, with 

other treatments declining in proportion to pre-drought 

covers. Recovery during favorable periods was more 

rapid in protected and conservatively grazed plots than 

intermediate and heavily grazed plots (3 times greater 

cover compared to 2 times, respectively). Low rainfall 

during 1935 again resulted in large declines in black 

grama on all grazing treatments. After recovery in 1937, 

basal cover of black grama increased 20, 68, 62, and 

42 percent on ungrazed, conservatively, intermediately, 

and heavily grazed plots, respectively. Previously high 

basal areas were reached only on conservatively grazed 

plots. Associated less drought tolerant species declined 

even more than black grama during drought periods but 

then recovered more rapidly than black grama during 

favorable growth periods. Another dry period ensued 

and by 1956 black grama had disappeared from all ungrazed plots, 67 percent of heavily grazed, 62 percent 

of intermittently grazed, and only 30 percent of the conservatively grazed. Recovery was slower when losses 

were extensive. Tobosa communities responded similarly to dry/wet conditions but were less sensitive to 

grazing intensities (see below). Paulsen and Ares do not 

report on temporal trends in shrub covers under various 

grazing intensities because their data were from small 

quadrats. They do, however, mention that the number 

of mesquite plants increased 125 percent from 1936 to 

1951 on even conservatively grazed range. Hennessy 

and others (1983) analyzed data from transects spanning over 2000 m that were sampled between 1935 and 

1980 on the Jornada Experimental Range. Changes that 

occurred over the 45 year time period did so both on 

areas grazed and ungrazed by livestock. Mesquite attained complete dominance, while black grama and 25 

percent of forb species had completely disappeared by 

1980. Species capable of inhabiting the inter-dune areas 

were mesa dropseed, fluffgrass, and broom snakeweed.

Rodents and rabbits are native herbivores that can 

have significant effects on plant communities and may 

interact with livestock if the domestic large herbivore 

has an influence on abundance of the small herbivores. 

Norris (1950) conducted one of the early studies of 

the relative effects of these herbivores in both black 

grama grassland and mesquite-snakeweed communities. Rodent and rabbit populations are relatively low 

in the black grama communities compared to mesquite 

type (3 1⁄2 times greater), whereby the small mammals 

had little effect on basal area or standing crops of perennial grass plants in the grassland community. Similarly, 

Havstad and others (1999) found only small effects of 

long-term rabbit exclosure on black grama grassland, 

and effects were only apparent after 50 years of protection. In contrast, standing crop in the shrub type was 

four to five times greater when protected from rodents 

and rabbits together (note, Norris [1950] often uses the 

term production when grazed-standing-crop was actually measured). Crown cover of snakeweed shrubs were 

somewhat greater in plots open to rodents or to both 

rodents and rabbits, but data was variable. The author 

suggests, however, that small herbivore grazing tends 

to favor shrub increases and prevent restoration of perennial grasses. Similar large effects of rodents on grass 

biomass was reported for the nearby College Ranch 

(Knox and others 1951, cited in Paulsen and Ares 1962). 

Gibbens and others (1993) reported similar shrub canopy 

cover to grass basal cover ratios in shrublands grazed or 

protected from rabbits for 50 years. However, changes in 

species composition were evident due to rabbit grazing. 

Although there was no difference in creosotebush, mesquite and tarbush decreased in grazed areas compared to 
USDA Forest Service RMRS-GTR-169. 2006. 47

exclosures, probably due to both browsing of the canopy 

and seedlings. Spike dropseed was 30-fold greater in exclosures, while fluffgrass was more abundant in grazed 

areas. Black grama and bush muhly did not respond to 

the long-term rabbit grazing treatments. Perennial grasses showed only small increases in basal cover 30 yrs after 

shrub removal treatment, similar to findings of Beck and 

Tober (1985). The potential role of small mammals in 

shrub invasions are discussed in the section below.

Buffington and Herbel (1965) described the shrub 

changes in a large 640 acre exclosure built across a 

mesquite sand dune / grassland transition zone in 1931. 

Even though grazing had been excluded, the mesquite 

expanded completely across the exclosure and out into 

grazed grassland on the other side. Based on visual observations, Hennessy and others (1983) indicated no 

difference in 1980 between grazed and ungrazed sections of the transect crossing through this exclosure.

Controlled grazing intensity studies of the recovery 

of “deteriorated” black grama range (“deteriorated” = 

areas where there were some remaining individuals) 

were undertaken in 1954 through 1963 at the College 

Ranch (Valentine 1970). The most severe drought in 

350 years occurred from 1951 to 1956. At each of three 

pastures at deteriorated locations near watering points, 

light, moderate, “proper”, and heavy grazing intensity 

treatment plots were established with goals of 20, 35, 

50, and 65 percent utilization of black grama herbage 

weight, respectively. Actual intensities averaged over 10 

years were 26, 35, 49, and 60 percent utilization. Light 

and moderate uses were generally similar and different 

from proper and heavy uses in terms of black grama 

seed-stalk height (a measure of “vigor”), average number and length of stolons, number of rooted buds (tillers 

initiating from stolons), and cover. Cover of black grama 

averaged 38 percent of that for “good-condition range” 

under both light and moderate use compared to 26 percent for proper and heavy use treatments. Recovery was 

just as rapid under moderate grazing as under light grazing. An interesting finding was that soil conditions at a 

particular site had large effects on grazing treatments. 

Recovery of black grama cover was 57 percent of that 

for good condition range where soils were a moderately 

deep sandy loam compared to only 8 percent recovery 

where soil was a shallow sandy loam with surface windscour, and 21 percent recovery where the soil was a very 

deep sandy loam.

An additional study of the extreme drought of the 

1950s was done at the adjacent Jornada Experimental 

Range (Herbel and others 1972). The importance of soil 

characteristics was again clearly seen, although the study 

did not address grazing as a factor. Prior to the drought, 

cover and yield of black grama did not differ among deep 

sandy, sandy flat, shallow sandy, low hummocky, flood 

plain, heavy sandy flat, or slope sites, and no site-location was particularly favored by precipitation during the 

drought. A direct relationship between drought damage 

to grasses and depth of soil was observed. Cover and 

biomass of grasses during the drought remained greater 

on shallow sandy sites and there were greater losses on 

the deep sandy and low hummocky sites. It was hypothesized that deep roots during a drought when moisture 

is primarily near the surface may be more susceptible 

to drought damage, and a shallow caliche layer holds 

more available moisture in a more readily obtainable 

depth. Black grama was so reduced on the deep sandy 

and low hummocky sites that there were not enough 

residual plants to provide a base for increase shortly 

following the drought. Many grass plants were killed 

by being covered by blowing sand, and erosion was 

greatest on sandy sites. The drought had greater impact 

on dropseeds and threeawns than on black grama, but 

those species do not have the constraint on recovery by 

seed like black grama does. Tobosa and burrograss on 

the flood plain sites were little affected by the drought. 

Possibly most important was that the drought favored 

mesquite at the expense of grasses. A later follow-up on 

this study reporting data through the late 1970s (Herbel 

and Gibbens 1996) showed that perennial grass cover 

remained low on deep sandy soils but increased on 

shallow soils over indurated caliche. A large response 

to protection from rodents, rabbits, or cattle was not observed. Mesquite cover was greater in areas protected 

from grazing, but grazing did not prevent an increase 

in mesquite.

Gil (1975) studied an extreme case of very heavy 

grazing compared to 12 years of exclosure to grazing. 

The exclosure had been built and grazed as a treatment 

during 1954 to 1963 in the Valentine (1970) study discussed above, but then left ungrazed. The exclosure 

was within 300 m of a stock tank. Therefore, the exclosure prior to 1954, and the adjacent grazed area, had 

been a “sacrifice” area. Areas near water are not only 

grazed heavily, but they are constantly trampled by animals coming to water and are also areas of very high 

nitrogen enrichment due to fecal and urine deposition 

(Senft 1983). While Atwood (1987) discounted the Gil 

study, it should simply be noted that the study does not 

represent general-area range conditions but a very heavily disturbed condition and the potential for recovery. 

Black grama was less than 1 percent of the grazed plant 

community compared to 64 percent of the ungrazed. 

Threeawns and dropseeds were 6 and 14 percent of the 

grazed and ungrazed communities, respectively. The 
48 USDA Forest Service RMRS-GTR-169. 2006.

majority of the heavily disturbed site was the half-shrub 

snakeweed (89 percent), while it was only 20 percent 

of the ungrazed community. Total biomass at the sites 

was 75 and 59 g/m2 for the ungrazed and grazed community, respectively, and bare ground cover was 69 and 

90 percent.

Comparisons were made between grazed and ungrazed treatments among 10 grasslands during the 

International Biological Program, with the Jornada being one of the sites (Sims and others 1978; Sims and 

Singh 1978 a, b; Grant 1971). More detailed information 

of the three years of sampling (1970 to 1972) specific 

to the Jornada site was published by Pieper and Herbel 

(1982) and Pieper and others (1983). Only one exclosure 

was sampled, but it was 10 ha. Sampling of shrubs by 

small quadrat methods was highly variable, but means 

across all years for mesquite and for yucca were higher 

in ungrazed than grazed treatment. Grazed areas were 

considered light to moderate at about 40 percent utilization. Based on peak live standing crop, black grama was 

the most abundant species on the ungrazed treatment, 

followed by Russian thistle. Russian thistle was the 

most abundant species on grazed treatments, followed 

by snakeweed. It is interesting to note that Russian thistle (tumble weed) is an exotic (but naturalized) annual 

forb that can be considered a colonizing ruderal species. 

Perennial grasses contributed 50 percent of the biomass 

of the ungrazed treatment compared with less than 30 

percent for the grazed.

A sequence in years of protection from grazing, and 

adjacent grazed sites, was sampled by Atwood (1987) 

by using exclosures at both Jornada experiment stations. 

Exclosures represented 17, 22, 32, and 48 years of protection from grazing. The 17-year site was the same one 

sampled during IBP (see previous paragraph), but sampling then was by standing biomass and this study used 

density (see Introduction for interpretation of methods). 

Years previous to and during the study had good rainfall, except for the 17-year site. Sites were far enough 

apart as to introduce a spatial factor into the temporal 

comparison, and there was only one site per duration 

of protection. The study does, however, provide one 

of the few temporal comparisons where sampling occurred at the same point in time, and sites were sampled 

for three years. Comparing grazed and ungrazed treatments within exclosure age, the density of black grama 

was equally abundant in the grazed and ungrazed 17-

year site, greatest in the grazed 22- and 48-year site, but 

greater in the ungrazed 32-year site (Table 6). The lack 

of a temporal pattern in black grama abundance suggests that differences among sites in local precipitation, 

soils, or initial conditions are more important than a 17 

to 48 year range in duration of protection. Dropseeds 

were generally favored by grazing. There were no consistent trends in forb, shrub, or annual species densities 

that could be attributed to grazing. Annuals were often 

greater on one treatment in one year or season and the 

other treatment in another year or season. Atwood concluded that properly managed grazing was sustainable 

in black grama communities. Again, however, the lack 

of a consistent temporal trend in direction of response 

may possibly be attributed to important interactions of 

grazing treatment with factors such as local precipitation, soils, or initial conditions.

Table 6—Plant density of ungrazed (Ungz) and grazed (Gz) plant communities at New Mexico mesquite 

savanna sites where the ungrazed treatment was protected for 17, 22, 32, or 48 years. Adapted from 

Atwood (1987).

Density (individuals/m2

)

17-Year 22-Year 32-Year 48-Year

Species Ungz Gz Ungz Gz Ungz Gz Ungz Gz

Black gramma 13.7 12.8 11.7 25.8 33.1 22.8 15.3 22.1

Dropseed spp. 0.8 9.5 1.8 4.4 0.3 0.3 1.1 3.4

seedlings 0.5 9.8 0.4 1.5 0 0 0.3 0.5

Other perennial grasses 0.2 0.8 0.9 11.4 0.8 17.6 1.0 1.2

seedlings 0 0.3 0.5 0.2 0 0.3 0 0.1

Annual grasses 8.7 34.9 1.6 9.0 2.0 1.6 8.4 3.8

seedlings 0 0 0.1 0 0 0 0.3 0

Perennial forbs 1.9 3.7 15.6 14.5 8.6 14.8 43.4 44.6

seedlings 0.3 1.7 6.6 13.3 0.1 0.4 37.9 34.8

Annual forbs 47.9 52.1 84.3 90.1 147.7 166.2 30.4 19.8

seedlings 9.8 14.4 42.5 50.4 1.9 4.1 37.2 21.7

Shrubs 1.9 1.4 1.3 1.5 0.8 0.7 1.2 1.8

seedlings 0.3 0.2 0.3 0.5 0.1 0 1.1 2.9
USDA Forest Service RMRS-GTR-169. 2006. 49

The permanent quadrat data for grass basal cover and 

forb density collected at the Jornada Experimental Range 

was summarized over a 64-year period by Gibbens and 

Beck (1988). Early years were reviewed above. For the 

black grama community type, long-term changes in 

perennial grass cover were similar for grazed and protected plots. Regardless of grazing treatment, however, 

perennial grass cover had not recovered by 1976 following the very severe drought of the 1950s. Highest cover 

of perennial grasses was recorded in the late 1920s to 

early 1930s at about 10 percent, fluctuated around 5 percent up to the 1950s drought, and then declined to below 

5 percent for the rest of the recorded time. Black grama 

disappeared from 77 percent of the original quadrats by 

1961 and was no longer the dominant perennial grass 

through the 1970s. Perennial forbs, annual grasses, and 

overall plant species richness declined through the 64 

year period. Shrubs dominated 42 percent of the quadrats by 1981 (Gibbens and Beck 1987). Similar results 

were recorded for quadrats in the poverty threeawn 

community type, except losses of perennial grasses and 

plant species richness through time were even more distinct. The loss in perennial grasses was not only due 

to drought, but also because mesquite had invaded and 

dominated the area, including the plot protected from 

grazing. Temporal patterns were not the same for tobosa 

or burroweed communities (see below).

A study of five exclosures (37 years old or greater) 

and adjacent grazed sites spanning both the Jornada 

Experimental Range and the College Ranch showed reduced perennial grass cover on the grazed sites but no 

effects of grazing on forb or shrub cover (Kerley and 

Whitford 2000). Black grama showed the largest decreases with grazing from 31 to 8 percent canopy cover. 

Fluffgrass was the only grass to increase with grazing, 

from 0.1 to 0.4 percent cover. Annuals were not affected by grazing treatments. No differences between 

grazing treatments were found for diversity or richness 

of either annuals or perennials. Nash and others (1999) 

also found only small effects of grazing on annual plant 

communities. Although there were no differences in 

shrub cover between treatments, Kerley and Whitford 

(2000) still suggest that grazing moves the plant community towards one more similar to a shrub community. 

Evidence is based on a DCA analysis showing some 

movement of ungrazed to grazed sites along axes towards shrub space. Additional evidence is cited from 

the Gibbens and Beck (1988) long-term study, but the 

material presented by these authors (reviewed above) 

indicates no difference in grazed and ungrazed quadrats 

in invasion by mesquite. Kerley and Whitford (2000) 

appear to imply that there is a sequence from long-lived 

black grama to short-lived dropseeds and threeawns 

and then to mesquite, but neither papers actually show 

this and neither show differences in mesquite between 

grazed and ungrazed treatments over reasonably long 

periods of time.

Herman (1988) sampled 15 exclosure and adjacent 

grazed areas that ranged in age of protection from 6 

to 74 years old. Only a few species showed time trends 

with increasing age of protection, and none were consistent among the 15 sites. Although some minor 

differences between grazed and ungrazed comparisons 

were evident, no differences between grazed and ungrazed treatments were observed for the four “problem 

shrubs” in the region. Similar results were observed for 

six mixed-brush grazed-ungrazed comparisons where 

the exclosures ranged from 5 to 52 years of protection.

Two other studies on the College Ranch (Nelson and 

others 1997; Winder and others 2000) are not considered 

here due to experimental design constraints in interpreting grazing intensity or grazing systems effects for the 

purpose of this review, although valid comparisons may 

be possible for other purposes. Sites previously in good 

condition were stocked lightly and pastures previously 

in fair condition were stocked more heavily, whereby 

site versus current intensity are confounding in comparisons. Additional studies at the College Ranch by Tembo 

(1990) and Holechek and others (1994) also have experimental design problems and unclear interpretation 

of results. Two different grazing systems were crossed 

during sampling of the moderately grazed site and the 

two grazing intensity treatments were past intensities, 

not current intensities as implied.

Jornada tobosa/burrograss community—Seral 

stages on various clay soils of the Jornada were studied 

from 1915 to 1929 by Campbell (1931). Basal cover was 

assessed after four years of above average precipitation, 

thereby representing species composition under favorable conditions for growth. A sodgrass, tobosa grass is 

the late seral-stage dominant or sub-dominant of clay 

soils on the Jornada. When not dominant it is often still 

the most important forage species. Tobosa is palatable 

to livestock only when green, and therefore, this plant 

community is often grazed during the growing season. 

This allows for deferment of grazing on black grama 

communities, which increases black grama because of 

its sensitivity to grazing during reproductive stages, and 

higher palatability during the dormant season. Campbell 

presents species composition and palatability of the four 

main seral stages but does not quantitatively address 

grazing treatments. A burrograss association is the second lower stage after the tobosa stage. Drought and too 

heavy of a grazing intensity reduce basal areas of both 
50 USDA Forest Service RMRS-GTR-169. 2006.

species, but tobosa recovers more slowly. Campbell 

suggested that tobosa could be grazed to 60 percent 

utilization without injury or lowering of seral stages 

in adjacent recovering communities. This high rate of 

utilization would suggest that previous very heavy grazing during the uncontrolled grazing of the late 1800s to 

early 1900s must have been extremely heavy to push 

communities to an annual, ruderal-weed stage.

Common, periodic cycling of tobosa grass through 

drought and relatively wet periods is similar to that for 

black grama described above, but fluctuations were reported as less pronounced by Paulsen and Ares (1961, 

1962) as well as Gardner (1950, see below). Somewhat 

different conclusions were reached by Herbel and others 

(1972) during the extreme drought of the 1950s. Tobosa 

and burrograss were reported to be little affected by the 

drought on flood plains sites. They indicate a capacity 

of tobosa to become completely dormant as soil moisture approaches the wilting point.

The Paulsen and Ares (1961, 1962) data showed 

highest basal cover of tobosa on intermittently grazed 

plots rather than conservatively grazed plots as for black 

grama. Average basal areas were 191, 461, 718, and 

294 cm2

/quadrat for protected, conservatively, intermittently, and heavily grazed plots, respectively. Protection 

from grazing resulted in accumulation of old-dead plant 

material in the clumps that resulted in “stagnation,” 

whereas heavy grazing resulted in thinning of the stand 

and increases in bare ground cover. Less palatable dropseeds were a larger proportion of the community under 

heavy grazing, and lower seral-stage burrograss was a 

larger proportion during dry cycles.

The 64-year permanent quadrat data for the Jornada 

Experimental Range included quadrats in tobosa and 

burroweed (Gibbens and Beck 1988). Although only one 

quadrat per type was located in an exclosure, no differences between grazing treatments were noted. Unlike 

the effect on the black grama or poverty threeawn type 

(see above), the severe drought of the 1950s did not result in an apparent permanent reduction in perennial 

grass basal cover in either type. Tobosa remained on all 

but three of the original 22 quadrats over the 64-year 

period, and shrubs became dominant on one. Tobosa 

retained reasonable cover during the drought and recovered rapidly after the drought. Burrograss maintained 

more cover through the drought than other perennial 

grasses and recovered rapidly following the drought. 

Of 12 quadrates, burrograss remained present on all 

but two over the 64 year period, and shrubs became 

dominant on only one. The tobosa and burrograss communities of heavier soils were more tolerant of drought 

and recovered more rapidly than the black grama or 

poverty threeawn types on the sandier soils. Herman 

(1988) sampled two 11-year-old, and a 29-year-old, exclosures in lowlands and found no difference in species 

frequency or richness with adjacent grazed areas.

Santa Rita Arizona mesa mesquite grassland—These 

plant communities occupy sandy loam soils receiving an 

average of 343 mm/yr precipitation, and are the lower 

elevation, drier of the two grassland types on the experimental range (Canfield 1948). Shrub invasion into these 

grasslands has also been very extensive (Brown 1950; 

Mehrhoff 1955). The Santa Rita Experimental Range is 

located in southeastern Arizona, south of Tucson.

Before focusing specifically on the quantitative data 

from the Santa Rita Range, we summarize the qualitative 

descriptions of range condition provided by Humphrey 

(1960) for sites in the same general area as the Santa 

Rita. Humphrey considered the extensive invasion of 

mesquite, burroweed, and cholla and prickly pear cactus 

into grasslands to be a result of too heavy of a grazing 

intensity and fire control. Three distinct sites were differentiated as valley-bottomlands, high-potential sites, 

and medium-potential sites. For valley-bottomlands, a 

decline from excellent to good condition was displayed 

by decreases in blue grama, sand dropseed, bluestems, 

and cane beardgrass with increases in Rothrock grama. 

Annual grasses such as six-weeks needle grama and 

six-weeks threeawn, and burroweed shrubs, are indicative of fair condition range. Erosion is evident, and even 

more so on poor condition range where even annual 

grasses are sparse. An excellent condition high-potential 

site supported sideoats, slender, hairy, and black gramas 

intermixed with mid-grasses like Arizona cottongrass, 

plains beardgrass, and plains lovegrass. Density of 

these grasses declines on good condition range and are 

replaced with increasing amounts of Rothrock grama, 

Santa Rita threeawn, burroweed, mesquite, and cholla. 

Poor condition range has only annual grasses and a well 

developed shrub component. The sequence of declining 

condition on medium-potential sites is somewhat similar, except the shrub component is made up of guajilla, 

various cacti, ocotillo, and desert hackberry.

On the Santa Rita, examples of general composition 

of plant communities with different grazing treatments 

based on extensive sampling of a large number 

of exclosures and grazed sites were presented by 

Canfield (1948). Although grazing intensities were 

not provided, large differences in species composition 

were observed between heavily grazed sites and those 

either conservatively grazed, protected for five years, or 

protected for 25 years (Table 7). The dominant species 

on heavily grazed sites was Rothrock grama, while black 
USDA Forest Service RMRS-GTR-169. 2006. 51

grama, and Arizona cottongrass shared dominance on 

long-term protected sites. The most uniquely different 

community among the four was the heavily grazed 

treatment. Black grama did not respond exactly the 

same as in the Jornada grasslands, where greatest covers 

were often found under conservative grazing and least 

under heavy grazing. It is not clear as to whether this 

was just because data are presented as composition and 

not absolute cover, but Canfield comments that black 

grama’s greater composition on heavily rather than 

conservatively grazed sites was because it was less 

susceptible to either very heavy grazing or drought. The 

highest composition of black grama was on long-term 

protected sites. Black grama was considered to increase 

after heavy grazing at greater rates under protection 

than under conservative grazing. At the Jornada black 

grama grasslands, the species was able to withstand 

drought better under conservative grazing compared 

to complete protection. Canfield does conclude that 

other species do almost as well under conservative 

grazing as under complete protection. The very high 

composition of Rothrock grama on deteriorated ranges 

was attributed to high seedling establishment during 

conditions of either high precipitation and/or low levels 

of competition from longer-lived, better competitor 

species of later seral stages. Arizona cottongrass 

was considered a good indicator of range condition, 

becoming depleted and confined to shrub canopies 

when range is very heavily grazed, and conspicuously 

increasing under conservative grazing or protection. A 

low abundance of bush muhly was another indicator of 

very heavy grazing. Shrubs were not considered a good 

indicator of range condition because of their lack of 

response to protection. Among Canfield’s conclusions 

were 1) rates of recovery are approximately equal under 

conservative grazing or protection if the range has not 

been abused for a long time and erosion is not excessive; 

2) the tall grasses such as Arizona cottongrass, sideoats 

grama, and black grama are major components of 

“climax” communities; and 3) deep-rooted shrubs that 

have invaded the grassland may be slow to respond to 

conservative grazing or protection. Once established 

“they become a relatively permanent part of the plant 

cover,” even though the grass understory improves with 

the protection. However, field observations suggested a 

more rapid invasion under heavy grazing.

The extreme expansion of shrubs into grasslands of 

the Santa Rita prompted a study of the effects of protection from cattle as well as protection from both cattle 

and rabbits-rodents (Brown 1950). The rodent exclosure 

was for rabbits and larger rodents, since chicken wire 

was used as fencing. The densities of 12 shrub or cacti 

species were sampled initially (in 1931), 10 years and 18 

years after fencing. Mesquite increased under all three 

treatments. However, the rate of increase was greater in 

the grazed community, and similar under cattle only and 

cattle plus rodent protection. Over the 18 year period, 

Table 7—Grass species composition and shrub cover for mesa-type mesquite savannas heavily 

grazed, conservatively grazed for five years, and protected for five or 25 years. Adapted from 

Canfield (1948).

Conservatively Protected Protected

Heavily grazed grazed 5 years 5 years 25 years

Grass species Composition (percent)

Threeawn species 10 23 24 14

Slender grama 5 11 0 1

Black grama 15 8 16 26

Rothrock grama 64 8 7 7

Sideoats grama trace 3 2 2

Tanglehead 3 2 7 3

Arizona cottongrass 2 22 19 26

Bush muhly trace 12 16 11

Other grasses 1 11 9 10

Shrub species Cover (percent)

Mesquite 4.7 9.8 9.2

False mesquite 0.1 0.2 2.1

Burroweed 8.2 9.0 6.6

Cacti species 1.5 0.7 2.7

Other shrubs 3.8 2.0 2.8
52 USDA Forest Service RMRS-GTR-169. 2006.

burroweed increased 248 percent in the grazed community, increased 189 percent under protection from cattle, 

and decreased 42 percent under protection from both 

cattle and rodents. Brown concluded: “The increase of 

all shrubs except burroweed under all treatments are at 

variance with the theory that grazing has directly affected the spread and increase of noxious plants in the 

desert grassland through breakdown of sod and release 

of shrubs from grass competition. Instead, it would 

seem that the shrubs, rather than the grass, were natural 

dominants of the area, and that the grass was present 

because of some factor that was unfavorable to the 

shrubs.” Brown suggested studies to determine whether fire was the factor missing from the current system. 

Even burroweed can eventually become dominant over 

large areas of grassland. The Santa Rita was ungrazed 

from 1903 until 1915, and burroweed spread over large 

portions of grassland during that period (Griffiths 1910, 

cited in Mehrhoff 1955).

A similar study to that of Brown (1950) above was 

conducted by Glendening (1952) on apparently different sites (two replications instead of one, slightly lower 

elevation and precipitation, sampling in 1932, 1935, and 

1949). In this study, mesquite more than doubled on 

treatments grazed moderately year-long, ungrazed by 

cattle for 17 years, and ungrazed by cattle, rabbits, and 

rodents for 17 years. However, increases in mesquite 

were slightly greater in both types of exclosures than in 

the grazed treatment. Since all shrubs were mapped on 

all plots, the establishment and mortality of individuals 

could be distinguished, rather than just density or increases in crown cover of the same individuals, thereby 

providing information concerning reproduction from 

seed. Cattle consume some mesquite seeds, and these 

could be transported to uninhabited grassland and germinate after passage through the animal. Small rodents 

could potentially cache seed that would then establish 

new individuals. New mesquite individuals were recorded on 70 percent of the grazed treatment and 79 percent 

of each of both exclosure treatments. Death loss overall 

was only 0.3 percent. The higher establishment under 

the two grazing exclosure treatments suggest that cattle 

were not an agent of dispersal in this case, but some very 

small rodents may have been able to get through or over 

chicken-wire-mesh size material. Glendening comments 

that the Merriam kangaroo rat is of a small enough size 

to get through the fencing and that it is known to bury 

seeds in a pattern whereby seedlings emerge in clusters 

of 2 to 10 or more individuals. Many such clusters were 

observed on all treatments, suggesting the kangaroo rat 

may be an important means of dispersal. Initial perennial grass densities differed among treatments, and a 

drought during 1935 thinned all stands. Differences in 

the percent loss in density due to the drought among the 

treatments would be difficult to attribute to treatments, 

since initial densities differed, and studies reported 

above for Jornada grasslands (Paulsen and Ares 1961, 

1962) indicated drought reduced basal covers to similar levels regardless of grazing treatment. Sampling 

was done only during the 1935 drought and again in 

1949, so the immediate post-drought conditions are 

not known. Interestingly, however, densities remained 

low long after the drought. This is similar to the lower perennial grass cover reported for the Jornada after 

the 1935 drought (see above Gibbens and Beck 1988). 

Glendening speculated that the increases in mesquite 

and/or grazing by rodents may have been responsible 

for the continued low levels of grasses. Rothrock grama 

was initially the dominant grass at the start of the experiment, but it was not at the end in the protected sites. 

Increases in cactus accompanied the increases in mesquite and decrease in grass abundance. Grass density at 

the final sampling was 9.5, 5.2, and 14.3 cm2/m2 for the 

grazed, cattle exclosure, and cattle plus rabbit/rodent 

exclosure, respectively, from initial values of 287, 214, 

and 146 cm2/m2. Glendening concluded that mesquite 

increases regardless of grazing treatment, and moderation in grazing would not prevent loss of grass cover.

Martin and Cable (1974) assessed effects of different seasons of grazing on grasslands of the Santa Rita 

(see management section below). Some sampling was 

also done in relation to a distance from water grazing 

intensity gradient. Sampling distances ranging from 

only 1/4, 5/8, and 1 mile did not result in large differences in utilization (48, 44, and 43 percent), but some 

differences in community species composition were 

noted. Jumping cholla and Rothrock grama were more 

abundant nearest water, while black grama was twice 

as abundant distant to water. Cholla is a tall cactus that 

can establish from branches that are knocked off and 

transported by cattle.

Santa Rita mixed-grama grassland—This grassland is generally similar to mesa mesquite grassland, 

but cacti and half-shrubs are rare, cover of grasses is 

double, and the relative composition of the grass species changes (Canfield 1948). Precipitation (470 

mm/yr) and elevation is higher than for mesa mesquite 

grassland, and soils are darker with higher organic matter content.

Changes in this community type with heavy grazing, 

conservative grazing, and short-term or long-term protection are somewhat similar to that described above for 

mesa mesquite grassland (Canfield 1948; Reynolds and 

Martin 1968). Differences include the dominance of 
USDA Forest Service RMRS-GTR-169. 2006. 53

depleted heavily grazed communities by slender grama, 

although long-term protected mixed-grama grassland 

also has a large component of slender grama (Table 8). 

Deteriorated range is in general characterized by shortlived grasses that reproduce abundantly by seed or 

short, sod-like grass that avoid herbivory because of the 

prostrate morphology. Long-term protected grassland is 

characterized by a dominance of Arizona cottongrass 

and also by a diversity of a number of other grass species that comprise small amounts of the composition on 

their own. A difference between this type and the mesa 

mesquite grassland is the difference between community species composition of short-term conservatively 

grazed and short-term protected sites. While there was 

little difference between the two treatments for mesa 

mesquite grassland, rates of recovery were greater under protection than under conservative grazing for the 

mixed-grama type. Mesquite did not respond to protection from grazing, although values for the heavily 

grazed treatment were not presented. There was some 

indication that burroweed declines with long-term protection. Canfield concluded that shrubs respond little 

to grazing management or protection once established, 

but he stated that field observations suggested a more 

rapid invasion of mesquite and burroweed under heavy 

grazing. Later consideration of a long-term data set 

for the Santa Rita indicate that burroweed and cactus 

abundance appear to be independent of cattle grazing 

intensity, and fire influenced only short-term dynamics 

(McClaran 2003).

Grazing intensity at the Santa Rita has declined 

through time, as perceptions of what was “conservative 

grazing” changed. When grazing intensities declined 

after 1942 to 52 to 58 percent utilization, and to 42 to 

49 percent utilization from 1962 to 1966, a substantial 

increase in perennial mid-grass composition was observed (Rivers and Martin 1980). The increases ranged 

from 14 to 35 percentage units, depending on site elevation/precipitation. Short grasses were the component 

that decreased, particularly for the short-lived Rothrock 

grama which showed up to a one-third reduction. 

Based on previous studies and additional more recent 

data, McClaran (2003) concluded that “grass dynamics appear to be more sensitive to varying intensities of 

livestock grazing and neighboring plants than the dynamics expressed by mesquite, burroweed, and cactus. 

The spread of the exotic sweet resin bush (hawk’s eye) 

was also not related to light to moderate grazing or ungrazed treatments (Howery and others 2003). Excellent 

reviews, more detailed than can be provided here, of 

the long-term monitoring of vegetation on the Santa 

Rita are found in McClaran (2003) and Ruyle (2003). 

The impact of rodents on vegetation is considered in the 

following section on Mesquite Savanna Conversion to 

Chihuahuan Desert, and in a review wildlife research 

on the Santa Rita (Krausman and Morrison 2003).

Lehmann lovegrass communities—Lehmann lovegrass is an exotic from South Africa introduced to 

southern Arizona in 1932 to prevent erosion during road 

construction and to provide forage for livestock after 

Table 8—Grass species composition and shrub cover for foothills-type mesquite savannas heavily 

grazed, conservatively grazed for five years, and protected for five or 25 years. Adapted from 

Canfield (1948).

Conservatively Protected Protected

Heavily grazed grazed 5 years 5 years 25 years

Grass species Composition (percent)

Threeawn species 5 11 15 9

Slender grama 49 44 27 15

Black grama 5 5 10 7

Hairy grama 2 3 3 3

Rothrock grama 12 3 2 2

Sideoats grama 6 14 18 10

Curlymesquite 13 4 2 1

Tanglehead 1 2 2 5

Arizona cottongrass 6 6 11 31

Other grasses 1 8 10 17

Shrub species Cover (percent)

Mesquite 6.63 8.28 6.42

False mesquite 4.06 4.95 5.92

Burroweed 2.81 2.10 0.38

Cacti 0.54 0.09 0.88

Other shrubs 0.51 4.31 4.93
54 USDA Forest Service RMRS-GTR-169. 2006.

loss of native grasses to drought or very heavy grazing 

(McClaran and Anable 1992; Robinett 1992). The species is well adapted to the environment and has more 

than doubled in area from that originally seeded. The 

species utilizes moisture stored in soil during winter 

more effectively than native species (Cable 1971) and 

produces large amounts of seed and maintains a large 

soil seed bank. The species provides an early spring 

forage, is very productive through the summer, and recovers rapidly from drought, fire, and grazing. Seeds 

are too small to be utilized effectively by small mammals and germinate profusely after canopy removal by 

fire or grazing. Diversity of birds and small mammals is 

lower in the near-monocultures of Lehmann lovegrass 

compared to native grassland (Bock and others 1986). 

Lehmann lovegrass is perceived as a major threat to native grasslands because it displaces natives and heavily 

dominates areas, but it may also be the only perennial 

grass capable of inhabiting areas invaded by shrubs and 

providing soil cover and forage for domestic and native 

consumers (Robinett 1992).

Robinett (1992) described the response of Lehmann 

lovegrass and native species to a drought in southern 

Arizona during winter 1988 to spring 1989. One-third to 

one-half of the area had dead patches of grass, particularly where the surface sandy loam soils were shallowest 

over a clayey subsurface. Utilization of Lehmann lovegrass and black grama appeared similar, and did not 

seem to affect mortality that also was similar between 

the two species. A transect read after rains fell indicated 

high germination of Lehmann lovegrass seedlings in 

the dead grass patches, and no germination of native 

species. One year after the drought, Lehmann lovegrass 

completely dominated the former dead patches. The native black grama, red threeawn, mesa threeawn, hairy 

grama, and sprucetop grama all declined at the expense 

of Lehmann lovegrass, with only aparejograss unaffected by the drought. Similar displacements were observed 

on the Santa Rita Experimental Range in response to 

fire (Cable 1965, 1971; reviewed by McClaran 2003). 

The fire killed 98 percent of the Lehmann lovegrass 

and 90 percent of the black grama, but only Lehmann 

lovegrass germinated immediately following the fire. 

Once established, Lehmann lovegrass soon became the 

dominant on the site.

The relationship between grazing intensity and 

Lehmann lovegrass abundance and rate of spread was 

assessed at the Santa Rita Experimental Range by 

McClaran and Anable (1992). Distance from water as 

a grazing intensity gradient, and ungrazed exclosures, 

were used to monitor native grass and Lehmann lovegrass density from 1972 to 1990. Lehmann lovegrass 

increased over time equally under all distance-fromwater grazing intensities or protection from grazing. 

Although grazing had no effect on Lehmann lovegrass, 

native grasses decreased at higher grazing intensities, 

whereby the relative proportion of natives to exotics decreased. Angell and McClaran (2001) reported on the 

same study but after another 10 years (1972 to 2000). 

They also examined differences between grazing management of frequent summer rest versus grazing every 

summer and at a heavier stocking rate (see Grazing 

Management section below). Although native grass 

density declined with increasing grazing intensity after 

18 years, there was no relationship with grazing intensity and native grass density after 28 years. Lehmann 

lovegrass showed no relationship to grazing intensity. 

There was no evidence that the faster, preemptive establishment of Lehmann lovegrass after drought or fire 

had changed over time. Some self-thinning had occurred, but the dominance by this species was expected 

to continue.

At the Page-Trowbridge Experimental Ranch in 

southern Arizona, Schmutz and Smith (1976) found 

greater basal cover of Lehmann lovegrass in areas protected from grazing for 28 years than areas “closely” 

grazed. In contrast, Brady and others (1989) reported 

greater canopy cover of Lehmann lovegrass in grazed 

compared to ungrazed communities at AppletonWhittell Research Ranch in south-central Arizona. 

However, Lehmann lovegrass was only a minor component of the plant community at both of these study plots. 

Bock (J. H., personal commun.) notes that lovegrass 

more permanently occupies sites following disturbance, 

compared to native, short-lived, bare-ground colonizers that play an important role in stabilizing a site after 

disturbance yet readily give way to natives of the next 

seral stage.

The invasion of native grasslands by Lehmann 

lovegrass does not appear to be affected by grazing intensity or whether a site is protected from grazing, or 

by specific grazing management systems (see Grazing 

Management section below).

Page-Trowbridge Experimental Ranch—Grazing at 

the Page-Trowbridge Experimental Ranch in southern 

Arizona was conservative from 1923 to 1941 and was 

ungrazed after that time (Schmutz and Smith 1976). 

An adjacent property was “overgrazed” at the time of 

Haskell’s (1945) first sampling in 1941. The “severelygrazed” site showed more erosion to the point that its 

runoff water deposited soil on a part of the conservatively grazed area. Both areas were pure grassland at 

one time, but by 1941 supported a community with mesquite, burroweed, and snakeweed. Large differences in 
USDA Forest Service RMRS-GTR-169. 2006. 55

species composition were observed between the two areas. Perennial three-awns comprised 45 percent of the 

basal cover of the conservatively grazed site and only a 

trace on the overgrazed site. Conversely, Rothrock grama 

comprised 38 percent of the overgrazed site and 17 percent of the conservatively grazed, and respective values 

for the non-palatable fluffgrass were 13 and 1 percent. 

Even though a greater proportion of the heavily grazed 

community, Rothrock grama attained greater cover on 

the conservatively grazed site. Once established, burroweed did not respond to a change to conservative 

grazing and remained at similar crown covers on the 

two sites. Mesquite was approximately the same density 

on both sites. Smith and Schmutz (1975) and Schmutz 

and Smith (1976) re-sampled the same sites in 1969. 

Actual grazing intensities were not provided but described as still conservative on the Research Ranch and 

“closely” grazed on the adjacent land. Some of the 1941 

data from Haskell were included, but differences in values suggests that only a portion of the original 8 blocks 

were sampled. Total cover increased on both sites from 

1941 to 1969 (Table 9). Differences between the treatments that changed since 1941 included the appearance 

of Arizona cottontop and the exotic Lehmann lovegrass 

with a much greater cover and composition on protected 

sites and a greater composition of poverty threeawns on 

grazed than on protected sites. With the exception of the 

exotic Lehmann lovegrass and possibly Arizona cottontop, differences between years may be due to temporal 

Table 9—Plant community characteristics for an Arizona mesquite savanna that had been lightly grazed from 

1923 to 1941 and the same area ungrazed from 1942 through 1969 or still heavily grazed through 1969. 

Adapted from Schmutz and Smith (1976).

Basal cover (percent) Composition (percent)

1941 1969 1941 1969

Grass species Light Heavy Ungz. Heavy Light Heavy Ungz Heavy

Arizona cottontop 0 0 0.07 0.08 0 0 23.3 7.1

Sideoates grama 0.12 0 0.53 0.01 7.9 0 17.7 0.9

Cane beardgrass 0.18 0.02 0.27 0.02 11.8 16.7 9.0 1.8

Poverty threeawns 0.15 T 0.37 0.32 9.9 T 12.3 28.6

Rothrock grama 0.30 0.06 0.35 0.52 19.7 50.0 11.7 46.4

Red threeawn 0.65 T 0.34 0.07 42.8 T 11.3 6.2

Lehmann lovegrass 0 0 0.30 0.01 0 0 10.0 0.9

Others 0.12 0.04 0.14 0.09 7.9 33.3 4.7 8.1

Total 1.52 0.12 3.00 1.12 100.0 100.0 100.0 100.0

Shrub species Crown cover (percent) Composition (percent)

Velvet mesquite 0.66 0.76 3.80 5.32 17.4 8.9 47.7 55.0

Burroweed 1.96 1.69 1.69 2.30 52.1 74.9 21.2 23.8

Wright buckweed 0.64 0.02 2.48 0.16 16.8 0.2 31.1 1.6

Sticky snakeweed 0 0.84 T 1.10 0 9.9 T 11.4

Others 0.52 0.52 T 0.79 13.7 6.1 T 8.2

Total 3.80 8.52 7.97 9.67 100.0 100.0 100.0 100.0

trends or to abiotic conditions during or just prior to the 

years of sampling, because only one year was sampled 

by each author.

Appleton-Whittell Research Ranch grama grass—

The Appleton-Whittell Research Ranch in southeastern 

Arizona has been an ungrazed National Audubon sanctuary since 1968 (Bock and Bock 2000). Adjacent 

moderately grazed sites are available for comparative 

purposes and are generally grazed moderately (Bock 

and Bock 1993). Precipitation is higher at this site compared with those above. Canopy cover is high at 86 

percent in ungrazed grassland, with a large amount of 

grass cover of 68 percent (Brady and others 1989). The 

flora of the site is mostly comprised of Madrean and 

Apachian elements similar to floras to the southeast in 

Mexico and southwestern Texas, even though the area 

has previously been classified as Plains and Great Basin 

Grasslands (McLaughlin and others 2001). Only 7.4 

percent of the flora are exotics, over half of which are 

grasses.

A comparison of initial conditions at the time of the 

removal of livestock from the sanctuary with conditions 

16 years later, as well as a cross-fence grazed versus 

ungrazed comparison after 16 years of protection, was 

conducted by Brady and others (1989). Total vegetation cover more than doubled over the 16 year period, 

and species richness tripled. However, neither cover nor 

species richness differed when compared across the 

fenceline in the same year. The very high current values 
56 USDA Forest Service RMRS-GTR-169. 2006.

on both grazed and ungrazed sites suggests that factors 

other than grazing were also responsible for the temporal changes. Blue grama increased the greatest over the 

period, and cover was similar on grazed and ungrazed 

areas along the fenceline. The relatively tall plains lovegrass increased greatly in both areas. Shrubs increased 

over time but were a small component of total cover and 

did not differ between the grazed and ungrazed sites. 

Based on the fenceline comparison, differences solely 

attributable to grazing were the increase in mid-grass 

cover (primarily plains lovegrass) and changes in forb 

composition.

Total vegetative cover in 1981 to 82 was 106 percent in the ungrazed compared to 73 percent in grazed 

sites (Bock and others 1984). Species composition did 

not vary greatly between treatments, but there were 10 

species that differed significantly with grazing (Table 

10). The dominant on both grazing treatments was blue 

grama, but it did not significantly differ with grazing. 

Sideoats grama was the second most abundant species 

on both sites and did not differ between treatments. This 

is contrary to most studies elsewhere showing increases in blue grama with grazing, but consistent with the 

Brady and others (1989) study at the same ranch. The 

cover of total grasses and total forbs was significantly 

greater on the ungrazed side of the fences, with values 

for forbs more than double that on the grazed portion of 

the study area. Total shrub cover did not differ between 

grazed and ungrazed sites. Separate measurements of 

shrub densities, heights, and maximum canopy widths 

did show some differences with grazing. Three out of 

five shrub species attained greater heights and canopy 

widths in ungrazed treatments. However, there were 

no significant differences between treatments in density, height, or width of mesquite, which occurred at an 

average of only 1 individual/400 m2. Burroweed was 

also not significantly different between grazed and ungrazed sites, although individual plants were larger in 

the ungrazed site. Bock and Bock (2000) indicated that 

burroweed population dynamics is much more related 

to fire and weather than to livestock grazing. Bock and 

others (1984) commented that the lack of response to 

grazing by the dominant grama grasses may be due to 

their co-evolution with large native herbivores (Mack 

and Thompson 1982; Milchunas and others 1988), 

but that other species were able to co-exist within the 

grazing resistant matrix of grama grasses within the 

sanctuary.

Eight cross-fence sites were intensively sampled for 

canopy cover by Bock and Bock (1993) in 1990, after 

22 years of protection from grazing in the sanctuary. 

Canopies were well developed at the time of sampling, 

and cattle were not yet present on grazed sites, whereby 

measurements were not confounded by current 

consumption (see Introduction section on methods). 

Total grass canopy cover was again greater without 

grazing. Out of ten grasses, the two stoloniferous grasses 

showed the greatest cover on grazed sites and the three 

tallest bunchgrasses showed the greatest increases with 

protection from grazing. A good relationship between 

“grazing release ratio” (ungrazed canopy cover divided 

by grazed canopy cover) and culm height (an index of 

undisturbed plant height) was observed for the 10 main 

species, indicating that taller species responded more 

to grazing than short species. While in agreement with 

global analyses (Milchunas and Lauenroth 1993), this 

is at the same time unusual because the species that 

increased was the generally grazing intolerant black 

grama. Blue grama, a general increaser with grazing in 

the Great Plains, was greater in ungrazed than grazed 

sites, although it was a greater composition of the 

grazed. Some very interesting points were raised by the 

Table 10—Basal cover (percent) of grama grassland in 

southeastern Arizona ungrazed for 14 years or grazed 

moderately. Adapted from Bock and others (1984).

Basal cover (percent)

Ungrazed Grazed

Grasses-total 80.4** 55.6

Cane bluestem 1.3 1.3

Reverchon three-awn 1.0 0.3

Purple three-awn 7.4** 2.1

Spider grass 4.6* 2.5

Sprucetop grama 1.5 3.1*

Sideoats grama 14.0 15.6

Hairy grama 3.5 4.5

Blue grama 22.0 20.1

Plains lovegrass 6.1** 0

Curly-mesquite 0.4 2.0*

Common wolfstail 8.1* 2.4

Vine mesquite 3.6** 0.6

Arizona cottontop 4.3* 0.1

Herbs-total 12.0** 5.6

Field bindweed 1.4 0.5

Leatherweed 3.5 2.0

Trailing fleabane 1.3 1.1

Silver dwarf morning-glory 1.6* 0.1

Woody plants-total 13.5 11.5

Yerba de pasmo 2.0 1.0

Cooley’s bundleflower 6.0* 3.6

Rayless tansyaster 1.0 1.9

Shrubby false mallow 3.0 3.3

Bare ground 17.6** 34.6

* Significantly greater than other treatment (P<0.05)

** Significantly greater than other treatment (P<0.001)
USDA Forest Service RMRS-GTR-169. 2006. 57

Bocks. First, they questioned the classification of this 

area as a grama grassland but as being an “artifact of 

over 100 years of livestock grazing.” Only one of five 

taxa that was a grama grass increased in response to 

release from grazing, and the study site is in a location 

that was not subjected to 10,000 years of grazing by 

bison that occurred in adjacent grasslands of the Great 

Plains. Second, they commented on the differences in 

response of their Arizona grassland with blue grama as 

a dominant and the shortgrass steppe of the Great Plains 

that is also dominated by blue grama. Changes in plant 

species composition were much greater for this study 

compared with that found in the Colorado shortgrass 

steppe by Milchunas and others (1989). This is confirmed 

by calculation of dissimilarity indices for the two studies. 

The average community dissimilarity (Whittaker 1952) 

for the eight sites in the Bocks study is 32 percent, 

compared with an average of 10 percent dissimilarity 

between ungrazed and grazed sites for the Milchunas 

study. The Bocks considered one possibility for the 

greater responsiveness of the Arizona site to protection 

from grazing was the greater precipitation (430 vs. 310 

mm/yr), whereby tall and mid-height species were more 

capable of increasing with protection at the Arizona site 

compared with abioticly maintained short grasses at the 

shortgrass steppe site. The Bocks also considered the 

possibility that, although the same species, blue grama 

responds differently to grazing in areas where it was not 

subjected to bison grazing for 10,000 years compared 

with areas where it was subjected to bison grazing. This 

is a very important consideration that will be examined 

in greater detail in the conclusions section below, based 

on responses of blue grama from studies throughout 

this review. Third, the Bocks raised the question as to 

whether protected areas may respond differently when 

they are large areas such as the sanctuary (3160 ha) 

compared to the generally small size of exclosures. 

Bock and Bock (2000) suggest that changes may still 

be taking place on the ungrazed sanctuary and that 

eventually taller grasses may displace blue grama and 

other shorter grasses.

Other locations—As with the Appleton-Whittell 

Research Ranch, black grama and blue grama co-occur 

at the Sevilleta National Wildlife Refuge Long-Term 

Ecological Research site, located in central New 

Mexico at a biome transition zone between Great Plains 

grasslands, the Chihuahuan Desert, and the Colorado 

Plateau shrub-steppe (Gosz and Gosz 1996; see Fig. 

2). Precipitation averages only 222 mm/yr in grassland 

communities. Domestic grazers were removed from 

this site when it was established as a refuge in 1973, 

and black grama has been increasing compared with 

cross fence visual observations and five years of data. 

Increases in blue grama were also observed, but they 

were not as great as that for black grama. A controlled 

clipping, burning, and native herbivory experiment 

demonstrated that black grama was more sensitive to 

all three factors than was blue grama in terms of plant 

height growth and production. Native herbivores did 

not, however, significantly affect productivity. The data 

suggested that the absence of both burning and grazing on the refuge could be attributed to the increases in 

black grama. Black grama also responded more to wetdry conditions than did blue grama.

A study established in 1976 and resampled in 1986 

and 1996 at the Sevilleta also showed very large temporal dynamics in plant communities, especially in 

the black grama communities (Ryerson and Parmenter 

2001). Although landscape-scale analyses showed no 

differences between inside and outside refuge boundaries between 1976 and 1996, site specific analyses inside 

the refuge showed stable perennial plant cover and 

increases in annual forb and litter cover while communities outside the refuge showed declines in perennial 

plant cover. However, changes in community similarities of the black grama communities from 1976 to 1996 

inside the refuge ranged from just under 80 to nearly 100 

percent different (5 sites), while those outside displayed 

a greater than 95 percent change over time. Temporal 

dynamics were extreme compared to herbivore regulated dynamics. The authors noted a very interesting black 

grama - blue grama relationship. Sites dominated by 

blue grama in 1976 were dominated by black grama in 

1996, and this occurred outside the refuge in the presence of livestock grazing. Blue grama, the ubiquitous 

increaser with grazing throughout the Great Plains, 

and a grazing tolerant species in the southwest, and the 

grazing intolerant black grama, must have responded 

more to climate or some other factor than to the presence of grazers. Perennial plant cover in the black 

grama community decreased over time in the presence 

of herbivores, while remaining the same in the absence 

of the grazers. In contrast, Burrograss communities increased in perennial grass cover outside the refuge and 

decreased inside the refuge. Short-term protection from 

grazing showed few effects, but these effects increased 

in importance with longer-term protection.

Blue grama also dominated at an upland site in 

southwestern New Mexico that had been protected for 

28 to 30 years or very heavily grazed (Gardner 1950). 

Precipitation at this site was approximately 370 mm/yr. 

Blue grama, ring muhly, and total grass basal cover 

(not density as the author calls it) in the protected area 
58 USDA Forest Service RMRS-GTR-169. 2006.

were at least twice as much as in the heavily grazed area 

(Table 11). Black grama, tobosa, and red threeawn were 

slightly greater in the grazed site. Although cross-fence 

tobosa lowlands in a grazed area were not available for 

direct comparison, swales in the grazed areas were cut 

by steep-sided arroyos while those in the protected area 

had some plant cover stabilizing the soils, but were predicted to require much more than 28 years to heal. Two 

years later after a drought in 1947, resampling of the 

same area showed a decline in total grass basal cover 

of 75 percent inside the exclosure compared to 78 percent outside. Although declines in cover due to drought 

were similar with either protection or very heavy 

grazing, some changes in species distributions were evident post-drought. Black grama became more abundant 

in the protected site and forbs increased greatly everywhere and were also greater in the protected site. The 

decrease in cover due to drought in the tobosa community was only about 50 percent. The greater resistance 

to drought of lowland tobosa communities compared to 

uplands agrees with those reported above for Jornada 

tobosa/burrograss communities.

Only one of the Potter and Krenetsky (1967) plots 

was in mesquite savanna. Although they report two in 

desert grassland, one was a site dominated by mesquite. 

Both sites showed an increase in grass cover and a decrease in shrub cover with protection. This is one of the 

few examples where mesquite was reported to increase 

greatly under grazing while decreasing under protection. Species richness increased over the 25 years in the 

exclosure, as did the cover of black grama. Total cover 

increased 60 percent and grass cover increased from 23 

to 70 percent of the vegetation in the exclosure.

Navarro and others (2002) estimated range condition 

during seven years spanning 1952 to 1999 at 41 BLM 

sites in six counties across southwestern New Mexico. 

Average grazing intensity at the sites was a conservative 34 percent utilization, and average precipitation 

ranged from 260 to 350 mm/yr. Range condition fluctuated greatly with fluctuations in precipitation but did 

not differ significantly between beginning and end of the 

sampling period. Cover of black grama, tobosa grass, 

and mesquite was not different between beginning and 

end of the sampling period, except mesquite increased 

dramatically on one of the 41 sites and creosote-bush on 

another site. The authors concluded that conservative 

grazing was sustainable in this part of New Mexico. 

Of the 41 sites during the 1998, 1999, and 2000 period, 

the percentage of sites grazed severely, heavily, moderately, conservatively, and lightly and ungrazed were 1, 

12, 35, 19, 28, and 5 percent, respectively, which is an 

approximately 50:50 split between what was considered 

sustainable and not.

Dutchwoman butte in central Arizona supports the 

only relict, geologic refuge for mesquite savanna available for this review. Precipitation was estimated at 

432 mm/yr, with 60 percent falling during the winter 

(Ambos and others 2000). Precipitation during the twoyear study period was above normal, but a long-term 

period of below average precipitation preceded sampling. The site is at the cool/moist end of the semiarid 

grassland gradient and is also at the transition between 

grassland and higher elevation woodlands of redberry 

juniper and turbinella oak. Dramatic differences were 

found between the plant community on the butte and 

surrounding grazing allotments on Tonto National Forest 

Table 11—Basal cover of herbaceous and canopy cover of shrubs (percent) and number 

of forbs/30.5 m transect in ungrazed and very heavily grazed grama grassland in 

southwestern New Mexico. Pre-drought was 28 years after protection and post-drought 

31 years after. Adapted from Gardner (1950).

Pre-drought Post-drought

Grasses Ungrazed Heavily grazed Ungrazed Heavily grazed

Red threeawn 0 0.23

Sideoats grama 0 0

Black grama 0.16 0.24 0.32 0.11

Blue grama 7.48 3.30 1.92 0.66

Hairy grama 0 0

Tobosa 0.51 0.79 0.03 0.19

Ringgrass 1.15 0.02 0.04 0.01

Other grasses 0.34 0 0.03 0.05

Total grasses 9.64 4.58 2.34 1.02

Half-shrubs 0 0.1 0.1 0.3

Shrubs 2.9 0 0.3 0.4

Forbs 0 0 0.33 0.17
USDA Forest Service RMRS-GTR-169. 2006. 59

(see below for productivity and soils). Sideoats grama, 

hairy grama, and plains lovegrass comprised 70 percent 

of the grass canopy on the butte (Table 12). The heavily grazed ranges were dominated (90 percent) by curly 

mesquite, which was only 5 percent of the grass canopy 

on the butte. This represents a complete change in dominant species on the butte compared with thousands of 

acres of adjacent rangeland. Sideoats grama (0.1 percent 

cover vs. 13.9 percent on butte) was the only other grass 

species to occur on the grazed area in more than trace 

amounts. In contrast, 12 species occurred on the butte 

in an abundance greater than 0.1 percent canopy cover. 

Canopy cover of palatable forage species was 55 percent 

on the butte compared to 23 percent on the grazing allotment. Abundant in the grazed areas, snakeweed was 

observed on the butte but not recorded in the sampled 

Table 12—Canopy cover (percent) of plant species and basal cover of ground components of a cental Arizona grassland never 

grazed by livestock (geologic refuge, mesa), protected from grazing for ~50 years, and a grazed allotment in Tonto National 

Forest. Adapted from Ambos and others (2000) and Ambos (unpublished data, personal communication).

Canopy cover (percent)

Never grazed Exclosure built

 Trees/shrubs/half-shrubs relict 1934 Grazed

White-ball Acacia Acacia angustissima 0.1

Catclaw Acacia Acacia gregii 1.0

Agave Agave parryi 1.2

False Mesquite Calliandra eriophylla 10.8

Hedgehog Cactus Echinocereus spp. 0.2

Wright Buckwheat Erigonum wrightii 4.7 14.1 7.3

Redberry Juniper Juniperus erythrocarpa 2.0 T T

Englemann Pricklypear Opuntia phaecantha 1.7 3.8 3.0

Banana Yucca Yucca baccata 3.1

Broom Snakeweed Gutierrezia sarothrae T 6.1

Littleleaf Ratany Krameria parvifolia T 0.4

Mimosa Mimosa biuncifera T 2.1

Velvet Mesquite Prosopis velutina T

Scrub Oak Quercus turbinella T T

Total Trees/Shrubs/Half Shrubs 24.8 17.9 18.9

 Forbs

Aster Aster sp. T

Sunflower Helianthus sp. T

Deer-vetch Lotus rigidus

Wright’s Thimblehead Hymenothrix wrightii 3.0 0.4

Plantain Plantago sp. 0.3 0.6

Total Frobs T 3.3 1.0

 Grasses

Threeawn Aristida sp. 1.2

Cane Beardgrass Bothriochloa barbinodis 0.3

Sideoats Grama Bouteloua curtipendula 13.9 0.4 0.1

Hairy Grama Bouteloua hirsuta 8.6

Plains Lovegrass Eragrostis intermedia 5.9

Tanglehead Heteropogon contortus

Curly Mesquite Hilaria belangeri 1.8 T 15.8

Prairie Junegrass Koeleria pyramidata 3.9

Green Sprangletop Leptochloa dubia 0.8

Bull Muhly Muhlenbergia emersleyi

Bottlebrush Squirreltail Sitanion hystrix 2.7 17.7 T

Sand Dropseed Sporobolus cryptandrus 0.2

Total Grasses 39.3 18.1 15.9

 Ground cover components Basal cover (percent)

Grass Basal Area 7.9 3.3 9.9

Tree/Shrub/Half-Shrub Basal Area 1.8 1.4 1.7

Litter > 1.2cm 20.7 20.9 5.8

Rock Fragments > 0.2 cm 32.8 14.0 10.0

Bare Soil 36.1 63.8 71.5
60 USDA Forest Service RMRS-GTR-169. 2006.

transects. There were 21 species encountered in sampling the relict compared to 12 in the exclosure and 13 

on the grazed area. Species sampled in the exclosure 

were generally the same species sampled in the grazed 

area, but a very different set of species was sampled 

on the butte. Basal cover of grasses was greatest on the 

grazed area due the sod-like structure on grazed areas 

compared with bunchgrass structure on the butte. Litter 

cover was similar in the butte and exclosure communities, which were three times greater than in the grazed 

area.

Primary Production

Primary production in mesquite savanna is highly 

variable from year to year, depending on precipitation. 

Paulsen and Ares (1962) estimated a 15 year average of 

perennial grass production of about 50 g/m2/yr in black 

grama grassland at the Jornada Experimental Range, 

with a minimum of 15 and a maximum of 80 g/m2/

yr. Tobosa type range produced 112 and 168 g/m2/yr 

for two years of measurement. At the nearby College 

Ranch, Holechek and others (1999) reports a 30-year 

average of 22 g/m2/yr forage production, with a range 

from near zero to over 80 g/m2/yr. Perennial plus annual 

grass production at the Santa Rita Experimental Range 

over a 10 year period averaged 22 g/m2/yr, with a minimum of 4 and a maximum of 33 g/m2

/yr (Martin and 

Cable 1974). Grass production was greater on coarse 

than on fine soils, although black and Rothrock gramas 

produced more on fine soils. Mesquite-free plots produced an average of 26 g/m2

/yr compared to 18 g/m2/yr 

for mesquite-infested plots. Annual grasses comprised 

71 percent of the total grass production, and this did not 

vary with mesquite cover. The annual grass component 

varied greatly among years compared to the relatively 

more stable perennial grass component. Production on 

low elevation, low rainfall sites heavily dominated by 

annuals can vary from 0.4 to 107 g/m2/yr in consecutive years, while values for productive sites with a good 

perennial grass component for the same two years were 

50 and 95 g/m2/yr (Martin 1975). While annual fluctuations are evident, previous as well as current-season 

precipitation was important in controlling perennial 

grass production.

Martin and Cable (1974) also estimated grass production over the 10 year-period for pastures grazed 

November to April, May to October, and year-long. 

Percentage differences for the last three years of study 

compared to the first three years indicated significant 

increases in perennial grass production for the yearlong grazing treatment with little change in the two 

seasonally grazed pastures. Concomitantly, annual 

grasses tended to decrease in pastures grazed year-long 

to a greater extent than those grazed in November to 

April or May to October. Fifteen different grazing 

systems involving different periods of seasonal or yearlong rest of 1 year rest over 3 years, 2 years rest over 3 

years, 3 years rest over 4 years, rested every year during 

winter, or never rested were tested over eight years at 

the Santa Rita (Martin 1973). Of the 15 treatments, only 

spring-summer rest 2 out of 3 years resulted in significantly greater production of perennial grasses compared 

to all other treatments. Weather greatly overrode any 

grazing management system. Rest schedules of the various treatments gave deferment percentages (time not 

grazed) of from zero to 67 percent. Density of perennial 

grasses other than Rothrock grama (a short-lived perennial) did, however, increase with the 44 and 67 percent 

rest schedules. None of the treatments involving periods 

of rest during the year improved production of annuals 

or Rothrock grama.

Valentine (1970) assessed different grazing intensity 

effects on yields of black grama at the College Ranch 

adjacent to the Jornada Experimental Range. Percent 

utilization of black grama during the dormant season 

averaged 26, 35, 49, and 60 percent over the 10 year 

study for what was considered light, moderate, proper, 

and heavy grazing. The sites used for the study were 

previously deteriorated due to proximity to water, and 

the objective of the experiment was to assess recovery 

of black grama. Average production of black grama was 

22, 21, 13, and 9 g/m2/yr, with plant production under 

light - moderate grazing being significantly greater than 

that under proper - heavy and not different within either 

of the two. Yield as an average of that for good-condition black grama range was 46 percent for the lighter 

grazed treatments. Light grazing compared to proper or 

heavy grazing resulted in greater seed-stalk height, stolons per plant, length of stolons, and rooted stolon buds 

per plant, while moderate grazing was greater than the 

heavier treatments only in length of stolons.

In a 10 to 11 year field clipping study, Canfield (1939) 

estimated production in both black grama and tobosa 

types. Clipping black grama at a height of one inch drastically lowered production to 9 to 11 g/m2

/yr regardless 

of clipping intervals of 2, 4, 6 weeks, or once at end-ofseason. Clipping at a two-inch height resulted in average 

11 year mean production of 20, 21, 32, and 34 g/m2/yr 

black grama production for clipping intervals of 2, 4, 6 

weeks, or once at end-of-season, respectively. Similar 

results were not obtained for the tobosa type until the 

last three years of treatment, at which time tobosa began to decline dramatically under severe clipping. Plots 
USDA Forest Service RMRS-GTR-169. 2006. 61

clipped weekly at four inches height produced 11 times 

as much biomass as plots clipped at two inches. A four 

week clipping interval represents very heavy defoliation, and production was 87 and 127 g/m2

/yr for the two 

inch and four inch height clipped plants, respectively.

Total primary production was estimated in grazed 

and ungrazed grassland for three years at the Jornada 

during the International Biological Program (Sims and 

others 1978; Pieper and Herbel 1982; Pieper and others 1983). Total primary production over the three years 

averaged 74 percent greater on ungrazed compared to 

grazed treatment (about 40 percent utilization), with 

significant differences in two out of three years (Table 

13). Warm-season grasses were always a greater percentage of the production of ungrazed treatment, and 

warm-season shrubs were always a greater proportion 

of the grazed. Holechek and others (2000) report 16 

and 13 g/m2/yr production for conservatively compared 

to moderately grazed treatments at the College Ranch, 

but there were experimental design layout problems 

for making this comparison (see Species Composition 

section).

At the relict, never grazed butte examined by Ambos 

and others (2000) (see Species Composition section 

above for description), herbage production was 127 g/

m2/yr in a good year and 53 g/m2/yr in a drought year. 

These authors indicate primary production on the butte 

was about four times that as similar grazed areas.

Root and Soil Responses

Although there is a considerable amount of research 

concerning aboveground plant responses to grazing 

on mesquite savannas, surprisingly very little research 

has been done belowground. Sims and others (1978), 

Pieper and Herbel (1982), and Pieper and others (1983) 

observed greater root biomass in ungrazed compared to 

grazed treatments at the Jornada in the 0 to 10 cm depth 

increment, and no effects of grazing at deeper depths 

to 30 cm. Average root biomass over the three years of 

study and all depths were 168 and 143 g/m2 for the ungrazed and grazed treatment, respectively (old estimates 

of root production based on various means of summation of peaks produce very biased values [Milchunas 

and Lauenroth 1992] so are not reviewed here). Other 

comparisons between grazed and ungrazed conditions 

involve extreme examples. Gill (1975) estimated 12 

and 6 g/m2 root biomass for ungrazed and grazed sites. 

However, the values are extremely low, possibly because the sites were established in “sacrifice” areas very 

close to water. The only conclusion from these data may 

be that relative root biomass does not recover from denuded areas after 12 years of protection from grazing. 

Ambos and others (2000) reported six-times the root 

biomass in surface horizons of the relict, never-grazed 

butte compared to similar nearby grazed allotments.

Soil compaction was compared by Ambos and others (2000) at the relict butte site, a long-term exclosure 

since 1934, and a grazed site, all of which were heavyloams. Soil bulk densities were 0.93, 0.98, and 1.33 for 

the butte, long-term exclosure, and grazed site respectively. Organic carbon contents of the butte site were 

very high relative to Arizona soils in general. Gil (1975) 

found no significant difference in bulk densities between 

ungrazed and grazed sites, but some improvement in 

soil organic matter after 12 years of recovery. Gardner 

(1950) found “no critical differences” in soil nutrients or 

organic matter between areas grazed or protected from 

grazing for 30 years.

Grazing Management

The free-range policy in the late 1800s to early 1900s 

represented a period of extremely heavy grazing. Early 

researchers thought that grazing systems with periods 

of rest would lead to improved range condition. Griffiths 

(1901, 1904, cited in Martin 1975) suggested that three 

year periods would be sufficient for improving condition to acceptable levels. This was based on observations 

that protected range recovered more rapidly than range 

where stocking rates were reduced (Wooton 1916, cited 

in Canfield 1948; Canfield 1948). Humphrey (1960) 

recommended rest during every July to August growing 

season for southern Arizona range in poor condition to 

allow palatable perennial grasses to grow and set seed. 

Rest every third year was thought to be adequate once 

the range reached good condition. Rest and particular 

periods of rest were thought to be important means by 

Table 13—Grass, forb, and shrub composition (percent) and 

aboveground primary production (g/m2/yr) of a New Mexico 

(Jornada) mesquite savanna ungrazed or moderately 

grazed by livestock. Adapted from Pieper and Herbel 

(1982).

Composition (percent)

Ungrazed Grazed

Warm season grasses 43.9 18.1

Warm season forbs 23.7 30.5

Cool season forbs 3.7 5.1

Warm season shrubs 22.2 40.3

Production (g/m2

/yr)

Total aboveground 148 109
62 USDA Forest Service RMRS-GTR-169. 2006.

which range managers could devise grazing systems 

that would do more towards improving range condition than just regulating year-long grazing intensity. A 

number of systems were devised that involved various, 

and sometimes complicated, combinations of deferment 

or rest and rotations among various numbers of pasture 

units.

Martin (1973) tested 15 different season/frequency 

of rest combinations on replicated but small plots at the 

Santa Rita Experimental Range for an eight year period. Seasonal or year-long rest for one year in three, two 

years in three, three years in four, and every year were 

compared to year-long grazing. Increases in the densities of long-lived perennial grasses occurred only when 

rest periods were 44 or 67 percent of the time (Fig. 6). 

Rest period totaling 33, 25, 22, or 11 percent of the time 

did not provide a significant increase over season-long 

grazing with no rest. A system with spring-summer rest 

in two years out of three showed the greatest increases 

in perennial grasses. Increases in perennial grasses other than Rothrock grama occurred after four years under 

this system, but this did not continue after additional 

years. None of the systems increased production of annual grasses or Rothrock grama. Martin concluded that 

rest for only a part of the year provided little increase 

in production over year-long grazing. In other studies at 

the Santa Rita, rest in some cases was shown to decrease 

perennial grasses and increase burroweed (Martin 

and Cable 1974). This occurred for annual rests from 

November through April and particularly May through 

October compared to year-long grazing. Martin and 

Ward (1976) further tested some of the promising treatments found for previously very heavily grazed sites 

(Martin 1973) on an additional three sites less intensely 

grazed. Four different alternate-year rest schedules, 

one seasonal every-year rest treatment, and continuous 

year-long grazing was compared for seven years. Yearto-year variability in perennial grass production was so 

great as to mask any treatment effects. The authors concluded that perhaps alternate-year rest did not provide 

sufficient periods of growing-season recovery. Based 

on the studies, however, Martin proposed a three-year, 

three-pasture system of spring-summer and winter rest 

two out of three years that would become known as the 

Santa Rita grazing system.

Grazing under the Santa Rita system would average 

the same 40 percent level over time as that for year-long 

grazing, but would be heavier in a pasture being grazed 

that year to make up for the amount not grazed when the 

pasture was rested. The heavy grazing followed by rest 

is the basis for a number of grazing systems that attempt 

to obtain more uniform utilization across species of different palatability and across the landscape, thereby 

more uniformly spreading the burden of defoliation. A 

13 year test of the Santa Rita system alongside year-long 

grazing at three sites representing low-, mid-, and highelevations resulted in no significant differences in grass, 

forb, or shrub densities or in shrub cover (Martin and 

Severson 1988). The lack of response to the rest-rotation grazing system was attributed to the good initial 

conditions of the study sites, and the moderate grazing 

intensities. The authors speculated that different results 

may have been found if the sites had been in a poor condition initially. In contrast, distance to water, in other 

words grazing intensity, resulted in significant differences in perennial grass production, particularly black 

grama.

Another long-term, 28-year study at the Santa Rita 

Experimental Range assessed some of the same pastures used previously by Martin and Severson (1988) 

that continued in the Santa Rita system compared to 

some of the previous pastures that had been switched 

to year-long heavier grazing than that used in the Santa 

Rita system pastures (Angell and McClaran 2001). 

The switch in grazing practices was for 15 years, but 

there was only one replicate per treatment. Results did, 

Figure 6—Density of perennial grasses other than Rothrock 

grama in relation to the percent of time grazing had been 

deferred over a seven year period in a mesquite savanna community. A total of 15 different season/frequency of 

rest combinations were examined on replicated but small 

plots at the Santa Rita Experimental Range. Adapted from 

Martin (1973).
USDA Forest Service RMRS-GTR-169. 2006. 63

however, confirm those reported originally by Martin and 

Severson. Grazing every summer compared to frequent 

rest during the growing season resulted in a decreased 

density of only one perennial grass (bush muhly). The 

lack of difference between grazing intensities/rest 

- no-rest treatments in density of the exotic Lehmann 

lovegrass was consistent with a previous study showing 

no effects of complete protection on the lovegrass.

On the Jornada, Paulsen and Ares (1962) recommended deferment of grazing on ranges with a high 

percentage of black grama. This was suggested as a 

means to allow black grama stolons to establish new 

plants. Grazing could be shifted to lowland tobosa 

communities, since tobosa could withstand moderate 

grazing, and was palatable during the growing season 

while black grama was palatable at any time of the year 

and more so during the dormant season. Based on three 

pastures with different percentages of grazing during 

winter-spring, land-area requirements decreased with 

increasing percentage of use during the non-growing 

season. Black grama increased with summer deferment in only one of three pasture replicates at the Santa 

Rita (Martin and Severson 1988), but the sites had only 

minor amounts of black grama. Paulsen and Ares also 

recommended placing salt one to four miles from water to obtain a more uniform distribution of pastures. 

Martin and Ward (1970) found lighter use near water 

if access to water was rotated in large pastures lightly 

or moderately stocked. While this may increase potential stocking rates, the heterogeneity of grazed to less 

grazed sites may be important to wildlife diversity.

Martin (1975) suggested there may be some advantage to grazing sheep and/or goats with cattle to balance 

pressure on forbs, shrubs, and grasses, based on some 

research in Texas. However, browsing by sheep or goats 

was not considered a likely means to reduce growth and 

spread of shrubs.

In general, the current view of range managers is that 

“various rotation grazing systems cannot overcome the 

rangeland degradation associated with chronic overstocking.” Grazing intensity is the overriding factor in 

range condition, although periods of rest longer than 

seasonal can be beneficial. Hughes (1980) suggested 

that low to moderate utilization rates were necessary 

for grazing systems to work in northwestern Arizona, 

but stocking rates were the only factor varied at the 

sites that were assessed. Martin (1975) suggested that, 

theoretically, there should be some advantage to rest 

in bunchgrass grasslands compared to sodgrass grasslands, simply because the former more often reproduce 

by seed and the latter by vegetative spread. Seed 

production is generally, but not always, reduced by 

current-year grazing (Obeso 1993). Longer-term restrotation systems may also better provide a diversity of 

community physiognomies across the landscape, which 

may be important for wildlife populations. An important 

aspect of management is the ability to reduce herd size 

during drought without seriously affecting the breeding herd, thereby protecting from overutilization when 

production is low (Paulsen and Ares 1962). Additional 

grazing systems studies in mesquite dominated systems 

are examined in the Chihuahuan Desert section below.

Mesquite Savanna Conversion 

to Chihuahuan Desert: Shrub 

Invasion, Livestock Grazing, 

Fire, Small Mammals, and 

Climate____________________

The invasion of grassland by shrubs has been extensive 

in many areas, and not extensive in other areas, although 

encroachment continues. For example, Mehrhoff (1955) 

and Humphrey and Mehrhoff (1958) indicate that threequarters of the Santa Rita Experimental Range was at one 

time grassland, but that at the time of writing less than 

one-fifth of the site remained in grassland. McClaran 

(2003) reviewed early studies at the Santa Rita and 

showed recent data that indicates cover, but not density, 

continues to increase through 2000. At the Jornada, a 

107 percent increase was observed in mesquite-dominated areas between 1915 and 1946, while grasslands 

declined 28 percent (Branscomb 1956, 1958; Buffington 

and Herbel 1965). In contrast, mesquite increased 

dramatically between 1952 and 1999 on only one of the 

41 sites, and creosote-bush on a single site, of the 41 sites 

assessed by Navarro and others (2002). Bahre (1991) 

suggested that the changes may not have been as extensive as some researchers have implied. However, the 

review above clearly indicates that 1) where shrubs have 

invaded, there have been changes in understory species 

composition, 2) the shrub invasions occur regardless of 

whether sites are currently grazed or not, 3) however, 

early studies and observations suggest that grazing can 

be a factor under conditions of very heavy grazing intensities now considered abusive, and 4) in some cases, 

light grazing may dampen post-drought effects on shrub 

increases. We here focus on mesquite/creosote when 

speaking of shrub invasion. Shrub invasion regardless 

of grazing management does not mean grazing is not a 

factor, since interactions such as grazing effects on fire 
64 USDA Forest Service RMRS-GTR-169. 2006.

and/or previous grazing may still be an operating explanation. A patchy spatial distribution of the phenomenon 

would perhaps suggest that no one single factor may be 

responsible, and/or there may be thresholds associated 

with particular or interacting factors, or there is a seed 

limitation or some other factor associated with a moving front of shrubs. Factors other than current direct 

effects of grazing that have been suggested to potentially be involved include: 1) previous very heavy grazing 

during the late 1800s to early 1900s, 2) current grazing 

effects on reducing fire fuel-loads and fire suppression 

practices, 3) grazing by small mammals, and 4) climate 

change or climate fluctuations and/or a reoccupation of 

sites that were previously a shrubland community. Soil 

types can provide the patchy framework for interactions 

with fire and climate, through productivity/fuel load 

relationships or precipitation use efficiency and infiltration depth dynamics. Before exploring how grazing may 

interact with the above factors, we first return briefly to 

the primary possibility that current grazing directly influences shrub invasion. We do this because there are 

recent authors that imply or suggest current grazing is 

still a factor in shrub invasions (Schlesinger and others 

1990; Bahre 1991; Bahre and Hutchinson 2001; Kerley 

and Whitford 2000). From the standpoint of “what can 

be done now?” and “what are the current management 

implications?”, it is important to differentiate between 

past effects of grazing, current direct effects of grazing, 

and current effects of grazing as an interaction factor. 

Residual effects of past grazing events without current 

effects of grazing become academic and do not have a 

bearing on the two questions.

The Schlesinger and others (1990) model of 

desertification was a high-impact publication in the 

journal Science and was further elaborated in Whitford 

(2002). It is an eloquent and logical formulation of a 

hypothesis for the formation and maintenance of the 

alternate-stable-state shrub condition. The model is 

based on local landscape distribution of soil nutrients 

and water. The homogeneous distribution of soil 

resources under grassland are disrupted when livestock 

grazing lowers black grama cover, resulting in erosional 

redistribution of soil and runoff redistribution of water. 

Some local areas receive increased water that percolates 

deeper and favors shrubs over shallow rooted grasses, 

while other areas receive less water and loose soil. 

Once established in microsites of increased soil water 

depth percolation, shrubs reinforce the heterogeneous 

pattern of soil nutrients and water because infiltration is 

greater under the litter of shrubs, and wind- and watermoved soil further accumulates under the shrubs over 

time creating “islands of fertility” that are commonly 

observed in shrublands throughout the world. The 

model is conceptually sound, and probably reflects the 

real world in many instances. However, 1) the model 

was presented as a hypothesis for the renewal proposal 

for the Long-Term Ecological Research program at the 

Jornada site, not as a statement of fact for the site in 

general. 2) past or current very heavy grazing could very 

well fit conditions necessary for the model to operate. 

However, some of the strongest data indicating that 

shrub invasion into grassland exclosures is comparable 

to grazed grassland comes from the same site. That 

does not detract from proposing to study processes that 

maintain the shrub stable state, or the conditions under 

which grazing could initiate such dynamics. Once 

established due to whatever cause, the Schlesinger 

and others model provides a good description of the 

dynamics helping to maintain the shrub dominated 

system (Martin and Cable 1974; Gibbens and others 

1993). Very heavy grazing could very possibly be a 

cause, but very heavy grazing is not the current situation 

at the Jornada or other research sites and is obviously 

not the current situation in exclosures. Based on the 

data reviewed above, grazing as practiced at these sites 

or protection from grazing does not differentially affect 

invasion outcomes. Similarly, the Kerley and Whitford 

(2000) data show only a slight directional shift in 

community trajectory towards an ordinated shrub space 

that requires a large extrapolation across Euclidian 

space to reach the shrub domain, while their own shrub 

data and the Gibbens and Beck (1988) and Gibbens and 

others (1993) long-term data do not support such an 

extrapolation. In a review of shrub encroachment in the 

southwest, Allred (1996) also questioned the emphasis 

the model placed on grazing, given the data from that 

site. Bahre (1991) and Bahre and Hutchinson (2001) 

do not clearly differentiate among historic past very 

heavy grazing, current grazing practices, and grazing 

interactions with other factors. There is no doubt that 

current heavy grazing and past very heavy grazing can 

and has had large effects on southwestern grasslands, 

but the evidence reviewed above indicates that current 

grazing intensities, or even protection from grazing, does 

not influence shrub invasion. This does not rule out other 

potential effects of various grazing intensities versus no 

grazing on understory plant species composition and 

structure. Because previous grazing intensities could 

very possibly lead to conditions described in Martin 

and Cable (1974) and the Schlesinger and others (1990) 

model, there is the potential for previous or current 

heavy grazing effects on soils to manifest with long 

time lags (Grover and Musick 1990). Soil formation 
USDA Forest Service RMRS-GTR-169. 2006. 65

is a long-term process, whereby loss or redistribution 

of soil can have very long-term consequences, all of 

which may not occur immediately. This topic will be 

discussed further, after reviewing other potential factors 

and interacting factors in shrub invasions.

A grazing effect on fuel loads resulting in reduced fire 

frequency or intensity is always a potential factor that 

may alter plant communities. While an extensive review 

of fire effects will not be attempted here, a few contrasting points of view should be mentioned. Some authors 

were of the view that productivity in these, or some of 

these communities, was low enough to reduce the importance of fire even under pristine conditions, and the 

potential for fire to reduce abundance of black grama 

(Reynolds and Bohning 1956 - “seriously damaged” and 

did not recover after 4 years; Cable 1965 - 90 percent 

mortality) is another factor arguing for a relatively lower 

importance of fire in some of these communities. Gosz 

and Gosz (1996) suggest that “the increased dominance 

of black grama reflects the absence of burning and grazing” on the Sevilleta National Wildlife Refuge. On the 

other hand, Kaib and others (1996) and Kaib and others (1999) indicate a fire frequency of 4 to 8 years that 

is more frequent than previously thought. Bock and 

Bock (1996) report that fires on the ungrazed AppletonWhittell Research Ranch have been somewhat regular. 

Black grama is not a component of that grassland, where 

fire was not a factor in grass mortality. For two or three 

growing seasons, fire reduced grass canopy cover, resulting in increased forb abundances for the same 

period, but had no negative long-term effect on shrub 

abundance except for burroweed. Mesquite resprouted 

from their bases. Importantly, changes in community 

physiognomy and seed production for the 2 to 3 years 

post-fire had some large effects on bird, rodent, and insect communities. Unfortunately, there are few large 

sanctuaries such as the Appleton-Whittell where fire 

frequencies may more closely resemble past conditions. 

The role of Indians in setting fires and affecting fire 

frequencies would be another aspect of debate. Would 

these be considered natural fires, and did these change 

through ecological time or did populations and social 

conditions remain constant enough and long enough to 

be considered an endogenous disturbance? While the 

Bocks found few long-term effects of fire, the observed 

short-term effects could historically be significant if fire 

return frequencies were short, whereby a patchy distribution of unburned and recently burned habitats was 

created across the landscape. Fire does have significant 

effects on grass and forb compositions for three to four 

years after the burn (Reynolds and Bohning 1956; Bock 

and Bock 1996).

The difficulty of killing mesquite with fire has led 

some researchers to question the role of fire in its distribution (Wright and others 1976). Some of the densest 

honey mesquite areas in Texas are in areas most subject 

to recurring fires (Allred 1949). Young velvet mesquite less than 1.3 cm (half inch) are more susceptible 

to mortality by fire in Arizona, but fire may kill only 

10, 30, to 60 percent of even the young (Glendening 

and Paulsen 1955; Reynolds and Bohning 1956; Cable 

1965). Even seedlings survive fire, although only onethird of one-year-olds survive (Cable 1961). Greater 

intensity fires kill more individuals, but even with over 

500 g/m2 grass fuel the mortality of young shrubs was 

low and black grama mortality was high under even less 

intense fire (Cable 1965). Fire on the Santa Rita significantly reduced burroweed, cholla, and prickly pear, but 

only reduced mesquite 9 percent (40 percent of young 

shrubs) and did serious damage to black grama (only 

37 percent recovery four years post-fire). Densities of 

mesquite on burned and unburned treatments do not 

differ 3 years (Martin 1983) or 13 years (Cable 1967) 

after prescribed fire or 8 years after wildfire (Gottfried 

and others 2003). Ffolliott and others (2003) reviewed 

evidence from the Santa Rita and concluded that the 

effects of fire on grasses lasts only a couple of years 

and shrubs and cacti come back quickly unless other 

control measures are also implemented, but the authors 

also concluded that the effects of fire are season and 

site specific. Fires that were not very hot and extensive 

have not been successful in controlling mesquite expansion, but fires hot enough to kill mesquite would also 

kill grass and forage species. On upland sites in Texas, 

seedlings of honey mesquite up to 1.5 years old were 

easily killed with moderate fires, severely damaged 

at 2.5 years of age, and very tolerant of fire after 3.5 

years of age. Short fire return intervals of 4 to 8 years 

(Kaib and others 1996; Kaib and others 1999), therefore, would have an effect on mesquite canopy cover, 

but would not be frequent enough to prevent invasion. 

This is counter to early perceptions of the role of fire 

and grazing-fire interactions (Humphrey and Mehrhoff 

1958; Branscomb 1958).

Rodents have often been cited as a potential factor 

in shrub invasions. Their very large effect on plant 

communities in the southwest has been well documented 

(Norris 1950; Brown and Heske 1990; Heske and others 

1993; Curtin and others 1999; review in Fagerstone and 

Ramey 1996). Where grassland and shrubland sites 

have been compared, smaller effects of small mammals 
66 USDA Forest Service RMRS-GTR-169. 2006.

have been observed in grassland compared to shrubland 

(Norris 1950). This is probably because most rodents 

prefer the shrublands, where they are 3 1⁄2 times more 

abundant. In the shrublands, small mammals can 

have large effects on grass abundance. Reynolds and 

Glendening (1949) and Wood (1969) showed that 

Merriam kangaroo rats prefer the shrub habitat and 

cache seeds that would later sprout. Wood found an 

average of nearly one kilogram of food stored in wood 

rat dens, much of which was mesquite beans. Further, 

Paulsen and Ares (1962) state that mesquite invasion 

occurs most rapidly in the transition zone between 

the shrubland and grassland. Therefore, a reasonable 

hypothesis can be generated whereby a front of shrub 

invasion is generated by rodents foraging out from the 

shrubland into the grassland border, consuming grass 

that reduces competition, and planting seed at the same 

time. Curtin and others (1999) and Curtin and Brown 

(2001) found that shrub cover increased from 1979 to 

1995 at an Arizona site, but the increase did not differ 

among kangaroo rat exclosures, all rodent exclosures, 

and control plots open to native mammals, even 

though rodents altered other components of vegetation 

composition. There are insufficient data from small 

mammal exclosures to reject this hypothesis as at least 

a potential contributing factor in shrub encroachment. 

However, Curtin and others (1999) and Curtin and 

Brown (2001) demonstrated a role of small mammals 

in vegetation patch-size dynamics, but the rodents 

countered the temporal increase in patch size associated 

with the temporal increase in shrubs. Gibbens and 

others (1992) observed a 79 percent mortality of marked 

mesquite seedlings due to lagomorphs or other small 

mammals, which suggested a role of these herbivores 

in limiting establishment. Based on kangaroo rat 

exclosures, Valone and Thornhill (2001) also observed 

a greater negative effect of herbivory on mesquite 

seedling establishment than the potential positive effects 

of seed-caching. The alternate hypothesis that small 

mammals inhibit spread of shrubs may also be posed. 

There are other shrubs and tree encroachment situations 

in many other southwestern communities in addition 

to the mesquite example. Rodents would need to be 

implicated in these as well for this to be a primary rather 

than a potential contributing factor. Although rodents 

other than kangaroo rats may also influence shrubland 

and grassland dynamics (Curtin and others 1999), it is 

interesting to note that Merriam kangaroo rats increase 

with grazing and fire (Bock and Bock 1996). Livestock 

can also consume and scatter mesquite seed. Hansen 

(1976) suggested that feral horses may also spread 

mesquite seed in their feces, because diets in southern 

New Mexico were 53 percent seed-pods in September. 

If there is a positive effect of livestock on seed planting 

rodent numbers and if both rodents and livestock are 

dispersal agents, both of these particular effects do not 

appear large enough to show significant differences in 

shrub encroachment between livestock exclosure sites 

and adjacent grazed sites. Reynolds (1954) observed 

that kangaroo rats consumed 98 percent of the seed 

they buried and transported seed a maximum of 32 m, 

suggesting that this would result in a dispersal rate of 

only 1.6 km in 500 years.

Climate change has many facets, some of which can 

have potentially important bearing on shrub encroachment in the southwest. Some aspects of climate change 

act at local and regional scales, and others at global 

scales. Changes include elevated concentrations of atmospheric CO2

, increases in temperature, and alterations in 

precipitation amounts, seasonal patterns, and temporal 

cycles. Because of differences in CO2 saturation points 

of the C3 versus C4 photosynthetic pathways, C3 plants 

can potentially increase growth with CO2 enrichment to 

a greater extent than C4 plants (Wand and others 1999). 

Atmospheric CO2 has been increasing since the industrial revolution (Crowley 2000), which coincides with 

shrub invasions in the southwest. This led to considerations of a possible link between the encroachment of 

the C3 shrubs into the C4 grass dominated communities 

(Polly and others 1996a, b). There is no direct evidence 

to support or reject this hypothesis, but the results of 

experiments in nearby semiarid systems have some implications. Although western wheatgrass (C3

) biomass 

increased more in response to elevated CO2 than blue 

grama (C4) when grown in pots, there was no difference in their response to elevated CO2

 when grown in 

the field at a semiarid shortgrass steppe site treated for 

five years (Morgan and others 2004). Needle and thread 

grass (C3) did, however, increase in biomass and community species composition in the elevated CO2

 field 

experiment, where CO2 treatment also resulted in better 

soil water relationships. How plants respond competitively to the soil water conservation effects of CO2

 in 

semiarid and arid systems can be more important than 

the relative carbon gain effects through photosythetic 

pathways. The results from the shortgrass steppe do, 

however, demonstrate that CO2 can change plant community species composition over short periods of time. 

Gill and others (2002) found that increases of CO2 from 

pre-industrial time levels to current-day levels altered 

plant communities in Texas, indicating that changes in 

plant community species compositions could already 

have occurred in response to CO2. Further, a recent study 
USDA Forest Service RMRS-GTR-169. 2006. 67

by Polley and others (2002) using a two-step correlative 

experimental approach showed that increased soil water 

with CO2 and reductions of grass roots or water addition 

could result in increased survival of mesquite seedlings. 

There are infinite possibilities for plant-herbivore interactions to alter the outcome of CO2

 induced competitive 

relationships, but the evidence showing no differences in 

shrub encroachment with long-term grazing treatments 

and protection from grazing in the southwest suggests 

that interactions with grazing are not important in this 

situation. Similar arguments could be made for increases in temperature if they do occur in the southwest.

Climate changes involving both temperature and 

precipitation have been proposed as causes of the 

community shifts from C4 grasses to C3 shrubs in the 

southwest (Neilson 1986). Although older simple 

assessments of annual precipitation did not show 

relationships with shrub encroachment (Humphrey and 

Mehrhoff 1958), new meteorological techniques and a 

focus on winter versus summer precipitation patterns 

coupled with plant-life form life cycles provides a 

general hypothesis describing shifts from C4 grasses to 

C3 shrubs. Neilson indicates that prior to 1900 there was 

a period characterized by cool weather and relatively 

dry winters and wet summers. The end of the “little 

ice age” about 1900 and a shift in climate to warmer 

conditions with relatively wetter winters coincided with 

shifts from grasses to shrubs. After 1940, years with 

favorable winter precipitation became more common. 

Warm season C4 species would be favored by summer 

precipitation and cool season C3 species by winter 

precipitation. C4 grasses would be photosynthetically 

more efficient in summer when temperatures were 

warm and relatively less efficient during winter and 

spring. C3 shrubs would be photosynthetically active 

in warm winters with favorable precipitation, and 

shut down during dry summers. Very heavy grazing 

in the late 1800s and early 1900s resulted in sufficient 

erosion and inhibited asexual reproduction by perennial 

grasses, without which there may have been sufficient 

“biological inertia” (sensu Harper 1977) to overcome 

the subtle changes in climate. There are some strengths 

and some weaknesses in this argument, but there are 

also some important points not presented by the author 

that support the hypothesized link between the changes 

in climate raised in the analysis and the grass-shrub 

shifts in plant community structure. The shifts in 

temperature are counter to what may be expected to 

favor cool versus warm season species, in other words, 

cooler conditions prior to 1900 would possibly favor the 

cool-season shrubs rather than the warm season grasses. 

However, the overriding factor in Mediterranean 

climates favoring the dominance by shrubs and annuals 

is the timing of precipitation in relationship to the larger 

annual amplitude in winter-summer temperatures. Any 

smaller oscillation in longer-term temperature would 

be minor in comparison to summer-winter oscillation, 

and the seasonal timing of precipitation on soil water 

dynamics would be the overriding factor. Greater 

precipitation falling during cooler seasons would not 

only photosynthetically favor cool-season species 

(including many Eurasian opportunistic species) but 

would also be less subject to rapid evaporative loss, 

percolate deeper into the soil profile, and favor deeper 

rooted shrubs regardless of their photosynthetic 

pathway. Temperature optimums occurring in spring 

would favor the cool season species, often before 

warm season species even become active. Fast growing 

annuals and relatively faster growing perennials and 

cool season shrubs utilize winter-stored water before 

warm season species become fully active. Sandy soils 

would accentuate the greater penetration into the soil 

profile of winter precipitation. It is interesting to note 

that sandy soils are preferentially invaded by shrubs 

(Gibbens and others 1983). Soil texture/rooting depth 

arguments were not invoked by Neilson, but further 

support his hypothesis. Rooting depths also explain 

the increase in shrubs after droughts. Years of drought 

and years of high moisture favor shrubs, as shrubs can 

access deep stored water to maintain longer through a 

drought and can access the deep water that percolate 

deep during years of high rainfall. On the other hand, 

years of moderate or frequent light rains favor grasses. 

Leopold (1951) and Cooke and Reeves (1976) reported 

that an analysis of precipitation from the latter quarter 

of the 1800s showed that large rainfall events were more 

frequent and small events less frequent, even though 

total precipitation was not changing.

Some evidence supporting climate change as an 

important factor in shrub invasion has been provided 

by Brown and others (1997) in a long-term study in 

southeastern Arizona near the transition between desert 

shrubland and arid grassland. A study site was fenced 

from livestock grazing in 1977, and adjacent areas remained grazed. At that time the area was largely an 

open grassland with only widely scattered large shrubs. 

Much higher winter precipitation occurred across the 

area in most years through 1995, which coincided with 

large increases in shrubs. Aerial photos of the ungrazed 

research site and adjacent grazed sites in 1979 and 1995 

indicated a 3-fold increase in shrubs on all sites regardless of whether it was grazed or protected from grazing. 
68 USDA Forest Service RMRS-GTR-169. 2006.

The shrub increase could not be attributed to drought 

as well. The rodent and ant populations that were being monitored at the site showed large effects of the 

shrub invasion. Large increases in shrubs did not occur 

in many other areas, some as close as 20 km from the 

research site, even though the large increases in winter precipitation occurred over a much wider area. The 

authors considered this as supporting Neilson’s (1986) 

hypothesis of shrub increases during periods of winter 

precipitation, but other factors must be operating that 

either produce changes through positive feedback or 

dampen changes through negative feedback. However, 

current grazing was found not to be the direct factor. 

McClaran (2003) provides long-term data on burroweed 

cycles that correlates abundance with wet winters.

While there may be many different factors interacting with short-term climate variability, fire has been 

shown to be one of the interacting factors associated 

with shrub invasion (Swetnam and Betancourt 1998). 

Decadal variability in climate is linked to fire occurrence across the southwest, and is associated with 

amplitude of the Southern Oscillation during the past 

three centuries. Fires are most likely to occur during 

drought, and rapid switches from wet to dry periods further enhances the fire potential. Plant mortality during 

drought was followed by pulses of shrub invasions not 

only in grasslands, but in shrublands and woodlands as 

well. The occurrence of these oscillations is further substantiated by independent tree-ring width, ice core, and 

coral isotope reconstructions.

Interactions between climate change and livestock 

herbivory were investigated by Curtin and Brown 

(2001) in association with the climate and rodent/climate interactions described above. Woody vegetation 

was assessed in the same region where increased winter precipitation was linked to shrub increases at the 

smaller study site, but was done at a larger landscape 

scale in 1) higher elevation oak savanna areas ungrazed 

since 1960 or continuously grazed, and 2) by comparisons across U.S. - Mexican borders where the Mexican 

side shows much more intensive grazing based on aerial 

photos. Aerial images showed no difference in shrub 

cover between grazed - ungrazed sides of the fence 

in 1946, 1957, or 1976, the latter being after 17 years 

of protection from grazing. Aerial photos from 1992 

indicated an increase in shrubs associated with the increased winter precipitation, as was also observed at the 

lower elevation rodent exclosure research site (discussed 

above). However, the increase in shrubs was not the 

same in grazed and ungrazed areas. Shrubs increased 

two-fold in the grazed area compared to six-fold in the 

ungrazed area. The contrast in shrub abundance across 

the international border also showed increases in shrubs 

between the 1979 and 1992 photos, and the increases 

were similar for upland and lowland sites. In this case, 

however, no differences were observed in shrubs between the more heavily grazed Mexican and the lighter 

grazed U.S. sides of the boarder. Curtin and Brown 

concluded that increased bare ground and decreased 

herbaceous understory cover with heavier grazing was 

not associated with increased shrub density.

Chihuahuan Desert__________

Chihuahuan Desert communities can be dominated 

by a variety of shrubs, two of which are mesquite species (see Plant Community Descriptions in Appendix 

1). Some of what may be said about mesquite dominated systems is contained in the Mesquite Savanna 

and the Mesquite Savanna Conversion to Chihuahuan 

Desert sections above. Chihuahuan Desert communities 

dominated by shrubs other than mesquite and long-term 

mesquite dominated systems are examined in this section. These shrub dominated communities produce 

less herbaceous understory vegetation in comparison 

with mesquite savanna, although mesquite savanna areas with low and moderate shrub cover may not show 

such reductions. However, dry grassland that has been 

invaded by shrubs actually store greater amounts of 

carbon in soil under the shrub condition than the grass 

condition, particularly in deeper depths (Jackson and 

others 2002). The shrub condition appears to be a selfperpetuating stable state (Schlesinger and others 1990, 

Whitford 2002). Although the above section focused 

on mesquite invasion into grassland, creosote bush and 

tarbush have also encroached into areas of what was 

previously grassland (Gardner 1951).

Species Composition

Much of the San Simon valley in southern Arizona 

was grassland in the late 1800s but had been heavily 

converted to mesquite-catclaw-greasewood shrub dominated communities by the early 1900s. Precipitation 

averages 350 mm/yr in a bimodal distribution, resulting 

in emergence of a winter annual group and a summer 

annual group of species. In this context, Kelt and Valone 

(1995) assessed the effects of a 16 year old livestock exclosure compared with adjacent grazed sites on annual 

plant communities composition. Total density of winter 

annuals was significantly greater in ungrazed compared 

to grazed treatment (166 versus 112), but only one species was significantly greater in the ungrazed treatment 
USDA Forest Service RMRS-GTR-169. 2006. 69

and there was no treatment effect on plant species diversity. No significant effects of grazing treatment 

were observed for total summer annual plant density or 

diversity. Six species were more abundant in the exclosure, but only one was significantly more abundant with 

grazing. A less conservative statistical approach would 

have resulted in additional significant effects of grazing, 

but the authors conclude that the few effects of protection from grazing suggest a very slow response in this 

community. The authors cite a personal communication 

with Chew as support for this, because he observed that 

it took 25 years of protection from grazing to see significant changes in perennial grass species at a site 10 km 

from this study site. Valone and Kelt (1999) compared 

fire treatments with the grazing treatments. Fire had 

greater, and independent, effects compared to grazing. 

These authors point out that summer annuals responded 

positively to fire and did not respond to grazing, and this 

group has affinities with communities to the east that 

more likely evolved with grazing and fire. In contrast, 

winter annuals displayed mixed and complex responses 

to the grazing and fire treatments, and this group has 

greater affinity with communities to the west in the 

Sonoran and Mojave Deserts.

The Chew (1982) study site mentioned above was 

in a creosote bush and a tarbush community. The exclosure was 170 m from a watering point, so grazing 

was often heavy although stocking rates varied with 

year. Eighteen to 19 years of protection did not alter the directional change in vegetation that started in 

this area in the late 1800s. In the creosote bush community, large increases in canopy cover of perennial 

understory species occurred from 1958 to 1977 in both 

grazed and protected sites. Only fluffgrass differed with 

grazing treatment, and it was greater inside the exclosure. Species diversity as estimated by H’ (an index of 

incorporating both richness and evenness components 

of diversity) was greater outside the exclosure due to the 

greater dominance of fluff grass inside. In the tarbush 

community, large increases in snakeweed occurred from 

1958 to 1977, but this happened both inside and outside 

the exclosure. Some grazing treatment differences were 

observed: fluffgrass increased over time outside but not 

inside the exclosure, and the opposite was seen for “other” perennial species. Black grama had large increases 

through time inside the exclosure but was not found in 

quadrats outside. Overall, changes with protection from 

grazing in these communities were considered very 

minor, even after 18 to 19 years of protection. Some increases in palatable perennials occurred, particularly in 

shallow washes. Based on general observations, Martin 

and Cable (1974) also commented that “mesquite could 

greatly retard or prevent improvement of deteriorated 

perennial grass stands...and mesquite effects might 

override the effects of other variables.”

More recently, however, a resampling of both the Kelt 

and Valone (1995) and Chew (1982) study sites showed 

that recovery in these systems is possible, but only after very long periods of time (Valone and others 2002). 

The site that had been ungrazed for 39 years (Chew’s 

site) now showed significantly greater cover of perennial grasses than the adjacent grazed site, with most 

changes occurring in the latter 20 years. The other site 

(Kelt and Valone’s site), that had at the resampling been 

ungrazed for 20 years, showed no indication of recovery 

compared to the adjacent grazed area. The difference 

in recovery between the 39 year old exclosure and the 

20 year old exclosure was apparently a length-of-protection and not a spatial phenomenon, because the two 

study sites were only 6 km apart. Further, the latter 20 

years did not receive above average summer precipitation, which could have potentially explained the greater 

recovery of perennial grasses. The authors argue that 

these arid systems respond very slowly to changes in 

grazing pressures, because conditions conducive to seed 

production and establishment of particular species are 

rare and episodic.

In an early study of creosote bush communities of 

the Rio Grande Valley, Gardner (1951) also commented 

on the very slow recoveries after protection from grazing, especially in areas where grasses were previously 

nearly eliminated due to historically uncontrolled very 

heavy grazing. A comparison of areas lightly grazed and 

heavily grazed showed very similar species composition. While no grasses were observed in heavily grazed 

areas, small amounts were found in the lightly grazed 

areas. Gardner suggested that recovery after protection 

was somewhat more rapid on north facing slopes and in 

areas where water concentrates. Throughout the valley, 

areas where gullies and arroyos had developed showed 

some increases in grass cover after only 10 years. An 

increase in half-shrubs was also noted as a first sign of 

recovery in areas devoid of grasses.

Beck and Tober (1985) assessed the effects of protection from cattle and rabbits and shrub removal after 

22 years at three creosote bush sites in southern New 

Mexico. Precipitation at the sites varied with averages 

of 207, 234, and 243 mm/yr, but the majority fell during the growing season at all sites. The treatments were 

established just after a drought. Responses to removal 

of cattle, rabbits, or shrubs was variable among sites, 

although there was a general increase for all herbaceous 
70 USDA Forest Service RMRS-GTR-169. 2006.

species combined (Table 14). The increases in cover 

were low for most species after removal of any of the 

three. The authors concluded that “range improvement 

will be a slow process” for most ranges dominated by 

creosote bush. Gibbens and others (1993) arrived at 

similar conclusions based on a 30 year shrub removal 

study, where perennial grasses showed only limited increases. Shrub to grass cover ratios did not change with 

rabbit grazing compared to 50 years of rabbit exclosure, 

but there were some large changes in species composition (see Species Composition of Jornada black grama 

and dropseed communities section above).

An additional three sites were studied by Roundy and 

Jordan (1988) in southeastern Arizona. Precipitation at 

the two Chihuahuan Desert sites averaged 248 and 284 

mm/yr, with most occurring in the summer. Changes in 

species composition or in life-form composition were 

small (Table 15). Native perennial grasses were not significantly different at one site protected from grazing 

for 19 years and were greater in the grazed treatment 

at the other site where the exclosure was 16 years old. 

Total overstory shrub cover did not differ between treatments at either site. These authors concluded “there is 

no evidence that exclusion from grazing alone or even 

in association with rootplowing will encourage reestablishment of native perennial grasses.”

In contrast to studies above indicating little or very 

slow change with protection from grazing, Potter 

and Krenetsky (1967) reported decreases in mesquite 

under protection and increases in grazed areas. Their 

one site in the Chihuahuan Desert type was dominated 

by mesquite and tarbush. The grazed area is, however, 

an example of very heavy grazing, since it is crossed 

with a frequently used cattle trail. After 25 years of 

protection from grazing, new species were found in the 

exclosure, while species were eliminated in the grazed 

area. Only fluffgrass remained in the grazed area, while 

black grama increased in the exclosure.

Minor influences of native mammalian folivores 

(cottontail rabbits, jackrabbits, mule deer, and collared 

peccaries) was observed after 11 years of exclosure in 

a southern Arizona site dominated by acacia, tarbush, 

and joint-fir (Guo and others 1995). In contrast, birds 

had a strong effect on both winter and summer annuals, 

rodents had a relatively weak effect compared to birds, 

and mammals had the least effect. Effects were primarily on large-seeded annuals. Excluding each of the bird 

and rodent groups did not produce the same combination 

of effects produced by excluding both at the same time; 

in other words, effects were not additive. The effects of 

mule deer on plant communities is generally not large in 

semiarid communities of the west (see Ponderosa Pine 

section), although there are situations where impacts 

are great (see Pinon-juniper section, Kaibab Plateau). 

Based on numbers in Wallmo and others (1981) and 

Teer (1996), assuming average deer densities of 4 to 

8 animals/km2 and consumption of 1 kg/day then deer 

would only consume 1.5 to 3.0 g/m2/yr of forage, which 

Table 14—Increase (+), decrease (-), or no change (0) in plant cover over a 21 year period in response to removal of rabbits, 

cattle, or shrubs from a Chihuahuan Desert site in southern New Mexico. Adapted from Beck & Tober (1973).

Site stocking rate

Low 80 ha/animal unit, Low initially, then 20 ha/ Low initially, then

then ungrazed 11 yrs animal unit continuous 26 ha/animal unit

Organism removed

Rabbit Cattle Shrubs Rabbit Cattle Shrubs Rabbit Cattle Shrubs

Grasses Plant species response

Threeawn spp. + – + – – +

Black grama +

Fluffgrass + 0 + +

Tobosa grass – – –

Bush muhly – + – – + – +

Burrograss + + + – – +

Alkali sacaton – + +

Other grasses – + + – – +

All Forbs + + + – – – + + +

All Herbaceous Species + + + + + + + + +

Half shrub

Broom snakeweed – + + – – +
USDA Forest Service RMRS-GTR-169. 2006. 71

Rodents, especially Dipodomys sp., have been 

considered “keystone” species in these systems (Brown 

and Heske 1990 reviewed above; Samson and others 

1992). Seed consumption of large seeded winter annuals 

reduced this group of plants and allowed persistence 

of small seeded annuals, thereby greatly influencing 

species diversity (Samson and others 1992). Ants also 

had impacts on the plant community, consuming smaller 

seeds than rodents. Ants and rodents affected each 

other differently and together produced large effects on 

densities of winter annuals (Davidson and others 1985). 

Reynolds (1958) observed that seed weight and plant 

height influenced preference for seeds by Merriam 

Table 15—Density of plants on Chihuahuan Desert sites in southeastern Arizona ungrazed for 16 or 

19 years or grazed by domestic livestock. The 16-year site was formerly desert grassland. Data 

from Roundy & Jordon (1988).

16-year site (San Simon) 19-year site (Bowie)

Species Ungrazed Grazed Ungrazed Grazed

Native perennial grasses Density (number/60 m2

)

Threeawn spp. 0.1 0 T O

Arizona cottontop T 0 N N

Fluffgrass 8.8 16.0 0 0

Bush muhly 15.8 15.1 0.9 0.2

Total native perennial grasses 24.7 31.1 0.9 0.2

Full-shrubs

Whitethorn acacia 0 0 N N

Fourwing saltbrush 5.8 5.0 N N

Longleaf jointfir 0.1 T N N

American tarwort 6.3 7.9 N N

Creosote-bush 4.3 2.7 28.7 27.8

Desert-thorn spp. 0.4 0.1 T 0

Honey mesquite 11.0 13.3 0.4 0

Total full-shrubs 27.9 29.0 29.1 27.8

Half shrubs

Rayless goldenhead N N N N

Broom snakeweed

Juveniles (<20 cm height) 128.7 159.8 3.4 0.1

Mature N N N N

Burroweed

Juveniles (<20 cm height) 0.1 0.7 T 0

Mature N N N N

Mariola 3.0 1.8 N N

Desert zinnia 11.1 4.9 0.2 0.1

Total half-shrubs 183.0 208.0 3.8 0.2

Total shrubs 210.9 237.0 32.9 28.0

Perennial forbs

Desert globemallow 2.2 2.8 0 0

Succulents

Walkingstick cactus N N N N

Devil cholla N N T 0.2

Purple pricklypear 0.2 T N N

0 = observed in area but not in transects

T = trace amount

N = none even observed in sampling area

is very low given even the low productivities for many 

southwestern plant communities. Water often restricts 

use of the landscape, as deer will generally not forage 

more than 2.4 km from water. In contrast, very high densities of white-tailed deer from 15 to 47 deer/km2

 have 

been reported in Texas (Teer 1996), and the Kaibab herd 

at the eruption peak would have averaged 34 deer/km2

(Rasmussen 1941). While there is generally not a lot 

of overlap in deer and cattle diets, deer are sensitive to 

drought in the Chihuahuan Desert and competition with 

cattle can increase under drought conditions. This then 

would be the period when one or both animal species 

may have a greater impact on the plant communities.
72 USDA Forest Service RMRS-GTR-169. 2006.

kangaroo rats, with short plants and large seeds 

favored. Some exceptions occurred because some small 

seeded species within reach of the rats were utilized 

by harvesting entire seed heads rather than individual 

seeds, and the seeds of some tall shrub species were 

harvested from the ground. Kangaroo rats were favored 

by shrub habitat, avoided habitat with large amounts of 

perennial grasses, and therefore increased with grazing 

by livestock. Reynolds (1950) reported that perennial 

grass, shrub, and cactus seeds were preferred, and 

the greatest impact of the kangaroo rats were found 

on these plants when comparing rodent grazed and 

ungrazed treatments. Small seeded perennial grasses 

were not affected by rodent grazing treatment. Chew 

and Chew (1970) estimated that rodents consumed over 

85 percent of seed production and only 1.1 percent of 

shoot production. Reynolds (1958) notes that re-seeding 

of small areas of rangeland in often unsuccessful due to 

the ability of rodents to easily locate large seeds planted 

up to one inch deep in the soil. In some cases, the seed 

catching activity of heteromyid rodents increases the 

probability of successful germination of seeds, even 

though many others are consumed. McAdoo and others 

(1983) found that germination of about 50 percent of 

seeds of Indian ricegrass in caches were greatly enhanced 

by rodent removal of the indurate lemma, palea, and 

pericarp. This was thought to be the primary means of 

stand renewal, since seeds of this species require some 

extreme form of treatment to break dormancy. The 

caching and shucking of mesquite seed by rodents acts 

as a burying, which improves vigor and survival relative 

to seeds on the surface of the soil (Reynolds 1958; 

Chew and Chew 1970). Browsing by lagomorphs and 

wood rats is another component of herbivory in these 

communities. Chew and Chew (1970) estimated that 

lagomorphs and wood rats kill about 4 percent of the 

shrub food resource, but consume only about 2 percent 

of it. However, this level of pruning and utilization has 

a large impact on some shrubs, reducing volumes of 

particular species up to 56 percent. Additional discussion 

of the role of small mammals in this system is covered 

in the section above on Mesquite Savanna Conversion to 

Chihuahuan Desert: Shrub Invasion, Livestock Grazing, 

Fire, Small Mammals, and Climate.

Although the number is small, the consensus of 

studies in shrub dominated Chihuahuan Desert communities is that recovery after even complete protection 

from mammalian grazing is a slow process, and the 

shrub dominance may place a limit on the extent of recovery. Small mammals, birds, and ants can have large 

influences on the plant communities.

Management

There are large declines in herbaceous production 

after conversion of mesquite savanna to shrub dominated Chihuahuan Desert communities. An extremely 

large body of literature exists on shrub control practices 

but will not be reviewed here. Interactions of grazing 

and mesquite control by other means is assessed below. 

The effects of fire on mesquite is briefly covered in the 

Mesquite Savanna Conversion to Chihuahuan Desert 

section above. However, fuel loads in moderate to dense 

stands of shrubs are often too low to carry a sufficiently 

intense fire for shrub control (Paulsen 1975). Paulsen 

also mentions that many perennial grasses associated 

with mesquite savanna of Chihuahuan Desert communities are sensitive to fire. While acknowledging this 

sensitivity as a consideration, Allen (1996) comments 

that experience in the Malpai area suggests that burning of black grama stolons is not as detrimental in this 

area of higher rainfall as it is in lower, drier desert 

elevations.

A couple of studies have assessed interactions between mesquite control and grazing systems. In Texas, 

Scifres and others (1974) studied herbicide spraying of 

mesquite with continuous, year-long heavy grazing and a 

moderately grazed deferred rotation system. Eight years 

after spraying, reinfestation of mesquite was greater in 

the moderately grazed deferred rotation system than 

with year-long heavy grazing. At the College Ranch adjacent to the Jornada, McNeely (1983) found reduced 

mesquite cover 20 years after herbicide treatment and 

greater herbaceous understory production even though 

basal area did not differ. Year-long grazing had greater 

herbaceous cover and cover of black grama than did 

a seasonal grazing system. While some differences in 

pastures could have confounded results, the author concluded that conservative grazing intensity was the most 

important factor in maintaining and improving range 

condition.

Great Basin Sagebrush 

Shrubsteppe________________

Much of the research on grazing in this community has been done at more northern sites. A couple of 

studies in southern Utah where climate is somewhat 

similar to northern New Mexico and Arizona are 

used in this review. A review of earlier research in the 

northern communities indicated that changes in grazing 

intensities resulted in changes in sagebrush abundance 

as well as herbaceous understory species, which tends 
USDA Forest Service RMRS-GTR-169. 2006. 73

to be a somewhat different situation than for many other 

shrub or woody communities discussed above (Ellison 

1960). Grazing in northern communities commonly 1) 

reduces palatable grasses and increases the sagebrush 

component, particularly when grazing is in the spring, 

2) promotes the spread of sagebrush into other communities, and 3) promotes spread of pinyon and juniper 

into sagebrush communities in locations where they occur in proximity. On the other hand, Ellison indicates 

that grazing in the sagebrush type in the fall reduces 

sagebrush and improves palatable herbaceous species, 

whereby “this is one of the few instances known to the 

reviewer where grazing pressure can be applied under 

range conditions to achieve improvement of the forage.” More recent assessments of grazing management 

in northern sagebrush communities support the capacity of fall grazing by sheep and complete protection to 

reverse the shrub dominance when the site has a good 

grass understory (Laycock 1987). However, sites heavily dominated by sagebrush with very little grass and 

herbaceous cover recover slowly or may not recover at 

all after complete protection for up to 46 years (Sanders 

and Voth 1983). West and others (1984) concluded that 

many sagebrush dominated sites are successionally 

stable. Laycock (1987) concluded that any system or 

season of grazing would not improve highly depleted 

sagebrush range. For ranges in fair or better condition, 

Laycock notes that various authors in different parts of 

the sagebrush range have come to different conclusions 

concerning positive and negative aspects of different 

grazing systems, and that different responses to grazing 

may occur in different sagebrush communities. Below 

are findings for southwestern communities. Grazing effects on a similar community type are presented above 

in the Great Basin Juniper-Sagebrush section.

Species Composition

Upland and lowland big sagebrush communities were 

studied using a 22-year-old exclosure in north central 

New Mexico (Holechek and Stephenson 1983). The site 

received 250 to 350 mm/yr precipitation and was grazed 

by cattle in late winter and spring at a 30 to 50 percent 

utilization level. Big sagebrush cover was greater inside 

the exclosure on the upland site and lower on the lowland site, although initial conditions were not known. 

The only other significant effect of grazing treatment on 

species was an increase in blue grama with grazing on 

both uplands and lowlands, and a decrease in western 

wheatgrass on grazed lowlands. The greater grass cover 

outside the exclosure on both sites was considered a positive effect of grazing at this particular site. The authors 

concluded that dominance of sagebrush and very low 

understory cover after 22 years of protection suggests 

similar slow to nonexistent recovery of depleted ranges 

as reported above for northern Great Basin sites.

Similar conclusions were reached by Hughes (1980) 

studying three 25-year-old exclosures, which included 

railed and seeded areas. The adjacent sites were grazed 

during the June-to-November growing season, which is 

considered to be the most detrimental period for grazing. 

Sagebrush increased over time both inside and outside 

the exclosures. The railed and seeded areas showed the 

largest increase in shrubs, from 10 percent after treatment to 80 percent after 25 years. Grass cover showed 

the opposite temporal trend. Only one of the three exclosures “maintained a better species composition,” but 

this is not described. The authors concluded that “grazing systems (rest-rotation, deferred) in arid sagebrush 

zones by all appearances seem to be a waste of money 

unless land treatment (chaining, burning, and so forth) 

is a recurring event with the grazing system to keep the 

sagebrush canopy very open and patchy.”

Measuring only grass and forb cover, Bonham and 

Trlica (1985) found very little difference in forb cover 

between a 48-year-old ungrazed and an area grazed at 

40 percent utilization in a sagebrush community. The 

site was in the New Mexico Four Corners area receiving 200 mm/yr precipitation. Grass cover of galleta 

and sideoats grama was only slightly greater in the ungrazed compared to the moderately grazed area. Total 

grass cover was 1.4 and 0.9 percent for the ungrazed 

and moderately grazed sagebrush communities, respectively, and total herbaceous cover differed by only 11.3 

compared to 11 percent, respectively. However, heavy 

grazing at 60 percent utilization resulted in greatly reduced grass and forb cover, with total cover only 1.8 

percent compared to the 11.3 percent in the ungrazed 

area. Similar ungrazed, moderately, and heavily grazed 

fourwing saltbush community comparisons showed 

total herbaceous covers of 19.9, 13.6, and 15.4 percent, respectively. This again suggests little effect of 

long-term protection. However, two species showed significantly different covers between long-term ungrazed 

and heavily grazed communities: galleta increased and 

bottlebrush threeawn decreased with grazing. A third 

winterfat community showed little difference in total 

cover between ungrazed and heavily grazed communities but an increase in blue grama with grazing and a 

decrease in sideoats grama and bottlebrush threeawn. 

However, data in this study appeared highly variable.

Short-term ungrazed exclosures (6 years) and longerterm ungrazed (30+ years) Canyonlands National Park 
74 USDA Forest Service RMRS-GTR-169. 2006.

in southeastern Utah were compared to adjacent grazed 

sites under a deferred system in fourwing saltbush 

communities. Canopy cover of perennial grasses was 

significantly greater in both the short- and long-term 

ungrazed compared to grazed communities, and annual 

forbs were greater in the grazed. Shrub density or cover 

did not differ between protected and grazed communities for either short- or long-term ungrazed sites.

A relict area on the north rim of the Grand Canyon 

receiving 305 mm/yr precipitation was studied by 

Schmutz and others (1967). Although isolated, a bridge 

allowed sheep to cross to Boysag Point between 1920 

and 1943 when weather and water availability permitted, with grazing generally occurring in March to April 

or April to May. The relict site was slightly lower in 

elevation than the grazed mainland site, and exposure 

and windswept conditions may have also caused differences in temperature and precipitation, while fire had 

not occurred at either site since the late 1800s. Large 

differences were observed in species richness and composition. A total of 88 species were found on the relict 

compared to 38 on the grazed mainland. The ungrazed 

relict had grass cover of 36 percent and an overstory 

cover of 60 percent, compared to the grazed site’s 6 

and 90 percent, respectively. Total vegetative cover did 

not differ at 23.5 and 21.5 percent for the ungrazed and 

grazed sites. Actual grass cover was 8.6 and 1.3 percent 

for the relict and grazed site, respectively, and shrub 

covers were 14.1 and 19.5 percent. Species increasing with grazing included blue grama and squirreltail, 

and those decreasing included black grama, galleta, 

muttongrass, and needlegrasses. Sagebrush was more 

than twice as abundant on the grazed site. Sagebrush 

in northern Arizona is able to maintain good levels of 

production with up to 80 percent simulated deer browsing of current annual growth (Sheperd 1971, cited in 

Carpenter and Wallmo 1981).

Primary Production

Over the entire range, West (1983a) gives total aboveground standing crop values of from 200 to 

1,200 g/m2 for the Great Basin sagebrush shrubsteppe, 

and a range in aboveground primary production from 

50 to 150 g/m2/yr. Belowground biomass is similar to 

aboveground values, and litter is about half the standing 

aboveground biomass.

Dense stands of sagebrush produce little understory 

forage, so early attempts at improving sagebrush range 

involved brush control efforts (Vale 1974). Grazed 

sagebrush in northwestern New Mexico at the southeastern edge of the community averaged 12 g/m2

/yr of 

usable livestock forage, with a range from 7 to 17 g/

m2/yr (Francis 1990 cited in Vincent 1992). Degraded 

sagebrush/grass range in this area produce only 10 to 15 

percent of the potential forage, but removing livestock 

from range in this condition for 12 years resulted in only 

minor reductions in sagebrush cover compared to areas with continued use by cattle (Dahl and others 1976, 

cited in Vincent 1992). Sagebrush is not considered a 

palatable forage for livestock, but some is utilized. Deer 

will browse up to 50 percent in their diet (volatile oils in 

large amounts can lower rumen microbial activity), and 

a mix of grass, forbs, and shrubs are considered good 

for both deer and pronghorn antelope (McEwen and 

DeWeese 1987). While early management efforts were 

focused on sagebrush eradication, current practices attempt to provide a mix of forage and browse for both 

livestock and wildlife.

Actual production estimates are available only for 

little-grazed relict and heavily grazed mainland sites 

from the Schmutz and others (1967) study. As mentioned above, differences between treatments may be 

due to site differences, and methods of clipping are not 

described but are called “annual production” and not 

standing crop. Values were 46 compared to 32 g/m2/

yr for the ungrazed and grazed communities, respectively. Grass production was 50 and 13 percent of the 

total for the ungrazed and grazed communities, respectively, while shrub production was 45 and 77 percent. 

Snakeweed and sagebrush were the shrubs producing 

more on the grazed site, whereas cacti and more palatable shrubs were more productive on the ungrazed site.

Management

Studies in Idaho, Utah, and Nevada all indicate that 

grazing too heavily or too early in the spring can result 

in loss of herbaceous understory and favor sagebrush 

(Laycock 1987). Vincent (1992) suggests this is true for 

New Mexico sagebrush communities as well, and grazing should occur only in the non-growing season. In 

degraded range in contrast, Hughes (1980) considered 

any grazing system to be ineffective since shrubs increased in abundance even in exclosures. However, one 

of the exclosures did show improvement in abundance of 

herbaceous species. Holechek and Pieper (1992) recommend 30 to 40 percent utilization of key forage species 

as a moderate level of grazing for sagebrush grassland. 

For the entire region, West (1983a) indicated stocking 

rates were originally 0.83 AUM/ha, declined to 0.27 

AUM/ha in the 1930s due to loss of perennial grasses, 

and were 0.31 AUM/ha by 1970.
USDA Forest Service RMRS-GTR-169. 2006. 75

Because of the cheatgrass invasion throughout much 

of the inter-mountain west, fire is not considered beneficial in sage control, since it highly favors the annual 

cheatgrass (Vincent 1992). In situations where cheatgrass is not a problem, Schmutz and others (1967) 

suggested that ungrazed relicts in the Great Basin juniper-sagebrush type (see above section; Jameson and 

others 1962) burning eliminates the pinyon and juniper and increases big sagebrush, whereas at his study 

site the absence of fire and grazing favored grasses and 

reduced shrub and tree composition. Considering the 

region from Navada and Utah to the southwest, West 

(1983a) suggests controlling wildfires in sagebrush 

communities because of the susceptibility of perennial 

grasses relative to annuals, loss of litter and nutrients, 

and a resulting decline in productivity.

In the Great Basin, a greater pressure on the shrub 

component in sagebrush shrubsteppe occurred during 

the Pleistocene by browsers such as camels and ground 

sloths (Martin 1970). While pronghorn antelope can 

consume relatively large amounts and deer up to 50 

percent, shrub intake by all present day ruminants are 

limited by the volatile oils that can inhibit rumen microbial activity when present in high amounts. West 

(1983a) suggested that reintroduction of a browser such 

as the saiga antelope may provide more efficient utilization than that obtained by livestock and native large 

herbivores. Although often in reference to more northern communities, early accounts of the herbaceous 

component in this system suggest a very sparse cover 

existed even before large numbers of livestock were in 

the area. West (1983a) comments that Jedediah Smith 

had difficulty finding forage for his horses in this community type during the 1820s and 30s. Livestock were 

first brought into sagebrush areas in Utah in the late 

1840s, and livestock depletion of the more northern upland communities began in the late 1800s.

There are few data to base conclusions concerning the 

effects of grazing in Great Basin sagebrush shrubsteppe 

when restricted to the southwestern region. The limited 

amount of data suggests very slow or no recovery 

in terms of lower shrub abundance even after longterm complete protection of more arid sites that were 

previously degraded by very heavy grazing. Weak (due 

to potential site differences) and limited data suggest 

relatively greater herbaceous and lower shrub cover on 

sites with little historic grazing and continued decline 

in condition on heavily grazed communities. Grazed 

versus protected sites can possibly show differences in 

herbaceous species composition.

Shinnery Oak/Sand Sage_____

No studies were found concerning grazing of this 

community type. Note that the Kleiner and Harper 

(1972, 1977) studies covered under the Temperate 

Grasslands section has a sand sage component and is 

considered by West (1983d) to be a Utah galleta-threeawn shrub steppe community.

Great Basin Temperate 

Desert_____________________

Very little research is available for these arid, low 

productivity communities. For blackbrush communities 

in southeastern Utah and northeastern Arizona, West 

(1983c) gives an average annual precipitation value of 

160 mm, with most occurring in late summer. In pristine 

sites never grazed by livestock, aboveground primary 

production ranges from 30 g/m2/yr in stands of nearly 

pure blackbrush to 169 g/m2/yr where soils allow a good 

understory of grasses. Grazing value of these communities is considered very low, and blackbrush itself is 

spiny, of low quality, and contains secondary compounds. Navajo Indians use this range for mixed herd 

goat and sheep grazing. Desert bighorn sheep were the 

primary native large herbivores utilizing the community. Aboriginal people made little use of the community, 

and livestock use as winter rangeland was confined to 

the more accessible areas with available water.

Similarly, much of the research in shadscale communities has been done far north of Arizona and New 

Mexico (West 1983b). Sheep grazing was the primary 

use, since there was little grass forage for cattle in these 

communities. As in the blackbrush type, shadscale type 

has never been suitable for livestock grazing and was 

little used by aboriginals. Based on research primarily 

in northern communities, West concluded that recovery 

from very heavy grazing is very slow with “progressive 

livestock management, and little can be done otherwise 

to accelerate improvement.” Intensive livestock management is not practical in this community type due to 

the very low primary production.

The limited number of studies in more southern communities are reviewed below.

Species Composition

Blackbrush communities—Thirteen years of grazing management at a demonstration ranch in southern 

Nevada, where intensity and season of grazing were 
76 USDA Forest Service RMRS-GTR-169. 2006.

controlled, resulted in no improvement in understory 

herbaceous production in a blackbrush community. 

Halliday (1957) reported “practically no forage produced” at the start of the study in 1938 and no change 

in production by 1955. At an associated grassland community, protection from summer grazing improved 

production of warm-season grasses, but April grazing was still preventing improvement in cool-season 

grasses.

A series of ungrazed and grazed sites were studied 

by Jeffries and Klopatek (1987) in northern Arizona 

and southern Utah with annual precipitation of about 

163 mm/yr. A relict (ungrazed by domestic livestock) 

on a mesa represented the pristine condition, although 

it was also inaccessible to wild ungulates as well as 

livestock. Another site experienced little or no grazing by livestock until 1981, due to lack of water, and 

light to moderate winter grazing after water development. A third site in the Navajo Indian Reservation was 

continually and heavily grazed by cattle horses, sheep, 

and goats since the 1880s. A fourth site represented a 

recovery from heavy grazing after it was isolated by the 

creation of Lake Powell in 1974, an island, similar to a 

10 year exclosure from grazing. Total herbaceous cover was four-times or more greater on the never-grazed 

geologic refuge than on any other site and tended to be 

higher on the long-term lightly grazed site than either 

heavily grazed or the 10-year protected site (Table 16). 

There was a general shift in grass species composition from Indian ricegrass in the relict site to galleta in 

the heavily grazed sites, although absolute amounts of 

galleta were similar among sites. Extremely large differences were found in cryptogram cover, where the relict 

or lightly grazed sites had at least 23 times the cover of 

the heavily grazed site or the site excluded from heavy 

grazing for 10 years. Shrub cover was also greater on 

the relict and lightly grazed site than the heavily grazed 

or recovering site. The authors concluded that recovery 

of cryptogram cover on sandy soils of the blackbrush 

type is much longer than that found for some heavier 

soils, that blackbrush is able to survive heavy grazing 

but is also the natural dominant climax on relict sites, 

that even light grazing impacts herbaceous vegetation 

in this community, and that recovery of the herbaceous 

community is also a slow process.

Very heavy grazing by feral burros in Grand Canyon 

National Park was reported to reduce cover and density 

of total vegetation, seed production, and diversity in both 

richness and evenness components (Bennett and others 

1980). Grasses were particularly heavily utilized and 

affected by burro grazing. At the same site, Carothers 

and others (1976) found canopy cover of 80 percent in 

an ungrazed site compared to 20 percent in the grazed 

site and a 30 percent increase in species richness on the 

ungrazed. Increased mistletoe infestation of mesquite 

appeared correlated with heavy grazing. The effects 

on vegetation translated into effects on small mammal 

communities, with much greater abundance and diversity on control compared to burro-grazed site. Douglas 

and Leslie (1996) provide a review of feral burro grazing in the western United States and the interaction of 

burros with desert bighorn sheep.

Shadscale communities—Three exclosures were 

studied by Hughes (1980) in Arizona that were built 

in grassland at the edge of a shadscale shrub community. Twenty-eight years later shadscale and snakeweed 

dominated the area. However, at one of the exclosures, 

shrubs attained a greater dominance outside than inside 

the exclosure. Snakeweed was twice as abundant outside 

as inside the exclosure, even though it also increased inside. Galleta and blue grama decreased at the expense of 

the shrub increase. At a second site, shadscale increased 

from nearly nothing to 80 percent of the species composition outside, replacing galleta and alkali sacaton. Galleta 

remained the dominant inside the exclosure even though 

shadscale invaded there as well. Conclusions from a 

third exclosure were confounded by contour furrowing 

outside the exclosure, but shrubs invaded this altered 

and grazed area while the exclosure remained nearly devoid of vegetation. These results in general may suggest 

differences in the response of shadscale communities 

compared to blackbrush communities, but may also just 

be due to invasion/expansion dynamics versus responses 

of long established communities.

Management

West (1983c) thought there was a low potential for 

grazing systems to effect change in blackbrush communities and that fire should not be used in management. 

Shrub control at a Nevada site resulted in much more 

erosion and instability in primary production. West 

considered feral burros as a threat to remote communities where livestock grazing had not previously depleted 

grasses. Halliday (1957) suggested that April grazing 

appears detrimental to cool-season species in an associated grassland community type, particularly Indian 

ricegrass, but there were no grazing management effects 

on the blackbrush type. Mixed-herd grazing would seem 

to be an area where future research may be useful.

The very limited number of studies in this 

community type in the southwest suggest that areas not 

previously grazed or only lightly grazed should remain 

ungrazed or lightly grazed. Recovery from grazing is 
USDA Forest Service RMRS-GTR-169. 2006. 77

slow. The relationship of grazing to shrub cover is not 

clear, although in some situations at least, high shrub 

dominance is the natural climax, while in others grazing 

may increase shrub abundance.

Sonoran/Mohavian Deserts___

Precipitation in the Sonoran Desert lies between that 

for Mojave and Chihuahuan Deserts in both the quantity 

and the percent that occurs during winter (MacMahon 

and Wagner 1985). Mean annual precipitation ranges 

from 75 to 300 mm/yr for the Sonoran Desert and 50 

to 275 mm/yr for the Mojave Desert, compared to 175 

to 425 mm/yr for the Chihuahuan Desert. The percentage of precipitation that occurs in winter is greatest for 

the Mojave, intermediate for the Sonoran, and least 

for the Chihuahuan Desert. MacMahon and Wagner 

(1985) illustrate good relationships between decreasing 

variability in precipitation with increasing amounts of 

precipitation, and increasing amounts of precipitation 

with increasing elevation. Shrub cover and standing 

biomass increases with increasing precipitation, with 

231 g/m2 standing live-plus-dead at a Mojave Desert 

site with 138 mm/yr and 747 g/m2 at a Sonoran Desert 

site with 279 mm/yr. Primary production at the same 

Mojave Desert site in southern Nevada ranged from 

a low of 20 g/m2/yr for perennials and near zero for 

annuals to about 60 g/m2/yr for each. Production at a 

Sonoran site in Arizona over eight years ranged from 

2.4 to 71.3 g spring grass/m2, 0.9 to 84 g spring forbs 

/m2, 3.6 to 39.2 g autumn grass /m2, and 12.5 to 30.8 g 

autumn forbs /m2 (Smith and LeCount 1976, cited in 

Wallmo and others 1981).

There is a wide variety of plant communities in both 

the Sonoran and Mojave Deserts, and within a site the 

Sonoran Desert is particularly diverse in terms of both 

floristics and physiognomy (MacMahon and Wagner 

1985). Temperature determines the appearance of the 

two distinct summer and winter species. The evolutionary origin of the summer species was Mexico and 

Central America, while winter species were primarily 

from the Pacific coast of California. Consumption by 

native herbivores is not a large proportion of primary 

production. MacMahon and Wagner speculate that consumption by livestock is also low because much of the 

vegetation is woody, allelopathic, or spinescent.

Humphrey (1960) also commented on the low potential for livestock utilization of Sonoran Desert 

communities because of the low grass production 

and unpalatability of the larger shrub component of 

production. As a result, there is little change in plant 

community species composition under long, continued 

use, little improvement in range condition under light 

use, and little noticeable deterioration under heavy grazing. On the more alkaline soils, saltbush can provide 

some good forage, but creosote bush has no forage value. Humphrey considered it difficult to classify this type 

into range condition classes. Sites heavily dominated by 

creosote bush or by paloverde and triangle bur-sage produce very little forage and cannot be converted to useful 

rangeland even with shrub control and seeding of herbaceous species. Thin soils with an indurated caliche layer 

may hold more moisture near the surface and produce 

some forage.

While the large pool of unpalatable perennial woody 

and cacti species may not be altered much with grazing, Shreve (1929) reported that the effects of grazing 

on small perennials may be large in some cases. A study 

of vegetation change after a 21-year period of protection indicated no consistent successional trend but an 

increase in small perennial plants from 33 to 164 percent on level sites and a gain of 7 percent to a loss of 

16 percent on hill slopes. However, the measurement of 

Table 16—Vegetation cover in a Great Basin temperate desert community dominated by blackbrush in northern Arizona/

southern Utah in a never grazed geologic refuge (mesa), lightly grazed, heavily grazed, and an island ungrazed for 

10 years after a dam caused formation of Lake Powell. Adapted from Jeffries and Klopatek (1987).

Grazing treatment

Cover categories None Light Heavy Ungrazed 10 years

Total shrub cover (m2/ha) 3645a* 2874b 1405c 1372c

Blackbrush cover (percent of total shrub) 87a 83a 92a 95a

Total herbaceous cover (m2/ha) 1047a 256b 102b 127b

Indian ricegrass cover (percent total herbaceous) 80a 22b 3b 76a

Galleta cover (percent total herbaceous) 2b 11b 32b 76a

Total cryptogamic cover (m2/ha) 2129a 1196a 70b 50b

Total cover (m2/ha) 6821a 4326b 1576c 1549c

* Values within a row not sharing the same letter are significantly different.
78 USDA Forest Service RMRS-GTR-169. 2006.

only two points in time, without adjacent grazed sites 

also being monitored, makes it difficult to separate grazing from climate effects. Climate effects in ungrazed 

refuge communities can be large. In a crater refuge site 

in Sonora, Mexico, Turner (1990) found that drought and 

unusually high periods of precipitation can profoundly 

affect community species composition. Creosote bush 

and paloverde populations declined by more than 50 

percent during the first half of the century, mesquite increased 200-fold in the playa-like center of the crater, 

and there were only three major peaks in recruitment of 

saguaro cacti over 170 years.

Prolonged periods of drought are common in the 

southwestern deserts, and grazing may interact with 

these periods of stress. Robinett (1992) reported no visible effect of grazing on drought mortality at a southern 

Arizona Sonoran Desert site in a grassland-desert transition zone receiving 305 mm/yr precipitation. However, 

recovery after a drought appeared to be promoted by 

the presence of litter that may have trapped seeds and 

lowered evaporative losses. This author recommended 

that ranchers grazing these communities be particularly 

aware of the need to de-stock during droughts whereby 

litter is left on the soil surface for subsequent recovery 

of the plant community.

Some unpalatable species may be susceptible to 

herbivores during the seedling stage. In the Sonoran 

Desert, the conspicuous and representative saguaro 

cactus and paloverde can be limited by herbivores to 

establishment in biotic or geologic refuges. Paloverde 

establishment is apparently limited by native mammalian herbivores to locations under unpalatable shrubs 

or among rock outcrops (McAuliffe 1986). Similarly, 

saguaro cacti mortality due to rodents and insects is 

much lower in rock outcrops, even though microenvironmental conditions are more unfavorable (Steenbergh 

and Lowe 1969). McAuliffe (1984) found that tree cacti 

protected two species of barrel cacti from rodents in 

the Sonoran Desert. While these studies were in areas 

ungrazed by domestic livestock and livestock find all 

unpalatable, trampling can be a factor restricting even 

unpalatable cacti species to refuges from large herbivores in other communities (Rebollo and others 2002). 

Abouhaidar (1992, cited in Wilson and others 1996) observed that young saguaro were much more abundant in 

areas where cattle had been excluded.

Some quantitative studies of the effects of grazing 

on Sonoran Desert communities have been reported. 

Blydenstein and others (1957) compared plant communities protected inside the Carnegie Desert Laboratory 

in southern Arizona for 50 years to an area outside 

that was grazed only lightly. Only minor differences in 

species composition were found, but the density of plants 

was significantly greater in the protected area (Table 

17). Increases in density with protection were primarily due to perennial grasses and one palatable shrub. The 

only species to increase on the grazed sites was a bursage. The dominant creosote bush showed no response 

to the different grazing regimes. Annuals did not appear 

to respond to protection, but the timing of sampling and 

climatic conditions at that time did not permit a good 

evaluation. Although the author commented that no invasions of new species occurred, there were 11 species 

sampled in the protected area that were not sampled in 

the grazed area, and only one sampled exclusively in the 

grazed area.

A decline in diversity with increasingly recent grazing was also observed by Waser and Price (1981) at 

sites inside and outside Organ Pipe Cactus National 

Monument in southern Arizona. The communities were 

dominated by creosote bush. The cover composition of 

winter annuals was sampled after 2 and 3 years of protection, after 16 and 17 years of protection, and on two 

adjacent heavily grazed sites. The decline in diversity 

with grazing was primarily due to loss of less common 

species. Total cover or the cover of the three most abundant species did not differ consistently among grazing 

treatments between years and replicates.

One of the three sites studied by Roundy and Jordan 

(1988) was considered Sonoran Desert. The site received an average of 222 mm/yr precipitation, and both 

the grazed and an area protected for 14 to 18 years were 

dominated by half-shrubs and large Ephedra and mesquite shrubs. Half-shrubs were slightly more abundant 

in the grazed area, and compositional differences were 

also somewhat small. Full-shrubs densities did not differ, but there were some compositional differences. 

Mesquite and Ephedra were more abundant in the protected area, and an Acacia was more abundant in the 

grazed. Perennial grasses and forbs were very low in 

abundance in both grazed and protected treatments. 

Annual forbs were greater in grazed than ungrazed 

areas, and there was no difference in annual grasses. 

Root-plowed treatments also showed only small increases in perennial and annual grasses. The authors 

concluded that “there is no evidence that exclusion from 

grazing alone or even in association with root-plowing will encourage reestablishment of native perennial 

grasses.”

A study of the effects of feral burro grazing on 

Mojave desertscrub communities found reductions 

in palatable perennial grasses, total plant cover and 

density, and plant diversity (Bennett and others 1980). 

These authors noted, however, that the abundance of 
USDA Forest Service RMRS-GTR-169. 2006. 79

Table 17—Relative frequency (Raunkaier’s frequency index) 

of a Sonoran Desert site in southern Arizona protected 

from grazing for 50 years or grazed lightly by livestock. 

A * represents a significant difference between the two 

treatment means. Adapted from Blydenstein and others 

(1957).

Frequency Index

Plant species or catagory Protected Grazed

Whitethorn acacia 9 8

Annuals 316 332

Threeawn spp. 13 9

Spider grass 3 ---

Dense ayenia 8 12

Red grama 29 * 2

Fairyduster 15 * 4

Saguaro 1 1

Yellow paloverde 2 7

Woody crinklemat 22 * 9

Condalia spp. 1 ---

Prairie clover spp. --- 1

Silver bush spp. 1 ---

Candy barrel cactus 1 ---

Ocotillo 13 * 3

Triangle burr ragweed 252 * 324

Slender janusia 2 ---

Sangre de cristo 9 4

White ratany 147 * 93

Creosote bush 73 64

Berlandier’s wolfberry 2 ---

Graham’s nipple cactus 1 ---

Rough menodora 2 ---

Bush mulhy 14 * 3

Prickly-pear cactus spp. 17 * 6

Cactus apple 24 19

Whiplash pappusgrass 1 ---

Mariola 2 ---

Paperflower spp. 18 * 5

Slim tridens 49 * 19

Low woollygrass 210 * 131

Zinnia spp. 1 ---

No Vegetation 57 * 91

species of low palatability that burros ignore tends to 

obscure impacts on the smaller palatable plants.

Deer densities in Sonoran and Mojave Desert habitat tend to be lower than for many other communities 

(Urness 1981). A review of diet studies showed that 

plants such as Wright buckwheat, mistletoe, cactus 

fruits, catclaw acacia, mesquite beans, and fairy duster 

leaves were important in deer diets in Sonoran Desert, 

and paloverde, desert ironwood, mesquite, and catclaw acacia important in Mojave Desert. The effects of 

deer browsing on the plant communities have not been 

reported.

Merriam’s kangaroo rats consumed nearly 7 percent of primary production at a Mojave Desert site in 

California (Soholt 1973). While this is a high level of 

consumption for a rodent population, it is small relative to the amounts generally consumed by livestock. 

However, the small mammals are much greater dietary 

specialists, consuming 75 percent seed material, and 

can therefore have relatively large impacts on plant communities. Redstem stork’s bill was a preferred food for 

kangaroo rats at this site, where utilization was over 95 

percent. Seed consumption was great enough to reduce 

density of that species by 30 percent. In southeastern 

Nevada, rodent consumption of seeds accounted for 30 

to 80 percent of losses, but no more than 25 percent 

of germination (Nelson and Chew 1977). These authors 

concluded that it takes “an exceptional coincidence of 

events...to cause a severe depletion of seeds reserves” 

and reported only slight evidence for decreases of largeseeded species due to rodent consumption.

Interactions of grazing with fire in these communities has in the past been considered as probably not 

important. Fuel loads are generally too sparse to carry 

fires over large areas (Schmid and Rogers 1988). These 

authors estimated that it would take 274 years for fires 

to affect the entire 391,000 ha desert portion of Tonto 

National Forest in Arizona. Allen (1996) considers the 

Sonoran Desert to be a system that did not evolve with 

frequent fires and many of the grasses associated with 

this community to be susceptible to fire injury and mortality. However, fuel loads can be high after periods or 

years of unusually high precipitation (McLaughlin and 

Bowers 1982), and the additional fuel load from introduced grasses and human-caused fires have increased 

the frequency of fires in many areas (Rogers 1985; 

Wilson and others 1996). Rogers (1985) suggests that 

saguaro cactus may be endangered over portions of its 

range when fire frequency is less than the 30 years necessary for them to reach reproductive maturity. Saguaro 

cacti are important to many cavity nesting birds, bats, 

and other wildlife and are the symbol of the southwestern desert, although they are of restricted distribution. 

Saguaro mortality after fires is related to age/size class, 

with younger classes displaying very high mortality 

rates of up to 100 percent (Cave and Patton 1984). Some 

large saguaros that survive fires and others that display 

delayed post-fire mortality provide a much diminished 

seed supply for regeneration (Cave and Patton 1984; 

Rogers 1985; Wilson and others 1996). Grazing can interact with fire in regeneration of saguaros because of 

its negative effect on seedling establishment (reviewed 

above).

The very limited number of studies in Sonoran/

Mohavian Deserts does not allow for strong 
80 USDA Forest Service RMRS-GTR-169. 2006.

conclusions to be drawn. Plant diversity appears to 

decline with grazing, although overall changes in 

species composition are relatively small. Grazing does 

not appear to have large effects on the major woody 

components, but impacts the less abundant and smaller 

species. Regeneration of saguaro cacti in areas of the 

Sonoran Desert can be negatively impacted. Recovery 

of a perennial herbaceous understory may be site 

dependent, and it is not clear whether the difference in 

recovery is due to site potential to support a herbaceous 

understory or due to the intensity of historic grazing.

Riparian and Wetland 

Communities_______________

Riparian and wetland communities represent a very 

small percentage of the land area in the southwest but 

are areas of high plant and animal diversity and productivity. Regulation of water flow and the elimination 

of winter-spring flooding, lowering of water tables due 

to irrigation, clearing, and wood cutting, as well as 

livestock grazing have and are currently resulting in 

large losses of all types of these habitats (Minckley and 

Brown 1982). Riparian areas and wetlands provide water and cover to animals that may be more associated 

with adjacent upland communities, including livestock, 

as well as many species that are obligate to these communities for part or all of their life cycle. These areas 

are probably more important to animals more associated with uplands in arid and semiarid regions because 

of the refuge from the harsh environment they can provide. Artificially built watering areas are often known 

as “sacrifice areas” because of the high trampling and 

utilization of forage near them. Resting grounds and 

watering areas are known to be centers of high nutrient 

concentration due to the deposition of feces and urine. 

Natural watering areas with greater production, a shaded favorable microenvironment, and scratching posts 

are most probably even more utilized by livestock than 

are artificial watering areas. Livestock impacts due to 

trampling effects on soil compaction are often evident 

on sparsely stocked unproductive upland communities 

so would be expected to be much greater in a heavy-use 

area, and wet soils are more susceptible to cutting and 

compaction than are dry soils. Unlike shrub invasion 

problems in upland communities, riparian zone overstories are often reduced by livestock grazing (Kauffman 

and Krueger 1984), and this strata provides cover and 

nesting for many land vertebrates and affects water temperatures for aquatic organisms. Stream-side vegetation 

exerts a large influence on bank and channel morphology through effects on flow velocities, cutting during 

flood conditions, and erosional inputs from uplands. 

There is a potential for these productive areas to be impacted by livestock to a relatively greater degree than 

adjacent less productive communities, but there is also 

the potential for more rapid recovery from disturbance 

because of faster growth rates of the vegetation.

Narrowleaf Cottonwood

Two narrowleaf cottonwood community types, in addition to two Broadleaf cottonwood community types, 

along the Mimbres River in southwestern New Mexico 

were examined by Boles and Dick-Peddie (1983). Early 

accounts of the appearance of the river in 1846 described 

a 4.6 m wide by 0.9 m deep river. Current conditions observed by Boles and Dick-Peddie were of a river with 

similar water course width, but much shallower water 

course and a much wider riverbed. This suggested that 

vegetation along the river bank had been reduced, because an effect of vegetation is to constrict and deepen 

a river channel. Vegetation along a river can be altered 

by grazing by livestock and/or changes in the historical 

water flow patterns and levels. Although no controlled 

grazing studies were accomplished, the authors present 

observations that suggest livestock grazing may be a primary cause of change along the river. Changes included 

factors that are commonly attributed to livestock grazing 

in other controlled studies: erosion, absence of reproduction by woody components of the vegetation, large 

cottonwoods uprooted by floods, and absence of moist 

alluvial microsites for seed germination. Although irrigation use could lower water tables, no permanent dams 

occurred on the river and signs of flooding were evident 

and substantiated by long-time residents. Data were collected showing an even-age-class distribution of obligate 

riparian woody species, and observations indicated severe grazing and browsing of most new woody shoots. 

However, this occurred primarily for obligate riparian 

species that included two broadleaf cottonwoods but not 

the narrowleaf cottonwoods. A much higher density of 

facultative riparian species (pines and junipers that also 

occur in the dry uplands) are associated with narrowleaf 

cottonwood communities than with broadleaf communities. This would suggest that the less mesic communities 

may have been favored. The Mimbres is a canyon-type 

drainage compared to the more broad, flat flood-plain 

type. Thus, while the authors report a wider riverbed, the 

vegetation favored is a canyon type. This would suggest 

that flash-flooding may be more common and a more favorable mesic flood plain condition less common than 
USDA Forest Service RMRS-GTR-169. 2006. 81

in the past, or that narrowleaf cottonwood is just less 

susceptible to grazing. Controlled grazing experiments 

would be needed to differentiate mechanisms, and to 

also clearly separate grazing from potential lower watertable influences on population structure and composition. 

The extreme grazing pressure observed on cottonwood 

sprouts, their emergence within biotic refuges (logs 

preventing access by grazers), and evidence of erosion 

would all point to grazing as at least a major factor.

Experimental evidence for some of the observations of Boles and Dick-Peddie (1983) were obtained by 

Skartvedt (2000). This author sampled 17 sites along 

the upper Mimbres River in what appears to be primarily narrowleaf cottonwood communities (dominants 

across all sites in decreasing order were ponderosa 

pine, alligator juniper, and narrowleaf cottonwood). 

Six of the 17 sites had been ungrazed for at least 10 

years. Sites were not paired grazed-ungrazed, but varied with elevation, stream direction, channel type, as 

well as grazing history. It is not clear from the authors 

description how confounded the various variables were, 

but principal component analyses were used to partition variance. Only one out of three graphs of PCA axes 

plotted against grazing showed any discernable correlations, where ungrazed sites showed a higher cover of 

less dominant woody species. Dominant constraints on 

plant communities were climate, hydrology, and elevation, with grazing a secondary factor. However, plots of 

the cover of four woody obligate riparian species showed 

extremely low values for grazed compared to ungrazed 

sites. Unfortunately, the degree of confounding factors 

due to other site differences was not discussed by the 

author, but two of the six protected sites were at a lower 

elevation Nature Conservancy site.

A 10-year exclosure was compared to a lightly 

grazed site in a narrowleaf cottonwood - gambel oak 

- ponderosa pine - Arizona walnut community along 

McKnight Creek in southwestern New Mexico (Medina 

and Martin 1988). Protection from grazing did not prevent channel erosion, a decline in herbaceous vegetation 

cover, or decreasing tree cover over time. No differences between grazing treatments was found for channel 

morphology or species composition of the plant community. The cause of high sediment transport, stream 

channel aggradation, and loss of plant cover was a large 

wildfire in the headwater section of the stream 24 years 

prior to erecting the exclosure and 34 years prior to 

the last reported measurements and subsequent storm 

events. Although grazing was very light and treatment 

differences minimal, the study demonstrates the very 

long-term nature of changes that can be brought about 

by an event associated with surrounding communities. 

While the authors considered the events to have “upset the dynamic equilibrium of the stream,” Asplund 

and Gooch (1988) considered flash flooding and stream 

degradation and aggradation to be the natural dynamics of cottonwood, disturbance-based communities (see 

below). Medina and Martin (1988) did comment that 

establishment of willows was greatest during years of 

greatest change in channel morphology. The disturbed 

or unstable condition of the streambanks was more conducive to establishment than were other soils. Unstable 

conditions of aggradation may be necessary for periods 

of establishment and disruption of even-aged conditions 

(Asplund and Gooch 1988).

Broadleaf Cottonwood Forest

No information within the Arizona - New Mexico 

area was found specific to broadleaf cottonwood forest 

community type 21 in Appendix 1. Information below 

pertains to broadleaf cottonwood forest community 

type 22.

In contrast to adjacent Madrean woodland and pinyon-juniper communities that displayed very slow 

recovery to protection from grazing, cottonwood-sycamore-alder riparian areas protected from grazing for 

10 years displayed significant increases in tree, shrub, 

and forb diversity compared to when the area had been 

grazed (Tolisano 1995). Additional measurements of areas where grazing continued indicates that regeneration 

of overstory riparian species was rare to nonexistent. 

Cottonwood and sycamore trees in the grazed areas 

were all 24 inches in diameter or greater compared to 

all size classes being represented in the ungrazed areas. Boles and Dick-Peddie (1983) also observed an 

even-age structure in Freemont cottonwood and lanceleaf cottonwood in two broadleaf community types 

along the Mimbres River (see Narrowleaf Cottonwood 

section above). A similar even-age structure of the narrowleaf cottonwood population was not observed. Boles 

and Dick-Peddie cite personal communications with 

Hilderbrant and Ohmart, and with Spellenberg, for 

similar observations along the Salt and Verde Rivers in 

Arizona and for sycamores in the Guatalupe Canyon in 

New Mexico, respectively. Based on visual observations 

along the same rivers, Davis (1982) suggested that the 

Santa Rita four-pasture rest-rotation grazing system with 

proper stocking management was favorable for cottonwood reestablishment goals, even though many authors 

in other parts of the country and Ames (1977) suggest 

that only complete protection is effective. Sponholtz 

(1997) observed 60 percent less vegetation, 50 percent 

greater channel width, 85 percent lower flow velocity, 
82 USDA Forest Service RMRS-GTR-169. 2006.

and 50 percent lower fish density at a grazed compared 

to ungrazed section of the Verde River.

The Sonoita creek in southeastern Arizona supports a 

broadleaf cottonwood-willow-ash-sycamore community 

that is surrounded by an oak-mesquite-juniper community (Glinski 1977). Measurements of the number and 

size-class of cottonwoods and sycamores were made in 

a segment of the stream ungrazed by livestock for eight 

or more years and in four areas grazed by cattle and 

horses, cattle, or horses. The ungrazed site was at the 

highest elevation, followed by the grazed sites sequentially downstream. Sycamore sprouts were the main 

form of reproduction, and cattle did not appear to inhibit reproduction by sprouting. In contrast, cottonwood 

reproduces only by seed, and trampling and grazing of 

seedlings by cattle appeared to severely limit reproduction. Horses did not graze seedlings but trampled some 

and stripped the bark from some saplings.

In contrast to the above studies showing a deleterious effect of grazing on cottonwood establishment, 

Asplund and Gooch (1988) found that recruitment of 

Fremont Cottonwood was a function of stream geomorphologic features and refugia from flood damage rather 

than on the absence of livestock grazing. Recruitment 

was found to be low in a “virgin” ungrazed canyon, 

and competition with mature trees was thought to be a 

factor. The authors conclude that recruitment primarily occurs where aggradation of materials was from 

upstream degradation and erosion. Flash flooding and 

stream degradation and aggradation were considered 

to be the natural dynamics of this disturbance-based 

community. Uncontrolled correlative studies of the effects of grazing on recruitment need to consider tree 

distributions over a long stretch of the watershed before 

eliminating the lack of natural geomorphic processes as 

the cause for the lack of recruitment. The authors argue 

that high sediment loads and upslope erosion are natural 

factors necessary for community persistence, and that 

management practices that emphasize stabilization of 

channels and floodplains are ignorant of the necessary 

natural processes required for recruitment.

An exclosure and an adjacent grazed plot in cottonwood-willow riparian areas along the southern 

Arizona-New Mexico border were sampled the year 

after establishment and over the subsequent 10 years 

(Hayward and others 1997). The area was at one time 

an 8 km long marsh until irrigation projects lowered the 

water table, so vegetation change due to this was overlaid on both grazing treatments. The primary objective 

of the study was to monitor small mammal populations, 

so vegetation data was limited and no description of the 

grazing regime was provided. Height of vegetation was 

greater in the exclosure in 13 of 18 measurement periods, and vegetation visually appeared more dense.

Minckley and Brown (1982) provide some general 

descriptions of understory plant community species 

composition changes that may be associated with grazing of Sonoran broadleaf cottonwood forest. Historically 

ungrazed communities would support a variety of native 

annual and perennial grasses, forbs, and several species 

of saltbushes in more saline areas. Grazed or otherwise 

disturbed communities would be likely to have an exotic component that can include filaree, mustards, red 

brome, and in open places Schismus and Bermuda grass. 

Other invasion problems also occur in overstory species, 

where saltcedar and tamarisk replace native vegetation. 

This has been linked to intermittent flooding during the 

growing season, which may or may not be indirectly 

associated with livestock grazing. Saltcedar is highly 

flammable, and fire can maintain a saltcedar disclimax 

scrub community.

One of the few studies measuring herbaceous understory vegetation as well as shrub and tree abundance 

and size distributions was conducted along Little Ash 

Creek in central Arizona in a Fremont cottonwood, velvet ash, Gooding willow community (Szaro and Pase 

1983). An area protected from grazing for four years 

showed “extremely limited” recovery compared to an 

adjacent grazed area. This conclusion appeared to be 

based on no significant grazing effects on overstory 

canopy cover or tree growth (average radial growth). 

There were large differences in size distributions of 

Fremont cottonwood between grazing treatments. No 

cottonwoods under 22.5 cm dbh were found on the 

grazed site while 48 percent of the trees on the ungrazed site were less than 22.5 cm. However, this would 

suggest there were initial differences in the sites prior 

to establishment of the exclosures. Size distributions of 

7.5 and 15 cm dbh were found on the ungrazed but not 

on the grazed site, yet it would not be possible for a tree 

to grow to a 15 cm diameter over four years. Total basal 

area of cottonwoods increased 10 percent in two years 

on the ungrazed treatment compared to 5.7 percent on 

the grazed treatment. Understory cover after four years 

of treatment was 5.2 and 2.2 percent for the ungrazed 

and grazed treatment, respectively. Differences in herbaceous understory species composition and diversity 

were also apparent. Ten species (one in trace amounts) 

were sampled in the ungrazed exclosure compared to 

seven (four in trace amounts) in the grazed treatment. 

Unfortunately, pre-treatment conditions were not sampled, and the possibility of prior differences based on 

cottonwood size distributions makes interpretation of 

understory responses difficult as well.
USDA Forest Service RMRS-GTR-169. 2006. 83

Ohmart (1996) provides photographic evidence for 

very large changes in vegetation structure along a portion of the San Pedro River in southeastern Arizona 

before and four years after livestock removal to create 

a National Riparian Conservation Area. Similar dramatic photos are shown of not-managed compared to 

24-year-managed sections of Date Creek in Arizona. 

Willows and cottonwood had developed along the creek 

in the managed section, and trees and herbaceous vegetation in this area survived a flood much better than 

in the not-managed section. However, the management 

regime was not described. Additional qualitative and 

quantitative data for the San Pedro River are presented 

in Krueper (1993, 1996), U.S. Fish and Wildlife Service 

(2002), and Krueper and others (2003). Herbaceous, 

shrub, and low and tall tree components of the vegetation was estimated as foliage height density prior 

to cattle removal, and two, five, and eleven years after 

cattle removal in Fremont cottonwood/Gooding willow 

riparian, mesquite grassland, and Chihuahuan desert 

communities. Herbaceous plant density doubled after 

two years in both riparian and grassland communities. 

Increases in herbaceous vegetation density continued 

from year two to year five after removal of livestock 

at lesser rates for low strata but at similar rates for upper strata, then remained similar from year five to year 

eleven. The shrub and tree components in both riparian and grassland communities showed little response 

over the period of study, and no significant responses 

to removal of livestock were ever observed for the 

Chihuahuan desert communities.

Only one other study assessed both understory and 

overstory vegetation responses to protection from grazing, and this study included three replicates along Trout 

Creek in Arizona (Reichenbacher 1984). The ungrazed 

treatments were geologic refuges formed by granite 

rubble, making this unique for riparian grazing studies. 

Reichenbacher describes the pristine ungrazed sites as 

“remarkable riparian luxuriance” unlike anywhere along 

the grazed stretches of the river. Understory species 

composition was drastically different between grazing 

treatments (Table 18). Ungrazed sites had dense stands 

of American bulrush, cattail, sweet clover, and tree 

saplings, while grazed sites were heavily dominated by 

the shrubby, opportunistic seepwillow. Total vegetative 

cover was nearly three-times greater, and density nearly 

seven times greater, on ungrazed compared to grazed 

areas. Diversity (calculated as H’, which incorporates 

both richness and evenness components of diversity) 

was nearly three-times greater in ungrazed compared to 

grazed sites (2.83 vs. 1.0, respectively). The high density 

Table 18—Plant density (#/464 m2) and cover (percent) along Trout Creek, Arizona, broadleaf cottonwood riparian communities 

in never grazed geologic refuges and grazed sites. The refuges were formed by granite rubble that prevents livestock entry 

(area for density represents smallest site). Values are an average of three sites/treatment. Adapted from Reichenbacher 

(1984).

Ungrazed refuge Grazed

Plant Density (#/464m2

) Cover (percent) Density (#/464m2) Cover (percent)

Shrubs and herbs

Mule’s fat 37 0.60 4,951 5.84

White clover 984 4.99 39 0.09

Salt cedar 26 0.01 144 0.02

Singlewhorl burrobrush 48 2.97 48 0.87

Sweetbush 0 0.00 39 0.64

Desert false indigo 32 0.53 0 0.00

sub total 1,127 9.14 5,220 7.46

Emergent aquatics

American bulrush 41,248 8.20 880 0.00

Pale spikerush 0 0.00 368 0.00

sub total 41,248 8.20 1,248 0.00

Trees

Green ash 5 1.46 3 3.23

Black willow 62 48.79 11 10.68

Fremont cottonwood 16 1.15 1 3.70

Soapberry 26 8.61 0 0.00

sub total 108 60.12 15 17.61

Total vegetation 42,483 69.25 6,483 25.07
84 USDA Forest Service RMRS-GTR-169. 2006.

of trees in ungrazed refuges was due to large numbers 

of seedlings and saplings.

Riparian Scrub

No information was found specific to this type. 

However, willow and alder are components of narrowleaf and broadleaf cottonwood forest types, and 

some species found there are also found in the riparian scrub type (see Appendix 1). General management 

considerations for riparian habitat are covered below in 

the Grazing Management section.

Other Riparian

Arizona willow is an imperiled species that has a 

very restricted range, from southern Utah, on one mountain in northern Arizona, and recently found in New 

Mexico (NatureServe Explorer 2002). Cattle and elk 

grazing in particular, and timber harvest and off-road 

vehicles are considered primary threats to this species. 

The importance of cattle versus native ungulate grazing of Arizona willow was assessed by Strohmeyer and 

Maschinski (1996) and Maschinski (2001) by exposing 

planted individuals to no grazing, open to cattle grazing only, or open to native grazers only for a one year 

period. Both native and domestic herbivores reduced 

mean plant height, number of branches, branch length, 

and total shoot length. Elk in particular selected for 

the willows, while similar evidence of deer or antelope 

herbivory were not observed. While having a similar 

effect, cattle selected for grasses amongst the willows 

and consumed willows only after other herbaceous forage had been consumed. The studies were conducted in 

both artificial and natural community settings and demonstrated that willow damage can be attributed to both 

wild and domestic herbivores additively. Maschinski 

(2001) noted that this species of willow is sensitive to 

herbivory and that it requires years to recover from 

defoliation. Stein and others (1992) also reported very 

heavy use of arroyo willow by elk, but this was of the 

nutritious resprouts after fire.

Rinne (1985, 1988) sampled some structural characteristics of the vegetation along the Rio de la Vacas in 

Sante Fe National Forest in conjuction with a study of 

fish populations in grazed and ungrazed sections of the 

third-order montane stream. No plant species descriptions were provided, but stream-bank vegetation cover 

averaged 8.4 and 1.0 percent for ungrazed and grazed 

treatments, respectively, and overhanging vegetation 

cover was 17 and 0 percent, respectively. An index of 

bank instability for grazed sections of the stream ranged 

from 20 to 100 percent, with an average of 64 percent, 

compared to 0 percent for all six ungrazed plots.

A riparian area along the Paria River in north-central Arizona, southern Utah, had been fenced from 

grazing for six to eight years with no establishment of 

new cottonwoods, even though a seed source was available (Bezanson and Hughes 1989). No description of 

the plant community was provided, but surrounding 

vegetation included saltbush - Indian ricegrass communities. Heavy grazing and flash-floods had previously 

degraded much of the vegetation. A grazing system 

whereby grazing occurred during November through 

January for two years, followed by a year of rest (6 

months grazed, 30 months rested), allowed cottonwood 

recovery after four years. The lack of sprouting of cottonwood during the protected period prior to implementing 

the grazing system would suggest that years suitable 

for reproduction of cottonwoods were not frequent, but 

this particular grazing system may have allowed for recovery when conditions were favorable. This was not a 

controlled experiment and no data were provided, but 

photographs indicate much improved conditions.

Another non-experimental observation was reported 

by Grette (1990) in an unidentified plant community 

along McCoy Gulch in south-central Colorado surrounded by pinyon-juniper communities. Woody 

riparian vegetation at the site was described as nonexistent due to unregulated winter grazing, and uplands 

were in poor condition even though stocking was moderate. When the grazing season was changed from 

winter to spring, cattle use shifted from willows to 

grass, resulting in an increase in willows. Although 

improving, the riparian area was fenced to allow better 

control of grazing. A grazing objective was established 

whereby willow leaders were allowed to be grazed only 

once (removing about 12 cm), thereby acting like a mild 

pruning. Utilization limit of preferred grasses was set 

at 80 percent, and supplemental protein was provided 

in the uplands. Photographs indicate good recovery of 

riparian woody species, and perennial grass and forb 

cover increased.

Tamarisk is a woody plant that was introduced into 

the United States in the early 1880s as an ornamental (Horton 1977). The plant spread throughout flood 

plains of the southwest and became a permanent vegetation type. Tamarisk uses more water than native species 

and also supports a less diverse faunal community. The 

plants can grow as much as 3.5 m in a year, and grazing 

of 50 percent of the foliage does little to prevent stands 

from becoming very dense.
USDA Forest Service RMRS-GTR-169. 2006. 85

Vegetation along streambanks can affect water flow 

and, therefore, sediment transport rates. A series of 

1- to 3-year-old exclosures was studied along the Rio 

Senorito in north central New Mexico, focusing on solute transport and retention (Sewards and Valett 1996). 

Even with short-term exclosure from livestock grazing, 

channel widths were restricted by vegetation growth in 

the exclosures, and channel depth was greater. These 

differences in channel structure resulted in a greater 

linear flow velocity in exclosed compared to grazed areas, but lateral areas with dense vegetation allowed for 

a four to five times greater retention time. The authors 

hypothesized that the lateral areas of increased residence 

time are important for biological processing of incoming material and could also influence sediment transport 

and deposition. Grazing of riparian areas could thereby 

increase occurrence of flash-flooding, and subsequent 

grazing of vegetation could slow recovery.

In contrast, Thibault and others (1999) measured 

greater nutrient retention in grazed sections of four 

streams in New Mexico due to greater in-stream vegetation. Greater in-stream vegetation and wider widths 

of streams slows water flow and allows for greater utilization of passing nutrients. However, ungrazed portions 

of the streams generally had greater cover of riparian 

vegetation. Two of the four streams had greater cover 

of herbaceous vegetation in ungrazed sections compared to grazed, one had greater in the grazed section, 

and one did not differ with grazing treatment. Two of 

the four streams had greater woody plant cover in the 

ungrazed sections. Implications of the differences in 

nutrient dynamics in streams due to grazing have not 

been investigated.

Dry Washes

Only four reports of grazing effects on dry wash communities were located, two of which were of feral burro 

grazing. Chew (1982) reported much greater response 

to protection for wash communities than for adjacent 

creosote bush or tarbush communities. Differences in 

standing biomass between inside versus outside exclosures (after adjustment for grazing loss) were about 

2, 4, and 73 g/m2 for creosote bush, tarbush, and dry 

wash communities, respectively. Differences between 

ungrazed and grazed washes were also large for density and basal cover of total vegetation excluding fluff 

grass. Chew concluded that washes, like riparian areas, 

have a potential production much greater than their 

area and than adjacent uplands, and recovery following protection from very heavy grazing can be much 

greater in these productive sites. Gardner (1951) also 

reported large grazing intensity effects on washes in 

the Rio Grande valley, New Mexico, although not as 

great as those reported by Chew. Gardner compared 

uncontrolled grazing with light grazing, and the lesser 

differences between this and a completely protected site 

can be due to heavy grazing of preferred washes even 

under general light stocking of the entire area. Grasses 

were common in lightly grazed areas and rare under 

uncontrolled grazing (0.06 percent vs. trace basal cover, 

respectively). Total shrub and forb covers did not differ much between grazing treatments, although species 

composition differences were evident. Creosote bush 

was a high proportion of shrubs in areas with uncontrolled grazing, and with two other species-dominated 

sites. Lightly grazed areas had more even proportions of 

seven to eight shrub species.

Grazing by feral burros in Grand Canyon National 

Park was observed to impact mesquite riparian scrub 

communities much greater than three other community types because of the relatively greater use (Bennett 

and others 1980). Plant cover, density, and diversity 

declined in burro impacted compared to control areas. 

Grazing by burros in secondary washes of Sonoran 

Desert ranges in Arizona was found to decrease total 

canopy cover from 8.6 to 2.8 percent, and of white bursage (burrobush) from 2.26 to 0.04 percent (Hanley and 

Brady 1977). Secondary washes were the most utilized 

plant community in the area, with some areas receiving 100 percent utilization of current-year’s growth. No 

species were observed to increase or invade under the 

heavy grazing.

Wetlands

The amount of wet meadows or cienagas in New 

Mexico and Arizona National Forests has been estimated at about 17,700 ha (Patton and Judd 1970). These 

communities occur within coniferous forests and may be 

important to wildlife populations because of their high 

productivity, even though they represent only 0.2 percent of the area. Productivity averaged approximately 

300, 150, and 20 g/m2/yr for the wet meadow, transition 

zone, and surrounding dry forest, respectively. Protein 

contents of meadow/transition community species were 

also greater than dry forest species. Arnold (1954, cited 

in Patton and Judd 1970) estimated that wet meadows produced four times the amount of herbaceous 

vegetation than associated ponderosa pine communities. Patton and Judd found the greatest plant species 

diversity in transition communities, with only 20 species recorded in the wet meadow compared to 43 in the 

moist transition and 28 in the dry forest communities. 
86 USDA Forest Service RMRS-GTR-169. 2006.

However, deer use of the forest edges was much greater than for wet meadow/transition communities, and 

elk displayed similar but less pronounced preferences. 

This was thought due to predator avoidance behavior, 

but also because the pellet count method of utilization 

probably underestimated the actual importance of the 

time spent foraging in meadows. In contrast, cattle use 

of the meadow/transition communities was from 4 to 10 

times more than for forest edge communities. Turkeys 

and ducks were observed to also preferentially use the 

meadow/transition communities during certain periods 

of the year.

At the site described above for Broadleaf Cottonwood 

Forest (Hayward and others 1997), an additional paired 

ungrazed-grazed plot was established in a cienaga dominated by sacaton grass. Height of vegetation was greater 

in the exclosure in 18 of 20 measurement periods, and 

vegetation visually appeared more dense.

In White Mountain riparian meadows dominated by 

Carex, Cyperus, and Juncus species, Neary and Medina 

(1996) found that trampling and grazing of native and 

domestic animals resulted in disruption of aggradation 

processes. Tearing of the root-soil matrix leads to loss 

of fine materials and down-cutting riffles. Grazing of 

aboveground material reduces the filtering and deposition of sediment. Introduced grasses such as Kentucky 

bluegrass, wheat grasses, or orchard grass do not have 

the same fibrous, thick root mats like native grasses, and 

break down easier under trampling. Eventually, downcutting of the channel can lead to changes in stream 

bank vegetation from aquatic to more mesic species.

Grazing Management

Kovalchik and Elmore (1992) provide a good review 

of grazing systems and management effects on northern, colder climate, willow dominated communities. A 

good review of grazing management of riparian habitat 

based on studies from throughout the western United 

States can be found in U.S. Fish and Wildlife Service 

(2002). Clary and Webster (1990) recommend spring 

grazing for intermountain ranges, as described above 

by Grette (1990). Differences between riparian and upland vegetation lushness and quality, and microhabitat 

differences (temperature, and so forth), are least during 

spring. Fall is recommended as a second-best option, 

while mid- to late-summer grazing should be avoided in 

most areas because the hot dry conditions make riparian habitat more preferred. A stubble height of grasses 

of 10 to 16 cm should be left to entrap sediment and 

protect streambanks during high flow. Other recommendations of Clary and Webster include: 1) limit use 

of herbaceous forage to 65 percent and remove animals 

while plants are still in a vegetative stage (this seems 

very high utilization compared to even use-levels of uplands reviewed above; note that Kovalchik and Elmore 

1992 indicate that livestock start utilizing willows when 

herbaceous forage use reaches about 45 % in Oregon), 

2) utilization should be less if grazed in summer, because of lower regrowth potentials, 3) fall grazing levels 

should be closely monitored, because no regrowth will 

occur, and use should be even lower at 30 to 40 percent, 

4) relatively sensitive habitat will require even more 

conservative use, 5) provide salt and alternative water 

away from riparian areas, and 6) rest of heavily degraded streambanks and/or woody vegetation may require 

from 1 to 15 years for recovery, followed by the above 

to maintain a good condition. Along the Verde River 

in Arizona, seasonal differences in use-preference by 

livestock described above were observed on an annual 

basis (Tonto National Forest unpubl. data cited in U.S. 

Fish and Wildlife Service 2002). Use of riparian areas 

was greater during dry winters, when upland vegetation was relatively less available than in good winters. 

In this case, use of woody vegetation by livestock increased after bud break in late February to early March. 

The above recommendations seem based on reasonably 

sound logical concepts in terms of proposed mitigation 

practices but, unfortunately, there is little hard scientific 

evidence to support them.

It seems rather surprising that, given the volumes of 

literature written on riparian systems in the southwest, 

such little actual research has been accomplished. This 

appears the general case for riparian research throughout 

the western United States (Larsen and others 1998). 

Some research is available on tree regeneration, with 

very little on understory vegetation. There are problems 

with statistical design and replication in riparian 

communities that are more difficult to overcome 

than in upland communities, due to the linear and 

elevational aspects of stream flow. Initial pre-treatment 

background measurements need to be considered 

of even greater importance in riparian systems than 

in upland communities due to especially high levels 

of variability in physical conditions along streams. 

Because of natural, continual aggradation/degradation 

processes that change in location along stream lengths 

where flooding and catastrophic events are common, 

there is an extreme need to consider sufficient length and 

replication of treatment areas. Although the congregation 

of livestock in riparian areas undoubtedly causes severe 

alterations in plant community species composition 

and structure in communities that historically have not 
USDA Forest Service RMRS-GTR-169. 2006. 87

experienced a grazing force by large herbivores, the 

experimental evidence showing the effects is weak. 

Based on a number of reports and a number of visual 

observations, cottonwoods that reproduce by seed are 

reduced by grazing, while root sprouting species are 

less affected. Channel morphology is altered, and the 

potential implications are great. Unfortunately, very 

little substantial work has been accomplished specific 

to southwestern riparian understory plant communities, 

but what information is available shows much greater 

effects of grazing than for upland, non-riparian 

communities (see also Comparison of Southwestern 

Community Types section below).
88 USDA Forest Service RMRS-GTR-169. 2006.

Relative Effects of Grazing Among 

Southwestern Community Types, 

and Comparison of SW with Great 

Plains Community Responses: Does 

Evolutionary Grazing History Matter?

The detailed assessments of plant community response 

to grazing examined above to some extent relies on 

each study author’s interpretation of their results. The 

compilation of these accounts within a particular community type to some extent depends on a reviewers 

interpretation, and a broader perspective of comparing 

among community types can be complex, qualitative, 

and difficult. Quantitative comparisons among community types using community-wide variables and indices 

often provide insight into relative sensitivities of different communities to controlled conditions or treatments 

and into factors contributing to the different sensitivities. A meta-analysis-like approach may sometimes 

bring order to what at first appears to be complex and 

unassociated responses. Even though methods and experimental designs differed, an assessment of 236 pairs 

of ungrazed-grazed treatment comparisons yielded explanations of factors that explained 54 to 69 percent of 

the variance in grassland or shrubland plant community species composition responses to grazing across 

communities throughout the world (Milchunas and 

Lauenroth 1993). Explanatory variables included evolutionary history of grazing, aboveground productivity 

or precipitation, intensity of grazing and years of protection and grazing treatment (see The Importance of 

Evolutionary History section above).

The objectives in this section are to examine all ungrazed-grazed comparisons reviewed in the sections 

above to see whether there are any discernable differences among community types in response to grazing, 

and then compare southwestern communities with short 

histories of grazing to communities in the adjacent 

Great Plains of North America with a long evolutionary 

history of grazing by native, large, generalist herbivores 

(bison).

Convergent selection pressures of herbivory and aridity in systems with a long evolutionary history of grazing 

and frequently low soil water availability are predicted 

to result in communities least responsive to grazing 

by introduced livestock. In contrast, divergent selection pressures of herbivory and an ability to compete 

for light in the canopy in more productive, subhumid 

environments are predicted to result in communities 

most responsive to grazing (Milchunas and others 1988, 

and History of Grazing by Native and Domestic Large 

Herbivores section above). Rapid switching from short 

species adapted to grazing to tall species adapted to 

competing in the canopy, and visa-versa, may occur 

with temporal and spatial changes in herbivore abundance in the latter. Intermediate in responsiveness are 

communities with a short evolutionary history of grazing. Productive communities with a short history of 

grazing by large herbivores do not have a suite of species capable of withstanding heavy grazing, whereby 

local species losses can occur and/or a few weedy ruderals eventually increase in abundance. Unproductive 

communities with a short history of grazing have some 

degree of tolerance/avoidance of herbivory due to adaptations to aridity but have not developed in the presence 

of the large herbivores. Therefore, the order of plant 

communities from least to most responsive to grazing 

can be hypothesized as unproductive/long history < 

unproductive/short history < productive/short history 

< productive/long history. The responsiveness does not 

possess a “good or bad” component, other than invasion of exotics or opportunistic species from outside the 

community would be most likely in the productive/short 

history situation followed by the unproductive/short history situation, and least likely in the communities with a 

long history of grazing.
USDA Forest Service RMRS-GTR-169. 2006. 89

Table 19—Examples of dissimilarity values calculated using the Whittaker Index of Community Association, and percent change in total species 

abundance, for comparisons of hypothetical ungrazed versus grazed communities. These hypothetical community examples may aid in 

interpretation of data for real ungrazed versus grazed community comparisons in figures that follow.

Comparison Example A Example B Example C Example D Example E Example F

Community Ungrazed Grazed Ungrazed Grazed Ungrazed Grazed Ungrazed Grazed Ungrazed Grazed Ungrazed Grazed

Species 1 50 50 50 0 90 90 90 900 98 91 98 5

Species 2 40 40 40 0 9 1 1 10 55 61 55 13

Species 3 30 30 0 50 1 9 9 90 10 13 10 61

Species 4 20 20 0 40 0 0 0 0 7 5 7 91

------- ------- ------- ------- ------- ------- ------- ------- ------- ------- ------- -------

Total 150 150 90 90 100 100 100 1000 170 170 170 170

Percent dissimilarity 0 100 8 8 5 79

Percent change 

abundance 0 0 0 900 0 0

Below, the Whittaker (1952) index of community 

association is used as a community-wide dissimilarity 

metric. A comparison of an ungrazed to an adjacent 

grazed community that yields a value of zero indicates 

that all species present in the ungrazed community are 

also in the grazed community, and every species is 

present in the same proportion (Table 19 example A). 

A dissimilarity value of 100 percent means there is not 

even one single species that was sampled in both communities (Table 19 example B). Index values are based 

on the proportions of all species in the two communities being compared and do not indicate differences in 

abundance but indicate proportional differences (Table 

19 example C compared to D). The percent difference 

in total vegetative cover is used as a second metric, to 

evaluate differences in abundance between the ungrazed 

and grazed communities being compared, since differences in abundance do not affect dissimilarity indices 

when proportions are similar. For example, differences in abundance between the ungrazed and grazed 

communities in example C in Table 19 are 0+8+8 divided by a total abundance of 90+9+1+90+9+1 which 

equals 8 percent. This is the same as the dissimilarity 

index of 8 percent because abundances are the same. 

However, the same calculation of difference in abundance for example D in table 19 is 810+1+89 divided by 

90+9+1+900+10+90 equals 82 percent which is not the 

same as the dissimilarity value of 8 percent. The dissimilarity measures do not place a value on particular 

species and quantitatively provide an overall index of 

change in community species composition between two 

treatments/sites based on all species but do not account 

for differences in abundances of the species.

Some information emerges for a simple comparison 

of community types for the southwest, without taking any other factor into consideration (Fig. 7). One 

possible conclusion may be that riparian community 

species compositions are consistently and very drastically affected by grazing (average dissimilarity of 84 

percent), although there are only four riparian comparisons from two studies. Grassland, shrubland, and forest 

communities averaged approximately the same differences in plant species composition between grazed 

versus ungrazed treatments, but desert communities 

were relatively less responsive to grazing. Great Basin 

temperate desert communities also have a high average 

dissimilarity, but there is very high variability within 

this community type. The two very high dissimilarity 

sites were both shallow-upland Bigelow sagebrush/galleta grass/blue grama sites where the ungrazed site 

had been protected for 73 years. Chihuahuan Desert, 

Sonoran/Mohavian Desert, interior chaparral, and Great 

Basin juniper/sagebrush communities appear particularly unresponsive to grazing/protection treatment. This 

may be due to the high shrub/non-forage component 

that varies little with grazing. Very small differences 

between ungrazed-grazed Chihuahuan Desert sites were 

also observed. Sites in Fig. 7 were sorted by increasing precipitation within a particular community type. 

Only pinyon-juniper and mesquite savanna community 

types have enough data to assess effects of increasing 

precipitation on grazing responses within a community 

type. No relationship of increasing effects of grazing 

with increasing precipitation was observed for the mesquite type (r2

 = 0.02). A slightly negative relationship 

was found for pinyon-juniper communities (r2 = 0.34, 

slope -0.002), possibly indicating an overriding effect 

of overstory control on the response to grazing.

Differences among communities in grazing response 

with respect to total vegetation abundance are less clear 

than that for species composition (Fig. 8). This may 

be due to the fact that method of measurement in the 
90 USDA Forest Service RMRS-GTR-169. 2006.

studies would likely have more influence on this variable. 

Estimates based on canopy cover or grazed, non-temporary-caged biomass versus ungrazed biomass would by 

definition of consumption due to grazing more likely 

show reductions. Estimates based on basal cover, productivity, density, or frequency may also show reductions, 

but not just due to current-year grazing unless immediate 

mortality occurs. Regardless of method, reduced plant 

abundance has implications for habitat structure for other consumer populations, even if it does not for long-term 

plant population dynamics (see Introduction section for 

additional interpretive definitions). In general, most studies and most community types showed declines in total 

vegetation abundance with grazing. Although limited by 

the number of studies, large and/or mostly reductions in 

abundance were observed in riparian (with one exception), temperate grassland, and Great Basin sagebrush, 

ponderosa pine, and pinyon-juniper types. Although only 

two comparisons, montane meadows showed increases in 

plant abundance with grazing. The most notable conclusion that can be drawn from Fig. 8 is the preponderance of 

negative values (67 negative, 28 positive). This is not the 

general direction of response for Great Plains grasslands 

(Milchunas and others 1989; Lauenroth and others 1994) 

or in other systems with long evolutionary histories of 

grazing (Milchunas and others 1988), where increases in 

cover of prostrate, grazing adapted species often occurs. 

Interestingly, blue grama consistently increases in Great 

Plains grasslands but does not always do so in southwestern communities (studies reviewed above). Blue grama 

actually decreased in two out of 13 community types assessed by Dahl and others (1976).

How do responses differ between plant communities 

of the southwest and Great Plains? Are there different 

multiple variables that may better explain variability in 

responses shown in Fig. 7, and do they differ between 

southwestern and Great Plains communities? First, over 

a low precipitation to similar high precipitation cutoff 

for Great Plains communities, average dissimilarity between ungrazed versus adjacent grazed communities 

was greater for Great Plains communities compared 

to southwestern communities (Fig. 9). The difference 

between ungrazed versus grazed communities increased with increasing precipitation for Great Plains 

Figure 7—Plant community species dissimilarity (grazed versus ungrazed percent) for plant community 

types for which ungrazed and grazed sites have been compared. Dissimilarity values calculated using 

the Whittaker (1952) index of community association. Community types are: 3 = mixed conifer forest, 

5 = ponderosa pine, 7 = pinyon-juniper, 9 = Great Basin juniper/sagebrush, 10 = interior chaparral, 12 

= montane meadow, 13 = temperate grassland, 14 = mesquite savanna, 15 = Great Basin sagebrush, 

17 = Great Basin temperate desert, 18 = Sonoran/Mohavian Desert, 19 = Chihuahuan Desert, 22 = 

Broadleaf cottonwood forest (M) riparian, 24 = dry washes. Community numbers correspond to the 

numbering system for community descriptions in Appendix 1.
USDA Forest Service RMRS-GTR-169. 2006. 91

communities and explained 22 percent of the variance 

among sites even though study methods, grazing intensities, and years of protection differed. In contrast, 

precipitation explained only eight percent of the variance for southwestern communities. The relationship 

between precipitation and change in species composition with grazing for the two regions can possibly better 

be seen by dividing studies into four groups and doing 

an analysis of variance. Large differences between arid 

and subhumid communities in the effects of grazing on 

plant species composition were seen in the Great Plains, 

with subhumid communities being very responsive and 

arid ones least responsive (Fig. 10). Also as predicted, 

southwestern communities were intermediate to the two 

types of Great Plains communities. However, subhumid 

southwestern communities were only slightly more 

responsive to grazing than arid southwestern communities, and arid southwestern communities only tended to 

be more responsive to grazing than arid Great Plains 

communities. The increase in change in species composition with increasing community productivity predicted 

by the Milchunas and others (1988) theoretical model, the Milchunas and Lauenroth (1993) quantitative 

global-community model, and various other studies relating increasing productivity to increasing response to 

grazing pressure (Milchunas and others 1989; Proulx 

and Mazumder 1998; Osem and others 2002, 2004) 

tended to hold within southwestern communities, but 

differences were not significant. Southwestern communities were more variable in response to grazing 

within each community type than were Great Plains 

communities.

There was a relationship between years of protection from grazing and plant community dissimilarity of 

grazed versus protected sites for Great Plains communities, but not for southwestern communities (Fig. 11). 

Even though the maximum years of protection from 

grazing for the studies available was 47 years for Great 

Plains grasslands compared to 73 years for southwestern communities, the r2

 values were 0.29 for the Great 

Plains compared to 0.03 for the southwest. As above, 

average dissimilarity between ungrazed versus grazed 

communities was greater for Great Plains communities 

(43 percent) compared to southwestern communities 

(33 percent), while average years of protection was 

greater for southwestern communities. Extremely 

Figure 8—Plant community total vegetation abundance (grazed / ungrazed percent) for plant 

community types for which ungrazed and grazed sites have been compared. Community 

types are: 3 = mixed conifer forest, 5 = ponderosa pine, 7 = pinyon-juniper, 9 = Great Basin 

juniper/sagebrush, 10 = interior chaparral, 12 = montane meadow, 13 = temperate grassland, 

14 = mesquite savanna, 15 = Great Basin sagebrush, 17 = great Basin temperate desert, 

18 = Sonoran/Mohavian Desert, 19 = Chihuahuan Desert, 22 = Broadleaf cottonwood forest 

(M) riparian, 24 = dry washes. Community numbers correspond to the numbering system for 

community descriptions in Appendix 1.
92 USDA Forest Service RMRS-GTR-169. 2006.

Figure 9—Plant community species dissimilarity (grazed versus ungrazed percent) in relationship to increasing precipitation (productivity) within southwestern and Great Plains plant 

communities for which ungrazed and grazed sites have been compared. Dissimilarity values calculated using the Whittaker (1952) index of community association. Values at the 

top of figure are average dissimilarity, average long-term precipitation at the study sites, 

and r2 for dissimilarity versus precipitation regression within geographic location.

Figure 10—Plant community species dissimilarity (grazed versus 

ungrazed percent) for all studies 

categorized into four groups based 

on southwestern United States 

versus Great Plains plant communities and arid versus subhumid 

environments. Arid versus subhumid communities were those above 

or below 410 mm/yr precipitation, 

which is used as a surrogate for level of primary productivity because of 

greater availability of precipitation 

than productivity data. Number of 

observations for each bar from left to 

right is 78, 35, 25, 44. Value above 

each bar is the standard deviation of 

that mean and represents individual 

differences in variance. See text for 

description of similarity data.
USDA Forest Service RMRS-GTR-169. 2006. 93

interesting was the difference between natural geologic, 

relict, never-grazed refuges compared to human-made, 

shorter-term, previously grazed exclosures. This was 

true for both Great Plains and southwestern communities. The average spread between exclosure versus relict 

comparisons was greater for southwestern communities 

than for Great Plains communities. However, an average dissimilarity in ungrazed versus grazed community 

species composition of 33 percent for human-made 

exclosures versus 64 percent for never grazed versus 

grazed communities for the southwest, and similarly 

43 percent versus 61 percent for the Great Plains, were 

very large differences due to type/length of protection. 

This may in some part be due to micro-climatic effects 

of refuges, but differences between never-grazed refuges and grazed communities that appear greater than 

for recent human-made exclosures are seen worldwide 

and in many cases cannot be attributed to micro-climate (Milchunas and Noy-Meir 2002). Separating 

micro-climatic effects from historical grazing effects 

of relict sites would seem to be a potential interesting 

avenue of investigation, but only a few studies (partially 

published comparison Ambos and others 2000 and 

Ambos and others unpublished data Table 12 personal communication, Milchunas and Noy-Meir 2004) 

attempt to compare or differentiate between natural refuges and human-made exclosures.

Averaged over all studies, grazing intensity effects 

on plant community species dissimilarity between ungrazed versus grazed communities shows a trend of 

increasing dissimilarity with increasing grazing intensity from light to moderate to heavy grazing for both 

southwestern and Great Plains plant communities (Fig. 

12). Communities in both geographic locations seem 

to in general show increasing differences between 

ungrazed versus grazed communities as grazing intensity increases from light, light/moderate, moderate, to 

moderate/heavy. Regression analyses for southwestern 

communities indicated grazing intensity explained 16 

percent of the variance over all communities. When 

the analyses was confined to grasslands alone, the r2

increased to 0.2. Grazing intensity was not related to 

Figure 11—Plant community species dissimilarity (grazed versus ungrazed percent) in relationship 

to increasing years of protection from grazing within southwestern and Great Plains plant communities for which ungrazed and grazed sites have been compared. Dissimilarity values calculated 

using the Whittaker (1952) index of community association. Values at the top of figure are average 

dissimilarity, average years of protection from grazing at the study sites, and r2 for dissimilarity 

versus years of protection regression within geographic location.
94 USDA Forest Service RMRS-GTR-169. 2006.

total vegetative abundance differences between ungrazed - grazed comparisons.

None of the single-factor models explained much 

of the variance associated with southwestern communities, although better relationships were observed for 

Great Plains communities. This could be due to other 

interacting factors, so multiple regression analyses 

were preformed on data for both the southwest and 

the Great Plains. For the southwest, analyses included 

both grassland and all communities combined. There 

were not enough data to examine shrub/forest communities separately. Great Plains communities were 

primarily grasslands. All models were run on dissimilarity data in coefficient form, not as percentages, so 

slopes reported below are based on dissimilarity values 

ranging from 0.0 to 1.0, not 0 to 100 percent. A good 

multivariate model could be developed for dissimilarity between ungrazed-grazed comparisons for Great 

Plains communities. Three factors (precipitation, grazing intensity, and years of grazing treatment/protection) 

together explained 48 percent of the variance among 

studies. Dissimilarity in species composition between 

grazing treatments was most influenced (greatest slope) 

by grazing intensity (0.09), followed by years of grazing 

treatment (0.0025), and precipitation (0.0006), with a 

model intercept of -0.17.

Models for southwestern communities were not at 

all as capable of explaining variance in the data as for 

the Great Plains. The most successful model for plant 

community species dissimilarity between grazing 

treatments was a single factor model for grasslands only. 

Twenty-three percent of the variance in grassland plant 

dissimilarity with grazing could be explained by years 

of grazing treatment/protection. The relationship was 

positive, where increasing years of protection resulted in 

greater dissimilarity between ungrazed and grazed sites 

(slope = 0.0033). Grassland plant community dissimilarity increased with increasing grazing intensity (slope 

= 0.063) and explained 20 percent of the variance.

Other models assessed percent change in total vegetative cover for grasslands and for all communities 

combined. For grasslands alone, only 15 percent of the 

variance could be explained by precipitation (slope = 

0.021) and years of grazing treatment/protection (slope 

= 0.005). For all communities combined, only 11 percent of the variance could be explained by precipitation 

(slope = 0.04) and years of grazing treatment/protection 

(slope = 0.0057).

It is interesting that good models can be produced 

for Great Plains plant community species composition 

response to grazing but not for southwestern communities given an even larger data set for the latter. One 

possible conclusion is simply that southwestern communities are just not very predictable. Large variability 

in response to grazing appears to occur both within 

and between community types relative to Great Plains 

plant communities (Fig. 9). The shrub invasions into 

some areas and not others reviewed in sections above, 

independent of grazing treatment, is another example 

of the more “random” southwestern phenomenon that 

are less clearly explained by grazing. Factors contributing to the response of Great Plains plant communities 

to grazing are twice as predictable as for southwestern 

communities. Alternatively, factors not considered 

here in analyses and not commonly reported in grazing studies may be explanatory variables. However, 

intensity of grazing and years of protection from 

grazing would be thought to be important within a 

Figure 12—Plant community species 

dissimilarity (grazed versus ungrazed percent) in relationship to 

increasing grazing intensity within 

southwestern and Great Plains 

plant communities for which ungrazed and grazed sites have 

been compared. Dissimilarity values calculated using the Whittaker 

(1952) index of community association. Values at the top of bars are 

the number of studies the value is 

based upon.
USDA Forest Service RMRS-GTR-169. 2006. 95

region, even though precipitation/productivity and 

evolutionary history of grazing are more important explanatory variables across global regions (Milchunas 

and Lauenroth 1993).

How does the southwest compare to the highly grazing-resistant shortgrass steppe, and to the rest of the 

world, with respect to change in plant community species composition with grazing? The shortgrass steppe 

has some species in common with some southwestern 

grasslands, and possibly the most closely related community types would be temperate grasslands (which 

includes some southwestern shortgrass steppe) and 

mesquite savanna. In general, the shortgrass steppe is 

among the most grazing tolerant of plant communities 

in the world, averaging nearly three times less change 

in plant community species composition than the rest 

of the world (Fig. 13). Temperate grasslands and mesquite savannas of the southwest were more responsive 

to grazing than shortgrass steppe but were near the 

mean dissimilarity for semiarid/arid communities 

with a short history of grazing in the rest of the world. 

Further, the mean dissimilarity between grazed and 

ungrazed sites for all arid southwest communities (Fig. 

3, 33 percent dissimilarity) was very close to the global 

mean for other semiarid/arid communities with a short 

evolutionary history of grazing (Fig. 4, 32 percent 

dissimilarity). In contrast, the mesquite savanna community type averaged 36 percent dissimilarity between 

grazed versus ungrazed sites, which is slightly higher 

than for other semiarid/arid communities with a short 

evolutionary history of grazing. However, precipitation 

for mesquite savannas averaged 374 mm/yr compared 

to 293 mm/yr for the other communities. On the more 

productive side, subhumid southwestern communities 

averaged 41 percent dissimilarity (Fig. 10) compared to 

a slightly higher 45 percent global mean for other subhumid communities with a short evolutionary history of 

grazing (Fig. 13). In general, southwestern communities 

were much more sensitive to grazing than the adjacent 

shortgrass steppe, but were similar in response to other 

semiarid/arid and subhumid communities with a short 

evolutionary history of grazing.

Figure 13—Plant community species dissimilarity (grazed versus ungrazed percent) for studies conducted in the shortgrass steppe with a long evolutionary history of grazing, contrasted with dissimilarity for southwestern temperate 

grasslands (13) with a mixed history of grazing and mesquite savanna (14) communities with a short evolutionary 

history of grazing, and communities from around the world based on data in Milchunas and Lauenroth (1993). Value 

above bars are the average for that particular community or group. Dissimilarity values calculated using the Whittaker 

(1952) index of community association.
96 USDA Forest Service RMRS-GTR-169. 2006.

All ecosystems are pulsed through time by many 

different forces, and the living components of the system are continually reacting to those disturbances. 

Because of their predictability and the adaptations of 

the species within a community, some disturbances 

become an integral part of the system. These types 

of disturbances stabilize the system and become a 

part of our management of the system should the historic natural source or frequency of that disturbance 

be altered by human activities or encroachment. For 

example, the removal of fire from southeastern longleaf pine communities would mean the loss of that 

community. Controlled burning of some communities 

may not exactly replicate historical patterns and frequencies but is generally a better approximation than 

no fire at all. Other disturbances are not within the 

genetic “memory” of the systems’ components and can 

cause changes in states that may or may not be permanent. Margalef (1968) termed these two types of 

pulses endogenous and exogenous disturbances. While 

conceptually appealing, defining what is an internal, 

stabilizing disturbance to a system and what is an 

external, destabilizing disturbance is most often very 

subjective. However, differentiating between these two 

classes of disturbance is what we attempt to do with 

respect to our approach to management. Usually it is 

not an either/or situation, but a matter of at what point 

does a particular disturbance become exogenous? If 

the disturbance is exogenous, then management strictly for conservation purposes would dictate its removal, 

unless the alternative results in an equal or greater exogenous disturbance. If the disturbance is exogenous 

and it remains, it may be considered “sustainable” only 

if an alternate state is not the permanent end state. If an 

alternate stable state is reached through deflection by 

some other force, then the previous reference point is 

no longer valid for evaluating the subject disturbance. 

What type of a disturbance is grazing in southwestern 

plant communities?

Some level and type of herbivory is endogenous to 

any ecosystem. However, the review of the evolutionary 

history of grazing of southwestern communities above 

indicates grazing by herds of large generalist herbivores 

is not an endogenous disturbance in southwestern communities, with the exception of the shortgrass steppe 

communities of far eastern New Mexico. That was based 

on geographic occupation by bison, but two lines of evidence presented in this review support that conclusion 

with respect to adaptations of species in the communities. First, dissimilarities in plant community species 

composition of ungrazed versus grazed treatments are 

greater for southwestern comparisons than for Great 

Plains comparisons (above section). This is predicted by 

the generalized model of Milchunas and others (1988), 

and differences are greater than the quantitative global 

comparison between semiarid systems with short versus long evolutionary histories of grazing (Milchunas 

and Lauenroth 1993). Second, the response of a species 

common to both Great Plains communities and some 

southwestern communities is not the same in the two 

locations, even under similar community productivities. 

Blue grama is a species that increases with grazing in 

abundance and composition throughout Great Plains 

plant communities with a long history of grazing by bison. When a component of the plant community, blue 

grama is also known as relatively tolerant to grazing in 

southwestern plant communities. However, decreases 

of this species were noted in 9, and increases in 8, of 

23 reports in the community response sections above. 

The “grazing lawn” (sensu McNaughton 1984, see 

Milchunas and Lauenroth 1989 for shortgrass steppe) 

structure of very heavily grazed communities was not 

noted above; rather, heavy grazing was often noted to 

result in relatively bare, non-vegetated conditions. Short 

grasses commonly evade grazers more effectively than 

taller bunch grasses (Milchunas and Lauenroth 1993; 

Diaz and others 2006), particularly those with large 

proportions of crown organs. Blue grama and other 

Conclusions
USDA Forest Service RMRS-GTR-169. 2006. 97

short grasses avoid excessive losses to grazers to some 

extent in southwestern communities as well, but there 

does not appear to be a similar capacity to increase 

under very heavy grazing. Blue grama is a genetically 

diverse species, and populations in mixed-grass prairie 

do not respond to defoliation treatments similar to even 

shortgrass steppe populations (Kotanen and Bergelson 

2000). Blue grama in the shortgrass steppe forms sodlike “grazing lawns” (Milchunas and Lauenroth 1989) 

while it appears as a bunchgrass under grazing in 

Arizona (J. H. Bock pers. commun.). While bison and 

blue grama co-occurred in eastern New Mexico, it is 

the edge of the range for bison, and the smaller southern 

bison may have been a more selective grazer than its 

large northern counterpart (see Evolutionary History of 

Grazing section). The species that provide the high degree of grazing tolerance of Great Plains communities 

do not appear to provide a similar degree of tolerance to 

southwestern ecotonal communities. To the degree that 

they do, the introduction of large domestic grazers into 

the southwest would only artificially expand the boundary of Great Plains associations. Thus, Bock and Bock 

(1993) questioned whether Appleton-Whittell Research 

Ranch site was a grama grassland, or a community 

created by grazing since the 1800s. Grazing by large 

generalist herbivores may be considered an exogenous 

disturbance in most southwestern communities.

Does livestock grazing lead to alternate stable states 

in plant communities of the southwest, or is it sustainable 

at some level? Although declining with periods of 

drought, the widespread increases in woody vegetation 

across many community types in the southwest, 

coincident with the very heavy grazing of the late 1800s 

and continuing through present, has for a long time been 

blamed on grazing by domestic livestock. From the 

review above, it is clear that the majority of evidence 

indicates that the encroachment occurs into ungrazed 

sites as well as grazed sites. There are four possibilities 

that could still invoke livestock grazing as a factor in 

this conversion of communities: 1) previous very heavy 

grazing during the unregulated grazing of the late 1800s 

resulted in redistribution of soil in a fashion favorable 

for mechanisms described in the Schlesinger and others 

(1990, see review above) and Whitford (2002) grazing/

desertification model (this has been proposed by 

Grover and Musick [1990]), 2) grazing interacts with 

fire, favoring woody encroachment, 3) there are highly 

erodible areas (see discussion of Davenport and others 

1998, pinyon-juniper section) where long-term studies 

have not been in place, where current grazing could 

initiate the mechanisms described in the Schlesinger 

and others model, and 4) grazing interacts with some 

other factor such as climate or rodents. All of these and 

some alternatives have been addressed in the review 

above but are briefly summarized here. The first is a 

remote possibility based on some relict, never-grazed 

geologic refuge studies reviewed above. Due to dispersal 

limitations of woody species, a time lag in occupation 

of local sites favorable and unfavorable to deep water 

penetration could eventually result in conditions 

described in the Schlesinger model. Based on exclosure 

studies reviewed above, current grazing does not result 

in dynamics described in the model in the locations 

studied. No management or policy options could alter 

outcomes if present dynamics were set in motion a 

century ago, and a century time-lag in dispersal would 

not seem realistic. Second, reductions in fuel loads 

due to removal by grazers, and fire suppression, would 

influence results of exclosure studies because both 

grazed and ungrazed treatments would be affected by 

the absence of fire. Very large ungrazed areas adjacent to 

grazed areas are needed to adequately test this possibility 

(Bock and Bock 1993). In some situations, however, a 

fire return interval shorter than the predicted natural 

interval would be necessary to cause enough mortality 

to prevent encroachment (see Mesquite Savanna 

Conversion to Chihuahuan Desert section). While a 

more frequent fire return interval in some southwestern 

communities may more closely mimic the endogenous 

fire disturbance, this may be a separate issue to the 

shrub encroachment one. Third, highly erodible soils 

are always more prone to irreversible vegetation change 

due to grazing. It is possible that the long-term study 

sites and the US/Mexican border analyses reviewed 

above were not in highly erodible areas. This is an 

area of comparison that would also make an interesting 

test of conditions necessary to initiate mechanisms in 

the Schlesinger and others model. However, given the 

number of study sites reviewed above, it would seem 

unlikely for this to on its own explain the extent of 

conversion that has been reported. Fourth, grazing can 

interact with other factors, and both climate and small 

mammals have been implicated as a potential factor 

in vegetation change in the southwest. Grover and 

Musick (1990) argue that all factors (grazing, fire, small 

mammals, and climate) interact complexly with positive 

feedbacks that can have long time-lags in producing 

the shift to shrublands. With respect to conversion 

of community types, however, changes in climate 

do not differ across fenceline comparisons. Further, 

climate interactions with livestock grazing and native 

herbivores can, in at least some situations, mitigate the 
98 USDA Forest Service RMRS-GTR-169. 2006.

effects of climate as a cause of shrub expansion (Curtin 

and Brown 2001, and see review above). There is an 

increasing amount of evidence implicating changes in 

the seasonal proportions of precipitation as a primary 

cause of community change in the southwest (reviewed 

above). Why only in the southwest, and what are the 

implications for assessing the impacts of livestock 

grazing in the southwest?

There are two reasons why the seasonal distribution 

of precipitation may affect community conversion in 

the southwest. Semiarid and arid Mediterranean plant 

communities have climates where precipitation falls in 

greater proportions in the winter and are characterized 

by a dominance by shrubs and annuals. In dry regions, 

deep rooted shrubs are favored by deep storage of winter precipitation. In contrast, relatively shallow rooted 

grasses are favored by precipitation falling when they 

are active and capable of intercepting the moisture. 

Annuals are capable of rapid utilization of moisture 

in spring, yet they do not require a season-long supply 

for survival, and remain as seed through dry summer 

periods. The southwestern United States is not a true 

Mediterranean climate, but much of the area does receive a relatively more equally proportional amount of 

winter/summer precipitation than a strictly continental 

climate. A second factor in the distribution of southwestern communities is the relatively diverse local 

topography, which plays an important role in the proximity of shrubland-grassland communities. Given these 

two factors, it would be easy to propose a teeter-totter 

hypothesis of expansion or contraction of shrub-grassland interfaces with only minor changes in the seasonal 

distribution of precipitation. Soil texture and rainfall 

event size distributions, as they influence the depth of 

percolation of soil moisture, could create the patchy, 

unpredictable spatial pattern of woody expansion that 

has been reported. Past very heavy grazing and current grazing of highly erodible soils could also be a 

factor in the spatial distribution of community conversion. A complication with this hypothesis is that shrub 

expansion also occurs in many parts of Texas (Archer 

1994), where the climate is relatively more continental 

that it is in the far west. No one factor alone appears 

to adequately explain the complex spatial pattern of 

community change throughout this region, but progress 

has been made in correlating local changes to climate. 

Regardless of changes in community types, seasonal 

and annual fluctuations in weather often have a greater 

influence on plant community species composition than 

do different grazing regimes. Based on long-term studies, annual fluctuations in weather have much larger 

effects on plant community species composition and 

production than do extreme differences in grazing pressures in shortgrass steppe of the Great Plains (Klipple 

and Costello 1960; Milchunas and others 1994). It is the 

author’s perception that the temporal variability in plant 

community species composition due to weather is much 

greater in southwestern than in Great Plains communities, based on long-term data sets such as those reported 

above from sites such as the Jornada and Santa Rita 

compared to long-term sites in the Great Plains such as 

the Central Plains Experimental Range. This may be 

due to the large influence in this area of El Niño and La 

Niña conditions and to greater orographic effects producing annual to decadal oscillations in amounts and 

seasonality of precipitation (Neilson 1986; Swetnam 

and Betancourt 1998), but the high temporal variability points to a particularly important need for long-term 

studies in this region for understanding the dynamics of 

vegetation within and among communities.

The implications of climate change/community conversion in assessing the effects of livestock grazing 

on plant communities are twofold. First, a different 

end-point for “recovery” with protection from grazing 

would need to be recognized; in other words, the previous reference point is no longer valid for evaluating the 

subject disturbance. Exclosures, and preferably large 

areas of ungrazed controls, are necessary to evaluate 

grazing effects. Grazing is simply overlain on changing 

communities. While the woody/grassland conversion 

has captured much attention over the past decade, other 

understory changes are also ongoing, as described in the 

review above and emphasized by Allred (1996). Some 

of these are grazing related and independent of community conversion, and others are overstory controlled. 

Second, if grazing is not related to change to woody 

communities, this does not at all preclude grazingrelated alternate stable states. Assessing whether this 

occurs is difficult when the reference state is changing. 

Do the differences between old exclosures and new 

exclosures converge? The uniquely large difference 

observed for natural geologic refuges never grazed by 

large herbivores and more recent human-made exclosures shown in Fig. 11 could be due to different abiotic 

conditions associated with the refuges or to divergent 

trajectories associated with grazing in the distant past 

(Milchunas and Noy-Meir 2004). Does current grazing 

result in the shrub state as predicted by the Schlesinger 

and others model even when winter precipitation is experimentally manipulated to disfavor shrubs, or when 

rain-event size-class distributions change? These types 

of questions would need to be addressed to answer 

questions concerning grazing-related alternate states if 
USDA Forest Service RMRS-GTR-169. 2006. 99

other factors are producing a changing reference state. 

This would appear difficult given the fact that there are 

few or no controlled studies of the effects of grazing 

for many of the broad community types covered above. 

Often the goal simply becomes managing to best maintain a resemblance to previous conditions. The even 

greater changes that may come with predicted future 

climate change will possibly push management decisions beyond even that baseline. Waiting for protected 

areas to return to some initial condition to demonstrate 

recovery is impractical/will never happen if the reference point is changing. We need to develop alternate 

bases for management guidelines.

Intense non-selective grazing is an exogenous disturbance in most southwestern communities. Changes 

in plant species composition between grazed and ungrazed communities are greater than in Great Plains 

grasslands, even when grazing at “moderate” levels is 

at a lower level of consumption than that considered 

“moderate” in the Great Plains. Grazing in the Great 

Plains is an endogenous disturbance to the system, both 

in terms of the plant community species composition 

(Milchunas and others 1989) and resistance to exotic 

invasion (Milchunas and others 1992) and in terms of 

endemic shortgrass fauna (Milchunas and others 1998). 

While grazing by cows may not be exactly like bison, 

they are in many ways similar in function and act as 

surrogates. From a conservation standpoint then, it 

would simply appear that there are places where grazing by livestock should be encouraged in the absence 

of the native grazer, and other places where livestock 

grazing should be discouraged or at least cautiously 

managed because similar grazers were not present historically. Things are never that simple, even if viewed 

from a purely biological perspective. The subdivision 

of large ranches into small acreage ranchettes throughout the western United States may be one of the largest 

threats to biodiversity conservation due to the land-use 

change it represents as well as effects through fragmentation (Curtin and others 2002). While information is 

still too scarce to fully evaluate the full extent of the 

tradeoffs, the elimination of fire, introduction of house 

pets, livestock pets, and exotic landscaping are just a few 

to consider. The fate of private lands will often closely 

be coupled to use of public lands, since many ranchers economically depend on grazing of public lands 

for their existence (Raish and McSweeney 2003). The 

review above clearly shows weaknesses in even basic 

understanding of the response of many major plant community types to grazing. For other community types, 

there is often contradictory results, suggesting there are 

factors responsible for differential sensitivities that we 

do not understand. However, possibly more critical at 

this time is the need for a consideration and an understanding of the alternatives if livestock are removed.
100 USDA Forest Service RMRS-GTR-169. 2006.

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USDA Forest Service RMRS-GTR-169. 2006. 117

1. Alpine tundra—synonyms: Rocky Mountain alpine tundra, krummholz.

2. Boreal forest—synonyms: Engelmann spruce [Picea engelmannii] province, spruce-fir forest, taiga, subalpine 

forest, bristlecone pine woodland, Petran subalpine coniferous forest, aspen.

3. Mixed conifer forest (C)—synonyms: upper montane forests, white fir/douglas-fir, pine/douglas-fir (Küchler 

1964), Petran conifer forest, aspen.

4. Mixed conifer forest (M)—synonyms: Madrean montane forest.

5. Ponderosa pine forest—synonyms: ponderosa pine [Pinus ponderosa] province, transition zone (Merriam), 

lower montane forest.

6. Madrean pine forest—synonyms: Chihuahuan pine [Pinus leiophylla] Series, Engelmann spruce [P. engelmannii] Series (Layser and Schubert 1979), Arizona pine forest (Küchler 1964), Pine-oak woodland (Marshall 

1957), Madrean montane conifer forests, aspen (rare).

7. Pinyon-juniper woodland—synonyms: Pygmy woodland, coniferous woodland, juniper-pinyon woodland, 

Upper Sonoran Zone (Merriam), Mexican pine [Pinus cembroides] province (Daubenmire 1978), Juniper 

savanna (Dick-Peddie 1993).

8. Madrean woodland—synonyms: encinal, open encinal, oak-juniper woodland, pine-oak woodland, cypress 

woodland on mountain slopes.

1. Gray Oak (Quercus grisea) woodland.

2. Arizona white oak (Quercus arizonica) woodland.

3. Emory oak (Quercus emoryi) woodland and savanna.

9. Great Basin juniper-sagebrush savanna—synonyms: Juniper steppe woodland, juniper savanna, Great Basin 

conifer woodland - juniper series.

10. Interior chaparral—synonyms: chaparral, evergreen chaparral, turbinella oak chaparral, turbinella oak scrub, 

oakbrush, Great Basin montane scrub.

11. Subalpine grassland—synonyms: Rocky Mountain montane grassland, forest steppe, high elevation parks, 

Thurber fescue meadows, subalpine meadows, mountain bunchgrass - Thurber fescue.

12. Montane meadow—synonyms: montane parks, forest steppe, mountain meadow, meadow, grasslands associated with ponderosa pine.

13. Temperate grassland—synonyms: shortgrass prairie, shortgrass steppe (Lauenroth and Milchunas 1991), 

grama-buffalo grass (Küchler 1964), mixed prairie, grama-galleta steppe (Küchler 1964), Galleta-three awn 

shrubsteppe (Küchler 1964), foothills grassland, plains-mesa grassland, tallgrass prairie (rare in the SW), 

mid-grass prairie, cold temperate grassland.

14. Mesquite savanna—synonyms: desert grassland, semi-desert grassland, mesquite-acacia savanna, gramatobosa shrubsteppe (Küchler 1964), Upper Sonoran (Merriam), Trans-Pecos shrub savanna (Küchler 1964), 

warm temperate grassland.

15. Great Basin sagebrush shrubsteppe—synonyms: Great Basin sagebrush, sagebrush-steppe (Küchler 1964), 

Western intermountain sagebrush steppe, Great Basin desert grassland.

16. Shinnery oak/sand sage—synonyms: Plains-Mesa sand scrub, shinnery scrub, shin-oak scrub.

17. Great Basin temperate desert—synonyms: cold temperate desert, Colorado Plateau semidesert, Great Basin 

desert scrub.

Appendix 1. Synonyms for the plant 

community classification system of Moir 

(2000).
118 USDA Forest Service RMRS-GTR-169. 2006.

18. Sonoran/Mohavian deserts—synonyms: Lower Sonoran Zone (Merriam), hot desert, creosote bush [Larrea 

divaricata] province of Daubenmire 1968.

19. Chihuahuan desert—synonyms: Lower Sonoran Zone (Merriam), hot desert, creosote bush [Larrea divaricata] province of Daubenmire 1968.

20. Narrowleaf cottonwood forest—synonyms: montane Cottonwood-Willow.

21. Broadleaf cottonwood forest (C)—synonyms: cold temperate cottonwood forest, flood plain-plains riparian 

forest, lowland plains broadleaf deciduous forest, Great Basin broadleaf deciduous forest, gallery forest, cottonwood-willow forest.

22. Broadleaf cottonwood forest (M)—synonyms: warm-temperate riparian forest, lowland interior southwestern broadleaf deciduous forest, Sycamore forest, Cottonwood forest, bosques, mesquite bosques, saltcedar, 

tamarisk, floodplains-plains riparian forest, Arizona walnut woodland.

23. Riparian scrub—synonyms: cold-temperate and Arctic-boreal riparian scrub, montane scrub, willow-alder 

shrub riparian, coyote willow shrub riparian.

24. Dry washes—synonyms: arroyo riparian, riparian scrub, desert riparian washes.

25. Wetlands—synonyms: wet meadows, herbaceous wet meadows, emergent wetlands, marshes, cienegas, 

strands, seeps, bogs (rare in the Southwest), playas.
USDA Forest Service RMRS-GTR-169. 2006. 119

Acacia Acacia spp.

Agave Agave spp.

Agoseris Agoseris spp.

Alderleaf mountain mahogany Cercocarpus montanus Raf.

Alkali sacaton Sporobolus airoides (Torr.) Torr.

Alligator juniper Juniperus deppeana Steud.

American bulrush Scirpus americanus Pers.

American tarwort Flourensia cernua DC.

Annual dropseed Muhlenbergia minutissima (Steud.) Swallen

Antelope bitterbrush Purshia tridentata (Pursh) DC.

Apache pine Pinus engelmannii Carr.

Apache plume Fallugia paradoxa (Don) Endl. ex Torr

Aparejograss Muhlenbergia utilis (Torr.) A.S. Hitchc.

Arizona alder Alnus oblongifolia Torr.

Arizona cottongrass Trichachne california (Benth.) Chase

Arizona cottontop Digitaria californica (Benth.) Henr.

Arizona cypress Cupressus arizonica Greene

Arizona fescue Festuca arizonica Vasey

Arizona pine Pinus arizonica Engelm.

Arizona ponderosa Pinus ponderosa var. arizonica (Engelm.) Shaw

Arizona sycamore Platanus wrightii Wats.

Arizona threeawn Aristida purpurea Nutt.

Arizona walnut Juglans major (Torr.) Heller

Arizona white oak Quercus arizonica Sarg.

Arizona willow Salix arizonica Dorn.

Arkansas rose Rosa arkansana Poter

Arroyo willow Salix lasiolepis Benth.

Aspen Populus tremuloides Michx.

Banana yucca Yucca baccata Torr.

Bearberry Arctistaphylos uva-ursi (L.) Spreng.

Beautiful fleabane Erigeron formosissums Greene

Bebb willow Salix bebbiana Sarg.

Berlandier’s wolfberry Lycium berlandieri Dunal

Big sagebrush Artemisia tridentata Nutt.

Bigelow sage Artemisia bigelovii Gray

Bigtooth maple Acer grandidentatum Nutt.

Bindweed spp. Calystegia spp.

Black dropseed Sporobolus interruptus Vasey

Black grama Bouteloua eriopoda Torr.

Black sagebrush Artemisia nova A. Nels.

Blackbrush Coleogyne ramosissima Torr.

Black willow Salix nigra Marsh.

Blue grama Bouteloua gracilis (H.B.K.) Lag.

Blue paloverde Cercidium floridum Benth. ex Gray

Appendix 2. List of scientific and common 

names of plants used in text.
120 USDA Forest Service RMRS-GTR-169. 2006.

Blue spruce Picea pungens Engelm.

Blueburr stickseed Lappula redowskii (Hornem) Greene

Border pinyon Pinus discolor Bailey & Hawksworth

Bottlebrush squirreltail Sitanion hystrix (Nutt.) J.G. Smith

Boxelder Acer negundo (L.)

Bristlecone pine Pinus aristata Engelm.

Brittle pricklypear Opuntia fragilis (Nutt.) Haw.

Broadleaf cattail Typha latifolia L.

Bromegrass Bromus spp.

Broom snakeweed Gutierrezia sarothrae (Pursh) Britt. & Rusby.

Buffalograss Buchloe dactyloides (Nutt.) Engelm.

Bull muhly Muhlenbergia emersleyi Vasey

Burrobush Ambrosia dumosa, (Gray) Payne

Franseria dumosa

Burrograss Scleropogon brevifolius Phil.

Burroweed Isocoma (Haplopappus) tenuisecta Greene

Bush muhly Muhlenbergia porteri Scribn.

Cactus apple Opuntia engelmannii Salm-Dyck

California pine Pinus californiarum D.K. Bailey

Cane beardgrass Amphilophis barbinodis (Lag.) Nash

Cane bluestem Bothriochloa barbinodis (Lag.) Herter

Candy barrel cactus Ferocactus wislizeni (Engelm.) Britt. & Rose

Canyon grape Vitis arizonica Engelm.

Carruther’s sagewort Artemisia carruthii Wood ex Carruth.

Catclaw acacia Acacia greggii Gray

Catclaw mimosa Mimosa biuncifera Benth.

Mimosa aculeaticarpa Ortega var. biuncifera 

(Benth.) Barneby

Ceanothus Ceanothus spp.

Cheatgrass Bromus tectorum L.

Chihuahuan pine Pinus leiophylla Schiede & Deppe

Chokecherry Prunus virginiana L.

Cholla spp. Opuntia spp.

Columbia needlegrass Achnatherum nelsonii (Scribn.) Barkworth

Common dandelion Taraxacum officinale Weber ex Wiggers

Common sotol Dasylirion wheeleri S. Wats.

Common wolfstail Lycurus phleoides Kunth

Condalia spp. Condalia spp.

Cooley’s bundleflower Desmanthus cooleyi (Eat.) Trel.

Creeping muhly Muhlenbergia repens (Presl) Hitchc.

Creosote-bush Larrea tridentata (DC.) Coville

Larrea divaricata auct. Non Cav.

Crested wheatgrass Agropyron desertorumq (Fisch.) Schult.

Crucifixion thorn Canotia holacantha Torr.

Curly mesquite Hilaria belangeri (Steud.) Nash

Deergrass Muhlenbergia rigens (Benth.) Hitchc.

Deer-vetch Lotus rigidus (Benth.) Greene

Dense ayenia Ayenia microphylla Gray

Desert ceanothus Ceanothus greggii Gray

Desert false indigo Amorpha fruticosa L.

Desert false-yarrow Chaenactis stevioides Hook. & Arn.
USDA Forest Service RMRS-GTR-169. 2006. 121

Desert globemallow Sphaeralcea ambigua Gray

Desert hackberry Celtis pallida Torr.

Desert ironwood Olneya tesota Gray

Desert tansy mustard Descurainia obtusa (Greene) O. E. Schulz 

Desert willow Chilopsis linearis (Cav.) Sweet

Desert zinnia Zinnia pumila Gray

Desert-thorn Lycium spp. L.

Devil cholla Opuntia stanyli var. stanyli Engelm.

Douglas-fir Pseudotsuga menziesii (Mirbel) Franco

Downy chess Bromus tectorum L.

Drummond’s willow Salix subcaerulea Piper

Salix drummondiana Barratt ex Hook.

Eastern Mojave buckwheat Eriogonum fasciculatum Benth.

Emory oak Quercus emoryi Torr.

Emory’s baccharis Baccharis emoryi Gray

Engelmann pricklypear Opuntia phaeacantha Engelm.

Engelmann spruce Picea engelmannii Perry ex Engelm.

Evergreen sumac Rhus choriophylla Wood & Standl.

Rhus virens Lindheimer ex Gray var 

cariophylla (Wood & 

Standl.) L. Benson

Fairyduster Calliandra eriophylla Benth.

False buffalo grass Munroa squarrosa (Nutt.) Torr.

False mesquite Calliandra eriophylla Benth.

Fendler sandwort Arenaria fendleri Gray

Fendler threeawn Aristida fendleriana Steud.

Fendler’s ceanothus Ceanothus fendleri Gray

Flatsedge Cyperus spp.

Fluffgrass Tridens pulchellus (H.B.K.) Hitchc.

Fluffgrass Erioneuron pulchella H.B.K.

Fourwing saltbrush Atriplex canescens (Pursh) Nutt.

Fremont cottonwood Populus fremontii Wats.

Fremont geranium Geranium caespitosum var. fremontii (Torr. ex Gray) Dorn

Fringed sagebrush Artemisia frigida Willd.

Galleta Hilaria jamesii (Torr.) Benth.

Gambel oak Quercus gambelii Nutt.

Gambel x Scrub live oak Quercus x pauciloba Rydb.

Garrya Garrya spp.

Goldenhills Encelia farinosa Gray ex Torr

Goodding’s willow Salix gooddingii Ball

Goosefoot Chenopodium spp.

Graham’s nipple cactus Mammillaria microcarpa Engelm.

Grama Bouteloua spp.

Gray oak Quercus grisea Liebm.

Green ash Fraxinus pennsylvanica Marsh.

Green joint-fir Ephedra viridis (Cutler) L. Benson

Green Sprangletop Leptochloa dubia (Kunth) Nees

Groundsel Packera spp.

Guajilla Calliandra humilis var. humilis Benth.

Hairy false goldenaster Heterotheca villosa (Pursh) Shinners

Hairy grama Bouteloua hirsuta Lag.
122 USDA Forest Service RMRS-GTR-169. 2006.

Hairy mountain-mahogany Cercocarpus montanus var. (Wats.) Martin

paucidentatus

Havard oak Quercus havardii Rydb.

Hawk’s eye Euryops multifidus or (Thunb.) DC.

Euryops subcarnosus DC. ssp. vulgaris B. Nord 

Hedgehog cactus Echinocereus spp.

Hoary tansyaster Machaeranthera pulverulehta (Nutt.) Greene

Honey mesquite Prosopis glandulosa Torr.

Honey mesquite Prosopis juliflora var. glandulosa (Torr.) Cockerell

Horse cinquefoil Potentilla hippiana Lehm.

Hymenoxys Hymenoxys spp.

Idaho fescue Festuca idahoensis Elmer

Indian ricegrass Achnatherum (Oryzopsis) hymenoides (Roemer & JA Schultes) 

Barkworth

James’ galleta Hilaria jamesii (Torr.) Benth.

Joint-fir Ephedra spp.

Joshua tree Yucca brevifolia Engelm.

Jumping cholla Opuntia fulgida Engelm.

Junegrass Koeleria cristata (L.) Pers.

Juniper Juniperus spp.

Kentucky bluegrass Poa pratensis L.

Lacy tansyaster Haplopappus spinulosus (Pursh) DC.

Leatherweed Croton pottsii (Klotzsch) Muell.-Arg.

Lehmann lovegrass Eragrostis lehmanniana Nees

Letterman needlegrass Stipa lettermanii Vasey

Limber pine Pinus flexilis James

Little bluestem Schzachyrium scoparium (Michx.) Nash

Littleleaf ratany Krameria parvifolia Benth.

Littleleaf sumac Rhus microphylla Engelm. ex Gray

Longleaf jointfir Ephedra trifurca Torr. ex S. Wats

Longtongue muhly Muhlenbergia longiligula Hitchc.

Longtongue muttongrass Poa longiligula Scribn. & Williams

Low woollygrass Dasyochloa pulchella (Kunth) Willd.

Lupines Lupinus spp.

Mariola Parthenium incanum Kunth

Mat muhly Muhlenbergia arsenei Hitchc.

Menodora Menodora spp.

Mesa dropseed Sprobolus flexuosus (Thurb. ex Vassey) Rydb.

Mesa threeawn Aristida hamulosa Henr.

Mesquite Prosopis spp.

Mesquitilla Calliandra eriophylla Benth.

Mexican blue oak Quercus oblongifolia Torr.

Mexican pinyon Pinus cembroides Zucc.

Mimosa Mimosa spp.

Mountain muhly Muhlenbergia montana (Nutt.) Hitchc.

Mountain snowberry Symphoricarops oreophilus Gray

Mountain-mahogany Cercocarpus spp. Kunth

Mule’s fat Baccharis salicifolia (Ruiz & Pavon) Pers.

Mutton bluegrass Poa fendleriana (Steud.) Vasey
USDA Forest Service RMRS-GTR-169. 2006. 123

Muttongrass Poa fendleriana (Steud.) Vasey

Narrowleaf cottonwood Populus angustifolia James

Narrowleaf willow Salix exigua Nutt.

Needle and thread grass Stipa comata Trin. & Rupr.

Needlegrasses Stipa spp.

Netleaf oak Quercus rugosa Nee.

New Mexico feathergrass Hesperostipa neomexicana (Thurb. ex Coult.) 

Barkworth

Nickleaf milkvetch Astragalus goniatus Nutt.

Nodding brome Bromus anomalus Rupr.ex Fourn.

Nodding wild buckwheat Erigonum cernuum Nutt.

Ocotillo Fouquieria splendens Engelm.

Oneseed juniper Juniperus monosperma (Engelm.) Sarg.

Orange sneezeweed Helenium hoopesii Gray

Orcutt’s threeawn Aristida orcuttiana (Vasey) Allred & 

Valdes-Reyna

Owl’s-claws Dugaldia hoopesii (Gray) Bierner

Pacific willow Salix lasiandra Benth.

Pale evening-primrose Oenothera pallida (Nutt.) Munz & Klein

Pale spikerush Eleocharis macrostachya Britt.

Paperflower spp Psilostrophe spp.

Park willow Salix monticola Bebb

Parry’s agave Agave parryi Engelm.

Parry’s oatgrass Danthonia parryi Scribn.

Peachleaf willow Salix amygdaloides Anderss.

Pennsylvania cinquefoil Potentilla pennsylvanica L.

Pine dropseed Blepharoneuron trichcolepis (Torr.) Nash

Pingue Actinea richardsonii (Hooker) Kuntze

Pingue rubberweed Hymenoxys richardsonii (Hook.) Cockerell

Pinyon ricegrass Piptochaetium fimbriatum (H.B.K.) Hitchc.

Plains beardgrass Andropogon hirtiflorus var. feensis (Fourn.) Hack.

Plains hiddenflower Cryptantha crassisepala (Torr. & Gray) Greene

Plains lovegrass Eragrostis intermedia Hitchc.

Plains prickly pear Opuntia polyacantha (Haw.)

Plantain Plantago spp.

Pointleaf manzanita Arctostaphylos pungens Kunth

Ponderosa pine Pinus ponderosa Laws.

Poverty threeawns Aristida divaricata Humb. & Bonpl.

Prairie clover spp. Dalea spp. L.

Prairie coneflower Ratibida columnaris (Sims.)

Prairie junegrass Koeleria macrantha (Ledeb.) Schultes

Praire pepperweed Lepidium densiflorum Schrad.

Prickly gilia Leptodactylon pungens (Torr.) Torr. ex Nutt.

Pricklypear Opuntia spp.

Pringle manzanita Arctostaphylos pringlei Parry

Purple picklypear Opuntia violacea Engelm.

Purple three-awn Aristida purpurea Nutt.

Pussytoes Antennaria spp.
124 USDA Forest Service RMRS-GTR-169. 2006.

Rabbitbrush Chrysothamnuns spp.

Rayless goldenhead Acamptopappus sphaerocephalus (Harvey & Gray ex Gray) 

Gray

Rayless tansyaster Machaeranthera grindelioides (Nutt.) Shinners

Red barberry Berberis haematocarpa Woot.

Red brome Bromus rubens L.

Red grama Bouteloua trifida Thurb.

Redstem stork’s bill Erodium cicutarium (L.) L’Hér. ex Ait. 

Red threeawn Aristida longiseta Steud.

Redberry buckthorn Rhamnus crocea Nutt.

Redberry juniper Juniperus erythrocarpa Cory

Redosier dogwood Cornus stolonifera Michx.

Redtop Agrostis alba L.

Reverchon three-awn Aristida purpurea var. nealleyi Nutt.

Ring muhly, ringgrass Muhlenbergia torreyi (Kunth.) Hitchc. ex Bush

Rio Grande saddlebush Mortonia scabrella Gray

Rocky Mountain iris Iris missouriensis Nutt.

Rocky Mountain juniper Juniperus scopulorum Sarg.

Rocky Mountain maple Acer glabrum Torr.

Rose pussytoes Antennaria rosea Greene

Rosy gilia Gilia sinuata Dougl.

Rothrock grama Bouteloua rothrocki Vasey

Rough menodora Menodora scabra Gray

Rubber rabbitbrush Ericameria nauseosa (Pallas es Pursh) Nesom 

& Baird

Rush Juncus spp.

Russian olive Elaeagnus angustifolia L.

Russian thistle Salsola iberica Semmem & Pau

Sacahuista Nolina microcarpa Wats.

Sagebrush Artemisia spp.

Saguaro Carnegia gigantea (Engelm.) Britt. & Rose

Saltbush Atriplex spp.

Salt cedar Tamarix chinensis Lour.

Sand bluestem Andropogon hallii Hack.

Sand dropseed Sporobolus cryptandrus (Torr.) Gray

Sandhill muhly Muhlenbergia pungens Thrub.

Sand sagebrush Artemisia filifolia Torr.

Sand wort Arenaria spp.

Sangre de Cristo Jatropha caraiophylla (Torr.) Muell.-Arg.

Santa Rita threeawn Aristida californica var. glabrata Vasey

Scarlet globemallow Sphaeralcea coccinea (Pursh)

Schott’s yucca Yucca schottii Engelm.

Screwleaf muhly Muhlenbergia virescens (H.B.K.) Kunth

Scrub oak Quercus turbinella Greene

Sedge Carex spp.

Sego lily Calochortus nuttallii Torr. & Gray

Shadscale saltbrush Atriplex confertifolia (Torr. & Frem.) Wats.

Sheep fescue Festuca ovina L.

Shrubby buckwheat Eriogonum fasciculatum Benth.

Shrubby false mallow Malvastrum bucuspidatum (S. Wats.) Rose

Sideoats grama Bouteloua curtipendula (Michx.) Torr.
USDA Forest Service RMRS-GTR-169. 2006. 125

Silver brush spp. Ditaris spp.

Silver dwarf morning-glory Evolvulus sericeus Sw.

Silverleaf oak Quercus hypoleucoides Camus

Silvery globe mallow Sphaeralcea leptophylla (Gray) Rydb.

Singlewhorl burrobrush Hymenoclea monogyra Torr. & Gray ex Gray

Six-weeks fescue Festuca octoflora (Walt.) Rydb.

Skunkbush sumac Rhus trilobata Nutt.

Slender janunsia Janusia gracilis Gray

Slender gilia Gilia leptomeria Gray

Slender grama Bouteloua filiformis (Fourn.) Griffiths

Slender wheatgrass Agropyron trachycaulum (Link) Malte

Slenderbush eriogonum Eriogonum microthecum Hook.

Slim tridens Tridens muticus (Torr.) Nash

Small wirelettuce Stephanomeria exigua Nutt.

Snakeweed Gutierrezia spp.

Soapberry Sapindus saponaria L.

Soaptree yucca Yucca elata (Engelm.) Engelm.

Soapweed yucca Yucca glauca Nutt.

Somoran scrub oak Quercus turbinella Greene

Southwestern white pine Pinus strobiformis Engelm.

Spider grass Aristida ternipes Cav.

Spike dropseed Sporobolus cryptandrus A. S. Hitchc.

Spike muhly Muhlenbergia wrightii Vasey

Spiny hackberry Celtis pallida Torr.

Splitleaf brickellbush Brickellia laciniata Gray

Sprucetop grama Bouteloua chondrosioides (Kunth) Benth. ex Wats.

Squirreltail Elymus elymoides (Sitanion hystrix) (Raf.) Swezey

Sticky snakeweed Gutierrezia Texana var. glutinosa (S. Schauer) M.A. Lane

Subalpine fir Abies lasiocarpa (Hook.) Nutt.

Sugar sumac Rhus ovata Wats.

Sun sedge Carex inops heliophila (Mackenzie) Crins

Sunflower Helianthus sp.

Sweetbush Bebbia juncea (Benth.) Greene 

Sweet resin bush Euryops multifidus or (Thunb.) DC.

Euryops subcarnosus DC. ssp. vulgaris B. Nord 

Tamarick Tamarix spp.

Tanglehead Heteropogon conlorius (L.) Richt.

Tarbush Flourensia spp.

Teddybear cholla Opuntia bigelovii Engelm.

Texas beardgrass Andropogon cirratus Hack.

Thimbleberry Rubus parviflorus Nutt.

Thin paspalum Paspalum stramineum Michx.

Thinleaf alder Alnus tenuifolia (Nutt.) Breitung

Threadleaf ragwort Senecio longilobus Benth.

Threeawn Aristida spp.

Thurber’s fescue Festuca thurberi Vasey

Timothy Pheum spp.

Tobosa grass Hilaria mutica (Buckl.) Benth.

Torrey joint-fir Ephedra torreyana S. Wats.

Toumey oak Quercus toumeyi Sarg.

Trailing fleabane Erigeron flagellaris Gray
126 USDA Forest Service RMRS-GTR-169. 2006.

Triangle burr ragweed Ambrosia deltoidea (Torr.) Payne

Trisetum Trisetum spp.

Tulip pricklypear Opuntia phaeacantha Engelm.

Tumblegrass Schedonnardus paniculatus (Nutt.) Trel.

Turbinella oak Quercus turbinella Greene

Turpentine bush Ericameria laricifolia (Gray) Shinners

Twoneedle pinyon Pinus edulis Engelm.

Utah juniper Juniperus osteosperma (Torr.) Little

Velvet mesquite Prosopis velutina Woot.

Velvetpod mimosa Mimosa dysocarpa Benth.

Vine mesquite Panicum obtusum Kunth

Walkingstick cactus Opuntia spinosior (Engelm.) Toumey

Wavyleaf oak Quercus x pauciloba Rydb. (pro sp.) gambelii ×

turbinella

Western thimbleberry Rubus parviflorus Nutt.

Western wheatgrass Agropyron smithii Rydb.

Western wheatgrass Pascopyrum smithii (Rydb.) A. Love

Western yarrow Achillea millefolium occidentalis DC.

Whiplash pappusgrass Pappophorum vaginatum Buckl.

White-ball acacia Acacia angustissima Torr. & Gray

White bursage Ambrosia dumosa (Gray) Payne

White clover Melilotus alba Desr.

White fir Abies concolor (Gord. & Glend.) Lindl. 

ex Hildebr.

White ratany Krameria grayi Rose & Painter

Whitethorn acacia Acacia constricta Benth.

Winterfat Krascheninnikovia lanata (Pursh) Meeuse & Smit

Wolftail Lycurus phleoides H.B.K.

Woody crinklemat Tiquilia canescens (D.C.) A. Richards.

Woolly cinquefoil Potentilla hippiana Lehm.

Wooly bunchgrass Elyonurus barbiculmis Hack.

Wooly Indian-wheat Plantago purshii R.& S.

Wright buckweed Eriogonum wrightii Torr.

Wright’s thimblehead Hymenothrix wrightii Gray

Wright’s silktassel Garrya wrightii Torr.

Yellow paloverde Cercidium microphyllum (Torr.) Rose & 

I.M. Johnston

Yerba de pasmo Baccharis pteronioides D.C.
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