Legacy of Robinia Pseudoacacia Invasion and Use of Ectomycorrhizal Fungi to Restore Pinus Rigida in the Albany Pine Bush Preserve, NY

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Dissertations and Theses

Spring 4-16-2019

Legacy of Robinia pseudoacacia invasion and use of ectomycorrhizal fungi to restore Pinus rigida in the Albany Pine Bush Preserve, NY.

Taylor Patterson [email protected]

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Recommended Citation Patterson, Taylor, "Legacy of Robinia pseudoacacia invasion and use of ectomycorrhizal fungi to restore Pinus rigida in the Albany Pine Bush Preserve, NY." (2019). Dissertations and Theses. 66. https://digitalcommons.esf.edu/etds/66

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LEGACY OF ROBINIA PSEUDOACACIA INVASION AND USE OF ECTOMYCORRHIZAL

FUNGI TO RESTORE PINUS RIGIDA IN THE ALBANY PINE BUSH PRESERVE, NY.

Taylor R. Patterson

A thesis submitted in partial fulfillment of the requirements for the Masters of Science Degree State University of New York College of Environmental Science and Forestry Syracuse, New York April 2019

Department of Environmental and Forest Biology

Thomas R. Horton, Major Professor Anthony Miller, Examining Committee Chairperson Melissa Fierke, Department Chairperson Scott S. Shannon, Dean, the Graduate School

Acknowledgements

This work would not have been possible without the support of Dr. Horton, my lab mates, the EFB department, or the college community at large. A McIntire-Stennis grant awarded to Dr. Horton funded the bulk of the work, but also instrumental were support from the Edna Bailey Sussman Foundation for an internship at the Albany Pine Bush Preserve, the Lowe-Wilcox Scholarship, as well as travel grants from OIGS and the SUNY ESF Graduate Student Association. I am grateful for the staff of the preserve for their expertise, feedback, and for facilitating my field work; especially Amanda Dillon, Director Neil Gifford, and Dr. Steven Campbell. The University of Minnesota Myco Journal group provided feedback on an early draft of this thesis which greatly improved it, though all faults remain my own. I am also grateful to the many fellow grad students, friends, and roommates who helped me grow as a scientist and provided welcome relief from academic and research work. The community of friends made in Syracuse, especially Daniel Nilsen, helped me grow as a person during my time there. My family and Abby Glauser were a constant support in all areas of my work and life, and I am immensely grateful to them for all they have done.

Table of Contents Acknowledgements ...... ii

List of Tables ...... v

List of Figures ...... vi

List of Appendices ...... vii

Abstract ...... viii

Introductory Literature Review ...... 1

Introduction ...... 1 Mycorrhiza overview ...... 1 Introduction...... 1 Arbuscular mycorrhizae...... 2 Orchid mycorrhizae ...... 3 Ericoid mycorrhizae...... 3 Ectomycorrhizae ...... 4 Functions and use of ectomycorrhizae ...... 4 Introduction...... 4 Nitrogen dynamics ...... 5 Plant-soil feedbacks ...... 6 Use in forestry ...... 7 Use in restoration ...... 8 The Albany Pine Bush Preserve ...... 10 Introduction...... 10 History...... 11 Soils ...... 12 Pinus rigida ...... 12 Robinia pseudoacacia invasion ...... 14 Conclusion ...... 15

Legacy of Robinia pseudoacacia invasion and use of ectomycorrhizal fungi to restore Pinus rigida in the Albany Pine Bush Preserve, NY...... 17

iii

Summary ...... 17 Introduction ...... 18 Materials and Methods ...... 19 Study Site ...... 19 Soil Bioassay ...... 21 Field Experiment ...... 22 Molecular Identification ...... 23 Data Analyses ...... 24 Results ...... 24 Soil Bioassay ...... 24 Field experiment ...... 31 Discussion ...... 34 Conclusion ...... 38

References ...... 40

Appendices ...... 52

Curriculum Vitae ...... 55

List of Tables Table 1. Top five BLAST results for each sample sequence or alignment of ITS sequences from root tips of Pinus rigida soil bioassay seedlings...... 26 Table 2. Field experiment seedling survival mixed effects logistic regression model parameters...... 32

List of Figures Figure 1. Mean proportion of seedlings colonized by ectomycorrhizal fungi identified on roots of Pinus rigida laboratory bioassay seedlings...... 29 Figure 2. Rarefaction of ectomycorrhizal richness on roots of soil bioassay Pinus rigida seedlings...... 30 Figure 3. Mean proportion of Pinus rigida seedlings surviving after 8 months in the field...... 33

List of Appendices Appendix 1. R code used for analyses ...... 52

vii

Abstract

T. R. Patterson. Legacy of Robinia pseudoacacia invasion and use of ectomycorrhizal fungi to restore Pinus rigida in the Albany Pine Bush Preserve, NY. 65 pages, 2 tables, 3 figures, 1 appendix, 2019. New Phytologist style guide used.

Invasive plants can leave lasting legacies on ecosystems, including changes to ectomycorrhizal fungi (EMF). Such legacies can impair restoration even after invaders are removed. A review of information on mycorrhiza fungi and the Albany Pine Bush Preserve (APBP) relevant to restoration of Pinus rigida is presented. The results of research investigating the impact of invasion by Robinia pseudoacacia and the efficacy of using local spore inoculum on P. rigida seedling survival follow. Bioassay seedlings grown in invaded soils had fewer (3) EMF species than non-invaded sites (5). One species was present in both. Invasion history had no effect on field seedling survival after 8 months. However, 72% of seedlings inoculated with live inoculum survived, compared to 31% inoculated with autoclaved spores. These results suggest the legacy of R. pseudoacacia does not limit restoration of P. rigida at the APBP, but that establishment improves when pines are inoculated with locally adapted fungi.

Key Words: ectomycorrhizal fungi, inoculation, Pinus rigida, restoration, Robinia pseudoacacia, invasion, Suillus, Rhizopogon

T. R. Patterson Candidate for the degree of Master of Science, April 2019 Thomas R. Horton, Ph.D. Department of Environmental and Forest Biology State University of New York College of Environmental Science and Forestry Syracuse, New York

viii

Introductory Literature Review

Introduction

This chapter provides a broad review of information on mycorrhizal fungi, their characteristics, functions and uses, and background on the Albany Pine Bush. Information on each topic is generally approached from the perspective of facilitating restoration of pitch pine, Pinus rigida Mill., in the Albany Pine Bush Preserve (APBP) following removal of the invasive black locust, Robinia pseudoacacia L.

Mycorrhiza overview

Introduction

The term “mycorrhiza,” first used by German botanist and mycologist A.B. Frank in 1885, is a combination of the Greek roots “myco” for fungus and “rhiza” for root (Frank, as cited in Trappe 2005). It describes a symbiosis between a plant and a fungus at the interface of fungal hyphae and a plant root tip, where “symbiosis” is defined in the sense that de Bary, drawing on Frank, used the term to refer to interspecific interactions that range from parasitic to mutualistic (Trappe, 2005; Oulhen et al., 2016). The basic activities of mycorrhizal symbioses are the uptake and supply of soil nutrients by the fungus to the plant which typically provides the fungus with photosynthetically-derived carbohydrates (Smith & Read, 2008; Courty et al., 2016). Mycorrhizal fungi may provide plants with up to 80% of their required nitrogen and up to 90% of their required phosphorus (van der Heijden et al., 2008). Mycorrhizal symbioses are usually thought of as mutualisms, but they are dynamic relationships that vary between and within partner plant and fungal species along a continuum from parasitism to mutualism (Johnson et al., 1997). A mutualistic interaction is one in which both organisms benefit from the interaction, while a parasitic interaction is one in which one partner benefits at the expense of the other. When the relationship is not mutualistic it is more common for plants to exploit or be parasitic towards the fungi than the reverse (Brundrett, 2002).

Mycorrhizal associations are common in nearly all environments and the majority of plant species engage in mycorrhizal symbioses (Smith & Read, 2008; Brundrett, 2009; van der Heijden et al., 2015; Tedersoo, 2017). There are four primary types of mycorrhizae: arbuscular mycorrhizae (AM), ectomycorrhizae (EM), orchid mycorrhizae, and ericoid mycorrhizae. These four mycorrhizal types are morphological categories primarily based on structural features of the mycorrhiza determined by the host plant, but these groupings map well in most cases on to both plant and fungal lineages (Brundrett, 2002, 2004). An estimated 67% of all vascular plant species associate with arbuscular mycorrhizal fungi (AMF), 2% with ectomycorrhizal fungi (EMF), 8% form orchid mycorrhizae, and 1% form ericoid mycorrhizae (Brundrett, 2009). Arbuscular mycorrhizae (AM) and ectomycorrhizae (EM) are the most abundant mycorrhizal relationships found worldwide (Tedersoo, 2017). There are two additional mycorrhizal types, arbutoid and monotropoid, which form between fungi that also form EM and some plants in the Arbutoideae or the Monotropoideae. Arbuscular, orchid, and ericoid mycorrhizae are briefly described here while EM, the focus of this research, are explored in greater depth.

Arbuscular mycorrhizae

Arbuscular mycorrhizae have the oldest fossil evidence, found in the Rhynie chert (c. 400 million years old) of Scotland and the Guttenberg Formation (c. 460 million years old) in Wisconsin (Redecker et al., 2000; Taylor et al., 2005; Strullu-Derrien et al., 2016). Arbuscular mycorrhizae are typically characterized by the formation of arbuscules, highly branched fungal structures within plant root cortical cells, and may form vesicles (Brundrett, 2004; Smith & Read, 2008). More than 200,000 plant species associate with AM fungi, including most grasses, herbs, some bryophytes, and many trees (Brundrett, 2009). AM plants are the dominant vegetation type in nearly all habitats (Brundrett, 2017). While plants forming AM are extremely diverse, AMF are represented by relatively few species. There are an estimated 350 – 1000 molecularly identified AMF taxa, exclusively within Glomeromycota (Smith & Read, 2008; Kivlin et al., 2011; Öpik et al., 2013). With such a broad array of plant symbionts and so few fungal symbionts, AMF are usually thought to exhibit low specificity in their associations (Brundrett, 2002). AMF are obligate symbionts, unable to complete their life cycle without a compatible plant partner (Smith & Read, 2008).

Orchid mycorrhizae

Orchid mycorrhizae, as the name suggests, are formed by plants in Orchidaceae. The fungi involved are primarily basidiomycetes in the families Tulasnellaceae, Ceratobasidiaceae, and Sebacinaceae, and some ascomycetes (Smith & Read, 2008; Jacquemyn et al., 2017). These associations are characterized by coils of hyphae called pelotons that penetrate the cell walls of the orchid roots cells (Brundrett, 2004; Smith & Read, 2008). The tiny dust seeds of orchids are dependent on early colonization of compatible fungi because they lack an endosperm (Dearnaley et al., 2012). Orchid seedlings, or protocorms, lack chlorophyll to photosynthesize and as such require carbon as well as nutrients from their mycorrhizal fungi – a reversal of the typical carbon flow in mycorrhizal relationships (Rasmussen & Rasmussen, 2009; Selosse et al., 2016; Jacquemyn et al., 2017). Orchids exhibit a range of nutrition strategies - from parasitism of their associated fungi (mycoheterotrophy) to more or less traditional autotrophic mutualism (Rasmussen & Rasmussen, 2009; Dearnaley et al., 2016). In mycoheterotrophy, the plant receives all or most of its required carbon and mineral nutrients from the fungi. The specificity of orchids and their mycorrhizal fungi tends to depend on the nature of the nutritional strategy of the orchid. Achlorophyllous mycoheterotrophs tend associate with narrow phylogenetic groups of otherwise ectomycorrhizal fungi (more specific) while green orchids may associate with a wide variety of saprotrophic fungi (less specific), though there are exceptions (Taylor et al., 2002; Brundrett, 2004; Mccormick & Jacquemyn, 2014; Jacquemyn et al., 2015).

Ericoid mycorrhizae

Ericoid mycorrhizae are formed by approximately 3,900 plants in the family Ericaceae and more than 150 fungal species (Smith & Read, 2008; Walker et al., 2011). The fungal symbionts are mainly ascomycetes of the order Helotiales, but also some basidiomycetes (Selosse et al., 2007; Smith & Read, 2008; Walker et al., 2011). Some fungi may form ericoid mycorrhizae with Ericaceae and ectomycorrhizae with other plants, similar to those which are parasitized by mycoheterotrophic plants (Villarreal-Ruiz et al., 2004). Plants in Ericaceae dominate heathland ecosystems and include the genera Erica, Rhododendron, and Vaccinium (Kohout, 2017).

Ericoid mycorrhizae are characterized by extensive colonization and intracellular hyphal coils in the epidermal cells of the thin roots of the plants in the Ericaceae (Brundrett, 2004; Smith & Read, 2008). They are thought to be moderately specific relationships, but more research is needed on this group, especially considering the widespread distribution of the plant symbionts in heathlands and elsewhere around the globe and their influence on the ecology of those systems (Read & Perez-Moreno, 2003; Read et al., 2004).

Ectomycorrhizae

Ectomycorrhizae form between basidiomycete as well as ascomycete fungi and a variety of angiosperm and conifer shrubs and trees (Smith & Read, 2008). Plant partners include commercially important trees and those that dominate in temperate and boreal forests (e.g. those in the Pinaceae and Fagaceae) (Wang & Qiu, 2006). The structure of an ectomycorrhiza is characterized by the formation of a hyphal mantle, which is a sheath that surrounds the plant root, and the Hartig net, a lattice of intercellular labyrinthine hyphae that form between cortical cells of the root but do not penetrate the plant cell walls (Brundrett, 2004; Smith & Read, 2008). Ectomycorrhizae have evolved independently as a strategy in the basidiomycetes multiple times from saprotrophic fungi between 100 – 200 million years ago (Hibbett et al., 2000; Brundrett, 2002; Tedersoo & Smith, 2013; Kohler et al., 2015). Some of the earliest fossil evidence indicates that by c. 50 million years ago a Rhizopogon or Suillus-like fungus was associating with pines (Lepage et al., 1997). There are an estimated 20,000 fungal species that form ectomycorrhizae (Rinaldi et al., 2008; Tedersoo et al., 2010). They associate with approximately 6000 plant species (Brundrett, 2009). This diversity relationship is the inverse of that of AM, and ectomycorrhizal plant and fungal species exhibit a range of partner specificity (Molina et al., 1992; Molina & Horton, 2015).

Functions and use of ectomycorrhizae

Introduction

Ectomycorrhizal symbioses are an important component of many ecosystems. In temperate ecosystems EM trees dominate. EM relationships may contribute to up to 80% of plant productivity, 80% of plant nitrogen acquisition, and ectomycorrhizal fungi (EMF) are often required for successful seedling establishment and survival (van der Heijden et al., 2015). Benefits of EMF include improved uptake of water and nutrients, protection of roots against drought, high temperatures, pathogens, heavy metals, high alkalinity, and salts. Here, the functions of EMF in nitrogen uptake and seedling establishment in different contexts are discussed, followed by a summary of their known influences and use in forestry and restoration.

Nitrogen dynamics

EMF differ widely in their ability to take up, use, and transfer different forms of nitrogen to host plants (Lilleskov et al., 2002b,a, 2011; Hobbie et al., 2008). Some EMF use organic nitrogen more effectively, usually in the form of amino acids, to the benefit of their partner plants (Abuzinadah & Read, 1986; Abuzinadah et al., 1986). Among inorganic forms of N, ammonium is the most readily used source; most EMF grow more in the presence ammonium than nitrate and discriminate against nitrate as an N source in favor of other sources (Finlay et al., 1992; Clemmensen et al., 2008). The differences between different EMF species in N uptake and transfer can be greater than differences in N uptake between EM and non-mycorrhizal plants (Jones et al., 2009). EMF are likely adapted to whatever N form is present in their undisturbed environments (Lilleskov et al., 2002b; Koide et al., 2014). As such, nitrogen availability (both amount and form) strongly influences EM fungal community composition and function. EM fungus diversity and distribution in any given system may reflect the diversity and heterogeneity of N forms available (Koide et al., 2011; Lilleskov et al., 2011; LeDuc et al., 2013). At elevated levels of N there is a negative effect on EM fungal communities. The threshold is estimated at 5-10 kg N ha-1 (Jarvis et al., 2013). Along a gradient of anthropogenic N deposition in Alaska, EMF richness decreased with increasing N (Lilleskov et al., 2002a). Increases in organic N also decreases richness and shifts communities, and may contribute to successional changes in EMF in forests (LeDuc et al., 2013; Cline et al., 2018). In natural or

5 normal (relatively low) ranges of nitrogen in soil, EM colonization is positively associated with greater nitrogen (Soudzilovskaia et al., 2015). One possible mechanism by which N form and availability indirectly affect EMF communities is through changes to the dynamics of their symbiotic interaction. Plants can take up nitrogen in organic and inorganic forms, but in the N-limited temperate and boreal forests EMF provide greater access to organic forms and allow these plants to succeed in otherwise N limited sites. When inorganic N is higher than 5-10 kg ha-1, EMF may depress seedling growth (Plassard et al., 2000). Plant C allocation to EMF is generally greater in N limited conditions than when N availability is high (Hobbie & Colpaert, 2003; Hobbie, 2006). Under higher levels of plant accessible inorganic N plants may suppress colonization or C transfer to fungi that are more effective at provisioning plants with nitrogen from organic sources (Taylor et al., 2000). However, the degree to which N supply by EMF and plant C allocation to them are coupled is unclear (Valtanen et al., 2014; Hortal et al., 2017; Bogar et al., 2019).

Plant-soil feedbacks

Plant communities often exhibit negative density dependence. Host-specific antagonists reduce seedling survival when close to mature plants of the same species (Johnson et al., 2012). However, host-specific mutualists (e.g. mycorrhizal fungi) follow the same pattern of abundance as antagonists with distance from host plants (Laliberté et al., 2015). These competing influences in the soil can be difficult to interpret in natural systems. EM host trees tend to facilitate conspecific recruitment through plant-soil feedbacks mediated by their mycorrhizal relationship (Horton & Van Der Heijden, 2008; Bennett et al., 2017). In addition to providing nutrients to seedlings, improving survival and recruitment, the EM hyphal sheath may also offer greater pathogen protection to young roots (Laliberté et al., 2015). In an inoculation and transplant study, Bennet et al. (2017) found that greater survival near conspecific trees appears attributable to protection from soil antagonists, which, under conspecific trees, would be directly or indirectly provided by their mycorrhiza. Under heterospecific trees anatogonist pressure was lower and prior inoculation with EMF did not benefit seedlings.

Use in forestry

Mikola (1970) reviewed the history and existing literature on mycorrhizal inoculation in forestry. He suggested that natural forests do not require management of, as he called them, “ectotrophic” mycorrhizae, the fungi being almost ubiquitous and locally adapted. In man-made forests, however, it was known as early as the 1930s that forest plantations may fail due to the absence of the appropriate fungal symbionts. Mikola noted records of failure outside the natural range of the tree and inside the natural range where the soil conditions had been drastically altered by anthropogenic activity. Soil was used successfully as inoculum as early as 1910 to establish pine plantations in Kenya where they had previously failed. Subsequent plantations on the African continent were started with the soil of successful plantations as inoculum. As pine plantations spread across the Southern Hemisphere early in the 20th century the same story played out; failure of nurseries until the addition of mycorrhizal inoculum in the form of top soil from established plantations or colonized nursery seedlings called “mother trees”. These repeated natural trials and nascent experimental research (e.g. Kessell, 1927) demonstrated the necessity of mycorrhizal inoculation for pines, at least outside their native range. Once fungi were identified as the primary driver of the advantages of inoculation by soil or mother trees, mycelial cultures, spore solutions and fruiting bodies were sometimes used as inoculum in the Philippines, Australia, the then Soviet Union, and elsewhere (Mikola, 1970). There are also examples of the practical application of EMF inoculum to forest regeneration within the natural range of pines from the early 20th century, particularly in previously agricultural areas. In Austria, Suillus species were missing from nurseries in former agricultural soils, so they were reintroduced as pure cultures of mycelium, with successful results (Moser, 1959 as reported in Mikola, 1970). Even where EMF inoculum was present, benefits were observed by inoculating with pine forest soils, likely due to the identity and specificity of fungal partners available. Research on and use of inoculation techniques in forestry has continued and flourished since the earlier review by Mikola (e.g. Amaranthus & Perry, 1987; Brundrett et al., 1996). One legacy of transplantation of fungi in the service of forestry is the now world-wide distribution of some EMF, including Pisolithus, Suillus, and Rhizopogon (Dunstan et al., 1998;

Giachini et al., 2000; Tedersoo, 2017). Suilloid fungi, those in the genera Suillus and Rhizopogon, have enabled the expansion and invasion of pines world-wide (Policelli et al., 2019). In Chile, for example, Pinus contorta Dougl. ex Loud. are dispersing and establishing away from a plantation established in the 1980s, co-invading with non-native EMF (Hayward et al., 2015b). The total richness of EMF species on trees established outside of the plantation was four species. Sampled trees greater than 250 m from the edge of the plantation and most isolated all harbored a single Suillus species, which Hayward et al. found to be sufficient to enable the invasion. This is compared to hundreds of EMF associating with Pinaceae species in their native range (Taylor et al., 2014). A similar situation is present in Argentina, where Pinaceae species are also invading with a subset of EMF present in plantations, notably suilloid species (Hayward et al., 2015a).

Use in restoration

The processes of forest regeneration as studied and exploited for commercial reasons can similarly be applied to restoration, defined as “assisting the recovery of an ecosystem that has been degraded, damaged, or destroyed.” (Clewell et al., 2004). The range of functions, activities, and influences of mycorrhizal symbioses increase their importance in restoration beyond simply their utility when restoration is defined as “actions to re-instate ecological processes, which accelerate recovery of forest structure, ecological functioning and biodiversity levels towards those typical of climax forest” (Elliott et al., 2013). Inoculation with EM is certainly beneficial when seedlings are planted in soils without suitable inoculum, as indicated by the above discussion in forestry. In restoration contexts, the benefits of inoculation are expected to be proportional to the degree to which degradation affects the capacity for natural colonization. Clearcut logging changes the EM fungal community through changes in biology and chemistry of the soil environment, but if tree roots are left intact and the forest floor is left relatively undisturbed the inoculum potential as measured by root colonization may not be affected (reviewed in Jones et al., 2003). In clearcuts, EM fungal propagules exist in the soil as spores, sclerotia, and root tips on dying stumps. Intact forest at the edge of clearcuts and retention patches can serve as a source of inoculum, though the influence of mature trees may only extend up to 10 m into the clearcut site (Jones et al., 2008; Baker et al.,

2013). With increasing distance from mature trees, Teste et al. (2009), found that seedlings had reduced EM fungal richness and diversity. In that study exclusion of mycorrhizal networks with mesh bags had no effect on mycorrhizal colonization, leading the authors to conclude that inoculum came from wind-borne and soil inoculum rather than intact mycorrhizal networks. Other studies have found that networks were important colonizers of seedlings, especially in primary succession (Nara, 2006, 2015). Regardless of the source of initial colonization, EM seedlings tend to benefit when connected to or in the presence of intact networks of older trees (reviewed in Simard et al., 2012; Horton, 2015). In systems that resemble clearcuts, nursery inoculation of seedlings may not be required for successful establishment if natural colonization is possible. Severe disturbances to soils and changes in vegetation type, especially to a different mycorrhizal type, have a negative effect on EM fungal communities to the detriment of seedling regeneration (Jones et al., 2003; Simard, 2009). Seeding with AM grasses, for example, can reduce the colonization of EMF on EM seedlings planted into those areas, relative to areas that host other EM plants (Amaranthus & Perry, 1994). Some invasive plants depend on mycorrhizal symbionts and experience positive feedback with co-invading fungi that can contribute to ‘invasional meltdown’ of invaded systems (Simberloff et al., 1999; Pringle et al., 2009; Dickie et al., 2017). Effects of non-EM or non-mycorrhizal plants can disrupt the mutualisms in a variety of ways. Grove et al. (2017) outline three mechanisms by which invasive plants disrupt native mycorrhizal dynamics: changes in host quantity or quality through plant-plant competition, changes to soil properties, and allelopathy. The loss or degradation of host species will, over time, change the fungal community and decrease the inoculum potential, making restoration without inoculation difficult. Similarly, as discussed above changes in the soil chemistry by changes to the dominant vegetation will affect mycorrhizal fungus communities. Changes to the vegetation will result in different plant-soil feedbacks, such as invasive plants promoting soil communities and chemistry that benefit them or other plants that may not be desirable in restoration projects. Invasion by Robinia pseudoacacia L. caused a change in vegetation composition that promoted other non-native plants in a pine-oak system in Cape Cod (Von Holle et al., 2013). Any of these more dramatic or long-lasting changes to vegetation might require inoculation of seedlings planted as part of restoration.

In the long term, a more diverse community of EMF in a restoration site is preferable. The positive effect of inoculation may be more likely with greater diversity of inoculum (Sousa et al., 2014). A more diverse inoculum might provide access to different nutrient pools or simply increase the likelihood of successful colonization. Colonization increased with an increase in the number of fungal species on Pinus rigida Mill. (Baxter & Dighton, 2005). The response to increasing EM fungal diversity in that study appeared to be dependent on the source and availability of nutrients. As a result, though there may be benefits of greater diversity in inoculum and high fungal diversity is an end of its own in restoration, the initial establishment may not require it. EMF frequently present in nurseries, Thelephora terrestris Ehrh. and Wilcoxina mikolae Yang & Korf can provide a benefit to outplanted seedlings (e.g. Jones et al., 1990, 2009). It bears consideration, however, because the nursery ‘contaminants’ can persist on seedling roots to the exclusion of indigenous fungi which may be more beneficial in the long term (Jones et al., 2003). In restoration a more particular set of EMF may be preferred. Inoculum from ecosystems similar to the target of restoration and with specific fungal species can improve colonization and outcomes of restoration with a slightly higher initial investment of time and resources (Maltz & Treseder, 2015). The choice of inoculum might include considerations of those fungi more likely absent in the site, those better adapted to the site or focal plant, or fungi able to perform another function such as providing heavy metal resistance (Hawkins et al., 2015).

The Albany Pine Bush Preserve

Introduction

The Albany Pine Bush Preserve (APBP) is a unique ecosystem on ancient sand dunes left after glacial Lake Albany drained. It is located between the cities of Albany and Schenectady, New York. The preserve is home to more than 70 protected animal and plant species, including the Karner Blue Butterfly. In 1988, recognizing the ecological importance and interest in the APB, the state passed Environmental Conservation Law article 46 to protect the site, manage it, and establish the Albany Pine Bush Preserve overseen by the APBP commission. The area was

10 proposed as a National Natural Landmark in 2013 and given that designation by the National Park Service in 2014 (Brickle et al., 2013). Approximately 3,300 acres of public and private lands are protected as part of the Preserve, which serves as an important natural and outdoor recreation area for the public. The natural history, ecology, present restoration challenges and goals are discussed below. An excellent book, Natural History of the Albany Pine Bush (Barnes, 2003) exists for a more thorough review and is the reference for information here unless otherwise noted.

History

The retreat of the Laurentide Ice sheet that covered the region began 18,000 years ago. As the glacier retreated the Capital Region of New York was then covered by glacial lake Albany. The melting glacier deposited sediments into the lake from 15,000 to 12,600 years ago. When the lake began to drain these sediments were exposed and wind formed dunes covered and moved through the area until between 9,000 and 5,000 years ago when vegetation began to stabilize the dunes. The xeric, acidic, well-drained sands came to host pine forests similar to the present-day site. During post-glacial warming, slightly warmer and drier than today, species from the great plains extended eastward and more southern Atlantic Coastal plains species extended northward and are still present in the area as relics of that period (Gill, 1997). The pine bush has long been influenced and managed by humans. Native American groups deliberately burned the pine bush to improve hunting and foraging in the area. Later, European settlers used fire to clear the land and improve the soil nutrients for agriculture. The pine barrens were maintained in part by these human activities, periodic accidental fires, as well as by logging. As settlement increased in the areas surrounding the pine bush fire suppression measures were put in place to protect property and the successional hardwood plant communities began to expand in area. The expansion and dominance of black locust, Robinia pseudoacacia L., during the last century has been especially problematic, and is discussed more below. Today, the Albany Pine Bush Preserve is maintained by prescribed burns as well as mechanical removal of invasive black locust, herbicide use, and mowing (APBPC, 2010).

Soils

The soils of the APBP are considered young because of the glacial history of the region. The surface material of the APBP consists of fine sandy loams deposited where glacial meltwater from Mohawk Valley emptied into Glacial Lake Albany between 15,000 to 12,600. As mentioned above, when the lake drained 12,000 to 13,000 years ago these sediments were exposed to wind and water erosion. The wind formed dunes in the sands and carried away the finest particles until they were stabilized by vegetation between 9,000 and 5,000 years ago. Three acidic, sandy soil series characterize most of the area of the APBP (Brown, 1992). The largest area is dominated by Colonie soils. Colonie soils are very deep, well to excessively well-drained soils formed in glaciolacustrine, outwash, and eolian deposits. They are composed of fine and very fine sand; the typical texture is classified as loamy fine sand. Colonie soils are usually found in the uplands and rises of the area. The second most common soil series found in the APBP area are Elnora soils. Elnora soils are very deep, moderately well drained soils similarly formed in sandy deltaic, eolian, and glacial lake sediments. Elnora soils are also loamy fine sands. They occupy the former beach ridges and sand bars of the Glacial Lake Albany plains. Stafford series soils are the third most common soils of the APBP. Stafford loamy fine sands are found on deltas and sand plains. They are characterized by very deep, somewhat poorly drained soils formed in sandy glaciolacustrine deposits. The soils of the Pine Bush have been disturbed by humans in a variety of ways. Large areas were plowed for agriculture, clearing and bulldozing for development, and the sands themselves have been harvested and used for a variety of industries including glassworks and concrete production.

Pinus rigida

The APBP is characterized by native pitch pine – scrub oak barren plant communities. This includes stands of pitch pine, Pinus rigida Mill., and scrub oaks, such as Quercus ilicifolia Wangenh., in varying densities in a matrix of grassland savannahs and thickets (Gebauer et al., 1996; Milne, 1985). The plant communities are maintained by and prone to frequent fires. Many plants and other organisms in pine barrens are fire-tolerant and fire dependent. P. rigida has

12 several adaptations to environments with frequent fires. It has thick bark, can resprout from basal or epicormal buds, and sometimes produce serotinous cones (Little, 1953). P. rigida seeds require bare mineral soil to germinate and seedlings are shade intolerant (Jordan et al., 2003). Lee et al. (2018) found that, in the APBP, P. rigida seedlings germinated and survived poorly in the shade and litter of fast-growing Q. ilicifolia but fared better in the low cover of Vaccinium. Fire or other disturbance is required to remove litter, understory plants, and to open canopies for successful regeneration of P. rigida. In the absence of fire, pitch pine – scrub oak barren communities typically transition to northern hardwood forests (Jordan et al., 2003). In the APBP, P. rigida may not have been regenerating naturally at a high enough rate to maintain populations or achieve restoration goals (Gill, 1997). However, prescribed burning began in the 1990s after establishment of the Preserve, so in the time since Gill (1997) natural recruitment may have improved due to the return of fire and other management practices. Personal observations of natural recruitment and reports from preserve managers are conflicting and varied for different sites within the preserve. In a similar system on Long Island, Landis et al. (2005) found that initial recruitment post-fire was limited by seed availability. This may also be limiting recruitment in the APBP. If natural regeneration is limited then transplantation of P. rigida may still be necessary in some areas, especially those areas most fragmented or disturbed such that there are few mature pines left. Gill (1997) transplanted seedlings into different areas of the Preserve to study edge effects. They found that 55% of P. rigida seedlings transplanted into undisturbed interior sites died within the first 18 months. Herbivory by mammalian herbivores was a primary cause of seedling mortality. It was unlikely that regeneration of P. rigida was limited by insufficient mycorrhizal inoculum in most areas of the preserve. In that study they found that most seedlings were colonized within 6 months of planting. There was a slight increase in mycorrhizal colonization rates and diversity of field seedlings from five to 18 months of growth. There was an increase in the mean number of fungal morphotypes (morphologically distinguished fungal root tips) per seedling from 1.69 ± 0.80 (± 1 SD) after five months to 2.39 ± 0.85 after 18 months. Colonization increased from 75% ± 24 to 80% ± 19. Gill did not find edge-related patterns of mycorrhizal colonization, but they did find lower colonization and fewer fungal morphotypes on P. rigida seedlings growing within R. pseudoacacia stands. Seedlings growing in R. pseudoacacia stands had high mortality (85%

13 mortality after one season) but those that survived were taller, which the author suggested may have been due to a nitrogen fertilization effect from R. pseudoacacia or simply greater allocation to upward growth to overcome light limitation. These results suggest that restoration of P. rigida in areas where R. pseudoacacia are present may be difficult. It is unclear what caused higher mortality among seedlings in R. pseudoacacia stands, but the reduced mycorrhizal colonization and reduced diversity of those stands may contribute. Additionally, it is unclear if the effects on survival, mycorrhizal colonization, and diversity persist after R. pseudoacacia has been removed. There may be a legacy effect which could impair P. rigida establishment and restoration in those areas.

Robinia pseudoacacia invasion

Robinia pseudoacacia is an N-fixing legume native to the central Appalachian and Ozark Mountains. It is an early successional, shade-intolerant species that reproduces and expands rapidly through vegetative stump and root sprouts as well as seeds (Boring & Swank, 1984). It began to invaded areas of the APBP in the 1960s, enabled by fire suppression, human disturbance and fragmentation, and deliberate planting for use as timber. It quickly expanded, sometimes forming large, mono-dominant stands which can force out and exclude native vegetation (Rice et al., 2004; Von Holle et al., 2006). Nitrogen fixing plants such as R. pseudoacacia can leave lasting legacies, especially when they invade relatively nutrient poor systems (Ehrenfeld, 2003; Corbin & D’Antonio, 2004). The changes to soil properties caused by Robinia invasion can be very strong. R. pseudoacacia stands can have 1 – 3 times greater soil nitrogen concentrations compared to non-invaded pitch pine – scrub oak sites (Rice et al., 2004). R. pseudoacacia can increase soil nitrogen by up to 75 kg N ha -1 yr -1 (Boring & Swank, 1984). In a study of pitch pine – oak forests in Cape Cod, Von Holle et al. (2013) found that soil N remained elevated for at least 14 years in soils of former R. pseudoacacia stands. Net nitrification and extractable soil nitrate concentration was highest in existing R. pseudoacacia stands, followed by former stands, and lowest in pine – oak stands. Additionally, non-native species cover increased in R. pseudoacacia stands and persisted in former stands. In the APBP, soil N and total net mineralization rates decrease to normal levels within two years of removal of R. pseudoacacia and net nitrification can remain elevated for up

14 to four years after removal (Malcolm et al., 2008). R. pseudoacacia stands generally have a low C:N ratio relative to non-invaded pitch pine – scrub oak barren sites. The ratio recovers to levels indistinguishable from pitch pine – scrub oak barren sites, but the carbon and nitrogen pools do not, in the short term (Jeff Corbin, pers. comm.).The differences between the duration of the impacts in Von Holle et al. (2013) and Malcolm et al., 2008) may be due to the differences in removal of R. pseudoacacia; in the former the trees were blown down by hurricane and roots remained in the soil, whereas in the latter roots and stumps were mechanically removed by APBP managers. R. pseudoacacia are AM hosts (Wang & Qiu, 2006; Brundrett & Tedersoo, 2018). In the framework of Grove et al. (2017) discussed above, of the three mechanisms by which invasive plants may disrupt native mycorrhizal dynamics (changes in host quantity or quality through plant-plant competition, changes to soil properties, and allelopathy) in the APBP, R. pseudoacacia has likely affected the native mycorrhizal communities by at least the first two mechanisms, and allelopathy of R. pseudoacacia has been demonstrated laboratory settings (Nasir et al., 2005). Changes similar to the increase in non-native plant cover in R. pseudoacacia stands in Cape Cod have likely occurred in the APBP, constituting likely changes in host quantity if not quality (Von Holle et al., 2013). The dramatically higher nitrogen levels in R. pseudoacacia stands relative to uninvaded areas constitute a change in soil properties that exceed the estimated threshold of 5-10 kg N ha-1 beyond which EM fungal communities are affected (Jarvis et al., 2013). EM fungal communities in R. pseudoacacia stands are, therefore, expected to be different from those of non-invaded pitch pine – scrub oak barren areas in the Pine Bush. Indeed, as mentioned above, Gill (1997) found lower mycorrhizal colonization and reduced fungal diversity on the roots of P. rigida seedlings in R. pseudoacacia stands. Restoration of these invaded areas is one of the goals of the Preserve Commission and restoring the EM fungal community should be considered, both as a means and an end to restoring the functioning of the pitch pine – scrub oak barren communities.

Conclusion

The variety of roles played by EMF should be of use in facilitating restoration of P. rigida in the APBP. The existing literature on the benefits of mycorrhizal inoculation to survival of pine

15 seedlings and the impacts of R. pseudoacacia suggest promising outcomes of using EMF the context of using this dynamic in a restoration context. Providing seedlings with the appropriate inoculum can improve seedling establishment. This has been long practiced in forestry and is beginning to be considered in restoration. It may only require a few specific fungi to improve results and local fungi from the APBP would likely be ideal. They are expected to be adapted to the site and less potentially disruptive than an outside inoculum source might be. It is hoped that this technique and information will aid the managers of the Albany Pine Bush Preserve in their efforts to restore protected areas to the native plant community in service of their goals, especially in sites impacted by invasion of R. pseudoacacia.

Legacy of Robinia pseudoacacia invasion and use of ectomycorrhizal fungi to restore Pinus rigida in the Albany Pine Bush Preserve, NY.

Summary • Invasive plants can leave lasting legacies, including changes to the ectomycorrhizal fungal (EMF) communities, that make restoration difficult even after invaders have been removed. Previous research has identified suilloid fungi, those in the genera Suillus and Rhizopogon, as important in early succession and sufficient to enable pine establishment in new areas, making them potentially useful in restoration. • To examine the legacy of invasion by N-fixing black locust (Robinia pseudoacacia) on resistant EMF propagule communities, we carried out a soil bioassay with field soil from sites where R. pseudoacacia had recently been removed and non-invaded sites with typical pitch pine (Pinus rigida) – scrub oak barrens plant communities. To test the hypothesis that suilloid fungi can improve survival and that R. pseudoacacia has a legacy effect on P. rigida establishment and survival, P. rigida seedlings were planted in a factorial field experiment (invasion history × inoculation treatment). • Our hypotheses were partially supported. Resistant EMF propagule communities in R. pseudoacacia invaded soils had three total EMF, while there were five in non-invaded sites, and one species was present in both. A single suilloid species, Rhizopogon pseudoroseolus was found only on seedlings grown in non-invaded soils. However, there was no difference in survival of field seedlings in invaded or non-invaded sites after 8 months. Inoculation with suilloid fungi improved seedling survival; 72% of seedlings inoculated with live spore solutions survived, compared to 31% of seedlings that received autoclaved control inoculum. • Legacies of invasion by R. pseudoacacia may not be limiting restoration of P. rigida in our study site at the Albany Pine Bush Preserve in New York. However, we suggest restoration should make use of locally adapted fungi as inoculum for the benefits they provide to seedling survival without introducing exotic species from commercial inoculum and to restore the disturbed belowground fungal communities.

Introduction

Plant invasions can be enabled by mycorrhizal fungi, disrupt native mycorrhizal interactions, and leave long-lasting soil chemical, physical, and biotic legacies even after invasive plants are removed (Pringle et al., 2009; Corbin & D’Antonio, 2012; Dickie et al., 2017). An important body of work has emerged describing the dynamics of mycorrhizae in both facilitating invasions and being affected by them, but less is known about the impacts of these invasion legacies on restoration or techniques for mitigating or overcoming invasion legacies. Restoration following invasion by nitrogen fixing plants may be especially difficult. N- fixers are frequent invaders and can dramatically affect ecosystem processes by increasing soil N and facilitating invasion by other non-native plants (Rice et al., 2004; Von Holle et al., 2006). These effects may persist for years even after removal of the N-fixing plants themselves (Von Holle et al., 2013; Grove et al., 2017b). Changes in soil nutrients and plant community brought about by invasions can impact the native ectomycorrhizal fungal (EMF) communities (reviewed in Grove et al., 2017a). Changes to the native EMF community are expected to be to the detriment of restoration when EMF compatible with native plants are absent or if inoculum potential is reduced. It may be especially important in restoration sites that resemble early successional environments due to either removal of the invaders or factors relating to the invasion itself. Such habitats are susceptible to re-invasion by the same or other non-native species, motivating rapid establishment of the desired vegetation. Inoculum from ecosystems similar to the target of restoration and with specific fungal species can improve colonization and outcomes of restoration with a slightly higher initial investment of time and resources (Maltz & Treseder, 2015). The choice of inoculum might include considerations of those fungi more likely absent in the site, those better adapted to the site or focal plant, or fungi able to perform another function such as providing heavy metal resistance (Hawkins et al., 2015). To the degree that restoration resembles early succession, early successional EMF may be particularly helpful in achieving this goal. EM conifers establishing after disturbance are frequently colonized by suilloid fungi (Rhizopogon and Suillus) (Horton et al., 1998; Baar et al., 1999). These same fungi enable pines to establish in previously uninhabited areas both within

18 and outside their native range (Ashkannejhad & Horton, 2006; Hayward et al., 2015a). These fungi produce resistant spore banks that can persist for years (Bruns et al., 2009; Nguyen et al., 2012). However, depending on the duration of an invasion they may not remain viable in soils long enough to support and promote restoration efforts. The objectives of this study are two-fold: first, to investigate the effects of invasion by an N-fixing tree, Robinia pseudoacacia L. (black locust), on resistant propagules of EMF in soil, and second, to assess the benefits of inoculation with previously identified critical partners, Suillus and Rhizopogon as a tool for restoration of Pinus rigida Mill. (pitch pine). To address these objectives, we first conducted a laboratory bioassay with dried field soil to survey potential resistant propagule inoculum in areas of differing invasion history; non-invaded pitch pine – scrub oak barrens and areas of recent R. pseudoacacia removal. We expected there to be a greater richness of resistant EMF present on seedlings grown in non-invaded soil than in invaded soil. Second, we experimentally planted P. rigida seedlings inoculated with live or autoclaved suilloid spores into those same areas and recorded their subsequent survival. We expected seedlings given live inoculum to survive in greater numbers than those given autoclaved control inoculum, and greater number of seedlings given control inoculum in non-invaded sites to survive than invaded sites due to an expected positive impact of greater EMF richness and existing hyphal networks supported by mature P. rigida in those sites.

Materials and Methods

Study Site

The Albany Pine Bush Preserve (APBP) is located at 42°43'07"N 73°51'53"W between the cities of Albany and Schenectady in Albany County, NY. The Albany pine bush is a unique ecosystem on ancient sand dunes left after glacial Lake Albany drained. It is home to more than 70 protected animal and plant species, including the Karner blue butterfly, Plebejus melissa samuelis (W.H. Edwards). In 1988, recognizing the ecological importance and interest in the pine bush, the state passed Environmental Conservation Law article 46 to protect the site, manage it, and establish the Albany Pine Bush Preserve overseen by the APBP Commission. The area was proposed as a National Natural Landmark in 2013 and given that designation by the

National Park Service in 2014 (Brickle et al., 2013). Approximately 3,300 acres of public and private lands are protected as part of the Preserve, which serves as an important natural and outdoor recreation area for the public. The area is characterized by native pitch pine – scrub oak barren plant communities. In the APBP, stands of P. rigida, and scrub oaks, such as Quercus ilicifolia Wangenh., occur in varying densities in a matrix with grassland savannahs (Gebauer et al., 1996; Milne, 1985). In the absence of fire, some areas of the APBP have been invaded by R. pseudoacacia, an N-fixing legume native to the south of New York state in the central Appalachian and Ozark Mountains. It is an early successional, shade-intolerant species that reproduces and expands rapidly through vegetative stump and root sprouts as well as seeds (Boring & Swank, 1984). It began to invade areas of the APBP in the late 1960s, enabled by fire suppression, fragmentation, and deliberate planting for use as timber (Barnes, 2003). It quickly expanded, sometimes forming large, monodominant stands. Restoration of areas invaded by R. pseudoacacia is one of the main challenges of the APBP. Some stands have been cut and roots mechanically removed by heavy equipment. In some sites, trees resprouting from remaining underground material are chemically treated with herbicide. Periodic burns have also been reintroduced throughout the preserve as a substitute for natural fire regimes. Mean annual temperature for the area is 9.1 °C and mean annual precipitation is 999.5 mm (NOAA, 2018). During the study period, August 2017 through April 2018, the site experienced a 2.1 °C warmer than normal autumn average temperature (normal 10.3 °C) with 97.5 mm less total precipitation; a 0.7 °C warmer than normal winter average temperature (normal -3.6 °C) with 8.1 mm less total precipitation; and a 5.2 °C cooler than normal spring average temperature (normal 8.3 °C) with 70.1 mm less total precipitation, though more than normal precipitation in the form of snow (635 mm more than normal total snow) (NOAA, 2018). The period from October 2017 to May 2018 was the 19th snowiest on record for Albany, receiving a total of 1963.4 mm snow (NOAA, 2018). Strongly acidic, largely sterile soils characterize most of the area of the APBP (Brown, 1992). The most common soil series, Colonie soils, are very deep, well to excessively well drained soils formed in glaciolacustrine, outwash, and eolian deposits. They are composed of fine and very fine sand; the typical texture is classified as loamy fine sand.

Study sites were selected based on invasion history. Five sites in four Preserve management units were selected to represent the invaded treatment. These sites previously supported monodominant R. pseudoacacia stands that had been clearcut during the winter and spring of 2016-2017. Five sites in five different management units were selected for non-invaded plots. Vegetation at these sites was identified as typical pitch pine – scrub oak communities by the APBP staff. None of the sites had been burned since the APBP had been established and managed by the Preserve Commission. A 12 m × 12 m experimental plot was established in each site at least 50 m from any roads, trails, management unit boundaries, or other experimental plots in the case of one invaded management unit which contained two plots.

Soil Bioassay

Three soil cores (5 cm diameter × 15 cm deep) were taken within each plot described above and pooled for an approximate total volume of 883 cm3 soil per plot and transported back to SUNY ESF in Syracuse. There, each sample was placed in a paper bag to dry for two weeks to select for resistant propagules that would survive drying. Soil was then sieved through 1 mm mesh before combining in equal proportions with sterilized peat and perlite. Peat and perlite were sterilized by autoclaving twice, 24 hours apart to ensure sterilization of heat stimulated spores that may have germinated after the first autoclave cycle. P. rigida seeds were harvested from cones collected in mature, non-invaded P. rigida stands at the APBP. Seeds were surface sterilized in 30% bleach for 10 minutes then scarified by stirring in room temperature water on a stir plate for 24 hours. Three seeds per pot were planted into 1:1:1 mixes of field soil, peat, and perlite in 164 ml cone-tainer pots (Stuewe & Sons, Inc., Corvallis, OR, USA). Soil mixes from each plot were kept separate. Ten pots were planted with the field soil mix of each plot, so that a total of 50 pots were planted for each site type (invaded or non-invaded). As seedlings grew they were culled to one per pot. Seedlings were grown in a laboratory growth chamber for six months under a 16 h : 8 h, light : dark photoperiod, with a temperature alternating between 20 °C during dark periods and 24 °C during light. Light levels were adjusted to approximate the full daylight sun in the APBP. Seedling racks were periodically rotated to avoid any confounding effects of the growth chamber on growth.

Six months after planting surviving seedlings were harvested and the whole root systems examined. Ectomycorrhizal root tips were categorized into morphological types (morphotypes) based on color, root branching, and other characters guided by Agerer (1987-2012) under a stereo microscope. Morphotype presence or absence on each seedling was recorded. Samples of each morphotype were collected for molecular analysis and kept separate by seedling to avoiding mixing genetic types with similar morphologies.

Field Experiment

Fruiting bodies of Rhizopogon and Suillus were collected from the APBP. Sequencing supported morphological field identifications of the fruiting bodies as R. pseudoroseolus and S. brevipes. Fruiting bodies were cleaned and used to prepare inoculum solutions following the protocol in Bruns et al. (2009). Combined inoculum solutions containing 107 spores ml-1 were prepared with equal concentrations of Rhizopogon and Suillus spores. Lamb and Richards (1978) found that spore concentrations of 106 Suillus and Rhizopogon spores were enough to achieve maximum levels of mycorrhization on Pinus radiata. Similarly, Ashkannejhad and Horton (2006) found that 86% of Pinus contorta seedlings had mycorrhizal roots when inoculated with spore solutions of 107 spores from deer fecal samples. Similar rates with spore solutions made from Rhizopogon fruiting bodies have been found in other studies (Castellano et al., 1985; Bruns et al., 2009). Spore concentrations were measured with a hemocytometer. A portion of inoculum was autoclaved twice over two successive days to control for non-mycorrhizal effects, such as fertilization, by the inoculum solution. P. rigida seeds from the APBP were prepared as above and planted in a 1:1 sand to perlite mix in cone-tainer pots. Sand and perlite were sterilized in the same manner as above. Seedlings were planted in March of 2016 and grown for 17 months in the same growth chamber conditions as the bioassay seedlings. Seedlings were randomly assigned to treatments with live or control inoculum. They were inoculated with 10 ml of live or control inoculum solution after 8 weeks of growth and again at 10 weeks. Two weeks prior to inoculation, seedlings were fertilized with modified Hoagland modified solution diluted 1:3 with distilled water (PhytoTechnology Laboratories LLC, Shawnee Mission, KS, USA).

A total 120 seedlings received live inoculum, 120 received control inoculum, and 20 received sterile water equal in amount and timing to those that received spore solutions. Twenty seedlings of each inoculum treatment were randomly chosen to be harvested and their roots examined to confirm presence or absence of mycorrhizal colonization. To assess contamination the 20 water-only control seedlings were harvested as well. In August 2017, the remaining 200 seedlings were transplanted into the APBP. Ten seedlings of each inoculation treatment (live or control inoculum) were randomly assigned to be planted into one of five replicate plots of each site type (invaded or non-invaded). Seedlings were planted in random order at least 3 m apart and at least 3 m from any existing P. rigida trees. A small cylindrical cage of poultry netting was placed around each seedling to deter browsing. The cages were staked to the ground together with a numbered tag to identify the seedling. In April 2018, after eight months in the field, seedling survival was recorded.

Molecular Identification

DNA was extracted from several representative root tip samples of each morphotype found in the bioassay, root tips of field experiment colonization-confirmation seedlings, fruiting bodies used to make inoculation spore solutions, and spore solutions themselves following Nuñez et al. (2013) and Hayward et al. (2015). The nuclear ribosomal internal transcribed spacer (ITS) region was amplified by PCR with standard Taq polymerase (New England Biolabs, Ipswich, MA, USA) using primers ITS1-F and ITS4 (Gardes & Bruns, 1993; White et al., 2014). Gel electrophoresis was used to determine which samples had successfully amplified. PCR products were run on 3% agarose gels then stained with ethidium bromide and imaged using a Gel Doc EZ System (Bio-rad, Hercules, CA, USA). Successful amplicons were digested with DpnII and HinfI restriction enzymes (New England Biolabs, Ipswich, MA, USA). Digested amplicons were run on 3% agarose gels, then stained with ethidium bromide, and restriction fragment length polymorphism (RFLP) band patterns imaged using the Gel Doc EZ System. RFLP patterns were compared to one another, fruiting body RFLPs, and lab references to group unique fingerprint patterns. ITS regions of two to four different root tip samples representing each different RFLP types were reamplified, quantified using a NanoDrop ND-2000 (Thermo Fischer Scientific Inc., Waltham, MA, USA) and PCR products were purified with a

Qiagen PCR Purification kit (Valencia, CA, USA) before being sent for sequencing by Eurofins Genomics following their recommended protocols and using the forward primer ITS1-F. Individual sequence chromatograms were manually screened for quality, sequences trimmed, and base-calls occasionally edited based on the chromatograms in CodonCode Aligner v. 8 (CodonCode Corporation, Dedham, MA, USA). The same program was used to construct alignments. Sequences and alignments were then subjected to a BLAST search against the GenBank database. Sequence alignments that could themselves be aligned with < 3% mismatch were considered the same Operational Taxonomic Unit (OTU). EMF were identified to species at > 97 % similarity with GenBank deposited sequences and to genus or higher at lower similarity. When BLAST searches returned conflicting or uncertain identifications even at high similarities a designation was made at higher taxonomic levels based on a consensus of BLAST results.

Data Analyses

The frequency of the EMF found on soil bioassay seedlings of each plot was used to calculate site-level average occurrences reported following Ashkannejhad & Horton (2006). R package iNEXT (version 2.0.19) was used to calculate richness, Shannon’s, and Simpson’s diversity estimates using site-level incidence frequencies in R (version 3.6.0) (Hsieh et al., 2019; R Core Team, 2019). Field experiment survival results were analyzed using a mixed effects logistic regression model with a binomial distribution. Survival outcome (surviving or dead at time of observation) was modeled as a function of fixed effects for site type (invaded or non-invaded) and inoculation treatment (live or control), their interaction, and a random effect of plot to account for the whole- plot error. The model was implemented using the lme4 package (version 1.1-21) in R (Bates et al., 2015).

Results

Soil Bioassay

Seedlings in 30 pots (of 50 planted) survived to harvest in soils from non-invaded sites (1 – 9 per source plot) and 28 seedlings survived in invaded soils (4-7 per plot). Visual inspection of root tips from bioassay seedlings resulted in 17 different root tip morphotypes, which were conservatively collapsed into 13 RFLP types represented by 36 samples sent for sequencing. Sequencing results further reduced the number of different EMF found to seven taxa (Fig. 1, Table 1 for top 5 BLAST results). Samples of a one morphotype appeared as a consistent RFLP type, but sequencing revealed an uncertain complex. BLAST returned Helotiales spp., Meliniomyces bicolor Hambl. & Sigler, and Cadophora finlandica (C.J.K. Wang & H.E. Wilcox) T.C. Harr. & McNew. This group is known to associate with roots of Ericaceae and Pinaceae and phylogenetic distinctions have been made, e.g. Hambleton & Sigler (2005), but for this study the ITS sequence data obtained did not allow finer classification to be resolved. As such, that morphotype is referred to here as Leotiomycetes sp. O.E. Erikss. & Winka. All invaded soil seedlings had at least one fungus present on roots, Pezizaceae sp. 1. Only four invaded soil seedlings had more than one taxon present. Total observed richness of invaded soil seedlings was three, and richness estimated by rarefaction asymptotically approached three (Fig. 2). Shannon diversity was 2.71 and Simpson diversity was 2.46. All non-invaded soil seedlings were colonized by an ascomycete as well, T. separans. Most seedlings had more than one EM fungus present. Total observed richness of non-invaded soil seedlings was five. Rarefaction estimated richness at five as well (Fig. 2). Shannon diversity was 4.97, and Simpson diversity was 4.94. Only the type identified as Leotiomycetes sp. was found on seedlings of both sites. It was more frequent on non-invaded soil seedlings (71 ± 18 %) than invaded soil seedlings (13 ± 10 %) (Fig. 1).

Table 1. Top five BLAST results for each sample sequence or alignment of ITS sequences from root tips of Pinus rigida soil bioassay seedlings.

Final Designation Query type Accession Alignment Max Total Query E value % Description (Number of length score score cover Identity (Abbreviated to ID sequences aligned) only) JN102365.1 594 1066 1066 98% 0 99% Pezizaceae sp. JN102381.1 603 1066 1066 100% 0 99% Pezizaceae sp. Alignment 1 (3) JN102422.1 594 1066 1066 98% 0 99% Pezizaceae sp. Pezizaceae sp. 1 JN102415.1 591 1061 1061 98% 0 99% Pezizaceae sp. (Alignments <3% MH794938.1 587 1059 1059 97% 0 99% Peziza sp. different from one JX434665.1 579 1064 1064 97% 0 99% Pezizales sp. another) JN102365.1 586 992 992 98% 0 97% Pezizaceae sp. Alignment 2 (2) JN102415.1 586 992 992 98% 0 97% Pezizaceae sp. JN102422.1 586 992 992 98% 0 97% Pezizaceae sp.

JN102374.1 585 987 987 98% 0 97% Pezizaceae sp. HM485385.1 610 1114 1114 99% 0 99% Tuber separans KU186911.1 600 1101 1101 97% 0 99% Tuber separans Alignment 1 (2) KC152255.1 611 1020 1020 99% 0 97% Tuber cf. separans T. separans KC152254.1 611 1013 1013 99% 0 97% Tuber cf. separans (Alignments <3% KT897484.1 612 1011 1011 99% 0 97% Tuber sp. different from one MG761623.1 566 902 902 100% 0 96% Fungal sp. another) MG761300.1 529 854 854 93% 0 96% Fungal sp.

Alignment 2 (2) MG761398.1 546 845 845 96% 0 95% Fungal sp. MH040299.1 567 839 839 100% 0 94% Pezizaceae sp. KM409439.1 562 824 824 99% 0 94% Peziza sp. JX030252.1 650 1146 1146 100% 0 99% Laccaria sp. KU878616.1 651 1138 1138 100% 0 98% Laccaria bicolor Laccaria laccata 1 Sequence KM067818.1 641 1133 1133 98% 0 99% Laccaria laccata KY777404.1 640 1129 1129 98% 0 98% Laccaria laccata var. pallidifolia JX504122.1 633 1127 1127 97% 0 99% Laccaria sp. AJ534899.1 421 778 778 100% 0 100% Laccaria sp. DQ149872.1 421 778 778 100% 0 100% Laccaria tortilis Laccaria tortillis Alignment (2) JF284353.1 421 778 778 100% 0 100% Laccaria sp. JF284355.1 421 778 778 100% 0 100% Laccaria sp. JQ724041.1 421 778 778 100% 0 100% Laccaria tortilis AJ810040.1 677 1179 1179 98% 0 98% Rhizopogon pseudoroseolus Rhizopogon Alignment (2) MF153075.1 642 1179 1179 93% 0 99% Rhizopogon pseudoroseolus pseudoroseolus MF153063.1 638 1175 1175 93% 0 99% Rhizopogon

pseudoroseolus MG773823.1 673 1171 1171 97% 0 98% Rhizopogon pseudoroseolus

AJ810041.1 678 1168 1168 98% 0 98% Rhizopogon pseudoroseolus JX507662.1 503 917 917 100% 0 99% Helotiales sp. JQ711936.1 503 900 900 100% 0 99% Meliniomyces sp. Alignment 1 (2) FN669230.1 503 896 896 100% 0 99% Meliniomyces sp. Leotiomycetes sp. KF428297.1 503 896 896 97% 0 99% Helotiaceae sp. (Alignments <3% AY394885.1 503 894 894 100% 0 99% Meliniomyces bicolor different from one KC007335.1 503 917 917 100% 0 99% Meliniomyces sp. another) KF428231.1 503 898 898 97% 0 99% Helotiaceae sp. Alignment 2(3) KF850368.1 503 889 889 100% 0 98% Cadophora finlandica FN669230.1 503 878 878 100% 0 98% Meliniomyces sp. JX507662.1 503 869 869 98% 0 98% Helotiales sp. KU924295.1 503 449 449 0.99 1E-122 1 Cenococcum sp. LC149749.1 503 449 449 0.99 1E-122 1 Cenococcum Cenococcum Alignment (2) geophilum geophilum LC149750.1 503 449 449 0.99 1E-122 1 Cenococcum

geophilum LC149751.1 503 449 449 0.99 1E-122 1 Cenococcum geophilum LC149752.1 503 449 449 0.99 1E-122 1 Cenococcum geophilum

1 0.9 0.8 0.7 0.6 Invaded 0.5 0.4 Non-invaded 0.3 0.2

0.1 Proportion of all seedlings colonized seedlings all of Proportion 0

Fungal taxa

Figure 1. Mean (± 1 SE) proportion of seedlings colonized by ectomycorrhizal fungi identified on roots of Pinus rigida laboratory bioassay seedlings grown in dried field soil from sites of former Robinia pseudoacacia invasion or from non-invaded sites. A total of 28 seedlings were sampled after 6 months of growth in invaded soil from 5 plots (red), and 30 seedlings were sampled after growing in non-invaded soil from 5 plots (blue).

Figure 2. Rarefaction of ectomycorrhizal richness on roots of soil bioassay Pinus rigida seedlings grown in dried field soil from five sites of former Robinia pseudoacacia invasion (red) or from five non-invaded sites (blue). A total of 28 seedlings were sampled after 6 months of growth in invaded soils, and 30 seedlings were sampled after growing in non-invaded soils. Shaded areas are 95% confidence intervals for estimates.

Field experiment

Only inoculation treatment was found to have a significant effect on survival of field seedlings (P < 0.001) (Table 2 for coefficients, Fig. 3). Mean proportion of surviving seedlings that received live inoculum in both site types was 0.72 ± 0.06 (mean ± 1 SE), while the mean proportion of surviving seedlings that received control inoculum was 0.31 ± 0.03. The live inoculum treatment seedlings had 4.2 greater odds of surviving than control inoculum seedlings. Mean proportion of surviving seedlings of both inoculum treatments did not differ by site, 0.53 ± 0.08 in non-invaded sites, and 0.50 ± 0.09 in invaded sites. The interaction of inoculation treatment and site type was not significant (P = 0.28), but it was kept in the model because we hypothesized that there would be a greater influence of inoculation treatment on survival in invaded sites than non-invaded sites due to reduced existing inoculum in invaded sites found in the bioassay. The random effect of plot was not significant (P = 0.06), but it was close to the chosen α of 0.05, and there was no significant difference between the full model and one without the random effect in a χ2 test of the difference in likelihood (χ2 = 0.17 (1), P = 0.683). The random effect of site was kept in the model to be conservative and account for random variances and heteroscedasticity between plots. Multiple EM root-tip morphotypes were found on field experiment live-inoculum confirmation seedlings harvested prior to out-planting. Although the inoculum solution contained spores from both Rhizopogon and Suillus fruiting bodies, only colonization by S. brevipes was positively identified by RFLP or sequencing. No mycorrhizal root tips were found on confirmation seedlings from the control inoculum treatment. One of the 20 water-only control seedlings had what appeared to be mycorrhizal root tips, but RFLP band patterns of DNA extracted and amplified from those root tips were weak and did not resemble anything found on field experiment seedlings or soil bioassay seedlings. Re-extraction, reamplification, and sequencing of DNA did not yield a useable sequence from that morphotype.

Table 2. Field experiment seedling survival mixed effects logistic regression model parameters. Survival was recorded for 200 total seedlings, 10 of each inoculation treatment (control or live spore inoculum) planted into 10 total sites (invaded or non-invaded, 5 replicates of each).

Fixed effects Coefficient

Log odds (LO) Odds Ratio (OR) LO SE OR SE z value Pr(>|z|) (Intercept) -0.58 0.56 0.31 1.37 -1.86 0.06 Site Type -0.48 0.62 0.46 1.58 -1.03 0.30 Inoculation Treatment 1.44 4.22 0.43 1.54 3.34 <0.01 Site × Inoculation 0.68 1.96 0.63 1.87 1.08 0.28 Random effects

Name Variance Std.Dev.

Plot (Intercept) 0.04787 0.2188

1 0.9 0.8 0.7 Invaded 0.6 0.5 Non-invaded 0.4 0.3

0.2 Proportion of seedlings surviving seedlings of Proportion 0.1 0 Control Live Inoculation treatment

Figure 3. Mean proportion of Pinus rigida seedlings surviving (error bars 1 SE) after 8 months in the field. Seedlings were treated with control spore inoculum or live spore inoculum. 10 seedlings per inoculum treatment were planted into 5 plots formerly invaded by Robinia pseudoacacia (red) and 5 plots in non-invaded sites (blue), so that each bar represents the proportion of 50 seedlings per treatment surviving.

Discussion

There was a greater richness of EMF inoculum present as resistant propagules in soils from non- invaded sites, as expected, and there was little overlap in fungal assemblages found in invaded and non-invaded soils. R. pseudoacacia are AM hosts, so would not support EMF compatible with pines (Wang & Qiu, 2006; Brundrett & Tedersoo, 2018). In the APBP, R. pseudoacacia has likely affected the native mycorrhizal communities directly by changes to soil nutrients and indirectly by changes to the quality or quantity of host plants in the invaded areas. R. pseudoacacia stands in the ABPB can have 1 – 3 times greater soil nitrogen concentrations, elevated net N-mineralization, and elevated N-nitrification rates compared to non-invaded pitch pine – scrub oak barren sites (Rice et al., 2004). The higher nitrogen levels in R. pseudoacacia stands relative to uninvaded areas constitute a change in soil properties that may, in part, be responsible for differences in the EMF community. Elevated levels of N can have a negative effect on EM fungal communities. R. pseudoacacia can increase soil nitrogen by up to 75 kg N ha -1 yr -1 (Boring & Swank, 1984). Along a gradient of anthropogenic N deposition in Alaska from 2.3 kg N ha -1 yr -1 to 14 kg N ha -1 yr -1, EMF richness decreased with increasing N (Lilleskov et al., 2002a). Increases in organic N also decreases richness and shifts communities, and may contribute to successional changes in EMF communities in forests (LeDuc et al., 2013; Cline et al., 2018). The vegetation of the APBP was not surveyed in this study, but it is reportedly different between invaded and non-invaded sites and even in the oldest R. pseudoacacia removal sites at the APBP. Sixteen years after removal of R. pseudoacacia the native plant community has not completely recovered even when fire has been reintroduced (Corbin, 2018 pers. comm.). R. pseudoacacia stands in a similar system in Cape Cod have increased non-native plant cover (Von Holle et al., 2013). Many of these non-natives are arbuscular mycorrhizal hosts (Von Holle et al., 2006). These together suggest a change in host quantity if not quality which may also influence EMF found in the bioassay. Resistant propagules found in the bioassay may be those able to persist in or disperse into invaded sites from non-invaded areas. However, given the small overlap between communities in invaded and non-invaded soils dispersal appears limited or may be influenced by site preparation. The soils of the previously invaded sites were partially mixed by heavy equipment

34 during the felling of R. pseudoacacia and extraction of their root systems which may have changed the distribution of wind or animal-borne propagules in soil profile such that spores near the surface were diluted through the soil column or were missed by sampling the upper 15 cm. Dispersal of spores from non-invaded sites may depend on mycophagus mammals, since dispersal distance from fruiting bodies is often limited and heterogeneous across landscapes (Galante et al., 2011; Peay & Bruns, 2014). This is especially true for hypogeous fungi like Rhizopogon, which produce buried or partially buried fruiting bodies. Whether animal dispersal of spores into R. pseudoacacia stands is limited is unknown. There are certainly mycophagus animals present in the area, white-tailed deer and small rodents. Anecdotally, fruiting bodies in non-invaded sites showed evidence of fungivory and droppings were found in the open after locust removal of invaded sites. The total richness of EMF in the bioassay was relatively low across both sites. A study of resistant propagules in soils of a Pinus muricata forest found up to 21 EMF types, including 6 suilloid types (Taylor & Bruns, 1999). In non-invaded sites the soils are undisturbed and there are many EM host individuals present, primarily oaks and pines. That only one suilloid, Rhizopogon pseudoroseolus, was found on bioassay seedling roots, is somewhat surprising, given observations of Suillus fruiting bodies in the APBP and the ability of suilloid spores to persist as long-live spore banks. However, it is consistent with the presence of existing EM host plants that it was only found on seedlings grown non-invaded soils if the previously invaded sites had reduced EMF and there is dispersal limitation from intact non-invaded areas to invaded sites. The differences in EMF community between invaded and non-invaded sites found in the bioassay apparently did not contribute to differences in survival of field seedlings planted into those same sites over the eight-month study period. Seedlings planted in non-invaded sites did not survive in greater numbers than those in invaded sites, despite the greater richness of EMF in non-invaded soils found in the soil bioassay or presumably intact and functioning hyphal networks. EMF richness in the soil not captured by the soil bioassay, which selected for species with resistant propagules may also play a role. Such bioassays do not capture the full diversity of fungi present (e.g. Taylor & Bruns, 1999). It is therefore possible that EMF symbionts undetected by the soil bioassays, but present in both the invaded and uninvaded sites, may have colonized and supported those seedlings. It would have been informative to survey the roots of

35 field seedlings to determine when control inoculum seedlings were colonized and by what EMF to compare to the bioassay seedlings. As expected, seedlings receiving live inoculum survived at a greater rate than those that received control inoculum. This difference in survival of inoculated vs. uninoculated seedlings is consistent with previous work (e.g. Castellano & Molina, 1989; Rincón et al., 2007; but see reviews by Castellano, 1996; Tomes, 2016). However, survival differences between control and live inoculum seedlings may in part be due to pre-transplant differences in health of seedlings attributable to inoculation. Additionally, live-inoculum seedlings had already allocated resources to their EMF prior to out-planting, while control inoculum seedlings would have to invest in them in the field. As such, caution should be used in drawing inferences about the efficacy of the specific composition of our inoculum and its comparisons to the presence or suitability of EM fungal inoculum in the APBP. Planting would typically be done in late spring, but transplanting was delayed until August due to R. pseudoacacia removal and site preparation. Transplant shock would have been exacerbated by planting at this time of year followed by the hotter, drier autumn the APBP experienced that year. These stresses may have had a greater effect on control inoculum seedlings if there were pre-transplant differences in health of seedlings. There are undoubtedly many processes affecting survival of seedlings in the field in addition to mycorrhizal dynamics, including differences in levels of light, temperature, soil moisture, nutrients, and pathogens. Lee et al. (2018) found that in the APBP, P. rigida seedlings germinated and survived poorly in the litter and shade of Q. ilicifolia but fared better in the low cover of Vaccinium. P. rigida is shade-intolerant and it may be that shading in the non-invaded sites decreased survival while the lack of competition for light in the recently cleared invaded sites allowed more seedlings to survive even if mycorrhizal colonization was reduced or with less favorable EMF. At the time of planting, non-invaded sites had much more undergrowth and Q. ilicifolia cover. Invaded sites had no tree canopy cover or undergrowth except rows of smaller native grasses sown by APBP managers after R. pseudoacacia was removed in the spring. While greater access to light may have benefited seedlings in invaded sites, those seedlings likely also experienced detrimentally higher soil temperature and lower soil moisture relative to those in non-invaded sites because of the reduced plant cover. As previously discussed, invaded sites have greater levels of soil N that may negatively affect EMF colonization but may positively affect the seedlings in the short term. After R.

36 pseudoacacia removal, soil N and total net mineralization rates decrease to normal levels within two years and net nitrification can remain elevated for up to four years (Malcolm et al., 2008). Within the first year of removal (i.e. during the study period) these levels were likely still elevated and could have supported the early survival of seedlings in invaded sites (Grove et al., 2017b). Non-invaded site seedlings presumably experienced greater interspecific competition in the more densely populated, more diverse plant communities of those sites than the invaded site seedlings, which were planted in the recently disturbed, relatively low-density, low plant richness sites. Control inoculum treatment seedlings in both sites would have experienced a delayed arrival of EMF symbionts, potentially obtaining them only after planting in the field, with some lag between planting and colonization. This delay would tend to place the pines at a disadvantage to their established competitors (Peay, 2018). In the invaded sites this delay and resulting competitive imbalance may have been less important or of smaller magnitude than in non-invaded sites. This could have contributed to the approximately equal survival between sites, even if control inoculum seedlings in non-invaded sites had more rapid colonization due to a greater richness of EMF and proximity to existing EM hosts including pines. Plant communities often exhibit negative density dependence. Host-specific antagonists reduce seedling survival when close to mature plants of the same species (Johnson et al., 2012). However, host-specific mutualists like EMF follow the same pattern of abundance as antagonists with distance from host plants (Laliberté et al., 2015). These competing influences in the soil can be difficult to interpret in natural systems. EM host trees tend to facilitate conspecific recruitment and recruitment of other EM hosts through plant-soil feedbacks mediated by their mycorrhizal relationship (Horton & Van Der Heijden, 2008; Bennett et al., 2017). The EMF provide nutrients to seedlings, improving survival and recruitment. In addition, EMF form a hyphal sheath that may offer greater pathogen protection to young roots (Laliberté et al., 2015). In an inoculation and transplant study, Bennet et al. (2017) found that greater survival near conspecific trees appears attributable to protection from soil antagonists, which, under conspecific trees, would be directly or indirectly provided by their mycorrhiza. In the present study there was no apparent positive plant-soil feedback evident from survival of seedlings in non-invaded areas. However, it may still be useful in interpreting these results. The antagonist pressure in non-invaded areas may have been of greater harm to control-inoculum seedlings than

37 access to greater EMF inoculum was beneficial. It seems likely that at least some of the success of live-inoculum seedlings would be attributable to protection from antagonists, but because of the caveat that there may have been pre-transplant differences in vigor of seedlings of different inoculation treatments it is unclear how important that may have been. All these factors may have contributed to the survival results, but together they suggest that legacies of R. pseudoacacia invasion (and the disturbance of removal) do not inhibit P. rigida survival in the short term. This may be expected, Grove et al. (2017b) found just a 5 % reduction in survival of Douglas-fir grown in soil invaded by N-fixing Cytisus scoparius relative to non-invaded soil after 15-18 months. Recruitment and survival of P. rigida can vary more than that due to factors not related to invasion (Landis et al., 2005). Inoculation with live suilloid spore solutions certainly benefitted the transplanted seedlings, but it is unclear in what ways or at what stage. As has been long practiced in forestry, providing seedlings with the appropriate inoculum can improve seedling establishment (Perry et al., 1987). Restoration efforts, including those by APBP managers, typically use nursery grown seedlings which are often colonized by at least a few EMF even if they are not deliberately inoculated (Castellano & Molina, 1989). How our inoculum would perform against those or other EMF would be useful information and a topic for future study.

Conclusion

Invasion by the nitrogen fixing Robinia pseudoacacia and its subsequent removal qualitatively changed the composition of resistant propagule EMF communities in comparison to non-invaded areas, but neither this change nor other legacies of invasion significantly affected survival of Pinus rigida seedlings planted into invaded or non-invaded areas. However, inoculation with local suilloid fungi significantly increased survival, which suggests a promising avenue for restoration. The inoculum chosen in this study represents local, early successional EMF that were not found in the resistant propagule community of invaded sites. It might be expected to out-perform nursery ‘contaminants’ or commercial inoculum and be less potentially disruptive than an outside inoculum source might be. It is hoped that this information will aid the managers of the Albany Pine Bush Preserve in their efforts to restore protected areas to the native plant community in service of their goals. Invasion legacies may not be limiting restoration of invaded

38 areas, but restoring the EM fungal community should be considered, both as a means and an end of restoring the functioning of the native communities.

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Appendices Appendix 1. R code used for analyses #R Code used in Analyses for LEGACY OF ROBINIA PSEUDOACACIA INVASION AND USE OF ECTOMYCORRHIZAL FUNGI TO RESTORE PINUS RIGIDA IN THE ALBANY PINE BUSH PRESERVE, NY. #Author: Taylor R. Patterson #April 2019

#Importing data and loading required packages ##### #Packages required: require(lme4) #This package is used for the mixed effects logistic regression model used to analyze survival data require(iNEXT) #This package is used for species accumulation curves and diversity estimates require(ggplot2) #This package is used to plot the rarefaction curve from iNEXT estimate output used in figure 2.

#Field experiment data surv <- read.csv("april2018fieldseedlingsurvival.csv", header = T) str(surv) #Check the structure of the imported data. Treatment factors need to be recognized as factors. Since the numeric codes of those factors are integers, we make them factors below: surv$Site <- as.factor(surv$Site) surv$SporetrmtNum <- as.factor(surv$SporetrmtNum) surv$PlantingtrmtNum <- as.factor(surv$PlantingtrmtNum)

#Bioassay data bioassay <- read.csv("bioassayincidencefrequencies.csv", header = T) str(bioassay) #Check structure of bioassay incidence frequency sheet. sitelevel <- aggregate(bioassay[,4:10], list(bioassay$Treatment, bioassay$Site), sum) #Collapse the seedling level presence/absence to site level sums of presences invaded <- subset(sitelevel, sitelevel$Group.1=="invaded", select = c(3:9)) #Subset the site level data to only invaded soil seedling

52 results for diversity estimates, only the columns 3 through 9 which are the species observed. invaded <- t(invaded) #Transpose the incidence frequencies as required by iNEXT invaded[which(invaded != 0)] <- 1 #Set all non-zero values to 1, turn site level sums to site level incidences. invadedincfreq <- as.incfreq(invaded) #Transform the transposed site level incidences the iNEXT incidence frequency table type noninvaded <- subset(sitelevel, sitelevel$Group.1=="noninvaded", select = c(3:9)) #Same procedure as above for the noninvaded soil seedling results noninvaded <- t(noninvaded) noninvaded[which(noninvaded != 0)] <- 1 noninvadedincfreq <- as.incfreq(noninvaded) bioassaylist <- list(Invaded=invadedincfreq, Noninvaded=noninvadedincfreq) #Make one list object (required by iNEXT) of the two incidence frequency tables

#Survival Model ##### full <- glmer(AprSurvNum ~ PlantingtrmtNum * SporetrmtNum + (1|Sitename), data = surv, family = binomial) #this is the full model. It incorporates the planting treatment (invaded or non-invaded), spore treatment (T = live spores, C = control, autoclaved spores), their interaction, and a random effect for the site they were planted in. anova(full) #ANOVA table of the full model summary(full) #Log odds coefficients and statistics of the full model used for Table 2. exp(summary(full)$coeff) #Exponentiated coefficients, which are the odds ratios (instead of log odds) used in Table 2. Only the Estimates and Standard errors make sense in this table, the exponentiated z values and p values do not.

#Diversity estimates ##### out <- iNEXT(bioassaylist, q=0, datatype = "incidence_freq") #run iNEXT function with hill number 0, input data type incidence frequencies) out$AsyEst #View the asymptotic diversity estimates

rarefaction <- ggiNEXT(out) + labs(x="Number of Plots Sampled", y="Richness") + theme(panel.background = element_blank(), axis.line = element_line(colour ="black"), text=element_text(size=16, family="serif")) + xlim(1,5) #Create the plot used for figure 2 rarefaction #Figure 2 plot

Curriculum Vitae Taylor R. Patterson 785.341.8694 [email protected]

Education M.S. Ecology ……………………………….…………………...………….………...……………2019 o State University of New York College of Environmental Science and Forestry, Syracuse, NY o GPA: 3.89 o Research Project: Legacy of Robinia pseudoacacia (black locust) invasion and use of ectomycorrhizal fungi to restore Pinus rigida (pitch pine) in the Albany Pine Bush Preserve o Major Professor: Dr. Tom Horton

B.S. Biology: Ecology and Evolutionary Biology, with Distinction ….………………….……….2013 and University Honors o University of Kansas, Lawrence, KS o GPA: 3.79 o Minor: Chemistry o Member of KU Honors Program

Employment and Professional Experience SUNY Research Foundation Research Assistant ……………………………...…..May 2017 – Present Albany Pine Bush Preserve Intern ……………………………………………………….Summer 2016 Research Assistant: KU Ecology and Evolutionary Biology …...………..……..May 2014 – July 2015 o Research assistant in the lab of Dr. John Kelly. Research Assistant: Kansas Biological Survey.…………… …...……………....May 2014 – July 2015 o Research assistant in the lab of Dr. Benjamin Sikes. Research Assistant: Monarch Watch, Monarch butterfly conservation organization ...... Summer 2013 o Student Hourly Employee

Teaching Experience SUNY ESF Graduate Teaching Assistant ……………………………...…...……..August 2015 – 2017 o Fall 2015: EFB 320, General Ecology Laboratory o Spring 2016: EFB 104, General Biology II Laboratory o Fall 2016: EFB 424/624, Limnology o Spring 2017: EFB 104, General Biology II Laboratory

Undergraduate Research Experience Independent research project under Dr. Orley Taylor of Monarch Watch ………………...Spring 2013 o Project title: Behavioral thermoregulation in Danaus plexippus fourth instar larvae and the effect on developmental duration NSF REU Participant ...…………………………………………………………………..Summer 2012 o Project Title: “Filamentous macroalgae as a biofuel feedstock” o Faculty advisor: Dr. Val Smith, KU Biology Department Research Experience Program certification ………………………………………………….May 2013 KU Center for Undergraduate Research.

Honors & Awards Graduate SUNY ESF Graduate Travel Grant ………………………………………………….…………….2018 o $500 to support travel to the 11th International Mycological Congress to be held July 16-21 in San Juan, PR. SUNY ESF Graduate Student Association Travel Grant …...……………………….…………….2018 o $250 to support travel to the 11th International Mycological Congress to be held July 16-21 in San Juan, PR. Edna Bailey Sussman Foundation Internship ………..…………….…………………….Summer 2016 o $7,350 funding an internship sponsored by the Albany Pine Bush Preserve investigating the role of mycorrhizal fungi in pitch pine restoration. Supervised by Neil Gifford, preserve director. Lowe-Wilcox Graduate Scholarship ………….…………………………………………….2015, 2016 o $3,000 scholarship in 2015, $1,750 in 2016 for SUNY ESF Environmental and Forest Biology graduate students working in the areas of forest pathology, mycology, wood microbiology, or plant physiology.

Undergraduate KU Summerfield Scholar ……………………………………………………………………2009-2013 NSF S-STEM Scholarship Recipient ………………………………………………………..2011-2013 o Recipients participated in a renewable energy outreach and education group. o Awarded under NSF EPSCoR Kansas Center for Solar Energy Research. KU College of Liberal Arts and Sciences Honor Roll ……………………………...... 2009-2013

Scientific Publications Prescribed fire reorganizes fungal communities and alters decomposition across a heterogeneous pine savanna landscape. (In revision) Authors: Tatiana A. Semenova-Nelsen, William J. Platt, Taylor R. Patterson, Jean Huffman, Benjamin A. Sikes

Presentations and Communication 11th International Mycological Congress ……………………………………………………. July 2018 o Symposium presentation: Improving pitch pine restoration outcomes in the Albany Pine Bush Preserve using ectomycorrhizal fungi o Authors: Taylor R. Patterson, Dr. Thomas R. Horton

Albany Pine Bush Preserve Research Symposium …………………...……………………..April 2017 o Poster Title: Improving pitch pine restoration outcomes in the Albany Pine Bush Preserve using ectomycorrhizal fungi. o Authors: Taylor R. Patterson, Dr. Thomas R. Horton

SUNY ESF Spotlight on Student Research Poster Presentation ……………………………April 2017 o Poster Title: Improving pitch pine restoration outcomes in the Albany Pine Bush Preserve using ectomycorrhizal fungi. o Authors: Taylor R. Patterson, Dr. Thomas R. Horton

Other Activities and Service SUNY ESF Graduate Student Association Undergraduate Association Representative …....2017-2018 SUNY ESF Academic Governance Awards Committee Graduate Student Member ……….2017-2018 SUNY ESF Graduate Student Association Secretary ………………………...………..…....2016-2017 56

Member of KU Cycling Club ………………………………………………………………2009-2013 Member of Sigma Lambda, Sophomore Honor Society, service organization ...…………... 2010-2011 Member of Global Scholars Program, First Cohort …………………………………………….…2011 Study Abroad: London Review Program ……………………………………………………….....2011