Drought exacerbates negative consequences of high

Drought exacerbates negative consequences of high-intensity cattlegrazing in a semiarid grassland

SARA SOUTHER,1,4 MATTHEW LOESER,2 TIMOTHY E. CREWS,3 AND THOMAS SISK1

1Landscape Conservation Initiative, Northern Arizona University, ARD Building #56, Suite 130, 1395 South Knoles Drive, Flagstaff,Arizona 86011-0001 USA

2Life Sciences Department, Yakima Valley College, P.O. Box 22520, Yakima, Washington 98907 USA3The Land Institute, 2440 East Water Well Road, Salina, Kansas 67401 USA

Citation: Souther, S., M. Loeser, T. E. Crews, and T. Sisk. 2020. Drought exacerbates negative conse-quences of high-intensity cattle grazing in a semiarid grassland. Ecological Applications 30(3):e02048.10.1002/eap.2048

Abstract. Grasslands managed for grazing are the largest land-use category globally, witha significant proportion of these grasslands occurring in semiarid and arid regions. In suchdryland systems, the effect of grazing on native plant diversity has been equivocal, some stud-ies suggesting that grazing reduces native plant diversity, others that grazing increases or haslittle impact on diversity. One impediment toward generalizing grazing effects on diversity inthis region is that high levels of interannual variation in precipitation may obfuscate vegetativeresponse patterns. By analyzing a long-term data set collected over a 20-yr period in a semiaridgrassland, we explicitly evaluated the role of climate in regulating the effect of cattle grazing onplant communities, finding that climate interacted with grazing intensity to shape grasslandcommunities. Community composition of plots that were intensively grazed varied consider-ably in response to climatic variation and native species richness was low relative to ungrazedand moderately grazed plots. Following a severe drought in 2002, exotic species richnessrapidly increased in the high-intensity grazing plots. While this pattern was mirrored in theother treatments, exotic species richness increased to a greater extent and was slower to returnto pre-drought levels in the high-intensity grazing plots. Overall, moderate grazing, even com-pared to grazing cessation, stabilized grassland communities through time, increased resilienceto drought, and maintained the highest levels of native plant diversity and lowest levels of exo-tic diversity. These findings suggest that grazing, at moderate levels, may support grasslandresilience to climate change in semiarid regions. However, grazing that exceeds tolerances, par-ticularly in combination with extreme climatic events, like drought, can alter plant compositionover relatively long timescales and possibly increase invasibility by nonnative species.

Key words: cattle grazing; climate change; drought; exotic species; plant diversity; resilience;southwestern United States.

INTRODUCTION

Within grassland systems, grazing by large vertebrateherbivores shapes plant community diversity (Floydet al. 2003, Harrison et al. 2003, Hayes and Holl 2003),structure (Adler et al. 2001), and function (Belsky andBlumenthal 1997, Hayes and Holl 2003, Belnap et al.2009), though the magnitude and direction of theseeffects varies considerably along environmental gradi-ents and depends on the strength of grazing pressure(Milchunas and Lauenroth 1993, Bakker et al. 2006,Beck et al. 2015, Irisarri et al. 2015, Liu et al. 2015,Herrero-J�auregui and Oesterheld 2018). Classic grazingtheory (i.e., Milchunas et al.’s [1988] “Generalized the-ory of the effects of grazing on plant community and

structure” and subsequent iterations), supported byempirical evidence, suggests that moderate levels of graz-ing will increase native plant diversity in highly produc-tive, mesic systems with long evolutionary histories ofgrazing (Milchunas et al. 1998, Cingolani et al. 2005).In terms of mechanism, such theories, which incorporateaspects of the Intermediate Disturbance Hypothesis(IDH), posit that a small number of plant species willoutcompete others for limiting resources and eventuallyattain dominance within the system in the absence ofdisturbance (Milchunas et al. 1988). Grazing at interme-diate levels reduces competitive exclusion of rare speciesby decreasing the performance of dominant species.Thus, in grassland systems with long evolutionary histo-ries of grazing, herbivory diversifies selective pressuresthrough time and space, allowing the persistence ofplants with alternative life history strategies (i.e., life his-tories that promote tolerance of herbivory, in additionto those adapted to rapidly acquire light, water, nutrients

Manuscript received 14 June 2019; accepted 21 October 2019.Corresponding Editor: John C. Stella.

4 E-mail: [email protected]

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or other limiting resources). When disturbance is toogreat, however, weedy or ruderal species displace thosethat lack adaptations to high-disturbance regimes, againreducing diversity.Levels of grazing that support high diversity in mesic

systems would be predicted to negatively impact nativediversity in drier, less productive systems lacking a longevolutionary history with large vertebrate herbivores,(Milchunas et al. 1988). Grazing in these systems, evenat intermediate levels, is expected to reduce diversity byeliminating species with low tolerance of herbivory,while providing a lesser benefit in terms of competitiverelease to other species, as competition is considered astronger determinant of community assemblage in highproductivity systems where greater biomass results in aunidirectional competition for light (Kondoh 2001, Har-pole et al. 2016). Within this framework, livestock graz-ing would be predicted to reduce diversity of semiaridgrassland plant communities in the southwestern regionsof the United States. These grasslands are characterizedby low overall annual precipitation and high rainfallvariability, resulting in frequent droughts (Belesky andMalinowski 2016). Evidence from the fossil record andother sources suggest that the number of large vertebrateherbivores was historically low until the introduction ofcattle and sheep by European settlers in the 1800s(Betancourt and Davis 1984, Schwinning et al. 2008).Indeed, 19th century overgrazing has been linked to con-version of grasslands to shrublands (Bestelmeyer et al.2011), soil erosion, desertification (Schlesinger et al.1990), and spread of noxious invasive plants, (Mack1981) when the co-occurrence of high-intensity grazingwith other stressors, particularly drought, sufficientlydegraded perennial plant cover to allow invasion or veg-etative state change. These past changes broadly supporttheories suggesting that high levels of grazing will nega-tively impact native plant diversity. While some contem-porary studies similarly document diversity declinesassociated with grazing in arid and semiarid systems(Floyd et al. 2003, Lezama et al. 2014), many find neu-tral (Jones 2000, Fensham et al. 2014) or even positiveeffects on plant diversity and ecosystem stability (Becket al. 2015, Alberti et al. 2017).One explanation for observed inconsistencies among

grazing studies in semiarid regions is that high levels ofinterannual variability in precipitation potentiallyobscure grazing effects on plant communities (Fuhlen-dorf et al. 2001). A study conducted in a high rainfallyear may yield dramatically different results than whenperformed in a drought year, due to alterations in graz-ing patterns (e.g., changes in forage selectivity and pro-portion of total aboveground plant biomass consumed)and competitive relationships among plants (Carmonaet al. 2012). In these systems, data must be collected overa sufficient time period and sampled with sufficient fre-quency to observe vegetative response to a robust rangeof climatic conditions and to distinguish long-term fromtransient responses of grassland communities. Meeting

these intensive data requirements is imperative, not onlyto better understand grazing impacts on plant diversityin semiarid grassland communities, but also to examinehow these effects may be modified as climate changes.Adaptation of management practices to ensure resis-tance/resilience are particularly important in semiaridand arid regions, where many species exist at the limitsof their physiological tolerances and where mismanage-ment of grazing could hasten conversion to alternative,potentially less desirable, states like shrublands or inva-sive-dominated systems (Schlesinger et al. 1990, VanAuken 2000, Peters et al. 2006, Knapp et al. 2008,D’Odorico et al. 2013, Kettenbach et al. 2017).Here, we examine a 20-yr dataset of plant community

response to experimental grazing in a semiarid grasslandon the Colorado Plateau in northern Arizona. We buildon a previous study, which documented a positive effectof moderate grazing on native plant diversity, whilesimultaneously observing a rapid increase of cheatgrassin intensively grazed treatments following drought (Loe-ser et al. 2007). Adding an additional 10-yr of data andapplying novel analytical techniques permitted by thisextended time horizon, we ask: (1) Do grazing and cli-mate interact to affect native and nonnative plant diver-sity in this semiarid grassland? and (2) How doesdrought shape southwestern plant community composi-tion in grazed systems? In particular, we track frequencyof cheatgrass and other invasive plant species to deter-mine whether previous documented increases in abun-dance resulted in an irreversible shift to an invasiveplant dominated system. Finally, we consider theseresults in the context of a changing climate in order toinform grazing management in arid and semiarid grass-lands.

METHODS

Study area

This study was conducted in a semiarid, high-eleva-tion (2,160 m) grassland bordered by ponderosa pine(Pinus ponderosa P. & C. Lawson) forests in north-cen-tral Arizona on the Colorado Plateau. In this region,precipitation generally occurs during the summer mon-soons (15 June–30 September) arriving from the eastern,tropical Pacific (Gulf of California/Gulf of Mexico) orduring the winter as frontal systems arriving from thenorthern or western Pacific (Gulf of Alaska) in approxi-mately equal amounts. Generally, the wettest month isAugust and the driest month is June (Schwinning et al.2008). During the 20-yr (1997–2017) study period, totalannual precipitation varied between 327 and 694 mmand average annual temperature varied between 6.9°Cand 9.7°C (Table 1). The study area is underlain by Ver-tisol soils, characterized as containing a high proportionof shrink-swell clay, which causes extensive mixing of theuppermost soil horizons. Prior to European settlement,grazing pressure on Colorado Plateau grasslands was

Article e02048; page 2 SARA SOUTHER ET AL.Ecological Applications

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believed to be low, given a paucity vertebrate herbivorefossils and lack of diversity and abundance of dung bee-tles, metrics that correlate with herbivore abundance(Betancourt and Davis 1984, Schwinning et al. 2008).Regional historic accounts indicate that extensive graz-ing by cattle and sheep occurred in Arizona during1870–1890 (Schwinning et al. 2008). There is no indica-tion that this site would have differed from other grass-lands in the region in terms of historical grazingpressure. Cattle stocking rates have steadily decreasedsince the late 1800s, and from the 1950s onward, havebeen maintained at low to moderate densities (seedescription of grazing treatments Experimental design).Other grazers occur red at the site, including mule deer(Odocoileus hemionus), elk (Cervus elaphus), and prong-horn (Antilocapra americana). While cattle fencing mayalso reduce elk presence within fenced areas (Gross andKnight 2000), pellet counts conducted at this site didnot indicate that visitation rates varied significantlyamong treatments (Loeser et al. 2007). However, pelletcounts were not conducted every year of the study, sograzing by wild ungulates within exclosures/enclosuresmay, at times, have been lower than ambient levels expe-rienced by the moderate use areas.

Experimental design

Grazing treatments were established on randomlyselected 1-ha plots located at least 100 m apart in 1997following initial vegetation surveys. At the time of plotestablishment, ranchers managed cattle so that they

consumed roughly 50% of the aboveground biomass,maintaining a stocking rate of approximately 0.5 cow-calf pairs/ha for 14 d/yr (~0.2 Animal Unit Months(AUM)/ha; Loeser et al. 2007). Grazing management inthe study area evolved from a blending of traditional,season-long rest-rotation techniques (Hormay andEvanko 1958), typical of the region, and a more rapidrotation schedule inspired, in part, by the ranchers’exposure to holistic management (Savory and Butter-field 1999; K. Metzger, personal communication). Thisambient grazing method became the moderate grazing-level treatment (MOD) in our study, because it wasintermediate in intensity to complete grazing cessationand the high-intensity treatment. We constructed three1-ha livestock exclosures to serve as a no-cattle grazingtreatment (EXC), and three 1-ha livestock enclosures toincrease grazing levels above ambient conditions. Withinthese high-grazing-level treatment (HIGH) plots, ranch-ers aimed to remove approximately 80% of the above-ground biomass in one short-duration, high-impactevent to simulate herding behavior, where impact ishighly concentrated in space and time, followed by longperiods (in our case 1 yr) with little or no grazing. Thestocking rate in HIGH plots was, on average, approxi-mately 200 cow–calf pairs/ha for 12 h/yr annually, withgrazing time never exceeding 24 h (~3.3 AUM/ha; Loe-ser et al. 2007). Periodically (e.g., in the drought year,2002), drought conditions reduced vegetation to such anextent that cattle were transported to another site andthus did not graze the MOD or HIGH plots. Actualstocking rates, therefore, varied between 0 and 0.5 cow–

TABLE 1. Climate conditions at the Reed Lake field site for all years in which plant surveys were conducted during the 20-yr studyperiod from 1997 to 2017.

Annual mean growingseason temperature

(°C)January mean

temperature (°C)July mean temperature

(°C)

Year Total annual precipitation (mm) Maximum Minimum Maximum Minimum Maximum Minimum PDSI June

1997 397.2 15.7 �0.7 4.0 �8.5 27.8 7.5 �2.91998 694.1 14.7 �0.9 6.5 �7.2 27.7 10.3 1.51999 400.2 16.7 �1.0 9.6 �7.6 24.7 11.0 �3.12000 391.3 17.1 0.2 8.4 �7.4 28.7 9.7 �4.52001 446.3 16.8 �0.2 4.1 �9.5 27.1 10.2 �1.82002 327.3 17.3 0.0 6.8 �7.9 29.1 13.0 �6.62003 454.1 17.0 0.7 10.7 �5.0 29.7 12.0 �4.62004 600.9 15.5 �0.3 4.4 �8.0 27.3 8.9 �4.62005 610.5 16.2 0.2 5.2 �5.7 29.2 10.0 4.62006 396.4 16.5 �0.3 7.7 �9.4 27.4 12.4 �5.32007 443.9 16.9 0.1 3.0 �10.3 28.6 12.0 �5.32008 479.1 16.6 �0.2 2.8 �11.3 27.4 12.0 �1.52011 525.8 16.2 �0.6 7.0 �10.4 27.5 10.6 �2.92014 525.4 17.5 1.0 10.4 �6.4 27.6 11.8 �4.82017 439.7 17.6 1.9 3.0 �5.6 27.4 12.1 �0.8

Notes: Annual climate summaries, as well as mean temperature conditions for the coldest (January) and warmest (July) month ofthe year, are shown. The final column displays the Palmer Drought Severity Index, (PDSI), a measure of relative dryness that variesbetween �10 (dry) and +10 (wet), for the historically driest month of the year, June. The year 2002 has set in boldface type to high-light the high temperatures and low precipitation that resulted in drought that growing season.

April 2020 GRAZING AND DROUGHTALTERGRASSLANDS Article e02048; page 3

calf pairs/ha for 14 d/yr in the moderate grazing treat-ment and 0–400 cow–calf pairs/ha for 12 h/yr in the high-intensity treatment. Cattle breached one of the threeexclosures twice during this 20-yr period. However, ineach case, the exclosure was grazed for no more than a24-h period, resulting in a grazing intensity below mod-erate levels.To monitor vegetation change through time, we estab-

lished Modified-Whittaker plots (Stohlgren et al. 1995).The Modified-Whittaker plot design included a series ofsubplots of exponentially increasing size, in this case 1,10, and 100 m2, nested within a 1,000-m2 plot, thusallowing researchers to examine how metrics of speciesdiversity varied as a function of sample area (Fig. 1).Plant surveys were conducted annually until 2008, whensurvey frequency was reduced to once every 3 yr. Vegeta-tion data were collected in July, following the onset ofthe monsoonal rains. While vegetation sampling at addi-tional time points during the year to capture intra-annual variation in vegetation due to species-specificphenology would be ideal, such a sampling regime wasnot tractable over this 20-yr study period. For this rea-son, sampling was timed to capture peak biomass pro-duction in the region, at a time when both cool andwarm season grasses were present. At plot sizes >1 m2,simple incidence data were collected in the form of spe-cies lists. At the 1-m2 subplot level, a 50-point frequencyframe was used to assess ground cover and to provide amore structured estimate of species occurrence. Werepeated vegetation surveys for three replicate blocks foreach treatment (Data S1).

Data analysis

Diversity metrics and cover analyses.—In order to ana-lyze data on species diversity through time, Hill numbersor the effective number of species of the diversity orderq = 0 (species richness), 1 (exponential Shannonentropy), and 2 (inverse Simpson index) were calculatedat the plot level for all years and all treatments using thepackage iNext in R (Hill 1973, Hsieh et al. 2016). Spe-cies were scored as present/absent within the 1-m2 sub-plots and incidence data were used to calculate diversityindices for each replicate plot. In addition to observedHill numbers, we extrapolated these parameters toasymptotic species diversity estimates using sample com-pleteness, which estimates false absences based on thenumber of rare species within an assemblage (Chao et al.2014). Standardizing diversity metrics by extrapolationmay improve comparisons among assemblages, becauseassemblages vary in terms of their species pools, anddiversity estimates are highly sensitive to sample size andcompleteness (Chao and Jost 2012).Once Hill numbers were calculated, we used linear

mixed effects models (R package lme4, function lmer) todetermine whether grazing treatment, year, or the inter-action of these two factors affected species diversity.Within these models, plot was treated as a random effect

and the equation structured to account for the repeatedmeasurement of these plots through time (n = 3). Mod-els were fit using restricted maximum likelihood(REML) due to small sample size. Because species rich-ness and percent cover calculations do not require ameasure of evenness, we were able to calculate richness/cover at the 1-m2 subplot level, and include these data asa subsample of the larger plots in the linear mixed effectsmodels (n = 3). For these analyses, model main effectsincluded treatment, year, and the interaction of thesefactors, with subplot nested within plot included as arandom effect. Pairwise comparisons of treatment effectswithin each year of study were conducted using Tukey’sHonestly Significant Difference (HSD) post hoc test (Rpackage lsmeans).We used multiple regression to determine if climatic

factors drove interannual variation in native and exoticspecies richness. In order to account for the dependencyof diversity on the vegetative state of the community theprevious year, we calculated the change in species rich-ness from the previous year for each treatment as thedependent variable for all analyses. Initial attempts tobuild explanatory models using backward and forwardmodel selection resulted in overly complex, uninforma-tive models. Based on ecological studies of the regionand given the observable, profound effects of the 2002drought on vegetation, we used the Palmer DroughtSeverity Index (PDSI), calculated for the driest monthof the year, June, as the explanatory variable in the finalmodels. The PDSI combines temperature and precipita-tion data to estimate relative dryness in relation to base-line data, in this case 1990 to 2000. PDSI is standardizedto vary between �10 (dry) and +10 (wet). For eachmodel, we explored both linear and polynomial fits. The2002–2003 drought-recovery transition was an outlierfor both native and exotic species richness for all treat-ment groups. Because we were interested in assessingdeterministic climatic drivers of species richness with thisanalysis, the 2002–2003 transition, which bracketed thisanomalous drought event, was excluded from models.

Community-level analyses of the 2002 drought event.—Asevere drought in 2002 had observable, negative impactson vegetation. In order to examine the effect of thisdrought event, we compared community compositionfor each treatment at three time points: (1) 2001, the pre-drought year, (2) 2003, the post-drought year, and (3)2017, the final year of the study. To test for communitycompositional differences, we pooled survey data (trea-ted as abundance data, see Royle and Nichols 2003) col-lected in the 10 1-m2 subplots to the plot level, andperformed a PERMANOVA analysis using 1,000 per-mutations and Bray-Curtis distance as a measure of dis-similarity in the R package vegan with the adonisfunction. To visualize differences in community compo-sition among treatments, we used nonmetric multidi-mensional scaling (NMDS; R package vegan), whichcondenses information on community composition into


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a pair of synthetic orthogonal axes. Within this ordina-tion space, proximity indicates similar species composi-tion. We calculated the contribution of individualspecies to the Bray-Curtis dissimilarity distance betweencommunities using function simper in R package veganto determine which species drove differences in pairwisecomparisons of plant communities.

RESULTS

Species diversity changes among grazing treatmentsthrough time

Interannual variation, presumably driven by climate,affected all metrics of native species diversity, includingrichness (q = 0; F14,134 = 21.162, P < 0.001; F14,134 =14.494, P < 0.001; observed and extrapolated, respec-tively), exponential Shannon entropy (q = 1; F14,134 =11.560, P < 0.001, F14,134 = 2,795.980, P < 0.001;observed and extrapolated), and inverse Simpson index(q = 2; F14,134 = 11.017, P < 0.001, F14,134 = 202.130,

P < 0.001; observed and extrapolated). Grazing, however,did not affect measures of native plant diversity, with theexception of observed species richness (q = 0), in whichthe effect of grazing on species richness depended on year(F28,134 = 1.897, P = 0.009). Overall, diversity metricsq = 0, 1, 2 measured at the plot level were highly corre-lated (Appendix S1: Fig. S1). We were able to measurespecies richness with greater precision than other diversitymetrics by sampling subplots in a structured manner.When subplots were treated as samples nested withinplots, the effect of grazing on native species richness againdepended on year (F28,134 = 1.793, P = 0.007; Fig. 2) aswas observed in plot-level estimation of species richness;however, a statistically significant main effect of grazingwas also detected (F2,134 = 4.637, P = 0.012) in additionto the main effect of year (F14,134 = 79.329, P < 0.001).The interactive effect appears to be driven by species rich-ness in the high-intensity grazing treatment, which variedin response to interannual differences to a greater extentrelative to the other treatments. When high-intensity graz-ing is removed from the model, main effects of grazing

High-intensity grazing

No ca�le grazing

Moderate grazing

20 m

5 m

5 m

2 m

50 m

20 m

2 m0.5 m

FIG. 1. Schematic illustrating Modified-Whittaker plot design for vegetation surveys, and the approximate location of plots in asemiarid grassland in Arizona. Within the inset plot design graphic, dashed lines on the left side of the figure indicate fence linesand solid lines indicate positioning of the survey plots. Plots were established in the center of the exclosure/enclosure areas to avoidedge effects. On the right side of the inset figure, solid lines demarcate survey areas, which included 10 1-m2, two 10-m2, and one100-m2 subplots nested within the entire 1,000-m2 survey area.


(F2,134 = 5.115, P = 0.032) and year (F14,134 = 22.574,P < 0.001) are still apparent, but not the interaction(F28,134 = 0.940, P = 0.516). Regardless of whether thehigh-intensity grazing treatment was included in themodel, native species richness varied dramatically amongyears, reaching a low in 2002 when an extreme droughtoccurred (Table 1), but rapidly rebounding to pre-2002levels the following year (Fig. 2). Generally, native speciesrichness was higher in the moderate grazing treatment, rel-ative to the exclosure and the high-intensity grazing treat-ment (Fig. 2).We also examined changes in exotic species richness

using data collected at the subplot level. As in the caseof native species, exotic species richness varied throughtime (F14,134 = 502.320, P < 0.001; Fig. 2). Again, theeffect of study year on exotic species was likely driven byclimate and depended on grazing treatment(F28,134 = 9.060, P < 0.001; Fig. 2). In the drought year,2002, exotic species richness across all grazing treat-ments was considerably reduced from the previous year.Following this drought, however, exotic species richnessreached a zenith across all grazing treatments, thoughthe magnitude of increase varied by treatment. Differen-tial rates of increase among treatments resulted in nearlytwice the exotic species richness in the exclosure andhigh-intensity plots relative to the moderately grazed

plots following the drought. Increase of exotic speciesrichness was greatest in the high-intensity grazing treat-ment plots, in which exotic species richness increasedby nearly 11 times the previous year’s average, followedclosely by increases in the exclosure plots of aroundseven times the 2002 average. In moderately grazedareas, exotic species richness increased at a significantlylesser rate, though this change represented a nearly five-fold increase from 2002 levels. While moderately grazedplots returned to pre-drought numbers of exotic speciesby 2004, the exotic species richness in the exclosure andhigh-intensity grazing plots remained elevated for anadditional 2 yr before declining. Notably, however, exo-tic species richness returned to pre-drought levels for alltreatments by 2006. Differential response through timeamong grazing treatments was again apparent in thelatter part of the study (2008–2017), during which timethe high-intensity grazing treatment increased in exoticspecies richness relative to both the moderate andexclosure plots. In general, moderate grazing plotsexhibited the lowest exotic species richness(F2,134 = 38.904, P < 0.001) and richness was less vari-able, whereas the high-intensity grazing plots had thehighest exotic species richness and experienced pulsesof exotics depending on, presumably, annual climaticconditions (Fig. 2).

b) Exotic

a) Native

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2 011 2014 2017

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

2

4

6

8

0

1

2

3Spec

ies

richn

ess

Treatment

EXC

HIGH

MOD

FIG. 2. (a) Native species richness and (b) exotic species richness varied by grazing treatment through time. Values are mean �SE. The y-axes for native and exotic species are on different scales, since native species were more abundant in the system. In 1997,the pretreatment year, native species richness did not differ among treatments, but exotic species richness was higher in the exclosure(EXC) plots by chance. By 2000, moderately grazed (MOD) plots had significantly higher native species richness relative to exclo-sure and high-intensity grazing (HIGH) plots, and maintained higher levels of richness relative to one or both of these treatments inmost years. In terms of exotic species richness, generally, moderate-grazing plots had the fewest exotics and the high-intensity graz-ing plots the most, while exclosure plots possessed numbers of exotic species intermediary to both grazing treatments.


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In order to speculate on response of plant communi-ties to increases in aridity projected by climate models,we related changes in native and exotic plant diversity tothe Palmer Drought Severity Index (PDSI). An index ofdrought, PDSI, calculated as the departure of evapora-tive demand in the month of June from mean conditionsat the study site, varied between a minimum of �6.6 to amaximum of 4.6 on a scale of �10 to 10, with negativenumbers indicating drier conditions (Table 1). JunePDSI appeared to be a driver of native, and to a lesserextent, exotic species richness in this system (Fig. 3). Inthe exclosure plots, relative dryness explained as muchas 73.5% of the variation in native species richness withdeviations from average conditions in the 1901–2000base period, either drier or wetter, reducing richness(Fig. 3). Generally, across all grazing treatments, decli-nes in native species richness were observed in responseto increasing relative dryness. Notably, climate haddecreasing predictive power as grazing intensityincreased, amplifying variability of native species rich-ness response to climate.

Community-level differences among grazing treatmentsthrough time

Given the effect of the 2002 drought on species rich-ness, we narrowed our analysis to specifically examinecommunity composition before and after the drought(years of analysis = 2001, 2003, pre- and post-droughtyears, and 2017, the final year of the study) for eachgrazing treatment. Pre- and post-drought plant commu-nities were not significantly different in the cattle exclo-sure and moderately grazed plots (F2,8 = 2.062,

P = 0.105, stress = 0.131; F2,8 = 1.204, P = 0.332,stress = 0.124, respectively; Fig. 4). However, droughtdid alter plant community composition in high-intensitygrazing plots (F2,8 = 2.826, P = 0.033, stress = 0.149,Fig. 4). As of 2017, compositional differences persisted,and the 2017 plant community aligned in an intermedi-ate position between pre- and post-drought years alongNMDS axes, suggesting a slow return to pre-droughtconditions. These differences in plant community com-position in the high-intensity grazing treatment preced-ing and following the drought were driven by changes inabundance of both native and exotic species. Exotic spe-cies, Sisymbrium altissimum L. and Bromus tectorum,explained 15.7% and 4.1% of the variation between the2001 and 2003 plant communities, respectively, and werethe first and third most important species driving com-munity differences over this time period. To put thesechanges into context, subplots contained an average ofzero S. altissimum and 9.3 B. tectorum detections in2001, with this number increasing to 116.7 and 39.7,respectively, in 2003. Over the long term, from 2001 to2017, changes in the abundance of the native grass Ely-mus elymoides (Raf.) Swezey, which decreased from anaverage count of 147.7 to 30 detections per subplot inthe final year of the study, explained 16.6% of the dis-similarities between communities (Table 2).

Changes in individual native species cover through timeamong grazing treatments

In order to better understand vegetative communitychange through time, we examined the role of grazingon individual species through time. Consistently across

R 2 = 0.74, P < 0.001

R 2 = 0.44, P = 0.05

R 2 = 0.70, P = 0.002

R 2 = 0.17,P = 0.39

R2 = 0.59, P = 0.01

R2= 0.42, P = 0.06

Exotic

EXC

Exotic

MOD

Exotic

HIGH

Native

EXC

Native

MOD

Native

HIGH

5.0 2.5 0.0 2.5 5.0 5.0 2.5 0.0 2.5 5.0 5.0 2.5 0.0 2.5 5.0

4

2

0

2

4

2

0

2

June PDSI

Mea

n c

han

ge

in s

pec

ies

rich

nes

s

FIG. 3. The relationship between change in species richness from one growing season to the next and the Palmer Drought Sever-ity Index (PDSI) in the month of June for native and exotic plant species in each grazing treatment.


treatments and through time, the dominant native spe-cies in this high-elevation grassland were the perennialC3 grass species Elymus elymoides and Pascopyrumsmithii (Rydb.) Barkworth & D. R. Dewey, the perennial

C4 grass Bouteloua gracilis (Kunth) Lag. Ex Griffiths,the perennial forb Artemisia caruthii Alph. Wood exCarruth, and the annual forb Heliomeris longifolia (B. L.Rob & Greenm.) Cockerell. For all five of these native

–0.4 0.0 0.4

–0.6

–0.2

0.2

EXC; NMDS axis 1

EX

C; N

MD

S a

xis

2a

2001

2003

2017

–0.5 0.0 0.5

–0.5

0.0

0.5

MOD; NMDS axis 1

MO

D; N

MD

S a

xis

2

b

2001

2003

2017

–0.4 0.0 0.2 0.4

–0.6

–0.2

0.2

0.4

HIGH; NMDS axis 1

HIG

H; N

MD

S a

xis

2

c

20012003

2017

FIG. 4. Species composition for (a) exclosure (EXC), (b) moderate (MOD), and (c) high-intensity (HIGH) grazing plots in2001, the year preceding the 2002 drought, 2003, the year immediately post-drought, and 2017, the final year of the study. The dot-ted line indicates the outer hull of the distribution of sites used in NMDS.

TABLE 2. Percent contribution to community dissimilarity of top 10 most influential species to differences in pairwisecomparisons of 2001, 2003, and 2017 plant communities in the high-intensity grazing treatment.

Mean richness

Species Comparison years Time point 1 Time point 2 Contribution (%) P

Sisymbrium altissimum 2001–2003 0 116.7 15.7 0.009Elymus elymoides 2001–2003 147.7 51 13.1 0.039Bromus tectorum 2001–2003 9.3 39.7 4.1 0.025Pascopyrum smithii 2001–2003 89 97.3 3.4 0.8571Artemisia caruthii 2001–2003 53.7 35.7 2.8 0.5644Brassica tourneforti 2001–2003 15.3 0 2.1 0.021Tragopogon dubius 2001–2003 0.7 14 1.8 0.017Bouteloua gracilis 2001–2003 10 6.3 1.6 0.8741Erigeron divergens 2001–2003 1.7 10.7 1.3 0.1009Heliomeris longifolia 2001–2003 6.7 1.3 0.7 0.9141Elymus elymoides 2001–2017 147.7 30 16.6 0.003Heliomeris longifolia 2001–2017 6.7 49.3 5.9 0.1538Pascopyrum smithii 2001–2017 89 125.3 5.5 0.1249Melilotus officinalis 2001–2017 3.7 19.3 2.3 0.0609Brassica tourneforti 2001–2017 15.3 0 2.2 0.006Bouteloua gracilis 2001–2017 10 15.3 2.1 0.3057Artemisia caruthii 2001–2017 53.7 67 1.9 0.8511Bromus tectorum 2001–2017 9.3 14.3 1.6 0.991Aristida purpurea 2001–2017 2.7 9.3 1.2 0.1089Gutierrezia sarothae 2001–2017 0.3 5.7 0.8 0.0569Sisymbrium altissimum 2003–2017 116.7 0 15.8 0.016Heliomeris longifolia 2003–2017 1.3 49.3 6.2 0.038Pascopyrum smithii 2003–2017 97.3 125.3 4.7 0.4306Artemisia caruthii 2003–2017 35.7 67 4.2 0.011Elymus elymoides 2003–2017 51 30 4 1Bromus tectorum 2003–2017 39.7 14.3 3.5 0.1199Melilotus officinalis 2003–2017 0.3 19.3 2.6 0.02Tragopogon dubius 2003–2017 14 0.3 1.9 0.009Bouteloua gracilis 2003–2017 6.3 15.3 1.8 0.7103Erigeron divergens 2003–2017 10.7 1.7 1.3 0.0999

Note: P values indicate the probability of estimating an equal or larger average contribution to community dissimilarity in 1,000random permutations of data.


Vol. 30, No. 3

species, the effect of grazing on their percent coverdepended on year (Fig. 5; Appendix S1: Table S2). Ely-mus elymoides declined markedly in the drought year,2002, in both grazing treatments. To a lesser extent,A. caruthii and B. gracilis (moderate treatment only,exclosure and high-intensity plots exhibited consistentlylow B. gracilis cover) also appear to have been negativelyimpacted by drought conditions, with proportion vegeta-tive cover of these species declining in the post-droughtyear, 2003. Unlike the other dominant grasses in this sys-tem, P. smithii cover increased in 2002 in the HIGHgrazing treatment and was unaffected by drought in theMOD or EXC treatments. Indeed, we begin to observedivergence of the high-intensity grazing treatment fromthe exclosure and moderate grazing treatments in the lat-ter portion of the study in terms of the percent cover ofP. smithii and A. caruthii, with P. smithii occurring inconsistently higher proportions and A. caruthii in lowerproportions in the high-intensity grazing treatment rela-tive to the exclosure and moderately grazed plots.Another notable pattern, the annual plant species,H. longifolia, which occurred in relatively low numbersin the years preceding the drought, rapidly proliferatedfollowing 2003, eventually comprising 40% of the totalvegetative cover in the HIGH treatment by 2011.

Changes in individual exotic species cover through timeamong grazing treatments

Individual exotic species responded idiosyncraticallyto grazing through time. Among exotic species, theresponse of Bromus tectorum, Melilotus officinalis (L.)Lam., Sisymbrium altissimum, and Tragopogon dubiusScop. to grazing depended on year (Fig. 6; Appendix S1:Table S3). This was not the case for Convolvulus arvensisL. and Lactuca serriola L., though there was a tendencyfor the proportion of these species to be influenced bygrazing alone (Appendix S1: Table S3). Bromus inermisLeyss., Erodium cicutarium (L.) L’H�er. ex Aiton, Linariadalmatica (L.) Mill., and Poa pratensis L. were observedin this system in surveys at the 10–1,000 m2 plot levelbut occurred in such low frequencies at the 1-m2 plotlevel as to render the data unanalyzable. Rather thanpersisting continuously in appreciable quantities, theproportion of exotic species increased in pulses and thenreturned to near 0 levels across all species and all grazingtreatments. During expansion events, the rate of increasein proportion of vegetative cover occupied by exoticswas greatest in the high-intensity treatment plots for allexotic species, except for M. officinalis, for which cover-age peaked in the exclosure plots. Moderately grazedplots consistently contained the lowest proportion ofexotic species relative to the exclosures and high-inten-sity grazing treatment plots. The 2002 drought affectedthe proportional abundance of some exotic species, butnot others. Bromus tectorum, S. altissimum, and T. du-bius increased following 2002, while M. officinalisremained virtually absent from the system. In 2011 and

2017, M. officinalis occupied around 15% of the vegeta-tive cover in the exclosure plots, driving the pulse of exo-tics observed in the latter half of the study. While theaverage percent cover of exotics across 30 subplotsreached as high as 30% of the total vegetative cover(Sisymbrium altissimum, HIGH treatment, Fig. 6), thespatial distribution of exotics was patchy, so that exoticswere virtually absent in some plots, but occurred in quitehigh quantities in others. For instance, B. tectorum,M. officinalis, S. altissimum, and T. dubius comprisedup to 75%, 79%, 74%, and 23% of the total vegetation ofindividual 1-m2 plots, respectively.

DISCUSSION

Grazing impacts in semiarid systems

Moderate levels of grazing had an overall positiveeffect on native species richness (q = 1) in this semiaridgrassland, though we did not detect an effect on otherdiversity metrics (q = 1 and 2); a pattern observed inother grazing studies (Lyseng et al. 2018). Species rich-ness weights rare and common species equally, asopposed to exponential Shannon entropy and inverseSimpson, which incorporate a measure of evenness andare therefore more sensitive to changes in abundant spe-cies. Given this, species richness measures may capturegrazing impacts when such effects manifest primarily inthe gain or loss of rare species. Notably, when specieswere analyzed individually, grazing did alter cover ofcommon species in this system, perhaps indicating thatsubsampling increased precision and reduced statisticalnoise for richness and cover measurements, thus permit-ting detection of grazing effects on these metrics.Plots that experienced moderate levels of grazing

exhibited consistently higher native and lower exotic spe-cies richness relative to plots exposed to high-intensitygrazing and from which cattle were excluded. Followingthe 2002 drought, plant communities in the moderategrazing treatment were resilient, native and exotic spe-cies richness quickly returning to pre-drought levels.Other studies conducted in semiarid grasslands supportthese findings, either documenting positive effects ofmoderate levels of grazing on native plant diversity (re-viewed in Jones 2000) and drought resilience (Sprinkleet al. 2006), or little or no increase in diversity followinggrazing cessation (Chew 1982, Valone et al. 2002);though grazing impacts in these systems appear to behighly site specific (Jones 2000). In productive, mesicsystems, an intermediate level of disturbance is pre-sumed to maintain diversity by limiting competitiveexclusion of rare species by species that attain domi-nance under stable environmental conditions (Milchunaset al. 1988, Cingolani et al. 2005, Yuan et al. 2016). Inarid and semiarid grasslands, particularly those that arenot believed to have evolved over long time scales in tan-dem with large vertebrate herbivores, grazing thatexceeds low levels is predicted to decrease diversity


Pascopyrum smithii

Heliomeris longifolia

Elymus elymoides

Bouteloua gracilis

Artemisia caruthii

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

0.00.10.20.30.4

0.000.050.100.15

0.0

0.2

0.4

0.00.10.20.30.4

0.10.20.30.40.5

Vege

tativ

e co

ver (

%)

Treatment

EXC

MOD

HIGH

FIG. 5. Response of the five most common native species in in this grassland to grazing from 1997 (pretreatment) to 2017.Values are mean � SE. Artemesia caruthii is a perennial forb; Bouteloua gracilis is a perennial C4 grass; Elymus elymoides and Pas-copyrum smithii are perennial C3 grasses; Heliomeris longifolia is an annual forb. The y-axes for individual species are on differentscales, since some species were more abundant in the system.

Tragopogon dubius

Sisymbrium altissimum

Melilotus officinalis

Bromus tectorum

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2011 2014 2017

0.000.050.100.150.20

0.00

0.05

0.10

0.15

0.0

0.1

0.2

0.3

0.00

0.02

0.04

0.06

Vege

tativ

e co

ver (

%) Treatment

EXC

MOD

HIGH

FIG. 6. Response of exotic species in this grassland to grazing from 1997 (pretreatment) to 2017 (�SE). The effect of grazing oncover of exotic species, Bromus tectorum, Melilotus officinalis, Sisymbrium altissimum, Tragopogon dubius, varied through time. They-axes for individual species are on different scales, since some species were more abundant in the system.


Vol. 30, No. 3

(Milchunas et al. 1988). As no universal scale for graz-ing intensity has been established, it is difficult to placeour study system in the context of Milchunas et al.’sgrazing intensity–diversity relationships. However, cattlegrazing at current levels in our system, in combinationwith grazing by wild ungulates, whose populations haveexpanded due to development of surface water sources,likely exceeds pre-European settlement grazing pressure(Schwinning et al. 2008). Notably, accurately character-izing prehistoric coevolutionary relationships betweengrassland communities and herbivorous megafauna maybe an impossible endeavor, with some suggesting thatthe intermountain west was likely subjected to similargrazing pressures as elsewhere in the continental UnitedStates (Burkhardt 1996).One explanation for the surprising tolerance of this

and other semiarid grasslands to grazing is that planttraits that confer fitness in arid environments, such assmall stature and drought deciduous leaves, also provideplants with a greater ability to avoid or tolerate her-bivory than that predicted by Milchunas et al. (1988;Coughenour 1985, Adler et al. 2004, Sprinkle et al.2006). A second explanation is that the introduction ofvast numbers of domesticated livestock by European set-tlers in the 19th century may have favored the assem-blage of plant communities more tolerant of grazing(Mack and Thompson 1982, Betancourt and Davis1984, Westoby et al. 1989, Cingolani et al. 2005, Sch-winning et al. 2008). Thus, grasslands in the semiaridwest may have previously transitioned to an alternativestate that differs from pre-European settlement plantcommunities, but is stable through time and resilient tograzing. Indeed, analyses of plant material preserved inpackrat middens suggest a decline in species particularlypalatable to livestock, such as winterfat (Kraschenin-nikovia lanata) and ricegrass (Oryzopsis hymenoides),with a concomitant increase in less palatable, better-defended species, like snakeweed (Gutierrezia sarothrae),rabbitbrush (Ericameria nauseosa), Russian thistle (Sal-sola tragus), and cheatgrass (Bromus tectorum), follow-ing European settlement of the western United States(Schwinning et al. 2008, Fisher et al. 2009). Cingolaniet al. (2005) built on previous theory, suggesting thatsystems with short evolutionary grazing histories wereless resilient to grazing, and hence, more prone to irre-versibly transition to alternative vegetation communitiesand structure. Perhaps, contemporary grazing in thisgrassland provided an overall positive effect to nativeplant diversity via the same mechanism observed in sys-tems with long evolutionary grazing histories: moderatelevels of grazing freed resources, likely water and soilnutrients, by removing abundant species, and preventingcompetitive exclusion of rare species. Generally, weobserved the greatest diversity and cover of exotic spe-cies and annual species (e.g., Heliomeris longifolia) in thehigh-intensity grazing treatment following the 2002drought, further supporting literature that suggests thatintensive grazing will advantage weedy species adapted

to high levels of disturbance (Coughenour 1985, Milchu-nas et al. 1988, Adler et al. 2004).While these findings suggest that livestock grazing, at

intermediate levels, may be an important component ofmanaging grassland plant communities for enhanceddiversity and stability in this region (discussed furtherherein), grazing methodology for achieving this outcomerequires further investigation. While targets for biomassremoval remained constant in this study, these targetswere met by varying both the density of cattle in a plotas well as grazing duration. Future research detanglingthe effects of cattle density and grazing duration ongrassland biodiversity could further improve manage-ment practices. We present findings from a single grass-land in northern Arizona. Examining among-sitevariation in ecological impacts of different levels of bio-mass removal would aid land managers in setting criticalupper and lower bounds on grazing intensity in thesouthwestern United States and permit the developmentof generalizable recommendations for grazing manage-ment in semiarid grasslands. Finally, the observed effectmay differ depending on livestock type and their specificdietary preferences (Li et al. 2017).

Climate and grazing impacts on vegetation

Grazing intensity and climate interacted to shapeplant communities. High-intensity grazing increasedvariability of grassland response to interannual climaticvariation in terms of both native and exotic species rich-ness and overall community composition. In years thatappeared to favor the expansion of exotic species (e.g.,1999, 2001, 2003, 2014; Fig. 2), exotic species richness inthe intensively grazed plots increased more compared tothe response within moderately grazed and ungrazedplots. While the number of exotic species consistentlyreturned to low baseline levels following expansionevents in the first half of the study, exotic richness hasremained elevated in the high-intensity plots relative toother grazing treatments for the last three censuses(2011–2017). It is still unclear whether observed higherexotic richness represents a permanent change in com-munity composition following 2002, or whether it is atransitory response. Droughts and other extremeweather events have been demonstrated to favor exoticsby reducing resistance of the native plant community toinvasion, sometimes producing soil-mediated legacyeffects on vegetation that persist even when rainfallreturns to normal levels (Jim�enez et al. 2011, Diez et al.2012, Meisner et al. 2013). However, abundance of exo-tic species in this system that proliferated following the2002 drought, namely cheatgrass (B. tectorum), tumblemustard (S. altissimum), and yellow salsify (T. dubius),returned to and remains near zero with the exception ofyellow sweet clover (M. officinalis), which has increasedin recent years. The increase in M. officinalis principallyoccurred in the exclosure plots, likely because, as a pre-ferred forage species, M. officinalis abundance was


depressed by grazing pressure in the grazed plots (Currieet al. 1977).Overall, this grassland has demonstrated remarkable

resilience following invasion by species like B. tectorum, aspecies notorious for attaining dominance in other west-ern grasslands (Mack 1981, Rimer and Evans 2006,Sperry et al. 2006, Bradley 2009). One explanation forobserved resilience may be this site’s robust perennialplant populations, which typically comprise 50–75% ofthe total plant cover, and has been shown to be the singlemost important factor in predicting resistance to B. tecto-rum invasion (Bradford et al. 2006, Chambers et al. 2007,2014, Condon et al. 2011). There is some evidence thatthe perennial plant community may be deteriorating inthe high-intensity grazing plots, where there has been areduction in squirreltail (E. elymoides), a native perennialC3 grass, and a concomitant increase in annual species,most notably the native species, longleaf false goldeneye(H. longifolia) and the nonnative species, yellow sweetclover (M. officinalis). This reduction in a native peren-nial grass, in combination with exotic species persisting inlow numbers, may allow for the rapid expansion of non-native species during future drought events.Climate certainly appears to drive grassland response

to grazing (Figs. 3, 4), but predicting plant communitydynamics under future climate conditions in complexsystems is challenging. While overall drier conditionsnegatively affect both native and exotic species, theimpact of aridity on nonnative species appears to be lesssevere, and thus increasing aridity may favor exoticexpansion events like that observed following the 2002drought. Generally, climate change is projected to favorinvasive species, which are often characterized by highphenotypic plasticity. Such plasticity allows exotics topersist in a dynamic climate under conditions that mayexceed the range of climatic variation tolerated by nativespecies (Willis et al. 2010, Davidson et al. 2011). Criti-cally, stochastic events, like the 2002 drought, have thepotential to interact synergistically with grazing torapidly transform grassland communities in ways notpredicted based on past climate–community relation-ships and to produce lasting effects on species composi-tion and abundance.

Climate change in rangelands; implications formanagement

Native and exotic species richness varied considerablyfrom year to year, presumably in response to climatic vari-ation. When change in native species richness wasregressed on a metric of drought, relative drynessexplained approximately 73% of the variation in diversityin the exclosure plots when the post-year drought, 2003,was excluded from the data set. This suggests that, in gen-eral, drier conditions may be expected to reduce nativespecies richness, though relative dryness explained lessvariation in the grazing treatments (particularly in high-intensity grazing plots) andwas a generally poor predictor

of exotic species richness. In semiarid systems, stochasticweather events, particularly droughts, which are projectedto increase as a consequence of climate change (IPCC2013), may be as important determinants of plant com-munity composition as mean changes in precipitation andtemperature. In this study, for instance, the 2002 droughtprecipitated an increase of exotic species diversity andincidence in intensively grazed grasslands, altering com-munity composition to the present day. Predicting preciseoutcomes of increasing drought frequency is challenging,however, as extreme events can shape plant communitiesin unexpected ways that are not simple extrapolations ofpast vegetation–climate relationships.Generally, we observed that moderate grazing main-

tained the highest levels of native species richness, thelowest levels of exotic species richness, and plant com-munities robust to the ecological impacts of drought.Adopting grazing strategies that promote resilience toclimatic change may prevent erosion of native speciesdiversity, while deterring the expansion of invasive spe-cies within semiarid systems. Critically, high-intensitygrazing may select for disturbance-adapted species, andhasten the spread of invasive or weedy species. While thehigh-intensity treatment in this experiment exceeded typ-ical levels managed for in this region, this intensity levelwas not incongruent with recommendations of holisticmanagement and likely occurs at grazing hotspots withinthe landscape (e.g., at water sources). In this way, areasthat experience intense biomass removal may act aspoints of establishment for invasive species, even in casesin which managers maintain stocking rates at moderatelevels. Land managers interested in preventing nonnativespecies spread should target invasive species controlefforts in areas of high-intensity grazing to prevent non-native species establishment and spread to other pas-tures. Grazing management in a dynamic anddirectionally changing climate must be adaptive andresponsive to field conditions (Jakoby et al. 2015).Specifically, stocking rates should be reduced in droughtyears to prevent conversion to exotic dominated grass-land plant communities. While reduction of stockingrates may result in short-term economic losses, theselosses will likely be recouped through enhanced long-term stability of grazing systems.

ACKNOWLEDGMENTS

We are grateful to the Diablo Trust, the Flying M ranch, andthe U.S. Forest Service for their support and collaboration onthis long-term effort. We thank the dozens of undergraduatestudents from Northern Arizona University, Prescott College,and Yakima Valley College that contributed to the fieldwork, aswell as many graduate students at NAU. This study has beensupported through funding from U.S. Department of Energy,the Environmental Protection Agency, U.S. Department ofEducation, Arizona Game and Fish Department, the Ecologi-cal Restoration Institute, the Merriam-Powell Center for Envi-ronmental Research, and the Olajos-Goslow Endowmentfor Environmental Science and Policy at Northern ArizonaUniversity.


Vol. 30, No. 3

LITERATURE CITED

Adler, P. B., D. G. Milchunas, W. K. Lauenroth, O. E. Sala, andI. C. Burke. 2004. Functional traits of graminoids in semi-arid steppes: a test of grazing histories. Journal of AppliedEcology 41:653–663.

Adler, P., D. Raff, andW. Lauenroth. 2001. The effect of grazing onthe spatial heterogeneity of vegetation. Oecologia 128:465–479.

Alberti, J., E. S. Bakker, R. van Klink, H. Olff, and C. Smit. 2017.Herbivore exclusion promotes a more stochastic plant commu-nity assembly in a natural grassland. Ecology 98:961–970.

Bakker, E. S., M. E. Ritchie, H. Olff, D. G. Milchunas, and J.M. H. Knops. 2006. Herbivore impact on grassland plantdiversity depends on habitat productivity and herbivore size.Ecology Letters 9:780–788.

Beck, J. J., D. L. Hern�andez, J. R. Pasari, and E. S. Zavaleta.2015. Grazing maintains native plant diversity and promotescommunity stability in an annual grassland. EcologicalApplications 25:1259–1270.

Belesky, D. P., and D. P. Malinowski. 2016. Grassland commu-nities in the USA and expected trends associated with climatechange. Acta Agrobotanica 69:1–25.

Belnap, J., R. L. Reynolds, M. C. Reheis, S. L. Phillips, F. E.Urban, and H. L. Goldstein. 2009. Sediment losses and gainsacross a gradient of livestock grazing and plant invasion in acool, semi-arid grassland, Colorado Plateau, USA. AeolianResearch 1:27–43.

Belsky, A. J., and D. M. Blumenthal. 1997. Effects of livestockgrazing on stand dynamics and soils in upland forests of theinterior west. Conservation Biology 11:315–327.

Bestelmeyer, B. T., et al. 2011. Analysis of abrupt transitions inecological systems. Ecosphere 2:art129.

Betancourt, J. L., and O. K. Davis. 1984. Packrat middensfrom Canyon de Chelly, northeastern Arizona: paleoeco-logical and archaeological implications. QuaternaryResearch 21:56–64.

Bradford, J. B., W. K. Lauenroth, B. John, and K. William.2006. Controls over invasion of Bromus tectorum: the impor-tance of climate, soil, disturbance and seed availability. Jour-nal of Vegetation Science 17:693–704.

Bradley, B. A. 2009. Regional analysis of the impacts of climatechange on cheatgrass invasion shows potential risk andopportunity. Global Change Biology 15:196–208.

Burkhardt, J. W. 1996. Herbivory in the Intermountain West:an overview of evolutionary history, historic cultural impactsand lessons from the past. In Idaho Forest, Wildlife andRange Experiment Station, College of Natural Resources,University of Idaho, Moscow, Idaho.

Carmona, C. P., F. M. Azc�arate, F. de Bello, H. S. Ollero, J.Lep�s, and B. Peco. 2012. Taxonomical and functional diver-sity turnover in Mediterranean grasslands: interactionsbetween grazing, habitat type and rainfall. Journal of AppliedEcology 49:1084–1093.

Chambers, J. C., B. A. Bradley, C. S. Brown, C. D’Antonio, M. J.Germino, J. B. Grace, S. P. Hardegree, R. F. Miller, and D. A.Pyke. 2014. Resilience to stress and disturbance, and resistanceto Bromus tectorum L. invasion in cold desert shrublands ofwestern North America. Ecosystems 17:360–375.

Chambers, J. C., B. A. Roundy, R. R. Blank, S. E. Meyer, andA. Whittaker. 2007. What makes Great Basin sagebrushecosystems invasible by Bromus tectorum? Ecological Mono-graphs 77:117–145.

Chao, A., N. J. Gotelli, T. C. Hsieh, E. L. Sander, K. H. Ma, R. K.Colwell, and A. M. Ellison. 2014. Rarefaction and extrapolationwith Hill numbers: a framework for sampling and estimation inspecies diversity studies. Ecological Monographs 84:45–67.

Chao, A., and L. Jost. 2012. Coverage-based rarefaction andextrapolation: standardizing samples by completeness ratherthan size. Ecology 93:2533–2547.

Chew, R. M. 1982. Changes in herbaceous and suffrutescentperennials in grazed and ungrazed desertified grassland insoutheastern Arizona, 1958–1978. American Midland Natu-ralist 108:159.

Cingolani, A. M., I. Noy-Meir, and S. D�ıaz. 2005. Grazingeffects on rangeland diversity: a synthesis of contemporarymodels. Ecological Applications 15:757–773.

Condon, L., P. J. Weisberg, and J. C. Chambers. 2011. Abioticand biotic influences on Bromus tectorum invasion and Arte-misia tridentata recovery after fire. International Journal ofWildland Fire 20:597–604.

Coughenour, M. B. 1985. Graminoid responses to grazing bylarge herbivores: adaptations, exaptations, and interactingprocesses. Annals of the Missouri Botanical Garden 72:852.

Currie, P. O., D. W. Reichert, J. C. Malechek, and O. C. Wallmo.1977. Forage selection comparisons for mule deer and cattleunder managed ponderosa pine. Journal of Range Manage-ment 30:352–356.

Davidson, A. M., M. Jennions, and A. B. Nicotra. 2011. Doinvasive species show higher phenotypic plasticity than nativespecies and, if so, is it adaptive? A meta-analysis. EcologyLetters 14:419–431.

Diez, J. M., et al. 2012. Will extreme climatic events facilitatebiological invasions? Frontiers in Ecology and the Environ-ment 10:249–257.

D’Odorico, P., A. Bhattachan, K. F. Davis, S. Ravi, and C. W.Runyan. 2013. Global desertification: drivers and feedbacks.Advances in Water Resources 51:326–344.

Fensham, R. J., J. L. Silcock, and J. Firn. 2014. Managed live-stock grazing is compatible with the maintenance of plantdiversity in semidesert grasslands. Ecological Applications24:503–517.

Fisher, J., K. L. Cole, and R. S. Anderson. 2009. Using packratmiddens to assess grazing effects on vegetation change. Jour-nal of Arid Environments 73:937–948.

Floyd, M. L., T. L. Fleischner, D. Hanna, and P. Whitefield.2003. Effects of historic livestock grazing on vegetation atChaco culture National Historic Park, New Mexico. Conser-vation Biology 17:1703–1711.

Fuhlendorf, S. D., D. D. Briske, and F. E. Smeins. 2001. Herba-ceous vegetation change in variable rangeland environments:the relative contribution of grazing and climatic variability.Applied Vegetation Science 4:177–188.

Gross, J. A., and J. E. Knight. 2000. Elk presence inside various-sized cattle exclosures. Journal of Range Management 53:287–290.

Harpole, W. S., et al. 2016. Addition of multiple limitingresources reduces grassland diversity. Nature 537:93–96.

Harrison, S., B. D. Inouye, and H. D. Safford. 2003. Ecologicalheterogeneity in the effects of grazing and fire on grasslanddiversity. Conservation Biology 17:837–845.

Hayes, G. F., and K. D. Holl. 2003. Cattle grazing impacts onannual forbs and vegetation composition of mesic grasslandsin California. Conservation Biology 17:1694–1702.

Herrero-J�auregui, C., and M. Oesterheld. 2018. Effects of graz-ing intensity on plant richness and diversity: a meta-analysis.Oikos 127:757–766.

Hill, M. O. 1973. Diversity and evenness: a unifying notationand its consequences. Ecology 54:427–432.

Hormay, A. L., and A. B. Evanko. 1958. Rest-rotation grazing:a management system for bunchgrass ranges. USDA ForestService California Forest and Range Experimental Station,Albany, California, USA.


Hsieh, T. C., K. H. Ma, and A. Chao. 2016. iNEXT: an R pack-age for rarefaction and extrapolation of species diversity (Hillnumbers). Methods in Ecology and Evolution 7:1451–1456.

IPCC. 2013. Climate change: the physical science basis. Cam-bridge University Press, Cambridge, UK.

Irisarri, J. G. N., J. D. Derner, L. M. G. Porensky, D. J. Augus-tine, J. L. Reeves, and K. E. Mueller. 2015. Grazing intensitydifferentially regulates ANPP response to precipitation inNorth American semiarid grasslands. Ecological Applica-tions 26:1370–1380.

Jakoby, O., M. F. Quaas, S. Baumg€artner, and K. Frank. 2015.Adapting livestock management to spatio-temporal hetero-geneity in semi-arid rangelands. Journal of EnvironmentalManagement 162:179–189.

Jim�enez, M. A., F. M. Jaksic, J. J. Armesto, A. Gaxiola, P. L.Meserve, D. A. Kelt, and J. R. Guti�errez. 2011. Extreme cli-matic events change the dynamics and invasibility of semi-aridannual plant communities. Ecology Letters 14:1227–1235.

Jones, A. 2000. Effects of cattle grazing on North Americanarid ecosystems: a quantitative review. Western North Ameri-can Naturalist 60:155–164.

Kettenbach, J. A., N. Miller-Struttmann, Z. Moffett, and C.Galen. 2017. How shrub encroachment under climate changecould threaten pollination services for alpine wildflowers: acase study using the alpine skypilot, Polemonium viscosum.Ecology and Evolution 7:6963–6971.

Knapp, A. K., et al. 2008. Shrub encroachment in North Amer-ican grasslands: shifts in growth form dominance rapidlyalters control of ecosystem carbon inputs. Global ChangeBiology 14:615–623.

Kondoh, M. 2001. Unifying the relationships of species richnessto productivity and disturbance. Proceedings of the RoyalSociety B 268:269–271.

Lezama, F., S. Baeza, A. Altesor, A. Cesa, E. J. Chaneton, andJ. M. Paruelo. 2014. Variation of grazing-induced vegetationchanges across a large-scale productivity gradient. Journal ofVegetation Science 25:8–21.

Li, W., F. Xu, S. Zheng, F. Taube, and Y. Bai. 2017. Patternsand thresholds of grazing-induced changes in communitystructure and ecosystem functioning: species-level responsesand the critical role of species traits. Journal of Applied Ecol-ogy 54:963–975.

Liu, J., C. Feng, D. Wang, L. Wang, B. J. Wilsey, and Z. Zhong.2015. Impacts of grazing by different large herbivores ingrassland depend on plant species diversity. Journal ofApplied Ecology 52:1053–1062.

Loeser, M. R. R., T. D. Sisk, and T. E. Crews. 2007. Impact ofgrazing intensity during drought in an Arizona grassland.Conservation Biology 21:87–97.

Lyseng, M. P., E. W. Bork, D. B. Hewins, M. J. Alexander, C. N.Carlyle, S. X. Chang, and W. D. Willms. 2018. Long-term graz-ing impacts on vegetation diversity, composition, and exoticspecies presence across an aridity gradient in northern temper-ate grasslands. Plant Ecology 219:649–663.

Mack, R. N. 1981. Invasion of Bromus tectorum L. into WesternNorth America: an ecological chronicle. Agro-Ecosystems7:145–165.

Mack, R. N., and J. N. Thompson. 1982. Evolution in steppe withfew large, hooved mammals. American Naturalist 119:757–773.

Meisner, A., G. B. De Deyn, W. de Boer, and W. H. van der Put-ten. 2013. Soil biotic legacy effects of extreme weather eventsinfluence plant invasiveness. Proceedings of the National Acad-emy of Sciences USA 110:9835–9838.

Milchunas, D. G., and W. K. Lauenroth. 1993. Quantitativeeffects of grazing on vegetation and soils over a global range ofenvironments. Ecological Monographs 63:327–366.

Milchunas, D. G., O. E. Sala, and W. K. Lauenroth. 1988. A gener-alized model of the effects of grazing by large herbivores on grass-land community structure. American Naturalist 132:87–106.

Milchunas, D. G., W. K. Lauenroth, and I. C. Burke. 1998.Livestock grazing: Animal and plant biodiversity of short-grass steppe and the relationship to ecosystem function.Oikos 83:65.

Peters, D. P. C., B. Bestelmeyer, J. E. Herrick, E. L. Fredrickson,C. H. Monger, and K. M. Havstad. 2006. Disentangling com-plex landscapes: new insights into arid and semiarid systemdynamics. BioScience 56:491–501.

Rimer, R. L., and R. D. Evans. 2006. Invasion of downy brome(Bromus tectorum L.) causes rapid changes in the nitrogencycle. American Midland Naturalist 156:252–258.

Royle, J. A., and J. D. Nichols. 2003. Estimating abundancefrom repeated presence–absence data or point counts. Ecol-ogy 84:777–790.

Savory, A., and J. Butterfield. 1999. Holistic management: anew framework for decision making. Second edition. IslandPress, Washington, D.C., USA.

Schlesinger, W. H., J. F. Reynolds, G. L. Cunningham, L. F.Huenneke, W. M. Jarrell, R. A. Virginia, and W. G. Whitford.1990. Biological feedbacks in global desertification. Science247:1043–1048.

Schwinning, S., J. Belnap, D. R. Bowling, and J. R. Ehleringer.2008. Sensitivity of the Colorado plateau to change: climate,ecosystems, and society. Ecology and Society 13:art28.

Sperry, L. J., J. Belnap, and R. D. Evans. 2006. Bromus tectoruminvasion alters nitrogen dynamics in an undisturbed aridgrassland ecosystem. Ecology 87:603–615.

Sprinkle, J., et al. 2006. Dutchwoman Butte revisited: examin-ing paradigms for livestock grazing exclusion. American Soci-ety of Animal Science 57:46–50.

Stohlgren, T. J., M. B. Falkner, and L. D. Schell. 1995. A modi-fied-Whittaker nested vegetation sampling method. Vegetatio117:113–121.

Valone, T. J., M. Meyer, J. H. Brown, and R. M. Chew. 2002. Time-scale of perennial grass recovery in desertified arid grasslands fol-lowing livestock removal. Conservation Biology 16:995–1002.

Van Auken, O. W. 2000. Shrub invasions of North Americansemiarid grasslands. Annual Review of Ecology and System-atics 31:197–215.

Westoby, M., B. Walker, I. Noy-meir, M. Westoby, B. Walker, andI. Noy-meir. 1989. Opportunistic management for rangelandsnot at equilibrium. Journal of Range Management 42:266–274.

Willis, C. G., B. R. Ruhfel, R. B. Primack, A. J. Miller-Rushing,J. B. Losos, and C. C. Davis. 2010. Favorable climate changeresponse explains non-native species’ success in Thoreau’swoods. PLoS ONE 5:e8878.

Yuan, Z. Y., F. Jiao, Y. H. Li, and R. L. Kallenbach. 2016.Anthropogenic disturbances are key to maintaining the biodi-versity of grasslands. Scientific Reports 6:22132.

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