The Ecological Impacts of the Emerald Ash Borer (Agrilus Planipennis): Identification of Conservation and Forest Management Strategies

THE ECOLOGICAL IMPACTS OF THE EMERALD ASH BORER (AGRILUS PLANIPENNIS): IDENTIFICATION OF CONSERVATION AND FOREST MANAGEMENT STRATEGIES

A dissertation submitted

to Kent State University College of Arts and Sciences

in partial fulfillment of the requirements for the

degree of Doctor of Philosophy

Constance E. Hausman

December, 2010

Dissertation written by

Constance E. Hausman

B.S., Bowling Green State University, USA, 1998

M.S., Bowling Green State University, USA, 2001

Ph.D., Kent State University, USA, 2010

Approved by

______, Chair, Doctoral Dissertation Committee Oscar J. Rocha

______, Members, Doctoral Dissertation Committee Andrea L. Case

______Daniel A. Herms

______Marilyn A. Norconk

______Alison J. Smith

Accepted by

______, Chair, Department of Biological Sciences James L. Blank

______, Dean, College of Arts and Sciences John R. D. Stalvey

TABLE OF CONTENTS

LIST OF FIGURES ...... VII

LIST OF TABLES ...... IX

DEDICATION...... X

ACKNOWLEDGMENTS ...... XI

CHAPTER 1 INTRODUCTION ...... 1

Host Plant: Genus Fraxinus ...... 6

Emerald Ash Borer Control Options ...... 9

Emerald Ash Borer Eradication Efforts ...... 12

Scope of Research ...... 15

Organization of the Dissertation ...... 16

CHAPTER 2 IMPACTS OF THE EMERALD ASH BORER (EAB)

ERADICATION AND TREE MORTALITY: POTENTIAL FOR A

SECONDARY SPREAD OF INVASIVE PLANT SPECIES ...... 19

ABSTRACT ...... 19

INTRODUCTION ...... 20

MATERIALS AND METHODS ...... 23

Study Site Description ...... 23

Experimental Design and Biotic Sampling ...... 24

Abiotic Sampling ...... 26

Statistical Analysis ...... 27

RESULTS ...... 28

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DISCUSSION ...... 44

ACKNOWLEDGEMENTS ...... 48

CHAPTER 3 DISTURBANCE FACILITATES A SECONDARY SPREAD OF

INVASIVE PLANT SPECIES: MANAGEMENT CONCERNS FOR

EMERALD ASH BORER INFESTED FORESTS ...... 50

ABSTRACT ...... 50

INTRODUCTION ...... 51

Research Objectives ...... 56

MATERIALS AND METHODS ...... 56

Site and Experimental Design ...... 56

Abiotic Sampling ...... 57

Biotic Sampling ...... 59

Seed Bank Survey ...... 59

Statistical Analysis ...... 60

RESULTS ...... 61

Abiotic: Light Environment ...... 61

Abiotic: Soil Compaction...... 65

Biotic: Species Diversity...... 65

Seed Bank ...... 78

DISCUSSION ...... 81

Forest Stand Maintenance ...... 85

CHAPTER 4 GENETIC STRUCTURE OF GREEN ASH (FRAXINUS

PENNSYLVANICA) IN THE STATE OF OHIO: IMPLICATIONS FOR

THE ESTABLISHMENT OF EX SITU CONSERVATION PROTOCOLS.87

ABSTRACT ...... 87

INTRODUCTION ...... 88

MATERIALS AND METHODS ...... 94

Study Species ...... 94

Study Site ...... 95

DNA Isolation ...... 97

Genetic Diversity Analysis ...... 98

RESULTS ...... 100

Allelic Diversity ...... 100

Genetic Variation Within Populations ...... 100

Genetic Differentiation Among Populations ...... 103

Sampling Design ...... 108

DISCUSSION ...... 110

Recommendations ...... 115

CHAPTER 5 DISCUSSION ...... 118

EMERALD ASH BORER IMPACTS...... 118

EMERALD ASH BORER ERADICATION...... 119

SCALE OF DISTURBANCE ...... 121

CONSERVATION EFFORTS FOR ASH...... 123

BROADER CONSERVATION IMPLICATIONS ...... 125

Management Concerns for Hazardous Trees ...... 125

Management Concerns of Deer Herbivory ...... 127

APPENDIX ...... 130

REFERENCES ...... 132

LIST OF FIGURES

Figure 1.1 Summary of the Fraxinus phylogeny adapted from Wallander 2008...... 7

Figure 1.2 EAB eradication locations in Northwest Ohio from 2003 to 2005...... 14

Figure 2.1 Example of Hemispherical pictures...... 32

Figure 2.2 Light environment as determined by Global Site Factor (GSF)...... 33

Figure 2.3 The number of sunflecks (a) and average duration of each sunfleck (b)

in Uncut and Cut plots...... 34

Figure 2.4 Soil compaction in Cut and Uncut plots at five different depths in the soil

profile from 2005...... 37

Figure 2.5 Shannon’s Diversity Index (H′) for understory plant community

in Uncut and Cut plots...... 38

Figure 2.6 Canonical Correspondence Analysis (CCA) ordination plot from 2006. ... 39

Figure 3.1 Light environment for the three disturbance treatments as determined

by GSF...... 62

Figure 3.2 The average number of sunflecks between the three disturbance levels. ... 63

Figure 3.3 The average duration of each sunfleck between treatments...... 64

Figure 3.4 Soil compaction as measured with a penetrometer...... 66

Figure 3.5 Soil compaction as measured by bulk density...... 67

Figure 3.6 Soil organic matter, as measured by loss on ignition (LOI)...... 68

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Figure 3.7 Shannon’s Diversity Index (H’) for the understory plant community

between disturbance treatments...... 69

Figure 3.8 Species richness between the three disturbance levels...... 71

Figure 3.9 Canonical Correspondence Analysis ordination plot...... 73

Figure 3.10 Total accumulated germination over an 8-week sample period...... 79

Figure 4.1 Distribution of the Toledo Area Metroparks in Lucas County, Ohio...... 96

Figure 4.2 Genetic relationship between green ash populations from the eight

Toledo Metroparks...... 104

Figure 4.3 Relationship of log geographical distance and genetic differentiation

Fst/(1-Fst) between populations as proposed by Rousset (1997)...... 106

Figure 4.4 Population structure of green ash populations from the eight Toledo

Metroparks as determined using STRUCTURE...... 107

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LIST OF TABLES

Table 2.1 Relative Density, Frequency, Coverage and Importance Values (IV) for

the 10 most important tree species across the entire forest...... 29

Table 2.2 IVs are calculated for the same top 10 tree species for each treatment. ... 31

Table 2.3 Invasive plant species in the understory of Uncut and Cut plots listed by

total percent cover from all 1 x 1 meter understory subplots...... 42

Table 3.1 Invasive plant species in the understory listed by total percent cover from

all 1 x 1 meter understory subplots within each treatment...... 75

Table 4.1 Allele summary of microsatellite loci for green ash

(Fraxinus pennsylvanica)...... 101

Table 4.2 Summary of Wright’s F-Statistics for all microsatellite loci...... 102

Table 4.3 Population summary of genetic variation...... 105

Table 4.4 Germplasm collection recommendations based on level of inbreeding,

number of mother trees sampled and number of seeds collected

from each mother...... 109

DEDICATION

An education is a privilege not to be taken for granted. I dedicate this work to all of the women in my family. To my mother Marlene Roman for her tireless commitment to our family, she always wanted more for her daughters. I am truly blessed to have her as a constant source of motherly inspiration. To my sisters Jennifer Anzalone and

Carolyn Roman, they have nurtured my family throughout this process and I am lucky and deeply grateful for their support. To my grandmothers Betty Roman and Penny

Kline, I am the first generation of women to go to college and the opportunities that I have been fortunate to experience would not have been possible if it weren’t for their sacrifices. I finally dedicate this work to my two beautiful daughters Lillian and

Charlene Hausman, who have been my inspiration and strength, may they grow up knowing there are no limits to their success. When the grind of long work hours would feel almost too much to bear, somehow they knew exactly when mommy needed a hug.

ACKNOWLEDGMENTS

My sincerest appreciation to my advisor Dr. Oscar Rocha for embracing a project that challenged us both. His encouragement has meant a lot to me and I have truly enjoyed working in his lab. I would like to acknowledge all of the members of my committee: Andrea Case, Dan Herms Marilyn Norconk and Alison Smith, I appreciate all of the support that they have provided. To Andrea Case, you have always been a professional source of inspiration and I thank you. To Dan Herms, you motivate me to be a better researcher and I thank you for supporting me from the very beginning.

I thank Barb Andreas who has not only been a professional mentor but a personal friend as well. I will miss the long chats in your office discussing everything from botany lectures to your next vacation adventure. May I someday grow up to be half the botanist you are. Every summer I had the privilege of working with John Jaeger, I thank you for sharing with me your love and passion for nature and the occasional Red Bull slush bomb.

I would like to acknowledge the following funding sources which provided much appreciated financial support throughout this research: The Metroparks of the Toledo

Area, Ohio Biological Survey, the Art and Margaret Herrick Research Grant, the Hobbs

Scholarship, the Kent State University Graduate Student Senate and the Kent State

University Fellowship. I would also like to thank all of the office and support staff in the

Department of Biological Sciences both past and present. xi

I have shared this experience with numerous fellow graduate students and have created many lasting memories and friendships. In particular I would like to thank

Maureen Drinkard, Jennifer Clark, Justin Montemarano, Raja Vukanti, Lisa Regula-

Meyer, Rajlakshmi Gosh, Mike Monfredi, Erin McNutt, Kelly Barriball, and Dylan

Stover. I especially thank Brendan Morgan, Michelle Lang and Mike Monfredi for putting in long hours in the field and lab; I greatly appreciated your research help.

I would like to thank my parents – my mom for sharing in the experience of a field season and my dad for providing me with the means and the value of an education.

Lastly, I share my love and appreciation for my husband Bryan. There are not words for the support and encouragement you have given me, you are my rock.

Constance E. Hausman

12/2010, Kent, Ohio

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CHAPTER 1

Introduction

The movement of exotic species has been the inevitable consequence of an expanding global society that relies heavily on international trade of goods and services.

The frequency of exotic introductions, as well as their establishment and subsequent spread, has increased through time with greater international trade and globalization (U.S.

Congress 1993). Consequently, human societies have radically changed the distribution of species by transporting them throughout the world (Drake et al. 1989, Williamson

1996, Ewel et al. 1999).

The outcome or success of a non-native species is dependent upon how well that species performs during the four stages of invasion: transportation, colonization, establishment and landscape spread (Theoharides and Dukes 2007). The exchange or transportation stage of invasion can occur intentionally or accidentally. Some species have been intentionally introduced for crops, for garden ornamentals or for forestry; while others were accidentally introduced as stowaways or hitchhikers (Elton 1958). In the United States this has resulted in the total introduction of the approximately 50,000 non-native species (Pimentel et al. 2005). However, only a small percentage of non- native species establish themselves in their new habitat; and an even smaller fraction of

1 2

these increase in abundance at the expense of native plant populations, communities and ecosystems (Pattison et al. 1998, Ellstrand and Schierenbeck 2000). For example, of the total 25,000 introduced plant species to the U.S. (Pimentel et al. 2005) approximately

5,000 have become established in natural ecosystems and have displaced native plant species (Morse et al. 1995).

There are both economical and ecological costs associated with the introduction of non-native species. Once established, the non-native species that are capable of producing environmental or economic damage or that represent a threat to human health are considered to be invasive (Rossman 2001). The economic impact of invasive species including damage and control efforts is approximately $120 billion annually for the

United States (Pimentel et al. 2005). Invasive species also have profound ecological impacts, and ecological costs are much harder to approximate with just dollars and cents.

The spread of non-native species is identified as the second greatest threat to the loss of biodiversity next to habitat destruction (Wilcove at al. 1998). In fact, 49% of the species that are listed as threatened or endangered are at risk due to competition or predation by non-native species (Wilcove at al. 1998).

Since invasive species have the potential to cause significant ecological damage, research efforts attempt to identify impacts as early in the invasion stage as possible.

There are several questions that should be addressed when determining a non-native species’ potential for becoming invasive. How fast will the non-native species spread?

Will this non-native species out compete or displace the native community members?

Does this non-native species have the potential to hybridize with native congeners?

These investigations focus on a range of biological levels from genetic variation to population and community structure upwards through ecosystem functioning and services. In order to quantify the effects of invasive species, field-based studies can be designed to compare invaded vs. non-invaded sites or to compare before and after invasion affects. For example, experimental field-based studies, such as an exclosure study, may assist with determining the impacts a non-native herbivore has on a native plant community. Early research on invasive species’ impact provides a framework for future control efforts as well as document patterns and processes of the invasion stages.

This research is significant as invasive species are capable of causing the functional extinction of native species; that is, a native species population can be reduced to such minimal numbers that the species no longer contributes to the biodiversity of the community and fails to support its previous ecosystem function.

Forest communities are affected by 360 exotic insect species and over 20 exotic pathogens (Liebold et al. 1995) that have significantly altered canopy composition. An example of such invasion is the chestnut blight disease that has nearly eliminated what was the most economically important hardwood species in the eastern U.S. forests, the

American chestnut tree. The American chestnut (Castanea dentata) once made up approximately 25% of the eastern deciduous forest (Campbell and Schlarbaum 1994); and the chestnut blight, since its introduction, has caused the death of nearly one billion trees which has also led to profound ecosystem changes in the eastern deciduous forest

(Primack 1993, McNeely 1999). The American elm (Ulmus americana) suffered a similar fate due to the introduction of the Dutch elm disease (McBride 1973). The

functional extirpation of the American chestnut and American elm has dramatically affected the composition and structure of native forest communities.

An unfortunate consequence of foreign trade is the increased risk for the spread of exotic species. The chestnut blight is just one example of the establishment and spread of invasive species that can have lasting impacts on the native landscape. Therefore, whenever possible conservation efforts should focus on earlier detection of probable invasive species with an emphasis on the identification of potential long-term ecological impacts. My research attempts to address these conservation issues by identifying the ecological impacts of an exotic, invasive beetle relatively soon after its discovery. Early detection of the emerald ash borer revealed its potential as a major threat to native forests.

The emerald ash borer (EAB), Agrilus planipennis Fairmaire (Buprestidae), is a wood-boring beetle native to Asia that relies on ash species (Fraxinus spp.) to complete its life cycle (Haack et al. 2002). Adults emerge from ash trees from May to

August/September and will feed on the foliage of the trees before mating. This foliage feeding causes relatively negligible damage to the tree. The female beetles will then lay approximately 60-90 eggs in the crevices of the bark. After 1 to 2 weeks the larvae hatch and chew into the tree where they feed on the conductive phloem tissue (Cappaert et al.

2005). As the larvae feed on the cambium layer of the ash, they interfere with the trees ability to translocate water and nutrients. Once an ash shows symptoms of canopy decline due to EAB the tree typically dies within two to four years (Herms et al. 2004).

The emerald ash borer typically has a one year life cycle; however, a 2-year larval

development has been documented in North America (Siegert et al. 2007, Tluczek et al.

2008).

EAB appears to demonstrate high levels of host specificity. Other tree species including Ulmus americana L., Juglans nigra L., Carya ovate (Mill.) K.Koch, Celtis occidentalis and Syringa reticulata have been tested to determine the colonization ability of EAB on other genera. Anulewicz et al. 2008 found that occasionally EAB will land on and oviposit on non-ash tree species; however, the larvae are not able to survive.

Successful EAB development only occurs on trees in the Fraxinus genus.

The detection of EAB-infested ash trees is difficult because a tree or stand of ash can be infested with EAB long before there is any visible indication that the tree is stressed. This is because EAB spends most of its life cycle, more than 300 days, in the juvenile stage under the bark of the tree. The external indicator symptoms (woodpecker holes, canopy dieback, bark fissures, epicormic branching and D-shaped exit holes at visible range) are not typically evident until an ash tree has a large EAB infestation

(Poland and McCullough 2006, McCullough et al. 2009).

The emerald ash borer first was identified in Detroit, Michigan in 2002; and it was likely the result of an accidental introduction that occurred at least 10-15 years prior

(Cappaert et al. 2005). EAB is thought to have been imported as larvae in ash wood crating or pallets originating in Asia (Stone et al. 2005). EAB-induced ash mortality was estimated in 2006 to be over 15 million trees in Michigan (Poland and McCullough

2006); and in 2007 ash mortality was estimated to be as high as 53 million across

Michigan, Ohio and Indiana (Smith et al., submitted for publication). By October 2010,

EAB had been found in 15 states: Michigan, Ohio, Indiana, Illinois, Pennsylvania,

Kentucky, Tennessee, Maryland, West Virginia, Virginia, Wisconsin, New York,

Missouri, Minnesota and Iowa as well as two provinces in Canada. Furthermore, it is projected that EAB has the ability to expand its range across 25 states in the next 10 years due to the expansive host tree range and a lack of effective control measures (Kovacs et al. 2009).

Host Plant: Genus Fraxinus

The genus Fraxinus is one of 24 genera within the Oleaceae or Olive family.

Fraxinus is a monophyletic genus (Wallander and Albert 2000) with 43 species that occur mostly in the temperate and subtropical regions of the Northern Hemisphere (Fig.

1.1). The geographical distribution of this genus is divided into 2 main regions with 20 species occurring in North America and another 20 occurring in eastern Asia (Wallander

2008). The remaining three species are found in Europe and western Asia. The spatial spilt in the number of species was likely attributed to intercontinental dispersal events that occurred during the Middle and Later Tertiary when there were several land mass connections between North America and Asia. The genus Fraxinus likely evolved in

North America followed by two migration events into Asia (Jeandroz et al. 1997). The origination of the Fraxinus genus in North America is further supported with fossil evidence from the Eocene (55-34 million years ago) in which the oldest known fossilized

Fraxinus fruits or samaras have been collected in southeastern North America (Call and

Dilcher 1992).

Figure 1.1 Summary of the Fraxinus phylogeny adapted from Wallander 2008.

There are 16 New World species of ash that are native to the United States which are identified by filled boxes.

The United States currently has 16 native species of Fraxinus that occur in 47 of the lower 48 states (USDA Plants 2010). These ash species range in stature from shrubs to canopy trees, have a variety of reproductive strategies (polygamous, monoecious, dioecious) and are either wind- or insect-pollinated with wind-dispersed seeds (Coder

2008; Wallander 2008). Individual species’ distribution can range from a single state

(Fraxinus gooddingii in Arizona) to 42 of the lower 48 states for Fraxinus pennsylvanica

(USDA Plants 2010). Thus far, there are 5 species with distribution ranges that overlap with EAB-infested areas. These 5 species — Fraxinus pennsylvanica (Green ash),

Fraxinus americana (White ash), Fraxinus quadrangulata (Blue ash), Fraxinus nigra

(Black ash) and Fraxnius profunda (Pumpkin ash) — are all susceptible to EAB (Smith

2006, Anulewicz et al. 2007).

While most North American Fraxinus species appear to be susceptible to EAB, there have been some species-specific host preferences identified. When the two species coexist, higher densities of EAB larvae are found on Fraxinus pennsylvanica (green ash) over Fraxinus americana (white ash) and a similar host preference pattern is observed for

Fraxinus americana (white ash) over Fraxinus quadrangulata (blue ash) (Anulewicz et al. 2007). Therefore in areas where multiple Fraxinus species occur, early detection of

EAB infestations may be determined by surveying the preferred host species (Anulewicz et al. 2007). Unfortunately, while some initial colonization preference may exist, it appears that ash mortality is ultimately influenced by the size of the EAB infestation which overwhelms any individual ash species selection preference. In addition, there does not appear to be a correlation between forest or landscape attributes and

susceptibility to EAB attack (Smith 2006). As a consequence, the emerald ash borer has the potential to cause significant changes to native forests due to an expansive host tree range with relatively few barriers limiting its spread.

In addition to restructuring the landscape due to the loss of several dominant tree species, there are a number of native fauna which also utilize ash. The future of these species becomes vulnerable to decline as EAB compromises the persistence of ash trees.

In all there are 44 other arthropod species (43 are native and one exotic) which are associated only with ash and are considered high risk for endangerment (Ghandi and

Herms 2010a). It is apparent that the emerald ash borer will have an ever increasingly negative impact on ecosystem structure and function; and as such, the future preservation of our native ecosystems will require active management interventions to minimize losses.

Emerald Ash Borer Control Options

As EAB continues to spread, new management strategies are being developed in an attempt to minimize the economic and ecological impact of this forest pest. It is estimated that ash tree dieback spreads outward from an EAB source infestation at a rate of 10.6 km per year (Smitley et al. 2008). Furthermore, if no containment measures are implemented, EAB will likely expand its range to 25 states in the next 10 years and could cause $10.7 billion in economic loss (Kovacs et al. 2009). Control measures for EAB include: biological control through parasitoids, chemical control through systemic

insecticides, and firewood control which restricts a significant but passive form of beetle dispersion through the movement of infested firewood.

There have been three parasitoid species introduced in the People’s Republic of

China that are promising biological control agents for EAB (APHIS 2009). Spathius agrili Yang is a larval ectoparasitoid in the Braconidae family (Yang et al. 2005) that has demonstrated high host specificity for EAB (Yang et al. 2008). Once the host ash tree is located, female Spathius locate the larvae under the bark by detecting mechanical vibrations caused during feeding (Wang et al 2010). Another parasitoid Tetrastichus planipennisi Yang is a larval endoparasitoid in the Eulophidae family (Liu et al. 2003,

Yang et al. 2006); and lastly, Oobius agrili Zhang and Huang is an egg parasitoid in the

Encyrtidae family (Zhang et al. 2005). While these parasitoids have demonstrated some promising potential in the laboratory, the large scale practicality of rearing is still under development, as well as, that of testing the non-target effects of these non-native parasitoids in the environment.

There have also been several native species of parasitoids found attacking EAB larvae; however, their parasitism rate on EAB is generally low 1-2% (Kula et al. 2010).

A native parasitoid wasp, Atanycolus sp. in the Braconidae family, has been found which has much higher parasitism rate (15-56%) on EAB larvae (Cappaert and McCullough, In press). These native parasitoids may not be causing widespread EAB mortality, but they are a natural enemy and may help regulate EAB populations. Biological control has also been tested and effectively demonstrated using the microbial Beauveria bassiana strain

GHA in both greenhouse and field studies (Liu and Bauer 2008). However, no large

scale control of EAB exists with biological control agents as research on their effectiveness and usage is still ongoing.

Chemical control of EAB has been tested with a number of different insecticides that target various life stages of the beetle. The systemic insecticides are applied to ash trees through soil drenches, trunk injections, or as a trunk spray to target the juvenile stage as the larvae feed on the vascular tissue; whereas the EAB adults are targeted through a cover spray that is applied to the trunk, branches and foliage (Herms et al.

2009). Imidacloprid and emamectin benzoate are the two most common active ingredients in pesticides for the control of EAB. Imidacloprid must be applied annually while emamectin benzoate has shown effective control of EAB larvae for two years after a single application (Herms et al. 2009). Little is known about non-target effects of these insecticides; however, Kreutzweiser et al. (2007) did find that imidacloprid caused significant mortality among aquatic decomposers when the insecticide enters the system via leaf fall or leaching. Studies to determine the effectiveness of these insecticides, as well as, their impacts on non-target species are ongoing.

Humans have been identified as an important vector in the long distance transport of EAB (BenDor et al. 2006). The unintentional spread and dispersal of EAB occurs when people transport firewood out of quarantined zones to uninfested areas. The movement of larvae-infested firewood has contributed to an unnatural range expansion that has interfered with EAB control measures. As such, techniques are currently being developed that will eliminate EAB from firewood. For example, there is a reduction in

EAB emergence, survival and size if firewood logs are cut early during larval

development (July /August); furthermore, splitting logs and storing them untarped also reduced adult emergence (Petrice and Haack 2006). Work continues to determine other methods of firewood sanitation (Goebel et al. 2009, Poland et al. 2008a, Poland et al.

2008b) in order to reduce the artificial spread of the beetle.

Emerald Ash Borer Eradication Efforts

There are six primary factors that contribute to the successful eradication of an invasive species: (1) Financial resources must be sufficient to fund the program until its conclusion; (2) The lines of authority must be clear, such that the lead agency is able to perform the necessary procedures across the entire affected area; (3) The biology of the target organism must be susceptible to the control procedures implemented; (4)

Reinvasions must be prevented; (5) The target pest must be detectable at relatively low densities; and (6) Eradication might require the restoration or management of the community or ecosystem once the target species is removed (Myers et al. 2000). The

United States has a history of eradication efforts including 42 arthropod pest species that resulted in failures, limited successes, or successful eradications (Klassen 1989). One of the biggest problems with eradication attempts is the vulnerability of reinvasion if eradication efforts are not completely successful (Myers et al. 2000).

In 2002 the USDA Animal Plant Health Inspection Service (APHIS), acting as the lead agency, instituted control efforts to minimize the spread of EAB. Eradication efforts of the emerald ash borer were deemed appropriate since EAB was thought to have been detected early in the invasion and to have limited distribution; additionally, control

procedures were identified to target the juvenile life cycle. These efforts included the establishment of quarantine areas with fines to prevent potentially infested ash trees, logs or firewood from being transported out of infested areas and an intensive eradication effort which called for the removal of every ash tree within a 0.8 km radius around each positively infested tree (USDA APHIS 2003, Stone et al. 2005).

Localized eradication efforts were initiated in Michigan when EAB was first discovered, and by 2005 several eradication efforts were being conducted at the border of

Michigan and Canada in St. Clair County and the borders with Ohio and Indiana, as well as, at outlier infestations across Michigan and the Upper Peninsula. Canada worked to create an “ash-free zone” in Essex County, Ontario, to act as a barrier to further EAB spread (Marchant 2004). The state of Indiana also began eradication efforts on a limited, localized scale in 2005. Much of the areas subjected to eradication efforts were either in urban environments (i.e., Detroit, Michigan and Toledo, Ohio) or in rural areas between

Michigan, Indiana and Ohio. As such, much of the eradication program involved removing street trees in residential and urban areas or removing ash from rural woodlots.

However, eradication efforts were conducted across many natural areas as well including

State Forests, State Parks and Metro Parks.

In 2003, Ohio detected its first EAB-positive tree which triggered the eradication efforts for the state. In all there were six isolated infestations identified in 5 Ohio counties: Lucas, Williams, Defiance, Wood and Franklin (Fig.1.2). The eradication efforts implemented at these sites removed a total of 48,016 ash trees within the 0.8 km radius of the EAB-positive tree. In 2004, there were 11 additional eradication sites

Figure 1.2 EAB eradication locations in Northwest Ohio from 2003 to 2005.

This map was produced by the Cooperative Emerald Ash Borer Program and details the specific locations and size of the EAB eradication efforts from 2003-2005.

identified: one in the town of Swanton located in Fulton County, one at Maumee State

Forest in Henry County, one in southern Wood County in North Baltimore and the remaining 8 eradication sites were scattered throughout Lucas County. Between March and July of 2005 eradication was conducted at 7 more locations and targeted the leading edge of the infestation in Lucas and Ottawa Counties and on a few satellite infestations that were fairly isolated (Wood and Delaware County). By 2006, the range of EAB was determined to exist well beyond any of the eradication zones; therefore, eradication efforts were no longer implemented. EAB has continued to spread throughout the state of Ohio; and as of September 2010, all 88 counties in Ohio were quarantined due to the discovery of EAB in the Wayne National Forest.

Scope of Research

There is a significant, growing body of literature on the emerald ash borer since its discovery in 2002. This research includes: identifying the beetles’ biology, behavior and ecology; exploring mechanisms for effective chemical and biological control; developing methods for early EAB survey and detection; and studying EAB host-plant relationships. My research falls under the latter category; however, questions addressed in my dissertation work focus on the impact EAB have on native plant communities rather than on the emerald ash borer itself. This research is one of the first attempts to determine the ecological impact of EAB including the unintentional effects of its attempted eradication. Furthermore, my research addresses forest management strategies

appropriate for EAB infested forests and also identifies the spatial, genetic structuring of one ash host species.

This dissertation research was initiated relatively early on in the emerald ash borer invasion in the United States. The intention and design of the research projects were to identify the ecological impacts of EAB and to determine long-term community changes to native deciduous forests. The three major objectives of this research are: (1) to identify consequences of EAB eradication efforts, (2) to determine altered community composition under different disturbance intensities (recommendations for tree removal management) and (3) to design effective ex situ conservation protocols for future ash tree preservation. The end result for all of these research areas is to devise effective forest management recommendations and prioritize conservation efforts for ash preservation.

Organization of the Dissertation

The main focus of my dissertation research is to address the ecological impacts of an invasive forest pest during the early stages of establishment. In the following chapters

I describe research projects that address relevant and timely conservation issues in order to provide a framework for future forest management strategies. The goal of each chapter is to recognize how EAB affects long-term forest community dynamics and to provide practical applied conservation strategies in order to minimize the disturbance caused by EAB. All of the research projects were conducted in Lucas County in

Northwest Ohio where EAB was first identified in 2003 and where eradication efforts were initiated at my study site in 2005. The dissertation is divided into three main

research chapters and a final synthesis chapter in which I summarize the major themes and conclusions of the research.

Chapter 2 focuses on the ecological impacts of the attempted EAB eradication protocol. This project was initiated in 2005 and identifies the shift in forest community structure (canopy and understory) that results from changes in the abiotic environment

(light and soil compaction) due to EAB eradication efforts implemented by the Ohio

Department of Agriculture. This chapter identifies the trade-offs between eradication efforts which cause large scale disturbance and the subsequent facilitation of a secondary plant invasion.

The research described in Chapter 3 builds off of the research conducted in

Chapter 2 in order to determine how various scales of disturbance affect the understory plant community. While Chapter 2 is a comparative field study between eradication areas and controlled areas with ash trees still standing, Chapter 3 is an experimental study that identifies the outcomes of ash tree removal as a forest management strategy. With the continued spread of EAB, ash tree removal remains a necessary forest management practice. However, it is important to determine an effective ash tree removal method while minimizing disturbance to the surrounding community. Therefore, the goal of this project is to determine whether invasive plant species’ establishment varies with different intensities of disturbance caused by ash tree removal.

Chapter 4 recognizes EAB’s likely persistence in the native forests and seeks to identify the loss of localized genetic diversity as ash trees continue to die. As EAB continues to spread, all species of ash encountered to date are susceptible; and once an

ash stand becomes infested, mortality is near 100% (Knight et al. 2007). With such potentially devastating impacts, we may be faced with the functional extinction of the entire Fraxinus genus from our native forests; and therefore, ash conservation efforts need prioritization. This study initiates ash conservation efforts by providing preliminary results for spatial genetic structuring. The research is designed to characterize the genetic diversity of green ash (Fraxinus pennsylvanica) from across Lucas County, Ohio. The objective for this project is to guide ash tree conservation efforts by developing an effective ex situ seed collection sampling protocol.

Chapter 5 is a synthesis of the project as a whole. The major conservation issues are addressed with recommendations established for the prioritization of forest management practices. Included are the broader conservation concerns that I feel are critical to long-term forest impacts with suggestions for future EAB research.

CHAPTER 2

Impacts of the emerald ash borer (EAB) eradication and tree mortality: potential

for a secondary spread of invasive plant species

(This chapter was published July 2010 in Biological Invasions)

ABSTRACT

Since the discovery of the emerald ash borer in 2002, eradication efforts have been implemented in an attempt to eliminate or contain the spread of this invasive beetle. The eradication protocol called for the removal of every ash tree within a 0.8 km radius around an infested tree. In 2005 this study was established to identify environmental changes attributed to the eradication program and to measure subsequent shifts in forest community composition and structure. I conducted this study in Ohio and compared areas that received the eradication treatment (ash trees cut down) to areas that were left uncut (ash still standing). The goal of this project was to identify how the plant community responded in these two areas. The eradication protocol accelerated the formation and size of gaps within the forest and thus increased the duration and intensity

19 20

of light penetrating through to the forest floor. In addition, the use of tracked vehicles for the removal of cut trees resulted in significant soil compaction. The resultant plant community had greater species diversity (H'). When specific species composition differences were compared, an increase in the establishment of invasive plant species was detected in areas that received eradication efforts compared to those that did not.

Invasive species accounted for 18.7% of the total herbaceous cover in this highly disturbed environment which included Cirsium arvense, Rhamnus cathartica and 2 species of Lonicera spp. In contrast, invasive species accounted for <1% of the total herbaceous cover in the undisturbed uncut areas.

INTRODUCTION

The introduction of invasive species has negative effects on biological diversity at the local level (Vitousek 1990, Sandlund et al. 1999, Pimentel et al. 2000) and can lead to severe disruption of ecological communities (Primack 1993, Johnson and Padilla 1996,

Moller 1996). It is estimated that the economic impact of invasive species on agriculture, forestry, fisheries and human health in the United States is at least $134 billion annually

(U.S. Congress 1993). Eradication efforts to combat the spread of an invasive species present themselves as a feasible alternative. However, the cost-benefit analyses of eradication programs typically underestimate the cost (both financially and ecologically) and overestimate the benefit (Myers and van Lear 1998).

The introduction of forest pests has a profound impact on native plant community structure due to the lack of natural enemies or the lack of plant defense mechanisms

(Mack 2000). The gypsy moth, Lymantria dispar (L.), is a well-documented forest pest that has had devastating impacts on native forests wherever established (Campbell and

Sloan 1977, Liebold et al. 1995, Fajvan and Wood 1996, Abrams 1998) including significant mortality among stressed trees during a single defoliation event (Davidson et al. 2001). Since the first documented outbreak of gypsy moth in 1865, eradication efforts have been heavily implemented to eliminate this pest. While eradication is no longer a feasible option, control mechanisms to manage the spread of the gypsy moth have limit gypsy moth populations at a cost of $11 million per year (Campbell and Schlarbaum

1994).

A recent example of a successful invasive species to the continental United States is the emerald ash borer (EAB), Agrilus planipennis Fairmaire (Buprestidae). This metallic-green, wood-boring beetle is native to Asia (Haack et al. 2002) and relies on ash species (Fraxinus spp.) to complete its life cycle. Larvae feed on the cambium layer of the ash interfering with the translocation of water and nutrients. Once infested, the ash tree typically dies within two to four years (Herms 2004).

The first EAB infestation in the United States was identified in Detroit, Michigan in 2002; but its introduction likely occurred at least 5-10 years prior (McCullough and

Roberts 2002). It is likely to have been imported as larvae in ash wood crating or pallets

(Stone et al. 2005). Since discovery, EAB has killed over 15 million ash trees in

Michigan (Poland and McCullough 2006) with even greater mortality estimated across all

other states to date. As of 2010, EAB was known to be established in 15 states and

Canada. There are 16 Fraxinus species native to the United States. A few of these species have protective status, including: Fraxinus profunda which is endangered in the states of New Jersey and Pennsylvania and is threatened in Michigan and Fraxinus quadrangulata which is threatened in Iowa and Wisconsin (USDA NRCS 2008). In Ohio where this study was conducted, there are five native ash species (F. americana, F. pennsylvanica, F. profunda, F. nigra and F. quadrangulata); and all are susceptible to and attacked by EAB (Smith 2006, Anulewicz et al. 2007). There are approximately 279 million ash trees in Ohio (USDA Forest Service 1991) which is about 6% of all trees by number and volume (Widmann 2008). No landscape or forest attributes have been identified to date that make a stand resistant to EAB attack (Smith 2006)

In 2002 the USDA Animal and Plant Health Inspection Service (APHIS) in conjunction with the Canadian Food Inspection Agency (CFIA) established a Science

Advisory Panel to develop a plan to contain and eventually eradicate EAB. Measures to control the outbreak included tree removal and the establishment of quarantine areas to regulate the potential movement of infested wood materials (USDA-APHIS 2003, Stone et al. 2005). The funding for this multi-state level eradication effort was provided by

APHIS. According to the APHIS protocol, after an infested tree was identified, every ash tree >2.5 cm diameter at breast height (DBH) within a 0.8 km radius was felled and chipped to destroy EAB larvae (Poland and McCullough 2006, Stone et al. 2005). The remaining stumps were treated with herbicide, and the chips were burned in an electricity co-generation plant (Poland and McCullough 2006). This eradication protocol, as it was

applied on site, was supposed to have as minimal an impact as possible. However, these measures created immediate environmental changes to the habitat with abrupt environmental consequences.

The eradication protocol may influence the composition of existing plant communities by unnaturally accelerating the formation of gaps in the forests due to the total removal of all ash trees. This may increase the duration and intensity of light reaching the forest floor. Second, the heavy, tracked vehicles used to remove trees may lead to soil compaction, as most soils become compacted during the first few passes of a vehicle (Lockaby and Vidrine 1984, Shetron et al. 1988). Our study performed a comparison of plant communities between areas that received an eradication protocol treatment to areas that received no treatment.

This project identified the biological consequences of implementing the emerald ash borer eradication program in Ohio. Specifically, I identified the shift in forest canopy dominance and the change in understory plant community dominance and distribution.

A secondary spread of invasive plants was predicted in these highly disturbed eradication cut areas.

MATERIALS AND METHODS

Study Site Description

This study was conducted at Pearson Metropark in Northwest Ohio (Lucas

County). Established in 1935, Pearson is 620-acres of the relic Great Black Swamp, a

glacial swamp forest. Situated on the eastern edge of the city of Oregon, it is surrounded by residential neighborhoods and agricultural fields. The forest in Pearson is characterized by dense hardwoods dominated by several ash, maple and hickory species

(Table 2.1). The soil substrate is composed of heavy clay, specifically >90% of the park is Latty silty clay (Soil Survey Staff 2008). In early spring of 2005 the Ohio Department of Agriculture identified three EAB-infested trees in Pearson Metropark which was identified as the leading eastern edge of the infestation and was selected to receive the most intensive eradication efforts.

Implementation of EAB eradication began in Pearson in April of 2005 and lasted approximately 3 months. Following the APHIS eradication protocol described above, self-propelled, tracked, hydro-axes were used to cut down ash trees within 0.8 km radius of an infested tree. Some tree-cutting truck chippers were used on-site to chip small branches and limbs, but skidders were primarily used to transport the trees out of the forest to staging areas. Logs were then transported to an off-site location for processing by a large drum chipper mill. Preexisting trails were used when possible with additional

“cut-roads” created as needed to access all ash trees. Approximately 5,000 trees, ranging from saplings to trees >40 cm DBH, were felled from a total of 83 acres within Pearson.

Experimental Design and Biotic Sampling

In April 2005 prior to the initiation of the EAB eradication protocol (tree removal), eight 20 x 25 meter plots were established using a block design to assess the status of the forest structure. Plot locations were randomly selected using global

positioning satellite (GPS) coordinates and included a 50-100 meter buffer between plots and a 50 meter buffer zone from any trail or forest edge. By July 2005 all ash tree- cutting eradication practices ceased due to a lack of funding; therefore, not all ash trees within Pearson Metropark were cut. None of the original 8 plots had ash trees removed and were, therefore, designated as Uncut control plots. Based on the total area that did receive the eradication protocol, only 6 additional 20 x 25 meter plots were able to be established and were designated as the Cut plots. These 6 Cut plots had buffer zones reduced to 50 meters between plots and 10-20 meters from trail or forest edge. A vegetation assessment of 14 plots (8 Uncut plots and 6 Cut plots) was conducted at the end of July 2005 and again in July 2006. Data collected included identifying canopy composition and structure and the herbaceous understory community. All trees, greater than 3 cm (DBH), within the 20 x 25 meter plots were counted, identified to species, and given a percent canopy cover. This information was used to determine the importance value of each tree species and was calculated by incorporating the number of trees per species that exist in an area, the overall frequency of each species in the forest and the total coverage, based on basal area, of each species (relative abundance + relative frequency + relative coverage = importance value) (Curtis and McIntosh 1951). Within each of the 14 plots, seven 1 x 1 meter subplots were established to determine the herbaceous understory community. Six of the subplots were placed around the perimeter of the 20 x 25 meter plot (4 corners and 2 at 12.5 meters along the edge) and an additional subplot was located in the center of the plot. Every plant encountered in these subplots was identified to species and given a % cover measurement. Species diversity

was assessed using Shannon’s Diversity Index (H’) for the understory plant community.

Voucher specimens of all plants were collected for proper species identification and deposited in the Kent State University Herbarium.

Abiotic Sampling

Abiotic measurements were taken to determine changes in the light environment and the effect of soil compaction. The light environment was assessed by taking six hemispherical photographs of the canopy from every 20 x 25 meter plot (4 corners and 2 from a center transect). Pictures were taken with a digital SLR Nikon COOLpix 950 camera with a fisheye lens attachment that was mounted to a self-leveling frame with a tripod base. These 180° fisheye images were taken from a height of 1 meter off the forest floor and show a 10 meter diameter view of the canopy (Oberbauer et al. 1993;

Whitmore et al. 1993). Fisheye images were imported into the software program

Hemiview version 2.1 and were used to calculate changes in solar radiation regimes

(Delta-T Devices Ltd. 1999). Hemiview was used to determine the Global Site Factor

(GSF), the proportion of global radiation (direct + diffuse) that occurs in the open sky compared to that which occurs under the forest canopy, and was used to identify the number and duration of sunflecks. A sunfleck occurs when direct radiation penetrates through to the forest floor. A sunfleck calculation scans the path of the sun from sunrise to sunset at 30 second intervals and determines start and stop times, as well as, the amount of radiation that occurs during each sunfleck.

Soil compaction measurements were taken in all fourteen plots. A Soil

Compaction Tester (soil penetrometer) (DICKEY-john Corp.) was used to determine the pressure in pounds per square inch (PSI) needed to penetrate through the soil profile.

These compaction measurements were converted to kilopascal (kPa). Six readings (4 corners and 2 from a center transect) were taken at five consecutive depths 7.6, 15.2,

22.9, and 38.1 centimeters in each plot. Soil temperature and percent moisture content were taken in the same six locations as compaction using a soil moisture meter (Aquaterr

T300).

Statistical Analysis

Multivariate analysis of variance was used to determine the effects of treatment

(Cut and Uncut), year (2005 and 2006) and their interaction on the light environment.

The global site factor (GSF) and average duration of each sunfleck were log transformed before analysis in order to meet the assumptions of normality. Multivariate analysis of variance was used to determine the effects of treatment (Cut and Uncut), depth of soil profile (7.6, 15.2, 22.9, 38.1 centimeters), soil moisture and the interaction of treatment and soil depth on soil compaction. One-way analysis of variance was used to examine the effect of treatment on species diversity. All stated analyses were performed using

JMP version 3.1 (SAS). Plant community composition was analyzed with ordination using canonical correspondence analysis (CCA) using PC-ORD version 5 (McCune and

Medford 1999). Ordinations were performed with percent cover data from 73 species and included environmental abiotic measurements (GSF, number of sunflecks, average

duration of sunflecks, soil compaction, soil moisture and soil temperature). Treatment

(Cut and Uncut) was also included as a covariate.

RESULTS

Ash (Fraxinus spp.) is the dominant tree species in the Pearson Metropark forest, comprising the highest Importance Value (IV) (86.2% in the Cut plots and 42.2% in the

Uncut plots) (Table 2.1). When ash species are eliminated from IV calculations, the IVs of all other tree species increase by an average of 39% in Cut plots and 16% in Uncut plots. The future projected dominant canopy species were similar in both treatments and include: Tilia americana, Acer rubrum, Acer saccharinum, and Ulmus rubra (Table 2.2).

Results from 2005 and 2006 show that removal of ash trees resulted in significant differences in the light environment (Fig. 2.1). These included an increase in duration and intensity of light reaching the forest floor. There was a greater GSF measurement found in Cut plots compared to Uncut plots (F1,162=56.87, P<0.0001) (Fig. 2.2). While

the difference in the number of sunflecks was not significant (F1,162=1.99, P=0.16),

approximately 1.5 more sunflecks occurred in Uncut plots compared to Cut plots (Fig.

2.3a). However, while fewer sunflecks occurred in the Cut plots, the duration of direct

sunlight lasted on average more than 1.5 minutes longer (F1,162=9.97, P=0.002). This is

the equivalent of 24 minutes more of direct radiation penetrating through to the forest

floor that day (Fig. 2.3b).

Table 2.1 Relative Density, Frequency, Coverage and Importance Values (IV) for the 10 most important tree species across the entire forest.

The last column projects the changes in dominant canopy structure when all of the ash trees are removed from IV calculations. Importance Values are calculated by incorporating the number of trees per species that exist in an area, the overall frequency of each species in the forest and the total coverage, based on basal area, of each species

(relative abundance + relative frequency + relative coverage = importance value) (Curtis and McIntosh 1951). Treatments are defined as: Cut — the Ohio Department of

Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy- tracked trucks and skidders; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

Importance Value Species Relative Relative Relative Importance without Density Frequency Coverage Value (IV) Ash Fraxinus spp. (Ash) 21.64 10.81 28.84 61.30 - Tilia Americana (Basswood) 13.70 11.71 12.53 37.94 48.22 Acer rubrum (Red Maple) 12.33 8.11 13.36 33.80 43.61 Ulmus Americana (Slippery Elm) 10.41 7.21 8.69 26.31 33.58 Acer negundo (Box Elder) 7.40 8.11 5.66 21.17 26.49 Acer saccharinum (Silver Maple) 6.58 4.50 11.68 22.76 29.86 Carya ovate (Shagbark Hickory) 4.66 6.31 3.14 14.11 17.43 Carya cordiformis (Bitternut Hickory) 3.84 6.31 2.43 12.57 15.38 Carpinus caroliniana (Blue Beech) 2.74 4.50 0.66 7.91 9.48 Crataegus spp. (Hawthorn) 2.74 4.50 0.84 8.09 9.73

Table 2.2 IVs are calculated for the same top 10 tree species for each treatment.

The first column is the IVs for the Uncut Control plots. The IVs for the Cut plots are calculated twice: first with ash trees present (calculated with DBH estimated as the shortest diameter of a cut stump) and second without ash trees (representing the immediate shift in canopy dominance after the eradication cuttings).

Species uncut plots cut plots cut plots without Ash with Ash Ash 42.22 86.18 - Basswood 38.95 36.68 53.29 Red Maple 40.96 24.41 35.87 Slippery Elm 34.53 15.68 21.80 Box Elder 25.79 15.05 20.84 Silver Maple 20.80 25.34 38.78 Shagbark Hickory 17.44 9.58 13.19 Bitternut Hickory 10.68 15.13 20.87 Blue Beech 4.76 12.32 16.24 Hawthorn 6.21 10.87 14.02 Totals: 242.35 251.24 234.89

Figure 2.1 Example of Hemispherical pictures.

Picture a. was taken from a portion of an Uncut plot which has 6 ash trees that comprise

40% of the canopy cover from within that plot. Picture b. was taken from a Cut plot that had 6 ash trees cut down during the EAB eradication. Treatments are defined as: Cut — the Ohio Department of Agriculture implemented EAB eradication protocol in which an

EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

Figure 2.2 Light environment as determined by Global Site Factor (GSF).

The proportion of radiation (GSF) that occurs under the forest canopy is greater in Cut than in Uncut plots (P<0.0001). Treatments are defined as: Cut — the Ohio Department of Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy- tracked trucks and skidders; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback. Error bars represent + 1 SE of the mean.

Figure 2.3 The number of sunflecks (a) and average duration of each sunfleck (b) in

Uncut and Cut plots.

The number of sunflecks did not differ between treatments, but the average duration of each sunfleck lasted longer in the Cut plots (P=0.0001). Treatments are defined as: Cut

— the Ohio Department of Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback. Error bars represent + 1 SE of the mean.

There was a significant year effect for GSF, number of sunflecks and average duration of sunfleck (F1,162=17.47, F1,162=13.37, F1,162=27.64, P<0.0003 respectively), such that all

light measurements were slightly greater in 2005 immediately following the tree cutting

eradication efforts.

There were significant differences in soil compaction between Cut plots and

Uncut plots (F1,404=44.61, P<0.0001) in 2005. The degree of soil compaction was greater

in the Cut plots with significant differences detected at each depth of the soil profile (Fig.

2.4). Soil compaction results remained similar in 2006 with greater compaction detected

in the Cut plots than in the Uncut plots (F1,404=106.80, P<0.0001).

Species diversity was assessed for the understory plant community using

Shannon’s Diversity Index (H′) in 2006. Analysis of variance revealed higher diversity

in the Cut plots (H′=0.56) where there was greater disturbance compared to the Uncut plots or undisturbed controls (H′=0.43) (F1,96=8.43, P=0.005) (Fig 2.5). CCA analysis

was performed to identify the species-specific distribution patterns. Differences in

species composition were detected between the Cut and Uncut plots (Fig. 2.6). The

Uncut and Cut plots were both strongly correlated with Axis 1 (both R2 = 0.574)

determining each end of the scale. In addition, GSF was also associated with Axis 1 (R2

= 0.229) in the direction of the Cut plots. The abiotic factors that correlated the highest with Axis 2 were as follows: number of sunflecks (R2 = 0.241), average duration of each

sunfleck (R2 = 0.256), and soil moisture (R2 = 0.305). While only 3% of the variance was

explained by either Axis, it was the location of non-native plant species that was of

Figure 2.4 Soil compaction in Cut and Uncut plots at five different depths in the soil profile from 2005.

Treatments are defined as: Cut — the Ohio Department of Agriculture implemented

EAB eradication protocol in which an EAB infested ash tree and all ash trees within a

0.8km radius of the infested tree were cut down using heavy-tracked trucks and skidders;

Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback. Error bars represent + 1 SE of the mean.

Figure 2.5 Shannon’s Diversity Index (H′) for understory plant community in Uncut and Cut plots.

Significantly greater diversity occurred in Cut plots (P=0.005). Treatments are defined as: Cut — the Ohio Department of Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

Error bars represent + 1 SE of the mean.

Figure 2.6 Canonical Correspondence Analysis (CCA) ordination plot from 2006.

CCA generated with the percent cover of 73 plant species from 1 x 1 meter subplots. An

X denotes location of 13 non-native invasive species. The black circle identifies the location of Fraxinus spp. seedlings. Treatments are defined as: Cut — the Ohio

Department of Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

interest. In Fig. 2.6 each dot represents a species or a cluster of species with non-native plants identified by black X’s. Out of 73 species total, 13 (17.8%) were non-native; all but one species were present within Cut plots including 10 that were found exclusively within Cut plots. The total percent cover by understory vegetation was not significantly different between the Uncut and Cut plots (P=0.81); however, the compositional difference of community members was significantly different. There were only 3 invasive plant species, which made up less than 1% of the total herbaceous vegetation cover for all of the Uncut plots combined (Table 2.3). Alliaria petiolata, the dominant invasive species in Uncut plots, was found in 7 of 8 plots but on average had 2.0% cover per plot. Rosa multiflora, the only species not found in the Cut plots, was found in one subplot of a single Uncut plot. In contrast, there were 12 invasive plant species in the Cut plots comprising 18.7% of the total herbaceous cover (Table 2.3). All species listed in

Table 2 are non-native but not all of them are considered invasive. However, 93% of the invasive species cover is attributed to 4 highly invasive plant species; and over half of that cover was attributed to a single species Cirsium arvense with Rhamnus cathartica and 2 species of Lonicera constituting another 40%.

Ash seedlings were found in the understory of all 8 Uncut and all 6 Cut plots.

They comprise 5.2% and 6.3% of the total herbaceous cover for those treatments respectively in 2006 (Fig. 2.6.).

Table 2.3 Invasive plant species in the understory of Uncut and Cut plots listed by total percent cover from all 1 x 1 meter understory subplots.

Included in the table is the number of plots where an invasive species was found. There are 8 total Uncut plots and 6 total Cut plots. Invasive species made up <1% of the total herbaceous cover for the Uncut plots and 18.7% for the Cut plots. Some of the non- native species listed are not generally considered invasive. Treatments are defined as:

Cut — the Ohio Department of Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

Species Total % Cover Number of plots where

(common name) species occur Uncut Alliaria petiolata Plots (Garlic Mustard) 0.8% 7 Urtica dioica (Stinging nettle) 0.1% 2 Rosa multiflora (Multiflora rose) 0.03% 1 Cut Plots Cirsium arvense (Canada thistle) 10% 6 Lonicera japonica (Jap. Honeysuckle) 3.2% 1 Rhamnus cathartica (Buckthorn) 3.0% 5 Lonicera spp. (Bush Honeysuckle) 1.1% 2 Taraxacum officinale (Dandelion) 0.7% 5 Arctium minus (Burrdock) 0.2% 2 Alliaria petiolata (Garlic Mustard) 0.2% 4 Glechoma hederacea (Gill-over-the-ground) 0.1% 1 Prunella vulgaris (Common Selfheal) 0.1% 2 Persicaria maculosa (Ladies Thumb Polygonum) 0.04% 1 Urtica dioica (Stinging nettle) 0.02% 1 Melilotus officinalis (Yellow sweet clover) 0.02% 1

DISCUSSION

Long term impacts of the EAB eradication protocol are unknown; however, differences in the abiotic environment were immediately detected after the implementation of the eradication protocol in 2005. The higher light environments caused by artificially created gaps and the greater degree of soil compaction created a disturbance in the forest which facilitated a secondary spread of invasive plant species. It should be noted that currently eradication is no longer a program objective of the USDA

APHIS Emerald Ash Borer Program as such tree removal is no longer funded.

Herbaceous plants are often used to signify habitat quality or health (Andreas et al. 2004) as plants are sensitive to disturbance events, natural or anthropogenic in nature

(Collins et al. 1985). Factors affecting the composition and diversity of the understory herbaceous community include canopy composition and light availability (Crozier and

Boerner 1984, Hausman et al. 2001). Light availability is often a limiting resource for plant growth. It has been shown that invasive plants species can leaf out earlier

(Harrington et al. 1989) and are capable of attaining higher photosynthetic capacity than native species (Pattison et al. 1998; McDowell 2002) thus outcompeting native plants for light. In forested ecosystems a short duration sunfleck (5-7 minute) can contribute 47-

68% of the potential light for understory seedlings and saplings to live (Canham et al.

1990). From this study, the disturbed Cut plots have significantly higher light environments with average sunfleck duration lasting over 9 minutes. In these high light plots there are greater numbers of invasive species which also make up a greater proportion of the understory cover. Non-native plants are often found in areas of

previous disturbance events. Small and McCarthy (2002) found a greater frequency of non-native species during the spring in resource rich areas with a previous clear-cut disturbance; whereas fewer non-native species were found in mature forests with greater canopy cover.

Consistent with our results, it has been shown that the highest degree of soil compaction occurs in the top 30 cm of the soil profile which is also where most plant root biomass occurs (Wingate-Hill and Jakobson 1982). The effects of soil compaction include a decrease in soil aggregates, soil porosity, and filtration capacity and an increase in soil bulk density, runoff and erosion. Severe soil compaction has been shown to inhibit germination and woody seedling growth (Kozlowski 1999). In addition to the interference with forest regeneration, soil compaction can also contribute to the mortality of standing trees. From a post-partial harvest study, proximity to skid trail was the most important determinant for elevated residual tree mortality with soil compaction identified as the principle cause (Thorpe et al. 2008). The soils in this study have high clay content; and therefore, soil compaction may take decades to recover (Shepperd 1993, Grigal

2000). As a result, this soil compaction will constrain forest regeneration and has the potential to restructure both the understory community, as well as, canopy composition.

It is important to note at this time that it is unclear whether the increase in invasive species is attributed to an increased light environment or to greater soil compaction or to the combination of factors. The continued spread of EAB will ultimately increase mortality of ash trees and thus increase gap formation and light environment. This ash-induced gap formation, however, occurs more slowly over a

period of about six years (Knight et al. 2007). I am currently identifying subsequent changes in the understory plant community and the degree of invasive species establishment under these slower EAB-induced ash diebacks. However, eradication efforts potentially caused a compounding negative effect by creating a rapid increase in the light environment and physically disturbing the soil which subsequently facilitated a secondary spread of invasive plant species. These results address the importance of the magnitude of scale or intensity of disturbance on community structure. Often invasive plant species have the ability to dominate and to outcompete native species in disturbed habitats. Ross et al. (2002) found that increases in anthropogenic disturbance in various fragment sizes reduces native plant species richness and increases invasive species richness. Furthermore, it is unclear whether the invasive species that I detected in this study are “new” arrivals from an external source or were already present in the seed bank.

It is possible that the source seeds of these invasive species were introduced as hitchhikers on the vehicles used during EAB eradication. Ongoing research compares differences in seed bank composition between eradication zones and control areas to determine potential invasion pathways. Addressing these concerns will identify patterns of invasive species establishment at various scales of disturbance, as well as, disturbance tolerance thresholds of the native plant species.

The future of ash trees in this forest depends upon the presence of ash in the seed- bank and continued regeneration. While ash saplings were found throughout the forest and within each of the treatments of this study, little is known about the longevity of ash seeds in the seed bank. The ultimate persistence of ash as a canopy tree based on the

presence of ash saplings may be overly optimistic. Ash saplings are under a secondary pressure caused by deer herbivory. Most of the ash saplings that occurred within the Cut plots showed a shrub-like growth. These ash saplings with stunted growth and multiple branching stems demonstrated characteristic signs of previous browse events (C.

Hausman, personal observations). Under continued deer herbivory pressure, one can only speculate as to the persistence of ash in this system. The eastern hemlock is also particularly susceptible to deer over browsing pressure (Frelich and Lorimer 1985). In areas infested with the hemlock wooly adelgid (HWA), Weckel et al. (2006) have documented a lack of hemlock sapling regeneration across a 40-year time span.

Furthermore, there was also a lack of regeneration of white ash (Fraxinus americana) in the HWA-infested areas. This study demonstrates the potential indirect effect invasive forest species have on non-target species. HWA and EAB represent parallel examples of high impact invasive pests capable of restructuring forest composition. The combined pressures of the forest pest along with deer browse may ultimately lead to the functional extirpation of these two trees from their respective forest systems.

As of 2006, the USDA APHIS had spent $100 million dollars on attempted eradication of EAB and on research and reforestation efforts (Lucik 2006). The emerald ash borer eradication efforts did not effectively control the spread of this beetle and may have unintentionally facilitated a secondary establishment of invasive plant species. The initial colonization of invasive plants exposes vulnerability in the habitat with potentially lasting and detrimental impacts for the native plant community. Once established, invasive plant species can alter ecosystem functioning and reduce native plant species

abundance and survival. For example, the invasive Amur honeysuckle (Lonicera maackii) negatively affects plant species richness and abundance of herbaceous and tree seedlings growing below L. maackii crowns (Collier et al. 2002). Even though lasting impacts on the environment are yet to be determined, preliminary evidence indicates that at least some of detrimental damage (i.e., soil compaction) caused by the eradication of

EAB may likely take decades or longer to recover (Shepperd 1993, Grigal 2000). This project exposes a potential and yet substantial concern for government agencies and land managers. With the likely continued spread of the emerald ash borer, one must question how ash tree removal should be implemented in sensitive natural areas. Thus far EAB eradication efforts have been expensive and labor intensive with localized but detrimental ecological consequences. The establishment of invasive species in the wake of eradication efforts adds a new variable for calculating future cost-benefit analysis of control or management objectives. Therefore, in the pursuit of finding suitable forest pest eradication measures, greater consideration should be given to determining disturbance intensity and the ecological impacts that influence invasive plant species establishment. Sometimes not implementing eradication is the best option.

ACKNOWLEDGEMENTS

The authors would like to thank Barb Andreas, Denny Cooke and two anonymous reviewers for comments and suggestions on a previous version of this manuscript. I thank

Tim Gallaher, Tim Schetter, Marty Overholt, Bob Jacksy and the rest of the Toledo Area

Metropark staff and volunteers for their support and assistance in the field. I also thank

Brendan Morgan, Mike Monfredi, Maureen Drinkard, Justin Montemarano, Dylan Stover and Wade Schock for their field support and assistance throughout this project. This project was financially supported with funding from Art and Margaret Herrick Research

Grants of the Department of Biological Sciences at Kent State University (KSU), Ohio

Biological Survey, Metropark District of the Toledo Area, and the KSU Graduate Student

Senate.

CHAPTER 3

Disturbance facilitates a secondary spread of invasive plant species: management

concerns for emerald ash borer infested forests

ABSTRACT

Disturbance events can create changes to the local environment that fundamentally alter the response of the plant community. When the emerald ash borer was first identified in the United States in 2002, eradication efforts were implemented to control the beetles’ spread. Preliminary research on the environmental impacts caused by the eradication program found accelerated canopy gap formation and thus greater light reaching the forest floor. In addition, tracked vehicles used during the eradication efforts caused significant soil compaction. The result of this habitat disturbance was a measurable shift in community composition, predominately attributed to the establishment of invasive plants (Hausman et al. 2010). Therefore, proper tree removal methods are needed to minimize the compounding negative effects on plant communities.

The current project seeks to understand patterns of and changes to plant community assemblages across a gradient of disturbance intensities. The disturbance gradient reflects different methodologies for managing ash tree removal in EAB-infested

50 51

forests. The high intensity disturbance was caused by the Ohio Department of

Agriculture-implemented EAB eradication efforts in which all ash trees were cut down using heavy-tracked trucks and skidders similar to those used in selective logging practices. An intermediate level of disturbance was created by a protocol in which ash trees were cut down; although these were felled by hand using chainsaws and the downed trees were left in place. The last disturbance level reflects a “natural” disturbance in which ash trees were not removed but allowed to progress through the characteristic stages of EAB infestation and natural canopy dieback. Results of this research will be used to indentify how the plant community responds to varying degrees of disturbance.

My ultimate objective is to determine an effective tree removal or silviculture practice for

EAB-infested forests in order to minimize the establishment of invasive plant species.

INTRODUCTION

The research presented in Chapter 2 described a comparative field study between areas that received EAB eradication efforts and areas that did not receive any eradication treatment and, therefore, still had ash trees standing. The research described here are the results of an experimental study that identified the outcome of ash tree removal as a management strategy for EAB-infested forests.

Disturbance events cause significant changes in ecosystem structure and function

(Pickett and White 1985, Gibson 2002, Potter et al. 2005). These events can be grouped into three categories: (1) physical disturbances, including: fires, hurricanes, tornadoes,

land-slides, flooding, erosion, and volcanic eruptions; (2) biogenic disturbances, including: the impact caused by herbivores, pathogens, and predators; and (3) anthropogenic disturbances, including: logging and deforestation, changes in land use

(agriculture and urbanization), pollution, and movement of exotic species to new locations outside of their native range. These disturbance processes can be characterized and assessed by measuring the severity, intensity, timing, frequency, and spatial distribution of a disturbance (Barnes et al. 1998, Spies and Turner 1999). Disturbance events not only change the physical characteristics of the environment and their resources, but they also alter population and community structure ultimately disrupting ecosystem functions.

Disturbance events caused by exotic forest pest insects have been shown to alter ecological processes through direct and indirect effects (Gandhi and Herms 2010b). The magnitude of the disturbance is determined by three characteristics of the pest species:

(1) mode of action, which identifies how the pest attacks its host (leaf defoliator, pathogenic fungus); (2) host specificity, which identifies whether a single or multiple host species are affected (species-specific, generalist); and (3) virulence, which identifies the degree of mortality (widespread, rate of mortality) (Lovett et al. 2006). These disturbance events have both short and long term effects on forest condition (Lovett et al.

2006).

The invasion of the emerald ash borer (EAB) into the United States and our response to it are creating forest ecosystem disturbances through a combination of biogenic and anthropogenic effects. EAB is a wood-boring beetle that has moderately

high host specificity for ash trees (genus Fraxinus). While the adult completes its life- cycle on the leaves of the ash tree, the most significant damage occurs as the larvae feed on the cambium tissue thus interfering with the trees ability to transport water and nutrients. EAB causes ash mortality within 2-4 years once a tree shows signs of infestation (Herms et al. 2004). Thus far, all species of ash with a range overlapping the

EAB invasion are susceptible to attack. Accordingly, these combined effects are likely to cause long-term structural change in the ecosystem. This disturbance potential stimulated the eradication efforts to contain the beetles’ spread.

In addition to the impact of disturbances caused directly by EAB, anthropogenic disturbance events have occurred through the EAB eradication efforts that were implemented soon after the beetles’ discovery. Even though the eradication efforts were on a smaller scale than most timber harvests, the environmental impacts were similar.

While the eradication protocol targeted only Fraxinus species, temporary roads comparable to skid roads created during logging were created to access sites (Feldpausch et al. 2005). The eradication protocol accelerated the formation and size of canopy gaps within the forest which increased the duration and intensity of light penetrating through to the forest floor. Furthermore, the vehicles used during the eradication efforts caused significant soil compaction. Significant soil compaction typically occurs when heavy machinery is used during timber harvesting (Whitman et al. 1997). This change in the abiotic environment resulted in an understory plant community with 18% of plant species richness as non-native (Hausman et al. 2010).

Anthropogenic disturbances, such as selective logging practices, are less severe than large scale clear-cutting; however, there is little consensus in the literature about the impacts selective logging have on species diversity and abundance (Sutherland and

Nelson 2010). Silviculture treatments such as thinning or selective logging of the forest canopy can alter the understory community by increasing the light environment and changing below ground resource availability. While thinning may stimulate herbaceous growth, the habitat disturbance created during the logging process may damage the plant community which may not be able to recover before the canopy closes again (Thomas et al. 1999; Berger et al. 2004).

Changes in habitat heterogeneity may be responsible for the way varying taxa respond to disturbance (Hamer and Hill 2000, Hill and Hamer 2004). Selective logging creates spatial mosaics on the landscape that can affect the understory plant community.

These changes are dependent upon the proportion of forest that is converted to gaps, as well as, the number of access roads and skidder tracks created during the logging process.

While there is variability in plant community response to silviculture treatments between sites, the non-native plant response in various spatial scales and in time-since-treatment applications is positively correlated with intensity of disturbance (Sutherland and Nelson

2010).

The establishment of non-native species depends upon the severity of the disturbance, the presence of non-native propagules and the species-specific competitive interactions between native and non-native plants (Lonsdale 1999). Land use subjected to high disturbance intensities (e.g., logging) can result in an increase of the non-native

plant contribution to the soil seed bank (Korb et al. 2005). Future disturbance events will, therefore, create a greater non-native propagule pressure. Invasive species are typically not found in the seed bank of the forest interior, but do occur within a limited 3 meter edge from the forest margin (Gehlhausen et al. 2000, Honnay et al. 2001,

Devlaeminck et al. 2005). Therefore, selective logging practices are creating skid roads which function as edge habitat and also function as a non-native seed source for disturbance events that occur within the forest interior. Kneitel and Perrault (2006) suggested that changes in species richness and composition are most important for facilitating invasions, and this may be the result of increased vulnerability to invasion in a post-disturbance community dominated by poor competitors. They also argued that understanding how species invasions are facilitated by community characteristics can provide further insight into identifying when communities are most vulnerable and how to prevent invasions.

EAB eradication is no longer an active program because it was determined that by the time an infestation was discovered and treated, the beetle had dispersed outside of the eradication zone. However, the spread of EAB continues and the removal of infested dead and dying ash trees remains a necessary practice to minimize safety hazards in active recreation areas. Understanding disturbance processes within forests allows for the recognition of structural and compositional patterns which is important for prioritizing management efforts (Newton 2007). Therefore, land managers will need to consider methods for proper tree removal to minimize any compounding effects that may adversely affect the native plant community.

Research Objectives

My objectives are: (1) to create various disturbance intensities to identify corresponding habitat changes, (2) to determine invasive plant establishment patterns across the disturbance gradient, (3) to identify whether invasive species exist in the seed bank or colonize from an outside source and (4) to propose land management practices to minimize the likelihood of invasive plant establishment.

This project compares high disturbance areas that received the eradication treatment and had all ash trees removed using tracked vehicles (Cut plots) to moderately disturbed areas with ash trees removed by hand using chainsaws to eliminate vehicle- caused soil compaction (Cut without Compaction) to “undisturbed” areas with ash trees still standing (Uncut plots). The Cut without Compaction treatment represents an intermediate level of disturbance and was added to isolate the impact of an increased light environment without the effects of soil compaction.

MATERIALS AND METHODS

Site and Experimental Design

This project was conducted in Pearson Metropark, Lucas County, in Northwest

Ohio. In order to address multiple disturbance intensities, this project builds off the experimental design described in Chapter 2. To summarize, the Cut and Uncut plots were established in 2005 at the time the Ohio Department of Agriculture (ODA) were

conducting EAB eradication efforts by cutting down all ash trees within a 0.8 km radius of an infested tree. Plots established in the EAB eradication zones represent the “Cut” treatment and all ash trees were removed using tracked vehicles as described in Chapter

2. Plots established in areas that did not have any ash removed represent the “Uncut” or control treatment. The Uncut treatment represents a natural dieback disturbance as these ash trees show progressive signs of EAB infestation with canopy dieback. In 2007 an intermediate disturbance treatment, designated as “Cut without Compaction,” removed all ash but without using vehicles therefore eliminating soil compaction. The ash trees were cut down using chainsaws and felled so that most of the tree biomass landed outside of the plot. The trunks of the trees were then cut into smaller sections and rolled out of the plot in order to minimize the downed woody debris. All ash trees within the plot and within a 10-meter buffer of the plot were cut down.

Each plot measured 20 x 25 meter and there were eighteen total plots (6 per treatment). Because the Cut without Compaction treatment was added in 2007, all comparisons between treatments are based on a “Response time” or the time from when the disturbance treatment was applied. Therefore, all measurements (biotic and abiotic) compared the data from the Cut and Uncut treatments in 2005 and with the Cut without

Compaction treatment in 2007.

Abiotic Sampling

The light environment was assessed using hemispherical photographs, and soil compaction was measured using a soil penetrometer as previously described in Chapter 2.

Soil temperature and soil moisture were also recorded as described in Chapter 2. In addition to the penetrometer readings, soil bulk density was measured directly. Soil bulk density (g/cm3) was measured by calculating the mass of dry soil (in grams) divided by

the volume of the sample. Five soil cores were collected within each plot (one from each

of the four corners and one from the center). Soil cores were collected using an AMS

Slide Hammer with a 2-inch diameter soil core liner. Each core weighed 200 grams on

average and was approximately 70 cm3 in volume. Once weighed, the cores were broken

up by hand and sifted through an 8 mm soil sieve to homogenize the sample. All

apparent root material was removed. From this mix, a 20-30 gram subsample was

weighed wet (fresh weight) and then dried in a drying oven at 65 oC for 48 hours. Dry

weight was measured and the difference between wet and dry weight was then multiplied

by the total number of subsamples in a complete soil core. This mass of dry soil

measurement was then divided by the soil core volume to determine bulk density. These

soil cores were also used to determine the percent of organic carbon, which was

calculated using the Loss-On-Ignition (LOI) method (Sparks et al. 1996). This technique

measures mass loss following high temperature combustion. Approximately 3 grams of

the dried soil from each core was incinerated in a muffle furnace at 400 oC for 16 hours.

Samples were then cooled to room temperature in a sealed desiccator with calcium

chloride (CaCl2) before weighing one last time. The LOI was then calculated as (Dry

weight-Furnace weight)/Dry weight *100.

Biotic Sampling

The herbaceous understory was determined using % cover of all species present from seven-1 m2 subplots within each plot. There were 4 subplots established in each of

the corners, 2 plots established at half the distance (12.5 m) along each of the 25 meter sides and the last subplot was located in the center of the plot. Vegetation measurements were conducted in the same manner as previously described in Chapter 2. All plant community measurements were collected annually for 3 years from 2007-2009.

Seed Bank Survey

In April 2007, ten soil cores measuring 4 inches in diameter and ~6 inches in depth were taken from each plot (180 soil cores total). The soil cores were collected using a 4-inch diameter PVC pipe that had a 45o blade edge at one end and a handle

attachment at the other end. The corer was hammered into the soil using a rubber mallet

to a depth of 6 inches, turned and pulled to extract the soil core. The cores were taken at

regular intervals across the entire 20 x 25 meter plot. Each core was labeled, placed into

resealable plastic storage bags and taken to the Kent State University greenhouse. Ten-

inch pots were partially filled with a general potting soil (Fafard 4P), and then each soil

core was broken up and evenly scattered across the top of the potting soil. The pots were

randomly placed within the greenhouse and were rotated once a week to avoid edge

effects. Germination was recorded 3 times per week for 8 weeks by placing color-coded

toothpicks near the germinating seedling. The pots were usually watered daily until germination occurred and then as needed throughout the duration of the experiment.

Statistical Analysis

Immediately following the disturbance application a nested analysis of variance was used to determine the effects of treatment (Cut, Cut without Compaction and Uncut), block and treatment nested within block on the light environment species diversity and richness, bulk density and % organic carbon. The global site factor (GSF) and average duration of each sunfleck were log transformed before analysis in order to meet the assumptions of normality. Multivariate analysis of variance was used to determine the effects of treatment (Cut, Cut without Compaction and Uncut), depth of soil profile (7.6,

15.2, 22.9, 38.1centimeters), soil moisture and the interaction of treatment and soil depth on soil compaction. All stated analyses were performed using JMP version 3.1 (SAS).

Plant community composition was analyzed with ordination using canonical correspondence analysis (CCA) using PC-ORD version 5 (McCune and Medford 1999).

Ordinations were performed with percent cover data from 84 species with abiotic measurements (GSF, number of sunflecks, total sunfleck duration, average duration of each sunflecks, soil compaction, soil moisture and temperature). Treatment (Cut, Cut without Compaction and Uncut) was also included as a covariate in the analysis.

RESULTS

Abiotic: Light Environment.

The over-all light environment was characterized by the Global Site Factor (GSF) which identifies the proportion of global radiation (direct + diffuse) that occurs in the open sky compared to that which occurs under the forest canopy. The GSF was significantly different among treatments (F2,89 = 25.16, P < 0.0001) (Fig. 3.1); upon

pairwise comparison using Tukey HSD, there was greater GSF in the Cut treatment but

no difference identified between the Uncut and the Cut without Compaction. Global Site

Factors for the three treatments were: 0.24 for Cut, 0.18 for Cut without Compaction and

0.15 for the Uncut. The number of sunflecks was found not to vary significantly between

treatments (F2,89 = 1.76, P = 0.18) even though there were typically more 2-3 more

sunflecks in the Uncut treatment compared to the Cut without Compaction or Cut (Fig.

3.2). However the average duration of each sunfleck and the total duration of sunflecks

were different among treatments (F2,89 = 20.12, P < 0.0001 and F2,89 = 17.74, P < 0.0001

and respectively). Based on Tukey HSD comparison, all treatments were different from

each other (Fig. 3.3). The Cut treatment had longest average sunfleck duration at 10 minutes and 55 seconds while the Cut without Compaction treatment had the shortest duration at 7 minutes and 12 seconds. The difference between the shortest total daily duration in the Cut without Compaction and the longest in the Cut treatment was 1 hour and 26 minutes

0.3

0.25

0.2

0.15

0.1

0.05 Global Site Factor (GSF) 0 Uncut Cut1 w/out Cut

Figure 3.1 Light environment for the three disturbance treatments as determined by

GSF.

The Global Site Factor (GSF) identifies the proportion of global radiation (direct + diffuse) of the open sky compared to that which occurs under the forest canopy. The light environment of the Cut plots was significantly (P < 0.0001) greater than that of the

Uncut and Cut without Compaction treatments. Treatments are defined as: Cut — the

Ohio Department of Agriculture implemented EAB eradication protocol in which an

EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Cut without Compaction (Cut w/out)

— ash tree were felled by hand using chainsaws and the downed trees were left on site;

Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback. Error bars represent ± 1SE of the mean.

32 31 30 29 28 27 26 # of Sunflecks 25 24 23 Uncut Cut1 w/out Cut

Figure 3.2 The average number of sunflecks between the three disturbance levels.

The average number of sunflecks did not vary between the three treatments (P = 0.18).

Treatments are defined as: Cut — the Ohio Department of Agriculture implemented

EAB eradication protocol in which an EAB infested ash tree and all ash trees within a

0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Cut without Compaction (Cut w/out) — ash tree were felled by hand using chainsaws and the downed trees were left on site; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback. Error bars represent ± 1SE of the mean.

12:58 11:31 10:05 08:38 07:12 05:46 04:19 02:53 01:26

Avg. Avg. sunfleck duration (min) 00:00 Uncut Cut1 w/out Cut

Figure 3.3 The average duration of each sunfleck between treatments.

There was a significant treatment effect (P < 0.0001) with longer sunflecks occurring in the Cut treatment. Treatments are defined as: Cut — the Ohio Department of

Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy- tracked trucks and skidders; Cut without Compaction (Cut w/out) — ash tree were felled by hand using chainsaws and the downed trees were left on site; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

Error bars represent ± 1SE of the mean.

Abiotic: Soil Compaction.

The use of tracked vehicles to remove ash trees caused significant soil compaction in the Cut treatment compared to the Uncut and Cut without Compaction treatment (F2,519

= 28.19, P < 0.001) with no difference detected between the Uncut and Cut without

Compaction treatment (Tukey’s HSD) (Fig. 3.4). These results were consistent at every

depth of the soil profile (7.6 - 38.1 cm). Similar compaction trends were identified using

soil bulk density even though those results were not significant, (F2,72 = 2.37, P = 0.09)

(Fig. 3.5). There was also no difference detected in the % soil organic carbon between treatments (F2,72 = 0.844, P = 0.1696) (Fig. 3.6).

Biotic: Species Diversity.

Within the first 3-4 months following the disturbance treatment, no difference in

diversity using Shannon’s Diversity Index (H’) was detected between treatments (F2,106 =

0.96, P = 0.386) (Fig. 3.7). However, by the following year (Year 2) there was a

treatment effect (F2,108 = P < 0.0001) with greater diversity detected in the Cut and Cut

without Compaction treatment. In the third response time period (Year 3) the results

were similar to Year 2 (F2,107 =, P < 0.01). While the Cut without Compaction treatment

did have greater diversity compared to the Uncut treatment, there was no increase in

diversity following disturbance as occurred in the Cut treatment. There was a steady

increase in species diversity over time for the Cut treatment (H’=0.45, 0.56, 0.60), while

350

300

250

200

150 Cut plots 100 Soil Compaction (PSI) Compaction Soil Cut w/out

50 Uncut plots

0 7.6 15.2 22.9 30.5 38.1 Depth (cm)

Figure 3.4 Soil compaction as measured with a penetrometer.

There was significantly greater compaction at every depth of the soil profile in the Cut plots (P < 0.001). Treatments are defined as: Cut — the Ohio Department of Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Cut without Compaction (Cut w/out) — ash tree were felled by hand using chainsaws and the downed trees were left on site; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback. Error bars represent ± 1SE of the mean.

1.95

1.9

1.85

1.8 Soil Bulk Density 1.75

1.7 Uncut Cut1 w/out Cut

Figure 3.5 Soil compaction as measured by bulk density.

Bulk density was not found to be statistically significant between treatments (P = 0.09).

Treatments are defined as: Cut — the Ohio Department of Agriculture implemented

EAB eradication protocol in which an EAB infested ash tree and all ash trees within a

9.6 9.4 9.2 9 Carbon

8.8 8.6 Organic 8.4 % 8.2 8 7.8 Uncut Cut1 w/out Cut

Figure 3.6 Soil organic matter, as measured by loss on ignition (LOI).

There were no differences in organic matter detected between treatments (P = 0.1696).

Treatments are defined as: Cut — the Ohio Department of Agriculture implemented

EAB eradication protocol in which an EAB infested ash tree and all ash trees within a

0.7

0.6 Uncut Cut w/out 0.5 Cut

0.4

0.3

0.2 Species Diversity (H') Diversity Species

0.1

0 Year 1 Year1 2 Year 3 Response Time

Figure 3.7 Shannon’s Diversity Index (H’) for the understory plant community between disturbance treatments.

Greater diversity occurred in the Cut plots (P < 0.000). Treatments are defined as: Cut

— ash tree were felled by hand using chainsaws and the downed trees were left on site;

Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback. Error bars represent ± 1SE of the mean.

diversity in the Uncut and Cut without Compaction treatment remained similar across all

3 years (Uncut: H’ = 0.47, 0.37, 0.47 and Cut without Compaction: H’=0.51, 0.54, 0.56).

While there was no statistical year effect detected for species diversity (F2, = 3.11,

P=0.05), each year following the disturbance there was an increase in diversity for only

the Cut plots. The higher diversity in the Cut without Compaction treatment is attributed

to greater species richness (6.5 species) compared to the Uncut treatment (5.3 species)

(Fig. 3.8).

The diversity indices changed little over time in the Uncut and Cut without

Compaction treatments likely because the understory plant community remained similar

over the three year study. The species richness, or total number of species present, in the

Uncut plots did not change over time (5.5 species in Year 1 to 5.3 species in Year 3).

Similarly, species richness in Cut without Compaction remained comparable (6.5 species

in Year 1 to 6.2 species in Year 3) (Fig. 3.8). Species diversity is greater in the Cut

without Compaction compared to the Uncut treatment and likely attributed to greater

species richness. The Cut treatment however nearly doubled the number of species that

colonized after the eradication disturbance (4.8 species in Year 1 to 8.8 species in Year 2)

(P<0.001). All of those establishing species are non-native and are persisting 3 years

after the disturbance was first performed.

In the year following the disturbance application (Year 2), there was a significant

increase in species richness for the Cut treatment but not for the Uncut or Cut without

Compaction treatment. This response signifies a potential shift in plant community

composition. Therefore, a CCA was performed to identify differences in species-specific

Figure 3.8 Species richness between the three disturbance levels.

There is a significant treatment effect (P<0.001) with the greatest number of species occurring in the Cut plots. Treatments are defined as: Cut — the Ohio Department of

Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy- tracked trucks and skidders; Cut without Compaction (Cut w/out) — ash tree were felled by hand using chainsaws and the downed trees were left on site; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

Error bars represent ± 1SE of the mean.

assemblages between treatments (Fig. 3.9). This ordination technique identified a distinctly different plant community in the Cut treatment. Axis 1 on the joint by plot is correlated with the Uncut treatment (R2 = 0.244) and soil moisture (R2 = 0.250) in one

direction of the scale and with the Cut treatment (R2 = 0.312) and soil compaction (PSI)

(R2 = 0.409) in the other. Axis 2 was also correlated with the Cut treatment (R2 = 0.548),

GSF (R2 = 0.257), soil temperature (R2 = 0.485) and with the Cut without Compaction

treatment (R2 = 0.258). In Fig. 3.9 each “+” represents a species or a cluster of species

with the location of a non-native species is represented by a “+” with an oval. Out of 83

species surveyed in this analysis, 17 or 20.5% of them are non-native species (Table 3.1).

There were a total of 4 non-native species in the Uncut treatment of which 2 of those species were found in all three treatments (Alliaria petiolata and Euonymus fortunei). Of

the remaining 2 species, Urtica dioica had a mere 0.1% total coverage and was also

found in the Cut plots. The last species (Lysimachia nummularia) was only found in a

single Uncut plot but ultimately contributed 92% of the total invasive species coverage

for the treatment. Without this species, invasive plants would contribute <1% of the total

herbaceous cover for the Uncut treatment. The Cut without Compaction treatment had

twice as many non-native species (8 spp.) that comprised 5.1% of the total vegetative

cover. There were 5 of these species that were also found in the Cut treatment. Three

species, however, were unique to this treatment with 2 of them, Rosa multiflora and

Solanum dulcamara, found each in only a single plot which made up <1% of the total

herbaceous cover. The last species (Viburnum opulus var. opulus) is a woody shrub and

was also only found in a single plot; however, this species comprised 62.3% of the

Figure 3.9 Canonical Correspondence Analysis ordination plot.

CCA ordination was generated with the percent cover of 83 plant species from the 1m x1m subplots. Each “+” represents a species or cluster of species. The “+” with ovals indicate the location of a non-native species. Treatments are defined as: Cut — the Ohio

Department of Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Cut without Compaction (Cut w/out) — ash tree were felled by hand using chainsaws and the downed trees were left on site;

Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

Table 3.1 Invasive plant species in the understory listed by total percent cover from all 1 x 1 meter understory subplots within each treatment.

There were a total of 6 plots per treatment. Included is the number of plots where an invasive species was found. Invasive species made up 18.7% for the Cut plots, 5.1% for the Cut without Compaction and 9.9% of the total herbaceous cover for the Uncut plots.

Some of the non-native species listed are not generally considered invasive. Treatments are defined as: Cut — the Ohio Department of Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Cut without Compaction (Cut w/out) — ash tree were felled by hand using chainsaws and the downed trees were left on site; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

Total %

Species Cover Number of plots where (common name) species occur

Uncut Plots Lysimachia nummularia (Moneywort) 9.1 % 1 Alliaria petiolata (Garlic Mustard) 0.7% 5 Urtica dioica (Stinging nettle) 0.1% 2 Euonymus fortunei (Winter creeper) 0.001% 1

Cut w/out Viburnum opulus var. Plots opulus (European cranberrybush) 3.1% 1 Alliaria petiolata (Garlic Mustard) 0.8% 6 Rosa multiflora (Multiflora rose) 0.7% 1 Euonymus fortunei (Winter creeper) 0.2% 2 Rhamnus cathartica (Buckthorn) 0.1% 2 Cirsium arvense (Canada thistle) 0.1% 1 Taraxacum officinale (Dandelion) 0.001% 1 Solanum dulcamara (Bittersweet nightshade) 0.001% 1

Cut Plots Cirsium arvense (Canada thistle) 10% 6 Lonicera japonica (Jap. Honeysuckle) 3.2% 1 Rhamnus cathartica (Buckthorn) 3.0% 5 Lonicera spp. (Bush Honeysuckle) 1.1% 2

Taraxacum officinale (Dandelion) 0.7% 5 Arctium minus (Burrdock) 0.2% 2 Alliaria petiolata (Garlic Mustard) 0.2% 4 Glechoma hederacea (Gill-over-the-ground) 0.1% 1 Prunella vulgaris (Common Selfheal) 0.1% 2 Persicaria maculosa (Ladies Thumb Polygonum) 0.04% 1 Euonymus fortunei (Winter creeper) 0.04% 1 Urtica dioica (Stinging nettle) 0.02% 1 Melilotus officinalis (Yellow sweet clover) 0.02% 1

invasive plant cover. The Cut treatment had the greatest number of non-native plant species (13spp) which contributed to 18.8% of the total vegetative cover. Seven of those species were only found in the Cut plots and included 2 honeysuckle species Lonicera spp. which made up >23% of the invasive species coverage. While Canada thistle

(Cirsium arvense) was found in both the Cut without Compaction and Cut treatments their coverage differed significantly. In the Cut without Compaction treatment, C. arvense was found in a single plot and comprised 1.8% invasive species coverage. While in the Cut treatment, C. arvense was found in every plot and comprised 53% of the total invasive species coverage.

Seed Bank

A total of 3,453 plants germinated from all pots combined. On average there were

27.3 plants per pot from the Cut plots as compared to 19.9 plants from Uncut plots and

10.3 plants from Cut without Compaction plots. Germination from soil cores in the Cut plots accounted for 48% of the overall total. Fig. 3.10a shows total accumulated germination over the course of the experiment. Upon closer inspection, all soil cores that were collected within a single Uncut plot had substantially greater germination compared to all other Uncut plots. The greater germination in this particular plot was attributed to a single species (Lysimachia nummularia L.). Commonly known as moneywort, this invasive species forms a creeping mat along the forest floor. As previously mentioned this plant is found within a single Uncut plot but was responsible for 9.1% of the 9.9% total invasive species cover for the entire Uncut treatment. The high rate of germination

Figure 3.10 Total accumulated germination over an 8-week sample period. a. Germination was scored 3 times a week for all three treatments (Cut, Cut without

Compaction and Uncut). Greater total germination occurred from the soil cores of the

Cut treatment. b. Total germination over the 3-week sample period with Lysimachia nummularia removed from one of the Uncut plots. Treatments are defined as: Cut — the

Ohio Department of Agriculture implemented EAB eradication protocol in which an

EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy-tracked trucks and skidders; Cut without Compaction (Cut w/out)

— ash tree were felled by hand using chainsaws and the downed trees were left on site;

Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

3.10a.

3.10b.

for this species is not representative for the Uncut treatment as a whole; and therefore when this Uncut plot is removed from the analysis, there is nearly identical germination for the remaining Uncut and Cut without Compaction plots (Fig. 3.10b). When the.

Lysimachia plot is removed, the reassessed average number of plants per pot becomes

10.8, which is again comparable to the Uncut treatment.

DISCUSSION

Plant communities are sensitive to and respond differently to disturbance events of varying intensity. Connell (1978) proposed the intermediate disturbance hypothesis that suggests the highest species richness within a community occurs at moderate levels of disturbance. While some disturbance regimes are often considered a critical element in maintaining native plant communities (historical fires, gaps in forests, etc.) (Davies et al. 2009), there is a reduction in biodiversity when disturbance levels are too high with fewer number of species and less diversity which can then alter ecosystem functioning and productivity (Hooper et al. 2005). Silviculture treatments, including thinning and selective logging, result in changes to the abiotic environment (increase in light and resource availability, decrease in plant competition) (Battles et al. 2001) that provides opportunities for invasive species to thrive (Hobbs and Huenneke 1992). This leads to the potential conflict between disturbance supporting native plant richness and disturbance facilitating the spread of invasive plant species.

The disturbance gradient created in this study represents significant environmental changes as a result of ash tree removal in EAB-infested forests. The experimental disturbance, the application of the ash tree removal protocol (eradication

Cut or Cut without Compaction), changes the forest composition and creates canopy gaps. Even though several ash trees were removed in the Cut without Compaction treatment (3.5 per plot on average), there were more than twice as many (9.2 per plot on average) trees removed during the eradication (Cut treatment). Average gap sizes in temperate hardwood forests range in size from 280-375 m2 (Frelich and Lorimer 1985);

whereas total plot size in this study was 500 m2. While the light environment, as

measured by GSF, was only higher in the Cut plots, the size of the gaps created by tree

removal for both Cut and Cut without Compaction treatments are likely still comparable

to average temperate forest gap sizes. The main environmental impact of the disturbance

treatments was the increase in soil compaction caused by the tracked vehicles during the

eradication efforts.

Soil compaction in the Cut plots was found to be significantly greater when using

a penetrometer but was not found to be greater when soil bulk density was assessed. The

discrepancy between these two measurements may be the result of the type of soil that is

found at Pearson Metropark. Pearson is characterized by having Latty soils which are

dense clay soils with an average bulk density measurement of 1.88 mg/m3 — a normal

soil bulk density measurement can range from 1.0-1.6 mg/m3 for clay soils (Aubertin and

Kardos 1965). While the differences in bulk density between our treatments were not

significant, the compaction trend was similar to that measured using the penetrometer.

The soil compaction resulting from the EAB eradication protocol is similar to the soil compaction that results from harvesting equipment during logging practices (Hatchell et al. 1970) which may take several decades to recover (Shepperd 1993, Grigal 2000).

Soil organic matter promotes the aggregation of soil particles which increases porosity and reduces bulk density. There is an inverse relationship between bulk density and organic matter such that it is reversible when organic matter is used as an amendment for compacted soils (Rivenshield and Bassuk 2007). In timber harvesting the impact on soil properties include compaction, nutrient loss and increased soil runoff and erosion

(Dyck et al. 1994). My measurements of organic matter showed no difference between treatments. However there was a minor decrease in % organic carbon throughout the disturbance gradient such that the high intensity disturbance Cut treatment had the least amount of % organic carbon.

Disturbance effects are variable and can have seemingly conflicting plant community impacts. Plant community response to disturbance events is dependent upon the scale of disturbance and the amount of invasive species propagule pressure. A comparison of understory plant composition before and after the creation of a canopy gap found nearly double the number of understory species after the canopy gap was created

(Aikens et al. 2007). In another study of low-intensity thinning there was not a significant change in herbaceous abundance, but there was an increase in flowering of some existing plant species (Lindh 2008). However, other studies have documented significant decreases in plant richness following harvesting disturbances (Roberts and

Zhu 2002). Selective logging practices have been determined to have directly caused species extinctions (Pimm and Raven 2000).

The results of this study identified a treatment - invasive plant species establishment effect. The treatments in this study represented three possible management scenarios against the spread of EAB in order to identify differences in forest community response. The Cut plots in the eradication zone reflected an intense level of anthropogenic disturbance, characterized by a high light environment and compacted soils. These abiotic changes resulted in the greatest vulnerability to invasive plant species establishment with invasive plants occupying ~19% of the total herbaceous cover.

The Cut without Compaction plots reflected an intermediate level of anthropogenic disturbance with a slightly higher light environment but without soil compaction. The understory plant community was less vulnerable to invasive plant establishment with only 5.1% invasive plant cover. The Uncut treatment had 9.9% invasive plant cover however 9% of was attributed to a single invasive species (Lysimachia nummularia) that was found in only one of the plots. Furthermore, there was no change in species richness over the three years of this study for either the Uncut or Cut without Compaction treatment.

The most problematic invasive species in this study were the invasive shrubs

Rhamnus cathartica (Buckthorn), Lonicera (Honeysuckle) and Viburnum opulus var. opulus (European cranberrybush). Invasive woody establishment has been previously documented to affect native plant populations by preventing further recruitment through germination (Ens and French 2008). However while the woody invasive species have

become established in the high disturbance areas, none of those species were found germinating from the seed bank. The lack of germination from the soil cores collected could indicate that those species are colonized after the disturbance event from an external source and were germinating immediately rather than being incorporated into the seed bank.

Previous studies of invasive species occurrence in seed banks have identified the patterns of a field-to-forest gradient with higher invasive seed densities found at the forest edge compared to the interior (Devlaeminck et al. 2005). This suggests that forest edges function as a barrier to seed dispersal and may become compromised as access roads or skid trails are created. A higher incidence of the invasive species Celastrus orbiculatus has been found along logging roads when compared to adjacent areas (Silveri et al. 2001). In addition, higher non-native plant cover was found along roadsides and skid trails when compared to the adjacent stand; although the authors attributed site differences in environmental conditions rather than propagule pressure for non-native abundance patterns (Nelson et al. 2008). While more research is needed, it is possible that as the eradication protocol penetrated into the forest interior, new “edges” were created which allowed for easier access of wind or bird dispersed invasive seed.

Forest Stand Maintenance

Understanding the effects of different levels of disturbance on species diversity and ecosystem function is fundamental to conservation planning, but our understanding of those effects is still inadequate. Such information would allow prediction of changes

in vegetation communities following disturbances. It is important to determine whether a forest is able to withstand or tolerate a particular anthropogenic disturbance (Newton

2007). Therefore, from a land management perspective an assessment of EAB-infested forests should identify the appropriate ash tree removal technique in order to minimize disturbance. The physical disturbance caused by the use of tracked vehicles during the eradication protocol may have facilitated a greater establishment of invasive plant species. If tree removal is necessary, cutting by hand with chainsaws has the least disturbance impact on the native plant community. Caution should be taken when cutting trees close to trails and edges as these are likely source locations for invasive plants.

Long-term forest stand maintenance should consider: (1) the ash tree removal technique necessary to be effective, (2) the impact each technique has on the native plant understory and (3) the proximity of the treatment area to potential invasive seed sources (i.e., edges).

CHAPTER 4

Genetic structure of Green Ash (Fraxinus pennsylvanica) in the State of Ohio:

Implications for the establishment of ex situ conservation protocols.

ABSTRACT

The emerald ash borer poses a significant threat to the long-term viability of ash populations across the state of Ohio, as well as, throughout their range. Native to Asia, the emerald ash borer (EAB) has recently been introduced to the United States. This beetle completes its life cycle on ash trees within the Fraxinus genus. There are 5 native species of ash native to Ohio. None of the species have protective status, and all of them appear to be susceptible to EAB infestation. Unfortunately, once an infested tree shows signs of decline mortality occurs within 2-4 years (Herms et al. 2004).

Understanding the genetic structure of a threatened or potentially threatened species is an integral component in constructing an ex situ conservation protocol. Proper design of a germplasm collection protocol requires information about the spatial genetic structure of the species of interest. Traditional germplasm protocols have been established and designed for plant species, locally adapted genotypes (landraces) and wild relatives of crop plants that are already rare and endangered. This study is one of

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the first attempts to recognize the potential loss of a genetic resource due to an exotic pest that may reduce a relatively common species to functional extinction. The conservation protocol developed contributes to the preservation of the targeted species (Fraxinus pennsylvanica) and may be extrapolated and applied to other species within the Fraxinus genus. While this study’s sampling area is geographically localized relative to the range of the species, the results indicate spatial, genetic structuring that should be incorporated in future ex situ collection efforts

INTRODUCTION

The goal of ex situ conservation is to capture and preserve the genetic variation, as well as, the biological and economic potential of species of interest. These species may be preserved by growing them in botanical gardens, greenhouses or in other live collections or by collecting and forming seed banks (National Research Council 1993).

An ex situ conservation strategy is relevant at times when there are major threats to the preservation of plant genetic resources due to changes in land use including: agricultural practices, habitat fragmentation and degradation, overexploitation or the attack of exotic pests and diseases. Most botanical ex situ conservation efforts are directed towards rare and endangered species, cultivated or ornamental plants and their wild relatives, and/or arboreal species of economical importance (Goodall-Copestake 2005, Zoro Bi et al.

1998, Feres et al. 2009, Khoury 2010, Xie 2010). However, existing ex situ germplasm collections are often challenging to manage because they can be expensive and they are

difficult to evaluate because of redundancies or duplications within the accessions

(Grenier et al. 2000, van Hintum et al. 2000).

For species that already have existing germplasm accessions, core collections are often created from them that are significantly smaller than the original collection and represent a significant portion of the diversity — at least 70% of the alleles (Brown

1989). In order to create this core collection, several strategies have been developed including the maximization strategy (M strategy) in which the number of observed alleles at each marker locus is maximized (Schoen and Brown 1993). Escribano et al. (2008) found the maximization strategy (M strategy) together with the use of simple sequence repeat marker data (SSR) to be the most efficient method for developing a core collection for a subtropical tree species Annona cherimola. Due to the utilization challenges that exist for established germplasm accessions, I propose designing a germplasm collection protocol that maximizes the genetic diversity represented in an area and minimizes duplications within the accession.

In order to preserve genetic variation it is important to understand potential spatial distribution. In general, it is known that genetic diversity in plant populations is not randomly distributed but is spatially structured (Allard 1988). Several studies have shown that environmental factors such as gradients in latitude (Clegg 1972) and altitude and changes in edaphological conditions (Nevo et al. 1981, Nevo et al. 1983) have some bearing on population genetic structure. In addition, life history traits and breeding systems are also important in defining the distribution of genetic variation at the landscape level (Loveless and Hamrick 1984). Knowledge about the structure of genetic variation is necessary to

decide how to best protect genetic diversity (National Research Council 1993, Engelmann and Engels 2002, Karp 2002).

DNA-based methods have revolutionized our ability to characterize genetic diversity, and can provide the information needed to guide collections efforts. Molecular markers are widely used to identify the genetic structure of plants and are recognized as a key tool when designing sampling protocols that maximize the proportion of genetic diversity captured in ex situ collections (Saura et al. 2008, but see Hamilton 1994).

Moreover, Karp (2002) identifies relevant questions for ex situ conservations that can be addressed using molecular markers: (1) determining which samples should be included

(or excluded) in a germplasm collection, (2) identifying where gaps occur in the collection, (3) identifying when there are duplicates, (4) determining how to multiply samples without loss of diversity, and 5) deciding how to choose material from the germplasm collection for utilization.

Adequate knowledge of the genetic structure could provide critical information when designing future reintroductions of currently threatened plants. For example, existing populations with limited size and distribution could be amended with novel genotypes in order to counteract the effect of inbreeding depression. Future restoration efforts may also utilize germplasm collections to create new populations from individuals with known genetic diversity in order to minimize founder effects. These restoration efforts would be effective only if adequate understanding exists about the genetic variability of a germplasm collection.

The optimal sampling procedure for adequate coverage of genetic variation within a species for ex situ conservation has been examined by other authors (Marshall and

Brown 1975, Yonezawa and Ichihashi 1989, Lawrence et al. 1995a, 1995b, Lawrence and Marshall 1997, Brown and Hardner 2000). Marshall and Brown propose that a sample of 50-100 plants is sufficient in most circumstances to capture most alleles in the population. Similarly, Hawkes (1980) proposed that a sample of 5,000 seeds (100 plants x 50 seeds) should be targeted when collecting from highly heterogeneous populations and 2,500 seeds (50 plants x 50 seeds) for homogeneous populations. Yonezawa and

Ichihashi (1989) developed a probabilistic model that estimates the likelihood of capturing all alleles in a population given their frequencies, number of maternal trees to be sampled and number of seeds to be collected per mother tree. They concluded that a representative sample does not have to come from a large number of maternal plants.

These results contrast with the number of plants previously proposed by Marshall and

Brown (1975), and Hawkes (1980). More recently, Lawrence et al. (1995a, 1995b) proposed that a sample of only 172 total plants is sufficed to capture all genetic variants of a given species; but Brown and Hardner (2000) pointed out that such a number may be too small and suggested that at least 15 trees should be sampled per population to collect alleles that exceed 0.10 in frequency. Furthermore, the Center for Plant Conservation

(CPC) provides guidelines when preparing rare plant collection for ex situ conservation efforts (CPC 1991). They recommend at least 50 individuals from each of up to 50 populations to maximize the genetic diversity represented in the collection (Guerrant et al. 2004, Trusty et al. 2009).

There are several recommendations for developing a germplasm collection; however, the utilization or usefulness of these collections is less evident. Reintroduction efforts of rare and endangered species that utilize ex situ collections are often founded from a limited gene pool. In fact, 79% of recent reintroductions in North America were established with propagules from a single source population (Husband and Campbell

2004). With a limited initial gene pool the long-term survival of these populations still remains uncertain. Therefore, future conservation efforts may benefit from an earlier recognition or detection of threats that compromise a species future before it becomes labeled as rare, threatened, or endangered. This study is one of the first attempts to recommend guidelines for the establishment of an effective ex situ germplasm collection for a plant species before there is permanent loss of a genetic resource.

The United States is currently experiencing a prolific invasion of the emerald ash borer (EAB). Native to Asia, Agrilus planipennis Fairmaire is a wood-boring pest of ash trees (Haack et al. 2002). EAB complete their life cycle on trees in the Fraxinus genus.

While the adult beetles cause relatively minimal damage while feeding on the foliage of ash trees, the larvae or immature stage feed on the inner bark of trees disrupting the ability to transport water and nutrients along the stem. Ultimately once infested, ash tree mortality occurs within two to four years (Herms et al. 2004). EAB was likely introduced as larvae in infested wood packing material (Stone et al. 2005) and was first identified near Detroit, Michigan, in the summer of 2002; although it was probably introduced 5-10 years before (Cappaert et al. 2005). After 8 years, the current

distribution of the emerald ash borer includes established populations in 15 states and

Canada.

While Fraxinus pennsylvanica is not rare, it is a host species for EAB which compromises its future and potentially the future of every species within the Fraxinus genus. Exotic forest pests magnify the negative impacts of an already stressed landscape from anthropogenic disturbance (i.e., habitat fragmentation). For example, the American chestnut (Castanea dentata) once made up approximately 25% of the eastern deciduous forest (Campbell and Schlarbaum 1994); however, since the introduction of the chestnut blight nearly one billion trees have died which also has led to profound ecosystem changes in the eastern deciduous forest (Primack 1993, McNeely 1999). The functional extinction of the American chestnut from native forests is the consequence of an exotic forest pathogen. Unfortunately, the exotic emerald ash borer now leaves the future of ash in question and in need of conservation efforts to ensure the preservation of ash trees.

At risk are the 16 native species of Fraxinus to the U.S. which include two species

(Fraxinus profunda and Fraxinus quadrangulata) that have protective status as either threatened or endangered in several states and 6 other species with geographic ranges limited to 1 - 4 states (USDA 2008). Since its discovery, EAB has killed tens of millions of ash trees within several states and throughout its range. In Ohio where this study was conducted, EAB was first identified in 2003. There are 5 native species of Fraxinus in

Ohio and it is estimated that there are approximately 250 million ash saplings and trees

(USDA Forest Service 1991). The economic impact of the complete loss of ash trees

throughout the urban landscape in the state of Ohio is estimated to be $7.5 billion

(Sydnor et al. 2007).

The objective of this study is to establish a sampling protocol for ex situ conservation of ash trees. Five microsatellite marker loci were used to determine the levels and distribution of genetic diversity of green ash trees (Fraxinus pennsylvanica) from nine Metroparks in the Toledo area. Using this information, I determined the genetic structure and variability within and among populations in order to define a germplasm sampling strategy that maximizes the genetic diversity preserved for the area sampled. As stated above, it is urgent to establish sampling priorities because of the continued spread of EAB, the near 100% ash mortality rate once infested and the potential for functional extinction of ash in the native forested landscape.

MATERIALS AND METHODS

Study Species

Fraxinus pennsylvanica Marsh. (green ash) was selected as the model species because of its distribution and dominance throughout the state of Ohio. These perennial, dioecious trees are wind pollinated and have wind-dispersed seeds. Reproductive maturity may take 10 to 25 years with abundant seed crops produced in 3 to 5 year cycles

(Bonner 1974). Fraxinus pennsylvanica has the broadest distribution of all Fraxinus species and is found across 42 of the lower 48 states (excluding most Western states) and

Canada (USDA, NRCS 2010).

Study Site

This project was conducted in Northwest Ohio in Lucas County. In the spring of

2005, eight populations of Fraxinus pennsylvanica were selected from eight Toledo Area

Metroparks. The distance between the populations (parks) ranges from 2.4 to 36.1 kilometers with an average distance of 16.8 km. These areas represent over 2,800 hectares of natural, historical and cultural parklands in Lucas County, including the Oak

Openings Region, the Great Black Swamp, the Maumee River, Ottawa River and Swan

Creek corridors (Fig. 4.1). These highly diverse habitats are representative of Ohio’s glaciation history.

In 2005, leaf tissue samples were collected from 15 trees within each of eight populations. There was a 15 - 20 meter minimum distance between selected trees. The distance between trees is typically used when sampling trees for germplasm collections

(National Academy of Science 1991, Guarino et al. 1995). Pole pruners were used to clip the tip of young leaf branches. The leaves were then placed into plastic freezer bags and kept at -20 oC until samples were transferred to the lab. Samples were then were kept at

-80 oC until DNA extraction was performed. Geographical coordinates for all trees were

taken using a Garmin eTrex Legend HCx GPS unit and a permanent metal tag with a distinct identification number was attached to the south-facing base of the tree.

Figure 4.1 Distribution of the Toledo Area Metroparks in Lucas County, Ohio.

The populations included in this study are identified by the name of the park where the trees were sampled. The parks that were included are Pearson, Wildwood, Secor, Swan

Creek, Oak Openings, Blue Creek, Fallen Timbers, and Side Cut.

DNA Isolation

Total genomic DNA was extracted from leaves of each ash tree using a modification of the CTAB protocol described by Cullings (1992) and Doyle and Doyle

(1987). The genetic structure was determined using 5 microsatellite loci: Femsatl1,

Femsatl4, Femsatl111, Femstal16, and Femsatl19. These markers were developed for a

European species of ash, Fraxinus excelsior and have been tested in fourteen other species within the genus including our study organism Fraxinus pennsylvanica (Lefort et al. 1999). PCR reactions were performed using a P-200 (MJ Research, Watertown, MA,

USA) in 20 l solution containing 50-80 ng of genomic DNA, 10mM Tris buffer, pH 8.0,

10 mM MgCl2, 0.2 mM dNTPs, 0.5 μM of each primer, and 1 unit of Taq polymerase

(Fermentas). PCR program used included: an initial denaturing step of 1 minute at 96°C,

followed by 35 cycles of 1 minute at 94°C, 30 seconds at either 45oC (Femsatl111 and

Femstal16), 50 oC (Femsatl4), 54 oC (Femsatl19) or 56 oC (Femsatl1) and 1 min at 72°C

and a final extension cycle of 2 min at 72°C. After the initial PCR, a second PCR

reaction was conducted using the leading universal M13 (-21) sequence labeled with

WellRED fluorescent dyes (D4-TGT AAA ACG ACG GCC AGT-3’) as primer

(Schuelke 2000). All PCR conditions were kept the same as described above, except for

a 48°C annealing temperature. Genotyping was conducted using capillary

electrophoresis on an automated genetic DNA analysis system (CEQ 8800, Beckman

Coulter, Fullerton, CA, USA). Four microliters of PCR product were mixed with 28 μl of formamide and 0.4 μl of 400 bp DNA size standard (Genolab) for capillary

electrophoresis. Fragments were identified on basis of their size and according to their mobility in relation to the size standard using a cubic function (Beckman Coulter 2002).

Genetic Diversity Analysis

Genetic variability within and between populations was analyzed using the program POPGENE 1.31 (Yeh et al. 1999) to determine observed and effective number of alleles and observed and expected heterozygosity for each locus and averaged over all loci, as well as, Nei’s genetic diversity (1978). In order to determine how genetic variation is distributed, the same software was further used to assess Wright’s inbreeding coefficient (FIS) and to determine genetic structure using hierarchical F-statistics.

Identified were the overall inbreeding coefficient (FIT) and the genetic variance among

populations (FST). Information about the genetic structure is important for the

development of conservation strategies, as it provides information necessary to capture

and maintain genetic diversity with a non-uniform configuration (National Research

Council 1990, 1991). Differences in allele frequency were assessed by performing a

Fisher exact test. The degree of relatedness between populations, based on Nei’s genetic

distances (1973), was used to construct a dendrogram with the Unweighted Pair Group

Method with Arithmetic Mean (UPGMA) method in the software TFPGA v. 1.3 (Miller

1997).

To identify if there was a relationship between the genetic relationship among

populations and their geographical proximity, a Mantel correlation test was performed

using a pairwise matrix of Fst/(1-Fst) and the logarithm of geographical distance

(km)(natural logarithm-scale). Further spatial structuring was determined using the software Structure (version 2.3.3; Hubisz et al. 2010) which is a Bayesian model-based algorithm that identifies groups of individuals based on their genotypes at multiple loci.

This method infers the presences of distinct populations by assigning individuals to a user-defined number of clusters or gene pools. The number of population clusters (K) is estimated by the probability that an individual belongs to a given cluster Pr(X/K)

(Pritchard et al. 2000).

Germplasm collection recommendations were established using a probability model developed by Yonezawa and Ichihashi (1989). This model calculates the probability of capturing all alleles in the population given their frequencies, the number of maternal trees to be sampled and the number of seed to be collected per mother tree. I estimated the sampling probability for a theoretical population where five loci have four alleles with two alleles found at high frequency (>0.45) and the other two alleles found at low frequency (0.005). These frequencies were established as a reasonable comparison to the measured allele frequencies identified in this study in which approximately 40% of the alleles frequencies are <1%. The probability that all alleles at each of the five loci (20 alleles in total) are captured in our theoretical sample is the product of the probability at each locus. Different numbers of mothers sampled and seeds collected per tree where considered to determine the best sampling strategy.

100

RESULTS

Allelic Diversity

There are significant levels of allelic diversity in Fraxinus pennsylvanica for all five microsatellite loci and across all populations included in this study. Unfortunately, one of the populations (Providence) was dropped from all future analyses due to insufficient sample size. It was determined that several of the trees from this population originated from planted nursery stock and several other samples repeatedly failed to amplify. There were a total of 82 alleles identified from 118 individuals. The observed number of alleles per locus ranged from 10 to 24; however, an average reduction of 54% occurred in the effective number of alleles per locus and ranged from (4.3 to 15), indicating that many alleles were found in low frequency (< 0.1) (Table 4.1).

Genetic Variation Within Populations

Within-population analyses revealed a high polymorphism rate with an average of

8 alleles per locus (Table 4.2). There was, however, on average 5 effective alleles per locus or a reduction of 36%. The observed heterozygosity was roughly 10% higher (HO

= 0.9075) than that expected (HE = 0.8054) based on a Hardy-Weinberg equilibrium.

Furthermore for all populations, Wright’s inbreeding coefficient (FIS) was negative with

an average value of -0.123 (Table 4.2).

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Table 4.1 Allele summary of microsatellite loci for green ash (Fraxinus pennsylvanica).

Note: N = sample size; na and ne are observed and effective number of alleles per locus,

respectively. The mean calculations include standard deviations which are shown in

parentheses.

Allele size Locus N n n range (bp) a e Femsatl 1 204 (188-206) 10 5.002 Femsatl 4 226 (172-198) 14 4.357 Femsatl 11 228 (192-244) 24 15.077 Femsatl 16 198 (188-226) 18 8.413 Femsatl 19 216 (180-232) 16 6.484 Mean 214 16.4 7.866 (stdev) (5.18) (4.322)

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Table 4.2 Summary of Wright’s F-Statistics for all microsatellite loci.

FIS, inbreeding coefficient; FIT , deviation from Hardy-Weinberg proportions in the total

population; FST, relative differentiation between populations. Loci are described in Table

4.1.

Locus FIS FIT FST Femsatl 1 -0.148 -0.079 0.06 Femsatl 4 -0.164 -0.093 0.061 Femsatl 11 -0.109 -0.027 0.074 Femsatl 16 -0.128 -0.009 0.105 Femsatl 19 -0.3165 -0.168 0.112 Mean -0.1711 -0.073 0.083

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Genetic Differentiation Among Populations

Genetic differentiation among populations was visualized with an Unweighted

Pair Group Method with Arithmetic Mean (UPGMA)-created dendrogram that identified

6 nodes based on Nei’s genetic distance (Fig. 4.2). Across all eight populations there is a moderate level of differentiation (FST= 0.083) and no detectable inbreeding (FIS= -0.171)

(Table 4.3). However, no significant relationship between genetic and geographical

distance was identified when comparing Fst/(1-Fst) ratio (Rousset 1997) for pairs of

populations and the natural logarithm of geographical distance (Mantel test P = 0.54, R2=

-0.056) (Fig 4.3).

The Structure analysis identified 4 distinct clusters (K). There was an increase in

log likelihood LnP(D) distribution until 4 clusters was reached (K=1: -2423.9, K=2: -

2342.2, K=3: -2307, K=4: -2281.5), then the likelihood distribution decreased again

between K=5: -2333.5, K=6: -2344, K=7: -2344.1 and K=8: -2388.9. Cluster 1 includes

most of the individuals from Blue Creek and Fallen Timbers, Cluster 2 includes Pearson

and Secor, Cluster 3 includes the Side Cut and Wildwood and Cluster 4 includes the Oak

Openings and Swan Creek (Fig 4.4). However, between several of these clusters there

appears to be some level of admixture occurring. Each population in Cluster 1 likely has

some level of gene flow with a neighboring population from a different Cluster. Fallen

Timbers (Cluster 1) has a fair amount of admixed individuals similar to the Side Cut

population of Cluster 3. Likewise, Blue Creek (Cluster 1) has some level of gene flow

with the Oak Openings population in Cluster 4.

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Blue Creek

Fallen Timbers

Wildwood

Pearson

Side Cut

Secor

Oak Openings

Swan Creek

Figure 4.2 Genetic relationship between green ash populations from the eight Toledo Metroparks.

Unweighted Pair Group Method with Arithmatic Mean (UPGMA) dendrogram is based on Nei’s (1973) standard genetic distances estimated from five polymorphic loci.

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Table 4.3 Population summary of genetic variation.

Mean number of alleles (Na), mean effective number of alleles per locus (Ne), mean observed heterozygosity (HO), and Levene's expected heterozygosity (HE) are given for

the five loci in the eight populations of Fraxinus pennsylvanica from Toledo Metroparks,

Toledo, Ohio. The standard deviations are shown in parentheses.

Population Na Ne HO HE FIS Blue Creek 7.4 5.351 0.927 0.819 -0.132 Fallen Timbers 9.2 6.330 0.960 0.840 -0.143 Oak Openings 7.2 4.279 0.848 0.758 -0.119 Pearson 7.2 4.751 0.873 0.793 -0.101 Side Cut 9.0 5.983 0.955 0.850 -0.124 Secor 7.6 4.111 0.889 0.784 -0.134 Swan Creek 6.6 4.110 0.904 0.757 -0.194 Wildwood 10.0 6.139 0.904 0.842 -0.074 Mean 8.03 5.132 0.908 0.805 -0.123 (stdev) (1.21) (0.94) (0.04) (0.38) (0.035)

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0.12

0.1

0.08

0.06 Fst/(1-Fst) 0.04

0.02

0 00.511.52 ln(distance)

Figure 4.3 Relationship of log geographical distance and genetic differentiation

Fst/(1-Fst) between populations as proposed by Rousset (1997).

(Mantel test, R2 = -0.056, P = 0.54).

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Cluster 1 Cluster 2 Cluster 3 Cluster 4

Figure 4.4 Population structure of green ash populations from the eight Toledo

Metroparks as determined using STRUCTURE.

STRUCTURE identified four distinct clusters: Cluster 1 (Blue Creek and Fallen

Timbers), Cluster 2 (Pearson and Secor), Cluster 3 (Side Cut and Wildwood) and Cluster

4 (Oak Openings and Swan Creek). Each pie chart illustrates the proportional assignment of individuals at each location to each of the four clusters.

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Sampling Design

The probabilistic model developed by Yonezawa and Ichihashi (1989) was used to determine an optimal germplasm sampling strategy for green ash. According to this model, recommendations for seed collection depended upon the number of mothers and the level of inbreeding in a population (Table 4.4). This study did not recognize significant levels of inbreeding; however, in order to avoid a potential sampling error, a minor level of inbreeding (F=0.05) was conservatively added to our estimates. With this level of inbreeding, a collection of 200 seeds from each of 5 mothers would have the highest likelihood of capturing all alleles (p >99%) present in a theoretical population where all loci have four alleles including two alleles at high frequency (>0.45) and two alleles at low frequency (0.005). Our estimates show that collecting more seeds or including more mothers does not significantly increase the sampling efficiency. In addition, Table 4.4 also includes the possible scenario for high levels of inbreeding

(F=0.80) which may be representative for other species in the Fraxinus genus that have highly fragmented populations, are locally rare or have limited geographical distribution.

In this scenario, the previously mentioned sampling design would only have a 79% probability of capturing all alleles present in the population. Therefore, the collection protocol for a population with high levels of inbreeding should include twice as many mothers but may include fewer seeds per mother (10 mothers and 50 seeds each) (p

>99%) (Table 4.4).

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Table 4.4 Germplasm collection recommendations based on level of inbreeding, number of mother trees sampled and number of seeds collected from each mother.

Note: Collection scheme established with 4 alleles, 2 at high frequency (0.48 and 0.51) and 2 at low frequency (0.005 and 0.005) at each of 5 independent loci. Bold columns indicate the recommended number of seeds collected per mother with the highest probability of capturing the genetic variability for an individual population depending on the assumption of low or high levels of inbreeding.

Low High inbreeding inbreeding (F=0.05) (F=0.80) Mother trees sampled 5 5 5 5 5 5 # Seeds collected 50 200 1000 50 200 1000 Prob. 0.818 0.993 0.994 0.739 0.792 0.890 Mother trees sampled 10 10 10 10 10 10 # Seeds collected 50 200 1000 50 200 1000 Prob. 0.998 0.999 0.999 0.994 0.996 0.998 Mother trees sampled 15 15 15 15 15 15 # Seeds collected 50 200 1000 50 200 1000 Prob. 0.999 0.999 0.999 0.999 0.999 0.999

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DISCUSSION

Ever since the introduction of the emerald ash borer, the future of ash trees in eastern deciduous forests has become compromised. F. pennsylvanica is a common and dominant tree across the deciduous forest in Ohio. However, long-term survival of a population is not guaranteed for even common species in a fragmented landscape (Van

Rossum et al. 2004). With tens of millions of ash trees killed already, the ecological as well as economical impacts are likely to escalate with the continued spread of this exotic beetle. Due to the rate at which an ash tree dies after infestation, there is now great concern about the loss of genetic variability, especially for the potential loss of any locally adaptive genotypes or unique rare alleles.

Even though the genetic structure of Fraxinus excelsior from across Europe

(Heuertz et al. 2001, 2004, Rudinger et al. 2008) and of Fraxinus mandshurica from across China (Hu et al. 2008, 2010) have been extensively studied, there is limited genetic information that exists about native American species of Fraxinus and most information is based on morphological and ecological attributes between species (Taylor

1971). This research is one of the earliest attempts to study of genetic structure for

Fraxinus pennsylvanica.

In general, I determined there is a high degree of allelic diversity which is similar to other deciduous temperate tree species that used microsatellites (Heuertz et al. 2001,

Streiff et al. 1998, Degen et al. 1999) and greater heterozygosity than expected. While there were common alleles, each population also had locally common alleles that contributed to the moderate level of genetic differentiation which occurred between

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populations. These findings support the need to develop effective sampling strategies that capture the diversity of these low frequency alleles.

My results also indicate that genetic diversity between populations cannot be explained solely on basis of physical distance between populations. While genetic structure has been shown to increase at larger spatial scales for a widespread tropical tree species (Rivera-Ocasio 2006), genetic isolation by distance, as proposed by Rousset

(1997) was not supported in this study. Similarly, Hu et al. (2008) did not find a significant correlation between population differentiation and geographical distance for

Manchurian ash (Fraxinus mandshurica) at population distances from 5 to 1,000 km. In contrast, isolation by distance has been reported in two studies of Common ash (Fraxinus excelsior) for population distances ranging from 0.7 to 400 km and from 5 to 2,500 km

(Heuertz et al. 2001, 2004). My study was conducted on a smaller scale with paired population distances ranging from 2.4 to 36 km suggesting that genetic distance for

Fraxinus pennsylvanica is not correlated with geographical distance at this scale. The fact that there was a lack of measurable isolation by distance but a moderate level of differentiation between populations (FST=0.083) suggests that genetic structure may be

influenced by restricted gene flow due to fragmentation by land-use practices or a relic of

glacial history.

The region of Ohio where the studied populations were found experienced the last

glacial event during the Pleistocene. This event shaped much of Northwest Ohio, and

Lucas County in particular, where Lake Erie once extended beyond its current shoreline.

During the various stages of glacial retreat and lake formation, there was likely a spatial

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and temporal difference in founder events of the surrounding plant communities. This geological history is reflected in the current vegetation community types across the region. The relic lake shore sand dunes are now characterized by oak savanna habitat and the heavy clay soils of the prehistoric lake bed are characterized by swamp forest, or the

Great Black Swamp, as it is known locally.

These different habitat types found across the study area appear to be related to the distribution of genetic diversity at the landscape level. The Structure analyses identified population clusters based on shared allelic composition; however, these clusters were determined to have some shared physical characteristic as well. For example, cluster 4 includes two populations that share the same typical sandy soils and oak savanna habitat. These populations (Oak Openings and Swan Creek) also form a distinct clade that sets them apart from the other populations using Nei’s genetic distance

(Fig. 8). Cluster 2 includes the Pearson population which is representative of the Great

Black Swamp with poorly drained Latty clay soils (Soil Survey Staff 2008). Unique soil types are also found in Cluster 3 where both populations have similar loam soils that occasionally flood.

Low levels of differentiation among populations within a floodplain have been documented for Fraxinus excelsior (Rudinger et al. 2008), as well as, for populations of

Fraxinus mandshurica in riparian habitats (Hu et al. 2010). Similar results were identified in this study for cluster 4 in which both populations (Oak Openings and Swan

Creek) are found along the Swan Creek floodplain. However, geographical distribution of the different clusters cannot be solely characterized using geological features. The

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populations (Blue Creek and Fallen Timbers) with individuals primarily assigned to cluster 1 are both found within the same watershed but have different soils and different vegetation composition; additionally both are relatively isolated by surrounding agriculture and/or development. This last cluster likely represents influences from previous land-use practices and disturbances events. It should be noted that Lucas

County, Ohio, is currently a highly fragmented landscape attributed to both localized urbanization surrounded by extensive agriculture.

Clear evidence for phylogeographic structure is not apparent in this study. While our clusters show some trends, these are not universal across all populations or between every cluster. These results may be attributed to our sampling design, as I did not seek to address phylogeographic relationships explicitly, or there may be a compounding influence from habitat fragmentation that has already altered any landscape attributes.

Fraxinus mandshurica is an endangered species across its range in China; and while there is genetic variation, there is also a lack of phylogeographic structure (Hu et al. 2010).

The lack of clear genetic structural relationship among populations makes future conservation efforts of Fraxinus species more challenging. Germplasm collection efforts are obviously necessary, but protocols for effective representation of the genetic diversity of a population are lacking.

This study marks a beginning step toward developing better sampling procedures to maximize the genetic variation collected during ex situ conservation strategies. The

CPC guidelines for the establishment of ex situ collections recommend collecting from 50 individuals from each population and up to 50 populations (Guerrant et al. 2004).

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Previous work has suggested greater sample sizes of 50 seeds from 100 individuals

(Hawkes 1980). Our recommendations for germplasm collection are determined by the number of mothers rather than the number of seeds per mother. These results are similar to the original computations performed by Yonezawa and Ichihashi (1989) for plant species with low levels of inbreeding. It is important to note that these smaller sample sizes are achieved because Fraxinus pennsylvanica is a dioecious, wind-pollinated plant with minimal levels of inbreeding. Therefore, a germplasm collection representing >99% likelihood of capturing all alleles in a population is achieved by collecting 200 seeds from only 5 individuals. An alternative sampling design was also established for other species of Fraxinus where there may be an increased likelihood of inbreeding due to localized rarity, to limited geographical distribution or for low density, high fragmented populations. Consequently, germplasm collections for species of Fraxinus with assumed higher levels of inbreeding (F=0.80) should, therefore, double the number of mothers with fewer seeds collected per mother such that 50 seeds from each of 10 mothers would accurately capture 99% of alleles in the population. It should be noted that increasing the number of mothers or seeds beyond these recommendations does not provide significantly greater representation of the genetic variability.

The number of trees recommended to be sampled from each location is lower than that proposed by Brown and Hardner (2000). They consider a sample of 15 trees should be enough to capture most of the variation that is relevant for the survival of the target species. Marshal and Brown (1975) proposed to group alleles on the basis of their frequency and their geographical distribution into four classes of alleles: (1) common and

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widespread, (2) common and local, (3) rare and widespread, and (4) rare and local. They also argued that common and local alleles deserve special attention for the design of sampling strategies, as they may represent alleles that may confer specific adaptations to local conditions. I argue that low frequency alleles (rare and widespread) may also be important when searching for novel genes that may translate to EAB resistance or tolerance. With respect to EAB, our native ash trees are naive to this pest and, therefore, have had no evolutionary history to warrant the positive effects of natural selection for an

“anti-herbivory” genetic attribute. The presence of EAB represents a selection pressure previous not experienced; therefore, any resistance or tolerance our native ash trees may have had would never have been selected; and if such a tolerance attribute exists, it would be present at very low frequency.

Recommendations

This study provides 3 levels of recommendation for the development of ex situ conservation efforts for the genus Fraxinus.

1.) Fraxinus pennsylvanica is a common species and has the most broadly distributed range of all ash species in the United States. Common species are an integral component to ecosystem functioning and sustainable biodiversity, but even common species can still be negatively affected by habitat fragmentation (Van Rossum et al. 2004). Fraxinus pennsylvanica has experienced an intensive, localized attack from EAB resulting in a rapidly fragmented distribution that may lead to permanent local population collapse and

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the loss of any regional genetic adaptation. Conservation efforts for common species need only include a few hundred seeds from just 5 mothers to capture most alleles present. Currently common, green ash trees are being killed by EAB within 2-4 years

(Herms et al. 2004) and EAB can attack saplings that are only 1 inch in diameter.

Unfortunately, ash trees can take 10 to 24 years before they reach sexual maturity

(Bonner 1974) which means that even a common species fails to become large enough to persist in a landscape with EAB present.

2.) In contrast, there are several other species of Fraxinus that have federal protective status throughout their range or have limited geographic ranges. These populations should receive special prioritization for ex situ collection efforts due to their restricted distribution. A more conservative collection protocol should be implemented in this situation by assuming that there is higher potential for inbreeding; and thus, collection efforts should focus on sampling from greater number of maternal plants.

3.) It should be pointed out that this sampling design considers ovule (maternal) frequencies are similar to pollen (paternal) frequencies; that is, sampling collections are conducted in natural populations with many trees. However, it is not uncommon to find single isolated trees from which one can collect seeds. In such cases, pollen contribution may come from just a few surrounding trees, and our ability to capture low frequency alleles would be different from that determined in this study. It is difficult to predict the consequences of collecting seeds from such trees, but it is likely that both ovule and pollen frequencies will be biased toward high frequency alleles in small clusters of trees.

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Unfortunately, just 4 years after leaf tissue was first collected, every ash tree included in this study, as well as, all ash trees within each of the populations, were showing signs of infestation and at various stages of decline or mortality (Hausman, personal observation). In addition, ash seedlings were abundantly observed in the understory of recently infested forested areas but within 4 years there is an obvious lack of new ash seedling recruitment and young samplings are subjected to multiple browse events by deer (Hausman, personal observation). These field observations, while anecdotal, further identify the potential for permanent loss of localized genotypes. The information obtained from this research identifies the extensive diversity that occurs across a relatively limited geographic distance and the temporal urgency for prioritizing germplasm collection efforts.

CHAPTER 5

DISCUSSION

EMERALD ASH BORER IMPACTS

The spread and establishment of invasive species pose a major threat to native ecosystems as they displace native species. In the Introductory chapter I discussed the mechanisms of non-native species introductions, the pathways of establishment or failure, and the spread of an invasive species at the expense of the native habitat. The range expansion by invasive species may be attributed to their release from control by natural enemies (enemy-release hypothesis). There may also be positive feedback where one exotic species facilitates the spread and establishment of another (invasional meltdown hypothesis). The invasional meltdown can be a synergistic relationship between two exotic species as has been shown for an exotic deer and a non-native tree species (Relva et al. 2010). Additionally, there have been relationships shown between invasive species establishment and land use change and climate change (Vitousek et al. 1996, Mack et al.

2000).

When introduced to a new range, an exotic forest pest can have a greater negative impact on the community compared to its native range due to a lack of competition, lack of natural predators, and a lack of plant defense mechanisms (Mack et al. 2000). 118 119

Similarly, the emerald ash borer does not appear to be a major pest species in its home range, but does demonstrate a significant negative impact in its introduced range. As such EAB’s rapid range expansion may support an enemy-release hypothesis. The North

American ash species are naive to EAB and appear to be more susceptible than the ash species that have coevolved with EAB in their native range (Rebek et al. 2008). A lack of coevolved predators and parasitoids and an abundant food source has created a seemingly limitless range for EAB expansion. While EAB may have been able to spread due to an enemy-release, its establishment has created an environment that has facilitated the establishment of invasive plant species thus generating an invasional meltdown.

EMERALD ASH BORER ERADICATION

Results of the dissertation work presented in Chapter 2 provide the first evidence in support of an invasional meltdown hypothesis. However, it should be noted that this research includes humans as a potential contributing factor inducing the invasional meltdown. EAB causes a disturbance to the native forest by creating large open canopies through ash tree dieback; yet, the EAB eradication efforts magnified the scale of disturbance through time and space which likely facilitated an immediate and significant establishment of invasive plant species.

Successful invasive plant species take advantage of the open habitat, establish new populations and outcompete the native plant species for resources (light, space, nutrients). There are several proposed reasons for their competitive advantage including:

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physical attributes, phenotypic plasticity, physiological mechanisms, and life history traits (Richards et al. 2006). Some invasive plant species are known to leaf-out earlier in the spring than native species (Trisel 1997, Silander and Klepeis 1999, McEwan et al.

2009). In addition, it has also been shown that some invasive plants have higher photosynthetic capacity and are able to maintain net photosynthesis for a longer period of time compared to their native congeners (McDowell 2002). Even when there are instances in which an invasive species and non-invasive congener have similar photosynthetic rates, the invasive species may still succeed because it has thinner leaves and thus a lower carbon cost per unit photosynthetic area (Pammenter et al. 1986). There are still other species that have life history strategies and structural morphologies that allow them to form dense monocultures (Phragmites and Purple Loosestrife) thus outcompeting native species for space.

Regardless of the specific attribute an individual invasive species may possess, the management concern is the fact that multiple invasive species with several “bad- habits” have established, have outcompeted the native community and have been continuing to reproduce and expand their range. What we know is that these invasive plant species will colonize when given the opportunity. What we need to identify, now in the wake of the EAB eradication disturbance, is what can be done to eliminate or reduce the scale of disturbance in order to minimize the establishment of invasive plants.

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SCALE OF DISTURBANCE

While Chapter 2 identifies the establishment of invasive plant species in the wake of EAB eradication, it inherently also includes the compounding influence of human disturbance. The research conducted in Chapter 3 intends to eliminate the obvious disturbance effect caused by the tracked vehicles during the eradication but still includes the instantaneous formation of a canopy gap. This design has been used to identify how the plant community responds to various degrees of disturbance in order to determine an appropriate management protocol for EAB-infested forests. Based on the intermediate disturbance hypothesis, the highest species richness should occur at moderate levels of disturbance (Connell 1978). However in a modern landscape, the function of disturbance is altered and the outcome provides opportunities for invasive species to thrive (Hobbs and Huenneke 1992).

There appears to be a secondary spread of invasive plants in EAB-infested forests of Michigan (Herms et al. 2008). These forests have had EAB infestations for at least 6 years with continued spread and dieback. In Ohio, the results of my disturbance research indicate that severely disturbed areas can experience dramatic shifts in plant community composition with significant establishment of “new” invasive plant species within a single year. The intermediate disturbance sites also had invasive plant species cover, but lacked a consistent pattern of invasive cover across the treatment as a whole. As such, the invasive plants in the intermediate disturbance were likely already established in the area and are now just exploiting the newly available light gap. Regardless of the degree of community change in the disturbed forests, over the 5 years of this study there has

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been no recognizable change in the community structure of the Uncut control treatment.

This treatment represents a “natural” scenario of EAB spread and continued infestation with slower change and dieback to the above ash canopy. So while there is concern now for the spread of invasive plants into EAB-infested forests, I would argue that there needs to be a distinction between new invasive species establishment and opportunistic spread by an already existing invasive cover. From a management perspective the knowledge of existing invasive cover relies heavily on forest managers recognizing the size and extent of problematic species. With this knowledge, prioritization of management efforts can focus on areas where ash stands are in close proximity to problematic areas (edges) or near established invasive populations. Proactive removal of an invasive species before

EAB infests a stand may be necessary.

The timing of ash tree removal is an additional consideration that has the potential to influence how the forest community responds to the disturbance. For example, Wolf et al. (2008) found that winter logged roads had greater number and percent cover of plant species with high coefficients of conservatism than summer logged roads. Winter logging may also have fewer negative environmental impacts. Regardless of the timing of logging (winter vs. summer), there was greater species richness near logging roads due to the number of alien species and to early successional plants that colonize these disturbed areas.

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CONSERVATION EFFORTS FOR ASH

Ash is a dominant tree in the native deciduous forests of the United States and occurs in 42 of the lower 48 states. There are 16 ash tree species native to the United

States with several having federal protective status and others with limited geographical distribution. It appears that all species of ash are susceptible and the effectiveness of existing EAB control measures remains uncertain. Since it was first identified in 2002,

EAB has spread to 15 states and 2 Canadian provinces. Perceptions must now incorporate EAB as a fixture in the landscape that will persist for many years to come.

Therefore, future research efforts should consider what the forest will look like after EAB and focus on regeneration of ash trees and reforestation efforts.

Chapter 4 of this dissertation contributes to these future conservation efforts by providing a guideline for ex situ collection methods. Ash seed collection efforts are performed by volunteers and by a few government agencies with limited funding and staff. Therefore, by understanding the diversity and spatial genetic structure of green ash

(Fraxinus pennsylvanica), I have been able to construct a seed collection protocol that will optimize the likelihood of capturing all alleles in a population while minimizing collection efforts.

There are currently 2 national efforts established for collecting ash seeds for long- term banking. The objectives of both are to collect ash seeds for long-term preservation; however, the methodology for collecting seeds varies decidedly between these efforts.

The U.S. Forest Service heads up one of the ash seed collecting programs. Their design uses Omernik Ecoregions Level III as the basic sampling unit and targets 15 individuals

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(25 to 50 preferable) from at least 10 locations, evenly spaced at least one mile apart

(Karrfalt 2010). There are 5 Level III Ecoregions in the state of Ohio. This U.S. Forest

Service National Seed Laboratory protocol keeps the collected ash seeds separated by maternal tree. The USDA Agricultural Research Service heads up the other main ash seed collection program. Their collection efforts are population-based in which seeds are collected from 8 to 12 mothers in a population. From this collection, 1,500 seed are bulked together from the combined mothers; any viable seed remaining is then kept separated by maternal plant (Widrlechner 2010).

Based on the results of Chapter 4, I have developed a sampling protocol that provides a more accurate representation of the genetic diversity across the study area.

Our methodology ensures not only that the common alleles are collected but also that the rare or unique alleles are represented as well. It is these rare alleles that occur in low frequencies that may have an adaptive or resistant trait that could be beneficial for coexistence with the emerald ash borer. It is possible that a relic allele occurs in the

Fraxinus genus and remains in all of the species such that even though our native ash trees lack recent coevolutionary history with EAB, an allele that codes for resistance or tolerance to the beetle may still remain in the genome. In this case, there would not be a selective force favoring this trait, since EAB, or other pest insect, with similar biology did not coevolve in North America; and as such the frequency would have become very low, but not gone. The ash seed collection protocol described in Chapter 4 was designed with the intention of increasing the likelihood of including low frequency alleles and thus preserving any potential resistance trait.

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It is likely that the U.S. Forest Service sampling unit is too large of an area with seeds collected from mothers that are too far apart in order to capture an adequate representation of the genetic diversity in a population. Their germplasm collection contains large quantities of seed from fewer individuals which likely represent only the most common alleles for any given region. Furthermore, a lack of phylogeographical structure has been documented in Fraxinus mandshurica in China (Hu et al. 2008); therefore, collection effort designs based on Omernick Ecoregions may not accurately represent regional genetic diversity. The Agricultural Research Service (ARS), on the other hand, does collect seeds from mothers within a localized population. Their targeted

1,500 seeds from 8-12 mothers do not adequately fulfill the recommendations that I have developed (200 seeds from 5 mothers). Unfortunately, the ARS protocol also specifies that collected seed be mixed in bulk which causes a great loss of maternal information.

BROADER CONSERVATION IMPLICATIONS

Management Concerns for Hazardous Trees

The emerald ash borer continues to spread across the landscape causing widespread ash mortality. Once dead, ash trees become brittle very quickly and do not stay standing for long. These hazardous trees will be a continual and mounting problem for cities, municipalities and natural areas as the removal of dead and dying ash trees will be dangerous and costly. As trees begin to fall or break, they affect the structure of the surrounding canopy through contact with falling stems which causes further canopy loss

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to neighboring downwind trees (Quine and Gardiner 2007). This collateral damage magnifies the gap opening in an otherwise closed forest making large open areas with exposed trees along the edges. Trees that occur on these “new” edges are vulnerable to further disturbance events such as windthrow. The gap opening in a canopy allows strong winds to penetrate which can generate instability and damage among the remaining trees (Gardiner 1995). This continued gap expansion has been shown to be an important process in several forested systems (Taylor 1990, Liu and Hytteborn 1991,

Rebertus and Veblen 1993). For example, Rebertus and Veblen (1993) found that 53% of gaps that occur in temperate deciduous forests were the result of multiple disturbance events. Thus the stability of that community and susceptibility to future disturbance will remain a concern for years to come.

While not explicitly measured in this research, natural disturbance such wind will likely play an important part in the future of forest structure and composition. Wind disturbance, often referred to as windthrow, occurs due the breakage or uprooting of trees by the wind (Mergen 1954). EAB can cause ash mortality within 2 to 4 years (Herms et al. 2004); however ash trees do not remain dead standing for long. Unlike other species of trees that can persist as a dead vertical snag, ash trees appear to be very brittle once dead. Therefore, an infested stand can experience an increased likelihood of branch failure, stem breakage or windthrow resulting in large amounts of downed woody debris on the forest floor. Dead ash trees are weak and unstable which make them a hazardous component of a forest stand. As such, wind disturbance has the potential to intensify

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collateral damage as falling trees compromise the structure of the surrounding canopy

(Quine and Gardiner 2007).

The long term impacts of dying ash trees will provide challenges to municipalities, parks and preserves in urban landscapes where trail maintenance, recreational activities and public safety may be compromise by hazardous trees. For larger state and national forests the long term impact of large quantities of dead, dying and windthrown ash trees will cause a build-up of downed woody debris. The ecological impact of this fallen debris does have the potential to suppress the understory plant community (Everham and Brokaw1996). Furthermore, windthrown trees add to the fuel load on the forest floor increasing the potential for fire (Myers and van Lear 1998). The excessive build up of fuel load increases the intensity with which a fire burns (Catchpole et al. 1982) and higher intensity fires can increase the amount of canopy tree mortality

(Gutsell and Johnson 2007).

Forest management practices should include a survey component to assess the structural integrity of ash trees within stands. Monitoring wind disturbance may be necessary to determine long term stand health, sapling regeneration, and fuel-load build up in order to identify the need for future management.

Management Concerns of Deer Herbivory

Canopy gaps occur in the forest and increase the light levels which create an opportunity for new recruitment and stimulate germination from the seed bank. These open gap areas also attract herbivores which may reduce gap-filling tree regeneration

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through browse pressure (Messier et al. 1999). Deer are large herbivores that are capable of dramatically affecting forest communities (Hobbs 1996). The continued browse pressure on forest tree seedlings and saplings can dramatically reduce tree regeneration

(Russell and Fowler 2004). There was a higher establishment of invasive plant species in the cut gaps created during the EAB eradication program. Therefore, in addition to plant establishment and competition for available resources, there is an additional selective pressure due to deer herbivory. Selective browsing on native plant species reduces population numbers with continued browse events potentially reducing the reproductive success of the plants. This effect may have magnified the invasive plant species establishment seen in the Cut treatment. Over-browsing pressure has been documented in other forested systems where an invasive forest pest occurs (Frelich and Lorimer

1985). The long-term effects of this kind of selective browsing can alter the entire forest community and change canopy dominance to browse-tolerant or less preferred species

(Gill and Beardall 2001).

Over the course of this research I observed hundreds of germinating ash seedlings immediately following the eradication cutting. However, after 5 years of continued survey there are virtually no new ash germination and the seedlings that emerged in year one were either outcompeted or have become subjected to multiple browse events for several years. The ash seedlings that still occur have a shrub-like stature from continual deer browse. Similar selective deer herbivory has occurred in stands where the hemlock wooly adelgid has infested hemlock trees (Frelich and Lorimer 1985). This negative impact has also been documented when deer are introduced to islands in which selective

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herbivory negatively affects the tree growth of a native species over an exotic tree species

(Relva et al. 2010).

This research further provides evidence for the invasional meltdown hypothesis in which EAB causes ash canopy dieback, opens the light environment, and stimulates seed recruitment by both native and non-native plants. Here the deer, while native but at population densities large enough to be considered invasive, are attracted to the new open areas and selectively feed on the native plants thus facilitating greater invasive plant establishment.

There were three major objectives of my dissertation research (1) to identify consequences of EAB eradication efforts, (2) to design effective ex situ conservation protocols for future ash tree preservation, and (3) to determine altered community composition under different disturbance intensities (recommendations for tree removal management). Each one of these objectives were thoroughly investigated and successfully completed. The end result provides effective forest management recommendations and prioritizes ash conservation.

APPENDIX

This appendix contains an aerial map of Pearson Metropark in Lucas, County

Ohio. Each plot measures 20 x 25 meters with the short side oriented N-S and the long side oriented E-W. Treatments are defined as: Cut — the Ohio Department of

Agriculture implemented EAB eradication protocol in which an EAB infested ash tree and all ash trees within a 0.8 km radius of the infested tree were cut down using heavy- tracked trucks and skidders; Cut without Compaction (Cut w/out) — ash tree were felled by hand using chainsaws and the downed trees were left on site; Uncut — no ash tree removal occurred and EAB progressively infested ash trees with natural canopy dieback.

Included are the GPS coordinates which were taken at the northeast corner of the plot.

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360

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130 131

Plot # Treatment Latitude Longitude

136 Cut 41.6423532 83.4476299 3 West Cut 41.6398852 83.4473094 8 Uncut 41.6406718 83.4452334 360 Cut Without 41.6415994 83.4445644 Compaction 3 East Uncut 41.6416485 83.4428851 350 Cut Without 41.6405354 83.4423884 Compaction 16 Cut 41.643965 83.4358784 34 Cut 41.64151 83.4375714 2 West Cut 41.6406408 83.4376369 4 Cut Without 41.6405988 83.4355234 Compaction 28 Cut 41.6384587 83.4355176 7 Uncut 41.6432591 83.4338179 5 Uncut 41.6420972 83.4327112 1 Cut Without 41.6406215 83.433156 Compaction 2 East Uncut 41.6395634 83.4329531 270 Cut Without 41.6388326 83.4321746 Compaction 6 Uncut 41.64012 83.4300391 260 Cut Without 41.6423965 83.4307884 Compaction

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