THE EFFECTS OF PLANT INVASION AND DROUGHT ON PLANT SOIL INTERACTIONS By CATHERINE FAHEY A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 2018
2018 Catherine Fahey
To Iris and James, the next generation of environmentalists
ACKNOWLEDGMENTS There are many people that I need to thank for suppo rting me on this journey. Firstly, my advisor, Luke Flory, has been critical in my development as a scientist and has provided me with amazing opportunities to expand my scientific horizons. I am extremely grateful for his guidance, for always being availa ble and for pushing me to be better but also understanding His enthusiasm, positivity, and critical eye made this work better at every stage. I am also thankful to my committee members Christine Angelini, D oria Gordon, and Tim Martin for their advice and feedback throughout this work. I am grateful for the financial support of Pedro Antunes and Kari Dunfield as well as the dedication of their time and expertise to improve the methodology for Chapter 3. That p iece would not have been possible without their support. Akihiro Koyama provided critical assistance with the bioinformatics in Chapter 3. I am also extremely grateful to H einke Jger for giving me the opportunity to work in one of the most interesting ec osystems I have ever known, the Galapagos. Although, that research is not included in this dissertation, her infectious enthusiasm motivated me through the last push to finish my dissertation. In addition to my mentors during my PhD, I want to thank all my previous mentors that Kitajima, who first got me hooked on tropical ecology. Her support was instrumental in developing the skills necessary to complete my PhD Mar tijn Slot also gave me a huge amount of support in my first years of graduate school. Rob York and Teresa Pawloska were incredible mentors during my undergraduate career and I greatly appreciate their taking the time to work with me and supporting me in co mpleting and publishing my honors thesis I am grateful to the School of Natural Resources and Environment for funding and support and particularly to Tom Frasier and Karen Bray. I also need to acknowledge the fantastic
work of numerous undergraduate assis tants that helped with the many laborious tasks involved in field ecology. I am also grateful to the entire Flory Lab crew I feel very fortunate to have had such supportive entertaining, and caring lab mates. Julia Maki, Deah Lieurance, James Estrada, Ch ris Wilson, Jules NeSmith, Drew Hiatt, Chrissy Alba, Emma Byerly, Taylor Clark, Whalen Dillon, Amy Kendig, and Tabitha Petri provided assistance advice, and good company throughout the years. I am especially thankful to Julia and Taylor for being such ama zing friends in addition to lab mates. Furthermore, a ll my Gainesville friends have made this an enjoyable time and helped me through the more difficult times. I am particularly indebted to my amazing friends Anya Brown, Caroline Storer, Lianne Allen Jacob son, Sarah Graves, and Verity Salmon who have provided incredible support in (partly) overcoming imposter syndrome and providing examples of strong female scientists that I aspire to be like. Finally, I would not have made it to this point without my lovin g family. M y parents, Lois and Tim, siblings, Beth, Bob, and Becky, siblings in law Ben and Beth, and niece and nephew Iris and James provided unwavering love, patience, and support throughout this journey F und ing for this work was provided by the Univers ity of Florida, Institute of Food and Agricultural Sciences (UF/IFAS); the Florida Forest Service, Florida Department of Agriculture and Consumer Services (Contract#21942) ; and the USDA/NIFA McIntire Stennis program (FLA AGR 005180), N ational Science Foun dation Division of Environmental Biology 1546638 and Natural Sciences and Engineering Research Council of Canada.
6 TABLE OF CONTENTS page ACKNOWLEDGMENTS ................................ ................................ ................................ ............... 4 LIST OF TABLES ................................ ................................ ................................ ........................... 8 LIST OF FIGURES ................................ ................................ ................................ ......................... 9 ABSTRACT ................................ ................................ ................................ ................................ ... 11 CHAPTER 1 INTRODUCTION ................................ ................................ ................................ .................. 13 Framew ork: Interacting Stressors ................................ ................................ ........................... 13 Microbial Communities ................................ ................................ ................................ .......... 18 Plant Soil Interactions ................................ ................................ ................................ ............ 22 2 GRASS INVASION AND DROUGHT INTERACT TO ALTER THE DIVERSITY AND STRUCTURE OF NATIVE PLANT COMMUNITIES ................................ .............. 27 Background ................................ ................................ ................................ ............................. 27 Methods ................................ ................................ ................................ ................................ .. 29 Experimental Design ................................ ................................ ................................ ....... 29 Vegetation Surveys ................................ ................................ ................................ .......... 31 Abiotic Measurements ................................ ................................ ................................ ..... 31 Statistical Analysis ................................ ................................ ................................ .......... 32 Results ................................ ................................ ................................ ................................ ..... 33 Per cent Cover ................................ ................................ ................................ .................. 33 Functional Groups ................................ ................................ ................................ ........... 34 Species Richness and Turnover ................................ ................................ ....................... 34 Diversity ................................ ................................ ................................ .......................... 36 Community Composition ................................ ................................ ................................ 36 Soil Moisture and PAR ................................ ................................ ................................ .... 37 Discussion ................................ ................................ ................................ ............................... 38 3 PLANT INVASION AND DROUGHT INTERACTIVELY STRUCTURE SOIL MICROBIAL COMMUNITIES ................................ ................................ ............................. 45 Background ................................ ................................ ................................ ............................. 45 Methods ................................ ................................ ................................ ................................ .. 48 Study System ................................ ................................ ................................ ................... 48 Experimental Design ................................ ................................ ................................ ....... 49 DNA Extraction, PCR, and Illumina Sequencing ................................ ........................... 50 Sequence Data Processing ................................ ................................ ............................... 51 Statistical Anal yses ................................ ................................ ................................ .......... 52
7 Results ................................ ................................ ................................ ................................ ..... 53 Bacterial Community ................................ ................................ ................................ ....... 53 Bacterial Taxa ................................ ................................ ................................ .................. 53 Fungal Community ................................ ................................ ................................ .......... 55 Fungal Taxa ................................ ................................ ................................ ..................... 55 Fungal Guilds ................................ ................................ ................................ .................. 56 Drivers of Microbial Community Structure ................................ ................................ .... 56 Discussion ................................ ................................ ................................ ............................... 57 4 COMPETITION AND SOIL LEGACIES ALTER THE ROLE OF SOIL MICROBES IN PLANT INVASION ................................ ................................ ................................ .......... 70 Background ................................ ................................ ................................ ............................. 70 Methods ................................ ................................ ................................ ................................ .. 73 Field Experiment ................................ ................................ ................................ ............. 73 Greenhouse Experiment ................................ ................................ ................................ .. 74 Statistical Analysis ................................ ................................ ................................ .......... 75 Results ................................ ................................ ................................ ................................ ..... 77 Discussion ................................ ................................ ................................ ............................... 79 5 CONCLUSIONS ................................ ................................ ................................ .................... 90 APPENDIX A CHAPTER 2 SUPPLEMENTAL INFORMATION ................................ .............................. 94 Experimental Plots ................................ ................................ ................................ .................. 94 Abiotic Data ................................ ................................ ................................ ............................ 95 Plant Functional Groups ................................ ................................ ................................ ......... 97 Tables ................................ ................................ ................................ ................................ ...... 98 B CHAPTER 3 SUPPLEMENTAL INFORMATION ................................ ............................ 100 Figures ................................ ................................ ................................ ................................ .. 100 Tables ................................ ................................ ................................ ................................ .... 101 C CHAPTER 4 SUPPLEMENTAL INFORMATION ................................ ............................ 103 Figures ................................ ................................ ................................ ................................ .. 103 Tables ................................ ................................ ................................ ................................ .... 104 LITERATURE CITED ................................ ................................ ................................ ................ 107 BIOGRAPHICAL SKETCH ................................ ................................ ................................ ....... 127
8 LIST OF TABLES Table page A 1 List of herbaceous species planted into the plots. ................................ ............................. 98 A 2 Results of the PERMANOVA of the main and interactive effects of invasion, drought, and date on plant community composition. ................................ ......................... 98 A 3 Results of the PERMANOVA of the main and interactive effects of invasion and drought on plant community composition by year. ................................ ........................... 99 B 1 Results of mixed effects model on bacterial richness, Shannon divers ity index, and evenness. ................................ ................................ ................................ .......................... 101 B 2 Results of mixed effects model on fungal richness, Shannon diversity index, and evenness. ................................ ................................ ................................ .......................... 101 B 3 Results of mixed effects model on arbuscular mycorrhizal fungal richness, Shannon diversity index, and evenness. ................................ ................................ ......................... 101 B 4 Results of PERMANOVA on weighted UNIFRAC distance of the bacterial c ommunity. ................................ ................................ ................................ ...................... 101 B 5 Results of PERMANOVA on unweighted UNIFRAC distance of the bacterial community. ................................ ................................ ................................ ...................... 101 B 6 Results of PERMANOVA o n Bray Curtis dissimilarity matrix of the fungal community. ................................ ................................ ................................ ...................... 102 B 7 Results of PERMANOVA on Bray Curtis dissimilarity matrix of the arbuscular mycorrhizal fungal community. ................................ ................................ ....................... 102 C 1 Results of mixed effects of soil legacy of invasion and drought, live or sterile inoculum, alone or in competition, and their interactions on total cogongrass biomass. ................................ ................................ ................................ ........................... 104 C 2 Results of mixed effects of soil legacy of invasion and drought, live or sterile inoculum, alone or in competition, and their interactions on total pine biomass. ........... 105 C 3 Results of mixed effects of soil legacy of invasion and drought, live or sterile inoculum, alone or in competition, and their interactions on total wiregrass biomass. ... 106 C 4 Li st of seedlings that died by treatment and were excluded from analysis. ................... 106
9 LIST OF FIGURES Figure page 1 1 Model diagram of possible interactions between two stressors (A and B). ...................... 26 2 1 Seasonal changes in vegetation cover over 4 years. ................................ ......................... 41 2 2 Plant community dynamics in re sponse to the individual and interactive effects of invasion and drought. ................................ ................................ ................................ ......... 42 2 3 Seasonal changes in diversity metrics in response to the individual and interactive effects of invasion and dro ught ................................ ................................ .......................... 43 2 4 Non metric multidimensional scaling (NMDS) ordination plots of summer plant community composition 2014 2017. ................................ ................................ ............... 44 3 1 Shannon diversity index of OTUs. ................................ ................................ .................... 62 3 2 Nonmetric multidimensional scaling ordination ................................ ............................... 63 3 3 Relative abundance of bacterial p hyla by plot in each invasion x drought treatment. ..... 64 3 4 Relative abundance of fungal phyla by plot in each invasion x drought treatment. ......... 65 3 5 Relative abundance of the most abundant bacterial phyla (>1% mean relative abundance) in response to invasion and drought treatments ................................ ............. 66 3 6 Relative abundance of the most abu ndant fungal families (>1% mean relative abundance) in response to invasion and drought treatments ................................ ............. 67 3 7 Relative abundance of fungi by trophic mode in response to invasion and drought treatm ents ................................ ................................ ................................ ........................... 68 3 8 Relative abundance of selected fungal guilds ................................ ................................ ... 69 4 1 Conceptual diagram of possible belowground interactions between an invader and native species. ................................ ................................ ................................ .................... 85 4 2 Relative difference in total biomass between live and sterile soil inoculum in soil with a history of native species or invasion for three species either alone or in competition ................................ ................................ ................................ ........................ 86 4 3 Relative competition intensity for each species under live or sterile soil inoculum with a history of invasion or no invasion ................................ ................................ ........... 88 4 4 Relative difference in root:shoot ratio between live and sterile soil inoculum treatments. ................................ ................................ ................................ .......................... 87
10 4 5 Effect of soil legacy of invasion and drought and live or sterile soil inoculum on total biomass production per plot in the competition treatment ................................ ........ 89 5 1 Synthesis diagram of the combined results of the experiments in this dissertation. ......... 93 A 1 Experimental plots at the Bivens Arm Research Site, Gainesville, Florida in May 2016, showing factorial combination of treatments. ................................ .......................... 94 A 2 Soil moisture and precipitation over the duration of the experiment ................................ 95 A 3 Percent soil moisture by depth ................................ ................................ .......................... 96 A 4 Percent avai lability of photosynthetically active radiation at ground level and 0.5 m height above the soil surface ................................ ................................ .............................. 96 A 5 Effects of invasion and drought on plant functional groups ................................ ............. 97 B 1 Gravimetric soil moisture in ambient and drought plots either invaded by Imperata cylindrica or uninvaded. ................................ ................................ ................................ .. 100 B 2 Root biomass in invaded and uninvaded plots with ambient precipitation or drought at the 5 15 cm depth ................................ ................................ ................................ ......... 100 C 1 Change in total biomass of cogongrass in live or sterile soil with soil legacies of invasion x drought grown alone or in competition with pine and wiregrass. .................. 103 C 2 Change in total biomass of longleaf pine in live or sterile soil with soil legacies of invasion x drought grown alone or in competition with wiregrass and cogongrass. ....... 103 C 3 Change in total biomass of wiregrass in live or sterile soil with soil legacies of invasion x drought grown alone or in competition with pine and cogong rass. ............... 104
11 Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy THE EFFECTS OF PLANT INVASI ON AND DROUGHT ON PLANT SOIL INTERACTIONS By Catherine Fahey August 2018 Chair: S. Luke Flory Major: Interdisciplinary Ecology Plant soil interactions are major drivers of plant community dynamics and are likely to be altered by anthropogenic global cha nge with consequences for ecosystem structure and function. I nteractions among global change factors have the potential to exacerbate ecological effects, but these interactions are notoriously difficult to predict. Plant invasion s are accelerating worldwid e with consequences for biodiversity, nutrient cycling, and disturbance regimes. Furthermore, p lant invaders will experience shifts in abiotic conditions associated with climate change such as increased frequency and severity of drought. Here, I present re search on the responses of plant and soil communities to interacting stressors and assess the potential consequences for ecosystem restoration. First, I assessed the response of longleaf pine forest plant communities to experimental invasion by Imperata cy lindrica and experimental drought imposed with rainout shelters over four years. I found that invasion caused severe declines in diversity and shifts in composition of the native plant community, while drought had moderate effects on diversity and shifted the dominant functional groups. S oil moisture under drought conditions with the invader was higher than without the invader and in combination the impacts of invasion and drought were lower than expected, indicating an ameliorating effect. Additionally I evaluated the effects of invasion and drought and the consequent shifts in plant community on the soil microbial
12 communities. On the whole, drought was a stronger driver of bacterial communities than invasion, wh ereas fungal communities were interactively affected by the treatments Functional groups of importance for plant communities including plant pathogens, mycorrhizal fungi, and nitrifiers were affected by both invasion and drought Finally, I assessed the impacts of these shifts in soil microbial co mmunities in response to invasion and drought on growth and competition of the dominant plant species in this ecosystem ( longleaf pine and wiregrass ) Interestingly, the effect of soil microbes on plant growth varied with competitive context. Additionally, soil legacy of invasion decreased the growth of wiregrass but not pine. Colle ctively, my research provides an evaluation of the role of plant invasion and drough t on interactions in the plant soil system and addresses the implications for native ecosystem s under future biotic and abiotic conditions.
13 CHAPTER 1 INTRODUCTION Framework: Interacting Stressors Many biotic and abiotic stressors can influence plant communities and the literature o n this topic is vast. Stress can be defined in different ways and t he chosen definition can influence the conclusions drawn from multiple stressor research One common definition is a condition that reduces performance or fitness of a species below optimal levels (Lichtenthaler 1996, Folt et al. 1999, Vinebrooke et al. 2004) This definition is useful when considering a single species but is difficult to apply at the community level because species have different optimal conditions (Thompson et al. 2018) F or the current discussion stress will include stimuli that cause a negative response at the level of organization assessed (i e. species, population, or community level). Abiotic stressors for plants include levels of environmental resources (e.g. water, light, nutrients) and conditions (e.g. temperature, soil pH), physical stressors associat ed with disturbance (e.g. fire, wind), and exposure to toxins (e.g. heavy metals, pollutants). Abiotic stressors can influence plant species composition, diversity, structure, and function in a community and influence s oil ecosystem function in turn (Thuiller et al. 2005, Kardol et al. 2010) For abiotic stressors that are resource based, plant communities are expected to be most strongly affected by reductions in the most limiting resource (van der Pl oeg et al. 1999) The limiting resource will not necessarily be the same for all species within a community; therefore, different abiotic stressors can differentially affect species performance, and in some cases allow species to coexist if they are most limited by different resources or have access to different resource pools (Chesson 2000, Nippert and Knapp 2007) Fluctuating resource levels and
14 occasional stress from di sturbance can also promote increased species diversity, as in the storage effect and intermediate disturbance hypothesis (Chesson 2000, Knapp et al. 2002) For plants, biotic stressors primarily incl ude herbivore or pathogen effects and plant plant competitive interactions. Biotic stressors such as specialist pathogens can influence species diversity by creating density dependence in plant populations and allowing for species coexistence (Chesson 2000, Bever et al. 2015) Herbivory can increase plant diversity in communities where undefended plant species associa te with highly defende d species. T he unpalatable species reduce herbivory on palatable species allowing them to persist (Rebollo et al. 2002, Rousset and Lepart 2002, Callaway et al. 2005) On the other hand, a generalist herbivore or pathogen ca n have the opposite effect on species coexistence, as the species with the greatest tolerance to herbivory/pathogen s will dominate. Finally, competition with non nat ive invasive plants can act as biotic stress, especially if the invader has enhanced compet itive ability compared to natives or novel weapons (Callaway and Ridenour 2004, Graebner et al. 2012) Vegetation responses to stress take place on different time scales and levels of ecological organization. Individual plants respond mos t immediately to stress through various physiological mechanisms triggered by hormone signaling (Atkinson and Urwin 2012) Acclimation can only occur in response to stress that is within the environmental tolerance of an individual plant, and plants with higher phenotypic plasticity may show greater ability to acclimate to stressful conditi ons. Over longer time scales plant populations can adapt to stressful conditions as conditions select for the most stress tolerant individuals. Changes in the abundance of different plant species depending on stress tolerance can lead to community reorder ing (Collins et al. 2012, Jones et al. 2016) Under severe or long term stress, communities begin to lose the most susceptible species, which can alter ecosystem function as well as the trajectory of the
15 community recovery if the stressor is removed (Jones et al. 2016) The effects of stress on ecosystem function generally are less severe in communities with greater functional diversity (Fry et al. 2013) Interacting stressors can have different effects on plant populations and communities tha n predicted from the individual stressors, i.e. synergistic or antagonistic responses (Figure 1 1) The terms for these responses have been used inconsistently, resulting in the over emphasis of synergy in ecological literature (Ct et al. 2016) Folt et al. (1999) proposed a framework wit h three potential null models for the interaction between two stressors. First, t he comparative effect model (or dominance model) predicts that with two interacting stressors, the effect will be equal to that of the greater stressor alone. Deviation from t his hypothesis results in increased or decreased effect in comparison to the effect of the strongest individual stressor. The comparative effect model should be applied when two stressors have similar mode of action and is often applied to limiting resourc es as in the L aw of the Minimum (van der Ploeg et al. 1999) Second, t he multiplicative effect model predicts that the effects of the combined stressors will be the product of the effects of the individual stressors Deviation from this hypothesis can be called (Folt et al. 1999) The multiplicativ e effect is expected when the response to stress is measured in terms of mortality (Ct et al. 2016) Finally, t he additive effect model hypothesizes that the effect of two stressors will be equal to the sum of the e ffects of the individual stressors. Effects greater than and less than those predicted are considered synergistic and antagonistic, respectively. The additive effect is expected when the stressors have physiologically different modes of action, such as wit h interacting biotic and abiotic stressors, and this is the null model I will use throughout my work (Schfer and Piggott 2018)
16 At the community level, stressor s affect species differently: what is stressful for one species may not be stressful for ano ther species. Thus, stress may alter competitive interactions and community composition (Schfer and Piggott 2018) The alteration of biodiversity in response to multiple stressors depend s on the type of stressors and how the tolerance of the species in that community to each of the stressors covaries (co tolerance) (Vinebrooke et al. 2004) If species responses to each of the stressors tend to be positively correlated, then a dominance null model would be more appropriate but if they are negatively corre lated then an additive model would be more appropriate (Vinebrooke et al. 2004, Schfer and Piggott 2018) Our ability to predict the effects of interacting stressors on plant communities depends on a better understanding of the applicability of the various nu ll models under different circumstances and the standardization of selecting null models (Piggott et al. 2015, Thompson et al. 2018) Abiotic stress and plant plant competition are highly interconnected. Abiotic physical stress and the strength of competitive interactions te nd to be negatively correlated (Grime 1977, Bertness and Callaway 1994) The stress gradient hypothesis predicts that positive interactions (i.e. facilitation) are more co mmon under stressful environmental conditions, and is well supported in the literature (Bertness and Callaway 1994, Pugnaire and Luque 2001, Lortie and Callaway 2006, He and Bertness 2014) and these harsh conditions may reduce the success of introduced species In contrast, low abiotic stress tends to favor strong competitive interactions, and disruption of these interactions, for example by disturbance, provides an opportunity for invasion by non native species (Melgoza et al. 1990) Resource addition into a site with previously low resource availability or high variability in resource supply rate over time can also promote invasion because invaders are provided the opportunity to take advantage of excess resources (Huenneke et al. 1990, Alpert et al. 2000, Davis et a l. 2000, Shea and Chesson 2002,
17 Koerner et al. 2015) Th us abiotic stress can influence the success and impacts of biological invaders. Invasion by non native plants often reduces species diversity (Alvarez and Cushman 2002, Yurkonis et al. 2005, Adams and Engelhardt 2009, Gaertner et al. 2009, Gooden et al. 2009, Hejda et al. 2009, Vil et al. 2011, Cook Patton and Agrawal 2014) ; however, the mechanis m of reduction of diversity is rarely addressed and may involve interaction s with abiotic stress (Levine et al. 2003) Impacts may occur throu gh direct competition, interference competition through production of allelopathic compounds (Hierro and Callaway 2003) or may be mediated by feedbacks with abiotic factors like nutrient availability or disturbance regimes Abiotic stress imposed by climate change is often expected to increase plant inva sion and exacerbate invasion effects on native communities (Bradley et al. 2010b, Diez et al. 2012) Severe stress such as ex treme drought can increase competitive effects of invaders on native species and prevent recov ery after the stress event (Caldeira et al. 2015) The majority of research in this area has focused on how stressors such as climate change will impact the ability of invaders to colonize and spread, but relatively little is known about what the specific impacts of these combined stressors will be at the community level Current understanding of how stress ors interact to influence plant communities is surprisingly limited despite the fact that increased stress due to climate change is occurring on a global scale. More explicit identification of the expected outcomes of combined stressors is needed to ident ify general patterns and trends. Additionally, a better integration of organism responses to stressors across disciplines, particularly between agricultural and ecological studies as well as between terrestrial and aquatic systems, would be helpful to pred ict future responses to
18 global change. Finally, a systematic application of null models across studies would improve synthesis of data across studies. Microbial Communities The same stressors that influence plant communities can also impact soil biota, bo th directly and indirectly. Soil biota respond directly to abiotic conditions but also to changes in plant community structure or function that result from abiotic stress. Biotic and abiotic drivers of soil microbial communities are complex, interactive, a nd context dependent. The abiotic environment strongly affects soil microbes because of their close association with the soil matrix. Abiotic conditions such as pH, nutrients, temperature, soil moisture, and O 2 concentrations alter microbial communities, b ut consensus has yet to be reached on which of these is most important under what circumstances and for which aspects of the microbial community (Ramirez et al. 2010, Evans and Wallenstein 2012, Shen et al. 2015) Moreover, it is difficult to distinguish the direct effects of the abiotic environment on microbial communities from indirect effects resulting from vegetation responses to environmental factors. I n a changing global environment these influences take on even greater complexity. Most research about drivers of microbial communities is confined to very broad grouping s ; however, the ecological function of microbial taxa at the higher taxonomic levels is highly variable (Philippot et al. 2010) Therefore, assessing finer scale variation in microbial communities can provide a better understanding of the ecological niche of microbial taxa At the most general level, global patterns in total microbial biomass in soils are driven prim arily by soil moisture and nutrients (S erna Chavez et al. 2013) While it is generally thought that bacteria are more sensitive to soil moisture conditions than fungi (Evans and Wallenstein 2012, Ochoa Hueso et al. 2018) Blankinship et al. (2011) found that bacterial abundance was influenced by tempe rature, while fungal abundance wa s more sensitive to precipitation. Within bacteria,
19 nitrifiers and gram negative bacteria are generally more susceptible to water stress than gram positive bacteria, and Actinomycetes are among the most tolerant (Manzoni et al 2011) Additionally, bacterial communities appear to adapt to frequent dry wet periods which would typically cause lysis in many soil microbes (Fierer et al. 2003) Warmin g alters soil microbial communities and can increase fungal:bacterial (F:B) ratios; however, it is difficult to distinguish independent warming effects from those driven by altered soil moisture conditions (Zhang et al. 2005) Soil chemistry conditions also play a major role in shaping microbial communities. F:B ratios are dependent on soil C:N ratios, and nitrogen addition reduces fungal compared to bacterial activity (Frey et al. 2004, Fierer et al. 2009) Nitrogen addition has been shown to alter bacterial commu nity composition (Coolon et al. 2013) and data suggest that changes in bacterial communities are due to direct effec ts of N availability rather than effects of N availability on pH or plant community shifts (Ramirez et al. 2010) Many studies have suggested that s oil pH is a strong predictor of bacterial community composition across large spatial scales and pH can influence soil nutrient availability (Fierer et al. 2009, Lauber et al. 2009, Kaiser et al. 2016) Bacterial community shifts could be related to changes in the abundance of copio trophs that thrive in high resource conditions (e.g., A Proteobacteria) and oligotrophs that dominate in low resource condition (e.g., Acidobacteria and Verrucomicrobia) in response to variation in resource supply (Fierer et al. 2007, Ramirez et al. 2012). Only certain compon ents of the microbial community are active at a particular time while many microbes become dormant under stressful conditions Total bacterial community composition varied most between locations associated with different soil types and less strongly with vegetation type, while only the active bacterial community, as measured with RNA sequencing, was altered by a
20 change in precipitation, possibly suggesting different dormancy responses of microbial groups (Felsmann et al. 2015) Vegetation composition and productivity plays a key role in shaping soil microbial communities (Burns et al. 2015) Some microbial groups are intimately associated with plant hos ts including mycorrhizal fungi, rhizobia, and plant pathogens, and these groups are strongly driven by plant abundance and species composition (A llen et al. 1995, Burrows and Pfleger 2002) Plant species also differ in the quality and quantity of resource inputs into the soil and many studies have shown increased activity of microbes in the rhizosphere compared with bulk soil (Van Der Krift et al. 200 1, Fierer et al. 2007) Plant productivity and soil organic matter are major drivers of microbial biomass (Fierer et al. 2009) Plant traits can also influence some aspects of microbial communities; for example, slow growing conservative plant species have fungal dominated microbial associates (Orwin et al. 2010) Furthermore, some evidence suggests that microbial diversity increases with plant diversity (Felsmann et al. 2015) Various lines of evidence suggest that biotic and abiotic drivers interact to shape microbial communities (de Vries et al. 2012) Abiotic conditions determine the suitability of a habitat for diff erent plant species and productivity of a species at a site. These plant communities can then have both direct effects (e.g. plants supply resources) and indirect effects (e.g. plants alter the physical environment) on soil communities. This interdependenc e makes it difficult to distinguish the direct effects of plant communities on microbial communities (Wardle et al. 2004) Many a biotic factors such as warming elevated CO 2 soil nutrients, pH, and precipitation have been shown to alter microbial communities; however, th ese effect s may be mediated by changes in plant growth o r physiology (Horner Devine et al. 2003, Zhang et al. 2005, Lesaulnier et al. 2008, Thomson et al. 2010, Shen et al. 2015) de Vries et al. (2012) showed that various
21 abiotic factors influence microbial communities including precipitation, soil nutrients, and pH, but that plant functional traits, independent of the abiotic site cha racteristics also contributed to microbial community structure Therefore, both biotic and abiotic factors concurrently and possibly interactively influence microbial communities, but more research is needed to parse out the importance of each under diffe rent scenarios. Soil microbes can play an important role in plant invasion and impacts of invasion on ecosystem processes. Invasive plants can alter microbial community composition directly or indirectly, either by changing the plant community composition or soil properties (Niu et al. 2007) Invaders can alter ecosystem process rates controlled by microbes, such as nutrient cycling and decomposition (Ehrenfeld et al. 2001, Kour tev et al. 2002a, Allison and Vitousek 2004) For example, Microsteg ium vimineum invasion changes N and P cycling, soil pH, base cations, and Al (Ehrenfeld et al. 2001) and these soil properties can then alter soil microbial communities (Kourtev et al. 2002b) While many studies have shown changes in microbial communities with plant invasion (Broz et al. 2007) this trend is not universal (Carey et al. 2015) therefore we need a better understanding of how invaders modify microbial communities, and how these responses depend on the environmental contex t. Additionally, interactions of invaders with the microbial community may change over time. For example, local species may adapt through time since in troduction of an invader allowing native pathogens to infect the invader (Nijjer et al. 2007) The time scale over which this might occur is unknown and in some cases appears not to occur even over a century (Day et al. 2015) Soil microbial communities can also influence invasion success. For example, escape from soil pathogens provides a possible mechanism for invasion in a new range (Mitchell and
22 Power 2003, Reinhart et al. 2003) Invasive plants can create positive plant soil feedbacks favoring their own growth or they can increase negative feedbacks for native species (Niu et al. 2007, Xiao et al. 2014) Invasive plants can accumulate native pathogens that negatively affect the native plants more than themselves (Malmstrom et al. 2005, Eppinga et al. 2006, Mangla et al. 2008) Invaders can also alter soil mutualist abundance (Kourtev et al. 2002b, Hawkes et al. 2006) Invaders such as Solidago canadensis and Centaurea maculosa reduce mycorrhizal fungal abundance and diversity in soils (Mummey and Rillig 2006, Zhang et al. 2010) Invaders can also alter and disrupt mycorrhizal associations in more dependent native species (Mummey et al. 2005, Wolfe and Klironomos 2005, Vogelsang and Bever 2009, Hagan et al. 2013b) C. maculosa appears to be able to take advantage of mycorrhizal networks to enhance its own growth at the expense of native species (Callaway et al. 2004, Carey et al. 2004) Alliaria petiolata is a non mycorrhiz al invasive notorious for inhibiting AM colonization of native species with allelochemicals ( Roberts and Anderson 2001 ) resulting in reduced native growth ( Stinson et al. 2006 ) In contrast, one study showed no inhibitory effect of this plant on native plants or on AMF diversity ( Koch et al. 2010 ) so even the effects of the most noxious invaders may be context dependent. Plant Soil Interactions Plants have been shown to modify the soil environment through physical processes (e.g. alterations in soil temperature or pH (Raich and Tufekcioglu 2000, Hinsinger et al. 2003) ), biogeochem ical processes (e.g. nutrient cycling (Hinsinger 2001, Yelen ik and Levine 2011) ), as well as by modifying soil biotic communities (Ehrenfeld et al. 2005) These altered soil conditions created by the plant can result in differences in fitness of subsequent plants growing in th at soil. Abiotic factors can also modify soil properties and biota to create soil legacies Soil legacies can be either abiotic, such as changes in nutrient availability, or biotic, such as changes
23 in microbial community composition. Biotic soil legacies t hat are most influential for plant growth are typically associated with pathogens and microbial mutualists like rhizobia and mycorrhizal fungi (Klironomos 2002) Plants serve as the princ ipal energy supply for these organisms, and plant fitness is influenced by their metabolic activities. Microbial communities can also indirectly affect plant fitness through regulation of biogeochemical processes. For example, nutrient availability for pla nts is strongly controlled by microbial decomposition and nutrient demand (Craine et al. 2007, Kuzyakov and Xu 2013). Plant soil feedback (PSF) experiments are a specific type of soil legacy experiment used to assess whether and how plant mediated changes in either soil abiotic conditions or soil microbial communities generate feedbacks to plant performance. In order to test for a PSF, both components (plant effects on soil and soil effects on plants) must be demonstrated. This is typically accomplished thr ough a 2 stage experiment. The first stage, or priming stage, involves growing a single plant species in soil for a period of time to alter the soil characteristics. The second stage, or growth stage, involves growing conspecific plants in that soil, calle conditions but in soil not self (Kulmatiski and Kardol 2008) To assess the reciprocal fe would be soil primed by the other species. At the end of the experiment, plant traits are measured to estimate fitness; most often biomass, but growth rate, reproductive output, or survival can also be used. Reciprocal PSF experiments can demonstrate the theoretical potential for plant species coexistence based on differential response to soil microbial communities generated by different plant species (Bever 2003) How ever, there are a variety of issues that can reduce the direct applicability of these studies. They are typically conducted under highly controlled conditions
24 (Bever 1994, Mangan et al. 2010) which are unrealistic representation s of processes that occur in nature making conclusions drawn from the m somewhat suspect. Additionally, d ifferent experimental approaches have been shown to yield different results indicating that better me thodological standardization is needed (Brinkman et al. 2010) Finally, in nature many plant root systems are ofte n interacting in the soil, so one individual plant species is unlikely to be exclusively altering the soil environment. In contrast using field experiments t o provide more realistic scenarios during the soil priming stage can improve the relevance of thes e experiments to natural systems. Because realistic field experiments are unlikely to have only one species present, they are not directly comparable to PSF experiments and cannot accurately show the potential for species coexistence but can provide an und erstanding of the influence of soil biotic and abiotic legacies on the growth of different plant species. Ideally, soil legacy experiments should be capable of distinguishing between biotic and abiotic changes in the soil. When trying to deconstruct th e m echanisms behind plant soil interactions differences in plant performance are often compared between sterilized and unsterilized (live) soil. The difference in performance of plants grown in live versus sterile soil indicates the direct effect of microbes on the plants. Because soil sterilization can release nutrients in soil, this is often controlled for by adding a small amount of live soil inoculum to a sterile growth mixture and assuming that abiotic effects of sterilization are masked because of the s mall volume of added soil (Kulmatiski and Kardol 2008) F ew studies look at interspecific competition or community effects of soil legacies (Kulmatiski et al. 2008, Suding et al. 2013) The spatial scale of soil alteration by each individual plant is another consideration when attempting to infer or model community dynamics (Levine et al. 2006) Soil communities can be altered by environmental conditions making it difficult to
25 extrapolate experimental results to a broad range of field conditions, and therefore measurements must be taken under alternate scenarios (Kolb et al. 2002, Carvalho et al. 2010) This limitation is partly due to the lack of understanding of the biology of the soil microbial communities. Additionally, different plant plant intera ctions may influence the outcome of legacy effects (Shannon et al. 2012) and therefo re it is important to place the concept of soil legacies in the larger context of plant plant interactions in order to evaluate its relative importance compared to direct competition, allelopathy, or other forms of indirect competition in driving plant com munity assembly (Bennett et al. 2011) Studying the net effect of microbes on plant communities provide s e still revealing its effects on plant communities. As methods for characterizing the soil microbial community have become more tractable on large scales the next step is to identify the major microbial players in plant soil interactions so that we can ex tend our understanding of how they work to broader general patterns (Batten et al. 2008) Based on this review, I identified a gap in our knowledge of how plant invasions and climate change interact to affect native ecosystems. Ecosystem responses w ill rely heavily on the responses of plant communities, soil communities, and their interactions as they drive many ecosystem functions. Specifically, I test the following questions : 1. How do plant invasion and drought individually and interactively affect native plant communities? 2. How do plant invasion and drought individually and interactively affect soil bacterial and fungal communities? 3. How do the legacies of changes in soil bacterial and fungal communities due to invasion and drought alter the perform ance of native plant species compared with the invader and competition between native and the invader ? To test these questions, I used longleaf pine forests of the southeastern US as a model system. This region is concurrently threatened by invasion by Imp erata cylindrica (cogongrass) a
26 rhizomatous C 4 grass as well as increased frequency and severity of drought I expected a negative effect of both invasion and drought on plant diversity and an additive effec t of invasion and drought in combination. Further more, I expected that changes in microbial communities would mirror changes in the plant communities. I expected that the legacy effects of the invader on soil microbial communities would benefit the invader more than native species and the legacy of droug ht would favor native species over the invader. Finally, I address the implications of this research for restoration of native plant communities. Figure 1 1. Model diagram of possible interactions between two stressors (A and B). A response equal to the sum of the individual stressors represents the null model. A more negative response designates a synergy between the two stressors while a less negative response indicates antagonism.
27 CHAPTER 2 GRASS INVASION AND DROUGHT INTERACT TO ALTER THE DIVERSITY AN D STRUCTURE OF NATIVE PLANT COMMUNITIES Background Plant community structure and function are determined by multiple biotic and abiotic drivers (Baruch and Jackson 2005, Gornish and Miller 2015) but a nthropogenic environmental cha nges may alter these drivers and their effects on plant communities (Alvarez and Cushman 2002, Knapp et al. 2002) Many global environme ntal changes are occurring simultaneously, but effects of multiple stressors are difficult to predict based on evaluation of individual stressors because it is unknown if interactions between stressors will occur In the absence of interactions, the effect s will be additive such that the combined effect will be equal to the sum of effects from individual stressors (Zavaleta et al. 2003, Ct et al. 2016) Interactions between stressors can occur when the combined effects are greater than (synergistic) or less than (antagonistic) the predicted additive effect (Ct et al. 2016) Synergistic interactions among g lobal change drivers have the potential to magnify effects on biodiversity and ecosystem function, while antagonistic interactions could partially ameliorate negative effects (Brook et al. 2008, Caldeira et al. 2015) Improved understanding of potentially complex interactions among stressors and their effects on native communities is necessary to predict long term outc omes of global environmental change (Alpert et al. 2000, Bellard et al. 2013) I nvasive plants can alter the structure and function of natural communities through changes in species interactions, biogeochemical cycling, and disturbance regimes (Brooks et al. 2004, Liao et al. 2008, Vil et al. 2011) At th e same time, c limate change is expected to increase extreme weather events such as prolonged drought, which can cause stress to native plant communities and exacerbate impacts of invaders (Knapp et al. 2008, Bradley et al. 2010b, 2010a, Diez et al. 2012) Although prior s tudies have evaluated the effects of various climate change
28 factors on invasive plant establishment and performance (Dukes and Mooney 1999, Dukes et al. 2011, Eskelinen and Harrison 2014, Manea et al. 2016) it is unknown how global change ffects on native species community dynamics. The response of invaded communities to abiotic stress depend s on the relative stress tolerance of native and invasive species and how stress influences species interactions such as competition and facilitation (Bertness and Callaway 1994, Tylianakis et al. 2008) For example, while invasive species are often expected to disproportionately invade high resource environments, some invaders have higher resourc e use efficiency than native species, allowing them to compete in low resource environments (Funk and Vitousek 2007) Invaders also may benefit when resources become available during extreme climate events or when competitive interactions are disrupted due to native species losses and reduced bio tic resistance (Huen neke et al. 1990, Alpert et al. 2000, Davis et al. 2000, Diez et al. 2012) In such cases, climate change factors can interact synergistically with invaders to suppress native species (Caldeira et al. 2015) Conversely climate change may inhibit invader s that are less tolerant of abiotic stress than native species (Bradley et al. 200 9, Sorte et al. 2013, Liu et al. 2017) thereby reducing the effects of invasion Moreover invader s may mitigate climate change effects on native communities if they moderate stressful abiotic conditions (Rodriguez 2006) The nature of these interactions, whether additive, synergistic, or antagonistic, may change in magnitude and even direction through time as either th e invader becomes increasingly dominant climate stress is increasingly severe or thresholds in the ability of native and invasive plants to persist are exceeded due to restricted resource access or extreme physical stress. Thus, experimental studies that assess how interactions between stressors change over time are needed to forecast their net effects on community diversity and structure.
29 C 4 grasses are problematic invaders across the globe 2004, Flory and Clay 2010, Hager et al. 2016) They typically have high drought tolerance and therefore could become mo re problematic as frequency or severity of drought increases; however, they also have high water use efficiency and may not draw down water resources as much as C 3 competitors (Sage and Monson 199 9, Ward et al. 1999) Therefore, it is difficult to predict how native plant communities will respond to C 4 grass invasion under predicted future changes in precipitation. To evaluate the individual and interactive effects of plant invasion and climate ch ange on plant communities and how they change over time, we established a factorial field experiment with invasion by Imperata cylindrica (cogongrass), a rhizomatous C 4 grass native to Southeast Asia (Estrada and Flory 2015) and chronic drought (simulated with rainout shelters Alba et al. 2017) I mperata cylindrica is a globally problematic inva der of warm temperate to tropical systems and a Federal Noxious Weed in the US with severe impact s on t hreatened longleaf pine ecosystems (Brewer 2008) C limate change predictions forecast more frequent and prolonged droughts in many regions including the s outheast ern US (Wang et al. 2010, Singh et al. 2013) I mperata cylindrica is predicted to increase in range and impacts in response to climate change (Bradley et al. 2010b) Our specific objectives were to 1) quantify the effects of drought on the invader and resident plant species; 2) compare the independent and interactive effects of invasion and drought on plant s pecies richne ss, diversity, evenness community structure, and dynamics; and 3) assess whether the relative effects of individual and interacting stressor s on plant communities change over time Methods Experimental Design To determine the effects of I. cylindrica inva sion and drought on native plant communities, we established a common garden field experiment at the University of Florida
30 Bivens Arm Research Site (BARS) in Gainesville, FL (29 37' N, 82 21' W; MAP 1300 mm, MAT 20.5C). Soils are primarily Bivans sand ( 75%; 5% 8% slope ) and Blichton sand (25%; 2% 5% slope ; Natural Resources Conservation Service, Web Soil Survey). To prepare the site, the area was mowed and tilled. Then, i n May 2012, we established native plant communit ies in each of 40 4 m x 4 m plot s sp aced 2.5 m apart with 20 bare root longleaf pine ( Pinus palustris ) seedlings (Florida Forest Service, Chiefland, FL) and 36 native perennial grass and forb seedlings (12 spp. x 3 individuals; The Natives Inc., Davenport, FL). By establishing replicate plan t communities, we controlled for initial plant community composition. Herbaceous seedlings were grown in growth chambers and then in a greenhouse for a total of four months prior to transplanting. Species were selected based on their occurrence in longleaf pine forests and suitability for the site (See Table A 1 for species list). Plots were not weeded to maintain composition and numerous other species recruited from the seed bank and surrounding environment during the study A blocked factorial combinatio n of I. cylindrica invasion and precipitation reduction was applied in spring 2013 (10 replicates per treatment combination) one year after the native plant co mmunities were established (Figure A 1) For the invasion t reatment, we planted nine I. cylindrica seedlings per invaded plot. Rhizomes were collected from an on site population and plants were grown from rhizomes in a greenhouse for six weeks before being transplanted into the experimental plots. The simulated drought treatment consisted of rainout shelters with 89% areal coverage of polycarbonate roofing with 89% light penetration (TUFTEX PolyCarb, Fredericksburg, VA), gutters connected to pipes to move precipitation off site, root impenetrable belowground plas tic barriers to 1 m depth to prevent subsurface flow of water into the plots, and aluminum flashing buried to 5 cm depth and
31 extending 10 cm above the soil surface to divert overland water flow (Alba et al. 2017) Based on a systematic study of effectiveness of rainout shelter design (Yahdjian and Sala 2002) we anticipated that high roof cover would be required to achieve a moderate level of reduction in soil moisture. As expected, the average reduction in soil moisture in our experiment ranged from 30 50% in the drought treated plots. We constructed structures over no drought control plots with 22% white shade cloth to account for shading by the rainout shelters. Vegetation Surveys To assess the effects of invasion and drought on plant communities, percent areal cover of all woody and herbaceous plant species in the plots was quantified beginning one year after the initiation of the treatments. Cover was recorded in July, October, and February from July 2014 to February 2018 July and Oc tober surveys represented mid and late growing season (rainy season) respectively and February represented the winter dormant season (dry season). During peak growing season in July, we were particularly interested in the species richness and community c omposition patterns, and therefore percent cover of all species was recorded. I n October and February, we were interested in changes in the dominant species and therefore only species with 5% or greater cover per subplot were recorded. Cover was evaluated within six 0.75 x 0.75 m sub plots in each plot by the same person (C. Fahey) over time for consistency. Because it was common for canopies of species to overlap, total vegetation cover often exceeded 100%. The USDA Plants Database (plants.usda.gov) was us ed to determine species functional group and life history strategy (annual/biennial or perennial). Abiotic Measurements Soil moisture was recorded every 2 4 weeks ( 61 total time points) using a HydroSense II soil water sensor paired with CS659 12 cm soil water probe (Campbell Scientific Inc., Logan, UT, USA). Measurements were taken in each of four quadrants per plot and averaged for each
32 plot Because many plant species can access deeper water reserves, s oil moisture data also w ere collected at six depths (10, 20, 30, 40, 60, and 100 c m) in one location in each plot using a PR2 profiler probe connected to an HH2 data logger (Dynamax, Houston, TX, USA) at 14 time points Photosynthetically active radiation (PAR) was measured every four weeks beginning in Ap ril 2015 using an ACCUPAR LP 80 ceptometer (Decagon Devices, Pullman, WA). PAR was measured in each of the four cardinal directions facing the center of the plot at 0.5 m and at ground level (Alba et al. 2017) Average percent light availability per plot was used for statistical analyses. Statistical Analysis To test if the drought treatment influenced I. cylindrica cover over time we used a mixed effects model with drought treatment and date as fixed effects and plot nested within block as a random effect. To test for effects of invasion and drought treatment combination s on the dependent variables we performed mixed effects models with invasion, drought, and date as fixed effects and plot within block as a random effect. Dependent variables included soil moisture, light availability, species richness, diversity, evenness, and percent cover (total, resi dent, perennial grasses, annual forbs, and perennial forbs) Plots with only one species present were excluded from evenness calculations. Diversity was calculated as the exponent of (Jost 2006) For analyses including all time points (i.e., all percent cover analyse s, diversity, and evenness) the July time points were subset to species with 5% cover or greater per subplot Analyses of richness, colonization, extinction, and community composition included all species Soil profiler probe data was averaged across the 14 time points and then differences between treatments were tested at each depth with ANOVAs. Mixed effects (Pinheiro et al. 2016)
33 Plant community composition was analyzed for July of each year (with I. cylindrica excluded) using ordination by nonmetric multidimension al scaling (NMDS) of the Bray Curtis dissimilarity matrix by plot We tested for differences in community composition with invasion, drought, date, and their interactions with a permutational multivariate analysis of variance (PERMANOVA) with 999 permutati ons and permutations constrained within each block. Community data w ere standardized relative to total plot cover (%) prior to calculating dissimilarity (Ok sanen et al. 2016) We tested for homogeneity of group dispersions with the multivariate analogue of Levene's test (Anderson and Walsh 2013) vegan package. PERMANOVA and homogeneity of group dispersions were calculated with the nctions in vegan. All statistical analyses were performed in R version 3.3.1 (R Development Core Team 2016) Results Percent Cover Total live vegetation cover was similar among plots in July and October, but plots with I. cylindrica had nearly twice as much live vegetation cover than uninvaded plots in February (February m ean SE; uninvaded = 29.4% 2.0, invaded = 55.3% 2.2; Figure 2 1 a ). Total cover of species other than I. cylindrica was strongly influenced by invasion (mean SE; uninvaded: 60.5% 1.8; invaded: 18.5% 1.4), an effect that was stronger in July and October than in February (date x invasion; F 1,436 =10.5, P=0.001; Figure 2 1 b). In the uninvaded plots, total cover of species other than I. cylindrica was not affected by the drought treatment in the first year but was lower in drought plots throughout th e following two years (invasion x drought; F 1,27 =5.2, P=0.03). Imperata cylindrica cover increased in the first year and a half of sampling, peaked at 76.3% cover in October 2015, and then declined. Imperata cylindrica cover declined
34 sharply in February 20 18 following a particularly severe cold weather event (date; F 1,218 =3.5, P=0.06; Figure 2 1 c). Imperata cylindrica cover was similar in ambient and drought plots throughout the first two years of sampling, after which I. cylindrica cover was slightly lowe r in drought plots than ambient plots (drought; F 1,9 =4.4, P=0.07). Functional Groups We grouped herbaceous species other than I. cylindrica into the three most abundant functional groups: perennial grasses annual forbs, and perennial forbs P ercent cover of all three groups was significantly lower in invaded plots regardless of drought treatment (mean cover: perennial grass: uninvaded 15.7%, invaded 3.6%; annual forbs: uninvaded 19.3%, invaded 3.7%; perennial forbs: uninvaded 7.5%, invaded 2.4%) h owever, functional groups respon ded differently to drought in the uninvaded plots. P erennial grass abundance did not differ significantly between drought and ambient plots (date x drought; F 1 436 = 3.1 P = 0.0 8). A nnual forbs were more abundant in drought than ambien t plots (mean cover; ambient: 15.5%, drought: 23.2%; drought; F 1,27 = 6.7 P=0.0 2) and perennial forbs were less abundant in the drought than ambient plots ( mean cover; ambient: 10.8%, drought: 4.3%; drought; F 1 27 = 9.0 P =0.01 ; Figure A 5 a c). Woody species other than longleaf pine made up less than 3% cover in the plots on average. Longleaf pine cover was significantly lower in uninvaded drought, invaded ambient, and invaded drought plots compared to uninvaded ambient plots and this difference increased ove r time (date x invasion x drought; F 1, 436 = 18.7 P<0.001 ). Species Richness and Turnover In July 2014, one year after initiation of treatments, species richness was similar across all treatments ( mean SE ; 15.3 0.4; Figure 2 2 a); however, across the sub sequent three years, richness was on average 58% lower in invaded plots than uninvaded plots ( mean SE ; uninvaded: 15.7 0.6, invaded: 6.6 0.4) S pecies richness in the invaded plots declined by 41%
35 from 2014 to 2015 (mean SE ; 15.6 0. 4 to 9.2 0.8 ). Richness continued to decline in invaded plots from 2015 to 2016 ( mean SE ; 9.2 0.8 to 4.5 0.4) but increased slightly in 2017 ( mean SE ; 6.1 0.4; date x invasion; F 1, 116 = 62.1 P < 0.00 1 ; Figure 2 2 a) In u ninvaded plots species richness was sig nificantly lower under drought ( mean SE ; ambient: 18.1 0. 6 drought: 13.0 0.5), and this effect increased over time as richness was 20% lower in 2015, 35% lower in 2016, and 38% lower in 2017 (date x drought ; F 1, 116 = 2.9 P =0.09); however, there was n o difference in richness with drought in the invaded plots (invasion x drought; F 1,27 = 38.1 P < 0.00 1 ; Figure 2 2 a ) Total number of species present across all plots of each treatment in 2017 was 65% lower in invaded plots than uninvaded plots (uninvaded: 4 9 species, invaded: 17 species), 22% lower in drought plots (38 species), and 51% lower in invaded drought plots (24 species). The number of colonizing species was consistently higher in uninvaded plots compared to invaded plots (mean SE ; 5.1 0.4 vs. 1 .9 0.2; invasion ; F 1 36 = 55.2 P <0.001; Figure 2 2 b), but drought treatment only affected colonization events from July 2016 to July 2017 in uninvaded plots (mean; ambient = 7.3, drought = 4.7; invasion x drought ; F 1 36 = 4.6 P =0.04). Number of extinction events was unaffected by drought treatment, but the effect of invasion decreased over time. From 2014 to 2015, regardless of drought treatment, the number of plot level extinctions was nearly twice as high in invaded plots as in uninvaded plots (mean; inv aded: 7.9 vs. uninvaded: 4.0). By 2016, cumulative number of plot level extinctions was 32% higher in invaded plots (mean; invaded = 13.5, uninvaded = 10.2). By 2017, cumulative plot level extinctions were similar in uninvaded plots and invaded plots (mean ; uninvaded: 13.6 vs. invaded: 14.7, date x invasion ; F 1 76 = 7.4 P =0.008; Figure 2 2 c).
36 Diversity In July 2014 plant species diversity (e ) was similar among all plots. F rom July 2014 to February 2015 diversity declined more rapidly in invaded plots th an un invaded plots and remained less than 2.1 for the duration of the study (date x invasion; F 1, 436 = 6.6 P = 0.01; Figure 2 3 a). D iversity was lower in the uninvaded drought than uninvaded ambient plots but no difference w as observed between invaded drough t and invaded ambient plots (invasion x drought; F 1,27 = 6.1 P=0.0 2; Figure 2 3 a ). Species evenness also decreased from 2014 to 2015 in the invaded plots but not in the uninvaded plots (invasion; F 1, 27 = 40.3 P < 0.00 1, date; F 1,408 =34.4, P<0.001 ); however, e venness was not significantly affected by drought (invasion x drought; F 1,408 =0.5, P=0.50 Figure 2 3 b). Community Composition Invaded and uninvaded plant communities became more dissimilar over time ( PERMANOVA date x invasion; Pseudo F 1,159 = 12.7 P<0.00 1 ; Table A 2 ) and a significant interaction between invasion and drought was observed (invasion x drought; Pseudo F 1,159 =7.7 P<0.001; Figure 2 4 ). Despite the significant interaction, the R 2 values show that the effects of invasion on community compositi on were much stronger than the invasion x drought effect (invasion: R 2 =0.41; invasion x drought: R 2 =0.02; Table A 2 ). To better understand the treatment effects, we analyzed plant communit ies for each time point separately. In July 2014, invasion and droug ht effects were each significant but there was no interaction ( invasion: Pseudo F 1,39 = 2.9 P <0. 0 1 ; drought : Pseudo F 1,39 = 2.3 P =0.01; Figure 2 4 ). During July 2015, there was a significant interaction between invasion and drought where uninvaded ambient p lots were more similar to invaded plots than uninvaded drought plots ( invasion x drought: Pseudo F 1,39 =2. 0, P=0.04 ) During July 2016 and 2017, only invasion had a significant effect on community composition (2016: Pseudo F 1,39 = 12.3, P<0.01; 2017: Pseudo F 1,39 = 9.3, P<0.01). Overall, t here
37 was also greater group dispersion in the invaded than uninvaded plots ( Pseudo F 1,158 = 16.9 P < 0.0 01 ; Figure 2 4 ) which was significant for 2015 2017 (2014: Pseudo F 1,3 8 = 1.8, P=0.2; 2015: Pseudo F 1,3 8 = 10.2, P=0.003; 2016: P seudo F 1,3 8 = 13.9, P=0.002; 2017: Pseudo F 1,3 8 = 14.9, P<0.001; Table A 3). This effect indicates that the invaded communities (with I. cylindrica excluded from analysis) became more dissimilar to each other than the uninvaded communities. PERMANOVA results a re insensitive to heterogeneity in dispersion for balanced designs such as ours (Anderson and Walsh 2013) so differences between treatments in the PERMANOVA can be confidently attributed to differences in group centroids rather than spread. Soil Moisture and PAR On average, s oil moisture in the top 12 cm was 41% lower in drought plots than ambient plots ( mean SE; 11.5% 0.2 vs. 19.5% 0.2). Soil moisture was similar in the ambient plots with and with out the invader (mean SE ; 20.1 % 0. 3 vs. 19. 0 % 0. 3 ) ; h owever, the invaded drought plots had higher soil moisture compared to the un invaded drought plots (mean SE ; 1 3 4 % 0. 3 vs. 9.6 % 0. 2). This interaction varied significantly by date ( date x invasion x drought; F 1 2595 = 9.5 P = 0.00 2), where the effect of the drought treatment increased with the duration of the experiment and did not appear to vary strongly with large natural variation in precipitation ( Figure A 2; Hoover et al. 2018) Soil moisture differed between ambient and drought treatments up to 40 cm depth but not at 60 or 100 cm depth (depth 10 60 cm: F 1 27 > 5.2 P<0.0 06; Figure A 3). Light availability at the ground was significantly lower in the invaded plots, especially in the winter and early in the growing season but was not affected by drought (date x invasion: F 18 646 = 16.7 P < 0.0 01; Figure A 4). Light availability at 0.5 m was lower in invaded plots and was also lower in drought vs. ambient uninvaded plots depending on date (date x invasion x drought: F 18 646 = 1.7 P = 0.0 4).
38 Discussion Our result s show that the biotic stress from an invasive grass dramatically altered n ative plant communit ies while the abiotic stress from chronic drought ha d moderate effects that were partially ameliorated by the invader. I nvasion by I. cylindrica had major effects on resident species cover, diversity, evenne ss, and community compositio n, while d rought had more moderate effects on plant communit ies, including alteration of dominant functiona l groups and a modest reduction in diversity. Together, the effects of invasion and drought combined were similar to invasion alone and lower than wo uld be expected in an additive model (Ct et al. 2016) Thus, we found an antagonistic interaction between invasion and drought, and we were uniquely able to identify the likely cause of the antagonism because we doc umented the ameliorating effect of invasion on soil moisture. These findings show that in uninvaded communities, drought had significant effects on plant community structure but when invasion and drought were combined, invasion was the primary driver of co mmunity structure regardless of drought. I nvasive plants generally have higher water use than natives species and have been shown in some cases to reduce water availability to native species (Levine et al. 2003, Cavaleri and Sack 2010, Caldeira et al. 2015) ; however C 4 species typically have higher water use efficiency (Sage and Monson 1999) W e observed higher soil moisture in invaded drought plots compared to uninva ded drought plots and, i nterestingly, this modulation of soil moisture by the invader only occurred under drought conditions. This effect was likely due to a combination of higher water use efficiency of I cylindrica and reduced soil surface temperature a nd increased humidity (Alba et al. 2017) potentially reducing transpiration and evaporative soil wa ter loss. These data indicate that higher soil moisture in invaded drought plots is a mechanism whereby invasion can moderate drought effects on community structure and suggest that I. cylindrica does
39 not dominat e via water competition Instead, this invad er likely outcompetes native species for light by maintain ing a dense live canopy and thick layer of thatch throughout much of the year Therefore, despite the potentially lower water consumption of C 4 invaders, they are likely to continue to be problemati c under increased drought due to high drought tolerance and competitive ability. Previous studies have suggested that some plant invaders alter community dynamics by suppressing colonization of native species, and that drought reduces diversity mainly thro ugh local extinction of rare species (Tilman and El Haddi 1992, Yurkonis et al. 2005) Our results showed that I cylindrica both prevented species from colonizing and prom ote d the loss of resident species from the invaded plots, but over time lower colonization became more important in driving differences in richness. Additionally, drought reduced species richness by 38%, which was almost exclusively due to lower colonizati on in drought treated plots, particularly in the fourth year of the experiment. These findings suggest that the balance of colonization and extinction changes over time and that it is critical to identify and remove invaders before species losses and recru itment limitation occur particularly in threatened longleaf pine forest s where endemic habitat specialists are strongly affected by I. cylindrica invasion (Brewer 2008) L osses of native species in invaded areas may not only lead to less biotic resistance (Fargione and Tilman 2005) but also loss of ecosystem function and lower habitat quality (Vil et al. 2011) Researchers have hypothesized that cl imate change will promote plant invasion s because many invasive species have high tolerance for environmental stress (Dukes and Mooney 1999, Theoharides and Dukes 2007) and high phenotypic plasticity (Davidson et al. 2011) Additionally, climate stress may lower native community resistance to invasion (Dukes and Mooney 1999, Diez et al. 2012) ; however, there is no consensus on whether empirical evidence
40 supports this hypothesis (Be llard et al. 2013, Sorte et al. 2013) The native communities in our study showed very low biotic resistance to invasion regardless of drought treatments. Additionally, drought had a minor effect on the inva der despite the relatively severe and long term drought we imposed. Thus, invaders that are highly resistant to climate stress should be the target of management and research regardless of their ameliorating effects on abiotic stress. Additionally, I. cylindrica is known to alter fire severity in longle af pine forests (Brooks et al. 2004) so other interactions in this system, such as between drought and fire, must be considered. Native ecosystems around the globe are concurrently threatened by plant invaders and increasingly severe effec ts of climate change (Baruch and Jackson 2005, Going et al. 2009) but little is known about how these factors interact at the community scale. Interactions between biotic and abiotic stressors can be complex and o ur results show that plant invasion and drought can interact in unexpected ways with profound consequences for resident plant communities. We demonstrated that I. cylindrica invasion and drought do not act synergistically; that is, drought did not promote invasion or exacerba te invasion impacts (Hamilton et al. 1999) Instead, the invader ameliorate d effects of the drought by maintaining soil moisture Regardless, high losses of diversity, likely due in part to intense light competition with the invader, indicate that even invaders with high wat er use efficiency will be problematic under future drought scenarios Moreover, even in the absence of synergistic or additive interactions combinations of multiple global change d r ivers can reduce biodiversity and alter community structure in threatened ecosystems. T herefore, although drought may not enhance susceptibility of th i s ecosystem to invasion low biotic resistance indicates that invasion s are a major threat to their persistence under current and future precipitation regimes. Removal of such hig h impact invaders should be a top priority of natural areas management.
41 Figure 2 1 Seasonal changes in vegetation cover over 4 years. Percent cover of (a) total live vegetation, (b) resident vegetation (i.e., species other than Imperata cylindrica ), an d (c) Imperata cylindrica (Mean SE; N=10). Ambient = Uninvaded plots with ambient precipitation, Drought = Uninvaded plots with experimental drought, Invasion = Plots invaded with I. cylindrica and ambient precipitation, Invasion+Drought = Plots invaded with I. cylindrica and experimental drought.
42 Figure 2 2. Plant community dynamics in response to the individual and interactive effects of invasion and drought. (a) Plant species richness; (b) cumulative number of plot level colonization and (c) extincti on events measured in the summer of each year ( Mean SE ; N=10 ) For (b) and (c), year indicates the end of the time period over which colonization and extinction were measured.
43 Figure 2 3. Seasonal changes in diversity metrics in response to the individ ual and interactive effects of invasion and drought. (a) D iversity (e ) and (b ) evenness per plot over time (Mean SE ; N=10 ).
44 Figure 2 4. Non metric multidimensional scaling (NMDS) ordination plots of summer plant community composition 2014 2017. O rdination based on Bray Curtis dissimilarity among plots with I. cylindrica excluded from the analysis Two dimensional NMDS stress : 2014 = 0. 187 2015 = 0.2 04 2016 = 0.16 4 2017 =0.1 70
45 CHAPTER 3 PLANT INVASION AND DROUGHT INTERACTIVELY STRUCTUR E SOIL MICROBIAL COMMUNITIES Background Soil provides essential ecosystem services such as nutrient cycling and water storage, and soil microbial communities can moderate provisioning of these services However, responses of soil microbes to interacting gl obal change factors, such as changes in precipitation, nitrogen deposition, or introduction of invasive species, remain difficult to predict. Some studies have found significant changes in microbial community composition in response to multiple global chan ge factors (Castro et al. 2010) whereas others have found no changes in fungal and bacterial community composition (Carey et al. 2015) or that global change impacts are overshadowed by seasonal variation (Matulich et al. 2015) Additionally, microbial communi ties may respond directly to abiotic changes such as increased temperature or drought (Sheik et al. 2011) but indirect effects via changes in plant communities also may occur (Zak et al. 2003, Berg and Smalla 2009, Lange et al. 2014) Plant invasion can cause profound shifts in plant species compositi on, driving associated changes in soil microbial communities (Kourtev et al. 2002b) D ecreased precipitation has also been shown to alter the soil microbiome (O choa Hueso et al. 2018) ; however, it is presently unknown how these factors interact to influence soil microbial communities. For example, invasive plants could alter the response of soil moisture to drought either through physiological (e.g., water use e fficiency) or biophysical (e.g., shading) effects. Conversely, drought could influence the survival or competitive success of invasive plants. Therefore, additional research is needed to tease apart responses of soil microbes to interacting biotic and abio tic global change drivers. Plant communities play a major role in shaping soil microbial communities through effects of root structure and root exudation as well as litter inputs and microclimate control (Zak
46 et al. 2003, Burns et al. 2015) Plant species can be particularly strong drivers of certain microbial groups such as plant pathogens and mutualists, which often have specialist plant hosts or variable host respons es (Bever 2002, Augspurger and Wilkinson 2007, Wang et al. 2012) Therefore, marked changes in plant community composition such as those that occur during plant invasion are likely to have major impacts on microbial communities and these effects may differ across microbial functional groups such as mycorrhizal fungi or plant pathogens ; however, few studies have assessed the response of different microbial taxa to invasion. Some invasive plants modify the soil microbiome in ways that have negative impacts on native plant species, including reductions in mutualists that benefit native species and accumulation of pathogens that disproportionately harm native species (Eppinga et al. 2006, Stinson et al. 2006, Batten et al. 2008, Barto et al. 2011) Thus, plant invaders have the potential to modify soil microbiomes in ways that benefit invader performance, but not all invaders display this abi lity (Del Fabbro and Prati 2015 ) and a better understanding of which microbial groups could drive these responses is needed. Climate change is likely to increase the global frequency and severity of droughts (Burke et al. 2006, Singh et al. 2013) with major consequences for microbial communities. Drought can cause rapid declin es in microbial activity as well as shifts in abundance of microbial taxa such as Actinobacteria, Chloroflexi, and Glomeromycota with consequences for ecosystem function and plant microbe interactions (Castro et al. 2010, Manzoni et al. 2011, Ochoa Hueso et al. 2018) However, plants can ameliorate the impacts of drought on microbes thro ugh changes in microclimate and plant water relations (Alba et al. 2017) Furthermore, certain micro bial groups have been shown to mediate plant response to drought (Aug 2001, Yang et al. 2009) Therefore,
47 plant microbe interactions are likely to play an important role in ecosyste m response to climate change induced drought. Microbial responses to environmental factors have historically focused on total microbial biomass, fungal:bacterial ratios, or individual microbial groups (Allen et al. 1995, Blankinship et al. 2011, Bragazza et al. 2015) ; however, limit our ability to understand the ecological interactions occurring belowground. High throughput sequencing methods allow for a more detailed picture of changes in microbial communities at different tax onomic levels as well as for the organization of microbes into general functional groups (Nguyen et al. 2016) These methods have yet to be used to evaluate the response of soil microbial communities to the interaction of plant invasion and climate change. In addition to assessin g microbial diversity and overall community composition, the response of particular microbial taxa and functional groups to changes in the abiotic and biotic environment now can be assessed (Matulich et al. 2015, Anthony et al. 2017) th ereby facilitating evaluation of how microbial responses to global change shape whole ecosystem function (Cline et al. 2018b) To evaluate the potential interactive effects of plant invasion and drought on soil microbial communities, we sampled soil from a long term fully crossed invasion x drought field experiment and sequenced bacterial and fungal communities (the dominant components of the soil microbiome) Our o verarching question was: How do plant invasion and drought independently and interactively affect bacterial and fungal diversity and microbial community composition? In Chapter 2, analysis of the effects of invasion and drought on resident plant communit ie s indicated 58 % decrease in richness in response to plant invasion compared with a 2 8 % decrease in response to drought. Because plant communities influence microbial diversity,
48 we hypothesized that microbial communities would respond to plant invasion and drought, including a more dramatic response to invasion and less extreme response to drought (Thakur et al. 2015) In particular, we expected this pattern to hold true for plant dependent species such as plant pathogens and mycorrhizal fungi. However, because bacteria are typically more vulnerable to drought than fungi, and certain bacterial and fungal groups are known to respond differently to wet versus dry conditions (Castro et al. 2010, Maestre et al. 2015, Zhang et al. 2016, Meisner et al. 2018) we hypothesized that drought would have a larger effect on these sensitive taxa. The interaction between invasion and drought is more difficult to predict because the invader in our system ( Imperata cylindrica ) is relatively resistant to drought and also moderated drought effects on soil moisture (Alba et al. 2017; Chapter 2) Thus, we hypothesized that microbial community responses to invasio n might be similar in the ambient and drought plots because the invasion may moderate the response of microbial communities to drought. Methods Study System Longleaf pine ( Pinus palustris ) dominated forests historically covered ~30 million hectares of the southeastern US, making up more than 50% of upland areas, with an additional 33% of upland areas in mixed stands including longleaf pine. Longleaf pine forests cover less than 3% of their original extent and major efforts are focused on restoring these ec osystems (Frost 2007) Longleaf pine forests have an open canopy and diverse understory plant communities maintained by fire but invasive species such as Imperata cylindrica (cogongrass) can greatly red uce diversity and threaten rare native plant species (Hardin and White 1989, Walker and Silletti 2007, Brewer 2008) Imperata cylindrica is a globally distributed invasive grass infesting over 500 million ha in tropical and subtropical regions worldwide. It was introduced to the United States from Asia
49 and has invaded the southeastern US from Florida to Virginia and westward to Texas (Estrada and Flory 2015) Climate models suggest that I. cylindrica invasion will accelerate under future climate chan ge (Bradley et al. 2010b) but experimental evidence of cogongrass response and impacts under climate change are lacking. The southeastern US is expected to experien ce increased frequency and duration of drought over the next several decades (Singh et al. 2013) and therefore it is of critical importance that the impacts of I. cylindrica invasion and drought on native ecosystems be evaluated. Experimental Design The study site was located at the University of Florida Bivens Arm Research Site (BARS) in Gainesville, Florida, USA (29 and precipitation are 20.5C and 1300 mm respectively. Soils are p rimarily Portsmouth sandy loam ( 78 % sand, 19 % silt, and 3% clay) composed of Blichton sand (25%; 2% 5% slope) and Bivans sand (75%; 5% 8% slope; Natural Resources Conservation Service, Web Soil Survey). The study area was mowed and tilled to prepare the si te. In May 2012, we established plant communities in 40, 4 m x 4 m plots with 36 native perennial grass and forb seedlings (12 spp. x 3 individuals; The Natives Inc., Davenport, FL) and 20 longleaf pine seedlings per plot (Florida Forest Service, Chiefland FL). I n spring 2013 a fully crossed combination of I. cylindrica invasion and precipitation ates per treatment combination) W e planted nine I. cylindrica seedlings into each invasion plot. The drought treatment consisted of rainout shelters with 89% polycarbonate roofing cover (TUFTEX PolyCarb, Fredericksburg, VA) and gutters connected to pipes to move the water off site Around each plot belowground plastic barriers were inserted to 1 m depth (to prevent subsurface
50 flow of water), and aluminum flashing extend ed 10 cm above the soil surfa ce (to reduce overland flow) We constructed structures over no white shade cloth to mimic shadin g by the rainout shelters. The experimental setup is described in detail in Alba et al. ( 2017) Soil samples were collected in May 2016 from the experimental plots using a 5 cm diameter hammer corer. Surface plant litter was removed, and three cores were extracted per plot to a depth of 15 cm. All roots were hand s orted from the soil and separated into fine (<1 mm diameter) and coarse (>1 mm) fractions, washed, dried at 60 C for 48 hours, and weighed. The 5 15 cm fraction of each of the three cores per plot w as homogenized through a 2 mm sieve and composited. Soils were then frozen a t 10 C until DNA extraction. A 5 g subsample of fresh soil was dried at 105 C for 72 hours and weighed to calculate gravimetric soil moisture. DNA Extraction, PCR, and Illumina S equencing Genomic DNA was extracted from each soil sam ple (0.25 g) using MoBio PowerSoil DNA extraction kit (MO BIO Laboratories, Inc., Carlsbad, CA, USA). The 16S rRNA and fungal ITS genes were amplified for each sample using primer sets of F515/R806 (Bates et al. 2011) and 5.8S FUN/ITS FUN (Taylor et al. 2016) respectively. The primers were modified for the Illumina platform by fusing CS1 and CS2 linker primers for forward and reverse primers, respectively. Polymerase chain reactions were conducted with 50 L assays. For the 16S amplification, 25 L of GoTaq Colorless Master Mix (Promega, Madison, Wisconsin, USA), 5 L of BSA (5 ng L 1 ), 1 L of each primer (10 M), 15 L of PCR gra de water, and 3 L of a genomic DNA template (5 ng L 1 ) were mixed in a 200 L PCR tube for each sample. The following thermal profile was used for PCR: an initial denaturation and enzyme activation step
51 of 94 C for 3 min, followed by 30 cycles of 94 C for 45 sec, 50 C for 60 sec, and 72 C for 90 sec, with a final extension of 72 C for 10 min. For the fungal ITS amplification, 25 L of GoTaq Colorless Master Mix, 5 L of BSA (5 ng L 1 ), 0.8 L of each primer (10 M), 18.4 L of PCR grade water, and 18.4 L of a genomic DNA template (5 ng L 1 ) were mixed in a 200 L PCR tube for each sample. The following thermal profile was used for the fungal PCR: an initial denaturation and enzyme activation step of 96 C for 2 min, followed by 30 cycles of 94 C for 30 sec, 58 C for 40 sec, and 72 C for 120 sec, with a final extension of 72 C for 10 min. Qualities of PCR products were evaluated by agarose gel electrophoresis. Additional rounds of PCR were performed to fuse CS1/CS2 linker primers to the indices and adapters before Illumina MiSeq sequencing at Gnome Qubec (Montral, Qubec, Canada). Sequence Data Processing QIIME 1.9.0 toolkit (Caporaso et al. 2010) was used to process the Illumina sequences. Chimeric 16S and ITS sequences were identified using reference based (May 2013 version of Greengenes database, (McDonald et al. 2012) ) and abundance based methods, respectively, via USEARCH (Edgar 2010) and removed for downstream analyses. Operational taxonomic units (OTUs) were determined at the 97% similarity level of the nucleotide sequences (Stackebrandt and Goebel 1994) using the open reference OTU picking option. Taxonomy was as signed to each OTU via Greengenes (McDonald et al. 2012) and UNITE (Kljalg et al. 2013) databases for 16S and ITS sequences, respectively. For the 16S sequences, de novo sequences, which accounted 90.1% of OTUs but only 18.6% of the total sequ ences, were removed for downstream analyses. After non bacterial sequences and singletons were removed, remaining 16S sequences were aligned using PyNAST (Caporaso et al. 2010) to build a phylogenetic tree using FastTree (Price et al. 2009) The
52 bacterial sequences were rarefied at 71,166 sequences per sample, and analyzed via Phylocom (Webb et al. 2008) (Oksanen et al. 2016) in R 3.4.1 (R Developmen t Core Team 2016) was used for dbRDA. For the fungal ITS sequences, de novo sequences, which accounted for 45.2% of OTUs but only 0.7% of the total sequences, were removed for downstream analyses. The fungal data were rarefied at 25,650 sequences per sam ple. Analyses were conducted for total fungal sequences and for arbuscular mycorrhizal (AM) fungi, a monophyletic group comprising the phylum Glomeromycota, as a subset of total fungi. FUNGuild was employed to categorize the OTUs into functional groups, in cluding pathogens, saprotrophs and mutualists (Nguyen et al. 2016) We used only taxa for which a unique role was identified in FUNGuild Statistical Analyses Statistical analyses were conducted in R 3.4.1 (R Development Core Team 2016) Mixed effect ANOVAs using the nlme package (Pinheiro et al. 2016) were conducted with the invasion and drought treatments as fixed effects and block as a random effect. We used OTU richness, (Chao 1984) as diversity indic ators for bacterial and fungal communities. Microbial community composition was analyzed separately for bacteria, whole fungal community, and AM fungal community using ordination by nonmetric multidimensional scaling (NMDS) For the bacterial community we used the weighted and unweighted UniFrac distance matrix to account for phylogenetic distance and for the fungal community we used the Bray Curtis dissimilarity matrix because UniFrac distance is not valid for the ITS region We tested for differences in community composition with invasion, drought, and their interactions with a permutational multivariate analysis of variance (PERMANOVA) with 999 permutations and permutations constrained within each block. NMDS scores were
53 ction of the vegan package (Oksanen et al. 2016) PERMANOVA was calcul in vegan. in vegan to determine correlations between environmental characte ristics and community composition. We used Mantel tests to determine the correlations between plant community function in vegan. Results Bacterial Community The invasi on and drought treatments significantly affected bacterial diversity (Figure 3 1 a). The drought treatment effects were evident for all measures of diversity, including decreased Shannon diversity index, evenness, OTU richness, and Chao 1 index (Table B 1) Invasion by I. cylindrica significantly increased Shannon diversity index and evenness of soil bacteria but had no effect on richness or Chao 1 index (Table B 1). Soil bacterial community structure was also significantly altered by I. cylindrica invasio n and drought. PERMANOVA of unweighted UniFrac distance showed a significant invasion x drought interaction where drought had the largest effect and invasion had a smaller effect and invasion+drought had an intermediate effect (Table B 5 ) O nly drought wa s a significant predictor of bacterial c ommunity structure based upon weighted UniFrac distance (Figure 3 2 a b; Table B 4 ) Bacterial T axa Relative abundance of bacterial taxa was fairly consistent across plots and treatments with Acidobacteria, Proteob acteria, and Verrucomicrobia as the dominant phyla (Figure 3 3) but individual phyla were affected by the treatments. Thus, we analyzed the effect of invasion and drought on the most abundant phyla and families (>1% relative abundance). Actinobacteria wer e
54 20% more abundan t in drought than ambient plots while Bacteriodetes, Nitrospirae, and Planctomycetes were 11 32% less abundant ( D rought (D) : F 1,27 > 5.4, P < 0.03) Verrucomicrobia were 15% more abundant in drought than ambient plots but 16% less abundan t in invaded plots than uninvaded plots ( I nvasion (I) : F 1,27 = 17.2, P = 0.0003; D: F 1,27 = 10.1, P = 0.004) Acidobacteria, Chloroflexi and Firmicutes were not significantly affected by the treatments (Figure 3 5). Four bacterial families had higher rel ative abundance under drought (Acidobacteriaceae, Bradyrhizobiaceae, Gaiellaceae, and Koribacteraceae), while three families had lower relative abundance under drought (Chitinophagaceae Ellin515, and Syntrophobacteraceae ) (D: F 1,27 > 5.2, P < 0.03) Chtho niobacteraceae relative abundance decreased by 21% in response to invasion and increased by 29% in response to drought (I: F 1,27 = 19.6, P = 0.0001 ; D: F 1,27 = 23.7, P < 0.0001 ), while Sinobacteraceae increased by 21% in response to invasion (I: F 1,27 = 5. 3, P = 0.03) Solibacteraceae were 44% more abundant in uninvaded drought plots than uninvaded ambient plots but no different in the invaded plots (I x D: F 1,27 = 7.2, P = 0.01) Five families showed no significant response to the treatments (Bacillaceae, Burkholderiaceae, Gemmataceae, Hyphomicrobiaceae, Rhodospirillaceae). Relative abundance of bacteria associated with nitrification including Nitrospira (primarily a nitrite oxidizer) and Nitosovibrio (an ammonium oxidizer) was interactively affected by the invasion and drought treatments (I x D: F 1,27 = 9.0, P = 0.006) Specifically, nitrifiers were similarly abundant in the ambient and drought uninvaded plots and the invaded drought plots, but more abundant in the invaded ambient plots. The relative abu ndance of nitrogen fixing bacteria including Rhizobium, Mesorhizobium, Bradyrhizobium, Frankia,
55 Methylobacterium, Azospirillum was 52% higher in invaded plots than uninvaded plots (I: F 1,27 = 8.3, P = 0.008) Fungal Community Fungal diversity was not sig nificantly affected by the invasion or drought treatments (Figure 3 1 b). While fungal OTU richness, diversity, and evenness were slightly higher in the invaded drought treatment than the other three treatments, differences were not statistically significa nt (Table B 2 ). I nvasion and drought interactively affected community composition of fungi (Table B 6 ; Figure 3 2 c) where the invaded ambient plots and uninvaded drought plots were different from the uninvaded ambient plots but the invaded drought plots were similar to the uninvaded ambient plots Fungal Taxa Relative abundance of fungal phyla was highly variable across plots and treatments ( Figure 3 4 ). We analyzed the effects of the treatments on the most abundant fungal phyla and families (>1% relati ve abundance) The only fungal phylum affected by the treatments was Glomeromycota (AM fungi), which were 40% less abundant under drought (D: F 1,27 = 5.0 6, P = 0.03 ). Among the fungal families, Atheliaceae were nearly 10 times more abundant in invaded ambi ent plots than uninvaded ambient plots, whereas they were only twice as abundant in invaded drought plots as uninvaded drought plots (I x D: F 1,27 = 4. 7, P = 0.04). Chaetomiaceae, Hypocreaceae, Nectriaceae and Trichocomaceae were more abundant under droug ht in uninvaded plots but unaffected by drought in invaded plots (I x D: F 1,27 > 5.5, P < 0.03). Thelephoraceae were 90% less abundant in invaded plots but unaffected by drought (I: F 1,27 = 4.3 7, P = 0.046). Chaetosphaeriaceae Cortinariaceae Lipomycetace ae Rhizophydiaceae were not significantly affected by the treatments (Figure 3 6).
56 There was a significant interaction between invasion and drought on AM fungal diversity, where diversity was 6% lower under drought in uninvaded plots but 10% higher under drought in invaded plots (I x D: F 1,27 = 6.3, P = 0.02: Figure 3 1). There were no significant effects of invasion or drought on AM fungal richness or evenness (Table B 3), but community composition of AM fungi was significantly affected by both invasion and drought (I: Pseudo F = 1.6, P = 0.04, D: Pseudo F = 2.6, P < 0.001; Figure 3 2). Fungal Guilds We classified fungal OTUs into trophic modes and guilds based on the FUNGuild database. In particular, we were interested in mycorrhizal fungi and plant pat hogens as these groups are likely to play a role in plant community dynamics. In the fungal dataset, 39.5% of OTUs were identified as having known guilds. Fungal guilds were interactively influenced by the invasion x drought treatments (I x D x Guild; F 3,1 35 = 6.2, P < 0.00 1). Pa thotrophs, which include plant and animal pathogens as well as fungal parasites, were significantly more abundant in invaded ambient than invaded drought plots but no t different in uninvaded ambient and drought plots (I x D; F 1,27 = 5. 9, P = 0.02 ; Figure 3 7 a ). These differences were driven mainly by plant pathogens (I x D; F 1,27 = 4.5, P = 0.04; Figure 3 8 c ). Saprotrophs were more abundant in uninvaded drought plots than uninvaded ambient plots but not influenced by drought in in vaded plots (I x D; F 1,27 = 8. 9, P = 0.006 ; Figure 3 7 b ). Symbionts as a whole were not significantly affected by the treatments (Figure 3 7 c ) and neither were ectomycorrhizal fungi (Figure 3 8 b ). Drivers of Microbial Community Structure Although the b acterial and fungal communities were significantly correlated (r = 0.19, P = 0.002), neither bacterial nor fungal communities were significantly related to the plant community composition based on a Mantel test of the Bray Curtis dissimilarity matrices (ba cterial: r = 0.06, P = 0.93; fungal: r = 0.04, P = 0.84). We assessed the role of several biotic
57 and abiotic environmental factors, including total plant cover, plant species richness, fine root biomass, coarse root biomass, bulk density, and mean percen tage of sand, silt, and clay in the soil, in structuring microbial communities. Fine root biomass was twice as high in invaded versu s uninvaded plots and 25% lower in drought than ambient plots (I: F 1,27 = 19.8, P = 0.0001; D: F 1,27 = 3.9, P = 0.06). Coars e root and rhizome biomass was interactively a ffected by invasion and drought where drought decreased coarse root biomass by 71% compared to ambient plots in uninvaded treatment while drought decreased coarse root biomass by only 43% compared to ambient i n invaded plots (I x D: F 1,27 = 7.4, P = 0.01; Fig ure B 2 ). On average, coarse root biomass was over six times higher in invaded than uninvaded plots. Fine root biomass and total plant cover were significant predictors of bacterial community composition ( fine root biomass: R 2 = 0.22, P=0.03; plant cover: R 2 = 0.16, P=0.01), while none of the environmental factors assessed predicted fungal community composition Discussion Our results demonstrate that soil bacterial communities in the longleaf pine ecosystem are primarily driven by drought with more moderate responses to invasion by I. cylindrica Fungal diversity was unaffected by the invasion and drought treatments but fungal community composition and some particular fungal taxa and functional groups respond ed interactively to invasion and drought. The greater response to drought of bacterial diversity than fungal diversity fits with our hypothesis and is consistent with previous studies that documented greater sensitivity of bacteria to soil moisture condit ions in a variety of systems (Gordon et al. 2008, Clark et al. 2009, Cregger et al. 2012, Barnard et al. 2013) T he lack of fungal diversity response to invasion did not fit our expectation whi ch was based on the idea that higher plant diversity would create greater fungal niche space (Cline et al. 2018a) However, fungal community composition changed in response
58 to invasion and drought which suggests that while the total number and relative abundance of OTUs remains relatively consistent, fungal community reordering occurs to favor different OTUs in response to the treatments (Lankau and Lankau 2014, Roy Bolduc et al. 2016) Invasion had dramatic effects on the plant community and total root biomass while drought had more moderate effects (Chapter 2; Figure B 2) ; hence, we expected that bacter ial and fungal communities would be strongly influenced by invasion and only moderately by drought Relative abundance of bacterial phyla responded strongly to drought. In particular, we found that Actinobacteria were favored under drought, while Bacterio detes, Nitrospirae, Planctomyces, and Proteobacteria were less abundant in drought plots. P revi ous studies have shown responses of bacterial taxa to drought that are not fully consistent across s tudies but Actinobacter ia appear to consistently be associat ed with drought or aridity (Battistuzzi and Hedges 2009, Acosta Martnez et al. 2014, Evans et al. 2014, Maestre et al 2015, Ochoa Hueso et al. 2018) Furthermore, m ost previous studies on the impacts of drought have been conducted in dry land systems (Clark et al. 2009, Cregger et al. 2012, Maestre et al. 2015, Ochoa Hueso et al. 2018) ; therefore more studies are needed to identify patterns of bacterial responses to drought within and across systems In our study, t he only phylum that exhibit ed a response to both drought and invasion was Verrucomicrobia, which w as more abun dant in drought plots and less abundant in invasion plots. From a functional standpoint, bacterial genera of particular significance for the nitrogen cycle respond ed strongly to invasion (Sy et al. 2001, Lee et al. 2012, Bahulikar et al. 2014) so that interactions with N cycling and I. cylindrica invasion are likely. Within phyla, bacteria are highly diverse in function and metabolism but studies have shown ecological coherence at the level of order and below (Philippot et al. 2010) Bacterial families were most often affected by drought with seven of the 15 most abundant families
59 responding only to drought however these families showed diverse ecological functions including aerobes and anaerobes, freeliving heterotrophs and endosymbionts and some families for which little is known One family responded only to invasion, one family to invasion and drought independently, and one family t o the interaction of invasion and drought. These results suggest that at multiple taxonomic levels, bacteria are more sensitive to soil moisture conditions than to plant invasion and the concurrent alteration of plant communities. While overall diversity of fungi was unaffected by the invasion and drought treatments, community composition and lower taxonomic levels responded to the treatments C ommunity composition of fungi in the uninvaded ambient plots was similar to the invaded drought plots but dissim ilar to the invaded ambient plots and uninvaded drought plots (Figure 3 2 d). This result suggests that invasion and drought may counteract each other in combination to alleviate their effects. At the level of phylum, only relative abundance of Glomeromyco ta (AM fungi) w as affected by the treatments. AM fungal relative abundance decreased in response to drought alone (Ochoa Hueso et al. 2018) while community composition responded to both invasion and drought. AM fungal diversity display ed an interactive response to drought and invasion where in uninvaded plots drought lower ed AM fungal diversity but in invaded plots drought drove grea ter AM fungal diversity (Figure 3 1 c ). Results from previous studies on AM fungal response to invasion have been mixed, with some studies showin g decreased AM fungal inoculum potential and diversity, as well as shifts in community composition, and others showing no response to invasion even in studies of the same invader (Roberts and Anderson 2001, Hawkes et al. 2006, Mummey and Rillig 2006, Stinson et al. 2006, Burke 2008, Barto et al. 2011, Ko ch et al. 2011) Furthermore, many of these studies have focused on a single non mycorrhizal invader, Alliaria petiolata and mycorrhizal invaders may have different effects on AM fungal communities.
60 I mperata cylindrica associates with AM fungi, so the ch ange in community composition but similarity in overall AM fungal abundance in invaded plots may indicate that the invader maintains mycorrhizal symbiosis while altering the AMF species composition to benefit its own growth. Multiple fungal families compos ed mainly of saprotrophs and pathogens showed the same interactive response to the treatments where their abundance was higher in uninvaded drought plots than uninvaded ambient plots, but no different across the invaded plots ( Chaetomiaceae, Hypocreaceae, Nectriaceae and Trichocomaceae ). This pattern may indicate that these taxa are drought tolerant and that the greater soil moisture associated with the invader in drought plots was sufficient to reduce the abundance of drought tolerant taxa (Chapter 2) Ad ditionally, total saprotroph abundance was higher in drought than ambient plots in the uninvaded treatment (Figure 3 7). This response may reflect increased detritus for decomposition under drought in uninvaded plots (Treseder et al. 2016) The lack of differen ce in the abundance of saprotrophs between the ambient and drought treatment in the invaded plots could be due to the low quality litter of I. cylindrica (Hagan et al. 2013a) or that the invader was more res istant to the drought treatment as seen in Chapter 2 Pathotrophs, and specifically plant pathogens, were most abundant in invaded ambient plots (Figure 3 7 & Figure 3 8) indicating that the invader can accumulate pathogens, but it is unknown whether these pathogens have differential effects on native plant species as has been seen for other invaders (Mangla et al. 2008) In the invaded drought plots, however, the abundance of plant pathogens was lower, indicating that the pathogens associated with invasion may be susceptible to drought. Symbiotrophs as a whole and ectomycorrhizal fungi were sl ightly less abundant in invaded plots but neither trend was significant. On the other hand, certain ectomycorrhizal families did show significant responses.
61 For example, Thelephoraceae were less abundant in the invaded plots. Conversely, Atheliaceae, which cont ains some ectomycorrhizal fungi but also saprotrophs, were more abundant in invaded ambient plots, but showed low abundance in all other treatment combinations. Therefore, even within functional groups, fungi still show taxa specific responses. T he drivers of soil microbiome structure are mixed and taxon specific While bacterial and fungal communities were strongly correlated, neither was significantly related to patterns in the plant community. This result differs from previous findings from Califo rnia grasslands suggesting both bacterial and fungal communities are correlated with plant communities (Matulich et al. 2015) Additionally, plant invasion was a weaker driver of bacterial communities than drought. Drought reduced all measures of bacterial diversity as well as bacterial community composition, while invasion increased diversity and evenness but not richness and C hao1 index The majority of abundant bacterial phyla and families responded to the drought treatment, though different taxa either increased or decrease d under drought. These observations indicate that abiotic conditions are stronger drivers of bacterial communities than invasion and consequent changes in plant communities However, bacteria associated with the nitrogen cycle were more abundant with invas ion. Fung al diversity was comparatively resilient to both invasion and drought; however, community composition and specific fungal taxa respond ed to both drought and invasion. Additionally, AM fungi were less abundant under drought, but plant pathogens wer e more abundant with invasion. Our results provide a better understanding of the drivers of soil microbiomes and their potential role in ecosystem function and plant community dynamics.
62 Figure 3 1. Shannon diversity index of a) baterial, b) all funga l, and c) arbuscular mycorrhizal fungal OTUs.
63 Figure 3 2. Nonmetric multidimensional scaling ordination of a) weighted UniFrac distance of the bacterial community (considers relative abundance), b) unweighted UniFrac distance of the bacterial community ( considers presence absence), c) Bray Curtis dissimilarity of the whole fungal community, d) Bray Curtis dissimilarity of the arbuscular mycorrhizal fungal community. Larger points with error bars indicate the mean SE of the 10 plots per treatment and sma ller background points show each plot NMDS values.
64 Figure 3 3. Relative abundance of bacterial phyla by plot in each invasion x drought treatment Each bar represents one plot.
65 Figure 3 4 Relative abundance of fungal phyla by plot in each invasion x drought treatment. Each bar represents one plot.
66 Figure 3 5. Relative abundance of the most abundant bacterial phyla (>1% mean relative abundance) in response to invasion and drought treatments (mean standard error) : a) Acidobacteria, b) Actinobacter ia, c) Bacteriodetes, d) Chloroflexi, e) Firmicutes, f) Nitrospirae, g) Planctomycetes, h) Proteobacteria, i)Verrucomicrobia.
67 Figure 3 6. Relative abundance of the most abundant fungal families (>1% mean relative abundance) in response to invasion and d rought treatments (mean standard error): a) Atheliaceae, b) Chaetomiaceae, c) Chaetosphaeriaceae, d) Cortinariaceae, e) Hypocreaceae, f) Lipomycetaceae, g) Nectriaceae, h) Rhizophydiaceae, i) Thelephoraceae, and k) Trichocomaceae.
68 Figure 3 7. Relati ve abundance of fungi by trophic mode : a) pathotrophs, b) saprotrophs, c) symbiotrophs, and d) other, in response to invasion and drought treatments (mean standard error). Trophic mode was assigned according to FUNGuild.
69 Figure 3 8. Relative abundanc e of selected fungal guilds (mean standard error): a) arbuscular mycorrhizal fungi, b) ectomycorrhizal fungi, c) plant pathogens Guilds were assigned according to FUNGuild.
70 CHAPTER 4 COMPETITION AND SOIL LEGACIES ALTER THE ROLE OF SOIL MICROBES IN PLA NT INVASION Background Invasive plants are problem atic because they can alter natural communities and ecosystems and are costly to land managers in time and resources (Mack et al. 2000, Flory and Clay 2010, Simao et al. 2010) When nonnative plants invade a community, they are subjected to novel interactions with species not present in their native range. These interactions can include interspecific competition, herbivory, pathogens, and mutualism s. The success of a plant species in a new range is partially dependent on the relative strengths of these interactions. Soil biota can play a critical role in mediating competition between native and invasive plants for example if invaders accumulate pat hogens that inhibit native species or degrade mutualistic networks (Wolfe and Klironomos 2005, Stinson et al. 2006, Mangla et al. 2008) These interactions may differ amo ng species with traits similar to the invader and those with different traits (Bever et al. 2010) For example, species that share common mycorrhizal types (i.e. arbuscular mycor rhizae (AM) or ectomycorrhizae (ECM) ) have the potential to share resources through common mycorrhizal networks (Robinson and Fitter 1999) Additionally, more closely related species are more likely to have common pathogens (Gilbert and Webb 2007) ; hence species that share traits with an invader could experience more negative interactions associated with soil microbes. Soil microbes can influence invasive plants and interactions with native plants bot h directly and indirectly (Wolfe and Klironomos 2005, In derjit and van der Putten 2010) Direct plant plant competition is related to the availability of resources and resource use by each competing plant species (Tilman 1982, Maron and Marler 2008; Figure 4 1 a c) and soil microbes can influence competition by altering availability of soil re sources to plants (Ehrenfeld et al. 2001) Specifically, saprotroph s drive ecosystem processes including decomposition and
71 nutrient cyc ling which affect soil nutrient availability to plants; however, invaders can disrupt these processes, for example via altered plant litter quality and quantity (Gordon 1 998, Holly et al. 2009, Lee et al. 2012; Figure 4 1 g) Additionally, soil mutualists such as mycorrhizal fungi can increase access to soil resources for associated plant species and can be involved in direct transfer of nutrients between plants (Robinson and Fitter 1999, He et al. 2006) Invaders may take advantage of or degrade mutualistic networks thus indirectly altering soil resour ce availability to native species (Marler et al. 1999, Mummey and Rillig 2006) Plants can also interact directly through interference competition via production of phytochemicals (Bennett et al. 2011; Figure 4 1 j l) Some microbes can breakdown phytochemicals thus reducing the ir effects on competing plant species (Li et al. 2 015) On the other hand, some invaders cause degradation of mycorrhizal networks by production of phytochemicals which indirectly affect native species (Roberts and Anderson 2001, Stinson et al. 2006, Cipollini et al. 2012) I nvaders and native species may respond differently to soil pathogens i f invaders are released from natural enemies in the introduced range (Mitchell and Power 2003) Furthermore, some invaders have been shown to accumulate pathogens that dispro portionately inhibit native species (Mangla et al. 2008; Figure 4 1 f) The complexity of these interactions has made it difficult to parse out the role of microbes in plant competition and invasion and an improved understanding of these interactions could improve outcomes for restoration of invaded ecosystems Plant soil interactions are likely to change under altered environmental conditions associated with climate change. For example, d rought is one of the strongest abiotic drivers of change in microbial community structure and function (Manzoni et al. 2011) and the effects of soil moisture on microbial commun ities can in turn influence plant performance and competition (Kaisermann et al. 2017, Fry et al. 2018) Drought may have multiple impacts on plant soil
72 interactions, includ ing effects on the availability of and plant access to soil resources (Fig ure 4 1 a c), the abundance and composition of soil biota (Evans and Wallenstein 2012; Figure 4 1 d i) and the production and movement of allelochemicals in the soil (Fig ure 4 1 j m) A ll of these effects may direct ly or indirect ly affect plant performance but it is unknown how they will interact to influence native versus invasive plants The aim of this study was to evaluate the p otential effects of soil legacy of invasion and drought on plant performance The invader, c ogongrass ( Imperata cylindrica ) is a warm season, rhizom atous, perennial grass native to Southeast Asia that has invaded 500,000 ha in Florida and over 500 million ha worldwide (MacDonald 2004, Estrada and Flory 2015) It is listed as a Federal Noxious Weed partly due to its impacts on threatened longleaf pine ecosystems in the southeastern US (Brewer 2008) Longleaf pine ( Pinus palustris ) and wiregrass ( Aristida stricta ) are foundation species in longleaf pine forests (Noss 1989) T he se species differ in their mycorr hizal associations ; cogongrass and wiregrass both associate with arbuscular mycorrhizal fungi, while pine is ectomycorrhizal. This difference in mycorrhizal associations is likely to influence how soil biota affect the interactions among these plants (Toju et al. 2014) The s pecific objectives of this study were to; 1) Determine the strength and direction of the impacts of microbial communities on performance of cogongrass and native species, 2) Evaluate how legacy effects of invasion and drought on soil microbial communities influence cogongrass and native species performance 3) Assess the role of soil microbes and legacy effects on competitive interactions between cogongrass and native species, and 4) Quantify the effect of soil microbes on net productivity of cogongrass and natives in competition. We expected that microbial effects on native species would be negative due to the effects of plant pathogens but neutral/positive for the invader because of enemy release. In Chapter 2 and
73 Chapter 3, we showed that invasion and dro ught both influence plant and soil microbial community composition; therefore we expected substantial soil legacy effects of both invasion and drought on native species Because of the drastic competitive effects of cogongrass invasion on native plants, w e hypothesiz ed that microbes promote cogongrass competitive ability compared to native plants. Finally, we hypothesized that overall microbes would limit plant productivity in competition due to pathogen effects, and that soil legacy effects of invasion a nd drought would limit productivity due to differences in pathogen and mutualist abundance observed in Chapter 3 Methods Field Experiment We conducted this study at a long term factorial invasion x drought field experiment at the University of Florida Bi vens Arm Research Site. Detailed methods for this experimental setup have been previously described ( Alba et al. 2017) Briefly, we established 40 plots, each 4 m x 4 m, in spring 2012. In each plot, we planted 20 longleaf pine seedlings and 36 native herbaceous seedlings (12 spp. x 3 individuals). In spring 2013, a crossed combination of drought treatme nt and cogongrass invasion was applied. The drought treatment was implemented by constructing rainout shelters with 89% roof cover, which reduced soil moisture by 40% on average over 5 years. Similar shelters were constructed over ambient rainfall plots wi th sh ade cloth instead of roofing to create similar light levels across the treatments (Chapter 2). The invasion treatment was applied by planting nine seedlings of cogongrass into the invasion plots, which resulted in 50% invader cover on average. The inv asion and drought treatments served as the soil legac y conditioning for this greenhouse experiment.
74 Greenhouse Experiment In a greenhouse at the University of Florida, we established an experiment to test the performance of three species; cogongrass longl eaf pine, and wiregrass, in response to soil microbial communities from the invasion x drought field experiment. In May 2017, soil samples were collected to provide inoculum by taking four soil cores in each plot with a 5 cm diameter x 15 cm deep hammer co rer. The soil corer was sterilized with 80% ethanol between each plot. The four cores per plot were composited and sieved through a 2 mm sieve. Roots were removed to reduce phytotoxicity effects. Background soil medium consisted of a 1: 1 mix of sand and l ocal topsoil collected near the field experiment. Topsoil was sieved through a 6 mm screen to remove rocks and large roots. Sand and topsoil were mixed in a cement mixer for ~2 minutes. The soil mixture was autoclaved for one hour on three successive days and then stored in a cold room until planting. Half of each soil sample from the field plots was sterilized by autoclaving in the same way and the other half was kept as live inoculum One gallon pots were filled with the sterile soil medium then covered with a thin layer ( 125 g ) of live or sterile inoculum (~5% of soil volume). This small volume of soil inoculum was used to reduce the effects of nutrient release from sterilization and abiotic effects of the treatments (Kulmatiski and Kardol 2 008) Finally, 5 cm of sterile soil was added to cap the inoculum Locke, NY ) were surface sterilized twice in 10% bleach for 15 minutes and rinsed. They were then soaked for 24 hours at room temperature in D I water. Excess water was drained and then seeds were stored for 7 days at 4 C before planting into sterilized sand for germination. Wiregrass seeds (The Natives Inc. Davenport FL) were surface sterilized for 5 minutes, rinsed with DI water and planted into
75 sterilized sand for germination. Cogongrass rhizomes were collected from a population near the field experiment Sheaths and fine roots were removed and then rhizomes were cut into three node segments and surface sterilized for two minutes. They were then rinsed and planted into sterilized sand. Seedlings of pine and wiregrass were transplanted into the pots in June 2018. These species were allowed to establish for one month and then cogongrass seedlings were transplanted into the pots. In total, we h ad four species combinations (each of the three species alone and all three species together in competition), two inoculum treatments (live/sterile), and four soil conditioning treatments (ambient uninvaded, ambient invaded, drought uninvaded, drought inva ded), with 10 independent replicates (corresponding with the 10 replicates in the field experiment) for a total of 320 pots. The pots were grouped into 10 blocks in the greenhouse matching the blocking structure of the field experiment. Seedlings that died within two weeks after transplanting were replaced with sterile seedlings of the same age. Seedlings that died after the first two weeks were also replaced to maintain competition effects but were excluded from the statistical analyses. Seven pine seedlin gs died after the first two weeks, all of which were in the sterile inoculum treatment. T wo wiregrass seedlings died and n o cogongrass seedlings died (Table C 4) In late November, above and belowground biomass was harvested. Roots were washed free of soi l, and for competition pots separated by species, and then above and belowground biomass was dried at 60 C for at least 72 hours before being weighed. Statistical Analysis To test for effects of live versus sterile soil inoculum, invasion and drought soi l legacy combinations, and competition, we used mixed effects models with inoculum, invasion, drought,
76 competition, and all possible interactions as fixed effects, and field plot nested within block as a random effect. Response variables included total pla nt biomass for each species, root: shoot ratio, and relative competition intensity. We calculated the relative competition intensity (RCI) index as the difference between the biomass when grown alone and biomass when grown in competition divided by the bio mass when grown alone (Weigelt and Jolliffe 2003) We also assessed the total biomass of all th ree species in the competition treatment in response to soil inoculum (live vs sterile) and soil legacy (invasion x drought) with a mixed model including inoculum, invasion, and drought as fixed effects and plot nested within block as a random effect. In order to determine the relative importance of direct competitive effects on the focal species of the biomass of each of the other species in the competition treatment, we ran a mixed model for each species in the competition treatment including soil inocul um and the biomass of each of the other species in the pot as fixed effects, and plot and block as random effects. A significant statistical interaction between the soil inoculum and soil legacy of invasion or drought indicate s that the response to the so il legacy differed in the live versus the sterile soil and therefore is referred to as a biotic soil legacy ( i.e. resulting f rom soil microbial communities). A significant individual effect of soil legacy with no significant interaction term indicate s that live and sterile soil inoculum treatments responded similarly and cannot be attributed solely to microbes and therefore is referred to as an abiotic soil legacy (i.e. not resulting from soil microbial communities) We expected abiotic effects to be minimi zed because we added inoculum that made up only 5% of total soil volume. While raw data were used for analyses, for visualization we also calculated the relative difference in total biomass and root:shoot ratio between the live and sterile treatment as (l ive sterile)/sterile. Because the inoculum soils were kept separate by plot and not pooled by
77 treatment, the plants formed natural pairs grown in live and sterile soil and replication was maintained when calculating th ese ind ices (Rinella and Reinhart 2018) Additionally, as detailed below there were no biotic effects of drought; thus, the ambient and drought data have been pooled in the figures. Figures with raw data are presented in Appendix C (Figure C 1 to Figure C 3). Results For cogongrass grown alone, plants grown in live soil had 12% greater total biomass than those in sterile soil, regardless of the soil legacy of invasion or drought. However, for cogongrass in competition, plants grown in live soil had 38% lower biomass than in sterile soil (inoculum x competition: F 1,108 = 25.7, P < 0.001; Fig ure 4 2 ). There was a legacy effect of drought on cogongrass performance such that cogongrass grown in soil with a history of drought did better in both live and sterile soils, indicating an abiotic legacy effect (drought: F 1,27 = 6.2, P = 0.02); however, the magnitude of this effect was relatively small (8% increase ; Table C 1 ). Pine seedlings exhibited responses that were the opposite of those just reported for cogongrass When pine seedlings were grown alone, they had 25% lower biomass in live soil than in sterile soil regardless of soil legacy (Figure 4 2) W hen grown in competition, however, pine seedling biomass was 17% higher when grown in live than sterilized soil (inoculum x competition: F 1,101 = 49.1, P < 0.001 ; Table C 2 ). W hether alone or in competition w ireg rass biomass was lower in live than sterilized soil (27% decrease in biomass) (Figure 4 2) Across live and sterile soil there was a negative effect of soil legacy of invasion on wiregrass (10% decrease) More specifically, wiregrass biomass was 15% lower when grown alone in live soils with a history of invasion than in live uninvaded soils (invasion x inoculum x competition: F 1,106 = 4.0, P = 0.047 ), indicating both abiotic and biotic effects of invasion, but only when this native grass was grown alone (T able C 3)
78 T reatments also modulated plant allocation to roots versus shoots i.e. r oot:shoot ratio (RSR) but the direction of treatment effects differed by plant species. For cogongrass, RSR was higher in live soil than sterile soil (inoculum: F 1,108 = 14.1, P = 0.0003). For pine, RSR was higher in sterile soil an effect that was larger when the pine seedlings were grown alone (inoculum x competition: F 1,101 = 12.5, P = 0.0006). For wiregrass, RSR was higher in sterile than live soil and higher in soi ls with a legacy of drought (inoculum: F 1,106 = 4. 3 P = 0.04; drought: F 1,27 = 4. 7 P = 0.04; Fig ure 4 3) but inoculum x drought was not significant for wiregrass suggesting an abiotic legacy of drought. The relative competition intensity (RCI), indicati ng the relative difference in total biomass when grown alone compared to competition, was higher for cogongrass in live soil (inoculum: F 1,36 = 25.8, P < 0.001) but higher for pine in sterile soil (inoculum: F 1,30 = 23.8, P < 0.001) and unaffected by soil inoculum for wiregrass (inoculum: F 1,3 3 = 0.05 P = 0.8 3 ; Fig ure 4 4 ). Soil legacy had no effect s on RCI. When the biomass of the other species in the competition treatment pots was taken into account using linear mixed effects models cogongrass was nega tively affected by live soil and by pine biomass but not wiregrass biomass (pine biomass: F 1,29 = 18.2, P = 0.0002; inoculum: F 1,29 = 19.6, P = 0.0001). However, pine was negatively affected by both cogongrass and wiregrass biomass and competition with co gongrass had a larger effect on pine biomass in the live soil inoculum (wiregrass biomass: F 1,29 = 45.4, P < 0.0001; cogongrass biomass x inoculum: F 1,29 = 6.6, P < 0.02). Wiregrass was only negatively affected by pine biomass and inoculum but not cogongra ss biomass (pine biomass: F 1,29 = 49.8, P < 0.0001; inoculum: F 1,29 = 21.8, P = 0.0001). We assessed total plant biomass ( sum of the three species) in the competition treatment to determine if there was an effect of soil microbes or soil legacy on total c ommunity production.
79 Biomass was 15% lower in live soil (inoculum: F 1,32 = 12.4, P = 0.001) and was a 10% lower in soils with a legacy of invasion when averaged across both live and sterile soil pots (invasion: F 1,27 = 8.4, P = 0.007 ; Figure 4 5 ). Discus sion Soil microbial communities play an important role, both directly and indirectly, in plant community dynamics and the maintenance of diversity (Van Der Heijden et al. 2008, Bever et al. 2010) In the context o f our conceptual diagram (Figure 4 1), we found evidence for pathways of interaction among native and invasive plants that were mediated by microbial communities and the biotic legacies of plant invasion but not drought We discovered that soil microbes stimulated growth bu t depressed both native species when grown alone. This pattern is consistent with the hypothesis that invader s escape from belowground enemies in the introduced range (Mitchell and Power 2003, Reinhart et al. 2003) Surprisingly, in our study, the effect of soil microbes switched from positive to negative for the invader when grown in competition with native species and switched from negative to positive for long leaf pine when grown in competition (Fig ure 4 2). Moreover, although the effect of soil microbes on wiregrass remained negative in competition, the magnitude of the effect was dampened. Therefore, our results indicate that competition between plant species can modify the interaction between plants and microbes and even change the direction of these effects (Abbott et al. 2015) Many plant soil interaction studies have assessed plant responses to soil legacies of different plant species individually but plant species in isolation are likely to have different interactions with soil microbial communities than plants in competition especially when invasive plants are present (Shannon et al. 2012, Crawford and Knight 2017) Furthermore, measured changes in the root:shoot ratio of these plants indicated that these complex effects are likely associated with differing responses of root allocation to inoculation between invasive cogongrass and the native
80 pine whereby cogongrass allocates more to roots in live soil and pine allocates less to roots in live soil (Fig ure 4 3). Addition of live i nocul um to sterilized soil could influence plant interactions with microbial communities in three major ways : introduction of pathogens that harm plants, promotion of mycorrhizal associations that usually benefits plants, and introduction of sapro trophic microbes. Plant species often show host specific responses to pathogens that can drive differences in plant fitness and may differentially affect native and non native species (Mitchell and Power 2003, Reynolds et al. 2003, Mangla et al. 2008) The effects of mycorrhizae on plant plant competition could be complex. Mycorrhizae can be involved in transfer of resources between plant species, resulting in altered com petitive dynamics (Robinson and Fitter 1999) ; they can increase a ccess to soil nutrien ts but can also act parasitically under certain conditions (Johnson et al. 1997) ; and the benefits provided by arbuscular mycorrhizae versus ectomycorrhizae differ (Smith and Read 2008) Saprophytic microbes can compete with plants for soil nutrients (immobilization) or they can convert nutrients to available forms (mineralization) dependi ng on resource ratios (Hodge et al. 2000) With this background, we can speculate on the potential causes of the switch in effect of microbes observed for the invader (co gongrass) and the native pine when grown alone compared with in competition (Figure 4 2). Soil microbes caused greater relative allocation to roots of the invader, but lower root allocation on average for both pine and wiregrass (Figure 4 3). Higher root:shoot ratios are often indicative of lower belowground resource availability or more intense belowground competition (Poorter et al. 2011) suggesting that cogongrass was more l imited by nutrients in live soil, while pine was more limited in sterile soil. Based on preliminary observation, we observed abundant mycorrhizal colonization of pine roots in the inoculation
81 treatment and minimal contamination with mycorrhizal fungi in th e sterile treatment (data not shown). We hypothesize that the ectomycorrhizal mutualism in pine in the inoculated soils allowed the species to compete effectively for nutrients with saprotrophic microbes and other plants (Figure 4 1 g i) In general, cogon grass is notorious for its high belowground allocation to rhizomes and strong competitive ability belowground, and our results indicate that soil microbes can further intensify this competitive imb alance, but may not benefit cogongrass in all situations. P revious studies have shown changes in allocation patterns in response to soil microbes ) and Te Beest et al. (2009) suggest that changes in allocation for an invader may relate to the evolution of increased competitive ability. Another possible contributing mechanism to the difference in microbial effect on root:shoot ratio for the invad er versus native species is limitation of root biomass by pathogen accumulation (de Kroon et al. 2012) Lower root allocation in live soil for either or both of the native species could potentially be caused by pathogen effects H owever, this mechanism seems unlikely for pine in competition where the effect of microbes on total biomass was positive. Alternatively, pathogens associated with longleaf pine might spillover to cogongrass when grown in competition, thereby creating a more negative effect on co gongrass and reduced competition with pine (Mordecai 2011) In addition to demonstrating a change in plant response to microbes under competition, we also showed that microbial communities can modify competitive interactions between plant species. Competiti on intensity (RCI) was higher for cogongrass in live soil but higher for pine in sterile soil, while inoculum did not affect competition intensity for wiregrass (Fig ure 4 4). Soil microbes have been shown to alter plant competitive interactions in some cas es (Hodge and Fitt er 2013, Abbott et al. 2015, Hortal et al. 2017) but the depth of understanding of these
82 interactions is limited. Furthermore, we found that pine responded to the biomass of the competing species and this response was modulated by soil microbes, while co gongrass and wiregrass responded directly to the inoculum treatment and to pine biomass but not to one another. Pine appeared to respond more to the changes in biomass of the other species in the competition treatment but not to the inoculum directly. Ther efore, as cogongrass and wiregrass were inhibited by the live soil treatment, pine experienced less competition and higher production. The presence of soil microbes appears to be essential to the competitive ability of pine and the presence of a robust mic robial community may prove to be important to the restoration of longleaf pine forests. Soil legacy of invasion and drought had surprisingly limited effects overall; however, there were a few significant effects worth noting. In the one instance of a signi ficant biotic effect of soil legacy, wiregrass was negatively affected by soil microbes from the invaded soils when grown alone, indicating that cogongrass may accumulate pathogens that have a negative effect on some native species (Mangla et al. 2008, Kelly et al. 2009, Flory and Clay 2013 ) This hypothesis is corroborated by the results in Chapter 3 where we found that invaded plots had higher relative abundance of fungal plant pathogens. As C 4 grasses, wiregrass and cogongrass are functionally more similar to each other than to pine, so cogongrass may accumulate pathogens that have more pronounced e ffect s on functionally similar species (Gilbert and Webb 2007) In Cha pter 3, w e also observed that invasion changed the community composition of AM fungi; therefore, it is possible that cogongrass legacy altered the AM fungal community causing a decrease in the benefit received by wiregrass and therefore a more negative net effect of microbes ( Figure 4 1 e f). While previous studies have observed legacy effects of drought on plant soil feedbacks (Kaisermann et al. 2017, Fry et al. 2018) we ob served no significant biotic
83 effects of drought on plant performance despite significant changes in the microbial community in response to drought in the field. In our system, microbial communities may recover rapidly upon rewetting in the greenhouse, or t he taxa that changed in abundance in the drought treatment may have minimal influence on the performance of these three plant species Additionally, i t should be noted that these are highly conservative results because we used a small volume of inoculum an d also more realistic than many studies because the initial soil legacy phase was conducted in the field (Kulmatiski and Kardol 2008, Kulmatiski et al. 2008, Heinze et al. 2016) We expected that using a small volume of soil inoculum would minimize abiotic effects of the soil legacy treatments; however, we observed some instances of significant abiotic soil legacy effects. Cogongrass growth was improved by soil legacy of drought and wiregrass root:shoot ratio was greater with soil legacy of drought. Additionally, we showed that microbial communities and invasion legacy can control plant productivity (Van Der Heijden et al. 2008) but the effect of invasion legacy appears to be at least partly abiotic as it occur red in both live and sterile soils. While the mechanism behind these abi otic effects is unclear, it is possible that allelochemicals in the invaded soil could reduce net biomass (Figure 4 1 j l) Previous research has suggested that cogongrass may release allelochemicals that reduce growth of wiregrass and pine species and inh ibit colonization by mycorrhizal fungi for both species (Hagan et al. 2013b) Many conservation eff orts are focused on habitat restoration, including the removal of invasive species and establishment of native communities; therefore, studies on the legacy effects of problematic invaders under different abiotic conditions are needed to inform management decisions for restoration (Smith Ramesh and Reynolds 2017) Our results show that soil microbes can fundamentally alter the competition betwee n native and invasi ve plants Therefore, it is critical that competitive interactions among plants be taken into account in
84 studies of microbial effects on plant performance (Shannon et al. 2012, Crawford and Knight 2017) Soil biotic and abiotic legacies had more m inor al though in some cases significant effects on relative perform ance of different plant species and total plant productivity. Hence, the microbial legacy of invasion may not be a substantial barrier to restoration of sites invaded by cogongrass ; however, degraded soil communities from management practices could inhibit restoration of longleaf pine More broadly, unders tanding the role of soil microbial communities in mediating plant competition could improve the success of restoration efforts for native ecosystems.
85 Figure 4 1. Conceptual diagram of possible belowground interactions between an i nvader and native species. Green circles represent competing plant species and brown circles represent aspects of the soil ecosystem that may mediate plant competition. Solid lines indicate direct effects and dashed lines indicate indirect effects.
86 Figur e 4 2. Relative difference in total biomass between live and sterile soil inoculum in soil with a history of native species or invasion for three species either alone or in competition (mean standard error). Points above the zero line indicate a positive effect of live inoculum compared to sterile inoculum and points below the zero line indicate a negative effect of live inoculum. Soil legac ies of ambient precipitation versus drought had no significant effects and are therefore combined in the figure. Bio tic legacy of invasion was only significant for wiregrass when grown alone. Competition and soil inoculum had a significant interactive effect on all species.
87 Figure 4 3. Relative difference in root:shoot ratio between live and sterile soil inoculu m treatments (mean standard error) Points above the zero line indicate a positive effect of live inoculum compared to sterile inoculum and points below the zero line indicate a negative effect of live inoculum. Drought had no significant effect on the d ifference between live and sterile soil and is combined in the figure. Cogongrass showed a significant response to inoculum. Pine showed a significant inoculum x competition interaction. Wiregrass showed significant inoculum and drought (not shown) effects
88 Figure 4 4 Relative competition intensity (RCI) for each species under live or sterile soil inoculum with a history of invasion or no invasion (mean standard error) RCI is calculated as the difference between the biomass when grown alone and bio mass when grown in competition divided by the biomass when grown alone so higher values indicate a greater effect of competition on plant biomass. Soil legacy of drought had no significant effect and is combined in the figure. The effect of live versus st erile inoculum was significant for cogongrass and pine but not wiregrass.
89 Figure 4 5. Effect of soil legacy of invasion and drought and live or sterile soil inoculum on total biomass production per plot in the competition treatment (mean stand ard error). Total biomass includes above and belowground biomass of three species (cogongrass, pine, and wiregrass). There was a significant effect of live versus sterile inoculum and of legacy of invasion.
90 CHAPTER 5 CONCLUSIONS The number of studies add ressing plant soil interactions has increased exponentially in recent years as the importan t role of soil biota in all aspects of ecosystem function and community structure ha s be come more apparent (Kulmatiski and Kardol 2008, van der Putten et al. 2013) Addition ally, the role of soil microb iomes in plant invasion has gained interest; however, studies of these interactions have focused on a few notable species that are known to have major impacts (Roberts and Anderson 2001) Furthermore, most studies of invader soil biota interactions assess plant performance in isolation under ideal conditions and do not consider the contex t dependency of these interactions. A s a model system, I used an ecologically understudied, but highly problematic invasive species, Imperata cylindrica, to make a comprehensive assessment of the interactive effects of invasion and drought on plant soil i nteractions (Estrada and Flory 2015) In Chapter 2, I found dramatic effects of invasion by Imperata cylindrica on plant diversity and community composition (Figure 5 1 a) Invasion reduced species richness by nearly 60% after just two years and caused the site level extinction of more than half the plant s pecies. Drought caused more moderate effects on plant diversity including a 30% reduction in richness and reordering of the dominant functional groups (Figure 5 1 b) Imperata cylindrica cover was not significantly affected by drought (Figure 5 1 c) Soil moisture was 30% higher in I. cylindrica invaded plots with the drought treatment compared to uninvaded drought plots thereby moderating the effects of drought on the plant community (Figure 5 1 d) This effect created an antagonistic interaction based on an additive model of interacting stressors (Figure 1 1) However, the major effects of invasion persisted indicating that regardless of future drought conditions, I. cylindrica will continue to be a problematic invader.
91 In contrast to the response s of th e plant community, in Chapter 3 I found that soil microbial communities overall showed greater response to drought than to invasion (Figure 5 1 e ,f ) Bacterial diversity and community composition responded strongly to drought, and most bacterial taxa resp onded to drought independently of the presence of the invader Invasion had moderate positive effects on bacterial diversity compared to the native dominated plant community (Figure 5 1 f,g). Fungal diversity w as more resilient to both treatments but comm unity composition responded interactively to invasion and drought Furthermore multiple f ungal taxa and functional groups responded interactively to invasion and drought including groups with prime significance for plants, such as AM fungi and plant patho gens. Finally, in Chapter 4, I found that t he direct effect of soil microbes on performance of the invader was positive (Figure 5 1 h), while the effect on native species was negative when grown alone (Figure 5 1 i). The striking result from this study wa s the switch in the effect of microbes on performance of the invader and the native pine when grown in competition. Microbes had a positive effect on I. cylindrica biomass when grown alone but a negative effect when grown in competition with pine and wireg rass. Pine showed the reverse effect where microbes had a negative effect when grown alone but a positive effect when grown in competition. Thus, microbes altered competitive interactions between the invader and native plants (Figure 5 1 j k ). Surprisingly the legacy effects of these changes in microbial communities resulted in only minor responses in plant performance. I showed that the indirect effect of soil biotic legacy of drought on the invader and the natives was not significant (Figure 5 1 l,m ), bu t that the biotic legacy of invasion had a small negative effect on biomass of wiregrass but not pine or the invader cogongrass (Figure 5 1 n,o ).
92 In summary, I assessed the ecological impacts of plant invasion and chronic drought on plant communities, soi l microbial communities, and how these effects feedback to performance of native versus invasive plants Overall, my results show that both invasion and drought have major consequences for native ecosystems, but they may act on different levels of the ecos ystem. Additionally, while indirect effects of soil legacies of invasion and drought may be minor, microbes play a major role in competition between native and invasive plants and should be considered in management of invasive species (Eviner and Hawkes 2008, Harris 2009, Elgersma et al. 2011) Recent studies have s uggested that reestablishing soil communities can increase the success o f ecosystem r estoration in highly degraded systems ; however, previous studies have found no benefit of soil community amendment (Harris 2009, Kardol et al. 2009, van der Bij et al. 2018) Furthermore, the r ole of plant soil interactions in restoring invaded systems i s highly context dependent and may be influenced by future climate change (Wolfe and Klironomos 2005, Harris et al. 2006, Bradley et al. 2009, Elgersma et al. 2011) Therefore, it ma y be unrealistic and potentially counterproductive to attempt to create a one size fits all recommendation for land managers across different settings (Eviner and Hawkes 2008) Here, I provide a compl ete picture of the effects of a widespread invasive plant und er current and possible future climate conditions on plant soil interactions that can inform managers dealing with restoration of I. cylindrica invaded pine forests Together, my results suggest that the direct competition from cogongrass will continue to be problematic to native communities under future precipitation regimes and the legacy of cogongrass invasion may have negative effects on restoration of wiregrass, but potentially more importantly pine requir es a robust microbial community regardless of soil legacy to compete effectively. This result sugge sts that
93 management such as top soil removal could inhibit longleaf pine restoration as mycorrhizal propagules decrease exponentially with depth (Genney et al. 2006) Furthermore, the comprehensive methodology used here could be adapted and applied to other plant invaders and alternative abiotic conditions. Figure 5 1. Synthesis d iagram of the combined results of the e xperiments in this dissertation. S olid arrows indicate direct effects while dashed arrows indicate indirect effects. Arrow widths are proportional to the size of the effect indicated. Red arrows indicate a negative effect, blue arrows indicate a positive e ffect and gray arrows indicate hypothesized interactions with no significant effect The direct effects of invasion and drought on native plants and soil microbes specifically refer to the effects on Shannon diversity index. The direct effect of drought o n the invader (specifically cogongrass) indicates change in percent cover. The direct and indirect effects of soil microbes and soil legacies on native plants and the invader refer to total biomass of longleaf pine and wiregrass, and cogongrass grown alone
94 APPENDIX A CHAPTER 2 SUPPLEMENTAL INFORMATION Experimental P lots Figure A 1 E xperimental plots at the Bivens Arm Research Site Gainesville, Florida in May 2016, showing factorial combination of treatments. Top left: u ninvaded plot with ambient precip itation. Top right: uninva ded plot with drought treatment. Bottom left: plot invaded with Imperata cylindrica and ambient precipitation. Bottom right: plot with inva sion and drought treatment s
95 Abiotic Data Figure A 2 Soil moisture and precipitation ove r the duration of the experiment. A) Monthly percent s oil moisture to 1 2 cm depth (Mean SE; N=10 ) ; B) sum of monthly precipitation measured at Gainesville Regional Airport, FL. Red line indicates 100 year average precipitation by month. Inset shows total annual precipitation by year. Black line indicates 100 year average of total annual precipitation and light gray shading indicates 1 S.D. from 100 year average.
96 Figure A 3. Percent soil moisture by depth (Mean SE). Points indicate averages across 10 replicates per treatment and across 14 sampling dates (9 26 14, 11 11 14, 11 24 14, 12 20 14, 1 21 15, 6 8 16, 6 22 16, 7 21 16, 8 22 16, 9 14 16, 9 30 16, 10 14 16, 1 6 17, 4 7 17). Soil moisture differed significantly between ambient and drought treatme nts to 40 cm depth but not at 60 or 100 cm depth (invasion x drought x depth; F 1,306 =5.0, P=0.03). Figure A 4 Percent availability of photosynthetically active radiation at ground level and 0.5 m height above the soil surface (Mean SE; N=10). Percent light availability was calculated as (ambient light light below canopy)/ambient light *100.
97 Plant Functional Groups Figure A 5. Effects of invasion and drought on plant functional groups. Percent cover of A ) perennial grasses (excluding I. cylindric a) B ) annual forbs, and C ) perennial forbs (Mean SE; N=10).
98 Tables Table A 1. List of herbaceous species planted into the plots. Species Functional group Andropogon brachystachyus grass Andropogon virginicus glaucus grass Aristida stricta gras s Eragrostis elliotti grass Eragrostis spectabilis grass Muhelenbergia capillaris grass Panicum anceps grass Carophephorus subtropicanus forb Elephantopus elatus forb Liatrus laevigata forb Pityopsis graminifolia forb Solidago fistulosa f orb Table A 2 Results of the PERMANOVA of the main and interactive effects of invasion, drought, and date on plant community composition. Df SumOfSqs Pseudo F R2 P D ate 1 3.86 31.58 0. 09 <0.01 I nvasion 1 17.74 145.24 0. 41 <0.01 D rought 1 0.59 4.85 0.0 1 <0.01 D ate : invasion 1 1.55 12.69 0.0 4 <0.01 D ate : drought 1 0.09 0.70 0.0 1 0.57 I nvasion:drought 1 0.93 7.65 0.0 2 <0.01 D ate:invasion:drought 1 0 .14 1.1 3 0.01 0.29 Residual 152 18.57 0. 43
99 Table A 3. Results of the PERMANOVA of the m ain and interactive effects of invasion and drought on plant community composition by year F values presented are pseudo F values. 2014 2015 Df SumOfSqs F P R2 SumOfSqs F P R2 Invasion 1 0.36 2.87 <0.01 0.07 2.12 8.70 <0.01 0.18 Drought 1 0. 28 2.25 0.01 0.05 0.39 1.61 0.11 0.03 Invasion:drought 1 0.11 0.85 0.49 0.02 0.49 2.00 0.04 0.04 Residual 36 4.46 8.79 2016 2017 Df SumOfSqs F* P R2 SumOfSqs F* P R2 Invasion 1 3.09 12.27 <0.01 0.24 2.63 9.25 <0.01 0.19 Drought 1 0.3 1 1.24 0.31 0.02 0.20 0.71 0.76 0.01 Invasion:drought 1 0.54 2.14 0.06 0.04 0.57 2.00 0.07 0.04 Residual 36 9.07 10.22
100 APPENDIX B CHAPTER 3 SUPPLEMENTAL INFORMATION Figures Figure B 1. Gravimetric soil moisture in ambient and drought plots either invaded by Imperata cylindrica or uninvaded. Figure B 2 Root biomass in invaded and uninvaded plots with ambient precipitation or drought at the 5 15 cm depth (mean SE). a) Fine root biomass (< 1 mm diameter), b) Coarse root and rhizome bioma ss (>1 mm diameter).
101 Tables Table B 1. Results of mixed effects model on bacterial richness, Shannon diversity index, and evenness. richness shannon evenness Treatment numDF denDF F value p value F value p value F value p value I nvasion 1 36 2.68 0.11 8.85 0.01 10.09 < 0.00 1 D rought 1 36 5.98 0.02 16.19 < 0.00 1 17.65 < 0.00 1 I nvasion:drought 1 36 2.15 0.15 0.86 0.36 0.34 0.56 Table B 2. Results of mixed effects model on fungal richness, Shannon diversity index, and evenness. ri chness shannon evenness Treatment numDF denDF F value p value F value p value F value p value I nvasion 1 36 0.06 0.81 0.62 0.44 0.75 0.39 D rought 1 36 0.42 0.52 1.16 0.29 1.15 0.29 I nvasion:drought 1 36 3.09 0.09 0.98 0.33 0.58 0.4 5 Table B 3. Results of mixed effects model on arbuscular mycorrhizal fungal richness, Shannon diversity index, and evenness. richness shannon evenness Treatment numDF denDF F value p value F value p value F value p value I nvasion 1 36 0.11 0.75 0.04 0.85 0.36 0.55 D rought 1 36 0.21 0.65 0.44 0.51 1.42 0.24 I nvasion:drought 1 36 2.57 0.12 6.35 0.02 2.77 0.11 Table B 4 Results of PERMANOVA on weighted UNIFRAC distance of the bacterial community. T reatment D f S um of sqs Pseud o F P R 2 D rought 1 0.03 2 4.251 0.002 0.10 I nvasion 1 0.01 3 1.708 0.088 0.04 Invasion: drought 1 0.009 1.246 0.261 0.03 Residual 36 0.270 Table B 5 Results of PERMANOVA on unweighted UNIFRAC distance of the bacterial community. T reatment D f S um of sqs Pseudo F P R 2 D rought 1 0.124 2.39 6 0.001 0.06 I nvasion 1 0.06 3 1.206 0.122 0.03 Invasion: drought 1 0.06 9 1.32 6 0.037 0.03 Residual 36 1.86 9
102 Table B 6 Results of PERMANOVA on Bray Curtis dissimilarity matrix of the fungal community T reatment D f S um of sqs Pseudo F P R 2 D rought 1 0.364 1.605 0.068 0.04 I nvasion 1 0.278 1.225 0.189 0.03 Invasion: drought 1 0.820 3.612 0.001 0.09 Residual 36 8.168 Table B 7 Results of PERMANOVA on Bray Curtis dissimilarity matrix of the arbuscular mycorrhizal fungal community. T reatment D f S um of sqs Pseudo F P R 2 D rought 1 0.75 1 2.639 0.001 0.06 I nvasion 1 0.45 3 1.592 0.043 0.04 Invasion: drought 1 0.41 6 1.46 1 0.085 0.04 Residual 36 10.242
103 APPENDIX C CHAPTER 4 SUPPLEMENTA L INFORMATION Figures Figure C 1. Change in total biomass of cogongrass in live or sterile soil with soil legacies of invasion x drought grown alone or in competition with pine and wiregrass. Figure C 2. Change in total biomass of longleaf pine in live or sterile soil with soil legacies of invasion x drought grown alone or in competition with wiregrass and cogongrass.
104 Figure C 3. Change in total biomass of wiregrass in live or sterile soil with soil legacies of invasion x drought grown alone or in com petition with pine and cogongrass. Tables Table C 1. Results of mixed effects of soil legacy of invasion and drought live or sterile inoculum, alone or in competition, and their interactions on total cogongrass biomass numDF denDF F value P value (Int ercept) 1 108 597.2597 < 0 .0001 Invasion 1 27 3.4757 0.0732 Drought 1 27 6.1531 0.0196 Inoc 1 108 0.7529 0.3875 Comp 1 108 2006.0846 < 0 .0001 Invasion:drought 1 27 3.8374 0.0605 Invasion:inoc 1 108 0.1311 0.718 Drought:inoc 1 108 0.0015 0.9695 Invasi on: comp 1 108 0.0433 0.8355 Drought: comp 1 108 0.7209 0.3977 Inoc: comp 1 108 25.6847 < 0 .0001 Invasion:drought:inoc 1 108 0.8142 0.3689 Invasion:drought: comp 1 108 0.6874 0.4089 Invasion:inoc: comp 1 108 0.0653 0.7988 Drought:inoc: comp 1 108 0.1521 0.6 973 Invasion:drought:inoc: comp 1 108 0.0004 0.9833
105 Table C 2. Results of mixed effects of soil legacy of invasion and drought, live or sterile inoculum, alone or in competition, and their interactions on total pine biomass. numDF denDF F value P value (Intercept) 1 101 849.4689 <.0001 Invasion 1 27 1.7429 0.1979 Drought 1 27 1.4933 0.2323 Inoc 1 101 13.9675 0.0003 Comp 1 101 343.2132 <.0001 Invasion:drought 1 27 0.9757 0.332 Invasion:inoc 1 101 3.5228 0.0634 Drought:inoc 1 101 2.2234 0.139 Inv asion: comp 1 101 1.4311 0.2344 Drought: comp 1 101 1.6539 0.2014 Inoc: comp 1 101 49.1315 <.0001 Invasion:drought:inoc 1 101 0.0292 0.8648 Invasion:drought: comp 1 101 0.0004 0.9841 Invasion:inoc: comp 1 101 0.0635 0.8016 drought:inoc:sp 1 101 0.5093 0.4 771 invasion:drought:inoc:sp 1 101 3.1502 0.0789
106 Table C 3. Results of mixed effects of soil legacy of invasion and drought, live or sterile inoculum, alone or in competition, and their interactions on total wiregrass biomass. n um D f d en D f F value P va lue (Intercept) 1 106 734.5888 < 0 .0001 Invasion 1 27 8.1932 0.008 Drought 1 27 2.2282 0.1471 Inoc 1 106 72.3774 < 0 .0001 Comp 1 106 683.5666 < 0 .0001 Invasion:drought 1 27 0.0302 0.8633 Invasion:inoc 1 106 2.1672 0.1439 Drought:inoc 1 106 1.0251 0.31 36 Invasion: comp 1 106 0.0002 0.9883 Drought: comp 1 106 0.39 0.5336 Inoc: comp 1 106 15.7356 0.0001 Invasion:drought:inoc 1 106 0.0208 0.8855 Invasion:drought: comp 1 106 0.1546 0.695 Invasion:inoc: comp 1 106 3.7407 0.0558 Drought:inoc: comp 1 106 0.57 38 0.4505 Invasion:drought:inoc: comp 1 106 1.3981 0.2397 Table C 4. List of seedlings that died by treatment and were excluded from analysis. species competition inoculum invasion drought dead count pine alone sterile uninvaded ambient 1 pine alone s terile invaded ambient 1 pine competition sterile uninvaded drought 2 pine competition sterile invaded drought 1 pine alone sterile invaded drought 2 wiregrass competition live uninvaded ambient 1 wiregrass alone sterile invaded ambient 1
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127 BIOGRAPHICAL SKETCH Catherine Fahey grew up in Ithaca New York. After graduating high school in 2006, she went on to study environmental science and international development at Cornell University and studied abroad in New Zealand and Chiapas, Mexico. She took part in research internships at Hubbard Brook E xperimental Forest in New Hampshire, Arnot Forest in New York, Amando Bermudez National Park in the Dominican Republic, and Blodgett Forest and Sequoia National Park in California where she did her honors thesis research with Dr. Robert York and Dr. Teresa Pawlowska. After earning a Bachelor of Science degree in 2010, she spent a year as a laboratory manager for Dr. Teresa Pawlowska in the Department of Plant Pathology at Cornell working on evolution of mycorrhizal fungi. She joined the Department of Biolog y at University of Florida with Dr. Kaoru Kitajima in 2011, working in the diverse tropical forests of Panama and received Department at the University of Florida to work on plant invasions in the southeast US In August 2018, she received her Ph.D. in Interdisciplinary Ecology.