The Effect of Plant Patch Size and Spatial Pattern on Biodiversity, Ecosystem Functions, and Grassland Community Structure

Shannon E. Seahra

A Thesis presented to The University of Guelph

In partial fulfilment of requirements for the degree of Doctor of Philosophy in Environmental Biology

Guelph, Ontario, Canada

ABSTRACT

THE EFFECT OF PLANT PATCH SIZE AND SPATIAL PATTERN ON BIODIVERSITY, ECOSYSTEM FUNCTIONS, AND GRASSLAND COMMUNITY STRUCTURE

Shannon E. Seahra Advisor: University of Guelph, 2015 Professor J. A. Newman

The interactions between plants that determine competition and coexistence are strongly influenced by their fine-scale spatial pattern. These interactions also ultimately influence diversity, function, and structure in plant communities. In areas that have undergone extensive land-use or anthropogenic degradation, such as North American grasslands, there is a critical need to understand how spatial patterning of plant species can be manipulated to maximize diversity and restoration success. The research presented in this thesis employed a novel planting strategy using conspecific patch sizes at seeding to spatially manipulate inter- and intraspecific interactions among native grassland plant species. I found that seeded patch size had strong effects on diversity maintenance, biodiversity effects, productivity, and invasion, and that the typical uniformly mixed seeding approach in biodiversity ecosystem function studies and restoration applications may not maximize these responses. Smaller patch plots tend to have strong selection effects from dominant forb species, although this may change over time. Additionally, there were species-specific and plant functional group-specific responses to patch size that should be considered in restoration of low-diversity sites. The initial fine-scale plant pattern of had measurable effects on the spatial abundance of species, functional groups, and invasion. Finally, I found that initial patch size had strong effects on the abundance several insect families that are ecologically relevant. Furthermore, the abundance of arthropod herbivores, parasitoids, and predators

were significantly influenced as well, with variable relationships to patch size. These findings help to further our understanding of how plant species spatial pattern affects biodiversity, ecosystem functioning, and community structure, as well as provide novel ideas for planting strategies in grassland restoration.

Acknowledgments

I will always have tremendous gratitude for my advisor, Jonathan Newman, for taking a chance on me, and inviting me into his lab to study biodiversity. His guidance, knowledge, and calm reassurance throughout the years were no doubt integral to my accomplishments.

Thank you to my committee member Kathryn Yurkonis, who always went above and beyond to help in all aspects of my research, despite being in another country. Her advice and encouragement, whether in person, online, or last minute, was a valuable part of my progress throughout the years.

Thanks to my parents who introduced me to science at a very young age (Billions and Billions!), and never stopped pushing me forward. Their immense support over the years, throughout all of my mistakes, will never be forgotten. Thanks to my sister, Nicole, for being a great friend, although to me you will always be little Coee.

To my partner, Rick Thompson, thank you for putting up with me, for always making dinner when I was too occupied, for all the morning coffees, and for making me a better person. Thank you as well for helping with biomass drying, and for learning to appreciate insects (and Collembola) the way I do. Thanks to our Lagomorpha babies/monsters for distracting me at the best of times, and the worst of times. My life would be incomplete without Jezebel, Rocco, and Garth.

Thank you to the Newman lab group, both past and present members, for their support and camaraderie. Dr. Kim Bolton was an esteemed supervisor both in the field and the lab, and hosted some of the best parties. Thanks to Aurora Patchett for being an important field member in nearly all projects, and for entertaining my insect interests, no matter how obsessive. Thanks to Dr. Heather Hager for sharing her expertise in multivariate stats. Thanks to Dr. Simone Harri, Dr. Emily Robinson, Dr. Gerry Ryan, and Kruti Shukla for welcoming me into the lab all those years ago.

Thank you to Neil Rooney, my co-advisor, for his outgoing help and support since my undergraduate years. Without him, I would have not applied to graduate studies at SES. Thank you to my committee member Alex Smith, for providing guidance in my arthropod research, and helpful feedback during the grueling writing stage.

Finally, thank you to my sweet Haze. I dedicate my thesis to him.

List of Tables

Table 2.1. Results from repeated measures ANOVA of seeded species patch size treatment effects on aboveground biomass (natural log transformed), proportion of invaders (arcsine square root transformed), overyielding ln(Di), biodiversity effects (selection and complementarity, square root transformed with sign preserved), and Simpson’s diversity. Linear contrasts between patch size treatments were based on the natural-log of the seeded patch edge to area ratio and excluded mixed seeding plots. Values are F-statistics and degrees of freedom, which were reduced for variables with two growing seasons of data.

Table 2.2. Results from repeated measures ANOVA of patch size treatment effects on interspecific association, the natural-log of total conspecific patches and the natural-log of total interspecific edges. Mixed plot treatment was not included in the analyses. Linear contrasts were based on the natural-log of the resident patch edge to area. Values are F-statistics and degrees of freedom.

Table 3.1. Species loadings from the RDA of relative spatial abundance of seeded species in relation to patch treatment and time for 2011-2012. Response variable (Resp.) coordinates correspond to axes in Fig 3.1 composition biplot.

Table 3.2. Species loadings from the RDA of relative spatial abundance of seeded species in relation to patch treatment for 2011. Response variable (Resp.) coordinates correspond to axes in Fig 3.2 composition biplot.

Table 3.3. Species loadings from the RDA of relative spatial abundance of seeded species in relation to patch treatment for 2012. Response variable (Resp.) coordinates correspond to axes in Fig 3.3 composition biplot.

Table 3.4. Standard least squares regression results from analysis of proportional change in number of conspecific clusters of seeded species and patch size, year, and patch × year.

Table 3.5. Effect tests from repeated measures ANOVA for relative spatial abundance of plant functional groups and patch size treatment (m) (patch trt) across 2011-2012.

Table 3.6. Standard least squares regression results from analysis of relative spatial abundance of functional groups and patch edge to area ratio (m/m2), year, and patch edge to area ratio × year.

Table 3.7. Standard least squares regression results from analysis of invader species relative spatial abundance and number of clusters using terms: patch edge to area ratio (m/m2), year, and patch edge to area ratio × year.

Table 4.1. Trophic group designations for arthropod taxon groups, based on published literature. Groupings were assigned based on the majority of described grassland species.

Table 4.2. Species loadings from the RDA of arthropod family (or taxon group) abundance in relation to patch treatment, block, and year. Response variable (Resp.) coordinates correspond to axes in Fig 4.1 composition biplot.

Table 4.3. Species loadings from the RDA of arthropod family (or taxon group) abundance in relation to patch treatment and block for year one (2011). Response variable (Resp.) coordinates correspond to axes in Fig 4.2 composition biplot.

Table 4.4. Species loadings from the RDA of arthropod family (or taxon group) abundance in relation to patch treatment, block for year two (2012). Response variable (Resp.) coordinates correspond to axes in Fig 4.3 composition biplot.

Appendix Table 1. Species list (n = 16) of grassland perennials selected for experiment, including authority, family, common name, seeds/g, plant functional and reproductive group, and distribution in Southern Ontario according to the Ontario Ministry of Natural Resources (Bradley, 2013). Seeds/g was calculated based on the mass of 100 seeds (n=10) for each species. All species are native to Southern Ontario except for S. arundinaceus.

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List of Figures

Figure 2.1. Effect of initial patch edge to area ratio (m/m2) of seeded species on a) aboveground biomass (g) (untransformed mean ± SE) over three growing seasons, and b) the selection effect (square root transformed mean with sign preserved ± SE), in the second (2011) and third (2012) growing seasons. Means with the same letter are not significantly different (LSD, P < 0.05). Plots were divided into patches 1, 0.5, 0.25 and 0.125 m on an edge, which corresponds to an edge to area ratio of 4, 8, 16 and 32 m/m2 for the seeded patches. Mixed seed plots (represented by the × symbol) developed an average patch edge to area ratio of 28.2 m/m2.

Figure 2.2. Effect of initial patch edge to area ratio (m/m2) of seeded species on a) Simpson’s diversity (mean ± SE) in the second (2011) and third (2012) growing seasons, and b) the proportion of non-seeded (invader) species (untransformed mean ± SE) over three growing seasons. Means with the same letter are not significantly different (LSD, P < 0.05). Plots were divided into patches 1, 0.5, 0.25 and 0.125 m on an edge, which corresponds to an edge to area ratio of 4, 8, 16 and 32 m/m2 for the seeded patches. Mixed seed plots (represented by the ×symbol) developed an average patch edge to area ratio of 28.2 m/m2.

Figure 2.3a) Patch treatment (m) and interspecific association (mean ± SE) in 2010- 2012, see Methods for measurement and calculation of intersection association. Means with the same letter are not significantly different (LSD, P<0.05). b) Patch treatment (m) and the natural logarithm of the total number of conspecific patches (mean ± SE) in 2010-2012. Means with the same letter are not significantly different (LSD, P<0.05). c) Patch treatment (m) and the natural logarithm of total interspecific edge (mean ± SE) in 2010-2012. Means with the same letter are not significantly different (LSD, P<0.05).

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Figure 3.1. Species composition biplot for relative spatial abundance of seeded species and patch treatment edge length (m) and time analysed with RDA, across 2011-2012. For species listing see Table 3.1.

Figure 3.2. Species composition biplot for relative spatial abundance of seeded species and patch treatment edge length (m) for 2011, analyzed with RDA. For species listing see Table 3.2.

Figure 3.3. Species composition biplot for relative spatial abundance of seeded species and patch treatment edge length (m) for 2012, analyzed with RDA. For species listing see Table 3.3.

Figure 3.4. Patch size (m) and proportional change in number of conspecific clusters of seeded species B. curtipendula, P. digitalis, and R. pinnata across years 2011-2012.

Figure 3.5. Patch size (m) and proportional change in number of conspecific clusters of seeded species (from left to right, down): A. gerardii, C. lanceolata, E. virginicus, H. helianthoides, M. fistulosa, O. rigidum, P. virgatum, and S. nutans separated by year.

Figure 3.6. Relative spatial abundance of C3 grass functional group (mean ± SE) and: patch treatment (m) (left), year (right). Means with the same letter and case are not significantly different (Tukey’s HSD test, P<0.05).

Figure 3.7. Relative spatial abundance of C4 grass functional group (mean ± SE) and patch treatment (m). Means with the same letter are not significantly different (Tukey’s HSD test, P<0.05).

Figure 3.8. Relative spatial abundance of C3 and C4 grass and patch edge to area ratio (m/m2) across years 2011-2012.

Figure 3.9. Mean relative spatial abundance of forbs functional group (mean ± SE) and patch treatment (m) × year. Means with the same letter are not significantly different (Tukey’s HSD test, P<0.05).

Figure 3.10. Relative spatial abundance of forbs and patch edge to area ratio (m/m2) separated by year 2011-2012.

Figure 3.11. Relative spatial abundance of legume functional group (mean ± SE) and patch treatment (m) × year. Means with the same letter are not significantly different (Tukey’s HSD test, P<0.05).

Figure 3.12. Relative spatial abundance of invader (non-seeded) species (mean ± SE) and patch treatment (m) × year. Means with the same letter are not significantly different (Tukey’s HSD test, P<0.05).

Figure 3.13. Relative spatial abundance of invader (non-seeded) species and patch edge to area ratio (m/m2) separated by year.

Figure 3.14. Total number of invader (non-seeded) species clusters and patch edge to area ratio (m/m2) separated by year.

Figure 4.1. Arthropod family relative abundance composition biplot in relation to patch size edge length (mix, 0.125, 0.25, 0.5, 1 m), and year analyzed with RDA. For species listing see Table 4.2.

Figure 4.2. Arthropod family relative abundance composition biplot in relation to patch size edge length (mix, 0.125, 0.25, 0.5, 1 m) for year one (2011), analyzed with RDA. For species listing see Table 4.3.

Figure 4.3. Arthropod family relative abundance composition biplot in relation to patch size edge length (mix, 0.125, 0.25, 0.5, 1 m) for year two (2012), analyzed with RDA. For species listing see Table 4.4.

Figure 4.4. Linear relationship between patch edge to area ratio (m/m2) and untransformed absolute abundance of a) Aphididae, b) Chalcidoidea, and c) Phoridae for 2011-2012. Plots were divided into patches 1, 0.5, 0.25 and 0.125 m on an edge, which corresponds to an edge to area ratio of 4, 8, 16 and 32 m/m2 for the seeded patches. Mixed seed plots developed an average patch edge to area ratio of 28.2 m/m2.

Figure 4.5. Linear relationship between patch edge to area ratio (m/m2) and untransformed absolute abundance of the a) herbivore and b) parasitoid trophic groups for 2012-2012. Plots were divided into patches 1, 0.5, 0.25 and 0.125 m on an edge, which corresponds to an edge to area ratio of 4, 8, 16 and 32 m/m2 for the seeded patches. Mixed seed plots developed an average patch edge to area ratio of 28.2 m/m2.

Figure 4.6. Relative abundance (mean ± SE) of predators and patch treatment (m) across both years of the study. Means with the same letter are not significantly different (Tukey’s HSD test, P<0.05).

Appendix Figure 1. Field and plot layout in randomized complete block design. Numbers on top indicate plot number, and codes on the bottom indicate treatment level of conspecific patch edge length (MC = monoculture plot). Plots are 4 m × 4 m in total area, separated by 2 m mowed aisles.

Appendix Figure 2. Seeded patch size treatments for experimental plots. The two- letter code for species represents first letters of the genus and specific epithet. All 16 species were planted in total area (1 m2) and density (1728 seeds/m2) for all treatment levels. Species were planted into a) one – 1 × 1 m patch, patch edge to

area ratio = 4 m/m2, b) four – 0.5 × 0.5 m patches, patch edge to area ratio = 8 m/m2, c) 16 – 0.25 × 0.25 m patches, patch edge to area ratio = 16 m/m2 or d) 64 – 0.125 × 0.125 m patches, patch edge to area ratio = 32 m/m2.

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TABLE OF CONTENTS

ABSTRACT ...... ii Acknowledgments ...... iv List of Tables ...... v List of Figures ...... viii Chapter 1: Introduction ...... 1 Dissertation Organization ...... 5 Chapter 2: Biodiversity and ecosystem function responses to plant spatial pattern in grasslands...... 7 Abstract ...... 7 2.1 Introduction ...... 8 2.2 Methods ...... 11 Data collection ...... 14 Data analysis ...... 15 2.3 Results ...... 15 2.4 Discussion ...... 17 Acknowledgements ...... 20 Figures ...... 21 Tables ...... 24 Chapter 3: Plant pattern affects spatial abundance of grassland species and functional groups ...... 26 Abstract ...... 26 3.1 Introduction ...... 27 3.2 Methods ...... 30 Data collection ...... 32 Data analyses ...... 33 3.3 Results ...... 34 3.4 Discussion ...... 36 Figures ...... 40 Tables ...... 50 Chapter 4: The effect of seeded patch size on arthropod abundance and community composition ...... 58 Abstract ...... 58 4.1 Introduction ...... 59 4.2 Methods ...... 61 Data collection ...... 63 Data analyses ...... 64 4.3 Results ...... 65 4.4 Discussion ...... 67

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Acknowledgements ...... 70 Figures ...... 72 Tables ...... 77 Chapter 5: General conclusions ...... 81 References ...... 86 Appendix ...... 95

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Chapter 1: Introduction

Grasslands are broad-ranging ecosystems that occur in zones with a distinct dry season (Axelrod, 1985). They typically occupy regions where the amount of precipitation and soil water is insufficient to support an arboreal canopy, but can support grass and forb dominated vegetation. The once extensive natural native grasslands throughout North America were fragmented and reduced to less than five percent of their original range (Knapp et al., 1999). Intensive agriculture and land-use during and post-European settlement drastically reduced native grassland soil biodiversity and food web diversity (Tsiafouli et al., 2015). In the eastern tallgrass prairie, grassland remnants are small patches throughout mostly agricultural land. Invasive species are now present throughout the ecosystem, which can aggressively displace native species.

Modern ecological research has shown that loss of biodiversity decreases the efficiency of ecosystem functions; evidence across various ecosystems, species, and trophic levels support this (Cardinale et al., 2012). There is also evidence that the stability of ecosystem functions is enhanced by greater biodiversity (Ives and Carpenter, 2007). More diverse communities tend to have greater productivity, as they have a greater probability of containing highly productive key species (the selection effect), and greater functional trait differences between species, leading to niche differentiation and functional facilitation (the complementarity effect). These two effects, collectively termed biodiversity effects (Loreau and Hector, 2001), are discussed in greater detail in Chapter 2. The loss of biodiversity affects primary productivity as much as global climate change drivers such as elevated CO2 and nutrient additions (Hooper et al., 2012; Tilman et al., 2012). Clearly, North American grasslands are in need of conservation and restoration due to its recent history of anthropogenic degradation.

Several grassland and prairie restoration projects are underway throughout North America in efforts to reverse biodiversity loss. Parks Canada restoration objectives are aimed at the ecological processes of grazing, fire, and plant succession with the ultimate goal of restoring and conserving grassland biodiversity to pre- settlement conditions (National Parks Directorate, 2011). In the USA, the large-scale Conservation Reserve Program (CRP) works to restore perennial land cover, reduce soil erosion, and preserve wildlife habitat (USDA and FSA, 2015). The ecological processes involved in these and other native restorations, such as coexistence and competition, are conceptually similar to those guiding community assembly and exotic invasions (Seabloom et al., 2003). Coexistence can occur among species when intraspecific competition is stronger than interspecific competition, or when interspecific competition occurs over shorter distances than intraspecific competition, known as heteromyopia (Murrell and Law, 2003; Murrell, 2010). Conversely, when two similar species competing for a limited resource cannot coexist and one species is eliminated, competitive exclusion is said to occur (Patten, 1961; Armstrong and McGehee, 1980; Cushing et al., 2004). Invading exotic species can dominate a native community when it is competitively superior; this may be caused by biological trait differences, or greater competitive ability due to the absence of natural enemies. Exotic invaders can also dominate native communities when anthropogenic disturbance creates environmental conditions that benefit the exotic, especially if competitively superior native species are rare, displaced, or recruitment-limited (Seabloom et al., 2003). The latter may occur when native ranges of perennial grassland species are significantly reduced, and seed production and establishment are relatively low. Moreover, if a competitive dominant species has a short dispersal range and low fecundity, an inferior competitive species with a longer dispersal range may invade (Bolker and Pacala, 1999).

Plants are sessile organisms, thus interspecific interactions are more likely to occur with neighbouring individuals than more distant ones (Tilman, 1994). The fine-scale patterning of plants may consequently have significant effects on the biodiversity and ecosystem functioning of a plant community. In 1947, Watt’s

publication on pattern and process influenced future research in plant communities as it initiated the concept of the plant community as a collective working mechanism, capable of self-maintenance and self-regeneration (Watt, 1947). Subsequently, spatial pattern in plant communities has been studied further, and pattern formation in grasslands has been attributed to local dispersal, species competition, and environmental heterogeneity (Seabloom et al., 2005). Exactly how plant spatial pattern affects community dynamics is not entirely clear, although studies have shown that planting arrangement can affect species coexistence, productivity, and invasion. Monzeglio and Stoll (2005) demonstrated that weaker competing species had greater productivity and fitness when intraspecific aggregation prevented interspecific interactions, but plant pattern effects on stronger competing species may be contingent on neighbouring species identity, traits, and non-spatial factors. Plant pattern can affect invasion resistance by limiting available low-richness sites for exotic species to invade and establish in a grassland community. Yurkonis et al. (2012) found that invasion by non-planted species was consistently higher in large species patch plots (or greater intraspecific aggregation) than plots planted with small species patches. Clearly, manipulations in fine-scale plant pattern have consequences on the interactions between establishing seedlings in communities, ultimately affecting larger plot-scale measures such as biodiversity and ecosystem functions.

The effects of fine-scale plant patterning on biodiversity and ecosystem functions (BEF) have valuable applications in grassland restoration and reconstruction. Planting strategies using large conspecific patches may initially limit or diminish competitive exclusion between weak competing species, however species more strongly affected by intraspecific competition may have decreased productivity under greater conspecific aggregation (Yurkonis et al., 2012). Furthermore, the long-term effects of species patterning on community dynamics and function are still unclear. Many BEF and restoration studies often employ the use of uniformly mixed species compositions (Fargione et al., 2007; Cadotte, 2013; Schittko et al., 2014), but this may not be the most effective method for maximizing

grassland biodiversity and restoration success, as I will discuss in my research. In the following chapters, I investigate the effects of fine-scale spatial patterning using varying seeded conspecific patch size treatments in experimentally reconstructed grassland plots.

Dissertation Organization

In Chapter 2 I explore how fine-scale species patterning at seeding can affect biodiversity and ecosystem functioning in a perennial grassland community. Species-rich communities typically have increased biodiversity effects, although the influence of species patterning on these effects is unclear. Plants interact over relatively short distances, thus the conspecific patch size at seeding likely affects the species interactions that drive biodiversity effects and ecosystem functions. Larger conspecific patch sizes at seeding may help to guard weaker competing species from stronger ones, delaying competitive exclusion. However, studies have shown that larger conspecific patches may allow for increased invasion to occur. To test the effect of patch size on biodiversity and ecosystem functioning, I used plots that were seeded with varying patch size treatments in a randomized complete block design. I found that smaller patch plots had initially increased productivity due to stronger selection effects from dominant species. Smaller patch plots also had greater resistance to invasion, but after three years the effects on productivity and invasion diminished. Compared to smaller patch and mixed plots, larger patch plots had greater seeded species diversity overtime, and thus may be more useful in restoration application of perennial grasslands.

In chapter 3 I examined how fine-scale spatial patterning affects plant species and functional group spatial abundance. Plant species and functional groups may differ in their inter- and intraspecific competitive abilities in both magnitude and distance. Spatial patterning at seeding should thus impact biotic assembly rules and community composition in establishing grasslands. Using landscape analyses over two years in my patch treatment plots, I found that forbs and C3 grasses were spatially more abundant in mixed and smaller patch plots, and may be stronger interspecific competitors, while C4 grasses were more abundant in larger patch plots, suggesting stronger intraspecific competitors. These results suggest that forbs dominate early establishing grassland communities, although this may change over time. Invading or non-seeded species were more abundant in larger patch plots,

with greater intraspecific aggregation. However, the number of clusters of invaders was greatest in smaller patch and mixed plots, indicating that subsequent establishment of invaders is hindered in plots with greater interspecific interspersion (or intermixing of species).

In native grasslands, the most diverse and abundant animals are insects, spiders, and mites, and they are known to vary spatially on very small spatial scales (Whiles and Charlton, 2006; Shorthouse and Larson, 2010). Therefore, in Chapter 4 I explore how plant patterning affects arthropod abundance and community composition. Larger or more concentrated resource patches may have greater herbivore abundance (resource concentration hypotheses), and more diverse plant communities may have greater predator and parasitoid abundance (enemies hypothesis), although these relationships do not always occur in experimental and observational research. Throughout my patch treatment plots, I sampled arthropods over two years and analysed family and trophic group abundance. Some economically- and ecologically-relevant families showed strong positive linear relationships between absolute abundance and patch size, as did the herbivore and parasitoid trophic groups. Furthermore, relative predator abundance was greatest in mixed plots. These results indicate that seeded patch size and thus plant spatial pattern has strong effect on not just plant, but arthropod community composition in establishing grasslands, and some patch sizes that increase the abundance of beneficial species may be valuable in restoration.

Chapter 2 has been modified from a manuscript (co-authored by K. A. Yurkonis and J. A. Newman) that has been tentatively accepted at the Journal of Ecology, pending minor revisions. Chapters 4 and 5 will be submitted to appropriate journals shortly. Chapter 5 discusses general conclusions and implications for future biodiversity-ecosystem function research and restoration efforts. Figures and tables are listed at the end of each respective chapter, except for the experimental design figures and table in the appendix, and all references are listed at the end.

Chapter 2: Biodiversity and ecosystem function responses to plant spatial pattern in grasslands

Abstract

Species interactions in diverse plant communities affect community-scale functions such as aboveground biomass production and invasion resistance. While the balance of inter- and intraspecific competition among the resident species can affect the strength of these interactions, it is unclear over what distances individuals typically interact in grasslands. Furthermore, it is unknown if species interactions at seeding can be effectively manipulated to improve these responses. In this study, I tested whether manipulating the size of seeded species patches affected aboveground biomass, diversity and invasion resistance in experimentally restored grassland plots (16 m2) using repeated measures ANOVA and linear contrasts. Plots were divided into patches that were 1, 0.5, 0.25 or 0.125 m on a side, and the equivalent of 1 m2 in each plot was seeded with one of 16 grassland species. A final treatment involved mixing and broadcasting all seeds into a plot to mimic typical restoration approaches. Using smaller conspecific seed patches resulted in plots that were less diverse and initially more productive than larger patch counterparts. Smaller patch plots also had a greater selection effect and experienced increases in resident species connectivity, suggesting that diversity declines and productivity gains resulted from the enhanced establishment and spread of more productive seeded species. Larger patch plots were initially more invaded than those seeded with smaller patches, but this effect diminished over time. This likely reflects the ability of the non-seeded species to more effectively colonize larger patches with poor seeded species establishment. Mixed seeding plots were most similar to the smallest patch plots in their metrics of resident species spatial pattern, diversity and invasion resistance. However, these plots were initially less productive and had a weaker selection effect. These results suggest that species interact over sub-meter scales in

establishing tallgrass prairie and, once established, their patterns are dynamic over time. Furthermore, structured seeding approaches should be considered over traditional mixed seeding approaches to control species dominance and preserve seeded species diversity within grassland systems.

2.1 Introduction

The field of biodiversity ecosystem functioning (BEF) research is founded on the hypothesis that ecosystem-level functions and services are determined, in part, by the abundance and richness of resident species. Biodiversity affects an ecosystem’s temporal stability of functions, as well as its ability to capture resources and produce biomass (Kennedy et al., 2002; Cardinale et al., 2012). A positive relationship typically arises between biodiversity and ecosystem functions as a result of species identities, life histories, and interactions (Cardinale et al., 2012; Hooper et al., 2012). However, we have much to understand about the scale of these interactions and whether they can be manipulated to affect desired ecosystem services and functions. This study investigates how the fine-scale spatial pattern among species can be manipulated to increase biomass production, plant diversity maintenance, and invasion resistance within a grassland restoration context.

Increasingly species-rich communities typically have greater biodiversity effects. The two pathways by which increasing biodiversity can benefit ecosystem functions have been generally labeled as the selection (or sampling) effect and the complementarity effect (Loreau and Hector, 2001). The selection effect, based on Price’s general theory of selection (Price, 1970; Price, 1972; Price, 1995), occurs when species with particular traits (yields) dominate a community (Loreau and Hector, 2001; Fox, 2005). Communities with positive selection effects are dominated by species that are also productive when grown alone in monoculture. Alternatively, a negative selection effect may occur when species that are less productive in monoculture dominate a mixture. Typically, positive selection effects explain the

positive diversity-productivity relationships that occur in the first few years of grassland experiments (Cardinale et al., 2007).

The complementarity effect involves resource partitioning and functional facilitation. It occurs when an interspecific variation in resource use leads to more complete resource use in mixtures than would occur in monoculture. This variation in resource use may be particularly beneficial when common resources are limited, reducing competition between species (Fargione et al., 2007). In plant communities, species resource use is varied spatially through different layers in the canopy aboveground or belowground due to root length (Wacker et al., 2009). A greater collective gain and use of resources can lead to higher productivity and ecosystem functions (Loreau, 2000). Conversely, negative complementarity effects indicate direct interference and competition among species (Wacker et al., 2008), which may result in species-level ‘underyielding’, or yields lower than expected based on monoculture yields. Complementarity typically increases over time in long-term biodiversity experiments (Fargione et al., 2007; Picasso et al., 2011). Collectively, positive selection and complementarity effects can explain increased productivity and reduced invasion in more species rich grasslands (Kennedy et al., 2002). Areas with high diversity typically have less successful establishment of novel or invading species due to fewer available resources for invaders to consume (Loreau and Hector, 2001; Kennedy et al., 2002).

While previous experiments have been able to examine responses on biodiversity and ecosystem functions, the influence of the fine-scale species patterning and the inter- and intraspecific interactions that fundamentally drive these responses cannot be evaluated with the traditionally used uniformly mixed planting method alone. Stronger interactions are more likely to occur among neighbouring individuals rather than among more distant ones (Tilman, 1994, Kennedy et al., 2002, Rayburn and Schupp, 2013), and there is increasing evidence that grassland plants interact over relatively short distances (Vogt et al., 2010). Since interactions may only occur at limited distances, the conspecific patch size, or

the intraspecific aggregation at seeding, likely affects the development of selection and complementarity effects that affect the emergent properties of diverse communities. Increased intraspecific aggregation may also reduce the frequency of inter- vs. intraspecific interactions, possibly delaying competitive exclusion and promoting species coexistence (Rayburn and Schupp, 2013, Stoll and Prati, 2001). Several studies have shown that inter- and intraspecific competition are affected by the fine-scale spatial pattern or arrangement of plant species in establishing grassland systems (Stoll and Prati, 2001, Yurkonis et al., 2012, Liao et al., 2014). For example, previous research by Yurkonis et al. (2012) showed that invasion by non-seeded species was greater in plots with larger conspecific patches of transplanted seedlings presumably due to decreased fine-scale heterogeneity. Other studies have manipulated density of seeded species in low-diversity grasslands, providing evidence that seed dispersal pattern can affect diversity and community structure (Houseman, 2014). This study is novel because it manipulated plant species pattern using conspecific patch sizes at seeding.

Manipulating species patch sizes at seeding can elucidate the scales over which individuals interact. Understanding how such manipulations affect community biomass production, diversity maintenance and invasion resistance would help to maximize these effects during grassland reconstruction efforts (Schittko et al., 2014, Cadotte, 2013, Fargione et al., 2007). In this study, I manipulated the balance between inter- and intraspecific interactions by experimentally constructing grassland communities with differing initial species patch sizes and subsequent edge (where interspecific > intraspecific interactions) to area (where interspecific < intraspecific interactions) ratios. I hypothesized that plots seeded with smaller conspecific patches would produce more aboveground biomass and be more resistant to invasion of non-seeded species than plots with larger patches due to greater biodiversity effects. I also hypothesized that plots with structured seeding would have increased aboveground biomass production and invasion resistance compared to plots seeded with a uniformly mixed seed mixture, as small seeded patches may help to reduce competitive exclusion during

establishment. My findings improve our understanding of the scale of plant species interactions that affect plant community biomass production and invasion resistance, and can be used to maximize these responses within grassland experimental and reconstruction efforts.

2.2 Methods

Study site

Experimental plots were planted in late May - early June 2010 at the University of Guelph Turfgrass Institute (GTI, Wellington County, Guelph, ON Canada; 43⁰ 32’ 56” N, 80⁰ 12’ 39” W). The GTI was established in 1987 on a site that had been variously disturbed for agricultural production. Soils at the study site consist of Guelph Sandy Loams (Brunisolic Grey-Brown Luvisol; Haplic Glossudalf) developed on loam till in the Guelph Drumlin Field. The location selected for the study was a low- maintenance lawn since the establishment of the research facility. Under such management, the site was mowed weekly during the growing season and occasionally sprayed for broadleaf weeds, though not in 2010. Vegetation prior to establishment of the study primarily consisted of Poa pratensis and Taraxacum officinale. In early May 2010, prior to seeding, the site was sprayed twice with Roundup WeatherMAX (glyphosate 540 g per L) at a rate of 13.4 mls per L (Roundup WeatherMAX, Monsanto, St. Louis, MO, USA). Debris was removed and the site was cultivated and leveled with a Blecavator (Specialized Turf Equipment Company, South Carolina, USA) to homogenize the vegetation and soil.

Experimental design

In May 2010, 28 plots (4 × 4 m with 2 m aisles) were seeded with 16 perennial grassland species in a randomized complete block design. Seeds of five C4 grasses [Andropogon gerardii Vitman, Bouteloua curtipendula (Michx.) Torr., Panicum virgatum L., Schizachyrium scoparium (Michx.) Nash and Sorghastrum nutans

(L.)(Nash)], three C3 grasses [Elymus canadensis L., Elymus virginicus L.

and Schedonorus arundinaceus (Schreb.) Dumort.], one legume [Desmodium canadense (L.) DC.] and seven forbs [Coreopsis lanceolata L., Heliopsis helianthoides (L.) Sweet, Monarda fistulosa L., Oligoneuron rigidum (L.) Small, Penstemon digitalis Nutt. ex Sims, Ratibida pinnata (Vent.) Barnh., and Symphyotrichum novae-angliae (L.) GL Nesom] were spread into a 1 m2 total area within each treatment plot at a rate of 1728 seeds/m2 based on the average weight of 100 seeds (n = 10) of each species (Appendix Table 1). This seeding density was based on regional pasture recommendations, and exceeds recommended tallgrass prairie seeding rates (Packard et al., 1997; Wilson, 2002). Many biodiversity field experiments plant species with either constant seed weight or constant seed number across treatments, although both seeding methods may introduce hidden treatments (Guo, 2011). Therefore, seed biomass differences across species (Appendix Table 1) were accounted for in determining relative expected yield (see below).

All selected species are native to Southern Ontario, except for S. arundinaceus, and co-occur throughout Ontario (USDA and NRCS, 2015). They were selected based on their availability and use in regional grassland restoration efforts. S. arundinaceus was included to test for effects of initial distribution on resistance to invasion of a species considered problematic in restored grasslands1 (Cully et al., 2003). This mixture of species represents typical functional diversity found in reconstructed tallgrass prairies (Stroud, 1941; Packard et al., 1997). S. arundinaceus (Cultivar: KY-31) seeds were obtained from the University of Kentucky and seeds of the remaining species were obtained from Wildflower Farm (Coldwater, Ontario, Canada). Seeds were dry-stored at 17°C for seven weeks prior to planting.

Four blocks of seven plots each (five treatment plots, one replicate of each seeded patch size, and two monoculture plots) were aligned in rows from East to West, and separated by 2 m mowed aisles (Appendix Fig 1). Experimental patch size

1 Future work with Dr. L. Cartwright and Dr. H. Hager will focus on invasion and dispersal of S. arundinaceus throughout this system

treatment plots (n = 20) were sown with increasingly larger conspecific patches in five different treatment levels. For each species, the 1 m2 total planted area was divided into the following treatments: one – 1 × 1 m patch, four – 0.5 × 0.5 m patches, 16 – 0.25 × 0.25 m patches or 64 – 0.125 × 0.125 m patches (Appendix Fig 2). Patch edge to area ratios for the above treatments were equivalent to 4, 8, 16 and 32 m/m2 respectively. Species were randomly assigned to patch locations within plots. All seeds for the fifth (mixed) treatment level were intermixed and hand broadcast. Broadcast seeding was chosen over drill seeding as broadcasting can result in greater seed-establishment, reduced new species establishment (Yurkonis et al., 2010), and a natural distribution of plants along existing soil heterogeneity patterns in ecological restoration (Wilson, 2002).

In addition to the treatment plots, each block included two 4 x 4 m monoculture arrays of the selected species at the study seeding rate (n = 4 per species except n = 3 for H. helianthoides which was reduced due to a seed shortage). Within these arrays, alternating 1 × 1 m areas (i.e., in a checkerboard design) were seeded with a single species so that species monocultures were separated by 1 x 1 m between edges. An alternating plot arrangement was used to minimize interspecific competition, and reduced monoculture plot sizes (1 x 1 m versus 16 x 16 m) were used to minimize the size of the experiment. Species within these plots experienced edge effects similar to those in the largest patch plots and their yields were expected to closely estimate species yields in the largest patch treatment plots.

Plots were watered lightly after seeding and straw soil erosion matting was placed over the seeded area to maintain soil moisture and minimize seed movement and granivory. Once seeds began to germinate in late June, the matting was carefully removed to avoid uprooting establishing seedlings. Non-seeded species were allowed to establish and persist within the treatment and monoculture plots.

Data collection

Productivity, invasion, and diversity

From September to November of 2010, 2011, and 2012, all aboveground biomass was clipped to 5 cm above the soil surface in treatment and monoculture plots, and sorted as seeded or non-seeded (invader). In 2011 and 2012, seeded biomass was sorted to species in treatment plots. Biomass was dried to constant mass at 60°C and weighed. Seeded species and invader relative yields were calculated as a proportion of the total yield for each plot in each year. The expected relative yield of species i in mixture or RYei, in treatment plots was calculated based upon the initial species proportion of the total seeded biomass. Simpson’s diversity index (1/D) (Wilsey et al., 2005) based on the seeded species relative yields was calculated for all treatment plots in 2011 and 2012.

Spatial analysis

In June 2011 and 2012 treatment plots were surveyed and mapped according to the dominant (≥ 50%) plant species within each cell of a continuous grid of 12.5 x 12.5 cm cells established over each plot (1024 cells/plot). Surveys were completed early in the growing season to minimize vegetation disturbance. Cells dominated by non- seeded species were classified as “other” and cells with > 50% bare ground were classified as “empty.” This effort generated a series of maps with 18 possible cover types (16 potential resident species, other, and empty) and the program QRULE (Gardner and Urban, 2007) was used to evaluate plot maps at seeding (mixed plots could not be evaluated at seeding) and over time. The QRULE association function was used to generate an association matrix for each map, which described the proportion of all possible neighbourships (based on a 4-neighbour rule of shared cell-cell boundaries) that occurred within (matrix diagonal) and among cover types. Pairwise cover type association values were summed to generate an interspecific association value for each map. I additionally determined the number of separate conspecific patches and the number of edges shared between cells of different cover types within each map from the analysis output. Finally, I used the QRULE patch

cumulative frequency distribution output, which includes the size and number of edges associated with each species patch, to calculate the average patch edge to area ratio (across all species) for each mixed seeding plot.

Data analysis

Aboveground biomass, biodiversity effects, Simpson’s diversity, the proportion of non-seeded invaders and species pattern response variables were analyzed with a repeated measures ANOVA (SAS ver. 9.3, SAS Institute, Cary, NC, USA) with block as a random factor and treatment, year and their interaction as model terms. Linear contrasts based on the natural logarithm of the seeded patch edge to area ratio (4, 8, 16 and 32 m/m2) for the treatment plots were used to test effects of decreasing initial patch size on the response variables. Least significant difference post-hoc tests were used to compare responses between mixed and structured seeding plots. Aboveground biomass, the number of patches and the number of edges were natural log transformed, biodiversity effects were square root transformed with sign preserved, and the proportion of invaders was arcsine square root transformed to meet normality assumptions.

2.3 Results

Species varied in their aboveground biomass production and establishment when seeded in monoculture. The most productive species in monoculture were H. helianthoides, E. virginicus, M. fistulosa, C. lanceolata and R. pinnata, which all yielded more than 600 g/m2 of aboveground biomass in monoculture averaged across all years. The least productive species in monoculture were B. curtipendula, S. nutans, S. scoparium and P. digitalis with less than 100 grams/m2 averaged across all years. Penstemon digitalis was not present in any replicate monocultures in 2010, and was present in only two of four plots in 2011. S. scoparium was present in two of four in 2010, one of four in 2011 and three of four replicate monocultures in 2012. Species establishment was also variable within the patch size treatment plots. Eleven of the sixteen species established consistently across all treatment plots (in

>95% of the twenty treatment plots) by 2012. The remaining five species were in less than 60% of the treatment plots by 2012 (A. novae-anglais, B.curtipendula, S. nutans, S. scoparium and P. digitalis). Schizachyrium scoparium established in two of twenty treatment plots.

Within the treatment plots, aboveground biomass production varied among years and was generally greater in plots with smaller patch size, or greater interspecific interspersion (intermixing of species patches), in the initial years (Table 2.1; Fig 2.1a). Mixed seed plots developed an average patch edge to area ratio of 28.2 m/m2 by 2011, closest to that of the smallest patch plots (32 m/m2). However, the mixed plots consistently produced less biomass than the smallest (0.25 m and 0.125 m patch edge plots with patch edge to area ratios of 16 and 32 m/m2 respectively; Fig 2.1a), but not the larger patch plots, although this effect diminished over time. Similar to productivity, overyielding increased with decreasing patch size (Table 2.1) and decreased from 2011-2012. This increase in productivity corresponded to an increase in the selection effect, but not complementarity, across years and across plots with greater interspecific interspersion (Fig 2.1b; Table 2.1). The selection effect in mixed seed plots was similar to that in the smaller (0.25 m and 0.125 m patch edge) patch (greater edge to area ratio) plots (Fig 2.1b). There was no evidence of complementarity in the second growing season (-5.18 ± 9.47), and in the third growing season complementarity was similarly negative across plot types (-31.74 ± 5.28; Table 1).

Diversity of the aboveground plant community increased over time and decreased across plots with greater interspecific interspersion (Table 2.1; Fig 2.2a). Mixed and the smallest patch plots were the least diverse by the end of the third growing season (Fig 2.2a). The proportion of invaders declined over time and plots with greater interspecific association were also more resistant to invasion (Table 2.1; Fig 2.2b). With exception of the largest patch plots (1 m on a side), mixed seed plots were just as effective as the patch size treatment plots in excluding, or resisting invasion from, non-seeded species (Fig 2.2b).

Species patterns within the seeded plots changed inconsistently across the patch size treatments over time. At seeding, as the seeded patch size decreased the proportion of interspecific associations decreased, the number of conspecific patches per plot increased and the total interspecific edges per plot increased (Table 2.2; Fig 2.3) across the patch size treatments. Over time, species patches coalesced in the small patch treatments and divided in the larger patch treatments (Fig 2.3).

2.4 Discussion

This plot-based field experiment tested the hypotheses that sub-meter scale species interactions affect aboveground biomass production, diversity maintenance and invasion resistance in establishing grasslands. Diversity within the first three growing seasons was most strongly affected by seeded patch size manipulation, indicating that fine-scale interactions likely lead to competitive exclusion during seedling establishment and that these can be minimized with structured seeding approaches. Biomass production initially increased across plots with successively smaller initial patches, a response indicative of the increased dominance and spread of productive species in smaller patch seedings. The largest patch plots had the highest and most consistent proportion of non-seeded species. This is consistent with Yurkonis et al. (2012) who also found that non-seeded species were more abundant in plots with larger patches of transplanted seedlings. Given these results, it appears that formative interactions occur over scales <0.5 m in developing grasslands and seeding species in patch sizes of 0.5 to 0.25 m on an edge could enhance restoration success over traditional mixed seeding approaches.

After three years, biomass production was structured by selection as opposed to complementarity effects in this experimental system. In diverse grasslands, complementarity effects typically increase over time as a result of increased facilitation among legume and non-legume species (Fargione et al. 2007). The one seeded legume in this system typically underyields in species mixtures and

does not appear to generate strong complementarity responses (K. A. Yurkonis, unpublished data), thus use of a different legume may have generated a different complementarity response. Given this outcome, it appears that complementarity effects may not be as pronounced within typical grassland restoration scenarios, and future efforts aimed at maximizing functional diversity and interspersion could help maximize these effects in restored grasslands.

Despite the lack of a strong biodiversity effect, productivity was initially affected the by spatial seeding structure. Increased productivity in smaller over larger patch plots likely reflects increased spread and competitive dominance by productive species, as these plots were less diverse had stronger positive selection effects. At least within this developing system, these results suggest that the outcome of localized species competitive interactions (manipulated by changing species patch sizes), will most strongly affect productivity responses. However, it is important to note that the present study was a legume deficient system and the productivity response may change under conditions with greater functional diversity and subsequent interspersion.

Spatially structured seeding notably affects development of species diversity within grassland systems. Houseman (2014) experimented with seed sowing method in an established grassland, and showed that when seed arrival in established grasslands was patchy instead of uniform, species diversity was increased. The heightened diversity was most likely due to decreased interspecific interactions between seedlings. Somewhat similarly, our results showed that species were more likely to be maintained at higher abundances in the system when seeded into initially larger patches or with greater aggregation. Furthermore, as also suggested by Houseman (2014), interspecific competition varies with the level of aggregation, and it appears that interspecific competition in the mixed and smaller patch plots may accelerate local competitive exclusion and increased dominance. Patches at least a half meter on a side appear to have enough interior spaces to free from neighbour competitive exclusion effects and maximize species establishment,

while those with 0.25 or 0.125 m on a side may be functionally all “edge”, and experience reduced success, during species establishment. Porensky et al. (2012) had similar results where intraspecific aggregation helped to maintain the diversity of an eight species plant community over three years. They suggested that “spatial priority” based on the initial intraspecific aggregation, or reduced interspecific interactions, helped to restrain aggressive or dominant species from outcompeting weaker ones. To our knowledge, this is the first study of its kind to use various levels of species aggregation to demonstrate the scales of formative interactions in a diverse perennial grassland setting. While these scales may vary across grassland systems, it is likely to be relatively consistent at least in perennial systems.

The traditional mixed species seeding methods used in grassland restoration may not fully maximize desired restoration outcomes. In this study, the mixed seeded plots most closely resembled the smallest patch plots in their resulting structure and function. Given the sometimes poor establishment from traditional mixed seeding approaches (Stevenson et al., 1995; Lawson et al., 2004; Houseman and Gross, 2006; Larson et al., 2011) where interspecific competition can be high during germination, it is not surprising that mixed seed and smallest patch seeding efforts resulted in similarly low diversity communities in my study. Moreover, a greenhouse experiment by Yurkonis and McKenna (2014) demonstrated that conspecific aggregation at seeding may help to separate slower establishing species from those that consume local resources faster. This suggests that the failure of mixed seeding approaches to produce desired diversity outcomes may be most attributable to species interactions during community establishment versus a general failure to provide appropriate germination and establishment conditions.

Based on these results, grassland reconstructions may benefit most from seeding with larger patches. Plots seeded with 0.5 m patches were more diverse, and just as productive and resistant as mixture plots. From a reconstruction perspective, larger patch seeding clearly improves diversity maintenance and representation among the seed species. Strip-seeding has been advocated as an

effective method for restoration in spatial pattern studies (Rayburn and Laca, 2013). However, it may require more time than patch-seeding to thoroughly colonize a reconstructed area since species dispersal rates from strips into the surrounding unseeded zones is still unclear, although often observed to be slow and requiring several seasons (Hedberg and Kotowski, 2010).

While a few previous studies have used transplants to assess pattern effects in grasslands (Wilson and Tilman, 1991; Ewing, 2002; Yurkonis et al., 2012; Rayburn and Schupp, 2013; Liao et al., 2014), this is one of the first to use patterned seeding at relatively large scales in a restoration context. It is clear that spatially structured seeding efforts most notably affect diversity maintenance in grassland systems and, at least in the initial stages of a typical reconstruction effort, species structure affects plant productivity and invasion resistance. Although seeding small (sub-meter) conspecific patches may seem challenging, especially at larger scales, the benefits may warrant the additional effort since costs associated with weed management and subsequent replanting efforts to improve diversity may be reduced.

Acknowledgements

P. Purvis, K. Carey and the staff at the Guelph Turfgrass Institute were immensely helpful in preparing and managing the site. K. Bolton, E. Drystek, M. Gillespie, J. Holdenried, L. Jasiuk, A. Patchett, J. Reynolds and K. Shukla helped with planting and sampling the plots. K. Cook, E. Palmer, P. Smith helped with surveying and biomass harvest. K. Yurkonis and J. Newman were invaluable helpful throughout the writing process. The project was funded through grants from the Ontario Ministry of Agriculture, Food and Rural Affairs (OMAFRA) and the Canadian Natural Sciences and Engineering Research Council (NSERC) to J. Newman.

Figures

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Tables

Table 2.1. Results from repeated measures ANOVA of seeded species patch size treatment effects on aboveground biomass

(natural log transformed), proportion of invaders (arcsine square root transformed), overyielding ln(Di), biodiversity effects (selection and complementarity, square root transformed with sign preserved), and Simpson’s diversity. Linear contrasts between patch size treatments were based on the natural-log of the seeded patch edge to area ratio and excluded mixed seeding plots. Values are F-statistics and degrees of freedom, which were reduced for variables with two growing seasons of data.

Source df Biomass1 Invasion1 df Overyielding2 Selection2 Complementarity2 Diversity2 Patch trt (P) 4, 42 4.64** 3.76* 4, 27 8.30** 11.25*** 1.88 27.91*** Linear contrast 1, 42 9.79** 12.00** 1, 27 30.34*** 29.10*** 0.27 47.30*** Year (Y) 2, 42 113.29*** 361.22*** 1, 27 14.43** 15.84** 7.92** 12.90** Y x P 8, 42 1.38 1.04 4, 27 0.63 1.49 0.19 0.38

1 = 3 years of data; 2 = 2 years of data; * P < 0.05; ** P < 0.01, *** P < 0.001

Source df Interspecific association ln(total patches) ln(total edges) Patch trt (P) 4, 39 65.56*** 612.54*** 266.62*** Linear contrast 1, 39 244.81*** 2284.47*** 940.05*** Year (Y) 2, 39 3.44* 34.94*** 161.36*** Y x P 7, 39 5.18*** 57.11*** 5.89***

* P < 0.05; ** P < 0.01, *** P < 0.001

Chapter 3: Plant pattern affects spatial abundance of grassland species and functional groups

Abstract

The small-scale inter- and intraspecific interactions that plants experience have larger consequences on the biotic assembly rules in a community. Plant species and functional groups will differ in their inter- and intraspecific competitive strengths, and these strengths may differ in magnitude and distance. Previous work has shown that when species differ in their relative inter- and intraspecific competitive abilities, manipulating fine-scale species patterning may affect diversity and ecosystem functioning. However, this has yet to be tested in diverse perennial plant communities, and with varying levels of aggregation. To address this, twenty 4 × 4 m plots were seeded with sixteen grassland species in a reconstructed grassland setting. Species were seeded in either conspecific patch sizes (0.125, 0.25, 0.50, or 1 m on edge) or uniformly mixed and broadcast. I used partial redundancy analyses to test the effect of patch size on the relative spatial abundance of seeded species over two years. Using repeated measures ANOVAs and regressions, I tested the relative spatial abundance (number of cells occupied over total plot) and proportional change in number of clusters of seeded species, functional groups, and invader (non- seeded) species in relation to patch size over two years. Elymus virginicus, Heliopsis helianthoides, and Monarda fistulosa had greatest relative spatial abundance in mixed plots and thus were relatively stronger interspecific competitors. A. gerardii, P. virgatum, and S. novae-angliae had increased relative spatial abundance in larger patch plots (0.5 m), suggesting that they were relatively weaker interspecific competitors (or stronger intraspecific competitors). Forbs and C3 grasses both had increased relative spatial abundance in mixed and small patch plots, and thus were stronger interspecific competitors; while C4 grasses had increased relative spatial abundance in larger patch plots (0.5 m) suggesting they were relatively stronger

intraspecific competitors. The one legume, D. canadense, had increased relative spatial abundance in 0.25 m plots. The relative spatial abundance of non-seeded species was higher in larger patch plots, although the number of clusters of invading species was highest in smallest patch and mixed plots, suggesting that subsequent establishment of invaders may be hindered by interspecific competition from dominant resident species (forbs). The results provide evidence that initial seeded patch size in grassland communities generates species-specific and functional group-specific responses. These findings are useful for restoration in nutrient-poor grasslands to favour increased forb or legume abundance. However, the relative competitive strengths and abundances observed in this study may only be applied to early successional or reconstructed grassland communities, since dominant species and functional groups may change over time.

3.1 Introduction

The spatial patterning of species in any community, regardless of overall diversity, can affect the inter- and intraspecific interactions, which may in turn influence larger-scale responses such as ecosystem function. As discussed in Chapter 2, grassland plants species most likely interact over relatively short distances (Vogt et al., 2010), and fine-scale patterning greatly influences inter- and intraspecific competition (Stoll and Prati, 2001; Yurkonis et al., 2012; Liao et al., 2014). Plant species may differ in their inter- and intraspecific competitive abilities, and the magnitude and distance over which these interactions occur may differ as well (Bolker and Pacala, 1999; Murrell and Law, 2003), even at the establishing seedling phase. Thus, these small-scale inter- and intra specific interactions impact biotic assembly rules (Götzenberger et al., 2012). Increasing the frequency at which individuals experience intraspecific interactions could have plot-scale effects on the overall performance (via productivity, stability, resistance) of a focal species (Murrell and Law, 2003) and its encompassing community. For species more strongly affected by interspecific competition, or weak competing species, larger aggregation or conspecific patches may initially help to prevent or delay competitive

exclusion (Stoll and Prati, 2001; Hart and Marshall, 2009). Yet, the initial interactions among seedlings could affect how plants disperse locally and subsequently establish through succession, how they interact at varying distances within a community, and the influence of environmental conditions on ecosystem functioning (Bolker et al., 2003; Murrell, 2010; Yurkonis et al., 2012).

Previous experiments conducted over short time periods have provided evidence for the hypothesis that when species differ in their relative inter- and intraspecific competitive abilities, manipulating fine-scale species patterning at equal coarse-scale species richness and evenness may influence diversity (Yurkonis and McKenna, 2014), as well as ecosystem functioning long-term. However, this hypothesis remains to be fully tested in diverse perennial plant communities (Bolker and Pacala, 1999; Bolker et al., 2003; Yurkonis et al., 2012). Porensky et al. (2012) showed that larger conspecific aggregations promoted species coexistence over three years, but the selected eight prairie species were planted in only two arrangements: interspersed via broadcast seeding, and aggregated in intraspecific sectors (one per species).

It remains to be seen how different levels of aggregation or patch size at seeding would affect community composition. Such responses may not only be species-specific, but could also impact functional groups in plant communities. Earlier ecological work has summarized that if competition is greatest among similar-trait and resource-exploiting species, functional divergence occurs in competition-related traits (Schleicher et al., 2011). Conversely, functional convergence occurs when highly competitive species competitively exclude weaker competitors (Tilman, 1990; Grime, 2006). Despite varying growth strategies, plants of different functional groups may compete with one another above or belowground, restricting available resources and impacting successional change (Kimball et al., 2014). Recent work by Kimball et al (2014) investigated seeding methods and success for native grassland functional groups, but in this study I further examine which patch size at seeding is most effective for each functional group. These results

should help to elucidate if there is an effect of spatial-pattern on functional competitive strategies.

In Chapter 2, I showed that plots with smaller patch size at seeding had lower diversity, yet these plots also had greater productivity and initial resistance to invasion. There was a greater selection effect in the smaller patch plots, which was driven by the dominant or higher yielding species. Conversely, previous grassland experiments conducted on similar scales have shown that diverse communities tend to have greater resistance to invasion (Kennedy et al., 2002; Fargione and Tilman, 2005; Frankow-Lindberg, 2012). I measured invasion using overall aboveground biomass of non-seeded species, but this response variable does not take into account relative spatial abundance. Invading species may have greater biomass yield in larger patch plots, but this does not explain spatial pattern or subsequent spread and dispersal, which may be limited by competition with seeded species. It is unclear how species interactions in an establishing perennial grassland community influence this effect. Consequently, patch size of species at seeding may not only affect overall plot scale biomass of invading species, it may also influence the magnitude and frequency of invasion spatially at the fine scale. Such information would be of use to small-scale restoration efforts, as well as biodiversity ecosystem function experiments that weed experimental plots (Roscher et al., 2005; Cadotte, 2013). Currently, it is unclear how initial conditions at seeding influence inter- and intraspecific interactions, which would be crucial to management of invasion resistance and restoration in invaded communities (Sheley and James, 2014).

Using varying sizes of seeded species patches in plots with equal initial diversity, I address the following questions: (1) Do native grassland species respond differently to patch size at seeding? (2) Which patch size results in the greatest establishment of each functional group (C3 grasses, C4 grasses, forbs, legumes)? and (3) Do smaller patch plots have less invasion spatially at the fine-scale? Species and functional groups that increase in spatial abundance in smaller patch plots may be stronger interspecific competitors, and those that increase in larger patch plots may

be stronger intraspecific competitors in early grassland communities. These results will help to clarify the relative strength of inter- and intraspecific competition between native grassland species and functional groups in early succession. As well, these results will help to elucidate how patch size at seeding can hinder (or facilitate) invasion spatially, potentially helping in the creation of optimal strategies for grassland restoration.

3.2 Methods

Study site

Experimental design

virgatum L., Schizachyrium scoparium (Michx.) Nash and Sorghastrum nutans

(L.)(Nash)], three C3 grasses [Elymus canadensis L., Elymus virginicus L. and Schedonorus arundinaceus (Schreb.) Dumort.], one legume [Desmodium canadense (L.) DC.] and seven forbs [Coreopsis lanceolata L., Heliopsis helianthoides (L.) Sweet, Monarda fistulosa L., Oligoneuron rigidum (L.) Small, Penstemon digitalis Nutt. ex Sims, Ratibida pinnata (Vent.) Barnh., and Symphyotrichum novae-angliae (L.) GL Nesom] were spread into a 1 m2 total area within each treatment plot at a rate of 1728 seeds/m2 based on the average weight of 100 seeds (n = 10) of each species (Appendix Table 1). This seeding density was based on regional pasture recommendations, and exceeds recommended tallgrass prairie seeding rates (Packard et al., 1997; Wilson, 2002). Many biodiversity field experiments plant species with either constant seed weight or constant seed number across treatments, although both seeding methods may introduce hidden treatments (Guo, 2011). Therefore, seed biomass differences across species (Appendix Table 1) were accounted for in determining relative expected yield (see below).

All selected species are native to Southern Ontario, except for S. arundinaceus, and co-occur throughout Ontario (USDA and NRCS, 2015). They were selected based on their availability and use in regional grassland restoration efforts. S. arundinaceus was included to test for effects of initial distribution on resistance to invasion of a species considered problematic in restored grasslands (Cully et al., 2003). This mixture of species represents typical functional diversity found in reconstructed tallgrass prairies (Stroud, 1941; Packard et al., 1997). S. arundinaceus (Cultivar: KY-31) seeds were obtained from the University of Kentucky and seeds of the remaining species were obtained from Wildflower Farm (Coldwater, Ontario, Canada). Seeds were dry-stored at 17°C for seven weeks prior to planting.

aisles (Appendix Fig 1). Experimental patch size treatment plots (n = 20) were sown with increasingly larger conspecific patches in five different treatment levels. For each species, the 1 m2 total planted area was divided into the following treatments: one – 1 × 1 m patch, four – 0.5 × 0.5 m patches, 16 – 0.25 × 0.25 m patches or 64 – 0.125 × 0.125 m patches (Appendix Fig 2). Patch edge to area ratios for the above treatments were equivalent to 4, 8, 16 and 32 m/m2 respectively. Species were randomly assigned to patch locations within plots. All seeds for the fifth (mixed) treatment level were intermixed and hand broadcast. Broadcast seeding was chosen over drill seeding as broadcasting can result in greater seed-establishment, reduced new species establishment (Yurkonis et al., 2010), and a natural distribution of plants along existing soil heterogeneity patterns in ecological restoration (Wilson, 2002).

Data collection

Spatial analysis

number of separate conspecific clusters and the number of edges shared between cells of different cover types (seeded and non-seeded species) within each map from the analysis output.

Data analyses

Species response to patch size

To evaluate if the sixteen selected species respond differently to initial patch size, the relative spatial abundance of each species was determined by finding the proportion of 12.5 cm2 cells (of 1024 total cells) occupied by seeded species or non- seeded species per plot using data collected in the vegetation survey. Using the software Canoco 5 (Microcomputer Power; Ithaca, NY, USA), a partial redundancy analysis (RDA) was used to test the effect of patch treatment on the relative spatial abundance of all plant species, employing hierarchical permutations for the split- plot repeated-measures design (499 unrestricted permutations). Proportional change in species clusters was calculated as the proportion of surveyed / number at seeding; thus, when change in conspecific clusters is greater than 1, the focal species increased in cluster number from seeding. The proportional change in number of clusters was chosen to theoretically represent points of establishment and subsequent dispersal of focal species. JMP (JMP ver. 11, Cary, NC: SAS Institute Inc.) was used to perform standard least squares regressions using the effects: patch size, year, and patch size × year to test for a linear response in proportional change in clusters of the seeded species across patch sizes.

Functional group response to patch size

To find which patch size treatment resulted in the greatest establishment of each functional group, the relative spatial abundances of each functional group were analysed in JMP using repeated measures ANOVAs in JMP with the model terms: block(random), patch size, block(random) × patch size, year, and patch size × year. Relative spatial abundances were calculated as above, and Box-Cox transformed prior to statistical analyses to meet the assumptions necessary for repeated

measures ANOVA. I also used standard least squares regressions with the effects: patch edge to area ratio, year, and patch size × year, to test for a linear response in the untransformed relative spatial abundances of functional groups across patch size.

Invasion and patch size

To assess if smaller patch plots have less invasion spatially, the relative spatial abundance of invader (non-seeded) species was analysed using repeated measures ANOVAs in JMP with the same model terms as above. Relative spatial abundances were calculated as above and Box-Cox transformed prior to statistical analyses to meet the assumptions necessary for ANOVA. I also used standard least squares regressions in JMP with the effects: patch edge to area ratio, year, and patch edge to area ratio × year, to test for a linear response in untransformed relative spatial abundance and number of clusters of non-seeded (invader) species across patch size treatments.

3.3 Results

Species response to patch size

In the RDA evaluating relative spatial abundance of plant species across both years, patch size and year explained 45.5% of the total variation (pseudo-F=3.2, P<0.005; axis 1,2 eigenvalues = 0.2810, 00667) (Fig 3.1, Table 3.1). While year did separate along axis-2, patch treatment did not cleanly separate along either axis. The relative spatial abundance of A. gerardii, P. virgatum, and S. novae-angliae was negatively correlated with axis-1, increasing in 0.5 m patch plots, while the relative spatial abundance of E. virginicus, H. helianthoides was negatively correlated with axis-1 and increased in the mixed plots. Time had a strong effect on relative spatial abundance of component species, thus 2011 and 2012 were analyzed separately to interpret the effect of patch treatment more clearly. In 2011, patch size accounted for 73.3% of the total variation (pseudo-F=4.7, P<0.005; axis 1,2 eigenvalues =

0.6573, 0.0355) (Fig 3.2, Table 3.2). Relative spatial abundances of E. virginicus and H. helianthoides were positively correlated with the axis-1, H. helianthoides again increasing in the mixed plots, and relative spatial abundance of P. virgatum was negatively correlated. Relative spatial abundance of S. scoparium was positively correlated with the axis-2. In 2012 patch size accounted for 88.8% of the total variation (pseudo-F=13.6, P<0.005; axis 1,2 eigenvalues = 0.7937, 0.0620) (Fig 3.3, Table 3.3). Relative spatial abundances of A. gerardii, B. curtipendula, D. canadense, E. canadensis, O. rigidum, and P. virgatum were negatively correlated with the axis-1, while relative spatial abundances of E. virginicus, H. helianthoides, and M. fistulosa were positively correlated, increasing in the mixed plots.

Based on a patch and edge analysis, the mixed seeded plots developed an average patch edge to area ratio of 28.2 m/m2 by 2011 (see Chapter 2), which was used in the regression analyses. In the regression analysis testing patch size, year, and proportional change in number of conspecific clusters, B. curtipendula, P.digitalis, and R.pinnata had a positive relationship with patch size (Fig 3.4, Table 3.4) across both years. A. gerardii, C. lanceolata, H. helianthoides, M. fistulosa, and P. virgatum had a positive relationship with patch size, although slopes varied between years, and D. canadense, E. canadensis, E. virginicus, O. rigidum, and S. nutans had a positive relationship with patch size in 2011, and negative in 2012 (Fig 3.5, Table 3.4).

Functional group response to patch size

Among the C3 grass functional group, mean relative spatial abundance generally increased in smaller patch plots, was greatest in the mixed plots, and there was an overall increase across all patch treatments in 2012 (Fig 3.6, Table 3.5). Mean relative spatial abundance of the C4 grasses was greatest in 0.5 m patch plots, and decreased in smaller patch plots (Fig 3.7, Table 3.5). Regressions showed that relative spatial abundance of C3 grasses had a positive relationship with patch edge to area ratio (or increased with decreasing patch size), and relative spatial

abundance of C4 grasses had a negative relationship with patch edge to area ratio across years (Fig 3.8, Table 3.6).

The forb functional group had greater mean relative spatial abundance in the smaller patch plots, and there was an interaction with patch treatment size and year (Fig 3.9, Table 3.5). While forb mean relative spatial abundance generally increased in 2012, most results had varying relationships with patch treatment size and year. Forb relative spatial abundance had a positive linear relationship with patch edge to area ratio (or increased with decreasing patch size) in both years (Fig 3.10, Table 3.6). The legume, D. canadense, generally had the greatest relative spatial abundance in 0.25 m patch plots (Fig 3.11, Table 3.5), but the results were variable as there was an interaction between patch treatment size and year, and relative spatial abundance of D. canadense overall increased markedly in the 0.25, 0.5, and 1 m patch plots in 2012. There was no linear relationship between patch size and legume abundance (Table 3.6).

Invasion and patch size

The mean relative spatial abundance of invader (non-seeded) species was affected by an interaction between patch treatment size and year (F4,15=14.7 P<0.0001) (Fig 3.12). The relationship increased with larger patch sizes, although to a lesser extent in 2011. Regression analysis confirmed the linear relationships in both years (Fig 3.13, Table 3.7). The number of clusters of invader (non-seeded) species increased with increasing patch edge to area ratio (or increased with decreasing patch size) in both years (Fig 3.14, Table 3.7).

3.4 Discussion

Species response to patch size

The results show that some of the selected native grassland species did have species-specific responses to patch size at seeding. E. virginicus, H. helianthoides, and

M. fistulosa in 2012, relatively dominant species throughout the study, all had greater relative spatial abundance mixed plots, suggesting that these species are strong interspecific competitors in establishing grassland communities. These dominant species also generally had greater conspecific clusters in smaller patch plots. H. helianthoides, C. lanceolata, and M. fistulosa had the greatest proportional change in number of conspecific clusters in larger patch plots, despite greater relative spatial abundance in smaller patch plots. While these dominant species may be able to disperse well in 1 m patch plots, its subsequent establishment and growth in abundance may be limited by intraspecific competition. A. gerardii, P. virgatum, and S. novae-angliae increased in relative spatial abundance in the 0.5 m patch plots, the second largest treatment size, indicating these species may be weaker interspecific competitors, and stronger intraspecific competitors in establishing grassland communities. If in the mixed plots E. virginicus, H. helianthoides, and M. fistulosa are stronger interspecific competitors, then these species may competitively exclude weaker interspecific competitors such as A. gerardii, P. virgatum, and S. novae-angliae. However, it is likely that competitive exclusion would be halted by disturbance, density-dependent factors, or niche differences of the dominant species (Connell, 1983; Adler et al., 2007). Furthermore, established eastern tallgrass prairies are typically dominated by C4 grasses (Wilson, 2002), such as A. gerardii and S. nutans (Silletti et al., 2004), and relative competitive strengths could shift between species throughout succession. This is substantiated by previous ecological research which shows that early dominating species in establishing grasslands (forbs in this case) decrease in proportional biomass production as communities undergo succession, and later successional species increase in biomass (Mangan et al., 2011).

Functional group response to patch size

All functional groups had a significant relationship between relative spatial abundance and patch size at seeding. C3 grasses had greatest spatial abundance in smaller patch and mixed plots where interspecific association was greater,

suggesting that this functional group may be stronger interspecific competitors.

Conversely, C4 grasses were highest in spatial abundance in 0.5 m patch plots, indicating stronger intraspecific competitors. Forbs were higher in relative spatial abundance in smaller patch plots, dominating these communities, suggesting that these native grassland forbs are most likely stronger interspecific competitors compared to other native functional groups at this stage. The one legume, D. canadense, had greater relative spatial abundance in 0.25 m patch plots. In nutrient- poor environments, restoration efforts may benefit from heightened legume abundance due to their ability to fix atmospheric nitrogen through rhizobia symbiosis. Furthermore, legumes may not have had sufficient time to establish after soil disturbance (Ritchie and Tilman, 1995), and so a planting strategy focused on increased legume abundance would be useful. As this study used only one legume, further examination may be required with a larger set of legumes, although grasslands generally have very low legume abundance and diversity (Ritchie and Tilman, 1995). As previously discussed, although forbs may be strong competitors at this stage of establishment, relative competitive strengths may shift over time, and dominant species and functional groups may likely change.

Invasion and patch size

The results from the analysis of invasion (or establishment) of non-seeded species generally provide evidence that the relative spatial abundance of invading non- seeded species in restored grassland plots is lower in smaller patch plots. Similarly, recall that results from Chapter 2 showed that the smallest patch plots (0.125 and 0.25 m) had the lowest invader biomass from non-seeded species in the first year of the study. However, the number of individual patches of non-seeded species was greatest in smaller patch plots. When considered with the relative spatial abundance and biomass results, this suggests that while fine-scale invasion may be more frequent, the subsequent establishment of invading species may be diminished in smaller patch plots. Invading species may have difficulty establishing and dispersing in smaller patch plots due to increasing interspecific competition from the dominant

species (forbs). Previous research in disturbed sites found that three daisy species of the genus Senecio achieved greater spread in uniformly distributed compared to patchy distributed sites (Bergelson et al., 1993; Hastings et al., 2005). The greater number of invading species clusters observed in the smaller patch and mixed plots of my study may thus indicate increased spread, or early dispersal. Identifying the species composition of the non-seeded group in future surveys would help to clarify individual competitive strengths of these species in early grassland establishment.

This experiment has provided evidence supporting the idea that native grassland species, and the functional groups they comprise, have different responses to initial seeded patch size due to varying inter- and intraspecific competition strengths. Results also indicate that the initial seeding pattern influences relative spatial abundance and establishment of invading or non-seeded species in a restored grassland ecosystem. These findings are of use to restoration strategies, as smaller patch size plots may result in decreased invasion and preferentially increased forb or legume abundance.

Figures

Figure 3.2. Species composition biplot for relative spatial abundance of seeded species and patch treatment edge length (m) for 2011, analyzed with RDA. For species listing see Table 3.2.

Figure 3.3. Species composition biplot for relative spatial abundance of seeded species and patch treatment edge length (m) for 2012, analyzed with RDA. For species listing see Table 3.3.

s 6

u l B. curtipendula

i P. digitalis

c R. pinnata

p 4

e 2

o 0 r 0.125 0.250 0.500 1.000

P Patch size (m)

Figure 3.4. Patch size (m) and proportional change in number of conspecific clusters of seeded species B. curtipendula, P. digitalis, and R. pinnata across years 2011-2012.

2011 2.0 2011 5 2012 2012

i 4 a

t 1.5

e e 3

n . 1.0

2 .

C 1 0.5

0 0.125 0.250 0.500 1.000 0.0 0.125 0.250 0.500 1.000 5 4 2011 2011 2012 2012 4

e 3

e 3 e

n 2

s 2

c 1 1

e 0 0 0.125 0.250 0.500 1.000 0.125 0.250 0.500 1.000

o 2011 6 2011

c 2 2012 2012

c 4

i i 3

v 1

n 1

t r 0 0 o 0.125 0.250 0.500 1.000 0.125 0.250 0.500 1.000

o r 5 P 2.0 2011 4 2011

a 1.5 2012 2012

d 3

u i

s i

1.0

O 2

0.5 1

0.0 0 0.125 0.250 0.500 1.000 0.125 0.250 0.500 1.000 2011 Patch size (m) 5 2012 2.0 2011 4 2012

m s 1.5

a a 3 t

v . 1.0

P 2 0.5 1

0 0.0 0.125 0.250 0.500 1.000 0.125 0.250 0.500 1.000 Patch size (m) Figure 3.5. Patch size (m) and proportional change in number of conspecific clusters of seeded species (from left to right, down): A. gerardii, C. lanceolata, E. virginicus, H. helianthoides, M. fistulosa, O. rigidum, P. virgatum, and S. nutans separated by year.

r 1.25

g a

f ab

o A

c ab n 1.20

b ab

a 1.15 b

R mix 0.125 0.250 0.500 1.000 2011 2012 Patch treatment (m) Year

a a

f ab

e 1.05

n ab

b b

t 1.00

e R 0.95 mix 0.125 0.250 0.500 1.000 Patch treatment (m)

Figure 3.7. Relative spatial abundance of C4 grass functional group (mean ± SE) and patch treatment (m). Means with the same letter are not significantly different (Tukey’s HSD test, P<0.05).

c C grass n 3

d 0.3 C4 grass

a 0.2

v 0.1

e R 0.0 10 20 30 Patch edge:area (m/m2)

Figure 3.8. Relative spatial abundance of C3 and C4 grass and patch edge to area ratio (m/m2) across years 2011-2012.

f 2011

o a

e a 2012

n 1.6

d b

b b

l bc a 1.4

i t cd

p cde de

s e de

i t 1.2

R mix 0.125 0.250 0.500 1.000 Patch treatment (m)

a 2011

n 2012

b 0.6

s 0.4

r 0.2

F 10 20 30

2 Patch edge:area (m/m )

Figure 3.10. Relative spatial abundance of forbs and patch edge to area ratio (m/m2) separated by year 2011-2012.

u 2011

f 2012

o a

c 1.05 ab

n ab

a ab

d ab

n ab ab u ab b

e 1.00

l e mix 0.125 0.250 0.500 1.000

R Patch treatment (m)

b 2011

a a

l 0.6 2012 a a ab i ab

p ab

i 0.4

t cd

c 0.2 e

e e

a mix 0.125 0.250 0.500 1.000

I Patch treatment (m)

d 2011

b 0.6 2012

e 0.4

e i 0.2

v 10 20 30

n I 2 Patch edge:area (m/m )

Figure 3.13. Relative spatial abundance of invader (non-seeded) species and patch edge to area ratio (m/m2) separated by year.

r 60

t 2011

l 2012

e 40

f 20

N 0 10 20 30 2 Patch edge:area (m/m )

Figure 3.14. Total number of invader (non-seeded) species clusters and patch edge to area ratio (m/m2) separated by year.

Tables

Species Resp.1 Resp.2 Andropogon gerardii -0.5829 -0.1645 Bouteloua curtipendula -0.4751 -0.1033 Coreopsis lanceolata 0.0187 -0.0225 Desmodium canadense -0.3101 0.0516 Elymus canadensis -0.3483 0.2781 Elymus virginicus 0.5770 0.0328 Heliopsis helianthoides 0.8243 0.0734 Monarda fistulosa 0.4169 0.1721 Oligoneuron rigidum -0.4766 0.3722 Panicum virgatum -0.7924 0.0810 Penstemon digitalis -0.3146 0.4102 Ratibida pinnata -0.3422 0.3033 Schedonorus arundinaceus 0.0929 0.1390 Schizachyrium scoparium -0.4026 -0.4233 Sorghastrum nutans -0.4554 0.0773 Symphyotrichum novae-angliae -0.7120 -0.0918

Species Resp.1 Resp.2 Andropogon gerardii -0.4678 0.0989 Bouteloua curtipendula -0.3489 0.3685 Coreopsis lanceolata 0.3622 0.3306 Desmodium canadense 0.1693 0.1316 Elymus canadensis 0.2111 -0.2381 Elymus virginicus 0.5049 0.4352 Heliopsis helianthoides 0.9395 -0.0521 Monarda fistulosa 0.4901 0.1104 Oligoneuron rigidum -0.3712 -0.1483 Panicum virgatum -0.5557 0.0567 Penstemon digitalis -0.0041 0.1397 Ratibida pinnata -0.4618 0.1252 Schedonorus arundinaceus 0.3645 0.4088 Schizachyrium scoparium -0.2805 0.5223 Sorghastrum nutans -0.2577 -0.3324 Symphyotrichum novae-angliae -0.3651 0.0564

Species Resp.1 Resp.2 Andropogon gerardii -0.5651 -0.0306 Bouteloua curtipendula -0.5187 -0.4732 Coreopsis lanceolata -0.1062 0.7071 Desmodium canadense -0.5113 0.1075 Elymus canadensis -0.5724 -0.0801 Elymus virginicus 0.7607 0.1527 Heliopsis helianthoides 0.9709 -0.0421 Monarda fistulosa 0.8503 0.3216 Oligoneuron rigidum -0.6242 -0.1578 Panicum virgatum -0.7095 -0.0323 Penstemon digitalis -0.4754 -0.1723 Ratibida pinnata -0.0848 0.8902 Schedonorus arundinaceus 0.0284 0.7252 Schizachyrium scoparium -0.2698 -0.0849 Sorghastrum nutans -0.3458 0.0062 Symphyotrichum novae-angliae -0.4581 0.1807

Table 3.4. Standard least squares regression results from analysis of proportional change in number of conspecific clusters of seeded species and patch size, year, and patch × year.

Term Std Err t Ratio Prob<|t| A. gerardii patch size 0.374 4.35 0.0002 * year 0.251 -1.11 0.275 (year-2011.5)*(patch size-0.46875) 0.748 -2.6 0.0147 *

B. curtipendula patch size 0.444 4.03 0.0004 * year 0.298 -1.41 0.1682 (year-2011.5)*(patch size-0.46875) 0.888 -1.71 0.0992

C. lanceolata patch size 0.122 3.97 0.0005 * year 0.082 -0.37 0.7136 (year-2011.5)*(patch size-0.46875) 0.244 -2.21 0.0358 *

D. canadense patch size 0.320 5 <.0001 * year 0.215 -1.27 0.2156 (year-2011.5)*(patch size-0.46875) 0.641 -5.41 <.0001 *

E. canadensis patch size 0.349 5.55 <.0001 * year 0.234 -1.39 0.1748 (year-2011.5)*(patch size-0.46875) 0.698 -6.52 <.0001 *

E. virginicus patch size 0.138 1.99 0.056 year 0.093 0.35 0.7305 (year-2011.5)*(patch size-0.46875) 0.276 -2.46 0.0201 *

H. helianthoides patch size 0.468 4.21 0.0002 * year 0.313 3.39 0.0021 * (year-2011.5)*(patch size-0.46875) 0.935 2.76 0.0101 *

M. fistulosa patch size 0.120 3.83 0.0007 * year 0.081 2.24 0.033 * (year-2011.5)*(patch size-0.46875) 0.240 -2.86 0.008 *

P. digitalis patch size 0.563 2.1 0.0449 * year 0.377 1.64 0.113 (year-2011.5)*(patch size-0.46875) 1.126 1.07 0.2946

P. virgatum patch size 0.418 3.61 0.0012 * year 0.280 0 0.9968 (year-2011.5)*(patch size-0.46875) 0.836 -2.76 0.0101 *

R. pinnata patch size 0.103 2.31 0.0288 * year 0.069 1.6 0.1217 (year-2011.5)*(patch size-0.46875) 0.206 -1.86 0.074

S. nutans patch size 0.261 2.09 0.0461 * year 0.175 -0.57 0.574 (year-2011.5)*(patch size-0.46875) 0.522 -2.42 0.0222 *

O. rigidum patch size 0.585 2.4 0.0235 * year 0.392 0.47 0.6439 (year-2011.5)*(patch size-0.46875) 1.169 -3.08 0.0046 *

* P<0.05

Table 3.5. Effect tests from repeated measures ANOVA for relative spatial abundance of plant functional groups and patch size treatment (m) (patch trt) across 2011-2012.

Source df F-Ratio Prob > F C3 grass patch treatment 4 3.735 0.034 * year 1 26.507 0.000 * patch treatment × year 4 1.291 0.318

C4 grass patch treatment 4 3.300 0.048 * year 1 0.030 0.864 patch treatment × year 4 2.335 0.103

Forb patch treatment 4 53.882 <.0001 * year 1 109.747 <.0001 * patch treatment × year 4 5.825 0.005 *

Legume patch treatment 4 0.941 0.474 year 1 12.911 0.003 * patch treatment × year 4 5.486 0.006 *

* P<0.05

Table 3.6. Standard least squares regression results from analysis of relative spatial abundance of functional groups and patch edge to area ratio (m/m2), year, and patch edge to area ratio × year.

Term Std Err t Ratio Prob>|t|

C3 grass patch edge:area 0.0006 3.43 0.0015 * year 0.0136 3.4 0.0017 * (patch edge:area-17.64)*(year-2011.5) 0.0012 -0.07 0.9477

C 4 grass patch edge:area 0.0004 -3.4 0.0017 * year 0.0097 -0.07 0.9481 (patch edge:area-17.64)*(year-2011.5) 0.0009 -0.81 0.4206

Forb patch edge:area 0.0009 11.7 <.0001 * year 0.0192 7.36 <.0001 * (patch edge:area-17.64)*(year-2011.5) 0.0018 3.96 0.0003 *

Legume patch edge:area 0.0002 -0.83 0.413 year 0.0045 1.82 0.0763 (patch edge:area-17.64)*(year-2011.5) 0.0004 -1.76 0.0871

*P<0.05

Term Std Err t Ratio Prob>|t| Invader species relative spatial abundance patch edge:area 0.001 -10.17 <.0001 * year 0.023 -8.46 <.0001 * (patch edge:area- 17.64)*(year-2011.5) 0.002 -2.57 0.0145 *

Invader species no. clusters patchedge:area 0.128 6.54 <.0001 * year 2.792 4.42 <.0001 * (patchedge:area- 0.255 1.65 0.1071 17.64)*(year-2011.5)

*P<0.05

Chapter 4: The effect of seeded patch size on arthropod abundance and community composition

Abstract

Arthropod predation, competition, and herbivory varies with different parts of the habitat, and these interactions are likely influences by the size and pattern of resource patches. In 1973, Root proposed two hypotheses to explain these effects: the resource concentration hypothesis predicted that larger or more concentrated resource patches should have greater abundance of specialist herbivore insects; and the enemies hypothesis predicted a positive relationship between plant species richness and enemy (predators and parasitoids) abundance. While some experimental and observational studies have supported these predictions, a general relationship remains unclear. Both of these hypotheses are based on the spatial distribution of resources, thus plant patch size should affect the arthropod community. I used an experimental approach to evaluate how seeded patch size affects the arthropod community at the family-level and across trophic groups in an early establishing reconstructed grassland located in Guelph, Ontario. Using plots of various seeded patch sizes (1 m, 0.5 m, 0.25 m, 0.125 m, mixed) in a randomized compete block design; I sampled arthropods in June-July of 2011-2012 with a Vortis suction sampler, identified individuals to family and assigned trophic groups based on published literature. A redundancy analysis showed that patch size had no overall effect on arthropod species community composition, however a few families and trophic groups had strong relationships. Aphididae, Chalcidoidea, and Phoridae absolute abundances all had positive linear relationships with seeded patch size, as did the absolute abundances of herbivores and parasitoids. A repeated measures ANOVA revealed that the relative abundance of the predator group was also strongly related to patch size, although not linearly, and was greatest in the uniformly mixed seed plots. Within a grassland restoration context, these results

provide evidence of the resource concentration hypothesis occurring at the arthropod level based on resource patch size, as well as suggests possible bottom-up effects on other trophic groups mediated by herbivores. Furthermore, patch size not only affects the abundance of certain arthropods, it seems to also influence the composition of the community, or the proportion of predators. Some patch sizes may be used to help maximize the abundance of beneficial or biological control arthropod species in reconstructed communities.

4.1 Introduction

Arthropod species differ in their degree of mobility, thus it is likely that plant spatial arrangement will impact the arthropod population. Some arthropod species maintain stable local populations, whereas others have relatively widespread distribution and experience local extinctions (Gripenberg and Roslin, 2007). Furthermore, herbivore, predator, and competitive interactions among arthropods vary among different parts of the landscape, and may differ due to the size and pattern of host or resource patches (Holyoak et al., 2005). Detection and location of resource patches by herbivorous insects occurs at various spatial scales using several possible mechanisms: contact, visual, and olfactory cues, which can vary between species (Andersson et al., 2013). Insects may even switch between cues to find a suitable host patch, and since these multiple cues act on multiple scales, the cue with the longest acting distance may cause greater arrival rates or species abundance in a resource patch (Andersson et al., 2013). Since foraging and predation behaviour varies among species and trophic groups, responses to resource patches should be heavily influenced by spatial scale (Levin, 1992; Andersson et al., 2013).

Previous work has shown that not just the abundance, but also the pattern of spatial aggregation of host plants affects the distribution and abundance of herbivorous insects. In 1973, Root proposed the resource concentration hypothesis, which predicts that specialist herbivorous insects should be more abundant in more

concentrated or larger patches of host plant species, where the probability of finding and remaining on resources is higher than in less dense or smaller host plant patches (Root, 1973; Grez and González, 1995). The resource concentration hypothesis also predicts that specialist herbivorous insects should be higher in abundance in monocultures or larger host-plant patches which are easier to locate (Bach, 1988; Andow, 1990; Borges and Brown, 2001), and that generalist predators are likely to be less abundant in monocultures (Grez and González, 1995). Root also proposed the “enemies” hypothesis (Murdoch et al., 1972; Root, 1973), based on premises by MacArthur (1955), which predicts that the abundance of predatory and parasitoid arthropods should be greater in more diverse plant communities, as more diverse plant communities support greater prey or host diversity. Based on these hypotheses, plant patch size should influence species abundance and the composition of the arthropod community.

The relationship between patch size and the herbivore community may seem self-explanatory. However, as discussed above, different arthropod species use different cues for resource-location and selection (Andersson et al., 2013), thus not all arthropods and trophic groups may respond the same to plant patch size. At the fine scale, we can characterize a habitat by: the size and number of host, resource, of habitat patches; as well as the edges and surrounding matrix of host patches. Host- plant patch size may have a negative relationship with arthropod diversity for species with risk-spreading movement behavior or species that spread their eggs over space (Fahrig and Jonsen, 1998). Furthermore, in cases where predator species dispersal ability is lower than that of its prey, patch size can also have a negative relationship with prey species richness (Fahrig and Jonsen, 1998). Edge and matrix effects from neighbouring plant species and resources also affect the arthropod community. Although edge effects may increase biodiversity due to the transition zone between different habitats, the survival of species common to the initial patch habitat may be negatively affected (Tscharntke and Brandl, 2004). Thus, arthropod species diversity may be determined in part by patch size and edge effects, which together may act at multiple scales (Debinski et al., 2001). Some studies have even

shown negative or natural effects of patch size on the arthropod abundance or diversity (Grez and González, 1995; Rhainds and English-Loeb, 2003; Crist et al., 2006). Thus, a universal understanding of patch size effects in natural grassland communities has not yet been reached (Perner et al., 2005).

My goal in this chapter was to examine how initial plant spatial patterning may influence arthropod diversity. Specifically, I evaluated if seeded patch size affected arthropod species and trophic group abundance. The resource concentration hypothesis may accurately predict patch size responses in some species or trophic groups, such as herbivores, but this response may not be consistent throughout the aboveground arthropod community. These results will help to clarify how plant spatial patterning in reconstructed grassland ecosystems affect arthropod diversity at the community and trophic level, and if there are planting strategies that help to maximize beneficial arthropod populations in natural communities.

4.2 Methods

Study site

(Roundup WeatherMAX, Monsanto, St. Louis, MO, USA). Debris was removed and the site was cultivated and leveled with a Blecavator (Specialized Turf Equipment Company, South Carolina, USA) to homogenize the vegetation and soil.

Experimental design

All selected species are native to Southern Ontario, except for S. arundinaceus, and co-occur throughout Ontario (USDA and NRCS, 2015). They were selected based on their availability and use in regional grassland restoration efforts. S. arundinaceus was included to test for effects of initial distribution on resistance to invasion of a species considered problematic in restored grasslands (Cully et al., 2003). This mixture of species represents typical functional diversity found in reconstructed tallgrass prairies (Stroud, 1941; Packard et al., 1997). S. arundinaceus (Cultivar: KY-31) seeds were obtained from the University of Kentucky and seeds of the remaining species were obtained from Wildflower Farm

(Coldwater, Ontario, Canada). Seeds were dry-stored at 17°C for seven weeks prior to planting.

Four blocks of seven plots each (five treatment plots, one replicate of each seeded patch size, and two monoculture plots used for expected yield calculations in Chapter 2) were aligned in rows from East to West, and separated by 2 m mowed aisles (Appendix Fig 1). Experimental patch size treatment plots (n = 20) were sown with increasingly larger conspecific patches in five different treatment levels. For each species, the 1 m2 total planted area was divided into the following treatments: one – 1 × 1 m patch, four – 0.5 × 0.5 m patches, 16 – 0.25 × 0.25 m patches or 64 – 0.125 × 0.125 m patches (Appendix Fig 2). Patch edge to area ratios for the above treatments were equivalent to 4, 8, 16 and 32 m/m2 respectively. Species were randomly assigned to patch locations within plots. All seeds for the fifth (mixed) treatment level were intermixed and hand broadcast. Broadcast seeding was chosen over drill seeding as broadcasting can result in greater seed-establishment, reduced new species establishment (Yurkonis et al., 2010), and a natural distribution of plants along existing soil heterogeneity patterns in ecological restoration (Wilson, 2002).

Data collection

Arthropods were sampled in all patch treatment plots using a Vortis Suction Sampler (Burkard Manufacturing Co. Ltd, Hertfordshire, UK) at approximately midday during late June – early July of 2011 and 2012. The sampling area of this apparatus is 0.2 m2, and the apparatus was rotated in a clockwise direction during

sampling. Samples were collected for a 60 s interval at five sample points per plot: at the centre of the plot and from a distance of 1m at 45 from each of the four plot corners to avoid edge effects. These sampling points were intersections of seeded plant patches in all plots (Appendix Fig 2). Collected samples were stored in a freezer at -23°C until samples were sorted. Using a binocular zoom microscope (Model: WF 10X/23, Fisher Scientific), arthropods were identified to family. In some cases arthropods were identified to superfamiliy, suborder, or order. Individuals that were too damaged to identify (missing or damaged anatomy vital for identification), or taxonomically indistinguishable larva and nymphs (<1% of all individuals collected) were not included in analyses. Hemipteran nymphs were identified to family. Juvenile holometabolous insects were not included in the analyses except for Lepidopterans (identified to order). Arthropods that were damaged prior to collection were most likely already dead, and were not counted to avoid spatiotemporal variability (e.g. could have been carried by wind, bird, predator). Species were assigned a general trophic grouping (e.g. detritivore, herbivore, omnivore, parasitoid, or predator) according to published literature and similar research, with special consideration of sampled species and known grassland species (Table 4.1).

Data analyses

Using the software Canoco 5 (Microcompouter Power; Ithaca, NY, USA), I conducted partial redundancy analyses (RDA) to test the effect of patch size treatment on arthropod community composition with absolute abundance of families, employing hierarchical permutations for the split-plot repeated-measures design (499 unrestricted permutations). Since multivariate tests can be skewed by low abundance values, I used a cut-off equal to the number of samples or observations in the dataset (n = 40) (McGarigal et al., 2000). I then used JMP (JMP ver. 11, Cary, NC: SAS Institute Inc.) to perform standard least squares regressions using Box-Cox (Box and Cox, 1964) transformed family abundances as the response variables and the effects: patch edge to area ratio, year, and patch edge to area ratio × year to test for a

linear response to patch size on individual arthropod families. The Shapiro-Wilk test (Shapiro and Wilk, 1965) was used to determine if abundances were from a normal distribution prior to and post data-transformation. Transformed abundances that were not normally distributed were not included in statistical analyses.

For trophic analyses, I used the absolute and the relative trophic group abundances in separate analyses. Relative abundance was calculated as the proportion of each species divided by the total number of individuals sampled per plot (n/N). I performed repeated measures ANOVA in JMP with arthropod trophic group abundances (Box-Cox transformed) as the response variables, using the effects block(random), patch size, block(random) × patch size, year, and patch size × year as model terms to determine if patch size had an effect on trophic groups. I also performed standard least squares regressions in JMP using the same effects as described above to test for a linear response to patch size in absolute and relative trophic group abundances (Box-Cox transformed).

4.3 Results

The 2011 (year 1) mean air temperature at the Guelph Turfgrass Institute was 7.06 °C, and total annual precipitation was 1090 mm, while in 2012 (year 2) the mean air temperature was 8.36 °C, and total annual precipitation was 1075 mm (Agricultural and Forest Meteorology Group, 2013). Based on a patch and edge analysis, the mixed seeded plots developed an average patch edge to area ratio of 28.2 m/m2 by 2011 (see Chapter 2), which was used in the regression analyses.

In year one, I collected 7722 individual arthropods that spanned 68 taxonomic families, and in year two I collected 14986 individuals that spanned 70 families. The following 16 arthropod families were present in year one but not year two: Apidae, Asteiidae, Bombyliidae, Cantharidae, Curtotonidae, Cymidae, Derbidae, Membracidae, Milichidae, Piophilidae, Sciomyzidae, Sepsidae, Syrphidae, Tephritidae, Thyreocoridae, and Trichoptera (order). By contrast, the following 18

families were present in year two but not year one: Aleyrodidae, Anthicidae, Berytidae, Bethylidae, Carabidae, Chilopoda, Cleridae, Cryptophagidae, Dictyopharidae, Galumnidae, Geocoridae, Lepidopsocidae, Mantidae, Neuroptera (order, non-Chrysopidae), Nitidulidae, Phymatidae, Reduviidae, and Rhyparochromidae.

Arthropod family abundances

In the RDA using patch size, year, and block, the explanatory variables explained approximately 55.1% of the total variation in arthropod family abundance (40 samples, 31 families) (pseudo-F=8.4, P<0.01) (Fig 4.1, Table 4.2). Year was separated along axis 1 (eigenvalue = 0.4864), and patch size was separated along axis 2 (eigenvalue = 0.035) (Fig 4.1, Table 4.2). Since year had such an overwhelming impact on arthropod abundances, I conducted separate RDAs for each year to explore the treatment effects. In year one, explanatory variables accounted for 18.9% of the within year variation in arthropod family abundance (axis 1, 2 eigenvalues = 0.0745, 0.0665), and the effects were not significant on overall arthropod family abundance (pseudo-F=0.9, P>0.05) (Fig 4.2, Table 4.3). Explanatory variables accounted for 22.7% of the variation in year two (axis 1, 2 eigenvalues = 0.1165, 0.0501), and again the effects were not significant on overall arthropod family abundance (pseudo-F=1.1, P>0.05) (Fig 4.3, Table 4.4). Patch size had a linear effect on the families Aphididae (Std Err=0.1644, t Ratio=-2.16, P<0.05), Chalcidoidea (Std Err=0.2704, t Ratio=-2.39, P<0.05, and Phoridae (Std Err=0.0433, t Ratio=-2.22, P<0.05) (Fig 4.4). These families increased in absolute abundance with smaller patch edge to area ratio (or larger patch size).

Arthropod trophic group abundances

Herbivore (Std Err=0.7108, t Ratio=-3.05, P<0.005) and parasitoid (Std Err=0.289734, t Ratio=-2.91, P<0.01) absolute abundance both increased with smaller patch edge to area ratio, or larger patch size (Fig 4.5). Patch size also had an effect on the relative abundance of predators (F4,12 Ratio=4.4276, P<0.05) (Fig 4.6).

Predator relative abundance was greatest in the mixed plots across both years, and lowest in the largest patch plots (1 m). Only these two patch size treatments (mixed and 1 m) had means that were significantly different from each other (see Tukey’s test, Fig 4.6).

4.4 Discussion

Several families and trophic groups showed strong relationships with patch size. The families Aphididae (Hemiptera) and Phoridae (Diptera), and the superfamily Chalcidoidea (Hymenoptera) had greater absolute abundance in larger patch size plots. Interestingly, each of these families is generally considered to be from separate trophic groups (herbivores, detritivores, and parasitoids respectively) (Table 4.1). As predicted by the resource concentration hypothesis (Root, 1973; Grez and González, 1995), herbivorous insects such as Aphididae would most likely benefit from increased resource area in larger patch plots. Conversely, some studies investigating patch size and, perhaps consequently, the resource concentration hypothesis have shown patch size to have a neutral or negative effect on several Aphididae species (Rhainds and English-Loeb, 2003). The Aphididae species in the various studies summarized by Rhainds and English-Loeb (2003) were agricultural, cruciferous, and other crop pests (MacGarvin, 1982; Grez and González, 1995). I did not identify to species in my research as family-level diversity is commonly used for arthropod studies, and in many cases considered sufficient for biodiversity assessment of terrestrial arthropods (Timms et al., 2013). Therefore, it is possible that not all Aphididae species follow the same pattern with resource patch size, and the relationship could be variable at finer taxonomic resolution. Furthermore, the Aphididae in the above studies were from Chile (Grez and González, 1995) and Great Britain (MacGarvin, 1982). Nevertheless, overall herbivore absolute abundance in my research was positively related to larger patch sizes as well, again consistent with the resource concentration hypothesis. Thus while some previous research has shown conflicting relationships, this research provides empirical evidence that within the first few years of a grassland reconstruction, herbivorous insects

generally prefer larger resource patches. This relationship may not be universal, however, and response to patch size may vary with time (Rhainds and English-Loeb, 2003).

Chalcidoidea species are parasitoids, although there are some herbivorous species such as seed chalcids (of the family Eurytomidae) (Crist et al., 2006). Their abundance could have increased as a bottom-up effect of greater herbivore (host) abundance, such as aphids. Previous research by Fenoglio et al. (2010) has shown plant patch size to have a positive effect of parasitoid species richness. Indirect effects of patch size, such as increased host abundance, could have driven the results. However, once the influence of host abundance was removed, parasitoid species richness was still significantly affected by patch size, perhaps due to direct effects on parasitoid foraging behaviour from increased host chemical cues and habitat-related visual cues (Fenoglio et al., 2010). It is possible that patch size had similar effects on parasitoid abundance in my research. Although some chalcid wasps may parasitize beneficial insects, many have economic importance and some are valuable pollinators (Gokhman, 2015). If the majority of parasitoid species sampled in my research are of considerable importance, again reconstruction and restoration strategies using larger seeded patches may be particularly advantageous. Furthermore, these parasitoids may be specialists, and thus would prefer a greater concentration of specific prey species, more likely to occur in the larger patch plots with greater host plant aggregation (resource concentration hypothesis).

Many Phoridae species are detritivores, although some are parasitoids and predators (Resh and Cardé, 2003). Parasitoid phorid flies are ecologically relevant and potential biological control agents of fire ants (Mathis and Philpott, 2012), which are serious invasive pests throughout the Southern American prairie. Should parasitoid phorid flies respond to patch size in the same manner as in my research, prairie reconstruction efforts throughout fire ant infested areas may benefit from using larger patch size plots. However, the mechanism explaining why Phoridae

increased in abundance in larger patch plots is unclear, especially since overall detritivorous arthropod abundance was not affected by patch size. Moreover, if the Phoridae sampled were actually mostly predaceous, predator abundance did not share the same relationship to patch size.

Predator relative abundance had a strong response to seeded patch size and was greatest in the mixed plots, although this effect was only significantly different from the 1 m patch plots. The 1 m plots in this study most closely resemble monoculture plots, composed of only one large patch of each seeded plant species (Appendix Fig 2). The majority of predators sampled in my study may have been generalist predators, which are predicted to be less abundant in monocultures, and more abundant in diverse communities (enemies hypothesis) (Grez and González, 1995; Borges and Brown, 2001; Haddad et al., 2001). However, absolute abundance of predators was not affected by patch size. It seems patch size not only influences the abundance of herbivores and parasitoids, it also affects trophic composition by increasing the proportional number of predators. Although Araneomorphae abundance was not affected in this experiment, complex patches offer better habitat for web-building spiders (Reid and Hochuli, 2007), and mixed plots may have greater fine-scale vegetation complexity than larger patch plots.

While this research has shown various responses by the arthropod community to patch size, it would be presumptuous to assume generality about these relationships. These experimental plots may be more similar to early successional communities rather than later seral stages, and the types and diversity of species sampled could be constrained by different stages of succession. Since terrestrial arthropod diversity has been found to increase with field age (particularly for herbivorous and parasitic arthropods), due to increases in plant diversity (Siemann et al., 1999), this study should be repeated in the following years to determine how plant patch size and successional dynamics influence the arthropod community. Furthermore, there may have been limitations in dispersal and colonization of native arthropod species from regional grasslands to the study

site, and so results presented here may not necessarily represent actual arthropod communities of North American grasslands.

Supplementary methods of arthropod surveying (e.g. sweep-nets, pitfall traps, malaise traps) could provide another level of arthropod diversity information that was neglected by the Vortis sampling method. Sweep-netting was rejected as a sampling method in this study as it can be destructive to surrounding vegetation, which consequently may have interfered with the plant biomass harvest results (Chapter 2). The Vortis sampler has the advantages of being portable, easy to operate, and relatively rapid. Additionally, it has the ability to capture smaller species and juvenile stages of arthropods (Borges and Brown, 2003). However, Borges and Brown (2003) recommend that suction and pitfall trapping both be used as complementary methods for sampling completion, as pitfall traps may sample chewing insects better than other sampling methods. In the future, it would be advantageous to include both sampling methods.

This research has several implications for future biodiversity research and restoration application. Patch size, when manipulated at seeding in grassland reconstruction, has significant effects on the arthropod community at several levels. Aphididae, Chalcidoidea, and Phoridae species, as well as arthropod herbivores and parasitoids were positively related to patch size. Mixed plots, which are traditionally used in restoration projects (Fargione et al., 2007; Cadotte, 2013; Schittko et al., 2014), had greater predator relative abundance. Resource patch size is clearly an important determinant in grassland arthropod diversity and abundance in early communities.

Acknowledgements

I wish to thank H. Kovacs, F. Cibula, and E. Palmer for their assistance with arthropod collection. For arthropod identification, I thank J. Holdenried, K. Shukla, and A. Patchett for their generous help. I also wish to thank C. Arsenault, R. Bisaillon,

and N. Castellano for their contribution to arthropod identification of the 2011 samples, and for choosing to help in this project to enrich their undergraduate research experience. Finally, I thank Dr. Heather Hager for her assistance with the multivariate statistical analyses.

Figures

Figure 4.1. Arthropod family relative abundance composition biplot in relation to patch size edge length (mix, 0.125, 0.25, 0.5, 1 m), and year analyzed with RDA. For species listing see Table 4.2.

a) b) e

200 n

a n 100

b u 150

a 80

b 100 60

o 40

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0 C 0 10 20 30 0 10 20 30 Patch edge:area (m/m2) Patch edge:area (m/m2) c)

a 15

s 10

r 5

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a) b)

n n 300

u u 100

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P 0 10 20 30 0 10 20 30 Patch edge:area (m/m2) Patch edge:area (m/m 2)

c a

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mix 0.125 0.250 0.500 1.000 Patch treatment (m)

Tables

Table 4.1. Trophic group designations for arthropod taxon groups, based on published literature. Groupings were assigned based on the majority of described grassland species.

Arthropod taxon group Trophic group Reference Aleyrodidae Herbivore (Marshall, 2006) (Resh and Cardé, 2003; Gillott, 2005; Anthocoridae Predator Koh and Holland, 2015) Anthomyzidae Herbivore (Keiper et al., 2002) Aphididae Herbivore (Borges and Brown, 2001) Araneomorphae Predator (Turnbull, 1973) Braconidea Parasitoid (Resh and Cardé, 2003) Cercopidae Herbivore (Hamilton and Whitcomb, 2010) Chalcidoidea Parasitoid (Resh and Cardé, 2003) Chloropidae Herbivore (Keiper et al., 2002) Chrysomelidae Herbivore (Strauss, 1988) Cicadellidae Herbivore (Hamilton and Whitcomb, 2010) Coccinellidae Predator (Joshi et al., 2012) Collembola Detritivore (Hopkin, 1997) Delphacidae Herbivore (Borges and Brown, 2001) Drosophilidae Detritivore (Gullan, 2014) Empididae Predator (Gillott, 2005) Ephydridae Predator (Keiper et al., 2002) Formicidae Predator (Koricheva et al., 2000) Galumnidae Omnivore (Behan-Pelletier and Kanashiro, 2010) Lepidoptera Herbivore (Chapman, 2009) (Resh and Cardé, 2003; Crist et al., Miridae Herbivore 2006) Nabidae Predator (Resh and Cardé, 2003; Gillott, 2005) Nematocera Detritivore (Curry, 1994) Parasitidae Predator (Behan-Pelletier and Kanashiro, 2010) (Koricheva et al., 2000; Crist et al., Pentatomidae Herbivore 2006) Phoridae Detritivore (Resh and Cardé, 2003) Proctotrupidae Parasitoid (Resh and Cardé, 2003) Rhyparochromidae Herbivore (Resh and Cardé, 2003) Sphaeroceridae Detritivore (Keiper et al., 2002) Thysanoptera Herbivore (Koricheva et al., 2000) Trombidiidae Predator (Zhang, 1998)

Arthropod taxon group Resp.1 Resp.2 Arthropod taxon group Resp.1 Resp.2 Aleyrodidae 0.7744 -0.0436 Ephydridae -0.0037 0.1326 Anthocoridae -0.8942 0.0156 Formicidae 0.5964 0.4275 Anthomyzidae -0.1672 -0.0257 Galumnidae 0.9214 0.0991 Aphididae 0.2996 0.4362 Lepidoptera -0.4158 0.2816 Araneomorphae 0.8977 -0.0559 Miridae -0.6046 0.3848 Braconidea -0.8587 0.201 Nabidae 0.3668 0.0005 Cercopidae -0.7401 -0.0207 Nematocera 0.0161 0.0734 Chalcidoidea -0.0749 0.3273 Parasitidae 0.8566 0.0138 Chloropidae -0.2487 0.2468 Pentatomidae 0.2233 0.2987 Chrysomelidae 0.7577 0.009 Phoridae -0.5101 0.2937 Cicadellidae -0.3119 0.2048 Proctotrupidae 0.3793 0.1619 Coccinellidae 0.3274 0.0989 Rhyparochromidae 0.8052 0.1678 Collembola 0.7087 0.0126 Sphaeroceridae -0.8007 0.2242 Delphacidae 0.286 0.1572 Thysanoptera 0.2024 0.1813 Drosophilidae -0.6428 0.1336 Trombidiidae 0.7009 -0.0709 Empididae -0.3898 0.0668

Arthropod taxon group Resp.1 Resp.2 Arthropod taxon group Resp.1 Resp.2 Anthocoridae 0.1192 -0.1702 Empididae -0.3208 0.5951 Anthomyzidae -0.026 -0.1414 Ephydridae -0.0726 0.6316 Aphididae 0.3429 0.1866 Formicidae 0.6773 -0.2429 Araneomorphae -0.0779 -0.0773 Lepidoptera -0.3508 -0.003 Braconidea 0.1584 0.3664 Miridae 0.2977 0.4718 Cercopidae -0.1618 0.0383 Nabidae 0.4794 -0.2732 Chalcidoidea 0.183 0.1452 Nematocera -0.0218 -0.1187 Chloropidae 0.3486 0.0264 Parasitidae 0.1115 0.1127 Chrysomelidae 0.1535 -0.0546 Pentatomidae 0.5992 0.1306 Cicadellidae 0.0793 -0.0124 Phoridae 0.1004 0.0985 Coccinellidae 0.2768 0.0959 Proctotrupidae -0.0159 0.0696 Collembola 0.189 -0.1347 Sphaeroceridae 0.1844 0.3483 Delphacidae 0.2349 0.0389 Thysanoptera 0.2741 -0.118 Drosophilidae 0.0414 0.0127 Trombidiidae -0.1079 -0.1055

Arthropod taxon group Resp.1 Resp.2 Arthropod taxon group Resp.1 Resp.2 Aleyrodidae -0.0517 -0.0128 Ephydridae 0.0872 -0.2683 Anthocoridae -0.0974 0.0021 Formicidae 0.5558 -0.1417 Anthomyzidae -0.0253 -0.5014 Galumnidae 0.4393 -0.1683 Aphididae 0.5891 0.3558 Lepidoptera 0.756 -0.0707 Araneomorphae -0.0949 -0.26 Miridae 0.4302 -0.2291 Braconidea 0.3931 -0.3882 Nabidae -0.2337 0.2173 Cercopidae 0.0817 0.2349 Nematocera 0.2539 -0.1744 Chalcidoidea 0.4522 -0.2336 Parasitidae -0.0022 0.0316 Chloropidae 0.2442 0.1759 Pentatomidae 0.0801 0.1389 Chrysomelidae -0.1262 -0.1255 Phoridae 0.4396 -0.3364 Cicadellidae 0.3137 0.0833 Proctotrupidae 0.2911 -0.0514 Coccinellidae -0.1164 -0.0649 Rhyparochromidae 0.4732 0.2595 Collembola -0.0484 0.1124 Sphaeroceridae 0.3155 -0.4852 Delphacidae 0.1574 0.0334 Thysanoptera 0.1625 -0.3244 Drosophilidae 0.4065 -0.1928 Trombidiidae -0.121 -0.2604 Empididae 0.2302 0.0333

Chapter 5: General conclusions

Biodiversity has been shown to have a positive relationship with ecosystem functions (such an productivity and resistance to invasion), as well as greater temporal stability of these functions (McCann, 2000; Tilman et al., 2006; Cardinale et al., 2012; Naeem et al., 2012). It is unclear how the fine-scale patterning of plant species affects the inter- and intraspecific interactions that drive these responses in biodiversity and ecosystem functions. However, research has shown that plants have stronger effects on their close neighbours, and fine-scale spatial pattern strongly regulates the magnitude and direction of inter- and intraspecific interactions between plants (Tilman, 1994; Kennedy et al., 2002; Vogt et al., 2010; Rayburn and Schupp, 2013). Moreover, spatial pattern may strongly influence species coexistence, competition, and community structure (Monzeglio and Stoll, 2005; Yurkonis et al., 2012).

There is a need to evaluate fine-scale plant patterning effects both to further understanding in the BEF field and for restoration applications. Current planting strategies for restoration projects usually employ uniformly mixed broadcast seeding compositions (Fargione et al., 2007; Cadotte, 2013; Schittko et al., 2014). This may not take full advantage planting patterns that balance inter- and intraspecific interactions in order to maximize biodiversity and ecosystem function. Uniformly mixed seeding communities may experience competitive exclusion due to high interspecific competition (Stoll and Prati, 2001). To empirically examine the effects of fine-scale spatial patterning on biodiversity, ecosystem functions, and species interactions in perennial tallgrass communities, I used varying seeded conspecific patch size treatments in experimentally reconstructed grassland plots.

In Chapter 2, I tested if initial species patterning affects the species interactions that determine productivity, diversity, and invasion resistance. Throughout the three years, productivity was highest in smaller patch plots, however these plots had greater competitive exclusion, and thus lower plant

diversity than the larger patch plots. In the spatially seeded larger patch plots, under greater intraspecific competition, plant diversity was maintained at higher levels. The selection effect greatly affected species interactions, especially in the smaller patch plots, instead of the complementarity effect. Based on a patch and edge analysis, the mixed seeded plots developed an average patch edge to area ratio of 28.2 m/m2 by 2011, and consistently produced less biomass than the smallest patch plots (patch edge to area ratios of 16 and 32 m/m2) but more than the larger patch plots (patch edge to area ratios of 4 and 8 m/m2). Invasion by non-seeded species was greatest in the largest patch plots, although invasion generally decreased throughout the years. Results were consistent with Yurkonis et al. (2012) where non-seeded species were more abundant in larger patch plots, although their experiment was conducted using transplanted seedlings.

In Chapter 3, I found that there were species-specific and functional group- specific responses to patch size at seeding by evaluating spatial relative abundance and patch or cluster number in landscape analyses. E. virginicus, H. helianthoides, and M. fistulosa, relatively dominant species throughout the system, had greater abundance in larger patch plots, indicating they were stronger interspecific competitors. A. gerardii, P. virgatum, and S. novae-angliae were higher in abundance in larger patch plots, suggesting they are weaker interspecific competitors, or stronger intraspecific competitors in establishing grassland communities. With time, the stronger interspecific competition species may competitively exclude the weaker, but this could be contingent on disturbance or density-dependent factors

(Connell, 1983). Forb and C3 grasses were more abundant in mixed and smaller patch plots (stronger interspecific competitors), while C4 grasses were greater in larger patch plots (stronger intraspecific competitors). Legumes, which may suffer in nutrient poor environments (such as restored prairie grassland), had greatest spatial abundance in 0.25 m plots.

Using landscape analyses in Chapter 3, I also considered how non-seeded species were invading and subsequently establishing in these plots. The spatial

abundance of invading non-seeded species in restored grassland plots was lower in smaller patch plots, although the number of clusters of non-seeded species was higher. It appears that while fine-scale invasion may have higher frequency in smaller patch plots, subsequent establishment was not affected, possibly due to greater interspecific competition from the planted forbs. In the future, it would be advantageous to identify the non-seeded species and track their spread, dispersal, and possibly coexistence and competition.

Potential limitations and confounding factors from the experiments in Chapters 2 and 3 were that some species had comparatively low establishment throughout the study (P. digitalis, S. scoparium, and S. novae-angliae), leaving those seeded patches relatively uncolonized and open for invasion from non-seeded species. In the larger patch plots, up to 1 m2 of patch area had little establishment of seeded species, resulting in a larger area that may not have been as easily colonized by neighbouring seeded species than in smaller patch plots. This may have had a confounding effect on invasion resistance, which was generally lower in the larger patch plots.

The final focus of my research in Chapter 4 was to assess how plant spatial pattern influenced the arthropod community. I found that the seeded patch size of native grassland species had strong linear effects on the abundance of several insect families. Aphididae, Chalcidoidea, and Phoridae abundance increased with larger patch sizes throughout both years of the study. The abundances of arthropod herbivores and parasitoids were also positively related to patch size, increasing in larger patch plots. Moreover, the relative abundance of arthropod predators was lowest in the largest patch plots, and highest in the mixed plots. These results suggest that plant spatial pattern and aggregation in an establishing grassland community strongly influences the abundance of common arthropod species, and affects trophic groups and community composition.

Temporal and climatic variation may have affected these establishing, early successional reconstructed communities, although yearly variations in arthropod communities are common and have been previously described (Andrewartha and Birch, 1954). This study should be repeated in the following years since arthropod diversity may increase with field age in response to increases in plant diversity (Siemann et al., 1999). Additional variables to be considered in the future are plant height or architecture of the plots, as some arthropod groups may be strongly influenced by vegetation structure (Greenstone, 1984). As recommended by Borges and Brown (2003), pitfall trapping should be considered as a complementary method to Vortis suction for arthropod sampling. Most of the limitations from this study were due to time constraints, and not arthropod identification, although higher power microscopes would have been ideal in reducing sorting and identification effort and time.

In addition to what this work offers to the field of BEF research, there are several implications from a restoration standpoint. Mixed plots generally had lower diversity and ecosystem function, and smaller patch plots had greater productivity (via aboveground biomass) throughout the three-year study. While greater productivity may seem advantageous, the smaller seeded patch plots did not promote the highest plant diversity. Higher productivity in the smaller patch plots was driven by dominance of higher yielding species, as evidenced by the greater selection effect. Thus, larger patch plots may be ideal for restoration application for improved diversity maintenance, although there may be a trade-off with productivity and invasion resistance during early establishment. Larger patch plots had greater invasion by non-seeded species, yet invasion generally decreased overall, and mean invasion among patch sizes was not significantly different in the third year of the study. Using an intermediate conspecific patch size such as 0.5 m at seeding may attain a balance between plant diversity maintenance and ecosystem function.

The fine-scale pattern of plant species may in part determine species competition and coexistence, ultimately affecting biodiversity, ecosystem function, and community composition. To examine fine-scale plant pattern, I manipulated the level of intraspecific aggregation of native species via conspecific patch sizes at seeding in reconstructed grassland plots. I determined that patch size: 1) has significant effects on biodiversity and ecosystem function, 2) generates species and plant functional group-specific responses affecting community composition, and 3) influences arthropod family and trophic group abundance. My future plans are to further test the effects of spatial plant patterning on the arthropod community at larger scales, and throughout various environments and habitats to determine if a general relationship may exist.

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Appendix

Figure 1. Field and plot layout in randomized complete block design. Numbers on top indicate plot number, and codes on the bottom indicate treatment level of conspecific patch edge length (MC = monoculture plot). Plots are 4 m × 4 m in total area, separated by 2 m mowed aisles.

Figure 2. Seeded patch size treatments for experimental plots. The two-letter code for species represents first letters of the genus and specific epithet. All 16 species were planted in total area (1 m2) and density (1728 seeds/m2) for all treatment levels. Species were planted into a) one – 1 × 1 m patch, patch edge to area ratio = 4 m/m2, b) four – 0.5 × 0.5 m patches, patch edge to area ratio = 8 m/m2, c) 16 – 0.25 × 0.25 m patches, patch edge to area ratio = 16 m/m2 or d) 64 – 0.125 × 0.125 m patches, patch edge to area ratio = 32 m/m2.

Table 1. Species list (n = 16) of grassland perennials selected for experiment, including authority, family, common name, seeds/g, plant functional and reproductive group, and distribution in Southern Ontario according to the Ontario Ministry of Natural Resources (Bradley, 2013). Seeds/g was calculated based on the mass of 100 seeds (n=10) for each species. All species are native to Southern Ontario except for S. arundinaceus.

Functional Rep. Distribution in Species Authority Family Common name Seeds/g group Group Southern Ontario Andropogon gerardii Vitman Poaceae Big bluestem 414 C4 grass monocot common Bouteloua curtipendula (Michx.) Torr. Poaceae Sideoats grama 316 C4 grass monocot very rare Coreopsis lanceolata L. Asteraceae Lance-leaved coreopsis 666 forb dicot common Desmodium canadense (L.) DC. Fabaceae Showy tick trefoil 226 Legume dicot common Elymus canadensis L. Poaceae Canada wildrye 177 C3 grass monocot common-widespread Elymus virginicus L. Poaceae Virginia wildrye 133 C3 grass monocot widespread Heliopsis helianthoides (L.) Sweet Asteraceae Ox eye sunflower 187 forb dicot uncertain Monarda fistulosa L. Lamiaceae Wild bergamot 2634 forb dicot widespread Oligoneuron rigidum L. Asteraceae Stiff goldenrod 1571 forb dicot rare Panicum virgatum L. Poaceae Switchgrass 922 C4 grass monocot common Penstemon digitalis Nutt. ex Sims Plantaginaceae Foxglove beardtongue 4074 forb dicot widespread Ratibida pinnata (Vent.) Asteraceae Yellow coneflower 1048 forb dicot rare Schedonorus arundinaceus (Schreb.) Poaceae Tall fescue- pasture; KY_31 476 C3 grass monocot introduced Dumort. Schizachyrium scoparium (Michx.) Nash Poaceae Little bluestem 1109 C4 grass monocot common Sorghastrum nutans (L.)(Nash) Poaceae Indiangrass 480 C4 grass monocot common Symphyotrichum novae- (L.) GL Nesom Asteraceae New England aster 3767 forb dicot widespread angliae