United States Department of Agriculture / Forest Service

Rocky Mountain Research Station

General Technical Report RMRS-GTR-285

August 2012

Climate Change in Grasslands, Shrublands, 

and Deserts of the Interior American West:

A Review and Needs Assessment
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Finch, Deborah M., ed. 2012. Climate change in grasslands, shrublands, and deserts 

of the interior American West: a review and needs assessment. Gen. Tech. Rep. 

RMRS-GTR-285. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, 

Rocky Mountain Research Station. 139 p.

Abstract

Recent research and species distribution modeling predict large changes in the 

distributions of species and vegetation types in the western interior of the United 

States in response to climate change. This volume reviews existing climate models that 

predict species and vegetation changes in the western United States, and it synthesizes 

knowledge about climate change impacts on the native fauna and flora of grasslands, 

shrublands and deserts of the interior American West. Species’ responses will depend 

not only on their physiological tolerances but also on their phenology, establishment 

properties, biotic interactions, and capacity to evolve and migrate. The volume is 

divided into eight chapters that cover the topics of carbon mitigation and adaptation. 

Current and likely responses of species and habitats to climate change are examined 

in relation to taxonomic group and ecoregion and with regard to other disturbances. 

The volume ends with a review of management decision support needs and tools for 

assessing vulnerability of natural resources and conserving and restoring ecosystems 

that are or may be impacted by climate change. 

Keywords: climate change, grasslands, shrublands, deserts, assessment

Acknowledgments

We are grateful to Julie McIntyre and Karen Wetherill for helpful discussions on 

pollinators. We thank David Hawksworth, Melinda Larson, Jennifer Jacobs, and 

Megan Friggens for their help in reviewing and formatting content. We are grateful 

to Karen Bagne, Megan Friggens, and California Game and Fish Department for 

reviews of the manuscript, and Kal Louks and MacKenzie Romero for facilitating 

financial arrangements. Funding was provided by USDA Forest Service Research and 

Development, Washington Office, and Rocky Mountain Research Station.

Cover photos: Upper left and lower photos, © Larry Jones. Upper right and lower right, USDA Forest Service.
Climate Change in Grasslands, Shrublands, 

and Deserts of the Interior American West:

A Review and Needs Assessment

Deborah M. Finch, Editor

United States Department of Agriculture

Forest Service

Rocky Mountain Research Station

General Technical Report RMRS-GTR-285

August 2012

Contents

Chapter 1 Modeling and Predicting Vegetation Response of Western USA 

Grasslands, Shrublands, and Deserts to Climate Change................... 1

Megan M. Friggens, Marcus V. Warwell, Jeanne C. Chambers, and 

Stanley G. Kitchen

Chapter 2 Restoring and Managing Cold Desert Shrublands for 

Climate Change Mitigation.............................................................. 21

Susan E. Meyer

Chapter 3 Climate Change and Arthropods: Pollinators, Herbivores, 

and Others...................................................................................... 35

Sandra L. Brantley and Paulette L. Ford

Chapter 4 Plant Vulnerabilities and Genetic Adaptation................................. 48

Bryce A. Richardson, Nancy L. Shaw, and Rosemary L. Pendleton

Chapter 5 Climate Change, Animal Species, and Habitats: Adaptation 

and Issues........................................................................................ 60

Deborah M. Finch1, D. Max Smith, Olivia LeDee , Jean-Luc E. Cartron, 

and Mark A. Rumble

Chapter 6 Disturbance and Climate Change in the Interior West................... 80

Paulette L. Ford, Jeanne K. Chambers, Sharon J. Coe, and Burton C. Pendleton

Chapter 7 Invasive Species and Climate Change............................................. 97

Justin B. Runyon, Jack L. Butler, Megan M. Friggens, Susan E. Meyer, and 

Sharlene E. Sing

Chapter 8 Decision Support: Vulnerability, Conservation, 

and Restoration............................................................................. 116

Megan M. Friggens, Jeremiah R. Pinto, R. Kasten Dumroese, and 

Nancy L. Shaw

USDA Forest Service RMRS-GTR-285. 2012. 1

Chapter 1

Modeling and Predicting Vegetation Response of Western USA 

Grasslands, Shrublands, and Deserts to Climate Change

Megan M. Friggens1, Marcus V. Warwell2, Jeanne C. Chambers3, and Stanley G. Kitchen4

1 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Sciences Laboratory, Albuquerque, New Mexico

2 U.S. Forest Service, Rocky Mountain Research Station; Forest and Woodland Ecosystems Program; Forestry Sciences 

Laboratory, Moscow, Idaho

3 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; Great 

Basin Ecology Laboratory, Reno, Nevada

4 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; Shrub 

Sciences Laboratory, Provo, Utah

Executive Summary

Experimental research and species distribution modeling predict large changes in 

the distributions of species and vegetation types in the Interior West due to climate 

change. Species’ responses will depend not only on their physiological tolerances but 

also on their phenology, establishment properties, biotic interactions, and capacity 

to evolve and migrate. Because individual species respond to climate variation and 

change independently and differently, plant assemblages with no modern analogs can 

be expected in areas where novel climate conditions develop. The capacity to accurately predict how species distributions and plant assemblages will change under future 

warming is essential for developing effective strategies for maintaining and restoring 

sustainable ecosystems.

The rate of predicted change in climate is unprecedented relative to the three centuries prior to industrialization. A conservative forecast for western North America 

(using 21 global climate models for the A1B scenario) is a linear change in mean 

temperatures and precipitation ranging from +2.1 °C to +5.7 °C and -3% to +14%, 

respectively, over the rest of the century. A number of recent studies that have used 

bioclimatic and related models predict drastic changes in vegetation types and native 

and invasive flora at regional scales in response to the change in climate. By the end of 

the century, 55% of future landscapes in the West likely will have climates that are incompatible with the vegetation types that now occur on those landscapes. Specifically:

• Suitable habitat for Rocky Mountain subalpine conifer forests and Rocky Mountain 

and Great Basin alpine tundra declines substantially (>97%) or disappears.

• Projected habitat suitable for Great Basin pinyon and juniper woodlands moves 

northward and upslope.

• Semi-desert grassland habitat expands northward and occupies an area nearly four 

times that of the present.

• Habitat suitable for Great Basin shrub/grassland decreases by 40% and becomes 

fragmented.
2 USDA Forest Service RMRS-GTR-285. 2012.

• Great Basin montane scrub habitat experiences moderate decline (69% decrease) 

and displacement through time.

• Climate changes appear to be most favorable for Mohave Desert, Sonoran Desert, 

and Chihuahuan Desert scrub vegetation types, which are all projected to expand.

Species’ specific analyses predict:

• Sagebrush (Artemisia tridentata), Joshua tree (Yucca brevifolia), saguaro (Carnegiea 

gigantea), and creosote bush (Larrea tridentata) shift northwards.

• Species with small distributions, such as smooth Arizona cypress (Cupressus arizonica ssp. glabra) and the endangered perennial MacFarlane’s four-o’clock (Mirabilis 

macfarlanei), experience complete climate disequilibrium early in the century.

• Invasive species, such as buffelgrass (Pennisetum ciliare), Lehmann lovegrass 

(Eragrostis lehmanniana), spotted knapweed (Centaurea biebersteinii), and leafy 

spurge (Euphorbia esula), expand under future climate regimes.

• The invasive annual grass cheatgrass (Bromus tectorum) shifts northward with increased risk in Idaho, Montana, and Wyoming but reduced risk in southern Nevada 

and Utah and an overall loss of 13% of suitable habitat.

Currently, the three primary approaches for projecting the ecological effects of climate change are experimental and observational studies, mechanistic modeling, and 

bioclimatic modeling. A framework for integrating these approaches could improve 

our ability to forecast changes in vegetation types and species. Such a framework could 

include long-term experiments (for specific species, locations, and climate change scenarios) designed to support more comprehensive mechanistic models and to provide 

for testing bioclimatic models, and could be used to both identify and address the following critical research needs:

• Obtain the necessary species-specific data regarding important climate variables, 

biotic interactions, genetic variation, and adaptive capacity to improve predictive 

capacity.

• Determine the interacting effects of climate and other global change processes such 

as extreme events, increasing CO2, nitrogen deposition, pests, and disease on species distributions and community composition to improve predictions and develop 

adaptation strategies.

• Evaluate the interacting effects of socioeconomic and biophysical factors on land use 

and land cover change and, consequently, on species distribution and community 

composition to improve predictions and develop adaptation strategies.

• Increase our knowledge of the effects of species diversity and functional groups on 

ecosystem processes as they relate to climate and other global change factors to 

improve predictions and develop adaptation strategies.

• Continue to advance modeling efforts that couple bioclimate analyses with models 

that are able to estimate feedback, competitive interactions, and disturbance effects.

• Continue to explore other modeling approaches, such as combining historic distribution data with contemporary distributions, to identify species-specific climate 

limitations that can then be used to parameterize models.

Introduction

Experimental research and species distribution modeling predict that large changes 

in the distributions of species and vegetation types will occur due to climate change. 
USDA Forest Service RMRS-GTR-285. 2012. 3

Species responses will depend not only on their physiological tolerances but also 

on their phenology, establishment properties, biotic interactions (Brown and others 

1997), and ability to evolve and migrate (Davis and Shaw 2001). The capacity of species and, thus, their distributions to respond to a warming environment also will be 

affected by changing disturbance regimes and other global change factors (Turner 

2010). Because individual species respond to climate variation and change independently and differently, plant assemblages with no modern analogs can be expected 

(Williams and Jackson 2007). New plant assemblages might also arise in areas where 

novel climatic conditions develop (Williams and Jackson 2007). Support for predictions of novel climate regimes and corresponding plant assemblages is found in studies 

examining relationships among paleo-climate and plant community reconstructions. 

As Williams and Jackson (2007) pointed out: (1) many past ecological communities 

are compositionally unlike modern communities; (2) the formation and dissolution of 

past “no-analog” communities appear to be climatically driven and linked to climates 

without modern analogs; (3) many future climate regimes will probably lack modern 

analogs; and (4) novel communities and surprises should be expected in the future. 

Novel climate conditions coupled with vegetation communities that lack modern analogs pose significant challenges for resource managers. Accurate predictions of how 

species distributions will change under future warming are essential for developing 

effective strategies for maintaining and restoring sustainable ecosystems (Harris and 

others 2006).

Several factors make predicting how species distributions and vegetation communities will change difficult. Global Circulation Models (GCMs) exhibit significant 

variation in forecasts of future temperature and especially precipitation (Christensen 

and others 2007). This variation is often amplified for topographically variable areas 

such as the Interior West (Rehfeldt 2006; Saenz-Romero and others 2010). In addition, information on species’ relationships to climate variables is often lacking and 

must be inferred from data on current species distributions. And other factors such as 

competitive interactions with other species and disturbance regimes often obfuscate 

interpretation of species climate profiles in projected future climate space.

In grassland, shrubland, and desert ecosystems, our understanding of likely changes 

in climate is limited. Also, we lack information on the climate profiles of the vast majority of species. Here, we provide (1) current forecasts for changes in climate over the 

remainder of the century and (2) available predictions for changes in regional vegetation types and individual species distributions. We then discuss the types of approaches 

that can be used to increase our predictive capacity and the research needs for these 

ecosystems.

How Is Climate Predicted to Change?

Global climate is predicted to continue to undergo substantial change through the 

current century. The rate of predicted change is unprecedented relative to the three 

centuries prior to industrialization (North and others 2006). The effect of greenhouse 

gas emissions on future temperatures is considered using a set of emission scenarios. 

These scenarios predict future CO2 concentrations given human activities. The most 

optimistic scenario (B1) involves rapid reductions in emissions as societies turn toward 

a global vision of sustainable land use. More commonly used in modeling efforts, however, are scenarios that assume a continued increase in human population growth and 

consumption of fossil fuel. In particular, scenarios that predict a moderately high CO2

output with continued focus on economic growth (A2) or a more optimistic scenario 

that includes an emphasis on multiple energy sources (A1B) are used in the models 
4 USDA Forest Service RMRS-GTR-285. 2012.

discussed here. Western North American is projected (using 21 global models for the 

A1B scenario) to experience a linear change in mean temperatures and precipitation 

ranging from +2.1 °C to +5.7 °C and -3% to +14%, respectively, over the rest of the 

century (Christensen and others 2007). Recent models indicate that actual CO2 concentrations are likely to be higher than those used in the A1B scenario resulting in 

temperatures at the warmer end of those ranges (Tom and Harte 2006; Füssel 2009).

Within the region, patterns of temperature and precipitation change are expected to 

vary in magnitude and direction and in seasonal timing across the landscape. Figures 

1-1 through 1-3 illustrate this variation for projected change in summer temperatures 

(degree days >5 °C), winter temperatures (mean temperature in the coldest month), 

and aridity (dryness index) across the interior western United States using three GCMs 

under the assumptions of the A2 emissions scenario, (see Rehfeldt and others 2009). 

In general, summer temperatures and aridity are projected to increase the most in the 

southwestern United States and at lower elevations throughout the Interior West (figs. 

1-1 and 1-3). Predicted patterns of occurrence vary broadly across the region by GCM. 

Winter temperatures are projected to increase by 2 °C to 9 °C (fig. 1-2). Changing 

balance of temperature and precipitation is projected to result in earlier spring snow 

run off (Stewart and others 2004), declines in snow pack in the northern and central 

Rocky Mountains (Plummer and others 2006) and in the Great Basin (Mote and others 

2005), and increased frequency, duration, and spatial extent of drought events (Seager 

and others 2007; Sheffield and Wood 2008). Climate change effects are also likely 

to change disturbance processes, leading to increases in insect and disease outbreaks 

(see Chapter 7) and extended fire seasons with more frequent and intense fires (see 

Chapter 6).

Modeling Vegetation Response to Climate Change

Three main types of approach exist for projecting the ecological effects of climate 

change, and each of these differs in type of information provided and the zone of 

inference.

•  Experimental and observations studies include climate manipulation experiments 

(e.g., Harte and Shaw 1995; Suttle and others 2007), observations across climate 

gradients (e.g., Peñualus and others 2007), and observations over time (e.g., Alward 

and others 1999). These studies generate detailed data and provide valuable information on the complex interactions between species- and system-level responses. 

Results are often specific to the study system and are primarily relevant at local to 

regional scales.

•  Mechanistic modeling has been used to scale local processes to regional scales 

(Araújo and others 2005; Berger and others 2008; Jeltsch and others 2008). 

Mechanistic modeling is a bottom-up approach that uses measurements or expert 

opinion about species’ life histories, physiology, and competitive interactions to 

predict distribution under changing climate conditions. The results of this approach 

differ depending on the emphasis of the model and are limited by whether the physiological constraints used to parameterize them hold at regional scales.

•  Bioclimatic envelope modeling (BEM) uses the relationships among regional distributions of species and physical variables (e.g., climate) to constrain the potential 

distribution of species or groups of species (Pearson and Dawson 2003). BEM is 

a top-down approach that relies on biogeography to project distribution changes 

(Hijmans and Graham 2006), but that does not directly consider species attributes. 

BEMs have rarely been experimentally validated or tested (but see Chew and others 

2004; Farber and Kadmon 2003). This general approach has been vastly improved 
USDA Forest Service RMRS-GTR-285. 2012. 5

Figure 1-1. Mapped predictions of increase in the yearly accumulation of daily temperature sums over 

5 °C for decades 2030, 2060, and 2090 relative to the contemporary climate (1961-1990) for the 

A2 scenario using Canadian Center for Climate Modeling (CGCM3), Geophysical Fluid Dynamics 

Laboratory (GFDLCM21) and Hadley Center/World Data Center (HADCM3) (see Rehfeldt and 

others 2009).
6 USDA Forest Service RMRS-GTR-285. 2012.

Figure 1-2. Mapped predictions of increase in mean temperature of the coldest month for decades 

2030, 2060, and 2090 relative to the contemporary climate (1961-1990) for the A2 scenario 

using Canadian Center for Climate Modeling (CGCM3), Geophysical Fluid Dynamics Laboratory 

(GFDLCM21) and Hadley Center/World Data Center (HADCM3) GCMs (see Rehfeldt and others 

2009).
USDA Forest Service RMRS-GTR-285. 2012. 7

Figure 1-3. Mapped predictions of increase in annual dryness index for decades 2030, 2060, and 

2090 relative to the contemporary climate (1961-1990) for the A2 scenario using Canadian Center 

for Climate Modeling (CGCM3), Geophysical Fluid Dynamics Laboratory (GFDLCM21) and 

Hadley Center/World Data Center (HADCM3) GCMs (see Rehfeldt and others 2009). 
8 USDA Forest Service RMRS-GTR-285. 2012.

upon by newer bioclimate modeling techniques (see Rehfeldt and others 2006; 

Rehfeldt and Jaquish 2010; Saenz-Romero and others 2010). However, lack of detailed species distribution data limits the development of these more sophisticated 

bioclimate models for use in the Rocky Mountains.

How are Vegetation Types and Species Distributions Predicted to Change?

Understanding how species distributions and vegetation types will change due to 

increased temperature and altered precipitation regimes is essential for developing 

effective management and restoration approaches. In order to make efficient and effective decisions, managers must be able to identify the best management trajectory under 

future climate changes. Habitat protection, restoration, and facilitated adaptation are 

options that vary widely in their implementation, approach, and ultimate objectives. 

The feasibility of one approach over another will vary depending upon the degree and 

type of change experienced by an area. Consequently, many successful management 

decisions will be dependent on efforts to define future conditions.

Here, we present predicted changes in distribution of vegetation types and both 

native and invasive plant species in the interior western United States that are based 

on BEM and related analyses. Our capacity to predict the future distributions of species and vegetation types depends on our ability to accurately model future climate 

and relate species occurrence to environmental characteristics. BEMs describe the 

fundamental climate associated with the occurrence of a species and can incorporate 

detailed information on loss of suitable habitat and succession over time (Guisan and 

Thuiller 2005; Guisan and Zimmermann 2000; Franklin 1995). BEMs can be used to 

forecast a single species distribution based upon a set of environmental variables or 

can be more broadly applied to evaluate the distribution of a group of species (e.g., 

niche-theory models) (Botkin and others 2007). BEMs are commonly coupled with 

GCM output using various emission scenarios to predict the future distribution of species contemporary climate profiles or predicted realized climate niches based on their 

current distribution (Guisan and Thuiller 2005). In this application, resulting projections inherently retain substantial uncertainty associated with output from GCMs using 

emission scenarios.

Vegetation Types

The most comprehensive analysis of changes in vegetation types for the western 

United States was conducted by Rehfeldt and others (2006). In general, their predictions are consistent with those for more localized analyses of vegetation types 

(Kupfer and others 2004) and individual species (e.g., Neilson and others 2005). 

Rehfeldt and others (2006) modeled the contemporary distribution of Brown and 

others’ (1998) biotic communities using climate variables and projected future distributions using the IS92a scenario (also known as the “business as usual” emission 

scenario) summarized for a combination of GCMs produced by the Hadley Center 

and Canadian Center for Climate Modeling and Analysis (table 1-1 and fig. 1-4). 

Rehfeldt and others (2006) estimated that by the end of the century, 55% of future 

landscapes in the West will have climates that are incompatible with the vegetation 

communities that now occur on those landscapes. In addition, 85% of the affected 

area is estimated to have a climate that does not match the climate profile of any contemporary western biotic communities. Figure 1-4 shows projections of the climate 

profile of Brown and others’ (1998) biotic communities and responses under climate 

change during this century. The contemporary climate profile of the closest matching 
USDA Forest Service RMRS-GTR-285. 2012. 9

biotic community is used. It is worthwhile to note several general trends(calculated 

as percent change between current and year 2060 projections):

• Suitable habitat for Rocky Mountain subalpine conifer forests that account for approximately 1% of current cover disappears almost completely (99.7%) by 2090. 

Suitable climate for Rocky Mountain and Great Basin alpine tundra disappears 

completely by 2060.

• Habitat with climate suitable for Rocky Mountain conifer forest shifts upslope, 

displacing habitat currently occupied by subalpine coniferous forests, and shows 

corresponding losses at low elevations. Localized extinctions are expected, especially in the south.

• Habitat suitable for Great Basin conifer woodlands is projected to move north and 

upslope with principal gains in Colorado and southwest Wyoming and losses in 

the Southwest. An equivalent spatial shift for this vegetation type would constitute 

a continuation of migratory trends documented from the late Pleistocene to the 

present (Betancourt 1987).

• The climate conducive for Plains grasslands is reduced slightly with losses occurring primarily in Colorado and New Mexico. Semi-desert grassland habitat 

expands northward from the Southwest into the Great Basin, Colorado Plateau, 

and southern Great Plains and occupies an area nearly four times that of the 

present. Habitat suitable for Great Basin shrub grassland decreases by 40% and 

becomes highly fragmented.

• In addition to semi-desert grasslands, projected climate changes through 2060 appear to be most favorable for Mohave Desert (85% increase), Sonoran Desert 

(79% increase), and Chihuahuan Desert (167% increase) scrub vegetation types. 

Between 2060 and 2090, Mojave and Sonoran Desert vegetation communities 

will continue to expand while Chihuahuan Desert vegetation will contract. In general, habitat suitable for the Sonoran Desert scrub type expands northward as the 

Mojave Desert scrub type migrates to the Great Basin and Snake and Columbia 

River Plains.

• Habitat favorable to Great Basin Desert scrub is projected to expand in the short 

term and then becomes fragmented as it contracts later in the century. Great Basin 

montane scrub habitat will experience moderate decline and displacement by 

Mojave Desert species through time (69% decrease) (fig. 1-4). Fragmentation of 

these and other habitats can exacerbate other negative effects of climate change 

and influence realized shifts and persistence of habitats and species.

Table 1-1. Mean and range of three climate variables for nine biomes (Brown and others 1998) of the Interior West. 

Coldest month mean

Degree-days >5 °C temperature (°C) Annual moisture indexa

Biome Data points Mean Range Mean Range Mean Range

Sonoran Desert 4,338 5849 2439–6379 11.2 2.2–12.8 0.47 0.18–1.00

Mojave Desert 5,963 4491 3041–6220 6.3 -6.5–9.4 0.43 0.20–0.68

Great Basin Desertscrub 29,047 2227 571–3088 -3.7 -10.8–11.9 0.19 0.11–0.29

Great Basin Shrub-Grassland 36,348 2030 1187–2933 -3.6 -81 – -0.4 0.14 0.07–0.25

Great Plains 16,153 2321 1498–3427 -2.8 -8.7–3.2 0.13 0.09–0.17

Pinyon-Juniper Woodlands 82,248 2189 322–3138 -2.1 -11.5–2.8 0.14 0.08–0.24

Rocky Mountain Montane Forest 52,735 1444 988–3425 -5.0 -8.1–3.2 0.07 0.04–0.09

Rocky Mountain Subalpine Forest 41,646 812 542–2149 -8.5 -11.4 – -3.4 0.02 0.01–0.03

Alpine Tundra 8,344 413 237–1711 -10.6 -13.2 – -3.4 0.02 0.01–0.03

a (degree-days >5 °C)0.5 / mean annual precipitation.
10 USDA Forest Service RMRS-GTR-285. 2012.

Figure 1-4. Modeled realized climatic niche (upper left) and their predicted future distributions 

for Brown and others (1998) biotic communities in the interior western United States 

(Rehfeldt and others 2006). Lake positions were omitted from future projections.
USDA Forest Service RMRS-GTR-285. 2012. 11

Native Species

Analyses specific to species in the Interior West are limited but provide insights 

into the effects of climate change on several of the region’s keystone species. 

Following are some examples of these analyses:

• Crookston (see http://forest.moscowfsl.wsu.edu/climate/) used methodology described by Refeldt and others (2009) to create a species climate profile (contemporary realized climate niche) projected in six future climates (three GCMs and two 

scenarios) for several western species. Among these, one seed juniper (Juniperus 

monosperma) (fig. 1-5), Utah juniper (Juniperus osteosperma) (fig. 1-6), two needle 

pinyon pine (Pinus edulis) (fig. 1-7), and single leaf pinyon pine (Pinus monophylla) (fig. 1-8) are presented. These predictions generally agree with those of Rehfeldt 

and others (2006) for Great Basin conifer woodlands and indicate a continuation of 

migratory trends documented from the late Pleistocene to the present (Betancourt 

1987).

• Neilson and others (2005) used a coupled bioclimate model to predict a decline in 

sagebrush habitat in the Great Basin due to the synergistic effects of temperature 

increases and fire and disease, and to displacement by species moving north from 

the Mojave Desert in response to the northward shift in frost lines.

• Bradley (2010) used risk analysis to assess the interactive impacts of land use conversion (woodland expansion), land use (roads, agriculture, etc.), and cheatgrass 

(Bromus tectorum) invasion on sagebrush (Artemisia spp.) habitat in Nevada. 

Changes to the climatic habitat of Artemisia spp. were estimated using an ensemble of 10 atmospheric-ocean GCMs and 2 BEMs. Sagebrush vegetation types in 

southern Nevada were at the greatest risk of losing suitable habitat due to climate 

change. Disturbance risk was greatest in the northern section of the state. Overall, 

Bradley (2010) supported the conclusion that climate change poses a substantial 

risk to sagebrush ecosystems in the Great Basin.

• Shafer and others (2001) used bioclimate modeling to predict that sagebrush 

(Artemisia tridentata) will shift northward and exhibit substantial range contraction due to increased summer moisture stress. They also predicted northward 

expansions of Joshua tree (Yucca brevifolia), saguaro (Carnegiea gigantea), and 

creosote bush (Larrea tridentata). Creosote bush is expected to expand northward 

into areas currently occupied by big sagebrush, matching Neilson and others’ 

(2005) prediction.

• Warwell and others (2008) used bioclimate modeling to conclude that the contemporary populations of species with small distributions such as smooth 

Arizona cypress (Cupressus arizonica ssp. glabra) and the endangered perennial 

MacFarlane’s four-o’clock (Mirabilis macfarlanei) are likely to be subjected to 

complete climate disequilibrium earlier in the century than more broadly distributed species.
12 USDA Forest Service RMRS-GTR-285. 2012.

Figure 1-5. Mapped realized climate niche of one seed juniper (Juniperus monosperma; green) for contemporary 

climate (1961-1990) and future climates for decades 2030, 2060, and 2090. Colors coding in future projections 

indicate occurrence of agreement among 3 GCMs and 2 scenarios. Dark red = complete agreement for all 6 

combinations; red = agreement for 5 combinations; dark orange = 4; orange = 3; yellow = 2; and tan = 1 (see 

Rehfeldt and others 2009).
USDA Forest Service RMRS-GTR-285. 2012. 13

Figure 1-6. Mapped realized climate niche of Utah juniper (Juniperus osteosperma; green) for contemporary climate 

(1961-1990) and future climates for decades 2030, 2060, and 2090. Colors coding in future projections indicate 

occurrence of agreement among 3 GCMs and 2 scenarios. Dark red = complete agreement for all 6 combinations; 

red = agreement for 5 combinations; dark orange = 4; orange = 3; yellow = 2; and tan = 1 (see Rehfeldt and others 

2009).
14 USDA Forest Service RMRS-GTR-285. 2012.

Figure 1-7. Mapped realized climate niche of two needle pinyon pine (Pinus edulis; green) for contemporary climate 

(1961-1990) and future climates for decades 2030, 2060, and 2090. Colors coding in future projections indicate 

occurrence of agreement among 3 GCMs and 2 scenarios. Dark red = complete agreement for all 6 combinations; 

red = agreement for 5 combinations; dark orange = 4; orange = 3; yellow = 2; and tan = 1 (see Rehfeldt and others 

2009).
USDA Forest Service RMRS-GTR-285. 2012. 15

Figure 1-8. Mapped realized climate niche of one seed juniper (Pinus monophylla; green) for contemporary climate 

(1961-1990) and future climates for decades 2030, 2060, and 2090. Colors coding in future projections indicate 

occurrence of agreement among 3 GCMs and 2 scenarios. Dark red = complete agreement for all 6 combinations; 

red = agreement for 5 combinations; dark orange = 4; orange = 3; yellow = 2; and tan = 1 (see Rehfeldt and others 

2009).
16 USDA Forest Service RMRS-GTR-285. 2012.

Invasive Species

There is increasing evidence that invasive species also are being affected by climate 

change (Dukes and Mooney 1999; Walther and others 2009; Bradley and others 2010). 

However, forecasting changes in invasive species distributions poses some unique 

challenges because these species may not be in equilibrium with their newly invaded 

habitat. Analyses specific to invasive species are limited for the Rocky Mountain region 

and focus on only a few species. Bradley and others (2009) used bioclimate models to 

identify the risk of spread of four invasive species in the western Unites States:

• Spotted knapweed (Centaurea biebersteinii) is expected to expand into new areas 

of California and Nevada. Its current distribution in the coastal states (California, 

Oregon, and Washington) showed little change (loss of suitability) under future scenarios, but climate change may result in a transition to higher elevations in interior 

states (e.g., Montana, Wyoming, Colorado, and Utah).

• Tamarisk (Tamarix spp.) distributions were not well constrained by environmental 

characteristics, and the authors found no evidence for temperature constraints. Not 

surprisingly, the risk of tamarisk spread, which occurs within riparian areas and 

elevated water tables across most of the western United States, was not shown to be 

influenced by the direct effects of climate change.

• In contrast, leafy spurge (Euphorbia esula) is expected to expand throughout the 

majority of northern states west of Mississippi and some of the rangelands west 

of the Rocky Mountains. However, leafy spurge is likely to retreat from Colorado, 

Nebraska, Oregon, and Idaho as result of warmer temperatures.

• Cheatgrass (Bromus tectorum) is constrained by multiple precipitation variables 

and winter maximum temperature. Climate change is expected to shift potentially 

susceptible habitats northward, creating an increased risk in Idaho, Montana, and 

Wyoming but reduced risk in southern Nevada and Utah. Currently infested areas 

of central Utah and southern and central Nevada did not appear to remain suitable 

to this species, which may indicate a future range contraction. As much as 13% of 

the current invaded lands may not be suitable for this species under global warming and replacement by other invasive species is probable (e.g., red brome [Bromus 

rubens]). This pattern may be repeated with other invasive species.

Research Needs

Developing a framework for integrating the different approaches could significantly 

improve our ability to forecast changes in vegetation types and species. Such a framework might include long-term experiments (e.g., for explicit species, locations, and 

climate change scenarios) designed to explore the local complexities of species and 

ecosystem responses to climate change that support more comprehensive mechanistic models and provide for testing BEMs. The integration of experimental work and 

modeling could further our understanding of the ecological consequences of climate 

change, improve our projections for the future, and provide information necessary to 

maintain ecosystem services through development of effective conservation and restoration strategies in the context of climate change. Frameworks for predicting vegetation 

response to climate change that include research experiments, mechanistic modeling, 

and species distribution modeling can be used to identify and address critical information gaps, including the following:

• Obtain the necessary species-specific data regarding important climate variables, 

biotic interactions, genetic variation, and adaptive capacity to improve model 
USDA Forest Service RMRS-GTR-285. 2012. 17

predictions. To date, the primary emphasis has been placed on tree species. For 

deserts, grasslands, and shrublands, an increased emphasis needs to be placed on determining the relationships among climate variables and establishment and growth 

and reproduction of grasses, forbs, and shrubs. And in these arid to semi-arid systems, climate variables other than temperature and precipitation, such as snowpack 

duration and water-balance deficit, may better reflect the resources that are available 

for plant growth (Stephenson 1990, 1998).

• Determine the interacting effects of climate and other global change processes such 

as extreme events, increasing CO2, nitrogen deposition, pests, and disease on species distributions and community composition to improve model predictions of 

the synergistic effects of these processes (Campbell and others 2000). Long-term 

research and regional monitoring may be necessary to distinguish the effects of climate from other change agents and identify the indirect effects of climate on species 

interactions (Pearson and Dawson 2003).

• Evaluate the interacting effects of socioeconomic and biophysical factors on land use 

and land cover change and, consequently, on species distribution and community 

composition (Hansen and others 2001).

• Increase our knowledge of the effects of species diversity and functional groups on 

ecosystem processes as they relate to climate and other global change factor.

• Continue to advance modeling efforts that couple bioclimate analyses with models 

that are able to estimate feedback, competitive interactions, and disturbance effects 

(e.g., Huntley and others 2010; Franklin 2010). Risk models can incorporate empirical, experimental, or observational information into a geospatial (GIS) framework 

to create spatially explicit predictions of threats and change (Bradley and others 

2010). 

• Continue to explore other modeling approaches, such as combining historic distribution data with contemporary distributions, to identify species specific climate 

limitations that may then be used to parameterize models (e.g., Arundel 2005).

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USDA Forest Service RMRS-GTR-285. 2012. 21

Chapter 2

Restoring and Managing Cold Desert Shrublands for 

Climate Change Mitigation

Susan E. Meyer1

1

 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrublands, and Desert Ecosystems Program; 

Shrub Sciences Laboratory, Provo, Utah

Executive Summary

The equation for slowing global warming includes decreasing carbon emissions 

into the atmosphere as well as increasing carbon sequestration in the biosphere. Many 

proposed schemes for increasing carbon sequestration, such as afforestation of nonforested lands, involve tradeoffs with other resource values, including water availability. An alternative idea is to restore native plant communities to a condition that 

maximizes carbon storage without the need for continued resource inputs. Cold desert 

shrublands are a particularly good choice for management for increased carbon sequestration for the following reasons:

• Because of deep rooting systems, high root:shoot ratios, and relatively high standing 

biomass and soil organic carbon reserves, intact cold desert shrublands can store 

impressive amounts of carbon, more than one-third as much as the average biome 

on a per-area basis, in spite of dramatically lower productivity.

• The low productivity of cold deserts makes them relatively unimportant from an 

agricultural economics perspective. In a world where carbon sequestration itself is 

destined to become a commodity, shrubland management for increased carbon storage could potentially become a source of revenue that would offset the investments 

necessary to achieve restoration and management goals.

• A large fraction of cold desert shrubland in the Interior West is in severely degraded 

condition, with immense acreages dominated by weedy annual grasses that turn 

these systems into net carbon sources rather than sinks. While this is unfortunate 

in the short term, it means immense potential for increasing carbon sequestration 

through restoration of these degraded systems.

Accomplishing the task of returning degraded cold desert shrublands to their status 

as net carbon sinks will not be easy. The success of large-scale restoration through 

direct seeding is hampered by low and unpredictable amounts of precipitation, and 

this problem will only be exacerbated as climate continues to warm. The relatively 

low success rate for cold desert shrubland seedings highlights the need for innovative shrubland restoration research. Ongoing research by Rocky Mountain Research 

Station (RMRS) scientists in the Grassland, Desert, and Shrubland Ecosystems (GSD) 

Program that addresses these issues includes:

• Developing novel biocontrol strategies for both weedy annual grasses and secondary 

dicot weeds.
22 USDA Forest Service RMRS-GTR-285. 2012.

• Developing the knowledge and technology to increase supplies of seeds of diverse 

native species, including nontraditional early seral species that may be better able to 

cope with current and future disturbance regimes.

• Developing improved technology for site preparation and seeding.

• Developing knowledge and technology needed to restore microbiotic soil crusts.

• Developing knowledge of genetic variation in traits important to climatic adaptation in key species such as big sagebrush to better match seed source to current and 

future site conditions.

• Increasing understanding of successional processes in desert shrublands, grasslands, 

and woodlands in the context of global climate change and disturbance regime.

Introduction

Deserts and semideserts occupy approximately 22% of the Earth’s land surface 

(Janzen 2004), yet because of their low productivity, they are generally assumed to be 

relatively minor players in the global carbon cycle. Schemes to mitigate global climate 

change have rarely considered that improving carbon sequestration in deserts could 

make a significant contribution at a global scale. Many ideas for increasing carbon 

sequestration, such as tree plantations in marginally suitable environments, involve 

tradeoffs with other resource values such as water use and quality (Jackson and others 

2005). In contrast, improving carbon sequestration in deserts by restoring degraded 

shrublands to a more functional state would address a broad suite of resource values, 

including improved air and water quality, wildland fire abatement, enhanced wildlife 

habitat, biodiversity conservation, and aesthetic and recreational values.

Because cold deserts store much of their carbon belowground and that carbon is 

stored in deeper soil layers, these deserts are likely to store more carbon per unit area 

than warm deserts with monsoonal moisture regimes. In addition, the desert shrublands 

of the interior West might be more appropriately classified as semideserts, as they 

generally have much higher standing biomass than the true deserts, for example, the 

Sahara Desert of North Africa, which is virtually plantless over large areas except in 

drainages (“wadis”). This combination of high belowground allocation and relatively 

high biomass production appears to make cold deserts exceptionally good candidates 

for carbon sequestration.

The question addressed here is whether restoration on a broad scale in the interior 

West could make a significant contribution to climate change mitigation. The premise 

is that restoration of degraded cold desert shrublands could result in sequestration of 

significant amounts of carbon and could also reduce the negative climatic effects of 

excessive windblown dust. The consumptive uses of these ecosystems (e.g., livestock 

grazing), which could potentially interfere with management for carbon sequestration, 

could be said to be relatively unimportant economically, at least in the Interior West. If 

the carbon credit market that is taking shape internationally becomes fully functional, 

well-managed cold deserts may be able to provide more revenue as carbon sinks than 

as grazing lands. In addition, management for carbon sequestration can also be viewed 

as management for maximum return in terms of many other ecosystem services and 

amenity resources.
USDA Forest Service RMRS-GTR-285. 2012. 23

Carbon Storage in Deserts

Examination of carbon (C) storage patterns in major biomes on a global scale reveals 

that deserts (including semideserts) are responsible for the storage of a substantial proportion of the terrestrial C pool. Stored carbon may be present as standing biomass or 

as soil organic carbon (SOC), which is generally considered to be the more stable and 

persistent form. It dominates the terrestrial carbon pool at about 80% of total stored 

C (table 2-1; Janzen 2004). The relative contribution of C as standing biomass versus 

SOC in deserts is even more strongly biased, with over 95% of the stored C as SOC. 

Standing biomass C in deserts is estimated to account for only 1.7% of global total, 

whereas desert SOC is estimated to account for 9.5%. Overall, deserts, which occupy 

about 22% of the total land area, account for about 8% of terrestrial C stocks (Janzen 

2004), indicating deserts are generally about a third as effective as the average biome 

at storing C on a per area basis. Given the intrinsically unproductive nature of deserts, 

these figures at first seem surprisingly high. It is hard to see how systems that support 

such low standing biomass can generate so much SOC. But the same factor that generally limits biomass production in deserts, namely lack of water during much of the 

year, particularly when temperatures are warm, also limits the rate of microbial respiration in soil, leading to accumulation and persistence of SOC (Jobbagy and Jackson 

2000).

The vertical distribution of C in deserts also helps explain how they can be effective 

carbon sinks (fig. 2-1). When compared with other temperate region biomes, standing 

biomass, particularly in cold deserts, is dominated by the belowground portion, with 

root:shoot ratios averaging between four and five (Jackson and others 1996; fig. 2-2). 

The maximum rooting depth is deeper for cold deserts than for any other biome examined (Canadell and others 1996), and less than 55% of root biomass is found in the 

upper 30 cm of soil (Jackson and others 1996). This contrasts with perennial grasslands, which have similar standing biomass and relatively high root:shoot ratios, but 

with >80% of the root biomass in the surface 30 cm. This pattern of deep and extensive 

rooting in cold deserts is probably related to the need to capture winter precipitation 

stored at depth during the ensuing growing season, which is usually quite dry. The pattern is not seen in warm deserts, where summer monsoonal moisture patterns dominate 

and root:shoot ratios average less than one (Jackson and others 1996).

Table 2-1. Estimated terrestrial global carbon stocks by biome (Janzen 2004) and estimated mean carbon 

stock per unit area for each biome.

Global carbon stocks (Pg)

Biome Area (109 ha) Plants Soil Total Carbon stock/area

Temperate Forests 1.04 59 100 159 152.9

Boreal Forests 1.37 88 471 559 111.6

Temperate Grasslands/Shrublands 1.25 9 295 304 89.3

Deserts and Semideserts1 3.04 8 191 199 58.2

Tundra 0.95 6 121 127 17.9

Croplands 1.60 3 128 131 81.9

Tropical Forests 1.76 212 216 428 243.2

Tropical Savannahs/Grasslands 2.25 66 264 330 108.1

Wetlands 0.35 15 225 240 68.6

Total (not including ice cover) 13.61 466 2011 2477 182

% of total in deserts/semideserts 22.3% 1.7% 9.5% 8.0%

1

Area and carbon stock per area estimates in Janzen (2004) for the desert/semidesert biome have been adjusted by 

removal of areas of ice cover.
24 USDA Forest Service RMRS-GTR-285. 2012.

In deserts and shrublands in general, SOC and standing belowground biomass follow similar distribution patterns, that is, with more SOC in deeper soil layers relative 

to the surface layer than is found in either grassland or forest vegetation (Jobbagy and 

Jackson 2000). The estimated proportion of total SOC found from 1 to 3 m in depth is 

higher for deserts (0.86) than for any other temperate ecosystem. In general, SOC has 

a deeper distribution in soil than roots, and this is especially true in ecosystems with 

lower precipitation. The most likely explanation for this is that SOC turnover at depth 

is very slow. Dominance of more slowly degrading forms of carbon, lower nutrient 

concentrations, and more resistant root tissues at depth contribute to SOC persistence 

(Jobbagy and Jackson 2000).

The ability of cold desert soils to retain SOC could be reduced by the effects of 

ongoing climate change. Aanderud and others (2010) showed in an 11-year rain manipulation study that near-surface (0 to 30 cm) SOC stocks in a sagebrush steppe 

(Artemisia tridentata) community were significantly reduced when precipitation was 

shifted from a winter pattern to a spring-summer pattern. They credited this loss to increased microbial activity in wet surface soil at warm temperatures. Shifts from winter 

to spring-summer rainfall patterns are predicted for many parts of the Interior West 

as climate continues to warm (Zhang and others 2007). Rainfall timing impacts on 

Figure 2-1. The carbon cycle in a cold desert ecosystem, showing fluxes to the atmosphere (plant respiration 

and animal/microbial respiration/decomposition), uptake from the atmosphere by plants (primarily shrubs 

and grasses; photosynthesis), standing plant biomass, and shallow and deep soil organic carbon (SOC). If C 

uptake exceeds C flux to the atmosphere, C sequestration to a net C sink takes place, whereas if flux to the 

atmosphere exceeds uptake, the system functions as a net carbon source. Deep SOC, the most stable form of 

stored C, dominates C storage in deserts and semideserts. 
USDA Forest Service RMRS-GTR-285. 2012. 25

deep SOC would be expected to be lower, however, because deep SOC is more buffered from seasonal temperature changes. This would tend to mitigate the effects of 

increased warm-season precipitation on soil C storage.

C cycling on U.S. rangelands has been the subject of several recent studies and 

reviews (e.g., Bird and others 2002; Hunt and others 2004; Schuman and others 2002; 

Svejcar and others 2008; Follett and Reed 2010; Brown and others 2010). Synthesis 

of information on carbon storage on rangelands is complicated by the fact that many 

different vegetation types occurring under many different climatic regimes fall under 

the rubric of rangelands. Hunt and others (2004), working in Wyoming, found that 

Figure 2-2. Quantity and distribution of biomass carbon in cold 

desert biomes contrasted with other temperate zone biomes 

(grassland, chaparral, deciduous forest, coniferous forest, and 

cropland): (A) total standing biomass, (B) total root biomass, 

(C) root:shoot ratio, (D) % root biomass in the top 30 cm, 

and E) maximum rooting depth (adapted from Jackson and 

others 1996).
26 USDA Forest Service RMRS-GTR-285. 2012.

mixed grass prairie vegetation was carbon-neutral, whereas sagebrush steppe vegetation was acting as a carbon sink. Schuman and others (2002) focused on the potential 

to increase carbon sequestration in rangelands through improved management, particularly grazing management—their emphasis was primarily on grasslands. Svejcar 

and others (2008) reported the results of an interesting six-year study on net ecosystem 

C exchange at eight rangeland sites across a range of habitats. They found that both 

sagebrush steppe (cold desert shrubland) sites and three of four perennial grassland 

sites generally acted as C sinks during the course of the study, whereas the two warm 

desert sites acted as C sources. Whether a site acted as a source or a sink varied across 

years and was closely tied to precipitation patterns. Drought years limited productivity 

and tended to make even the most productive sites temporary C sources because even 

though C loss was also lowered by dry conditions, much reduced productivity shifted 

the C balance to net loss over the short term. The potential for increased C sequestration on rangelands, especially cold desert shrublands, is high, but the dynamics of the 

C cycle in rangelands are complex and temporally variable.

Shrubland Degradation and Carbon Storage

Historically, intact desert ecosystems were most likely in a steady state relationship 

with regard to carbon budgets, acting in the long term neither as sources nor sinks. But 

two sets of factors have been operating to disturb this steady state, and these factors 

generally operate in opposing directions. First, woody “encroachment” of former desert and other temperate grasslands is often thought to have shifted the carbon balance 

in these ecosystems to make them net carbon sinks. Whether conversion from perennial grassland to woody vegetation results in a net increase in C sequestration is the 

subject of considerable debate, however. Jackson and others (2002) found that whether 

woody encroachment of perennial grasslands resulted in an increase or decrease in 

SOC depended on precipitation. There was substantial loss of SOC with woody encroachment in more mesic environments, a loss sufficient to more than counterbalance 

the increase in standing biomass C resulting from the conversion to dominance by 

woody species. At the dry end of the spectrum, on the other hand, conversion from 

perennial desert grassland to shrubland resulted in increases in both standing biomass 

C and SOC. Most land managers regard woody encroachment as a form of degradation, but its causes are complex and, in many cases, not completely understood. 

Climate change may be driving woody encroachment in some ecosystems, for example, in the northern Chihuahuan Desert, where creosote bush (Larrea tridentata) 

and tarbush (Flourensia cernua) are actively invading desert grasslands (Van Auken 

2000). Changes in historic fire regimes, poor grazing management, and other factors 

may contribute to woody encroachment in other semiarid ecosystems, for example, the 

invasion of juniper (Juniperus spp.) species into sagebrush steppe in the Interior West.

The second process that has had a major impact on carbon storage in the deserts of 

western North America is the displacement of desert shrubs by invasive annual grasses 

through increased frequency of fire following destruction of the perennial herbaceous 

understory through improper grazing management. The conversion to annual grassland 

results in a transformation from slow to rapid C cycling, the cessation of C deposition in deeper soil layers, and the direct and rapid transfer of aboveground biomass C 

to atmospheric C associated both with the initial loss through fire of standing shrub 

biomass C and with subsequent increased fire frequency. This phenomenon has not received the attention of carbon brokers that has been given to woody encroachment, but 

it potentially has more impact on carbon budgets as it is very likely in the process of 
USDA Forest Service RMRS-GTR-285. 2012. 27

converting large portions of the Great Basin and surrounding areas into carbon sources. 

This possibility was apparently first noted by Bradley and others (Bradley and Mustard 

2005; Bradley and others 2006). Using sophisticated remote sensing technologies, the 

authors conservatively estimated that as of 2006, the area of former salt desert and 

shrub steppe vegetation in the Great Basin alone that has been converted through repeated burning to cheatgrass monocultures was on the order of 20,000 km2

. In addition, 

cheatgrass is not the only invasive annual grass that is having major impacts in western North America. Medusahead wildrye (Taeniatherum caput-medusae) and North 

Africa grass (Ventenata dubia) are major invaders in the Interior Northwest, while 

red brome (Bromus rubens) has become a driver of frequent large-scale fires in the 

Mojave Desert. Many of these fires are occurring in fire-intolerant shrub communities, 

for example, blackbrush (Coleogyne ramosissima) shrublands, that had very low preinvasion probabilities of burning (Brooks and others 2004).

Bradley and others (2006) also carried out an on-the-ground assessment of carbon stocks in cold desert shrublands versus cheatgrass monocultures. They measured 

aboveground carbon stocks and SOC in the near-surface soil horizon in burned and unburned salt desert shrubland (one site) and Wyoming big sagebrush steppe (two sites). 

They demonstrated a 3- to 30-fold decrease in standing aboveground carbon stocks as 

a consequence of type conversion to cheatgrass (figs. 2-3 and 2-4).

Figure 2-3. Standing biomass carbon in intact cold desert 

shrubland communities versus adjacent areas that 

have been converted to cheatgrass (Bromus tectorum) 

monocultures at Rye Patch, Nevada (salt desert 

shrubland), Button Point, Nevada (sagebrush steppe), and 

Jingo, Nevada (sagebrush steppe). Aboveground biomass 

data from Bradley and others (2006); belowground and 

total biomass estimated from independent root:shoot 

ratio data. 
28 USDA Forest Service RMRS-GTR-285. 2012.

While Bradley and others’ (2006) study did not include any assessment or estimate 

of root biomass C, root:shoot ratio information for the dominant species obtained from 

other studies can provide at least a rough estimate of root biomass C in these communities. A root:shoot ratio estimate of two was used for cheatgrass in the calculations that 

follow. This estimate is undoubtedly high; in greenhouse and field studies, root:shoot 

ratios greater than one for this species are rarely encountered, but a conservative estimate was chosen to avoid exaggeration of pre and post-burn differences (Meyer 

unpublished data). The estimate of six for the root:shoot ratio of Atriplex shrubs is 

based on estimates by Brewster (1968), while the estimate of four for the root:shoot 

ratio of Artemisia is similar to the estimates for cold desert shrublands in Jackson 

and others (1996). By revising the carbon stock data of Bradley and others (2006) to 

include these rough estimates, it can be demonstrated that the loss of belowground biomass carbon has the potential to contribute greatly to the effect of burning on carbon 

storage in these shrublands (fig. 2-2). Using these estimates, the biomass carbon stocks 

in the salt desert shrubland were reduced eight-fold through burning and conversion to 

annual grasslands, while those of sagebrush steppe were reduced from at least 6-fold 

to over 50-fold.

It is true that belowground C from shrub roots is still present for some undetermined 

length of time post-conversion, after the large pulse of CO2 emission from the combustion of the aboveground shrub biomass. But ultimately, this carbon will be released to 

the atmosphere, and without actively growing shrubs to replenish this belowground 

stock, the effect will be conversion of this formerly carbon-efficient system into a longterm source of atmospheric C. Estimates of biomass C loss from the study of Bradley 

and others (2006) ranged from 1.1 to 6.5 metric tons per ha for aboveground biomass 

C, 8.6 to 26.4 metric tons per ha for belowground biomass C, and 9.8 to 32.8 metric 

tons per ha for total biomass C.

Bradley and others (2006) combined their estimates of the areal extent of conversion to cheatgrass monoculture in the Great Basin with their estimates of reduction 

in aboveground biomass C stocks as a consequence of this conversion to calculate 

total biomass C released to the atmosphere (table 2-2). They estimated that about 8 

teragrams of C have been released to the atmosphere through shrubland conversion to 

annual grassland in the Great Basin as of 2006, and the potential for continuing type 

conversion and carbon release is immense. Adding estimated long-term belowground 

biomass carbon stock reduction resulted in an estimate of 29 to 60 teragrams of C that 

will ultimately be released to the atmosphere as a consequence of type conversion from 

shrubland to annual grassland that has already occurred in the Great Basin.

Invasive annual grass monocultures are not only very poor at carbon sequestration 

in terms of standing biomass relative to shrublands but also tend to concentrate their 

SOC near the surface and facilitate rapid turnover of both soil C and nitrogen (Norton 

Figure 2-4. Estimated loss of biomass carbon 

resulting from conversion from cold desert 

shrubland to cheatgrass (Bromus tectorum) 

monoculture at three Nevada sites (adapted from 

Bradley and others 2006).
USDA Forest Service RMRS-GTR-285. 2012. 29

and others 2004). This is perhaps one reason why it has been difficult to demonstrate 

direct losses of SOC following annual grass invasion or conversion to annual grass 

dysclimax (Gill and Burke 1999, Ogle and others 2004, Bradley and others 2006). 

Most of these studies have examined only the near-surface soil where SOC under annual grasslands is concentrated. The technology for the study of deep SOC remains 

cumbersome, so that information on this fraction of the C pool is not readily obtained.

Shrubland Degradation and Windblown Dust

Another consequence of anthropogenic disturbance on a landscape scale in arid and 

semiarid regions is a large increase in the load of windblown dust. This increased 

dust load results in negative health effects for the human population and also has the 

potential to exacerbate climate change by causing mountain snow cover to melt more 

quickly in the spring, reducing yield from the mountain watersheds that are the main 

source of water for human use. Restoring and managing cold desert shrublands for carbon storage also has the potential to reduce these negative effects of windblown dust 

by stabilizing the soil surface.

To examine the magnitude of the windblown dust effect, Neff and others (2008) 

analyzed rates of sediment accumulation in mountain lakes in southwestern Colorado 

over the last 5000 years. Results showed that the rate of sediment accumulation peaked 

very sharply in the second half of the Nineteenth Century, a time frame that corresponds with a massive increase in the scope and intensity of livestock grazing in the 

arid and semiarid regions to the west. The authors further demonstrated using mineralogical analysis that these sediments were not of local origin, but instead represented 

deposits of windblown dust from the valleys to the west of the watershed.

Livestock grazing and other human activities that disturb the surface soils of deserts generate dust by removing herbaceous plant cover and, often more importantly, 

through destruction of the cryptobiotic soil crust that stabilizes the surface in many 

desert regions (Neff and others 2005). These effects are further exacerbated by annual 

grass invasion and associated frequent fire. Annual grass cover provides some protection against wind erosion relative to bare ground but it prevents cryptobiotic crust 

recovery, resulting in increased dust generation, especially when these areas burn. The 

Milford Flat fire of 2007 was the largest wildfire in Utah’s history (Miller and others 

2011). An enduring legacy of this fire has been massive dust storms that have swept 

windborne dust into the urban areas of northern Utah and onto mountain watersheds. 

In addition to direct impacts on air quality and human health, this windborne dust exacerbates the effects of climate change through its effect on snow melt rates.

Snow cover has the highest albedo (light reflecting ability) of any natural land surface, and this ability to reflect light also reduces heat loading and melting rate (Flanner 

and others 2009). When particulate matter, such as dust or carbonaceous pollutants, is 

Table 2-2. Estimated biomass carbon loss as a consequence of conversion from cold desert shrubland to cheatgrass 

(Bromus tectorum) monocultures in the Great Basin as of 2006 (adapted from Bradley and others 2006).

Salt desert shrubland Sagebrush steppe Total

Aboveground biomass C loss (tons/km2) 110 250-650 360-670

Estimated total biomass C loss (tons/km2 ) 1000 1500-3200 2500-4200

Estimated area burned (km2) 2000 18,000 20,000

Estimated aboveground biomass C loss (teragrams) 0.2 4.5-11.7 4.7-11.9

Estimated total biomass C loss (teragrams) 2 27-58 29-60
30 USDA Forest Service RMRS-GTR-285. 2012.

deposited along with snow, it lowers the albedo of the remaining snow cover as the 

snow melts because the dark particles are concentrated near the surface of the snow. 

While it is true that particulate matter in the air lowers insolation and heat load on 

snow at the surface, this “dimming” effect is more than compensated by the reduction 

in snow albedo from these particles once they are deposited (“darkening effect”). This 

effect is especially pronounced in spring when large areas are snow covered and incident solar radiation is high. Flanner and others (2009) found that progressively earlier 

snow melt dates observed in Europe over the last few decades are almost as much due 

to this snow darkening effect of pollutants from fossil fuel combustion as to longterm increases in spring temperature caused by global warming. Moreover, the positive 

feedback from earlier snow melt caused by darkening created warmer spring temperatures independently of the effects of global warming, thus compounding the problem.

Though not as potent a darkening agent as carbonaceous pollutants, windborne 

dust can also significantly increase snow melt rates (Painter and others 2007). Spring 

dust storms in the desert region to the west of the mountain study area in southwestern Colorado resulted in several dust-on-snow deposition events per year, with more 

events in a drought year (2006, eight events) than in an average moisture year (2005, 

four events). These dust-on-snow deposition events resulted in snow cover durations 

that were decreased by 18 to 35 days. Shortened snow cover duration has measurable 

ecological impacts at the local scale in alpine and subalpine areas (Steltzer and others 

2009). More importantly, it also has the potential to significantly reduce water yields 

from mountain watersheds. Given that most of the agricultural and culinary water supplies in the Interior West are closely tied to mountain snowfall, and that the thickness 

and duration of the snow pack and its rate of melting have a strong impact on the ability 

to harvest this water supply, the fact that desert dust storms can shorten the duration of 

snow cover in mountainous areas downwind by a month or more should be of grave 

management concern (Painter and others 2007).

Managing Desert Shrublands for Climate Change Mitigation

Successful restoration of cold desert shrubland ecosystems on a large scale and subsequent management for ecosystem integrity could help to mitigate the adverse effects 

of climate change both through increased C sequestration and decreased effects of 

windblown dust. Climate change mitigation through desert shrubland management has 

the goal of maintaining or restoring adapted native shrubland vegetation that produces 

maximum carbon storage in the long term by exploiting all available niches, thereby 

maximizing productivity. It is likely that the vegetation that evolved in response to the 

selective forces in a particular environment will be best able to exploit its resources. This 

vegetation includes the woody shrub overstory, the herbaceous understory, and also the 

cryptobiotic crust community that occupies the interspaces. All these components are 

essential for long-term stability, including surface stability, and sustained carbon storage 

capacity.

An intact shrubland community is much more likely to be resilient in the face of continued climate change and other disturbances than “shrub plantations” analogous to the 

tree plantations being proposed and implemented for carbon sequestration. Emphasizing 

shrubs to the exclusion of other community components in a short-sighted effort to maximize carbon storage would probably result in vegetation that would require intensive 

management to be sustained. Annual grass weed invasion of the bare interspaces and 

consequent shrub loss through fire would be a constant threat. A more realistic goal, and 

one that is bound to be more effective in the long term, is to manage for intact shrubland 

communities that can rebound even from disturbances such as prolonged drought and 
USDA Forest Service RMRS-GTR-285. 2012. 31

fire without high risk of conversion to annual grass dysclimax. Both prevention of further 

degradation and restoration of degraded shrublands are part of this management scenario.

Cold desert shrublands in the Interior West currently exist in one of three states along 

a continuum of ecological condition:

(1) Some sites have relatively high-condition shrubland, with native understory and 

cryptobiotic crust still intact.

(2) Many more sites, perhaps most of the area still occupied by shrubs, are in some 

intermediate condition, with native perennial understory and/or cryptobiotic crust 

damaged or absent and with annual weed invasion in the understory. These sites are 

often at high risk of conversion to the third state.

(3) Loss of the shrub overstory through fire and post-burn dominance by annual 

grass weeds. Shrublands in these different states present different challenges 

and opportunities for management for carbon sequestration and windblown dust 

abatement.

The most important consideration for high-condition shrublands is prevention of degradation. This means keeping the cryptobiotic crust and the herbaceous understory in 

the best possible condition, which minimizes the probability of massive annual grass 

expansion after fire and maintains surface stability to minimize dust generation. Direct 

protection from invasion, for example, by controlling nearby weed infestations that could 

be propagule sources, is another way to maintain ecosystem integrity, as is providing 

priority protection in the event of wildfire. Even though occasional wildfire was a natural 

occurrence before settlement, especially in sagebrush steppe, protection from burning 

under current conditions is a top priority because of the threat of annual grass invasion.

Shrublands in intermediate condition often present more problems than opportunities 

in terms of improvement for climate change mitigation. Protection from further disturbance may result in little improvement in these shrublands. Loss of the seed bank of 

native understory species limits recruitment, and the cryptobiotic crust often cannot recover because of the heavy litter from annual grass invasion. In addition, a common 

occurrence, especially in sagebrush steppe, is shrub stand thickening or shrub canopy 

closure in response to loss of understory vegetation. The site at Jungo (Bradley and others 2006) seems to represent such a scenario. Sagebrush standing biomass was very high, 

and the understory was completely dominated by cheatgrass. Such a site could be described as “walking dead” in terms of the risk of conversion to annual grassland because 

a shrub-destroying fire is nearly inevitable. Natural shrub recovery after fire is often 

nil for dominant shrub species such as sagebrush and shadscale (Atriplex confertifolia), 

which cannot resprout after fire and rarely establish from seed in areas of high annual 

grass competition. Active management of shrublands with an understory dominated by 

cheatgrass will necessitate the development of effective tools to eliminate cheatgrass; 

reduce shrub cover, if necessary; establish understory species; and encourage cryptobiotic crust recovery, all with a minimum of surface disturbance. At present, such tools are 

largely unavailable.

Shrublands that have been converted to annual grass dysclimax communities have 

usually been given up for lost because of the futility of seeding into dense annual grass 

stands. But these annual grass dysclimax communities present the most hopeful scenario 

for increased carbon sequestration. If restoration of these communities is successful, 

substantial gains in carbon storage can be achieved. There should therefore be a strong 

emphasis on research aimed at increasing restoration success in areas that no longer 

support perennial vegetation. Many of the same tools needed to improve degraded 

shrublands will be needed to restore areas that no longer support shrubs, namely innovative methods for annual grass weed control and new approaches to improving seeding 
32 USDA Forest Service RMRS-GTR-285. 2012.

success in environments with low and variable precipitation. At present, many seedings 

in these environments fail, which may seem discouraging. But this points the way toward 

developing new approaches that, while they may be more expensive up front, could result in greatly improved seeding success and therefore a much better cost:benefit ratio for 

shrubland restoration in the long run. It is our challenge as researchers to develop these 

new approaches. With climate change mitigation as the goal rather than management of 

these shrublands for consumptive uses such as livestock grazing, scientists will be inspired to think outside the box and devise the methodology needed to make Interior West 

shrublands a significant C sink. This includes changing the paradigms for the timing 

of seedings to better take advantage of changed seasonal rainfall patterns. Even better, 

along with our partners in management, we will have the opportunity to simultaneously 

enhance the many other ecosystem services and amenity resources provided by these 

landscapes.

Research Needs and Activities

Scientists in the GSD Program at RMRS are actively involved in many areas of research that directly address the problems of desert shrubland restoration, and the potential 

exists for expansion into new areas of research that could yield a high return in useful 

knowledge for a relatively small investment of resources. Some examples of innovative 

research projects underway that will improve chances of restoration success are:

• Development of novel biocontrol strategies for eliminating the seed bank of cheatgrass 

and other invasive annual grasses through the use of augmentative mycoherbicidal 

approaches with naturally occurring fungal pathogens and endophytes.

• Development of classical biocontrol strategies with specialized herbivores from the 

native range for control of secondary dicot weeds, such as knapweeds and rush 

skeletonweed.

• Development of the knowledge and technology needed to produce seed supplies for 

diverse native species, including nontraditional early seral species such as annual and 

perennial forbs and shrubs that readily establish on disturbances.

• Increased understanding of the importance of seed microsites and development of new 

site preparation and seeding technology to increase probability of seeding success.

• Understanding of the biological importance of microbiotic soil crusts and development 

of technology for soil crust restoration.

• Knowledge of patterns of genetic variation in traits important to climatic adaptation 

in key species such as big sagebrush to better match seed source to current and future 

site conditions.

• Understanding of successional processes in desert shrublands, grasslands, and woodlands in order to separate the effects of climate change from the effects of disturbance 

regime and to potentially predict stable future plant communities to use as targets for 

restoration.

Some examples of new areas of research where rapid progress could be made if resources were made available are:

• Determination of limiting factors that restrict successful seedling emergence and 

establishment in restoration seedings in cold desert ecosystems, including soil crusting, absence of safe sites, pathogen attack, absence of beneficial symbionts, use of 

inappropriate seed sources, and uncontrolled competition from preexisting weeds, 

and devising strategies to overcome these limiting factors.
USDA Forest Service RMRS-GTR-285. 2012. 33

• Genetic characterization of a suite of native species potentially useful in cold desert 

shrubland restoration in terms of traits important in climatic adaptation in order to 

develop seed supplies specifically adapted to establish and persist in stable future 

plant communities on different site types.

• Development and field testing of seed sources for this suite of native species using a 

new set of criteria centered on ability to establish, compete, and persist on disturbed 

sites rather than on traditional agronomic criteria such as forage production and seed 

production.

Literature Cited

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rainfall reduces soil organic carbon storage in a cold desert. Ecosystems 13: 673-682.

Bird, S. B., J. E. Herrick, M. M. Wander and S. F. Wright. 2002. Spatial heterogeneity of 

aggregate stability and soil carbon in semi-arid rangeland. Environmental Pollution 

116: 445-455.

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land cover change: cheatgrass in the Great Basin. Remote Sensing of Environment 

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shrub type. Thesis, Brigham Young University, Provo, Utah.

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Effects of invasive alien plants on fire regimes. BioScience 54: 677-688.

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estimates of rangeland carbon sequestration potential in the U.S. Southwest. 

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1996. Maximum rooting depth of vegetation types at the global scale. Oecologia 

108: 583-595.

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warming and reduced snow cover from carbonaceous particles. Atmospheric 

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Follett, R. F. and D. A. Reed. 2010. Soil carbon sequestration in grazing lands: social 

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at three sites in the semiarid United States. Oecologia 121: 551-563.

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Reiners. 2004. Estimation of carbon sequestration by combining remote sensing and 

net ecosystem exchange data for northern mixed-grass prairie and sagebrush–steppe 

ecosystems. Environmental Management Vol. 33, Supplement 1: pp. S432-S441.

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Ecosystem carbon loss with woody plant invasion of grasslands. Nature 418: 

623-626.
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Jackson, R. B., J. Canadell, J. R. Ehleringer, H. A. Mooney, O. E. Sala and E. D. Schulze. 

1996. A global analysis of root distributions for terrestrial biomes. Oecologia 108: 

389-411.

Jackson, R. B., E. G. Jobbagy, R. Avissar, S. W. Roy and others. 2005. Trading water 

for carbon with biological carbon sequestration. Science 310: 1944-1947.

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Jobbagy, E. G. and R. B. Jackson. 2000. The vertical distribution of soil organic carbon 

and its relation to climate and vegetation. Ecological Applications 10: 423-36.

Miller, M., R. Reynolds, M. Reheis, H. Goldstein and J. Yount. 2011. Wind erosion and 

post-fire rehabilitation strategies—lessons from the Milford Flat fire. Proceedings 

of the Sixteenth Wildland Shrub Symposium: Threats to Shrubland Ecosystem 

Integrity. May 18-20, 2010. Logan, Utah.

Neff, J. C., P. Ballentyne, G. L. Farmer, N. M. Mahowald and others. 2008. Increasing 

eolian dust deposition in the western United States linked to human activity. Nature 

Geoscience 1: 189-195.

Neff, J. C., R. L. Reynolds, J. Belnap and P. Lamothe. 2005. Multi-decadal impacts of 

grazing on soil physical and biogeochemical properties in southeast Utah. Ecological 

Applications 15: 87-95.

Norton, J. B., T. A. Monaco, J. M. Norton, D. A. Johnson and T. A. Jones. 2004. Soil 

morphology and organic matter dynamics under cheatgrass and sagebrush-steppe 

plant communities. Journal of Arid Environments 57: 445-466.

Ogle, S. M., D. Dojima and W. A. Reiners. 2004. Modeling the impact of exotic annual 

brome grasses on soil organic carbon storage in a northern mixed-grass prairie. 

Biological Invasions 6: 365-377.

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disturbed desert soils on duration of mountain snow cover. Geophysical Research 

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potential carbon sequestration by rangelands. Environmental Pollution 116: 391-396.

Steltzer, H., C. Landry, T. H. Painter, J. Anderson, and E. Ayres. 2009. Biological 

consequences of earlier snowmelt from desert dust deposition in alpine landscapes. 

Proceedings of the National Academy of Science 106: 11629-11634.

Svejcar, T., R. Angell, J. A. Bradford, W. Dugas and others. 2008. Carbon fluxes on 

North American rangelands. Rangeland Ecology and Management 61: 465-474.

Van Auken, O. W. 2000. Shrub invasions of North American semiarid grasslands. 

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human influence on Twentieth-Century precipitation trends. Nature 448: 461-465.
USDA Forest Service RMRS-GTR-285. 2012. 35

Chapter 3

Climate Change and Arthropods: Pollinators, Herbivores, 

and Others

Sandra L. Brantley1 and Paulette L. Ford2

1 Museum of Southwestern Biology, University of New Mexico, Albuquerque, New Mexico

2 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Sciences Laboratory, Albuquerque, New Mexico

Executive Summary

The Interior West is rich in arthropod diversity because of its varied topography, 

which provides a wide range of elevations and levels of isolation for these small animals (Parmenter and others 1995). Some taxa are known rather well, such as butterflies 

and tiger beetles, but we have little information on many groups, which are known only 

from a few locations although they are probably more widespread. Arthropods live at 

large to small scales (e.g., migrating butterflies crossing countries to habitat specialists on rock outcrops or sand dunes). They may be generalists or specialists, vagile or 

sedentary, and have immature life stages that are similar or different from the adult 

(Triplehorn and Johnson 2005).

Predicted climate changes for the interior of North America, particularly the western 

portion, include:

• drier summers,

• increased precipitation outside the summer season, and

• increased frequency of extreme events such as heat waves.

Arthropods are ectothermic, so the temperature increases associated with global 

warming directly affect their development time, usually by decreasing the time needed 

for immature stages to become adults. This allows not only more generations per year 

in a given habitat, but it also opens new habitat for colonization because minimum 

temperatures no longer exclude arthropods (Crozier 2003; Robinet and Roques 2010).

The ecological roles of arthropods are important, even critical, and should be included when monitoring and predicting effects of global warming. Although knowledge of 

many arthropod species is lacking (Cane and Tepedino 2001), some groups have been 

studied in many areas and for long periods, such as butterflies, grasshoppers, some 

bees, and some ants. While these groups do not represent all arthropods, data about 

them give us a place to start in understanding their responses to climate change.

Research needs for arthropods include:

• There is a lack of data for many species and scientists are urged to develop longterm monitoring efforts with as many taxa as possible. There is also a great need to 

identify species interactions as climate changes, whether between insects and host 

plants or among insects and their competitors and predators.

• Several studies have shown the value of museum collections and data from natural 

history surveys. There is increasing need and emphasis from funding sources on 
36 USDA Forest Service RMRS-GTR-285. 2012.

grantees to publish specimen holdings online, including images, in order to build a 

more comprehensive understanding of species distributions.

• In the western United States, there is much state and Federal land that could be used 

for surveys and other interdisciplinary efforts. Opportunities are needed for collaborations among state and Federal agencies, universities, long-term ecological sites, 

state natural history surveys, and other natural heritage programs.

Introduction

The large number of arthropod (insects and their relatives) species in the world often 

means they are underrepresented in diversity assessments because of their “challenging taxonomy,” meaning that they can be hard to identify and many species have yet 

to be described. The Interior West is rich in arthropod diversity because of its varied 

topography, which provides a wide range of elevations and levels of isolation for these 

small animals (Parmenter and others 1995). Some taxa are known rather well, such as 

butterflies and tiger beetles, but we have little information on many groups, which are 

known only from a few locations although they are probably more widespread.

The ecological roles of arthropods are important, even critical, and should be included when monitoring and predicting effects of global warming. Important ecological 

“services” provided by insects include:

• Pollination

• Herbivory

• Seed dispersal

• Soil aeration and mixing

• Plant/animal decomposition

• Predation/parasitism

• Food source for many vertebrate species

They live at large to small scales (e.g., migrating butterflies crossing countries to 

habitat specialists on rock outcrops or sand dunes). They may be generalists or specialists, vagile or sedentary, and have immature life stages that are similar or different from 

the adult (Triplehorn and Johnson 2005).

Arthropods are ectothermic, so the temperature increases associated with global 

warming directly affect their development time, usually by decreasing the time needed 

for immature stages to become adults. This allows not only more generations per year 

in a given habitat, but it opens new habitat for colonization because minimum temperatures no longer exclude arthropods (Crozier 2003; Robinet and Roques 2010).

Many of these species depend on plants as food for immature and adult stages; an 

expected problem is the different effects of climate change on plants and insects, resulting in mismatches in timing between insects and their host plants (e.g., budburst, 

flowering time, and seed production). Species living at high elevations may continue 

moving up until there is no place left to go, and specialists may lose habitat or other 

resources to climate change and other effects such as land use change.

Some of the current shifts in distributions give the impression that arthropods can 

accommodate temperature increases, but they cannot continue shifting indefinitely, especially if projected temperature increases become larger. In addition, it is important 

to remember that not all species will respond the same way to changes in temperature, precipitation, and/or shifts in vegetation. We present research from western North 

America when possible and expand to Europe, the Northern Hemisphere, or the globe 

when no more specific information is available.
USDA Forest Service RMRS-GTR-285. 2012. 37

Key Issues

Predicted climate changes for the interior of North America, particularly the western 

portion include:

• Drier summers

• Increased precipitation outside the summer season

• Increased frequency of extreme events, such as heat waves (Easterling and others 

2000; Seager and others 2007; Diffenbaugh and Ashfaq 2010)

Over the Twentieth Century, average temperatures have increased about 0.6 °C, 

not only as higher maximum temperatures but also as higher minimum temperatures 

(Easterling and others 2000; Parmesan 2006) (i.e., there may be less difference between daily highs and lows). Seager and others (2007) suggested that one cause for 

these changes is a shift in atmospheric circulation cells toward the poles that is allowing sub-tropical dry zones to expand northward. Even if greenhouse gas levels do not 

increase above current levels, overall temperature increases will probably be 2 to 4 °C 

(Diffenbaugh and Ashfaq 2010).

Throughout the Midwest and the Southwest, the number of extreme precipitation events has increased over the Twentieth Century (Easterling and others 2000). 

Diffenbaugh and Ashfaq (2010) reported model results looking toward 2039, predicting increased winter precipitation but drying during the summer, so that soil moisture 

will decrease and evapotranspiration will increase. The authors also found that the 

Southwest is likely to experience summer heat waves every second or third year. 

Drought events from the El Nino Southern Oscillation system will also continue to occur and make the region drier still during those years (Seager and others 2007).

Within these broad-scale predictions, local conditions will vary, especially in arid 

regions where large daily temperature changes and variable precipitation virtually define the area (Noy-Meir 1973). Western North America contains numerous mountain 

ranges, which collect precipitation and nutrients such as nitrogen (N) and phosphorus 

(P) from air pollution. The complex topography can either retain these resources or 

transport them far from the source through watersheds (Seastedt and others 2004).

Increased winter precipitation in southeastern Arizona over 20+ years (Brown and 

others 1997) has resulted in species change in vegetation, insects and small mammal communities. Shrub establishment was improved so that they became a larger 

proportion of the vegetation cover, while grass species declined. Such a pattern had 

also been attributed to grazing effects or drought, but those factors were not involved 

at Brown and others’ (1997) site. The number of arthropod species present at the site 

did not change significantly but species composition did. The seed harvester ants 

Pogonomyrmex rugosus and P. desertorum, originally dominant species, became locally extinct, while Pheidole xerophila, also seed collectors, did not change, showing 

the independent response of species, even within one subfamily (Myrmicinae).

Chapin and others (2001) examined possible impacts on biodiversity across world 

biomes from five drivers: climate change, land use change, N deposition, biotic exchange 

(e.g., invasive species), and increased atmospheric CO2. The authors found that land use 

change produced impacts as large as or larger than climate change for most biomes (results summarized in Sala and others 2000). In many areas, land use change has already 

altered biodiversity extensively, so that climate change has not produced much additional 

change. As an example for deserts, Huenneke (2001) described the Chihuahuan Desert 

of southern New Mexico as an area further desertified by agricultural and grazing use. 

For grasslands, Sala (2001) noted that conversion to agriculture is the major impact and 

that climate change effects are expected to be moderate. Other human-caused factors of 
38 USDA Forest Service RMRS-GTR-285. 2012.

importance to grasslands are pollution from nearby cities and loss of water to cities and 

agriculture.

In desert systems, increased precipitation is usually thought of as beneficial, but if it 

arrives at the wrong time of year and as extreme events, as is predicted by Easterling and 

others (2000), the effects can be harmful, washing away soils, and ruining seed banks 

and insect eggs, and affecting water quality in rivers and lakes. In particular, specialist 

arthropods will be impacted because their host plants or substrates tend to be patchy in 

distribution in arid systems (Huenneke 2001).

Arthropods vary widely in their abilities to avoid or tolerate these kinds of changes. 

The distributions of many of the thousands of species are poorly understood, making 

it difficult to know how or if they will be affected by land use and changing climate. 

Because insects are ectothermic, temperature increases that most commonly characterize climate change will undoubtedly affect them in various ways such as more rapid 

development through life stages, more generations per year, and an increase in favorable 

habitat that was formerly too cold.

There may also be contraction of species’ ranges as some areas become too hot 

(Robinet and Roques 2010). If species are forced to move out of habitats that have become too hot, they are seen as “rescued” by having moved northward or up in elevation 

(Parmesan 2006). However, a given species may not be able to maintain populations in 

these new areas because of substrate or food needs and because there will already be 

other species established there. It is not clear that the shifting populations are undergoing 

genetic change over the short term to better fit them to the new habitats (Parmesan 2006). 

A likely negative impact will be a mismatch or asynchrony with host plants for larval growth or maintenance of adults during their reproduction periods (Parmesan 2007; 

Robinet and Roques 2010).

Following, we discuss current knowledge of representatives of pollinators, herbivores, 

omnivores, and detritivores (scavengers).

Pollinators: Butterflies

By far, the best known pollinators are butterflies because of their rather large size, 

relative ease of field identification, and attractiveness to the general public. They have 

become the representatives of all insects in climate change studies because datasets are 

available for them (Inouye 2007). However, butterflies do not necessarily respond as 

other insect taxa do (Hickling and others 2006), highlighting a need for monitoring studies of other groups.

In the western United States, many studies have been done for the Pacific Coast (e.g., 

McLaughlin and others 2002; Crozier 2003; Preston and others 2008), but fewer have 

focused on the interior of the continent. In two mountain ranges in Nevada, Fleishman 

and MacNally (2003) studied species richness of butterflies from 1996 to 2002, testing how well a series of short sampling periods or “snapshots” would detect possible 

climate change effects. Their results showed that richness between the mountain ranges 

and even among sites within a range were greater than any changes across the six years. 

Species living successfully in arid climates already had life history strategies for habitat 

variability in abiotic and biotic resources (what the authors called “tough-tested”). In an 

earlier study modeling species changes in six mountain ranges (one in California, five 

in Nevada), Fleishman and others (2001) used a climate scenario that increased temperatures enough to move plant species up in elevation by 500 m. The butterflies were 

expected to move as much also, assuming their dispersal capabilities were good. Under a 

moderate temperature increase, the Great Basin area would probably not lose many butterfly species, but each mountain range had its own patterns of species richness making 

it impossible to generalize outcomes to other areas.
USDA Forest Service RMRS-GTR-285. 2012. 39

Climate change patterns not only show an increase in maximum temperatures but in 

minimum temperatures as well, so that some areas experience fewer days of freezing or 

very cold weather. These minimum temperatures are increasing at a faster rate than the 

maximum temperatures (Easterling and others 2000; Parmesan 2006) and are important 

to monitor for their effects on arthropod distributions. Crozier (2003) tracked the northward movement of the sachem skipper (Atalopedes campestris) from its historical range 

in the southern and western United States and Mexico as it moved up the Pacific Coast to 

Eugene, Oregon, in 1967; Portland, Oregon, in 1985; Tri-Cities (Richland, Kennewick, 

and Pasco) area of Washington in 1993; and Yakima, Washington, in 1999. Her work 

showed that the skipper is limited by low temperature and can readily take advantage 

of increasing temperatures to move into new areas. The host plants for this species are 

grasses (Brock and Kaufman 2003), including crabgrass and Bermuda grass that are 

planted in cities and towns, so not only climate change but also conversion of land to 

urban settings favor the spread of this species. The author encouraged conservationists to 

not only save current habitat but also areas that may become suitable in the future.

A European butterfly dataset from the past 30 to 100 years found similar shifts 

(Parmesan and others 1999). For species whose entire range was known, the authors 

found 22 extended distributions to the north, 2 extended distributions to the south, and 

11 remained stable. These changes reflect the 0.8 °C increase in temperature over the last 

century, suggesting that a further increase of 2 to 4 °C could leave some species with no 

further possible movement northward or upward in elevation. Poor dispersers would be 

particularly at risk because of the added factor of increasing habitat fragmentation, leaving them in shrinking “islands” of climatically suitable habitat.

Forister and others (2010) investigated the combined effects of climate change and 

habitat loss, the two most important drivers of species change (Sala and others 2000), on 

butterflies in the Sierra Nevada in California. The data covered 159 species over 35 years 

and an elevation gradient of 0 to 2775 m. The elevational gradient allowed the authors 

to determine whether climate or land use produced species changes. At lower elevations where human populations were higher, the greater impact on butterflies came from 

habitat loss. At intermediate and high elevations, climate change was the likely cause of 

species change. At middle elevations, species moved up, and at the highest elevations 

(alpine), habitat specialists declined as conditions warmed, but overall species and abundance increased because of the influx of intermediate elevation species. An unexpected 

result was the decline in ruderal species at low elevations. These widespread species are 

generally predicted to do well in the face of climate change. In this case, however, habitat 

destruction was severe enough to remove both larval host plants and nectar sources for 

adults, showing that even generalist species may also need conservation protection in 

some areas. The authors suggested that one reason generalists do well in Europe is that 

they have had centuries to adapt to land use change; what we see there today are the species that are successful in highly disturbed and managed habitats.

Generalist butterfly species were predicted to do well in Canada, based on models 

by White and Kerr (2007), for 102 species. The authors built a long-term dataset from 

museum records (some extending back to the Nineteenth Century) and added geographic features such as elevation and changes in land use, temperature, and precipitation. 

Temperature and elevation were the best predictors of butterfly species richness; precipitation was an important factor only in the driest areas, the prairies of south-central 

Canada. As is common in other studies, the authors found a decline in specialists due 

more to loss of forest habitat than climate change. High richness was found in two areas 

but for different reasons: in high elevations, habitat heterogeneity provided many suitable microhabitats; but in lower elevations, which are dominated by people, diversity 

was high because of land use change. The increased expansion of already widespread 
40 USDA Forest Service RMRS-GTR-285. 2012.

generalist species has led to what the authors called a “homogenization” of species across 

the southern part of the country. Though the diversity is high, it is being maintained by 

disturbance and agriculture.

From a pair of studies, we can compare two closely related butterflies, the Quino 

checkerspot (Euphydryas editha quino) and the Bay checkerspot (Euphydryas editha 

bayensis) in California. The Quino checkerspot is a well-studied endangered subspecies that is a habitat specialist in the shrublands of southern California. Preston and 

others (2008) modeled its possible population changes under temperature and precipitation changes. With a temperature increase of 0.6 °C and no change in precipitation, the 

butterfly would probably do well. Under temperature increases of 1.7 and 2.8 °C and precipitation changes of -50% or +150%, suitable habitat would be reduced by 98 to 100%. 

In this Mediterranean-type habitat, drought can be common, so the modeled reduction 

in precipitation would understandably produce habitat loss, but it is interesting to note 

that precipitation increases also virtually wiped out the habitat. In both cases, vegetation 

species and cover would presumably change so much that the butterfly could no longer 

be supported. Larval host plants are dwarf plantain, penstemon, and Indian paintbrush 

(Brock and Kaufman 2003). As a specialist, it is unlikely to move readily to grasslands 

or forests even if they are relatively close.

The Bay checkerspot butterfly (Euphydryas editha bayensis) is also a specialist on the 

same plants as the Quino checkerspot (Brock and Kaufman 2003) but lives in grasslands 

of the San Francisco Bay area of California. McLaughlin and others (2002) investigated 

the factors leading to the extinction of two populations (in 1991 and 1998) of this threatened subspecies. The most important factor was a change in the pattern of precipitation: 

increased variability after 1971, including more severe weather events. The larvae and 

their host plants were also affected, with a shortened period of overlap and a resulting 

increase in larval mortality. A second impact on the populations was the loss of metapopulations nearby to recolonize the area because of habitat fragmentation and urban 

growth.

Pollinators: Bees

Although bees are major pollinators of many wild and managed plants, they are less 

well-known than butterflies (Inouye 2007); many species are difficult to identify in the 

field. In North America, there are few data to document a decline in these pollinators 

because of high interannual variation in abundance, the effort and expense needed to 

adequately monitor and identify bee diversity, and our lack of knowledge about suitable 

habitat for many species (Cane and Tepedino 2001). Studies show varying amounts of 

response to climate or disturbance impacts; each local area has its own combination of 

factors that influence richness and abundance.

In 1997, the U.S. National Research Council published a report on the status of pollinators in North America in agricultural settings. Along with the honeybee (Apis mellifera), 

common introduced species that are extensively managed for crop production include 

several bumblebee species (Bombus spp.) and a leafcutter bee (Megachile rotundata). All 

are susceptible to parasites, pathogens, and pesticides, as are native species. In addition, 

these introduced generalist species frequently outcompete native species by reducing 

overall nectar availability in some areas. At present, the authors concluded that the effects 

of transgenic corn with incorporated Bacillus thuringiensis (Bt) poses a small threat to 

native bee species because Bt targets herbivorous caterpillars. However, a greater cause 

for concern is that crops with increased herbicide resistance allow farmers to treat their 

fields for weeds without damaging their crop species. Those weeds are nectar sources for 

many native bee species. If more cropland is developed with these resistant plants, a possible indirect result will be the loss of native pollinators on non-crop plants.
USDA Forest Service RMRS-GTR-285. 2012. 41

Little information about bees is available for western North America (Cane and 

Tepedino 2001), but two studies in the Rocky Mountains of Colorado illustrate shifts 

in the relationships between bumblebees and their host plants. Inouye (2007) has been 

monitoring bumblebees for several decades in the area surrounding Rocky Mountain 

Biological Laboratory in Crested Butte. At least one bumblebee species has moved 

upward some 457 m in elevation with a corresponding 1.4 °C increase in temperature 

during this time. Further temperature increases of 2 to 4 °C are predicted for the western 

United States over the next 30 years (Diffenbaugh and Ashfaq 2010), and it is not clear 

if bumblebees will be able to keep up with such change. Inouye (2007) noted that host 

plants may not be moving at the same rate as their pollinators (also in Parmesan 2007) 

because their needs also include soil moisture and substrate type.

In the second study in western Colorado, Thomson (2010) studied blooming and fruit 

set of a lily, Erythronium grandiflorum, which is pollinated by bumblebees. Some members of the lily population start blooming early as soon as snow melts, which may vary by 

a month from one year to the next. There were occasional frosts even after snowmelt, and 

even when early and mid-period blooming individuals survived frosts, they were often 

limited by the small number of pollinators available (they are ectothermic and limited 

by very cold temperatures). Later-blooming individuals had the advantage of a greater 

number of bumblebee colonies. Thomson reported that the lack of synchrony between 

lily and bumblebees has worsened in recent years. If winter precipitation patterns and 

extreme storm events increase as predicted (Easterling and others 2000), early blooming 

lilies will be severely limited by frosts and the lack of bumblebees.

In Carlinville, Illinois, Marlin and LaBerge (2001) compared bee data from surveys in 

the 1970s with records from Charles Robertson’s surveys from 1884 to 1916. Of the 214 

species that Robertson collected, Marlin and LaBerge found 140 (65%) in their surveys, 

as well as 14 species that Robertson did not collect. Since the 1880s, the area has lost 

and gained forest cover and has increased the amount of land converted from prairie to 

agriculture, yet the bee fauna was largely intact. The authors attributed this to a diversity 

of remnant habitats between agricultural fields. Winfree and others (2009) for bees, and 

Forister and others (2010) for butterflies, showed that insects managed relatively well 

below a threshold of extreme disturbance; the Illinois bee fauna in the 1970s had perhaps 

not experienced such severe disturbance. These studies show that in some cases, pollinators can cope with low to moderate levels of disturbance.

A more recent European study documented broader patterns and explanations for declines in bees. Biesmeijer and others (2006) took advantage of a good historical record of 

bee distributions in Britain and the Netherlands to compare changes in species richness in 

10 km x 10 km map grid cells before and after 1980. For grid cells with enough data, bees 

in Britain declined in 52% of cells and increased in 10%, while in the Netherlands, bees 

declined in 67% of cells and increased in only 4%. The species with the largest declines 

were specialists on certain flowers or habitats, had only one generation per year, and did 

not migrate. The authors also found that plant species with specialist pollinators declined, 

wind-pollinated species increased, and self-pollinating species showed no change. In 

Britain, species increases were for those that were already widespread (as in White and 

Kerr [2007] for Canadian butterflies). In both countries, species became less evenly distributed, that is, fewer species made up a larger proportion of those present.

Winfree and others (2009) used datasets from the world in a meta-analysis of bees and 

human disturbance. Although their questions did not include climate change, their work 

showed the importance of land use change in altering bee communities. The biological 

aspects they looked at were managed versus wild species and solitary versus social species, grouping genera into Apis, Bombus, and other. Disturbance factors included habitat 

loss/fragmentation, pesticide use, fire, deforestation, and grazing. Habitat loss was the 
42 USDA Forest Service RMRS-GTR-285. 2012.

primary factor reducing bee species and abundance, but only when such loss was extreme. This was also the case with Sierra Nevada butterflies (Forister and others 2010) 

and their results agree with the predictions of Sala and others (2000). By dividing the 

bee data by genus and life history features, it became clear that disturbance impacts 

were varied (e.g., tree-nesting species would be negatively affected by deforestation, but 

the increase in open land might favor ground-nesting species). However, even with the 

combined power of multiple studies, it is not possible to extrapolate these results to all 

bees in all areas.

Pollinators: Flies

The true flies, Order Diptera, are a hyperdiverse group with over 21,000 species (86% 

endemic) in the Nearctic (Bio Systematic Database of World Diptera). They are usually 

thought of as disease vectors (mosquitoes), crop pests (leafminers), or livestock pests 

(stable flies), but they are important parasitoids of other arthropods and are pollinators as 

well. As pollinators, many species are generalists, but the syndrome called sapromyophily (flowers producing appearances or scents of decaying flesh as attractants) shows 

that plant-fly interactions have existed for long periods to produce such specializations 

(Kearns 2001). Flies are common at high elevations, pollinating a variety of arctic and 

alpine plants (Kearns 2001). Although they are often not the target of studies of plantinsect interactions in these habitats, large numbers of individuals may be collected but 

not prepared or identified because of time/budget constraints. Their identification can be 

difficult, but depositing specimens in museums can provide material for future research. 

Syrphidae, also known as flower flies or hoverflies, is a family of common and important 

pollinators (Triplehorn and Johnson 2005). In Britain and the Netherlands, Biesmeijer 

and others (2006) studied syrphid fly records before and after 1980 in 10 km x 10 km 

map grid cells as previously discussed for bees. In Britain, syrphid richness increased in 

25% of cells but decreased in 33%, while in the Netherlands, richness increased in 34% 

of cells and decreased in 17%. The authors interpreted these changes as shifts in species’ 

distributions in many cases. The study showed larger declines of specialists and those 

with only one generation per year. Compared with bees in the same study, the syrphid 

flies did not decline as much, perhaps in part because their larvae are more varied in food 

sources (some are predators, others are detritivores, and others are herbivores, while bee 

larvae feed only on pollen and nectar) (Triplehorn and Johnson 2005).

Herbivores

Two important groups of insects that feed on foliage are grasshoppers (Orthoptera) 

and moths and butterflies (Lepidoptera). Grasshoppers feed on foliage in all life stages, 

while lepidopterans feed on foliage in the larval or caterpillar stage. As adults, moths and 

butterflies are often important pollinators; (butterflies are discussed in the “Pollinators” 

section). There is little information on responses of moth pollinators to climate change. In 

Missouri, Forkner and others (2008) studied 15 families of moths whose caterpillars feed 

on oak leaves. Among these species were those with one generation per year or several, 

seasonal feeders or those feeding most of the year, and those that were mobile or that 

fed in leaf rolls or mines. The study questions involved variability in population density 

as a way of predicting which life history patterns might be more vulnerable to climate 

change. Those with the highest variability were spring feeders that were not mobile, suggesting that they may be most vulnerable to a mismatch in timing between caterpillar 

development and oak budburst. Species that fed over a longer period and that could move 

within or between trees would be able to compensate for plant/insect timing mismatches.
USDA Forest Service RMRS-GTR-285. 2012. 43

Melanoplus sanguinipes, the lesser migratory grasshopper, is found over all of the 

United States (except for Florida and western California) and into southern Canada 

(Capinera and Sechrist 1982), and because it can be an outbreak species on agricultural 

crops, it has been monitored well in some areas. Olfert and Weiss (2006) used records 

from Saskatchewan from 1931 to the present to model its possible response to climate 

warming. The most favorable habitats for the grasshopper were mixed grassland and 

moist mixed grassland. Under conditions of a 2 °C temperature increase, the species 

became a possible outbreak pest in 17.3% of Canada. With a 4 °C increase, the species colonized new habitats such as the Boreal Plains and Boreal Shield, and Canada’s 

susceptibility to outbreaks increased to 28.2%. With a 6 °C increase, the species would 

be able to live in most of Canada. In all scenarios, M. sanguinipes had the potential to 

become a major pest in cereal crops.

Nufio and others (2009) made use of a collection of grasshopper data at the Colorado 

Museum of Natural History by comparing Gordon Alexander’s collections from 1958 

to 1960 with the present. Alexander surveyed grasshoppers at different elevations in 

the Rocky Mountains near Boulder, Colorado. Preliminary results from the presentday comparison showed that species are hatching and reaching adulthood 15 to 30 

days earlier than in Alexander’s time. Work will continue through 2012 (see www.

ghopclimate.colorado.edu).

Omnivores: Ants

On a global scale, ants are thermophiles, with highest diversity in the tropics. In North 

and South America, their richness, both past and present, is well explained by temperature (Dunn and others 2009), but under current climate conditions, fewer species were 

found than expected in North America. Dunn and others (2009) suggested that current 

richness represents a loss due to past climate change as far back as the Eocene, after 

which North America began to cool more than South America. Some North American 

fossil species represent taxa that are now found only to the south in more tropical habitats. It is likely that increased temperatures will allow not only current species to move 

north but also allow species that were formerly in North America to return.

In western North America, Kaspari and others (2000) studied ants in numerous habitats: desert shrubland, grassland, coniferous forest, and tundra. They accounted for 

70% of ant abundance through positive correlations with temperature, plant productivity, and seasonality. From this, the expectation was that increased regional temperatures 

would favor the activity and spread of ant species, but the authors suggested that the 

ants have also benefited from cold winter temperatures through lower metabolic costs. 

A shorter winter season would increase the amount of time available for foraging but 

there might not be increased plant productivity, in which case the ants would not recover their energy investment through seed harvesting or predation on herbivorous insects.

Detritivores and Predators

Very little work has been published on the effects of climate change on arthropods 

that have little impact on human activities. However, in Britain, Hickling and others (2006) compared occurrence records from 1960 to 1975 with those from 1985 to 

2000 for many taxa. Species that were included were those found only in southern 

or lower elevations in Britain in the earlier years to see if they moved northward in 

the later years. On average, species shifted northward 31 to 60 km or upward 25 km. 

These shifts were significant for grasshoppers, butterflies, long-horned beetles, ground 

beetles, soldier beetles, metallic wood-boring beetles, millipedes, isopods, spiders, and 
44 USDA Forest Service RMRS-GTR-285. 2012.

dragonflies. Shifts were not significant for lacewings or harvestmen. These taxa cover 

a range of ecological roles, including predators, detritivores, and herbivores, with implications for local changes in ecosystem processes. Species responded independently; 

there was no overall pattern within a taxon or trophic group. The species considered 

here were moving at sometimes different rates than better-monitored species, such as 

butterflies. We should not assume that information on only a few taxa will be a good 

substitute for all arthropods.

Synthesis

A number of general patterns emerge from the information presented:

• Land use change (habitat loss/fragmentation) is as important as climate change as a 

driver for altering insect populations and communities. In some areas, species richness has already been affected by habitat loss, so climate change may not produce 

much more of an impact.

• Many insects have already expanded northward or upward in elevation in response 

to the 0.6 °C temperature increase of the last century. An additional increase of 2 to 

4 °C as predicted may mean a permanent loss of habitat for some species.

• Many insect species are resilient to low to moderate amounts of disturbance. These 

changes favor some species and disfavor others.

• The mismatch between insects and their host plants is increasing for some species. 

Species most at risk are host specialists, active early in the growing season, poor 

dispersers, and/or have only one generation per year. Generalist species are likely to 

become even more widespread. There is a lack of information on interactions as a 

result of species shifting into new areas.

• There are not enough long-term datasets on most insect species. The predictions 

from models need to be validated with more field data—model results can vary 

widely. Insect species respond independently to environmental changes, so studying 

only a few species will not predict what the rest will do. Monitoring must involve 

interdisciplinary efforts to integrate data on species, climate, and other environmental factors.

Research Needs

• Almost all of the researchers cited in this chapter have noted the lack of data for 

many species and urged current workers to develop long-term monitoring efforts 

with as many taxa as possible. There is also a great need to look at species interactions as climate changes, whether between insects and host plants or insects and 

their competitors and predators.

• Several studies have shown the value of museum collections and data from natural 

history surveys. There is increasing emphasis from funding sources on grantees to 

publish specimen holdings online, including images, to build a more comprehensive 

understanding of species distributions.

• In the western United States, there is much state and Federal land that could be 

used for surveys and other interdisciplinary efforts. We can create opportunities for 

collaborations among state and Federal agencies, universities, long-term ecological 

sites, state natural history surveys, and other natural heritage programs.
USDA Forest Service RMRS-GTR-285. 2012. 45

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48 USDA Forest Service RMRS-GTR-285. 2012.

Chapter 4

Plant Vulnerabilities and Genetic Adaptation

Bryce A. Richardson1, Nancy L. Shaw2, and Rosemary L. Pendleton3

1 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; Shrub 

Sciences Laboratory, Provo, Utah

2 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Aquatic Sciences Laboratory, Boise, Idaho

3 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Sciences Laboratory, Albuquerque, New Mexico

Executive Summary

The biogeography of plant species and population genetic structure within species 

is principally governed by climate. The association between climate change and plant 

distributions has been well documented since the last ice age, and recent studies have 

shown contemporary climate changes can create landscape-scale die-offs or movement 

of plant taxa. Terrestrial ecosystem conservation and restoration success hinges on 

understanding the vulnerabilities imposed by climate on plant taxa. Successful conservation and restoration under a changing climate will require:

• Recognizing whether climate is the root cause of changes in biotic communities.

• Determination of which species and populations are most vulnerable and at the highest risk to extirpation.

• Accurate prediction of future displacement and movement of plant communities.

• Fostering regeneration or assisting the movement of appropriately adapted plant 

materials.

Genetic and ecological research can provide critical components to meet these goals. 

GSD scientists are focusing on the following areas of research:

• Plant species-climate relationships: a means to understand how climate shapes distribution of species on the landscape. This research provides a means of predicting 

how species distributions could be affected by climate change.

• Adaptive genetic variation: research aimed at quantifying plant responses from different populations in a common environment. This research provides the necessary 

component to develop seed transfer guidelines for plant species.

• Genetic diversity and structure: use of molecular markers to identify areas of high 

or low diversity and how genetic variation is structured across the landscape. This 

research provides a means to identify physical or biological barriers to gene flow 

and at-risk populations with low genetic diversity.

• Ecological interactions: research aimed at identifying biological interactions critical to the success and persistence of native plants. Plant movement, either natural 

or human mediated, may require other organisms for pollination, seed dispersal, or 

seedling establishment.
USDA Forest Service RMRS-GTR-285. 2012. 49

Plant Species Distributions and Climate

Plant biogeography is principally governed by climate, exhibiting strong responses 

with both temperature and precipitation (Brown and Gibson 1983). The relationship 

between climate and plant species distributions has been demonstrated through several lines of research. First, contemporary predictions of plant species distributions have 

been shown to be highly accurate based on models using climate variables as predictors 

of species presence or absence (Iverson and Prasad 1998; Rehfeldt and others 2006; 

Friggens and others, Chapter 1 this volume). Second, past range shifts in plant distributions have been documented through patterns of genetic variation (e.g., Richardson and 

others 2002; Petit and others, 2003; Davis and others 2005; Richardson and Meyer 2012) 

and in records from packrat middens (e.g., Betancourt and others 1990; Thompson and 

Andersen 2000). These range shifts have been mainly attributable to climate change and 

associated glacial and interglacial oscillations during the Pleistocene and early Holocene. 

Third, studies monitoring vegetation have shown that range shifts are ongoing for a number of plant taxa and most are likely attributable to climate warming (Soja and others 

2007; Beckage and others 2008; Kelly and Goulden 2008; Thomas 2010). Given this 

close association between climate and plant biogeography, the predicted rapid change in 

climate by human-made greenhouse gas emissions should be the impetus for developing 

the knowledge base regarding seed transfer guidelines and other activities that mitigate 

this change.

In western North America, climate has been implicated as a factor in recent vegetation 

die-offs. The drought of 2002-2003 has been associated with the widespread mortality 

of pinyon (Pinus edulis) and juniper (Juniperus spp.) in the southwestern United States. 

Breshears and others (2005) showed that this drought coincided with warmer temperatures not seen in previous droughts during the Twentieth Century. Similarly, Rehfeldt 

and others (2009) used weather station data to show that the changes in climate variables 

important in predicting aspen (Populus tremuloides) distribution were also associated 

with stands of aspen die-off caused by sudden aspen decline. As with the pinyon die-off, 

drought and higher-than-average temperatures in 2002-2003 have been implicated as 

causative factors in sudden aspen decline. Based on predictions from GCMs, the aspen 

die-off may represent the trigger for a range shift, wherein many of these dead or dying 

stands will not recover (Rehfeldt and others 2009).

While die-offs of landscape dominant plant species, such as aspen and pinyon, may represent a fundamental change in ecosystem processes, climate change could also threaten 

other regionally distributed or endemic plant taxa with extinction. The varied topography 

and soils of western North America create isolated, discontinuous patches of habitat for 

plant specialization. Endemics are commonly found on exposed substrates such as shale 

or gypsum, or are associated with cliffs or shaded slopes of isolated mountain ranges 

(Johnston 1977; Meyer 1986; Sivinski and others 1996). These disjunct distributions limit 

the colonization pathways of potential habitat in future climates, especially in plant species with limited seed dispersal capabilities. For example, environments of river canyons 

create one source of endemism. These microclimates with perennial water sources create 

hanging gardens, supporting numerous endemic plants (Welsh 1989). The persistence of 

these microhabitats is dependent on sustaining ground water. If droughts in these regions 

become more common, the hanging gardens could dry up. Another example is provided 

by the isolated nature of mountain ranges in the southwest. These regions support some 

of the highest levels of plant endemism in North America (Warshall 1995; Anonymous 

2007). The Madrean Archipelago of southern Arizona, southwestern New Mexico, and 

adjoining states of Sonora and Chihuahua Mexico consists of some 40 isolated mountains 

known as “sky islands.” The effects of climate change may be particularly pronounced 
50 USDA Forest Service RMRS-GTR-285. 2012.

in these isolated ranges as vegetation zones shift upward and high-elevation habitat is 

lost. Bioclimatic modeling of Mexican spruce (Picea mexicana), which is isolated in 

small subalpine habitats, predicts the disappearance of contemporary habitat by 2030. 

However, re-emergence of suitable habitat for this species occurs 500 km to the south, a 

distance impossible for natural dispersal (Ledig and others 2010). Mexican spruce and 

other endemics will likely need human-assisted dispersal to areas of suitable habitat. To 

limit the vulnerabilities to climate-caused extirpation of plant taxa, bioclimatic analyses 

are of value in identifying new locations of suitable habitat that may emerge under climate change.

Biotic Interactions and Climate Change

In addition to direct effects, global drivers of climate change may affect plant distribution, abundance, and fitness through biotic interactions (Tylianakis and others 2008). 

Ecological disturbances creating large-scale plant mortality, such as insect and disease outbreaks, could be symptomatic of underlying plant stress due to climate change 

(Dale and others 2001). Temperature and moisture have been demonstrated to be critical components in the interaction between plants and insects or diseases, and climate 

change-caused stress can predispose plants to insect and disease outbreaks. However, 

determining whether climate change affects the intensity, geographic distribution, or 

longevity of an insect or disease outbreak is a complex task (Garrett and others 2006). 

Studies must take into account historical records of climate and outbreaks, spatial patterns of climate variables, host distributions, and other factors that affect the host and the 

disturbance agent interactions. Despite this complexity, some studies have shown a correlation between disturbance agent outbreaks and temperature or precipitation variability 

outside of the historical norms. Berg and others (2006) associated spruce beetle outbreaks 

in Alaska and the Yukon Territory with high summer temperatures. Likewise, Woods 

and others (2005) implicated climate change in an epidemic outbreak of Dothistroma 

needle blight. Widespread mortality of pinyon pine, previously mentioned, was further 

linked to outbreaks of the pinyon ips bark beetle (Ips confuses; Breshears and others 

2005). Drought-stressed trees were unable to produce sufficient resin to ward off beetle 

attacks. A fungus carried by the ips may also have been a factor (P.L. Ford, personal 

communication).

Disruption of mutualistic relationships, such as plant-pollinator interactions, may also 

occur due to climate change. Many flowering plant taxa require animal pollinators for 

reproduction (Brantley and Ford, Chapter 4 this volume). Corresponding declines in 

pollinators and insect-pollinated plants have been found in northern Europe, whereas 

wind-pollinated plants were unaffected (Biesmeijer and others 2006). Other studies have 

shown that climate change could disrupt plant-pollinator interactions by changing floral 

phenology (Memmott and others 2007). Rising CO2 levels and increased summer temperatures have been linked to changes in flowering phenology (Cleland and others 2007; 

Springer and Ward 2007; Crimmins and others 2009). Decoupling of plant flowering and 

pollinator availability could result in reproductive failure for both plant and pollinator. 

For example, flowering in ocotillo (Fouquieria splendens) is currently timed to synchronize with the northern migration of hummingbirds (Waser 1979). A lack of high-quality 

nectar resources could lead to a decline in hummingbird populations. Thus, mitigating 

the impacts of climate change on flowering plants that are dependent on pollinators has 

additional complexity and will require increased knowledge. Movement of plants to suitable habitats without recognizing the importance of pollinators could lead to failure. 

Other biotic interactions may also be affected, including inter- and intra-specific competition, herbivory, dispersal agents, mycorrhizae and other fungal mutualistic relationships 
USDA Forest Service RMRS-GTR-285. 2012. 51

(Tylianakis and others 2008). The effects of climate change on these interactions have 

received very little attention.

Climate Change and Altered Fire Regimes

Disturbance resulting from improper grazing practices, off-road vehicle use, and other 

anthropogenic disturbances have contributed to the widespread invasion of exotic annual 

grasses, primarily Bromus spp. in the Great Basin and, more recently, Bromus spp. and 

Schismus spp. in the Mojave and Sonoran Deserts (Esque and Schwalbe 2002; Ford and 

others, Chapter 6 this volume). Resulting changes in wildfire regimes (shorter fire intervals and longer fire seasons) and increasing temperatures have combined to accelerate 

the further spread of annual and perennial exotics, deplete native seed banks, simplify 

community structure and species associations, and reduce landscape patchiness (Brooks 

and Pyke 2001; Esque and Schwalbe 2002). Ecosystem resilience declines with disruption of critical functions such as snow or water catchment, nutrient cycling, and loss of 

microbiotic crusts and mycorrhizae. As a consequence, the future of entire communities 

and their component species are at risk due to the direct impacts of wildfire and invasives, climate change, habitat fragmentation, and resulting bottlenecks to plant migration 

(D’Antonio and Vitousek 1992).

The status of fire-intolerant sagebrush and its communities is threatened not only by 

wildfire and the incursion of exotic annuals, but also by the encroachment of native conifers, in part due to fire control and northerly movement of Mojave vegetation in response 

to warming temperatures (Bradley 2009, 2010; Ford and others, Chapter 6 this volume). 

Nielson and others (2005) simulated climate change impacts on potential future distribution of the sagebrush ecosystem. The greatest warming scenario reduced the system 

to 20 percent of its current area within the Twenty-First Century. Currently, about 350 

species of conservation concern are associated with the sagebrush ecosystem (Wisdom 

and others 2005) and 20 percent of the systems flora and fauna are considered imperiled 

(Center for Science, Economics and Environment 2002). Thus, major species losses can 

be expected if current trends continue.

Proposed research and management to meet these threats are: expanded research on 

biocontrol and other control methods for cheatgrass and other invasives (Runyun and others, Chapter 7 this volume); adaptation of native plants coexisting with invasives (Mealor 

and others 2004; Leger 2008), species specific seed zones (Erickson and others 2004); 

deployment of pooled seed sources, including accessions that will pre-adapt vegetation 

to expected changes in climatic conditions (Johnson and others 2010); management to 

reduce bottlenecks to species migration, and assisted succession (Friggens and others, 

Chapter 8 this volume).

Genetic Responses to Climate Change

Plant fitness, the ability to produce viable offspring, is often dependent on attunement 

to climate. Three types of biological responses from an organism can direct attunement: 

phenotypic plasticity, dispersal, and genetic change. These responses are fundamentally 

different biological processes occurring at different temporal scales, but all responses 

could potentially interact with each other and impact an organism’s capacity to genetically adapt to climate change (Davis and others 2005; Visser 2008; Reed and others 2010).

• Phenotypic plasticity is defined as the capacity of a particular genotype to produce 

varied phenotypes in response to different environments (Pigliucci 2001). Phenotypic 

plasticity operates within a generation, the shortest time scale of the three responses. 

It can be temporary (non-heritable) or inherited through some epigenetic mechanisms, 
52 USDA Forest Service RMRS-GTR-285. 2012.

the changes in gene expression by means other than DNA substitutions (Bird 2007). 

Epigenetic mechanisms could have important implications on how we assess plant 

vulnerabilities to climate change. For example, if a plant species possesses phenotypic 

plasticity in climatically adapted traits, this species may be more resilient under climate change.

• Dispersal of propagules (e.g., seed and pollen) can lead to gene flow. These dispersal processes can create a shift in gene frequencies and introduce novel genotypes 

from different populations, potentially affecting fitness. Dispersal occurs on a multigeneration temporal scale, yet the rate and distance propagules travel can have 

implications on plant species capacity to respond to a changing climate.

• Genetic change, the process of creating novel genes by mutation in the coding or regulatory DNA sequence that undergoes natural selection, occurs at the longest temporal 

scale. In this sense, a mutation could also include gene or whole genome duplications 

(i.e., polyploidy), which are common in some angiosperm lineages (Fawcett and Van 

de Peer 2010). These temporal scales are all dependent on the life history characteristics of the particular species. Species with short generation times will likely respond at 

a faster rate than those that have longer generation times (i.e., annuals are more likely 

to have higher fitness that perennials). Unfortunately, this benefits many invasive plant 

species that are annuals.

Genetic changes that affect the fitness of an organism can shape ecological processes, 

such as fecundity, mortality, and dispersal (reviewed in Carroll and others 2007; Kinnison 

and Hairston 2007). For example, a novel genotype that conveys fitness advantage to 

drought tolerance would increase in frequency in a climate that becomes arid, leading to 

higher survival, fecundity, and dispersal rates for individuals possessing this genotype. 

If arid conditions persist, strong selection pressure for drought tolerance could change 

the gene frequencies in a population over several generations. Conversely, if no drought 

tolerance genes exist, mortality and low fecundity could create a smaller population size. 

Genetic drift, stochastic changes in gene frequencies, could lead to the loss of genetic 

diversity, increasing the risk of extirpation of the population. A similar scenario was 

recently shown empirically in the flower phenology of field mustard (Brassica rapa). 

Franks and others (2007) demonstrated that flower phenology in this species responded 

to a five-year drought by earlier flowering and that the ancestors (i.e., pre-drought plants) 

had significantly reduced survival rates compared to contemporary plants.

Natural selection caused by climate is spatially dependent, especially for widely distributed species. The process can create variable plant trait responses or ecotypes across 

the landscape (Turesson 1925). Therefore, climatic selection has been demonstrated to 

be a major factor in intraspecific genetic adaptation (Langlet 1936; Savolainen and others 

2007). Much of this research, completed in common garden studies of trees species, has 

shown that the adaptive strategies among species are not the same. For example, Douglasfir (Pseudotsuga menziesii) is highly variable in growth phenology and populations have 

a narrow threshold for cold hardiness. The variability in this trait is mainly associated 

with winter temperatures and frost dates (St. Clair and others 2005). In contrast, western 

white pine (Pinus monticola) exhibits little adaptive variation to cold hardiness despite 

having a similar species distribution as Douglas-fir. In western white pine, cold hardiness 

is apparently a phenotypically plastic trait. However, for western white pine, variation in 

growth is strongly associated with the amount of growing season precipitation (Rehfeldt 

and others 1984; Richardson and others 2009). These genecological studies illustrate that 

ecologically similar plant species may have vastly different adaptive strategies to cope 

with climatic stresses and that populations within these species will respond differently 

under climate change.
USDA Forest Service RMRS-GTR-285. 2012. 53

Genetic Tools for Addressing Climate Change

Knowledge of intraspecific and population-level responses to climate change are 

essential for the restoration of plant species. Once associations between genetic traits 

and climate variables can be established, as previously discussed, genecological modeling can be used to define contemporary seed zones and predict future seeds zones 

using GCM scenarios (e.g., Rehfeldt and others 2002; Rehfeldt 2004; St. Clair and 

Howe 2007). While GCMs scenarios are not without assumptions (IPCC 2007), modeling intraspecific adaptive genetic variation in the future is instructive for identifying 

geographical regions of current and future on-site and off-site conservation and planning for the potential movement of desirable genotypes. However, most genecology 

research has been completed for forest trees and little knowledge exists for desert or 

species.

Molecular genetic approaches have served a different purpose to that of common 

garden trials of adaptive traits. Since molecular markers are generally neutral (i.e., not 

influenced directly by natural selection), they are valuable in assessing range-wide 

genetic diversity and structure. Measures of genetic diversity provide a means to assess the relative fitness of populations and the level of inbreeding (Reed and Frankham 

2003). Therefore, low levels of genetic diversity suggest the effective population size 

is or has been experiencing a bottleneck. Further losses to population size could result 

in leading to a greater risk of extirpation.

Molecular markers can also reflect past biogeographic distributions and demographic changes. Isolated populations from past glacial cycles are often inferred from 

patterns of genetic variation. Therefore, molecular markers are indirectly tied to past 

climate events. For example, organellar DNA markers were used to elucidate putative glacial refugia and post-glacial colonization in whitebark pine (Pinus albicaulis). 

Richardson and others (2002) showed that Holocene warming provided whitebark pine 

with opportunities to establish contact zones and genetic introgression from previously isolated populations. Other studies have shown that present-day genetic patterns 

are supported by bioclimatic modeling of predicted Pleistocene climate (Rebernig and 

others 2010). Much of the research using molecular markers on temperate plants show 

similar patterns of post-glacial colonization, such that the higher proportions of genetic diversity and structure are found in the lower latitudes of a species’ distribution 

(Hampe and Petit 2005). The skewed distribution of genetic diversity may make species more vulnerable to climate warming since lower latitude populations are generally 

at higher risk of extirpation.

Popular population genetic techniques such as microsatellites and amplified fragment 

length polymorphisms have been instrumental in identifying vulnerable plant taxa and 

populations by elucidating genetic diversity and structure. However, these techniques 

have limitations in the amount of detectable genetic variation and applications to other 

areas of genetic research. Next-generation sequencing technology has the capacity to 

sequence millions of 50 to 500 base pair fragments of genomic DNA or RNA that can 

uncover hundreds of thousands of single nucleotide polymorphisms (Mardis 2008). 

Researchers can develop these polymorphisms into thousands of molecular markers 

for discerning genetic relationships. The capabilities of next-generation sequencing 

have been compared to existing data generated from traditional sequencing methods 

in Fishers (Martes pennanti). Knaus and others (2011) discovered distinct populations 

of this rare carnivore in California that was unresolved in the previous genetic data. 

Such findings will a have an impact on how this species will be managed. In addition to providing more accuracy in discerning genetic relationships, next-generation 

sequencing offers greater versatility to address questions in plant-climate interactions. 
54 USDA Forest Service RMRS-GTR-285. 2012.

Two different molecular approaches have the capability to elucidate the relationships 

between genotype and phenotypes, bridging the gap between physiological/phenotypic 

plant responses and the underlying genotypic mechanisms.

• RNA sequencing: a process of transcriptome profiling (i.e., sequencing all expressed 

RNA transcripts in a sample) using next-generation sequencing technology. In this 

approach, RNA sequences are decoded and levels of gene expression can be estimated by transcript copy number (Wang and others 2009)

• Association genetics: an approach to identify a gene or genes involved with a phenotype by determining a pattern of presence and absence between molecular markers 

and a trait using groups of plants or plant populations with differing phenotypes 

(Neale and Savolainen 2004; Ingvarsson and Street 2010)

Research involving next-generation sequencing can improve our understanding of 

processes and mechanisms involved with creating variable traits and physiological 

responses important to climatic adaptation.

Potential Plant Vulnerabilities

Based on the available research and the previous discussion, inferences can be 

drawn about the general plant characteristics that may increase vulnerabilities to climate change. While the details on criteria for assessing and scoring vulnerabilities are 

discussed elsewhere (Friggens and others, Chapter 8 this volume), we provide a list of 

characteristics that are potentially influential to the success or failure of plant species 

under climate change.

• Habitat specialists: Many habitat specialists and narrow endemics could be at high 

risk to climatic extirpation. Examples include the hanging garden flora previously 

discussed. These plants species are limited in their movement because similar habitats are rare or nonexistent. For managers, long-term seed banking may be the only 

option for preservation of these species.

• Plant dependencies on other organisms: Plants that rely on other organisms for their 

reproduction (e.g., pollinators and seed dispersers) or survival (e.g., mycorrhizal 

fungi) could be more vulnerable to climate change (Brantley and Ford, Chapter 3 

this volume). Understanding the relationships of these plants and their associates is 

an important step to increase the chances for successful restoration or movement of 

plant species.

• Life history characteristics: Generation time, fecundity, and dispersal capabilities are 

important life history characteristics when considering vulnerability. Species with 

long generation times, low seedset, and limited dispersal capabilities are typically 

vulnerable to climate change. Unfortunately, these life histories also describe many 

of our nativet species (Runyon and others, Chapter 7 this volume).

• Adaptive genetic strategies: Plants have different adaptive genetic strategies. Plants 

that have more narrowly defined adapted populations will require more attention 

to the movement of seed to the appropriate seed zones. Therefore, plants with narrowly defined adapted populations are more likely to experience maladaptation with 

climate change.

• Genetic diversity: The extant genetic structure and diversity within a species can 

be a major influence on vulnerability to climate change. Genetic diversity affects 

a spectrum of adaptive responses a plant species can possess. Therefore, reduced 

diversity could impact fitness, making a species more vulnerable to climate change.
USDA Forest Service RMRS-GTR-285. 2012. 55

Research Needs

The impacts of climate change on the flora inhabiting grassland, shrublands, and 

desert ecosystems of western North America remain largely unknown. Accurate assessments of plant species’ potential vulnerabilities to climate change hinge on 

research needs spanning multiple disciplines to resolve their complexity (table 4-1). 

Understanding of these vulnerabilities can then be developed into plans and technologies for conservation and restoration (Friggens and others, Chapter 8 this volume). 

At the heart of these complex problems are several key principals: (1) recognizing 

changes in plant species and communities and whether or not climate is the root cause, 

(2) understanding species vulnerabilities under climate change, (3) accurate prediction 

of the movement of plant communities to plan for the future, and (4) mitigating these 

changes by fostering regeneration or assisting the dispersal of appropriately adapted 

plant materials.

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60 USDA Forest Service RMRS-GTR-285. 2012.

Chapter 5

Climate Change, Animal Species, and Habitats: 

Adaptation and Issues

Deborah M. Finch1, D. Max Smith1, Olivia LeDee 2, Jean-Luc E. Cartron3, and Mark A. Rumble4

1 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Sciences Laboratory, Albuquerque, New Mexico

2 Department of Forest and Wildlife Ecology, University of Wisconsin, Madison, Wisconsin

3 Dryland Institute, New Mexico Office, Albuquerque, New Mexico

4 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; Forest 

and Grassland Research Laboratory, Rapid City, South Dakota

Executive Summary

Because the rate of anthropogenic climate change exceeds the adaptive capacity 

of many animal and plant species, the scientific community anticipates negative consequences for ecosystems. Changes in climate have expanded, contracted, or shifted 

the climate niches of many species, often resulting in shifting geographic ranges. In 

the Great Basin, for example, projected increases in temperature could decrease ability of sagebrush to compete with warm-desert shrubs, leading to a shift in community 

composition.

Individualistic plant responses to climate change such as altered germination, flowering, and fruiting phenologies influence timing of reproduction and migration of 

animal species. Changes in timing of production of seeds, fruits, insects, and other 

prey items may result in a disconnect between timing of food availability and timing of 

migratory bird arrival or peak period of animal reproduction.

Increasing water scarcity, river and wetland drying, and further disruption of water 

flow regimes in the Interior West are likely to become overriding conservation issues, 

particularly given that riparian areas are known to have unusually high biological diversity and often harbor endangered species. Because of its dependence on ground 

and surface water, riparian vegetation is sensitive to hydrological effects of climate 

change. Changes in timing and magnitude of snowmelt have already been documented 

in western mountains. Water scarcity and river drying will negatively impact riparian 

habitats and wetlands, causing changes in vegetation composition and habitat structure 

and altering, in some cases eliminating, wildlife habitat.

Wildfire frequency is likely to increase due to changes in temperature and precipitation and invasion of combustible exotic species such as cheatgrass (Bromus tectorum). 

The interrelation among temperatures, moisture, biological invasions, and fire could 

trump direct impacts of climate change, leaving species and ecosystems with even less 

time to adapt.

Research priorities include the need to assess vulnerability of wildlife species, habitats, and ecosystems to climate change and to identify habitats and ecosystems where 

impacts are high and conservation actions are necessary. Improved understanding of 

climate change effects on habitats prone to fire, invasive species, insects, and diseases 
USDA Forest Service RMRS-GTR-285. 2012. 61

will enable habitat management planning and conservation. New methods and decision 

support tools are needed for protecting vulnerable populations and habitats and/or or 

assisting their migration in relation to shifting ecosystems.

We describe the current and potential future responses of terrestrial animal species 

and habitats to climate change in the Great Basin, Great Plains, and Southwest. With 

its extensive ecological, landscape-scale research background, RMRS and its partners 

and cooperators have an opportunity to play a leading role studying the complex interactions of climate change and other stressors affecting the ecosystems and species 

within these regions.

Introduction

Greenhouse gas emissions will continue to increase for at least the next few decades 

as fossil fuels remain the primary source of global energy (IPCC 2007a), resulting in 

further warming of our climate system. Depending on the emissions scenario, the IPCC 

projects an end of century increase in average global surface temperature of 1.8 °C B1 

(low emissions scenario) to 4.0 °C A1FI (high emissions scenario) (IPCC 2007a).

Because the rate of anthropogenic climate change exceeds the adaptive capacity of 

many animal and plant species, the scientific community anticipates “predominantly 

negative consequences” for ecosystems (IPCC 2007b).

For plants and animals, the impacts of climate change may be direct (e.g., water 

stress) or indirect (e.g., change in interspecific interactions). Species with a strong life 

history linkage to ambient conditions are susceptible to the direct impacts of climate 

change. For these species, altered weather-climate conditions are of concern: advanced 

timing of spring conditions, spatial shift in climate niche, heat stress, drought, and 

frequent, heavy precipitation events; for species that are specialists, dependent upon 

narrow habitat requirements or few interspecific interactions, the indirect impacts of 

climate change are of particular concern (LeDee and others 2011). These species are 

susceptible to changes in the availability and quality of habitat and the intensity and 

duration of interactions with other species. In turn, these changes may alter their behavior, distribution, development, reproduction, and/or survival.

We describe the current and potential future responses of terrestrial animal species 

and habitats to climate change in the Great Basin, Great Plains, and Southwest.

The Great Basin

Overview of Climatic Conditions and Habitats

In the Interior West, the Great Basin hydrographic region occupies most of Nevada, 

much of the western half of Utah, and a portion of southeastern Idaho. The physiographic region identified as the Great Basin Desert extends beyond these boundaries 

to include portions of Wyoming, Arizona, and New Mexico (Shreve 1942; MacMahon 

1979). Observed changes in climate and hydrology of the Great Basin include 5 to 10% 

decreases and increases in annual average precipitation in parts of Nevada and Idaho, 

5 to 20% increases in precipitation in Utah, and a 20-day advance in spring snowmelt 

date (U.S. GCRP 2009). IPCC climate models predict average temperature to increase 

by 1.7 to 3.3 oC in Nevada and Utah (EPA 1998a, 1998b). Following is a discussion 

of animal species vulnerability and adaptation to these changes in several Great Basin 

habitats.

Sagebrush uplands. The big sagebrush (Artemisia tridentata) species complex 

characterizes at least 45 percent of the land area in Great Basin (West 1983). Sagebrush 
62 USDA Forest Service RMRS-GTR-285. 2012.

shrublands provide nesting sites for breeding passerine birds, breeding and wintering habitat for diurnal raptors, foraging sites for pronghorn (Antilocapra americana), 

and year-round habitat for small mammal communities (Hanley and Hanley 1982; 

Ryser 1985; Katzner and Parker 1997; Bradford and others 1998; Wisdom and others 

2005). Wisdom and others (2005) identified 13 wildlife species as sagebrush-obligates 

or sagebrush associates. Two of these species, the greater sage-grouse (Centrocercus 

urophasianus) and pygmy rabbit (Brachylagus idahoensis), have been petitioned for 

Federal listing as threatened or endangered. Both feed on sagebrush during certain 

periods of the year and rely on plant or soil structure provided by sagebrush stands for 

nests and burrows (Wallestad and others 1975; Green and Flinders 1980; Katzner and 

Parker 1997; Gabler and others 2001; Himes and Drohan 2007).

Riparian zones. Riparian vegetation, found in areas where surface or groundwater is available to support phreatophytic species, covers a very small percentage of 

the Great Basin and can take many forms, such as woodlands, shrublands, and grasssedge meadows (Thomas and others 1979; Dobkin and others 1995, 1998). Numerous 

species of birds use western riparian vegetation for nesting, migrating, and wintering habitat (Knopf and others 1988). Riparian vegetation is particularly important for 

Neotropical migrant birds, many of which have declining populations (Ryser 1985; 

Ammon 2002; Arsenault 2002). Several riparian-obligate mammal species, such as 

beaver (Castor canadensis) and muskrat (Ondatra zibethicus), declined due to trapping or habitat loss in the Nineteenth Century, but populations have recently recovered 

in some areas (Thomas and others 1979; National Audubon Society 2000).

Wetlands and Lakes. Great Basin streams drain to inland bodies of water that form 

productive wildlife habitat. The relative scarcity of water in the region makes wetlands and lakes attractive to wildlife species, most notably migratory waterbirds (Ryser 

1985; Oring and Reed 1996). These bodies of water vary in depth, salinity, and presence of aquatic vegetation, but all play important roles in wildlife conservation. Many 

lakes and wetlands are identified as globally important breeding habitat for waterbirds, 

including the American white pelican (Pelecanus erythrorhynchos); waterfowl; and 

shorebirds, including species of concern such as the snowy plover (Charadrius alexandrines, Cornely 1982; Knopf and Kennedy 1980; Jehl 1986; Neel and Henry 1996; 

Warnock and other 1998; Manning and Paul 2003). Wetlands and lakes are used by 

impressive numbers of migratory waterbirds for stopover and staging areas (Boula 

1985; Jehl 1986; Oring and Reed 1996; Warnock and other 1998; Jehl and others 2002; 

Manning and Paul 2003) and are an important component of bald eagle (Haliaeetus 

leucocephalus) winter habitat (Isaacs and Anthony1987).

Key Future Issues Associated With Climate Change and Adaptation

Sagebrush-steppe/salt desert scrub. Sagebrush is more frost tolerant than woody 

species growing at lower elevations and latitudes in the arid west (Beatley 1975). 

Projected increase in temperatures could decrease ability of sagebrush to compete 

with warm-desert shrubs, leading to a shift in community composition. Under climate 

change projections, sagebrush shrublands are expected to recede in range as warmdesert vegetation encroaches from the north and woodlands encroach from higher 

elevations due to increases in precipitation (Neilson and others 2005; Rehfeldt and others 2006; Friggens and others, Chapter 1 this volume). In addition, wildfire frequency 

is likely to increase due to changes in temperature and precipitation and invasion of 

combustible exotic species such as cheatgrass. This change in fire regime facilitates 

a shift from native shrubland to exotic grassland (Knapp 1996; Chambers and others 

2007), which has already been shown to have negative consequences for shrubland 
USDA Forest Service RMRS-GTR-285. 2012. 63

reptiles, birds, and mammals (Reynolds 1979; Knick and Rotenberry 2002; Knick and 

others 2003; Wisdom and others 2005). In addition to fire, climate change has the 

potential to increase frequency of insect outbreaks, which could alter vegetation composition as well (Chambers 2008).

Little is known about potential for natural adaptation to climate change by Great 

Basin upland wildlife (Chambers 2008). What is certain, however, is that populations 

of sagebrush-obligate species, such as greater sage-grouse and pygmy rabbits, would 

not be capable of adapting to loss of sagebrush in their range because it provides essential food and cover, and would therefore be restricted to areas where sagebrush is likely 

to persist (Aldridge and others 2008). Sage-grouse also avoid areas where woodlands 

have encroached on shrublands (Atamian and others 2010; Doherty and others 2010). 

Humans could facilitate adaptation to climate change impacts by maintaining landscape connectivity to ensure that sagebrush-obligate species can move from unsuitable 

habitat and colonize available sites.

The impact of interactions between climate change and other disturbance agents can 

result in increased stress on ecosystems. For example, sagebrush ecosystems stressed 

by climate change in combination with cheatgrass invasions may be less resistant and 

more susceptible to fire with consequences for sagebrush-dependent species.

Riparian zones. Because of its dependence on ground and surface water, riparian 

vegetation is sensitive to hydrological effects of climate change. Changes in timing 

and magnitude of snowmelt have already been documented in western mountains 

(Cayan and others 2001; Mote and others 2005; Stewart and others 2005; Barnett and 

others 2008; Pierce and others 2008). These changes disrupt flood regimes and riparian plant reproduction in the Great Basin (Melack and others 1997; Webb and other 

2007). Along with changes in hydrology, increased temperatures and fire frequencies 

increase mortality of native riparian trees such as cottonwood (Populus spp.) and aspen 

(Populus tremuloides), which provide important foraging and nesting sites for birds 

(Dobkin and others 1995; van Mantgem and others 2009; Allen and others 2010; Liu 

and others 2010).

Riparian-nesting birds have been extirpated from Great Basin riparian zones following losses of native vegetation from grazing and streamflow regulation (National 

Audubon Society 2000; Wright and Chambers 2002; Webb and others 2007). In many 

cases, exotic vegetation, which is more tolerant of altered conditions, has replaced 

native species along watercourses (Knopf and others 1988). Though some species of 

migratory and breeding birds have successfully adapted to use exotic vegetation, others require native vegetation for nesting (Hunter and others 1987; Stoleson and Finch 

2001; Sogge and others 2008; van Riper and other 2008; Walker 2008). As the majority 

of research has focused on birds, information is clearly needed to determine adaptation 

potential of riparian amphibians, reptiles, and mammals to climate change in the Great 

Basin.

Wetlands and lakes. Water levels of Great Basin wetlands and lakes regularly fluctuate in response to temperature, precipitation, and water diversion (Neel and Henry 

1996; Manning and Paul 2003). These and other stressors can lead to reductions in 

aquatic prey for waterbirds (Sada and Vinyard 2002). Long-term changes in precipitation patterns would affect water levels, which would impact many migratory bird 

species (Ryser 1985). Rising temperatures and increasingly frequent droughts, predicted under climate change scenarios for the western United States (RMRS 2009), 

could increase salinity in many lakes, altering availability of invertebrate prey (Boula 

1985). Alternately, precipitation totals and storm intensity are expected to increase as 

well (Neilson and others 2005; Chambers 2008), which could decrease salinity and 

impact availability of invertebrate prey during critical periods for migratory birds. 
64 USDA Forest Service RMRS-GTR-285. 2012.

Animals using Great Basin lakes and wetlands have few opportunities for adaptation 

if bodies of water disappear or fluctuate beyond levels of optimal salinity, apart from 

changing migratory paths or breeding locations. Future research into natural or humanfacilitated adaptation is needed to conserve migratory bird populations in the face of 

climate change.

The Great Plains

Overview of Climatic Conditions and Habitats

Since the 1960s, average temperature in the Great Plains has increased 0.8 °C (U.S. 

GCRP 2009). Extreme high temperatures are more frequent and extreme low temperatures less frequent than in previous decades (DeGaetano and Allen 2002); in addition, 

heat waves (≥3 days extreme high temperatures) increased in frequency over the last 

50 years (DeGaetano and Allen 2002). Precipitation in the Great Plains, sensitive to 

global land surface-atmosphere-ocean interactions, is characteristically variable; this 

interannual variation is particularly high in spring and summer (Ruiz-Barradas and 

Nigam 2005; Ruiz-Barradas and Nigam 2010). The high natural variability in precipitation across years and seasons decreases the capacity to detect changes in this 

climate parameter. For example, in the Central Great Plains, after two decades of dry 

conditions (1961 to 1980), annual precipitation increased 12% in the subsequent two 

decades (Garbrect and others 2004). Although patterns in temperature and precipitation 

are of interest to natural resource managers, a climate condition of particular concern 

in the region is drought. Widespread, severe, multi-year droughts, such as those of 

the 1930s and 1950s, are a regular element of the Great Plains system (Woodhouse 

and Overpeck 1998; Schubert and others 2004). Altered wind patterns, an increase in 

drought affected areas, and sea-level rise are among a list of changes associated with 

the enhanced greenhouse gas effect (IPCC 2007a).

By mid-century, average temperatures in the Great Plains may increase 1.1 °C to 

3.3 °C; by the end of the century, global climate models project average temperatures 

to increase 1.4 °C to 7.2 °C (U.S. GCRP 2009). The projected increases in average 

temperature are greatest under scenarios with higher emissions of greenhouse gases 

(U.S. GCRP 2009). Climate projections indicate that the Northern Great Plains will 

experience the greatest warming in the region (U.S. GCRP 2009). In the Northern 

Great Plains, precipitation is likely to increase (5 to 20% lower emissions scenario; 5 to 

>40% higher emissions scenario); the Southern Great Plains may experience precipitation declines from 5 to 20% lower emissions scenario; 5 to >40% higher emissions 

scenarios (U.S. GCRP 2009). Higher maximum temperatures, high minimum temperatures, more hot summer days, and more heavy precipitation events are very likely by 

the end of the Twenty-First Century (Easterling and others 2000). Drought conditions 

are projected to increase in frequency, severity, and spatial extent (IPCC 2007b), including in the vulnerable Great Plains region (Woodhouse and Overpeck 1998). Again, 

the enhanced greenhouse gas effect is associated with a suite of climatic changes (i.e., 

increased storm intensity, fire, and flooding).

Habitats. The Great Plains are vast with topographic features ranging from steep 

mesas to rolling hills (Vinton and Collins 1997). The region is characterized by diverse 

terrestrial habitats. While shortgrass prairie is the dominant native vegetation in the 

arid western Great Plains, tallgrass prairie is dominant in the mesic, eastern region 

of the Great Plains (Vinton and Collins 1997). Cool-season grasses and forbs are the 

dominant native vegetation in the northern Great Plains (Vinton and Collins 1997).

These native grasslands host a diverse array of birds, mammals, reptiles, and invertebrates. In the northern Great Plains, thousands of shallow wetlands also dot the 
USDA Forest Service RMRS-GTR-285. 2012. 65

landscape; this swath of wetlands, called the Prairie Potholes, is an important source 

of stopover and breeding habitat for waterfowl. In the river drainages of the northern 

Great Plains, riparian woodlands provide important, structurally complex habitat for 

wildlife in the region (Rumble and Gobeille 2004, 1998). Although they cover less 

than 2% of the landscape, riparian woodlands host more than 50% of the animal species in the region (Theobald and others 2010).

Key Future Issues Associated With Climate Change and Adaptation

Climate-induced changes in vegetation will impact the distribution and abundance 

of animals on the landscape. For the Great Plains, projections of future vegetation do 

not indicate major declines in the total amount of grassland (see Friggens and others, Chapter 1 this volume), but the distribution may shift northwestward (Rehfeldt 

and others 2006). The projected increase in drought conditions will likely alter grassland composition and productivity, disturbance requirements, and erosion (Clark and 

others 2002). For example, woody vegetation in mixed and shortgrass prairie only 

occurs in areas of moisture compensation (Mack 1981; Girard and others 1989). These 

changes, particularly drought conditions, will likely exacerbate the major declines in 

grassland-dependent bird species (George and others 1992). Wetlands are highly sensitive to changes in hydrology and elevated temperatures; climate change alters the 

timing and availability of water resulting in changes in wetland species and habitat 

structure (Burkett and Kusler 2000; Winter 2000). For this reason, the Prairie Pothole 

region is considered highly vulnerable to future climate conditions (Johnson and others 

2010). The likely result is major declines in waterfowl production (Sorenson and others 

1998), particularly in the western Prairie Pothole region (Johnson and others 2010). In 

riparian systems of the western United States, reduced precipitation, early snowmelt, 

and higher temperatures will alter the timing and magnitude of stream flow; the result 

will likely be decreases in vegetation cover and particular species requiring moist soil 

conditions (Theobald and others 2010). In turn, this will lead to declines in the diverse 

animal species dependent on the riparian zone for reproduction and survival.

Earlier spring conditions (e.g., snowmelt, warm temperatures, and ice-out) cue some 

species to initiate migration and reproduction earlier in the year (Parmesan and Yohe 

2003). For example, in a study of 44 migratory bird species, researchers found that 

more than 50% arrive earlier to breed in the Northern Prairie region; the trend is correlated with local climate and is most notable for species associated with aquatic habitats 

(e.g., blue-winged teal [Anas discors], northern shoveler [Anas clypeata], and Wilson’s 

phalarope [Steganopus tricolor]) (Swanson and Palmer 2009). Large, rapid fluctuations in temperature are characteristic of spring; early arrival may expose migrants 

to more extreme cold snaps, low food availability, and other inhospitable conditions. 

Shifts in timing of resource availability in response to changes in rainfall patterns may 

also affect migrating birds. Altered snow conditions may also affect migrations of elk 

(Benkobi and others 2005) and deer and elk attendance by displaying sage-grouse 

(Bradbury and others 1989; Swanson 2009).

The geographic region where temperature and precipitation patterns are suitable 

for a species to persist is called its climate niche (Pearson and Dawson 2003). While a 

climatic niche describes “suitable” habitat, it does not mean that a species will not be 

able to inhabit an area that is considered climatically unsuitable in the future. Modeled 

shifts in a species climatic niche are not necessarily a perfect representation of the 

species future distribution. Changes in climate may expand, contract, or shift the climate niches of species (Parmesan and Yohe 2003). Under future climate conditions, 

bird species in the Great Plains, such as the Baird’s sparrow (Ammodramus bairdii), 
66 USDA Forest Service RMRS-GTR-285. 2012.

may experience considerable reductions in their potential distributional area (Peterson 

2003). Although some species may shift their distribution, agriculture and other landscape alterations will likely restrict these range shifts (Vos and others 2008).

For some species, high ambient temperatures may lead to heat stress or death. The 

regal fritillary (Speyeria idalia) is an endangered butterfly native to the mixed and 

tallgrass prairies of the central United States. Like most insects, it is sensitive to subtle, 

brief changes in temperature. High summer temperatures, especially in combination 

with intense solar radiation, may be lethal for eggs and larvae (Kopper and others 

2000). The greater prairie-chicken (Tympanuchus cupido) is “especially vulnerable” 

to climate change (North American Bird Conservation Initiative 2010) and heat stress 

may result in juvenile mortality (Flanders-Wanner and others 2004). Heat stress may 

also impact reproduction (Finch 1983) and survival in song birds.

Warmer temperatures and changes in precipitation will likely increase the frequency 

and severity of disease outbreaks (Harvell and others 2002). Type C and E botulism, a 

disease of increasing concern for waterfowl, is more prevalent is shallow, warm waters 

(Rocke and Samuel 1999). Under such conditions (projected for the Great Plains), 

there may be more frequent outbreaks that result in massive waterfowl mortality.

Water availability determines the distribution and abundance of all living organisms; few species tolerate dry conditions. The northern leopard frog (Rana pipiens), 

a species of conservation concern in the Great Plains and western United States, is in 

widespread decline across its range (Smith and Keinath 2007). The local extinction of 

the species in montane regions of Colorado is attributed to severe drought conditions 

in the mid-1970s (Corn and Fogelman 1984). Under future climate scenarios, drought 

conditions are projected to increase in frequency, severity, and spatial extent (IPCC 

2007a) with potentially large declines in amphibian populations (McMenamin and others 2008).

Flooding is a relatively common cause of reproductive failure in birds, particularly 

waterbirds (Burger 1982). In the Platte River area of the Great Plains, flooding in response to heavy precipitation is an established source of nest failure and chick mortality 

for the threatened Great Plains piping plover (Charadrius melodus) and endangered 

interior least tern (Sterna antillarum athalassos); in 1990, all surveyed nests for both 

species were inundated and resulted in extensive mortality (Sidle and others 1992). For 

populations already in decline, additional mortality is cause for management concern.

Although climate change and its associated impacts (e.g., conversion of land for 

biofuel production) will exacerbate declines in many animal populations and initiate major shifts in vegetation communities, some animal and plant species will likely 

benefit. Individual plants and animals may exhibit behavioral or phenotypic plasticity, 

altering their behavior or phenotype to minimize negative impacts or benefit from the 

novel conditions (e.g., Charmantier and others 2008; Kearney and others 2009). In 

addition, some animal and plant populations with high genetic diversity will exhibit 

flexibility, adapting to altered climatic conditions (Rice and Emery 2003). Given the 

varying responses of species, monitoring, research, and vulnerability assessments will 

be needed to prioritize management activities and allocate limited time and resources 

to meet conservation goals.

The American Southwest

Overview of Climatic Conditions and Habitats

The American Southwest is already experiencing pronounced climate change impacts that can be measured: 93% of New Mexico’s watersheds have experienced 

increasing annual trends in moisture stress during 1970 to 2006; snowpack in New 
USDA Forest Service RMRS-GTR-285. 2012. 67

Mexico’s major mountain ranges has declined over the past two decades in 98% of 

all sites analyzed; and the timing of peak streamflow from snowmelt in the state is an 

average of one week earlier than in the mid-Twentieth Century (Enquist and others 

2008).

The Southwest is characterized by general water scarcity compounded by cyclical 

droughts. It is also experiencing rapid human population growth resulting in increasing 

water use conflicts. As a result of all these factors, many southwestern natural communities are particularly at risk from global climate change (e.g., Hurd and Coonrod 

2007). Most climate models for the Southwest predict pronounced temperature increases, particularly in the summer, coupled with more frequent, more severe droughts 

and possibly an overall drier climate from reduced winter precipitation (Meehl and 

others 2007; Seager and others 2007). Water is expected to become increasingly scarce 

(in part from increased evaporative rates), with lower soil moisture, reduced base river 

flow, reduced river basin recharge (Serrat-Capdevilla and others 2007), and decreased 

reservoir levels (Milly and others 2008). Climate and hydrological models also show 

reduced snowpack and earlier snow melt, which, in turn, will further alter river flow 

regimes (Barnett and others 2008; Pierce and others 2008; Bonfils and others 2008; 

Enquist and others 2008).

The ecological importance of riparian areas and wetlands is disproportionate to their 

size on the southwestern landscape, as they concentrate much of the regional wildlife 

(e.g., Carothers and others 1974; Ohmart and Anderson 1982; Rosenberg and others 1982; Hunter and others 1988; Krueper 1996; Cartron and others 2000). These 

naturally fragile ecosystems have already been heavily impacted by river regulation, 

groundwater pumping, woodcutting, overgrazing, mining, and invasive non-native 

plants (e.g., Knopf and others 1988; Unitt 1987; Ohmart 1994; Cartron and others 

2000; Finch and Stoleson 2000). While some of the wildlife may increasingly rely on 

riparian areas, further shifting predation and competition equilibria, diminishing river 

flows, and river basin recharge may strongly threaten the regeneration of native riparian trees and other vegetation (Lytle and Merritt 2004), leading to large-scale loss and 

fragmentation of riparian and wetland ecosystems.

Key Future Issues Associated With Climate Change and Adaptation

Although the size of southwestern grasslands, shrublands, and deserts may not decrease over the long term (Rehfeldt and others 2006), their floristic composition may be 

drastically altered from individualistic species responses to climate change (Rehfeldt 

and others 2006) and from biological invasions occurring at an increasing pace (see 

Archer and Predick 2008).

Lizard species richness is unusually high in the southwestern United States, and 

consequently, many conservationists are alarmed about the potential for population extinctions in the Southwest because lizards comprise a group of species already showing 

severe impacts of climate change worldwide (Sinervo and others 2010). For instance, 

12% of local populations of 48 lizard species in Mexico have already disappeared because of warmer temperatures, and up to 20% could go extinct by 2080 if these trends 

continue. Surveys of 200 Mexican locations indicated the temperature has changed 

too rapidly for the lizards to keep pace (Sinervo and others 2010). Similar results were 

found at other locations around the world. By modeling extinction risk, the research 

team learned that climate change is occurring too rapidly for lizards to compensate 

with physiological adaptations to higher body temperatures.

Severe droughts, in combination with increased incidences of fire, are predicted to 

accelerate the pace of invasions by certain non-native plant species (e.g., tamarisk) 
68 USDA Forest Service RMRS-GTR-285. 2012.

in southwestern environments. Ecosystems that are not adapted to fire (e.g., riparian 

woodlands and the Sonoran Desert) are particularly at risk. Risk of conversion of some 

ecosystems to monocultures of non-native plants is high, with resulting impacts on 

associated wildlife.

Of particular concern in the Sonoran Desert has been the spread of buffelgrass 

(Pennisetum ciliare) and Lehmann lovegrass (Eragrostis lehmanniana), introducing 

wildfires into an ecosystem not adapted to them (Finch and others 2010). The spread 

of buffelgrass and Lehmann lovegrass, fostered by drought and a positive feedback 

response to fire, illustrates the potential for climate change, biological invasions, and 

fire to interact synergistically (e.g., Cox and Ruyle 1998). The potential consequences 

for invasions by these exotic species are changes in food resources and structure of 

wildlife habitats.

Individualistic responses to climate change and seasonal latitudinal and/or elevational movements of wildlife species have the potential to create temporally mismatched 

phenologies between flowering plants and pollinators (Memmott and others 2007), an 

important issue for hummingbirds and bats in the Southwest.

As water availability, plant species composition, and habitat types in the Southwest 

change in response to warming, there will likely be matching changes in accessibility 

of some ecosystem services as well as changes in conservation targets. In a dual context of water scarcity and human population growth, climate change impacts are likely 

to lead to increasingly frequent clashes between economic interests and conservation 

efforts. Conservation and restoration efforts to mitigate or minimize climate change 

along southwestern rivers will likely be increasingly limited due to higher priorities 

in water allocation, such as irrigation for agriculture. Water allocations and natural 

flows are critical for supporting aquatic and riparian habitats in the Southwest. Wildlife 

populations associated with floodplain riparian areas and wetlands are expected to become increasingly impacted by drier weather and associated habitat changes coupled 

with higher water demands from growing human populations (Finch and others 2011). 

Increased movements of wildlife populations seeking to adapt to changing conditions 

and scarcity of water may increase conflicts with humans in urban-wildlife interfaces 

or in recreational areas. Drought typically promotes reduced crops of seeds, berries, 

and other sources of food for wild animals, resulting in a higher frequency of conflicts 

between humans and wildlife (e.g., bears) as wildlife expand their search efforts by 

entering campgrounds and urban fringe areas.

Amphibian and fish species may be especially vulnerable to climate change because of their dependency on vulnerable aquatic ecosystems (Bagne and others 2011). 

Further loss, degradation, and fragmentation of riparian areas, particularly riparian 

woodlands, may not only affect breeding and wintering populations of many bird species but may also disrupt migration (loss of stopover habitat) and precipitate further 

population declines of species such as the endangered southwestern willow flycatcher 

(Empidonax traillii extimus), which requires moist habitats (Finch and Stoleson 2000), 

and yellow-billed cuckoo (Coccyzus americanus), which requires large patches of 

suitable riparian wooded habitat. Increased aridity of rivers and streams is known to 

foster the spread of tamarisk, an exotic woody plant that is replacing southwestern riparian habitats dominated by native cottonwoods and willows. Increasing aridity also 

promotes increased fire risk. Tamarisk rebounds more quickly after fire than native 

species, leading to habitat conversion to a tamarisk-dominated community (Smith and 

others 2009) with a different habitat structure that promotes a different wildlife species 

composition (Smith and others 2007).

Endemic and rare species and remnant faunas found in patchy habitats are prevalent 

in the southwestern United States. As air temperatures increase and water availability 
USDA Forest Service RMRS-GTR-285. 2012. 69

and soil moisture decrease, remnant and isolated habitats (e.g., the Madrean oak and 

Sky Islands of southeastern Arizona and southwestern New Mexico) are especially 

vulnerable to impacts. If local habitats wink out, sedentary animal species may be unable to travel to similar habitats elsewhere and risk local extinctions (Coe and others 

2012; Davison and others 2012).

Management Practices in the Interior West

How should we manage wildlife under a changing climate? The novelty of climate 

change means that we do not have tried-and-true adaptation strategies. Furthermore, 

because local conditions define what is appropriate and feasible, we should not expect 

a universal answer. However, the emerging consensus among scientists, managers, and 

stakeholders is that proper stewardship can maintain the resilience of natural systems 

to change (Halpin 1997). Although some argue this is a business as usual approach 

(Heller and Zavaleta 2009), such strategies will reduce existing stressors and increase 

the ability for animals and plants to cope with the direct and indirect impacts of climate 

change (table 5-1).

To manage wildlife and habitats faced with climate change, Federal, state, and private natural resources managers will need new information that projects where habitats 

and populations will change and will persist. Assessments of vulnerability to climate 

change at the species, habitat, ecosystem, and landscape levels will enable managers to 

identify target areas and species for conservation and management actions (Glick and 

others 2011; Bagne and others 2011).

Table 5-1. Best practices to maintain ecosystem resilience to climate impacts and examples of implementation in the Great Plains. 

Best practice Regional examples

Reduce existing threats The Conservation Reserve Program provides more than 10 million acres of suitable 

habitat for many grassland species in the region. The program provides habitat for 

species displaced by conversion to agriculture, the dominant regional land use.

Practice adaptive management Grazing in sensitive vegetation communities can compact soils and reduce regeneration 

of species like green ash or cottonwood. Managers can monitor effects of grazing and 

modify grazing management plans to ensure forage and community resources are not 

over utilized.

Evaluate land acquisition, including The U.S. Geological Survey North Dakota Gap Analysis Project is an example of 

connectivity habitat-based planning for species that are poorly represented in an existing reserve 

network.

Consider natural disturbance regimes Current grazing management guidelines include annual adjustments based on 

precipitation patterns. Research will be needed to predict the effects of increased 

reoccurring drought cycles on resource availability.

Build capacity via agency and university Landscape Conservation Cooperatives are building partnerships with regional agencies,

partnerships universities, and stakeholder to address climate change and other stressors in the region.

Engage the public in education and Youth environmental education programs provide a venue to engage the public

outreach activities on the topic of climate change. For example, the U.S. Forest Service sponsors the Kids 

in the Woods Program that provides conservation training. It also publishes a youth 

magazine called Natural Inquirer and a children’s series known as The Investigator. 

Environmental education can offer positive opportunities to individuals who want to 

engage on the climate change issue (e.g., developing urban wildlife habitat or pollinator 

gardens).
70 USDA Forest Service RMRS-GTR-285. 2012.

Top Research Priorities

• Assess vulnerability of wildlife species, habitats, and ecosystems to climate change 

in grasslands, shrublands, and deserts of the Interior West (Glick and others 2011; 

Bagne and others 2011).

• Identify “hot spots” of management concern where high concentrations of vulnerable species are likely to be impacted by climate change, and partner with managers 

to develop action plans.

• Develop methods for protecting vulnerable populations and habitats and/or or assisting their migration relative to shifting ecosystems.

• Evaluate and identify critical sagebrush, grassland, and desert habitats and corridors 

that are likely to persist in coming decades so these lands can be managed for obligate, endemic, and endangered species and for maintaining biological diversity.

• Improve understanding of climate change effects on habitats prone to fire, invasive 

species, insects, and diseases to enable habitat planning and conservation (Chambers 

2008).

• Determine whether and how wildlife species and communities can adapt to climateinduced changes in vegetation composition and prey availability.

• In riparian areas and other fragile ecosystems, evaluate changing wildlife species 

assemblages (including migration) and competition and predation pressures as a 

function of the interactions between climatic and hydrological conditions, spread of 

invasive plants, and fire frequency and intensity.

• Evaluate changing distributions and phenologies of hummingbird (and bat) nectar 

resources within the context of climate change and invasive species (top priority 

identified by Western Hummingbird Partnership 2010).

• Develop decision support tools, models, and website applications for conserving 

vulnerable wildlife and their habitats and biological diversity.

Strengths of RMRS Research Background for Proposed Research

The USDA Forest Service RMRS has developed a climate change strategy specifying priorities in research aimed at water, wildland fire, and terrestrial ecosystems 

(USDA FS RMRS 2010). In relation to this strategy, RMRS researchers are uniquely 

poised to address wildlife, habitat, and adaptation issues in response to climate change. 

Station scientists have developed a system for assessing vulnerability of species to 

climate change known as SAVS (Bagne and others 2011) that can be modified and 

adapted for broader application. Current and past research projects have focused on 

changes in habitat composition, distributions of invasive and native species, wildlife 

habitat use, population demographies, and populations of sensitive and endangered 

species under climate change scenarios. Station researchers in the Great Basin specialize in the development of native plant materials useful for habitat restoration and for 

sustaining critical plant species and ecosystems. Studies on the effects of fire, invasive 

species, and other disturbances on sagebrush, grassland, and riparian ecosystems comprise a major focal area of the RMRS GSD Ecosystems Program.

Partners, Collaborators, and Sponsors

Current and potential cooperators include Federal, state, and municipal agencies 

such as USDA Forest Service Regions 1, 2, 3, and 4 and western research stations, 

USGS Climate Change and Science Centers, Bureau of Land Management, U.S. Fish 
USDA Forest Service RMRS-GTR-285. 2012. 71

and Wildlife Service, Bureau of Reclamation, state wildlife management agencies, 

state conservation districts, and city open space divisions. Science and management 

collaborations with Department of the Interior Landscape Conservation Cooperatives 

have become an important RMRS focus. Developing research partnerships with 

USFWS Joint Ventures has potential to address climate change impacts on wildlife. 

Collaboration with Mexican researchers, governmental agencies, and non-Government 

organizations will serve several needs: (1) climate conditions and species assemblages in tomorrow’s Southwest may have their analogs in today’s Mexico; (2) some 

compounding threats to southwestern ecosystems may spread from Mexico (insect 

diseases; invasive plants originally from Africa and other tropical or subtropical regions), particularly those elevating the risk of fire; and (3) some conservation tools 

and strategies could be tested first in Mexico. RMRS is already working with The 

Nature Conservancy’s Southwest Climate Change Initiative to identify opportunities to address global climate change adaptation challenges in New Mexico, Arizona, 

Utah, and Colorado. Other non-Government partners where relationships have already 

been established include National Wildlife Federation, Sky Island Alliance, Western 

Hummingbird Partnership, Partners in Flight, Western Elk Foundation, Nature Serve, 

National Wild Turkey Foundation, National Wildlife Foundation, and some industry 

partners such as wind energy corporations.

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Chapter 6

Disturbance and Climate Change in the Interior West

Paulette L. Ford1, Jeanne K. Chambers2, Sharon J. Coe1,3, and Burton C. Pendleton1

1 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Sciences Laboratory, Albuquerque, New Mexico

2 U. S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Sciences Laboratory, Reno, Nevada

3 School of Renewable Natural Resources, University of Arizona, Tucson, Arizona

Executive Summary

Within the continental United States, average annual temperature increased during 

the Twentieth Century by approximately 0.65 oC. The most extreme warming occurred 

throughout the northern and western United States (IPCC 2007a; Williams and others 

2010). Disturbances such as fire, drought, grazing, urbanization, and energy development are predicted to have a heightened impact on the western United States under a 

changing climate. For example, extreme drought conditions accompanied by rising 

temperatures characterized the American Southwest during the past decade (Floyd and 

others 2009). Key future issues to consider are:

• Longer and more severe droughts.

• Longer periods of dry conditions leading to increased potential for “mega fires.”

• Increased ecosystem susceptibility to insect pests and disease.

• Increased susceptibility of ecosystems to invasion of non-native species.

• Changes in the relationships among plants, water, nutrients, and soils on grazed 

lands.

• Increased effects of grazing on greenhouse gas (GHG) levels.

• Increased conflicts over use of freshwater for native ecosystems and for growing 

human populations.

• Changes in precipitation runoff and potential for changes in flooding patterns.

• Increased concern over the potential for fire as human development extends the 

wildland-urban interface.

• Increased disturbance associated with domestic fuel exploration and production.

Some examples of research the RMRS GSD Program is conducting to address many 

of these key issues are:

• Studies of the response of Great Basin upland and riparian plant communities to 

climate change and anthropogenic disturbance and the implications for future responses to climate and management activities.

• Methods for maintaining and restoring riparian communities and their ecosystem 

services based on an understanding of the effects of climate change.

• Understanding the interacting effects of climate and fire on successional processes 

at landscape scales.
USDA Forest Service RMRS-GTR-285. 2012. 81

• Plant-climate relationships in the western United States and the effects of climate 

change on those relationships and on the landscape distribution of key species and 

communities.

• The interacting effects of climate change and livestock grazing on rangeland resources of the United States.

• The role of climate variation on historic fire regimes at local to sub-continental scales.

• The effects of climate and land use factors on successional dynamics and ecological threshold crossings in grassland ecosystems and the use of this information for 

state-and-transition models.

• The effects of climate on pinyon-juniper woodlands in western landscapes over the 

last 150 years.

However, there are gaps in our knowledge base. Some major research needs are:

• Improved methods for post-fire restoration of arid ecosystems and of ecosystems 

dominated by fire-adapted exotic grasses.

• Ecosystem recovery after fire, including effects of livestock grazing and re-seeding 

and re-planting of plant species.

• Relationships of weather and climate patterns to past and present fire regimes at local scales.

• Regional climate change predictions and improved capacity to predict ecological 

responses.

• Interactions among the primary climate change variables: precipitation regimes, elevated CO2

 and increased temperature.

• The most appropriate and least impactful sites for energy development.

• Appropriate plant materials for use in reclamation and restoration of disturbed sites 

under expected climate change scenarios.

• Effects of large solar generating facilities and wind-power developments on migration corridors, seed movement, and potential off-site effects of chemical dust control 

agents.

Our research sponsors, collaborators, and partners include: universities, The National 

Forest Service, Bureau of Land Management, Agricultural Research Service, Natural 

Resources Conservation Service, U.S. Geological Survey, Bureau of Indian Affairs, 

U.S. Fish and Wildlife Service, National Park Service, National Science Foundation, 

Long Term Ecological Research Network, state and municipal governments, tribal governments, Joint Fire Science Program, Department of Defense, Department of Energy, 

Bureau of Reclamation, non-Governmental organizations and other RMRS programs.

Introduction

We examine the interactions between climate change and natural and anthropogenic disturbances in grassland, shrubland, and desert ecosystems of the Interior West. 

Disturbances such as fire, drought, grazing, urbanization, and energy development are 

predicted to have a heightened impact on the western United States under a changing 

climate. We identify key issues related to these predictions, and provide examples of 

research that the RMRS GSD Ecosystems Program is conducting to address many of 

these predictions, as well as identify future research needs.
82 USDA Forest Service RMRS-GTR-285. 2012.

Interaction of Climate Change and Fire

Climate change is predicted to have multiple effects on fire regimes in grassland, 

shrubland and desert ecosystems. Current climate change models predict increased 

temperatures that will result in higher evaporative potential and heat stress across the 

western United States (Bates and others 2008). Predicted changes in precipitation are 

less certain, but it is likely that the southwestern United States and central Rockies will 

exhibit decreased annual precipitation.

Key Issues

Increases in spring and summer temperatures of ~0.9 oC and a one to four week earlier melting of snowpacks have been related to increased fire activity across the West 

(Westerling and others 2006). Since 1986, the length of the active wildfire season has 

increased by 78 days and the average burn duration of large fires has increased from 

7.5 to 37.1 days (Westerling and others 2006). Earlier snowmelt at high elevations and 

lack of snow retention at mid elevations in GSD ecosystems (Mote and others 2005) 

is resulting in longer periods of dry conditions and, when coupled with hot, dry, and 

windy weather, provides the conditions for wildfire.

The number of days of high fire danger (based on the energy release component 

index) is increasing and the Great Basin and the Southwest are among the regions most 

affected (Brown and others 2004). Extremely dry conditions result in more erratic 

fire behavior, with large flame lengths, torching, crowning, and blowups (Brown and 

others 2004). Mega fires, or large-scale fires that significantly exceed those of recent 

decades, are now occurring in GSD ecosystems as evidenced by the 264,260-ha burn 

in the Murphy Complex wildfire in Idaho and Nevada and 146,901-ha burn in the 

Milford Flat wildfire in Utah in 2007 (Chambers and Pellant 2008).

Relationships among climate and fire regimes differ among the Bailey’s ecoregions 

(Bailey 1995) due to a diversity of vegetation types and historic fire regimes.

• In most mountainous ecoregions, low precipitation, high temperatures, and a high 

Palmer’s Drought Severity Index in seasons prior to and including the fire season 

are associated with higher burn areas due to low fuel moisture, increased probability 

of ignition, and higher potential for fire spread (Westerling and others 2006; Littel 

and others 2009).

• In southwestern and arid ecoregions, precipitation in seasons prior to the fire season 

is more highly associated with burn area than warmer temperatures or drought the 

year of fire due to the importance of fine fuel production (Littel and others 2009).

• In the Great Basin, prior year relationships are even more important and burn area 

is associated solely with warm, wet conditions in winter or spring one or more 

years prior to the fire season. A warming climate with less snowpack and a higher 

proportion of rain at low to mid elevations in mountainous ecoregions may result in 

increases in area burned at these elevations due to decreased water availability and 

longer growing seasons (Littel and others 2009).

• In southwestern and arid ecoregions, increasing aridity may result in a decrease in 

area burned due to a reduction in fine fuels.

Current ecological conditions within different ecoregions influence the responses to 

climate-induced changes in fire regimes. In arid and semi-arid shrublands and deserts, 

invasion of exotic grasses with higher flammability and fire spread than native species 

is increasing both fire frequency and extent.
USDA Forest Service RMRS-GTR-285. 2012. 83

• In the Mojave and western Sonoran Deserts, invasion of the winter annual grasses 

Bromus madritensis subsp. rubens and Schismus barbatus is resulting in frequent 

and large fires in ecosystems that seldom burned in the past (Brooks and McPherson 

2006).

• In the eastern Sonoran Desert, invasion of exotic perennial grasses such as Eragrostis 

lehmanniana and Pennisetum ciliare are resulting in increased fire frequency and 

size (Brooks and McPherson 2008).

• In cold desert shrublands, invasion of exotic annual grasses, especially cheatgrass 

(Bromus tectorum), is increasing fine fuels and fire spread and resulting in conversion of diverse shrublands to near monocultures with greatly increased fire frequency 

and size (Knapp 1996).

• In many arid and semi-arid grassland ecosystems, expansion and dominance of 

woody species is resulting in significant decreases in fire frequency and lower resilience to fire (Van Auken 2000).

In these topographically diverse ecosystems, a warmer and drier climate will likely 

increase the susceptibility of higher-elevation and more mesic ecosystems to invasion 

by exotic grasses that are currently constrained to lower elevation systems, thus, increasing the risk of wildfire (Chambers and others 2007). Lower-elevation ecosystems 

may become less susceptible to invasion due to hotter and drier conditions that exceed 

the ecological amplitude of the invader. As a consequence, active restoration may be 

necessary (Bradley and others 2009).

Research Needs

• Relationships of weather and climate patterns to past and present fire regimes at local scales.

• Ecological factors and management strategies that increase ecological resistance to 

exotic species invasions.

• Ecological factors and fire and fuels management strategies that increase ecological 

resilience to fire at landscape scales and prevent threshold crossings.

• Effects of changes in fire regimes and fuels management on watershed function (erosion, water quality and quantity, and biotic organisms).

• Effects of fire and fuels management on riparian ecosystems, stream channels, and 

aquatic habitat.

• Improved methods for post-fire restoration of arid ecosystems and ecosystems dominated by fire-adapted exotic grasses.

• Effects of livestock grazing on ecosystem recovery after fire.

Effects of Drought

Recent climate change predictions indicate increasing periods of drought for the 

western United States (Bates and others 2008). Decreased annual precipitation in the 

southwestern United States and central Rockies coupled with increased evapotranspiration due to higher temperatures will increase the potential for longer and more severe 

droughts across most of the western United States. In addition, the Interior West is one 

of three areas where heat waves are predicted to become more frequent, more intense, 

and longer lasting (IPCC 2007b). Since 1950, the western United States has already 

experienced up to four more days per decade with extreme high temperatures.
84 USDA Forest Service RMRS-GTR-285. 2012.

Key Issues

Differential effects of climate change on GSD ecosystems may increase susceptibility of some ecosystems to the effects of drought and heat waves. For example, earlier 

snowmelt and a lower proportion of precipitation falling as snow will decrease the 

extent of snow cover and result in lower snow water equivalent at mid elevations in 

mountainous regions (Mote and others 2005). A decrease in snow water equivalent and 

increase in the amount of solar radiation absorbed on snow-free surfaces (Mote and 

others 2005) may exacerbate the effects of drought and heat waves on ecosystems at 

these elevations.

Increases in the severity and frequency of drought have the potential to significantly 

increase the level of stress experienced by native ecosystems. Drought typically results 

in reductions in plant primary production and water use, mortality of immature plants 

(seedlings) and animals, and increased susceptibility to insects or disease (Hanson and 

Weltzin 2000).

Severe or prolonged drought can eventually result in mortality of adult plants despite 

adaptations to arid conditions (Breshears and others 2005). Changes in biogeochemical 

cycling can occur, and reductions in plant cover can significantly increase soil erosion 

(Rapport and Whitford 1999).

Drought can increase the susceptibility of ecosystems to invasion of non-native 

species, especially under elevated CO2 conditions (Smith and others 2009). However, 

prolonged drought or drought that occurs at the margin of a species range can result in 

mortality of non-native invaders such as Bromus tectorum (Jeanne Chambers, personal 

observation). Drought also can alter fire regimes but effects differ among ecoregions 

(Westerling and others 2006; Littel and others 2009).

• In most mountainous ecoregions, drought conditions in seasons prior to and including the fire season are associated with larger burn areas.

• In southwestern and arid ecoregions, precipitation in seasons prior to the fire season 

is more highly associated with burn area than warmer temperatures or drought the 

year of fire due to the importance of fine fuel production (Littel and others 2009). 

However, prolonged drought can decrease production of fine fuels and may decrease the probability of fire.

The ecological conditions and anthropogenic stressors of GSD ecosystems can influence their responses to drought and heat waves. Many GSD ecosystems, including 

hot desert grasslands (Rapport and Whitford 1999), cold desert shrublands (Chambers 

and Wisdom 2009) and hot desert shrublands (Smith and others 2009), have undergone 

significant transformations as a result of overgrazing by livestock, agriculture, mining, 

energy production, urban, suburban and exurban development, recreation, and road development. A high proportion of these ecosystems have crossed ecological thresholds 

or are at risk of crossing thresholds to less desirable alternative states than existed prior 

to Anglo-American settlement. Severe or prolonged drought has the potential to result 

in additional threshold crossings and novel alternative states.

The potential for large-scale die-offs of key species and the consequences for 

ecosystems can be illustrated by recent, widespread mortality of Pinus edulis in the 

Southwest. In 2002 to 2003, 40 to 80% of P. edulis, a dominant, overstory tree species, 

suffered mortality in a four-state region of the Southwest after 15 months of soil water 

depletion (Breshears and others 2005). Mortality of the trees resulted from subcontinental drought and anomalously high temperatures. The proximal cause of mortality 

was apparent infestation by bark beetles, which is often related to drought-induced 

water stress (Allen and Breshears 1998). The die-off was not constrained to P. edulis; 
USDA Forest Service RMRS-GTR-285. 2012. 85

a dominant herbaceous species Boutelua gracilis exhibited a 50% reduction in live 

basal cover at one of the study sites. Interrelated effects included large changes in 

near-ground solar radiation (Martens and others 2000), runoff and erosion (Allen and 

Breshears 1998), and genetic structure of Pinus edulis on the landscape (Mitten and 

Duran 2004).

The potential for regional vegetation die-off under global-change-type drought remains a pivotal uncertainty in projections of climate change impacts (IPCC 2001). 

Regional mortality of key species has the potential to rapidly alter vegetation composition, associated ecosystem properties, and land surface conditions for decades.

Research Needs

• Regional climate change predictions and improved capacity to predict ecological 

responses.

• Experiments that examine the interactions of the primary climate change variables: 

precipitation regimes, elevated CO2, and increased temperature.

• Patterns of response of key species over environmental gradients.

• Interactive effects of soil moisture and temperature on competitive interactions.

• Effect of precipitation and drought on secondary succession and carbon sequestration.

• Role of drought in increasing/decreasing susceptibility to invasion by exotic plant 

species.

• Role/importance of precipitation seasonality, serial correlation, and extremes on ecosystem structure and function.

• Role of precipitation on nutrient cycling and feedbacks to plant production.

• The level of drought that kills key native perennial species.

Grazing and Climate Change

Rangelands are defined as lands where native vegetation is predominantly grasses, 

shrubs, or open woodlands, and includes grasslands, shrublands, savannas, and most 

desert, tundra, meadow, wetland, and riparian ecosystems (Kauffman and Pyke 2001). 

Rangelands provide important ecosystem services such as biological diversity, wildlife 

habitat, soil conservation, and GHG sequestration (Ritten and others 2010). Globally, 

livestock grazing is an important factor affecting ecosystem integrity as it occurs on 

more area than any other land use. It has been considered the most widespread influence 

on native ecosystems in western North America (Fleischner 1994). Grazing can reduce 

biological diversity and ecosystem integrity through removal of vegetation; trampling 

of soils, vegetation, and biological soil crusts; redistribution of nutrients; and dispersal 

of exotic plant species and pathogens (Kauffman and Pyke 2001). Many arid grasslands evolved in the presence of large mammalian herbivores (Brown and MacDonald 

1995), and thus the intensity of grazing is an important consideration when evaluating past, current, and future effects of grazing on plant species in grasslands (Loeser 

and others 2007). Also, these effects can be influenced by drought (Loeser and others 

2007); therefore, drought should be considered when evaluating the results of grazing 

studies. In addition, carefully managed grazing may provide opportunities for reducing the density of non-native plant species if such species are consumed by livestock 

(Kleppel and LaBarge 2011). In another example, targeted cattle grazing on rangeland 

dominated by cheatgrass reduced cheatgrass cover and, in turn, reduced rate of spread 

of prescribed fire in the northern Great Basin (Diamond and others 2009).
86 USDA Forest Service RMRS-GTR-285. 2012.

Key Issues

A variety of studies have addressed issues related to grazing under a changing climate, with topics ranging broadly from quantifying the impact of grazing on GHG 

levels (Wolf and others 2010) to evaluating potential direct impacts on animals from 

nutritional stress (Craine and others 2010). Relatively few studies have addressed the 

impacts of both grazing and climate change on native species of animals and plants. 

Research results are more readily available that address how grazed lands may change 

in their ability to support grazing as a livelihood in the future. Furthermore, there is a 

shortage of work aimed specifically at arid and semi-arid lands in the western United 

States. Some of the issues that are expected to face managers of public lands where 

grazing occurs are:

• Changes in the relationships among plants, water, nutrients, and soils on grazed 

lands are expected to occur under climate change.

• Determining appropriate stocking rates for maintenance of ecosystem function and 

long-term conservation of rangelands will become more difficult.

• Determining how grazing may affect GHG levels.

Hatfield and others (2008) identified several ecological processes that have the potential to be affected by climate change that relate to grazing: growing season length 

and plant phenology, net primary production, species composition, and nitrogen cycle 

feedbacks.

Effects are expected to vary by region. The spatial and temporal distribution of precipitation is a primary driver in the ecology of rangelands (Hatfield and others 2008). 

The water content of soil and the relationship of water to plants are influenced by levels 

of CO2 and ambient temperature. Warming is expected to lead to an earlier onset of 

green-up in spring and a longer growing season (Badeck and others 2004), but species 

are expected to vary in their responses to these phenomena. In some areas, modeling 

efforts have suggested that the spring growing season may shorten (Cullen and others 

2009). Furthermore, precipitation variability and increased CO2 may cause deviations 

from the overall patterns set by temperature.

Climate change and increased atmospheric CO2 concentrations are considered to 

have had an influence on the expansion of shrublands over the past couple hundred 

years (Hatfield and others 2008). Shrub encroachment into grasslands reduces forage 

for livestock and wildlife (Burkinshaw and Bork 2009). On the other hand, estimates 

of future suitable climate in western states covered in this review suggest that changes 

will favor an increase in area possessing conditions that would support semi-desert 

grasslands (Friggens and others, Chapter 1 this volume). The threat of soil erosion on 

grazed lands increases in periods of years with reduced precipitation (WashingtonAllen and others 2010). Herbivory by native herbivores can alter plant community 

responses to warming (Post and Pedersen 2008), yet experimental manipulations aimed 

at the potential consequences of climate change on rangelands often have not incorporated native herbivory.

Overall, nutrient and water cycles on grazed lands are complex. Del Grosso (2010) 

pointed out that the results from various studies demonstrate that “generalizations 

about how changes in climate and land management affect element cycling are often 

confounded.”

Concerns exist over how to make appropriate stocking decisions in the face of 

climate change (Ritten and others 2010; Tietjen and Jeltsh 2007). The long-term sustainability of rangelands depends on appropriate range management. Variable range 

production, caused in large part by stochastic precipitation, is already complex, with 
USDA Forest Service RMRS-GTR-285. 2012. 87

stocking decisions often made before growing season precipitation is known (Ritten 

and others 2010). Semi-arid and arid regions show high variability in mean annual 

rainfall (references in Tietjen and Jeltsh 2007). Rainfall directly influences primary 

productivity, which, in turn, significantly influences the carrying capacity of a given 

system for livestock. If net primary productivity is reduced, the carrying capacity is reduced and overgrazing is exacerbated (Tietjen and Jeltsh 2007). An increase in grazing 

pressure can lead to the reduction of palatable grasses and herbs and an increase in unpalatable grasses and herbs and woody plants (references in Tietjen and Jeltsh 2007).

Craine and others (2010) indicated that “predictions of how forage quality will be 

affected by changes in temperature and precipitation are varied and conflicting.” The 

increased threat of drought severity, frequency, and duration under climate change increases the complexity of making sound stocking rate decisions (Ritten and others 

2010; Popp and others 2009). Sustainable management of arid and semi-arid rangelands is expected to depend heavily on how well rangeland managers understand 

important ecological processes (Popp and others 2009).

Livestock production contributes to GHG emissions. On a global level, livestock 

are considered to produce 37% of the global anthropogenic methane (CH4) through 

digestive processes (Neely and others 2009), though others have put that value closer 

to 50% (Morrison 2009). Nitrous oxide, also a GHG, is released mostly as a side-effect 

of using nitrogen fertilizer or animal manure (Morrison 2009), although high levels of 

grazing are thought to increase nitrous oxide emissions because grazing increases the 

rate of nitrogen cycling (Del Grosso 2010).

Recent evidence suggests that grazing can reduce such emissions in semi-arid grasslands that experience soil freeze-thaw cycles in spring (Wolf and others 2010). Options 

exist in terms of more efficient use of grazing lands and manure to reduce GHG emissions (Gill and others 2010).

The value of rangeland soils in carbon sequestration has been identified as important on a national and an international level (Brown and others 2010; Follett and Reed 

2010; Neely and others 2009), particularly as degradation of rangeland reduces the 

ability of the soils to sequester carbon. Worldwide, grazing lands have been estimated 

as reducing approximately 20% of the CO2 released annually into the atmosphere from 

global deforestation and land use changes (Follett and Reed 2010).

Research Needs

• Continue efforts to improve predictions about future trends of vegetation dynamics 

under climate change and various management alternatives on grazed lands.

• Increase the scale at which field experiments are conducted (i.e., more large-scale 

studies) and ensure that rigorous experimental design is employed. Craine and others (2010) noted that climate change experiments are generally too small to allow 

grazing to be adequately characterized.

• More research on understanding the mechanisms driving observed changes in rangelands with respect to temperature, precipitation, soil cycles, etc. Nutrient cycles on 

rangelands (carbon, nitrogen, temperature, water, and soil) are critically important 

as they affect plant growth and quality of forage, which, in turn, influence stocking 

rates.

• Increase research focus beyond grasslands to grazing in deserts and shrublands.

• Incorporate interactions of invasive plant species and fire behavior in models and 

field studies.
88 USDA Forest Service RMRS-GTR-285. 2012.

• Develop decision-support tools to help managers evaluate how stocking rates may 

need to be adjusted under the threat of climate change.

Population Growth, Development, and Climate Change

The population of the United States continues to increase. In 2000, the population 

of the United States was 281 million people (U.S. Census Bureau 2000a), and between 

1990 and 2000, the U.S. population increased by 33 million people, the largest censusto-census increase recorded in the census history. Growth rates in the West (Montana, 

Wyoming, Colorado, New Mexico, and the seven states to the west) were higher than 

the national rate during this period at 20%.

Increasingly, human populations are concentrated in cities. Globally, almost all of 

the expected population growth is projected to occur in urban areas (Alig and others 

2004). In the United States, according to the 2000 census, 79% of the population lived 

in urban areas (defined as all “urbanized areas” of 50,000 or more people, and “urban 

clusters” of 2500 to 49,999 people; U.S. Census Bureau 2000b). In each of the 12 

states considered by this assessment, there was an increase in the percentage of the 

population that lived in urban areas between 1960 and 1990.

Population growth creates increased demands on natural resources, such as water 

and wood products, as well as energy products (coal, oil, natural gas, etc.). Clumping 

development into urban areas can reduce certain kinds of energy demands (e.g., housing and work locations may be closer) but can negatively impact agricultural output 

from the loss of farmlands and can lead to fragmentation of forests and other adjacent 

landscapes (Alig and others 2004).

Key Issues

There are a variety of ways in which climate change will interact with factors that arise 

from a growing population that increasingly will be concentrated in cities, including: 

changes in the availability of freshwater, changes in precipitation runoff and potential 

for changes in flooding, increased energy demands, and ex-urban development.

Increases in population place increasing demands on freshwater resources. While 

predictions vary as to future changes in the timing and total amount of precipitation according to model and region, the IPCC predicts an increase in precipitation in 

northern areas and a decrease in southern areas of the western half of North America 

(Christensen and others 2007). Thus, in southern areas, inputs of freshwater from precipitation are likely to decrease, and increasing demands for freshwater for human 

consumption (from increased population) will worsen existing impacts on native flora 

and fauna.

• Summer and autumn flows in river basins supplied by snowmelt water are predicted 

to be reduced (Kundzewicz and others 2008).

• Alterations to streamflow will differ by region due to variation in the degree to which 

precipitation and temperature change as well as variation in the characteristics of 

individual hydrologic basins (DeWalle and others 2000).

Changes to the timing of peak flows in spring are likely to alter freshwater availability for municipalities and exacerbate issues of water storage in reservoirs (Bartolino 

and Cunningham 2003). The High Plains Aquifer lies under portions of eight states and 

has supported irrigation of agricultural lands; in some areas, water levels have dropped 

greater than 30 m since pre-pumping periods (Bartolino and Cunningham 2003). In 
USDA Forest Service RMRS-GTR-285. 2012. 89

Tucson and Phoenix, groundwater levels have dropped 91 to 152 m in some places and 

riparian vegetation has been lost; similar problems exist in Las Vegas, Nevada.

An increase in population leads to an increase in the demand for energy. In 2006, 

the three top sources of energy in the United States were petroleum, natural gas, and 

coal (McDonald and others 2009). McDonald and others (2009) evaluated the landuse intensity of different types of energy production that are expected to increase in 

the United States based on the Energy Information Administration forecast for energy 

production by 2030, particularly if legislation passes to implement a system of capand-trade to regulate emissions:

• The type of energy development expected to impact the most total area (km2) is 

biofuel development. Biofuels and biomass burning of energy crops take up the 

most space per unit power (320 to 375 km2/TW hr/yr, and 433 to 654 km2/TW hr/

yr, respectively).

• Wind and solar power have intermediate values of space required per unit of power. 

Wind power was estimated to account for a relatively small percentage of the new 

land area impacted between 2006 and 2030.

• Nationally, future energy development is likely to impact temperate deciduous forests and temperate grasslands the most.

• Regardless of whether cap-and-trade policy is implemented, MacDonald and others 

(2009) estimated that by 2030, the total new area affected by energy production in 

the United States will be greater than 206,000 km2 (more than the area of Nebraska).

This loss of land to energy development will directly impact native species of flora 

and fauna. Furthermore, some forms of energy production require substantial amounts 

of water, which leads to additional indirect impacts.

Urban residents worldwide tend to produce less GHG than residents elsewhere in the 

same countries (Dodman 2009). In the United States in 2001, urban households drove 

less on average; for households with vehicles, urban households traveled on average 

~21,000 miles, whereas rural households traveled on average ~28,000 miles (U.S. EIA 

2005). However, in cities that experience a substantial heat island effect, residents 

may use more energy in summer to maintain cool temperatures indoors. Increasing 

development reduces natural areas that are likely to be critical as plant and animal 

communities change due to temperature and precipitation changes. The impacts of fire 

as human development extends into suburbs and the wildland urban interface need to 

be considered along with impacts from population growth on natural areas adjacent to 

large population centers.

Research Needs

• Continue research on regional-scale predictions of changes in precipitation under 

climate change.

• Identify streams, rivers, and associated riparian areas at greatest risk from altered 

streamflow and changes in water quality due to human-induced freshwater consumption and reduced precipitation.

• Identify policies that could be implemented to conserve water, either through reduced 

consumption and/or re-use by industrial and residential consumers.

• Identify regionally specific recommendations on locating new development to reduce 

energy consumptions and minimize impacts to natural areas.
90 USDA Forest Service RMRS-GTR-285. 2012.

Energy Development and Climate Change

There is an almost universal consensus among the international scientific community that GHG must be reduced in order to slow the rate of global warming (Abbasi and 

Abbasi 2000; Dincer 2000). As fossil fuel costs and the impact of their use have become 

more readily apparent, there has been increased industry investment and Government 

support for the development of renewable energy sources (Kunz and others 2007). 

Renewable alternative energy sources include nuclear power, biomass energy (Abbasi 

and Abbasi 2010), solar power, and wind power. Traditional oil, coal, and natural gas 

exploration and production occur within the boundaries of GSD and will continue for 

the foreseeable future.

In the western United States, the sprawl of energy development will have a potentially greater impact than in other areas of the country due to several factors, including 

the availability of large tracts of Government-owned land, high solar radiation, a 

fast-growing population, space available for new transmission corridors, and public 

acceptance of alternative energy production. Climate change velocities are also expected to be highest in low-elevation communities such as deserts, temperate grasslands, 

and shrublands (Loarie and others 2009). Development of alternative energy sources 

has far surpassed our understanding of their environmental impacts. It is important that 

we understand the potential environmental impacts of alternative energy development 

in order to mitigate undesirable effects.

Key Issues

Traditional energy production using fossil fuels can be expected to continue well 

into the Twenty-First Century. The geographic area covered by RMRS includes significant sources of these fuels as well as uranium deposits. Disturbances associated with 

fuel exploration and production include: drill pads, roads, pipeline corridors, open pits, 

mine tailings, and wind turbine sites.

While considerable research has already been focused on techniques for reclaiming 

these sites, the current goal of restoring historic communities may no longer be viable. 

Given the predicted ecotone and community shifts under climate change projections, 

large areas within the western states may no longer support present ecosystems (Saxon 

and others 2005), significantly complicating the selection of plant materials for restoration (Hufford and others 2010). Researchers must identify and develop new plant 

materials that better tolerate increasing drought and high temperatures and are competitive with invasive species (Hufford and others 2010).

Energy production from biomass, termed “biofuels,” has been touted as a way to 

reduce reliance on fossil fuels and reduce GHG emissions. Biofuels are considered 

carbon neutral because they use recently fixed carbon as opposed to the ancient carbon 

of fossil fuels (Abbasi and Abbasi 2010). Key considerations in the conversion of large 

acreages to biofuel production are:

• Biomass production is not “nutrient neutral,” and large production may not be sustainable due to soil degradation and depletion of nutrients (Abbasi and Abbasi 2010).

• Use of crop plants and agricultural lands traditionally used for food production for 

biofuels competes with feed and food demands, increasing world prices and decreasing supplies of foodstuffs.

• Alternatives include the use of non-food energy crops, such as crop residues, hydrocarbonrich plants, organic waste products, weeds, and fast-growing grasses and woody 

species (Schmer and others 2008; Abbasi and Abbasi 2010).
USDA Forest Service RMRS-GTR-285. 2012. 91

• Additional mitigation may come from use of degraded rangeland and abandoned 

cropland for biofuel production rather than diverting existing agricultural lands 

needed for food production (Fargione and others 2010).

Construction of alternative energy-generating facilities is a high priority for utility 

companies, state and Federal governments, and the public. The arid Southwest contains 

a large amount of Federal lands that also receive significant amounts of solar radiation, and the majority of large-scale solar facilities are slated for development on these 

lands. Six new solar facilities were approved on Federal land in Nevada and California 

in October 2010 (USA Today 2010). The largest is 2830 ha (greater than 10 square 

miles) and larger facilities are being considered (USGS 2010). The solar facilities are 

approved for construction on the fast track, though extensive environmental assessment will not be completed prior to breaking ground (USGS 2010). While new solar 

facilities are significantly more efficient than traditional power plants and produce far 

fewer CO2 emissions, they are not without environmental impacts in their construction 

and operation. Limited experience with large-scale solar facilities means that there is 

little on-the-ground experience regarding these potential impacts (Tsoutsos and others 

2005; McDonald and others 2009). The U.S. Department of the Interior has identified 

24 solar energy zones in six western states that have high potential for solar energy 

development and have limited environmental impacts. The zones will facilitate more 

rapid solar energy development. The final number and size of the solar energy zones 

will be determined in 2012 (BLM 2011). Some issues to consider are:

• Facilities must be kept free of vegetation in order to prevent the possibility of heatinduced wildfires, resulting in large areas devoid of vegetation.

• Dust emissions from construction, and resulting un-vegetated areas and roads will 

likely increase. Little is known regarding potential off-site movement of chemical 

dust-control agents or their effect on surrounding vegetation.

• The impact of solar energy development on wildlife has not been well researched. 

Effects on migration corridors as well as effects on resident species are largely unknown (USGS 2010).

• Service roads may serve as transmission corridors for exotic plant species.

• Attention during the planning, construction, and operation phases could minimize 

potential environmental impacts (Tsoutsos and others 2005).

Wind energy is increasingly important as a means of alternative energy production. 

As with solar installations, new and much larger wind farms are planned or under 

development. Wind-generating turbines are installed in linear arrays on ridge tops, in 

canyons, along coastlines, and offshore. Smaller-scale development occurs in more 

urban settings. The technology to develop energy from wind on a large scale is more 

mature than that of large-scale solar development. Several studies have been conducted 

on wildlife impacts, but questions still remain.

• Habitat types most likely affected by wind energy production are temperate conifer 

forests and temperate grasslands (McDonald and others 2009).

• As with fossil fuel production, impacts on wildlife from wind turbines will largely 

come from habitat fragmentation, species avoidance behavior, and bird and bat mortality rather than from direct effects from installation (McDonald and others 2009).

• Our understanding of potential impacts of proposed large-scale wind farms, especially offshore, is severely lacking (Drewitt and Langston 2006).

• Recent monitoring studies indicate large numbers of bat fatalities associated with 

wind energy facilities (Kunz and others 2007).
92 USDA Forest Service RMRS-GTR-285. 2012.

• Potential effects on birds include collision mortality, displacement due to disturbance, altered migration flyways or local flight paths, and habitat change or loss 

(Drewitt and Langston 2006).

Research Needs

• Background research to help managers and companies select the most appropriate 

and least impactful sites for energy development.

• Rigorous, long-term studies of the environmental impacts of large-scale power development, including pre-construction assessments where possible.

• Development of appropriate plant materials for use in reclamation and restoration of 

disturbed sites under expected climate change scenarios.

• Identification of biomass materials suitable for use in small-scale rural applications 

as well as development of high-yield, non-food energy crops.

• Research to examine the impact of large, solar generating facilities and wind farms 

on migration corridors and seed movement and potential off-site effects of chemical 

dust control agents.

• Multi-year monitoring and hypothesis-based research on causes of bat and bird mortality (Kunz and others 2007), particularly in the Southwest where data are lacking.

• Development of education programs to teach citizens the economic and ecological 

value of energy efficiency and conservation as a step in reducing energy development sprawl in the western United States.

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USDA Forest Service RMRS-GTR-285. 2012. 97

Chapter 7

Invasive Species and Climate Change

Justin B. Runyon1, Jack L. Butler2, Megan M. Friggens3, Susan E. Meyer4, and Sharlene E. Sing1

1 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Sciences Laboratory, Bozeman, Montana

2 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; Forest 

and Grassland Research Laboratory, Rapid City, South Dakota

3 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Sciences Laboratory, Albuquerque, New Mexico

4 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; Shrub 

Sciences Laboratory, Provo, Utah

Executive Summary

Invasive species present one of the greatest threats to the health and sustainability 

of ecosystems worldwide. Invasive plants, animals, and diseases are known to have 

significant negative effects on biological diversity and the ecological structure and 

functions of native ecosystems. Moreover, the economic cost imposed by invasive 

species is enormous—the damage inflicted to natural resources and costs of control 

measures is an estimated $137 billion each year in the United States. Climate change 

can fundamentally alter the behavior, spread, and harm caused by invasive species and 

the effectiveness of control methods. If we are to keep pace with and effectively limit 

the spread and damage caused by invasive species, it is critical to understand and predict how climate change will affect species invasions and the efficacy of the tools used 

to combat these invasions. To better identify research needs, we review the current 

state of knowledge pertaining to climate change impacts on several key topics, including invasive plants, their biocontrol, and wildlife disease.

Climate change is expected to alter the distribution and spread of invasive plants 

but in largely unknown ways. The climate models used to predict future distributions 

of invasive plants (e.g., cheatgrass) are limited by a lack of knowledge about, and consideration of, the ecology, genetics, etc., of the plant. Moreover, climate change may 

favor and convert non-native species considered benign today into the noxious weeds 

of tomorrow, but we are unable to predict which species might be favored. There is 

evidence that the physiology and competitive ability of invasive plants will be favored 

more than native plants, particularly in arid ecosystems, but in complex ways that are 

poorly understood.

Biological control—one of few tools proven effective against widespread invasive 

plants—will be affected by climate change in a number of ways. Like invasive plants, 

the range and spread of biocontrol insects are likely to be altered. Of particular concern 

is the potential for climate change to disrupt the temporal or spatial synchrony between 

biocontrol agents and their invasive host plants. There is also evidence that climateinduced changes in plant chemistry will alter plant-insect interactions in important 

ways (e.g., how much an insect eats) and could ultimately affect population levels of 

insects and invasive plants. However, we know next to nothing about the potential 
98 USDA Forest Service RMRS-GTR-285. 2012.

consequences of these predicted changes for successful use of this powerful management tool.

Climate change can alter wildlife disease dynamics with potentially severe consequences for the affected species and entire ecosystems. For example, increased 

temperatures and altered precipitation patterns can increase the range and abundance 

of vector species (e.g., mosquitoes and ticks) and thus the frequency of vector borne 

disease outbreaks, including West Nile Virus. Climate-induced changes can further 

promote disease by affecting host susceptibility to infection. However, climate change 

could limit the spread of some pathogens. Research is needed to understand how 

climate change will impact disease emergence and spread to prioritize and inform 

management actions.

Climate change will modify invasive species and the tools used to manage them. 

Our understanding of how and in what direction climate change will drive such changes is insufficient to adequately predict and respond. However, climate-induced changes 

are likely to be complex and will need to be examined on a case by case basis until 

more generalized frameworks can be developed. This review will guide development 

of important research questions, the answers to which will better position us to devise 

and apply meaningful management options to address invasive species in both present 

and future climates.

Introduction

Biological invasions threaten the integrity of many ecosystems and are considered 

second only to habitat destruction in their effects on biodiversity and on landscapes 

as a whole. For example, invasive plants currently infest approximately 100 million 

acres of land in the United States and are spreading at the rate of several million acres 

per year. However, climate change has the potential to greatly alter the behavior of 

invaders and their interactions with other organisms, with important consequences for 

invaders’ management. A better understanding of what these climate-induced changes 

will be is critical to adapt, develop, and successfully apply management strategies to 

control invasive species.

Invasive species are defined here as any native or non-native species that causes or is 

likely to cause social, economic, or ecological harm. This chapter is divided into three 

sections: the first covers climate change effects on invasive plants and their competitive 

interactions with other plants. The second deals with potential impacts on biocontrol 

of invasive plants, including direct effects on biocontrol agents and their interactions 

with invasive plants. The third summarizes climate change effects on wildlife disease.

Climate Change Effects on Invasive Plants

Non-native plants are now a common theme of many ecosystems throughout the 

United States (Pimentel and others 2000), including places established specifically to 

protect native species and communities (Allen and others 2009). Although the exact 

number of plants introduced into the United States is unknown, published estimates 

range from 4000 (Stein and Flack 1996) to 5000 species (Morse and others 1995). 

In comparison, there are approximately 17,000 species of native plants in the United 

States (Stein and Flack 1997). At present, only a small proportion of the 4000 to 5000 

non-native species are classified as invasive, adversely impacting native species, communities, and ecosystems (Hiebert 1997; Skinner and others 2000). The overall impact 
USDA Forest Service RMRS-GTR-285. 2012. 99

of existing invasive species may increase or decrease under several scenarios of global 

change driven by greenhouse-gas influenced climate change, increasing carbon dioxide (CO2

) concentration, increasing nitrogen (N) deposition, and altered disturbance 

regimes (Dukes and Mooney 1999; Bradley and others 2010). Furthermore, the larger proportion of non-native species considered benign and maintained artificially as 

ornamentals represent a substantial pool of potentially damaging species should environmental conditions shift in their favor as a result of one or more of the elements of 

global change (Sutherst 2000).

Although alterations of climatic patterns will undoubtedly exert a major influence 

on the distribution of invasive plants (Blumenthal and others 2008; Hellmann and others 2008), studies that report the responses of invasive plants to climate change in an 

ecosystem context are limited (Dukes and Mooney 1999). However, Bradley and colleagues (Bradley 2009; Bradley and others 2009) have examined the effects of variable 

future climate scenarios on cheatgrass (Bromus tectorum L.) for the interior western 

United States using bioclimatic envelope modeling approaches based on AtmosphereOcean General Circulation Models (AOCGM). Cheatgrass is an invasive winter annual 

brome species that, along with red brome (Bromus rubens L. or Bromus madritensis

ssp. rubens [L.]), dominate millions of hectares of former shrublands in semi-arid and 

arid habitats throughout the Interior West (Brooks and others 2004; Chambers and others 2007). Both species facilitate their own increase and spread through their effect on 

fire return frequency (Bromus spp. invasions provide abundant, highly flammable fine 

fuel); cycles of frequent burning can create near-monocultures of these grasses over 

large areas (Mack and D’Antonio 1998; Brooks and others 2004). Using the climate of 

the area currently occupied as a guide, Bradley (2009) and Bradley and others (2009) 

identified major variables that predict cheatgrass presence: annual precipitation, summer precipitation, and spring precipitation, with lesser effects of winter temperature. 

Modeling potential distribution given current climatic conditions indicated that sizeable areas (Wyoming, for example) may become available as climatically suitable 

habitat. AOCGM models generally predict that the Great Basin region will experience 

both reduced total precipitation and reduced spring precipitation, but the models vary 

widely in their predictions for summer precipitation. Consequently, the outcome of 

bioclimatic envelope modeling of future cheatgrass distribution is very much dependent on the AOCGM model used; both major increases (45%) and major decreases 

(70%) in future climatically suitable area have been predicted (Bradley 2009). The 

high uncertainty of these largely precipitation-based predictions is, in part, due to the 

inherent difficulty in projecting future precipitation (compared to temperature).

Bioclimatic envelope modeling is a valuable tool for predicting species response 

to climate change, but these models assume that species distribution is static under 

a given set of climatic conditions (see Friggens and others, Chapter 1 this volume). 

Bradley and colleagues (2009) used a presence-only approach that makes no a priori 

assumptions about the climate unsuitability of areas not known to be occupied, and 

their model does not take into account the fact that cheatgrass is known to be rapidly expanding its range into novel habitats. These researchers have documented range 

expansion into montane, extreme salt desert, and warm desert environments, many 

of which are likely outside the defined bioclimatic envelope (e.g., Ramakrishnan and 

others 2006; Leger and others 2009; Scott and others 2010). Furthermore, the bioclimatic envelope model as used by Bradley and others (2009) does not take into account 

another very important fact, namely that not all cheatgrass is created equal. Research 

has confirmed that the invasion of novel habitats involves unique biotypes (inbreeding 

lines) of cheatgrass that have suites of adaptive traits that preadapt them to specific 

environments (Ramakrishnan and others 2004, 2006; Merrill and others, in review). 
100 USDA Forest Service RMRS-GTR-285. 2012.

This genetic variability, combined with highly dispersible seeds, creates the possibility 

that biotypes preadapted to areas with changed climates could rapidly replace cheatgrass biotypes that have become locally maladapted. This shift would be transparent to 

a casual observer, as there are few or no morphological markers for identifying these 

unique biotypes. In essence, these shifts involve evolutionary change in response to 

climate change, a possibility that is rarely considered and difficult to include in modeling scenarios.

At the plant level, increased atmospheric CO2 concentration has a positive impact 

on plant photosynthesis and growth, although the relative response varies considerably among species (Hunt and others 1991). Previous research indicates relatively 

strong responses of invasive plants to elevated CO2 in competition-free environments 

(Ziska 2003; Dukes 2000), and studies are available that demonstrate higher rates of 

photosynthesis for invasive plant species when compared to their native counterparts 

(Pattison and others 1998; Baruch and Goldstein 1999; Durand and Goldstein 2001). 

However, Dukes (2000) concluded after an extensive review that, while a large number of invasive species respond strongly to elevated CO2

, the overall responses were 

not statistically different from the responses of non-invasive plants within the same 

functional group.

Studies that experimentally compare the physiological responses of invasive plants 

to elevated CO2 to their native constituents are limited (Sasek and Strain 1991; Dijkstra 

and others 2010a; Song and others 2009). In an environmentally controlled experiment 

involving monocultures of three invasive species and three co-occurring native species, 

Song and others (2009) reported that elevated CO2 resulted in significantly higher photosynthetic rates and increased biomass for the invasive species compared to the native 

species tested. Similarly, Sasek and Strain (1991) found that Japanese honeysuckle

(Lonicera japonica), a non-native species, produced significantly more biomass under 

CO2 enrichment than did coral honeysuckle (Lonicera sempervirens), a related native 

species. These authors suggested that this may convey an advantage for invasive species under increasing atmospheric CO2

, although such responses may differ in diverse, 

competitive environments (Bazzaz and McConnaughay 1992). For example, Dijkstra 

and others (2010) evaluated the effects of elevated CO2 on a mix of five native semiarid grassland species and one invasive plant species under greenhouse conditions. 

They found that species’ responses to elevated CO2 and supplemental water depended 

on whether the species were grown in a monoculture or in a mixture. Specifically, 

Linaria dalmatica, the invasive species used in that study, responded positively to elevated CO2

 when grown in a monoculture but negatively when grown in mixtures that 

included the five native species.

In mixed species competition, the response of a functional group, including invasive 

plants, to elevated CO2 may depend upon how CO2 indirectly alters competition for 

other resources (Bazzaz and McConnaughay 1992; Patterson 1995). Increased CO2 

also decreases transpirational water loss, which can improve season-long soil moisture 

conditions, and invasive species that can take advantage of the additional moisture 

may become more abundant (Dukes and Mooney 1999). Elevated CO2-induced increases in water use efficiency are predicted to be greater in arid ecosystems because 

of the increased availability of a very limited resource (Smith and others 2000; Morgan 

and others 2004). The direct effect of atmospheric CO2 enrichment on annual bromes 

(cheatgrass and red brome) has been investigated in both greenhouse and free air carbon enrichment (FACE) studies (Smith and others 1987, 2000). The general conclusion 

from these studies is that annual bromes have a larger positive growth response to CO2 

enrichment than associated native herbaceous species, which should increase annual 

brome competitive advantage as atmospheric CO2 levels increase.
USDA Forest Service RMRS-GTR-285. 2012. 101

Elevated CO2-induced increases in soil moisture can enhance plant N uptake, which 

is important for sustained increases in plant productivity associated with increased 

atmospheric CO2 (Dijkstra and others 2008, 2010a). However, Dijkstra and others 

(2010b) reported that N availability decreased with increasing CO2-induced soil moisture, while warming increased soil inorganic N and plant N uptake.

Anthropogenic-caused increases in N deposition is a component of global change 

that receives limited attention with respect to its potential impact on non-forested ecosystems (Vitousek 1994) where slow-growing native species adapted to low nutrient 

poor soils are likely to be the most impacted by increasing N (Milchunas and Laurenroth 

1995; Tilman 1987). Fertilization experiments in N-limited ecosystems demonstrate 

that increased N concentrates plant diversity into one or a few N-responsive plants 

(Tilman 1987; Vitousek 1994 and references within), which likely includes many 

non-native species found in several North American grasslands. Stohlgren and others 

(1999) reported that much of the variation in non-native species richness could be explained by the total percentage N in the soil; however, more recent research indicates 

that elevated CO2 could ameliorate the negative effect of N enrichment on species 

richness (Reich 2009).

In summary, predicting the specific effects of global change on current and potential 

invasive plants is a significant challenge because the complex matrix of interacting elements associated with global change often produces inconsistent patterns (Bradley and 

others 2010) and also because invasive plants already represent a significant component 

of global change (Vitousek 1994). Overall, the evidence indicates that invasive plants 

will be favored by many of the elements associated with global change. This may be 

especially true for non-forested ecosystems of the Interior West that are typically water 

and nutrient limited. However, applying general patterns to specific species invading 

specific sites is difficult, and more site-specific research is needed in non-forested ecosystems that simultaneously evaluate multiple elements of global change, including 

the effects of extreme events (Bradley and others 2010). For example, more research is 

needed on the population genetic structure of both cheatgrass and red brome, including 

the initiation of large reciprocal seeding experiments to determine whether differences 

in adaptive phenotypes associated with inbreeding lines that have distinctive marker 

genotypes result in differential establishment, survival, growth, and fecundity in contrasting environments. Additionally, GSD Ecosystem researchers need to respond to 

restoration opportunities where the distribution and abundance of invasive plants are 

contracting because of global change.

Climate Change Effects on Biocontrol

Direct Impacts on Biocontrol Agents

Biocontrol intentionally reunites, for the purposes of pest population regulation in 

an invaded or adopted range, the species targeted for control with co-evolved and hostspecific natural enemies originating from their common native range. The establishment, 

abundance, and distribution of invasive plants (weeds) and the co-evolved, hostspecific herbivorous insects known as their biological control (or biocontrol) agents are 

fundamentally mediated by habitat suitability, competition, and natural enemies (Holt 

and Barfield 2009). Long-term shifts in seasonal temperature and coupled precipitation 

patterns anticipated under climate change in turn have the potential to alter species abundance and distribution (Walther and others 2009). Studies specifically evaluating how 
102 USDA Forest Service RMRS-GTR-285. 2012.

climate change may affect weed biocontrol are scarce (but see Sims-Chilton and others 2010; Watt and others 2010). Under these circumstances, extrapolation from generic 

(“herbivorous insects” or “invasive species”) or multi-trophic (e.g., “host-specific invasive pest” or “rare indigenous species”) case studies can be useful. In addition, results 

reported from studies of invasive or conservation-targeted species are effective for inferring many direct and indirect potential climate change impacts on non-native weeds and 

their arthropod biocontrol agents.

Bale and others (2002) provided a comprehensive review of key direct effects—insect 

herbivore development, survival, range, and abundance—as influenced by the increased 

temperatures predicted under global climate change. Hellmann and others (2008) identified probable consequences of climate change for invasive species and developed testable 

hypotheses for invasive species responses to specific climate change consequences. 

Monitoring for changes in the geographic distribution of host-specific herbivorous insect 

species across latitudinal and elevational gradients was identified as a robust methodology for accurately sensing climate change impacts (Andrew and Hughes 2005; Hodkinson 

and Bird 1998).

Because many of the plant species targeted for classical weed biocontrol in temperate 

regions of the United States originated in Eurasia, the simplistic assumption might be 

made that increasing temperatures in North America will generally benefit biocontrol 

agents that have been collected from a comparatively warmer native range. Studies evaluating ecological sorting along altitude-influenced thermal gradients indicate that insect 

physiological responses to temperature, specifically thermal tolerance thresholds, can 

profoundly affect agent demographics via temperature-mediated fecundity (Dangles and 

others 2008) and distribution (Hodkinson and Bird 1998). The results of a meta-analysis 

of insect species range margins suggests that although genetic diversity tends to decline 

during colonization of new habitats, there could be a positive feedback between range 

expansion and an increase in traits that accelerate range expansion through adaptations 

specifically affecting dispersal, metabolic rate, and changes in habitat associations (Hill 

and others 2011).

Increased temperatures projected under climate change could substantially extend 

the core area and edge-of-range distributions for both weed biocontrol agents and their 

host plants, especially for Eurasian species established in northern temperate locales. 

Thomas and others (1999) determined that a 2 to 3 °C rise in mean spring and summer 

temperature increased available suitable habitat and the length of time that successional 

habitat could be occupied, and decreased the effective distance between suitable habitat 

patches in northern temperate locations. Davis and others (1998) asserted that accurate 

predictions of species range and abundance cannot be based on physiological response 

to temperature alone (called “climate mapping”) but should also consider climatic influences on species dispersal and inter-specific interactions. Bradley and others (2009) 

cautioned that climate change could result in contractions of invasive plant ranges as well 

as range expansions.

The probability that a species will reach locations that have, under the influence of 

global warming, changed from unsuitable to suitable habitat patches depends on the 

interaction of dispersal ability and behavioral responses with environmental structural 

components (Gaston 2009). Fox and others (1999) determined that positive effects of 

winter warming on St. John’s wort (Hypericum perforatum L.) would be ephemeral and 

readily neutralized by a projected increase in summertime herbivory if winter warming 

was followed by summer drought; the authors concluded that St. John’s wort would 

not likely benefit from the warmer temperatures predicted under global warming, particularly at the northern extent of its range. The guild structure of herbivores colonizing 

Acacia falcata growing within or transplanted at sites 208 km beyond its current range 
USDA Forest Service RMRS-GTR-285. 2012. 103

was similar even though the transplant sites were 1.2 and 5.5 °C warmer; this was not the 

case for the guild structure of herbivores collected from a related host species A. leptostachya growing at transplant sites outside the current range (Andrew and Hughes 2007).

The internal temperature of ectotherms, including all plants and most insect species, is regulated by ambient environmental temperature (Gutierrez and others 2008). 

Ectothermic metabolic rate, dictated by body mass and body temperature, is therefore 

highly susceptible to alterations in habitat thermal properties (Dillon and others 2010; 

Gillooly and others 2001). Response to changing climatic conditions is restricted to 

dispersal, phenotypic plasticity, or adaptation (Holt 1990). Phenological alterations 

resulting from the increased length of temperate growing seasons can facilitate the development of asynchrony in key interspecific interactions (van Asch and others 2007; 

Cobbold and Powell 2010; Fabiana and others 2010). Hegland and others (2009) pointed out that asynchrony in insect-plant interactions can be temporal or spatial in nature, 

and, in extreme cases, could lead to trophic decoupling and food web scale disruptions 

characterized by a mismatch in abundance of consumers and their food sources.

Much of the research conducted in Interior West grasslands, shrublands, and desert 

ecosystems concerns function and productivity in dryland habitats. The interaction of 

environmental (e.g., climate change) and anthropogenic (e.g., management practices) 

drivers profoundly affects dryland function and productivity (Chambers and Pellant 

2008). Productivity, in terms of increased plant photosynthesis, biomass, and water use 

efficiency, is predicted to increase, especially for alien invasive species, in U.S. arid 

ecosystems under higher atmospheric levels of CO2 associated with climate change 

(Ziska 2003; Smith and others 2000). Species interactions, and not only direct effects 

of climate change, were shown to influence grassland productivity and species diversity 

(Suttle and others 2007). Drylands are susceptible to dominance by non-native transformer species (i.e., species capable of significantly altering ecosystems over a wide 

area) (Richardson and others 2000). Transformer species such as Bromus tectorum

(cheatgrass; downy brome) and Linaria dalmatica (Dalmatian toadflax) (D’Antonio 

and others 2004) are targets of biocontrol research by scientists in the GSD Ecosystems 

Program of RMRS. Biocontrol, used alone or as part of an adaptive integrated weed 

management strategy, may significantly reduce unintended negative, non-target impacts to surrounding desirable vegetation. Conventional herbicide treatments have 

been correlated with secondary invasions of weeds such as cheatgrass (Pearson and 

Ortega 2009) and an increase in the proportion of bare or unvegetated ground (Barnes 

2007); the ability of dryland vegetation communities to rebound after herbicide applications is likely to be compromised under climate change and may increase the 

frequency, intensity, and persistence of desertification. Verstraete and others (2009) 

paraphrased the United Nations Convention to Combat Desertification definition of 

desertification as: “any form of degradation in drylands...where degradation refers to 

a persistent reduction in the bundle of services provided to humans by the ecosystem 

under considerations, thus encompassing both social and biophysical considerations.”

Impacts on Plant-Herbivore Interactions

The success or failure of biocontrol is largely determined by the outcomes of interactions between individual plants and biocontrol agents (herbivores). That is, herbivory at 

the individual level can have negative effects on a plant’s growth and reproduction, which 

can, in turn, impact the abundance, distribution, and dynamics of entire plant populations 

(Maron and Crone 2006). For biocontrol to be deemed successful, individual herbivory 

must lead to population-level reductions in the target weed. However, climate change has 

the potential to fundamentally alter interactions between plants and herbivores, which 
104 USDA Forest Service RMRS-GTR-285. 2012.

could alter broader population-level outcomes and the success of biocontrol. Another 

primary route by which climate change is predicted to affect herbivory is by modifying 

plant chemistry—the central factor regulating plant-herbivore interactions. Major climate change factors, most notably elevated CO2

 and temperature, can potentially affect 

the susceptibility or resistance of plants to herbivory (e.g., by altering leaf nutrients and 

defensive compounds); however, little is known about how these changes may affect 

individual plant-insect interactions or broader community dynamics. A better understanding of how climate change will impact relationships between invasive plants and 

their biocontrol agents is needed to predict and advance biocontrol efficacy in a rapidly 

changing climate.

The most obvious way climate change can affect interactions between plants and insects is by altering the basic nutritional value of plants. Most studies show that herbivores 

consistently respond to CO2-induced changes in their host plants by consuming more 

foliage (Coviella and Trumble 1999). For example, elevated CO2 generally causes an 

increase in plant growth (the “fertilizer effect”) and increases in the ratio of C:N in plant 

tissues, which reduces the nutritional quality for N-limited insects (Coviella and Trumble 

1999). As a result, insects must eat more to compensate for less N content (Coviella and 

Trumble 1999; Dermody and others 2008; Johnson and McNicol 2010). Another route 

by which increased CO2 can affect insect feeding is by increasing sugars in plant leaves. 

For example, soybean plants grown in elevated CO2 conditions contain 31% more sugars 

than plants grown in ambient air and, as a result, Japanese beetles (Popilla japonica) 

prefer and consume twice as much foliage from high-CO2 soybeans (Hamilton and others 

2005). Drought stress—something predicted to increase dramatically in western North 

America—can affect virtually every plant nutrient and may encourage herbivore outbreaks (Mattson and Hauk 1987). Such climate-induced increases in insect damage are 

expected to negatively impact agricultural production by off-setting potential gains in 

plant productivity due to the fertilizer effect (DeLucia and others 2008). However, the 

implications for biocontrol, though potentially far reaching, are unknown.

Climate change can also significantly impact plant nutritional value by altering 

chemical defenses against herbivores. Plant defensive chemistry can have important 

consequences for plant fitness and populations (Baldwin 1998) and can drive cycles in 

herbivore populations (Underwood 1999). Elevated CO2, temperature, ozone (O3), and 

ultra-violet (UV) light are each reported to affect levels of plant secondary chemicals 

(Bidart-Bouzat and Imeh-Nathaniel 2008). However, available information is limited 

and dependent on the plant and insect species involved as well as the class of chemicals examined (e.g., C-based versus N-based defenses). For example, elevated CO2

, 

temperature, O3, and UV light can each either increase, decrease, or have no effect on 

plant defensive chemistry (Bidart-Bouzat and Imeh-Nathaniel 2008). These conflicting studies have hindered attempts to develop general predictions about how climate 

change will affect plant defensive chemistry and point to the involvement of a host of 

interacting factors. However, climate-induced changes in chemical defenses can have 

important consequences for plants and herbivores. For example, quaking aspen trees

(Populus tremuloides) grown in elevated CO2 and O3, singly and in combination, had altered physical and chemical leaf defenses that led to increased populations of herbivores 

and pathogens (Percy and others 2002). It was recently discovered that elevated CO2 

can disrupt herbivore-induced plant defenses, specifically the production of proteinase 

inhibitors that interfere with insect digestion, resulting in poorly defended leaves and increased growth and development of herbivores (Zavala and others 2008). Moreover, the 

reduction of proteinase inhibitors can further reduce plant fitness by increasing herbivore 

attack on younger leaves, which contributes disproportionately to plant growth (Zavala 

and others 2009).
USDA Forest Service RMRS-GTR-285. 2012. 105

The production and release of volatiles—airborne chemical compounds emitted by 

plants—is another important hallmark of plant-herbivore interactions that is expected 

to be affected by climate change (Yuan and others 2009). These airborne chemicals 

can significantly impact the fitness of plants and insects by serving as foraging cues 

for organisms that are beneficial to plants, such as predators and parasites of herbivores 

(De Moraes and others 1998) or by directly repelling herbivores (De Moraes and others 2001). Volatiles can also convey information about a plant’s identity and location 

to harmful organisms (or beneficial organisms in the case of biocontrol insects) such 

as herbivores (De Moraes and others 2001; Runyon and others 2006). Climate change 

can alter the biological functions of plant volatiles with largely unknown consequences 

(Yuan and others 2009). For example, drought and elevated CO2 can increase emission 

of volatile terpenoids, which could increase plant apparency to herbivores or perturb 

attraction of herbivore natural enemies (Mattson and Hauk 1987; Himanen and others 

2009). Increases in CO2 have been shown to increase volatile production by soybean 

plants, which seemingly act as a super stimulus that may elicit an exaggerated feeding 

response in herbivores (O’Neill and others 2010). Conversely, O3 can react with and 

rapidly degrade certain volatiles in the atmosphere (Pinto and others 2007). Elevated 

temperature has long been known to increase plant volatile emission rates (Guenther 

and others 1993), indicating a general increase in plant volatiles under a warmer climate 

with unknown but likely profound impacts on ecological interactions between plants and 

insects (Yuan and others 2009). It is unlikely these changes could lead biocontrol insects 

to shift to non-target plant species; biocontrol insects are highly host-specific and rely on 

species-specific chemical cues (e.g., the presence/absence of compounds) to locate and 

feed. However, our poor current state of knowledge about such potential climate change 

impacts does not allow us to rule this possibility out.

In summary, we know alarmingly little about how climate change will impact the relationship between plants and insects, despite the profound implications for agriculture and 

biocontrol. The limited knowledge available indicates that responses are highly variable 

and dependent on the species involved. Many herbivores will alter how much they eat 

in response to climate-induced changes in plant nutrition and plant defensive chemistry; 

yet, we know next to nothing about what this might mean ecologically or economically. 

If biocontrol is to keep pace with and remain effective in a changing climate, increased 

funding is needed to:

• Determine how climate change will affect the nutrient content of invasive plants and 

what impact these changes will have on biocontrol agents.

• Evaluate effects of climate change on the defensive chemistry of invasive plants and 

determine how these changes impact biocontrol agents.

• Develop a conceptual framework to understand and predict how climate-induced 

changes will alter broader population- and landscape-level outcomes of biocontrol.

• Develop effective tools and techniques to best use and adapt biocontrol to manage 

invasive plants in a changing environment.

This is an opportunity to advance our basic understanding of the ecology of plantinsect interactions and the conditions under which herbivory translates into meaningful 

changes in plant populations—fundamental ecological questions that hold great promise 

for managing invasive plants in present and future environments. Moreover, because 

climate-induced changes in western North America—much of which is dominated 

by grasslands, shrublands, and deserts—have generally outpaced change elsewhere 

(Overpeck and Udall 2010), biocontrol is likely to be affected first and most severely 

there. As such, the GSD Ecosystems Program is well positioned to take on the research 

needed to adapt and respond to future environmental changes.
106 USDA Forest Service RMRS-GTR-285. 2012.

Climate Change Effects on Wildlife Disease

Disease is a natural process in wildlife communities that, under normal circumstances, acts to regulate communities and interactions. However, when disease dynamics are 

altered by wildlife manipulations (translocations, hunting, and commercial trade) and 

lead to new pathogen-host interactions, disease often becomes a destructive force. In 

general, invasive species, and in particular, vectors and pathogens, can destabilize natural communities and irrevocably change ecosystem structure and function with severe 

economic and environmental consequences (Crowl and others 2008). Wildlife managers and conservation biologists have become increasingly concerned with the rise in 

emergence of many serious diseases, including plague (Yersinia pestis) encephalitis, 

canine distemper, and West Nile virus (Daszak and others 2001; Deem and others 

2000; Gubler and others 2001).

Disease introductions in wildlife populations usually occur at domestic-wildlife interfaces or are related to translocation efforts. Wildlife disease outbreaks are commonly 

associated with increased proximity to humans and domestic animals (Deem and others 2000). Multispecies land use, such as occurs in buffer zones where domestic and 

wildlife share grazing lands, is thought to facilitate disease spread (Daszak and others 

2001; Deem and others 2000). However, the primary mechanism for disease spread 

is translocation (Deem and others 2000). Translocations have multiple consequences 

for disease management, including an increased risk of exposure of wildlife to new 

diseases present in new location or unintentional introductions of disease vectors or 

carrier species (Deem and others 2000). Recent translocations of the white-tailed deer 

(Odocoileus virginanus) resulted in the spread of the lungworm Parelaphostronglyus 

tenuis to Wassa Island, Georgia (Davidson and others 1996). Similarly, reintroduction 

of a confiscated desert tortoise (Gopherus agassizii) infected with mycoplasmosis in 

Las Vegas Valley led to the spread of that disease in native populations (Jacobson and 

others 1995), and the transport of infected carcasses was associated with the spread of 

canine distemper in the Southwest (Davidson and others 1992; Deem and others 2000). 

The zebra mussel (Dreissena polymorpha) is an invasive species whose impact has 

been amplified through its status as a carrier of the roundworm parasite, Bucephalus 

polymorphus, which can also infect many freshwater cyprinid fish (Crowl and others 

2008). Among the invasive diseases currently known to occur in the Interior West, 

plague, canine distemper, brucellosis (Brucella spp.), chronic wasting disease, bovine 

tuberculosis (Mycobacterium bovis), whirling disease (Myxobolus cerebalis), and West 

Nile virus have considerable negative impacts on wildlife population (Bengis and others 2002; Clinton and others 2010; Githecko and others 2000; Longstreth and Wiseman 

1989; Mason 2008).

Global warming will impact many wildlife disease patterns, particularly vectorborne diseases (Daszak and others 2001; Harvell and others 2002; Patz and others 

2000; Keesing and others 2006; Rosenthal 2009). Local climatic conditions are 

thought to play an important role in determining disease emergence (Githeko and 

others 2000; Harvell and others 2002; Hofmeister and others 2010; Lafferty 2009), 

and global warming is predicted to lead to range expansions of many vector species 

and increase the frequency of vector borne disease outbreaks (Epstein 2001; Harvell 

and others 2002). Issues associated with wildlife translocations and their roles in disease emergence are also expected to increase under future climate scenarios (Deem 

and others 2000). Among other effects, increases in temperature are expected to positively influence the spread of disease by decreasing overwinter mortality of many 

arthropod vectors and parasites (Harvell and others 2002) and increasing vector and 

pathogen developmental rates (Wilcox and Gubler 2005). Within the Interior West, 
USDA Forest Service RMRS-GTR-285. 2012. 107

the spread of diseases such as West Nile virus and Lyme disease may be encouraged 

by increases in mean temperature and humidity, which will open up new zones for 

their mosquito and tick vectors (Deem and others 2000). In higher elevations, warmer temperatures may contribute to the spread of chytrid fungus (Batrachochytirum 

dendrobatidis) among amphibians (Pounds and others 2006; Rohr and Raffel 2010). 

Furthermore, amphibian host susceptibility may increase as heat stress and high 

UV-B affect immune response (Harvell and others 2002; Martin and others 2010). 

Rising water temperatures are likely to exacerbate ongoing issues with the introduced agent of whirling disease, Myxobolus cerebalis (Mason 2008; Longstreth and 

Wiseman 1989). Increases in host breeding season may contribute to the spread of 

Ophryosystis elektroscirrha (protozoal infection) in monarch butterflies (Harvell and 

others 2002). Of the exotic diseases present in the Interior West, plague and West 

Nile virus are the most susceptible to climate-related changes in distribution and 

incidence. Disease distribution and incidence may also increase if wildlife is translocated as part of assisted migration projects being developed to lessen negative 

climate change impacts (McDonald-Madden and others 2011).

Changes to precipitation regimes will influence the availability of favorable habitat for vectors that rely on water bodies and, in turn, influence the establishment of 

disease in new areas (Gubler and others 2001; Patz and others 2000). Reduced precipitation might limit breeding sites for many mosquito vectors but could also lead to 

conditions, such as overcrowding at limited water holes, increased water temperatures, 

and more organic matter, that are conducive to disease spread, particularly for waterborne diseases such as avian cholera and pox (Friend and Franson 1999). Precipitation 

also affects intermediate and reservoir hosts by influencing important food resources 

(Dazak and others 2001). Changes in the overlap of species due to phenological effects of changing temperature and precipitation regimes have consequences for host, 

reservoir, and vector populations and may lead to new disease issues (Harvell and others 2002; Hofmeister and others 2010; Patz and others 2000). Changes to the timing 

of host migrations may also influence the risk of disease exposure for some wildlife 

populations and lead to novel host-pathogen interactions.

However, climate change may not always lead to an expansion of disease and 

could, in fact, decrease some disease threats. Increased temperatures can reduce adult 

survivorship of vectors (Harvell and others 2002) and limit transmission of certain 

diseases such as plague that have upper critical temperature thresholds (Githeko and 

others 2000). Entomophatogenic fungi in insects (some biocontrol agents), coldwater 

disease in salmon, and avian cholera in waterbirds (Pasteurella multocida) may decline as temperatures rise (Harvell and others 2002). Warming may limit the spread of 

chytrid and iridoviruses, which rely on cool, moist conditions, in amphibian populations in warmer climates. Hot, dry conditions may impede fungal development and 

enhance insect immune response (Martin and others 2010). Perhaps to the benefit of 

some biocontrol efforts, warmer temperatures are expected to reduce the cold-induced 

mortality of the nuclear polyhedrosis virus Lymantria dispar on gypsy moth egg cases 

(Harvell and others 2002). However, the direction of disease response depends on local conditions and the inherent limitations of the disease agents (Githeko and others 

2000). Commonly, pathogens limited by winter mortality show predicted range contractions in the south and corresponding expansions to northern areas (Haile 1989). 

Ultimately, many diseases may shift rather than experience an absolute change in their 

area (Lafferty 2009). Identifying if and how climate change will impact specific ecosystems, populations, and pathogens is a critical step toward informing management 

agency actions with respect to disease emergence.
108 USDA Forest Service RMRS-GTR-285. 2012.

Wildlife disease becomes a Forest Service issue when it affects threatened and endangered species and human use of Forest Service lands and when management of 

forest lands can mitigate the introduction or expansion of invasive disease. We need to 

implement research and build expertise to address invasive disease issues and facilitate 

actions that allow us to manage rangelands for biological diversity, health, and sustained and enhanced use by our stakeholders. Research needs to focus on mitigating the 

potential effects of invasive disease on threatened and endangered species to prevent 

further population declines, identify populations at risk due to inherent susceptibilities 

or increased exposure (migrating species, species in buffer zones), and identify the 

important interactions (climate, species interactions, and land use practices) that affect 

ecosystem integrity and invasibility. In addition, wildlife disease has socioeconomic 

effects when it relates to human use of land and to human health issues (e.g., zoonotic 

diseases such as plague, hanta virus, Lyme disease, and West Nile virus).

Specific research needs that address the strategic goals of the Forest Service mission 

are:

• Identify disease threats to threatened and endangered species.

• Identify disease risk for critical habitats, particularly breeding and migratory stopover 

sites. Analyze risk factors for species and management units and for translocation 

or assisted migrations.

• Assess economic and socioeconomic issues related to emergent wildlife disease.

• Identify and monitor susceptible/at risk populations.

• Determine how current and proposed management (e.g., restoration) activities affect 

disease invasibility of ecosystems or basic health parameters of wildlife populations.

• Determine if certain practices are more or less likely to favor the spread of disease.

• Evaluate the effects of management actions for mitigating disease impacts.

Many diseases affect species present or dependent upon grassland habitats and 

new diseases will further threaten these populations. Many species, such as the desert 

tortoise and sagebrush grouse, are already endangered and attention must focus on preventing further population decline. The GSD Ecosystems Program is able to address 

many relevant research questions with expertise in human resources, disease, soil and 

invasive species. The diversity of sites and ecosystems available in the GSD region 

puts scientists in this program in a unique position to address some if not all of these 

issues and to be able to satisfy recent calls for regional-level monitoring (see Crowl 

and others 2008) and analysis of disease emergence and spread. If applied successfully, these efforts might also be used effectively in other systems and ranger stations. 

Current activities with the black footed ferret (Mustela nigripes) (Rapid City Lab), 

which are highly susceptible to plague and canine distemper, are one way the GSD 

Program can develop a more aggressive and comprehensive disease-oriented research. 

Furthermore, many activities related to current RMRS research, particularly as they 

pertain to grazing; human-domestic, animal-wildlife interfaces; and restoration activities are also highly relevant to disease prevention and management issues.

In addition to building upon its own expertise and initiating new research, researchers in the GSD Program need to engage opportunities to work with other programs 

and agencies with ongoing research relevant to species of interest to the Forest Service 

(e.g., USGS/BLM SAGEMAP project for the greater sage-grouse, which is threatened 

by West Nile virus). The recent spread of white-nose syndrome illustrates the devastating impact of introduced disease and points to the critical importance to establish 

networks and cooperation in anticipation of disease emergence. The recent finding 

of white-nose syndrome in the cave bat (Myotis velifer) in Oklahoma is a troubling 

predictor that this western bat species could soon bring white-nose syndrome to the 
USDA Forest Service RMRS-GTR-285. 2012. 109

western United States. As this disease potentially impacts over half the species endemic to United States, the inevitable spread of this disease is an impending crisis requiring 

immediate action. Forest Service researchers should have a role in identifying research 

needs and intervention strategies managing white-nose syndrome.

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116 USDA Forest Service RMRS-GTR-285. 2012.

Chapter 8

Decision Support: Vulnerability, Conservation, 

and Restoration

Megan M. Friggens1, Jeremiah R. Pinto2, R. Kasten Dumroese2, and Nancy L. Shaw3

1 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Science Laboratory, Albuquerque, New Mexico

2 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Forestry Science Laboratory, Moscow, Idaho

3 U.S. Forest Service, Rocky Mountain Research Station; Grassland, Shrubland, and Desert Ecosystems Program; 

Aquatic Sciences Laboratory, Boise, Idaho

Executive Summary

Current predictive tools, management options, restoration paradigms, and conservation programs are insufficient to meet the challenges of climate change in western 

North America. Scientific and management capabilities and resources will be sapped 

trying to identify risks to genetic resources and ecosystems and determine new approaches for mitigating and managing changing environments. Developing new tools 

will require innovative research, improvement and creation of predictive models, continuous evaluation of management outcomes, and integration with social scientists and 

economists.

Climate change threatens the biodiversity of grasslands, shrublands, and deserts at 

scales ranging from the gene to complex ecosystems. The rate of climate change may 

overcome normal ecosystem resilience, disrupting ecosystem functioning and provision 

of critical services. Guidelines for identifying and conserving at-risk species through 

a variety of experimental methods are available and being utilized. Nonetheless, these 

approaches and models for predicting future risks are evolving and not universally accepted or applicable.

Elements used to identify species or systems vulnerable to climate change include 

effects of exposure to climate change, sensitivity or the level to which the organism or 

system is altered, and its capacity to adjust to the change. Vulnerability assessments 

focus on unique variables or combinations of variables for comparison of organisms, 

natural systems, or human systems and range widely in their objectives; all rely on 

projections of future conditions. These assessments aid in planning adaptation strategies and prioritizing management. Available assessment tools include: vulnerability 

indices, process simulations, evaluation of shifts in species or community distribution, 

and integrated models. Research must focus on improved climate change predictions, 

species and habitat response models, identification of new community compositions, 

and management options.

Selection of appropriate plant materials for restoration necessitates an understanding of genetic variation and structure across the landscape. Species-specific seed zones 

are available for commercial trees but only for a few other species. A number of bioclimatic tools are used to delineate provisional seed zones, and broadly adapted seed 
USDA Forest Service RMRS-GTR-285. 2012. 117

sources are being developed for selected species and zones. Although western plant 

communities have constantly reassembled over time in response to changing climatic 

conditions, rapid climate change will increase fragmentation and cause appropriate 

habitat to appear in new locations. Assisted migration of native plants, a form of ex situ 

conservation, involves moving pre-adapted genotypes into remaining portions of the 

species range or moving a species into new but remote habitat. This approach remains 

controversial from biological and sociological standpoints. There is an urgent need to 

better understand future climate scenarios and appropriate transfer of genetic material 

and to provide analysis and discussion of natural and assisted future redistributions of 

species.

Climate change impacts on grassland, shrubland, and desert species and ecosystems 

are expected to increase but are difficult to predict for many areas. There is an immediate need for improved tools and approaches for assessing vulnerabilities at all levels, 

conserving diversity, and developing new techniques for selecting appropriate native 

plant materials for restoring disturbed areas and for moving genetic materials to new 

locations as climatic conditions change. Resources needed to accomplish these goals 

include genecologists, modelers, nursery and plant materials specialists, biologists, social scientists, and economists.

Introduction

Conservationists and land resource managers are gravely concerned about the impact of climate change because it will involve large numbers of species in diverse 

ecosystems, climate change interactions with ecosystems are wrought with complicated uncertainties, and our response will be limited by available human resources. 

Managers require effective tools now to manage natural resources under current climatic conditions. Managers will also need new methods and tools to help identify 

species and ecosystems at greatest risk of harm due to climate change and how to 

mitigate, or exploit, that change. To focus limited resources in the most effective and 

efficient manner, these tools should identify potential management intervention points 

(e.g., identify how systems are likely to be harmed) and address uncertainties in future 

conditions modeled by climate model projections and species’ responses to those future conditions.

This chapter has three main topics:

• First, we discuss the ramification of the interactions of biodiversity and climate 

change and why conserving biodiversity is paramount.

• Second, we discuss how biodiversity, either from a species or ecosystem standpoint, 

can be assessed for its vulnerability to climate change. Vulnerable species or systems can then be identified and targeted for restoration.

• Third, we discuss how appropriate genetic material of vulnerable plant species or 

systems is currently transferred and may need to be transferred in the future to ensure successful restoration.

Climate Change and Biodiversity

Biodiversity affords ecosystems the plasticity to respond to natural disturbances, 

including naturally changing climate (Risser 1995). Climates are, however, changing at a rate faster than observed historically, thereby compromising these natural 
118 USDA Forest Service RMRS-GTR-285. 2012.

biological responses (Hughes 2000; Parmesan and Yohe 2003). It is therefore critical 

to identify conservation efforts at all scales (genetic, population, species, and ecosystem) in order to maintain plasticity and ecosystem function (Hannah and others 2002).

The general research consensus is that biodiversity (genetic variation, population variation, species richness, and ecosystem complexity) is threatened by climate 

change (Hannah and others 2002; Midgley and others 2002; Schwartz 1992; 

Schwartz and others 2006). Climate change has the potential to reduce valuable 

ecosystem services (such as production of food, pharmaceuticals, timber, and clean 

water) can contribute to floods and droughts, and can disrupt biogeochemical cycles 

(Daily 1997; Hughes and others 1997). Climate change may compromise ecosystem 

resiliency by reducing or eliminating plant and animal species (Thomas and others 

2004) through range shifts in plant distributions (Beckage and others 2008; Soja 

and others 2007; Thomas 2010), increases in invasive species pressure (Smith and 

others 2000), and associated changes in disturbance regimes (McKenzie and others 

2004). Significant habitat loss, disturbance, and increased habitat fragmentation also 

threaten native species’ genetic diversity through inbreeding depression (Holt 1990; 

Johnson and others 2010; Thomas and others 2004). In grassland, shrubland, and 

desert ecosystems of the Great Basin, climate change effects have been forecasted 

and documented (Friggens and others, Chapter 1 this volume) as they relate to rare 

and vulnerable species (Fleishman 2008) and water resources, agriculture, native 

ecosystems, biodiversity, and recreation (summarized in Chambers 2008). The direct 

pressures on grassland, shrubland, and desert ecosystem biodiversity in the West are 

varied. Higher-elevation ecosystems are expected to shrink or vanish (Ledig and 

others 2010); ephemeral riparian and wetland systems may vanish (Hurd and others 1999); and highly invasive species may negatively affect native species through 

competition or altered fire regimes (Ziska and others 2005).

Conversation surrounding the loss of biodiversity due to climate change is contentious. Although we have clear guidelines, both globally (NatureServe) and nationally 

(Endangered Species Act; ESA), for identifying species at risk of extinction and 

conserving them (e.g., the black-footed ferret [Mustela nigripes], the models and 

assessments used for predicting future biodiversity losses of species still relatively 

abundant have yet to gain wide acceptance (Botkin and others 2007; Hannah and others 2002). In addition, the appropriate conservation strategies (i.e., in management 

areas [in situ], via assisted colonization [ex situ], or via germplasms, botanical gardens, or captive breeding programs [in vitro]) are under scrutiny (Hoegh-Guldberg 

and others 2008; Ricciardi and Simberloff 2009). Even so, germplasm of plants 

critically imperiled on a global level are currently being conserved in vitro (cryogenic storage of germplasm) by the Forest Service’s National Seed Laboratory and 

Agricultural Research Service, while land managers work to protect plant and animal 

species under the ESA. However, efforts may not be adequate or sufficiently proactive to mitigate species and genetic losses due to climate change (Hoegh-Guldberg 

and others 2008).

To be sufficiently proactive, we need to identify, develop, and use appropriate vulnerability assessment tools to predict climate related increases in the risk of species 

extinction and population bottlenecking. To preserve biotic diversity, these assessments must provide potential management actions and refine research needs. These 

tools also must integrate bioclimatic modeling, genecology, and climate interactions 

with disturbance, invasive species, and species autecology. It is only through the 

development of these tools that we will be able to accurately assess and identify effective management actions for the preservation or restoration of critical habitats and 

biodiversity.
USDA Forest Service RMRS-GTR-285. 2012. 119

Assessing Species Vulnerability to Climate Change

Vulnerability is commonly defined as a function of exposure, sensitivity, and adaptive capacity (IPCC 2007; Stein and others 2011) and how these elements relate to the 

likelihood that species or systems are affected by climate change, the degree to which 

they are impacted by change, and their capacity to deal with change. Vulnerability 

assessments, using models, scoring systems, and comprehensive synthesis of the literature, determine which species or systems are most likely to be affected by climate 

change. Assessments usually target a unique variable or set of variables that act as the 

measure of vulnerability. Biodiversity and degree of expected change in microclimate 

are common measures to compare habitats, whereas vulnerability comparison among 

species depends on exposure levels and the possession (or lack thereof) of specific 

characteristics. Assessments may focus narrowly on species in select habitats or be 

global in perspective, but all evaluate the potential sensitivity, exposure, and adaptive 

capacity of their targets and all rely on projections of future conditions.

Climate change vulnerability assessments include a broad array of documents and 

analyses that synthesize many predictions and projections, and may take the form of 

qualitative evaluation of species traits or ecosystem function or involve statistical analysis of the relative influence of various parameters on population trends. Climate change 

vulnerability assessments vary in their objectives and can target human systems, natural 

systems, and processes of both (Füssel and Klein 2005). Vulnerability assessments are 

often the first step in planning adaptation strategies and management. By providing information on susceptibility to climate change impacts, assessments help identify targets 

for mitigation, enable managers to prioritize management activities and resources, and 

assist with implementing adaptive strategies (Füssel and Klein 2005).

Although many types of assessment tools are used, most fall into four broad categories: (1) vulnerability indices; (2) simulated processes; (3) community distribution 

shifts; and (4) complex, integrated models.

Vulnerability Indices

Several assessments rely on an indicator or index of vulnerability, which is used to 

compare the relative vulnerability of plant or animal species or systems. For example, 

the NatureServe Climate Change Vulnerability Index (http://www.natureserve.org/

prodServices/climatechange/ccvi.jsp), designed for both plant and animal species, was 

used in Nevada and Massachusetts (Galbraith and O’Leary 2011; Young and others 

2011); and SAVS, a System for Assessing Vulnerability of Species to climate change 

(Bagne and others 2011), was used to assess terrestrial vertebrate species in New 

Mexico (Finch and others 2011; Friggens and others, in prep.) and Arizona (Bagne and 

Finch 2010; Coe and others, in prep.). On a broader scale, the Environmental Protection 

Agency has a scoring system that has been applied nationally to assess the combined 

impact of climate change and non-climate related vulnerabilities for threatened and 

endangered species (Galbraith and Price 2011; U.S. EPA 2009). Other regional assessments incorporate indices and other projection tools (e.g., Czúcz and others 2009; 

Tremblay-Boyer and Anderson 2010). For example, an analysis of climate effects for 

the Pacific Northwest includes the use of sensitivity indicators (traits), downscaled 

climate projections, and dynamic global vegetation models (Case and Lawler 2011; 

Lawler and others 2009). This approach is commonly used to prioritize intervention 

or management actions or to identify research needs. It may also serve to identify new 

management targets when assessments reveal significant impact for targets not currently of concern.
120 USDA Forest Service RMRS-GTR-285. 2012.

Process Simulation

Assessments that use models that simulate processes, most commonly biogeochemical models or dynamic global vegetation models, can provide important information for 

management and policy decision. For example, WaterSim (Gober and Kirkwood 2010) 

estimates water shortages in the Phoenix area under different scenarios of population 

growth; the MAPSS biogeography model (Hanson and others 2001) projects biome response to climate changes in forests; and Hauer and others (1997) simulated impacts of 

climate change on freshwater ecosystems in North America’s Rocky Mountains. These 

types of models formed the basis for species distribution analyses used by Glick and 

Wilson (2011), Lawler and others (2009), and Rehfeldt and others (2006). Assessments 

based on these methods are strongly influenced by the quality of data used in generating 

output, including the projections for future conditions.

Shifts in Species or Community Distributions

A number of assessments use estimates of shifts in species or community distributions to infer climate change impacts, which takes the form of an occupancy or niche 

modeling effort that relates future species distribution to climate or other abiotic conditions based on current environmental conditions. Future conditions are estimated 

based on climate projections created from downscaled GCMs or future expectations for 

biogeochemical processes predicted from computational models. Rehfeldt and others 

(2006) showed that changes in biotic community and individual plant species distributions for the western United States will be great under a number of different climate 

scenarios (Friggens and others, Chapter 1 this volume). In the western hemisphere, 

they predicted 90% of nearly 3000 vertebrate species will be lost from certain habitats 

(Lawler and others 2009), with some species experiencing declining distribution (e.g., 

fresh water fishes; Eaton and Scheller 1996) and some experiencing expanded distribution (Humphries and others 2002; Meyer and others 1999; Shutter and Post 1990). 

These efforts are data intensive but are able to provide scenarios for a potential future. 

They can be used to infer potential loss of habitat suitability for species or communities.

Complex Analyses

The most complex analyses attempt to integrate adaptive strategies with vulnerability 

assessments to gauge how actions influence relative susceptibility to climate change 

impacts. One analysis incorporates sensitivity scores with an analytical framework to 

create output relevant to both management and policy decisions (Luers 2005) whereas 

others integrate regional assessments, adaption planning frameworks, and a number of 

climate modeling tools (Enquist and Gori 2008, described in McCarthy and Enquist 

2011; NatureServe Vista found at: http://www.natureserve.org/prodServices/vista/overview.jsp).

Current Assessment Tools

We list, albeit not comprehensively, many widely and freely available tools 

for managers to assess species or ecosystem vulnerability to climate change (table 8-1). Other syntheses of assessment tools can be obtained from the U.S. Forest 

Service, Pacific Northwest Research Station (http://www.fs.fed.us/nw/corvallis/mdr/

mapss) and the Nairobi Work Programme, under the United Nations Framework 

Convention on Climate Change (http://unfccc.int/adaptation/nairobi_workprogramme/

knowledge_resources_and_publications).
USDA Forest Service RMRS-GTR-285. 2012. 121

Table 8-1. Examples of the types of tools and data commonly used to assess vulnerability to 

climate change.

Type Name Description Target/	
  

Scope

Sources/Websites

Scoring	
  tools Typically	
  quantify	
  vulnerability	
  

through	
  a	
  tally	
  of	
  traits	
  or	
  

characteristics	
  associated	
  with	
  

increased	
  risk	
  of	
  negative	
  impact

1.	
  NatureServe	
  Climate	
  

Change	
  Vulnerability	
  Index	
  

Classifies	
  species	
  into	
  six

categories:	
  six	
  possible	
  scores	
  are	
  

Extremely	
  Vulnerable,	
  Highly	
  

Vulnerable,	
  Moderately

Vulnerable,	
  Not	
  

Vulnerable/Presumed	
  Stable,	
  Not	
  

vulnerable/Increase	
  Likely,	
  and	
  

Insufficient	
  Evidence

Animal	
  and	
  

plant	
  species

www.natureserve.org/prodS

ervices/climatechange/Clima

teChange.jsp

2.	
  System	
  for	
  Assessing	
  

Vulnerability	
  of	
  Species	
  SAVS

Uses	
  a	
  questionnaire	
  format	
  

create	
  a	
  score	
  indicating	
  relative	
  

vulnerability	
  to	
  expected	
  changes	
  

in	
  future	
  conditions

Terrestrial	
  

vertebrate	
  

species

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

ssland-­‐shrubland-­‐

desert/products/species-­‐

vulnerability/savs-­‐climate-­‐

change-­‐tool/

3.	
  EPA	
  Framework T&E	
  Species EPA/600/R-­‐09/01

4.	
  Vulnerability	
  Surface Uses	
  a	
  three-­‐dimensional	
  

analytical	
  surface	
  to	
  determine	
  

relative	
  vulnerability

Applicable	
  to	
  

variety	
  of	
  

systems

Luers	
  2005;	
  Luers	
  and	
  others	
  

2003

	
  Habitat	
  and	
  species	
  distribution	
  

(e.g.,	
  bioclimatic)	
  models

Use	
  biophysical	
  measures	
  to	
  

define	
  climate	
  space	
  of	
  species	
  or	
  

communities.

Typically	
  

vegetation	
  

communities

1.	
  Climate	
  surface	
  models	
  for	
  

plant	
  species*

Use	
  climate	
  surfaces	
  and	
  

observed	
  species-­‐climate	
  

relationships	
  to	
  predict	
  species	
  

distributions

Plant	
  

communities	
  

and	
  species

Rehfeldt	
  and	
  others	
  2006;

http://forest.moscowfsl.wsu.

edu/climate/customData/ind

ex.php

3.	
  Genetic	
  Algorithm	
  for	
  

Rule-­‐Set	
  Prediction	
  (GARP)	
  

niche	
  model

Uses	
  spatial	
  data	
  on	
  temperature,	
  

rainfall,	
  and	
  elevation	
  with	
  point	
  

data	
  on	
  species	
  range	
  to	
  estimate	
  

potential	
  range

Native	
  and	
  

non-­‐native	
  

species

nhm.ku.edu/destopgarp

4.	
  Maximum	
  entropy	
  

(Maxent)	
  Habitat	
  model

Uses	
  set	
  of	
  environmental	
  

variables	
  and	
  georeferenced	
  

occurrence	
  locations	
  to	
  produce	
  

models	
  of	
  species'	
  ranges

Animal	
  or	
  

plant	
  species	
  

Philips	
  and	
  others	
  2006;	
  Elith	
  

and	
  others	
  2011;

http://www.cs.princeton.edu

/~schapire/maxent/

5.	
  Random	
  Forest	
   Classification	
  system	
  that

produces	
  robust	
  estimates	
  of	
  

species	
  presence.	
  Used	
  in	
  

Rehfeldt	
  and	
  others	
  2006.

Various Breiman	
  2001;	
  Cutler	
  and	
  

others	
  2007;

http://www.stat.berkeley.ed

u/~breiman/RandomForests/

cc_home.htm

7.	
  Climate	
  FVS* Models	
  species	
  climate	
  profiles.	
  

Users	
  input	
  species	
  profile	
  and	
  

elevation	
  to	
  get	
  projected	
  

distributions	
  under	
  a	
  variety	
  of	
  

climate	
  scenarios

Forests/tree	
  

species

http://www.fs.fed.us/fmsc/f

vs/description/climate-­‐

fvs.shtml

Biogeochemical	
  models	
   Model	
  changes	
  in	
  climate	
  

parameters,	
  including	
  

Table 8-1. Examples of the types of tools and data commonly used to assess vulnerability to climate change.

See text
122 USDA Forest Service RMRS-GTR-285. 2012.

Table 8-1. Examples of the types of tools and data commonly used to assess vulnerability to 

climate change.

Type Name Description Target/	
  

Scope

Sources/Websites

temperature	
  and	
  relative	
  

humidity.	
  Often	
  inform	
  

parameterization	
  of	
  the	
  above	
  

class	
  of	
  tools.

1.	
  Instantaneous	
  canopy	
  flux	
  

model	
  (PnET)	
  

Merge	
  of	
  three	
  computational	
  

models	
  that	
  simulate	
  carbon,	
  

water,	
  and	
  nitrogen	
  dynamics

Forest	
  

ecosystems

Aber	
  and	
  Federer	
  1992;

http://www.pnet.sr.unh.edu

/download.html

2.	
  Soil	
  Organic	
  Matter	
  Model	
  

(CENTURY)

Simulates	
  nutrient/hydrological	
  

flows	
  and	
  includes	
  fire/harvest	
  

frequency

Watershed www.nrel.colostate.edu/proj

ect/century5/

6.	
  BIOCLIM	
  (BIOMAP) Prediction	
  systems	
  that	
  uses	
  

mean	
  monthly	
  climate	
  estimates	
  

to	
  predict	
  energy	
  and	
  water	
  

balances	
  at	
  specified	
  location

Area	
  defined	
  

by	
  user

software.infromer.com/getfr

ee-­‐bioclim-­‐download-­‐

software

3.	
  Mapped	
  Atmosphere-­‐

Plant-­‐Soil	
  systems—MAPSS

	
  Equilibrium	
  model	
  that	
  calculates	
  

plant	
  available	
  water	
  and	
  

temperature	
  thresholds	
  according	
  

to	
  climatic	
  zone,	
  life	
  form,	
  and	
  

plant	
  type.

Area	
  defined	
  

by	
  user

See	
  Bachelet	
  and	
  others

2001;	
  	
  

www.fs.fed.us/pnw/corvaliis

/mdr/mapss

Coupled	
  models

Dynamic	
  global	
  vegetation	
  

models

Incorporate	
  vegetation	
  

projections	
  and	
  general	
  

circulation	
  models	
  (GCMs)	
  with	
  

the	
  purpose	
  to	
  inform	
  climate	
  

dynamics	
  (e.g.,	
  albedo	
  and	
  water	
  

evaporation	
  rates)	
  

Botkin	
  and	
  others	
  2007

1.	
  MC1 Combines	
  CENTURY	
  and	
  MAPSS http//www.fsl.orst.edu/dgv

m

Hydrological	
  Models Model	
  changes	
  in	
  ground	
  water,	
  

stream	
  flow,	
  evaporation,	
  etc.

Christensen	
  and	
  others	
  2008

1.	
  Regional	
  Hydro-­‐Ecologic	
  

Simulation	
  System	
  (RHESSys)	
  

GIS	
  based	
  hydro-­‐ecological	
  model	
  

simulates	
  water,	
  carbon	
  and	
  

nutrient	
  flow

Watershed fiesta.bren.ucsb.edu/~rhessy

s/setup/downloads/downloa

ds.html

2.Sea	
  Level	
  Affecting	
  

Marshes	
  Model-­‐	
  SLAMM

Models	
  processes	
  dominating	
  

wetland	
  conversion	
  and	
  shoreline	
  

modification

Coastal	
  areas

Glick	
  and	
  others	
  2010;	
  

http://www.slammview.org

Others

1.	
  The	
  Terrestrial	
  

Observation	
  and	
  Prediction	
  

System	
  (TOPS)

Simulation	
  framework—links	
  

historical	
  climate	
  data,	
  remotely	
  

sensed	
  data,	
  climate	
  projections,

and	
  response	
  models

Nemani	
  and	
  others	
  2009;

http://gcmd.nasa.gov/record

s/NASA_ARC_TOPS.html

2.	
  Program	
  to	
  Assist	
  in	
  

Tracking	
  Critical	
  Habitat	
  

(PATCH)	
  

Models	
  species	
  vulnerability	
  by	
  

linking	
  landscape	
  pattern	
  and	
  

species	
  traits

Ideal	
  for	
  

habitat	
  

specialists

www.epa.gove/wed/pages/n

ews/03June/schumaker.htm

Statistical	
  decision	
  support Statistical	
  methods	
  to	
  estimate	
  

potential	
  response	
  of	
  targets	
  to	
  

risk	
  factors	
  and	
  uncertainty.	
  

Bernliner	
  and	
  others	
  2000;	
  

Prato	
  2009

Table 8-1. Continued.

m

s

news/03June/schumaker.htm
USDA Forest Service RMRS-GTR-285. 2012. 123

Table 8-1. Examples of the types of tools and data commonly used to assess vulnerability to 

climate change.

Type Name Description Target/	
  

Scope

Sources/Websites

1.	
  Bayesian	
  Analysis	
  

Toolkit	
  

Software	
  package	
  that	
  allows	
  

users	
  to	
  compare	
  model	
  

predictions	
  to	
  data,	
  test	
  model	
  

validity,	
  and	
  extract	
  values	
  of	
  free	
  

parameters	
  of	
  models.

http://www.mppmu.mpg.de

/bat/

2.	
  Treeage	
  Pro	
   Decision	
  support	
  software	
  that	
  

uses	
  various	
  methods	
  to	
  

distinguish	
  between	
  models	
  and	
  

decisions	
  options

www.treeage.com/products.

index.html

3.	
  Delphi	
  Decision	
  Aid	
  

site	
  

Data	
  gathering	
  tool	
  for	
  

forecasting	
  purposes

armstrong.wharton.upenn.e

du/delphi2/

Conceptual	
  models Qualitative	
  descriptions	
  and	
  

diagrams	
  of	
  attributes	
  and	
  

processes	
  of	
  concern

Species,	
  

habitats	
  or	
  

ecosystems

Heemskerk	
  and	
  others	
  2003;

www.fileheap.com/sofware/

conceptual_data_model.htm

l

Data	
  sources

1.	
  U.S.	
  Geological	
  

Survey’s	
  Gap	
  Analysis	
  

Program	
  (GAP)

Online	
  tool	
  to	
  aid	
  in	
  analysis	
  and	
  

retrieval	
  of	
  species	
  distribution	
  

data

Land	
  cover	
  

and	
  

vertebrate	
  

species

http://www.nbii.gov/portal/s

erver.pt/community/gap_onl

ine_analysis_tool/1851

2.	
  National	
  Atlas Provide	
  GIS	
  format	
  data	
  on	
  land	
  

cover,	
  land	
  use,	
  hydrography,	
  

climate,	
  digital	
  elevation	
  models	
  

Varies

http://www.nationalatlas.go

v/atlasftp.html

3.	
  Multi-­‐Resolution	
  Land	
  

Characteristics	
  

Consoritum

Landcover	
  databases Bioregions

http://www.mrlc.gov/mrlc2k

_nlcd.asp

4.	
  Vegetation/Ecosystem

modeling	
  and	
  analysis	
  

Project—VEMAP

Uses	
  historical	
  and	
  future	
  

projected	
  climate	
  data,	
  soils	
  and	
  

vegetation	
  maps,	
  and	
  a	
  number	
  

of	
  process	
  models	
  (Century,	
  

biome-­‐bgc,	
  gtec,	
  lpj,	
  mc1,	
  tem)	
  to	
  

project	
  communities	
  across	
  the	
  

globe

Vegetation	
  

types/biome

s

Kittel	
  and	
  others	
  1995,	
  1996;

http://www.cgd.ucar.edu/ve

map/

5.	
  ClimateWizard Estimates	
  historical	
  and	
  future	
  

temperature	
  and	
  precipitation	
  

changes	
  as	
  absolute	
  or	
  percent	
  

change

Climate	
  

variables

http://www.climatewizard.or

g/

Table 8-1. Continued.

l

g/

s
124 USDA Forest Service RMRS-GTR-285. 2012.

The selection of an appropriate assessment tool depends upon stakeholder objectives 

(see Glick and Stein 2011). Each assessment tool described in table 8-1 varies in how 

it may be applied (spatial and temporal scales) to systems and used for adaptation planning. Assessments meant to inform policy makers need to be focused on a key outcome 

as influenced by multiple stressors (e.g., outcome-based approach described in Luers 

2005), whereas assessments that describe biological-based vulnerabilities or encompass 

multiple outcome variables are likely to be more informative from an ecological and 

research perspective.

Tools that rank species or habitats can provide relatively quick methods for assessing climate change vulnerabilities. However, summarizing the complexity of climate 

change impacts into a single variable may limit the application of these methods (Patt 

and others 2009). Those that rely on species distribution models allow users to visualize 

potential future conditions and responses, which can aid in adaptation planning. Such 

modeling efforts often inform the creation process for indices of sensitivity (Bagne and 

others 2011; Young and others 2011). Caution must be used when selecting and applying 

these models because estimates of future distributions can be biased and users should 

be aware of the limitations of scope of chosen tools (Graham and others 2004). Still, 

those ecosystems that are projected to incur the greatest change should be most vulnerable to climate change. Similarly, ecosystems or species that persist under high annual 

variations in climate, which can be estimated from some of these analyses, should be 

more resilient to climate change. Mechanistic models form the basis of many distribution modeling efforts and are useful for projecting future climate conditions relevant to 

species presence. These tools, as well as those commonly used to guide decision making processes (e.g., conceptual models and statistical decision trees), are often critical 

components of the assessment process.

Assessment Work Within the U.S. Forest Service

The following is research by RMRS and cooperators relevant to the assessment of 

biodiversity and ecosystem function in grassland, shrubland, and desert ecosystems of 

the western United States:

• The U.S. Forest Service is mandated by the Renewable Resources Planning Act (RPA, 

1974) to conduct periodic assessments of forest and rangeland resources; since 1990, 

this includes a requirement to address climate change. RMRS provides technical assistance and analysis for each RPA assessment. The 2000 RPA assessment focused on 

climate impacts to forest systems, and the 2010 assessment was expanded to include 

climate impacts on water and wildlife. To see a complete list of RPA climate change 

publications or further description of ongoing projects, see http://www.fs.fed.us/

rmrs/climate-change/assessments or http:/www.fs.fed.us/rm/landscapes/Research/

Climate.shtml.

• RMRS scientists are developing an index to assess potential effects of climate change 

on biodiversity and wildlife habitat. Contact Linda Joyce (ljoyce@fs.fed.us) or Curt 

Flather (cflather@fs.fed.us) for more information.

• RMRS developed a scoring tool, System for Assessing Vulnerability of Species 

(SAVS) to climate change (Bagne and others 2011), to assess vulnerability of terrestrial vertebrates to climate change. Using this system, managers can prioritize 

actions for species conservation and management. This scoring tool is available at 

http://www.fs.fed.us/rm/grassland-shrubland-desert/products/species-vulnerability/

savs-climate-change-tool/.

• Rehfeldt and others (2006) produced maps (http://forest.moscowfsl.wsu.edu/climate/) 

of current and future vegetation species and biotic communities for North America.
USDA Forest Service RMRS-GTR-285. 2012. 125

Research Needs

In order to develop and improve application of vulnerability assessment tools and 

frameworks to the grasslands, shrublands, and deserts of the western United States, 

research areas should focus efforts to:

• Continue to refine our capacity to identify new community composition; this work 

has the highest priority because of its relevance to inform future management needs 

and best courses of action.

• Improve accuracy of models and methods used to generate climate change predictions 

and habitat suitability maps. This includes continued development and improvement of habitat response models (both mechanistic and correlative) for animal and 

plant species. In addition, distribution models for forest and rangeland habitats and 

species should incorporate dispersal mechanisms.

• Develop and refine systems for assessing plant species vulnerability.

• Develop physiologically based models of species occurrence (see Glick and Stein 

2011).

• Identify measures of species adaptive capacity (Czúcz and others 2009).

• Build tools to identify synergistic effects of climate change, species interactions, and 

other disturbances.

• Integrate management scenarios with scenarios for climate change.

• Identify the appropriate framework for analyzing vulnerability with respect to adaptation strategies, including potential application of existing frameworks (e.g., 

National Center for Ecological Analysis and Synthesis).

• Identify need to develop new frameworks for creating adaptation strategies that integrate vulnerability with management decision processes.

• Complete cost benefit analyses that incorporate multiple scenarios, including the 

validity of inaction as an option. Passive restoration techniques may be more cost 

effective and feasible for many areas (Birch and others 2010) and should be considered among management options.

• Identify and implement methods to make tools more available and useful for decision makers.

Plant Conservation and Restoration

We discuss some of the specific methods and tools used for selecting, collecting, and 

deploying native plant materials to ensure proper conservation of genetic resources. 

These activities provide foundation for, and are particularly relevant to, our future 

capacity under climate change to manage and restore lands with appropriate genetic 

materials.

Approaches and challenges for selecting native plant material

Historically, restoration activities made use of “off the shelf,” agronomically developed, introduced plant materials to fill specific needs (see Monsen and others 2004). 

This was particularly true in the Intermountain West where semi-arid and arid lands 

were often especially challenging sites (Monsen and Shaw 2001). These introduced 

species were developed through selection and breeding programs for improved germination, establishment, reproduction, and quality (e.g., palatability or erosion control) 
126 USDA Forest Service RMRS-GTR-285. 2012.

(Monsen and others 2004). Consequently, native plants that often had complex germination requirements and unique establishment criteria were discriminated against with 

little research completed on them. The U.S. Forest Service and other Federal agencies 

are, however, mandated to use genetically diverse, locally adapted native plants to 

maintain or restore self-sustaining ecosystems to protect the services (e.g., soil stabilization, clean water, and forage) they provide (Johnson and others 2010; USDI and 

USDA 2002; USDA 2008). With realization of this mandate, emphasis is now being 

placed on research that identifies functional traits contributing to native plant competitive ability; improves availability of plant materials; reduces plant materials cost; 

improves techniques for identifying and describing site conditions suitable for native 

plants; identifies appropriate species or combinations of species for planting; and identifies effective planting strategies (Call and Roundy 1992; James and Svejcar 2010; 

Johnson and others 2010; Sheley and James 2010).

Paramount for appropriate use of native plants to meet legislative mandates is an 

understanding of the patterns of genetic (adaptive) variation and structure in the morphology, phenology, and reproduction of native plants across varied landscapes. For 

commercial tree species, this is relatively well known, but only a paucity of information exists for most other native plants despite a growing need to better manage them 

(Hufford and Mazer 2003; Johnson and others 2004; Lesica and Allendorf 1999).

When genetic variation and structure are understood, species-specific genetic transfer zones (commonly referred to as “seed zones”) can be mapped and transfer guidelines 

can be developed to describe how far plant materials can be moved from their point of 

origin and the risks associated with that movement. To properly understand this genetic 

variation and structure, researchers enlist genecological studies that entail collecting 

germplasm representing the variety of climatic and environmental conditions present 

within a large portion or the entire range of the species. These collections are grown in 

common gardens and evaluated for survival, growth, and reproduction characteristics. 

The described genetic diversity is correlated to climatic variation among collection 

sites through regression models and is mapped to provide seed zones. Although seed 

zones for western conifer species are provided by Rehfeldt (1986), genecological studies and subsequent seed zones for grasses, forbs, and shrubs are more recent and have 

been achieved for only a handful of native species in the western United States (e.g., 

Darris and others 2008; Doede 2005; Erickson and others 2004; Horning and others 

2010; Johnson and others, submitted; Kitzmiller 2009; Wilson and others 2008). This 

research is difficult, time consuming, and expensive, so genetic information is lacking 

for many native plants of interest to land managers.

When genetic information is lacking, however, the current management paradigm 

is to use plant materials proximal to their point of origin. This “local is best” prescription is supported by a plethora of studies (Johnson and others 2010; Rice and Knapp 

2008), but a major disadvantage is defining “local” (McKay and others 2005) and often 

this paradigm is more conservatively restrictive than needed. Fortunately, a number of 

climatic and biogeographic tools can be used as surrogates to aid in matching available plant materials to environmental conditions at the planting site. Referred to as 

provisional seed zones, these estimates of genetic appropriateness do not address the 

specificity of adaptation that can vary greatly among species. Thus, provisional seed 

zones are not expected to provide a best fit for all or any species. Their development, 

based on climate and ecological factors, can, however, provide interim science-based, 

decision-making support for land managers until empirical knowledge of adaptive 

variation is obtained and translated into seed zones and transfer guidelines for individual species (Bower and others 2010).
USDA Forest Service RMRS-GTR-285. 2012. 127

Commonly used surrogates to species-specific seed zones are:

• Ecoregion maps (Bailey 1995, 2009; Omernik 1987) that consider floristic regions, 

soils, and other parameters. Subdivision level can be selected to provide broad or 

narrow zones.

• USDA Cold Hardiness Zones are useful for species with distributions limited by 

minimum temperatures (Cathey 1990).

• Plant adaptation region maps (Vogel and others 2005) that combine ecoregions with 

USDA Cold Hardiness Zones.

• Climatic models (Bower and others 2010) that combine multiple climatic variables.

• Focal point models that combine biogeoclimatic characteristics of a region and indicate degree of similarity between potential seed collection and planting sites.

• The Data Extraction Tool (Gearrard and others 2006) that permits users to extract 

information from a number of data layers.

• The Center for Forest Provenance Data, an online database that archives data from 

long-term provenance tests and seedling genecology tests. The database currently 

includes only tree data but may eventually be expanded to include other species (St. 

Clair and others 2010).

• Seed Zones for Native Plants, an online mapping application for provisional and 

species specific seed zones for plant materials development, gene conservation and 

native plant restoration (USDA FS WWETAC 2011).

• An online seed transfer decision-support tool (e.g., Seedlot selection tool http://

www.fs.fed.us/ccrc/tools/seedlot.shtml) to aid in selecting appropriate seedlots that 

can be applied to multiple species using multiple climatic variables and various climate change scenarios (B. St. Clair, personal communication).

• Online databases, such as the Web Soil Survey (USDA NRCS 2009), the Ecological 

Site Information System (USDA NRCS 2010), and climatic databases such as 

PRISM (Prism climate group 2010), can aid in describing biotic and abiotic characteristics of seed origin and planting sites.

In the Interior West, current genecological-based and provisional seed zone mapping efforts illustrate the climatic complexities associated with western ecosystems; 

they are much more complex than those found in the eastern half of the United States. 

Therefore, impacts of climate change and resulting efforts to manage plant communities will be more difficult in the West, particularly the Intermountain West, as 

boundaries on seed zone maps diverge from the environmental conditions used to create them. Success where underlying conditions are most complex, however, should 

readily translate to less complicated systems.

Managing Collections of Genetic Materials Within Species-Specific and 

Provisional Seed Zones

An ecological approach to providing plant materials for use within species-specific 

or provisional seed zones requires that multiple seed collections of a species be made 

from diverse locations within the zone, each representing multiple parent plants. Once 

pooled, the progeny from these collections provide genetically broad-based materials, 

maximizing the likelihood that some seeds will be pre-adapted to planting site conditions and capable of adapting to future environmental fluctuations, including climate 

change. This approach also minimizes the potential for inbreeding and outbreeding 

depression (Johnson and others 2010; McKay and others 2005; Withrow-Robinson 

and Johnson 2006).
128 USDA Forest Service RMRS-GTR-285. 2012.

Selecting Native Plant Material Under Future Climate Scenarios

Climate change may also require movement of genetic material to locations where 

it currently does not exist. Such anthropogenic movement, referred to as assisted colonization, assisted migration, or managed relocation, may be necessary because climate 

change is occurring more rapidly than species can adapt and/or disperse along environmental gradients (Warren and others 2001), or anthropogenic activities have narrowed 

or disrupted natural dispersion corridors (Marris 2008; Minteer and Collins 2010).

Assisted Colonization

Assisted colonization can be accomplished at two levels: (1) moving discrete genetic resources of a species into a new area already occupied by that species (e.g., 

moving seeds of warmer ecotypes into areas currently occupied by colder ecotypes), 

or (2) moving genetic resources of a species into areas where that species does not 

currently exist.

The first scenario attempts to augment current genetic diversity. For example, a collection of seeds from a high-elevation seed zone could be augmented with seeds from 

a lower-elevation seed zone in anticipation that the higher-elevation site will become 

warmer because of climate change. This approach leverages diverse genetic mixtures 

from within seed transfer zones by incorporating genetic material from adjacent seed 

transfer zones; leading to the expression of new desired traits.

The second scenario, introducing species into areas where they currently do not 

exist in order to facilitate their continued existence in response to climate change, has 

become a lightning rod among ecologists and conservationists. Opponents of assisted 

colonization cite potential for unintended and unpredicted consequence on the recipient ecosystem, such as creation of new invasive species, disruption of evolutionary 

and ecological processes at the reintroduction site, and negative genetic interactions 

between relocated and native populations (Ricciardi and Simberloff 2009; Seddon and 

others 2009; Vitt and others 2009). Fazey and Fischer (2009) argued that assisted colonization is a short-term fix that ignores causal reasons for plant extinction, and Sandler 

(2010) stated that ethical, philosophical, and socioeconomic values may not be a justifiable method for preserving, through assisted colonization, the value of a species. 

Proponents purport that such harmful consequences are overstated, can be managed 

(Sax and others 2009; Schlaepfer and others 2009), and exceed the consequences of 

species extinction. In fact, such movement is obligate under the Endangered Species 

Act (Shirey and Lamberti 2009). Indeed, many scientists see assisted colonization as 

one part of a multi-faceted solution to conserve and preserve genetic diversity, and 

decision-support matrices have been suggested for such implementation (HoeghGuldberg and others 2008; Hunter 2007; Richardson and others 2009; Vitt and others 

2009).

Assembled Ecosystems

This assisted colonization debate, unfortunately, often fails to recognize the transitory 

nature, in terms of species composition, of functional ecosystems; current ecosystems 

have no historic analogs and will, under climate change, probably not persist (Williams 

and Jackson 2007). Thus, land managers perhaps need not only contemplate moving species to ensure their survival, but contemplate assembling new “ecosystems” 

representing novel species compositions in order to provide ecosystem function and 

vital delivery of ecological services (e.g., clean water, fiber supply, and healthy soil) 
USDA Forest Service RMRS-GTR-285. 2012. 129

necessary to civilization (Minteer and Collins 2010). In addition, climate change, species introductions, and human activities may cause shifts in land use patterns, thus 

requiring land managers to conduct adaptive ecosystem management of drastically 

altered sites (domesticated or severely degraded) back to a naturally sustainable state 

(Hobbs and others 2006; Hobbs and others 2009; Seastedt and others 2008). Both of 

these management activities would require a holistic evaluation to maintain a sustainable suite of symbiotic flora, and fauna are present to ensure sustainability.

Genetic Transfer Work Within the Forest Service

The following is ongoing research by RMRS and cooperators in plant materials 

development and use in grassland, shrubland, and desert ecosystems of the western 

United States:

• Delineation of provisional seed zones based on biogeoclimatic factors.

• Genecological studies of widespread native grass and forb species.

• Increase of genetically diverse, locally adapted stock seed of native forbs and grasses 

for provisional and species-specific zones.

• Evaluation of native species existing in long-established stands of exotic species as 

potential competitive native plant materials (rapid evolution research) (Leger 2008; 

Mealor and others 2004).

• Identification of selective climatic gradients of importance to big sagebrush distribution and development of climate responsive seed zones for the entire range of big 

sagebrush.

• Design of a website tool for managers to match big sagebrush seed sources to restoration sites.

Research Needs

The following are research areas for developing genetic transfer guidelines to mitigate climate change impacts in grasslands, shrublands, and desert ecosystems of the 

western United States:

• Develop risk assessment tools for selecting seeding and planting sites to reduce negative impacts and the incidence of failures.

• Continue development of provisional and species-specific seed zones and seed transfer guidelines.

• Refine tools for identifying and mapping future environments suitable for these 

species.

• Provide recommendations for developing seed production areas of genetically diverse 

populations pre-adapted to climatic change and other environmental perturbations.

• Examine autecology and adaptive characteristics of key restoration species and species at risk from climate change and other biotic and abiotic stressors (species that 

are long-lived, inbreeding, or characterized by small or disjunct populations or species with low genetic variation and rare species).

• Research and develop approaches for managing genetic variation to influence plant 

response to climate change; enhance and conserve genetic diversity within seed 

zones; and promote natural migration, gene flow (establish outlier populations) and 

assisted migration.
130 USDA Forest Service RMRS-GTR-285. 2012.

• Examine completed research on native species and species specific seed zones for 

generalizations regarding such areas as specificity in environmental requirements, 

capacity for in situ adaptation to climate change, and potential rates of migration.

• Provide for ex situ and in situ conservation.

• Develop a simple, readily accessible tool for nursery managers, seed producers, 

and land managers to help them move plants across the landscape in a genetically 

appropriate manner to conserve genetic diversity, facilitate current management decisions, and provide a foundation for reaction to climate change.

• Investigate the intersection of socioeconomic, environmental, and philosophical debate toward a better understanding of the difficult decisions associated with assisted 

colonization of plants and animals to new locations. A decision support matrix that 

conceptualizes and quantifies the advantages and disadvantages of assisted colonization is required. Use paleobotanic and paleoclimatic data to further understand 

and model plant community evolution from the last glaciation to contemporary associations, and how those processes can be leveraged toward ensuring development 

of new, non-analogous ecosystems under evolving climate conditions.

RMRS Expertise and Partners

The GSD Program includes a cadre of scientists and their collaborators working 

with wildland restoration from plant selection, to seed increase, nursery stock production, outplanting, monitoring, and management. Multiple partners are essential for 

progress due to the large number of plant species and variety of landscapes involved 

as well as the multidisciplinary nature and immense time commitment of the research. 

Forest geneticists in RMRS and Pacific Northwest Research Station are now providing 

leadership for non-conifer genecology research and plant response to climate change. 

Our current sponsors, collaborators, and partners include: U.S. Forest Service National 

Forest System Regions 1, 2, 3, 4, 6, and 8, Research and Development, and State 

and Private Forestry; USDA Agricultural Research Service and Natural Resources 

Conservation Service; USDI Bureau of Land Management, National Park Service, and 

U.S. Fish and Wildlife Service; Department of Defense; U.S. Geological Survey; universities; state departments of natural resources; state and private crop improvement 

associations and foundation seed programs; non-profit organizations; and the native 

plant and seed industries.

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