A Synthesis of Climate andVegetation Cover Effects onBiogeochemical Cycling inShrub-Dominated Drylands

Marie-Anne de Graaff,1* Heather L. Throop,2 Paul S. J. Verburg,3

John A. Arnone III,4 and Xochi Campos1

1Department of Biological Sciences, Boise State University, 1910 University Drive, Boise, Idaho 83725-1515, USA; 2Department ofBiology, New Mexico State University, Las Cruces, New Mexico 88003, USA; 3Department of Natural Resources and Environmental

Science, University of Nevada, Reno, Nevada 89557, USA; 4Division of Earth and Ecosystem Sciences, Desert Research Institute,

Reno, Nevada 89512, USA

ABSTRACT

Semi-arid and arid ecosystems dominated by shrubs

(‘‘dry shrublands’’) are an important component of the

global C cycle, but impacts of climate change and ele-

vated atmospheric CO2 on biogeochemical cycling in

these ecosystems have not been synthetically assessed.

This study synthesizes data from manipulative studies

and from studies contrasting ecosystem processes in

different vegetation microsites (that is, shrub or her-

baceous canopy versus intercanopy microsites), to as-

sesshowchanges inclimateandatmosphericCO2 affect

biogeochemical cycles by altering plant and microbial

physiology and ecosystem structure. Further, we ex-

plore how ecosystem structure impacts on biogeo-

chemical cycles differ across a climate gradient. We

foundthat: (1)ourability toprojectecological responses

to changes in climate and atmospheric CO2 is limited by

a dearth of manipulative studies, and by a lack of

measurements in those studies that can explain bio-

geochemical changes, (2) changes in ecosystem struc-

ture will impact biogeochemical cycling, with

decreasing pools and fluxes of C and N if vegetation

canopy microsites were to decline, and (3) differences

in biogeochemical cycling between microsites are pre-

dictable with a simple aridity index (MAP/MAT),

where the relative difference in pools and fluxes of C

and N between vegetation canopy and intercanopy

microsites is positively correlated with aridity. We

conclude that if climate change alters ecosystem struc-

ture, it will strongly impact biogeochemical cycles, with

increasing aridity leading to greater heterogeneity in

biogeochemical cycling among microsites. Additional

long-term manipulative experiments situated across

dry shrublands are required to better predict climate

change impacts on biogeochemical cycling in deserts.

Key words: climate change; elevated atmospheric

CO2; semi-arid and arid ecosystems; biogeochemi-

cal cycles; meta analysis; spatial heterogeneity.

INTRODUCTION

Drylands (arid and semi-arid ecosystems) are an

important, but poorly understood, component of

the global carbon (C) cycle. Although above- and

belowground C concentrations are typically rela-

tively low in drylands, these systems contain a

Received 27 November 2013; accepted 23 February 2014

Electronic supplementary material: The online version of this article

(doi:10.1007/s10021-014-9764-6) contains supplementary material,

which is available to authorized users.

Author Contributions: All authors contributed to conceiving the idea,

compilation of the data and preparation of the manuscript. Analyses were

performed by MA de Graaff.

*Corresponding author; e-mail: marie-annedegraaff@boisestate.edu

EcosystemsDOI: 10.1007/s10021-014-9764-6

� 2014 Springer Science+Business Media New York

substantial amount of total terrestrial C (for

example, nearly 20% of the global soil C pool)

because they cover almost 40% of Earth’s land area

and account for 30–35% of terrestrial net primary

productivity (NPP) (UNDP/UNSO 1997; Field and

others 1998; Lal 2004). Changes in biogeochemical

processes in these systems therefore have the po-

tential to strongly influence regional and global

biogeochemical cycles. For example, many dryland

systems contain a high proportion of shrub cover

and evidence suggests that increases in woody

vegetation in these ecosystems may account for a

substantial, albeit highly uncertain, ecosystem C

sink (Pacala and others 2001). However, despite

their global importance, the responses of biogeo-

chemical cycling in semi-arid and arid shrublands

(hereafter ‘‘dry shrublands’’) to global change fac-

tors such as elevated CO2, precipitation (PPT), and

temperature have not been synthetically assessed.

Understanding the role of dry shrublands in the

global C cycle is particularly critical in light of

anthropogenic stressors acting disproportionally

strongly on these systems. Climate factors in North

American dry shrublands, including temperature

and PPT, may be particularly sensitive to elevated

atmospheric CO2 (Overpeck and Udall 2010) and

ecological processes in these systems may be highly

susceptible to climate change (Weatherly and oth-

ers 2003; Jasoni and others 2005; Weiss and

Overpeck 2005; Notaro and others 2012). Although

dry shrublands across the globe are characterized

by water limitation of ecological processes, pre-

dicting ecological responses in dry shrublands to

climate change is complex due to large regional

differences in PPT and temperature patterns. For

example, the four major deserts in North America,

(Chihuahuan, Sonoran, Mojave and Great Basin

Deserts) differ strongly in timing of PPT (cool vs.

warm season inputs) and in the magnitude of

seasonal temperature extremes. Climate change in

these four deserts is projected to push systems in

different directions with regard to seasonality of

precipitation, but with a nearly common trajectory

of greater water limitation. For example, in the

Chihuahuan and Sonoran Deserts, climate change

is predicted to increase mean annual temperature

(MAT) by 4–5�C and decrease mean annual pre-

cipitation (MAP), with a marked increase in

drought severity and frequency (Seager and others

2007; IPCC 2007a, b; Cayan and others 2010;

Fawcett and others 2011; Notaro and others 2012).

In the Great Basin Desert, PPT is expected to tran-

sition from a winter to a fall/spring dominated re-

gime (NAST Report 2000; IPCC 2007a, b) with

other models suggesting a 10–15% increase in cold

season precipitation (IPCC 2007a, b). These chan-

ges could thus shift the amount and timing of

available moisture. As biological processes in these

drylands are typically already strongly constrained

by moisture availability, changes in biologically

available moisture have strong potential impacts

(Weatherly and others 2003; Weltzin and others

2003; Austin and others 2004; Huxman and others

2004; Aanderud and others 2010).

Climate change has the potential to affect eco-

logical processes through a combination of direct,

climate-mediated changes in plant and microbial

physiology, and more indirectly through climate-

mediated changes in ecosystem structure by

changing the composition and spatial heterogene-

ity of the plant community (Figure 1). Plant

physiological and soil microbial processes may re-

spond very rapidly to changes in abiotic conditions,

and these changes may cause a substantial and

Figure 1. Conceptual model of pathways by which glo-

bal change influences biogeochemical pools and pro-

cesses. A Over short time periods, climate change

(changes in temperature, precipitation, and CO2) may

cause localized changes in biogeochemical pools and

processes by changing plant and microbial physiology. In

spatially heterogeneous dry shrublands, the extent to

which biogeochemical processes respond to these func-

tional changes will likely differ among vegetation mi-

crosites (for example, shrub canopies and intercanopy

grass or bare patches). Functional responses can be ex-

plored using relatively short-term manipulative experi-

ments. B Over longer time periods, prolonged changes in

biogeochemical cycles of C and N following climate

change may alter the proportional cover of the landscape

occupied by different vegetation microsites (‘‘ecosystem

structure’’). C These structural changes may in turn

influence biogeochemical pools and processes. These

patterns may be explored by assessing biogeochemical

differences among vegetation microsites across regional

climate gradients.

M.-A. de Graaff and others

sustained alteration in C and nutrient pools and

dynamics (for example, de Graaff and others 2006).

In dry shrublands, increases in PPT and atmo-

spheric CO2 are predicted to promote plant CO2

assimilation (Housman and others 2006; Ignace

and others 2007; Polley and others 2010; Throop

and others 2012) and subsequent plant and litter

production (Smith and others 2000; Billings and

others 2003; Bachman and others 2010), as well as

enhance soil C and N mineralization rates

(Weatherly and others 2003; Aanderud and others

2010; Sorensen and others 2013). Over longer time

scales (years to decades), climate and atmospheric

CO2 changes can alter C and nutrient fluxes in dry

shrublands through changes in ecosystem structure

in terms of the composition of the plant (Dukes and

Mooney 1999; Smith and others 2000; Bates and

others 2006; Thuiller 2007) or soil microbial com-

munity (Cregger and others 2012; Sorensen and

others 2013). For example, a recent cross-site

analysis of long-term vegetation change in the

Sonoran Desert in southern Arizona showed dif-

ferential responses among plant species and func-

tional types to increasing temperature and drought

conditions (Munson and others 2012). Among the

responses were reductions in perennial grass and

forb cover in mesquite savannas and a reduction in

shrub cover in Larrea dominated shrublands. In

addition to changing the proportional cover of

existing species, changes in PPT and temperature

may cause additions or deletions of species present

in the community through range shifts (Bradley

and others 2009; Bradley 2009). These changes in

the composition of plant communities are expected

to impact biogeochemical cycles of C and nutrients

at a landscape scale, as community composition

strongly regulates biogeochemical cycles (Gill and

Burke 1999; Potter 1999; Norton and others 2004;

Hooker and Stark 2008; Rau and others 2011).

Climate change and elevated atmospheric CO2

may be particularly important in affecting C and

nutrient fluxes in dry shrublands if the relative

abundance of shrubs and shrub-free intercanopy

microsites (‘‘vegetation microsites’’) is altered. In

contrast to the relatively continuous canopy cover

typical of mesic systems, drylands are characterized

by discontinuous vegetation cover, with landscapes

composed of a matrix of shrub canopy and inter-

canopy microsites (Klopatek 1987; Schlesinger and

others 1990). Intercanopy microsites contain an

often temporally dynamic mixture of bare ground,

grasses, forbs, and biological soil crusts (Klopatek

1987; Schlesinger and others 1990; Weatherly and

others 2003). Vegetation microsites differ in

microclimate, with shrub microsites typically have

lower temperatures, greater soil moisture, and

lower solar radiation (Breshears and others 1997;

Scholes and Archer 1997; Schlesinger and Pilmanis

1998), erosion/deposition dynamics, and litter in-

puts to soil surfaces (Hibbard and others 2001) than

bare or herbaceous microsites. This structural het-

erogeneity in vegetation cover enhances the role of

erosional processes and causes strong spatial het-

erogeneity in biogeochemical processes (Schle-

singer and others 1996; Okin and others 2006; Ravi

and others 2007; Throop and Archer 2007).

Developing a predictive, mechanistic understand-

ing of biogeochemical responses to global change in

dry shrublands at the landscape-scale must, there-

fore, include explicit recognition of the strong

spatial heterogeneity in these systems.

Detection of responses of dry shrublands to envi-

ronmental perturbations can be challenging given

the relatively low NPP and slow growth rates of

many dryland plant species. We suggest that a pre-

dictive framework for assessing global change im-

pacts must combine emphasis on physiological and

ecosystem structural responses. Physiological re-

sponses may be deduced from short-term manipu-

lative climate change studies that focus on assessing

climate change impacts on ecosystem function ra-

ther than structure. In contrast, structural responses

may be deduced from comparative studies among

vegetation microsites, climate gradient studies or

long-term manipulative studies that take into con-

sideration the effects of changes in plant community

composition on ecosystem processes.

The objectives of this study were to synthesize

available data to assess how future changes in cli-

mate and atmospheric CO2 may affect biogeo-

chemical cycling in dry shrublands. We evaluate

these data from several perspectives: (1) we use

data from manipulative global change studies to

evaluate how global change alters plant and

microbial physiology, (2) we use data from studies

that contrast ecosystem pools and processes among

vegetation microsites to quantify climate change

impacts on biogeochemical cycles mediated by

changes in ecosystem structure, (3) we develop

predictive relationships between climate parame-

ters and ecosystem structure impacts on biogeo-

chemical cycles. We hypothesized that climate

change significantly alters biogeochemical cycles

via changes in plant and microbial physiology in

the short-term. We further hypothesized that if

climate change alters ecosystem structure by

changing the proportion of area occupied by

different vegetation microsites, biogeochemical

cycling will be altered because pools and processes

are greater under plant canopy microsites than in

A Synthesis of Climate and Vegetation Cover Effects

intercanopy areas. Finally, we hypothesized that

the relative difference in pools and processes

between vegetation canopy and intercanopy

microsites increases with the relative aridity of an

ecosystem. As a result, if climate change causes

areas to become more arid, biogeochemical cycling

in canopy and intercanopy sites would diverge

while the opposite is true if sites become less arid.

METHODS

Literature Compilation

We compiled published studies on biogeochemical

pools and processes in dry shrubland ecosystems

using a variety of search engines: ISI Web of Sci-

ence, Google Scholar, and online search engines

from libraries at Boise State University and the

University of Nevada. Additional relevant studies

referenced in those returned by the search engines

were also included in the literature compilation.

We restricted studies to those in dryland ecosys-

tems (defined here as MAP £ 500 mm and MAP/

MAT £ 65 mm/�C) where shrubs were the dom-

inant vegetation at the time of study. Although

there was no intentional geographic bias, the

majority of the data were derived from the four

major North American deserts: Chihuahuan, Son-

oran, Mojave, and Great Basin. Each study in-

cluded in our analysis presented data on one or

more commonly measured biogeochemical pool

and/or process. Biogeochemical pool measure-

ments included: soil C content (C; in some cases

this was measured as total soil C, but in soils with

inorganic C, data were typically reported as soil

organic C), soil organic matter content (SOM), soil

total N (TN), soil C:N ratio (C:N), microbial biomass

N (mbN), soil NO3 (NO3-), and soil NH4 (NH4

+).

Biogeochemical process measurements were: leaf

net CO2 assimilation rate (PS), soil respiration

(Cresp), N mineralization (Nmin), leaf litter decom-

position (litter decomp), root decomposition (root

decomp), and root production (root prod).

From the master database of 164 published

studies in arid shrublands, we accumulated fitting

the above criteria, we used 72 studies for further

analyses that met the additional criteria of either

(1) including a global change manipulation of

temperature, PPT/soil moisture, or atmospheric

CO2 concentration that allowed us to assess short-

term functional responses, or (2) presenting data

from measurements made on two or more con-

trasting vegetation types (for example, shrubs and

herbaceous versus intercanopy microsites) within a

study location (hereafter ‘‘microsite analysis’’) that

can be used to explore responses to ecosystem

structural changes. When compiling data from

these two subsets of studies, we took values directly

from published tables or the text whenever possi-

ble. When necessary, we estimated values from

graphical data by hand or using image analysis

software (ImageJ, National Institutes of Health,

Bethesda, MD).

To meaningfully compare experiments, we ap-

plied restrictive criteria to each of the response

variables. For all soil pool data (that is, soil C, TN,

mbN, CN, SOM, NO3-, NH4

+), we included mea-

surements for soil layers ranging in depth from 0–5

to 0–20 cm. When data were reported for several

depths, we included only the results that best rep-

resented the entire top 20 cm of the soil. Data for

mbN were obtained by the fumigation–extraction

method (Vance and others 1987) or the substrate-

induced respiration technique (Anderson and

Domsch 1978). Data on NO3- and NH4

+ pools were

obtained by extraction of soil samples with KCl in

the laboratory or in situ using cation and anion

exchange resin techniques. We included data for C

and N fluxes (that is, Cresp, Nmin) that were obtained

from either laboratory or in situ incubation studies.

Laboratory studies typically estimated potential C

mineralization rates, generally using temperature

and moisture conditions assumed to be optimal for

microbial activity. These measurements were made

in closed microcosms with flux rates estimated

from two or more repeat measurements of head-

space gas concentrations. In situ studies used static

or flow-through chambers to measure CO2 flux

rates from soil surfaces, and thus would include

both microbial heterotrophic and root (autotroph-

ic) respiration (Holland and others 1999). Above-

and belowground litter decomposition data were

obtained from studies that used mesh litterbags to

measure mass loss through time (Harmon and

others 1999), with measurements made at two or

more points in time. For all biogeochemical pool

and process studies from which data were available

from multiple measurement times, we averaged

data across measurement times prior to incorpora-

tion into our final database.

Plant and Microbial PhysiologicalResponses to Global Change

To assess the prevalence of manipulative global

change experiments and their general effects, we

compiled the number of studies for each response

variable and each of the three categories of

manipulative global change experiments (temper-

ature, PPT, CO2). Many studies measured more

M.-A. de Graaff and others

than one response variable and/or multiple soil

types or dominant vegetation (hereafter ‘‘cases’’).

We tabulated the direction of response to the

manipulative treatments for each case in each

study. The limited number of studies did not allow

for a quantitative meta-analysis on these data.

Ecosystem Structure Impacts onBiogeochemical Cycles

To assess how differences in vegetation microsites, as

could occur from ecosystem structure responses to

climate change, affects biogeochemical pools and

processes, we synthesized data from 46 studies that

included measurements from at least two vegetation

microsites (shrub, herbaceous, and bare patches).

We further defined microsites as canopy or inter-

canopy, with canopy microsites consisting of shrub

or herbaceous patches. Within individual studies,

we compared biogeochemical pool and process val-

ues between microsites. Results from microsite

contrasts under different treatments within the same

published study (for example, N, burning, grazing,

and irrigation treatments; plant species or commu-

nities; soil types) were considered independent

measurements. To assess how canopy microsite

vegetation type affects biogeochemical pools and

fluxes, we further categorized the canopy microsite

by dominant vegetation type (shrub or herbaceous)

and contrasted those with their respective inter-

canopy microsites (herbaceous or bare).

We analyzed microsite impacts on biogeochemi-

cal pools and processes with meta-analysis (for

example, Curtis and Wang 1998; Hungate and

others 2009) using MetaWin 2.0 (Rosenberg and

others 2000). For each measurement, we calculated

the response ratio (r) as the percentage change in a

response variable (that is, soil C content, soil or-

ganic matter content, soil total N, soil C:N ratio,

microbial biomass N, soil NO3-, soil NH4

+, soil

respiration, N mineralization, litter decomposition)

between microsites (Eq. 1) .

r ¼ response for canopy microsite

response for intercanopy microsite� 1

� �� 100

ð1Þ

Thus, for response variables where there was no

change between canopy and intercanopy micro-

sites r would equal 0, those with greater values for

response variables in canopy microsites than in-

tercanopies would have positive values for r, and

those with smaller values for response variables in

canopy microsites than intercanopies would have

negative values for r.

Conventional meta-analyses weigh each indi-

vidual observation by the reciprocal of the mixed

model variance (Curtis and Wang 1998). However,

such an analysis requires that the standard devia-

tions of individual studies are known. These data

were not available for a large proportion of the

studies used in our analysis. Thus, we weighted

individual values included in the analysis by

experimental replication (Hedges and Olkin 1985;

Adams and others 1997), assuming that better

replicated experiments did not result in data with

greater variance. We choose this metric because

well-replicated studies provide more reliable esti-

mates of the response of individual variables (for

example, of soil C and N responses; see Hungate

and others 1996; Hungate and others 2009). We

used bootstrapping to calculate confidence inter-

vals on mean effect size estimates for the whole

data set and for categories of studies (that is, shrub

versus herbaceous types; 4999 iterations, Adams

and others 1997). We considered microsite effects

significant if the 95% confidence intervals did not

overlap with zero. In addition, we considered re-

sponse ratios for different microsite types (shrub or

herbaceous plant cover) different from each other if

they varied significantly at the P £ 0.05 level.

Vegetation Microsite Impacts as aFunction of Climate

We further assessed whether relative influence of

canopies on biogeochemical pools and processes

differed across climate gradients. Using simple lin-

ear regression, we regressed the natural log of the

response ratio for microsite contrasts against: (1)

MAP alone, and (2) a simple aridity index (MAP/

MAT in mm/�C) for biogeochemical response

variables with adequate sample size. Because many

papers did not report MAP and/or MAT for the

study sites, we used estimates of 1971–2000 site-

specific means generated by the PRISM model

(PRISM Climate Group 2013). For consistency, we

used PRISM data even when means were reported

for sites; PRISM and reported empirical means

across all sites were within 10% of each other.

Although other more complex aridity indices have

been developed (for example, Middleton and oth-

ers 1997), we elected to use a simple index to limit

the needed climate data to readily available vari-

ables. Climate variables such as potential evapo-

transpiration that are required with more complex

aridity indices are very infrequently reported in

field studies, and there is limited high spatial-res-

olution data available for these variables.

A Synthesis of Climate and Vegetation Cover Effects

RESULTS

Functional Responses to Global Change

Our literature compilation yielded few manipula-

tive global change experiments that have been

carried out in dry shrublands. Of the 164 studies

compiled, 27 studies were in situ manipulative cli-

mate change experiments (Table 1). Carbon

assimilation and leaf decomposition studies were

the most commonly reported response variables in

manipulative global change experiments, with 11 C

assimilation studies and 7 leaf or root decomposi-

tion studies including experimental field manipu-

lations of PPT or CO2. One additional C assimilation

study assessed the response to temperature

manipulations. There were four cases in which C

assimilation responded positively to altered PPT

and four cases with no significant response to the

manipulation. One case showed a negative re-

sponse to PPT. Elevated CO2 enhanced C assimila-

tion in 9 cases, with no response in 1 case. Of 10

cases total, decomposition rate responded positively

to PPT in 7 cases, negatively in 1 case, and there

was no response to PPT in 2 cases. The only other

response variable with at least three studies was

root production in response to elevated CO2. Out of

7 elevated CO2 cases, root production increased in

3 cases, decreased in 2 cases, and did not respond in

the other 2 cases. It is important to note that al-

though we found 4 different published studies on

root production responses to elevated CO2, all of

these studies reported results from the same free air

carbon enrichment (FACE) study conducted in the

Mojave Desert.

Ecosystem Structure Impacts onBiogeochemical Cycles

Results from the meta-analysis showed strong

vegetation microsite impacts on biogeochemical

pools and processes. Many biogeochemical pools

were significantly greater in canopy (shrub or

herbaceous) compared to intercanopy microsites

(herbaceous or bare): soil C (+41%), SOM (+52%),

TN (+59%), microbial biomass N (+59%), and

NH4+ (+144%) (Figure 2A). In contrast, NO3

-

stocks did not differ between canopy and inter-

canopy microsites (Figure 2A). Several biogeo-

chemical process variables related to C and N fluxes

such as C respiration (+37%) and N mineralization

(+83%) were also greater in canopy relative to in-

tercanopy microsites (Figure 2A). However, the

presence of a plant canopy reduced leaf litter

decomposition rates (-47%). We found that sep-

arating plant canopies into ‘‘shrub’’ and ‘‘herba-

ceous’’ categories did not significantly alter the

results of the analysis (Figure 2B, C). The exception

was TN, where shrubs had a significantly greater

TN content relative to their adjacent intercanopy

microsites compared to herbaceous plants in rela-

tion to their intercanopy microsites; P £ 0.05.

Ecosystem Structure Impacts onBiogeochemical Cycles as a Function ofClimate

Canopy microsite influences on the biogeochemical

pool and process response variables were sensitive

to climate in many cases. The aridity index (MAP/

MAT) generally was more strongly related to

Table 1. Summary of Results from the Literature Compilation of Global Change Manipulation Studies

Altered PPT Elevated CO2

Studies Cases Studies Cases

N + - NS N + - NS

Leaf decomp 5 7 1 2 1 0 0 2

Root decomp 1 1 1 0 0 – – –

Root prod./biomass 1 – – 2 4 3 2 2

Soil organic C 1 0 0 8 1 2 2 0

C mineralization 1 2 0 0 1 3 1 0

Soil N 1 0 0 1 1 3 1 0

Microbial biomass N 1 0 0 1 2 6 2 0

C assimilation 6 4 1 4 5 9 0 1

‘‘N’’ is the number of studies assessing each ecosystem process variable for a manipulative experiment of one of the three global change factors. Many studies measured morethan one response variable and/or multiple soil types or dominant vegetation; each of these is referred to as a ‘‘case.’’The number of cases with a positive response to the manipulation is indicated by ‘‘+,’’ cases with a negative response are indicated by ‘‘-,’’ and cases with no significantresponse are indicated by ‘‘NS.’’ PPT manipulations included both additions and reductions, so we categorized ‘‘+’’ as a positive correlation between PPT and the responsevariables. Altered temperature is not included as there was only one dry shrubland warming experiment.

M.-A. de Graaff and others

response ratio than MAT alone, hence we present

aridity index data for all subsequent results. For soil

C pools (Figure 3), vegetation microsite influence

on both soil C and SOM declined with decreasing

aridity (greater MAP/MAT, or relatively wetter

systems). Microsite influence on N pools and pro-

cesses were similarly responsive to changes in

aridity (Figure 4), with TN, Nmin, and NH4+

response ratios all declining with decreasing aridity.

In contrast, there was no significant relationship

between response ratio for NO3- and aridity.

Available data for Nmin, NH4+ and NO3

- were re-

stricted to more arid sites (Figure 4B–D), whereas

TN data spanned a much greater range of climate

conditions (Figure 4A). An inadequate amount of

data from sites across a range of aridity values

prevented meaningful analysis of vegetation

influence on C assimilation, soil respiration,

microbial biomass N, and litter decomposition in

relation to climate.

DISCUSSION

Physiological Responses to GlobalChange

Our literature compilation reveals that projecting

ecological responses to climate change in dry

shrublands is currently severely limited by a dearth

of mechanistic studies that explicitly consider cli-

mate and atmospheric controls over biogeochemi-

cal processes. Above- and belowground biomass

production were rarely measured, although chan-

ges in biomass production, particularly below-

ground, typically lie at the root of any changes in

biogeochemical cycles (Rasse and others 2005). On

the other hand, C assimilation and decomposition

was measured fairly frequently. Perhaps due to the

short-term nature of the manipulative studies,

there was no more than one study reporting results

-100

-50

0

50

100

150

200

C Cresp mbN C:N SOM TN Nmin NO3- NH4

+ decomp

0

50

100

150

200

0

50

100

150

200

Canopy vegetation type: shrub and herbaceous

Canopy vegetation type: shrub

Canopy vegetation type: herbaceous

Per

cent

res

pons

e to

veg

etat

ion

pres

ence

2222

37

27

24

2972

41

79

49

14 14

30

2155 64

41

88

7 817

15

8

a

b

c

Figure 2. Percent difference in total soil C contents (C),

microbial respiration (C-resp), microbial biomass N

(mbN), soil C:N (C:N), soil organic matter content

(SOM), soil total N (TN), N mineralization (Nmin), soil

NO3- (NO3

-), soil NH4+ (NH4

+), and litter decomposition

(decomp) between canopy versus intercanopy microsites:

A shrub or herbaceous microsite versus intercanopy

microsite; B shrub microsite versus intercanopy micro-

site; C herbaceous microsite versus intercanopy micro-

site. Numbers above the bars indicate the number of

observations included in the meta analysis.

Figure 3. Relationship between aridity and the relative

influence of vegetation microsites on soil response vari-

ables: A soil C, and B soil organic matter. The relative

influence of vegetation microsites was calculated as the

natural log of the response value in a plant canopy mi-

crosite divided by that in a corresponding intercanopy

microsite. The aridity index is calculated as MAP divided

by MAT.

A Synthesis of Climate and Vegetation Cover Effects

from experimental manipulations for any of the

pool response variables.

Biological processes are typically sensitive to

temperature, which is confirmed by several syn-

theses of warming impacts on temperate systems

indicating that warming has generally positive

influences on aboveground biomass, NPP, respira-

tion, and photosynthesis (Rustad and others 2001;

Wu and others 2011). However, the relevance of

these results to dry shrublands is unclear as a

majority of warming studies to date have taken

place in high latitude or high elevation sites where

water availability is infrequently limiting biological

activity (Zelikova and others 2012). Warming may

be of primary importance in drylands during the

cool season as biological processes are often effec-

tively shut down during hot, dry periods (Phillips

and others 2006; Verburg and others 2013). A

comprehensive understanding of how ecological

processes respond to temperature changes will re-

quire systematic, mechanistic studies from tem-

perature manipulations in dry shrublands taking

into account different seasonal moisture regimes.

Given that biological activity in arid shrublands

is predominantly limited by water availability, it is

not surprising that more studies have assessed

biological influences of changing PPT (that is, ei-

ther a decrease, increase, or a change in the tim-

ing of PPT), compared to changing temperature.

Most of these studies see significant impacts of

altered PPT regimes on biogeochemical processes

(for example, Austin and others 2004; Huxman

and others 2004; Aanderud and others 2010; Reed

and others 2012; Zelikova and others 2012) sup-

porting the notion that systems with low MAP

respond most strongly to decreased PPT (Wu and

others 2011). Our literature compilation reveals

that with the exception of one study, studies have

focused on PPT reductions rather than additions,

presumably because PPT reductions in drylands

are often logistically much easier to implement

than PPT additions. Furthermore, many studies

have manipulated PPT by proportional changes

relative to ambient PPT (for example, rainout

shelters; Throop and others 2012) or by adding a

predefined amount of PPT (Verburg and others

Figure 4. Relationship between aridity and the relative influence of vegetation microsites on soil response variables:

A total N, B N mineralization, C NH4+, and D NO3

-. The relative influence of vegetation microsites was calculated as the

natural log of the response variable in a plant canopy microsite divided by that in a corresponding intercanopy microsite.

The aridity index is calculated as MAP divided by MAT.

M.-A. de Graaff and others

2013). Fewer studies have manipulated timing of

PPT inputs, which may be a major consequence of

climate change in drylands (NAST Report 2000;

IPCC 2007a, b).

The generally positive response of litter decom-

position to increased PPT suggests that decomposi-

tion is frequently limited by PPT in dry shrublands

as would be expected based on microbial responses

to changes in water availability (for example,

Skopp and others 1990). This finding contrasts with

recent suggestions that dryland litter decomposi-

tion is largely decoupled from PPT and strongly

controlled by abiotic processes such as UV photo-

degradation (Austin 2011). Decomposition re-

sponses to PPT did not lead to corresponding

changes in SOC, with all of the 8 cases of soil C pool

response to PPT manipulations showing no signif-

icant change. However, soil C pool response data

were all from a single published study that assessed

soil C response following 6 years of changes in the

amount and timing of PPT in a semi-arid shrubland

in the Great Basin Desert (McGonigle and others

2005). The lack of a soil C response may be related

to the relatively short duration of manipulations

from published experiments. An increase in the

number of long-term experiments (>15 years)

assessing changes in biogeochemical cycling under

a variety of PPT scenarios is warranted. In addition,

soil C in these systems is typically low, whereas

spatial variability is high making it challenging to

detect relatively small climate-induced ecosystem-

scale changes in soil C stocks. Although bulk soil C

stocks typically respond fairly slowly to environ-

mental change, more labile fractions such as

aggregates or particulate organic matter may be

more appropriate metrics for assessing short-term

changes (Six and others 2000).

Similar to the paucity of temperature and pre-

cipitation field manipulations, limited information

about biogeochemical responses to atmospheric

CO2 manipulations exist for dry shrublands. The

Nevada Desert Free Air CO2 Enrichment (FACE;

Jordan and others 1999) is the only field site in a

dry shrubland in which atmospheric CO2 has been

manipulated. Responses to elevated CO2 appear to

depend on the spatial and temporal scales of

observations and whether data were collected

during time periods (for example, years) of average,

below-average, or above-average precipitation,

which challenges mechanistic interpretation of

long-term exposure effects. For example, consistent

elevated CO2-induced reductions in leaf stomatal

conductance of dominant shrubs, and increased

leaf-level net CO2 assimilation (Naumburg and

others 2003), measured when plants were under

no water stress (for example, Barker and others

2006) may lead to increased individual shoot

growth in the short term (Smith and others 2000;

Housman and others 2006). However, physiologi-

cal responses of vascular plants may only be tran-

sitory and not result in long term changes in

biomass production either aboveground (that is,

after 10 years; Newingham and others 2013) or

belowground (Ferguson and Nowak 2011; New-

ingham and others 2013; Sonderegger and others

2013). These data are largely congruent with field

results from mesic systems on the impacts of ele-

vated CO2 on soil C and N cycling (de Graaff and

others 2006; Luo and others 2006; van Groenigen

and others 2006).

Our analysis focused on vascular plant responses

to climate change, but biological soil crusts are

prominent in some dry shrublands and may re-

spond strongly to changes in climate and atmo-

spheric CO2. For example, a common moss in the

Colorado Plateau was not affected by increasing

temperature, but experienced higher mortality

when rainfall was delivered in smaller pulses that

caused net carbon loss (Reed and others 2012). In

addition, reductions in the abundance of the

cyanobacterial component of crusts observed after

10 years of exposure to high CO2 (Steven and

others 2012) and precipitation-dependent, variable

responses of the moss, lichen, and cyanobacterial

crust components (Wertin and others 2012) could

portend possible long-range alternations in C and N

cycling in these ecosystems. Given the prominence

of intercanopy microsites in dry shrublands, we

suggest that continued research on climate change

impacts on biotic crusts would improve our overall

understanding of climate change impacts on bio-

geochemical cycling in dry shrublands.

Ecosystem Structure Impacts onBiogeochemical Cycles

Global change has the potential to strongly influ-

ence ecosystem processes if environmental condi-

tions affect the type and relative distribution of

microsites. Climate models coupled with biocli-

matic envelope modeling suggest a decline in veg-

etation cover and an accompanying sharp decline

in vegetation C with projected warming and drying

in the southwestern United States (Munson and

others 2012; Notaro and others 2012). In contrast,

in northern dry shrublands, a transition in vege-

tation structure from a shrub to grass-dominated,

largely driven by invasive grasses such as cheat-

grass (Bromus tectorum), has been linked to elevated

atmospheric CO2 and climate change (Bradley and

A Synthesis of Climate and Vegetation Cover Effects

others 2009; Bradley 2009; Smith and others

2000). Increases in temperature are expected to

result in a northern migration of the northern and

southern limits for many sagebrush species with

resulting increased northern migration of Mojave

desert species such as Larrea tridentata (Chambers

and Pellant 2008). Increases in winter precipitation

could result in expansion of woody species and

shifts from shrublands and grasslands to woodlands

and forests whereas decreases in precipitation may

cause decreases in vegetation productivity resulting

in decreases in vegetation cover (Chambers and

Pellant 2008). A major challenge in predicting

vegetation changes especially in northern arid

lands is the co-occurrence of ecosystem responses

to climate change and introduction of invasive

species. The presence of invasive species has pro-

found effects on arid landscapes by affecting fire

cycles and subsequent elimination of native shrubs

but these effects are intermingled with responses to

climate change. In many arid areas in the western

United States plant invasions and woody

encroachment may represent management, rather

than climate-driven changes. Still, our meta-ana-

lysis indicates that changes in vegetation cover can

strongly influence ecosystem processes, particularly

when these changes lead to large increases or de-

creases in the area covered by woody canopies.

Thus, both management and climate change in-

duced shifts in vegetation cover should be taken

into account when discussing future trends in

biogeochemical cycling in arid lands.

Results from our meta-analysis of vegetation

influence on biogeochemical pools suggest that

both shrub and herbaceous microsites have greater

soil C, N, SOM, microbial biomass N and NH4+

pools than intercanopy microsites. The relative

impact of shrub and herbaceous canopy microsites

on soil pools and processes when compared to

their intercanopy microsites was equivalent, except

for total N, where the impact of shrub canopy

microsites was significantly greater than that of

herbaceous canopy microsites. Further, our meta-

analysis showed a trend toward a relatively smaller

impact of herbaceous canopy microsites on SOM

and C stocks when compared to shrub canopy

microsite impacts. These data indicate that a change

in vegetation types can significantly alter biogeo-

chemical pools and cycles. Nonetheless, reported

impacts of woody encroachment on biogeochemi-

cal pools and processes is varied, with studies

reporting increases, decreases, and no change in

soil C and N pools (reviewed in Wessman and

others 2004; Barger and others 2011; Eldridge and

others 2011). Part of this inconsistency may be due

to failure to account for woody plant age, sub-

canopy gradients in SOC concentrations, and land

use history (Houghton 2003; Throop and Archer

2007).

Our meta-analysis results indicate that with an

increase in shrub and herbaceous canopy micro-

sites, N mineralization and soil C respiration rates

will increase, whereas litter decomposition rates

will decrease (Figure 2A). Greater C respiration and

N mineralization in shrub microsites may be related

to greater root biomass. Decreased decomposition

in shrub canopy microsites may reflect changes in

abiotic conditions, litter quality, and/or decom-

poser communities among microsites. Shrub mi-

crosites receive large aboveground litter inputs

composed of relatively labile leaf litter from her-

baceous and woody plants and more recalcitrant

woody material (Hibbard and others 2003). In

addition, shrub canopies can alter the microclimate

for decomposition by reducing temperature and UV

exposure, and altering patterns of soil-litter mixing

(Throop and Archer 2009), all of which may con-

tribute to lower decomposition rates.

Our analyses separated intercanopy microsites

from herbaceous or woody canopies, but did not

discriminate between intercanopy microsites of

bare soil and those with biological soil crust cover.

The composition and cover of these biocrusts are

sensitive to precipitation change (Reed and others

2012; Zelikova and others 2012). Because biocrusts

can play an important role in biogeochemical cy-

cling, infiltration rates, and soil stability (Caldwell

and others 2008) future consideration of composi-

tion changes in intercanopy microsites may be

important.

Ecosystem Structure Impacts onBiogeochemical Cycles as a Function ofClimate

Biogeochemical responses of dry shrublands to

climate change will largely depend on the degree of

initial moisture limitation and other factors that

affect moisture availability (for example, changes in

evapotranspiration as a function of temperature;

Wu and others 2011). A largely unexplored ques-

tion is whether the impacts of a reduction in PPT on

soil C inputs and outputs and the resulting net

ecosystem C uptake are regulated by similar, pre-

dictable mechanisms across drylands, and whether

the magnitude of these changes can be predicted

based on intrinsic climate characteristics. Although

many strong relationships between climate and

ecosystem processes have been reported in more

mesic systems (for example, Sala and others 1988;

M.-A. de Graaff and others

McCulley and others 2005), relationships based on

MAP typically break down in drylands (for exam-

ple, see reported data where MAP < 600 mm for

Knapp and others 2008; Barger and others 2011;

Eldridge and others 2011). However, when we as-

sessed ecosystem process responses relative to a

simple aridity index rather than MAT alone, we

found significant correlations between aridity and

the relative differences in C and N cycling between

shrub and intercanopy microsites. For example,

shrub microsites typically had greater total soil C

and N contents than intercanopy microsites, and

the magnitude of this difference increased with

aridity. This increase in magnitude at more arid

sites may be explained by the fact that intercanopy

areas in more arid climates often have less inter-

canopy herbaceous vegetation. This pattern is

exacerbated by greater surface erosion from inter-

canopy patches with low herbaceous cover, leading

to greater redistribution and accumulation of litter

in shrub patches (Schlesinger and Pilmanis 1998;

Okin and Gillette 2001). Because shrub influence is

calculated as subcanopy value versus intercanopy

value, absolute soil C and N values may still decline

with increasing aridity. Increasing aridity in dry-

lands as a result of climate change will function to

enhance patterns of spatial variability in biogeo-

chemcial pools and processes.

Our simple aridity index provides a crude

assessment of mean annual moisture conditions

and thus provides strong support for relationships

between aridity and differences in microsite im-

pacts on biogeochemical processes. More complex

aridity indices that use additional climate variables

may provide a more nuanced view of these rela-

tionships. Furthermore, incorporation of seasonal

patterns of aridity may provide important insight

into how shifts in temporal patterns of PPT with

climate change may influence biogeochemical

cycling.

Challenges for the Future

In contrast to most experimental manipulations,

natural systems will be influenced simultaneously

by multiple global change factors, and changes are

likely to be dynamic through time. Manipulative

experiments are costly to implement and maintain,

yet they provide the only means to assess climate

change impacts that are relevant at the ecosystem

scale. A common perception has been that arid

ecosystems may not be as responsive to climate

change relative to mesic systems. However,

manipulative studies in dry shrublands have pro-

vided surprising results including interactions

between global change factors and plant invasions

(for example, Smith and others 2000) and decou-

pling of aboveground and belowground physio-

logical responses to climate change factors (for

example, Barker and others 2006; Verburg and

others 2013). Some of these responses were

unanticipated and not represented in ecosystem

models, thereby limiting our ability to predict fu-

ture responses of these globally significant ecosys-

tems to climate change. These data show that

multifactorial experiments with combined tem-

perature and precipitation manipulations in dry

shrublands could provide important insights in the

potential impacts of climate change on biogeo-

chemical cycles. Different scenarios should consist

of a variety of rainfall manipulations, including

changes in the quantity (additions and reductions)

and timing of precipitation, as well as a combina-

tion of both scenarios across time periods and with

varying pulse size (for example, Huxman and

others 2004; Potts and others 2006; Cable and

others 2008).

Climate change is likely to result in a cascade of

responses in dryland ecosystems with rapid (days–

months) microbial and plant physiological re-

sponses to environmental changes such as moisture

pulses (Figure 1). Above- and belowground NPP

are key processes that drive biogeochemical cycles

and ultimately ecosystem structure responses to

changes in climate. Climate-induced changes in the

production, turnover, and activity of plant roots

will play an especially pivotal role in further reg-

ulating the short-term impacts of climate change

on soil C cycling, because roots are a main conduit

of C to soils (Norby and Jackson 2000; Rasse and

others 2005), yet our analysis shows that the im-

pact of climate change on root function and feed-

backs to other ecosystem processes has rarely been

evaluated. Changes in available soil moisture and

atmospheric CO2 have resulted in altered standing

root biomass and root morphology across soil pro-

files in forest ecosystems (Meier and Leuschner

2008; Iversen and others 2008; Iversen 2009). If a

change in climate would increase root branching,

we may expect an increase in the surface area of

roots that actively contributes to inputs of rhizo-

deposits to soil (Groleau-Renaud and others 1998).

An increase in input of root-derived C compounds

can either promote soil C sequestration (Rasse and

others 2005), or could facilitate microbial activity

and priming of SOC (Hoosbeek and others 2004;

Phillips and others 2012) and this response will

depend on the depth of the soil evaluated (de

Graaff and others 2013). Because drylands are

strongly C limited, changes in the quantity of C

A Synthesis of Climate and Vegetation Cover Effects

shunted belowground by plant roots will mediate

the response of soil microorganisms to abiotic

changes and impact changes in soil C and N cycles

perhaps to a greater extent than what has been

found to date in mesic ecosystems. Yet, the re-

sponse of this process to changes in climate so far

has not consistently been evaluated (but see: Phil-

lips and others 2006; Sonderegger and others 2013;

Verburg and others 2013).

Changes in NPP and associated changes in bio-

geochemical cycling can cause longer-term re-

sponses (years–decades) such as changes in

vegetation composition, relative abundances of

canopy versus intercanopy microsites and associ-

ated changes in total soil C and N contents (Fig-

ure 1). So far, very little information is available

with regard to these changes partly because the

duration of manipulative experiments has not been

long enough to address these changes. Conse-

quently, most predictions regarding these processes

are inferred from modeling analyses (for example,

Bradley and others 2009; Bradley 2009; Munson

and others 2012). Future manipulative climate

change studies in dry shrublands should quantify

how changes in climate will affect the key processes

of NPP, and in particular belowground NPP, if we

are to predict accurately how drylands will con-

tribute to the global C cycle in the future. Under-

standing how communities change will be an

important part of these predictions, due to the

impact of microsites on biogeochemical pools and

processes.

ACKNOWLEDGMENTS

We thank Brenda Nieto, Amanda Sills, and Richard

Jasoni for assisting with data collection. This work

was supported by the National Science Foundation

EPSCoR Program (EPS-0814387) with additional

support from NSF DEB 0953864.

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