AN EXPERIMENTAL APPROACH TO UNDERSTANDING HOW BROMUS TECTORUM WILL RESPOND TO GLOBAL CLIMATE CHANGE IN THE SAGEBRUSH-STEPPE by Christian Douglas Larson A thesis submitted in partial fulfillment of the requirements for the degree of Master of Science In Land Resources and Environmental Sciences MONTANA STATE UNIVERSITY Bozeman, Montana November 2016 ©COPYRIGHT By Christian Douglas Larson 2016 All Rights Reserved ii DEDICATION I would like to dedicate this thesis to my wife Michele, whose love, patience, and support kept me going throughout this adventure. I would also like to dedicate this thesis to my advisor Lisa for giving me this opportunity and for her support throughout. iii ACKNOWLEDGEMENTS I would first like to acknowledge my advisor Dr. Lisa Rew. Without her continual support, patience, and hard work in the field, lab, and with the writing of this thesis, I would not have been able to complete such a momentous undertaking. I would like to thank Dr. Erik Lehnhoff for his assistance in developing and setting up my projects and Dr. Bruce Maxwell for helping design my projects. I am indebted to Chance Noffsinger for all his hard work and comradery in the lab, greenhouse, and field, without him I don’t know what I would have done. I also owe a debt of thanks to Noelani Boise and Kaylee Schmitz for their tireless work in the greenhouse, lab, and field. Finally, I would like to thank Drs. Kim Taylor and Tim Seipel for their assistance with ‘R’ and my statistical analyses. iv TABLE OF CONTENTS 1. INTRODUCTION TO THESIS ................................................................................................. 1 Plant Responses to Climate Change ................................................................................... 2 Plant Responses to Precipitation and Water Availability ...................................... 3 Plant Responses to Temperature .......................................................................... 5 Bromus tectorum .............................................................................................................. 10 History and Range ................................................................................................ 10 The Impacts of the Bromus tectorum Invasion in the Intermountain West ........ 11 Bromus tectorum Response to Fire and Nutrient Availability ............................. 12 Limiting Factors of the Bromus tectorum-Fire Cycle ........................................... 13 Native Plant Community Response to Fire .......................................................... 14 Bromus tectorum and the Native Plant Community............................................ 14 Climate, Native Plant Community, and Bromus tectorum ................................... 15 Effects of Precipitation and Soil Moisture on Bromus tectorum ......................... 16 Bromus tectorum and Temperature .................................................................... 19 Climate Change and Projections for Western North America .......................................... 20 The Projected Expansion of Bromus tectorum .................................................... 21 Plant Responses to Elevated Atmospheric CO2 Concentrations ....................................... 22 CO2 and Plant Growth and Competition .............................................................. 22 Bromus tectorum and CO2 ................................................................................... 25 An Experimental Approach to Understanding How Global Climate Change Will Affect Bromus tectorum .................................................................. 26 References ........................................................................................................................ 29 2. THE EFFECTS OF INCREASED TEMPERATURE, DECREASED PRECIPITATION, AND A PRESCRIBED BURN ON BROMUS TECTORUM IN A MONTANA SAGEBRUSH-STEPPE PLANT COMMUNITY .......... 46 Introduction ...................................................................................................................... 46 Methods ............................................................................................................................ 49 Site Description .................................................................................................... 49 Experimental Design ............................................................................................ 51 Climate Manipulation Designs ............................................................................. 52 Open Top Chambers ............................................................................... 52 Rainout Shelters ...................................................................................... 53 Sampling Methods ............................................................................................... 54 Data Analysis ........................................................................................................ 54 Results ............................................................................................................................... 56 Discussion ......................................................................................................................... 64 The Effects of Climate and Bromus tectorum on the Native Community ........... 67 The Effects of Fire on Bromus tectorum and the Native Grass Community ........ 69 References ........................................................................................................................ 73 v TABLE OF CONTENTS - CONTINUED 3. THE EFFECTS OF INCREASED TEMPERATURE, ALTERED RESOURCE AVAILABILITY, AND ELEVATED ATMOSPHERIC CO2ON THE COMPETITION BETWEEN BROMUS TECTORUM AND PSEUDOROEGNERIA SPICATA ................................................ 81 Introduction ...................................................................................................................... 81 Methods ............................................................................................................................ 87 Experimental Design ............................................................................................ 87 Experiment 1: Decreased Water, Increased Temperature And Increased Nutrient Availability ........................................................ 87 Experiment 2: Increased Atmospheric CO2 Concentration And Decreased Water Availability at an Elevated Temperature ............ 89 Statistical Analysis ................................................................................................ 90 Results ............................................................................................................................... 91 Effects of Competition, Decreased Water, Elevated Temperature, And Increased Nutrient Availability ..................................................................... 91 Effects of Competition, Increased Atmospheric CO2, and Decreased Water Availability Under an Elevated Temperature ................... 97 Discussion ....................................................................................................................... 102 Conclusions ........................................................................................................ 106 References ...................................................................................................................... 108 4. CONCLUSION TO THESIS ................................................................................................. 116 Conclusions and Future Work ......................................................................................... 123 References ...................................................................................................................... 125 REFERENCES CITED ................................................................................................................ 131 APPENDIX A: Supplemental Figures (Chapter Two) .............................................................. 152 vi LIST OF TABLES Table Page 2.1 Mean temperature and precipitation data for Norris, MT climate station (1908-2016) ........ 50 2.2 Temperature response to climate treatments recorded at Red Bluff Research Station 2014-2016. ................................................................................... 53 2.3 Akaike Information Criterion (AIC) table for Bromus tectorum cover ..................................... 57 2.4 Results of the best linear mixed-effects models assessing the Bromus tectorum response to the burn and climate treatments, and native grass cover ..... 59 2.5 Results of the best linear mixed-effects models assessing the Pseudoroegneria spicata response to the burn and climate treatments, and Bromus tectorum cover .................................................................................................... 62 3.1 Results of linear mixed-effects models conducted to assess the effects of competition, water availability, temperature, and nutrient availability on Bromus tectorum and Pseudoroegneria spicata biomass and growth .............................. 93 3.2 Results of linear mixed-effects models conducted to assess the effects of competition, water availability, temperature, and nutrient availability on Bromus tectorum and Pseudoroegneria spicata relative yield .......................................... 95 3.3 Results of linear mixed-effects models conducted to assess the effects of competition, water availability, temperature, and nutrient availability on Bromus tectorum and Pseudoroegneria spicata proportion of total pot biomass ............ 96 3.4 Results of linear mixed-effects models conducted to assess the effects of elevated atmospheric CO2 concentration and decreased water on Bromus tectorum and Pseudoroegneria spicata biomass and growth ................................... 98 3.5 Results of linear mixed-effects models conducted to assess the effects of elevated atmospheric CO2 concentration and decreased water on Bromus tectorum and Pseudoroegneria spicata relative yield ............................................... 99 3.6 Results of linear mixed-effects models conducted to assess the effects of elevated atmospheric CO2 concentration and decreased water on Bromus tectorum and Pseudoroegneria spicata proportion of total pot biomass ............... 101 vii LIST OF FIGURES Figure Page 2.1 Bromus tectorum cover within the six climate-burn treatments ............................................ 56 2.2 Bromus tectorum cover response to native grass cover by climate treatment ...................... 58 2.3 Bromus tectorum individual fecundity for the burned and unburned treatments ................. 60 2.4 Relationship between Pseudoroegneria spicata cover and Bromus tectorum cover by climate treatment ........................................................................ 61 2.5 Native grass cover within each of the six climate-burn treatments ........................................ 63 2.6 Relationship between Bromus tectorum cover and total species richness and native species richness .................................................................. 63 3.1 Effects of Pseudoroegneria spicata competition, water availability, and nutrient availability on Bromus tectorum height ............................................................. 91 3.2 Effects of intraspecific competition, water availability, and temperature on Pseudoroegneria spicata height ............................................................ 92 3.3 Effects of competition with Pseudoroegneria spicata, the combination of elevated temperature and decreased water, and increased nutrient availability, on Bromus tectorum and Pseudoroegneria spicata relative yield .......................................... 94 3.4 Effects of competition with Pseudoroegneria spicata, and the combination of decreased water and increased nutrient availability, on Bromus tectorum and Pseudoroegneria spicata proportion of total pot biomass ..................................................... 96 3.5 Effects of competition with Pseudoroegneria spicata, and atmospheric CO2 concentration on Bromus tectorum and P. spicata relative yield within the watered treatment and the dry treatment ..................................................................... 100 3.6 Effects of competition with Pseudoroegneria spicata, and atmospheric CO2 concentration on B. tectorum and P. spicata proportion of total pot biomass .................... 100 3.7 Monoculture B. tectorum and P. spicata biomass and height responses to elevated atmospheric CO2 concentration ......................................................................... 102 viii ABSTRACT Global climate change, including elevated atmospheric CO2 concentrations, increases in global surface temperatures, and changes in resource availability, has significant consequences for global plant communities, one of which is the expansion of invasive species. The invasive grass species Bromus tectorum dominates areas of the North American sagebrush-steppe. In these areas, B. tectorum responds positively to elevated nutrients after fire and a positive feedback with fire has been initiated. Bromus tectorum dominance and its positive response to fire are limited by cold and moist climates. Global climate change is predicted to expand the climate suitability for B. tectorum dominance, as well as that of its response to fire. Using a field study and controlled setting experiments, I investigated this prediction. In a cold and moist southwestern Montana sagebrush-steppe, my field experiment assessed the response of B. tectorum and the native plant community to increased growing season temperatures, decreased growing season precipitation, and a prescribed burn. We found that both B. tectorum and a dominant native perennial grass, Pseudoroegneria spicata, responded negatively to experimental warming, and warming and drying. Bromus tectorum’s response to fire was limited to an increase in individual fecundity across the climate scenarios and compensatory growth in warm and dry conditions. In controlled settings, using differing densities of B. tectorum and P. spicata, I performed replacement series experiments that altered temperature, water availability, nutrient availability, and, secondly, atmospheric CO2 concentration and water availability. Bromus tectorum competitiveness was enhanced by warmer and drier conditions and elevated nutrient availability. When grown in monoculture, both species responded positively to elevated CO2. When grown in competition, elevated CO2 increased P. spicata’s already significant suppressive effect on B. tectorum. This effect was magnified when soil moisture was limited. Due to B. tectorum’s significant negative response to the field climate treatments, its limited response to fire, and the significant suppressive effect of the native grasses in both experiments, especially in elevated CO2, I conclude that similar future climate scenarios will not promote the expansion of B. tectorum dominance and its positive response to fire within the cold and moist northern region of the sagebrush-steppe. 1 CHAPTER ONE INTRODUCTION TO THESIS Since the inception of the industrial revolution the composition of the atmosphere has undergone an unprecedented change (IPCC 2013); it is projected that the 2015-2016 atmospheric CO2 concentration will be the highest on record and concentrations will surpass and remain above 400 ppm for the entire year (Betts et al. 2016). The rise in atmospheric CO2 and other greenhouses gasses has had widespread effects that include increasing global surface temperatures and altering global precipitation patterns (IPCC 2013). Plant competition and community composition is affected by CO2 concentration (Bazzaz 1990; Owensby et al. 1999; Williams et al. 2007), temperature (Luo 2007; Williams et al. 2007; Mueller et al. 2016), and resource availability (Wilson and Tilman 1991; Wilson and Tilman 1993). As a result, global climate change has had significant ecological consequences for global plant communities (McCarty 2001; Gordo and Sanz 2010; Belesky and Malinowski 2016). One of the most significant ecological effects has been facilitating the spread and increasing the impact of non- native invasive plant species (Dukes and Mooney 1999; McCarty 2001; Walther et al. 2002). Non-native invasive plant species are those species that have been introduced to new areas and have self-sustaining populations that are spreading both spatially and in density, with the potential to spread across considerable distances (Richardson et al. 2000). This can often result in them growing in unwanted places or at unwanted densities, which can have significant ecological and economics effects (Richardson et al. 2000). These effects are broad, affecting both abiotic ecosystem attributes, nutrient cycling (Kourtev et al. 2002; Ehrenfeld 2003), hydrology (Levine et al. 2003; Boxell and Drohan 2009), disturbance regimes (D’Antonio et al. 2 1992), and biotic factors, often reducing the growth, production, and abundance of either crops or dominant native species, and reducing native biodiversity (Vitousek et al. 1996; Richardson et al. 2000; Chornesky and Randall 2003; Maron and Marler 2008; Vilà et al. 2011; Pyšek et al. 2012). Invasive plant species often have high growth rates and are highly competitive with native species when they are planted together (Levine et al. 2003; Van Kleunen et al. 2010). However, established and intact native plant communities are often resistant to plant invasions unless they experience a disturbance, such as grazing or fire, which can facilitate invasions to a greater or lesser degree depending on ecosystem attributes (D’Antonio et al. 1992; Hobbs and Huenneke 1992). An ecosystem’s resilience to disturbance and resistance to plant invasions have been shown to be affected by resource availability, including atmospheric CO2 (Dukes and Mooney 1999; Weltzin et al. 2003b), soil nutrients, and water (Davis et al. 2000; Seabloom et al. 2003), as well as temperature and precipitation (Alward et al. 1999; Dukes and Mooney 1999; Walther et al. 2002). Plant Responses to Climate Change Climate has long been associated with global species distributions (Von Humboldt and Bonpland 2010) and more recent work has provided physiological evidence supporting the importance of climate, specifically precipitation and temperature, for plant growth and determining global species distributions (Woodward and Williams 1987). Changes to global temperature and precipitation regimes have effected significant changes to both the biotic and abiotic aspects of global ecosystems (Chapin et al. 2000; Walther et al. 2002; Parmesan and Yohe 2003; Chen et al. 2011; Westerling et al. 2011). 3 Plant Responses to Precipitation and Water Availability Precipitation acts as the primary control of global plant production (Knapp and Smith 2001; Weltzin et al. 2003a; Huxman et al. 2004; Luo et al. 2008; Wu et al. 2011); this is especially true for semi-arid grassland and sagebrush ecosystems (Graetz et al. 1988; Knapp and Smith 2001; Weltzin et al. 2003a; Huxman et al. 2004; Dalgleish et al. 2011; Cherwin and Knapp 2012; Xian et al. 2012; Ren et al. 2015). In a study comparing six North American grasslands, Knapp et al. (2015) found an inverse relationship between sensitivity to drought and mean annual precipitation; arid and semi-arid grasslands were the most sensitive to the effects of drought. Consistent with this, a study in the Colorado shortgrass-steppe demonstrated that when precipitation was experimentally decreased using rainout shelters production was reduced by up to as much as 51 percent (Cherwin and Knapp 2012). Reductions in semi-arid production have also been shown to accompany increases in water availability. In a long-term rainfall manipulation with Mediterranean and semi-arid sites, production was reduced in both the drought and the irrigated treatments (Hänel and Tielbörger 2015). Hänel and Tielbörger (2015) concluded that the reduction in production within the irrigated treatment was a result of increased competition. Altered precipitation and water availability are important factors influencing plant-plant interactions and plant community composition (Whisenant and Uresk 1989; Anderson and Inouye 2001; Everard et al. 2010; Dalgleish et al. 2011; Chambers et al. 2014a; Chambers et al. 2014b; Prevéy and Seastedt 2014). Changes in precipitation along with land-use in the Southwest United States have shifted communities from being arid grasslands to being shrub dominated sites (McCarty 2001; Peters et al. 2015). Similarly, a study that used historical data for a location within the Idaho sagebrush steppe, demonstrated that plant community 4 composition was significantly different between drought years and those with average or above average precipitation (Anderson and Inouye 2001). Likewise, in an arid system, the reduction in precipitation over a five-year period resulted in the reduction of the dominant grass, a shift in the plant community functional abundance, and species composition, which ultimately led to the formation of a new plant community (Báez et al. 2013). The timing of precipitation is also an important factor affecting plant growth and plant community dynamics (Fay et al. 2000; Heitschmidt et al. 2005; Bates et al. 2006; Prevéy and Seastedt 2014). For example, a grassland experiment utilizing rainout shelters to alter the timing of precipitation found that reductions in rainfall quantity alone had little to no effect, however, when rainfall was reduced at crucial times, the effects were significant for the plant community (Fay et al. 2000). A Montana study, which also utilized rainout shelters, found that total production, dominated by cool season perennial grasses, was reduced by 20-40 percent when spring (April, May, June) precipitation was experimentally limited (Heitschmidt et al. 2005). Contrasted with these findings are those of Miranda et al. (2011), which found few consistent results from their decreased precipitation experiment in their Spanish shrub-steppe system. They concluded that responses were species specific and their semi-arid system demonstrated great resilience to precipitation manipulation. In general, changes in precipitation patterns have been observed to decrease ecosystem resilience to community change and reduce resistance to non-native invasive species (Milchunas and Lauenroth 1995; Dukes and Mooney 1999; Weltzin et al. 2003b; Everard et al. 2010; Hoeppner and Dukes 2012; Prevéy and Seastedt 2014; Prevéy and Seastedt 2015). In a greenhouse study, Robinson and Gross (2010) found that both the emergence and growth of two different non-native plant species responded to precipitation variability. An experiment in 5 the Colorado shortgrass-steppe found that experimentally increased winter precipitation, in the form of rain, increased non-native invasive grass species abundance and, when this change was accompanied with decreased summer precipitation, there was a concurrent reduction in native grass cover (Prevéy and Seastedt 2014). Conversely, they also found that when winter precipitation was limited and summer precipitation was increased, non-native invasive species were reduced and native grass species were favored. Change in precipitation patterns, even if they are temporally restricted, can have lasting impacts for plant communities. Another study within the Colorado shortgrass-steppe demonstrated that sixteen years after the cessation of a five-year water manipulation experiment, in which water was added to plots and added with nutrients, found a significant non-native species presence in both treatments compared with no non-native species presence in the control plots (Milchunas and Lauenroth 1995; Dukes and Mooney 1999). A 19th century drought is thought to be a significant factor in the replacement of a California perennial bunchgrass ecosystem by one dominated by invasive annual grasses (Corbin and D’Antonio 2004; Suttle and Thomsen 2007; Everard et al. 2010). These findings contrast with those of a Montana garden experiment, which simulated the invasion of native species monocultures by several different invasive exotic species (Maron and Marler 2008). Maron and Marler (2008) found the invasive species’ to be competitively dominant over the native species regardless of water availability. Plant Responses to Temperature Different plant species have different temperature optima for photosynthesis and growth (Lambers et al. 2008), as such, plant responses to experimentally elevated temperatures have been negative, positive, and neutral (Rustad et al. 2001; Luo 2007). The responses demonstrate the importance of these different optima, as well as the importance of the initial 6 physical and chemical environment of ecosystems (Shaver et al. 2000; Luo 2007). Physiologically, plants can respond to elevated temperatures in different ways and, unless ecosystem conditions are optimal and there is enough of the requisite substrates for the stimulated physiological processes, they become increasingly inefficient and decline in net photosynthesis, thereby limiting production (Luo 2007; Lambers et al. 2008). Conversely, plant responses to increased temperatures can be enhanced through soil nutrient fertilization (Rustad et al. 2001). Plant community responses to experimental warming are significantly shaped by ecosystem water availability (Shaver et al. 2000; Rustad et al. 2001; Luo 2007). Elevated temperatures increase evapotranspiration and can cause soil drying (Shaver et al. 2000; Rustad et al. 2001; Luo 2007), which, if water availability is already limiting, can increase desiccation, heat stress, and reduce aboveground plant growth (Shaver et al. 2000; Rustad et al. 2001; Luo 2007; Lambers et al. 2008; Bloor et al. 2010; Wu et al. 2011). Conversely, if experimental warming is accompanied by sufficient water, plant production is significantly enhanced (Wu et al. 2011). A meta-analysis on ecosystem-level responses to experimental warming, alterations to precipitation, and the combination of the two, found that experimentally warming and drought adversely affected plant production, while increased precipitation and warm conditions increased plant production (Wu et al. 2011). The ambient temperature of an ecosystem can also affect plant responses to elevated temperatures (Rustad et al. 2001). Plant growth in cold ecosystem or during colder times of the year often responds positively to experimental warming (Rustad et al. 2001; Hollister et al. 2005; Walker et al. 2006; Bloor et al. 2010). A meta-analysis of the results from warming experiments across the tundra biome found plant production responded rapidly and largely positively 7 (Walker et al. 2006). Similarly, a separate meta-analysis found that, while there was considerable variation in plant responses to experimental warming, plant productivity in cold regions responded more positively than it did in warmer regions (Rusted et al. 2001). Furthermore, in an upland grassland system, Bloor et al. (2010) found that positive responses in plant production to experimental warming were restricted to the coldest times of the year. However, there have also been cases where warming within cold regions has had negative effects on the plant community; at high elevations, experimental warming can decrease snow pack and increase exposure of high elevation plants to extreme cold events and increase risk of damage due to freezing (Sierra-Almeida and Cavieres 2010). Similarly, the reduction in snowpack and loss of insulation has been shown to negatively affect germination and establishment of plants at higher latitudes (Bokhorst et al. 2013). Regardless of the direction and magnitude of plant production responses to elevated temperatures, experimental warming can favor some species, functional groups, or plant growth forms, and is associated with changes to plant communities (Harte and Shaw 1995; De valpine and Harte 2001; Zavaleta et al. 2003a; Zavaleta et al. 2003b; Walker et al. 2006; Luo 2007). For example, an experiment in Colorado’s sagebrush-steppe found experimental warming increased shrub abundance and decreased forb abundance (Harte and Shaw 1995). Another study in a temperate steppe found that experimental warming reduced grass species richness and community cover (Yang et al. 2011). However, experimental warming has found to increase overall graminoid growth (Hollister et al. 2005; Walker et al. 2006). Given the difference in physiology, elevated temperatures differentially affect C3 and C4 species, which can shape plant community responses to elevated temperatures (Lambers et al. 2008); experimental warming in a Great Plains grassland community enhanced C4 species dominance 8 (Luo et al. 2009). Consistent with the plant community responses to experimental warming, the literature has demonstrated that, among other effects, increasing global temperatures have altered global plant communities and species range distributions (McCarty 2001; Walther et al. 2002; Parmesan 2006; Chen et al. 2011; Wolkovich et al. 2012; Edwards and Henry 2016). A Canadian study found that between 1970-2012 the daily maximum and minimum temperatures of their semi-permanent site had risen by ~1.6 °C and ~2.5 °C, respectively (Savage and Vellend 2015). As a result, over the same time period, the abundance weighted mean elevation of their plant species distributions increased by 9 m/decade. They also found community composition shifts for both their high and low elevation sites, although the high elevation communities experienced a greater change. Interestingly, this change resulted in a loss of beta diversity between their high and low elevation communities. A study within the European Alps comparing current plant species distributions with historical data found an upward shift in range for nine different species consistent with the effects of warming (Grabherr et al. 1994). A California study found no difference in total plant cover along their elevation gradient between 1977 and 2006-2007, however, the distributions of individual species changed significantly; consistent with the effects of observed temperature increases, the mean elevation of nine out of ten focal species significantly increased in elevation, with an average increase in elevation across all the species of 64.7 m (Kelly and Goulden 2008). Global vegetation models predict significant changes to ecosystem plant communities in response to global climate change (Cramer et al. 2001; Lenoir and Svenning 2015). Niche based models simulate significantly negative effects of climate change on species’ ranges in alpine, dry and hot, or Mediterranean climates (Thuiller et al. 2008). Consistent with these community effects and range shifts, it is 9 expected that climate change will increase species turnover in native communities, which may create favorable conditions for invasion by non-native species (Thuiller et al. 2008). Global climate change and the associated increase in global surface temperatures are believed to favor non-native invasive plant species and exacerbate their impacts on global ecosystems (Rejmánek and Richardson 1996; Dukes and Mooney 1999; Dukes 2002; Walther et al. 2002; Hellmann et al. 2008; Thuiller et al. 2008; Willis et al. 2010; Dukes et al. 2011; Polley et al. 2013; Sheppard et al. 2014; Wu et al. 2016). A recent study using a de Wit replacement series design evaluated the effects of elevated temperatures on the competitive dynamic between a non-native and native species (Wu et al. 2016). They found that warming significantly reduced native biomass and increased non-native biomass, and warming significantly reduced the relative yield of the native, especially when planted with higher densities of the non-native species (Wu et al. 2016). Another study using historical data from the Colorado shortgrass- steppe demonstrated that increasing temperatures resulted in a reduction in the dominant grass species and an increase in exotic forb density (Alward et al. 1999). Similarly, another study demonstrated that since 1900 exotic species richness has increased in southern Switzerland, while the number of frost free days has decreased (Walther et al. 2002). Species distribution models relating species’ distribution data with environmental conditions are used to predict how species will respond to climate change (Sheppard et al. 2014). They have predicted range expansions for invasive species due to climate change (Kriticos et al. 2003; Gritti et al. 2006; Bradley 2009; Kleinbauer et al. 2010; Sheppard 2013; Taylor et al. 2014) as well as range contractions (Peterson et al. 2008; Bradley 2009; Bradley et al. 2009; Bourdôt et al. 2012; Gallagher et al. 2013). However, they all indicate that invasive plant species’ ranges, as well as natives, are responsive to increasing global surface temperatures and global 10 climate change (Lenoir and Svenning 2015). Differing responses indicate the importance of the individual species modelling and the importance of field experiments that can provide insight into the factors that might affect the spread of invasive plant species (Sheppard et al. 2014). Bromus tectorum History and Range Bromus tectorum is a Eurasion winter annual grass that was first identified within the United States in Pennsylvania and New York in 1861 (Morrow and Stahlman 1984); however, it is thought that it was separately introduced into the Pacific Northwest as a grain contaminant in 1889 (Mack 1981; Knapp 1996). In the late 19th and early 20th centuries there was widespread grazing throughout the Western United States and, as a grain contaminant, B. tectorum was dispersed throughout the American West via transportation lines (Mack 1981; Knapp 1996). Once established in localized areas, it was dispersed throughout the region, primarily by livestock and feral horses, and by the 1920’s it was ubiquitous throughout the region (Mack 1981); it can currently be found throughout the entire North American continent (Morrow and Stahlman 1984). Despite its ubiquity, B. tectorum has come to dominate vast tracts of land throughout the sagebrush steppe biome of the North American West (Knapp 1996). Bradley and Mustard (2005) estimated that B. tectorum dominates 20,000 km2 within the Great Basin alone, with moderate to high probably of B. tectorum presence predicted for 281,000 km2 of the entire Intermountain West region (Meinke et al. 2009; Miller et al. 2011); invasion by B. tectorum poses a significant threat to the sagebrush-steppe (Miller et al. 2011). There are several reasons that B. tectorum has successfully invaded the Intermountain West: it filled the unoccupied niche of dominant native annual grass within the region (Knapp 11 1996); its high phenotypic plasticity makes it able to withstand and thrive in the highly variable environments of the region (Mack and Pyke 1983; Knapp 1996); it has demonstrated higher relative growth rates than native perennial grasses (Arredondo et al. 1998; Mangla et al. 2011); and its life history strategy as a winter annual provides it with a competitive advantage (Mack and Pyke 1983; Knapp 1996; Chambers et al. 2007). As a winter annual, it has the ability to germinate in the fall, winter, or spring depending on resource availability (Knapp, 1996) and B. tectorum roots can grow in colder temperatures than the roots of its native perennial bunchgrass competitors (Harris 1967). This affords it a competitive advantage: B. tectorum roots establish and grow throughout the winter and are established by the time the early spring precipitation and warmer temperatures ensue, whereas native perennial seedlings must germinate and establish before they can take advantage of the spring conditions (Harris 1967; Mack and Pyke 1983; Knapp 1996). The Impacts of the Bromus tectorum Invasion in the Intermountain West The effects of the B. tectorum invasion in the Intermountain West have been substantial. Bromus tectorum invasion has not only reduced agricultural productivity (Morrow and Stahlman 1984), it has also been associated with significant changes in plant community composition (Mack 1981; Knapp 1996), losses in native biodiversity (Knapp 1996; Bansal and Sheley 2016), alterations to ecosystem nutrient availability and cycling (Belnap and Phillips 2001; Norton et al. 2004), changes to ecosystem water availability and hydrology (Morrow and Stahlman 1984; Boxell and Drohan 2009), and significant disruptions to ecosystem’s fire regimes (Whisenant 1990; Brooks et al.2004). 12 Bromus tectorum Response to Fire and Nutrient Availability B. tectorum’s fall and winter germination often leads to early establishment, growth, and senescence (Mack 1981). The early life cycle provides B. tectorum seedlings an advantage over neighboring perennial grass seedlings (Mack 1981). It also provides ecosystems with an abnormal abundance of highly flammable fine fuels early in the dry summer season (Whisenant 1990; Brooks et al. 2004); B. tectorum sites were ~250% more likely to burn than native sagebrush-steppe sites (Germino et al. 2016 but see Balch et al. 2013). This has resulted in larger fires that return more frequently (Whisenant 1990; Brooks et al. 2004). It has often been observed that B. tectorum responds positively to fire (Billings 1994; West and Yorks 2002; Davies et al. 2007; Mata-González et al. 2007; Davies et al. 2009; Gucker and Bunting 2011). A study with sites along elevation gradients in both Nevada and Utah demonstrated that experimental burning increased B. tectorum biomass and seed production by two to six times (Chambers et al. 2007). This fire facilitation and positive response to fire often begins what has been referred to as the positive feedback grass-fire cycle, where B. tectorum promotes fire and fire promotes B. tectorum, often resulting in dense near monoculture B. tectorum stands (D’Antonio et al. 1992; Knapp 1996). One mechanism thought to be behind this positive feedback is the effect fire has on ecosystem resources (D’Antonio et al. 1992; Blank et al. 2007). The fluctuating resource hypothesis argues that when there is an increase in unused resources, community invasibility increases (Davis et al. 2000). Disturbance has been closely linked with resource availability and invasion (Elton 2000; Crawley 1987; Lodge 1993; Chambers et al. 2007). Specifically, ecosystem nutrient availability, especially nitrogen availability, often increases after fire (Blank et al. 2007; Rau et al. 2008). Bromus tectorum has been associated with modified soil resource availability 13 (Norton et al. 2004) and been shown to respond positively to nitrogen fertilization (Morrow and Stahlman 1984 and references therein). While those areas that have tighter coupling between nutrient availability and plant uptake are more resistant to invasion by B. tectorum (Blank et al. 2007). The results of studies into the effects of fire on B. tectorum’s resource uptake, have suggested that fire enhances its ability to take up several different forms of nitrogen (Grogan et al. 2000), including more recalcitrant nitrogen (Johnson et al. 2011). Furthermore, nitrogen additions have been shown to disproportionately favor B. tectorum’s growth over native species, which has increased its competitive ability, especially with native grasses (Vasquez et al. 2008; He et al. 2011). It has been suggested that variation in ecosystem nutrient availability explains observed variation in B. tectorum density and variation in ecosystem resistance to B. tectorum invasion (Beckstead and Augspurger 2004) and that increased ecosystem nutrient availability enhances B. tectorum invasion (He et al. 2011), especially after fire (Grogan et al. 2000; Johnson et al. 2011). Limiting Factors of the Bromus tectorum-Fire Cycle Despite the numerous studies documenting the positive response to fire by B. tectorum, there have also been numerous studies that have not observed this response (Antos et al. 1983; Cook et al. 1994; Davies et al. 2012; Taylor et al. 2014). The climate where the fire takes place has been shown to affect B. tectorum’s response. Taylor et al. (2014) analyzed 18 studies where B. tectorum demonstrated positive or negative responses to fire and, using a model they developed from climate data, they identified climate factors that best constrained B. tectorum’s positive response to fire. They found that a B. tectorum positive response to fire was most likely in areas with higher annual temperatures and lower summer precipitation. This responsiveness situates B. tectorum’s response to fire in a community context where its response is mediated 14 by ecosystem attributes. Native Plant Community Response to Fire Climate has long been affiliated with shifts in plant communities (Robertson 1939; Weaver 1943; Everard et al. 2010); furthermore, moisture stress after fire has been shown to strongly affect native plant recovery (Redmann et al. 1993; Pylypec and Romo 2003; Prieto et al. 2009; Pratt et al. 2014). In a native grassland community dominated by Festuca, Stipa, and Agropyron species, post-fire production decreased with low soil moisture and increased with elevated soil moisture (Redmann et al. 1993; Pylypec and Romo 2003). In a western South Dakota upland grass community, dominated by Bouteloua gracilis, Stipa comata, Agropyron spp., native grassland production demonstrated significant sensitivity to moisture availability; if there was ample post-burn moisture availability production was elevated compared to unburned levels, whereas, if post-burn water was limited, production was depressed compared with unburned levels (Whisenant and Uresk 1989). In addition to these observational studies, experimentally decreased precipitation and increased temperatures after a burn have been associated with lower community resilience and post-fire recovery by native perennial communities (Prieto et al. 2009; Enright et al. 2014). Bromus tectorum and the Native Plant Community Native plant recovery after fire is important for ecosystem resistance to B. tectorum invasion because robust native perennial grass communities are highly competitive with B. tectorum, often either preventing seedling establishment, or limiting growth after establishment (Orloff et al. 2013; Reisner et al. 2013; Prevéy and Seastedt 2015; Brummer et al. 2016). A greenhouse study investigating the role size plays in the competitive dynamic between B. 15 tectorum and a common native perennial competitor (Pseudoroegneria spicata), found that larger, more established, P. spicata avoided suppression from B. tectorum, whereas smaller and younger P. spicata individuals were out competed and significantly suppressed (Orloff et al. 2013). A large-scale survey of 555 plant communities within sagebrush biome found that native grass cover was the best biotic predictor variable affecting the extent and severity of B. tectorum invasion; native grass canopy cover of 25% or greater was associated with little or no B. tectorum cover (Brummer et al. 2016). Similarly, a study using data from 75 Great Basin sites found that native bunchgrass community structure, abundance, and composition played an important role in limiting B. tectorum dominance (Reisner et al. 2013). Reisner et al. (2013) concluded that disturbance exacerbates B. tectorum invasion through its adverse effects on the native plant communities, which, in turn, decreases ecosystem resistance to invasion. Climate, Native Plant Community, and Bromus tectorum A recent conceptual model has proposed a connection between native community responses to fire, climate, and B. tectorum invasion (Chambers et al. 2014a). Temperature and precipitation regimes are important to an ecosystem’s resistance to B. tectorum invasion because they inform plant community response to disturbance, which, as previously mentioned, is central in defining B. tectorum landscape distribution (Brummer et al. 2016). The model demonstrates that ecosystem resilience and the post-fire recovery by the native plant community and ecosystem resistance to B. tectorum decreases along a temperature and precipitation gradient. According to the model, post-fire recovery by native plant communities will be more complete in cool and wet ecosystems, which would inhibit B. tectorum invasion, while post-fire recovery would be less in warm and dry areas, resulting in a greater chance of post-fire invasion. 16 These model predictions were tested in a large-scale manipulative study, which experimentally burned and mowed at sites within the sagebrush-steppe of five western states (Idaho, Utah, Nevada, Oregon, Washington) (Chambers et al. 2014b). Chambers et al. (2014b) found that B. tectorum’s response differed according to the dominant vegetation of the site, which was associated with a climate regime. Bromus tectorum responded most positively to their disturbance treatments in the warm and dry Wyoming big sagebrush, most negatively in the cool and wet mountain big sagebrush, and demonstrated moderate negative responses at the warm and moist Wyoming big sagebrush sites. Similarly, an Oregon study found that 11 years after a stand-replacing fire, native perennial species dominated the cool and moist areas, while invasive annual grasses, including B. tectorum, dominated the areas with a higher heat load index and climate moisture deficit (Dodson and Root 2015). Consistent with the literature on B. tectorum’s response to fire, climate has also been identified as a significant factor defining those areas within its North American distribution where it has come to dominate and effected significant change in plant communities (i.e. areas within the Intermountain West) (Bradford and Lauenroth 2006; Bradley et al. 2016; Brummer et al. 2016). Knapp (1996) suggested that one reason B. tectorum has come to dominate vast tracts of land within the Intermountain West is the that climate of the Intermountain West resembles the climate of its native range in central and south-western Asia. Effects of Precipitation and Soil Moisture on Bromus tectorum Bromus tectorum germination, establishment, and growth are associated with available soil moisture; when there is sufficient soil moisture, establishment, germination and growth are facilitated (Mack and Pyke 1983; Miller et al. 2006; Bradley et al. 2016), whereas, if soil moisture 17 is limiting these variables are negatively impacted (Chambers et al. 2007; Bradley et al. 2016). An arid-land site demonstrated that B. tectorum only responded in wetter years and concluded that the invasion process would be stunted because there could only be net B. tectorum population growth in favorable water years (Meyer et al. 2001). Furthermore, a study that utilized Landsat and Advanced Very High Resolution Radiometer (AVHRR) found that rainfall patterns effected significant yearly variation in the area dominated by B. tectorum; increased precipitation amplified B. tectorum growth (Bradley and Mustard 2005). While precipitation is certainly important for B. tectorum success, the seasonality and timing of precipitation strongly influences B. tectorum establishment, growth and reproduction (Hulbert 1955; Mack and Pyke 1983; Bradford and Lauenroth 2006; Bradley et al. 2016). A study that utilized a soil water model, in addition to an individual plant based model, to examine the effects of climate, soil type, competition, and seed availability on B. tectorum establishment and success, found climate (specifically winter precipitation) to be the most important factor defining areas susceptible to B. tectorum invasion (Bradford and Lauenroth 2006). A recent in situ field study concluded that changes to precipitation patterns, in the form of increased winter rain, could facilitate B. tectorum expansion along its high elevation range margin (Concilio et al. 2013). Consistent with this finding, the results of two Colorado field studies demonstrated that when winter-spring precipitation in the form of rain was increased B. tectorum responded positively, and when it was reduced B. tectorum responded negatively (Prevéy and Seastedt 2014; Prevéy and Seastedt 2015). They also found that when winter-wet conditions were coupled with summer-dry conditions, the positive B. tectorum response was concurrent with a significant negative response by the native plant community (Prevéy and Seastedt 2014; Prevéy and Seastedt 2015). 18 The positive response by B. tectorum to winter-wet and summer-dry conditions in conjunction with the negative response to these same conditions by neighboring plant communities is consistent with literature on the competition between B. tectorum and perennial neighbors under altered soil moisture conditions. In a greenhouse study, Evans (1961) investigating the competitive interactions between B. tectorum and a common perennial neighbor, found that B. tectorum has an increased ability to extract water from drier soils and is, therefore, more competitive in dry conditions. These same conclusions were reached by Harris (1967) in a study investigating competition between B. tectorum and one of the most common perennial bunchgrasses of the sagebrush biome, Pseudoroegneria spicata. Harris grew the two species in monoculture and mixed pots and noted that when grown in competition, P. spicata roots very minimally entered soil having very low moisture content, whereas, B. tectorum roots penetrated and extracted water from dry soils. Another study determined that B. tectorum required only 66% as much water to produce one gram of biomass as a perennial neighbor; B. tectorum’s high water use efficiency was one of its competitive advantages over other common grasses of the sagebrush-steppe (Hull 1963). While B. tectorum may individually benefit from wetter winter and spring conditions, it has the capability to out compete its perennial neighbors when soil moisture is limiting, which enhances its relative competitiveness and success in those conditions. Consistent with these studies, Bradley (2009), using regional presence B. tectorum data based on a remote sensing map developed by Bradley and Mustard (2005) and average precipitation and temperature data from the PRISM dataset (39 climate variables for the time- period 1971-2000), constructed a bioclimatic envelope model for those areas where B. tectorum is considered invasive. She found that June precipitation and summer (June-September) 19 precipitation, were the climate factors that best constrained B. tectorum’s invasiveness. Similarly, in a large-scale survey of the landscape distribution of B. tectorum Brummer et al. (2016) found the most significant climate variables defining its landscape position were decreased summer precipitation and warm mean temperatures within the driest quarter. Bromus tectorum and Temperature Bromus tectorum tolerates a broad range of temperatures (Thill et al. 1979; Aguirre and Johnson 1991; Nasri and Doescher 1995); at higher temperatures, it has demonstrated greater germination (Thill et al. 1979), growth and reproduction than common perennial competitors (Aguirre and Johnson 1991; Nasri and Doescher 1995), while also demonstrating a capability to tolerate cold winter temperatures and freezing events (Bykova and Sage 2012). However, it is still limited by cold temperature within the western United States (Chambers and Pellant 2008; Brummer et al. 2016); it’s growth rate ceases below 3°C (McCarlie et al. 2001). Cold temperatures have generally decreased B. tectorum germination rates (Thill et al. 1979; Roundy et al. 2007; Bradley et al. 2016), establishment, growth, and reproduction (Chambers et al. 2007; Bradley et al. 2016). Similarly, across a Utah-Nevada elevation gradient, Roundy et al. (2007) found, when ample water was available, spring soil temperature was the best predictor of B. tectorum germination. Chambers et al. (2007) also found that along a similar Utah-Nevada elevation gradient B. tectorum establishment, growth, and reproduction was limited by cold temperatures. B. tectorum’s responsiveness to temperature along its cold range margins, has been studied using experimentally increased temperatures in these areas. A recent Wyoming study that utilized infrared heaters to experimentally warm their plots year-round found that warming, regardless of soil moisture availability, increased B. tectorum biomass and seed 20 production (Blumenthal et al. 2016). In a Utah study at a site close to the lower bounds of the B. tectorum temperature range, Compagnoni and Adler (2014b) found their experimentally warmed plots increased B. tectorum fecundity and survival. Similarly, a study that experimentally increased temperatures at sites across a Utah elevation gradient found that experimental warming significantly increased B. tectorum population growth rate at the cold high elevation site (Compagnoni et al. 2014a). At the warmer mid and low elevation sites, the effects on B. tectorum, while positive, were not significant (Compagnoni and Adler 2014a). Despite these positive responses resulting from the fact that precipitation and soil moisture during the years of this experiment were average and above average, Compagnoni and Adler (2014a) speculated that if soil moisture had been limited, especially at the lower elevation sites, B. tectorum might have responded negatively to the warming. This was what Zelikova et al. (2013) found in their warming study on the Colorado Plateau: B. tectorum responded positively to experimental warming when there was ample soil moisture but when soil moisture was limiting B. tectorum responded negatively. Climate Change and Projections for Western North America Anthropogenic effects on the global climate have significantly increased the temperatures of Intermountain West over the last century (Bonfils et al. 2008; Pederson et al. 2010). Using daily and monthly temperature data from nine reliable western Montana meteorological stations for the last 100+ years, Pederson et al. (2010) found that the number of extremely hot days (≥32.2 °C) per year since 1980 has tripled compared with the previous 80+ years, rising from an average of 5 per year to an average of 15. Concurrently, there has been a decrease in the number of cold (≤0 °C) days and extremely cold (≤-17.8 °C) days per year 21 (Pederson et al. 2010). Over the 1900-2006 period, this northern portion of the sagebrush biome experienced a mean annual temperature increase of 1.3 °C, and its increase between 1900-2005 was 1.8 times greater than the rise in global temperatures over the same period (Pederson et al. 2010). There are many global climate models which project the future climate given different emission scenarios (Mote and Salathé 2010). While there is variability in the temperature projections of the different models, they all consistently project that the northern portion of the western United States will experience an increase in temperatures over the next century (Chambers and Pellant 2008; Mote and Salathé 2010; Polley et al. 2013); mean annual temperatures are projected to increase 2.0 °C - 5.7 °C by the end of the century (Chambers and Pellant 2008; Mote and Salathé 2010; Polley et al. 2013). Projected changes to the precipitation regimes for the region have more overall and seasonal variation (Chambers and Pellant 2008; Mote and Salathé 2010; Polley et al. 2013). However, while projections of annual precipitation changes vary from significant increases to significant decreases (Mote and Salathé 2010), they have been relatively consistent in their seasonal predictions: winter precipitation will increase (and could fall as rain not snow) and summer precipitation will decrease, especially at higher latitudes (Mote and Salathé 2010; Polley et al. 2013). The potential ecological consequences of these changes are significant (Chambers and Pellant 2008; Polley et al. 2013) and models have found that the predicted future climate change will cause shifts in landscape scale vegetation dynamics (Keane et al. 2008). The Projected Expansion of Bromus tectorum One of the likely shifts in vegetative dynamics brought on by global climate change is the spread of invasive species within the region (Chambers and Pellant 2008; Polley et al. 2013), 22 especially B. tectorum (Chambers and Pellant 2008; Bradley 2009; Abatzoglou and Crystal 2011; Taylor et al. 2014; Bradley et al. 2016). A recent observational study in Rocky Mountain National Park site has found that B. tectorum is increasing along its high elevation range margin (Bromberg et al. 2011). Consistent with these studies and how B. tectorum has responded to altered precipitation and temperature, the Bradley (2009) bioclimatic envelope model projects that with increased temperatures and decreased summer precipitation B. tectorum will increase into those areas of the North American West where it had been previously limited by the climate, especially at higher latitudes. Likewise, the model developed by Taylor et al. (2014) projects that the area suitable for the positive B. tectorum-fire feedback will increase as temperatures increase and summer precipitation decreases. These changes will not take place in a vacuum and will be affected by concurrent changes in atmospheric CO2 concentration. Plant Responses to Elevated Atmospheric CO2 Concentrations CO2 and Plant Growth and Competition Photosynthesis fixes the carbon required for plant growth (Lambers et al. 2008). Carbon dioxide (CO2) is the primary source of this carbon and increasing atmospheric CO2 levels has long been used to increase plant growth in controlled settings; however, its effects depend on plant and ecosystem attributes (Bazzaz 1990). Plants have different photosynthetic pathways, which have different CO2 requirements and mechanisms for obtaining their requisite carbon (Lambers et al. 2008). As such, plants of these different pathways respond differently to elevated CO2 concentrations, with plants of the C3 pathway experiencing the greatest benefit (Poorter and Navas 2003; Wang et al. 2012). Other plant attributes, such as functional group, reproductive/life history strategy, and relative growth rate, also affect plant responses to 23 elevated CO2 levels (Zangerl and Bazzaz 1984; Poorter and Navas 2003; Ziska and George 2004; Polley et al. 2012; Wang et al. 2012). For example, in separate meta-analyses, Wang et al. (2012) and Poorter and Navas (2003) found that responses varied between functional group, and Polley et al. (2012) demonstrated that grass production was favored under elevated CO2 conditions. Zangerl and Bazzaz (1984) concluded that annual species may display greater production responses than perennials and the Poorter and Navas (2003) meta-analysis concluded that fast- growing species benefit more from elevated CO2 than do slow growing species. Ecosystem attributes such as plant density and competition have been shown to affect plant responses to elevated CO2 concentrations (Bazzaz and Carlson 1984; Bazzaz 1990; Bazzaz et al. 1992; Ackerly and Bazzaz 1995; Dukes 2002). CO2 can magnify the effects of competition and can enhance the growth of the dominant individuals (Bazzaz 1990; Manea and Leishman 2011). For example, in single species studies Bazzaz et al. (1992) and Wayne et al. (1999) found that responses to elevated CO2 were depressed by increasing density and intraspecific competition. Dukes (2002) found Centaurea solistitialis responded positively to elevated CO2 concentrations when it was grown in monoculture, however, in a community setting it failed to respond significantly. Ecosystem resource levels, especially water availability, also inform plant responses to elevated CO2 (Bazzaz 1990; Polley 1997; Shaw et al. 2002; Morgan et al. 2004; Smith et al. 2014; Zelikova et al. 2015). Elevated CO2 concentrations reduce stomatal density and conductance, which decreases transpiration and increases water use efficiency (Bazzaz 1990; Polley 1997; Morgan et al. 2004). This has resulted in increased plant production in dry conditions (Bazzaz 1990; Polley 1997; Morgan et al. 2004). For example, using open top chambers and rainout shelters to evaluate the interactive effects of warming and drying on plant communities within 24 ambient and elevated CO2 conditions, Dermody et al. (2007) found that elevated CO2 mitigated the negative effects of the warming and drying treatment. The results of seven year free-air CO2 enrichment (FACE) Wyoming field study demonstrated that elevated CO2 had positive effects on the plant community aboveground biomass during dry years; however, this effect was reduced to negligible levels in those growing seasons with high soil moisture and precipitation (Mueller et al. 2016). The interaction between elevated CO2 and soil moisture can affect community composition and plant-plant interactions (Owensby et al. 1999; Morgan et al. 2004; Polley et al. 2012). In their study, Polley et al. (2012) found that the competitive effect of tall-grass species on mid-grass species intensified when soil moisture content was increased in conjunction with elevated CO2 concentrations. Invasive non-native plant species as a group often have high phenotypic variation, high growth rates, and the ability to adapt to rapidly changing conditions, which is why they are believed to be favored by elevated atmospheric CO2 concentrations (Dukes and Mooney 1999; Weltzin et al. 2003b; Moore 2004; Ziska and George 2004; Sorte et al. 2013). In a meta-analysis of invasive species responses to elevated CO2 concentrations, Ziska and George (2004) found that, while both native and invasive species responded positively to elevated CO2, invasive responses were significantly greater than native responses. Ziska and George (2004) did note that the studies they analyzed were primarily monoculture growth studies and there was a general dearth in elevated CO2-invasive species community and competition studies. Over the last decade more competition and community studies investigating native- invasive responses to elevated CO2 concentrations have been conducted and results have been mixed. In a controlled environment competition experiment between two C3 grasses, one native and one non-native, Hely and Roxburgh (2005) demonstrated that the non-native was 25 always able to suppress the native regardless of atmospheric CO2 concentrations. Similarly, Manea and Leishman (2011) evaluated competition between 14 pairs of native-non-native invasive species and found that while the invasive species were more competitive regardless of CO2, the strength of the competitive effects were heightened in elevated CO2. Contrasting with these findings are those of two studies focused on how two non-native species, Centaurea solistitialis and Chenopodium album, responded to elevated CO2 concentrations (Taylor and Potvin 1997; Dukes 2002). Both these studies found that when the respective non-native species was grown in monoculture it responded positively, while both failed to respond significantly when grown in a community setting (Taylor and Potvin 1997; Dukes 2002). Non-native invasive species responses to elevated CO2 have also been shown to be responsive to other site factors. For example, in a long-term FACE study, Smith et al. (2014) found that the dominant invasive annual grass species of their arid study site responded positively to elevated CO2, but its response changed through time and depended on water availability. Bromus tectorum and CO2 As an annual C3 grass species, with a high growth rate, which is highly competitive with its neighbors when both are grown from seedlings (Orloff et al. 2013), B. tectorum has many of the traits that have been shown to be favored under elevated CO2 concentrations. Consistent with this working hypothesis, several studies have found that B. tectorum has responded positively to elevated atmospheric CO2 concentrations. Single factor, monoculture studies have demonstrated that a doubling in atmospheric CO2 positively affected B. tectorum aboveground biomass (Smith et al. 1987; Poorter 1993; Ziska et al. 2005), height, leaf area, and number of basal stems, as well as increased its water use efficiency (Smith et al. 1987). Increases in atmospheric CO2 concentrations have also been shown to increase the C:N ratio of B. tectorum 26 tissue, making it not only less palatable but also more flammable (Ziska et al. 2005; Blank et al. 2011). These results have led some to suggest that the effects of increased atmospheric CO2 have already affected B. tectorum growth, establishment, and facilitated its positive feedback with fire (Ziska et al. 2005) Similarly, others have suggested that increased B. tectorum productivity and changes in tissue composition, accompanying increasing atmospheric CO2, will likely continue to increase B. tectorum induced fire frequency and the extent of such fires (Chambers and Pellant 2008). As previously established, invasive species’ responses to elevated atmospheric CO2 within a community setting are significantly different from their responses when grown in monoculture (Taylor and Potvin 1997; Dukes 2002). There has only been one study that has researched B. tectorum’s response to elevated CO2 in either a community or competitive setting. A two-year southern Wyoming FACE study found that in their five elevated CO2 plots B. tectorum failed to demonstrate a response to the treatment and when elevated CO2 was accompanied by warming it only demonstrated a slight increase in density, which was attributed to the effects of the warming treatment (Blumenthal et al. 2016). As the only study that has researched the effects of elevated CO2 on B. tectorum in a community, its limited duration (2 years) and number of replicates (n=5) leaves plenty of room for further research. An Experimental Approach to Understanding How Global Climate Change Will Affect Bromus tectorum Using climate data from B. tectorum’s current range, it has been predicted that under future climate conditions, B. tectorum, its ecological dominance, and its positive feedback to fire will increase in the northern areas of the sagebrush-steppe where it is currently limited by the climate (Bradley 2009; Taylor et al. 2014). The validity of bioclimatic envelope models, which 27 uses the climate-space of a species and the projected change to predict how a species will respond to climate change scenarios (Davis et al. 1998a), has been called into question because they do not incorporate species-species interactions (Davis et al. 1998a; Davis et al. 1998b). Therefore, in situ field experiments and controlled setting studies still hold value as they can be used to test and inform the predictions of the bioclimatic envelope models. Previous B. tectorum research has demonstrated that it has responded positively to increased temperatures along its high elevation range margin (Compagnoni and Adler 2014a; Compagnoni and Adler 2014b) and within a southeastern Wyoming mixed prairie (Blumenthal et al. 2016); however, to date, no study has experimentally tested the prediction that both B. tectorum (Bradley 2009) as well as its positive response to fire (Taylor et al. 2014) will expand into the northern sagebrush-steppe where they are currently limited by the climate. Chapter two describes how I used a manipulative field study, repeated over three years, within a sagebrush-steppe plant community of southwestern Montana to address this gap in knowledge. I hypothesized: (1) Experimental warming, and decreased growing season precipitation would positively affect B. tectorum’s abundance within a Montana sagebrush community; (2) B. tectorum would respond positively following fire in an experimentally warmed and dried Montana sagebrush-steppe; (3) Fire, experimental warming, and decreased growing season precipitation would increase B. tectorum’s competitive effect on the dominant perennial bunchgrass, Pseudoroegneria spicata, and the native grass community. There have been studies that have investigated B. tectorum’s response to altered water and nutrient availability (Evans 1961; Hull 1963; Harris 1967; Dakheel et al. 1994; Adair et al. 2007; Orloff et al. 2013) and temperature (Harris 1967; Aguirre and Johnson 1991; Dakheel et al. 1994; Nasri and Doescher 1995). However, there have not been any studies investigating the 28 effects of altered water and nutrient availability under ambient and elevated temperatures between B. tectorum and established P. spicata. Nor, have there been any studies investigating the effects that altered water availability will have on B. tectorum while in competition with an established native perennial grass in ambient and elevated atmospheric CO2 concentrations. Accordingly, Chapter three describes two experiments designed to fill in these gaps in knowledge. By altering temperature, available soil moisture, and nutrient levels, the first goal of these studies was to determine if, in a controlled setting, the dynamic between B. tectorum and an established native perennial bunchgrass (P. spicata) was responsive to these factors. The second goal of the controlled setting experiments was to determine if competition between established P. spicata and B. tectorum was responsive to elevated CO2 concentrations and if decreased water availability impacted this response. Specifically, I hypothesized that: (1) Increasing the temperature and reducing the available soil moisture would alter the competitive dynamic between established P. spicata individuals and B. tectorum in favor of B. tectorum; (2) B. tectorum would respond positively to increased nutrient availability and this would increase its competitiveness with the established P. spicata individuals; (3) B. tectorum would respond positively to elevated CO2 concentrations, which would also facilitate an increase in competitiveness with established P. spicata individuals; (4) Decreasing soil water availability would augment B. tectorum’s competitiveness within elevated CO2 conditions. 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Introduction The sagebrush-steppe biome covers more than 43 million hectares and is one of the largest ecosystems in North America (Rowland et al. 2010). The sagebrush-steppe provides productive rangelands (Rowland et al. 2010), acts as an important carbon sink (Gilmanov et al. 2006) and fosters biodiverse native communities that provide habitat for threatened species (Klebenow 1969; Miller et al. 2011). This region has an extensive history of disturbance (grazing, fire, development) (Knapp 1996; Rowland et al. 2010), which continues today. Understanding the effects of these disturbances in a changing climate is important for maintaining their diversity and productivity. One of the most significant results of disturbance within this region has been its role in the spread of non-native invasive plant species, which negatively impact its biodiversity and productivity (Rowland et al. 2010). The plant species that has had the most negative impact and poses the most significant threat to the sagebrush-steppe is the invasive annual grass Bromus tectorum (Suring et al. 2005). Bromus tectorum, accidently introduced in the 1880’s, was ubiquitous throughout western North America by the 1920’s (Mack 1981; Billings 1994; Knapp 1996) and is currently naturalized throughout North America (Morrow and Stahlman 1984). Its invasion has been closely tied with anthropogenic disturbance; grazing facilitated its dispersal and early establishment, and a positive feedback with fire has led to its ecological dominance in some areas (Hull 1965; Mack 1981; Knapp 1996; Taylor et al. 2014). This dominance has largely been 47 constrained to the Great Basin and Columbia Plains regions (Mack 1981; Knapp 1996; Bradley 2009; Brummer et al. 2016). Research into the mechanisms behind this dominance and what has constrained it to these regions has illuminated an interesting dynamic between native grass communities, disturbance, climate, and ecosystem resistance to B. tectorum invasion (Chambers et al. 2014a; Chambers et al. 2014b). Robust native perennial grass communities are the single most important factor limiting B. tectorum (Chambers et al. 2007; Compagnoni and Adler 2014a; Prevéy and Seastedt 2015; Brummer et al. 2016). However, if the native grassland communities have been disturbed, by either fire or grazing, B. tectorum presence has often increased (Mack 1981). Therefore, how a native community responds to disturbance informs ecosystem susceptibility to B. tectorum invasion (Chambers et al. 2014a). It has been demonstrated that sagebrush-steppe community resilience to disturbance is responsive to both temperature and precipitation (Chambers et al. 2014a); warmer and drier ecosystems being less resilient to disturbance than those that are cooler and have more abundant available moisture (Chambers et al. 2007; Chambers et al. 2014a; Chambers et al. 2014b). The lower resilience of warmer and drier ecosystems has made them more susceptible to B. tectorum invasion (Chambers et al. 2007; Chambers et al. 2014a; Chambers et al. 2014b; Taylor et al. 2014). Consistent with these studies, warm and dry summer climate conditions is a key abiotic factor defining B. tectorum distribution throughout the sagebrush biome (Bansal and Sheley 2016; Brummer et al. 2016) Climate models for western North American project temperatures to increase by 2-4 oC by 2100 and summer precipitation to decrease for Intermountain West locations higher in latitude (Mote and Salathé 2010; Pederson et al. 2010; Polley et al. 2013). Accordingly, species distribution models have projected an expansion of B. tectorum dominance at both higher 48 elevations and higher latitudes (Bradley 2009; Taylor et al. 2014). Consistent with these models, field studies conducted in cool and moist areas of B. tectorum’s range have demonstrated that experimentally increased temperatures have positively affected B. tectorum (Compagnoni and Adler 2014a; Compagnoni and Adler 2014b; Blumenthal et al. 2016). Similarly, other studies have found B. tectorum to be expanding along its high elevation range margin (Bromberg et al. 2011). Experimental changes to precipitation regimes have likewise increased B. tectorum seed production and cover (Prevéy and Seastedt 2015). Similarly, an investigation into the climate conditions that define the geographical limits of the B. tectorum-fire cycle concluded that those locations where B. tectorum has demonstrated a positive response to fire were warmer and had less summer precipitation (Taylor et al. 2014). Accordingly, Taylor et al. (2014) forecasted an increase in suitable climate for the B. tectorum-fire cycle throughout the American West under warmer and drier conditions. In Montana, B. tectorum is naturalized and a problematic agricultural weed (Mcvay et al. 2009), however, in this cold and moist northern region, there have not been any documented cases of B. tectorum dominating natural ecosystems by forming dense monocultures nor any cases of the B. tectorum-fire cycle (Taylor et al. 2014). A change in climate may not only facilitate the spread of B. tectorum within this region but could also effect a change from its current subordinate community role to what has been referred to as a ‘transformer’ and initiating the B. tectorum-fire cycle (Richardson et al. 2000; Hellmann et al. 2008; Taylor et al. 2014). At the time of this study, while small scale field studies have addressed the possibility of a B. tectorum range shift, no studies have addressed the potential for climate change to induce a B. tectorum community role change, particularly at a latitude where it has been limited by cold temperatures. Thus, my goal was to assess the response of B. tectorum and the native plant 49 community through combinations of experimentally increased growing season temperatures and decreased precipitation, in addition to a spring prescribed burn. My hypotheses were: (1) Experimental warming, and decreased growing season precipitation will positively affect B. tectorum’s abundance within a cool and moist northern range margin. (2) B. tectorum will respond positively to fire in a cool and moist sagebrush-steppe after it has been experimentally warmed and dried. (3) Fire, experimental warming, and decreased growing season precipitation will increase B. tectorum’s impact on the dominant perennial bunchgrass Pseudoroegneria spicata. Methods Site Description The field site was a sagebrush-steppe rangeland located 56 kilometers west of Bozeman, MT, at the Montana State University Red Bluff Agricultural Research Station in Norris, MT, USA (5049898.184 N., 451464.866 E. (UTM)) at an elevation of 1600 m. Although the site has a history of low density grazing, our experiment was fenced off and had a robust plant community. The vegetation of the site was dominated by the native species Ericameria nauseosa and Artemisia species in the shrub layer, and P. spicata, Stipa comata, and Lupinus argenteus in the herbaceous layer. Invasion of B. tectorum at the site made it the most abundant non-native species. Other non-native species included: Allysum desertorum, Sisymbrium altissimum, and Tragopogon dubious (using nomenclature of Lesica et al. (2012)). The soils of the Red Bluff site were part of the Nuley-rock outcrop complex; sandy loam (0-10 cm), sandy clay loam (10-28 cm), gravelly sandy loam (28-61 cm) (http://websoilsurvey.nrcs.usda.gov/app/websoilsurvey.aspx). 50 Historical temperatures (Norris climate station, 16 kilometers south of the Red Bluff research station, Table 2.1; NCEI 2016) for the site were similar but cooler than the temperatures for the years of our experiment (Table 2.2). For the last 30 years (1984-2013), April-June have been the wettest months, receiving an average of 46% of the annual precipitation (NCEI 2016; Table 2.1). For the same period (1984-2013), January-March have been the driest months, receiving an average of 12% of the annual precipitation (NCEI 2016; Table 2.1). December, January, and February are the coldest months with averages ranging between -2.90 oC and -1.0 oC. The warmest months are July and August, with monthly averages of 21.05 oC and 20.34 oC, respectively (NCEI 2016). Table 2.1. Mean temperature and precipitation data for Norris, MT climate station (1908-2016). Data was downloaded from the National Centers for Environmental Information (http://www.ncdc.noaa.gov). Not all 2016 data was available. Year Mean (°C) Minimum (°C) Maximum (°C) Total precipitation (mm) Annual 1908-1983 8.10 1.80 14.40 451.10 1984-2013 9.09 2.82 15.28 424.15 2014 9.54 3.73 15.33 611.64 2015 10.81 4.88 16.79 563.36 Jan.-Mar. 1908-1983 -1.15 -6.24 3.94 63.85 1984-2013 0.85 -4.17 5.84 50.85 2014 1.06 -3.48 5.57 128.53 2015 4.72 -0.56 9.98 46.98 2016 4.07 -0.76 8.91 58.17 Apr-June 1908-1983 11.52 4.71 18.32 196.76 1984-2013 12.24 5.55 18.89 194.90 2014 12.41 6.11 18.68 217.17 2015 13.59 7.18 19.98 232.67 July-Sept. 1908-1983 18.57 10.45 26.67 113.06 1984-2013 19.35 11.33 27.34 106.36 2014 19.29 11.94 26.60 170.69 2015 20.02 12.39 27.74 137.41 Oct.-Dec. 1908-1983 3.33 -1.90 8.55 78.26 1984-2013 3.53 -1.57 8.60 70.07 2014 5.59 0.34 10.44 95.25 2015 4.93 0.50 9.44 146.30 51 Experimental Design The questions were addressed using a growing season experiment with two burn treatments (burn and unburned) and three different climate treatments (control, warming, warming and drying) that were implemented between April and July for 2014, 2015, and 2016. A 30m x 60m grid with uniform aspect and slope with natural densities of both P. spicata and B. tectorum was established, in which, the six treatments were replicated ten times (2014) and six times (2015 and 2016). Plot locations were randomly assigned within the established grid at the beginning of each growing season. In 2015, an additional 6 replicates were assigned to each treatment and left through the winter. However, due to intense small mammal soil disturbance and herbivory in these plots their data were excluded. In accordance with local fire restrictions, the prescribed fire took place in early April for all three years. Consistent with prescribed burning literature (Sirois 1993; Kral et al. 2015) a propane torch was used (Red Dragon model VT21/2 vapor propane torch by Flame Engineering Corporation, La Crosse, Kansas) in conjunction with a temporary circular fire block created out of aluminum flashing material around each plot (Jones et al. 2015). Climate manipulation structures were established in April after the prescribed burn had taken place. Delmhorst soil moisture measuring systems (model KS-D1) were installed in the center of each plot at a depth of 10 cm (Aho and Weaver 2008); in 2014 and 2016 soil moisture readings were taken weekly and in 2015 readings were taken bi-weekly. Temperatures were recorded using Maxim Integrated thermochron I-button devices (DS1921G; -40Co: - 85Co) that were deployed in each plot on the north side of a stake 20 cm above the ground surface. Temperature was recorded every three hours. 52 Climate Manipulation Designs Open Top Chambers. Open top chambers (OTC) were used to increase the temperature of the warming and warming and drying treatments, following the cone chamber design (Molau and Molgaard 1996; Marion et al. 1997). They were constructed out of Sun-Lite HP (1mm thick) (Solar Components Corporation, Manchester, New Hampshire). The fiberglass material has a relatively high solar transmittance in the visible wavelengths (86%) and a low transmittance in the infra-red range (<5%). The basal diameter of the cone was 1.48 m, with the diameter of the top opening 1.01 m. The chambers were 40 cm tall with a 60o incline (Marion et al. 1997). Yearly temperature data were analyzed separately using linear mixed-effects models; Julian day were treated as random effects and the climate and burn treatments as fixed effects (Marion et al. 1997; Dabros et al. 2010). The burn treatment did not affect temperatures. The OTCs significantly increased both the average mean and maximum temperatures of the warming, and warming and drying treatments, while the warming and drying treatment significantly increased minimum temperatures (Table 2.2). 53 Table 2.2. Temperature response to climate treatments recorded at Red Bluff Research Station 2014-2016. April temperatures were not recorded in 2014. Year data were analyzed separately and the superscripts indicate statistical significance (p<0.05). Year Treatment Mean (°C) Minimum (°C) Maximum (°C) April 2015 Ambient 7.19a -2.78k 19.77v Warming 8.80 b -2.77k 24.23w Warming & Drying 9.10b -1.75l 23.51w 2016 Ambient 9.48c 1.26m 21.68x Warming 10.89 d 0.98m 25.56y Warming & Drying 11.09d 1.66n 24.96y May-July 2014 Ambient 17.00e 5.86o 29.77z Warming 18.40 f 5.85o 34.00aa Warming & Drying 18.68f 6.53p 32.57bb 2015 Ambient 18.08g 7.76q 31.62cc Warming 19.96 h 7.80q 36.25dd Warming & Drying 20.15h 8.74r 35.05dd 2016 Ambient 18.69i 6.59s 33.19ee Warming 20.47 j 6.26t 37.59ff Warming & Drying 20.58j 7.29u 36.36ff Rainout Shelters. Rainout shelters following the design specified in Yahdjian and Sala (2002) were placed over OTCs to decrease the precipitation for the warming and drying treatment. They had wooden frames supporting gutters made of corrugated clear polycarbonate material (Suntuf). Suntuf clear polycarbonate was chosen as the material because of its high light transmittance 90%, its toughness and flexibility to withstand high winds and inclement weather. The gutters extended 0.10 m beyond the OTC on the high side and 0.20 m on the low side to maximize interception and to prevent capillary action of soil water from inadvertently watering the vegetation. To maximize their effectiveness, the rainout shelters were oriented southwest towards the prevailing winds. Soil moisture readings from the 54 Delmhorst gypsum block sensors were converted to soil water potential using the conversion equation of Aho and Weaver (2008). The soil moisture data for all three years were analyzed separately using linear mixed-effects models, where Julian day was treated as a repeated measure. The results provided substantial evidence that the mean soil water availability for all three years (2014, 2015, 2016) was significantly reduced by the rainout shelters (p=2.22e-12, p=3.96e-11, and p=4.46e-07, respectively). Soil water availability was not affected by either the burn treatment or the warming treatment. Sampling Methods Percent cover of the target species (B. tectorum and P. spicata) and all other species, as well as litter, rock and bareground was assessed for a 0.75 m2 plot centered within the plots. In addition to percent cover, average height and density of flowering tillers were assessed for both target species. For these additional metrics, in 2015, subsets of the B. tectorum plot populations were taken using three rings with a diameter or 11 cm placed randomly over the populations within each plot, while in 2014 and 2016 the metrics were evaluated for the entire 0.75 m2 plot. Sampling began early May and continued weekly in 2014 and biweekly in 2015 until June 30th, at which time the final community assessments were taken. In 2016, cover, height, and density were only evaluated on June 30. For all years, biomass samples were destructively taken for both B. tectorum (June 30th) and P. spicata (July 14th). Data Analysis Linear mixed-effects models were used to assess the effects of warming, warming and drying, the burn treatment, and other explanatory variables on the B. tectorum response variables assessed at maximum biomass. Climate manipulation, burn treatment, native grass 55 cover, and forb cover were treated as fixed effects, while year was treated as a random effect. To evaluate relative importance of the different predictor variables (climate and burn treatments, as well as forb and native grass cover) on B. tectorum’s cover, an Akaike Information Criterion (AIC) table was developed. The relative importance of each variable in accounting for the variation in B. tectorum cover values was established by comparing AIC values for different linear mixed-effects models; the difference in AIC values for each model producing the importance of the tested factor. Using an AIC stringency of any change above two, I used a stepwise approach to remove predictor variables from the models that did not explain significant variation in response variables (Prevéy and Seastedt 2014). Linear mixed-effects models were also used to assess the effects of warming, warming and drying, the burn treatment, and B. tectorum cover on the P. spicata response variables. Using a stepwise approach, I again removed those predictor variables that did not account for significant variation in the response variables and identified the best model by comparing their AIC values (lowest value signifying the best model). The effects of the climate and burn treatments on native grass cover were also assessed using a linear mixed-effects model. Significant differences between predictor variables and response variables at the p<0.05 level were calculated from F statistics based on Satterthwaite’s approximations of degrees of freedom for linear mixed-effects models (Kuznetsova et al. 2014). Data were analyzed using the lme4 package (Bates et al. 2011) and the lmerTest package (Kuznetsova et al. 2014) in the statistical analysis program R (R Development Core team 2015). 56 Results Bromus tectorum cover responded negatively to the warming and drying treatment when in plots that were not burned. There were significant differences between warming and drying unburned B. tectorum cover values and both the unburned and the burned ambient treatments (p=0.04 and 0.03, respectively; Fig. 2.1). There was no evidence supporting differences between B. tectorum cover values of the other treatments (Fig. 2.1). Figure 2.1. Bromus tectorum cover within the six climate-burn treatments. Letters indicate significant differences (p<0.05). To satisfy model assumptions, cover data were natural log transformed. Of the variables that explained the variation in the B. tectorum cover, native grass cover was the most important variable (its exclusion resulting in an AIC change of 39.17; Table 2.3). Climate was the second most important explanatory variable (AIC change of 3.11; Table 2.3). 57 Table 2.3. Akaike Information Criterion (AIC) table for B. tectorum cover models. Delta AIC indicates the relative importance of the explanatory variable when included. To satisfy model assumptions, data were transformed and the exact form is illustrated below. Model Tested Factor AIC Δ AIC •Full model NA 338.12 0 •Full-Forb Forb 336.84 -1.28 •Full-Native grass Native grass 377.29 39.17 *Climate + Veg Burn treatment 337.4 -0.72 *Burn + Veg Climate treatment 341.23 3.11 *Climate + Burn + Veg Climate:Burn interaction 336.54 -1.58 •Full model=lmer (ln(B. tectorum cover)~(climate x burn status) + native grass + forb + (1|year) *Veg=Native grass + Forb Results of the most parsimonious B. tectorum cover model provided substantial evidence supporting a negative relationship between B. tectorum cover and native grass cover (p=1.51e-10; Table 2.4; Fig 2.2). There was also substantial evidence that B. tectorum cover was significantly reduced by the warming and drying treatment (p=6.50e-03; Table 2.4; Fig. 2.2) and considerable evidence that the warming treatment also negatively affected B. tectorum cover (p=0.03; Table 2.4; Fig. 2.2). B. tectorum cover was not significantly affected by the burn treatment (p=0.10; Table 2.4). 58 Figure 2.2. Bromus tectorum cover response to native grass cover by climate treatment. B. tectorum was negatively affected by native grass cover and both climate treatments. B. tectorum seed production per plot and reproductive tiller density per plot demonstrated negative relationships with native grass cover (p=5.22e-06 and 1.02e-08, respectively; Table 2.4) and were negatively affected by both warming (p=0.02 and p=9.95e-03, respectively; Table 2.4) and warming and drying (p=4.80e-03 and p=3.89e-04, respectively; Table 2.4). Neither response variable was significantly affected by the burned treatment (p=0.12 and p=0.06; Table 2.4). B. tectorum biomass was negatively affected by native grass cover and the warming and drying treatment (p=4.43e-07 and p=0.02, respectively; Table 2.4), while it demonstrated a marginal response to the warming treatment and no response to the burned treatment (p=0.08 and 0.34, respectively; Table 2.4). 59 Table 2.4. Results of the best linear mixed-effects models assessing the Bromus tectorum response to the burn and climate treatments, and native grass cover. To satisfy model assumptions, the B. tectorum data were natural log transformed. P-values and degrees of freedom (df) were calculated using Satterthwaite approximation and values in bold indicate statistically significant differences (p<0.05). Response variables were assessed on a per plot scale except Ind. fecundity, which is mean individual fecundity. Seed prod. is seed production. Density is number of reproductive tillers. Nat. grass is native grass cover. Fixed effects Random effects Response Predictor Est. SE df t value p(>) Variance Year Residual Cover Intercept 3.65 0.43 4.50 8.54 6.05E-04 0.34±0.59 0.70±0.84 Burned -0.25 0.15 120.98 -1.63 0.10 Warming -0.39 0.18 120.99 -2.14 0.03 Warm & Dry -0.50 0.18 120.97 -2.77 6.50E-03 Nat. grass -0.06 0.01 121.21 -7.00 1.51E-10 Biomass Intercept 3.08 0.45 5.52 6.78 7.12E-04 0.43±0.66 0.82±0.90 Burned 0.16 0.17 125.01 0.95 0.34 Warming -0.35 0.20 125.02 -1.77 0.08 Warm & Dry -0.45 0.19 125.00 -2.32 0.02 Nat. grass -0.05 0.01 125.23 -5.33 4.43E-07 Seed prod. Intercept 8.21 0.89 5.58 9.26 1.36E-04 1.67±1.29 2.11±1.45 Burned 0.43 0.27 120.00 1.58 0.12 Warming -0.77 0.32 120.00 -2.36 0.02 Warm & Dry -0.92 0.32 120.01 -2.87 4.80E-03 Nat. grass -0.09 0.02 120.12 -4.77 5.22E-06 Density Intercept 6.80 0.60 7.51 11.54 4.81E-06 0.61±0.78 1.47±1.21 Burned -0.41 0.22 124.95 -1.86 0.06 Warming -0.69 0.26 124.97 -2.62 9.95E-03 Warm & Dry -0.95 0.26 124.94 -3.65 3.89E-04 Nat. grass -0.08 0.01 125.23 -6.14 1.02E-08 Ind. fecundity Intercept 2.02 0.47 3.88 4.32 1.00E-02 0.57±0.76 0.26±0.51 Burned 0.53 0.09 120.01 5.62 1.29E-07 Warming -0.12 0.11 120.02 -1.08 0.28 Warm & Dry -0.18 0.11 120.02 -1.58 0.12 Nat. grass -0.01 0.01 120.06 -2.14 0.03 B. tectorum individual fecundity (seeds produced per stem) was not significantly affected by climate treatment (Table 2.4), but was negatively affected by native grass cover (p=0.03; Table 2.4), and responded positively to the burned treatment (p=1.29e-07; Fig. 2.3; Table 2.4). 60 Figure 2.3. Bromus tectorum individual fecundity for the burned and unburned treatments. Letters indicate significant differences (p<0.05). To satisfy model assumptions, data were natural log transformed. B. tectorum negatively affected P. spicata cover (Fig. 2.4), biomass, seed production per plot, and reproductive density (p=1.24e-03, p=0.01, p=0.02, and p=0.01, respectively; Table 2.5). P. spicata cover and biomass responded negatively to the warming and drying treatment (p=0.01 and p=5.38e-03; Table 2.5), however, there was no evidence of that either response variable was affected by the warming treatment or the burned treatment (Table 2.5). Both P. spicata seed production per plot and reproductive density were negatively affected by the warming treatment (p=0.04 and p=0.02, respectively; Table 2.5) and the warming and drying treatment (p=7.00e-05 and p=3.36e-05, respectively; Table 2.5) but not the burned treatment (Table 2.5). There was no evidence that P. spicata individual fecundity was affected by either climate treatment, the burned treatment, or B. tectorum cover. 61 Figure 2.4. Relationship between Pseudoroegneria spicata cover and Bromus tectorum cover by climate treatment. P. spicata was negatively affected by B. tectorum cover and both climate treatments. The analysis of the effects of our climate and burn treatments on the native grass community demonstrated an interesting trend (Fig. 2.5); native grass cover response was significantly lower in the warming and drying burned treatment compared with its unburned pair (p=1.49e-03) and the ambient unburned treatment (p=9.15e-03), but it did not differ from other treatments. The unburned warming and drying treatment did not differ from the ambient unburned, but was greater than the warmed unburned (p=0.05). 62 Table 2.5. Results of the best linear mixed-effects models assessing the Pseudoroegneria spicata response to the burn and climate treatments, and Bromus tectorum (B. tect.) cover. To satisfy model assumptions, the P. spicata response variables were natural log transformed, as were the B. tectorum cover data. P-values and degrees of freedom (df) were calculated using Satterthwaite approximation and values in bold indicate statistically significant differences (p<0.05). Response variables were assessed on a per plot scale except Ind. fecundity, which is mean individual fecundity. Seed prod. is seed production. Density is number of reproductive tillers. Nat. grass is native grass cover. Fixed effects Random effects Response Predictor Est. SE df t value p(>) Variance Year Residual Cover Intercept 2.55 0.28 4.81 9.20 3.11E-04 0.13±0.37 0.47±0.68 Burned -0.08 0.12 121.03 -0.66 0.51 Warming -0.28 0.15 121.03 -1.89 0.06 Warm & Dry -0.39 0.15 121.1 -2.61 0.01 B. tect. cover -0.21 0.06 122.96 -3.31 1.24E-03 Biomass Intercept 1.82 0.40 2.99 4.57 0.02 0.38±0.61 0.47±0.68 Burned -0.14 0.12 121.04 -1.14 0.26 Warming -0.26 0.15 121.04 -1.71 0.09 Warm & Dry -0.53 0.15 121.07 -3.56 5.38E-04 B. tect. cover -0.16 0.06 122.36 -2.54 0.01 Seed prod. Intercept 6.10 0.93 3.75 6.53 3.55E-03 1.82±1.35 3.85±1.96 Burned -0.31 0.35 121.06 -0.89 0.37 Warming -0.88 0.43 121.07 -2.05 0.04 Warm & Dry -1.76 0.43 121.11 -4.12 7.00E-05 B. tect. cover -0.43 0.18 122.85 -2.41 0.02 Density Intercept 3.30 0.54 5.37 6.10 1.34E-03 0.63±0.79 1.21±1.10 Burned -0.19 0.20 125.05 -0.99 0.33 Warming -0.57 0.24 125.05 -2.35 0.02 Warm & Dry -1.03 0.24 125.09 -4.30 3.36E-05 B. tect. cover -0.26 0.10 127.11 -2.60 0.01 Ind. fecundity Intercept 2.75 0.10 26.61 27.98 <2.0E-16 0.0054±0.07 0.11±0.33 Burned -0.01 0.07 107.66 -0.19 0.85 Warming -0.09 0.08 107.52 -1.17 0.24 Warm & Dry -0.08 0.06 107.37 -1.29 0.20 B. tect. cover -0.02 0.03 91.96 -0.67 0.5 63 Figure 2.5. Native grass cover within each of the six climate-burn treatments. Letters indicate statistically significant differences (p<0.05). To satisfy model assumptions, cover data were natural log transformed. Species richness was not affected by climate nor burn treatments, but there was a negative relationship between total and native richness values and B. tectorum cover (p=5.58e- 04 and p=5.58e-06, respectively; Fig.2.6). (a) Total (b) Native Figure 2.6. Relationship between Bromus tectorum cover and (a) total species richness and (b) native species richness. 64 Discussion To evaluate how B. tectorum and its response to fire would respond to global climate change, my field experiment experimentally increased the growing season temperatures and reduced growing season precipitation, and imposed a fire treatment. B. tectorum metrics responded negatively to native grass cover, increased temperatures, and the combination of increased temperatures and reduced precipitation, while B. tectorum individual fecundity responded positively to the spring burn. Native perennial grasses are important in shaping B. tectorum’s landscape position and community role (Brummer et al. 2016), therefore, this experiment also investigated how the native grass community responded to the experimentally increased temperatures and drier conditions, as well as fire. A dominant native perennial sagebrush-steppe grass species, P. spicata, responded negatively to increased temperatures, the combination of increased temperatures and reduced precipitation, as well as B. tectorum cover. While the burn treatment did not affect P. spicata, total native grass cover responded negatively to the spring prescribed burn when conditions were warmer and drier. One of the most documented and significant ecological effects of increasing temperatures associated with global climate change has been its impact on species range distributions; many species ranges have shifted up in elevation and poleward in latitude (McCarty 2001; Walther et al. 2002; Parmesan 2006; Lenoir and Svenning 2014). A meta-analysis of studies that experimentally manipulated climate factors, found that experimental warming stimulated photosynthesis and plant growth (Wu et al. 2011). Similarly, field experiments have demonstrated that B. tectorum has responded positively to experimentally increased temperatures (Zelikova et al. 2013; Compagnoni and Adler 2014a; Compagnoni and Adler 65 2014b; Blumenthal et al. 2016). My results were inconsistent with these results, as my populations of B. tectorum responded negatively to experimental warming. There were significant differences between my study and the other in situ studies (Blumenthal et al. 2016; Compagnoni and Adler 2014a; Compagnoni and Adler 2014b; Zelicova et al. 2013). Plant responses to experimental warming have been consistently positive when the warming has taken place in either cold locations or during cold times of year (Rustad et al. 2001; Hollister et al. 2005; Walker et al. 2006; Bloor et al. 2010). Consistent with this, the other B. tectorum field studies that experimentally increased temperatures warmed their plots for the entire year (Compagnoni and Adler 2014a; Compagnoni and Adler 2014b; Blumenthal et al. 2016). This resulted in increased fall germination and increased survival over the winter in experimentally warmed plots (Blumenthal et al. 2016) and alterations to the snowpack, timing of snowmelt and water availability (Compagnoni and Adler 2014a; Compagnoni and Adler 2014b), which have been shown to affect B. tectorum (Bates et al. 2006; Griffith and Loik 2010; Concilio et al. 2013; Compagnoni and Adler 2014a; Compagnoni and Adler 2014b; Prevéy and Seastedt 2015). My study contrasted with these studies because it was performed during the active growing season of B. tectorum (April-July) in this region. I did have a year-round warming treatment however, my over-winter OTCs were colonized by small mammals and their plots experienced significant herbivory and soil disturbance, making their data potentially unrepresentative and biased so it was not included in the analysis. Bromus tectorum sensitivity to soil moisture and precipitation patterns has been correlated with the winter precipitation of the in situ studies that have experimentally increased temperatures. Compagnoni and Adler (2014a, 2014b), which found positive effects of experimentally increased temperatures on B. tectorum fecundity and survival, resulting in 66 increased population growth rate, were conducted in locations that received considerably more precipitation over the winter compared with our site. The three sites of Compagnoni and Adler (2014a) averaged 85 mm, 128 mm, and 122 mm of total precipitation between January and March (wrcc.dri.edu/summary) (Compagnoni and Adler (2014b) was conducted at the mid-site location of Compagnoni and Adler (2014a)), whereas our site only received 47 mm and 58 mm in 2015 and 2016, respectively (Table 1). The B. tectorum response during these years was significantly negative, whereas, in 2014, when our site received a similar amount of January- March precipitation (128 mm) as the Compagnoni and Adler sites, B. tectorum responded neutrally to the warming treatment. This same trend has been found in other studies investigating B. tectorum’s response to experimental warming. Zelicova et al. (2013) found that B. tectorum responded positively to their experimental warming treatment when there was ample (110mm) precipitation between January and March, however, when there was considerably less January-March precipitation (18mm, 25mm, 45mm) B. tectorum responses were largely either negative or neutral. Bromus tectorum establishment, survival, growth and reproduction, in areas with limited fall/winter germination, has been found to be highly dependent on spring precipitation (Mack and Pyke 1983; Meyer et al. 2001; Bradford and Lauenroth 2006; Concilio et al. 2013; Zelikova et al. 2013). Due to limited winter precipitation and the cold temperatures of our site, B. tectorum fall germination and winter survival was minimal and its effective growing season was largely constrained to April-June. The condensed growing season made precipitation during this period vital for B. tectorum growth. Therefore, it wasn’t surprising that B. tectorum responded negatively when we limited precipitation during this critical growth period in our region. 67 Despite the demonstrated importance of climate, native perennial grass communities have been found to be even more important for B. tectorum growth (Brummer et al. 2016). Similarly, in situ manipulative climate studies have found that native perennial grass community abundance better explains B. tectorum abundance than either temperature or precipitation (Compagnoni and Adler 2014a; Prevéy and Seastedt 2015). Even after disturbance, which has been vital for invasion of the sagebrush-steppe (Mack 1981; Knapp 1996), B. tectorum response has been shown to be dramatically affected by competition from surrounding native vegetation; B. tectorum abundance increased significantly after a burn when competition with neighboring perennial grasses was minimal, while the burn had little effect when there were higher levels of competition (Chambers et al. 2007). Consistent with these studies, my findings demonstrated that it was the native grass cover that most affected B. tectorum abundance. The Effects of Climate and Bromus tectorum on the Native Community Negative plant responses to experimental warming have often been attributed to reduced soil moisture availability (De valpine and Harte 2001; Rustad et al. 2001; Bloor et al. 2010). In our study, however, there was no evidence that the warming treatment dried the soil. We surmise that our experimental warming caused heat stress and reduced the production of our native grass community. Plant species have different photosynthetic temperature optima that when exceeded can adversely affect production (Luo 2007). Consistent with the decreased production in response to experimental warming, others have shown the ambient conditions prior to and during the experiment influence the temperature response (Shaver et al. 2000). Negative responses to experimental warming have occurred when implemented in warmer ambient 68 conditions (Rusted et al. 2001; Bloor et al. 2010; Shaver et al. 2000). Therefore, we conclude that our OTC warming treatment was successful and during the warm spring and summer months I induced heat stress in the native plant community. Native grass production has been directly correlated with precipitation (Sims and Singh 1978). Equally, when reduced precipitation has either been observed or manipulated within the sagebrush/cool season perennial grasslands it has been associated with decreased perennial grass cover and abundance (Anderson and Inouye 2001; Heitschmidt et al. 2005). Further manipulative studies in the Colorado sagebrush-steppe, have illustrated that, when in conjunction with increased temperatures, experimentally reduced precipitation reduced both total and graminoid growth (Harte and Shaw 1995; Cherwin and Knapp 2012). Similarly, it has previously been shown that P. spicata is unable to extract water from extremely dry soils and responds negatively to drought conditions, thus it has a limited capacity to respond to the combination of reduced water and increased temperatures (Harris 1967; Cline et al. 1977; Fraser et al. 2009). Therefore, it was not surprising that our warming and drying treatment negatively affected the overall cover of the native perennial grass community and the cover and abundance of P. spicata. Temperature and water availability affect competition (Whisenant and Uresk 1989; Harte and Shaw 1995; Everard et al. 2010; Prevéy and Seastedt 2014) and climate changes have been associated with significant community shifts (Anderson and Inouye 2001; Zavaleta et al. 2003a; Luo 2007; Everard et al. 2010; Dalgleish et al. 2011; Gherardi and Sala 2015). Furthermore, it is generally thought that shifts in climate will favor invasive species (Dukes and Mooney 1999; Walther et al. 2002; Vilà et al. 2007). For example, a Colorado sagebrush-steppe study found that since 1970 the mean average temperature had risen by 1.3oC, which has 69 resulted in the reduction of the dominant grass and a concurrent increase in exotic forb density (Alward et al. 1999). Similarly, a 19th century drought is thought to have effected a community change from a native perennial grass system to one dominated by invasive annual grasses (Corbin and D’Antonio 2004; Suttle and Thomsen 2007; Everard et al. 2010). Bromus tectorum has similarly responded positively to experimentally increased temperatures (Blumenthal et al. 2016) and has a shallow diffuse root system, making it drought tolerant and more competitive with perennial sagebrush-steppe grasses, including Pseudoroegneria spicata, in dry conditions (Evans 1961; Hull 1963; Eissenstat and Caldwell 1988; Link et al. 1990). Therefore, I hypothesized that our climate treatments would interact with B. tectorum and enhance its competitiveness with the dominant perennial grass of my native community, P. spicata. While I found that P. spicata responded negatively to B. tectorum and both climate treatments, their effects were independent of one another and my findings were inconsistent with the literature and failed to support my hypothesis. The Effects of Fire on Bromus tectorum and the Native Grass Community B. tectorum provides those sagebrush systems it has invaded and dominates with a dense, highly flammable litter layer early in the fire season (Knapp 1996). This has affected the fire regimes of these areas, often by significantly reducing the fire return intervals (Whisenant 1990). Meanwhile, it has also been shown that B. tectorum recovers rapidly and is highly competitive with native grasses after fire (Melgoza et al. 1990; Whisenant 1990; D’Antonio et al. 1992; Reed-dustin et al. 2016). Consistent with this, B. tectorum fecundity responded positively to my burn treatment. In warmer and drier portions of the sagebrush biome, the positive response by B. tectorum to fire has created a positive feedback between the two (Chambers et 70 al. 2007; Taylor et al. 2014). However, there was no evidence that B. tectorum responded positively to my burn treatment when I increased the temperature and reduced the precipitation. However, B. tectorum cover after the prescribed burn within the warming and drying treatment lacked the ill effects of the warm and dry conditions, which the unburned B. tectorum demonstrated. Therefore, I posit that in warm and dry conditions the burn compensated for the negative effects of the climate treatment. Native perennial grasses typically decrease in abundance in the first year after fire (Bailey and Anderson 1978; Whisenant and Uresk 1989; West and Yorks 2002; Davies et al. 2007; Davies et al. 2009; Davies et al. 2012; Reed-dustin et al. 2016). Consistent with these findings the native grass community in my study responded negatively to the spring burn the same growing season. However, similar to other studies investigating factors affecting native perennial grass growth after fire (Antos et al. 1983; Redmann et al. 1993; Pylypec and Romo 2003; Davies et al. 2007; Prieto et al. 2009; Pratt et al. 2014), the post-fire production of our native perennial grasses demonstrated a sensitivity to climate conditions following the burn; if the warming and drying treatment was removed from the analysis, there was no evidence that the burn negatively affected the native grass community; the effects of the burn treatment within the warm and dry treatment were important. The initial negative response of native grasses to fire was consistent with several other studies. In a native grassland community dominated by Festuca, Stipa, and Agropyron species, post-fire production decreased with low soil moisture and increased with elevated soil moisture (Redmann et al. 1993; Pylypec and Romo 2003). In a western South Dakota upland grass community, dominated by Bouteloua gracilis, Stipa comata, Agropyron spp., Whisenant and Uresk (1989) demonstrated the effects of an interaction between spring burn and precipitation 71 on native grassland production; if there was ample water for growth after a burn, production was elevated compared to unburned production, whereas, if post-fire water was limited, production was depressed compared with unburned levels. In addition to these observational studies, that experimentally decreased precipitation and increased temperatures after a burn have shown reduced native perennial production and lower community resilience and post-fire recovery to a community structure consistent with pre-fire conditions (Prieto et al. 2009; Enright et al. 2014). The B. tectorum invasion of the sagebrush-steppe of western North America has demonstrated an interesting dynamic between native plant communities, disturbance, and climate. In a study that spanned across the sagebrush-steppe of five western states, Chambers et al. (2014b) demonstrated that those sagebrush-steppe sites with warmer and drier conditions were more susceptible to B. tectorum invasion after fire than those sites with cooler temperatures and more available soil moisture. Furthermore, B. tectorum distribution models utilizing temperature and precipitation parameters have predicted an expansion of B. tectorum and the area suitable for the B. tectorum-positive feedback fire cycle along its cold and wet range margins (Bradley 2009; Taylor et al. 2014). Several factors within our findings are consistent with the previous studies and provide support for the distribution models predicting expansion into cooler and moister regions of the sagebrush biome. First, B. tectorum fecundity responded positively to fire. Second, B. tectorum cover demonstrated a positive compensatory response to fire within the warming and drying treatment. Lastly, the native grass community responded negatively to fire under warmer and drier conditions. When taken together, these factors demonstrate decreased ecosystem resilience to disturbance by the native grass community, resulting in reduced resistance to B. 72 tectorum invasion, accompanied by the increased threat of the B. tectorum-positive feedback fire cycle being initiated in these areas due to increased fecundity following fire. This study only investigated the immediate effects of changes in climate parameters and fire and the effects of these factors that could continue years into the future; two to three years is a common time period for native communities to recover from a fire and drought conditions (Bailey and Anderson 1978; Perryman et al. 2002; Heitschmidt et al. 2005; Gucker and Bunting 2011; Davies et al. 2012), although the effects of some droughts can be seen for a much longer time period and can be responsible for considerable change within the plant communities (Suttle and Thomsen 2007). 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Ecol Evol 3:1374–87. doi: 10.1002/ece3.542 81 CHAPTER THREE THE EFFECTS OF INCREASED TEMPERATURE, ALTERED RESOURCE AVAILABILITY, AND ELEVATED ATMOSPHERIC CO2 ON THE COMPETITION BETWEEN BROMUS TECTORUM AND ESTABLISHED PSEUDOROEGNERIA SPICATA Introduction Atmospheric CO2 concentrations have increased at an unprecedented rate since the beginning of the industrial era (IPCC 2013). It is projected that the 2015-2016 atmospheric CO2 concentration growth rate will be highest on record and concentrations will surpass and remain above 400 ppm for the entire year (Betts et al. 2016). The increase in atmospheric CO2 has effected global climate change, altering all components of the earth’s climate, including surface temperatures and precipitation patterns (IPCC 2013). As plant growth is directly affected by temperature, precipitation (Woodward and Williams 1987) and atmospheric CO2 concentrations (Bazzaz 1990), these changes have consequences for global plant communities (Chapin et al. 2000; Shaver et al. 2000; Cramer et al. 2001; Walther et al. 2002; Parmesan and Yohe 2003; Chen et al. 2011). Plant responses to elevated CO2 concentrations vary between functional groups (Poorter and Navas 2003; Wang et al. 2012), life-history strategies (e.g. annual species) (Zangerl and Bazzaz 1984), and between plants with different growth rates (Poorter and Navas 2003). Elevated CO2 concentrations have been shown to magnify the effects of competition (Bazzaz 1990; Manea and Leishman 2011); in controlled settings increasing density and competition depressed plant responses to elevated CO2 concentrations (Bazzaz et al. 1992; Wayne et al. 1999). Plant community and competitive responses under elevated CO2 concentrations have 82 been shaped by resource availability, especially water availability (Bazzaz 1990; Morgan et al. 2004). For example, the results of seven year free-air CO2 enrichment (FACE) Wyoming study demonstrated that elevated CO2 had positive effects on the plant community aboveground biomass during dry years, however, this effect was reduced to negligible levels in those growing seasons with high soil moisture and precipitation (Mueller et al. 2016). Similarly, in their study, Polley et al. (2012) found that the competitive effect of tall-grass species on mid-grass species intensified when soil moisture content was increased in conjunction with elevated CO2 concentrations. Non-native invasive species are often fast growing, highly competitive species, which as a group are believed to benefit from elevated atmospheric CO2 concentrations (Dukes and Mooney 1999; Weltzin et al. 2003; Moore 2004; Ziska and George 2004; Sorte et al. 2013). Indeed, a meta-analysis of largely monoculture studies, found non-native invasive species to have greater responses to elevated CO2 concentrations than native species (Ziska and George 2004). A study evaluating competition between 14 native - non-native invasive species pairs under elevated CO2 concentrations, found non-native invasive species to be the better competitors and the strength of their competitive effects were heightened by elevated CO2 concentrations (Manea and Leishman 2011). Contrasting with these findings, studies focusing on the non-native invasive species Centaurea solistitialis and non-native Chenopodium album found that both responded positively to elevated CO2 concentrations when they were grown in monoculture, however, when grown in a community setting they failed to respond significantly (Dukes 2002; Taylor and Potvin 1997). Furthermore, the responses of invaded communities have been found to change through time and depend on water availability (Smith et al. 2014). The differing responses led Dukes (2000) to conclude that invasive species’ responses to elevated 83 CO2 concentrations were context based, being influenced by ecosystem attributes, especially water availability and the identity of the species involved. Reduced water availability has been shown to significantly reduce plant growth (Knapp and Smith 2001), favoring annual species over perennial species (Everard et al. 2010), and resulting in reduced abundance of dominant grass species and significant community shifts (Báez et al. 2013). Similarly, experimental warming has altered competitive dynamics between annual and perennial species (Alward et al. 1999) and has favored certain functional groups over others, resulting in significant changes in community and competitive dynamics (Harte and Shaw 1995). Increased temperatures and reduced water availability have been associated with reduced ecosystem resilience to community change and resistance to invasion by invasive species (Milchunas and Lauenroth 1995; Everard et al. 2010; Chambers et al. 2014; Prevéy and Seastedt 2014), and global climate change is thought to favor the spread of invasive species (Dukes and Mooney 1999; Weltzin et al. 2003; Thuiller et al. 2008; Hoeppner and Dukes 2012). An example of this is a 19th century drought which has been suggested as a significant factor in the replacement of a California perennial bunchgrass system with an invasive annual grass system (Corbin and D’Antonio 2004; Suttle and Thomsen 2007; Everard et al. 2010). Likewise, a mean annual temperature increase of 1.3 oC since 1970 in a Colorado sagebrush-steppe resulted in the reduction of the dominant native grass and a concurrent increase in non-native invasive species abundance (Alward et al. 1999). Similarly, since 1900 the Swiss Alps have seen a decrease in frost free days and an increase in non-native species richness (Walther et al. 2002). However, species distribution models have predicted both range expansions for invasive species due to climate change (Kriticos et al. 2003; Gritti et al. 2006; Bradley et al. 2009; Kleinbauer et 84 al. 2010; Sheppard 2013; Taylor et al. 2014) as well as range contractions (Peterson et al. 2008; Bradley 2009; Bourdôt et al. 2012; Gallagher et al. 2013; Bradley et al. 2016). The differing responses indicate the importance of the species being modelled and the importance of field experiments, which can provide insight into the factors affecting the spread of invasive plant species (Sheppard et al. 2014). Warming within western North America has been greater than projected by climate models (Polley et al. 2013); over the last 100 years the Pacific Northwest and the Intermountain West region has warmed by 0.5 °C - 1.5 °C and it is projected to warm another 2 °C - 5 °C by the end of the century (Chambers and Pellant 2008). One commonly accepted ecological effect that climate change will have on western ecosystems is that it will facilitate the spread of invasive plant species throughout the region (Chambers and Pellant 2008; Abatzoglou and Crystal 2011; Polley et al. 2013). Bromus tectorum, an invasive winter annual grass introduced in the 1880’s, was ubiquitous throughout western North America by the 1920’s (Knapp 1996; Mack 1981; Billings 1994). B. tectorum is naturalized throughout North America (Morrow and Stahlman 1984), however, it’s ecological/ecosystem dominance has historically been constrained to the Columbia, Bonneville, Lahontan, and Lower Snake river basins (Brummer et al. 2016). The mechanisms behind and the limitations of the B. tectorum invasion into western North America informs how it will respond to changes in global temperatures, precipitation patterns, and atmospheric CO2 concentrations. Disturbance to native perennial sagebrush-bunchgrass communities by both grazing and fire facilitated the invasion (Mack 1981; Knapp 1996; Chambers et al. 2007) and established native perennial bunchgrass communities, in addition to climate, have been shown to be the major limiting factors in its expansion (Brummer et al. 85 2016). Both fire and grazing have been associated with increasing soil nutrient availability (Hobbs and Huenneke 1992 and references therein; Blank et al. 2007) and B. tectorum responds positively to increased levels of soil nutrients, especially nitrogen (Meyer et al. 2001; Vasquez et al. 2008; He et al. 2011; Orloff et al. 2013). Thus, increased nutrient availability is thought to contribute to B. tectorum’s positive-feedback with fire and its invasion of western North America (D’Antonio et al. 1992; Meyer et al. 2001; Vasquez et al. 2008; He et al. 2011; Orloff et al. 2013). Observational and experimental studies have demonstrated B. tectorum invasion and growth to be associated with temperature and soil water relations (Bradford and Lauenroth 2006; Chambers et al. 2007; Compagnoni and Adler 2014a; Compagnoni and Adler 2014b; Prevéy and Seastedt 2015; Blumenthal et al. 2016; Brummer et al. 2016). However, the effects of these factors can be mediated by the presence and response of native grass communities to these same conditions. It has consistently been demonstrated that robust native perennial grass communities are the single most important factor limiting B. tectorum (Chambers et al. 2007; Compagnoni and Adler 2014a; Prevéy and Seastedt 2015; Brummer et al. 2016). It has been proposed that an intricate dynamic exists between climate, native perennial grass communities, disturbance, and ecosystem resistance to B. tectorum invasion: native community resilience to disturbance and resistance to B. tectorum decreases along a climate gradient, being high in cool moist systems and low in warm and dry systems (Chambers et al. 2007; Chambers et al. 2014a; Dodson and Root 2015). As stated above, elevated CO2 concentrations have been found to favor invasive species (Poorter 1993; Polley 1997; Weltzin et al. 2003; Ziska 2003; Ziska and George 2004; Manea and Leishman 2011), and because elevated CO2 concentrations increase water use efficiency (WUE) 86 and this is especially true for fast growing C3 species (Bazzaz 1990; Morgan et al. 2004). Single factor CO2-B. tectorum monoculture studies have demonstrated that B. tectorum responds positively to elevated CO2 concentrations (Smith et al. 1987; Poorter 1993; Ziska et al. 2005). However, when studied in a community setting, B. tectorum has responded neutrally to elevated CO2 concentrations (Blumenthal et al. 2016). While B. tectorum growth and invasive success has been associated with temperature, soil moisture, available nutrients, and competition with native perennial grasses, to date no research has studied the combination of all these factors in a controlled setting. Therefore, one of my goals was to determine if the competitive dynamic between B. tectorum and an established native perennial bunchgrass (Psueodoroegnaria spicata) was responsive to altered temperature, available soil moisture, and nutrient levels in a controlled chamber environment. Given the general paucity of literature on how B. tectorum responds to elevated CO2 concentrations (3 monoculture studies and 1 community study), my second goal was to determine if competition between B. tectorum and established P. spicata individuals was responsive to elevated CO2 concentrations and if decreased water availability impacted this response. I hypothesized that in both experiments increasing proportional density of established P. spicata individuals would have the greatest limiting effect on the establishing B. tectorum individuals. My second hypothesis was that decreased water availability and increased temperature would favor B. tectorum, and that increased nutrient availability would further heighten B. tectorum competitiveness. Finally, I hypothesized B. tectorum would respond positively to elevated CO2 concentrations and the combination of elevated CO2 and decreased soil water availability. 87 Methods Experimental Design Experiment 1: Decreased Water, Increased Temperature and Increased Nutrient Availability. I used a replacement series design, with five different density combinations of B. tectorum and established P. spicata individuals, and two level temperature, water, and nutrient treatments, giving a total of 80 pots per trial. The experiment was performed in two growth chambers at the Plant Growth Center, Montana State University (MSU), Bozeman, MT. The experiment was replicated twice (April—June and July—September, 2014). The ambient treatments were designed to represent the mean growing season temperature, precipitation and day length, as experienced in our area in June; the temperature in the low chamber was set at 23.3 °C for the daylight period (14 hours of 100 micromoles of PAR) and 6 °C for the night (10 hours) period. The elevated temperature chamber was two degrees higher for both the day and night periods (25.3 °C and 8 °C, respectively). To control for chamber effect, the chamber temperatures were switched every two weeks and the plants were moved. The water treatment was designed to simulate ambient June precipitation and a reduction of this amount by 50%; the watered treatment (water (+)) was watered 3 times a week with 300 ml, while the decreased water treatment (water (-)) was watered 3 times a week with 150 ml. Due to poor growth during the first trial, after 5 weeks the water treatments were upped from 300 to 500 ml, and from 150 to 250 ml. The same watering patter and total amount was used for the second trial. For the elevated nutrient treatment (NPK+), 173 mg, 75.50 mg, 17.30 mg of slow release N, P, K, was added, (respectively). 88 The total target density for each pot was 50 plants pot-1 (988 plants m-2); the five density combinations of B. tectorum and P. spicata were 50, 37-13, 25-25, 13-37, 50 plants pot-1 (988, 731, 494, and 256, B. tectorum plants m-2, respectively). These densities were based on B. tectorum densities found locally at the Montana State University (MSU) Red Bluff Research Station near Norris, MT. Seeds were sown into circular pots (25.4 cm diameter) filled with equal parts loam soil, washed concrete, and sphagnum peat moss. The soil was aerated steam pasteurized at 70 °C for 60 min. Seeds were randomly sown within a grid with 2 cm spacing and, to account for edge effect, no seeds were sown closer than 4 cm to the sides of the pot. To simulate B. tectorum invasion of an established P. spicata community, the P. spicata plants were planted one month before the B. tectorum in a greenhouse with temperatures of 21.1 °C during the day (14 hrs) and 18.3 °C during the night (10 hrs). The P. spicata seeds were the ‘Goldar’ variety obtained from the USDA Natural Resources Conservation Service, Aberdeen Plant Materials Center (Aberdeen, ID). The B. tectorum seeds were hand collected at the MSU Red Bluff Research Station. Seeds of both species were assessed for viability prior to sowing, and seeding rates were adjusted accordingly. After the B. tectorum was sown, the pots were watered evenly to facilitate germination and moved immediately to the temperature controlled chambers; the water and nutrient treatments were then implemented. To prevent damage to the seedlings, plant measurements were not taken for first two weeks after B. tectorum had been sown. Sub-samples individual plant height in each pot were taken weekly from the second week until the termination of the experiment, when biomass was clipped, dried, and weighed. The mean of the sub-samples for each species of each pot were obtained and analyzed after each trial. The trials lasted 70 and 67 days after B. tectorum was sown. To combat an aphid infestation during the second trial the pots 89 were treated with pesticides. Experiment 2: Increased Atmospheric CO2 Concentration and Decreased Water Availability at an Elevated Temperature. This experiment also utilized a replacement series design with five different density combinations of B. tectorum and established P. spicata. In addition, this experiment utilized two atmospheric CO2 concentrations, ambient (400 ppm) and elevated (800 ppm), as well as a two-level water treatment. The experiment was repeated twice (January-April 2015 and May-August 2015) and there were 6 replicates per trial of the density (5-level), CO2 (2-level), and water (2-level) combinations for a total 120 pots per trial. The growth chambers in which the experiment took place had the temperature set to 25.3 °C and 8 °C, day and night, respectively. The seeds were sown into square pots with sides of 11 cm. The total target density for each pot was 12 plants pot-1 (1000 plants m-2); the five density combinations of B. tectorum-P. spicata were 12, 9-3, 6-6, 3-9, 12 plants pot-1 (1000, 750, 500, and 250, B. tectorum plants m-2, respectively). Germination, seeding rates and techniques, as well the soil and seeds used, were the same as experiment one. The water treatment was again designed to simulate local ambient June precipitation conditions and a reduction of this amount by 50%; the watered treatment (water (+)) was watered 3 times a week with 100 ml, while the decreased water treatment (water (-)) was watered 3 times a week with 50 ml. These water levels provided the same amount of water per unit area as the previous experiment. Again, beginning with the second week, plant height was measured weekly until the termination of the experiment. The experiments were run for 69 and 54 days after B. tectorum was sown in the first and second trial, respectively. Upon the termination of the experiment, final height was taken and aboveground biomass was clipped, dried, and weighed. 90 Statistical Analysis The effects of the treatments on response variables of height, biomass, and relative yield of both B. tectorum and P. spicata were conducted using linear mixed-effects models for both experiments. Relative yield was calculated using the proportion of the species in the mixture (P) and the mean plant biomass of the species in mixture (Amix) and monoculture (Amon): P(Amix/Amon) (Cousens and Neill 1993). All models were fit using the treatments as fixed effects and trial as a random effect. Treatments for the first study included: water (ambient, decreased), temperature (ambient, increased), nutrient (ambient, increased) and the proportion of P. spicata within each pot. Treatments for the second experiment were: CO2 concentration (ambient, increased), water (ambient, decreased) and the proportion of P. spicata within each pot. To satisfy model assumptions of normality and heteroscedasticity, response data were transformed when appropriate; exact transformations used are shown below. Initial models included interactions and were reduced to the most parsimonious model with experimental treatments still included. The analyses were conducted using the statistical program R (version 3.2.2, R Development Core Team, 2015). Linear mixed-effects models were constructed using the lme4 package (Bates et al. 2011) and the lmerTest package (Kuznetsova et al. 2014). Significant relationships between the treatment effects and the response variables were calculated at the p<0.05 level from F-statistics based on Satterthwaite’s approximations of degrees of freedom for linear mixed-effects models (Kuznetsova et al. 2014). 91 Results Effects of Competition, Decreased Water, Elevated Temperature, and Increased Nutrient Availability. Bromus tectorum biomass responded negatively to the interspecific competition with the established P. spicata individuals (p<2.0e-16) and decreased water availability (p=4.71e-03), while it responded positively when nutrients were added (p=6.88e-05) (Table 3.1). Similarly, B. tectorum height (Fig. 3.1) responded negatively to interspecific competition (p<2.0e-16) and decreased water availability (p=9.39e-09), and positively to increased nutrient availability (p=5.01e-04) (Table 3.1). (a) Water treatment (b) Nutrient treatment Figure 3.1. Effects of Pseudoroegneria spicata competition, (a) water availability, and (b) nutrient availability on Bromus tectorum height. Temperature did not affect B. tectorum height. NPK represents nutrient availability. 92 P. spicata biomass was negatively affected by decreased water availability (p=6.14e-09), increased temperature (p=3.82e-05), and there was an interaction between decreased water and increased temperature (p=1.00e-03). P. spicata biomass was positively affected by both added nutrients and increased P. spicata proportion (p=3.50e-07 and p<2.0e-16, respectively; Table 1). P. spicata height (Fig 3.2) responded negatively to decreased water availability (p=5.33e-15), increased temperature (p=9.94e-04), and when P. spicata proportion was increased (p=1.22e-04) (Table 3.1). (a) Water treatment (b) Temperature treatment Figure 3.2. Effects of intraspecific competition, (a) water availability, and (b) temperature on Pseudoroegneria spicata height. Nutrient availability did not affect P. spicata height. 93 Table 3.1. Results of mixed-effects models conducted to assess the effects of competition, water availability, temperature, and nutrient availability on Bromus tectorum and Pseudoroegneria spicata biomass and growth. To satisfy model assumptions data were transformed, as necessary, and the exact form is shown below. P-values and degrees of freedom (df) were calculated using the Satterthwaite approximation and values in bold indicate statistically significant differences (p<0.05). Water (-) represents decreased water availability, Temp (+) represents elevated temperature, and NPK (+) represents increased nutrient availability. Fixed effects Random effects Response Predictor Est. SE df t value p(>) Variance Trial Residual B. tectorum •Biomass Intercept 1.72 0.15 1.31 11.39 0.03 0.038±0.20 0.12±0.34 Water (-) -0.18 0.06 111.00 -2.88 4.71E-03 Temp (+) -0.04 0.06 111.03 -0.67 0.51 NPK (+) 0.26 0.06 111.01 4.14 6.88E-05 P. spicata -10.79 0.48 111.05 -22.49 <2.0E-16 Water: P. spicata 1.96 0.68 111.00 2.86 5.07E-03 Height Intercept 19.85 1.54 1.20 12.86 0.03 4.18±2.05 8.22±2.87 Water (-) -3.50 0.56 100.00 -6.27 9.39E-09 Temp (+) -0.01 0.56 100.02 -0.02 0.98 NPK (+) 2.01 0.56 100.01 3.60 5.01E-04 P. spicata -40.24 2.96 100.44 -13.58 <2.0E-16 P. spicata Biomass Intercept 8.31 1.59 1.06 5.23 0.11 4.85±2.20 2.36±1.54 Water (-) -2.50 0.4 111.00 -6.30 6.14E-09 Temp (+) -1.22 0.28 111.01 -4.29 3.82E-05 NPK (+) 2.16 0.40 111.01 5.42 3.50E-07 P. spicata 15.39 1.55 111.03 9.92 <2.0E-16 Water (-): NPK(+) -1.92 0.57 111.00 -3.38 1.00E-03 Height Intercept 34.84 0.63 102.00 55.15 <2.0E-16 0.00±0.00 11.37±3.37 Water (-) -5.99 0.65 102.00 -9.18 5.33E-15 Temp (+) -2.21 0.65 102.00 -3.39 9.94E-04 NPK (+) 1.10 0.65 102.00 1.69 0.09 P. spicata -13.48 3.37 102.00 -4.00 1.22E-04 •lmer(sqrt(B. tectorum biomass) ̴temp + NPK + water x P. spicata + (1|trial)). Bromus tectorum’s relative yield was positively affected by both the elevated 94 temperature (p=0.04; Table 3.2; Fig. 3.3a) and the increased nutrient treatment (p=2.55E-03; Table 3.2; Fig. 3.3b). While it was not affected by reduced water availability (p=0.17; Table 3.2) and was negatively affected by proportion of P. spicata within the pot (p<2E-16; Table 3.2; Fig. 3.3). P. spicata relative yield was negatively affected by the reduced water treatment and the elevated temperature treatment (p=3.84E-04 and p=0.03, respectively; Fig 3.3a) those these variables interacted significantly, with P. spicata responding negatively to the combination of ambient temperature and decreased water at higher densities of pot P. spicata (Table 3.2). It was not affected by the elevated nutrient treatment and was positively affected by the proportion of P. spicata within the pot (p=0.70 and p<2E-16; Table 3.2; Fig 3.3). (a) Temperature-water treatment (b) Nutrient treatment Figure 3.3. Effects of competition with Pseudoroegneria spicata, (a) the combination of elevated temperature and decreased water (Temp (+):Water (-)), and (b) increased nutrient availability (NPK+), on Bromus tectorum and P. spicata relative yield. 95 Table 3.2. Results of mixed-effects models conducted to assess the effects of competition, water availability, temperature, and nutrient availability on Bromus tectorum and Pseudoroegneria spicata relative yield. To satisfy model assumptions data were transformed, as necessary, and the exact form is shown below. P-values and degrees of freedom (df) were calculated using the Satterthwaite approximation and values in bold indicate statistically significant differences (p<0.05). Water (-) represents decreased water availability, Temp (+) represents elevated temperature, and NPK (+) represents increased nutrient availability. Fixed effects Random effects Response Predictor Est. SE df t value p(>) Variance Trial Residual B. tectorum •Relative yield Intercept -3.30 0.32 1.30 -10.20 0.03 0.17±0.42 0.40±0.63 Water (-) -0.19 0.14 79.00 1.40 0.17 Temp (+) 0.29 0.14 79.04 2.11 0.04 NPK (+) 0.43 0.14 79.02 3.12 2.55E-03 P. spicata -13.90 0.65 79.36 -21.50 <2.0E-16 P. spicata2 -1.29 0.64 79.16 -2.02 0.05 P. spicata ○Relative yield Intercept -0.22 0.09 79.00 -2.41 0.02 0.00±0.00 0.15±0.39 Water (-) -0.43 0.12 79.00 -3.71 3.84E-04 Temp (+) -0.26 0.12 79.00 -2.22 0.03 NPK (+) -0.03 0.08 79.00 -0.38 0.70 P. spicata 9.73 0.39 79.00 25.15 <2.0E-16 P. spicata2 -0.82 0.39 79.00 -2.11 0.04 Water (-): Temp (+) 0.36 0.17 79.00 2.13 0.04 •lmer(sqrt(B. tectorum rel. yield) ̴water + temp + NPK + poly(P. spicata, 2) + (1|trial)). ○lmer(logit(P.spicata rel.yield) ̴water x temp + NPK + poly(P. spicata, 2) + (1|trial)). There was strong evidence that the proportion of P. spicata (p<2.0e-16), decreased water availability (p=9.01e-03), and increased nutrient availability (p=3.37e-03) affected the proportion of the total biomass of the two species (Table 3.3; Fig 3.4). There was minimal evidence that increased temperature had a significant effect on the proportion of the total biomass of the two species (p=0.07; Table 3.3). 96 Figure 3.4. Effects of competition with Pseudoroegneria spicata, and the combination of decreased water and increased nutrient availability (water (-):NPK (+), on Bromus tectorum and P. spicata proportion of total pot biomass. Table 3.3. Results of linear mixed-effects models conducted to assess the effects of competition, water availability, temperature, and nutrient availability on Bromus tectorum and Pseudoroegneria spicata proportion of total pot biomass (Prop. biomass). To satisfy model assumptions data were transformed, as necessary, and the exact form is shown below. P-values and degrees of freedom (df) were calculated using the Satterthwaite approximation and values in bold indicate statistically significant differences (p<0.05). Water (-) represents decreased water availability, Temp (+) represents elevated temperature, and NPK (+) represents increased nutrient availability. The values in the table are for B. tectorum; the P. spicata values are the inverse. Fixed effects Random effects Response Predictor Est. SE df t value p(>) Variance Trial Residual •Prop. biomass Intercept -2.18 0.60 1.09 -3.61 0.15 0.68±0.83 0.49±0.70 Water (-) 0.40 0.15 80.00 2.68 9.01E-03 Temp (+) 2.28 0.15 80.01 1.83 0.07 NPK (+) 0.46 0.15 80.01 3.02 3.37E-03 P. spicata -10.23 0.72 80.12 -14.30 <2.0E-16 •lmer(logit(prop. biomass) w̴ater + temp. + NPK + P. spicata +(1|trial)). 97 Effects of Competition, Increased Atmospheric CO2, and Decreased Water Availability Under an Elevated Temperature. Bromus tectorum biomass and height responded negatively to both interspecific competition with P. spicata (p<2.0e-16 and p=2.88e-12, respectively) and elevated CO2 treatment (p=0.01, p= 0.05, respectively) although there was an interaction between these variables (Fig 3.3; Table 3.3). In addition, decreased water negatively affected B. tectorum biomass (p=0.01), but height was unaffected (p=0.09) (Table 3.4). In contrast, P. spicata biomass and height responded positively to elevated CO2 (p<2.0e-16, and p<2.0e-16, respectively; Table 3.4). Similarly, as the proportion of P. spicata increased, its biomass responded positively (5.02e- 12), although its height was unaffected (p=0.26) (Table 3.4). When the water availability was decreased both P. spicata biomass and height responded negatively (p=7.28e-08, p=1.43e-04, respectively; Table 3.4). Again there was an interaction between P. spicata and CO2, due to switch in target species response at higher proportions of P. spicata (Table 3.4). 98 Table 3.4. Results of mixed-effects models conducted to assess the effects of elevated atmospheric CO2 concentration and decreased water on Bromus tectorum and Pseudoroegneria spicata biomass and growth. To satisfy model assumptions, data were transformed and the exact form is shown below. P-values and degrees of freedom (df) were calculated using the Satterthwaite approximation and values in bold indicate statistically significant differences (p<0.05). Water represents decreased water availability, CO2 (+) represents elevated atmospheric CO2 concentration. Fixed effects Random effects Response Predictor Est. SE df t-value p(>) Variance Trial Residual B. tectorum •Biomass Intercept -0.38 0.51 1.06 -0.74 0.59 0.49±0.70 0.69±0.83 CO2(+) -0.31 0.12 178.00 -2.53 0.01 Water(-) -0.30 0.12 178.00 -2.47 0.01 P. spicata -15.55 1.30 178.07 -11.97 <2.0E-16 P. spicata2 3.74 0.12 178.07 2.60 0.01 P. spicata:CO2(+) -7.33 1.73 178.03 -4.24 3.62E-05 *Height Intercept 2.60 0.26 1.03 9.88 0.06 0.13±0.37 0.10±0.31 CO2(+) -0.09 0.05 177.00 -1.97 0.05 Water(-) -0.08 0.05 177.00 -1.73 0.09 P. spicata -3.55 0.47 177.02 -7.51 2.88E-12 P. spicata:CO2(+) -1.96 0.64 177.01 -3.07 2.47E-03 P. spicata ○Biomass Intercept 1.38 0.38 1.01 3.60 0.17 0.29±0.54 0.08±0.28 CO2(+) 0.57 0.04 178.00 13.48 <2.0E-16 Water(-) -0.23 0.04 178.00 -5.62 7.28E-08 P. spicata 2.88 0.39 178.00 7.41 5.02E-12 P. spicata2 -1.87 0.39 178.01 -4.85 2.65E-06 P. spicata:CO2(+) -2.32 0.58 178.00 -3.98 1.01E-04 P. spicata2:CO2(+) 1.59 0.59 178.00 2.70 7.61E-03 Height Intercept 24.47 2.17 1.07 11.27 5.00E-02 8.95±3.00 14.54±3.81 CO2(+) 10.52 0.56 181.00 18.71 <2.0E-16 Water(-) -2.17 0.56 181.00 -3.89 1.43E-04 P. spicata 5.86 5.22 181.03 1.12 2.60E-01 P. spicata:CO2(+) -28.16 7.72 181.02 -3.65 3.49E-04 •lmer (ln(B. tectorum biomass) w̴ater + CO2 * poly(P. spicata, 2) + (1|trial)). *lmer (ln(B. tectorum height) w̴ater + CO2 * P. spicata +(1|trial)). ○lmer (sqrt(P. spicata biomass) w̴ater + CO2 * poly(P. spicata, 2) + (1|trial)). Bromus tectorum relative yield responded negatively to the increase in CO2 and was unaffected when the water level was decreased (p=2.22E-16 and p=0.34, respectively; Table 3.5). It demonstrated no interaction between the two treatments; within both water treatments, B. tectorum relative yield was negatively affected by the CO2 treatment (Fig 3.5). 99 Pseudoroegneria spicata relative yield responded negatively to the decreased water treatment and demonstrated no response to the elevated CO2 treatment (p= 1.11E-03 and p=0.23, respectively; Table 3.5). There was evidence supporting a significant interaction between the two treatments (p=8.16 E-04; Table 3.5). Further investigation of this interaction demonstrated that within the watered treatment (water (+)) elevated CO2 had no effect (p= 0.29; Fig 3.5a), however, when water levels were reduced (water (-)), the elevated CO2 concentration had a significantly positive effect on P. spicata (p= 1.81E-08; Fig 3.5b). Table 3.5. Results of mixed-effects models conducted to assess the effects of elevated atmospheric CO2 concentration and decreased water on Bromus tectorum and Pseudoroegneria spicata relative yield. To satisfy model assumptions, data were transformed and the exact form is shown below. P-values and degrees of freedom (df) were calculated using the Satterthwaite approximation and values in bold indicate statistically significant differences (p<0.05). Water represents decreased water availability, CO2 (+) represents elevated atmospheric CO2 concentration. Fixed effects Random effects Response Predictor Est. SE df t value p(>) Variance Trial Residual B. tectorum •Relative yield Intercept -2.98 0.96 1.03 -3.10 0.19 1.80 ±1.34 0.86 ±0.92 CO2(+) -1.52 0.16 134.00 -9.39 2.22E-16 Water(-) 0.15 0.16 134.00 0.96 0.34 P. spicata -12.65 0.96 134.02 -13.20 <2.0E-16 P. spicata ○Relative yield Intercept 0.57 0.02 8.62 33.80 1.82E-10 6.15E-05 ±7.84E-03 8.67E-03 ±0.09 CO2(+) 0.03 0.02 132.00 1.20 0.23 Water(-) -0.07 0.02 132.01 -3.33 1.11E-03 P. spicata 1.90 0.10 132.00 19.72 <2.0E-16 P. spicata2 -0.51 0.09 132.45 -5.50 1.93E-07 CO2(+):Water(-) 0.10 0.03 132.02 3.43 8.16E-04 •lmer(logit(B. tectorum rel. yield) ̴CO2 + water + P. spicata + (1|trial). ○lmer(sqrt(P. spicata rel yield) ̴CO2 x water + poly(P. spicata, 2) + (1|trial). 100 (a) Water (+) (b) Water (-) Figure 3.5. Effects of competition with Pseudoroegneria spicata, and atmospheric CO2 concentration on Bromus tectorum and P. spicata relative yield within the (a) watered treatment (Water (+)) and the (b) dry treatment (Water (-)). Bromus tectorum relative contribution to total pot biomass was negatively affected by P. spicata pot proportion as well as increased atmospheric CO2, while P. spicata proportion of total pot biomass was positively affected by both factors (Table 3.6; Figure 3.6). There was an interaction between the polynomial form of P. spicata and CO2, due to the pattern at higher pot proportions of P. spicata (Table 3.6). Species proportion of the total pot biomass, was not affected by decreased water (Table 3.6). Figure 3.6. Effects of competition with Pseudoroegneria spicata, and atmospheric CO2 concentration on B. tectorum and P. spicata proportion of total pot biomass. 101 Table 3.6. Results of the mixed-effects model conducted to assess the effects of competition (P. spicata), elevated atmospheric CO2 concentration, and water availability on Bromus tectorum and Pseudoroegneria spicata proportion of total pot biomass (prop. biomass). To satisfy model assumptions, data were transformed and the exact form is shown below. P-values and degrees of freedom (df) were calculated using the Satterthwaite approximation and values in bold indicate statistically significant differences (p<0.05). Water represents decreased water availability, CO2 (+) represents elevated atmospheric CO2 concentration. The values in the table are the B. tectorum values; the P. spicata values are the inverse. Fixed effects Random effects Response Predictor Est. SE df t value p(>) Variance Trial Residual •Prop. biomass Intercept -1.38 0.32 1.45 -4.34 0.08 0.15±1.15 0.39±1.07 CO2(+) -1.66 0.19 131.00 -8.82 6.22E-15 Water(-) 0.15 0.18 131.00 0.80 0.43 P. spicata -11.29 1.74 131.61 -6.50 1.54E-09 P. spicata2 5.69 1.84 131.06 3.10 2.39E-03 P. spicata: CO2(+) 3.81 2.42 131.33 1.57 0.12 P. spicata2: CO2(+) -6.89 2.38 131.12 -2.89 4.46E-03 •lmer(logit(prop. biomass) w̴ater + temp. + NPK + poly(P. spicata, 2) +(1|trial)). Interestingly, the analysis of B. tectorum in monoculture, yielded evidence supporting significantly positive effects of elevated CO2 on its biomass (p=4.00e-15; Fig 3.7a) and height (p=1.95e-07; Fig 3.7b), a direct contrast to the results when grown in competition. Pseudoroegneria spicata monoculture biomass and height also responded positively to elevated CO2 concentrations (p=2.87e-08 and p=1.14e-10; Figs 3.7c and 3.7d, respectively). 102 (a) Height (b) Biomass Bromus tectorum (c) Height (d) Biomass Pseudoroegneria spicata Figure 3.7. Monoculture B. tectorum and P. spicata biomass and height responses to elevated atmospheric CO2 concentration. Discussion Established native perennial grasses, including P. spicata, are highly competitive with B. tectorum (Orloff et al. 2013; Prevéy and Seastedt 2015) and have been found to be the most significant biotic factor limiting B. tectorum distribution in the sagebrush biome (Brummer et al. 2016). Consistent with the literature and as expected, our study found that B. tectorum was 103 more limited by the interspecific competition with established P. spicata individuals than the other tested variables. The first goal of the study was to determine if this competitive dynamic would respond to decreased water availability, increased temperature, or the combination of the two treatments. Bromus tectorum responded negatively to decreased water availability but showed no response when the temperature was increased, while P. spicata demonstrated negative responses to both, as well as the combination of the two treatments. While previous studies have demonstrated that both species respond negatively to decreased water availability (Cline et al. 1977; Chambers et al. 2007; Fraser et al. 2009; Prevéy and Seastedt 2015), B. tectorum’s shallow diffuse root structure lends it a greater ability to extract water from extremely dry soil, thereby increasing its competitiveness with its native perennial neighbors when water is limiting (Hull 1963; Harris 1967; Eissenstat and Caldwell 1988; Link et al. 1990). This dry site competitiveness was illustrated in the specific relative yield responses and the specific proportion of the total biomass results we observed. These results, in addition to P. spicata’s individual negative responses to both treatments, supported our hypothesis that increased temperature and decreased water would alter the competitive dynamics in favor of the invasive B. tectorum. Studies have found that non-native invasive plant species out compete native plants in high resource environments, while native plants are more successful in low resource areas (Davis et al. 2000; Daehler 2003). Others have found that increased resources have no effect on invasive species competitive effects (Maron and Marler 2008), or that invasive species have enough phenotypic plasticity that they can do equally well in both high and low resource areas (Funk and Vitousek 2007; Funk 2008). Bromus tectorum is an invasive species that has been 104 associated with areas where elevated nutrients are present (Norton et al. 2004), it has been found to be a very strong competitor for available soil nutrients, especially nitrogen (Booth et al. 2003). It has demonstrated increased growth and competitive success associated with increased nutrient levels, and its successful invasion and positive feedback with fire have been tied with increased availability of soil nutrients (D’Antonio et al. 1992; Meyer et al. 2001; Vasquez et al. 2008; He et al. 2011; Orloff et al. 2013). Consistent with the literature, the addition of nutrients had positive effects on all B. tectorum response variables (biomass, height, and relative yield). Similarly, the addition of nutrients combined with the decreased water treatment increased the B. tectorum proportion of the total pot biomass. There was generally a lack of response by P. spicata to the nutrient addition. Consistent with our hypothesis, this suggests that, while interspecific competition with the larger P. spicata still limited B. tectorum, added nutrients did increase B. tectorum’s competitiveness and this effect was exaggerated under drier conditions. Elevated CO2 concentrations have consistently been associated with increased growth, especially for C3 species (Bazzaz 1990; Poorter 1993; Ackerly and Bazzaz 1995; Polley 1997; Poorter and Navas 2003), and B. tectorum (Smith et al. 1987; Poorter 1993; Ziska et al. 2005). A mechanism through which atmospheric CO2 concentrations facilitate plant growth is by increasing plant water use efficiency (Bazzaz 1990), therefore soil water relations will mediate and affect how plant communities respond to increasing CO2 (Bazzaz et al. 1992; Morgan et al. 2004; Smith et al. 2014). The importance of this interaction for annual grasses was illustrated by a long term free-air carbon dioxide enrichment (FACE) plant community study (Smith et al. 2014). Smith et al. (2014) demonstrated that responses by the annual grass species, Bromus rubens, to elevated CO2 were highly contingent on soil moisture. Thus, the third goal of our study was to assess the impact of elevated atmospheric CO2 concentrations on the competition 105 between P. spicata and B. tectorum and to determine if these effects were responsive to a 50% reduction in water availability. When B. tectorum and P. spicata were grown in monoculture, elevated CO2 had positive effects, however, when grown in competition, the established P. spicata responded positively, while the younger B. tectorum plants responded negatively. The effective changes in specific proportion of total pot biomass and relative yield clearly demonstrated that elevated CO2 provided the established P. spicata with an even greater competitive edge. The decreased water treatment had no effect on B. tectorum’s response to elevated CO2 and it was clear that when soil moisture was reduced, which has been shown to favor B. tectorum, P. spicata greatly benefitted from the elevated CO2 concentration. The monoculture findings were consistent with the results of the other B. tectorum-CO2 studies that have been done in a controlled setting (Smith et al. 1987; Ziska et al. 2005). However, when grown in competition, my findings were inconsistent with the previously mentioned monoculture studies as well as the findings of Blumenthal et al. (2016), who studied B. tectorum’s response to elevated CO2 in a native Wyoming mixed prairie community. Blumenthal et al. (2016) found that neither B. tectorum nor the native plant community was responsive to elevated CO2 concentrations. Competitive responses to elevated CO2 have been shown to be significantly different than monoculture responses and are often responsive to other factors in addition to CO2 concentration (Ackerly and Bazzaz 1995; Shaw et al. 2002; Smith et al. 2014). There are two likely mechanisms underlying B. tectorum’s response to elevated CO2. B. tectorum is strongly limited by interspecific competition with established P. spicata (Orloff et al. 2013) and the positive effects of the elevated CO2 on the growth of the established P. spicata individuals exaggerated their size and competitive advantage; the effects of increased interspecific 106 competition overwhelmed any positive effects the increased CO2 had on B. tectorum. The second mechanism potentially underlying B. tectorum’s response to elevated CO2 while in competition, could be the indirect effects of the CO2 on available N. Elevated CO2 commonly reduces N availability (Luo et al. 2004), which could moderate the positive effects that elevated CO2 might hold for invasive species (Sorte et al. 2013; Blumenthal et al. 2016), especially those that are responsive to heightened nutrient availability, such as B. tectorum. B. tectorum is highly competitive with native perennial grasses when moisture is limiting (Hull 1963; Harris 1967; Eissenstat and Caldwell 1988). However, one of the physiological effects of elevated atmospheric CO2 concentrations is increased water use efficiency (WUE) for C3 species (Bazzaz 1990). Although we did not quantify the WUE for either species’, P. spicata’s relative yield response to the interaction between elevated CO2 and water availability demonstrated that the CO2 treatment had a positive effect on its WUE. This increase in WUE translated into heightened competitiveness with B. tectorum when water availability was limited. Such a positive response by one of B. tectorum’ perennial bunchgrass competitors could limit invasion success of B. tectorum under future climate conditions and CO2 concentrations. Conclusions Our results confirmed that larger, established P. spicata, individuals exert a strong suppressive effect on invading B. tectorum individuals. However, this strong suppressive effect was mediated by temperature, soil water relations, nutrient availability, and atmospheric CO2 concentrations; decreased water availability, increased temperature, and increased nutrient availability increased B. tectorum competitiveness. However, elevated atmospheric CO2 concentrations magnified the suppressive effect and reduced B. tectorum’s competitiveness, especially in dry conditions. 107 Climate conditions of the American West are projected to become warmer, with more variable precipitation (Bradley, 2009; Chambers and Pellant, 2008; Mote and Salathé, 2010), thus it has been postulated B. tectorums’ range will expand into elevations and latitudes where the previous climate was unsuitable (Bradley 2009; Bradley et al., 2015; Taylor et al., 2014). Furthermore, elevated atmospheric CO2 concentrations are expected to favor invasive species (Dukes and Mooney, 1999; Thuiller et al., 2007; Ziska and George, 2004; Weltzin et al., 2003), including B. tectorum (Smith et al., 1987; Ziska et al., 2005), which could further facilitate B. tectorum range expansion within the sagebrush-steppe biome. The results of my first experiment, with elevated temperatures, reduced water and a nutrient treatment, were consistent with the projections that B. tectorum will expand north in latitude into those areas where its invasiveness is limited by climate. However, the significant suppressive effect of the established P. spicata on B. tectorum, especially when the atmospheric CO2 concentration was increased, contrast with the projections. 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To understand and appropriately moderate these effects it is imperative to increase our knowledge of how particular species and communities will shift under future climate conditions. The Bromus tectorum invasion of western North America’s sagebrush-steppe has caused significant ecological as well as agricultural damage (Mack 1981; Knapp 1996), however, the damage has been restricted by robust native communities and cold and wet climates (Chambers et al. 2014a; Taylor et al. 2014; Brummer et al. 2016). One of the most damaging effects of the B. tectorum invasion has been the alteration of the fire regime in those areas where it has come to dominate (D’Antonio et al. 1992; Knapp 1996; Brooks et al. 2004). In these areas, B. tectorum has promoted fire, which, in turn, has facilitated its expansion and domination (D’Antonio et al. 1992; Brooks et al. 2004). Similar to B. tectorum’s ecological dominance, its response to fire has been limited by climate (Taylor et al. 2014). As the climate of the North American Intermountain West is projected to become warmer and drier in the summer (Chambers and Pellant 2008; Mote and Salathé 2010; Polley et al. 2013), and native plant community resilience to 117 disturbance and resistance to B. tectorum decrease under these conditions (Chambers et al. 2014a; Chambers et al. 2014b; Dodson and Root 2015), models have predicted that B. tectorum and its response to fire will expand up in elevation and latitude (Bradley 2009; Taylor et al. 2014). The goal of this thesis was to experimentally test the prediction that as the climate becomes warmer and drier in the late spring/summer, B. tectorum’s ecological dominance and its response to fire will expand north into latitudes of the sagebrush-steppe where they have previously been limited by a cold and wet climate. In order to address this goal, the objectives of this thesis were to assess under field conditions (1) B. tectorum’s response to increased temperatures, and the combination of increased temperatures and decreased growing season precipitation, in a cold and moist southwestern Montana sagebrush-steppe plant community; (2) B. tectorum’s response to fire in the same experimental treatments and at the same location; (3) If fire, experimental warming, and decreased growing season precipitation increase B. tectorum’s impact on the dominant perennial bunchgrass Pseudoroegneria spicata; and under more controlled settingsassess (4) how the competitive dynamic between B. tectorum and an established native perennial grass, P. spicata, would be affected by increased temperature and altered resource availability (water and nutrients), and (5) the same competitive dynamic at an elevated temperature, with altered atmospheric CO2 concentrations and soil moisture availability. Established native perennial grasses, including P. spicata, are highly competitive with B. tectorum (Orloff et al. 2013; Prevéy and Seastedt 2015) and have been found to be the most significant biotic factor limiting B. tectorum distribution in the sagebrush biome (Brummer et al. 2016). Even B. tectorum’s response to disturbance, which has been vital for the B. tectorum 118 invasion of the sagebrush-steppe (Mack 1981; Knapp 1996), has been restricted by competition with perennial grasses (Chambers et al. 2007). Consistent with these studies, both of our studies demonstrated that competition with native grasses proved to be the most strongly suppressive factor for B. tectorum growth and abundance. Due to the importance of the relationship between B. tectorum and the native perennial grass community, understanding how their relationship is affected by climate variables and resource availability is central to understanding how sagebrush-steppe plant communities will respond to climate change. Elevated temperatures affect competition, increasing the competitiveness of non-native invasive plant species over native species (Wu et al. 2016). Plus, elevated temperatures have been associated with shifts in plant communities (Harte and Shaw 1995; De valpine and Harte 2001; Zavaleta et al. 2003a; Zavaleta et al. 2003b; Walker et al. 2006; Luo 2007), including shifts from native dominated communities to invasive dominated ones (Alward et al. 1999; Walther et al. 2002; Thuiller et al. 2008). Furthermore, B. tectorum consistently grows faster and produces more biomass at higher temperatures than does its common competitors (Aguirre and Johnson 1991; Nasri and Doescher 1995; Hardegree et al. 2010) and, at lower latitudes, B. tectorum has responded positively to experimentally increased temperatures in the field (Zelikova et al. 2013; Compagnoni and Adler 2014a; Compagnoni and Adler 2014b; Blumenthal et al. 2016). With this understanding we hypothesized that B. tectorum would respond positively and would be more competitive with P. spicata at experimentally increased temperatures. In the field setting as well as the controlled setting, P. spicata responded negatively to experimental warming. Yet, instead of taking advantage of this response, B. tectorum also responded negatively to the experimental warming in the field and demonstrated no response to elevated 119 temperatures in the controlled setting. One significant difference between our field study and those where positive responses to experimental warming were observed, was the length and duration of the warming: they warmed year-round, while we only warmed for the active part of the growing season (April-July) because our year-round warmed plots experienced considerable soil disturbance and herbivory from small mammals. Native and non-native invasive species have different soil water use patterns (Cavaleri and Sack 2010) and reductions in soil moisture availability have resulted in the expansion of non-native invasive species (Milchunas and Lauenroth 1995; Dukes and Mooney 1999) and community shifts from native perennial grass communities to non-native invasive annual grass communities (Corbin and D’Antonio 2004; Suttle and Thomsen 2007; Everard et al. 2010). Similarly, B. tectorum’s shallow diffuse root structure lends it a greater ability to extract water from extremely dry soil, making it highly competitive with native perennial grass competitors in dry conditions (Hull 1963; Harris 1967; Eissenstat and Caldwell 1988; Link et al. 1990). However, the results of my water reduction treatment alone did not demonstrate this heightened competitiveness; in a controlled setting however, when it was coupled with other factors (temperature and nutrient availability) B. tectorum did experience an increase in competitiveness against established perennial neighbors. The combination of decreased water and increased temperature elicited different results in our field and controlled setting experiments. Bromus tectorum distribution and competitiveness with native perennial grasses has been associated with warm and dry climates (Chambers et al. 2007; Chambers et al. 2014b; Dodson and Root 2015). Consistent with this and my hypothesis, in my controlled setting study we found increased temperature combined with decreased water availability enhanced B. tectorum competitiveness. This contrasted with the 120 results of my field study, in which B. tectorum was negatively affected by the combination of warmer and drier conditions. While inconsistent with my controlled setting experiment, my field result wasn’t entirely inconsistent with the prior climate change-B. tectorum studies. Bromus tectorum sensitivity to soil moisture and precipitation patterns has been correlated with the winter precipitation of the in situ studies that have experimentally increased temperatures. Compagnoni and Adler studies (2014a, 2014b), which demonstrated positive B. tectorum responses to experimentally increased temperatures, were conducted in locations that received considerably more precipitation over the winter, when compared with my site; the three sites of Compagnoni and Adler (2014a) averaged 85 mm, 128 mm, and 122 mm of total precipitation between January and March (wrcc.dri.edu/summary) whereas my site only received 47 mm and 58 mm in 2015 and 2016, respectively. The B. tectorum response during these years was significantly negative, whereas, in 2014, when my site received significantly more January-March precipitation (128 mm), B. tectorum responded neutrally to the warming treatment. This same trend of B. tectorum responding positively under moist winter conditions versus negatively or neutrally under drier winters (<60 mm) has been observed in one other study (Zelikova et al. 2013). These data suggest that differences in winter precipitation amounts when combined with warming elicit different responses: warmer and drier winter conditions generate negative to neutral response by B. tectorum, whereas warmer and wetter winter conditions provide neutral to positive responses by B. tectorum. Robust native plant communities are the most important biotic factor limiting B. tectorum invasion (Brummer et al. 2016). Therefore, how these communities respond to fire is fundamental to how resistant they are to invasion. Increased temperature and decreased precipitation affect how native communities respond to fire (Whisenant and Uresk 1989; 121 Redmann et al. 1993; Pylypec and Romo 2003; Prieto et al. 2009; Pratt et al. 2014) and a recent model has proposed a connection between native community responses to fire, climate, and B. tectorum invasion (Chambers et al. 2014a). Chambers et al (2014) posited that ecosystem resilience to disturbance and ecosystem resistance to B. tectorum decreases along a temperature and precipitation gradient; native plant community recovery after disturbance (e.g. fire) is more complete in cool and wet ecosystems, and is reduced in warm and dry areas. Therefore, cool and wet ecosystems have higher ecosystem resistance to invasion than do warm and dry systems. These model predictions were exemplified in a large-scale manipulative study, which experimentally burned and mowed sites within the sagebrush-steppe of five western states (Idaho, Utah, Nevada, Oregon, Washington) (Chambers et al. 2014b). Chambers et al. (2014b) found that B. tectorum responded differently according to the dominant vegetation of the site, which was associated with a climate regime. Bromus tectorum responded most positively to their disturbance treatments in the warm and dry Wyoming big sagebrush, most negatively in the cool and moist mountain sagebrush, and demonstrated moderate negative responses at the warm and moist Wyoming big sagebrush sites. Similarly, an Oregon study investigating plant community responses to a stand replacing fire found that 11 years after the fire native perennial species dominated the cool and moist areas, while invasive annual grasses, including B. tectorum, dominated the areas with a higher heat load index and climate moisture deficit (Dodson and Root 2015). Findings from both of my studies provided evidence supporting the model and these other studies. As previously mentioned, within my controlled setting study, B. tectorum was more competitive in warmer and drier conditions, it responded positively to increased nutrient 122 treatment, and its competitiveness in drier conditions, likewise, increased with added nutrients. The evidence in support of the Chambers et al. (2014a) model from the field study was threefold: (1) B. tectorum fecundity responded positively to fire, (2) B. tectorum cover demonstrated a positive compensatory response to fire within the warming and drying treatment, (3) the native grass community responded negatively to fire under warmer and drier conditions. Increased global surface temperatures and altered precipitation regimes that accompany global climate change do not occur in a vacuum; their ecological consequences will always be mediated by elevated atmospheric CO2 concentrations (Zavaleta et al. 2003b; Norby and Luo 2004; Luo et al. 2008). Furthermore, it has been shown that temperature and altered resource availability, especially water availability, interact to alter plant responses to CO2 (Bazzaz 1990; Polley 1997; Shaw et al. 2002; Zavaleta et al. 2003b; Morgan et al. 2004; Smith et al. 2014; Zelikova et al. 2015). Elevated atmospheric CO2 concentrations are expected to favor invasive species (Dukes and Mooney, 1999; Thuiller et al. 2007; Ziska and George 2004; Weltzin et al. 2003), including B. tectorum (Smith et al. 1987; Poorter 1993; Ziska et al. 2005). The previous B. tectorum-CO2 studies have largely been monoculture studies (Smith et al. 1987; Poorter 1993; Ziska et al. 2005), similarly, in my study when B. tectorum was grown in monoculture it responded positively to increased CO2. However, mixed-species or community responses to elevated CO2 differ from monoculture responses (Ackerly and Bazzaz 1995; Shaw et al. 2002; Smith et al. 2014), and in my study when B. tectorum was grown in competition with established P. spicata individuals, the significant suppressive effect of the larger P. spicata was magnified, and its biomass and growth suffered. The positive effect that elevated CO2 had on P. spicata’s competitiveness was heightened in dry conditions, where B. tectorum normally has the 123 competitive edge. Our findings also differed from the only study to have researched B. tectorum’s response to elevated CO2 while in a community setting. Blumenthal et al. (2016) studied B. tectorum’s response to elevated CO2 in a native Wyoming mixed prairie community and found that neither B. tectorum nor the native plant community was responsive to elevated CO2 concentrations. Blumenthal et al. (2016) attributed the lack of response by B. tectorum to their elevated CO2 concentrations to the indirect effects of elevated CO2 on soil N availability, which is important for species, such as B. tectorum, that are highly sensitive to N availability. I did not measure or alter nutrient availability in my CO2 experiment so it is difficult to attribute B. tectorum’s response to elevated CO2 to this mechanism. I attribute the difference between the results of my study and the other greenhouse and field studies, that have tested the effects of elevated CO2 on B. tectorum, to the initial size difference between the P. spicata individuals and the B. tectorum individuals. The growth of the older P. spicata individuals was stimulated by the elevated CO2, which resulted in increased aboveground and belowground competition with the establishing B. tectorum individuals. Conclusions and Future Work Climate conditions of the American West are projected to become warmer, with more variable precipitation (Bradley 2009; Chambers and Pellant 2008; Mote and Salathé 2010); thus, it has been proposed that B. tectorum’s range will expand into elevations and latitudes where the climate was previously unsuitable (Bradley 2009; Bradley et al 2016; Taylor et al. 2014). The goal of this thesis was to provide insight into how B. tectorum and the native grass community will respond to fire and projected climate changes in a cold and moist region of the sagebrush- steppe biome. We accomplished this by altering the temperature, soil moisture availability, and resource availability in both field and controlled settings. 124 The results of chapters 2 and 3 provided limited evidence in support of the conclusion that B. tectorum will expand its northern border of ecological dominance and that global climate change will promote the initiation of the B. tectorum-positive feedback fire cycle in these areas. However, the negative responses by B. tectorum to warming, warming and drying, its limited response to fire, as well as the substantial suppressive effect of the perennial grasses, especially in elevated CO2 concentrations, suggest that it is unlikely that B. tectorum dominance and its positive feedback with fire will expand north in latitude, without a considerable disturbance that lowers the abundance of native perennial grasses in the community. There were some disparities between my field and controlled setting experiments, and between my findings and those of other studies. The other field studies have been performed in sagebrush and grassland ecosystems to the south of mine, they employed their warming throughout the winter, and during their experiments they had a greater proportion of precipitation during the winter. At my more northerly site, either because soil moisture is frozen and unavailable or because of a lack of winter precipitation, there is rarely sufficient available soil moisture for fall and winter germination and growth, therefore the growing season is condensed to the higher precipitation months of April-June. Further experimental work to evaluate if this timing and variability of precipitation is driving the differences we observed may be warranted. Should funding be obtained an experiment evaluating the effect of the timing and variability of precipitation within and between seasons, along with warming and elevated CO2 treatments, would aid our understanding of the potential expansion of B. tectorum and the response of the native grass community in these colder and wetter regions. 125 References Ackerly DD, Bazzaz FA (1995) Plant growth and reproduction along CO2 gradients : non-linear responses and implications for community change. Glob Chang Biol 1:199–207. 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Relationship between the Simpson’s diversity index of the Red Bluff climate treatments and Bromus tectorum cover. To satisfy model assumptions B. tectorum cover was natural log transformed. Data were analyzed using a linear mixed-effects model. No statistically significant differences were found between climate treatments. B. tectorum cover negatively affected plot diversity (0.03). Figure A. 8. The burn treatment effect on the percent cover of litter in the Red Bluff plots (linear mixed-effects model, p=0.04). 157 Figure A. 9. The burn treatment effect on the percent cover of bareground in the Red Bluff plots (linear mixed-effects model, p=9.11E-09). Figure A. 10. Species rank abundance in the ambient climate treatment of the Red Bluff plots. 158 Figure A. 11. Species rank abundance in the warming climate treatment of the Red Bluff plots. 159 Figure A. 12. Species rank abundance in the warming and drying climate treatment of the Red Bluff plots.