Vegetation, Coarse Woody Debris, and Cerambycid Communities in Riparian Forests Invaded by the Emerald Ash Borer

VEGETATION, COARSE WOODY DEBRIS, AND CERAMBYCID COMMUNITIES IN RIPARIAN FORESTS INVADED BY THE EMERALD ASH BORER

Patrick Engelken

A THESIS

Submitted to Michigan State University in partial fulfillment of the requirements for the degree of

Entomology – Master of Science

2019

ABSTRACT

VEGETATION, COARSE WOODY DEBRIS, AND CERAMBYCID COMMUNITIES IN RIPARIAN FORESTS INVADED BY THE EMERALD ASH BORER

Patrick Engelken

This thesis, presented in three chapters, focused on evaluating the effects of emerald ash borer (EAB) on forest overstories, regeneration dynamics, and associated woodboring cerambycid beetles in riparian forests across Michigan.

In chapter one, forests bordering first order streams were evaluated in three watersheds, representing a temporal gradient of the EAB invasion across southern Michigan. Canopy gap overstories were originally dominated by ash species but >85% were killed by EAB across all sites. Coarse woody debris (CWD) was most abundant in southeastern sites, with the longest history of EAB invasion. Gap regeneration was dominated by ash but seedlings were rare and understories were dominated by dense sedge mats (Carex spp.).

In chapter two, I conducted trapping surveys of cerambycid beetles in the sites from chapter one. Beetle assemblages differed among invasion stages, and between traps at different heights. Several species were more abundant in southeast sites than southcentral and southwest sites and appear to be exploiting the abundant CWD in southeast Michigan.

In chapter three I evaluated riparian forests bordering economically important northern lower Michigan rivers. Canopy gaps from EAB killed ash trees comprised 15-20% of the riparian buffer within 100 m of river banks. Prior to the EAB invasion, green and black ash dominated riparian forests, but >95% had died. Most dead ash trees were standing and ash CWD was not yet abundant. In gaps, green ash dominated regeneration, but black ash was uncommon.

As in southern Michigan, dense sedges dominated gap understories, and seedlings were rare.

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ACKNOWLEDGEMENTS

I thank my advisor, Dr. Deborah G. McCullough, for her guidance, leadership and invaluable insight throughout this project. Deb introduced me to the world of forest insects, for which I’ll always be grateful. Thank you for not only broadened my horizons but for imparting to me the value of being a more inquisitive researcher. I also would like to thank my guidance committee members: Dr. Therese Poland, Dr. M. Eric Benbow, and Dr. Michael Walters. Thank you for your encouragement and assistance along the way.

These projects would not have been possible without the help and support from my fellow lab mates and undergraduate students in the Forest Entomology Lab. A special thanks goes to Justin Keyzer and Ryan Rupp for the countless hours spent collecting data in the field.

No matter how unpleasant the conditions were or how thick the mosquitoes got, these two never wavered and were always great to work with. Additionally, I would like to thank James

Wieferich for lending insight and advice in both the lab and the field.

I thank my parents for loving me, supporting me and always pushing me to be my best self. Thank you for teaching me to love nature, and for always believing in me. Without you guys I would undoubtedly not be where I am now.

I thank my fiancé, Angelica for enduring me through my highs and my lows, and for supporting and loving me throughout this process. You never stopped believing in me and kept me motivated even when I was at my worst. I love you unconditionally and you know that you’re my favorite.

Funding for this research was provided through the Evaluation Monitoring program of the USDA Forest Service, Forest Health Protection, Northeastern Area. Partial funding to enable

iii presentation of this research was provided by The College of Agriculture and Natural Resources, and the Ray and Bernice Hutson Memorial Entomology Endowment Travel Funds.

PREFACE

Emerald ash borer (EAB) (Agrilus planipennis Fairmaire) (Coleoptera: Buprestidae), a phloem-boring beetle native to Asia was first detected in the greater metropolitan area of Detroit,

Michigan, USA in 2002. The range invaded by EAB has continued to progress and now encompasses much of the eastern U.S. and Canada. Previous studies have quantified mortality of ash (Fraxinus spp.) trees following EAB invasion, particularly in forests in southeast Michigan and Ohio, but little is known about the impacts of EAB in riparian forests bordering streams and rivers. In the northeastern U.S., black ash (F. nigra Marsh.) and green ash (F. pennsylvanica

Marsh.) are abundant in riparian forests and are both highly preferred and vulnerable EAB hosts.

In riparian forests where ash trees were abundant prior to the EAB invasion, mortality could result in cascading effects in both terrestrial and aquatic environments.

My research, which is presented in three chapters of this thesis, focused on evaluating effects of EAB on forest overstories, regeneration dynamics, and associated woodboring cerambycid beetles in riparian forests of Michigan. Each chapter is intended to be submitted as a manuscript for publication in scientific journals.

In Chapter One, I assessed the overstory trees, regeneration, shrubs, herbaceous plants, coarse woody debris, and photosynthetically active radiation in canopy gaps formed from EAB killed ash trees and in adjacent forests in riparian forests bordering first order streams. Sites were located in three distinct watersheds across southern Michigan and were selected to represent a temporal gradient of the EAB invasion. In canopy gaps along streams, green ash (F. pennsylvanica) and black ash (F. nigra) dominated the overstory before the EAB invasion but more than 85% were killed by EAB in all sites, resulting in canopy gaps. In southeastern sites

v with the longest EAB invasion history, dead ash trees had begun to fall and coarse woody debris

(CWD) volumes were higher than in areas of more recent EAB invasion, where most of the dead ash remain standing. Regeneration in canopy gaps was dominated by ash but seedlings were rare and gap understories were dominated by dense sedge mats (Carex spp.).

In Chapter Two, I captured cerambycid beetles (Coleoptera: Cerambycidae) in cross vane panel traps baited with (R) 3-hydroxyhexan-2-one, an aggregation pheromone attractive to several genera of cerambycids. Traps were deployed in the canopy and at ground level at the perimeter of the canopy gaps described in Chapter 1. Trapping took place during summer months of 2017 and 2018, and captured beetles were collected at 2-3 week intervals. During both trapping seasons, the majority of species were initially captured early in the summer. Beetle captures were similar among watersheds and between canopy and ground level traps, but canopy traps captured more species than ground traps. Several species captured in high abundance were notably more common in different watersheds, and are perhaps responding to forest changes in the aftermath of the EAB invasion.

In Chapter Three, I evaluated riparian forests along three rivers in northwest lower

Michigan that are economically important for recreation, provide habitat for spawning Great

Lake trout and salmon and drain into reservoirs that empty into Lake Michigan. Forests were evaluated in a similar manner as in Chapter One. In these northern forests, canopy gaps from

EAB killed ash trees comprised 15-20% of the forest area near river banks. Prior to the EAB invasion, green ash (F. pennsylvanica) and black ash (F. nigra) dominated overstories in all gaps, but >95% of those trees had died. Most dead ash trees remained standing and ash CWD was limited in all canopy gaps. Green ash saplings and recruits dominated the advance

vi regeneration in canopy gaps, but black ash regeneration was minimal. As in southern Michigan, seedlings were rare in gaps, where understory vegetation was dominated by dense sedge mats.

Overall, results from this research provide a baseline of the post-EAB invasion status of riparian forests across Michigan. Green ash and black ash have been functionally lost in the overstory, and canopy gaps now make up a large component of the riparian corridors. While ash regeneration is present in canopy gap understories, dense sedge mats may inhibit future seedling recruitment. If the remnant ash cohort is unable to persist, large areas where ash was abundant prior to the EAB invasion may transition away from forests towards sedge dominated meadows.

This may present long term changes in terrestrial and aquatic conditions. Ash mortality appears to be altering assemblages of cerambycid species, indicating EAB can indirectly affect other taxa. Effects of EAB related changes in riparian forests on aquatic ecosystems are as of yet largely unknown.

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TABLE OF CONTENTS

LIST OF TABLES ...... x

LIST OF FIGURES ...... xiii

CHAPTER 1: LEGACY EFFECTS OF EMERALD ASH BORER ON RIPARIAN FOREST VEGETATION AND STRUCTURE ...... 1 Introduction ...... 1 Materials and Methods ...... 6 Study Sites ...... 6 Estimate of Gap Formation ...... 6 Survey Design ...... 6 Statistical Analysis ...... 9 Results ...... 12 Canopy Gap Formation ...... 12 Overstory Vegetation ...... 12 Overstory Ash Condition ...... 14 Coarse Woody Debris ...... 15 PAR ...... 17 Saplings ...... 18 Seedlings ...... 18 Herbaceous Plants ...... 19 Invasive Plant Species...... 20 Species Compositions of Pre-EAB Overstories and Current Saplings ...... 21 Discussion ...... 23 APPENDIX ...... 31

CHAPTER 2: SPECIES DIVERSITY AND ASSEMBLAGES OF CERAMBYCIDAE IN THE AFTERMATH OF THE EMERALD ASH BORER INVASION IN RIPARIAN FORESTS IN SOUTHERN MICHIGAN...... 48 Introduction ...... 48 Materials and Methods ...... 51 Study Sites ...... 51 Site Surveys ...... 51 Cerambycid Trapping ...... 52 Collection and Identification...... 53 Statistical Analysis ...... 53 Results ...... 56 Overall Beetle Captures ...... 56 Seasonal Activity ...... 56 Watershed Comparisons ...... 58 Canopy vs. Ground Trap Comparisons ...... 58

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Comparison of Cerambycid Species Assemblages Among Watersheds ...... 58 Comparison of Assemblages in Canopy vs. Ground Traps ...... 60 Discussion ...... 61 APPENDICES ...... 66 APPENDIX A. Tables and Figures ...... 67 APPENDIX B. RECORD OF DEPOSITION OF VOUCHER SPECIMENS...... 80

CHAPTER 3: RIPARIAN FOREST CONDITIONS ALONG THREE NORTHERN MICHIGAN RIVERS INVADED BY THE EMERALD ASH BORER (AGRILUS PLANIPENNIS) ...... 83 Introduction ...... 83 Materials and Methods ...... 88 Study Sites ...... 88 Survey Design ...... 88 Overstory...... 89 Regeneration ...... 89 Coarse Woody Debris ...... 89 Statistical Analysis ...... 90 Results ...... 92 Gap Size and Frequency ...... 92 Overstory Vegetation ...... 92 Overstory Ash ...... 94 Coarse Woody Debris ...... 95 Herbaceous Plant Cover ...... 95 Seedlings ...... 96 Saplings ...... 96 Recruits ...... 97 Species Composition of Pre-EAB Overstories and Current Regeneration ...... 98 Discussion ...... 100 APPENDIX ...... 105

LITERATURE CITED ...... 114

LIST OF TABLES

Table 1.1. Mean (± SE) basal area of live and dead overstory trees of all species and ash (Fraxinus spp.), species richness and volume of coarse woody debris represented by ash and by all species combined recorded in four canopy gaps and surrounding forests in three watersheds in Michigan. (n = 12 sites) ...... 32

Table 1.2. Number of overstory trees (DBH > 10 cm), mean (± SE) Diameter at breast height (DBH), basal area, and relative importance values (RIVs) for the five most dominant overstory species recorded in four canopy gaps and surrounding forests in each of three watersheds (n=12 sites) ...... 33

Table 1.3. Mean (± SE) of pre-EAB and current ash component1 of total basal area, current ash component, percent of ash BA killed by EAB, three ash (Frazinus spp.) species tallied in canopy gaps and surrounding forests in the three watersheds in Michigan (n=12 sites) ...... 34

Table 1.4. Mean density of seedling species (no. per m2) tallied in canopy gaps and surrounding forest in three watersheds in Michigan. Species are presented in descending order of overall density...... 35

Table 1.5. Mean (± SE) density of saplings (no. ∙ ha-1) in four canopy gaps and surrounding forests in three watersheds across southern Michigan. Species listed comprise ≥ 5 % of the total number of saplings tallied are listed in descending order of overall densities (n=12 sites) ...... 36

Table 1.6. Frequency and mean (± SE) percent cover of the 10 most abundant herbaceous plant species in canopy gaps and surrounding forests recorded in 12 sites in three watersheds of southern Michigan ...... 37

Table 1.7. Mean (± SE) density of the four most abundant invasive plant species recorded in 12 sites in three watersheds in Michigan ...... 38

Table 1.8. Results of PERMANOVA analysis testing for differences in forest compositions among the three watersheds and between canopy gaps and surrounding forests and the associated pairwise comparisons ...... 39

Table 1.9. Group identifier from non-metric multidimensional scaling ordination, indicator values (IV) and P values for significant indicator species of trees in canopy gaps and surrounding forests ...... 40

Table 2.1. Year of canopy gap formation, mean (± SE) live basal area of current overstory, percent of total basal area that was ash prior to EAB invasion, percent of ash basal area represented by dead trees, forest overstory species richness, Dominant overstory species and CWD volume in canopy gaps and surrounding forests in three watersheds across southern Michigan ...... 67

Table 2.2. Number of individuals of cerambycid species captured in canopy traps and ground traps by subfamily and tribe ...... 68

Table 2.3. Mean (±) number of beetles collected (N), species richness, and Chao 2 species richness estimates for cerambycid species captured in traps in the Clinton, Grand and Kalamazoo river watersheds by year and total (n=12 sites) ...... 71

Table 2.4. Results of PERMANOVA analysis testing for differences in cerambycid captures among the three watersheds and between the two trap locations (canopy versus ground traps) and pairwise comparisons ...... 72

Table 2.5. Group identifier from non-metric multidimensional scaling ordination, Indicator values (IV) and P values for significant indicator species for cerambycid species captured in three watersheds in traps located in forest canopies or at ground level ...... 73

Table A.1. Quantity and preservation method of Cerambycid species turned in as voucher specimens to the Albert J. Cook Arthropod Research Collection...... 80

Table 3.1. County, date of emerald ash borer quarantine, total linear length of river, number of canopy gaps caused by EAB-killed ash trees, average (± SE) frequency and size of canopy gaps, and current percentage (± SE) of forest comprised of canopy gaps in 100 m of the three north rivers surveyed ...... 106

Table 3.2. Mean (± SE) basal area and stem density of live and dead overstory trees (> 10.2 cm DBH), pre-EAB and current ash, and volume of coarse woody debris represented by ash and by all species in nine canopy gaps and nine forested areas along three northern rivers in Michigan ...... 107

Table 3.4. Mean density of seedling species (no. per m2) tallied in nine canopy gaps and nine forested areas along three rivers in northern Michigan. Species are presented in descending order of overall density...... 109

Table 3.5. Mean (± SE) density of recruits and saplings (no. ∙ ha-1) in three canopy gaps and forested areas along three rivers in northern Michigan. Species listed comprise ≥ 5 % of the total number of saplings tallied and are listed in descending order of overall densities ...... 110

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LIST OF FIGURES

Figure 1.1. Location of the 12 riparian sites in the three watersheds selected to represent the temporal progression of the emerald ash borer across southern Michigan ...... 41

Figure 1.2. Canopy gaps and surrounding forests were surveyed using fixed radius plots comprised of a macroplot (11.4 m radius) where overstory trees were recorded, a subplot (8 m radius) where saplings and percent cover of shrubs were recorded, and a microplot (1 m2 area) where seedlings and herbaceous vegetation were tallied. Overstory trees and coarse woody debris were recorded in (50x2 m or 100x2 m) linear transects ...... 42

Figure 1.3. Mean (± SE) volume of coarse woody debris (m3 ∙ ha-1) in four canopy gaps and surrounding forests in each of three watersheds in Michigan (n = 12 sites) ...... 43

Figure 1.4. Linear relationship between total volume of coarse woody debris (m3 ∙ ha-1) in canopy gaps and the first year that overstory ash mortality was apparent in aerial imagery ...... 44

Figure 1.5. Linear relationship between percent cover of herbaceous plants and (A) seedling density and (B) the density of saplings (<2.5 cm diameter) in four canopy gaps and surrounding forests in each of three watersheds in Michigan (n=12 sites) ...... 45

Figure 1.6. Non-metric multidimensional scaling ordination (two-dimensional) of species composition in pre-EAB forest overstories and in the current sapling stratum. Polygons were overlaid to easily visualize groupings of sites of canopy gaps and surrounding forests. The successional arrows connect pre-EAB overstories to the corresponding saplings in the same site to examine stand level changes in species composition. Sites are labelled according to watershed, an identification number of the site (1-4) in a watershed and whether it was a canopy gap or surrounding forest. A small “s” refers to the sapling stratum of a site. For example: a site labelled “C1Gs” would denote that the site is in the Clinton River watershed, is a canopy gap was the first of four sites surveyed in this watershed and references the sapling stratum of the site. Final stress = 14.57; axis 1 R2 = 58.2%; axis 2 R2 = 26.2 %; cumulative R2 = 86.4 %...... 46

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Figure 1.7. Multivariate regression tree (MRT) comparing square root transformed species compositions of pre-EAB overstory and current sapling strata in canopy gaps and surrounding forests of three watersheds. Cross validation of 1000 trees suggested a 3 branched tree the most often (748). Total variance explained: 78.3% ...... 47

Figure 2.1. Locations of 12 trapping sites in three watersheds selected to represent the temporal gradient of the emerald ash borer invasion ...... 74

Figure 2.2. Seasonal activity, representing the period when individuals of each species (with ≥ 10 captured over the course of both years) were captured. Activity is displayed as the percentage of the total collected individuals during a range of accumulated GDD (base 10 °C). The monthly timeline along the bottom indicates the average accumulated GDD for each of the three watersheds on the 1st of each month during 2017 and 2018 ...... 75

Figure 2.3. Numbers of cerambycid species captured by accumulated growing degree days (base 10 ̊ C) and dates in 2017 (A) and 8-10 June 2018 (B). Degree days are recorded beginning on 1 March annually ...... 76

Figure 2.4. Individual based rarefaction estimates comparing cerambycid captures by traps suspended from a mid-canopy branch vs. ground traps hung on 1.5 m tall rebar and among watersheds (Clinton/Grand/Kalamazoo) during (A,B) 2017, (C,D) 2018 and for both trapping seasons combined. (E,F)...... 77

Figure 2.5. Nonmetric multidimensional scaling output (2-dimensional) of cerambycid species assemblages captured by traps in sites in the three watersheds. Two site variables with weak correlation with either axis are overlaid (Fresh CWD and total CWD). Final stress=12.59; axis 1 R2 = 60.8%; axis 2 R2 = 27.5%; cumulative R2 = 88.3% ...... 78

Figure 2.6. Nonmetric multidimensional scaling output (2-dimensional) of cerambycid species assemblages captured by canopy and ground traps. Final stress=11.81; axis 1 R2 = 22.1%; axis 2 R2 = 67%; cumulative R2 = 89.1% ...... 79

Figure 3.1. Location of surveyed rivers and associated watersheds in northern lower Michigan ...... 111

Figure 3.2. Linear relationship between percent cover of herbaceous plants and density of (A) seedlings, (B) saplings, and (C) recruits in three canopy gaps and three forested plots in each of three surveyed rivers in northern Michigan ...... 112

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Figure 3.3. Multivariate regression tree (MRT) comparing log transformed species compositions between pre-EAB overstory and current sapling and recruit strata in canopy gaps and adjacent forests of three northern Michigan rivers. Cross validation of 1000 runs yielded a 10 branched tree most frequently (360). Explained 46.9% of total variance ...... 113

CHAPTER 1: LEGACY EFFECTS OF EMERALD ASH BORER ON RIPARIAN FOREST VEGETATION AND STRUCTURE

Introduction

More than 450 nonindigenous forest insect species are established in North America

(Aukema et al. 2010). While most of these species have relatively minor effects, a subset has caused serious damage, affecting an array of ecological processes and often resulting in substantial economic costs (Aukema et al. 2011, Gandhi and Herms 2010, Liebhold et al. 2017,

Lovett et al. 2016). In contrast to most natural disturbances, which indiscriminately remove trees regardless of species, approximately two-thirds of the nonindigenous insects in the US are monophagous, colonizing plants in a single genus, or oligophagous, colonizing hosts in a single family (Aukema et al. 2010, Niemela and Mattson 1996). Non-host trees in mixed species stands may benefit from reduced competition if a dominant overstory species is killed by an invader, leading to changes in forest composition and affecting successional dynamics in the understory

(Costilow et al. 2017, Ford et al. 2012, Gandhi and Herms 2010, Lovett et al 2006, Morin and

Liebhold 2015, Orwig and Foster 1998). Over the long term, loss of a major overstory species can generate cascading ecological impacts, further altering forest composition, nutrient dynamics, and wildlife habitat (Hoven et al. 2017, Flower et al. 2013, Lovett et al. 2006, Morin and Liebhold 2015). Regeneration and the ability of young trees to persist and grow into the overstory largely determines whether an affected species persists as a functional component of a forest following a major pest invasion (Burr and McCullough 2014, Flower et al. 2013, Herms and McCullough 2014, Klooster et al. 2014, Orwig and Foster 1998).

Subcortical insects that feed on phloem and/or wood represent a relatively small proportion of the nonindigenous forest insect species established in the USA (Aukema et al.

2010), but account for disproportionately high economic costs relative to forest defoliators and

sap-feeders (Aukema et al. 2011). Regulations designed to reduce new introductions of wood- and phloem-borers into the USA include mandatory treatment of solid wood packing material such as crating, pallets, dunnage and related materials used to protect an array of commodities transported as cargo (Haack and Brockerhoff 2011). While these efforts have been somewhat successful at reducing interceptions of larvae in wood packing material (Haack et al. 2014,

Leung et al. 2014, Lovett et al. 2016), they are probably offset to some degree by continued increases in international trade (Aukema et al. 2011, Brinkerhoff et al. 2014, Lovett et al. 2016).

Emerald ash borer (EAB) (Agrilus plannipenis Fairmaire) (Coleoptera: Buprestidae), a phloem-boring beetle native to eastern Asia, was first detected in 2002 in the greater Detroit metropolitan area in southeast Michigan and nearby Windsor, Ontario, Canada (Cappaert et al.

2005). Since 2002, infestations of EAB have been detected in 35 U.S. states and five Canadian provinces (EAB.info 2019) and hundreds of millions of ash trees in forests and landscapes have been killed (Herms and McCullough 2014). Populations of this invasive pest, which has already become the most destructive and costly forest insect to ever invade North America (Aukema et al. 2011, Herms and McCullough 2014, Lovett et al. 2016), continue to spread.

Dendrochronological evidence showed EAB populations were established by the early

1990s and killing ash (Fraxinus spp.) trees in localized areas of southeast Michigan by the mid to late 1990s (Siegert et al. 2014). Progression of the EAB invasion across an east-west gradient in southern and central Michigan is reflected by previous studies that showed ash mortality rates in forested areas were inversely related to the distance of sites from the original EAB epicenter near

Detroit (Burr and McCullough 2014, Klooster et al. 2014, Smith et al. 2015). Although populations of EAB are established in all counties in the lower peninsula of Michigan, forests in

southeastern, central and southwestern Michigan represent different invasion histories and different stages of post-invasion recovery.

Although all or nearly all North American ash (Fraxinus spp.) species are likely to support EAB development (Herms and McCullough 2014, Poland and McCullough 2006), interspecific differences in EAB host preference and host resistance are consistently observed

(Anulewicz et al. 2007, 2008, Rebek et al. 2008, Tanis and McCullough 2015, Villari et al.

2016). Reported mortality rates for green ash and black ash, which are highly preferred and vulnerable EAB hosts, have ranged from 79 to 99% in southeast Michigan forests (Burr and

McCullough 2014, Kashian et al. 2018, Klooster et al. 2014, Knight et al. 2013, Smith et al.

2015). Green ash (F. pennsylvanica Marsh.) and black ash (F. nigra Marsh.) are notably well suited to hydric forests and one or both species are often abundant in riparian corridors along streams and rivers or in bogs or swamps (Gucker 2005 a,b, Kennedy 1990, Wright and Rauscher

1990). White ash (F. americana) appears to be an intermediate EAB host, with mortality rates in southeast Michigan ranging from 99% to less than 20% of the trees or basal area (Klooster et al.

2014, Robinett and McCullough 2019, Smith et al. 2015, Tanis and McCullough 2012). In contrast, healthy blue ash (F. quadrangulata) trees growing on upland fertile soils are not preferred EAB hosts and appear likely to persist in post-invasion forests (Tanis and McCullough

2012, 2015). Like most phloem- and wood-borers, EAB larvae cannot complete their life cycle on seedlings, but they can develop on trees as small as 2.5 cm in diameter (Cappaert et al. 2005).

Long term prospects for ash regeneration and persistence, therefore, remain uncertain in areas with catastrophic levels of ash mortality.

Riparian forests are the forested areas adjacent to waterways where periodic inundation occurs. These forests are functionally linked to the aquatic systems they border, and may be

particularly affected by the EAB invasion in areas where ash is a dominant and abundant component of the overstory. Riparian forests regulate transfer of energy to the forest floor and adjacent waterways via inputs of nutrients in leaf litter and coarse woody debris and affect light and temperature through filtration of sunlight (Lovett et al. 2004, Nisbet et al. 2015). Losing a major overstory species in riparian zones affects not only overstory structure and species composition, but can also impact leaf litter quality, soil and water chemistry, light and temperature (Huddleston 2011, Lovett et al. 2004, Nisbet et al. 2015, Wallace et al. 1997). These changes may be especially important along first order streams, which likely have limited buffering capacities to changes in their surroundings. Additionally, in a river network, communities of consumers occupying first order streams rely the heaviest on the surrounding vegetation (Vannote et al. 1980). As stream size and order increase, species adjustments occur to capitalize on processing inefficiencies or nutrient “leaking” from upstream consumers. This results in a continuum of species replacement as stream size increases with communities adapted to break down smaller and finer organic material occupying waterways the more downstream from a headwater stream they occur (Vannote et al. 1980). Because of this, changes in vegetation in riparian forests bordering headwater streams may result in largescale differences in organic material availability downstream, cascading across multiple trophic levels and influencing resource availability throughout the entire river network.

While previous studies have quantified mortality rates of ash species in stands invaded by

EAB (Burr and McCullough 2014, Klooster et al. 2014, Knight et al. 2013, Robinett and

McCullough 2019, Smith et al. 2015), little is known about EAB effects in riparian forests. In this study, I investigated conditions in riparian forest sites representing different stages of recovery from the EAB invasion. I assessed overstory vegetation, regeneration, shrubs and

herbaceous vegetation, quantified light interception, and evaluated coarse woody debris in canopy gaps created by EAB-caused ash mortality and surrounding forests in sites along first order streams. I hypothesized that in gaps resulting from overstory ash mortality, density of ash regeneration would be inversely related to the progression of the EAB invasion across southern

Michigan, such that ash regeneration would be lower in sites with a longer history of invasion, while regeneration in forests surrounding the gaps would largely reflect composition of the non- ash overstory. Volume of coarse woody debris (CWD), particularly fallen ash trees, was predicted to be substantially higher in gaps than in surrounding forests, but whether CWD would vary across the east-west invasion gradient in southern Michigan was unknown.

I additionally hypothesized that invasive plant species were likely to exploit light available in canopy gaps resulting from overstory ash mortality and would be more abundant or diverse in southeast Michigan sites with the longest history of EAB establishment than in more recently invaded sites further west. Establishment of invasive plant species, which often form dense monocultures in disturbed habitats, can suppress regeneration of native forest species

(Cipollini et al. 2009, Collier et al. 2009, Stinson et al. 2006). Additionally, changes resulting from a major invasive pest such as EAB have been theorized to facilitate invasion by other species, i.e., an invasional meltdown (Hoven et al. 2017, Klooster et al. 2018, Simberloff and

Von Holle 1999). Evaluating riparian forests at varying stages of recovery from the EAB invasion can provide useful information for projecting forest composition and regeneration dynamics and could contribute to developing a framework for future restoration efforts.

Materials and Methods

Study Sites: I identified three watersheds in southern Michigan that represent the temporal progression of the EAB invasion across an east-west gradient (Figure 1.1). Watersheds included the Clinton River watershed in the southeast, which was invaded relatively early by EAB, the recently invaded Kalamazoo River watershed in the southwest, and the Grand River watershed in south central Michigan, which is intermediate in both location and EAB invasion history. I used the Michigan Atlas and Gazetteer (DeLorme 2016) to identify first order streams in forested areas on public lands in each watershed, then examined aerial photos to identify major canopy gaps in the midst of largely intact forest along the streams. Areas with canopy gaps were visited to determine if the gaps resulted from EAB-caused ash mortality and if so, GPS coordinates were recorded at the midpoint of the gap. I identified four study sites in each of the three watersheds

(12 sites total) (Figure 1.1). All sites were in lowland hardwood forests that included a substantial, albeit patchy, ash component in the overstory prior to the EAB invasion.

Estimate of Gap Formation: Aerial images from the U.S. Geological Survey

(earthexplorer.usgs.gov) and Google Earth (Google Earth Pro V 7.3.2) acquired during summer

(leaf-on) between 2000 and 2017 were examined chronologically to assess annual canopy conditions. These images along with the county quarantine dates, allowed us to identify the two to three year period when substantial canopy dieback and ash mortality became apparent in each canopy gap. These periods corresponded to observations of EAB-related mortality by other scientists who had worked in these areas and were also confirmed by site managers.

Survey Design: Light, vegetation and coarse woody debris (CWD) were sampled in gaps and in the surrounding forest using a combination of linear transects and fixed radius plots (Figure 1.2).

A 100 x 2 m linear transect centered on the midpoint and running parallel to the stream, 2-5 m

from the stream edge, was established in each canopy gap. A second, 100 x 2 m linear transect was established in the surrounding forest, 5-10 m from the perimeter of the canopy gap, and running parallel to the streamside transect. Two additional linear transects, each 50 x 2 m, were established at both the upstream and downstream gap-forest transitions, 2-5 m from the stream bank. These transects extended 25 m into the canopy gap and 25 m into the adjacent forest.

I established eight fixed radius plots in each site, including four in the canopy gap and four in the surrounding forest (Figure 1.2). Two plots were positioned in the gap/forest transition zones, both upstream and downstream of the center point. Two additional plots were randomly located in the canopy gap in the area bounded by the two 100 m linear transects. The final two plots were positioned in the surrounding forest, at least 10 m from the perimeter of the gap. Each fixed radius plot was comprised of three concentrically organized plots, including a macroplot

(11.4 m radius), a subplot (8 m radius), and a square microplot (1 m2) (Figure 1.2).

Overstory: I recorded species and DBH of all live and dead overstory trees (DBH ≥10 cm) encountered along transects and in macroplots. Canopy dieback and transparency of live trees were visually assessed in 10% increments, with 10% indicating a nearly full, healthy canopy and

90% indicating a severely declining tree.

Variables recorded in transects and macroplots in gaps and the surrounding forests were standardized per ha. Overstory basal area (m2 ha-1) of live and dead trees was calculated by species and for all species combined. Relative importance values (RIV) were calculated for overstory tree species in gaps, surrounding forests and overall for each site. For a given species, the RIV was calculated by summing its relative frequency (frequency of a species in plots as a proportion of the total frequency of all species), relative density (number of stems of a species as

a proportion of stems of all species) and relative dominance (basal area of a species as a proportion of the total basal area for all species) in the stand (Kent and Coker 2011).

Regeneration: In subplots, saplings were tallied by species and percent cover was estimated for shrub species. I defined saplings as woody stems with a DBH of 2.5 – 10 cm and ≥ 0.5 m tall. In microplots, tree seedlings, (woody stems < 0.5 m tall and <2.5 cm diameter), were tallied by species or genera, as necessary and percent cover of herbaceous plants was recorded.

Coarse Woody Debris: I recorded species (when possible), diameter, decay class, and length of coarse woody debris (CWD) pieces ≥ 7.6 cm in diameter in all linear transects. Decay was classed as (1) bark mostly firm and attached with few or no small areas of loose or sloughed outer bark; and wood solid; (2) bark loose or detached from one or more areas; outer wood beginning to decay with areas that were spongy or broke in response to pressure; (3) bark mostly loose or sloughed off, outer wood spongy and some broken but most inner wood solid; and (4) little or no bark remaining, wood spongy and mostly crumbling. Number, size and volume of

CWD pieces were summed and standardized per ha and by decay class for canopy gaps and adjacent forests. None of the coarse woody debris classified as decay class 3 or 4 (advanced decay) had any evidence of EAB larval galleries or adult exits. This material was therefore assumed to have fallen prior to the onset of EAB-related ash mortality.

PAR: Photosynthetically active radiation (PAR) was recorded in canopy gaps and surrounding forest between 10 AM and 2 PM using light meters (Licor Model LI-250A) held 1.5 m aboveground. Light was measured in each cardinal direction around the perimeter of the macroplots, at 20 m intervals along the two 100 m linear transects, and at 10 m intervals along the two 50 m linear transects (48 recordings per site). Simultaneously recorded PAR measurements were collected in unshaded locations near each site to quantify the percentage of

available PAR that penetrated the overstory. Recordings were taken in late fall or winter (snow absent) to assess light interception by standing trees (live or dead) and in mid-summer when leaves were fully expanded.

Statistical Analysis: I evaluated differences in vegetation between canopy gaps and surrounding forests and among the three watersheds, representing the progression of the EAB invasion, using a two-factor split-plot ANOVA (Proc GLIMMIX; SAS Institute Inc. 2015). For PAR data, a three factor ANOVA (Proc GLIMMIX; SAS institute Inc. 2015) was used to assess variation among watersheds, between canopy gaps and surrounding forest, and between summer (leaf-on) and fall/winter. All variables were checked for normality and heterogeneity of variance using histograms, normal probability plots, and side by side boxplots of residuals. Variables that did not meet normality assumptions, including basal area of snags (standing dead trees), volume of ash coarse woody debris, seedling densities, and the percentage of light reaching the understory, were normalized using a log transformation. When ANOVA results were significant, least square means were separated using a Students t-test (pdiff). Linear relationships between understory herbaceous cover and densities of seedlings, and saplings, as well as coarse woody debris volume since gap formation were evaluated with simple linear regression (Proc Reg; SAS

Institute Inc. 2015).

To evaluate compositional differences between pre-EAB forest and canopy overstories, and to project future overstory compositions of stands, nonmetric multidimensional scaling

(NMS) and multivariate regression trees (MRT) were applied to square root transformed data of overstory tree and current sapling stem densities. NMS were performed using PC-ORD (PC-

ORD version 6.08, MJM Software, Gleneden Beach, OR, US) and MRT was performed using the mvpart package (De’ath 2011) in R version 3.5.1 (R Core Team 2018).

To conduct non-metric multidimensional scaling (NMS) in PC-ORD, I initially used autopilot mode with Sorenson (Bray-Curtis) distance on the “slow and thorough” setting to determine the correct dimensionality for the NMS output. This setting uses a random starting configuration and performs 250 runs of real and randomized data with a maximum of 500 iterations for six axes. Correct dimensionality was established based on the output of the NMS scree plot. The number of interpreted axes was determined by attempting to maximally reduce stress while not distorting the output over too many axes. Non-metric multidimensional scaling ordinations with stress >25 are usually deemed uninterpretable, and an ideal final stress of 5-15 is preferred (Peck 2016). After dimensionality was determined, I again ran the ordination for the set number of axes. Pre-EAB overstories and associated regeneration were connected with successional arrows to indicate changes in species composition in sites. The use of successional arrows in NMS outputs to visualize temporal changes in site community compositions is a useful tool in assessing potential long term impacts of forest changes in forest structure and has been used to compare pre disturbance forest overstories to current regeneration and understory herbaceous communities prior to and following a disturbance (Fornwalt et al. 2018, Kayes and

Tinker 2012). Because seedling density was low, or seedlings were absent in several canopy gaps, NMS outputs for forest overstory-seedling comparisons provided little information and were excluded from further consideration.

To test for significance in communities among watersheds, between overstories and saplings, and between canopy gaps and surrounding forests, PERMANOVA analysis was performed using PC-ORD. When PERMANOVA results were significant, pairwise comparisons of treatments were applied. For grouping variables which had significant PERMANOVA results,

indicator species analysis (ISA) was performed in PC-ORD to determine which tree species were the most common and abundant.

Multivariate regression trees are hierarchical clustering models that group and successively split data by similarity of assemblages into homogenous groupings based on a number of explanatory variables (De’ath 2002). MRT analysis have few underlying assumptions about data distribution and are widely applicable to and often fit ecological assemblage data well

(De’ath 2002). I used MRTs to further explore species composition in canopy gaps and surrounding forests and to compare species present in the pre-EAB overstory with current sapling species. I selected the final tree based on cross validation results by pruning the smallest branched tree in one standard deviation of the minimum cross-validated relative error.

Results

Canopy Gap Formation: Mortality of individual overstory ash trees was apparent by 2006 in aerial images from the Clinton watershed sites and by 2008, the four canopy gaps I surveyed were evident. Progression of ash mortality varied among sites in the Grand River watershed. In three of the four Grand River sites, some dead and declining ash were evident in 2011 and by

2012, nearly all overstory ash had died and canopy gaps were obvious. The fourth Grand River site in Shiawassee County, however, was invaded earlier. Ash were declining in 2006 imagery and most ash trees in the canopy gap I surveyed were dead in 2008. In three sites in the

Kalamazoo River watershed, a few ash trees with substantial dieback were observed in 2012 imagery, but canopy gaps resulting from extensive mortality of the overstory ash were not apparent until 2015. In one Kalamazoo site, I noted some declining ash as early as 2010 and all overstory ash trees in this site were dead in 2012.

Overstory Vegetation: I tallied 2037 overstory trees (DBH ≥ 10 cm) representing 38 species in the 12 sites. This included 683 trees representing 27 species in the canopy gaps and 1354 trees representing 34 species in the forests surrounding the gaps. Overall, 77.5% of the trees were alive (1579 trees), while the rest (458 trees) were snags (i.e., standing dead trees). I recorded 410 and 1169 live trees in canopy gaps and surrounding forests, respectively. Species richness was lower in canopy gaps than in surrounding forests (F=15.64; df=1,9; P=0.0033) but did not differ among the three watersheds (F=1.39; df=2,9; P=0.3) (Table 1.1). More snags were present in canopy gaps (273 dead trees) than surrounding forests (185 dead trees) (F=5.41; df=1,18;

P=0.032), but snag abundance was similar among the three watersheds (F=1.12; df=1,9;

P=0.37). Densities of all trees (live and dead) did not differ among the three watersheds (F=2.09; df=2,9; P=0.18), averaging (± SE) 364.7 ± 45.4, 414.5 ± 29.3 and 474.0 ± 21.8 trees per ha for

the Clinton River, Grand River and Kalamazoo River watersheds, respectively. Total tree density in canopy gaps averaged 278.1 ± 24.1 stems per ha and was lower than tree density in surrounding forest, which averaged 557.4 ± 37.3 stems per ha (F=40.95; df=1,9; P=0.0001).

Similarly, total basal area of all overstory species (live and dead trees) did not differ among the three watersheds (F=0.72; df=2,9; P=0.51) (Table 1.1). Basal area of live trees was also similar among watersheds, (F = 0.19; df = 2,9; P=0.82), with live trees accounting for an average of 77.9 ± 8.1%, 70.5 ± 5.3% and 83.1 ± 2.4% of the total basal area in sites in the

Clinton, Grand and Kalamazoo watersheds, respectively (Table 1.1). Not surprisingly, total basal area in canopy gaps, which averaged 16.1 ± 1.5 m2.ha-1, was lower than in surrounding forests, which averaged 36.8 ± 2.9 m2.ha-1 (F=38.89; df=1,9; P=0.0002). Similarly, live basal area was lower in canopy gaps, which averaged 9.9 ± 1.5 m2.ha-1, than in surrounding forest, which averaged 33.8 ± 2.7 m2.ha-1 (F=62.30; df=1,9; P<0.0001). Additionally, there were no differences in total basal area among watersheds for gaps (F=2.45; df=2,9; P=0.14) nor for forests surrounding the gaps (F=0.73; df=2,9; P=0.51) when analyzed separately.

While overstory species richness and species composition varied among sites in the three watersheds for both canopy gaps and surrounding forests, several species were consistently present regardless of watershed. In canopy gaps, American elm (Ulmus americana L.) and

American basswood (Tilia americana L.) ranked in the top five species in relative importance values (RIV) for all watersheds (Table 1.2). Black walnut (Juglans nigra L.) was common in gaps in Clinton and Grand River sites, and swamp white oak (Quercus bicolor Willd.) was commonly encountered in Clinton and Kalamazoo river sites (Table 1.2). In surrounding forests, black cherry (Prunus serotina Ehrh.) and northern red oak (Quercus rubra L.) had high RIVs for all watersheds. Red maple (Acer rubrum L.) ranked in the top three in species RIV in forests for

both the Clinton and Kalamazoo watersheds (Table 1.2). In watersheds, some species were relatively dominant in both canopy gaps and surrounding forests. For example, swamp white oak and American basswood had relatively high RIV in canopy gaps as well as surrounding forests in the Clinton River watershed, American elm and sugar maple (Acer saccharum Marsh.) had high

RIVs in the Grand River watershed, and red maple, swamp white oak and northern red oak had high RIVs in gaps and forests in the Kalamazoo River watershed. While maple and oak species were generally common in all sites, high RIV’s for maple predominantly reflected high stem densities whereas high oak RIV’s represented a few large trees that dominated basal area estimates.

Overstory Ash Condition: Overall, 328 standing ash (Fraxinus) trees (> 10 cm DBH), including

44 live ash trees, were recorded. Relative importance values for live ash species were uniformly low (Table 1.2). At least one live ash was present in 11 of the 12 sites; no live ash trees were encountered in one site in the Grand River watershed (Table 1.3).

Ash consistently ranked among the top three overstory species in stem density and basal area in all three watersheds when live and dead standing trees were summed. Estimates of the pre-EAB ash overstory component, comprised of live ash, standing dead ash, and relatively fresh ash CWD with visible EAB galleries, showed that ash trees accounted for a large percentage of overstory stems and basal area in the current canopy gaps in all three watersheds, but were uncommon in the forests surrounding canopy gaps (Table 1.3). More than 75% of the ash basal area was dead in canopy gaps and forests in all sites and the proportion of ash basal area killed by EAB did not differ among watersheds (F=0.39; df=2,9; P=0.69) nor between canopy gaps and adjacent forests (F=1.72; df=1,9; P=0.22) (Table 1.3).

Overall, ash comprised 61.7% (284 trees) of the standing dead trees recorded across sites.

Ash represented a higher percentage of dead trees in canopy caps than in the surrounding forests

(F=15.09; df=1,18; P=0.0011) (Table 1.3). In canopy gaps, the percentage of standing dead trees that were ash differed significantly among the three watersheds (F=16.75; df=2,9; P=0.001). A lower percentage of dead trees in the Clinton watershed were ash (45%) compared to gaps in the

Grand (89.8%) (t=5.68; df=9; P=0.0003) and Kalamazoo watersheds (75%) (t=-3.79; df=9;

P=0.004), which did not differ from each other. To estimate the proportion of dead ash trees still standing, basal area of dead ash trees was added to the basal area of the relatively fresh, non- deteriorated ash CWD (decay classes 1-2). The percentage of ash trees still standing was significantly affected by overstory structure (F=6.96; df=1,16; P=0.018), watershed (F=4.10; df=2,16; P=0.037) and the interaction of the two factors (F=3.75; df=2,16; P=0.046). Slicing the interaction by the watershed factor showed that the percentage of ash still standing differed between canopy gaps and surrounding forest only in sites in the Clinton River watershed

(F=13.28; df=1,16; P=0.002), where lower proportions of dead ash remained standing in canopy gaps than in forests. Slicing by the overstory structure factor showed that the percentage of dead ash trees still standing in canopy gaps varied among watersheds (F=7.05; df=2,16; P=0.006), but was similar in the forested areas. In canopy gaps, a significantly lower percentage of ash trees were still standing in the Clinton watershed gaps (36.0 ± 9.0%) than in gaps in the Grand (68.2 ±

18%) (t=-3.68; df=16; P=0.002) and Kalamazoo watersheds (50.0 ± 12.0%) (t=-2.48; df=16;

P=0.025), which did not differ from each other.

Coarse Woody Debris: Coarse woody debris (CWD) was recorded in all 12 sites. I tallied 554 pieces of CWD with diameters ranging from 7.6 cm to 50 cm and averaging (± SE) 15.2 ± 0.51 cm. Diameter of CWD did not differ among watersheds (F=0.09; df=2,9; P=0.92) or between

canopy gaps and surrounding forests (F=0.09; df=1,9; P=0.77). Volume of CWD averaged 70.0

± 8.49, 31.6 ± 7.10 and 33.7 ± 8.83 m3. ha-1 in the Clinton, Grand, and Kalamazoo River watersheds, respectively, and averaged 44.8 ±7.75 and 33.2 ± 6.98 m3 ha-1 in the canopy gaps and the surrounding forests, respectively. Main effects of watershed (F=4.17; df=2,9; P=0.052) and overstory structure (gap versus forest) did not affect total CWD volume (F=3.8; df=1,9;

P=0.08). There was a significant interaction between the two factors (F= 4.89; df=2,9;

P=0.037), reflecting higher CWD volume in canopy gaps than in surrounding forests in the

Clinton watershed but not in the Grand or Kalamazoo watersheds (Figure 1.3). A contrast to compare CWD volume between the Clinton watershed and the grouped Grand and Kalamazoo watersheds indicated CWD volume in the Clinton watershed sites in the southeast was significantly higher than in the more recently invaded sites (t = 2.88; DF = 2,9; P = 0.023).

Additionally, simple linear regression showed a strong positive relationship between CWD volume in canopy gaps and the time since gap formation (R2=0.698) (Figure 1.4). When basal area of snags (all standing dead overstory trees) and relatively fresh coarse woody debris (CWD in decay classes 1-2) were added to that of live trees, total basal area in canopy gaps was similar to that in forested areas (F=3.37; df=1,9; P=0.081).

I could identify 54.5% of the total CWD volume to species. At least 23 tree species were represented by this CWD but advanced decomposition precluded identification of the remaining logs. The proportion of CWD volume identified as ash ranged from 21 to 32% and did not differ among watersheds (F=0.57; df=2,9; P=0.58).

Proportion of CWD represented by relatively fresh ash logs varied between canopy gaps and adjacent forests. Across all sites, an average (± SE) of 12 ± 3%, 25 ± 4%, 35 ± 5% and 28 ±

6% of the CWD volume was assigned to decay classes 1-4, respectively. In canopy gaps, ash

comprised a significantly higher proportion of the total CWD in early and intermediate stages of decay (classes 1-2) than any other species. Ash accounted for 44.8% of the fresh CWD in canopy gaps compared to 11% in the surrounding forests (F=13.24; df=1,9; P=0.005). Volumes of CWD in classes 3 and 4, which presumably fell before EAB invaded the sites, were pooled for analysis.

I used 2-factor ANOVA to assess proportions of CWD in early (stage 1), intermediate (stage 2) and advanced (stages 3 and 4) stages of decay among the three watersheds and between canopy gaps and surrounding forests. No significant differences were detected among watersheds, but canopy gaps had a higher proportion of CWD in the intermediate decay class (F=6.25; df=1,9;

P=0.034) and a lower proportion of CWD in advanced stages of decay (F=5.31; df=1,9;

P=0.047) than surrounding forests.

To further evaluate CWD distribution, I used one-way ANOVA to assess whether the proportion of CWD volume in the three decay classes in canopy gaps and in forests differed among the watersheds. In forested areas, there was no difference for decay class 1 (F=0.35; df=2,9; P=0.72), 2 (F=0.27; df=2,0; P=0.77) or classes 3-4 (F=0.02; df=2,9; P=0.98). In canopy gaps, however, a significantly higher proportion of CWD volume was fresh (decay class 1) in the

Clinton watershed sites in the southeast than in sites in the Grand and Kalamazoo watersheds

(F=4.67; df=2,9; P=0.041). Watersheds did not affect volume of CWD that was assigned to intermediate (class 2) (F=0.99; df=2,9; P=0.41) or advanced decay classes (F=3.81; df=2,9;

P=0.063) in canopy gaps.

PAR: Analysis of PAR (photosynthetically active radiation) indicated that the proportion of full sun reaching the understory (measured 1.5 m above ground) differed between canopy gaps and surrounding forest (F=118.21; df=1,9; P<0.0001), and between winter and summer measurements (F=92.82; df=1,18; P<0.0001) but did not differ among watersheds during either

winter (F=0.69; df=2,9; P=0.526) or summer months (F=0.91; df=2,9; P=0.438). The interaction between time of measurement (summer/winter) and overstory structure (canopy gap/intact forest) also was significant (F=68.24; df=1,18; P<0.0001), with similar measurements of PAR being recorded in canopy gaps during winter (65.9 ± 4%) and summer (62.4 ± 8%) (F=0.15; df=1,9;

P=0.708) but higher PAR values were recorded in winter (50.0 ± 3%) than in summer (7.4 ± 1%) in forested areas (F=135.95; df=1,9; P<0.0001). Additionally, during winter months PAR recordings taken in canopy gaps (65.9 ± 4%) and forested areas (50.0 ± 3%) were similar

(F=5.08; df=1,9; P=0.052).

Saplings: Saplings (2.5-10 cm diameter and ≥ 0.5 m tall) represented established regeneration above the herbaceous line. I recorded 1253 saplings representing 23 tree species across the 12 sites. Densities did not differ between canopy gaps and surrounding forests (F=3.06; df=1,9;

P=0.11) or among watersheds (F=0.55; df=2,9; P=0.59). Ash accounted for three of the five most abundant species of this regeneration stratum (Table 1.5), was abundant in all sites and comprised 40.7% of the total saplings recorded. Ash were especially abundant in canopy gaps, where they accounted for 66.3% of the regeneration. Green ash was the most abundant species in canopy gaps and the second most abundant species in surrounding forests. Red maple was most abundant in forested areas and was especially abundant in the Kalamazoo River watershed, where it was the most abundant species in the forested areas. Black cherry was present in all watersheds but was only common in sites in the Clinton watershed. In contrast to the seedling stratum, there was no linear relationship between densities of saplings and herbaceous plant cover (R2=0.08) (Figure 1.5 B).

Seedlings: I recorded 414 tree seedlings representing 17 species in the microplots. Both seedling density (F=17.44; df=1,9; P=0.002) and species richness (F=40.33; df=1,9; P<0.0001) were

lower in canopy gaps than in adjacent forests, but differences among watersheds did not affect density (F=1.06; df=2,9; P=0.39) nor richness (F=0.49; df=2,9; P=0.63). Densities of seedlings and understory herbaceous plants were inversely related (R2=0.67) (Figure 1.5 A), and in two canopy gaps where herbaceous plant percent cover exceeded 95%, no seedlings were encountered. Black cherry was the most abundant seedling species in both canopy gaps and forested areas (Table 1.4). Sugar maple seedlings were the second most abundant species overall, but were only encountered in forests in the Grand River watershed, where they occurred in high densities. Green ash seedlings were also abundant and were recorded in all watersheds, but densities in canopy gaps were lower than in surrounding forests (F=7.34; df=1,9; P=0.035).

Green ash comprised only 17.5% of seedlings tallied in canopy gaps despite the original abundance of ash in the overstory in these areas prior to the EAB invasion. Only three black ash seedlings were recorded in one site in the Clinton River watershed; two were in the canopy gap and one was in the adjacent forest. I did not observe any white ash seedlings in any sites.

Herbaceous plants: Percent cover of herbaceous plant species were similar among the three watersheds (F=3.08; df=2,9; P=0.09) but densities were much higher in canopy gaps (82.3 ±

4.6%) than in surrounding forests (23.4 ± 3.4%) (F=263.5; df=1,9; P<0.0001). In canopy gaps, the most prevalent herbaceous plants within all watersheds were sedge species (Carex spp.), primarily tussock sedge (Carex stricta Lam.), retrorse sedge (Carex retrorsa Schwein.) and awl fruited sedge (Carex stipata Muhl.), which collectively comprised >75% of the herbaceous plant cover in the gaps (Table 1.6). Other minor species encountered in canopy gaps included goldenrod (Solidago spp.), stinging nettle (Urtica dioica L.) and eastern skunk cabbage

(Symplocarpus foetidus Salisb.) (Table 1.6). These species were present in gaps of all watersheds but none represented more than 10% of the herbaceous cover in any gap. In forests surrounding

canopy gaps, herbaceous plant densities were uniformly low. Species commonly encountered in plots in the forests included fowl mannagrass (Glyceria striata (Lam.) Hitchc.), cinnamon fern,

(Osmunda cinnamomeum L.), poison ivy (Toxicodendron radicans (L.) Kuntze), smooth sweet- cicely (Osmorhiza longistylis (Torr.) DC.), common jewelweed (Impatiens capensis Meerb.), and trillium (Trillium grandiflorum (Michx.) Salisb.) (Table 1.6). However, no individual species of herbaceous plants comprised more than 25% cover of any forested plot, and no forested plot contained more than 50% herbaceous cover.

Invasive Plant Species: A total of 12 invasive plant species were encountered overall and invasive plants were tallied in 11 sites, including 11 of the 12 canopy gaps and nine adjacent forests. No invasive plant species were recorded in either the gap or the surrounding forest in one site in the Kalamazoo River watershed. Overall species richness of invasive plants averaged

3.75 ± 0.85, 3.25 ± 0.75 and 1.75 ± 0.63 for the Clinton, Grand and Kalamazoo watersheds, respectively, and 2.17 ± 0.46 and 1.67 ± 0.43 for gaps and surrounding forests, respectively.

Total density of invasive plant species did not differ among watersheds (F=0.31; df=2,9;

P=0.74) nor between canopy gaps and forests (F=0.65; df=1,9; P=0.65). Similarly, species richness was similar among watersheds (F=0.34; df=2,9; P=0.72) and between gaps and surrounding forests (F=0.67; df=1,9; P=0.44). The most frequently encountered invasive plants were multiflora rose (Rosa multiflora Thunb.) and common buckthorn (Rhamnus cathartica L.), which were both encountered in a total of six sites and in all three watersheds. Distribution of individual invasive plant species was patchy and varied considerably. Percent cover or density of the four most abundant invasive plant species did not differ between gaps and forested areas or among watersheds (Table 1.7). Other undesirable invasive plant species detected in sites were bull thistle (Cirsium vulgare (Savi) Ten.), Chinese yam (Dioscorea polystacha Turcz.), garlic

mustard (Alliaria petiolata (M. Bieb.) Cavara and Grande), Japanese barberry (Berberis thunbergii DC.), oriental bittersweet (Celastrus orbiculatus Thunb.), and sweet autumn clematis

(Clematis terniflora DC.). While these species were detected they only occurred in 1-4 sites, were not a dominant species in any site, and did not appear to be correlated with the EAB invasion. Additionally, while presence of invasive plant species did not seem to be affected by the EAB invasion, in all sites where an invasive species occurred it was found in EAB caused canopy gaps but were not always found in the surrounding forest.

Species Composition of Pre-EAB Overstories and Current Saplings: To visually assess and interpret the output from the non-metric multidimensional scaling (NMS), I outlined sites in canopy gaps and surrounding forests with polygons, and added arrows to indicate successional associations between species composition of pre-EAB overstory and current saplings.

PERMANOVA analysis suggested that while assemblages did not differ among the three sampled watersheds, communities of species did differ between canopy gaps and surrounding forests (Table 1.8). Pairwise comparisons of PERMANOVA and visual assessment of NMS successional arrows between overstories and regeneration in sites revealed that canopy gap overstories were similar to communities found at the sapling stratum of gaps (Table 1.8) (Figure

1.6). Results of ISA suggested that ash was the dominant species in canopy gaps prior to the

EAB invasion and is still both highly abundant and the dominant species in the gap sapling stratum currently (Table 1.9). Conversely, forests surrounding canopy gaps were highly variable and forest overstory compositions were not similar to the communities of saplings in either forested areas or canopy gaps (Table 1.8) (Figure 1.6). A number of species were significant indicators of forests surrounding canopy gaps, with the strongest indicators being black cherry, red maple and northern red oak (Table 1.9).

Cross validation of 1000 multivariate regression trees based on species composition of pre-EAB overstories and corresponding saplings in the three watersheds yielded a 3-branched tree most often (748) (Figure 1.7). The first split occurred between canopy gaps and surrounding forests, and the second split occurred between pre-EAB overstory and sapling species in forests surrounding canopy gaps (Figure 1.7). Similar to results from NMS and ISA, MRT results suggested that ash was the dominant species in the overstory and the sapling stratum in canopy gaps. The MRT further refined the results from NMS analysis by showing that variability in forests occurred mainly between the pre-EAB overstory and the current saplings. Dominant species in the overstory of forests surrounding canopy gaps included northern red oak, followed by red maple and black cherry. Saplings in these forests were characterized by abundant ash, black cherry and red maple but northern red oak was relatively rare (Figure 1.7).

Discussion

In the riparian forests I evaluated, the current canopy gaps were dominated primarily by green ash, and in most cases, some amount of black ash, prior to the EAB invasion. Both green ash and black ash are highly preferred and vulnerable EAB hosts (Anulewicz et al. 2007, Tanis and McCullough 2015, Robinett and McCullough 2019) and both species are frequently abundant in mesic sites with poorly drained soils (Gucker 2005a, 2005b, Kennedy 1990,

Youngquist et al. 2017). Sites that were evaluated generally represented the temporal gradient of the EAB invasion for southeast to southwest Michigan. Canopy gaps in the Clinton River watershed in southeast Michigan, near the epicenter of the North American EAB invasion

(Siegert et al. 2014), were apparent by 2008. In sites in the central Grand River watershed and the southwestern Kalamazoo River watershed, gaps became obvious by 2012 and 2015, respectively. This timeline mirrors that previously reported by Burr and McCullough (Burr and

McCullough 2014, Burr et al. 2018), who recorded nearly 80% mortality of green ash basal area in forests in southeast Michigan in 2011, with progressively lower levels of mortality in forested sites in south central (~40%) and southwestern areas (~20%). Impacts of EAB in the riparian sites I surveyed have been substantial; 85-95% of the ash basal area has been killed by EAB and ash has been effectively lost as a functional overstory tree in the riparian forests across southern

Michigan. My data are consistent with other reports indicating extensive ash mortality in a local area occurs in four to six years after external signs of EAB infestation are recognized (Knight et al. 2013, McCullough and Mercader 2012, Poland and McCullough 2006, Smith et al 2015).

While high rates of ash mortality extend across southern Michigan, I observed several differences in forests across the temporal gradient of the EAB invasion. The most notable difference was in the status of dead ash trees and coarse woody debris (CWD). In southeastern

sites in the Clinton River watershed, all of which are within 54 to 62 km of the epicenter of the

EAB invasion in North America (Siegert et al. 2014), most ash snags had fallen, resulting in large influxes of relatively fresh CWD in the canopy gaps. These fallen and broken ash trees had abundant EAB galleries and exits and were clearly killed by EAB. In contrast, most dead ash trees remain standing in sites in the Grand River and Kalamazoo River watersheds. In canopy gaps in these more recently invaded watersheds, CWD was mostly decaying material that pre- dated the EAB invasion. While CWD is substantially lower in nutrients than leaf litter, it increases understory heterogeneity (Harmon et al. 1986), can aid in seedling establishment

(Beach and Halpern 2001), and provides habitat and nutrients for a wide array of plants and animals (Gandhi and Herms 2010, Harmon et al. 1986). A large pulse of CWD from dead ash trees will presumably affect plant and animal communities in these sites. An influx of CWD, similar to what has occurred in the southeast sites, will undoubtedly occur in the coming years as dead ash break and fall in south central and southwest sites. Future research to track the corresponding response of invertebrate and vertebrate species in these communities as ash trees continue to fall and decay would be valuable.

Although the decay status of dead ash was perhaps not surprising, these results illustrate several unforeseen effects of EAB on riparian forests. Despite the approximate 10-year progression of ash mortality across southern Michigan, regeneration was notably similar among the sampled watersheds. Ash seedlings were rare in all sites, especially in canopy gaps, where ash previously dominated the overstory. Ash seeds are not known to persist long in the seed bank

(Kashian and Witter 2011, Klooster et al. 2014) and previous studies have reported the near complete absence of newly germinated seedlings once overstory trees die (Burr and McCullough

2014, Kashian and Witter 2011, Klooster et al. 2014). Green ash is known to produce seed

relatively early in its development with trees as small as 3-4 inches (8-10 cm) in diameter observed producing seeds (Gucker 2005 B, Kashian 2016). Conversely Black ash requires 30- 40 years to reach seed bearing age, and even when mature produce seeds sporadically, with intervals of five or more years often occurring between abundant seed years (Gucker 2005 A). While this might result in differences in regeneration dynamics among ash species, the rarity of mature ash trees in gaps and surrounding forests indicates seed sources are now uncommon and the likelihood that new ash seedlings will be become established is low. This is further exacerbated by the apparent exclusion of seedlings in canopy gaps by competing sedges and vegetation.

Persistence of ash in riparian forests appears to be limited to saplings, i.e. young trees that are 2.5 to 4 cm diameter and recruits, previously termed the “orphaned cohort” of ash regeneration

(Herms and McCullough 2014, Klooster 2012, Klooster et al. 2018). Multivariate (NMS and

MRT) analysis comparing pre-EAB forest overstories and current saplings showed that species composition of saplings in canopy gaps is similar to that of the overstory before the EAB invasion. In other words, ash dominates the sapling stratum in gaps much as ash dominated the pre-EAB overstory in the gaps. This suggests that if the ash saplings in canopy gaps survive long enough to be recruited into the overstory, resulting forests could closely mirror those that existed prior to the EAB invasion. I estimated total phloem area of pre-EAB ash trees following methods of McCullough and Siegert (2007), and determined that among all gaps, an average of 96.2 ±

1.3% of the overstory (DBH >10 cm) ash phloem area has died, resulting in an almost complete lack of suitable material for EAB larval development. This radically reduces the carrying capacity for EAB and densities may be low enough for young ash regeneration to persist and mature. Parasitoids introduced for classical biological control of EAB may be mostly successful at locating and parasitizing EAB larvae in small trees (Duan et al. 2017). Healthy ash trees are

relatively resilient and can tolerate low densities of EAB larvae without declining (Anulewicz et al. 2007, Herms and McCullough 2014, McCullough et al. 2019). Furthermore, while ash mortality rate was high in all areas, some overstory ash did remain. I saw between 12-17 live mature trees of black, green, and white ash across the total 4.64 ha I surveyed. This remnant may fit the description of “lingering ash” (Knight et al. 2014) and may demonstrate some level of resistance to EAB. If introduced and native natural enemies of EAB can collectively regulate

EAB populations at low densities, ash regeneration may persist, mature and eventually re- establish ash in the seed bank (Klooster et al. 2018).

Water is rarely a limiting resource in riparian forests, but increased light availability resulting from relatively sudden openings in the forest canopy provides an important resource which may initially reduce competition for light in the understory (Davis et al. 1998) and may shift regeneration dynamics to favor early successional disturbance-adapted species. Light availability during the growing season was ~55% higher in canopy gaps than in the surrounding forests.

While many species of ash will successfully colonize disturbed habitats (Gucker 2005), relatively rapid increases in light as ash trees decline and succumb to EAB presumably favors both establishment and rapid growth of light and disturbance adapted plant species. In the canopy gaps I surveyed, wetland adapted sedges (Carex L.) have taken advantage of this resource and many of these previously ash-dominated forests are transitioning to sedge meadows below the sapling stratum of the understory. These dense sedge meadows appear to limit establishment of woody seedlings, to the extent that I detected no tree seedlings in three of the 12 gaps. This pattern is similar to observations reported by Abrams et al. (1985), who found disturbances in jack pine forests in northern lower Michigan led to the formation of sedge meadows in the forest understory. They labelled this phenomenon as “regressive succession” and

commented that the apparent stability of the sedge meadows may arrest or inhibit further succession, effectively converting jack pine stands to sedge meadows with open canopies.

Similarly, in black ash swamps of Minnesota and upper Michigan, simulated EAB mortality from girdling all black ash trees in a stand resulted in increased area and height of wetland herbaceous plants (Kolka et al. 2018). This could similarly represent the future for previously ash-dominated riparian forests, particularly if EAB populations rebound and kill most of the regenerating ash cohort as it matures. Continued monitoring of regeneration in these forests is needed to determine the likelihood of ash seedling establishment.

While invasive plant species were present in almost all sites, I found no evidence to indicate ash mortality and subsequent canopy gaps were facilitating the establishment or spread of any invasives, or that their abundance reflected the time since EAB invasion. These results would suggest that there were not substantial positive interactions between EAB and other invasive animal or plant species that could result in an “invasional meltdown”. While invasive plants such as multiflora rose (Rosa multiflora Thunb.) were abundant in some areas, distributions were relatively patchy in all sites and invasives never dominated the understory of gaps nor surrounding forests. I noted, however, that in sites where invasive plants occurred, they were consistently present in EAB-induced canopy gaps but were not necessarily present in the forests bordering these gaps. These findings are in contrast to those by Hoven et al. (2017), and

Hausman et al. (2010), who reported a positive growth response of invasive plant basal area and percent cover in response to ash mortality. Their study sites, however, were located in upland hardwood forests rather than in riparian forests and competitive interactions among understory species could easily differ. Perhaps more importantly, Hausman et al. (2010) felled all ash trees down > 2.4 cm DBH in selected plots then compared vegetation unmanipulated plots. Ash

mortality from EAB occurs more gradually and it seems unlikely that the tree felling generated the same conditions plants would experience in forests invaded by EAB. Hausman et al. (2010) additionally concluded that their findings may have been at least partially confounded by soil compaction resulting from the tree removal process in cut plots. It is unclear why I did not see invasive plant species such as amur honeysuckle, oriental bittersweet and multiflora rose, all opportunistic vine and shrub species known to dominate disturbed habitats, respond more strongly to the formation of EAB-caused canopy gaps. I suspect that wetland adapted sedges responded to the pulse of light more quickly and were able to outcompete most other plant species, including invasives, as the gaps formed. This is consistent with data showing seedling exclusion in canopy gaps and supports the resilience and stability of these sedge meadows, as suggested by Abrams et al. (1985). These findings raise questions about competitive interactions among species in riparian forests and may require further research to fully understand.

Nearly complete mortality of overstory ash trees and the loss of the annual influx of ash leaf litter will likely have adverse effects on nutrient availability in riparian forests as well as in adjacent streams. Ash leaves are considered to be high quality leaf litter, have efficient turnover rates and contribute important soil nutrients such as nitrogen, organic carbon and exchangeable cations (Mg2+,Ca2+) (Finzi et al. 1998, Langenbruch et al. 2012, Melillo et al. 1982). In riparian forests where ash formerly contributed a large proportion of the annual leaf litter, these changes could further alter regeneration and competitive interactions among shrubs and herbaceous vegetation, including invasive species in the understory.

In the forests surrounding canopy gaps, ash was not a dominant species before the EAB invasion but nearly all the overstory ash that were present are now dead. Forest overstories surrounding the canopy gaps are largely dominated by red oak, black cherry and maple species.

Lateral ingrowth from surrounding trees has closed any gaps resulting from mortality of scattered ash, a response previously noted by Burr and McCullough (2014). Nevertheless, ash saplings were still abundant in the forests in all watersheds. Ash seedlings are relatively shade tolerant but become less so as they grow into saplings and access to full or nearly full light is necessary for trees to reach the overstory (Gucker 2005). The low light environment in these forests is unlikely to favor recruitment of the remnant ash regeneration into the forest canopy.

Rather, compositional changes in forest sapling assemblages from MRT analysis suggest many of these sites may be transitioning from red oak dominated forests to red maple and black cherry dominated forests, a pattern similarly reported in other eastern hardwood forests (Abrams 1992,

Lorimer 1984, Olson et al. 2014). A proposed reason for this shift is forest mesophication resulting from the rapid reduction of fire frequency during the 20th century, leading to higher density in forest stands, decreasing light availability in forest understories and increasing soil moisture content (Nowacki and Abrams 2002). Consequently, recruitment rates of shade adapted, fire-sensitive species such as black cherry and red maple are increasing, while fire adapted, shade-intolerant species such as oak and hickory are less successful. Additionally, red maple leaves have less lignin than oak leaves and decay rapidly (Lovett 2004), reducing fuel loads in forest understories and feeding into a positive feedback loop by further inhibiting fire

(Nowacki and Abrams 2002). Long term monitoring of stand dynamics in riparian forests will be needed to evaluate whether forest mesophication is occurring in these forests.

Stream changes in response to loss of ash in riparian forests could parallel those seen in forests invaded by the invasive insect hemlock woolly adelgid (HWA) (Adelges tsugae Annand), where high levels of eastern hemlock (Tsuga canadensis (L.) Carr.) mortality have resulted in similarly extensive canopy openings in riparian corridors. Large scale mortality of eastern

hemlock has altered stream temperatures, chemistry, light availability and nutrient poor hemlock litter was replaced by higher quality hardwood litter (Huddleston 2011,), with In the riparian forests I surveyed, loss of high quality, nutrient rich ash litter nutrient (Finzi et al. 1998,

Langenbruch et al. 2012, Melillo et al. 1982) could have opposite effects on aquatic macroinvertebrate communities. Such impacts could either be compounded or diluted as stream size and buffering capacity increase. The frequency and extent of ash canopy gaps along other first order streams as well as larger rivers and associated effects on aquatic systems warrant further investigation.

APPENDIX

Watershed Overstory structure

Basal Area (m2.ha-1) Clinton Grand Kalamazoo P value Canopy gap Forest P value

Live trees 21.3 ± 3.7 23.2 ± 1.1 21.1 ± 1.6 0.82 9.9 ± 1.5 33.8 ± 2.7 < 0.0001

Live ash 0.5 ± 0.2 0.5 ± 0.3 0.2 ± 0.1 0.46 0.5 ± 0.2 0.3 ± 0.1 0.252

Dead trees 4.8 ± 1.8 5.8 ± 1.2 3.0 ± 0.6 0.3 6.2 ± 0.9 2.9 ± 0.6 0.033

Dead ash 2.5 ± 1.2 5.0 ± 1.4 1.7 ± 0.3 0.2 4.7 ± 1.4 1.4 ± 0.4 0.013

Species Richness 9.2 ± 0.25 10.1 ± 0.52 10.8 ± 0.32 0.30 8.6 ± 0.5 11.5 ± 0.6 0.003

Coarse woody debris volume (m3 ha-1) Total logs 70.0 ± 8.49 31.6 ± 7.10 33.7 ± 8.83 0.052 57.0 ± 14.0 33.2 ± 6.98 0.083

Ash logs 34.8 ± 12.14 8.7 ± 7.24 6.8 ± 1.03 0.09 31.3 ± 11.37 2.2 ± 0.63 0.011

Table 1.2. Number of overstory trees (DBH > 10 cm), mean (± SE) diameter at breast height (DBH), basal area, and relative importance values (RIVs) for the five most dominant overstory species recorded in four canopy gaps and surrounding forests in each of three watersheds (n=12 sites).

Canopy Gaps Surrounding Forests No. DBH Basal area RIV1 No. trees DBH Basal area RIV1 trees (cm) (m2·ha-1) (cm) (m2∙ha-1) Clinton River Clinton River Quercus bicolor 29 24 ± 4 3.9 ± 3.1 62.5 Prunus serotina 52 28 ± 7 3.6 ± 0.9 47.8 Ulmus americana 35 16 ± 4 1.04 ± 0.4 52.7 Acer rubrum 65 22 ± 2 3.04 ± 1.7 36.3 Tilia americana 10 29 ± 7 2.25 ± 1.7 38.3 Quercus rubra 26 41 ± 9 8.2 ± 4.8 34.4 Juglans nigra 11 36 ± 3 2.11 ± 0.9 27.6 Tilia americana 35 27 ± 4 5.58 ± 3.4 33.7 Acer negundo 7 16 ± 5 1.34 ± 0.5 23.2 Quercus bicolor 28 31 ± 7 4.08 ± 3.2 23.8 Grand River Grand River Ulmus americana 41 16 ± 2 1.4 ± 0.4 90.4 Quercus rubra 70 38 ± 3 12.55 ± 2.6 61.3 Acer saccharum 9 38 ± 8 1.3 ± 0.5 76.3 Acer saccharum 86 29 ± 6 10.11 ± 4.2 52.1 Juglans nigra 21 13 ± 11 1.69 ± 1.5 40.2 Fagus grandifolia 14 43 ± 7 6.47 ± 0.8 26.7 Tilia americana 6 11 ± 7 0.43 ± 0.2 14.5 Prunus serotina 30 25 ± 5 3.13 ± 2.6 25.1 Celtis occidentalis 6 27 ± 18 0.3 ± 0.3 12.2 Ulmus americana 30 16 ± 1 0.96 ± 0.2 21.9 Kalamazoo River Kalamazoo River Forest Ulmus americana 43 18 ± 2 1.54 ± 0.7 63.9 Acer rubrum 163 21 ± 2 9.38 ± 3.1 89.8 Acer rubrum 33 19 ± 4 1.64 ± 0.9 45.5 Quercus rubra 68 33 ± 4 7.93 ± 2.6 57.2 Tilia americana 29 18 ± 1 1.13 ± 0.2 44.8 Populus grandidentata 53 30 ± 4 4.31 ± 2.4 33.4 Quercus bicolor 16 29 ± 5 2.66 ± 0.5 39.1 Prunus serotina 40 23 ± 3 3.12 ± 0.8 28.4 Quercus rubra 10 33 ± 12 1.39 ± 0.6 22.1 Quercus bicolor 17 25 ± 4 1.69 ± 0.3 12.2 1Relative importance values (RIVs) represent the sum of the frequency, density and dominance of a specific species relative to all other species.

Watersheds Percent of Pre- Percent of Mortality of ash F. nigra F. pennsylvanica F. americana EAB BA (%) Current BA (%) basal area (%) Live Dead Live Dead Live Dead Clinton 36.2 ± 8.9 5.9 ± 3.2 86.8 ± 10.2 9 44 8 35 11 2 Canopy gap 61.6 ± 15.5 9.2 ± 6.7 97.1 ± 0.8 5 33 5 18 3 1 Forest 10.9 ± 2.3 2.5 ± 2.0 76.6 ± 21.0 4 11 3 17 8 1 Grand 32.8 ± 5.4 3.3 ± 2.0 87.6 ± 8.3 1 22 4 76 6 11 Canopy gap 59.7 ± 12.7 5.5 ± 3.1 92.6 ± 3.9 0 16 3 60 2 3

Forest 6.0 ± 2.0 1.2 ± 1.0 82.5 ± 14.1 1 6 1 16 4 8 Kalamazoo 25.4 ± 1.7 1.5 ± 0.7 95.6 ± 2.6 4 62 1 28 2 0 Canopy gap 45.9 ± 5.0 2.8 ± 1.5 96.7 ± 1.4 4 56 1 20 0 0

Forest 4.8 ± 1.7 0.2 ± 0.2 94.4 ± 5.6 0 6 0 8 2 0 1Percentage of the total forest basal area comprised of a species

Table 1.4. Mean density of seedling species (no. per m2) tallied in canopy gaps and surrounding forest in three watersheds in Michigan. Species are presented in descending order of overall density.

Watershed Clinton Grand Kalamazoo Total Species Gap Forest Gap Forest Gap Forest Recorded Prunus serotina 0.2 2.8 0.0 2.1 1.6 1.8 135 Acer saccharum 0.0 0.0 0.0 5.6 0.0 0.0 90 Fraxinus pennsylvanica 0.3 1.1 0.3 1.8 0.2 0.3 62 Quercus rubra 0.1 0.0 0.0 0.9 0.4 0.4 27 Acer rubrum 0.3 0.1 0.0 0.3 0.0 0.9 25 Quercus bicolor 0.0 0.4 0.1 0.0 0.1 0.5 17 Sassafras albidum 0.0 0.0 0.0 0.0 0.0 1.0 16 Ostrya virginiana 0.0 0.0 0.0 0.3 0.3 0.2 13 Carya ovata 0.0 0.2 0.0 0.0 0.0 0.4 9 Ulmus Americana 0.1 0.1 0.0 0.2 0.0 0.0 6 Fagus grandifolia 0.0 0.0 0.0 0.1 0.0 0.1 3 Fraxinus nigra 0.1 0.1 0.0 0.0 0.0 0.0 3 Crataegus spp. 0.0 0.2 0.0 0.0 0.0 0.0 3 Quercus alba 0.0 0.1 0.0 0.0 0.1 0.0 2 Tilia americana 0.0 0.0 0.0 0.0 0.1 0.0 1 Populus tremuloides 0.0 0.0 0.0 0.0 0.0 0.1 1 Robinia pseudoacacia 0.0 0.0 0.0 0.1 0.0 0.0 1 Total 0.9 5.1 0.3 11.3 2.7 5.6 414

Watershed Clinton Grand Kalamazoo Species Gap Forest Gap Forest Gap Forest Fraxinus pennsylvanica 248.8 ± 109.9 112.0 ± 66.4 161.7 ± 76.9 236.4 ± 89.0 31.1 ± 4.7 6.2 ± 3.9 Acer rubrum 15.6 ± 9.3 146.2 ± 101.2 0 ± 0 65.3 ± 65.3 3.1 ± 3.1 295.5 ± 109.0 Fraxinus nigra 307.9 ± 115.9 40.4 ± 36.4 15.6 ± 11.8 0 ± 0 99.5 ± 34.1 0 ± 0 Prunus serotinus 21.8 ± 21.8 205.3 ± 82.9 0 ± 0 65.3 ± 27.5 28 ± 16.4 62.2 ± 10.2 Fraxinus americana 24.9 ± 15.2 84 ± 35.7 15.6 ± 7.8 65.3 ± 49.3 115.1 ± 76.6 56 ± 33.9 Ulmus americana 102.6 ± 36.8 96.4 ± 45.2 65.3 ± 22.4 46.7 ± 26.1 24.9 ± 24.9 21.8 ± 9.3 Acer saccharum 0 ± 0 0 ± 0 24.9 ± 24.9 236.4 ± 116.0 0 ± 0 21.8 ± 21.8 Other 40.4 ± 20.6 161.7 ± 56.8 124.4 ± 124.4 152.4 ± 75.2 68.4 ± 16.5 74.6 ± 20.3 Total 762 ± 247.0 845.5 ± 240.6 407.4 ± 144.9 867.7 ± 139.5 370.1 ± 46.1 538 ± 138.3

Watershed Clinton Grand Kalamazoo Canopy Gaps Frequency % Cover Frequency % Cover Frequency % Cover Carex retrosa 1 18.9 ± 11.1% 1 19.3 ± 6.9% 0.5 16.2 ± 16.2 Carex stipata 0.75 20.4 ± 14.6% 1 34.8 ± 18.2% 1 12.9 ± 3.4 Carex stricta 1 27.1 ± 21.5% 1 14.9 ± 10.2% 1 28.9 ± 8.2 Eutrochium maculatum 0.5 2.8 ± 2.8% 0.25 3.0 ± 3.0% 0 0 Juncus effusus 0.25 2.0 ± 2.0% 0 0 0.25 2 ± 2 Onoclia sensibilis 0 0 0.5 9.9 ± 9.9% 0.75 4.7 ± 3.1 Pilea pumila 0 0 0.25 2 ± 2% 0.25 2.0 ± 2.0 Solidago spp. 1 4.1 ± 2.0% 0.5 4.3 ± 4.3% 0 0 Symplocarpus foetidus 1 5.0 ± 5.0% 0.75 4.2 ± 2.9% 0.5 2.9 ± 2.9 Urtica dioica 0 0 0.5 4.9 ± 4.9% 1 5.2 ± 1.9 Total 80.3 ± 6.6% 97.3 ± 1.1% 74.8 ± 7.9%

Surrounding Forests Asarum canadense 0 0 0.25 2.0 ± 2.0% 0 0 Clematis terniflora 0.25 2.0 ± 2.0 0 0 0 0 Glyceria striata 1 12.3 ± 6.9% 0.75 5.9 ± 4.8% 0.75 5.2 ± 3.9% Impatiens capensis 0.5 2.0 ± 2.0% 0.5 2.0 ± 2.0% 0 0 Onoclea sensibilis 0 0 0 0 0.25 2.0 ± 2.0% Osmunda cinnamomeum 0.25 5.0 ± 5.0% 0 0 0.75 6.1 ± 4.9% Osmorhiza longistylis 0.5 2.0 ± 2.0% 0.25 2.0 ± 2.0% 0 0 Parthenocissus quinquefolia 0.25 2.0 ± 2.0% 0 0 0 0 Toxicondendron radicans 1 8.2 ± 6.8% 0.5 6.3 ± 6.3% 0 0 Trillium grandiflorum 0 0 0.25 2.0 ± 2.0% 0.25 2.0 ± 2.0% Total 33.5 ± 4.3% 21.5 ± 6.8% 15.3 ± 2.7%

Table 1.7. Mean (± SE) density of the four most abundant invasive plant species recorded in 12 sites in three watersheds in Michigan.

Species Rosa multiflora Rhamnus spp. Elaeagnus angustifola Lonicera maackii Watersheds (m2 per ha-1) (stems per ha-1) (m2 per ha-1) (m2 per ha-1) Clinton Gap 50.0 ± 50.0 32.4 ± 14.5 43.8 ± 35.9 18.8 ± 18.8 Forest 0 0 406.5 ± 366.0 250.2 ± 176.9 Grand Gap 70.4 ± 70.4 16.8 ± 9.3 225.1 ± 216.9 531.6 ± 309.3

Forest 350.2 ± 273.2 1.3 ± 1.3 150.1 ± 112.3 106.3 ± 61.6 Kalamazoo Gap 429.9 ± 219.8 31.1 ± 31.1 164.2 ± 100.0 0

Forest 50.0 ± 50.0 18.1 ± 18.1 93.8 ± 74.4 0 Watershed 0.19 0.59 0.88 0.23 P value

Gap/forest 0.86 0.77 0.65 0.53 P value

Main effect df F P Gap/Forest 1 8.50 0.0002 Watershed 2 1.12 0.290

Pairwise comparisons t P Gap overstory vs Forest overstory 3.28 0.0002 Gap overstory vs Gap saplings 0.947 0.473 Gap overstory vs Forest regeneration 2.42 0.0004 Forest overstory vs Gap regeneration 3.84 0.0002 Forest overstory vs Forest regeneration 2.09 0.0008 Gap regeneration vs Forest regeneration 2.70 0.0002

Table 1.9. Group identifier from non-metric multidimensional scaling ordination, indicator values (IV) and P values for significant indicator species of trees in canopy gaps and surrounding forests.

Species NMS Group Indicator value P

Fraxinus spp. Canopy gap 78.5 0.0002 Fagus grandifolia Surrounding forest 33.2 0.0066 Carya cordiformis Surrounding forest 25.0 0.020 Prunus serotina Surrounding forest 80.3 0.0002

Quercus rubra Surrounding forest 62.3 0.0014 Acer rubrum Surrounding forest 73.2 0.0006 Sassafras albidum Surrounding forest 39.6 0.0028 Carya ovata Surrounding forest 30.1 0.0346

Figure 1.1. Location of the 12 riparian sites in the three watersheds selected to represent the temporal progression of the emerald ash borer across southern Michigan.

Macroplot

Subplot

Microplot

140 A

) Gap

1 120

ha Forest

3 100 80 B 60 B 40

(m Volume CWD 20 0 Clinton Grand Kalamazoo Watershed

Figure 1.3. Mean (± SE) volume of coarse woody debris (m3 ∙ ha-1) in four canopy gaps and surrounding forests in each of three watersheds in Michigan (n = 12 sites).

- 160

140 y = -14.257x + 28715 R² = 0.6983

120

100

60 40

Volume Coarseof Woody Debris (m3 ha 20

0 2005 2007 2009 2011 2013 2015 Year of gap formation

Figure 1.4. Linear relationship between total volume of coarse woody debris (m3 ∙ ha-1) in canopy gaps and the first year that overstory ash mortality was apparent in aerial imagery (n=12 sites).

100

90 80 y = -8.478x + 83.976 R² = 0.6689 70 60

40 30 20 Herbaceous density cover)(% 10

0 0.00 2.00 4.00 6.00 8.00 10.00 Seedling density (per m2)

100 y = -0.0249x + 68.596 R² = 0.0774 90 80

70 60 50 40 30

20 Herbaceous density cover)(% 10 0 0 200 400 600 800 1000 1200 1400 1600 1800

Sapling density (per ha-1)

Figure 1.6. Non-metric multidimensional scaling ordination (two-dimensional) of species composition in pre-EAB forest overstories and in the current sapling stratum. Polygons were overlaid to easily visualize groupings of sites of canopy gaps and surrounding forests. The successional arrows connect pre-EAB overstories to the corresponding saplings in the same site to examine stand level changes in species composition. For example: a site labelled “C1G” would denote that the site is in the Clinton River watershed, is a canopy gap and was the first of four sites surveyed in this watershed. Final stress = 14.57; axis 1 R2 = 58.2%; axis 2 R2 = 26.2 %; cumulative R2 = 86.4 %.

Canopy Gaps Surrounding forests

Fraxinus T. americana F. grandifolia U. americana P. grandidentata C. cordiformis P. serotina J. nigra

S. nigra A. negundo O. virginiana Q. rubra A. rubrum P. resinosa S. albidum C. ovata A. saccharum

Q. bicolor R. pseudoacacia Q. alba P. strobus

Saplings Overstory

(n=24)

(n=12) (n=12)

CHAPTER 2: SPECIES DIVERSITY AND ASSEMBLAGES OF CERAMBYCIDAE IN THE AFTERMATH OF THE EMERALD ASH BORER INVASION IN RIPARIAN FORESTS IN SOUTHERN MICHIGAN

Introduction

Mortality of overstory trees resulting from invasion by a destructive pest is likely to affect multiple variables in forest ecosystems, including native invertebrates. Phloem- and woodboring insects may be particularly impacted by the relatively rapid influx of declining, dying and dead trees. Larval stages of these insects, which are predominantly in the orders

Coleoptera, Hymenoptera, and Lepidoptera, feed on phloem and/or sapwood (Drooz 1985).

Some of these insects are damaging invaders, including emerald ash borer (Agrilus plannipennis

Fairmaire (Coleoptera: Buprestidae), Asian longhorned beetle (Anoplophora glabripennis

Motschulsky) (Coleoptera: Cerambycidae), and the Sirex woodwasp (Sirex noctilio Fabricius)

(Hymenoptera: Siricidae). In their native habitats, however, these and most other phloem and woodboring insects are secondary feeders on native plant species, typically colonizing severely stressed, dying, recently killed or recently cut trees, logs or branches (Drooz 1985, Evans et al.

2004, Spradberry and Kirk 1978). Native longhorned beetles (F. Cerambycidae) in North

America, for example, rarely attack healthy trees and typically play important roles in wood decomposition and nutrient cycling in forests (Evans et al. 2004, Gandhi et al. 2009, Ulyshen

2016).

I was interested in the response of native cerambycids to the North American invasion of the emerald ash borer (EAB). Native to parts of Asia, EAB is a phloem-boring beetle that has caused massive mortality to ash trees in eastern North America (Herms and McCullough 2014,

Klooster et al. 2014, Poland and McCullough 2006, Pugh et al. 2011). While EAB is a secondary pest in its native range in China, EAB can also colonize and kill healthy North American ash

trees (Herms and McCullough 2014, Wei et al. 2004). Interspecific variation in host preference or host resistance has been documented (Anulewicz et al. 2007, 2008, Rebek et al. 2008, Tanis and McCullough 2015, Villari et al. 2016), but all North American ash species are likely at some risk (Herms and McCullough 2014, Poland and McCullough 2006).

Emerald ash borer was first detected in the Detroit metropolitan area in southeast

Michigan and nearby Windsor, Ontario, Canada in 2002 (Cappaert et al. 2005, Poland and

McCullough 2006), but dendrochronological reconstruction showed EAB was killing ash trees by the mid 1990s (Siegert et al. 2014). Populations of EAB have been identified in 35 U.S. states and five Canadian provinces as of May 2019 (emeraldashborer.info 2019). In areas with the longest invasion history in southeast Michigan, ash mortality rates of 80-99% have been observed (Burr and McCullough 2014, Klooster et al. 2014, Knight et al. 2013, Smith et al.

2015). Despite establishment of numerous satellite populations resulting from human transport of infested ash nursery trees, logs or firewood (Siegert et al. 2014), the EAB invasion generally progressed from southeast to southwest Michigan over a ten year period. As recently as 2011, mortality rates of green ash (Fraxinus pennsylvanica Marsh.) ranged from 79% in southeast

Michigan forests to less than 20% in southwest Michigan forests (Burr and McCullough 2014).

In riparian forests, black ash (Fraxinus nigra Marsh.) and green ash are often abundant and dominant in mesic or hydric forests bordering bogs, swamps and along riparian corridors

(Crocker et al. 2006, Gucker 2005, Nisbet et al. 2015, Poland and McCullough 2006). Black ash and green ash are both highly preferred EAB hosts and highly vulnerable to EAB colonization

(Klooster et al. 2014, Tanis and McCullough 2015). In a recent survey of 12 riparian forests across southern Michigan, nearly all overstory ash trees had been killed by EAB, but overstory mortality in southwest Michigan occurred nearly ten years after that in southeast Michigan

(Chapter 1). While >85% ash mortality was observed across the state, the temporal east to west gradient of the EAB invasion was manifested in differences in standing dead ash trees and coarse woody debris in these stands (Chapter 1). For example, in early-invaded southeast Michigan stands, most dead ash trees have fallen and relatively fresh coarse woody debris is abundant

(Chapter 1). In contrast, in more recently invaded riparian forests in south central and southwestern Michigan, most dead ash trees remain standing. Ash mortality following the EAB invasion in riparian forests and the subsequent availability of dead trees or coarse woody debris could potentially provide hosts or habitat for an array of xylophagous insects (Perry and Herms

2016, Ulyshen et al. 2011). I surveyed woodboring cerambycid species in riparian forest sites selected to represent the temporal gradient of the EAB invasion across southern Michigan.

Cerambycid beetles are a large and diverse family with many xylophagous species in the eastern

U.S. Previous research has shown a lure originally identified as a pheromone for Neoclytus mucronatus mucronatus (Lacey et al. 2007) is broadly attractive to many cerambycid species

(Hanks and Millar 2013), providing a means to assess assemblages of these beetles. Cerambycid species captured in baited traps were identified and related to stand variables, including overstory composition and coarse woody debris abundance. Because many cerambycid species are semivoltine (Haack 2017), changes in populations and local species assemblages may temporally lag a major disturbance. I hypothesized, therefore, that cerambycid species assemblages would differ among riparian forest sites along the temporal gradient of EAB invasion. I predicted that beetle captures and species richness would be higher in sites invaded early by EAB than in similar sites that were more recently invaded. I was also interested in evaluating the diversity and general activity periods of cerambycid species attracted to traps in these sites.

Materials and Methods

Study Sites: Sites were located in three watersheds selected to represent the temporal gradient of the EAB invasion that progressed from the epicenter in southeast Michigan to the south central region and eventually southwest Michigan. The Clinton River watershed in southeast Michigan was invaded early by EAB. This watershed lies in counties that were quarantined by 2002 and extensive ash mortality was evident in aerial images by 2006 (Table 2.1). Sites along the Grand

River watershed in central Michigan represented an intermediate stage of the EAB invasion.

Sites were in counties quarantined in 2003-2004 and ash mortality became evident by 2011

(Table 2.1). Sites along the Kalamazoo river in southwest Michigan were invaded later.

Counties in this area were quarantined in 2004-2006 and ash mortality first became apparent in aerial images in 2012 (Thesis ch.1) (Table 2.1). Specific sites for trapping were identified by using an atlas to locate first order streams running through public land in each watershed. Aerial images were then examined to identify substantial canopy gaps in intact forest bordering the streams. These areas were then visited to determine whether the observed canopy gap was due to ash mortality caused by EAB. Four sites where all or nearly all ash trees had been killed by EAB were chosen in each watershed (12 sites total) (Figure 2.1).

Site Surveys: Vegetation and coarse woody debris were characterized using fixed radius plots and linear transects. I established eight circular fixed radius plots (11.4 m radius) and four linear transects (two 100 x 2 m transects and two 50 x 2 m transects) in each site. Four of the plots were located in canopy gaps and the others were in the adjacent forest, 25-50 m from the perimeter of the gap. One 100 x 2 m transect was located in the canopy gap, running parallel to the stream and

2-5 m from stream’s edge. A second 100 m transect was located 5-10m in the surrounding forest, running parallel to the gap perimeter. Two 50 x 2 m transects were established along the

stream at the upstream and downstream transition between the gap and intact forest. Each transect extended 25 m into both the canopy gap and surrounding forest. Species and DBH of standing live and dead trees (>10.2 cm DBH) were recorded in the linear transects and fixed radius plots. For each piece of coarse woody debris encountered in the linear transects (CWD)

(>7.6 cm diam), I recorded the length, midpoint diameter and decay class. Each log was classified as either fresh (outer bark still intact, wood solid) or decayed (outer bark sloughing off or absent, wood spongy). All variables were standardized per hectare for analysis. A more detailed description of survey methods is available in (Chapter 1).

Cerambycid Trapping: Traps were baited and monitored from early summer to fall in 2017 and

2018. In each site, I deployed two cross-vane panel traps (1.2 m height by 0.3 m width) (Contech

Enterprises, Inc., Victoria, British Columbia, Canada) at the perimeter of the canopy gap. One trap was deployed at ground level and one was suspended directly above it in the canopy of a living hardwood tree. Ground traps were hung from an L-shaped section of rebar embedded into the soil such that the collection cup was ~1 meter above the ground. A braided nylon line was launched using a Big Shot® (Sherrill Tree, Greensboro, NC) over the middle of a lower canopy branch. I secured the line to the top and bottom of the trap, then raised the trap, ensuring it was not impeded by the tree bole. All trap surfaces were coated in Fluon (Fisher Scientific,

Pittsburgh, PA), which renders trap surfaces more slippery, reducing escape of insects attracted to the traps (Graham et al. 2010).

Each trap was baited with a bubble cap lure containing (R) 3-hydroxyhexan-2-one

(Synergy Semiochemicals Corp.) and a pouch of ultra-high release ethanol (release rate 0.5 g·day−1; Alpha-Scents, Inc. Portland, OR). The compound in this lure, also known as 3R-ketone, is a primary component of the pheromones for many members of the subfamily Cerambycinae

and is attractive to multiple species (Hanks et al. 2007). Collection cups were filled to a depth of

~ 5 cm with non-ethanol propylene glycol (RV & Marine Anti-Freeze -50°F Burst Proof,

National Automotive Parts Association [NAPA], Atlanta, GA, USA) as a preservative. Traps were deployed from 30 May - 1 June in 2017 and 8 - 10 May in 2018. Traps were checked at 2-3 week intervals and lures were replaced in mid-summer. Traps were retrieved on 10-12 October

2017 and 25-27 September 2018.

Collection and Identification: Captured insects were kept frozen until they were sorted, then were stored in 70% ethanol until they were identified using dichotomous keys (Lingafelter

2007). At least two representative individuals of each species were submitted as voucher specimens to the Alfred J. Cooke Arthropod Research Collection at Michigan State University.

Statistical Analysis: To estimate seasonal activity, beetle captures from both years were pooled and activity periods were calculated for species with ≥10 individuals collected over the course of both trapping years. Accumulated GDD (base 10 ° C) were obtained for each year from the nearest MSU Enviro-Weather station (enviroweather.msu.edu). Cumulative GDD at the start and at the end of each collection period were recorded for beetle species trapped in each site. An activity chart was constructed to indicate the proportion of the total number of beetles collected during each interval between trap checks.

Annual and total cerambycid catches and species richness were analyzed using one-factor

ANOVA (Proc GLIMMIX; SAS Institute Inc. 2015) to assess differences between canopy vs ground traps and among sites in three watersheds (Clinton, Grand, Kalamazoo). Variables were checked for normality and heterogeneity of variance using histograms, normal probability plots, and side by side boxplots of the residuals. If ANOVA results were significant (P < 0.05), least

square means were separated using a Student’s t test. Analysis was performed separately for each year and again with all beetle captures pooled to detect differences between years.

Individual based rarefaction, and Chao 2 nonparametric occurrence-based species richness estimators were used to examine cerambycid richness. For species richness estimates, beetle captures were pooled across trap types and watersheds. Individual based rarefaction curves were calculated for the two trap types (canopy, ground) using Analytic Rarefaction

(Holland, S. M. 2003) in increments of ten individuals. Individual based rarefaction was used to estimate the number of species encountered per number of beetles captured, and allowed us to compare species collected in traps that were in different watersheds or at different vertical heights. Chao 2 richness estimates (Chao 1987) were calculated using EstimateS (Colwell 2013).

Chao 2 estimates use the ratio of species encountered in a single site (singletons) to species found in two sites (doubletons) to estimate the number of missing species. Chao 1 and Chao 2 species richness estimators use the same formula to estimate the total species in a site. They differ in that Chao 1 is primarily used when species abundance data are analyzed and Chao 2 is used for analysis of occurrence data. Given that the lure used in traps attracts a wide range of cerambycid species but not all species, I did not assume that captured species represented the total abundance or distribution of all species. Because of this, I used the Chao 2 occurrence- based estimator with 100 randomizations of data.

To examine similarities and differences in assemblages of beetle species captured among the three watersheds and between canopy and ground traps, non-metric multidimensional scaling

(NMS) was performed (PC-ORD). I began by using autopilot mode with Sorenson (Bray-Curtis) distance on the “slow and thorough” setting to determine the ideal dimensionality for the NMS output. The “slow and thorough” setting uses a random starting configuration and completes 250

runs of real and randomized data with a maximum of 500 iterations for two to six axes. Correct dimensionality was then determined using the output of the NMS scree plot. Number of interpreted axes was determined by maximally reducing stress while attempting to not distort the output over too many axes. Non-metric multidimensional scaling ordinations with stress >25 are usually deemed uninterpretable, and an ideal final stress of 5-15 is preferred (Peck 2016). After dimensionality was determined, I again ran the ordination for the set number of axes. Once the final ordination was achieved, I overlaid correlational variables from the related surveys of vegetation and CWD if they had at least a weak correlation with either axis (r2 > 0.20). To test for significance among watersheds and between canopy/ground traps from each NMS ordination,

PERMANOVA analysis was performed (PC-ORD). When PERMANOVA results were significant, pairwise comparisons of treatments were applied.

To test for potential drivers of differences in species assemblages, indicator species analysis (ISA) was performed for each ordination (PC-ORD). Indicator species analysis (ISA) allows data exploration to determine which species are responsible for differences observed in grouping variables based on the constancy or distribution of the abundance of each species. An indicator value (IV) was calculated for each species based on its abundance and constancy in each level of the grouping variable used in the ordination. Statistical significance is then evaluated by randomly reassigning sample units to groups, reiterating IV calculations many times, then calculating the proportion of IV values for each species that are ≥ the maximum IV permutation calculated during bootstrapping. A low P-value indicates a species is more abundant and constant in a group than would be expected by chance, and is said to “indicate” its assigned group.

Results

Overall Beetle Captures: Over the course of both trapping years, I collected 3645 beetles representing 65 species and five subfamilies of cerambycids. This included 1875 beetles representing 50 species captured in 2017 and 1770 beetles from 46 species that were captured in

2018. Canopy traps captured 1730 beetles from 57 species and ground traps captured 1915 beetles from 42 species (Table 2.2). The most abundant species collected were from the subfamily Cerambycinae, with Clytus ruricola (Olivier) (10.3%), Neoclytus mucronatus (F.)

(15.1%) and Xylotrechus colonus (F.) (40.2%) combining to make up 65.6% of the total beetles collected. Across both years, 29 species of Cerambycinae comprised 90% of the total beetles collected.

For the majority of species, relatively few individuals were captured by the baited traps.

Across both years, 45 of the 65 total species were represented by ten or fewer beetles and 31 species were represented by only one or two specimens (Table 2.2). Trap height clearly influenced captures, particularly for uncommon species. I captured 23 species (47 beetles) only in canopy traps while eight species (11 beetles) were caught only in ground traps (Table 2.2). Of the species caught exclusively in one trap type, only Eburia quadrigeminata (Say) was represented by 10 or more individuals and all 12 were captured in canopy traps (Table 2.2). The only non-native species captured was Phymatodes testaceous (L.). This beetle, known as the tanbark borer, was originally introduced from Europe in the mid-1800s and is widely established across much of the U.S. (Lingafelter 2007).

Seasonal Activity: During 2017, I collected 26 species (51% of the total species) the first time I checked traps on 13-15 June, corresponding to 410 cumulative growing degree days (base 10 ̊

C). By the fourth trap check on 9-10 August (corresponding to 1015 GDD), 48 of the total 50

species (96% of total) had been collected (Figure 2.3 A). In 2018, traps were set up earlier, and by the first trap check on May 29-31 (corresponding to 290 GDD), 17 species (37% of total) were collected. As in 2017, nearly all species (44 of the total 46 species (95.7%)) in 2018 were collected by the fourth trap check on 30 July-2 August (corresponding to 1015 GDD) (Figure 2.3

B). In both years, two species were collected only after 1000 GDD, but they were not the same species. In 2017, Parelaphidion aspersum (Haldeman) and Lepturges confluens (Haldeman), were collected between 1350 and 1600 GDD, while in 2018, Clytoleptus albofasciatus

(Castelnau and Gory) and Hyperplatys maculata (Haldeman) were collected between 1300 and

2150 GDD. Less than ten individuals of these four species were collected across all sites. Both of the species collected after 1000 GDD in 2017 were collected only during 2017. The two species collected only after 1000 GDD during 2018, however, were also collected in 2017.

Individuals of Hyperplatys maculata were collected as early as 6 July 2017 (~ 650 GDD).

Fifteen of the 20 species that were represented by 10 or more captured individuals belonged to the Cerambycinae, while four species belonged to the Lamiinae, and one species is in the Lepturinae. These 15 species varied in their seasonal phenology and no clear trends were observed among subfamilies (Figure 2.2). Most species with short, ephemeral activity periods were captured early in the season, between early May and mid-June (~200-500 GDD) (Figure

2.2). Examples include Megacyllene caryae (Gahan), Anelaphus villosus (F.), and Phymatodes aereus (Newman). More than 80% of the beetles in these species were captured by mid-June

(~500 GDD). Other species, including Xylotrechus colonus and Astylopsis macula (Say), were active throughout the season (Figure 2.2). Similarly, 60% or more of the individuals from 13 of the 20 species represented by at least 10 specimens were captured during a single period. Two species, Eburia quadrigeminata and Aegomorphus modestus (Gyllenhal), were active mostly

during mid-summer (July-August, ~500-1000 GDD) and more than 60% of Neoclytus mucronatus beetles were captured in the late summer (August-September, ~1000-1300 GDD).

Watershed Comparisons: Neither mean species richness nor total beetle captures varied significantly among watersheds for either year or as a whole when captures from both years were combined. Nonparametric techniques, however, suggested traps in sites in the Grand River watershed captured fewer species than traps in sites in the other two watersheds. Individual based rarefaction curves indicated fewer species were captured in the Grand River watershed sites in 2018 and across both years compared to sites in the Clinton River and Kalamazoo River watersheds (Figure 2.4 D, F). Additionally, Chao 2 estimates of species richness among watersheds were lowest in 2018 and for both years combined for sites in the Grand River watershed (Table 2.3).

Canopy vs. Ground Trap Comparisons: While similar numbers of beetles were captured by canopy and ground traps, on average, canopy traps collected more species than ground traps in

2017 (F=5.7; df=1,22; P=0.026), 2018 (F=6.03; df=1,22; P=0.023), and when both years were combined (F=8.5; df=1,22; P=0.008). Additionally, individual based rarefaction curves suggested that canopy traps consistently detected more species than ground traps in terms of both trapping effort and on a per beetle basis for each year individually and both years combined

(Figure 2.4 A, C, E). Chao 2 estimates of species richness were also higher for canopy traps than ground traps (Table 2.3).

Comparison of Cerambycid Species Assemblages Among Watersheds: Analysis of cerambycid assemblages in watersheds using PERMANOVA and NMS indicated that cerambycid presence and species abundance varied among watersheds. Species assemblages in the Clinton River watershed differed from those in the Grand and Kalamazoo river watersheds which had similar

assemblages (Figure 2.5) (Table 2.4). I plotted species abundances against site specific variables from previous surveys, which had at least a weak association (r2 > 0.2) with either axis. Results showed that total CWD volume (r2=0.45) and fresh CWD volume (r2=0.31) were positively correlated with axis 1 and with cerambycid species assemblages in the Clinton River watershed.

Results of indicator species analysis (ISA) showed that three species were significant indicators of the Clinton River watershed (Table 2.5). While little is known about the ecology of these species, all are known to feed in a variety of dead and declining hardwoods and Clytus ruricola reportedly colonizes decaying logs (Evans 2014). Clytus ruricola was by far the most prevalent of these three species. I captured 225 beetles in sites in the Clinton River watershed compared to

65 and 90 in sites in the Grand and Kalamazoo River watersheds, respectively. Fewer than 100 individuals of Euderces picipes (F.) and Sarosesthes fulminans (F.) were captured overall, but

70% and 65% of the individuals of each species, respectively, were captured in the Clinton watershed sites. Indicator species analysis did not suggest any species were significant indicators for the sites in the Grand River or Kalamazoo River watersheds, which had similar assemblages.

Abundance of two species, however, Neoclytus acuminatus acuminatus (F.) and Xylotrechus colonus, varied between traps in Clinton River watershed and traps in the sites in the Grand

River and Kalamazoo River watersheds. Over both trapping seasons, I collected 160 Neoclytus a. acuminatus beetles, including 27 beetles (17%) that were collected in traps in the Clinton River watershed, 70 beetles (44%) collected in traps in the Grand River watershed and 63 beetles

(39%) collected in traps in the Kalamazoo River watershed. Xylotrechus colonus, the most abundant species captured (1544 beetles total), was also less abundant in traps in the Clinton

River sites than in sites in the other watersheds. I captured 289 (19%), 546 (35%), and 709 (46%)

X. colonus beetles in traps in sites in the Clinton, Grand and Kalamazoo River watersheds,

respectively. Neoclytus a. acuminatus and X. colonus both reportedly feed on a wide range of hardwoods, but both can develop in ash ((Drooz 1985, Evans 2014, Gosling 1984, Lingafelter

2007, Vlasak and Vlasakova 2002). Additionally, X. colonus will reportedly colonize freshly killed trees (Drooz 1985).

Comparison of Assemblages in Canopy vs. Ground Traps: Results from PERMANOVA and

NMS analysis revealed that despite substantial overlap, species assemblages captured by canopy and ground traps differed significantly (Figure 2.6) (Table 2.4). The ISA showed that 11 species were responsible for the differences; eight species were indicators of canopy taps and three species were indicators of ground traps (Table 2.5). Significant indicator species included the three most abundant species collected; Xylotrechus colonus and Clytus ruricola were indicators for ground traps while Neoclytus mucronatus was an indicator for canopy traps.

Discussion

Cerambycid species assemblages varied among riparian forests selected to represent a gradient of EAB invasion history across southern Michigan. Approximately 85% of the ash trees

(>10.2 cm DBH), representing 40-60% of the total pre-EAB forest basal area, have died in these sites (Table 2.1). This relatively rapid change in live tree cover did not necessarily affect cerambycid beetle abundance or overall species richness, but species assemblages differed among the three watersheds. Different EAB invasion histories in the watersheds where beetles were trapped reflected the length of time since ash mortality occurred and the dynamics of dead ash. In sites in the Clinton River watershed, with the longest history of EAB invasion, three cerambycid species appear to be responding to the influx of coarse woody debris (CWD) as dead ash trees break and fall. All three of these species reportedly colonize and develop on dead or declining hardwoods (Drooz 1985, Evans 2014, Gosling 1984, Lingafelter 2007, Vlasak and

Vlasakova 2002). Clytus ruricola, which may feed in decaying wood as well as fresh CWD

(Evans 2014), was collected in particularly high densities in the Clinton River watershed sites where dead ash logs are in advanced stages of decay. In contrast, captures of two species,

Neoclytus acuminatus acuminatus and Xylotrechus colonus, were lower in the Clinton River watershed sites than in more recently invaded sites in the Grand and Kalamazoo River watersheds where dead standing ash trees and fresh CWD are abundant. Other invertebrate groups have responded to ash mortality or inputs of dead wood following EAB invasion.

Ulyshen et al. (2011) found a variety of arthropods were captured in higher abundances near freshly fallen ash logs than away from them. Gandhi et al. (2014) suggested ground beetles

(Coleoptera: Carabidae) responded to canopy gap size and ash mortality in forests of southeast

Michigan, but responses varied among species. More recently, however, Perry and Herms (2016)

reported changes in carabid species composition were ephemeral and concluded that carabids are resilient to EAB related disturbance. Whether arthropod species or even community level changes are an ephemeral response to ash mortality or represent a long-term shift in these forest ecosystems remains to be seen. Further research to evaluate cerambycid species along a controlled gradient of dead ash volume and decay over time would be valuable for further understanding community level changes and projecting long term effects of the EAB invasion on this important group.

Although the lure I used to bait the traps is broadly attractive to numerous cerambycid species (Hanks and Millar 2013, 2016), my results may represent only a portion of the total cerambycid communities in southern Michigan forests. Overall, the cerambycid species I captured were similar to reports from other studies that used the 3R-ketone lure to assess longhorned beetles in eastern U.S. hardwood forests (Graham et al. 2012, Handley et al. 2015,

Hanks and Millar 2013, Hanks et al. 2014). Species in the subfamily Cerambycinae dominated my collection, accounting for 90% of the total beetles captured and Xylotrechus colonus,

Neoclytus mucronatus, and Clytus ruricola were particularly abundant. The active component of the lure (R)-3-hydroxyhexan-2-one (3R-ketone), is a highly conserved aggregation pheromone in

Cerambycinae species (Hanks et al. 2007). Several trapping surveys in eastern U.S. hardwood forests have shown these three species to be attracted to 3R-ketone even when baited as a single component lure (Graham et al. 2012, Hanks and Millar 2013). Additionally, trapping surveys for cerambycids which included 3R-ketone as either a lure or a lure component consistently report high abundances of Xylotrechus colonus and Neoclytus mucronatus in trap captures, indicating they are likely both attracted to the lure component and an abundant species in surveyed forests

(Handley et al. 2015, Graham et al. 2012, Hanks and Millar 2013, Hanks et al. 2014, Schmeelk et

al. 2016). Conversely, while Clytus ruricola is attracted to 3R-ketone, it is typically captured in low numbers, often comprising less than 1% of the total beetle captures (Handley et al. 2015,

Graham et al. 2012, Hanks and Millar 2013, Hanks et al. 2014, Schmeelk et al. 2016). In my survey, Clytus ruricola was uncommon in areas of intermediate and recent EAB invasion, but it accounted for nearly 20% of the beetles captured in southeast Michigan sites. These higher than typical capture rates further indicate Clytus ruricola may be exploiting the abundance of decaying wood present in the Clinton River watershed.

While activity periods varied among species, most beetles were captured between mid

May and mid June (100-400 GDD10 C) and by mid August, corresponding to 1000 accumulated

GDD10 C, >95% of all species had been captured in both years. Other studies that involved multiple lures and trap combinations (Handley et al. 2015, Hanks and Millar 2013, Hanks et al.

2007,2014) have similarly reported most cerambycid species were captured by mid-summer.

This pattern has implications for trapping programs designed for detection or surveys of cerambycids. Traps deployed early in the season (May-June) are likely to capture more species of cerambycids compared to traps established later in the summer.

Trap height was the major driver of species assemblages in sites. Overall, 31 of 65 species were exclusively found only in canopy traps (23 species) or ground traps (8 species) and canopy traps consistently captured more species than ground traps in all watersheds. Non-metric multidimensional scaling and indicator species analysis determined that 11 species were indicators of either canopy traps (8 species) or ground traps (3 species). This included the three most abundant species I trapped. Xylotrechus colonus and Clytus ruricola were indicators of ground traps and Neoclytus mucronatus was an indicator of canopy traps. These results are consistent with previous studies (Dodds 2014, Graham et al. 2012, Schmeelk et al. 2016,

Ulyshen and Hanula 2007, Vance et al. 2003, Wermlinger et al. 2007) that reported that cerambycid assemblages differed between vertical strata of the forest. In past studies,

Xylotrechus colonus was consistently captured in ground level traps (Graham et al. 2012,

Schmeelk et al. 2016), while Eburia quadrigeminata and Neoclytus mucronatus were usually captured by traps in forest canopies (Graham et al. 2012, Vance et al. 2003). While it is not clear why different species are more prevalent in different vertical strata, larval development niches may help to drive species distribution (Graham et al. 2012, Wermlinger et al. 2007). For example, species with larvae that can feed in decaying wood such as Clytus ruricola and

Orthosoma brunneum, may be more likely to be captured in ground traps that are near CWD.

Species that girdle or develop in the upper branches or twigs, such as many Lamiinae species, may be captured more frequently in traps suspended in forest canopies (Graham et al. 2012,

Wermelinger et al. 2007). Differences in species occurrence and abundance among different levels of the forest canopy have been observed in a wide range of forest insects across several orders and families (Dodds 2014, Ulyshen and Hanula 2007, Wermelinger et al. 2007). These results reinforce the importance of establishing traps in different vertical strata to maximize species detection of forest dwelling cerambycids.

This study indicates that the flux of ash coarse woody debris from the EAB invasion has at least changed species distributions of several cerambycid species and has perhaps resulted in population level changes of these species. This adds to the literature demonstrating indirect effects of EAB on communities of arthropods. While assemblages differed among watersheds, we did not see a loss of biodiversity, even in areas with extended history of EAB. Many species of cerambycids are known to be polyphagous, with varying host ranges (Wong 2016). This adaptation may allow for polyphagous species including those known to develop in ash such as

Neoclytus acuminatus to be resilient and to persist albeit perhaps in lower abundances following the loss of a major host species. Despite our findings however, the long term impacts of the EAB invasion on assemblages of cerambycids is unknown. Information on the ecology of many species is absent or incomplete, leaving the long term implications of changes in their populations difficult to predict.

APPENDICES

APPENDIX A. Tables and Figures

Table 2.1. Year of canopy gap formation, mean (± SE) live basal area of current overstory, percent of total basal area that was ash prior to EAB invasion, percent of ash basal area represented by dead trees, forest overstory species richness, and CWD volume in canopy gaps and surrounding forests in three watersheds across southern Michigan.

Watershed Year of Current Pre EAB Mortality Species CWD volume (m2 ha-1) gap overstory ash basal of ash Richness formation BA (m2 ha-1) area % basal area (%) Clinton 2006-2008 Fresh Total Canopy Gap 12.7 ± 3.5 61.6 ± 15.5 97.1 ± 0.8 7.5 ± 1.0 25.7 ± 7.2 108.5 ± 20.2 Surrounding forest 29.9 ± 6.1 10.9 ± 2.3 76.6 ± 21 11 ± 1.3 2.6 ± 1.9 31.5 ± 12.9 Grand 2011-2012 Canopy Gap 6.9 ± 2.5 59.7 ± 12.7 92.6 ± 3.9 8.3 ± 0.5 5.3 ± 4.5 34.2 ± 19.5 Surrounding forest 39.5 ± 2.4 6.0 ± 2.0 82.5 ± 14.1 12 ± 1 3.3 ± 1.2 29.0 ± 7.4 Kalamazoo 2012-2015 Canopy Gap 10.0 ± 1.5 45.9 ± 5.0 96.7 ± 1.4 10.0 ± 0.5 5.2 ± 2.1 28.3 ± 5.9 Surrounding forest 26.7 ± 3.3 4.8 ± 1.7 94.4 ± 5.6 11.5 ± 0.9 4.7 ± 2.8 39.1 ± 17.2

Table 2.2. Number of individuals of cerambycid species captured in canopy traps and ground traps by subfamily and tribe. Subfamily/Tribe Species Canopy Ground Total Cerambycinae

Anaglyptini Cyrtophorus verrucosus (Olivier) 109 63 172 Bothriospilini Knuliana c. cincta (Drury) 1 0 1 Phymatodes aereus (Newman) 77 46 124 Phymatodes amoenus (Say) 3 1 4 Callidini Phymatodes testaceous (L.) 23 6 29 Phymatodes varius (F.) 1 0 1 Clytoleptus albofasciatus (Castelnau and Gory) 0 3 3 Clytus ruricola (Olivier) 96 280 376 Megacyllene caryae (Gahan) 46 14 60 Neoclytus a. acuminatus (F.) 61 102 163 Clytini Neoclytus mucronatus (F.) 397 152 549 Neoclytus scutellaris (Olivier) 2 0 2 Sarosesthes fulminans (F.) 54 21 75 Xylotrechus colonus (F.) 481 986 1467 Xylotrechus convergens LeConte 0 1 1 Eburiini Eburia quadrigeminata (Say) 12 0 12

Anelaphus parallelus (Newman) 51 11 62 Anelaphus pumilus (Newman) 43 3 64 Anelaphus villosus (F.) 28 14 42 Elaphidion mucronatum (Say) 1 0 1 Elaphidiini Micranoplium unicolor (Haldeman) 1 0 1 Parelaphidion aspersum (Haldeman) 2 0 2 Parelaphidion incertum (Newman) 16 1 17 Stenosphenus notatus (Olivier) 5 0 5 Ibidionini Heterachthes quadrimacualtus (Haldeman) 1 0 1 Molorchini Molorchus b. bimaculatus (Say) 1 0 1 Stenopterini Callimoxys sanguinicollis (Olivier) 1 0 1 Tillomorphini Euderces picipes (F.) 21 33 54 Trachyderini Purpuricenus paraxillaris (Macrae) 4 0 4

Table 2.2. (cont’d)

Lamiinae Species Canopy Ground Total Astylopsis collaris (Haldeman) 0 1 1 Astylopsis macula (Say) 85 69 154 Hyperplatys maculata (Haldeman) 7 1 8 Leptostylus transversus (Gyllenhal) 4 1 5 Lepturges angulatus LeConte 1 0 1 Lepturges confluens (Haldeman) 1 0 1 Lepturges symmetricus (Haldeman) 2 0 2 Acanthocinini Liopinus alpha (Say) 2 0 2 Liopinus punctatus (Haldeman) 1 0 1 Urgleptes foveatocollis (Hamilton) 1 1 2 Urgleptes querci (Fitch) 12 46 58 Urgleptes signatus LeConte 3 4 7

Urographis fasciatus (DeGeer) 3 3 6 Urographis triangulifer (Haldeman) 2 1 3 Acanthoderini Aegomorphus modestus (Gyllenhal) 13 9 22 Oplosia nubila LeConte 1 1 2 Eupogonius pauper LeConte 1 2 3 Desmiphorini Eupogonius subermatus LeConte 1 1 2 Psenocerus supernotatus (Say) 18 5 33 Dorcaschematini Dorcaschema cinereum (Say) 1 0 1 Lamiini Microgoes oculatus Leconte 3 1 4 Pogonocherini Ecyrus dasycerus (Say) 1 1 2 Saperdini Saperda imitans (F.) 4 1 5 Saperda tridentata (Olivier) 1 0 1

Table 2.2. (cont’d)

Lepturinae Species Canopy Ground Total

Analeptura lineola (Haldeman) 1 1 2 Bellamira scalaris (Say) 1 0 1 Brachyleptura rubrica (Say) 0 1 1

Centrodera decolorata (Harris) 3 0 3 Lepturini Gaurotes cyanipennis (Say) 11 11 22 Stenocorus vittiger (Randall) 1 0 1 Stictoleptura canadensis (Olivier) 1 0 1 Strophiona nitens (Forster) 0 2 2 Trigonarthris minnesotana (Casey) 0 1 1 Necydalini Necydalis mellita (Say) 0 2 2 Parandrinae

Parandrini Neandra brunnea (F.) 6 2 8 Prioninae Prionini Orthosoma brunneum (Forster) 0 1 1

Total 1730 1915 3646

Watershed/Trap location Beetles collected Species richness Chao 2 estimate

Clinton 2017 72.5 ± 9.9 17 ± 2.3 45.7 2018 80.5 ± 22.2 17.5 ± 3.2 49

Total 153 ± 31 24 ± 3.3 68 Grand 2017 73.8 ± 11.3 15 ± 1.8 49.2 2018 56.8 ± 14.8 13.5 ± 0.9 24.3 Total 130.5 ± 22.3 19 ± 1.9 48.7 Kalamazoo 2017 88.3 ± 10 19.5 ± 0.6 48.1 2018 84.1 ± 8.1 19 ± 2 48.1 Total 172.4 ± 9.1 27.5 ± 1.6 88.3 Canopy

2017 74.3 ± 9.5 13.8 ± 1.1 84.4 2018 70.0 ± 11.7 14.2 ± 1 64.5 Total 144.3 ± 18.5 20.2 ± 1.4 109.6

Ground 2017 82.1 ± 8.2 10.7 ± 0.8 39.9 2018 77.6 ± 8.7 10.8 ± 0.9 51.8 Total 159.7 ± 14.3 14.5 ± 1.3 75.1

Main effect df F P Watershed 2 2.54 0.013 Trap location 1 5.8 0.001 Pairwise comparisons t P Clinton vs Grand 1.67 0.022 Clinton vs Kalamazoo 1.91 0.006 Grand vs Kalamazoo 1.05 0.315

Species NMS Indicator P Group value Indicators of watershed assemblages Clytus ruricola Clinton 59.1 0.01 Euderces picipes Clinton 60.9 0.005

Sarosesthes fulminans Clinton 56.8 0.037

Indicators of trap location assemblages Analaphus parallelus Canopy 74.5 0.011 Analaphus pumilus Canopy 46.7 0.0348 Sarosesthes fulminans Canopy 72.7 0.0084 Eburia quadrigeminata Canopy 66.7 0.002

Megacyllene caryae Canopy 68.8 0.0344

Neoclytus mucronatus Canopy 73.4 0.0104 Parelaphidion incertum Canopy 55.1 0.016 Phymatodes testaceous Canopy 61.1 0.0272 Clytus ruricola Ground 67.8 0.0402 Urgleptes querci Ground 67.1 0.039

Xylotrechus colonus Ground 68.4 0.0044

Figure 2.1. Locations of 12 trapping sites in three watersheds selected to represent the temporal gradient of the emerald ash borer invasion

Cyrtophorus verrucosus

Phymatodes aereus

Phymatodes testaceous

Clytus ruricola

Megacyllene caryae

Neoclytus a. acuminatus

Neoclytus mucronatus

Sarosesthes fulminans

Xylotrechus colonus

Cerambycinae Eburia quadrigeminata

Anelaphus parallelus

Anelaphus pumilus

Anelaphus villosus

Parelaphidion incertum

Euderces picipes

Astylopsis macula

Urgleptes querci

Aegomorphus modestus

Lamiinae Psenocerus supernotatus

Gaurotes cyanipennis

Leptur.

Figure 2.2. Seasonal activity, representing the period when individuals of each species (with ≥ 10 captured over the course of both years) were captured. Activity is displayed as the percentage of the total collected individuals during a range of accumulated GDD (base 10 °C). The monthly timeline along the bottom indicates the average accumulated GDD for each of the three st watersheds on the 1 of each month during 2017 and 2018.

2017 A 50 45 40 35 30 25 20 15 Canopy 10 5 Ground Number Species of Number 0 May June July Aug. Sept. Oct. 0 500 1000 1500 2000 May June July Aug. Sept.

Accumulated GDD (°C)

2018 B 50 45 40 35 30 25 20 15 Canopy 10 5 Ground

0 Number Species of Number 0 500 1000 1500 2000 May June July Aug. Sept. Oct. Accumulated GDD (°C)

A B

C D

E F Figure 2.4. Individual based rarefaction estimates comparing cerambycid captures by traps suspended from a mid-canopy branch vs. ground traps hung on 1.5 m tall rebar and among watersheds (Clinton/Grand/Kalamazoo) during (A,B) 2017, (C,D) 2018 and for both trapping seasons combined. (E,F).

APPENDIX B. RECORD OF DEPOSITION OF VOUCHER SPECIMENS

RECORD OF DEPOSITION OF VOUCHER SPECIMENS

The specimens listed below have been deposited in the named museum as samples of those species or other taxa, which were used in this research. Voucher recognition labels bearing the voucher number have been attached or included in fluid preserved specimens.

Voucher Number: 2019-04 Author: Patrick Engelken Title of Thesis: Vegetation, Coarse Woody Debris, and Cerambycid Communities in Riparian Forests Invaded by the Emerald Ash Borer Museum(s) where deposited: Albert J. Cook Arthropod Research Collection, Michigan State University (MSU) Specimens: Table A.1 Quantity and preservation method of Cerambycid species turned in as voucher specimens to the Albert J. Cook Arthropod Research Collection Family Genus-Species Life Stage Quantity Preservation Cerambycidae Aegomorphus modestus Adult 2 Pinned Cerambycidae Analeptura lineola Adult 2 Pinned Cerambycidae Anelaphus parallelus Adult 2 Pinned Cerambycidae Anelaphus pumilus Adult 1 Pinned Cerambycidae Anelaphus villosus Adult 1 Pinned Cerambycidae Astylopsis collaris Adult 1 Pinned Cerambycidae Astylopsis macula Adult 2 Pinned Cerambycidae Bellamira scalaris Adult 1 Pinned Cerambycidae Brachyleptura rubrica Adult 2 Pinned Cerambycidae Callimoxys sanguinicollis Adult 1 Pinned Cerambycidae Centrodera decolorata Adult 2 Pinned Cerambycidae Clytoleptus albofasciatus Adult 2 Pinned Cerambycidae Clytus ruricola Adult 2 Pinned Cerambycidae Cyrtophorus verrucosus Adult 2 Pinned Cerambycidae Dorcaschema cinereum Adult 1 Pinned

Table A.1 (cont’d) Cerambycidae Eburia quadrigeminata Adult 2 Pinned Cerambycidae Ecyrus dasycerus Adult 2 Pinned Cerambycidae Elaphidion mucronatum Adult 1 Pinned Cerambycidae Euderces picipes Adult 2 Pinned Cerambycidae Eupogonius pauper Adult 2 Pinned Cerambycidae Eupogonius subarmatus Adult 1 Pinned Cerambycidae Gaurotes cyanipennis Adult 2 Pinned Cerambycidae Heteracthes quadrimaculatus Adult 1 Pinned Cerambycidae Hyperplatys maculata Adult 2 Pinned Cerambycidae Knulliana cincta Adult 1 Pinned Cerambycidae Leptostylus transversus Adult 2 Pinned Cerambycidae Lepturges angulatus Adult 1 Pinned Cerambycidae Lepturges confluens Adult 1 Pointed Cerambycidae Lepturges symmetricus Adult 2 Pinned Cerambycidae Liopinus alpha Adult 1 Pinned Cerambycidae Liopinus punctatus Adult 1 Pointed Cerambycidae Megacyllene caryae Adult 2 Pinned Cerambycidae Microgoes oculatus Adult 2 Pinned Cerambycidae Micranoplium unicolor Adult 1 Pinned Cerambycidae Molorchus bimaculatus Adult 1 Pinned Cerambycidae Neandra brunnea Adult 2 Pinned Cerambycidae Necydalis mellita Adult 2 Pinned Cerambycidae Neoclytus acuminatus Adult 2 Pinned Cerambycidae Neoclytus mucronatus Adult 2 Pinned Cerambycidae Oplosia nubila Adult 2 Pinned Cerambycidae Orthosoma brunneum Adult 1 Pinned Cerambycidae Parelaphidion aspersum Adult 2 Pinned

Table A.1 (cont’d) Cerambycidae Parelaphidion incertum Adult 2 Pinned Cerambycidae Phymatodes aereus Adult 2 Pinned Cerambycidae Phymatodes amoenus Adult 2 Pinned Cerambycidae Phymatodes testaceus Adult 2 Pinned Cerambycidae Phymatodes varius Adult 1 Pinned Cerambycidae Psenocerus supernotatus Adult 1 Pointed Cerambycidae Purpuricenus paraxillaris Adult 2 Pinned Cerambycidae Saperda imitans Adult 2 Pinned Cerambycidae Saperda tridentata Adult 1 Pinned Cerambycidae Sarosesthes fulminans Adult 2 Pinned Cerambycidae Stenocorus vittiger Adult 1 Pinned Cerambycidae Stenosphenus notatus Adult 2 Pinned Cerambycidae Stictoleptura canadensis Adult 1 Pinned Cerambycidae Strophiona nitens Adult 2 Pinned Cerambycidae Trigonarthris minnesotana Adult 1 Pinned Cerambycidae Urgleptes foveatocollis Adult 1 Pointed Cerambycidae Urgleptes querci Adult 1 Pinned Cerambycidae Urgleptes signatus Adult 1 Pinned Cerambycidae Urographis fasciatus Adult 2 Pinned Cerambycidae Urographis triangulifer Adult 2 Pinned Cerambycidae Xylotrechus colonus Adult 2 Pinned

CHAPTER 3: RIPARIAN FOREST CONDITIONS ALONG THREE NORTHERN MICHIGAN RIVERS INVADED BY THE EMERALD ASH BORER (AGRILUS PLANIPENNIS)

Introduction

Since its accidental introduction into North America, emerald ash borer (EAB) (Agrilus planipennis Fairmaire (Coleoptera: Buprestidae), a destructive phloem-boring beetle native to parts of Asia, has continued to spread in North America. This invasive pest was initially identified in 2002 as the cause of ash decline in the Detroit, Michigan USA metropolitan area and in nearby Windsor, Ontario, Canada (Cappaert et al. 2005). Dendrochronological evidence, has shown that EAB was established in the Detroit area by the early 1990s and was killing ash

(Fraxinus spp.) trees in localized areas of southeast Michigan in the mid to late 1990s (Siegert et al. 2014). Populations of EAB are currently known to be established in 35 U.S. states and five

Canadian provinces (EAB.info 2019). Despite its fairly short invasion history, EAB has killed hundreds of millions of ash trees in eastern North America and has become the most destructive and costly forest pest to ever invade North America (Aukema et al. 2011, Herms and

McCullough 2014, Lovett et al. 2016, Morin et al. 2017).

Substantial research on the effects of EAB on North American forests has been conducted, particularly in areas of southeast Michigan and Ohio that were invaded by EAB relatively early (Burr and McCullough 2014, Flower et al. 2013, Kashian and Witter 2011,

Klooster et al. 2014, Robinett and McCullough 2019). Mortality rates of overstory ash trees in these areas have ranged from 79% to 99% (Burr and McCullough 2014, Kashian et al. 2018,

Klooster et al. 2014). Catastrophic levels of ash mortality following EAB invasion may alter forest composition, carbon cycling, and regeneration dynamics in ash dominated forests in much

of the eastern United States (Burr and McCullough 2014, Flower et al. 2013, Klooster et al.

2014, Costilow et al. 2017).

While all species of North American ash (Fraxinus spp.) may be suitable for EAB development (Herms and McCullough 2014), consistent interspecific differences in EAB host preference and host resistance have been observed (Anulewicz et al. 2007, 2008, Rebek et al.

2008, Tanis and McCullough 2015, Villari et al. 2016). White ash, a widely distributed species common in upland sites and mixed species stands, appears to be a moderately preferred EAB host (Robinett and McCullough 2019, Tanis and McCullough 2015), while blue ash is the least preferred species encountered by EAB in North America to date (Tanis and McCullough 2012).

Conversely, black ash (Fraxinus nigra Marsh.), and green ash (Fraxinus pennsylvanica Marsh) are highly preferred and vulnerable EAB hosts (Tanis and McCullough 2015). Black ash occurs over much of the north-eastern U.S. and eastern Canada (Gucker 2005a, Wagner and Todd 2015,

Wright and Rauscher 1990). While it is an uncommon species in upland forest habitats, black ash is often a dominant species in poorly drained forests such as lowland bogs and riparian corridors

(Gucker 2005a, Wagner and Todd 2015, Wright and Rauscher 1990). It can grow in mixed species stands with other trees that occur on hydric sites such as American elm, red maple, and northern white cedar, but also occurs in monocultures on sites with wet soils (Gucker 2005a,

Wagner and Todd 2015, Wright and Rauscher 1990). Green ash is the most widely distributed species of ash in North America, ranging from Texas to much of eastern Canada (Wagner and

Todd 2015). A highly adaptable species, green ash grows in a wide range of cover types and in habitats ranging from well-drained upland forests to hydric lowlands (Gucker 2005b, Kennedy

1990, Wagner and Todd 2015). It tends to be most abundant, however, in lowland forests

including poorly drained or swampy sites and in riparian forests bordering streams, rivers and lakes (Gucker 2005b, Kennedy 1990).

Loss of overstory ash trees as a result of the EAB invasion may be particularly significant in riparian forests and bogs, impacting not only terrestrial forest conditions but also affecting the aquatic environments which are functionally linked to these forests (Nisbet et al. 2015). Trees in riparian forests provide nutrients and substrate structure to the forest floor and adjacent waterways through inputs of leaf litter and coarse woody debris and affect light and water temperatures through shading (Lovett et al. 2004, Nisbet et al. 2015). In riparian forests where ash are dominant overstory species, cascading effects following EAB invasion may alter ecological functions in riparian forests as well as aquatic systems (Nisbet et al. 2015,

Kreutzweiser 2018, Youngquist et al. 2017). To date, however, little research has examined riparian forest ecosystems in the aftermath of the EAB invasion.

In riparian forests bordering first order streams, I surveyed canopy gaps resulting from green ash and black ash mortality in sites representing an east to west progression of EAB invasion across southern Michigan (Chapter 1). In these sites, dead ash trees were beginning to break and fall in higher numbers in sites in southeast Michigan with the longest history of EAB invasion than in more recently invaded sites in south central and southwest Michigan.

Additionally, in southern Michigan riparian forests, high light levels in canopy gaps affected understory vegetation dynamics including the apparent exclusion of tree seedlings in understories dominated by wetland sedge (Carex spp.) species. Despite nearly complete mortality of overstory ash trees and minimal seedling establishment, the canopy gaps included high densities of ash saplings and recruits, presumably established before overstory ash trees

were killed. Whether this advanced ash regeneration will survive and mature, given the persistence of EAB populations, remains unknown.

Small headwater streams are important tributaries to larger rivers and support several functions affecting larger bodies in river networks. Tributaries contribute nutrients, particulate organic matter and provide habitat for numerous invertebrates and vertebrates, including spawning areas for fish (Meyer et al. 2007). Communities in small streams such as headwater streams heavily rely on their surrounding environment for nutrient inputs and largely feed on coarse organic material such as leaf litter (Vannote et al. 1980). As stream and river size increase, species replacement and community adjustments occur to minimize energy loss and to capitalize on processing inefficiencies of upstream organisms (Vannote et al. 1980). This creates a species continuum where, as stream order increases, communities adjust to feed on smaller and finer organic material (Vannote et al. 1980). The origin of a large component of the fine organic material in larger streams and rivers therefore originates from leaf litter inputs into small streams

(Vannote et al. 1980). Mortality of ash trees bordering headwater streams may, therefore, have indirect impacts on larger rivers depending on size of drainage basins and channels, and the number of tributaries that feed it. Moreover, the riparian corridor bordering rivers will directly affect the aquatic environment (Broadmeadow and Nisbet 2004).

Several rivers in northwestern Michigan provide important habitat for aquatic organisms, including spawning habitat for Great Lake salmon and trout, and are heavily used for recreational fishing, kayaking and canoeing. In 2016, sportfishing alone contributed an estimated

64 million USD to northwest Michigan economies (michiganoutofdoors.com 2019).

Additionally, these rivers drain several large state and national forests and feed directly into the

Great Lakes. Many riparian forests along the rivers in northwest Michigan likely included an

ash component, but to date, conditions following EAB invasion of these areas have not been evaluated. I evaluated riparian forests along three ecologically and economically important rivers in northwest Michigan, all of which were invaded by EAB in the past decade. Objectives addressed in this study included evaluating the current and pre-EAB abundance of overstory ash and other tree species in these riparian forests. I was interested in determining if ash mortality from EAB resulted in canopy gaps, and if so, I wanted to characterize the frequency and extent of those canopy gaps along each river. I also assessed regeneration to determine whether ash species are likely to persist or if the overstory will be dominated by other species in these riparian forests.

Materials and Methods

Study Sites: Rivers selected for this study included the Betsie river, the Platte river and the Little

Manistee river (Figure 3.1). These rivers were selected based on their important role for Great

Lakes salmon fisheries and the economic value of water-based recreational activities, including kayaking, canoeing and tubing. All three rivers also include substantial stretches on public land, providing us with access to riparian forests. I used the Michigan Atlas and Gazetteer (DeLorme

2016) to identify the headwaters and termination of each river. The three rivers ranged considerably in linear length (30-100 km) but average channel widths were similar (~10 m) and they all drain into reservoirs that empty directly into lake Michigan.

I used time lapsed aerial images from Google Earth (Google Earth Pro 7.3 2019) to identify canopy gaps likely resulting from EAB-caused ash mortality in a 100 m buffer along the bank of the entire span of each river. Canopy gaps were identified by the appearance of an increasing proportion of dead trees visible during summer months (leaf on) over time, that occurred after the initial EAB quarantine for the given county. The perimeter of each canopy gap was outlined in Google Earth and uploaded into ArcMap (ESRI ARCGIS Desktop 10.6.1 2018) where a polyline was overlaid along the river, and river distance, gap area, and gap frequency were calculated.

Survey Design: I selected a 3-5 km stretch of each river where multiple canopy gaps occurred.

Using aerial images, I identified and recorded GPS coordinates of three canopy gaps: one located near the near the upstream beginning of the selected stretch of river, one near the middle, and one near the end of the selected stretch. In between each canopy gap, I identified and recorded

GPS coordinates of a segment of uninterrupted forest. Overall, the three canopy gaps alternating with the three areas of forest were as equidistant as possible along the selected length of each

river. All canopy gap and forest locations were separated by 200 to 800 m. I then accessed the locations along each river using kayaks. In each selected canopy gap or forested segment, I established a center point (25-50 m from the bank of the river) for concentric fixed radius plots including a macroplot (16.1 m radius), a subplot (8 m radius) and a microplot (1 m2). I also established three 25 x 2 m linear transects which extended from the center point in random but unduplicated directions. In addition to the surveyed canopy gaps, I entered and examined multiple canopy gaps along the accessed stretch of each river to determine whether gaps were formed from EAB caused ash mortality.

Overstory: I recorded species, DBH and canopy condition of all live and dead overstory trees (>

10.2 cm) in macroplots then standardized data by ha for analysis. Overstory live and dead basal area (m2 ha-1) was calculated for individual tree species and for all species combined. Relative importance values (RIV) were also calculated for overstory tree species. For a given species, the

RIV was calculated by summing its relative frequency (frequency of a species in plots as a proportion of the total frequency of all species), relative density (number of stems of a species as a proportion of stems of all species) and relative dominance (basal area of a species as a proportion of the total basal area for all species) in the surveyed stand (Kent and Coker 2011).

Regeneration: In subplots, I tallied species and number of recruits (diameter 2.4-10.2 cm and height >2 m), and saplings (diameter < 2.4 cm and height < 2 m). In microplots, tree seedlings

(woody stems < 2.4 cm diameter and <0.5 m height) were tallied by species or genera when necessary. Percent cover of herbaceous plant species in microplots was also recorded.

Coarse Woody Debris: Dead wood pieces ≥ 7.6 cm diameter were recorded by species (when possible) along linear transects and volume standardized by m3 per ha. For each piece of CWD, I measured length of the piece in the transect and diameter at the midpoint of the piece in the

transect. Decay status of each piece was visually estimated as follows: (1) bark firm, with few or no small areas of loose outer bark and wood solid; (2) bark loose or beginning to detach from one or more areas of log; outer wood in early stages of decay with areas spongy in response to pressure; (3) bark mostly loose or sloughed off, outer wood spongy and some broken but most inner wood solid; and (4) little to no bark remaining, wood spongy and crumbling.

Statistical Analysis: Differences in variables between canopy gaps and forests and among the three sampled rivers were analyzed using a 2 factor ANOVA in SAS 9.4 (Proc GLIMMIX; SAS institute Inc. 2015). All variables were tested for normality and heterogeneity of variance using histograms, normal probability plots, and side by side boxplots of residuals. Variables that did not meet normality assumption were normalized using data transformations. Pre-EAB basal area of ash was log transformed, and total current live basal area was square root transformed. When

ANOVA results were significant, least square means were compared using a Student’s t test. Due to the small number of comparisons made for each analysis, multiple comparison procedures were not utilized in order to reduce the chance of type 2 error. Relationships between herbaceous plant cover and densities of seedlings, saplings, and recruits were evaluated with simple linear regression (Proc Reg; SAS Institute Inc. 2015).

To examine and assess species compositions of forest overstories prior to the EAB invasion and to compare pre-EAB overstory species to the current composition of regeneration, a multivariate regression tree (MRT) analysis was performed using the mvpart package (De’ath

2011) in R version 3.5.1 (R Development Core Team, 2018). Multivariate regression trees are hierarchical clustering models which divide variables successively into branches or groupings based on maximum homogeneity of the number of explanatory variables (De’ath 2002).

Advantages of this analysis include relatively few underlying assumptions about data distribution

and it often fits ecological and assemblage data well (De’ath 2002, Larsen and Speckman 2004).

While I compared pre-EAB overstory species composition in canopy gaps and adjacent forests to current recruit and sapling species, seedlings were rarely encountered in any of the canopy gaps and were therefore excluded from the MRT analysis. Data were log transformed prior to analysis to correct for normality. A final regression tree was selected based on cross validation results, where the regression tree was pruned to the smallest tree that fell in one standard deviation of the minimum cross-validated relative error.

Results

Gap Size and Frequency: River length varied substantially, ranging from the Little Manistee river (100.6 km) followed by the Betsie (82.6 km) and the Platte river (34.1 km) (Table 3.1). I identified and outlined a total of 419 canopy gaps, ranging from 0.14 to 25 ha in area, along the rivers. Average frequency of canopy gaps per linear km varied but did not differ among rivers

(F=2.19; df=2,16; P=0.144) (Table 3.1). Average gap size, however, was different among rivers

(F=17.01; df=2,16; P<0.001). Largest gaps occurred along the Platte river, while the Betsie and the Little Manistee rivers had smaller and similarly sized gaps. When I examined areas delimited by a 100 m buffer along both banks of the rivers, 417 of the 419 gaps fell completely in this range, with only 2 gaps having a portion of their area extending past 100 m of the bank . In 100 m of the river bank, the percentage of this area represented by canopy gaps resulting from dead ash trees did not differ among rivers (F=0.94; df=2,16; P=0.41). These canopy gaps comprised an average of 20.7 ± 3.6%, 20.4 ± 5.8%, and 13.3 ± 2.6% of the riparian forests bordering the

Betsie, Platte, and Little Manistee rivers, respectively. When I extended past the 100 m buffer no new canopy gaps were identified, and only two of the 419 streamside gaps extended past the 100 m buffer, indicating that ash distribution was largely confined to the river corridors.

Overstory Vegetation: I recorded 358 trees representing 19 species in the nine canopy gaps and nine forested areas I surveyed along the three rivers. Overstory basal area (standing live and dead trees) (F=0.8; df=2,12; P=0.47) and stem density (F=0.08; df=2,12; P=92) were similar among rivers and between canopy gaps and forests (F=4.09; df=1,12; P=0.066) (F=3.99; df=1,12;

P=0.07), but live basal area (F=11.82; df=1,12; P=0.005) and stem densities (F=45.7; df=1,12;

P<0.0001) were substantially higher in forests than in canopy gaps (Table 3.2). Conversely, while basal area of standing dead trees did not differ among rivers (F=3.28; df=2,12; P=0.073),

stem densities of dead trees were higher along the Betsie river, than in the Platte and Little

Manistee rivers (F=8.14; df=2,12; P=0.01). Additionally, dead trees were more abundant in canopy gaps (8.3 ± 0.9 m2 · ha) than in adjacent forests (2.4 ± 0.7 m2 · ha) (F=29.72; df=1,12;

P<0.0001). This was primarily due to the dead ash trees, which accounted for 78.0% of the 158 standing dead trees in gaps versus 66% of the 52 standing dead trees in forests. Species richness of live trees did not differ between canopy gaps and the adjacent forests (F=0.72; df=1,12;

P=0.413), but was different among rivers (F=6.26; df=2,12; P=0.014). I tallied fewer species along the Betsie river (2.5 ± 0.8 species) than in either the Platte (5.7 ± 0.7 species) or Little

Manistee rivers (5.2 ± 0.6 species).

Species composition varied among rivers, but maples (Acer spp.) were consistently among the top three dominant species based on RIVs in canopy gaps and forested plots along all rivers. Northern white cedar (Thuja occidentalis L.) had a high RIV in forests of all rivers and in canopy gaps along the Platte and Little Manistee rivers (Table 3.3). Eastern white pine (Pinus strobus L.) was encountered in the forests along the Platte and Little Manistee rivers, but was only abundant in forests adjacent to gaps along the Little Manistee river (Table 3.3). Silver maple (Acer saccharinum L.) was by far the dominant species in the forests along the Betsie river with an RIV score of 230 (a score of 300 is a monoculture). Silver maple was also a dominant species in canopy gaps along the Betsie river, but three other species were also common (Table 3.3). Platte river forests were dominated by sugar maple (Acer saccharum

Marsh.), while eastern hemlock (Tsuga canadensis (L.) Carrière) and bigtooth aspen (Populus grandidentata Michx.) were somewhat common. American basswood (Tilia americana L.) dominated canopy gaps along the Little Manistee river, while quaking aspen (Populus tremuloides Michx.) dominated canopy gaps along the Platte river (Table 3.3).

Overstory Ash: Overall, I recorded 138 standing ash trees (>10.2 cm DBH) in the nine canopy gaps and 39 ash trees in forests between gaps. This included 35 black ash, 129 green ash, and 13 white ash (Fraxinus americana L.). Size of standing ash trees was similar among species with average DBH of 19.0 cm, 19.0 cm and 18.6 cm for black, green and white ash respectively. I recorded 13 green ash trees with a DBH greater than 30 cm but none of the black or white ash trees were larger than 30 cm DBH. Only six live ash trees were tallied, and all were white ash.

Five live white ash were recorded in one Platte river canopy gap and one live white ash occurred in a plot in the Little Manistee river forest.

Estimates of pre-EAB conditions, including the sum of live and dead ash trees and fresh ash CWD with EAB galleries, indicated that ash species comprised an average of 71.0 ± 14.6%,

46.2 ± 1.3%, and 65.2 ± 7.5% of the total basal area in canopy gaps along the Betsie, Platte, and

Little Manistee rivers, respectively, while accounting for less than 10% of the pre-EAB basal area in the forests between gaps. More than 95% of the estimated pre-EAB ash basal area along the three rivers was dead. While pre-EAB ash basal area was similar among sites along the three rivers (F=3.01; df=2,12; P=0.087), pre-EAB ash basal area was consistently higher in canopy gaps (13.6 ± 3.5 m2.ha-1) than in the adjacent forests (2.8 ± 1.1 m2.ha-1) (F=11.56; df=1,12;

P=0.005).

Most dead ash trees remain standing in all the sites I surveyed. On average, 71.1 ±

10.0%, 90.9 ± 7.0%, and 55.9 ± 11.2 % of dead ash basal area is still standing along the Betsie,

Platte, and Little Manistee rivers, respectively, The proportion of dead ash trees still standing did not differ between canopy gaps (66.8 ± 11.8%) and adjacent forests (82.9 ± 7.5%) (F=1.79; df=1,12; P=0.205) but did differ among rivers (F=4.15; df=2,12; P=0.043). Sites along the Little

Manistee river had a lower proportion of ash trees still standing than sites along the Betsie and

Platte rivers.

Coarse Woody Debris: Coarse woody debris was encountered in transects in all plots along each river. I tallied 155 pieces of CWD, with diameters ranging from 7.6 cm to 57.6 cm and averaging

16.0 ± 2.3 cm. Volume of CWD was similar among rivers (F=0.93; df=2,12; P=0.423) and between canopy gaps and adjacent forests (F=0.10; df=1,12; P=0.752) (Table 3.2). On average,

CWD was less decayed in canopy gaps, where 77.4 ± 9.9% of the volume was relatively fresh

(decay classes 1-2), than in adjacent forests where 79.0 ± 10.2% of CWD volume was classified as advanced decay (decay classes 3-4).

I was able to identify 74.1% of CWD pieces to species and tallied at least 12 tree species.

In canopy gaps, CWD was primarily ash (67.6 ± 14.0% of volume), while ash CWD was not common in forested plots (7.4 ± 3.8% of volume). The proportion of CWD represented by ash was lower in sites along the Platte River (30.2 ± 9.8%) than in sites along the Betsie (52.5 ±

20.3%) and Little Manistee rivers (46.4 ± 19.7%) (F=7.03; df=2,12; P=0.01). Canopy gaps had a higher proportion of CWD represented by ash than in forested plots (F=200.19; df=1,12;

P<0.0001).

Herbaceous Plant Cover: Percent cover of herbaceous plants differed among rivers (F=10.38; df=2,12; P=0.0024). Plots along the Betsie (64 ± 16% percent cover) and Little Manistee (47 ±

17% cover) rivers had denser understory vegetation than plots along the Platte river (27 ± 12% cover). Additionally, vegetation in canopy gaps (77 ± 7% cover) was much more dense than in the forested plots (14 ± 7% cover) (F=88.43; df=1,12; P<0.0001). In canopy gaps, understories were dominated by species of sedge (Carex spp.), primarily awl fruited sedge (Carex stipata

Muhl.) and woolly sedge (Carex pellita Muhl.). In forests adjacent to canopy gaps, the majority

of plots had low densities of plant cover (five of the nine forest plots had ≤ 5% cover). In forested plots, common species encountered included interrupted fern (Osmunda claytoniana L.), cinnamon fern (Osmunda cinnamomeum L.), stinging nettle (Urtica dioica L.), and wintergreen

(Gaultheria procumbens L.). However, no individual herbaceous plant species comprised more than 10% of the understory in any forested plot.

Seedlings: I recorded 131 seedlings representing nine species of trees along the three rivers

(Table 3.4). Seedling densities were higher in forests (13.6 ± 5.5 per m2) than in canopy gaps

(1.3 ± 0.6 per m2) (F=5.99; df=1,12; P=0.01), but were similar among rivers (F=1.42; df=2,12;

P=0.28). Species richness of tree seedlings did not differ between canopy gaps and the adjacent forests (F=3; df=1,12; P=0.109), but on average, sites along the Betsie river had fewer species

(0.5 ± 0.2) than the Platte (2.3 ± 0.5) and Little Manistee rivers (1.7 ± 0.7) (F=7.58; df=2,12;

P=0.007). Densities of seedlings were inversely related to herbaceous plant cover (R2=0.308;

P=0.017) (Figure 3.2 A) and in five of the nine canopy gaps, herbaceous cover exceeded 90% and no tree seedlings were observed. Only 13 seedlings were recorded in the nine canopy gaps.

Five were green ash, which was the most common species and was the only species found along all three rivers. In forests, maple seedlings were most abundant. Silver maple accounted for

53.8% of the total number of seedlings recorded in forests between gaps but was encountered only along the Betsie river. Red maple accounted for 36.1% of all seedlings in forests and was recorded along all three rivers. Green ash seedlings were occasionally recorded in forests along the Platte and Little Manistee rivers, but accounted for just 4.2% of the total seedlings recorded.

Black ash was nearly absent, with only a single seedling observed in a Platte river canopy gap.

Saplings: I tallied 159 saplings (diameter >2.4 cm, height <2 m) representing 12 species along the three rivers (Table 3.5). Densities of saplings were similar among rivers (F=2.96; df=2,12;

P=0.091) but were much higher in canopy gaps (685.2 ± 251.5 per ha) than in the adjacent forests (193.4 ± 54.6 per ha) (F=5.28; df=1,12; P=0.04). Species richness of saplings, however, was similar between canopy gaps and forests (F=2.06; df=1,12; P=0.176) and among rivers

(F=3.52; df=2,12; P=0.063). There was no evidence of a linear relationship between herbaceous cover and sapling densities (R2=0.012; P=0.672) (Figure 3.2 B). Saplings were consistently dominated by green ash in canopy gaps and forests. Green ash accounted for 49.1% of all saplings recorded but was more than four times more abundant in canopy gaps (431.0 ± 201.3 per ha) than in forests between canopy gaps (99.4 ± 45.4 per ha). Black ash and black cherry were also tallied in gaps, but each accounted for only 6.5% of all saplings recorded. In forests between canopy gaps, green ash was the most abundant species along the Betsie and Platte rivers while eastern white pine was the most abundant sapling in forests along the Little Manistee river.

Recruits: I tallied 154 recruits (2.5-10.2 cm diameter and height >2 m) from 19 tree species along the three rivers. Recruit densities did not differ between canopy gaps (331.6 ± 90.0 per ha) and forests (463.6 ± 117.9 no. per ha) (F=1.04; df=1,12; P=0.328) or among sites along the

Platte (522.2 ± 73.5 per ha) the Betsie (297.6 ± 85.2 ha) and Little Manistee rivers (373.0 ±

193.8 per ha) (F=1.04; df=2,12; P=0.384). Species richness of recruits was similar between canopy gaps and the forests between gaps (F=0.96; df=1,12; P=0.346), but on average, sites along the Platte river had more species (4.7 ± 0.6) than sites along the Betsie (1.7 ± 0.3) and

Little Manistee rivers (2.2 ± 0.5) (F=10.73; df=2,12; P=0.002). Similar to saplings, there was no linear relationship between recruit densities and herbaceous plant cover (R2=0.1836; P=0.076)

(Figure 3.2 C). Green ash and red maple (Acer rubrum L.) recruits were encountered along every river and green ash was the most abundant species overall, accounting for 27% of tallied recruits.

Green ash recruits were especially abundant in canopy gaps along all rivers, accounting for 53%

of the recruits, and was the most abundant species in canopy gaps. In forests between gaps, silver maple recruits were the most abundant in sites along the Betsie river, and sugar maple recruits were the most common in sites along the Platte river. Eastern white pine recruits were abundant in forests between gaps along the Little Manistee river, accounting for 79.2% of all recruits tallied in these sites.

Species Composition of Pre-EAB Overstories and Current Regeneration: To assess species composition of pre-EAB forests and to gain insight into potential future forest conditions, I used multivariate techniques to compare pre-EAB overstory composition and the corresponding recruit and sapling regeneration in the sites surveyed along the three rivers. Cross validation of

1000 multivariate regression trees yielded a 10 branch tree most often (360), which explained

46.9% of the variability in the data (Figure 3.3). The first two nodes (a node is any point where a split occurs in a regression tree) separated the three rivers and explained 16.6 % of the variability. The most pronounced differences between pre-EAB overstory vegetation and current regeneration occurred among the three sampled rivers and between canopy gaps and forested plots along each river.

In the Betsie river branch of the regression tree, there were two nodes that split into three terminal nodes or (leaves). The first separation occurred between canopy gaps and forests and the second split divided the overstory and regeneration (recruits and saplings) in intact forests.

Canopy gaps species compositions were dominated by green ash in the Pre-EAB overstory and in current recruits and saplings. The pre-EAB forests along the Betsie river were dominated by silver maple, black ash and green ash in the overstory. Understory regeneration consisted mostly of silver maple and green ash, while black ash regeneration was substantially less common than in the pre-EAB overstory.

The Little Manistee river branch of the regression tree consisted of two nodes; the first between the pre-EAB overstory and current recruits and saplings, and the second between pre-

EAB overstories in canopy gaps and intact forests. Regeneration was dominated by eastern white pine and green ash in canopy gaps and in intact forests. The pre-EAB overstories in canopy gaps were comprised largely of green ash, black ash and American basswood, while overstories in intact forests were generally dominated by eastern white pine and red maple, with occasional green ash trees.

The Platte river branch of the regression tree consisted of three nodes. The first delineated pre-EAB overstories from current regeneration (recruits and saplings) and the second and third nodes separated species compositions of overstory and regeneration between canopy gaps and adjacent forests. Regeneration in intact forests along the Platte river was comprised of sugar maple and green ash, while regeneration in canopy gaps was predominantly green ash. Pre-

EAB overstories in forests between canopy gaps consisted of sugar maple, eastern hemlock and bigtooth aspen, while pre-EAB overstories in canopy gaps included substantial green ash, along with quaking aspen and red maple.

Discussion

Riparian forest compositions along three northern Michigan rivers were distinctly different, varying in both species richness and species assemblages, but all included frequent pockets dominated by green and black ash prior to the EAB invasion. More than 95% of those ash trees have died. All of the canopy gaps I identified and attributed to EAB-caused ash mortality using time lapsed aerial images proved to be accurate, and this may be a useful tool for monitoring ash decline in future surveys. Ash distribution clearly followed the riparian corridors in these river systems. My estimates suggest 13-21% of the basal area with the riparian forests along the rivers now consists of dead ash trees. Ash trees occurred occasionally in the forests between the gaps but nearly all of these trees were also dead.

Mortality of overstory ash trees represents a substantial disturbance of streamside forest habitat that likely threatens the integrity of riparian buffers. Buffers along riparian forest corridors stabilize stream banks and maintain channel depth by preventing erosion, provide continual nutrient flux via annual leaf litter, moderate water temperatures through shading and increase habitat and structure through inputs of coarse woody debris (Broadmeadow and Nisbet

2004). In pre-EAB riparian corridors dominated by eastern hardwood species, annual leaf fall from ash trees reportedly contributes a significant proportion of the total litter into ecosystems

(Kreutzweiser et al. 2018). High quality ash leaves are considered to be an important source of litter because of their low lignin: nitrogen ratios and nutrient turnover rates in soil are faster for ash than for other common hardwood species (Melillo et al. 1982). Additionally, increased amounts of ash litter in forests have been correlated with increased availability of soil nitrogen, organic carbon, and exchangeable cations (Mg2+,Ca2+) (Langenbruch et al. 2012). Given that ash represented up to 70% of the pre-EAB basal area in the current canopy gaps I evaluated, the loss

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of ash litter has likely affected nutrient availability in terrestrial soils and perhaps in the adjacent rivers. In riparian corridors, macroinvertebrates utilize leaf litter as both a food source and as microhabitat (Wallace and Webster 1996). Additionally, the bulk of in-stream organic material originates from inputs of leaf litter (Vannote et al. 1980). Therefore, the loss of a major contribution of the annual leaf litter in riparian forests may negatively affect invertebrate abundance and diversity (Wallace et al. 1997).

Reduced shading of streams is associated with increased water temperatures in riparian channels (Broadmeadow et al. 2011) and could affect commercially important species of salmonid fish who spawn in the rivers I surveyed. Increased water temperature in spawning streams of salmonids has been correlated with delayed migrations, reduced viability of eggs, and increased egg mortality (Richter and Kolmes 2005). Additionally, increased shading in streams and rivers can limit algal primary productivity potentially reducing eutrophication (Burrell et al.

2013, Halliday et al. 2016). To date, I am aware of little research addressing potential indirect effects of EAB invasion on aquatic environments and the implications of changes in riparian forest conditions are largely unknown.

Most (75%) of the dead ash trees killed by EAB remain standing along the three rivers and the ash CWD I encountered was relatively fresh. In riparian forests along first order streams across southern Michigan, I observed a linear relationship between the length of the post-EAB invasion period and CWD accumulation (Chapter 1). Abundance of CWD will increase annually along the three northern rivers as dead ash trees fall and if patterns in these northern rivers are similar to those in southern Michigan, most of the dead trees will fall in the next 5-6 years.

Accumulation of CWD in post-EAB invasion forests may influence abundance, activity and species compositions of ground dwelling insects (Perry and Herms 2016, Ulyshen et al. 2011)

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and woodboring cerambycid beetles (Chapter 2). Some of the ash in the riparian forests are likely to fall into the rivers, which could impede canoes and kayaks. On the other hand, coarse woody debris in streams provides complex habitat, refuge from swift moving currents, and retains organic matter, all of which may benefit some invertebrate and fish species (Wondzell and

Bisson 2003). In riparian forests invaded by hemlock wooly adelgid (HWA) (Adelges tsugae

Annand), increased volume of coarse woody debris volume as dead eastern hemlock trees fell into streams was correlated with higher invertebrate abundance and diversity (Pitt and Batzer

2015). Similar dynamics may occur in these rivers as ash trees decompose and fall.

Regeneration dynamics along the three northern Michigan rivers were notably similar to those observed along first order streams in southern Michigan (Chapter 1). In canopy gaps, green ash dominated advanced regeneration, i.e., saplings and recruits that were presumably established before EAB killed the ash overstory trees. Tree seedlings, however, were rare, and in many sites, I encountered no seedlings in my plots, nor did I observe seedlings in the vicinity of the plots. As in the canopy gaps along the headwater streams in southern Michigan, tree seedlings in gaps along the rivers appear to have been excluded by thick mats of wetland sedges

(Carex spp.). Prevalence of sedge meadows may also reflect higher water tables following overstory ash mortality. In black ash dominated wetlands, water tables rose significantly, and understory composition shifted towards wetland sedges when the overstory ash were clearcut or girdled in an effort to simulate EAB impacts (Slesak et al. 2014, Van Grinsven et al. 2017). In these studies, however, overstory trees were killed or felled during a single winter, which differs considerably from the multi-year decline and eventual mortality of ash trees following EAB invasion. Nevertheless, mortality of overstory ash trees and the associated reduction in evapotranspiration, combined with higher light availability in the canopy gaps I evaluated could

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affect both water table levels and understory vegetation dynamics. In the absence of restoration efforts, remnant ash in the sapling and recruit strata represent the only ash regeneration along the rivers I surveyed. Whether these trees persist and mature despite the continued presence of EAB may determine whether ash remains a functional component of the riparian forests and may also represent the last chance for re-establishment of overstory trees in the current canopy gaps.

In forests adjacent to canopy gaps, species composition of seedlings generally reflected the tree species in the overstory. Seedlings in the intact forests were dominated by maple species along the Betsie and Platte rivers, while eastern white pine dominated regeneration in forests along the Little Manistee river. Green ash saplings were relatively abundant in forests between the canopy gaps along the three rivers but whether the young saplings will reach the overstory remains to be seen. Green ash recruits occurred in forests along the Platte and Little Manistee rivers, but were completely absent in the forested areas between current canopy gaps along the

Betsie river. Whether green ash recruits or saplings will persist long enough to mature and produce seed is unknown. Emerald ash borer can colonize and kill trees down to 2.5 cm in diameter (Cappaert 2005), and while biocontrol efforts have shown some promise in protecting small trees (Duan et al. 2017), remnant ash remain vulnerable. In riparian forests already impacted by EAB, the last hope for maintaining ash as a viable overstory species may require protecting advance ash regeneration as it matures.

In many northern wetland sites across the Great Lakes region of the U.S., black ash is considered a foundational species and is often a dominant or a codominant species (Youngquist

2017). Data from the current canopy gaps I surveyed showed that prior to EAB invasion, black ash, while not as abundant as green ash, was a major overstory species along the Betsie and

Little Manistee rivers, comprising an average of 26 ± 1% and 18% ± 2% of the pre-EAB

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overstories, respectively. Black ash regeneration, however, was rare in canopy gaps and the forests between gaps along all three rivers. I recorded no black ash recruits, only eight saplings and a single black ash seedling. My data indicate that black ash is likely to be lost from the riparian forests I evaluated. Given that black ash is a highly preferred and vulnerable EAB host

(Tanis and McCullough 2015), the high mortality rates and lack of regeneration documented here and in other studies suggest the future of this species across its range in North America is uncertain.

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APPENDIX

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Table 3.1. County, date of emerald ash borer quarantine, total linear length of river, number of canopy gaps caused by EAB- killed ash trees, average (± SE) frequency and size of canopy gaps, and current percentage (± SE) of forest comprised of canopy gaps in 100 m of the three north rivers surveyed.

River County EAB River length No. of Gap Average gap Percent of area comprised quarantine (km) Gaps (N) frequency size (Ha) of canopy gaps in 100 m imposed (No. ∙ km-1) buffer Betsie Benzie 2005 82.6 199 6.7 ± 0.9 1.6 ± 0.2 a 20.7 ± 3.6% Platte Benzie 2005 34.1 45 3.6 ± 1.4 3.2 ± 0.8 b 20.4 ± 5.8% Little Manistee Manistee 2004 100.6 175 7.1 ± 1.2 1.0 ± 0.2 a 13.3 ± 2.6%

Total 217.3 419 3.8 ± 0.4 1.8 ± 0.3 18.7 ± 2.3%

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Basal Area (m2 ha-1) Betsie Platte Little Manistee P value Canopy gap Forest P value

Live trees 12.7 ± 6.3 12.0 ± 2.5 15.4 ± 4.0 0.77 6.4 ± 1.1 20.3 ± 3.6 0.005

Dead trees 7.2 ± 1.5 3.7 ± 1.7 5.0 ± 1.5 0.073 8.3 ± 0.9 2.4 ± 0.7 <0.0001

Pre-EAB ash 9.4 ± 2.5 10.2 ± 2.2 12.2 ± 5.4 0.087 13.6 ± 3.5 2.8 ± 1.1 0.005

Current ash 0 0 0 -NA- 0 0 -NA-

Stem density (no. ha-1)

Live trees 271.9 ± 46.9 302.9 ± 60.7 257.9 ± 58.0 0.68 135.1 ± 20.4 353.4 ± 34.7 <0.0001

Dead trees 206.7 ± 41.1 b 83.9 ± 25.5 a 139.2 ± 47.0 a 0.01 215.6 ± 25.9 71.0 ± 23.3 <0.0001

Pre-EAB ash 311.4 ± 64.6 373.4 ± 75.8 510.8 ± 233.3 0.11 514.2 ± 142.3 82.8 ± 32.9 <0.0001

Current ash 0 0 0 -NA- 0 0 -NA-

Coarse woody debris volume (m3 ha-1) Fresh ash logs 8.3 ± 3.4 8.4 ± 0.4 19.1 ± 9.9 0.076 17.6 ± 6.5 0.9 ± 0.9 0.0003 (Decay 1-2) Fresh total logs 9.9 ± 2.9 10.6 ± 6.4 25.4 ± 9.7 0.11 25.0 ± 6.5 5.7 ± 2.6 0.01 (Decay 1-2) Decayed total logs 20.4 ± 10.8 13.2 ± 4.9 14.0 ± 4.7 0.55 6.8 ± 2.8 24.9 ± 6.5 0.01 (Decay 3-4)

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Table 3.3. Number of live overstory trees (DBH > 10.2 cm), mean (± SE)* diameter at breast height (DBH), total live basal area, and relative importance values (RIVs) for the five most dominant overstory species recorded in nine canopy gaps and nine forested areas in each of three rivers in northern lower Michigan. Canopy Gaps Forests No. DBH (cm) Basal area RIV1 No. DBH (cm) Basal area RIV1 trees (m2·ha-1) trees (m2∙ha-1) Betsie River Betsie River Acer saccharinum 5 28.0 1.6 85.1 Acer saccharinum 53 28.8 ± 6.9 21.0 ± 10.9 229.7 Thuja occidentalis 7 15.1 0.5 70.5 Acer rubrum 6 23.0 1.1 31.3 Ulmus americana 5 17.1 0.5 60.1 Ulmus americana 2 13.0 0.1 20.4 Betula alleghaniensis 3 20.9 0.4 50.0 -NA- -NA- -NA- Platte River Platte River Populus tremuloides 13 33.3 ± 2.5 4.9 ± 1.4 110.1 Acer saccharum 44 17.9 ± 2.5 5.3 ± 2.3 90.8 Acer rubrum 8 14.9 ± 1.5 0.7 ± 0.3 46.7 Tsuga canadensis 11 28.8 ± 4.1 2.9 ± 0.7 44.6 Acer saccharum 7 19.3 1.5 41.9 Populus grandidentata 16 21.1 2.7 43.0 Thuja occidentalis 4 21.1 0.9 30.0 Acer rubrum 9 19.3 1.2 26.5 Tsuga canadensis 4 19.8 0.6 29.7 Thuja occidentalis 8 19.7 1.2 25.6 Little Manistee River Little Manistee River Tilia americana 17 23.6 ± 3.6 4.0 ± 1.9 122.7 Pinus strobus 44 25.0 ± 5.3 9.8 ± 2.4 107.0 Acer rubrum 6 24.3 1.2 48.9 Thuja occidentalis 17 29.1 ± 1.8 4.9 ± 0.7 51.1 Thuja occidentalis 3 36.1 1.3 40.6 Acer rubrum 15 26.0 ± 3.1 3.5 ± 1.3 48.1 Betula payrifera 3 21.8 0.5 30.7 Quercus bicolor 2 35.6 0.8 16.9 Tsuga canadensis 2 24.9 0.5 19.2 Betula papyrifera 3 16.5 0.3 15.6 1Relative importance values (RIVs) represent the sum of the frequency, density and dominance of a specific species relative to all other species.

*Standard error bars are provided where applicable. Some canopy gaps contained no live trees and not all species occurred in every plot.

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Betsie River Platte River Little Manistee River Species Gap Forest Gap Forest Gap Forest Total Recorded Acer saccharinum 0 21.3 0 0 0 0 64 Acer rubrum 0 4 0 1.3 0 9 43 Fraxinus pennsylvanica 0 0 1.3 0.3 0.3 1.3 10 3 Prunus serotina 0 0 0.3 0.3 0 0 Ostrya virginiana 0 0 0.6 0.3 0 0 3 Pinus strobus 0 0 0 0.3 0 0.6 3 Quercus rubra 0 0 0 0 0 0.7 2 Betula alleghaniensis 0 0 0.7 0 0 0 2 Fraxinus nigra 0 0 0.3 0 0 0 1 Total 0 25.3 3.2 2.5 0.3 11.6 131

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Betsie River Platte River Little Manistee River Recruit Species Gap Forest Gap Forest Gap Forest Total Fraxinus 39 281.8 ± 158.2 0 ± 0 232.1 ± 59.8 49.7 ± 49.7 156.5 ± 59.6 66.3 ± 43.9 pennsylvanica Pinus strobus 0 ± 0 0 ± 0 16.6 ± 16.6 33.2 ± 33.2 0 ± 0 563.7 ± 346.2 37 Acer saccharum 0 ± 0 0 ± 0 33.2 ± 16.6 165.8 ± 33.2 0 ± 0 0 ± 0 12 Betula alleghaniensis 16.6 ± 16.6 82.9 ± 82.9 0 ± 0 82.9 ± 82.9 0 ± 0 0 ± 0 11 Ostrya virginiana 0 ± 0 0 ± 0 116.1 ± 116.1 0 ± 0 0 ± 0 16.6 ± 16.6 8 Acer rubrum 0 ± 0 66.3 ± 66.3 0 ± 0 33.2 ± 33.2 0 ± 0 16.6 ± 16.6 7 Fraxinus americana 0 ± 0 0 ± 0 99.5 ± 99.5 16.6 ± 16.6 0 ± 0 0 ± 0 7 Acer saccharinum 0 ± 0 97.8 ± 28.8 0 ± 0 0 ± 0 0 ± 0 0 ± 0 5 Other 49.7 ± 49.7 0 ± 0 198.9 ± 86.2 82.7 ± 33.2 49.7 ± 28.7 33.1 ± 33.1 23 Total 348.1 ± 151.9 247.0 ± 103.1 580.2 ± 82.9 464.2 ± 129.5 216.3 ± 53.1 697.7 ± 304.4 154 Sapling Species Fraxinus 96 pennsylvanica 314.8 ± 217.3 149.1 ± 86.1 478.1 ± 135.8 132.6 ± 108.7 325 ± 118.9 16.6 ± 16.6 Pinus strobus 0 ± 0 0 ± 0 0 ± 0 33.2 ± 33.2 182.5 ± 158.3 166.0 ± 87.9 23 Prunus serotina 0 ± 0 0 ± 0 132.6 ± 132.6 16.6 ± 16.6 0 ± 0 0 ± 0 9 Fraxinus nigra 82.8 ± 16.6 0 ± 0 49.7 ± 28.7 0 ± 0 0 ± 0 0 ± 0 8 Other 99.4 ± 75.9 32.3 ± 32.3 215.5 ± 108.7 33.1 ± 33.1 0 ± 0 0 ± 0 23 Total 497.0 ± 216.6 182.2 ± 116.0 776.0 ± 551.9 215.5 ± 116.1 482.5 ± 158.3 182.5 ± 92.4 159

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Figure 3.1. Location of surveyed rivers and associated watersheds in northern lower Michigan.

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100 y = -3.7038x + 64.558 90 R² = 0.3084 80 Gap 70 Forest 60 50 40 30 20 10 0 0 5 10 15 20 2

Percent coverhervaceousof plants Seedling density (No. per m )

B. y = 0.007x + 42.769 100 R² = 0.0115 90 Gap 80 Forest 70 60 50 40 30 20 10 0 0 500 1000 1500 2000 2500

Percent coverherbaceousof plants Sapling density (No. per ha-1) C. 100 y = -0.0523x + 66.627 R² = 0.1836 90 80 Gap 70 Forest

60 50 40 30 20 10 0

Percent coverherbaceousof plants 0 500 1000 1500 Recruit density (No. per ha-1) Figure 3.2. Linear relationship between percent cover of herbaceous plants and density of (A) seedlings, (B) saplings, and (C) recruits in three canopy gaps and three forested plots in each of three surveyed rivers in northern Michigan.

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Betsie River Platte River, Little Manistee River River< 1.5 River>=1.5

Little Manistee River Platte River Cover< Forest 1.5 Cover>=1.5Gap River>=2.5 River< 2.5

Recruits, Overstory Status>=1.5Saplings Status< 1.5 Recruits, Recruits, Saplings Overstory Saplings Overstory 209(n=9 : n=9) Status>=1.5 Status< 1.5 Status>=1.5 Status< 1.5

110 : n=5 34.9 : n=3 (n=6) (n=3)

Forest Gap Forest Gap Forest Cover< 1.5 Cover>=1.5 Cover< 1.5 Cover>=1.5 Cover< 1.5 Cover>=1.5Gap

194 : n=9

(n=12) (n=6) (n=6) 202 : n=5 231 : n=6 83.5 : n=3 59.8 : n=3 98.2 : n=3 41 : n=3 (n=3) (n=3) (n=3) (n=3) Error : 0.531 CV Error : 0.941 SE : 0.069 Figure 3.3. Multivariate regression tree (MRT) comparing log transformed species compositions between pre-EAB overstory and current sapling and recruit strata in canopy gaps and adjacent forests of three northern Michigan rivers. Cross validation of 1000 runs yielded a 10 branched tree most frequently (360). Explained 46.9% of total variance.

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