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1 PLANT SOIL INTERACTIONS IN COGONGRASS ( Imperata cylindrica ) IMPACTED SOUTHERN PINE ECOSYSTEMS By DONALD LEE HAGAN 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 2012
2 2012 Donald Lee Hagan
3 To my famil y
4 ACKNOWLEDGMENTS This project would not have been possible without the support and guidance of my graduate committee: Drs. Shibu Jose, Kimberly Bohn, Francisco Escobedo, Andy Ogram and Greg MacDonald I am very thankful for the assistance provided by Vincent Morris and the staff at the Withlacoochee Forestry Center, particularly for their help with GIS, the establishment of fi eld plots and the identification of unknown species. Moshe Doron Dr. Abid Al Agely and Dr. Hee Sung Bae were tremendously helpful with the mycorrhizal analyses Dr. Chung Ho Lin provided invaluable assistance with HPLC MS/MS analyses and interpretation T he efforts of Dr. Michael Andreu, Dr. Ed Barnard, Wayne Bell, Gerardo Celis, James Colee, Nick Fray, Althea Hotaling, Bob Querns, Dr. Don Rockwood Dr. Jason Smith Dr. Ajay Sharma, Lisa Stanley and Joel Zak who assisted in various capacities, are greatly appreciated. Funding was provided by a competitive grant from the Florida Exotic Pest Plant Council (Julia Morton Invasive Plant Research Program ) as well as an Alumni Fellowship from the University of Florida (UF) and scholarships from the UF Graduate S chool and School of Forest Resources and Conservation.
5 TABLE OF CONTENTS page ACKNOWLEDGMENTS ................................ ................................ ................................ .. 4 LIST OF FIGURES ................................ ................................ ................................ .......... 9 LIST OF ABBREVIATIONS ................................ ................................ ........................... 10 ABSTRACT ................................ ................................ ................................ ................... 11 CHAPTER 1 INTRODUCTION ................................ ................................ ................................ .... 13 Plant Invasions ................................ ................................ ................................ ....... 13 Invasive Alien Plants and Soil Properties ................................ ................................ 14 The Legacy of Invasion ................................ ................................ ........................... 17 A Case for Cogongrass ................................ ................................ ........................... 18 Objectives and Hypotheses ................................ ................................ .................... 20 2 NOVEL RHIZOSPHERE CHEMI STRY OF COGONGRASS: IMPLICATIONS FOR THE PERFORMANCE OF NATIVE PINE SAVANNA SPECIES IN THE SOUTHEASTERN US ................................ ................................ ............................ 24 Background ................................ ................................ ................................ ............. 24 Mate rials and Methods ................................ ................................ ............................ 27 Greenhouse Study ................................ ................................ ........................... 27 Isolation and Characterization of Putative Allelochemicals ............................... 30 Statistical Analysis ................................ ................................ ............................ 31 Results ................................ ................................ ................................ .................... 32 Biomass Production, Allocation and Root Morphology ................................ ..... 32 Mycorrhizal Inoculation and Infected Root Length ................................ ............ 33 Chemical Profiling of Leachates ................................ ................................ ....... 33 Discussion ................................ ................................ ................................ .............. 34 Summary and Implications ................................ ................................ ...................... 40 3 COGONGRASS INVASION AND ERADICATION: IMPLICATIONS FOR SOIL BI OGEOCHEMICAL PROPERTIES IN A FIRE MAINTAINED FOREST ECOSYSTEM ................................ ................................ ................................ ......... 47 Background ................................ ................................ ................................ ............. 47 Materials and Methods ................................ ................................ ............................ 49 Study Area ................................ ................................ ................................ ........ 49 Experimental Design ................................ ................................ ........................ 50 Soil Chemistry and Nutrient Pools ................................ ................................ .... 52 Soil Nutrient Availability ................................ ................................ .................... 52
6 Litter Decomposition and Nutrient Mineralization ................................ ............. 53 AM Fungal Spore Quantification ................................ ................................ ....... 53 Soil AM Fungal DNA Extraction, PCR, Cloning and Sequencing ..................... 54 Sequence Processing and Anal ysis ................................ ................................ 55 Statistical Analysis ................................ ................................ ............................ 56 Results ................................ ................................ ................................ .................... 57 Organic Matter and pH ................................ ................................ ..................... 57 Nitrogen ................................ ................................ ................................ ............ 58 Phosphorus ................................ ................................ ................................ ...... 58 Tissue Quality, Decomposition and Mineralization ................................ ........... 59 Arbuscular Mycorrhizal Fungi ................................ ................................ ........... 59 Discussion ................................ ................................ ................................ .............. 60 Influence of Cogongrass on Soil Chemistry ................................ ...................... 61 The Post Eradication Legacy of Cogongrass on N and P Cycling .................... 63 Arbuscular Mycorrhi zal Fungal Dynamics ................................ ........................ 65 Summary and Implications ................................ ................................ ...................... 66 4 PATTERNS OF SECONDARY SUCCESSION FOLLOWING COGONGRASS ERADICATION IN A LONG LEAF PINE SANDHILL ECOSYSTEM ........................ 80 Background ................................ ................................ ................................ ............. 80 Materials and Methods ................................ ................................ ............................ 83 Study Area ................................ ................................ ................................ ........ 83 Experimental Design ................................ ................................ ........................ 84 Sampling Protocol ................................ ................................ ............................ 85 Results ................................ ................................ ................................ .................... 87 Cover, Richness and Diversity ................................ ................................ ......... 87 Relative Groundcover by Growth Habit and Persistence ................................ 88 Dominant Species ................................ ................................ ............................ 88 Multivariate Analyses ................................ ................................ ....................... 89 Longleaf Pine Regeneration ................................ ................................ ............. 90 Non native Species ................................ ................................ ........................... 90 Discussion ................................ ................................ ................................ .............. 91 Understory Community Assembly ................................ ................................ .... 91 Patterns and Environmental Covariates of Species Colonization ..................... 92 Summary and Implications ................................ ................................ ...................... 94 5 CONCLUSIONS ................................ ................................ ................................ ... 104 APPENDIX A RELATIVE ABUNDANCE OF EACH OF THE 31 AM FUNGAL OTUS, BY TREATMENT ................................ ................................ ................................ ........ 109 B SPECIES L IST ................................ ................................ ................................ ...... 110 LIST OF REFERENCES ................................ ................................ ............................. 114
7 BIOGRAPHICAL SKETCH ................................ ................................ .......................... 127
8 LIST OF TABLES Table page 2 1 Native pine savanna species used in a study of the allelopathic effects of cogongrass le achate ................................ ................................ .......................... 41 2 2 Effects of aqueous and chloroform extracts of leachates (native, cogongrass, DI water control) on the germination (%) of lettuce seeds.. ................................ 42 2 3 Species wise comparisons of the effects of a cogongrass leachate treatment vs. the effects of a native leachate treatment. ................................ ................ 43 2 4 Mean chemical composition o f leachates collected from the rhizosphere of greenhouse grown cogongrass monocultures and native polycultures ...... 44 3 1 Number of study plots in each in each treatment x block x replication ............... 68 3 2 Initial mean tissue chemistry of herbicide treated cogongrass rhizomes and foliage, along with calculated k coefficients for mass loss and N and P mineralization ................................ ................................ ................................ ..... 69 3 3 Pairwise treatment comparisons of AM fungal community assembly ................. 70 4 1 Number of study plots in each in each treatment x block x replication ............... 96 4 2 Landscape and soil variables for a canonical discriminant analysis (CDA), used along with the relative covers of the 23 most dominant plant species in reference plots ................................ ................................ ................................ .... 97 4 3 Mean values for the Shannon Wiener ( D ) Indices, species richness and total plant cover ................................ ................................ 98 4 4 Mean relative cover of the 23 most prevalent understory species in reference plots, compared to their relative covers in plots where cogongrass was eradicated three, five and seven years prior ................................ ....................... 99 4 5 Top 5 dominan t species in plots where cogongrass was eradicated three five and seven years prio r ................................ ................................ ....................... 100 A 1 Relative abundance of each of the 31 arbuscular mycorrhizal fungal OT Us ... 109 B 1 Complete list of all plant species identified ................................ ....................... 110
9 LIST OF FIGURES Figure page 1 1 Map of the curre nt distribution of cogongrass in the southeastern US ............... 23 2 1 Difference in percent mycorrhizal colonization and total mycorrhizal root length for four native species watered with cogongrass le achate, relative to those watered with leachate from native species ................................ ............... 45 2 2 Speculated chemical structure of a novel alkaloid identified in cogongrass leachate, along with ion chromatography ................................ ........................... 46 3 1 S chematic diagram of a replicate ................................ ................................ ....... 71 3 2 Mean soil organic matter content (%) ................................ ................................ 72 3 3 Mean water extractable soil pH ................................ ................................ .......... 73 3 4 Mean total soil nitrogen (TKN method), resin adsorbed soil ammonium and nitrite+nitrate ................................ ................................ ................................ ....... 74 3 5 Mean Mehlich 1 extractable phosphorus and resin extracted soil phosphorus .. 75 3 6 Patterns of mass loss and N and P mobilization/immobilization for foliage and rhizomes of cogongrass treated with glyphosate and imazapyr herbicides ........ 76 3 7 Summary statistics (Chao1 richness, Shannon index, 1 arbuscular mycorrhizal fungal comm unities ................................ ........................ 77 3 8 Phylogenetic tree of arbuscular mycorrhizal fungal SSU rRNA genes ............... 78 3 9 Weighted UniFrac PCA biplot ................................ ................................ ............. 79 4 1 Mean relative cover (%) by growth habit type in plots where cogongrass was eradicated three, five and seven years prior, along with uninvaded native reference plots ................................ ................................ ................................ .. 101 4 2 Canonical discriminant analysis (CDA) biplot s of the patterns of compositional similarity between native reference plots and plots where cogongrass was eradicated three, five and seven years prior .......................... 102 4 3 Longleaf pine stems per m 2 in plots where cogongrass was eradicated three, five and seven years prior, along with uninvaded native reference plots 103
10 LIST OF ABBREVIATIONS ACN Acetonitrile ACU Acaulosporaceae AM Arbuscular mycorrhizal APCI Atmospheric pressure chemical ionization BA Basal area BLAST Basic Local Alignment Search Tool CDA Canonical Discriminant Analysis EcM Ectomycorrhizal EM Ericoid mycorrhiz al GIG Gigasporaceae GIS Geographic Information Systems GLO Glomeraceae HPLC High performance liquid chromatography IA Invasive alien M1 Mehlich 1 extractable MS Mass spectrometry OM Organic matter OTU Operational taxonomic unit PAR Paraglomeraceae PCR Pol ymerase Chain Reaction PPM Parts per million SSU rRNA Small subunit ribosomal RNA (18S) TKN Total Kjeldahl Nitrogen
11 Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requi rements for the Degree of Doctor of Philosophy PLANT SOIL INTERACTIONS IN COGONGRASS ( Imperata cylindrica ) IMPACTED SOUTHERN PINE ECOSYSTEMS By Donald Lee Hagan August 2012 Chair: Shibu Jose Cochair: Francisco Escobedo Major: Forest Resources and Conse rvation The objective of this project was to assess how cogongrass ( Imperata cylindrica ) affects soil and ecosystem processes in southern pine ecosystems. In a greenhouse study (Chapt er 2), I evaluated whether cogongrass impedes native pine savanna species through the release of allelopathic compounds. In a field study (Chapter 3), I assessed pre and post eradication nitrogen, phosphorus and arbuscular mycorrhizal fungal dynamics in p ine sandhill stands severely impacted by cogongrass. In another field study (Chapter 4) I describe d the patterns (and potential drivers) of secondary succession following cogongrass eradication in these same stands. There was an allelopathic effect of co gongrass, although it varied by species. A ruderal grass and an ericaceous shrub were unaffected by cogongrass soil leachate, while a mid successional grass and pine were negatively affected. C hemical analyses revealed 12 putative allelopathic compounds i ncluding a novel alkaloid in cogongrass leachate. The concentrations of most of these compounds were significantly lower in the native leachate. Compared to a native reference treatment, c ogongrass invasion had no
12 effect on soil chemical properties, altho ugh significant but temporary changes (increases in pH and available nitrate, decreases in available phosphorus) occurred post eradication. While invasion resulted in the development of a novel arbuscular mycorrhizal (AM) fungal community, AM fungal commun ity structure returned to a reference state within five years. Displaced native plant communities, however, were slow er to recover following cogongrass eradication. S imilar levels of plant species richness and diversity were observed by year seven but com position remained markedly different from reference S oil properties ( e.g. organic matter, mycorrhizal spore counts and pH) covaried with successional patterns. These findings provide insight into the ecology of southern pine ecosystems impacted by cogon grass. Differences in leachate chemistry between cogongrass and native species may imply that the competitive ability of cogongrass is augmented by reference state re latively quickly following eradication is encouraging. The recovery of soil properties before native plant communities suggests that belowground processes and/or dispersal limitations may influence ecological succession following eradication.
13 CHAPTER 1 INTRODUCTION Plant Invasions overcome barriers to long distance dispersal and are able to persist, reproduce and spread in new areas (Richardson et al. 2000). Whi le only a small subset of introduced plant taxa meet these criteria, these (relatively) few species greatly threaten the productivity and functioning of terrestrial ecosystems (Simberloff 2005). In the United States, some 5000 IA plant species have become established, with spread rates into forests, grasslands and other natural areas estimated at 700,000 hectares per year (Pimentel et al. 2005). In natural systems, alien plant invasions can cause dramatic shifts in plant community assembly, with their succe ss often occurring at the expense of diverse assemblages of native species. For these reasons, it has frequently been reported that IA species (including plants) have become a leading cause of biodiversity loss, second only to habitat destruction (Simberlo ff 2005). Consequently, the desire among scientists to understand and predict these transformative effects has led to intense speculation on the underlying drivers and mechanisms of successful plant invasions. What are the causes and effects of plant inva sions, and why do some alien plants become invasive and others do not? These are the primary questions that have motivated IA plant research over the last two decades. Williamson and Fitter (1996), in an effort to develop a predictive framework for biotic which states that 1/10 of all introduced alien species escape, 1/10 of those that escape become established and 1/10 of those that become established become invasive. While
14 this is more of a generalization than an actua l scientific rule, it illustrates that invasion is a multistep process, with many barriers that must be overcome in order for an alien plant to become invasive. Moreover, it implies that interactions between an alien plant and its environment are a major d eterminant of whether or not it establishes and becomes invasive. Unfortunately most studies of these interactions have focused on the primary producers ( i.e. the plants themselves), typically viewing invasion as the end result of a plant plant interaction in which an introduced alien species successfully outcompetes established natives to become invasive. Furthermore, most have been aboveground centric, with few researchers attempting to elucidate the complex suite of interspecific interactions that take p lace in the rhizosphere (Wolfe and Klironomos 2005) As more studies are conducted, the role of belowground processes in invaded systems is becoming clearer, as are the changes to the soil communit y that occur following invasion and the implications they have for ecological succession ( Ehrenfeld and Scott 2001; Bais et al. 2003; Ehrenfeld 2003; Callaway and Ridenour 2004; Wolfe and Klironomos 20 05 ) Invasive Alien Plants and Soil Properties A prima ry way that IA plants alter soil properties by differing f rom natives in the quantity and/or quality of biomass that they produce (Ehrenfeld et al. 2001; Ehrenfeld 2003). Since the carbon cycle is intrinsically linked to other element cycles, these changes can greatly impact the availability of soil nutrients p articularly macronutrients such as nitrogen and phosphorus which are frequently limiting ( Vitousek et al. 1987; Ehrenfeld 20 03 ). While invasive plants may or may not produce more litter than natives, most studies have found these inputs to be of higher qua lity (lower C:N, lignin:N and C:P ratios) This, in turn, may result in net mineralization, faster turnover rates, altered
15 nutrient pools and an increase in nutrient availability. The opposite trend, however, has also been observed (Ehrenfeld 2003 and citat ions therein). By differing from natives in terms of belowground architecture and nutrient uptake patterns, invasive species may also affect the distribution of mineral nutrients in the soil profile. An IA plant with a deep and/or highly prolific root syst em, for example, may act as a nutrient mine effectively capturing nutrients in and depositing them at or near the soil surface as litterfall (Lambers et al. 2008; Perkins et al. 2011). The loss of a deeply rooted native species in favor of an invasive, however, would have the opposite effect. The mechanisms behind altered nutrient cycling by invasives are not strictly limited to biomass production and litter quality. Invasives have also been shown to alter soil pH (Ehrenfeld 2003 and citations therein), which, aside from affecting the solubility of soil organic matter, has implications for nitrification, NH 4 volatilization and phosphorus complexation reactions (Brady and Weil 2002). There does not, however, appear to be a characteristic trend of pH altera tion, as decreases as well as increases have been reported (Ehrenfeld 2003 and citations therein). While changes in litter quality may contribute to alterations in soil pH, diffe rences in exudate chemistry ( Bais et al. 2006), altered nitrification rates (E hrenfeld et al. 2001 ) and differential uptake of nitrate vs. ammonium may also be factors (Ehrenfeld 2003). Invasive plants can also alter the nutrient dynamics of an ecosystem indirectly through their effects on mycorrhizal communities (Pringle et al. 20 09). Since many invasive plants form only weak associations with mycorrhizae, the density and thus the efficacy of these important mutualists may decline following invasion (Vogelsang and Bever 2009). This reduces the competitive ability of native spec ies, and due to the
16 differences described above (morphology, tissue chemistry, etc .), leads to the alteration of nutrient cycling processes (Pringle et al. 2009) A similar pattern likely occurs when invasion results in a change in the functional group com position of a plant community (Pringle et al. 2009). The replacement of a woody species by a grass, for example, may result in a shift in mycorrhizal community structure to favor arbuscular mycorrhizae over ectomycorrhizae (Vosatka et al. 1991). Difference s in root architecture between natives and invasives, coupled with differences in nutrient uptake efficiency between different types of mycorrhizae (Jones et al. 1998), may in turn affect the nutrient dynamics of an ecosystem. The effects of altered nutri ent cycling regimes are often exacerbated by the fact that invasive plant species often establish dense monocultures. This is an interesting phenomenon considering that the same species in their native habitat typically coexist with other species (Callawa y and Aschehoug 2000). This suggests that certain invasives are not only superior competitors, but are also capable of using additional mechanisms ( i.e. ) that exploit the lack of co evolved tolerances among natives (Hierro and Callaway 2003 ). Allelopathy, the inhibition of one plant by another by the release of phytotoxic compounds ( i.e. allelochemicals), has been suggested as such a mechanism (Hierro and Callaway 2003; Callaway and Ridenour 2004). Allelochemicals include a diverse array of secondary metabolites and can be released in various forms, including root exudates, and litter, bark and seed leachates. Some of these chemicals rapidly volatize or degrade, while others may persist in the soil (Reigosa et al. 1999). Alone or in combinati on, these substances can inhibit seed germination and root elongation (Hierro and Callaway 2003) and in some cases lead to
17 the partial or complete death of the root systems of susceptible plants (Bais et al. 2003). Many allelochemicals also have microbicid al properties, which suggests that they might impede the formation and/or efficacy of important symbioses and associations, such as those involving symbiotic nitrogen fixing bacteria and mycorrhizal fungi (Wardle et al. Klett et al. 2007). This could be an additional disadvantage for native species, especially when phosphorus which tends to be poorly mobile in soils is the limiting resource (Smith and Read 1997). The Legacy of Invasion The body of k nowledge on the effects of invasive species on belowground processes while limited, has increased greatly in recent years. Comparatively less attention, however, has been paid to the legacies that invasives leave behind once they have been eradicated. Ind eed only a handful of studies have incorporated the eradication of an IA species and subsequent monitoring of nutrient cycling processes (Maron and Jeffries 2001; Yelenik et al. 2004). This, however, is an area that deserves more consideration, as the rest oration of native plant communities following the eradication of invasives is a high priority among land managers (Miller et al. 2010). Much like any other disturbance, if soil processes and properties are altered by an invader, these effects will likely p ersist for some time after the invader is eradicated subsequent alteration of immobilization/mineralization processes, may further alter soil biogeochemistry following the eradication of the invasive. These changes in turn, may have implications for the invasibility of the new community as well as its suitability for revegetation with native plant species.
18 A Case for Cogongrass Cogongrass ( Imperata cylindrica ( L ) P. Beauv ) is a rapidly growing C 4 perennial grass that readily invades natural ecosystems and disturbed sites. With invasions problematic invasive plant species. In total, some 5 00 million hectares worldwide have some degree of cogongrass infestation ( MacDonald 2004 ) In the US, several hundred thousand hectares are infested (MacDonald 2004), with its current range overlapping much of the historic range of longleaf ( Pinus palustris Mill.) and slash pine ( Pinus elliottii Engelm) (Figure 1 1) The sparse canopy that is characteristic of these forests, in concert with frequent fire, allows for high levels of understory diversity, but also makes them very susceptible to transformative i mpacts from cogongrass (Holzmueller and Jose 2011). Since cogongrass is becoming a significant problem in forest systems of the Southeast its ecology and management have been the subjects of considerable research interest among forest ecologists in recent years. Perhaps the most dramatic characteristic of cogongrass invasion is the density of the resultant monoculture and the amount of biomass produced. This creates significant pressure not only for space, but also for soil resources. According to Ramsey et al. (2003), cogongrass produces over three times more foliar biomass and up to ten times more root/rhizome biomass than native vegetation growing on the same site. Fresh weights of up to 10 metric tons/ha for shoots and up to 40 metric tons/ha for rhizo mes have been reported in some sites (MacDonald 2004 and citations therein ). Tissue quality is also an important consideration. Daneshgar and Jose (2009 a ), for example, found cogongrass to be very effective at competing for soil nitrogen, but since it prod uced so much biomass, its tissue nitrogen concentrations were considerably lower
19 than those of native vegetation. Secondary organic compounds (Koger and Bryson 2004) and silica crystals ( MacDonald 2004) in cogongrass tissue may also reduce its palatability to herbivores and soil microbes. Combined, these factors suggest that cogongrass invasion results in the production of recalcitrant nutrient pools, likely leading to nutrient immobilization, decreased nutrient availability and reduced nutrient pool turnov er rates. Cogongrass has been observed to alter soil chemistry in forest ecosystems. Collins and Jose (2008), for example, observed seasonal reductions in extractable NO 3 N and K, increases in Mg and decreases in pH in cogongrass invaded pine sites, compa red to non invaded sites in the same forests. No significant differences in organic matter, P, or Ca, however, were observed between invaded and uninvaded sites. Despite these observations, however, our understanding of the effects of cogongrass invasion o n soil chemistry is far from complete. The effects of invasion on the overall nitrogen cycling in a system (not simply NO 3 N), for example, deserve consideration in acidic forest soils where nitrification may be inhibited (Chapin et al. 2002). No studies h ave evaluated whether or not nutrient dynamics in cogongrass invaded forest ecosystems return to pre invasion conditions following eradication. Cogongrass is known to form associations with AM fungi (Brook 1989), which undoubtedly contributes to its super ior competitive ability in nutrient poor soils. The formation of a cogongrass monoculture, therefore, likely results in a decrease in non AM fungal propagule density in the soil ( i.e. ericoid and ectomycorrhizal fungi) (Korb et al. 2003). This in turn may magnify the selective pressure against obligate non AM plant species. Chemical eradication of cogongrass, which typically involves the use of one or
20 more systemic herbicides ( MacDonald 2004), may depress AM fungal density as well, effectively killing the s ymbiont by eliminating its host Since most plants growing in low pH, phosphorus fixing forest soils are highly dependent on mycorrhizal symbioses (Smith and Read 1997), the recovery of the mycorrhizal community likely plays a key role in the reestablishme nt of the desired plant species following cogongrass eradication. Several authors have suggested that the competitive ability of cogongrass is augmented by the production of allelopathic compounds. Putative allelochemicals (mostly phenolics) have been extr acted from cogongrass tissues and from soils in the vicinity of cogongrass patches (Abdul Wahab and Al Naib 1972; Hussain and Abidi 1991; Inderjit and Dakshini 1991, Xuan et al. 2009) and some of these compounds have been shown to have inhibitory effects o n test plants (Koger and Bryson 2004, Xuan et al. 2009). The current body of research on cogongrass allelopathy in natural systems, however, should be considered inconclusive, as single compound bioassays on weed and crop species may not be an accurate rep resentation of the complex interaction between live plants that occurs in nature (Mallik 2000). No studies to date have assessed the effects of cogongrass allelopathy on the performance of native understory species like those that it readily displaces. Obj ectives and Hypotheses I conducted these studies to elucidate the role of belowground processes in southern pine ecosystems impacted by cogongrass and to describe the patterns of secondary succession following cogongrass eradication. The specific objective s were to:
21 Assess whether or not allelopathic compounds are present in biologically significant concentrations in the cogongrass rhizosphere and to determine the effects of these compounds on a suite of species native to southeastern pine savannas. Analyz e how invasion by cogongrass affects soil N and P dynamics and arbuscular mycorrhizal fungal communities in fire maintained longleaf pine sandhill stands. Quantify soil N and P dynamics and assess changes in arbuscular mycorrhizal community assembly in th e years following cogongrass eradication. Describe the patterns of secondary succession following the eradication of cogongrass in a longleaf pine sandhill ecosystem. I hypothesized that rhizosphere water collected from cogongrass invaded soils would adve rsely affect the growth, root morphology and mycorrhizal colonization of native species. Additionally, I expected that compounds present in the cogongrass rhizosphere would not be present in the rhizospheres of native plants or they would be present at mu ch lower concentrations Cogongrass invasion was expected to decrease the availability of soil N and P, likely through reductions in pH and/or changes to the soil carbon cycle (Brady and Weil 2002). I expected these changes in N and P cycling to persist fo llowing eradication, perhaps due to the slow decomposition rates of low quality cogongrass foliage and rhizomes after herbicide treatment (Ehrenfeld 2003). I expected that cogongrass invasion would result in the development of a novel AM fungal community, and additional modifications to AM fungal community structure would arise following eradication. I hypothesized that formerly invaded sites would, by year seven begin to regain many of the characteristics of native reference sites. Specifically, I expecte d to see increases in total plant cover, increases in species richness and diversity, decreases in dominance and increases in the relative cover of desirable native species such as wiregrass ( Aristida stricta Michx. var. beyrichiana Ward). Shifts in
22 commun ity assembly, I hypothesized, would be associated with changes in soil resource availability and alterations to arbuscular mycorrhizal fungal community structure. I expected that the elimination of cogongrass and other competing vegetation would facilitate the establishment of longleaf pine seedlings, but would also lead to a secondary invasion of alien plant species, particularly fast growing ruderals that are readily able to take advantage to a post eradication resource flux.
23 Figure 1 1 Map of the current distribution of cogongrass in the southeastern US. Adapted from EDDmapS (2012)
24 CHAPTER 2 NOVEL RHIZOSPHERE CH EMISTRY OF COGONGRAS S: IMPLICATIONS FOR THE PERFORMANCE OF NATIV E PINE SAVANNA SPECI ES IN THE SOUTHEASTERN US Background Numerous theories and hypotheses have been proposed to explain the success of IA plant species in their new environments, the majority of which are based on resource competitive ability (Bakker and Wilson 2001). Indeed, there is ample evidence to suggest that resource based mechanisms such as competition play a major role in successful plant invasions. Many IA plant species, for example, grow fast and are highly efficient in the uptake, use and allocation of limiting resources (Daehler 2003) qualities which undoubtedly help explain how they are able to cause such dramatic alterations to the community assembly of the sites they invade. Th e inherent competitive ability of these species is also likely augmented by the release from co evolved specialist enemies (Maron and Vil 2001) and in some cases may be an evolutionary response in which alien species, over time, allocate less photosynthat e to defense and more to growth and reproduction (Blossey and Notzold 1995; Hnfling and Kollmann 2002). While competition for resources likely plays a major role in most alien plant invasions, resource based mechanisms alone may not adequately explain th e success of some IA plant species (Hierro and Callaway 2003). The propensity of certain plant species to form dense monotypic stands, for example, suggests that additional interactions may also be involved (Hierro and Callaway 2003). Allelopathy, the inhi bition of one plant by another via the release of phytoinhibitory chemical compounds ( i.e. allelochemicals), is one such mechanism (Hierro and Callaway 2003; Callaway and
25 Ridenour 2004). Allelochemicals include a diverse array of secondary metabolites and can be released in various forms, although root exudates and litter probably constitute the primary sources (Wardle et al. 1998). Some of these chemicals rapidly volatize or degrade, while others may persist in the soil (Reigosa et al. 1999). The breakdown products of exuded allelochemicals often retain some bioactivity (Blum 1998; Blum et. al. 2000). Alone or in combination, these substances can inhibit seed germination and root elongation (Hierro and Callaway 2003) and in some cases lead to the partial or complete death of the root systems of susceptible plants (Bais et al. 2003). Many allelochemicals also have microbicidal properties, which suggests that they might impede the formation and/or efficacy of important symbioses and associations, such as those involving symbiotic nitrogen fixing bacteria and mycorrhizal fungi (Wardle et al. the mycorrhizal symbiosis (Frey Klett et al. 2007). The naivety of native communiti es to the novel allelochemicals produced by IA plants may make them particularly susceptible to transformative impacts ( Callaway and Ridenour 2004). While there is some compelling anecdotal evidence for the role of allelopa thy in invaded natural systems, our ability to empirically assess the role of allelopathy in plant plant interactions has been hindered by some major methodological limitations (Mallik 2000; Hierro and Callaway 2003). Conclusions about allelopathy drawn ex clusively from P etri dish bioassays, in which seedlings or seeds are watered with a leachate extracted artificially from dead plant tissues (Richardson and Williamson 1988; Hierro and Callaway 2003; Gmez Aparicio and Canham 2008), or pot studies in which plant
26 residues are incorporated into the soil medium (Singh et al. 2005; Norsworthy 2003), should be interpreted with some skepticism, as the chemical composition of these materials may be qualitatively or quantitatively different from the allelochemicals exuded by live plants and their litter. Additionally, biologically significant concentrations of allelopathic compounds have rarely been isolated from the rhizosphere soil of IA plants. Due to limitations such as these, evidence of allelopathic interferenc e in most cases cannot be considered conclusive. With invasions reported on six continents, cogongrass ( Imperata cylindrica ( L ) P. Beauv.) total, some 500 million hectares wor ldwide have some degree of cogongrass infestation (MacDonald 2004), with dense monotypic stands widely reported in tropical and subtropical forests, savannas, grasslands, pastures and agricultural fields (MacDonald 2004). In the southeastern US, cogongrass has been observed to dramatically alter the species and functional composition of native pine ( Pinus spp.) ecosystems by displacing native groundcover species (Jose et al, 2002; Collins et al. 2007) and inhibiting the performance of sapling trees (Daneshg ar and Jose 2009 a ; Holzmueller and Jose 2011). The tremendous success of cogongrass in its expanded range has been attributed, in part, to a suspected allelopathic ability (Koger and Bryson 2004; MacDonald 2004) and several putative allelopathic compounds have been isolated from cogongrass tissues and from soils in the vicinity of cogongrass patches (Abdul Wahab and Al Naib 1972; Hussain and Abidi 1991; Inderjit and Dakshini 1991, Xuan et al. 2009). Some of these compounds have been shown to have inhibitory effects on agricultural species (including weeds) (Koger and Bryson 2004) and other IA plants (Xuan et al. 2009). To
27 date, however, no studies have assessed the effects of cogongrass allelopathy on native wildland plant species. The reliance on phytotoxic compounds artificially extracted from cogongrass tissues, rather than exudates and their breakdown products, is also a limitation of previous research. I conducted this study to assess whether or not allelopathic compounds are present in biologically sig nificant concentrations in the cogongrass rhizosphere and to determine the effects of these compounds on a suite of plant species native to southeastern pine savannas. I hypothesized that rhizosphere water collected from cogongrass invaded soils would adve rsely affect the growth, root morphology and mycorrhizal colonization of native species. Additionally, I hypothesized that compounds present in the cogongrass r hizosphere would not be present in the rhizospheres of native plants, or they would be present a t much lower concentrations. Materials and Methods Greenhouse S tudy For this study, I employed a greenhouse protocol, in which seedlings of four native species were irrigated collected fro m pot grown monocultures of cogongrass or from polycultures of native species. The latter treatment, while not a true control, was treated as such since it consisted of conspecifics, congenerics and functionally similar species that naturally co occur and compete with native test species in pine savannas in the southeastern US. A DI water control was used in the second season to verify that any treatment effects were due to a negative influence of cogongrass, rather than a facilitative effect from the nativ e species. The four test species (Table 2 1) included an arbuscular mycorrhizal (AM) ruderal grass ( Andropogon arctatus Chapm.), an AM mid successional grass
28 ( Aristida stricta Michx. var beyrichiana (Trin. and Rupr.) D.B.Ward ) an ericoid mycorrhizal (EM) shrub ( Lyonia ferruginea (Walter) Nutt) and an ectomycorrhizal ( EcM ) tree ( Pinus elliottii Engelm.). Predominant species in the native polycultures, in order of decreasing cover, were A. stricta Andropogon virginicus L., Vaccinium myrsinites Lam., Gaylus sacia frondosa (L.) Torr. and A. Gray, Gaylussacia dumosa (Andrews) Torr. and A. Gray, Pinus elliottii and Smilax spp. Three 11.4 liter pots for each of the two leachate treatments were established in March 2009 and 2010 with vegetative plugs and rhizomes obtained from local sources. Native seedlings were planted from surface sterilized pre germinated seeds in 200 mL Ray Leach tubes (Stuewe and Sons, Tangent Oregon). Seedlings were planted in two cohorts, with A. arctatus and P. elliottii established in Ju ne of each year, and A. stricta and L. ferruginea established in August. For both cohorts, each of three plots contained 20 tubes ( five for each species x leachate treatment). Due to events beyond my control I was unable to successfully produce L. ferrugin ea seedlings in 2009. All plants, both in the leachate pots and in the seedling tubes, were grown in a Sparr fine sand (loamy, siliceous, subactive, hyperthermic Grossarenic Paleudult), one of the predominant soil series in north central Florida (United St ates Department of Agriculture 1985). The soil collection site was heavily vegetated with native AM, EM and EcM plant species and thus the planting medium was assumed to contain a diversity of compatible mycorrhizal inoculum. Soils were thoroughly homogeni zed prior to filling pots and tubes. For each species, t here were three replications (blocks) each containing five seedling tubes from each of the two leachate treatments In the second year I included an additional five seedling tubes to account for the DI water control. Tubes were arranged
29 in strips, randomly assigned by treatment, to prevent cross contamination while watering Twice weekly, one leachate pot from each treatment was watered with 1 .3 .L of distilled water, which allowed me to collect appro ximately 1 L of raw leachate from the bottom of each pot. The other four pots were watered to field capacity (approximately 300 mL). Pots were rotated so that, under this procedure, leachate was collected from each at approximately 10 day intervals. Fresh leachates were filtered twice through Whatman # 5 filter to remove debris, fungal spores and sporocarps. The first collection and application of leachate was timed to correspond with the planting of the first seedling cohort (early June). Each seedling tube received approximately 15 mL of filtered leachate from either the native or the cogongrass treatment (or DI water control). Seedlings were harvested after 8 weeks. Upon harvest, plants were separated into their above and belowground components. Dry weig hts for shoots were obtained after drying them for 48 hours at 70C. Since fresh roots were needed for mycorrhization and root length analyses (see below), root dry weights were calculated by multiplying fresh weight by a weight conversion factor, which wa s determined by drying three seedlings not included in the analyses for each species x treatment combination. Root lengths were determined using the modified line intercept method described by Tennant (1975). Roots were prepped for mycorrhizal analysis usi ng standard clearing and staining procedures (Manoharachary and Kunwar 2002) and analyzed for mycorrhizal colonization. For AM and EM species (grasses and L. ferruginea ) root segments were assessed on each sample for the presence or absence fungal coloniza tion as well as the degree of
30 colonization ( i.e. light: 0 33%, moderate: 33 67% and heavy: 67 100%). Percent mycorrhizal colonization was quantified for each sample as the mean of 10 microscope fields (10X magnification) with each field assigned eith er 0 or the midpoint of each colonization class ( i.e. 16%, 50% or 84%). For EcM P. elliottii the line intercept method (Tenant 1975) was used to determine percent mycorrhizal colonization, using a dissecting microscope. Isolation and C h aracterization of P utative A llelochemicals After observing evidence of bioactivity in raw cogongrass leachate, steps were taken to identify active compounds and determine their concentrations. Polar fractions were separated from n on polar fractions using a chloroform extraction method and a standard lettuce seed bioassay was used to determine the bioactivity of each fraction (4 replicates of 10 per treatment, plus control) (Table 2 2). Active (polar) fractions in cogongrass leachat e, along with polar fractions of native leachate, were then concentrated in an N 2 vortex evaporator. Samples were analyzed by injecting 25 L of the concentrated extract into a Shimadzu SCL 10Avp high performance liquid chromatography system (HPLC) (Columb ia, MD), compounds were separated using a silica based Columbus C8 column (4.6 mm x 250 mm, 5 m; Phenomenex, Torrance, CA) and eluted with a two 1 Mobile phase A consisted of 0.1% H 3 PO 4 buffer (pH =2. 1) and mobile phase B was 100% ACN. The gradient started at 10% A, ramped linearly to 40% A at 30 min, 75% A at 40 min, 10 % A at 45 min, and was held at 10% for 14 min. The chemical profiles and concentrations of each analyte were determined by comparing the retention times of a reference library of 45 chemical standards (10 ppm) with known or suspected allelopathic properties. The concentrations of identified compounds were further
31 confirmed by a Thermo Finnigan TSQ7000 triple quadrupole mass spectrometer (HPLC/MS/MS, Thermo Electron Corp., Waltham, MA) using electrospray (+ and ionization modes) or HPLC MS atmospheric pressure chemical ionization (APCI). For an unknown compound that appeared to be present in the cogongrass leachate at high concentration s, retention time and ion fragmentation patterns, coupled with library matching, were used to generate a tentative structure. Statistical A nalysis T he effect of cogongrass leachate on the performance of each native species was assessed via compar isons with the native leachate treatment. Statistical comparisons between these two treatments were done using the MIXED procedure in SAS 9.2 (SAS Institute 2007 ) within the framework of a randomized complete blocks design, replicated in time. Y ear and bl ock (year) were treated as a random effect s For the different response variables (aboveground biomass, belowground biomass, total biomass, root length, specific root length, % mycorrhizal colonization and total mycorrhizal root length), differences between treatments were declared statistically significant at P < 0.05. Additional analyses were done with data from year two, using t test for post hoc comparisons with the DI water control (SAS Institute 2007). The Kenward Roger calculation was used t o estimate denominator degrees of freedom (Schaalje et al. 2002) 1 Some non statistically significant trends are reported in cases where there may be some ecological significance ( e.g when non significant relationships add evidence to inferences drawn fro m significant relationships). Concentrations of the various compounds (adjusted by the concentration factor) were 1 This method can result in non integer va lues for denominator degrees of freedom.
32 compared with the nonparametric Wilcoxon Mann Whitney test using the NPAR1WAY procedure (SAS Institute 2007). Differences between treatments w ere declared statistically significant at P < 0.05. Results Biomass P roduction A llocation and R oot M orphology Allelopathic interference from cogongrass leachates had variable effects on the biomass production and a llocation patterns for the four native seedlings. While no species had significant differences in total biomass between leachate treatments, aboveground biomass for A. stricta was 35 .7 % lower in the cogongrass leachate treatment than in the native leachat e treatment ( F (1, 40.28 ) = 15.04 P = 0. 0 004 ). This difference corresponded with a 2 2 2 % reduction in total root length ( F (1, 41. 02 ) = 4.86 P = 0. 0 331 ) and a 22.9 % reduction in specific root length ( F (1, 41.18) = 17.28 P = 0. 00 02 ). No such effects were o bserved for A. arctatus P. elliottii or L. ferruginea (Table 2 3 ). C omparisons made with the DI water control using year two data provide supporting evidence that the observed differences were due to the negative effects of cogongrass leachate. In all of the above cases where treatment effects were observed, the native leachate treatment was within 4.8 % of control ( t ( 1, 21) 0.35 P = 0. 9187; t (1, 21) 0.12 P = 0. 9 041 ; t (1, 21) 0.14 P = 0. 8872) for aboveground biomass, total root length and specific root length, respectively). Comparisons between the cogongrass leachate treatment and the DI water control, however, showed more substantial differences. Aboveground biomass was 37.7 % lower than control ( t (1, 21) 2.84, P = 0. 00 19 ), total root length was 29% lo wer ( t (1, 21) 2.22, P = 0. 0 695 ) and specific root length was 18.1% lower ( t (1, 21) 2.77 P = 0. 0 217 ).
33 Mycorrhizal I noculation and I nfected R oot L ength Differential treatment effects were also observed for plant mycorr hizal fungi associations. For P. elliottii EcM fungal inoculation (% mycorrhizal colonization) was 19. 4 % lower in the cogongrass leachate treatment than in the native leachate treatment ( F (1, 47.54 ) = 12.11 P = 0. 0 011 ). Reductions in total mycorrhizal ro ot length were observed for both A. stricta (23. 4 % ; F (1, 4 1 20 ) = 3.79 P = 0. 0 280 ) and P. elliottii ( 21.8 %; F (1, 4 7 62 ) = 4.96 P = 0. 03 07 ). For A. stricta this reduction is likely associated with the reduction in total root length observed. No such tren ds were observed for either A. arctatus or L. ferruginea Percent mycorrhizal colonization and total mycorrhizal root length were higher for A. arctatus in the cogongrass treatment, but these differences were not statistically significant (Figure 2 1). Aga in, comparisons with the DI water control in year 2 suggest that these differences were due to the negative influence of cogongrass leachate. In all cases, the differences between the cogongrass treatment and control were more substantial than those betwee n the native treatment and control. For P. elliottii both EcM colonization and total mycorrhizal root length in the native leachate treatment were within 11 % of control ( t (1, 2 8 ) 1.07 P = 0. 4604 and t (1, 2 8 ) 0.97; 0. 5226 respectively). In the cogongra ss leachate treatment, however, P. elliottii EcM colonization was 25. 6 % lower ( t (1, 2 8 ) 3.24 P = 0. 00 59 ) and total mycorrhizal root length was 22.1 % lower ( t (1, 2 8 ) 1.95 P = 0. 1079 ). Total mycorrhizal root length for A. stricta in the native leachate t reatment was 3 .6% lower than control ( t (1, 21) 0.18 P = 0. 9783 ), while the cogongrass treatment was 26.1% lower ( t (1, 21) 1.30 P = 0. 3461 ). Chemical P rofiling of L eachates The chemical profile of cogongrass leachate was qualitativ ely and quantitatively different from that of native leachate. Eleven potentially allelopathic organic compounds
34 in the cogongrass leachate were identified in the initial HPLC analysis, most of which were found at significantly lower concentrations or no t found at quantifiable levels in the native leachate. Phenolic acids were the predominant class of allelopathic compound in the cogongrass leachate. The phenolic compound with the highest concentration was gallic acid ( 3 .03 ppm), followed by caffeic aci d (0.85 ppm), salicilyc acid (0.61 ppm) and sinapinic acid (0.33 ppm). The other compounds, which included a carboxylic acid (benzoic acid), an anthraquinone (emodin) and a dihydroxy benzene (resorcinol) all had concentrations less than 0.16 ppm. No compou nds in the native leachate had concentrations greater than 0.09 ppm. Five of the compounds (caffeic, benzoic, cinnamic, ferulic and chlorogenic acid) have been positively identified in previous studies of cogongrass alle l ochemistry (Abdul Wahab and Al Naib 1972; Hussain and Abidi 1991; Xuan et al. 2009) A complete list of the identified compounds, along with statistical comparisons of their concentrations is provided in Table 2 4. Along with the confirmation of the compounds described above, the HPLC MS a nalysis also suggested that a novel alkaloid compound was present in the cogongrass leachate. Based on fragmentation patterns and library matching, the speculated structure is hexadecahydro 1 azachrysen 8 yl ester ( C 23 H 33 NO 4 ) (Figure 2 2) It appeared to b e present at fairly high levels, although it was not possible to estimate its actual concentration due to lack of commercial reference standards. This compound was not found in the native leachate treatment. Discussion Uren (2007) compiled a list of over 100 secondary compounds thought to be exuded by plant tissues. This list includes an array of sugars, polysaccharides, amino acids, organic acids, fatty acids, sterols, growth factors, enzymes, flavonones,
35 nucleotides and other chemicals, many of which ar e suspected to be involved in mediating belowground interactions with plants and/or soil fauna. Effectively assessing the role of these substances on ecological processes, however, is dependent upon an understanding of their composition and significance in plant soil systems (Mallik 2000; Uren 2007) an area where research is sorely lacking. W hen using live plants in a natural soil medium as I did, however, it is undoubtedly very difficult to isolate the effects of these compounds from the confounding infl uences of water extractable matrix solutes, microbes, microbial compounds, root degradation products and other substances (Uren 2007) Th ese shortcomings however, may be offset by the fact that test species were exposed to a biologically realistic mixture of allelochemicals and their breakdown products. While allelopathy is commonly suspected to be a driving force behind alien plant invasions into natural areas, only a small number of studies have reported the presence and bioactivity of exudates in the r hizospheres of IA plants. The bulk of the studies of allelopathy in natural systems have focused on spotted knapweed ( Centaurea maculosa ), a species native to Europe and western Asia that has transformative effects on ecosystems throughout North America. S oils in the vicinity of this problematic invader have sometimes been shown to contain high concentrations of the flavonoid secondary metabolite () catechin (Perry et al. 2007). In controlled experiments, ( ) catechin has been shown to have significant inh ibitory effects on the germination, growth and overall health affected species, along with having microbicidal properties (Vivanco et al. 2004). In plants, this compound is believed to work by causing the production of reactive oxygen species at the root m eristem, which initiate a series of
36 biochemical and genetic alterations (Bais et al. 2003, but see partial retraction 2010). A similar mechanism is suspected in closely related C. diffusa (Hierro and Callaway 2003). Little is known about the exudate chemi stry of cogongrass, but my findings suggest that unlike C. maculosa no single compound is likely responsible for its apparent allelopathic effect. This is probably typical for most allelopathic species, as allelopathic interference is generally thought to result from combinations of allelochemicals and their breakdown products, interacting simultaneously and sometimes synergistically, with multiple physiological processes in the affected organism (Einhellig 1995). The concentrations of individual compounds are usually below a bioactivity threshold, but their effects can be additive (Chung et al. 2002). Phenolics are among the most common classes of allelopathic compounds exuded by grasses (Snchez Moreiras et al. 2003), and my findings suggest that cogongra ss is no exception. Inderjit and Dakshini (1991) isolated 18 nonspecific phenolic fractions from cogongrass tissues and soils, but it is impossible to confirm if any of the same compounds were present in my cogongrass leachates Some of the phenolic compou nds I described have been identified and (in some cases) shown to have phytotoxic activity in studies of cogongrass tissue extracts (Abdul Wahab and Al Naib 1972; Hussain and Abidi 1991; Xuan et al. 2009). However, many of the phenolics identified by the a bove authors were not found in this study. Additionally, Xuan et al. (2009) identified several long chain fatty acids ( e.g stearic acid and myristic acid) and miscellaneous compounds (coumaran and pantolactone) in cogongrass roots and rhizomes that were n ot present in my cogongrass leachates. A few of the non phenolic
37 compounds that I identified ( e.g. emodin, resorcinol) have not been reported in other studies of cogongrass allelopathy. Overall, despite highlighting the near ubiquity of certain phenolics, these differences reinforce the notion that tissue extracts may not be a biologically realistic proxy for allelopathic exudates. While the specific mechanism of action that brought about the observed reductions in growth and mycorrhization for A. stricta a nd P. elliottii is unclear, it has been proposed that phenolics such as cinnamic, benzoic and ferulic acids all of which were present in the cogongrass leachate have general toxicity and can interfere with phytohormone interactions, cell membrane struc ture and function, photosynthesis, enzymatic reactions and carbon flow, among other important physiological processes in plants and soil organisms (Einhellig 1995). Alkaloids have received comparatively less research attention, but common alkaloid compound s have been reported to affect DNA synthesis, respiration and electron transport (Einhellig 2002). An alkaloid similar to the one I described in this study has been identified in studies involving Sorghum bicolor and appears to function as a nitrification inhibitor in soils ( Chung Ho Lin, personal communication ). Emodin, an anthraquinone that is also found in the invasive plant Japanese knotweed ( Polygonum cuspidatum Siebold and Zucc.) has been shown to reduce root and shoot growth and alter the availabili ty of mineral nutrients in the soil (Izhaki 2002). Resorcinol, which was present in very low concentrations, does not appear to be directly phytotoxic (Seal et al. 2004) but has been shown to have antifungal properties (Suzuki et al. 1996). An intriguing aspect of these findings is the fact that the allelopathic influence of cogongrass appears to vary by species. Others have speculated on the possible
38 species specificity of allelopathic interference from IA plants (McCarthy and Hanson 1998; Abhilasha et al 2008), including cogongrass (Xuan et al. 2009) but I believe that this is the first study that has focused on the effects of cogongrass on native species. The root length, morphology and mycorrhizal measurements provide insight into some possible explan ations behind the observed differences. For A. stricta reductions in total root length and concurrent decreases in specific root length suggest that allelopathic interference inhibits root elongation and/or branching. This in turn likely creates less oppo rtunity for mycorrhizal colonization. Bluestem grasses and/or their associated belowground symbionts may be resistant to allelochemicals exuded by cogongrass, but the underlying mechanism is unclear. Among the two woody species, only the EcM tree ( P. ellio ttii ) appeared to be affected. Low concentrations of phenolic mixtures have been shown to inhibit EcM fungi (Souto et al. 2000); perhaps this is the mechanism at play here. Ericoid mycorrhizal fungi, like those that colonize L. ferruginea may have an inhe rent tolerance to allelochemicals, since the root systems of these species proliferate in litter and organic soil horizons where plant exuded phenolics and their breakdown products are typically present in relatively high concentrations (Bending and Read 1 997). Ericoid mycorrhizal fungi have some ability to degrade phenolics (Bending and Read 1997), which may enable access to labile organics and nitrogen formerly complexed with these compounds (Bending and Read 1996) However I observed no evidence of enhan ced L. ferruginea growth in the cogongrass leachate treatment. The species specificity of the apparent allelopathic response, which was perhaps mediated by interactions with mycorrhizal fungi, may help explain the patterns of invasion that have been obser ved both in cogongrass impacted pine ecosystems and in
39 experimental mesocosms. Aristida stricta which appeared to be negatively affected by cogongrass soil leachate, rarely persists in sites invaded by cogongrass (Hagan, personal observation; Jose et al. 2002). Cogongrass invasion also inhibits pine regeneration (Daneshgar et al. 2008). While the above effects have most often been attributed to competition (Brewer 2008; Daneshgar and Jose 2009 a ), fire feedbacks (Lippincott 2000) and/or physical interferenc e (Holly and Ervin 2006), my findings suggest that these interactions may be compounded by allelopathic activity as well. In the acidic nutrient poor flatwoods soils characteristic of the environments where these species are typically found, reductions in root length and/or mycorrhizal colonization will likely represent a major fitness disadvantage for affected native species. In a mesocosm study that looked into the effects of native species diversity and identity on cogongrass invasion, broomsedge blueste m ( Andropogon virginicus L. ) performed significantly better than other native herbaceous species when grown in close association with the cogongrass (Daneshgar and Jose 2009 b ). Perhaps this was due in part to an inherent resistance to allelopathic interfer ence, as my findings with closely related A. arctatus suggest. I know of no studies that have looked into the performance of Lyonia in cogongrass invaded systems, but research on other species has suggested that the resistance to allelopathy conferred to e ricaceous plants by their symbiotic fungi enhances their ability to persist in environments dominated by allelopathic invaders. Ericads, for example, are among the few species apparently capable of overcoming the powerful allelopathic influence of Casuarin a (Reed 1989) a commonly invasive genus known to exude phenolics and other phytotoxic compounds (Sayed et al. 2002).
40 Summar y and Implications This study represents the first attempt to assess the effects of allelopathy from cogongrass o n native species in an ecologically relevant setting. Overall, my findings support the hypothesis that novel and potentially allelopathic compounds are present in the cogongrass rhizosphere. Substantial reductions in biomass production and/or mycorrhizal c olonization, along with altered root morphology, which were observed for 2 of 4 native species treated with the cogongrass leachate, suggest that these compounds are present in biologically significant concentrations. It is likely, therefore, that the tran sformative nature of cogongrass in its invaded range can be attributed, at least partially, to the effects of allelopathic interference. Additional research should seek to shed light on the bioactivity of the alkaloid, as well as explore possible alleloche mical tolerance mechanisms possessed by L. ferruginea and A. arctatus It would also be beneficial to assess the presence and concentration of the observed compounds in the field
41 Table 2 1. Native pine savanna species used in a study of the allelopathic effects of cogongrass leachate. a Denotes the mycorrhizal fungal symbiont (AM, arbuscular mycorrhizal; EM, ericoid mycorrhizal; EcM, ectomycorrhizal). Scientific name Common name Family Symbiont a Aristida stricta Michx. var beyrichiana (Trin. and Rupr.) Wiregrass Poaceae AM Andropogon arctatus Chapm. Pinewoods bluestem Poaceae AM Lyonia ferruginea (Walter) Nutt. Rusty lyonia Ericaceae EM Pinus elliottii Engelm. Slash pine Pinaceae EcM
42 Table 2 2. Effects of aqueous and chloroform extracts of le achates (native, cogongrass, DI water control) on the germination (%) of lettuce seeds. Four replicates of each treatment x extract combination, each with 10 seeds, were used. hoc t test. Pairs of means with asterisk s are significantly different at P < 0.05. Extract Native Cogongrass Control Aqueous 90 75 92.5 Chloroform 85 87.5 87.5
43 Table 2 3. Species wise comparisons (means and standard errors) of the effects of a cogongrass leachate trea tment vs. the effects of a native leachate treatment on biomass production, allocation, root length and specific root length for four native pine savanna species. Pairs of means with asterisks are sig nificantly different at P < 0.05. Species Treatment Abo veground (g) Belowground (g) Total (g) Root length (cm) cm root/g A. stricta Cogon 0.009 (0. 00 6 ) 0. 03 5 (0. 0 10 ) 0. 04 4 (0. 0 15 ) 4 5 70 ( 13.92 ) 13 4 9 75 ( 58.33 ) A. stricta Native 0.014 (0. 00 6 ) 0.036 (0. 0 10 ) 0.049 (0. 0 15 ) 5 8 71 ( 13.99 ) 17 5 1 96 ( 6 5 56 ) A. arctatus Cogon 0. 03 1 (0. 00 5 ) 0.111 (0. 0 17 ) 0. 14 3 (0. 0 17 ) 10 3 90 ( 9.38 ) 1064.02 ( 123.33 ) A. arctatus Native 0. 02 3 (0. 00 6 ) 0. 10 9 (0. 0 20 ) 0. 1 4 0 (0. 0 26 ) 93.61 ( 8.11 ) 107 1 2 3 ( 125.50) P. elliottii Cogon 0.159 (0. 0 0 8 ) 0. 23 1 (0. 0 17 ) 0. 3 89 (0.0 20 ) 37. 44 ( 6.2 5 ) 169.96 ( 21.52 ) P. elliottii Native 0.184 (0. 0 0 8 ) 0. 2 69 (0. 0 17 ) 0. 4 51 (0. 0 20 ) 37. 88 ( 6.24 ) 14 2 64 ( 21.47 ) L. ferruginea Cogon 0 00 5 (0.00 1 ) 0. 01 2 (0.002) 0. 01 7 (0. 00 3 ) 7. 62 (1. 30 ) 662.11 ( 36.16 ) L. ferruginea Native 0 00 5 (0. 00 1 ) 0. 01 3 (0.002) 0. 01 7 (0. 00 3 ) 7. 8 4 (1. 28 ) 701.60 ( 29.13 )
44 Table 2 4. Mean chemical composition of leachates (ppm) collected from the rhizosphere of greenhouse grown cogongrass monocultures and native polycultures. Compounds identified in previous studies are denoted, alo ng with the source of the extract. Compound Family Previously reported? Retention (min) Concentration (ppm) Wilcoxon Mann Whitney Cogon Native Gallic acid Phenolic acid 2.50 3.03 0.09 < 0.05 Caffeic acid Phenolic acid 1 SH 3.77 0.85 0.03 < 0.05 Salicylic acid Phenolic acid 12.41 0.61 0.05 < 0.05 Sinapinic acid Phenolic acid 5.60 0.33 0.01 < 0.05 Benzoic acid Carboxylic acid 3 RH, RO 10.30 0.16 0.03 NS Emodin Anthraquinone 24.51 0.16 BQ < 0.05 Cinnamic acid Phenolic acid 3 RO, 15.92 0.12 0.01 < 0.05 Ferulic acid Phenolic acid 1 SH, 3 RH, RO 6.10 0.11 0.01 NS 4 hydroxyphenylacetic acid Phenolic acid 4.09 B Q B Q -Cholorogenic acid Phenolic acid 1 SH, 2 RO 2.50 B Q B Q -Resorcinol Dihydroxy benzene 4.13 B Q B Q -1 Abdul Wahab and Al Na ib (1972); 2 Hussain and Abidi (1991); 3 Xuan et al. (2009); SH = extracted from shoots; RH = extracted from rhizomes; RO = extracted from roots; *< 0.05, indicates that differences between treatments were statistically significant; NS indicates that differ ences between means were not statistically significant; BQ, below limits of quantification; NS -, indicates that comparisons not possible due to concentrations in both leachates being below limits of quantification.
45 Figure 2 1. Difference in percent mycorrhizal colonization (A) and total mycorrhizal root length (B) for four native species watered with cogongrass leachate, relative to those watered with leachate from native species. Means and standard errors. Differe nces with asterisks are statistically different at P < 0.05. A B
46 Figure 2 2. Speculated chemical structure of a novel alkaloid (hexadecahydro 1 azachrysen 8 yl ester) identified in cogongrass leachate (A), ion chromatography of the alkaloid, indicating t he retention time (19.06 minutes) (B) and the mass spectrum (m/z 372.18) (C). A B C
47 CHAPTER 3 COGONGRASS INVASION AND ERADICATION : IMPLICATIONS FOR SOI L BIOGEOCHEMICAL PROPE RTIES IN A FIRE MAINTAINED FOREST EC OSYSTEM Background Invasive alien (IA) p lants typically grow fast, rapidly colonize new sites, compete favorably with native species and have few natural enemies outside of their home range. While theories abound as to the specific mechanisms alien plants use, or traits they possess, that enable them to be successful invaders (Davis et al. 2000; Maron and Vil 2001; Bakker and Wilson 2001; Callaway and Ridenour 2004), a near universal characteristic is their ability to alter often dramatically the species composition of ntonio and Vitousek 1992; Gordon 1998; Hejda et al. 2009). Differences in resource uptake patterns, as well as the changes in litter quality and quantity that accompany these vegetative shifts, in turn, can greatly alter the nutrient cycling dynamics of an ecosystem (Kourtev et al. 2002; Ehrenfeld 2003; Allison and Vitousek 2004). The extent of change depends on how different the invader is from the species that it replaces with respect to traits such as life history, physiology, size, above and belowgroun d architecture, tissue chemistry, photosynthetic pathways, symbiotic relationships and other factors (Ehrenfeld 2003). The body of knowledge on the effects of IA species on soil nutrient cycling processes while limited, has increased greatly in recent ye ars (Ehrenfeld 2003). Less attention, however, has been paid to post eradication effects (Maron and Jeffries 2001; Yelenik et al. 2004), although researchers have speculated on the potential for legacy tonio 2004; Renz and Blank 2004; Yelenik et al. 2004). This, clearly, is an area that deserves more consideration, as the restoration of native plant communities following the eradication of IA species is a high
48 priority among land managers (Zavaleta et al 2001; Hartman and McCarthy 2004 ; Miller et al. 2010 ). Much like any other disturbance, if soil processes and properties are altered by an invader, these effects will likely persist for some time after the invader is removed rdan et al. 2008). The decomposition of plant biomass following treatment might cause further alterations to soil properties through its effects on soil organic matter (OM) pH and the mineralization and immobilization of nitrogen (N) and phosphorous (P). These alterations may in turn affect nitrification rates (Raison 1979; Chapin et al. 2002), or phosphorus complexation reactions (Brady and Weil 2002). Novel mycorrhizal fungal communities may also persist (or develop ) following eradication, but this is an area that has yet to receive much research attention. In concert, these changes might impede the re establishment of desirable native species and/or increase the potential for re invasion by either the same or new alien species (Kourtev et al. 2003). An improved understanding of how IA plant species alter soil properties, and how novel soil properties persist /develop following eradication, is essential in order to develop effective long term restoration strategies for invaded forest communities. Toward th ese ends, I undertook this study to assess nutrient dynamics in forest stands severely impacted by cogongrass ( Imperata cylindrica ( L ) P. Beauv. ) a C 4 rhizomatous invader that affects tropical and subtropical ecosystems on six continents (MacDonald 2004 ). I chose a fire maintained longleaf pine ( Pinus palustris Mill.) sandhill ecosystem as the study site, as these forests are frequently targeted in restoration efforts (Walker and Silletti 2006) and are commonly invaded by cogongrass (Jose 2002; Daneshgar and Jose 2009 a ). Since nitrogen and/or phosphorus availability often drive
49 ecological succession following disturbance (Tilman 1985; Vitousek et al. 1993), I focused on processes that affect their availability and uptake. The three primary objectives of t his study were: Analyze how invasion by cogongrass affects soil N and P pools, fluxes and associated processes in fire maintained longleaf pine sandhill stands. Quantify soil N and P dynamics in the years following cogongrass eradication. Assess the effec ts of cogongrass invasion and eradication on arbuscular mycorrhizal (AM ) fungal communities I hypothesized that cogongrass invasion would decrease the availability of soil N and P, likely through reductions in pH and/or changes to the soil carbon cycle I also expected these changes in N and P cycling to persist following eradication, perhaps due to the slow decomposition rates of low quality cogongrass foliage and rhizomes after herbicide treatment. Additionally, I hypothesized that cogongrass invasion wou ld result in the development of a novel AM fungal community, and that additional modifications to AM fungal community structure would arise following cogongrass eradication. Materials and Methods Study A rea The study area was an uneven aged, naturall y regenerated longleaf pine forest in the Croom Tract of Withlacoochee State Forest in Hernando County, Florida (2836'19.99"N, 8216'19.73"W). The tract is near one of the original points of cogongrass introduction in the United States and has a long hist ory of invasion. Efforts in recent years to chemically eradicate most cogongrass infestations in the tract have been successful and there are numerous areas in various stages of recovery throughout, although some untreated cogongrass patches remain. The un invaded areas are characterized by high levels of understory species richness and diversity, as is
50 typical of an actively managed, frequently burned longleaf pine sandhill community. Predominant understory species in uninvaded areas were wiregrass ( Aristid a stricta Michx. var. beyrichiana (Trin. and Rupr.) D.B.Ward), along with various native trees, graminoids, forbs, shrubs and vines ( Chapter 4 ). Soils in the study area were predominantly deep, well drained to excessively drained sands of the Lake and Cand ler series (hyperthermic coated Typic Quartzipsamments and hyperthermic uncoated Lamellic Quartzipsamments, respectively). Small inclusions of the Arredondo series (Loamy, siliceous, semiactive, hyperthermic Grossarenic Paleudults) comprising less than 2 0% of the total area were also present (US Department of Agriculture 1977). Mean overstory basal area for the study site was 10 .7 m 2 ha. Longleaf pine constituted approximately 88% of total basal area. Experimental D esign I used invaded and uninv aded sites across 4 longleaf pine sandhill stands in the study area to assess the effect of cogongrass invasion on soil N and P dynamics. Additionally, since some sites within these stands had cogongrass eradicated in previous years, I changes in N and P cycling following eradication. Sites selected for the chronosequence treatments were treated in the late summer/early fall approximately three five and seven years prior respectively with a t ank mix solution (sprayed to the point of runoff) consisting of 2 % Roundup (41% glyphosate plus surfactant) and 0. 4 % (28.7% imazapyr ). Glyphosate and imazapyr tank mixes such as this are among the most common and effective methods of chemical control for cogongrass (Mac Donald
51 2004) 1 All sites hereafter referred to as plots, were identified and selected using Geographic Information Systems (GIS) and with the help of state forest personnel. Native reference plot s randomly selected using GIS f rom the surrounding uninvaded area, were ground truthed to verify that they were not currently invaded and did not fall on disturbed or degraded sites ( e.g. roads, bicycle trails, abandoned rock mines formerly invaded sites ) Plot s with live cogongrass we re estimated to be 1 2 years old, which is approximately the same age that those in the recovery chronosequence were when they were treated Cogongrass was the dominant species in these plot s The study was laid out as a complete blocks design replicated twice. Each replicate consisted of an adjacent pair of 259 ha stands (blocks) with similar, but asynchronous burn histories (both burned approximately every 4 years, but usually staggered 2 years apart). One block in each replicate was burn ed last in June 2009 and the other was burned in June 2007. Each block typically contained 2 3 plot s from each of the five treatments : reference ( i.e. uninvaded), invaded three years since eradication, five years since eradication, and seven years since eradiation (Tabl e 3 1). Within each plot three subplots were randomly selected, each being at least 2.5 meters from the other at least 8 m from the edge and distant from any cogongrass re sprouts (where applicable) (Figure 3 1) Additionally, 4 recently treated cogongra ss patches were selected as locations for a litterbag decomposition study (described below). 1 A single herbicide treatment does not always completely eradicate a cogongrass patch. However, for young ( 2 year old) patches in the Croom tract, > 95% control is typical. For the purposes of this study,
52 Soil C hemistry and N utrient P ools Soil samples from the top 15 cm of the profile were collected from subplots in June 2010 using a stand ard 1 piece soil probe. Each sample was a composite of five subsamples one taken at plot center and the remaining four collected one meter away at each of the four cardinal directions. Samples were transported to the lab in a cooler and then moved to a f reezer, where they remained at minus 4C until analysis. Soil OM content was determined by acid digestion and pH was measured using a 1:2 soil:water ratio. Total N (TKN method) and potentially available P were quantified using an Apkem autoanalyzer and a M ehlich 1 (M1) extraction, respectively (Mylavarapu 2002). Soil N utrient A vailability The availability of N and P in the different treatments was assessed with mixed cation anion exchange resin bags, incubated in situ (Standish et al 2004; Harpole and Tilman 2007) during the 2010 growing season. By integrating microenvironmental factors ( e.g. water availability, flow and plant uptake) during the incubation period (Binkley 1984), this method provides additional information about N and P cycling in terrestrial systems (Binkley et al. 1986; Gibson 1986; Feller et al. 2003). Prior to incubation, resins were washed in sodium chloride (NaCl) and sodium hydroxide (NaOH) per the procedure outlined in Thiffault et al. (2000). Bags consisted of approximately 10 g (moist weight) of washed resin (Dowex Marathon MR 3), cinched in a square of acid washed nylon lycra mesh with a plastic zip tie to make a firm, spherical bag (Thiffault et al. 2000). In early May and September of 2010, a bag was buried at a depth of five cm in e ach subplot Bags were removed after 33 day incubation periods Upon return to the lab, they were gently rinsed in de ionized (DI) water to remove adhering soil particles, and then shaken in a 2 M NaCl solution for two hours. Ext racts
53 were analyzed using an autoanalyzer for total adsorbed nitrite+nitrate N (NO 2 +NO 3 ) and ammonium N (NH 4 ) and by inductively coupled plasma mass spectrometry (ICP MS) for total adsorbed P (Thiffault et al. 2000; Harpole and Tilman 2007). Litter D ecomposition and N utrient M ineralization Initial nutrient cycling following cogongrass eradication may be affected by the decomposition of dead biomass. Therefore, to determine the rate s of cogongrass decomposition and nu trient mineralization/immobilization, 40, 1 mm fiberglass mesh litterbags (20 filled with five g of air dried cogongrass rhizomes and 20 filled with five g of air dried cogongrass foliage) (Ashton et al. 2005) were incubated in clusters of five in four ran dom locations in recently treated (fall 2009) cogongrass patches scattered across the two blocks. These tissues were collected from an adjacent area that was treated two weeks prior with the glyphosate and imazapyr tank mix In December 2009, rhizome bags were buried to a depth of five cm and foliage bags were left on the soil surface. After 31, 90, 192, 373 and 544 days (approximately 1, 3 6, 12 and 18 months), eight bags were collected ( one per tissue type, per location) and transported to the lab, where their contents were carefully removed, freed from adhering soil and dried at 65 C Subsamples were ground to <1 mm and analyzed for total C, N and P per the procedure outlined in Bray et al. ( 2005 ) Mass loss and nutrient mineralization/immobilization, relative to pre incubation values, were then determined for each sampling date (Allison and Vitousek 2004). AM F ungal S pore Q uantification Arbuscular mycorrhizal spores from soil samples (approximately five g) taken from each plot were iso lated using a standard wet sieving, decanting and glucose centrifugation technique (Daniels and Skipper 1982). The final product was transferred
54 to a test tube and brought to a volume of 5 mL. A 0.5 mL aliquot was transferred to a piece of filter paper cut to the size of a microscope slide for spore quantification. Spore counts were converted to spores/mL, then spores/gram for statistical analysis. Soil AM F ungal DNA E xtraction PCR, C loning and S equencing Composite s oil samples ( two per treatment = 10 total ) were prepared by thoroughly homogenizing four samples from randomly selected plot s (one plot per block) from each of the five treatments. An exception was made for the invaded treatment, since one of the blocks co ntained only one invaded plot. Two of the samples in this composite sample came from the same block Each of the four samples in a composite sample was, in itself, a composite of soil samples from three subplots in a plot Soil DNA was extracted from compo site instructions using a MO BIO UltraClean Soil DNA Isolation Kit. This DNA was used in a PCR reaction using the AM fungal SSU rRNA specific primers AML1 ( 5 ATC AAC TTT CGA TGG TAG GAT AGA 3 ) and AML2 ( 5 GAA CCC AA A CAC TTT GGT TTC C 3 ) (Lee et al. 2008) using the following regime: 15 min initial denaturation at 94C, 36 cycles at 94C for 30 sec, 58C for 40 sec, 72C for 55 sec and a final extension at 72C for five min. PCR products (10 L) were verified by gel e lectrophoresis and purified as needed, with a MO BIO Ultra Clean PCR clean up kit, following the Purified PCR products were cloned according to the TA Cloning kit with On e Shot electrocompetent cells and kanamycin selective plates. Colonies were incubated for overnight at 37C. Selected colonies were then transferred to 96 well plates in 200 L of kanamycin selective LB medium and incubated at 37C for an additional 24 ho urs.
55 Clones were sequenced on an Applied Biosystems Model 3130 Genetic Analyzer using the T7 and R24 sequencing primers. Sequence P rocessing and A nalysis Sequences were aligned using CLUSTALX2 (Larkin et al. 2007) and a consensus nei ghbor joining phylogenetic tree was generated with Mega V. 5 (Tamura et al. 2011 ), using default settings, 1000 bootstrap replications and representative sequences from Genbank. The MOTHUR program (v. 1.23.0) was used to assign sequences to operational tax onomic units ( OTUs ) at a 3 % cutoff, using the cluster (furthest neighbor algorithm) and bin.seqs commands (Schloss et al. 2009). A BLAST search was conducted on representative sequences from each OTU using the nr/nt nucleotide database. Since the different AM fungal families may have different functional strategies or ecological niches (L ekberg et al. 2007), BLAST hits were binned by their respective families, by treatment. Fungal OTU richness (Chao1 richness estimator) for each sample was estimated using the summary.single command in group mode in MOTHUR. Summary.single was also used to calculate two common measures of OTU diversity. The Shannon Wiener index ( H ) index was calculated by MOTHUR as follows: w here S obs is the number of observed OTUs, n i is the number of individuals in OTU i and N is the total number of individuals in the community (Schloss et al. 2009). The Simpson index ( D ) was calculated by MOTHUR as follows:
56 w here n i is the number of OTUs with i individuals (all other parameters are th e same as in H ). The Jaccardian pairwise similarity index (summary.shared; MOTHUR) was used to determine the proportion of individuals between each combination of treatments that belong to shared OTUs (Schloss et al 2009). A Mantel test which tests the n ull hypothesis of no relationship between matrices (McCune and Grace 2002), were used to assess the relationship between a matri x of fungal OTUs and the soil variables reported in this study. Statistical A nalysis The three subplot v alues from the soil resin bag and spore analyses were averaged to obtain plot level estimates For the resin bags, plot level estimates for the May and September deployments were also averaged to generate single growing season estimates of NO 2 +NO 3 NH 4 and P availabilit y. The data were analyzed using the MIXED procedure in SAS 9.2 (SAS Institute, Inc.). Replicate and block( replicate ) were treated as random effects. The Kenward Roger calculation, a preferred method for unbalanced mixed models (Spilke et al. 2005) was use d to estimate denominator degrees of freedom 2 Differences between means were declared statistically significant at P hoc test was used for pairwise comparisons. Some non statistically significant trends are reported in cases where there may be some ecological significance ( e.g when non significant relationships add evidence to inferences drawn from significant relationships). Decomposition rates ( k coefficients) for cogongrass rhizomes and foliage were calculated based on a negativ e exponential model following Bray (2005). Treatment means (Chao1, Shannon Wiener and Simpson ) for mycorrhizal 2 This method can result in non integer values for denominator degrees of freedom.
57 analyses were compared using the GLM procedure in SAS 9.2. The weighted UniFrac Significance Test, a Monte Carlo procedure (100 permutations) wa s used to compare the community structures of the five treatments. This method measures the fraction of branch length in a phylogenetic tree that is unique to each treatment, and accounts for the proportional representation of each treatment in each branch (Schloss 2008). It generates P values which are used to assess if the differences in genetic distance between pairs of communities is greater than would be expected by chance alone (Lozupone et al. 2007 ; Schloss 2008 ). Differences between treatments with UniFrac P were declared statistically significant. A principal components analysis (PCA) biplot was generated in UniFrac to help visualize treatment separation in variable space ( Lozupone et al. 2007 ). Results Organic M atter and p H Like many of the other measured soil properties, soil OM was highly variable among the five treatments, ranging from 0.83 to 3 .17% with a mean of 1.84%. Differences between treatments were significant ( F ( 4, 43.94 ) = 2.71 P = 0. 0417 ) with OM contents be ing highest in invaded plots ( 2.22%) and lowest in plots where cogongrass was eradicated seven years prior (1.6 2 %) ( Figure 3 2) Soil pH in the different treatments ranged from 4.87 to 6.00 with a mean of 5 .44. Differences in pH between treatments were sign ificant ( F ( 4, 43.47 ) = 7.34, P = 0. 0 001 ) with pH in five year plots ( 5 6 3 ) being significantly higher than that of the native reference plot s ( 5 .40) and currently invaded plot s ( 5 2 3 ) (Figure 3 3 ).
58 Nitrogen Total N contents ranged from 666. 7 to 1400.0 m g/kg, with a mean of 990.1 mg/kg. Differences between treatments were not significant ( F ( 4, 43.52 ) = 1.38 P = 0. 2 5 57 ). R esin adsorbed NH 4 followed a similar pattern rang ing from <0.01 to 0.08 mg/bag, with a mean of 0.02 mg/bag and no significant differen ces between treatments ( F ( 4, 43.38 ) = 0.26 P = 0.9000) Resin adsorbed NO 2 +NO 3 did however vary significantly between treatments ( F ( 4, 43.15 ) = 12.81 P < 0.0001). Contents ranged from <0.01 to 0.10 mg/bag, with a mean of 0.02 mg/bag. Soil NO 2 +NO 3 levels were highest three years after eradication (0.05 mg/bag) and lowest in the reference and invaded treatments (0.01 and 0.01 mg/bag, respectively. Levels decreased after this initial spike and were not significantly different from the reference treatment se ven years following cogongrass eradication (Figure 3 4). Phosphorus Soil M1 extractable P contents ranged from 50.8 to 347. 3 mg/kg with a mean of 120. 7 mg/kg. Differences between treatments were significant ( F ( 4, 43.92 ) = 3.76, P = 0. 01 02 ), with M1 P con tents being lowest three and five years after cogongrass eradication (9 8.2 and 102. 3 mg/kg, respectively), and highest in cogongrass invaded plot s ( 164.4 mg/kg). Resin adsorbed P ranged from 0.02 to 0.78 mg/bag with a mean of 0.20 mg/bag. It followed a sim ilar pattern as M1 extractable P ( F ( 4, 44.04 ) = 3.06 P = 0. 02 63 ), with the highest values being found in invaded plot s (0.33 mg/bag) and the lowest values found in plot s where cogongrass was eradicated three and five years prior (0.13 and 0.14 mg/bag, res pectively) (Figure 3 5 ).
59 Tissue Q uality D ecomposition and M ineralization Cogongrass rhizomes and foliage differed with respect to tissue chemistry and they exhibited different patterns of mass loss during the 18 mon th field incubation period. Decomposition rates ( k coefficients) were 1.01 and 0.44 for rhizomes and foliage, respectively (Table 3 2 ). N utrient mineralization occurred fairly rapidly for cogongrass foliage, with 43. 7 and 20. 5 % of initial N and P remaining respectively, after 18 months. A different trend, however, was observed for rhizome tissues. Following an initial spike in immobilization, in which tissue N levels were more than 2. 5 times greater than initial values, tissue N dropped to 70.0% of initial levels after 18 months. In contrast with N, rapid P mineralization occurred for cogongrass rhizomes, with 15.4% remaining after 18 months (Figure 3 6). Arbuscular M ycorrhizal F ungi Arbuscular mycorrhizal spore counts ranged from 1.89 t o 33.93 spores/g with a mean of 12.72 spores/g. There were no significant differences between chronosequential treatments ( F (4, 43.65 ) = 0.72 P = 0. 5 838 ) nor were there any apparent trends or patterns. The PCR protocol generated the expected ca. 800 bas e pair amplicons. At the 97% cutoff, clone library coverage averaged 92% and was highest in the seven year treatment (97%) and lowest in the invaded treatment (84%). The 304 sequences were classified into 31 OTUs. BLAST searches suggested that nine OTUs we re in the Acaulosporaceae family, five were Gigasporaceae, twelve were Glomeraceae and five were Paraglomeraceae (Appendix A) Across treatments, Chao1 richness ranged from 13.33 to 21.75, H D from 0.80 to 0.87. None of these indic es differed significantly between treatments ( F (4,5) = 0.89, P = 0. 53 ; F (4,5) = 0.86, P = 0. 5 4;
60 and F (4,5) = 1.76, P = 0. 27) for Chao1, H D respectively) and there were no clearly evident trends, but there was considerable between treatment variab ility (Figure 3 7 ). Many fungal OTUs were shared between treatments, as indicated by high similarity indices (mean 72.8, range 56.6 80.4). The Mantel test did not indicate a strong relationship between fungal and soil matrices ( P = 0.270). A condensed p hylogenetic tree (Figure 3 8) generated from a subset of aligned sequences, was consistent with AM fungal SSU trees constructed previously (Redecker 2002; Redecker and R aab 2006) Pairwise comparisons made using the UniFrac Significance Test (using a phyl ogenetic tree made from all sequences) indicated significant differences in the community structure between the different treatments. The reference treatment was significantly different from the three year and invaded treatments ( P = 0. 0 0 and 0. 0 0 respect ively). The three year treatment was also significantly different from the seven year treatment ( P = 0.01) and the invaded treatment was significantly different from the five year treatment ( P = 0.05) (Table 3 3 ) A plot of the first two axes of the weight ed UniFrac PCA, which accounted for > 67 % of total variation between treatments, illustrated a similar pattern. While there was substantial within treatment variability, points for the five and seven year treatment s generally were closer to reference than w ere the points from the three year and invaded treatments ( Figure 3 9 ). Discussion My findings suggest that no substantial changes in soil properties occurred directly as a result of cogongrass invasion in the longleaf pine sandhill ecosystem. This does not support my first hypothesis and it stands contrary to the findings of other researchers, who have suggested that cogongrass alters soil chemistry in southern pine
6 1 eco systems (Collins and Jose 2008; Daneshgar and Jose 2009 a ). Since changes in soil nutri ent cycling due to cogongrass invasion were not evident, the persistence of such effects which I proposed in hypothesis 2 was not possible. Some temporary alterations to N and P dynamics did, however, develop in the years following eradication. These t rends, I propose, are most readily explainable by decomposition, mineralization and nitrification processes, in concert with differences in pH along the experimental chronosequence. Invasion and eradication had substantial but again temporary implicati ons for AM fungal community structure. Influence of C ogongrass on S oil C hemistry Like many IA plants, cogongrass grows faster and produces more biomass than the native understory species that it displaces (Jose et al. 2002). C ogongrass is also suspected to have high nutrient use efficiency (Daneshgar and Jose 2009 a ), which contributes to the production of low quality tissues that decompose slowly (Bray 2005) Since the carbon cycle is intrinsically linked to other elemental cyc les, alterations in biomass production may affect the cycling of soil nutrients particularly macronutrients such as N and P that are frequently limiting. These differences in biomass production and tissue chemistry would seemingly lead to elevated soil O M levels (Ehrenfeld 2003), although this was not clearly evident in this study. It is difficult to assess why the expected trends were not observed but it is possible that the invaded plot s used for this study had not been impacted long enough for substan tial alterations to occur. There does not appear to be a characteristic trend of pH alteration in invaded systems, as studies show both decreases (Gremmen et al. 1998; Grierson and Adams 2000; Collins and Jose 2008) and increases (Hector et al. 1999; Ehren feld et al. 2001). The apparent lack of effect of cogongrass invasion on soil pH in this study, however, is
62 contrary to the findings of other cogongrass researchers. Collins and Jose (2008), for example, reported pH values in cogongrass invaded forest site s to be nearly one quarter unit lower than in uninvaded sites. While differential NH 4 and NO 3 uptake is commonly cited as an explanation for such effects (Ehrenfeld et al. 2001; Hewins and Hyatt 2010), the fact that I did not observe differences in pH, NO 2 +NO 3 or NH 4 availability between invaded and reference plot s suggests that this may not be the case for cogongrass. Cogongrass tissues have been shown to contain or exude a variety of different organic acids (Inderjit and Dakshini 1991; Koger and Bryson 20 04, Chapter 2 ) that may increase soil acidity but again this was not evident in this study. While the superior competitive ability of cogongrass (Collins and Jose 2008; Daneshgar and Jose 2009 a ; Holzmueller and Jose 2011) would seemingly lead to reductio ns in soil nutrient levels in invaded areas, total N, M1 P and resin adsorbed N and P in this study were not lower in invaded plot s relative to reference plot s In the case of N, it is likely that both systems have highly conservative cycles, in which inpu ts are limited (due to frequent fire and the preponderance of non N fixing vegetation) and that intense competition for this frequently limiting resource leads to internalized N cycling, and the maintenance of low levels of soil N (Chapin et al. 2002). It is possible that cogongrass more effectively captures soil P than do native species, perhaps by as has been observed with other species (Lambers et al. 2008; Perkins et al. 2011). H owever my findings, which showed no significant difference in soil P between invaded and reference plot s, do not support this assertion.
63 The P ost E radication L egacy of C ogongrass on N and P C ycling The different patterns of N and P mineralization that I o bserved in the litter bag study can be attributed to differences in tissue chemistry. The decomposition rates of dead plant tissues are largely controlled by their C:N ratios ( i.e. tissue quality) (Chapin et al. 2002). For low quality tissues ( e.g. C:N >25 :1), N immobilization occurs, for a time, and slow decomposition rates may cause litter to accumulate. For higher quality tissues, N mineralization occurs and decomposition is more rapid (Ehrenfeld 2003). Carbon:phosphorus ratios have less of an effect on decomposition rates, but like N, organic P immobilization and mineralization are governed by a tissue quality threshold (between 200 and 300:1) (Brady and Weil 2002). While P mineralized quite rapidly, for both above and belowground tissues, a substantial amount of N taken up by cogongrass remained immobilized after 18 months. Daneshgar and Jose (2009 a ) proposed that cogongrass establishes and maintains dominance in forest ecosystems by monopolizing the soil N pool and storing it in belowground tissues, wh ich constitute the bulk of total biomass. My findings partially support this claim, and further suggest that much of this N remains sequestered for an extended period of time following is often done, releases previously immobilized N into the atmosphere, potentially magnifying an N limitation (Daneshgar and Jose 2009 a ), although there may be a temporary increase in available forms of soil N (Certini 2005). Since the early stages of suc cession are often driven by N availability (Vitousek et al., 1993), an increased N limitation due to immobilization in cogongrass biomass (and atmospheric losses from burning) could affect the establishment and growth of desirable nitrophilic species immed iately after eradication. The spike in NO 2 +NO 3 at three years is
64 promote nitrification ( e.g. OM mineralization, elevated pH and temperature) along with decreases in fine root biomass (Attiwill and Adams 1993), likely resulted in a temporary increase in net nitrification. The subsequent decline was likely due to a combination of leaching, plant uptake and the development of conditions less conducive to nitrification. The ra pid decline in M1 and resin adsorbed P in the first three years after eradication is enigmatic, given the poor mobility of P in most soils ( Chapin et al. 2002), but it could be explained by the exploitation of this newly available pool by overstory pines in the absence of most competing understory vegetation. While little is known about post eradication nutrient cycling processes in invaded systems, it can be assumed that they are strongly tied following the decomposition of dead tissues to the effec ts of soil OM (Attiwill and Adams 1993; Tiessen et al. 1994). In terrestrial systems, the OM pool is constantly turning over, with measurable OM contents representing the balance between inputs ( e.g. litter) and outputs ( e.g. carbon mineralization) at a gi ven point in time (Chapin et al. 2002). In the sandy soils typical of longleaf pine systems, soil OM constitutes a major pool of potentially available nutrients (Wilson et al. 1999). My findings suggest that soil OM levels either were not significantly aff ected by invasion and eradication, or they equilibrated to near reference levels within three years of eradication In highly leached, acidic forest soils of humid regions, slight changes in pH can greatly alter the chemical form, solubility and mobility of N and P (Attiwill and Adams 1993). Soil pH has been shown to vary predictably with changes in vegetation during forest succession, with the highest pH values typically found in intermediate
65 successional seres (Christensen and Peet 1984). In this study, soil pH increased for five years following cogongrass eradication before declining to near reference levels by seven years. Since nitrification is inhibited at low pH (Chapin et al. 2002), the significant increases in pH observed for five years following eradication could have contributed to the elevated levels of resin adsorbed NO 2 +NO 3 observed across plot s Perhaps this increase in pH is also associated with burning (Raison 1979), as all five year plot s had been burned at least once since cogongrass erad ication. Arbuscular M ycorrhizal F ungal D ynamics While cogongrass has been shown to form associations with AM fungi ( Brook 1989 ), its mycorrhizal characteristics ( e.g. host/symbiont specificity, dependence) relative to the native species that it displaces is not known. Because of this, It was difficult to make informed assumptions about the effect that cogongrass invasion might have on AM fungal spore availability Research on other IA plants has shown that invasion can result in a reduction in AM fungal in oculum ( Roberts and Anderson 2001; Vogelsang and Bever 2009 ; Busby et al. 2012 ), but this was not evident in this study. If reductions occurr ed following eradication, the se effects were apparently short lived Since mycorrhizal fungal s pores are readily tr ansported by a variety of abiotic and biotic vectors (Sylvia 1986; Warner et al. 1987 ), spore counts in the recovery chronosequence may indicate rapid re dispersal following eradication This would support the findings of Anderson et al. (2010), who found that AM fungal inoculum potential rebounded soon after the eradication of IA plant populations. Spore longevity for AM fungi is not well known, but it is also possible that a persistent spore bank remained following eradication, even when host vegetation w as largely absent (Chapter 4).
66 Additionally, m y findings suggest that the r ichness and diversity of the AM fungal community were unaffected by cogongrass invasio n This is contrary to the findings of Busby (2011), who reported substantial reductions in AM fungal richness and diversity in semi arid steppe communities invaded by cheatgrass ( Bromus tectorum L.). To my knowledge, this study is the first to use molecular methods to assess post eradication effects on AM fungal richness and diversity in ecosystem s impacted by IA plants I f there was an effect of eradication, it was only temporary, as differences in richness and diversity were not evident by year three While the invaded and three year treatments were characteri zed by novel assemblages of AM fungal sequences, structural convergence occurred by year five The diversity and richness of the plant community in these same treatments however, converged later (year seven ) and remained markedly different from the reference in terms of species composition a nd community structure (Chapter 4). The fact that the mycorrhizal community, along with other soil properties, returned to a reference state prior to the plant community doing so suggests that belowground recovery might be a prerequisite for aboveground re covery (Anderson et al. 2010) Indeed, it is increasingly accepted that mycorrhizal fungi can play a significant role in terrestrial plant community succession (Janos 1980; Hartnett and Wilson 1999), particularly in resource poor soils (Gange et al. 1993) Summary and Implications This study represents one of the first attempts to assess the effects of both invasion and eradication of alien grasses on soil nutrient cycling processes. With the exception of alterations to AM fungal community struc ture my findings suggest that the effects of cogongrass invasion on soil properties in longleaf pine sandhill ecosystems are not substantial. However, considerable albeit temporary alterations to soil N and
67 P cycling as well as additional changes to the AM fungal community did occur in the years following eradication. Since the restoration of formerly invaded sites is a high priority among land managers, a logical next step is to determine the ecological significance of these effects as they pertain t o the re establishment either naturally or artificially of desirable native plant species within an acceptable time frame. This includes improving our understanding of how altered soil properties affect the potential for re invasion by cogongrass, as w ell as other problematic IA plant species.
68 Table 3 1. Number of study plots in each in each treatment x block x replication combination in longleaf pine sandhill communities in Hernando County, FL, USA. Treatment Replicate 1 Replicate 2 Block 1 Block 2 Block 1 Block 2 Reference 3 3 3 3 Invaded 3 3 1 3 3 years 2 3 3 2 5 years 3 2 3 3 7 years 3 3 3 3
69 Table 3 2 Initial mean tissue chemistry (standard deviations in parentheses) of herbicide treated cogongrass rhizomes and foliage, along with calcul ated k coefficients for mass loss and N and P mineralization over time. Tissue %N %P %C C:N C:P k biomass Foliage 0.94(0.29) 0.22(0.05) 43.43(0.56) 48.70(12.38) 207.70(44.72) 0.44(0.14) Rhizomes 0.43(0.14) 0.29(0.04) 45.06(0.60) 122.73(64.25) 147.91(5. 76) 1.01(0.09)
70 Table 3 3 P airwise treatment comparisons of AM fungal community structure generated using the weighted UniFrac significance test. Differences with P values < 0.05 are statistically significant. Reference Invaded 3yr 5yr 7yr Referenc e Invaded 0.00 3 yr 0.00 0.11 5 yr 0.06 0.05 0.07 7 yr 0.34 0.36 0.01 0.59
71 Figure 3 1. Schematic diagram of a replicate (not to scale), with formerly and currently invaded plot s indicated as dashed circles. R eference plot s three pe r block are randomly scattered in the uninvaded area among the other plot s Close up view of a plot indicates the location and arrangement of subplots Blocks are located in longleaf pine sandhill communities in Hernando County, FL, USA.
72 Figure 3 2. Mean soil organic matter content (%) in native reference plot s plot s currently invaded by cogongrass and plot s where cogongrass was eradicated three five and seven years prior in longleaf pine sandhill communities in Her nando County, FL, USA. Means and standard errors. Means having different lowercase letters are statistically different at P < 0.05.
73 Figure 3 3. Mean water extractable soil pH in native reference plot s plot s currently invaded by cogongrass and plot s where cogongrass was eradicated three, five and seven years prior in longleaf pine sandhill communities in Hernando County, FL, USA. Means and standard errors. Means having different lowercase letters are statistically diff erent at P < 0.05.
74 Figure 3 4. Mean total soil nitrogen (TKN method) (A), resin adsorbed soil ammonium (B) and nitrite+nitrate (C) in native reference plot s plot s currently invaded by cogongrass and plot s where cogongrass was eradicated three, five and seven years prior in longleaf pine sandhill communities in Hernando County, FL, USA.. Means and standard errors. Means having different lowercase letters are statistically different at P < 0.05. A B C
75 Figure 3 5. Mean Mehlich 1 (M1) extractable phosphorus (A) and resin extracted soil phosphorus (B) in native reference plot s plot s currently invaded by cogongrass and plot s where cogongrass was eradicated three, five and seven year s prior in longleaf pine sandhill communities in Hernando County, FL, USA. Means and standard errors. Means having different lowercase letters are statistically different at P < 0.05. A B
76 Figure 3 6. Patterns of mass loss and N and P mobilization/immobili zation for foliage (A) and rhizomes (B) of cogongrass treated with glyphosate and imazapyr herbicides. Means and standard deviations. A B
77 Figure 3 7. Summary statistics (Chao1 richness, Shannon index, 1 ex), for arbuscular mycorrhizal fungal communities generated in native reference plots plots currently invaded by cogongrass and plots where cogongrass was eradicated three, five and seven years prior in longleaf pine sandhill communities in Hernando Coun ty, FL, USA. Estimates were generated using the summary.single command in MOTHUR (cutoff = 0.03) : Chao1 Richness (A), Shannon Wiener index (B) and 1 Means and standard errors. Means having different lowercase letters are statistically different at P < 0.05 A B C
78 Figure 3 8. Condensed p h ylogenetic tree of arbuscular mycorrhizal fungal SSU rRNA genes based on a subset of 61 randomly selected sequences Interior nodes with bootstrap values less than 50 were collapsed. S equences from Genbank ( preceded with accession numbers) included for reference Sequences not assigned a genus and specific epithet had < 97% similarity with sequences in Genbank.
79 Figure 3 9. Weighted UniFrac PCA biplot illustrating separation in variable space between the mycorrhizal fungal community structure of 2 composite soil samples from each of the five treatments ( r eference = R; invaded = I; three y ea r = 3; five y ea r = 5; seven y ea r = 7) in longleaf pine sandhill communities in Hernando County, FL, USA.
80 CHAPTER 4 PATTERNS OF SECONDAR Y SUCCESSION FOLLOWI NG COGONGRASS ERADICATION IN A LON GLEAF PINE SANDHILL ECOSYSTEM Background The effects of I A plants on natural communities have been well documented in the ecological literature. These impacts include both alterations to plant community 1992; Gordon 1998; Mack et al. 2001; Hejda et al. 2009). Considerable research attention has also been paid to the species traits that confer invasiveness (Rejmanek and Richardson 1996) and the community characteristics that impart resistance or susceptibility to invasion (Elton 1958; Davis et al. 2000). On the management side, advancements in herbicide chemistry, coupled with research and field trials on integrated approaches have contributed to the development of highly effective species specific control strategies (Miller et al 2010). Relatively little attention, however, has been given to the recovery of native plant communities following the removal of problematic IA plant populations from the landscape. Successional theory suggests that the manipulation or reintroduction of natural processes that control disturbance, colonization and species performance can promote the development of robust native communities following the eradication of IA plant populations (Sheley et al. 2006; Sheley and Krueger Mangold 2011). This is a cr itical assumption, because the control of IA plants in natural areas is merely the first step in a restoration process that should also include the re establishment of desirable native plant species (Ogden and Rejmanek 2005, Miller et al. 2010). In some ca ses, recruitment of desirable species may occur with little to no assistance. Assuming there are no barriers to establishment, temporal shifts in resource availability should cause
81 early colonizers such as ruderals and legumes to ultimately give way to lat er successional species (Tilman 1985) like those that are typically targeted in restoration efforts. Unfortunately, empirical evidence to suggest that ecosystems can recover on their own following the eradication of an invasive plant species remains scant. To explain this, authors have suggested that legacy factors such as altered soil chemical or biological properties may lead to novel successional trajectories following the eradication of IA species (Yelenik et al. 2004; Wolfe and Klironomos 2005; Malcolm et be also hindered by dispersal limitation (Seabloom et al. 2003), particularly where seed longevity is short and/or remnant native populations are lacking (Clark et al. 2007). Cogongrass ( Imperata cylindrica ( L .) P. Beauv ), a pyrogenic C 4 rhizomatous 1977). In total, some 500 million hectares worldwide have some degree of cogongras s infestation. In the US, several hundred thousand hectares are infested (MacDonald 2004), with the current range overlapping much of the historic range of longleaf ( Pinus palustris Mill.) and slash pine ( Pinus elliottii Engelm) The sparse canopy that is characteristic of these forests, in concert with frequent fire, allows for high levels of understory diversity, but also makes them very susceptible to transformative impacts from cogongrass (Holzmueller and Jose 2011). Within a few years of invasion, near monocultures of cogongrass can dominate longleaf pine understories where species richness previously exceeded 20/m 2 (Hagan personal observation). Cogongrass also impedes pine regeneration (Richardson et al. 2007; Daneshgar et al. 2008).
82 Standard rates of common forestry herbicides have proven effective at controlling cogongrass in longleaf pine systems (Jose 2002), but it is not known if desirable native species will recolonize formerly invaded sites in the years following eradication. In a study of poten tial legacy effects (Chapter 3) I fou nd that alterations to soil N and P cycling processes develop following cogongrass eradication and persist for up to five years. Concurrent with these changes, I also observed changes in the assembly of the arbuscular mycorrhizal fungal community. Secondary invasions are also cause for concern following eradication (Symstad 2004; Yelenik et al. 2004). The objective of this study was to assess the patterns and possible drivers of secondary succession following the eradic ation of cogongrass in a longleaf pine sandhill ecosystem. Specifically, I sought to answer the following four questions: Do native species recolonize formerly invaded sites within seven years, or do novel community characteristics persist? What soil and environmental factors covary with the observed successional patterns? How does cogongrass eradication affect longleaf pine regeneration? Are formerly invaded sites susceptible to invasion by other alien plant species in the years following cogongrass erad ication? I hypothesized that formerly invaded sites would, by year seven begin to regain many of the vegetative characteristics of native reference sites. Specifically, I expected to see increases in total plant cover, increases in species richness and di versity, decreases in dominance and increases in the relative cover of desirable species such as wiregrass ( Aristida stricta Michx. var. beyrichiana Ward) and other pyrogenic native plants Shifts in community assembly, I hypothesized, would be associated with changes in soil resource availability. I expected that the elimination of cogongrass and
83 other competing vegetation would facilitate the establishment of longleaf pine seedlings, but would also lead to a secondary invasion of alien plant species, part icularly fast growing ruderals that are readily able to take advantage to a post eradication resource flux. Materials and Methods Study Area The study area was an uneven aged, naturally regenerated longleaf pine forest in the Croom Tract of Withlacoochee S tate Forest in Hernando County, Florida (2836'19.99"N, 8216'19.73"W). The tract is adjacent to one of the original points of cogongrass introduction in the United States and has a long history of cogongrass invasion. Efforts in recent years to chemically eradicate most cogongrass infestations in the tract have been successful and at the time of this study there were hundreds of areas in various stages of recovery throughout. The uninvaded matrix was characterized by high levels of understory species richn ess and diversity, as is typical of an actively managed, frequently burned longleaf pine sandhill community. Soils in the study area were predominantly deep, well drained to excessively drained sands of the Lake and Candler series (hyperthermic coated Typi c Quartzipsamments and hyperthermic uncoated Lamellic Quartzipsamments, respectively). Small inclusions of the Arredondo series (Loamy, siliceous, semiactive, hyperthermic Grossarenic Paleudults) comprising less than 20% of the total area were also pre sent (US Department of Agriculture, 1977). Mean basal area for the study area was 10.02 m 2 ha. Longleaf pine constituted approximately 89% of total basal area.
84 Experimental Design Across 4 stands in the study area, I used sites where cogongrass was eradic ated in previous years as a recovery chronosequence to assess temporal changes in plant community assembly following eradication. Uninvaded native understory sites, randomly selected from across these same stands, were used as a reference treatment. Sites selected for the chronosequence treatments were treated in the late summer/early fall approximately three five and seven years prior respectively with a tank mix solution (sprayed to the point of runoff) consisting of 2% Roundup plus surfactant) and 0. 4 such as this are among the most common and effective methods of chemical control for cogongrass (MacDonald 2004) 1 Study sites hereafter referred to as plots, were ide ntified using Geographic Information Systems (GIS) and with the help of state forest personnel. Native reference plot s were ground truthed to verify that they were not currently invaded and did not fall on disturbed or degraded sites ( e.g. roads, bicycle t rails, abandoned rock mines, formerly invaded sites). The study was laid out as a complete blocks design, replicated twice. Each replicate consisted of an adjacent pair of 259 ha stands (blocks) with similar, but asynchronous burn histories (both burned ap proximately every 4 years, but usually staggered 2 years apart). One block in each replicate was burned last in June 2009 and the other was burned in June 2007. Each block contained 2 3 plot s from each of the 4 treatments: reference ( i.e. uninvaded), three years since eradication, five years since 1 A single herbicide treat ment does not always completely eradicate a cogongrass patch. However, for young ( 2 year old) patches in the Croom tract, > 95% control is typical. For the purposes of this study,
85 eradication, and seven years since eradiation (Table 4 1). Within each plot three 3 m 2 vegetation sampling subplots were randomly selected, each being at least 2. 5 meters from the others at least 8 m from the ed ge of the plot and distant from any cogongrass re sprouts (where applicable) 2 Sampling Protocol In September 2010, plants in each vegetation sampling subplot were identified and the cover of each species was visually estimated. Coverage values were then used to compute relative cover ( P i ) for each species. Relative cover is defined as the percentage that a species contributes to the total cover of all species in a given location (Bazzaz 1975). Additionally, I quantified longleaf pine stem density (seedlin gs + saplings) in each subplot C over data were also used to determine species richness (mean number of species in three 3 m 2 subplots) and to calculate commonly used indices of community diversity and evenness. The Shannon Wiener diversity index ( H ), was calculated as follows: D ) was calculated as follows: For both indices, S represents the total number of species (Wilsey and Potvin 2000). For the purposes of this study, I expressed the Simpson index as 1 D so that higher values ( i.e. approaching 1) indicate greater species evenness (Jones et al. 2009). 2 Vegetation sampling subpl ots were centered on the soil sampling subplots from Chapter 3.
86 Species were classified by functional group ( i.e. forb, graminoid, tree, shrub, vine) using the growth habit categories from the USDA Plants Database ( U S Department of Agricultur e 2012 ). The same database was also used to determine the nativity status and persistence of the identified species. Plot values for plant cover ( including associated indices, described below) along with longleaf pine seedling and sapling stem counts, wer e calculated as the mean of the three subplot s Longleaf pine stem counts were converted to stems/m 2 for analysis. Data that met parametric assumptions were analyzed in SAS 9.2 (SAS Institute, Inc.) using the MIXED procedure. Replicate and block( replicate ) were treated as random effects. The Kenward Roger calculation, a preferred method for unbalanced mixed models (Spilke et al. 2005) was used to estimate denominator degrees of freedom 3 For the five treatments, differences between means were declared sta tistically significant at P hoc test was used for pairwise comparisons. Due to their non parametric nature ( i.e. the preponderance of zero or one values), comparisons of longleaf pine seedling/sapling stem count s, annual cover and n onnative cover were done with the Kruskal Wallis test using the Wilcoxon Mann Whitney test ( P < 0.05) for pairwise comparisons (SAS 9.2; NPAR1WAY procedure ). A Mantel test (PC ORD 5 ; McCune and Grace 2002) was used to compare a matrix of species P i values to a secondary matrix of selected soil and environmental variables (Table 4 2) from Chapter 3 This procedure tests the null hypothesis of no relationship between matrices (McCune and Grace 2002) After the Mantel test indicated a relationship ( P = 0. 0 25 ) I compared plant community assembly between treatments using a canonical discriminant analysis (CDA; JMP 9) based on the 3 This method can result in non integer values for denominator degrees of freedom.
87 P i values of the predominant species ( P i 1%) in reference plot s Variables from the secondary matrix were included in this analysis. The CDA is an eigenanalysis technique in which variables are used to predict group membership (McCune and Grace 2002). In this case, the groups were the four trea tments. Dominant variables (standardized scoring coefficients > |1|) that influenced the two most important canonical axis are reported. For a detailed description of the soil sampling and analysis protocol for the variables used in the Mantel test and CDA see Chapter 3 Results Cover, Richness and Diversity Total percent plant groundcover showed a significant increase ( F ( 3, 35.07 ) = 15.6 2 P < 0.0001) in the years following cogongrass eradication. Mean cover averaged 20.2 % after three years, and rose to 36. 9 % and 44.9% after five and seven years, respectively. By year seven cover was not significantly different from the reference treatment (52.3%). Shannon Wiener diversity indices ( ) showed a pattern of increasing understory plant diversity with increa sing time since cogongrass eradication ( F ( 3, 35.31 ) = 23.35 P < 0.0001). Specifically, values were significantly lower than reference (2.01) three and five years post eradication (1. 3 2 and 1. 3 9 respectively) before recovering to near reference levels by seven years (2.07). The distinct increase in that occurred between years five and seven was statistically significant A similar trend was observed for D ), with the values observed after three and five years (0.64 and 0.64, respe ctively) being significantly lower (less evenness) than the seven year (0.80) and reference (0.77) treatments ( F ( 3, 35.35 ) = 11.10 P < 0.0001). Species richness also increased with time, again being lowest at three and five years (6. 02 and 7 1 4 species /pl ot ) and increasing significantly by year seven (13.05 species/plot ) ( F ( 3,35.10 ) =
88 53.35, P < 0.0001). The latter was not significantly different from the reference treatment (14.71) (Table 4 3 ). Relative Groundcover by Growth Habit and Persistence Forb co ver, as a percent of total cover, was highest five years after eradication ( 61.8 %), followed by three years ( 51.2 %), seven years (38.2%) and reference (24.8%) ( F (3, 35.10) = 12.80 P < 0.00 0 1). Relative graminoid cover was highest after three years ( 3 5 8 %) followed by seven years (36. 5 %), five years (22. 7 %) and reference (12.8%) ( F ( 3, 35.09 ) = 8.27 P = 0.0003 ) Understory tree cover, on the other hand, showed a different trend, increasing from 3.8 % at three years to 12.1 % and 14.9% and five and seven year s, respectively. All of these coverage values, however, were significantly lower than reference (48.0%) ( F ( 3, 35.21 ) = 36.27 P < 0.00 0 1). Shrub cover also increased after eradication (4. 5 %, 2.9 % and 8.4% for three five and seven year plots, respectively; ( F ( 3, 35.18 ) = 3.89 P = 0. 0 168 ), the latter not being significantly different from reference (12. 7 %). Vine cover averaged 1.9% and did not vary significantly between treatments (Figure 4 1 ). Relative cover of annual vegetation was lower in the native ref erence plot s (2.2%) than in formerly invaded plot s (8.1, 8.2 and 6. 7 % in three five and seven year plot s respectively) but these differences were not statistically significant ( chi 2 = 4.954 P = 0.1 752 ) Dominant Species A total of 100 plant species wer e identified in the understory of the study plot s Seventy eight of these species were found in the native reference plot s Of these 78, 23 had relative groundcovers greater than or equal to 1%. Ten of these species were forbs, five were graminoids, five w ere trees and three were shrubs. All 23 species were perennials. Bluejack oak ( Quercus incana W. Bartram) was the most dominant species
89 in reference plot s ( Table 4 4). The majority of the dominant species in formerly invaded plot s were graminoids and forbs one of which (S etaria corrugata (Elliot) (prevalent in three five and seven year plot s ) was an annual. Dogfennel ( Eupatorium capillifolium (Lam.) Small ex Porter and Britton) was the most dominant species three and five years following eradication and S corrugata was the dominant species after seven years (Table 4 5 ). Multivariate Analyses The CDA analysis was quite robust (0% treatment misclassification) and indicated considerable variability between treatments ( P < 0.0001). The first two canonical axe s cumulatively accounted for 94. 7 % of the total variation (77.0 and 17. 7 % respectively). For axis 1, the most influential variables, based on standardized discriminant function coefficients, were Quercus incana ( 2.14), organic matter (1.78), Q. laurifolia ( 1.52), TKN ( 1.43), Desmodium sp. (1.21), Q. margaretta ( 1.11), pH (1.05) and M1 P ( 1.03). For axis 2, the most influential variables were arbuscular mycorrhizal spores ( 1.90), pH (1.50), Polygala sp. ( 1.20), Diospyros virginiana (1.19) and M1 P (1. 16). Overall, there was no clear indication that formerly invaded plot s begin to approach a reference state over time (Figure 4 2A ). A different pattern, however, was observed when woody species (which managers may consider undesirable) were eliminated fro m the analysis. This modification still resulted in a high degree of treatment separation ( P < 0.0001 ), but distance be tween treatments particularly between reference plot s and formerly invaded plot s was much less distinct. Axes 1 and 2 cumulatively ac counted for 93.4% of total variation (70.6 and 22.8%, respectively). For axis 1, the most influential variables were Polygala sp. (1.58), arbuscular mycorrhizal spores (1.57), M1 P ( 1.48), Paspalum
90 sp. (1.37) and organic matter (1.05). For axis 2, the mos t influential variable was Eupatorium capillifolium (1.07) (Figure 4 2B ). Longleaf Pine Regeneration Longleaf pine regeneration (seedling and sapling stems /m 2 ) varied greatly among the four treatments. Stem counts were highest five and seven years followi ng cogongrass eradication ( 0.37 and 0. 26/m 2 respectively) and lowest in reference plot s (0. 02 ). T hese differences were statistically significant ( chi 2 = 17.84 P = 0.000 5 ) Stem counts in three year plot s (0. 14 ) were not significantly different from the o ther treatments (Figure 4 3 ). Non n ative Species All 23 of the most prevalent species in the study plot s were native, but non native plant species were found in all treatments. A total of two non native plant species, both legumes (white clover ( Trifolium repens L.) and hairy indigo ( Indigofera hirsuta L.)), were identified. Relative nonnative cover (both species combined) was 1.4% in treated plots and 0.1% in reference plot s Mean relative cover of T. repens ranged from 0 to 11.9%, with a mean of 0. 7 %. Wh ile T. repens cover was highest in plot s where cogongrass was eradicated three years prior (1.9% vs. 0. 7 0. 5 and 0.01% for five year, seven year and reference plot s respectively), these differences were not statistically significant ( chi 2 = 3.5214 P = 0 .2678). For I. hirsuta relative cover ranged from 0 to 6. 3 % with a mean of 0. 3 %. This species was only found in seven year and reference plot s (0.1% and trace, respectively) and differences between these treatments were not statistically significant ( chi 2 = 0.4943 P = 0. 4820 ).
91 Discussion Understory Community Assembly The late summer/early fall application of a glyphosate + imazapyr herbicide tank mix effectively eliminated nearly all vegetation (including remnant natives) from cogongrass invaded longleaf pine sandhill sites (Hagan, personal observation). It can be assumed, therefore, that the plants observed in treated plot s originated from the soil seed bank, recruitment/encroachment from adjacent unimpacted areas, or the persistence of scattered individ uals not killed as a by product of cogongrass treatment. Steady increases in plant cover following eradication can be attributed to these factors, along with the growth and spread of newly established individuals (Huston and Smith 1987). The fact that tota l understory plant cover for the first five years remained significantly lower than in the reference treatment suggests that there was still available space for additional recruitment and expansion. It is also arguably the most readily observable indicatio n that the effects of cogongrass invasion and eradication on native sandhill plant communities persist for several years. Along with the availability of sites for establishment, species availability and species performance are ultimately what dictate which species colonize following disturbance (Pickett et al. 1987). Typically, early successional seres are characterized by low species diversity and richness and high species dominance (low evenness) (Huston and Smith 1987). This is consistent with the post e radication findings from the first five years of recovery. The significant increase in diversity by year seven was associated with an increase in species richness and an increase in evenness, but the ecological explanation for this spike is unknown Simila r levels of diversity, evenness and richness between the seven year and reference plot s suggest that the complexity
92 though not composition of formerly invaded sites approaches reference levels over time. Patterns and Environmental Covariates of Specie s Colonization Successional theory suggests that shifts in understory composition in the years following cogongrass eradication reflect differences in plant functional strategies and changes in resource availability (Tilman 1985; Grime 1985). The graminoi ds and forbs which dominated the early stages of succession were likely the species that were best suited to rapidly capitalize on a post eradication resource flux. After seven years, many of these species were still dominant, perhaps due to multiple burns which helped to maintain a subclimax state. While some recruitment of trees and shrubs occurred following eradication, the significant reductions in woody species cover in formerly invaded plot s particularly non pyrogenic species like oaks may be view ed favorably by restoration ecologists (Provencher et al. 2001; Walker and Silletti 2006). The relative lack of wiregrass cover in treated plot s (cover < 1.1% three five and seven years post eradication), however, is problematic, since it is considered a keystone species for the longleaf pine ecosystem (Noss 1989; Mulligan et al. 2002). Reductions in fuel connectivity, caused by the decreased herbaceous component and reductions in total groundcover, may have altered the behavior of the low intensity ground fires that are considered essential for the maintenance of these systems (Landers 1989). Additionally the loss of nutrient rich woody browse may be detrimental to populations of native ungulates (Pearson and Sternitzke 1976). Multivariate analytical tech niques such as CDA provide a useful index to visually and quantitatively assess complex ecological questions and have proven useful in chronosequence studies (Matthews 1979; Stylinski and Allen 1999; Frouz et al. 2008).
93 In this study while there was disti nct separation between the four treatments, there was little evidence that formerly invaded plot s began to approach a reference state in the first seven years following eradication, especially when woody species are included in the model. Stylinski and All en (1999) reported a similar trend in severely degraded shrubland ecosystems in California. These authors attributed the lack of native species recovery to the severity of disturbance (a combination of anthropogenic soil disturbance and alien plant invasio ns) that their study sites had experienced. The passing of a resistance threshold, they proposed, lead to the development of an alternative stable state characterized by novel species assemblages. While my study plot s had not been subject to severe disturb ance, perhaps the invasion and subsequent eradication of a functionally novel grass had a similar effect on successional processes. Indeed, substantial but temporary, alterations to soil biogeochemistry were shown to develop in the years following cog ongrass eradication ( Chapter 3 ) and the results of the multivariate analyses suggest that these alterations to the soil environment may play a role in determining successional trajectories Most studies on longleaf pine regeneration have focused on the e ffects of competition from overstory trees (Brockway and Outcalt 1998; McGuire et al. 2001). In uninvaded areas throughout the Croom Tract, large numbers of longleaf pine seedlings and young saplings can typically only be found in canopy gaps where substan tial reductions in basal area have occurred due to lightning or disease. My l ongleaf pine s tem counts, however, highlight the role of understory competition High seedling and sapling stem counts for longleaf pine in formerly invaded plot s relative to r ef erence plot s suggest that the conditions suitable for regeneration are enhanced following
94 cogongrass eradication. Since juvenile longleaf pines are known to be poor competitors for water, nutrients and light (Jose et al. 2003) it is likely that the elimi nation of cogongrass and most other competing vegetation helped facilitate the establishment of this desirable overstory species. N o evidence of cogongrass re invasion was observed in the study plot s or in other formerly invaded sites across the Croom Tra ct. Relative covers of other nonnative species were generally (albeit not significantly) higher in treated plot s than in reference plot s but this does not appear to constitute a secondary invasion. Both T. repens and I. hirsuta are naturalized throughout Florida ( US Department of Agriculture 2012 ) and neither is listed by the Florida Exotic Pest Plant Council as species that is likely to substantially alter native plant communities ( FLEPPC 2009). Both species were introduced to the study area many years pr ior as forage crops (Vincent Morris, personal communication) and hardseededness and the presence of dispersal agents ( e.g. deer, horses, etc .) likely permitted their spread outside of the original areas of cultivation (Sulas et al. 2000). Their presence in formerly invaded plot s may simply reflect the successional status of these plot s, as legumes are one of the functional groups most characteristic of early secondary succession in longleaf pine systems, partly due to a nitrogen limitation exacerbated by fr equent fire (Lajeunesse et al. 2006). A companion study (Chapter 3 ), however, showed little evidence of a post eradication nitrogen limitation This could indicate that dispersal limitations impede the reestablishment of nitrophilic plant species. Summary and Implications Cogongrass is becoming a major threat to the ecological integrity of longleaf pine ecosystems in the southeastern US. As such, there is a growing need to develop
95 effective strategies for the restoration of native understory comm unities following cogongrass eradication. By shedding light on the successional dynamics of longleaf pine sites formerly invaded by cogongrass, the seven year post eradication chronosequence provides valuable information about the feasibility of passive re generation as a restoration option. However, the results of this study only partially support the original hypotheses. While the diversity and complexity of formerly invaded plots increased in the years following eradication, the species composition remain ed markedly different from that of native reference plots If recovery is occurring, it is likely proceeding at a slower than desirable pace. The substantial reductions in woody species cover observed following cogongrass eradication may, however, be viewe d favorably by some land managers. As expected, differences in community assembly were, to an extent, associated with variability in soil properties. The presence of large numbers of longleaf pine seedlings and saplings in formerly invaded plot s is a posit ive sign and suggests that the removal of cogongrass (and most other understory vegetation) alleviates a substantial recruitment limitation. The lack of a secondary invasion, which is contrary to one hypothesis, is also encouraging. Future studies should s eek to determine the specific soil and environmental factors that limit the dispersal or recruitment of desirable native species. Manipulative studies and a longer term chronosequence would also be beneficial.
96 Table 4 1. Number of study plot s in each i n each treatment x block x replication combination in longleaf pine sandhill communities in Hernando County, FL, USA. Treatment Replicate 1 Replicate 2 Block 1 Block 2 Block 1 Block 2 Reference 3 3 3 3 3 years 2 3 3 2 5 years 3 2 3 3 7 years 3 3 3 3
97 Table 4 2. Landscape and soil variables for a canonical discriminant analysis (CDA), used along with the relative covers of the 23 most dominant plant species in reference plot s to assess the patterns secondary succession in plot s formerly invaded by cogongrass. Abbreviations provided. Variable Abbreviation Mean (SE) Basal area (m 2 ) BA 10.02 (0.63) Soil pH pH 5.49 (0.03) Total soil P (mg/kg) M1 P 113.18 (4.67) Total soil N (mg/kg) TKN 955.25 (22.31) Soil organic matter (%) OM 1.67 (0.08) Mycorrhizal inoculum (spores/g) Spores 12.54 (1.22)
98 Table 4 3. Mean values for the Shannon Wiener ( D ) Indices and species richness ( mean number of species in three 3 m 2 subplots ) in plot s where cogongrass was eradicated three five and seven years prior along with uninvaded native reference plot s in longleaf pine sandhill stands in Hernando County, FL, USA. Means and standard errors. Means having different lowercase letters are statistically different at P < 0.05. Treatment H' 1 D Richness T otal cover (%) 3 years 1. 3 3 a 0.64 a 6. 02 a 20.20 a (0.08) (0.03) (0. 64 ) ( 3.62 ) 5 years 1. 3 9 a 0.64 a 7. 15 a 36. 91 b (0.08) (0.03) (0. 61 ) (3. 4 6 ) 7 years 2.07 b 0.80 b 13.05 b 44.92 bc (0.08) (0.03) (0. 5 8 ) (3. 2 9 ) Reference 2.01 b 0.77 b 14.71 c 52.33 c (0.08) (0.03) (0. 5 8 ) (3. 29 )
99 Table 4 4. Mean relative cover of the 23 most prevalent understory species (relative cover > 1%) in reference plot s compared to their relative covers in plot s where cogongrass w as eradicated three five and seven years prior, in longleaf pine sandhill stands in Hernando County, FL, USA. CDA abbreviations provided. Species CDA Growth habit Treatment 3 y r 5 y r 7 y r Ref Quercus incana W. Bartram QuI Tree 0.00 1.37 3.88 20.40 Quercus margaretta Ashe ex Small QuM Tree 0.00 0.57 1.58 10.95 Quercus laurifolia Michx. QuA Tree 1.35 1.49 2.78 10.04 Morella cerifera L. MoC Shrub 0.74 0.00 2.84 6.65 Pteridium aquilinum (L.) Kuhn var. pseudocaudatum (Clute)Clute PtA Forb 0.48 3.43 5. 45 5.22 Quercus virginiana Mill. QuV Tree 0.40 0.73 0.92 4.29 Aristida stricta Michx. var. beyrichiana Ward ArS Graminoid 0.15 0.73 1.08 2.65 Rubus argutus Link RuA Shrub 2.06 0.25 1.66 2.30 Dichanthelium laxiflorum (Lam.) Gould DiL Graminoid 4.09 1.53 3.10 2.15 Paspalum sp. Pas Graminoid 9.01 2.99 5.68 2.11 Elephantopus carolinianus Raeusch. ElC Forb 2.32 0.51 1.92 1.97 Eupatorium capillifolium (Lam.) Small ex Porter and Britton EuC Forb 13.68 36.62 9.82 1.89 Dyschoriste oblongifolia (Michx.) Kuntz e DyO Forb 1.49 1.07 1.59 1.81 Dichanthelium aciculare (Desv. ex Poir.) Gould and C.A.Clark DiA Graminoid 1.38 0.85 2.41 1.65 Diospyros virginiana L. DiV Tree 0.00 1.25 0.78 1.57 Rhus copallinum L. RhC Shrub 0.12 1.75 0.31 1.34 Desmodium sp. Des Forb 0 .31 0.87 1.36 1.31 Stillingia sylvatica L. StS Forb 0.20 0.17 0.83 1.23 Sorghastrum secundum (Elliott) Nash SoS Graminoid 0.44 2.81 1.18 1.18 Eupatorium pilosum Walter EuP Forb 7.36 0.24 0.55 1.16 Desmodium floridanum Chapm. DeF Forb 0.14 0.19 0.23 1.0 6 Polygala sp. Pol Forb 2.31 0.54 0.53 1.02 Galactia regularis (L.) Briton et al. GaR Forb 1.56 0.23 0.83 1.01
100 Table 4 5. Top 5 dominant species in plot s where cogongrass was eradicated three five and seven years prior in longleaf pine sandhill stands in Hernando County, FL, USA. Abbreviations and r elative covers (%) in parentheses. See Table B 1 ( A ppendix ) for full names. Rank Treatment 3 years 5 years 7 years 1 Eupatorium capillifolium (F, P) (13.68) Eupatorium capillifolium (F, P) (36.62) Setar ia corrugata (G, A) (10.03) 2 Paspalum sp. (G, P) (9.01) Pinus palustris (T, P) (6.11) Eupatorium capillifolium (F, P) (9.82) 3 Setaria corrugata (G, A) (7.93) Setaria corrugata (G, A) (4.57) Andropogon arctatus (G, P) (8.11) 4 Eupatorium pilosum (F, P) (7.36) Rhynchosia michauxii (F, P) (3.81) Paspalum sp. (G, P) (5.68) 5 Andropogon arctatus (G, P) (5.61) Pteridium aquilinum (F, P) (3.43) Pteridium aquilinum (F, P) (5.48) A = annual; F = forb; G = graminoid; P = perennial; T = tree
101 Figure 4 1 Mean relative cover (%) by growth habit type in plot s where cogongrass was eradicated three, five and seven y ears prior, a long with uninvaded native reference plot s in longleaf pine sandhill stands in Hernando Co unty, FL, USA. Means having different lowercase letters are statistically different at P < 0.05. ab a bc c ab a a b a a a b b a a a b a a a a Time since eradication
102 Figure 4 2 Canonical discriminant analysis (CDA) bi plot (with 95% confidence circles) of the patterns of compositional similarity between native referen ce plot s and plot s where cogongrass was eradicated three, five and seven years prior (A). Additional CDA with woody species removed (B). C ircles represent the 95% confidence intervals for each treatment. Each point represents one plot Abbreviations are pr ovided in Tables 4 2 and 4 4. A B
103 Figure 4 3 Longleaf pine stems per m 2 in plot s where cogongrass was eradicated three, five and seven years prior, along with uninvaded native reference plot s in longleaf pine sandhill s tands in Hernando County, FL, USA. Means and standard errors. Lowercase letters denote the results of the post hoc Wilcoxon Mann Whitney pairwise comparisons. Pairs with different lowercase letters are statistically different at P < 0.05
104 CHAPTER 5 CONCLU SIONS Cogongrass ( Imperata cylindrica ( L ) P. Beauv. ) invasion is a significant and growing threat to the integrity of pine ecosystems in the southeastern United States. This fast growing C 4 rhizomatous grass readily displaces native understory forbs, gram inoids and shrubs (Daneshgar et al. 2008) impedes the regeneration of commercially and ecologically valuable overstory trees (Daneshgar et al. 2009 a ) and alters fire behavior (Lippincott 2000) Little is known, however, about the belowground mechanisms th at might help explain the transformative success of this species Our understanding of post eradication legacy factors and successional processes is also lacking. In light of these deficiencies, I conducted a series of studies to bolster our understanding on the effects of cogongrass invasion and eradication on soil properties and to assess the patterns and possible drivers of native plant community recovery following eradication. Cogongrass is suspected to have allelopathic properties ( Abdul Wahab and Al Naib 1972; Hussain and Abidi 1991; Inderjit and Dakshini 1991; Koger and Bryson 2004; Xuan et al. 2009 ) but the specific compounds it produces, and their mechanisms of action on susceptible plants, were previously unknown In Chapter 2, a greenhouse study, I hypothesized that rhizosphere leachate collected from cogongrass pot cultures would adversely affect the growth, root morphology and mycorrhizal colonization of native species (relative to leachate collected from mixed natives). Additionally, I hypothes ized that compounds not present in a native savanna rhizosphere would be present in the cogongrass rhizosphere. My results indicated an apparent allelopathic effect from cogongrass, although it varied by species. A ruderal grass ( Andropogon
105 arctatus Chapm. ) and ericaceous shrub ( Lyonia ferruginea (Walter) Nutt. ) were unaffected by the cogongrass leachate, while mid successional grass ( Aristida stricta Michx. var. beyrichiana (Trin. and Rupr.) D.B.Ward ) and the tree ( Pinus elliottii Engelm. ) were negatively affected. For A. stricta I observed a 35. 7 % reduction in aboveground biomass, a 22.2 % reduction in total root length, a 2 2 9 % reduction in specific root length and a 23. 4 % reduction in total mycorrhizal root length, relative to the native leachate treatme nt. For P. elliottii there was a 19. 4 % reduction in percent mycorrhizal colonization and a 21.8 % reduction in total mycorrhizal root length. Comparisons made with a DI water control in the second year support the notion that the observed differences were due to the negative effects of cogongrass leachate. My chemical analyses identified 12 putative allelopathic compounds (mostly phenolics) in the cogongrass leachate. The concentrations of most of these compounds were significantly lower if they were found at quantifiable levels, in the native leachate. One compound was a novel alkaloid. The speculated structure was hexadecahydro 1 azachrysen 8 yl ester ( C 23 H 33 NO 4 ) and it appeared to be present at fairly high levels. This compound was not found in the nativ e leachate treatment. Researchers have suggested that cogongrass alters soil nutrient dynamics in southern pine ecosystems ( Collins and Jose 2008; Daneshgar and Jose 2009a ) In Chapter 3 I assessed pre and post eradication soil biogeochemical dynamics i n longleaf pine sandhill stands severely impacted by cogongrass. Across a seven year post and native reference plot s I analyzed soils for total N (TKN), potentially available P (Mehlich 1), pH and organic matter content. I also used a resin bag technique to assess
106 fluxes of plant available N and P in the soil solution. Since nutrient cycling following eradication may be influenced by the turnover of herbicide treated biomass, I used litterbags to monitor the decomposition and nutrient mineralization patterns of rhizomes and foliage. I also used spore counts and molecular techniques (PCR, cloning and sequencing) to characterize changes to the AM fungal community. My results indic ate similar total N and M1 P contents in invaded and reference plot s with levels of M1 P being lower than in invaded plot s for five years following eradication. Soil organic matter content was highest in cogongrass invaded plots an d lowest seven years fol lowing eradication Resin bag analyses suggest that cogongrass invasion did not affect soil 2 +NO 3 occurred in the first three years following eradication. No such trends were observed for ammon ium. Resin adsorbed PO 4 was lowest three years following eradication and pH was highest five years following eradication. The litterbag study showed that approximately 55% of foliar biomass and 23% of rhizome biomass remained 18 months after herbicide trea tment. Substantial N immobilization was observed in rhizomes for the first 12 months, with slow mineralization occurring thereafter. Rapid P mineralization occurred for both tissues, with 15.4 and 20. 5 % of initial P remaining after 18 months in rhizomes an d foliage, respectively. Neither cogongrass invasion nor eradication affected AM fungal diversity, richness or spore counts Substantial alterations to AM fungal community assembly, however, occurred due to invasion, with novel community characteristics pe rsisting for an additional three years following eradication I suggest that f uture research should assess the extent to which the sum of these changes affect the re establishment of
107 desirable native species, as well as the potential for re invasion b y cog ongrass or other IA plant species. The re establishment of native plant cover following IA plant removal is considered essential for the long term control of cogongrass in natural areas (Miller et al. 2010). Natural regeneration, if effective, may be an a ttractive option for many land managers. In Chapter 4, I used the post eradication chronosequence to assess patterns of secondary succession following cogongrass eradication. I hypothesized that the plant community assembly of formerly invaded plot s would begin to approach reference state within seven years. R esults revealed a general pattern of increasing richness, diversity and evenness in the years following eradication. For all three measures, there was a distinct and statistically significant change be tween years five and seven By year seven cover diversity and richness were not statistically different from the native reference treatment, Despite this apparent recovery, there was no clear evidence that the composition of formerly invaded plot s even after seven years approached a reference state, unless woody species were removed from the analysis. Soil properties ( e.g. organic matter, mycorrhizal spores and pH) appeared to correlate with successional patterns. Dogfennel ( Eupatorium capillifolium (Lam.) Small ex Porter and Britton) was the most dominant species three and five years following eradication and Setaria corrugata (Elliott) was the dominant species after seven years. Bluejack oak ( Quercus incana W. Bartram) was the most dominant species in reference plot s Longleaf pine regeneration was enhanced following eradication ( 0.37 and 0. 26 stems/ m 2 five and seven years post eradication vs. 0. 0 2 in reference ). Nonnative
108 legumes were found in all treatments, but it does not appear that a secondary invasion occurred following cogongrass eradication. My findings provide insight into the ecological dynamics of southern pine ecosystems impacted by cogongrass. The differences in leachate chemistry between cogongrass and native species coupled with the negative effects observed on wiregrass and slash pine suggest that allelopathy contributes to the alterations in plant community assembly that have been observed in cogongrass invaded southern pine ecosystems T he fact that soil chemical and arbuscular m ycorrhizal properties return to a reference state within five to seven years of cogongrass eradication is encouraging, as it indicates that post eradication legacy effects are short lived. The recovery of soil properties before native plant communities sug gests that belowground processes may influence ecological succession following eradication. Dispersal limitation of desirable species, however, is a possibility that should be addressed in the future, perhaps via manipulative studies that evaluate the perf ormance of reintroduced native plants
109 APPENDIX A RELATIVE ABUNDANCE O F EACH OF THE 31 AM FUNGAL OTUS, BY TREATMENT Table A 1. Relative abundance of each of the 31 arbuscular mycorrhizal fungal OTUs (operational taxonomic units) identified from soil s col lected from plot s where cogongrass was eradicated 3, 5 and 7 years prior, currently invaded plot s and uninvaded native reference plot s in longleaf pine sandhill stands in Hernando County, FL, USA. ACA* ACA ACA ACA ACA ACA ACA ACA ACA GIG GIG GIG GIG GIG GLO GLO GLO GLO GLO GLO GLO GLO GLO GLO GLO GLO PAR PAR PAR PAR PAR 8** 9 10 11 12 13 14 15 16 27 28 29 30 31 17 18 19 20 21 22 23 24 25 26 32 33 1 2 3 4 7 Reference 0.00 0.02 0.02 0.00 0.15 0.02 0.02 0.00 0.00 0.00 0.12 0.00 0.00 0.17 0.12 0.00 0.00 0.00 0.06 0.06 0.00 0.02 0.00 0.00 0.03 0.02 0.00 0.18 0.00 0.00 0.02 Invaded 0.03 0.00 0.03 0.03 0.07 0.00 0.00 0.00 0.00 0.03 0.14 0.00 0.03 0.04 0.01 0.01 0.00 0.01 0.00 0.01 0.00 0.00 0.01 0.00 0.07 0.04 0.00 0.40 0.00 0.03 0.00 3 yr 0.00 0.00 0.00 0 .00 0.00 0.02 0.00 0.00 0.00 0.00 0.13 0.00 0.00 0.00 0.17 0.02 0.03 0.02 0.02 0.00 0.02 0.03 0.00 0.05 0.00 0.05 0.02 0.40 0.00 0.03 0.00 5yr 0.00 0.00 0.00 0.00 0.11 0.02 0.00 0.02 0.02 0.00 0.15 0.02 0.00 0.08 0.11 0.00 0.03 0.00 0.04 0.02 0.00 0.00 0. 00 0.00 0.00 0.04 0.00 0.28 0.02 0.06 0.00 7yr 0.00 0.00 0.00 0.00 0.13 0.02 0.01 0.00 0.00 0.00 0.08 0.00 0.00 0.05 0.15 0.01 0.07 0.00 0.00 0.00 0.00 0.00 0.00 0.06 0.05 0.05 0.02 0.28 0.00 0.04 0.00 *Family abbreviation: ACU= Acaulosporaceae GIG=Giga sporaceae, GLO=Glomeraceae, PAR=Paraglomeraceae ** Numeric OTU ID provided by MOTHUR
110 APPENDIX B SPECIES LIST Table B 1. Complete list of all plant species identified in plot s where cogongrass was eradicated 3, 5 and 7 years prior and uninvaded native refe rence plot s in longleaf pine sandhill stands in Hernando County, FL, USA. Species Family Growth habit Aeschynomene viscidula Michx. Fabaceae Forb Ambrosia artemisiifolia L. Asteraceae Forb Andropogon arctatus Chapm. Poaceae Graminoid Andropogon virgin icus L. var. glaucus Hack. Poaceae Graminoid Aristida stricta Michx. var. beyrichiana Ward Poaceae Graminoid Asclepias tuberosa L. Apocynaceae Forb Asimina pygmea (W.Bartram) Dunal Annonacee Shrub Astragalus obcordatus Elliott Fabaceae Forb Astragalus villosus Michx. Fabaceae Forb Baccharis halimifolia L. Asteraceae Tree Balduina angustifolia (Pursh) B.L.Rob. Asteraceae Forb Chamaecrista fasciculata (Michx.) Greene Fabaceae Forb Clitoria fragrans Small Fabaceae Forb Cnidoscolus stimulosus (Michx.) Engelm. and A.Gray Euphorbiaceae Forb Crotalaria rotundifolia Gmelin Fabaceae Forb Croton argyranthemus Michx. Euphorbiaceae Forb Croton michauxii G.L.Webster Euphorbiaceae Forb Desmodium floridanum Chapm. Fabaceae Forb Desmodium paniculatum (L.) DC. Fabaceae Forb Dichanthelium aciculare (Desv. ex Poir.) Gould and C.A.Clark Poaceae Graminoid Dichanthelium laxiflorum (Lam.) Gould Poaceae Graminoid Diospyros virginiana L. Ebenaceae Tree Dyschoriste oblongifolia (Michx.) Kuntze Acanthaceae Forb
111 Tab le B 1 Continued Species Family Growth habit Elephantopus carolinianus Raeusch. Asteraceae Forb Eupatorium capillifolium (Lam.) Small ex Porter and Britton Asteraceae Forb Eupatorium pilosum Walter Asteraceae Forb Galactia regularis (L.) Britton et al Fabaceae Vine Galium pilosum Aiton Rubiaceae Forb Gelsemium sempervirens (L.) Aiton F. Gelsemiaceae Vine Helianthus hirsutus Raf. Asteraceae Forb Hieracium megacephalon Nash Asteraceae Forb Houstonia procumbens (J.F. Gmelin) Rubiaceae Forb Hyperic um hypericoides (L.) Crantz Clusiaceae Shrub Hypericum punctatum Lam. Clusiaceae Shrub Ilex opaca Aiton Aquifoliaceae Tree Indigofera hirsuta L. Fabaceae Forb Itea virginica L. Iteaceae Shrub Lespedeza hirta (L.) Hornem. Fabaceae Shrub Licania michau xii Prance Chrysobalanaceae Shrub Lobelia homophylla E.Wimm. Campanulaceae Forb Lygodesmia aphylla (Nuttall) de Candolle Asteraceae Forb Mimosa quadrivalvis L. var. angustata (Torr. and A.Gray) Barneby Fabaceae Vine Morella cerifera L. Myricaceae Shrub Opuntia humifusa (Raf.) Raf. Cactaceae Shrub Parthenocissus quinquefolia (L.) Planch. Vitaceae Vine Passiflora incarnata L. Passifloraceae Vine Pinus clausa (Chapm. ex Engelm.) Vasey ex Sarg. Pinaceae Tree Pinus palustris Mill. Pinaceae Tree Pityops is graminifolia (Michx.) Nutt. Asteraceae Forb Plantago major L. Plantaginaceae Forb
112 Table B 1 Continued Species Family Growth habit Pseudognaphalium obtusifolium (L.) Hilliard and B.L.Burtt Asteraceae Forb Pteridium aquilinum (L.) Kuhn var. pseudo caudatum (Clute) Clute ex A.Heller Dennstaedtiaceae Fern Pterocaulon pycnostachyum (Michx.) Elliott Asteraceae Forb Quercus incana W. Bartram Fagaceae Tree Quercus laevis Walter Fagaceae Tree Quercus laurifolia Michx. Fagaceae Tree Quercus margaretta Ashe ex Small Fagaceae Tree Quercus nigra L. Fagaceae Tree Quercus sp Fagaceae Tree Quercus virginiana Mill. Fagaceae Tree Rhus copallinum L. Anacardiaceae Shrub Rhynchosia michauxii Vail Fabaceae Forb Rubus argutus Link Rosaceae Shrub Rudbeckia hi rta L. Asteraceae Forb Ruellia caroliniensis (J.F.Gmel.) Steud. Acanthaceae Forb Setaria corrugata (Elliott) Schult. Poaceae Graminoid Sisyrinchium angustifolium Mill. Iridaceae Forb Smilax spp. Smilacaceae Vine Solanum chenopodioides Lam. Solanaceae Forb Sorghastrum nutans (L.) Nash Poaceae Graminoid Sorghastrum secundum (Elliott) Nash Poaceae Graminoid Sporobolus junceus (P.Beauv.) Kunth Poaceae Graminoid Stillingia sylvatica L. Euphorbiaceae Forb Trichostema setaceum Houtt. Lamiaceae Forb Trif olium repens L. Fabaceae Forb Unknown Andropogon Poaceae Graminoid Unknown aster Asteraceae Forb
113 Table B 1 Continued Species Family Growth habit Unknown Carex Cyperaceae Graminoid Unknown Cyperus 1 Cyperaceae Graminoid Unknown Cyperus 2 Cyperacea e Graminoid Unknown Cyperus 3 Cyperaceae Graminoid Unknown Cyperus 4 Cyperaceae Graminoid Unknown Desmodium sp. Fabaceae Forb Unknown grass 1 Poaceae Graminoid Unknown grass 2 Poaceae Graminoid Unknown Hypericum sp. Clusiaceae Shrub Unknown Ipomoea sp. Convolvulaceae Vine Unknown Paspalum sp. Poaceae Graminoid Unknown Polygala sp. Polygalaceae Forb Unknown Pseudognaphalium sp. Asteraceae Forb Unknown Rhynchospora sp. Cyperaceae Graminoid Unknown Scleria sp. Cyperaceae Graminoid Unknown Solidago sp. Asteraceae Shrub Vaccinium arboreum Marshall Ericaceae Shrub Vaccinium darrowii Camp Ericaceae Shrub Vaccinium myrsinites Lam. Ericaceae Shrub Vaccinium stamineum L. Ericaceae Shrub Vitis cinerea (Engelm.) Engelm. ex Millardet var. floridana Muns on Vitaceae Vine Vitis rotundifolia Michx. Vitaceae Vine
114 LIST O F REFERENCES Abdul Wahab AS, Al Naib FAG (1972) Inhibitional effects of Imperata cylindrica (L) PB. Bull lraq Nat Hist Mus 5:17 24 Abhilasha D, Quintana N, Vivanco J, Joshi J (200 8) Do allelopathic compounds in invasive Solidago canadensis sl restrain the native European flora? J Ecol 96:993 1001 Allison SD Vitousek PM ( 2004 ) Rapid n utrient c ycling in l eaf l itter from i nvasive p lants Oecologia 141 : 612 619 Anderson RC e t al. ( 2010 ) Effect of removal of garlic mustard ( Alliaria petiolata Brassicaceae) on arbuscular mycorrhizal fungi inoculum potential in forest soils. Open Ecol J 3:41 47 Ashton IW, Hyatt LA Howe KM Gurevitch J Lerdau MT ( 2005 ) Invasive s pecies a ccele rate d ecomposition and l itter n itrogen l oss in a m ixed d eciduous forest. Ecol Appl 15 : 1263 1272 Attiwill PM, Adams MA ( 1993 ) Tansley Review No 50 Nutrient c ycling in f orests New Phytol 124 : 561 582 Bais HO, Vepachedu R, Gilroy S, Callaway RM, Vivanco JM (2 003) Allelopathy and exotic plant invasion: From molecules and genes to species interactions. Science 301:1377 1380 Bais HP, Weir TL, Perry LG, Gilroy S, Vivanco JM (2006) The role of root exudates in rhizosphere interactions with plants and other organi sms. Annu Rev Plant Biol 57:233 266 Bakker J, Wilson S (2001) Competitive abilities of introduced and native grasses. Plant Ecol 157:119 127 Bazzaz FA (1975) Plant Species diversity in old field successional ecosystems in southern Illinois. Ecology 56:48 5 488 Bending GD, Read DJ (1996) Nitrogen mobilization from protein polyphenol complex by ericoid and ectomycorrhizal fungi. Soil Biol Biochem 28:1603 1612 Bending GD, Read DJ (1997) Lignin and soluble phenolic degradation by ectomycorrhizal and ericoid m ycorrhizal fungi. Mycol Res 101:1348 1354 Binkley D ( 1984 ) Ion e xchange r esin b ags: f actors a ffecting e stimates of n itrogen a vailability Soil Sci Soc Am J 48 : 1181 1184
115 Binkley D, Aber J Pastor J Nadelhoffer K ( 1986 ) Nitrogen availability in some Wiscons in forests: comparisons of resin bags and on site incubations Bio l Fert Soils 2 : 77 82 Blossey B, Notzold R (1995) Evolution of increased competitive ability in invasive nonindigenous plants: a hypothesis. J Ecol 83:887 889 Blum U (1998) Effects of microbi al utilization of phenolic acids and their phenolic acid breakdown products on allelopathic interactions. J Chem Ecol 24:685 708 Blum U, Staman KL, Flint LJ, Shafer SR (2000) Induction and or selection of phenolic acid utilizing bulk soil and rhizosphere b acteria and their influence on phenolic acid phytotoxicity. J Chem Ecol 26:2059 2078 Brady NC Weil RR ( 2002 ) The Nature and Properties of Soils 13 th ed n. Pearson, New Jersey Bray S R ( 2005 ) Interactions between plants and soil mi crobes in Florida communit ies: Implications for invasion and ecosystem ecology Dissertation, University of Florida Brewer S (2008) Declines in plant species richness and endemic plant species in longleaf pine savannas invaded by Imperata cylindrica. Biol Invasions 10:1257 1264 Br ockway DG, Outcalt K (1998) Gap phase regeneration in longleaf pine wiregrass ecosystems. Forest Ecol Manag 106:125 139 Brook RM (1989) Review of literature on Imperata cylindrica (L.) Raeuschel with particular reference to South East Asia. Trop Pest Ma nag 35:12 25 Burns RM, Honkala BH (1990) Silvics of North America: 1 Conifers Agriculture Handbook 654 US Department of Agriculture, Forest Service, Washington, DC Busby RR (2011) Interactions between plants and soil mi crobes in Florida communities: Implications for invasion and ecosystem ecology Dissertation, Colorado State University Busby RR, Paschke MW, Stromberger ME, Gebhart DL (2012) Seasonal variation in arbuscular mycorrhizal fungi root colonization of cheatgrass ( Bromus tectorum ) an invasi ve winter annual J Ecosyst Ecog http://dx.doi.org/10.4172/2157 7625.S8 001 Callaway RM, Aschehoug ET (2000) Invasive plants versus their new and old neighbors: A mechanism for exotic invasion. S cience 290:521 523 Callaway RM, Ridenour WM (2004) Novel weapons: invasive success and the evolution of increased competitive ability. Front Ecol Environ 2:436 443
116 Certini G (2005) Effects of fire on properties of forest soils: a review Oeocolgia 143:1 10 Chapin FS, Matson PA Mooney HA ( 2002 ) Terrestrial N utrient C ycling In: Chapin FS, Matson PA, Mooney HA (eds) Principles of Terrestrial Ecosystem Ecology Springer, New York, p 197 223 Christensen NL Peet RK ( 1984 ) Convergence d uring s econdary f orest s ucc ession J Ecol 72 : 25 36 Chung IM, Kim KH, Ahn JK, Chun SC, Kim CS, Kim JT, Kim SH (2002) Screening of allelochemicals on barnyardgrass ( Echinochloa crus galli ) and identification of potentially allelopathic compounds from rice ( Oryza sativa ) variety hull e xtracts. Crop Prot 21:913 920 experimental test using cogongrass. Biol Invasions 9:433 443 Collins AR Jose S (2008) Imperata cylindrica an exotic invasive grass, changes so il chemical properties of forest ecosystems in the Southeastern United States In: Kohli RK, Jose S, Singh HP, Batish DR (eds) Invasive Plants and Forest Ecosystems CRC Press, Boca Raton, pp 237 247 CM ( 2004 ) Effects of e xotic s pecies on s oil n itrogen c ycling: i mplications for r estoration Weed Technol 18 : 1464 1467 Daniels BA and Skipper HA (1982) Methods for the recovery and quantitative estimation of propagules from soil. In Methods and Principles of Mycorrhizal Research. Ed. N C Sc henck. pp. 29 35. American Phytopathological Society, St. Paul, MI, USA CM, Vitousek PM ( 1992 ) Biological invasions by exotic grasses, the grass/fire cycle, and global change Annu Rev Ecol Syst 23 : 63 87 Daehler CC (2003) Performance comparison s of co occurring native and invasive alien plants: implications for conservation and restoration. Annu Rev Ecol Evol Syst 34:183 211 Daneshgar P, Jose S (2009 a ) Imperata cylindrica an invasive alien grass, maintains control over nitrogen availability in an establishing pine forest. Plant Soil 320:209 218 Daneshgar P, Jose S (2009b) Role of species identity in plant invasions: experimental test using Imperata cylindrica Biol Invasions 11:1431 1440 Daneshgar P, Jose S, Collins A, Ramsey C (2008) Cogongrass ( Imperata cylindrica ), an invasive alien grass, reduces survival and productivity of an establishing pine forest. Forest Sci 54:579 587
117 Davis MA, G rime JP Thompson K ( 2000 ) Fluctuating resources in plant communities: a general theory of invasibility J E col 88 : 528 534 EDDMapS (2012) Early Detection & Distribution Mapping System. The University of Georgia Center for Invasive Species and Ecosystem Health. Ehrenfeld JG (2003) Effects of exotic plant invasions on soil nutrient cycling processes. Ecosystems 6:503 523 Ehrenfeld JG, Kourtev P Huang W ( 200 1) Changes in s oil f unctions f ollowing i nvasions of e xotic u nderstory p lants in d eciduous f orests Ecol Appl 11 : 1287 1300 Ehrenfeld JG, Scott N (2001) Invasive species and the soil: effects on organisms and ecosystem processes. Ecol Appl 11:1259 1260 Einhellig FA (1995) Mechanism of action of allelochemicals in allelopathy. In: Allelopathy ACS Symposium Series American Chemical Society, pp 96 116 Einhellig FA (2002) The physiology of allelochemicals actio n: clues and views In: Reigosa, M, and Pedrol, N (eds) Allelopathy: From Molecules to Landscapes Science Publishers, Inc pp 1 24 Erskine Ogden JA and Rejmanek M (2005) Recovery of native plant communities after the control of a dominant invasive plant spe cies, Foeniculum vulgare : Implications for management. Biological Cons erv 125:427 439 Feller IC, McKee KL Whigham DF JP ( 2003 ) Nitrogen vs phosphorus limitation across an ecotonal gradient in a mangrove forest Biogeochemistry 62 : 145 175 FLEPP C (2011) List of Invasive Plant Species. Florida Exotic Pest Plant Council. Internet: http://www.fleppc.org/list/11list.htm or Wildland Weeds 14:11 14. Summer/Fall 2011 Frey Klett P, Garbaye J, Tarkka M (2007) The mycorrhiza helper bacteria revisited. New Phytol 176:22 36 Frouz J et al (2008) Interactions between soil development, vegetation and soil fauna during spontaneous succession in post mining sites. Eur Jou r Soil Bio l 44:109 121 Gange AC, Bro wn VK Sinclair GS (1993) Vesicular arbuscular mycorrhizal fungi : a determinant of plant community structure in early succession Funct Ecol 7:616 622 Gibson D ( 1986 ) Spatial and temporal heterogeneity in soil nutrient supply measured using in situ ion exch ange resin bags Plant Soil 96 : 445 450
118 Gmez Aparicio L, Canham CD (2008) Neighbourhood analyses of the allelopathic effects of the invasive tree Ailanthus altissima in temperate forests. J Ecol 96:447 458 Gordon D ( 1998 ) Effects of invasive, non indigenou s plant s pecies on ecosystem processes: lessons from Florida Ecol Appl 8 : 975 989 Gremmen NJM, Chown SL Marshall DJ ( 1998 ) Impact of the introduced grass Agrostis stolonifera on vegetation and soil fauna communities at Marion Island, sub Antarctic Biolog ical Conserv 85 : 223 231 Grierson PF, Adams MA ( 2000 ) Plant species affect acid phosphatase, ergosterol and microbial P in a Jarrah ( Eucalyptus marginata Donn ex Sm) forest in south western Australia Soil Biol Biochem 32 : 1817 1827 Grime JP ( 1988 ) The C S R model of primary plant strategies: origins, implications, and tests In: Gottlieb LD, Jain SK (eds) Plant Evolutionary Biology Chapman and Hall, London pp 371 393 Hnfling B, Kollmann J (2002) An evolutionary perspective of biological invasions. Tren ds Ecol Evol 17:545 546 Harpole WS, Tilman D ( 2007 ) Grassland species loss resulting from reduced niche dimension Nature 446 : 791 793 Hartman KM McCarthy BC ( 2004 ) Restoration of a f orest u nderstory a fter the r emoval of an i nvasive s hrub, A mur h oneysuckle ( Lonicera maackii ) Restor Ecol 12 :1 54 165 Hartnett DC, Wilson GWT (1999) Mycorrhizae influence plant community structure and diversity in tallgrass prairie. Ecology 80:1187 1195 Hector A et al ( 1999 ) Plant d iversity and p roductivity e xperiments in E urope an g rasslands Science 286 : 1123 1127 V ( 2009 ) Impact of invasive p lants on the species richness, diversity and composition of invaded communities J Ecol 97:393 403 Hewins D, Hyatt L ( 2010 ) Flexible N uptake and assimilation mech anisms may assist biological invasion by Alliaria petiolata Biol Invasions 12 : 2639 2647 Hierro JL, Callaway RM (2003) Allelopathy and exotic plant invasion. Plant Soil 256:29 39 Holly DC, Ervin GN (2006) Characterization and quantitative assessment of int erspecific and intraspecific penetration of below ground vegetation by cogongrass ( Imperata cylindrica (L) Beauv) rhizomes. Weed Biol Manag 6:120 123
119 Holm LG, Pucknett DL, Pancho JB, Herberger JP (1977) Weeds Distribution and Biology Un iv Press of Hawaii, Honolulu, HI Holzmueller EJ, Jose S (2011) Invasion success of cogongrass, an alien C4 perennial grass, in the southeastern United States: exploration of the ecological basis. Biol Invasions 13:435 442 Hussain F, Abidi N (1991) Allelo pathy exhibited by Imperata cylindrica (L). Beauv P Pak J Bot 23:15 25 Huston M, Smith T, (1987) Plant succession : life history and competition Am Nat 130:168 198 Inderjit, Dakshini KMM (1991) Investigations on some aspects of chemical ecology of cogo ngrass, Imperata cylindrica (L) Beauv. J Chem Ecol 17:343 352 Izhaki I (2002) Emodin a secondary metabolite with multiple ecological functions in higher plants. New Phytol 155:205 217 Janos DP (1980) Mycorrhizae influence tropical succession. Biotropica 12:56 64 Jeffries P, Gianinazzi S, Perotto S, Turnau K, Barea JM (2003) The contribution of arbuscular mycorrhizal fungi in sustainable maintenance of plant health and soil fertility. Biol Fert Soils 37:1 16 Jones MD, Durall DM, Tinker PB (1998) A comp arison of arbuscular and ectomycorrhizal Eucalyptus coccifera : growth response, phosphorus uptake efficiency and external hyphal production. New Phyt ol 140 :125 134 Jones PD, Edwards SL, Demarais S and Ezell AW (2009) Vegetation community responses to different establishment regimes in loblolly pine ( Pinus taeda ) plantations in southern MS, USA. Forest Ecol Manag 257:553 560 Jordan N, Larson D Huerd S ( 2008 ) Soil modification by invasive plants: effects on native and invasive species of mixed grass p rairies Biol Invasions 10 : 177 190 Jose S, Cox J, Miller DL, Shilling DG, Merritt S (2002) Alien plant invasions: the story of cogongrass in southeastern forests. J Forest 100:41 44 Jose S, Merritt S, Ramsey CL (2003) Growth, nutrition, photosynthesis and transpiration responses of longleaf pine seedlings to light, water and nitrogen. Forest Ecol Manag 180:335 344 Koger CH, Bryson CT ( 2004 ) Effect of c ogongrass ( Imperata cylindrica ) e xtracts on g ermination and s eedling g rowth of s elected g rass and b road leaf s pecies Weed Technol 18 : 236 242
120 Korb JE, Johnson NC, Covington WW (2003) Arbuscular mycorrhizal propagule densities respond rapidly to ponderosa pine restoration treatments. Journal of Appl Ecol 40:101 110 Kourtev PS, Ehrenfeld JG, Hggblom M (200 2) Exotic plant species alter the microbial community structure and function in the soil. Ecology 83:3152 3166 Kourtev PS, Ehrenfeld JG, Hggblom M (2003) Experimental analysis of the effect of exotic and native plant species on the structure and function of soil microbial communities. Soil Biol Biochem 35:895 905 Lajeunesse SD et al (2006) Ground layer carbon and nitrogen cycling and legume nitrogen inputs following fire in mixed pine forests. Am J Bot 93:84 93 Lambers H et al (2008) Plant nutrient acqui sition strategies change with soil age. Trends Ecol Evol 23:95 103 Landers JL, Byrd NA, Komarek R (1990) A holistic approach to managing longleaf pine communities In Farrar Jr RM (ed) Proceedings of the symposium on m anagement of longleaf pine, USDA Fore st Service General Technical Report SO 75 Southern Forest Experiment Station, New Orleans, LA Larkin MA et al (2007) Clustal W and Clustal X version 2.0. Bioinformatics 23:2947 2348 Lee J, Lee S, Young JPW (2008) Improved PCR primers for the detection and identification of arbuscular mycorrhizal fungi. FEMS Microbiol Ecol 65:339 349 Lekberg Y et al (2007) Role of niche restrictions and dispersal in the composition of arbuscular mycorrhizal fungal communities. J Ecol 95:95 105 Li Q, Allen HL, Wilson CA ( 2003) Nitrogen mineralization dynamics following the establishment of a loblolly pine plantation. Can J Forest Res 33:364 374 Lippincott CL (2000) Effects of Imperata cylindrica (L) Beauv (Cogongrass) invasion on fire regime in Florida sandhill (USA). Nat Area J 20:140 149 diversity measures lead to different insights into factors that structure microbial communities. Applied Environ Microb 73:1576 1585 MacDonald GE (2004) Cogo ngrass ( Imperata cylindrica ) Biology, ecology, and management. Crit Rev Plant Sci 23:367 380 Dynamics by Exotic Plants: A Case Study of C4 Grasses in Hawaii. Ecol App l 11:1323 1335
121 M alcolm GM, Bush DS, Rice SK (2008) Soil Nitrogen Conditions Approach Preinvasion Levels following Restoration of Nitrogen Fixing Black Locust ( Robinia pseudoacacia ) Stands in a Pine Oak Ecosystem. Rest Ecol 16:70 78 Mallik AU (2000) Challenges and oppor tunities in allelopathy research: a brief overview. J Chem Ecol 26:2007 2009 Manoharachary C, Kunwar IK (2002) Root clearing techniques and quantification of arbuscular mycorrhizal fungi In: Mukerji KG, Manoharachary C, and Chamola, BP (eds) Techniques in Mycorrhizal Studies Klewer Academic Publishers, Netherlands, pp 231 248 Maron JL Vil M (2001) When do herbivores affect plant invasion? Evidence for the Natural Enemies and Biotic Resistance hypotheses Oikos 95:361 373 Maron JL, Jeffries RL (2001) Rest oring enriched grasslands: effects of mowing on species richness, productivity and nitrogen retention. Ecol Appl 11:1088 1100 Matthews JA (1979) A study of the variability of some successional and climax plant assemblage t ypes using multiple discriminant analysis J Ecol 67:255 271 McCarthy BC, Hanson SL (1998) An assessment of the allelopathic potential of the invasive weed Alliaria petiolata (Brassicaceae). Castanea 63:68 73 McCune B, Grace JB (2002) Analysis of Ecological Communities MJM Software Des ign, Gleneden Beach, Oregon McGuire JP et al (2012) Gaps in a gappy forest: plant resources, longleaf pine regeneration, and understory response to tree removal in longleaf pine savannas. Can J For Res 31:765 778 Miller JH, Manning ST, Enloe SP (2010) A Management Guide for Invasive Plants in Southern Forests. United States Department of Agriculture, Forest Service, Southern Research Station Mulligan MK, Kirkman LK, Mitchell RJ (2002) Aristida beyrichiana (Wiregrass) Establishment and Recruitmen t: Implications for Restoration. Rest Ecol 10:68 76 Mylavarapu RS Moon DL ( 2002) UF/IFAS extension soil testing laboratory (ESTL) analytical procedures and training manual. Circular 1248 Soil and Water Science Department, Florida Cooperative Extension Service, Institute of Food and Agriculture Science, University of Florida Norsworthy JK (2003) Allelopathic potential of wild radish ( Raphanus raphanistrum ). Weed Technol 17:307 313
122 Noss RF (1989) Longleaf pine and wiregrass keystone components of an end angered ecosystem Nat Areas J 9:211 213 Perkins L, Johnson D, N owak R ( 2011 ) Plant induced changes in soil nutrient dynamics by native and invasive grass species Plant Soil 1 10 Perry LG, Thelen GC, Rienour WM, Callaway RM, Paschke MW, Vivanco JM (2007) Concentrations of the allelochemical (+/ ) catechin in Centaurea maculosa soils. J Chem Ecol 33:2337 44 Pickett STA, Collins SL, Armesto JJ (1987) Models, Mechanisms and Pathways of Succession. Bot Rev 53:335 371 Pimentel D, Zuniga R, Morrison D (2005) Update on the environmental and economic costs associated with alien invasive species in the United States. Ecol Econ 52:273 288 Pringle A, Bever JD, Gardes M, Parrent JL, Rillig MC, Klironomos JN (2009) Mycorrhizal symbioses and plant invasions. Annu Re v Eco Evol Sys 40:699 715 species environmental change and management and health Annu Rev Environ Resour 35:25 55 Raison RJ (1979) Modification of the soil environment by vegetation fires, with particular refe rence to nitrogen transformations: a review. Plant Soil 51:73 108 Ramsey CL, Jose S, Miller DL, Cox J, Portier KM, Shilling DG, Merritt S (2003) Cogongrass response to herbicides and disking on a cutover site and in a mid rotation pine plantation in so uthern USA. Forest Ecol Manag 179:195 207 Redecker D (2002) Molecular identification and phylogeny of arbuscular mycorrhizal fungi. Plant Soil 244:67 73 Redecker D, Raab P (2006) Phylogeny of the Glomeromycota ( arbuscular mycorrhizal fungi ): recent develo pments and new gene markers Mycologia 98:885 895 Reed ML (1989) Ericoid mycorrhizas of styphelieae: intensity of infection and nutrition of the symbionts Aust J Plant Physiol 16 :155 160 Reigosa MJ, Snchez Moreiras A, Gonzlez L (1999) Ecophysiological ap proach in allelopathy. Crit Rev Plant Sci 18:577 608 Rejmanek M, Richardson DM (2012) What attributes make some plant species more invasive ? Ecology 77:1655 1661 Renz MJ, Blank RR ( 2004 ) Influence of perennial pepperweed ( Lepidium latifolium ) biology and plant soil relationships on management and restoration Weed Technol 18:1359 1363
123 Richardson D et al (2007) Human impacts in pine forests: past, present, and future. Annu Rev Ecol Evol Syst 38:275 297 tta FD, West CJ (2000) Naturalization and invasion of alien plants: concepts and definitions. Diversity Distrib 6:93 107 Richardson DR, Williamson GB (1988) Allelopathic effects of shrubs of the sand pine scrub on pines and grasses of the sandhills. Forest Sci 34:592 605 Roberts KJ Anderson RC (2001) Effect of garlic mustard [ Alliaria petiolata (Beib. Cavara & Grande)] extracts on plants and arbuscular mycorrhizal (AM) fungi. Am Midl Nat 146:146 152 Snchez Moreiras A, Weiss O, Reigosa Roger M (2003) Alle lopathic evidence in the Poaceae. Bot Rev 69:300 319 SAS Institute ( 2007) The SAS system for Windows Release 9 20 SAS Institute, Cary, NC Sayed WF, El Sharouny HM, Zahran HH, Ali WM (2002) Composition of Casuarina leaf litter and its influence on Franki a Casuraina symbiosis in soil. Folia Microbiol 47:429 434 Schaalje GB, McBride JB, Fellingham GW (2002) Adequacy of approximations to distributions of test statistics in complex mixed linear models. J Agric Biol Envir S 7:512 524 Schloss PD (2008) Evaluati ng different approaches that test whether microbial communities have the same structure. ISME J 2:265 275 Seabloom EW et al (2003) Competition, seed limitation disturbance and reestablishment of California native annual forbs Ecol Appl 13:575 592 Sea l AN, Pratley JE, Haig T (2004) Identification and quantitation of compounds in a series of allelopathic and non allelopathic rice root exudates. J Chem Ecol 30:1647 1662 Sheley RL, Krueger Mangold J (2011) Principles for restoring invasive plant infested rangeland. Weed Sci 51:260 265 Sheley RL, Mangold JM, Anderson JL (2006) Potential for successional theory to guide restoration of i nvasive plant dominated Rangeland. Ecol Monogr 76:365 379 Simberloff D (2005) Non native species DO threaten the natura l environment! J Agr Environ Ethic 18:595 607
124 Singh HP, Batish DR, Pandher JK, Kohli RK (2005) Phytotoxic effects of Parthenium hysterophorus residues on three Brassica species. Weed Biol Manag 5:105 109 Smith SE, Read DJ (1997) Mycorrhizal Symbiosis. Else vier Academic Press, St. Louis Souto C, Pellissier F, Chiapusio G (2000) Allelopathic effects of humus phenolics on growth and respiration of mycorrhizal fungi. J Chem Ecol 26:2015 2023 Spilke J, Piepho HP, Hu X (2005) Analysis of unbalanced data by mixe d linear models using the MIXED Procedure of the SAS System. J Agron Crop Sci 191:47 54 Standish RJ, Williams PA, Robertson AW, Scott NA, Hedderley DI (2004) Invasion by a perennial herb increases decomposition rate and alters nutrient availability in war m temperate lowland forest remnants. Biol Invasions 6:71 81 Stylinski CD, Allen EB (1999) Lack of native species recovery following severe exotic disturbance in southern Californian shrublands. J Appl Ecol 36:544 554 Suzuki Y, Esumi Y, Hyakutake H, Kono Y, Sakurai A (1996) Isolation of 5 heptadecenyl) resorcinol from etiolated rice seedlings as an antifungal agent. Phytochemistry 41:1485 1489 Sylvia DM (1986) Spatial and temporal distribution of vesicular arbuscular mycorrhizal f ungi associated with Uniola paniculata in Florida foredunes Mycologia 78:728 734 Symstad AJ (2004) Secondary invasion following the reduction of Coronilla varia (Crownvetch) in sand prairie Am Midl Nat 152:183 189 Tamura K et al (2011) MEGA5: Molecular evolutionary geneti cs analysis using maximum likelihood evolutionary distance, and maximum parsimony methods. Mol Biol Evol 28:2731 2739 Tennant D (1975) A test of a modified line intersect method for estimating root length. J Ecol 63:995 1001 Thiffault N et al (2000) Washin g procedure for mixed bed ion exchange resin decontamination for in situ nutrient adsorption. Commun Soil Sci Plan 31:543 546 Tiessen H, Cuevas E Chacon P ( 1994 ) The role of soil organic matter in sustaining soil fertilit y. Nature 371 : 783 785 Tilman D (19 85) The resource ratio hypothesis of plant succession. Am Nat 125:827 852
125 Uren NC (2007) Types, amounts, and possible functions of compounds released into the rhizosphere by soil grown plants. In: Pinton, R, Varanini, Z, and Nannipieri, P (eds) The Rhizosp here: Biochemistry and Organic Substances at the Soil Plant Interface 2 nd edn. CRC Press, Boca Raton, FL US Department of Agriculture (1977) Soil Survey of Hernando County, Florida US Department of Agriculture, Natural Resources Conservation Service U S Department of Agriculture ( 1985) Soil survey of Alachua County, Florida US Department of Agriculture, Natural Resources Conservation Service US Department of Agriculture (2012) The PLANTS Database ( http://plants.u sda.gov 22 May 2012). National Plant Data Team, Greensboro, NC 27401 4901 USA Vitousek PM et al (1993) Nutrient limitations to plant growth during primary succession in Hawaii Volcanoes National Park. Biogeochemistry 23:197 215 Vitousek PM, Walker LR, Wh iteaker LD, Mueller Dombois D, Matson PA (1987) Biological invasion by Myrica faya alters ecosystem development in Hawaii. Science 238: 802 804 Vivanco JM, Bais HP, Stermitz FR, Thelen GC, Callaway RM (2004) Biogeographical variation in community respo nse to root allelochemistry: novel weapons and exotic invasion. Ecol Lett 7:285 292 Vogelsang KM, Bever JD (2009) Mycorrhizal densities decline in association with nonnative plants and contribute to plant invasion. Ecology 90:399 407 Walker JL, Silletti A (2006) Restoring the groundcover of longleaf pine ecosystems In: Jose S, Jokela E, Miller DL (eds) The Longleaf Pine Ecosystem: Ecology, Silviculture, and Restoration Springer Science, New York, pp 297 333 Wardle DA, Nilsson M, Gallet C, Zackrisson O (1 998) An ecosystem level perspective of allelopathy. Biol Rev Camb Philos Soc 73:305 319 Warner NJ, Allen MF, MacMahon JA (1987) Dispersal agents of vesicular arbuscular mycorrhizal fungi in a disturbed arid e cosystem. Mycologia 79:721 730 Willamson M, Fitt er A (1996) The varying success of invaders. Ecology 77:1661 1666 Wilsey BJ, Potvin C (2000) Biodiversity and ecosystem functioning: importance of species evenness in an old field. Ecology 81:887 892 Wolfe BE, Klironomos, JN (2005) Breaking new ground : s oil communities and e xotic plant invasion BioScience 55:477 487
126 Xuan DT, Toyama T, Fukata M, Khanh TD, Tawata S (2009) Chemical interaction in the invasiveness of cogongrass ( Imperata cylindrica (L) Beauv). J Agric Food Chem 57:9448 9453 Yelenik SG, Sto ck WD, Richardson DM (2004) Ecosystem level impacts of invasive Acacia saligna in the South African Fynbos. Restor Ecol 12:44 51 Zavaleta ES, Hobbs RJ Mooney HA ( 2001 ) Viewing invasive species removal in a whole ecosystem context Trends Ecol Evol 16 : 454 459
127 BIOGRAPHICAL SKETCH the sandhills, swamps and bayous of his native Florida Panhandle. In 2002, He earned e nvironmental s tudies from the University o f West Florida. From 2004 to 2006, he served as a Peace Corps agroforestry extensionist in Ecuador, where he worked with landowners to preserve some of the last remnants of coastal dry tropical forest. Upon returning t o the U.S., he enrolled in the i nterdi sciplinary e cology program at the University of Florida graduating in 2008 From 2008 to 2012 he was a Ph.D. A lumni F ellow in the School of Forest Resources and Conservation at the University of Florida