Research Article
Print
Research Article
Importance of soil and plant community disturbance for establishment of Bromus tectorum in the Intermountain West, USA
expand article infoA. Joshua Leffler, Thomas A. Monaco§, Jeremy J. James|, Roger L. Sheley
‡ South Dakota State University, Brookings, SD, United States of America
§ Forage and Range Research Laboratory, USDA-ARS, Logan, UT, United States of America
| University of California-Davis, Browns Valley, CA, United States of America
¶ Eastern Oregon Agriculture Research Center, USDA-ARS, Burns, OR, United States of America
Open Access

Abstract

The annual grass Bromus tectorum has invaded millions of hectares in western North America and has transformed former perennial grass and shrub-dominated communities into annual grasslands. Fire plays a key role in the maintenance of B. tectorum on the landscape but the type of disturbance responsible for initial invasion is less well understood. We conducted an experiment in a perennial shrub/grass/forb community in eastern Idaho, USA to examine the roles of plant community and soil disturbance on B. tectorum emergence and establishment prior to state-changing fires. Our experiment consisted of a plant community disturbance treatment where we (1) removed the shrub component, (2) removed the grass/forb component, or (3) removed all shrubs, grasses, and forbs. We followed this treatment with seeding of B. tectorum onto the soil surface that was (1) intact, or (2) disturbed. Each experimental plot had an associated control with no plant community disturbance but was seeded in the same manner. The experiment was replicated 20 times in two sites (high and low aboveground biomass). We measured emergence by counting seedlings in late spring and establishment by counting, removing, and weighing B. tectorum individuals in mid-summer. We also examined the influence of plant community disturbance on the soil environment by measuring extractable NH4+ and NO3– four times each summer. Soil disturbance greatly influenced the number of B. tectorum individuals that emerged each spring. Plant community disturbance, specifically disturbance of the grass/forb component, increased N availability in the late growing season and biomass of B. tectorum the following summer. We conclude that soil disturbance and plant community disturbance interact to promote the initial invasion of B. tectorum in Intermountain West valley ecosystems.

Keywords

Cheatgrass, germination, emergence, nitrogen, sagebrush, Artemisia tridentata

Introduction

Disturbance is widely appreciated as one of the critical factors leading to invasion by non-native plants worldwide (D’Antonio and Vitousek 1992, Lonsdale 1999, Davis et al. 2000, Jauni et al. 2015). Ecological theory suggests that disturbance ‘resets’ succession, leading to establishment by r-selected species (Elton 1958, Jauni et al. 2015), many of which are non-native. Disturbance however is a feature of all ecosystems and it is critical in maintaining the ‘typical’ structure of some ecosystems such as grasslands (Anderson 2006). Consequently, the timing or nature of the disturbance, relative to the historical disturbance regime, is likely more important than disturbance alone in triggering non-native plant invasion (Sher and Hyatt 1999).

Cheatgrass (Bromus tectorum L.) is one of the most widespread invasive plants in North America often replacing communities dominated by sagebrush (Artemisia tridentata) and other perennial grass and forb species (Stewart and Hull 1949, Knapp 1996, Chambers et al. 2007). While the annual grass B. tectorum is found throughout North America, it is invasive primarily in the Great Basin of the Intermountain West (Knapp 1996, Chambers et al. 2007) and may become problematic in the western Great Plains due to climate change (Concilio et al. 2015). Invasion by B. tectorum often results in a state change from a perennial-dominated system with infrequent fire (> 100 year return interval) to one with abundant annuals and a fire-return interval as little as 3–5 years (Knapp 1996, Mensing et al. 2006, Chambers et al. 2007, Bagchi et al. 2013, Chambers et al. 2014). Restoration of B. tectorum invaded systems is exceedingly challenging because of changes in fire frequency, loss of native perennial species, and altered nutrient cycling (Chambers et al. 2007, Hooker et al. 2008, Brabec et al. 2015, Stark and Norton 2015).

The invasion–fire cycle is well understood as the primary paradigm of B. tectorum dominance on the landscape (D’Antonio and Vitousek 1992, Chambers et al. 2007). However, the type of disturbance that triggers the initial stage of B. tectorum invasion – seed arrival and colonization (Theoharides and Dukes 2007) of perennial sagebrush communities prior to fire – is less well understood. Grazing by domestic livestock is thought to contribute to colonization (Reisner et al. 2013) although A. tridentata communities vary in susceptibility regardless of livestock presence (Chambers et al. 2007). Grazing may disperse seeds (Schiffman 1997), reduce abundance of native grasses that compete with B. tectorum (Briske and Richards 1995), decrease biological soil crust cover (Ponzetti et al. 2007), or create safe sites for establishment (Fowler 1988, Ponzetti et al. 2007). Regardless of the exact mechanism or the role played by livestock, disturbance to the plant community and soil surface is important for invasion by B. tectorum (Reisner et al. 2013).

Several studies document the importance of intact perennial communities and soils in resisting B. tectorum invasion. Large gaps between perennial bunchgrasses (Rayburn et al. 2014) and low perennial grass cover (Chambers et al. 2007) can promote B. tectorum establishment. Presence of the native grass Elymus elymoides, even with aboveground biomass removal, can enhance resistance to invasion (McGlone et al. 2011). Cover by B. tectorum is highest in areas with low cover and species richness of biological soil crusts (Ponzetti et al. 2007), and lichen crusts can reduce B. tectorum abundance by 85% (Deines et al. 2007). In integrating samples from 75 study sites in Oregon, Reisner et al. (2013) demonstrated the importance of diverse bunchgrass cover and intact soil crusts in conferring resistance to invasion. Intact plant communities can provide competition with B. tectorum while bare soils may provide an ideal substrate for germination.

Invasion by non-native species is often linked to changes in resource availability following disturbance (Davis et al. 2000, Davis and Pelsor 2001). Disturbance of vegetation, such as biomass removal, slows N acquisition by plants and can result in high N soils favored by short-lived annual species (Davies et al. 2007). Mowing A. tridentata communities as a restoration technique increased B. tectorum density and nutrient availability and sequential removal of functional groups from rangelands sequentially increased available NO3– and NH4+ (Davies et al. 2007, Davies et al. 2011). Annual grasses such as B. tectorum more rapidly use soil N than native perennial grasses (Leffler et al. 2011, 2013) and are more capable at exploiting pulses of N (James et al. 2006). Moreover, NO3– and NH4+ forms of N appear to affect growth of B. tectorum seedlings differently (Monaco et al. 2003).

We examine the importance of two disturbance types, soil surface and plant community disturbance, in the initial establishment of B. tectorum in a mixed sagebrush/perennial grass system in the Intermountain West. We address the initial stages of B. tectorum invasion, before fire removes the perennial component of the system and causes a state change to a B. tectorum dominated landscape. Specifically, we ask if soil and plant community disturbance influence (1) the number of B. tectorum individuals that germinate and emerge and (2) the biomass of established individuals in mid-summer. We also ask if plant community disturbance influences soil N resources that may contribute to B. tectorum establishment. We hypothesize that these disturbances act in concert and that both are necessary for initial invasion by B. tectorum.

Methods

We conducted this experiment at the United States Department of Agriculture Sheep Experiment Range near Dubois in eastern Idaho, USA (44.3° N, 112.7° W, elevation 1800 m) in an Intermountain West valley ecosystem. The study area is a mesic sagebrush (Artemisia tridentata) and perennial grass (Festuca idahoensis) community widespread in northern Intermountain West (i.e., sagebrush-steppe, West and Young 2000). The site has a long history of light grazing for experimental purposes. The system has warm summers (July mean 20.6 °C) and cool winters (January mean –7.2 °C) and receives ca. 330 mm of annual precipitation, primarily during winter (ca. 225 Oct.–May). Soils are fine-loamy or loamy-skeletal of the Maremma (Pachic Haploxerolls), Pyrenees (Typic Calcixerolls), and Akbash (Pachic Argixerolls) series with basalt parent material and loam or clay loam surface texture (Moffet et al. 2015). While parent material differs throughout the Intermountain West, soil texture in valley bottoms throughout the region tends toward finer particles observed here. Individuals of B. tectorum are present at the Experiment Range, but large, monospecific stands of this invasive species are not present. Lichen, moss, and cyanobacterial soil crusts were evident beneath dense vegetation but were not quantified.

We established two study sites, separated by ca. 4.5 km and 80-m elevation, at the Experiment Range that differed in standing biomass and vegetation height to generalize our study to multiple Intermountain West communities. The ‘Low’ site had an average dry mass of ca. 109 g m-2 while the ‘High’ site had an average dry mass of ca. 129 g m-2. Vegetation height was 37 and 61 cm in the Low and High sites, respectively. The largest difference, however, was dominance by A. tridentata; at the High site, 81% of the dry mass was A. tridentata, while at the low site 53% of the dry mass was A. tridentata. Other common species included the forb sulpher-flower buckwheat (Eriogonum umbellatum), the grass Idaho fescue (Festuca idahoensis), and the shrub bitterbrush (Purshia tridentata).

Within each study site, we randomly selected 60 study plots in June and July 2010. Plots were placed with a random-point generator prior to visiting the field sites but potential plots were rejected if cover and species composition in the surrounding 1–2 m was not representative of the community. Each plot (3 m × 1.5 m) consisted of two directly adjacent sub-plots (1.5 m × 1.5 m), one designated as the community treatment and randomly assigned, the other as the control. Three community treatments were imposed: (1) ‘Shrub’ – removal of the woody-shrub component of the plant community; (2) ‘Forb’ – removal of the grass and forb (i.e., the non-woody) component of the plant community; and (3) ‘All’ – removal of all plant material. We removed plants in 2010 by clipping to the ground and followed up clipping by targeted application of glyphosphate herbicide (Roundup, Monsanto Co., Creve Coeur, MO) to grass and forb species when any re-growth occurred; no shrub growth was observed following clipping and clipping did not disturb the soil surface. Plots were maintained with additional clipping and herbicide application during summer 2011.

In Autumn 2010 and 2011, we added seed of B. tectorum to microplots within each sub-plot at both study sites. Six microplots (10 cm × 10 cm) in each sub-plot received 100 seeds (locally collected, germination > 90%) yielding 1440 microplots between the two sites. Prior to seeding, one of the microplots in each sub-plot was scraped with a laboratory spatula to a depth of 5 cm to remove any vegetation and litter, and provide a bare substrate for seed germination. Removed material was collected and returned to a greenhouse to monitor for background germination of B. tectorum, which was minimal (data not shown).

Data were collected in 2011 and 2012. We visited plots in mid-spring to count B. tectorum individuals that emerged, and again in early summer to remove individuals that established. All individuals removed were dried and weighed for biomass. We collected soil for measurement of inorganic N content (NH4+ and NO3–) four times each summer (in 2011; mid-June, mid-July, mid-August, late-September: in 2012, mid-May, late-June, late-July, mid-September) to describe the changes in available N following our plant community treatments; microplots were too small and numerous for soil inorganic N analysis. Soils were collected using 2” diameter steel conduit to 15-cm depth. Ions were extracted from soils using 2M KCl, shaking, and filtration (Mulvaney 1996). All extractions took place in the field within 8 hours of sample collection. The filtrate was analyzed for NO3– by reduction of NO3– to NO2– with VCl3, which was subsequently captured with the Griess reagents (Doane and Horwáth 2003). Absorption at 540 nm was measured with a benchtop spectrophotometer after eight hours of color development. We analyzed the filtrate for NH4+ using the Berthelot reaction (Rhine et al. 1998) and measurement of absorption at 660 nm after one hour of color development.

Data were analyzed using mixed and zero-inflated Poisson models. All mixed models included main effects of study site (High or Low), plant community treatment (All, Shrub, Forb), and their interaction. Analysis of biomass of B. tectorum included an effect for soil treatment (i.e., Intact or Disturbed microplot) and the interaction of soil treatment with study site and soil treatment with plant community treatment. Analysis of soil NH4+ and NO3– content included a time effect (multiple measurements each summer), and interactions of time and plant community treatment, and time and study site. We did not examine three-way interactions due to difficulty of interpretation. Plot was treated as a random effect for all analyses. Data were transformed as necessary to satisfy normality. Seedling counts of B. tectorum were analyzed with a zero-inflated Poisson model. Counts of individuals follow a Poisson distribution rather than a normal distribution and typical methods of analysis include Poisson regression. However, our seedling establishment data included numerous zeros, which can result in a highly biased result. A zero-inflated model (Martin et al. 2005) combines a Poisson (i.e., discrete counts) with a binomial (i.e., presence/absence) to improve parameter estimation. We examined all models for significance using a likelihood-ratio test (Zurr et al. 2009). We initially fit a full model for each response with all main effects and interactions then removed effects to determine if model fit significantly declined. Confidence intervals (95%) of parameter estimates were obtained by bootstrapping the full model. Models were fit separately for 2011 and 2012 data. All analyses were conducted with packages NLME and PSCL in the R statistical computing environment (R Development Core Team 2013).

Results

The number of B. tectorum individuals emerging each spring following autumn addition of seed was highly influenced by soil disturbance and less so by plant community disturbance and study site (Table 1). In both years, the strongest effect in the model was soil disturbance but the other main effects and most interactions were statistically significant. In 2012, emergence of B. tectorum was nearly four-fold greater in microplots that received soil disturbance and emergence both years in disturbed soil was clearly lower in plots from the high site compared to the low site (Fig. 1). Moreover, sub-plots that received no community disturbance had lower B. tectorum emergence when soils were intact in both years, and when soils were disturbed in 2012. Emergence in 2011 was greater than emergence in 2012, although this difference was not statistically examined.

Figure 1.

The number of B. tectorum seedlings and biomass during two years of sampling following seeding of B. tectorum the previous autumn. Values represent bootstrapped median and 95% confidence intervals of a zero-inflated Poisson model (number of seedlings) or mixed-model (seedling biomass) analysis. Note difference in scale for seedling count between 2011 and 2012.

Analysis of number of B. tectorum individuals using a zero-inflated Poisson model.

2011 2012
Effect df* X2 p X2 p
Soil 5 6243 < 0.001 2356 < 0.001
Treatment 9 4183 < 0.001 93.25 < 0.001
Site 5 1833 < 0.001 19.12 < 0.001
Treatment*Site 3 226.9 < 0.001 16.83 < 0.001
Soil*Site 1 0.0139 0.993 7.057 < 0.001
Treatment*Soil 3 105.7 < 0.001 13.35 < 0.001

The biomass of B. tectorum in early summer was most strongly influenced by soil disturbance but plant community disturbance and study site were also significant effects (Table 2). In both years, greater biomass was observed in the soil disturbance site but the role of plant community disturbance was only obvious in 2012 (Fig. 1). During this second year of the experiment the least biomass of B. tectorum was observed when the plant community was intact while the most biomass was observed when both the grass/forb and the shrub community was removed. Removal of the shrub component of the plant community resulted in only slight increases in B. tectorum biomass, but removal of both components greatly enhanced B. tectorum growth. Despite more B. tectorum individuals observed in 2011, similar biomass was observed both years.

Analysis of B. tectorum biomass using a mixed-model.

2011 2012
Effect df* X2 p X2 p
Soil 5 138.3 < 0.001 138.4 < 0.001
Treatment 9 21.46 0.011 121.4 < 0.001
Site 5 12.94 0.024 19.92 0.001
Treatment*Site 3 3.567 0.312 4.096 0.251
Soil*Site 1 0.013 0.910 7.786 0.005
Treatment*Soil 3 0.246 0.970 8.659 0.034

Extractable NH4+ was influenced by plant community disturbance in 2011 but not in 2012, and NH4+ declined each growing season and differed among study sites in both years (Table 3). In 2011, the treatment effect clearly indicates high NH4+ concentration associated with disturbance of both the grass/forb and shrub components of the plant community at the Low site (Fig. 2). At the High site, disturbance had a lesser effect except in September when shrub and grass/forb removal increased [NH4+] by ca. 50% over the control. In 2012, the most pronounced experimental difference was between the high and low study sites in the late spring and early summer. The high site had ca. 60% more NH4+ than the low site in mid-May.

Figure 2.

Extractable N as NH4+ in soils four times each year of observation. Values represent bootstrapped median and 95% confidence intervals of a mixed-model analysis.

Analysis of extractable NH4+ using a mixed-model.

2011 2012
Effect df* X2 p X2 p
Month 15 221.5 < 0.001 947.5 < 0.001
Treatment 15 77.30 < 0.001 14.99 0.453
Site 7 87.48 < 0.001 112.9 < 0.001
Month*Treatment 9 29.79 < 0.001 10.25 0.331
Site*Treatment 3 19.60 < 0.001 2.138 0.544
Month*Site 3 67.29 < 0.001 55.41 < 0.001

Extractable NO3– differed through time, among plant community treatments, and between study sites in both years of the experiment (Table 4). In both years, high soil NO3– was observed in mid to late summer and consistent differences among plant community treatments were evident in July (Fig. 3). As with NH4+, when differences among treatments were clear, the highest NO3– was observed when both the grass/forb and shrub components were removed and the lowest NO3– was observed in the control and shrub removal plots. In the absence of disturbance, NO3– remained below 1 µg N g-1 soil throughout both years and at both sites.

Figure 3.

Extractable N as NO3– in soils four times each year of observation. Values represent bootstrapped median and 95% confidence intervals of a mixed-model analysis.

Analysis of extractable NO3– using a mixed-model.

2011 2012
Effect df* X2 p X2 p
Month 15 401.1 < 0.001 676.8 < 0.001
Treatment 15 401.6 < 0.001 433.9 < 0.001
Site 7 110.9 < 0.001 322.7 < 0.001
Month*Treatment 9 217.1 < 0.001 232.1 < 0.001
Site*Treatment 3 7.227 0.065 11.08 0.011
Month*Site 3 69.48 < 0.001 229.1 < 0.001

Discussion

Invasion is a complex process with many stages and each stage may be driven by different ecological factors. We examine the early stages of B. tectorum invasion and demonstrate the relative importance of both plant community and soil surface disturbance in promoting establishment of this annual grass. While our statistical tests suggest the importance of both disturbance types for emergence and subsequent growth, each disturbance appears to play a distinct role in invasion. The soil disturbance likely created ‘safe sites’ (Fowler 1988) for germination of B. tectorum while disturbance of the plant community, specifically the grass and forb functional groups, appears to be most important for post-emergence growth. Our data suggest increased N availability, especially in autumn when germination occurs, may be responsible for enhanced growth of this invasive grass. Consequently, at our sites in the northern Great Basin, two different disturbances appear to interact to promote the establishment of B. tectorum before fire results in a state change.

Disturbance of the soil surface results in bare ground with good seed-substrate contact, allowing an emerging radicle to rapidly reach critical water and N resources. Soil disturbance removes litter and breaks up soil crusts. Litter can promote establishment if it acts primarily to protect seedlings from frost, full sun, or excessive water loss (Evans and Young 1972, Rotundo and Aguiar 2005, Loydi et al. 2013). An approximately 1.5-fold increase in litter depth resulted in a 2 to 4-fold increase in B. tectorum density (Davies et al. 2012) and other species often respond positively to litter cover although the importance of litter varies with environmental conditions (Loydi et al. 2013). Litter can also hinder establishment if it prevents seed contact with the soil (Rotundo and Aguiar 2005, Loydi et al. 2013). In disturbed forests, deep litter hindered B. tectorum recruitment and growth (Pierson and Mack 1990). Beckstead and Augspurger (2004) found litter enhanced establishment of B. tectorum but was neutral at other life stages.

Disruption of biological soil crusts (BSC) clearly promotes invasion by B. tectorum. We did not examine BSC coverage in this study but lichen, moss, and cyanobacterial crusts were present and our soil surface disturbance removed these BSC. Lichen crusts can reduce the abundance of B. tectorum by 85% possibly through reducing germination percentage or inhibiting root penetration of soil (Serpe et al. 2006, Deines et al. 2007). Disruption of BSC was likely more important than litter presence in hindering B. tectorum establishment in this study although the interaction of BSC removal and litter presence is largely unexplored.

Once B. tectorum was established the role of plant community disturbance became clear. We observed the greatest biomass increase by B. tectorum over controls when both the shrub and grass/forb plant communities were removed. However, independent removal of these components demonstrates that removal of the grass/forb component was most important. Our results are broadly consistent with numerous studies showing intact perennial communities can resist invasion by B. tectorum (e.g., Chambers et al. 2007). Cover by E. elymoides, a native bunchgrass, at as little as 20% can nearly exclude B. tectorum (Booth et al. 2003b) and there is generally a negative relationship between B. tectorum and perennial cover at another site in eastern Idaho (Anderson and Inouye 2001). Despite finding greater importance of the grass/forb component here, others have demonstrated that removal of A. tridentata from the system promotes invasion by B. tectorum and non-native forbs in southeastern Idaho (Prevéy et al. 2010). Moreover, mowing tall A. tridentata communities as a restoration technique to improve grass/forb dominance appears to promote B. tectorum invasion despite accomplishing the stated goal of the treatment (Davies et al. 2011, Davies et al. 2012). Consequently, site-specific factors are likely important in local invasion and broadly speaking, a complete perennial community composed of grasses, forbs, and shrubs in the Intermountain West will be most invasion resistant (Pokorny et al. 2005, Leffler et al. 2014).

Previous research strongly links increased soil water and N to invasion by B. tectorum. Soil water made available by removal of A. tridentata enhanced B. tectorum abundance (Prevéy et al. 2010), and soil N and B. tectorum success were positively correlated (Jones et al. 2015). Furthermore, numerous studies show B. tectorum to respond more to high N availability than native species (Monaco et al. 2003, James et al. 2008, Leffler et al. 2011, 2013). Here, we show that disturbance of A. tridentata and the grass/forb component of plant communities can differentially influence soil inorganic N resources. As with our observation of biomass, disturbing both components yielded a similar increase in N as disturbing the grass/forb component, and this increase was larger than that produced by the removal of the woody shrub component alone. We primarily observed this pattern in the late summer or early autumn which coincidentally is when this annual grass germinates (Knapp 1996, Chambers 2007). Furthermore, B. tectorum dominated communities can maintain high soil N among years (Booth et al. 2003a, Hooker et al. 2008, Stark and Norton 2015). This raises the possibility that B. tectorum may facilitate its own persistence on the landscape if its early senescence in summer increases N availability (due to lack of N uptake by other species) in autumn when it germinates (Leffler et al. 2011, Jones et al. 2015). A similar enhancement of soil water was observed in B. tectorum compared to perennial communities in western Utah (Ryel et al. 2010).

The interaction between soil and plant community disturbance as a mechanism for initial B. tectorum establishment likely applies broadly to valley ecosystems of the northern Intermountain West of the USA where the A. tridentata/F. idahoensis association is widespread. We conducted this experiment simultaneously in two plant communities at the Experiment Station and the influence of soil and plant community disturbance was qualitatively similar at both sites. In both cases, soil disturbance enhanced emergence and disturbance of the grass/forb component resulted in enhanced soil N and biomass.

The Intermountain West was historically an ecosystem that received infrequent disturbance. The fire return interval was likely greater than 100 years and may have reached 500 years in some locations (Mensing et al. 2006) and frequent fires appear to hasten B. tectorum spread on the landscape (D’Antonio and Vitousek 1992, Chambers et al. 2007). Additionally, the plant communities of the Intermountain West do not have a recent evolutionary history of grazing by large mammals such as the Great Plains does with bison (Mack and Thompson 1982, Anderson 2006). Consequently, removal of vegetation and soil surface disturbance differs from the historical regime. A recent meta-analysis suggests that soil disturbance may influence non-native plant diversity, but that plant removal did not influence diversity or abundance (Jauni et al. 2015). While we recognize our result is a single case study, our finding of the importance of the interaction between soil and plant disturbance indicates that these disturbances should not be examined in isolation.

Conclusion

Invasion is often described as a multi-stage process and different factors influence invasion at each step (Theoharides and Dukes 2007). Numerous processes interact for a non-native species to become invasive. For B. tectorum, the early stages of invasion require an interaction between soil and plant community disturbance and the broader state change on the landscape requires an interaction with fire. Theoretically, preventing further spread requires removing only one of the interacting processes although invasion resistance was not complete in the absence of soil or plant community disturbance. For B. tectorum invasion, limiting soil disturbance can suppress emergence, limiting plant community disturbance can suppress growth, and preventing fire can slow spread.

Acknowledgments

This study is a contribution of the USDA-ARS Area-Wide Ecologically Based Invasive Plant Management Program. We thank USDA-ARS Sheep Experiment Station and G. Lewis for permission to conduct this project. Students including B. Pasbt, J. Killpack, W. Packer, S. Felix, H. Holland, and M. Hirsch made the fieldwork possible. Special thanks to J. Williams for excellent assistance in the field and laboratory.

References

  • Anderson RC (2006) Evolution and origin of the central grasslands of North America: climate, fire, and mammalian grazers. The Journal of the Torrey Botanical Society 133: 626–647. doi: 10.3159/1095-5674(2006)133[626:EAOOTC]2.0.CO;2
  • Bagchi S, Briske DD, Bestelmeyer BT, Wu XB (2013) Assessing resilience and state-transition models with historical records of cheatgrass Bromus tectorum invasion in North American sagebrush-steppe. Journal of Applied Ecology 50: 1131–1141.
  • Beckstead J, Augspurger CK (2004) An experimental test of resistance to cheatgrass invasion: limiting resources at different life stages. Biological Invasions 6: 417–432. doi: 10.1023/B:BINV.0000041557.92285.43
  • Booth MS, Caldwell MM, Stark JM (2003) Overlapping resource use in three Great Basin species: implications for community invasibility and vegetation dynamics. Journal of Ecology 91: 36–48. doi: 10.1046/j.1365-2745.2003.00739.x
  • Booth MS, Stark JM, Caldwell MM (2003) Inorganic N turnover and availability in annual- and perennial-dominated soils in a northern Utah shrub-steppe ecosystem. Biogeochemistry 66: 311–330. doi: 10.1023/B:BIOG.0000005340.47365.61
  • Brabec MM, Germino MJ, Shinneman DJ, Pilliod DS, McIlroy SK, Arkle RS (2015) Challenges of establishing big sagebruch (Artemisia tridentata) in rangeland restoration: Effects of herbicide, mowing, whole-community seeding, and sagebrush seed sources. Rangeland Ecology & Management 68: 432–435. doi: 10.1016/j.rama.2015.07.001
  • Briske DD, Richards JH (1995) Plant responses to defoliation: a physiological, morphological, and demographic evaluation. In: Bedunah DJ, Sosebee RE (Eds) Wildland Plants: Physiological Ecology and Development Morphology. Society for Range Management, Denver, CO, 625–710.
  • Chambers JC et al. (2014) Resilience and resistance of sagebrush ecosystems: implications for state and transition models and management treatments. Rangeland Ecology & Management 67: 440–454. doi: 10.2111/REM-D-13-00074.1
  • Chambers JC, Roundy BA, Blank RR, Meyer SE, Whittaker A (2007) What makes great basin sagebrush ecosystems invasible by Bromus tectorum? Ecological Monographs 77: 117–145. doi: 10.1890/05-1991
  • Concilio AL, Prevéy JS, Omasta P, O’Connor J, Nippert JB, Seastedt TR (2015) Response of a mixed grass prairie to an extreme precipitation event. Ecosphere 6: 172. doi: 10.1890/ES15-00073.1
  • D’Antonio CM, Vitousek PM (1992) Biological invasions by exotic grasses, the grass/fire cycle, and global change. Annual Review of Ecology and Systematics 23: 63–87. doi: 10.1146/annurev.es.23.110192.000431
  • Davies KW, Bates JD, Nafus AM (2011) Are there benefits to mowing Wyoming big sagebrush plant communities? An evaluation in southeastern Oregon. Environmental Management 48: 539–546. doi: 10.1007/s00267-011-9715-3
  • Davies KW, Bates JD, Nafus AM (2012) Mowing Wyoming big sagebrush communities with degraded herbaceous understories: has a threshold been crossed. Rangeland Ecology & Management 65: 498–505. doi: 10.2111/REM-D-12-00026.1
  • Davies KW, Pokorny ML, Sheley R, James JJ (2007) Influence of plant functional group removal on inorganic soil nitrogen concentrations in native grasslands. Rangeland Ecology & Management 60: 304–310. doi: 10.2111/1551-5028(2007)60[304:IOPFGR]2.0.CO;2
  • Davis MA, Grime JP, Thompson K (2000) Fluctuating resources in plant communities: a general theory of invasibility. Journal of Ecology 88: 528–534. doi: 10.1046/j.1365-2745.2000.00473.x
  • Davis MA, Pelsor M (2001) Experimental support for a resource-based mechanistic model of invasibility. Ecology Letters 4: 421–428. doi: 10.1046/j.1461-0248.2001.00246.x
  • Deines L, Rosentreter R, Eldridge DJ, Serpe MD (2007) Germination and seedling establishment of two annual grasses on lichen-dominated biological soil crusts. Plant and Soil 295: 23–35. doi: 10.1007/s11104-007-9256-y
  • Doane TA, Horwáth WR (2003) Spectrophotometric determination of nitrate with a single reagent. Analytical Letters 36: 2713–2722. doi: 10.1081/AL-120024647
  • Evans EA, Young JA (1972) Microsite requirements for establishment of annual rangeland weeds. Weed Science 20: 350–356.
  • Fowler NL (1988) What is a safe site? Neighbor, litter, germination date, and patch effects. Ecology 69: 947–961. doi: 10.2307/1941250
  • Hooker TD, Stark JM, Leffler AJ, Peek M, Ryel R (2008) Distribution of ecosystem C and N within contrasting vegetation types in a semiarid rangeland in the Great Basin, USA. Biogeochemistry 90: 291–308. doi: 10.1007/s10533-008-9254-z
  • James JJ (2008) Leaf nitrogen productivity as a mechanism driving the success of invasive annual grasses under low and high nitrogen supply. Journal of Arid Environments 72: 1775–1784. doi: 10.1016/j.jaridenv.2008.05.001
  • James JJ, Aanderud ZT, Richards JH (2006) Seasonal timing of N pulses influences N capture in a saltbush scrub community. Journal of Arid Environments 67: 688–700. doi: 10.1016/j.jaridenv.2006.03.014
  • Jauni M, Gripenberg S, Ramula S (2015) Non-native plant species benefit from disturbance: a meta-analysis. Oikos 124: 122–129. doi: 10.1111/oik.01416
  • Jones RO, Chambers JC, Board DI, Johnson DW, Blank RR (2015) The role of resource limitation in restoration of sagebrush ecosystems dominated by cheatgrass (Bromus tecotrum). Ecosphere 6: 107. doi: 10.1890/ES14-00285.1
  • Knapp PA (1996) Cheatgrass (Bromus tectorum L) dominance in the Great Basin Desert. Global Environmental Change 6: 37–52. doi: 10.1016/0959-3780(95)00112-3
  • Leffler AJ, James JJ, Monaco TA (2013) Temperature and functional traits influence differences in nitrogen uptake capacity between native and invasive grasses. Oecologia 171: 51–60. doi: 10.1007/s00442-012-2399-4
  • Leffler AJ, Leonard ED, James JJ, Monaco TA (2014) Invasion is contingent on species assemblages and invasive species identity in experimental rehabilitation plots. Rangeland Ecology & Management 67: 657–666. doi: 10.2111/REM-D-13-00140.1
  • Leffler AJ, Monaco TA, James JJ (2011) Nitrogen acquisition by annual and perennial grass seedlings: testing the role of performance and plasticity to explain plant invasion. Plant Ecology 212: 1601–1611. doi: 10.1007/s11258-011-9933-z
  • Loydi A, Eckstein RL, Otte A, Donath TW (2013) Effects of litter on seedling establishment in natural and semi-natural grasslands: a meta-analysis. Journal of Ecology 101: 454–464. doi: 10.1111/1365-2745.12033
  • Mack RN, Thompson JN (1982) Evolution in steppe with few large, hooved mammals. The American Naturalist 119: 757–773. doi: 10.1086/283953
  • Martin TG et al. (2005) Zero tolerance ecology: improving ecological inference by modelling the source of zero observations. Ecology Letters 8: 1235–1246. doi: 10.1111/j.1461-0248.2005.00826.x
  • McGlone CM, Sieg CH, Kolb TE (2011) Invasion resistance and persistence: established plants win, even with disturbance and high propagule pressure. Biological Invasions 13: 291–304. doi: 10.1007/s10530-010-9806-8
  • Mensing S, Livingston S, Barker P (2006) Long-term fire history in Great Basin sagebrush reconstructed from macroscopic charcoal in spring sediments, Newark Valley, Nevada. Western North American Naturalist 66: 64–77. doi: 10.3398/1527-0904(2006)66[64:LFHIGB]2.0.CO;2
  • Moffet CA, Taylor JB, Booth DT (2015) Postfire shrub cover dynamics: A 70-year fire chronosequence in mountain big sagebrush communities. Journal of Arid Environments 114: 116–123. doi: 10.1016/j.jaridenv.2014.12.005
  • Monaco TA et al. (2003) Contrasting responses of Intermountain West grasses to soil nitrogen. Journal of Range Management 56: 282–290. doi: 10.2307/4003820
  • Mulvaney RL (1996) Nitrogen–Inorganic Forms. In: Sparks DL, Page AL, Helmke PA, Loeppert RH (Eds) Methods of Soil Analysis Part 3–Chemical Methods. Soil Science Society of America, American Society of Agronomy, Madison, WI, 1123–1184.
  • Pierson EA, Mack RN (1990) The population biology of Bromus tectorum in forests: effect of disturbance, grazing, and litter on seedling establishment and reproduction. Oecologia 84: 526–533. doi: 10.1007/BF00328170
  • Pokorny ML, Sheley RL, Zabinski CA, Engel RE, Svejcar TJ, Borkowski JJ (2005) Plant functional group diversity as a mechanism for invasion resistance. Restoration Ecology 13: 448–459. doi: 10.1111/j.1526-100X.2005.00056.x
  • Prevéy JS, Germino MJ, Huntly NJ (2010) Loss of foundation species increases population growth of exotic forbs in sagebrush steppe. Ecological Applications 20: 1890–1902. doi: 10.1890/09-0750.1
  • R Development Core Team (2013) R: A language and environment for statistical computing, reference index version 3.2.1. R Foundation for Statistical Computing, Vienna, Austria.
  • Rayburn AP, Schupp EW, Kay S (2014) Effects of perennial semi-arid bunchgrass spatial patterns on performance of the invasive annual cheatgrass (Bromus tectorum L.). Plant Ecology 215: 247–251. doi: 10.1007/s11258-013-0293-8
  • Reisner MD, Grace JB, Pyke DA, Doescher PS (2013) Conditions favoring Bromus tectorum dominance of endangered sagebrush steppe ecosystems. Journal of Applied Ecology 50: 1039–1049. doi: 10.1111/1365-2664.12097
  • Rhine ED, Sims GK, Mulvaney RL, Pratt EJ (1998) Improving the Berthelot reaction for determining ammonium in soil extracts and water. Soil Science Society of America Journal 62: 473–480. doi: 10.2136/sssaj1998.03615995006200020026x
  • Rotundo JL, Aguiar MR (2005) Litter effects on plant regeneration in arid lands: a complex balance between seed retention, seed longevity and soil–seed contact. Journal of Ecology 93: 829–838. doi: 10.1111/j.1365-2745.2005.01022.x
  • Ryel RJ, Leffler AJ, Ivans C, Peek MS, Caldwell MM (2010) Functional differences in water-use patterns of contrasting life forms in Great Basin steppelands. Vadose Zone Journal 9: 548–560. doi: 10.2136/vzj2010.0022
  • Schiffman PM (1997) Animal-mediated dispersal and disturbance: driving forces behind alien plant naturalization. In: Luken JO, Thieret JW (Eds) Assessment and Management of Plant Invasions. Springer-Verlag, New York, 87–94. doi: 10.1007/978-1-4612-1926-2_8
  • Serpe MD, Orm JM, Barkes T, Rosentreter R (2006) Germination and seed water status of four grasses on moss-dominated biological soil crusts from arid lands. Plant Ecology 185: 163–178. doi: 10.1007/s11258-005-9092-1
  • Sher AA, Hyatt LA (1999) The disturbed resource-flux invasion matrix: a new framework for patterns of plant invasion. Biological Invasions 1: 107–114. doi: 10.1023/A:1010050420466
  • Stark JM, Norton JM (2015) The invasive annual cheatgrass increases nitrogen availability in 24-year-old replicated field plots. Oecologia 177: 799–809. doi: 10.1007/s00442-014-3093-5
  • Stewart GS, Hull AC (1949) Cheatgrass (Bromus tectorum L.)-An ecologic intruder in southern Idaho. Ecology 30: 58–74. doi: 10.2307/1932277
  • Theoharides KA, Dukes JS (2007) Plant invasion across space and time: factors affecting nonindigenous species success during four stages of invasion. New Phytologist 176: 256–273. doi: 10.1111/j.1469-8137.2007.02207.x
  • West NE, Young JA (2000) Intermountain valleys and lower mountain slopes. In: Barbour MG, Billings WD (Eds) North American Terrestrial Vegetation, 2nd edn. Cambridge University Press, Cambridge, UK, 255–284.
  • Zurr A, Leno EN, Walker N, Saveliev AA, Smith GM (2009) Mixed effects models and extensions in ecology with R. Springer, New York. doi: 10.1007/978-0-387-87458-6