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AFLP, mtDNA, AND MICROSATELLITE ANALYSIS OF
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ASIA AND NORTH AMERICA

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ALICIA MARIE BRAY

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5/08 KIProj/AodPres/ClRCIDateDuohdd

AFLP, mtDNA, AND MICROSATELLITE ANALYSIS OF EMERALD ASH BORER
POPULATION STRUCTURE FROM ASIA AND NORTH AMERICA

By

Alicia Marie Bray

A DISSERTATION
Submitted to
Michigan State University

in partial fulfillment of the requirements
for the degree of

DOCTOR OF PHILOSOPHY

Entomology and
Ecology, Evolutionary Biology and Behavior

2009

ABSTRACT

AFLP, mtDNA, AND MICROSATELLIT ANALYSIS OF EMERALD ASH BORER
POPULATION STRUCTURE FROM ASIA AND NORTH AMERICA

By
Alicia Marie Bray

Emerald ash borer (EAB), Agn'lus planipennis Fairmaire (Coleoptera:
Buprestidae), is a devastating invasive pest of North American ash trees
(Fraxinus spp.) that was first discovered outside of its native range of
Northeastern Asia in 2002 (Haack et al. 2002). With the unintended assistance
from human movement of infested ash material, EAB spread swiftly from its initial
discovery in the Detroit area of Michigan and Windsor, Ontario to currently
include 13 states in the United States and two provinces in Canada.
Understanding the population biology of an invasive species such as EAB could
provide valuable information on geographic origin and location of possible
effective biological control agents, estimate host range potential, and provide
evidence of the main mode of spread. This study had three main goals: 1) obtain
samples from throughout the native and introduced ranges of EAB, 2)
characterize the genetic population structure of EAB in its native and introduced
range using mitochondrial partial gene sequencing and DNA fingerprinting using
amplified fragment length polymorphisms, and 3) develop and characterize
microsatellite loci to further assess population dynamics and invasion history.

To accomplish the first goal, a network of collaborators, including myself,
was developed to obtain samples throughout the native of China, South Korea,

and Japan and introduced ranges in the United States, Canada, and western

Russia. This effort yielded a collection consisting of 1799 specimens from across
7 states in the United States, 114 in Ontario, Canada, 12 in Moscow, Russia, 274
specimens in China, 17 in South Korea, and 3 in Japan.

To accomplish the second goal, individual insects were characterized with
partial mtDNA cytochrome oxidase subunit I sequence (481 bp) and four AF LP
primer pair combinations yielding 273 loci. COI sequences detected one
common haplotype found in China, South Korea and all samples in N. America,
as well three unique haplotypes in China, and four haplotypes from South Korea
that all differed from the common sequence by 2-4 nucleotides. In addition, a
single EAB from Japan differed from the common sequence by 22 nucleotide
changes (3.7%). The majority of the AFLP genetic variability was within
populations and not among populations. Very weak genetic structure was
detected. Average pairwise (Dpt across all populations in N. America revealed
the lowest population differentiation between Dagong and Tangshan, China ((Dpt
= 0.0877 and 0.0848, respectively). Over 67%individual beetles from N. America
Were assigned either to Dagong or Tangshan, China by assignment tests.

The final goal to develop and characterize microsatellite loci evaluated 41
primer pair combinations that successfully amplified in EAB, however, none of
these loci yielded variation to assess within and between populations in Asia.
Microsatellite loci variation of EAB is presented on a subset of samples in this
study on two loci developed by a Dr. Jenny Cory (Simon Fraser University, BC,
Canada). These data did not provide enough information to distinguish between

single or multiple EAB introductions into N. America.

Dedicated with love to my husband and parents for their unwavering support and
guidance through all the obstacles that life gives, and to my daughter for her
amazing smile that brightens each and every day.

iv

ACKNOWLEDGEMENTS

There are numerous factors that all fit together that helped me accomplish
this degree. First, I am deeply grateful to my major professors, Dr. James J.
Smith and Dr. Anthony Cognato, whose knowledge, guidance, and patience are
the reason I have reached this accomplishment. I would also like to extend a
special thank you to Dr. James R. Miller, who not only helped guide me to
complete a Master’s degree, was willing to continue in my development being on
my doctoral committee, but was always there for personal listening ear to help
me put things in perspective. I am also grateful to Dr. Mark Scriber, and Dr. Leah
Bauer for their support and insight as members of my committee.

I would like to thank the Department of Entomology, MSU Lyman Briggs
College, and the Center for Integrated Studies in Biology for the opportunity to
teach throughout my tenure at MSU. l have had the opportunity to teach for and
learn from some of the best instructors at MSU, including Dr. Kendra Cheruvelil,
Dr. Cori Fata-Hartley, Dr. Larry Besaw, Dr. Todd Tarrant, Dr. Charles Elzinga, Dr.
Jerry Urquhart, Dr. James J. Smith and Dr. Doug Luckie. I would also like to
thank Dr. Gabe Ording for his guidance in my development as a teacher; he was
always available to discuss ideas and provide advice.

Barb Stinnett deserves a very special thank you for helping me find out
how exciting and rewarding Outreach and Extension is in my career.

Thanks to all the support staff in the Department of Entomology for the

countless time throughout my time at Michigan State. There is truly no better

team at a university; you all contribute to why this department is so welcoming
and efficient.

I would like to thank all those who helped me during my project by directly
or indirectly collecting insect samples in Asia and North America, without which I
would not have had samples to conduct this research. In China - Leah Bauer
PhD, Houping Liu PhD, Therese Poland PhD, Roger Fuester (USDA-ARS), Deyu
Zou (Chinese Academy of Agricultural Sciences), and Tonghai Zhao, Zhong-qi
Yang (Research Institute of Forest Ecology, Environment and Protection,
Chinese Academy of Forestry). In Japan - Paul Schaefer PhD (USDA-APHIS),
Dr. Naoto Kamata (University of Tokyo, Japan), Minemitsu Kaneko (Japan
Wildlife Research Center), and Takaharu Hattori (private collector). In Russia -
Bob Haack PhD with Oleg Kulinich PhD. In South Korea - Ho Yul Choo PhD,
Gun Young Han, Seok Jun Lee, Engene Lim, Sung Wook Kim, Sung Gi Moon
(all from Gyeongsang National University), Dong Woon Lee PhD(Sangju National
University), and David Williams (USDA-APHIS PPQ). In North America - Leah
Bauer PhD, Houping Liu PhD, and Debbie Miller (USDA-F8), Deb McCullough
PhD, David Smitley PhD, Nate Siegert PhD, and Andrea Anulewicz of the
Department of Entomology at MSU and several employees of the Michigan
Department of Agriculture collected individual beetles from populations
throughout Michigan. Lucy Hunt (Ohio Department of Agriculture), Mark
Cinnamon (Illinois Department of Agriculture), Benjamin West (Indiana
Department of Natural Resources), and Dick Bean (Maryland Department of

Agriculture) assisted in beetle collection in Ohio, Illinois, Indiana, and Maryland,

vi

respectively. Ches Caister, Plant Protection officer Southwest region, collected
specimens in Windsor, Ontario. R.M. Turcotte (USDA-FS) collected specimens
in West. Virginia. Nate Siegert PhD and Sven-Erik Spichiger (Pennsylvania
Department of Agriculture) collected specimens in Pennsylvania. I am indebted
to these and many others for helping me obtain samples.

I am very grateful for the support of my good friends who always seemed
to have time to listen. Although this list is certainly not an exhaustive, I would like
to single our Dr. Susannah Cooper, Rachel Olsen, Megan Fritz, Dr. Kendra
Cheruvelil, Dr. Cori Fata-Hartley, Dr. Mollie McIntosh, Anna Fiedler, Erik Foster,
Ryan Kimbirauskas, Kristi Zurawski, Bob McCIowry, and Kristin Bott who were all
integral in my accomplishment and keeping my sanity.

Finally, I would like to thank my family for their unwavering support and
belief in my ability. My parents, Dennis and Diana King, and Darcy Todd, my
mother-in-Iaw Carol Bray, and my brother Jeff King. The support and
understanding provided by my husband, Kevin, is beyond words, he is truly a
blessing in my life. I also want to thank my daughter Isabel for her bright smile
helping me press-on.

Support for this project was primarily funded by the USDA Forest Service,
North Central Research Station and Project GREEEN (Generating Research and
Extension to meet Economic and Environmental Needs). Funding was also
provided by a predissertation fellowship from the MSU international center, a
student grant provided by the Hutson Endowment in the MSU Department of

Entomology, and several fellowships from the MSU Graduate School.

vii

TABLE OF CONTENTS

LIST OF TABLES ................................................................ x
LIST OF FIGURES .............................................................. xii
CHAPTER 1
INTRODUCTION .................................................................. 1
Invasive species ......................................................... 1
Ecological impact of invasive species ..................... 2
Control strategies of invasive species ..................... 4
Detecting evolutionary change associated with invasions ..... 6
Variation within a population .................................. 6
Molecular markers to assess genetic variation .......... 7
Genetic structure of populations in the native ranges. 10
Geographic structure of introduced species ............. 12
Genetic makeup of introduced populations ............... 14
Population change after introduction ....................... 14
Geographic origin of invasive species ..................... 15
Number of introductions of invasive species ............. 18
Hybridization of introduced populations with native
species ............................................................. 1 9
Emerald Ash Borer ....................................................... 20
Control Methods of EAB in North America ......................... 22
Detection of EAB ................................................ 23
Chemical control of EAB ....................................... 23
Biological control ................................................. 24
Objectives .................................................................. 26
CHAPTER 2
GEOGRAPHIC DISTRIBUTION OF EMERALD ASH BORER ........ 27
Introduction .................................................................. 27
Host range ................................................................... 31
EAB life history ............................................................ 33
Methods for EAB collection in introduced and native range... 34
EAB collection with collaborators ............................ 34
Collection of North American EAB samples ............... 40
Collection of Asian EAB samples ............................ 41
Results of emerald ash borer collection ............................. 49
Discussion .................................................................. 56
CHAPTER 3
AFLP AND mtDNA ANALYSIS OF EAB POPULATIONS IN ASIA AND
NORTH AMERICA ................................................................ 57

viii

Introduction ................................................................ 57

Materials and methods .................................................. 63
Sample collection ................................................ 63
DNA isolation ...................................................... 68
Mitochondrial DNA gene sequence ......................... 68
Amplified fragment length polymorphism (AFLP) ........ 70
Results ....................................................................... 74
mtDNA variability ................................................. 74
AFLP variability ................................................... 80
Discussion .................................................................. 93
CHAPTER 4
MICROSATELLITE ANALYSIS OF EMERALD ASH BORER ......... 100
Introduction ................................................................ 1 00
Materials and methods ................................................. 101
Sample collection and DNA isolation ...................... 101
Microsatellite marker development ......................... 103
Data analysis ..................................................... 104
Results and discussion ................................................. 105
APPENDICES ..................................................................... 113
Appendix A: COI sequence haplotypes ..................................... 114
Appendix B: Deposition of voucher specimens ............................ 126
LITERATURE CITED .................................... , ........................ 129

ix

LIST OF TABLES

Table 2.1 .a. Field location survey records for trees examined for the presence of
Emerald Ash Borer in South Korea .......................................................... 42

Table 2.1.b. Field location survey records for trees examined for the presence of
Emerald Ash Borer in China .................................................................. 44

Table 2.1.c. Field location survey records for trees examined for the presence of
Emerald Ash Borer in Japan ................................................................... 47

Table 2.2. Emerald ash borer specimens obtained/collected in North America
from 2003-2008 ................................................................................... 50

Table 2.3. Emerald ash borer specimens obtained/collected in Asia from 2003-
2008 .................................................................................................. 54

Table 3.1. Emerald ash borer specimens collected/obtained from 2003-2008...64

Table 3.2. Mitochondrial DNA variability of emerald ash borer in Asia and North
America .............................................................................................. 75

Table 3.3. AFLP and mtDNA COI results from AMOVA analysis of Asian and
North American collection locations of emerald ash borer ............................. 79

Table 3.4. Pairwise FST between populations of mtDNA COI haplotypes from
emerald ash borer based on Kimura 2—parameter distance ........................... 81

Table 3.5. Descriptive statistics for Asian and North American EAB collection
locations with (N) number of individuals using 273 total AFLP Ioci .................. 82

Table 3.6. Pairwise (Dpt between Asian populations with AF LP data set of 273 loci
from emerald ash borer .......................................................................... 84

Table 3. 7. Pairwise (1th between North American populations with AFLP data set
of 273 loci from emerald ash borer ........................................................... 85

Table 3.8. Pairwise (Dpt between Asian and North American populations with
AFLP data set of 273 loci from emerald ash borer ....................................... 86

Table 3.9. Structure analysis results ......................................................... 92

 

Table 3.10. Assignment test of individuals of the introduced range of emerald ash
borer in North America against collection locations from the native range in
Asia .................................................................................................. 94

Table 4.1. Locality information for EAB individuals included in the mocrosatellite
analysis ........................................................................................... 102

Table 4.2. Characteristics of 2 microsatellite loci isolated from emerald ash borer
of 162 individuals from Asian and North American populations ..................... 106

Table 4.3. Genetic diversity of two microsatellite loci in emerald ash borer
populations from Asia and North America ............................................ _...107

Table 4.4. Genetic differentiation (pairwise Fst) of emerald ash borer populations
in China and North America .................................................................. 108

xi

LIST OF FIGURES

Figure 2.1. Emerald ash borer introduced distribution range in North America
reported as of July 2009 ......................................................................... 29

Figure 2.2. Emerald ash borer distribution in China, South Korea, and Japan
based on literature recordings and field surveys conducted from 2003-
2006 .................................................................................................. 30

Figure 2.3. Fraxinus spp. host tree distribution of emerald ash borer in China...32

Figure 2.4. Collection locations of emerald ash borer in the native range of Asia
......................................................................................................... 35

Figure 2.5.a. Collection locations for emerald ash borer in North America ........ 37

Figure 2.5.b. Expanded view of collection locations for emerald ash borer in
North America ..................................................................................... 38

Figure 3.1. Strict consensus of six parsimonious trees for 9 emerald ash borer
haplotypes ......................................................................................... 77

Figure 3.2. Haplotype network for 326 emerald ash borer individuals collected in
Asia and North America based on partial sequences (481 bp) of the
mitochondrial cytochrome oxidase l .......................................................... 78

Figure 3.3. Neighbor-joining analysis of EAB AFLP data set based on pairwise
(Pm values ............................................................................................ 88

Figure 3.4. Neighbor-joining analysis of EAB AFLP data set based on pairwise
(Dp. values for the native range in Asia and introduced range in North
America .............................................................................................. 89

Figure 3.5. Structure analysis results representing the most probable number of
populations of EAB, K = 4 through K = 7 in Asia .......................................... 90

Figure 4.1. Neighbor-joining analysis of EAB microsatellite loci (two) based on
pairwise FST values for Asian and North American populations ................... 110

Figure 4.2. Test for isolation by distance in the North American EAB populations
based on the analysis of microsatellite markers (2 loci) ............................... 111

xii

 

CHAPTER I:
INTRODUCTION

Invasive Species

Introduced and established exotic species are an increasing challenge
throughout the world, given enhanced globalization and trade. Invasive species
can have serious implications on the global environment (Elton 1958; Simberloff
2005; Perrings et al. 2005). There are countless examples of human action and
movement causing unprecedented movement of organisms to novel habitats
having adverse impacts in the new environment, ranging from examples such as
the near extinction of the American chestnut due to the invasive plant disease
chestnut blight imported accidentally from Asia to North America (Elton 1958) to
the extinction or endangered status of most birds on the island of Guam due to
the accidental introduction of the brown tree snake from Australia/New Guinea
(Burdick 2005). Many introductions of species into North America are intentional,
and a few are even beneficial, such as the crops and livestock that currently yield
98% of the US food supply and generate $800 billion per year (Pimental et al.
2000). However, many non-native species, either accidentally or intentionally
introduced, have become invasive costing, approx. $137 billion per year in
losses, damage, and management. An invasive species is characterized as an
established, widespread exotic species and in sufficient numbers to threaten
native species (Colautti and Maclsaac 2004). While the financial responsibility to
detect and eradicate an introduced species rests solely on the country in which

they are newly discovered (Simberloff 2005), many only become noticed when

they become invasive in the new geographic range. The potential for species
introduction increases with the volume of trade to or from a particular region
(Semmens et al. 2004; Levine and D'Antonio 2003), increasing the probability of
introducing a potentially invasive organism.
Ecological impact of invasive species

Once an introduced species becomes invasive, a variety of ecological
consequences may occur (Kenis et al. 2009). Successful invaders tend to out-
compete native species for space and resources; possibly due to fewer direct
competitors, predators and prey defenses causing displacement, or introgression
or extinction (Mooney and Cleland 2001; Sakai et al. 2001). For example,
introduced purple Ioosestrife, Lythrum salican'a Linnaeus, grows rapidly in
wetlands throughout its introduced range, creating monospecific stands in which
native wetland species are unable to compete for space or food, and thus
become rare or eliminated (Strefeler et al. 1996). This has long-lasting effects
not only for the plant life, but also for the animal life that use the native species
for shelter or food. If animals are unable to adapt or disperse to new habitat their
population will decline. Hybridization between introduced species and wild native
relatives is another consequence of introductions and can lead to the transfer of
invasive traits (Strefeler et al. 1996). Mallard ducks (Anus platyrhynchos
Linnaeus) hybridize with a variety of other species in their introduced range
(Hawaii, Australia, and New Zealand), including the Hawaiian duck (Anus
wyvilluana Sclater) or the New Zealand gray duck (Anas superciliosa superciliosa

Gmelin) (Rhymer and Simberloff 1996). This hybridization has led to the

reduction in populations of the New Zealand gray and Hawaiian duck
populations.

Argentine ants (Linepithema humile Mayr), now widely distributed in North
and South America, Africa, and Europe, displace native ants species presumably
because they are better at food acquisition (T sutsui et al. 2001). The traits that
have led to the success of the ants are thought to have been caused by a genetic
bottleneck in the introduced ranges leading to a reduction of intraspecific
aggression and formation of supercolonies (T sutsui et al. 2000; Tsutsui and Case
2001). Due to the increased land needed to maintain supercolonies, Argentine
ants have changed the landscape where they are found. The red imported fire
ant, Solenopsis invicta Buren, accidentally imported into the US. from South
America in the early 1900’s, also changes the landscape in introduced regions
causing a variety of ecological impacts including displacement of native species,
killing poultry and reptiles, and medical complications due to painful bites (Ross
et al. 2007; Shoemaker et al. 2006). Significant ecological changes have also
been observed due to the intentional introduction of the gypsy moth, Lymantria
dispar Linnaeus, from France to North America in 1869 for silk production.
Repeated defoliation causing tree decline and death has occurred across
hardwood forests in eastern N. America (Reineke et al. 1999). This has reduced
native animal species, including mammals, birds and insects, dependent on tree
resources for survival (Elkinton and Liebhold 1990).

Ecological damage can occur during range expansion and invasion of

species to new continents in cropping systems. For example, the Colorado

potato beetle (Leptinotarsa decemlineata Say) (CPB), native to Mexico and
feeding on buffalo bur, expanded its range north into the US. in the mid 1800’s
where it feeds on various solanaceous plants including the potato (Grapputo et
al. 2005). CPB was also accidentally introduced into Europe in the 1920’s. It
quickly spread through the continent. The spread of this pest has caused severe
loss in crop yield both in North America and Europe, exacerbated by extensive
insecticide resistance. The western com rootworm (Diabrotica virgifera virgifera
LeConte) (WCR) has a similar range expansion in N. America and invasion in
Europe to the pattern observed in CPB. It is thought to have originated in
Nebraska, Colorado, and Kansas expanding its range north feeding on cultivated
corn (Kim and Sappington 2005). It was accidentally introduced into Yugoslavia
in the 1990’s and has since been found in France, Italy, Switzerland, Belgium,
the United Kingdom and the Netherlands. Outbreaks of the western corn
rootworm have caused severe crop loss in both North America and Europe (Kim
and Sappington 2005; Miller et al. 2005). Increased land use for host crops and
the speed of spread and development of insecticide resistance in both CPB and
WCR cause severe damage to crop systems.
Control strategies of invasive species

Given the potential for major environmental consequences, considerable
effort is placed on eradication or control of introduced species that become
invasive. Successful eradication projects have common characteristics and are
generally more cost effective than the mechanical, chemical or biological control

methods later (Simberloff 2005). The little fire ant, Wasmannia auropunctata

Roger, had displaced and reduced native ant populations on Marchena Island,
Galapagos since being introduced in the 1900’s. A comprehensive and
widespread eradication project was initiated, that included a toxic commercial
pesticide AMDRO specific to the fire ant to reduce and eliminate it from all
regions it was found on the island. Within two years, the fire ant was almost
eliminated and native ant species began to recolonize the area (Causton et al.
2005). Eradication of three of four invasive fruit fly species (Bactrocera spp.),
devastating mango and breadfruit production, was successful in the Republic of
Nauru. This program was implemented by using protein bait application, male
annihilation technique, managing fallen fruit, and developing more stringent
quarantine regulations to reduce the risk of re—introduction (Allwood et al. 2002).
These two programs (W. auropunctata and Bactrocera spp.) may have been
successful for a variety of reasons. Both programs were on islands, which would
limit the spread and geographic range expansion potential of the invading
species, and in each case the eradication effort was developed and implemented
fairly quickly after the problem was discovered. In each case there was also a
widespread and comprehensive eradication plan that included a variety of
collaborators. All of these qualities increased the likelihood of successful
eradication. However, even in the absence of one of these factors, successful
eradication may be possible if designed well and enough labor and financial

support are in place (Simberloff 2005).

Detecting evolutionary change associated with invasions
Variation within a population

Diversity is a key component to every environment and population.
Because habitat is not continuous over a species’ range, many populations are
composed of metapopulations, where many subpopulations fluctuate in isolation
from each other (Hanski 1999). Subpopulations randomly become extinct and
recolonize through time as populations drift and migrate (McCullough 1996).
Genetic variation within a species population is thought to be the foundation for
its continued survival in an ever-changing climate (Frankham 1996).

Invasive species have special characteristics in their population dynamics
compared to populations in the native regions since they typically originate from
a relatively small population when introduced (Sakai et al. 2001; Kenis et al.
2009). Many ecological hypotheses have been suggested to explain why
introduced species become invasive and proliferate, such as, absence of natural
enemies, absence from competitors, and/or open ecological niches in the new
environment (Orians 1986).

There could also be a genetic basis to explain why certain species
become invasive. Most introduced species have experienced a population
bottleneck at the time of introduction; i.e. the number of individuals in the
introduced population is relatively small compared to the native population, thus
the genetic composition of the founder population is greatly reduced (Allendorf
and Luikart 2007; Sakai et al. 2001). Therefore, the probability of reduced

genetic variability, measured by the amount of heterozygosity and loss of rare

alleles, is increased as the founding population size and number of founding
populations decrease. This reduction of variation could lead to founder effects
with the genetic composition in the introduced population changed from the
native range. Reduced genetic variation could also increase the rate of change
in allele frequency leading to genetic drift (Allendorf and Luikart 2007). On the
other hand, there could be an increase in genetic variation (increased
heterozygosity and numbers of alleles) in an introduced population if multiple
introductions from different source populations or if the single founding
population was from multiple source populations (Baker 1992; Kolbe et al. 2004;
Allendorf and Luikart 2007). This could occur if a mixture of populations occurred
in the native region, such as combining goods and products at a port before
shipment. Determining the genetic structure and composition of an invasive
species is important since an increase or decrease in variation could influence
management practices. For example, introduction of a natural enemy from the
single source population would have a higher probability of controlling the
invasive species with little genetic variation than a population with high genetic
variation (Conord et al. 2006).
Molecular markers to assess genetic variation

Molecular markers, such as mtDNA gene sequences, nuclear gene
sequences, amplified fragment length polymorphic DNA (AFLP), and
microsatellites, are powerful tools to assess genetic variation in populations and
identify sources of biological invasions (Roderick 2003). Fonseca et al. (2001)

used multiple molecular markers to study the invasion of Aedesjaponicus

japonicus Theobald into the USA. The authors used random amplified
polymorphic DNA (RAPD) and mtDNA sequencing of NADH dehydrogenase
subunit 4 to compare the genetic diversity of insects in the introduced range and
the native range in Asia, and to identify the source population responsible for the
introduction into the USA. Grapputo et al. (2005) utilized a similar approach
using both mtDNA sequences and variation at AF LP loci to pinpoint the source
population of Colorado potato beetle into Europe from N. America. These
studies are examples of how to use multiple molecular markers to identify the
source of the invading population and assess genetic variation between native
and introduced populations.

Mitochondrial DNA has been studied extensively in animal systems,
including insects (see Harrison 1989; Caterino et al. 2000 for reviews). Absence
of recombination within mtDNA and its maternal mode of inheritance (exceptions:
Wallis 2000; Andolfatto et al. 2003) make mtDNA useful in detecting population-
bottleneck events caused by an introduction incident. Numerous primers are
available for amplifying insect mtDNA genes, making the amplification and
sequencing of mtDNA a straightforward process using PCR (Simon et al. 1994).
Mitochondrial DNA alleles found in a population can be sequenced and analyzed
via phylogenetic analysis.

AF LP genotyping (Vos et al. 1995), or genetic fingerprinting, has come
into use for studying population structure and differentiation, estimating
population genetic parameters (Reineke et al. 1998; Cardoso et al. 2000), and

studying closely related species (Albertson et al. 1999). The AF LP technique

yields information from many loci distributed throughout the nuclear genome
(Mueller and Wolfenberger 1999). AF LP data can be analyzed both within a
phylogenetic framework and from a population genetics perspective. AF LP data
can also be used to carry out assignment tests (Campbell et al. 2003).

Microsatellite loci, also known as short tandem repeats (STRs), are
characterized by the presence of multiple copies of a di—, tri-, or tetranucleotide
DNA sequences (e.g., (CA)... or (ATGC)n). Microsatellite loci are distributed
throughout the genome (Bruford and Wayne 1993), and analysis of microsatellite
variation is used to study a variety of problems (Balloux & Lugon-Moulin 2002).
Microsatellites are widely used to assess the genetic structure of populations
(e.g., Paupy et al. 2004) and study the geographic origin and subsequent spread
of introduced species (e.g., Miller 2001). Microsatellite markers have some
distinct advantages over other types of molecular markers, such as DNA
sequences of mitochondrial or nuclear genes, and various genotyping or DNA
fingerprinting techniques (e.g., RAPDs and AFLPs). Microsatellites are
codominant markers (both alleles in heterozygous diploid individuals can be
scored) and alleles at microsatellite loci can be scored accurately and fairly
quickly via a polymerase chain reaction-based strategy.

One important use of molecular data to assess an invasive species is to
determine the geographic origin of the putative source population(s) thought to
be responsible for an introduced population. Population assignment tests are an
informative method that attempts to assign an individual to the most likely

population from which it originated. These tests are used to answer a variety of

questions including: identifying dispersal patters in a metapopulation (Berry et al.
2004), confirming population structure of native and/or invasive populations
(Elderkin et al. 2004; Paupy et al. 2004), assessing population structure to design
eradication strategies (Abdelkrim et al. 2005), estimating the origin of invasive
species (Bonizzoni et al. 2001; Genton et al. 2005; Tsutsui et al. 2001), and to
assess if individuals are from an established population or recently migrated to
an area (Kim et al. 2006). These tests are typically performed using
microsatellite data, however, AF LP’s are considered a useful alternative for
species where microsatellite markers are unavailable or logistically difficult to
obtain (Campbell et al. 2003).
Genetic structure of populations in the native ranges

Bernardi et al. (1993) used both nuclear and mtDNA genes as well as
restriction fragment length polymorphism (RFLP) loci to understand the
population structure of the native teleost fish, Fundulus heteroclitus Linnaeus, in
North America ranging from Canada to Florida. Phylogenetic relationships of the
populations were compared using maximum parsimony analysis for all markers.
Each marker yielded similar results by detecting genetic separation between
northern and southern populations of teleost fish with a mixture of populations in
the region of overlap. The authors concluded that the different markers were
useful to obtain independent support for the separate populations. Chong et al.
(2000) found geographic structure among five populations of Malaysian river
catfish, Mystus nemurus Cuvier & Valenciennes, by analyzing samples with

AFLP and RAPD techniques. Fifteen to thirty individuals from each population

10

(one population on the island of Borneo and the remaining four populations from
the west side of Peninsular Malaysia) were characterized using both techniques
to determine the similarity of individuals within a population, genetic distance
between populations, and to carry out phylogenetic analyses based on calculated
similarity indices and genetic distance. The authors found that both AF LP and
RAPD analysis provided similar geographic structure, with the Borneo Island
population clustering alone. However, only AFLP analysis was able to detect
subpopulation structure of the populations from Peninsular Malaysia into three
identifiable subgroups. The authors concluded AF LP analysis was more efficient
in detecting structure in populations that are closely related than RAPD.
Population structure of gypsy moth collected in the native range from four
cities in western Japan was estimated with microsatellite markers (Koshio et al.
2002). The authors identified three microsatellite markers that yielded high
polymorphism with unique alleles detected in three distinct populations.
However, there was variation within a population as well as between populations
and the authors concluded that more loci would be needed to resolve populations
into separate groups. Population structure also was evaluated for Aedes aegypti
Linnaeus in Cambodia using multiple molecular methods, AFLP, microsatellite,
and isozyme markers (Paupy et al. 2004), in part to determine if all markers
yielded similar population structure measured by genetic differentiation (F...) The
authors determined all molecular markers were able to detect two distinct
populations in Cambodia when compared by an estimation of genetic

differentiation, Fst, however, the strength of differentiation was up to five times

11

higher in AFLP data compared to microsatellite and isoenzyme markers. In this
study, as seen before, it was important to evaluate different genetic markers to
ensure the estimate of population structure was valid and consistent patterns of
structure were detected.

Geographic structure of introduced species

Molecular markers are also useful to study biological introductions.
Genetic analysis can provide insight on a variety of questions pertaining to
invasive species, including 1) determine the geographic structure in native and
introduced ranges, 2) genetic make-up of the introduced population, 3) genetic
change after introduction, 4) geographic origin, 5) number of introductions, and 6)
assess hybridization of introduced populations with native species. Determining
the answers to these questions will provide valuable information on the biology
and ecology of the invasive species as well as possibly impacting the
development of eradication or management strategies (Allendorf and Lundquist
2003; Caldera et al. 2008)

Grapputo et al. (2005), with mtDNA COI and AFLP with two primer pair
combinations, evaluated the genetic diversity of the CPB (L. decemlineata)
throughout its native range in North America and introduced range in Europe.
The mtDNA variability was high through N. America with 20 unique haplotypes,
however, there was a single fixed haplotype found in Europe providing evidence
of a bottleneck event in the introduced population. AFLP analysis revealed
higher genetic diversity in N. America than in the introduced population in Europe

with the average level of polymorphism measured 67.47 and 47.98%,

12

respectively and average expected heterozygosity of 0.354 and 0.175,
respectively. The authors also found that specimens from North America and
Europe separate into two groups according to their continent of collection based
on Nei’s genetic distance (D). They concluded the reduced mtDNA and AFLP
variability in the introduced range in Europe compared to the native environment
from N. America could have been caused by one introduction or several small
introductions from the same source population.

D. virgifera virgifera, provides a unique situation in that it is historically
native to Nebraska, Colorado and Kansas but increased its range and became
invasive to the east coast within 60 years (Kim and Sappington 2005). It was
also accidentally introduced from N. America into Europe where it was initially
detected in 1992 (Miller et al. 2005). Kim and Sappington (2005) employed
seven microsatellite loci to characterize WCR population structure from the
native and range expansion populations in the United States using a total of 595
individuals sampled from ten populations across the current range. Genetic
differentiation was not detected among populations sampled with the
microsatellite loci used in the study with very little genetic differentiation detected
across the range (global Fst = 0.006). There was also very little differentiation
among populations with the majority of painrvise Fst values not detecting
significant variation. The authors attributed the lack of differentiation to continued
migration throughout the current range to maintain gene flow or there was
insufficient time for population divergence to be detected at the seven loci used

in the study.

13

Genetic makeup of introduced populations

Assessing the genetic structure of an introduced population may provide
important baseline information on the genetic makeup of the invading population
(Sakai et al. 2001; Allendorf and Lundquist 2003). For example, genetic
relatedness was evaluated for the invasive bark beetle, Tomicus piniperda
Linnaeus, of pine (Pinus spp.) to determine if introduced populations were
genetically similar (Carter et al. 1996). lsozyme loci and RAPD were used to
assess variation at eight localities in five states (Illinois, Indiana, Michigan, Ohio
and New York). RAPD analysis revealed the Illinois population separated from
the other four locations with a mean genetic distance of 0.895 while the other
four locations having a mean genetic distance of 0.595 among populations when
Illinois was excluded. lsozyme data analyzed with maximum parsimony analysis
supported the RAPD analysis with the Illinois population isolated from the rest of
the introduced populations. Given both RAPD and isozyme markers discovered
a separate population in Illinois differed from the remaining introduced
population, it is likely there was a separate introduction event in Illinois. This
information could be useful in developing management strategies since the same
control method may not be successful in the separate intrOduced lineages
(Allendorf and Lundquist 2003).
Population change after introduction

After introduction into a novel environment, the introduced population may
undergo rapid genetic change in response to the local environment (Mooney and

Cleland 2001). Drosophila subobscura Collin is an excellent example of an

14

introduced population evolving in a new environment (Huey et al. 2000). D.
subobscura was introduced to North and South America during the 1970’s from
its native range in the Old World. Although populations throughout the native
range exhibited an increase in body size with latitude, populations in the
introduced range did not exhibit this trend at the initial time of introduction (Huey
et al. 2000). The authors collected specimens from varying latitudes in both the
introduced range In N. America and in the native range in Europe two decades
after introduction and reared each population in a common garden for up to six
generations to ensure differences between populations were due to genetic
differences rather than unknown environmental causes. Twenty years after the
initial introduction into North America, D. subobscura populations had evolved to
exhibit a latitudinal cline; wing length increases with latitude similar to the Old
World populations. This study signifies the importance to re-examine the
variation in an introduced population through time to understand how the
populations may be responding to local environmental conditions.
Geographic origin of invasive species

Molecular markers provide researchers with powerful tools to identify
sources of biological invasions (Roderick 2003). Microsatellite markers, mtDNA
and satellite telemetry were used to determine the source populations of Canada
Geese, Branta Canadensis Linnaeus, in new colonies in Greenland (Scribner et
al. 2003). Assignment tests based on likelihood analysis were used to show the
Greenland population arose from East Ungava, and refuted the hypothesis that

the population arose from the North Atlantic population as inferred from flight

15

patterns. Cognato et al. (2005) used mtDNA cytochrome oxidase subunit I gene
(COI) partial sequence to determine the geographic origin of a destructive insect,
red turpentine beetle (Dendroctonus valens LaConte) introduced into China from
North America. Parsimony and statistical parsimony analyses of the haplotypes
determined populations in China were more closely related to populations from
the Pacific Northwest than to populations from Mexico and Michigan. The
population diversity was also unusually high in China suggesting that there was a
large founding population or multiple introductions.

Navia et al. (2005) evaluated the geographic origin and ancestral host
association of the invasive coconut mite, Acen'a guerreronis Keifer, with mtDNA
16S and nuclear ribosomal internal transcribed spacer sequences. They
compared individuals from populations throughout the current range, 12
countries from North and South America, 8 countries in Africa as well as Sri
Lanka and India for genetic variability and haplotype diversity. The authors
determined the genetic variability from populations in North and South America
was high, but only a single mtDNA 16$ haplotype in Africa, Sri Lanka, and India.
The level of genetic variation measured by % nucleotide diversity observed in the
ITS region was also higher in S. America than in Africa and southern Asia (1.90,
0.70, and 0.49, respectively). This lower genetic variation led the authors to
hypothesize that these countries contain the introduced populations while North
and South America are most likely the native range. This result also indicated a
recent host switch of the mite onto coconut since the plant is native to the Indo-

Pacific region and not from North or South America (Navia et al. 2005).

16

The Argentine ant provides another example of the utility of molecular
markers to estimate the geographic origin of an invasive species. The Argentine
ant has become invasive throughout most of the world currently inhabiting 15
countries on six continents. Determining the source of introduced populations
would provide valuable information on where to focus research for development
of biological control. Tsutsui et al. (2001) used mtDNA cytochrome b and seven
microsatellite loci to compare samples from the native range in Brazil and
Argentina to samples from introduced areas including California, Hawaii, Maui,
Louisiana, Bermuda, Australia, South Africa, Chile and Italy. Population
assignment tests were used to estimate the log-likelihood that each individual
genotype was found in potential source populations and employed to make
inferences about the geographic localities of source populations. Similar mtDNA
haplotypes were observed in all of the introduced populations, all native
populations in Argentina, and a single population in Brazil, while distinct
haplotypes were observed in other Brazilian individuals that separated from the
other haplotypes using a maximum parsimony analysis. Microsatellite analysis
provided a finer resolution than mtDNA between populations in the native and
introduced ranges. In the native range, pairwise genetic distances increased as
geographic distance increased forming a cluster of individuals from Brazil and
three separate clusters defining populations from Argentina (northern Rio
Parana, southern Rio Parana, and Rio Uruguay). The only native site that
grouped with introduced populations based on pairwise genetic distances was a

population from Rosario, Argentina. Assignment tests based on microsatellite

17

loci also determined the majority of the samples throughout the introduced range
were assigned to samples collected in Rosario, Argentina. The authors
concluded this location was the most likely source of the majority of introduced
populations throughout the world. This multi-Iocus approach provided strong
evidence of the geographic origin of introduced populations and should be
incorporated in future studies of invasive species whenever possible.
Number of introductions of invasive species

Miller et al. (2005) compared WCR populations from the native range in N.
America to introduced populations in Europe to determine the source of the
introduced populations in Europe. They used data from eight microsatellite loci
to compare genetic variation from populations on both continents and tested
hypotheses for the pattern of introduction based on computer simulations with
approximate Bayesian computation analysis (ABC). ABC analysis is able to test
different introduction scenarios by developing a linear regression model of
computer-generated parameter values using simulated summary statistics and
then replacing the observed summary statistics into the equation (Beaumont et
al. 2002). The authors determined there was strong evidence for at least three
separate introductions of WCR into Europe (Paris, France, northwestern Italy,
and central and southeastern Europe) from N. America. The analysis of the
number of introductions such as this has strong implications for management
strategies since eradication could only be successful if the pathways of

introduction are eliminated to stop future invasions of WCR.

18

Hybridization of introduced populations with native species

Population genetic variation has also been studied for a variety of invasive
species to determine effects of introductions on the genetic makeup of related
native species. For example, Strefeler et al. (1996) studied hybridization
between native and introduced L. salicaria. They proposed a low level of genetic
introgression, as revealed by isozyme data. However, Houghton-Thompson
(2001) found no evidence for hybridization in purple loosestrife when she
evaluated the species by AFLP.

For the invasive tamarisk plant (Tamarix chinensis Laureiro and Tamarix
ramosissima Ledebour), Gaskin and Schaal (2002) used an intron of a nuclear
gene, phosphoenolpyruvate carboxylase (PepC) to determine population
dynamics. The variable intron region revealed novel hybrid plants in the
introduced range in the United States, but few hybrids were found in the native
range in Eurasia. They continued to describe how this information can be helpful
in developing control strategies; in this case, the hybrid species with novel allele
combinations may have changed the biology of the plant enough to render
current biological control agents ineffective.

As noted above, molecular markers provide important information for
understanding the dynamics of invasive species. Incorporation of genetic
analysis to assess the population biology of invasive species may be critical to
develop practical management strategies for control (Sakai et al. 2001). Some
of these methods will be used in this dissertation to characterize the population

dynamics of an invasive species introduced into North America.

19

Emerald ash borer

The United States and Canada are currently battling a devastating
invasive pest, the emerald ash borer (EAB), Agrilus planipennis Fairmaire, which
is native to Eastern Asia. The pest was detected in North America in 2002
(Haack et al. 2002), and was initially found in southeast Michigan, northern Ohio,
and in southwestern (Windsor) Ontario (Scarr et al. 2002). Based on
dendrochronological reconstruction, Siegert et al. (2007) postulated that EAB
was introduced in the early to mid 1990’s into Wayne County, MI. Since the
initial collection in 2002 in the Detroit area of Michigan and Windsor, Ontario,
EAB has been discovered over a wide area including 15 states and two
Canadian provinces: Illinois, Iowa, Indiana, Kentucky, Maryland, Michigan,
Minnesota, Missouri, New York, Ohio, Pennsylvania, Virginia, West Virginia,
Wisconsin, Ontario and Quebec, Canada (Cooperative Emerald Ash Borer
Project 2009).

EAB is native in a vast region of Asia, including northern China, Japan,
Korea, Mongolia, the Russian Far East, and Taiwan (Y u 1992; Chinese Academy
of Sciences 1986). Several species were synonymized to what is currently
accepted to be EAB, A. planipennis, including A. marcopoli Obenberger in China,
A. molco-poli, A. feretrius Obenberger in Taiwan, and A. marcopoli subsp. ulmi
Kurosawa in Japan (Jendek 1994). This has implications on possible host range
of EAB since the host range in China includes various Fraxinus spp. (Oleaceae)
(Y u 1992; Chinese Academy of Sciences 1986), while a broader host range is

found in Korea, Mongolia and Japan that includes some species of walnut

20

(Juglandaceae) (Jug/ans spp.), elm (Ulmaceae) (Ulmus spp.), and wingnut
(Juglandaceae) (Pterocarya spp.), in addition to Fraxinus spp. (Akiyama and
Ohmomo 1997). Given the various host species in Asia, Anulewicz et al. (2008)
conducted a study to determine the extent of the host range in the introduced
region of N. America to better predict the potential extent of the damage that may
be caused. They determined EAB in Michigan successfully oviposited and
developed on a variety of N. America ash species, however, did not oviposit
and/or develop in other tree species tested, including Syringe reticulata
(Oleaceae) Japanese tree lilac, Jug/ans nigra (Juglandaceae) black walnut,
Ulmus americana (Ulmaceae) American elm, Carya ovata (Juglandaceae)
shagbark hickory, and Celtis occidentalis (Ulmaceae) hackberry.

Ash trees (Fraxinus spp.) are very common in the eastern hardwood
forests and urban landscapes of N. America. Ash was one of the primary tree
types used to replace American elms lost during the Dutch Elm disease epidemic
in the latter half of the 20th century (MacFarlane and Meyer 2005). It has been
estimated that Michigan has 693 million ash trees throughout its forests and I
urban environments (estimated value > $18.92 billion) that are susceptible to
EAB (USDA-APHIS 2003). Ash trees in the N. America, including Michigan,
have a variety of economic values including nursery sales, timber, baseball bats,
hockey sticks, and recreational use in state forests. Estimated loss in the
Michigan nursery industry was over $2 million in revenue the first year EAB was
detected (USDA-APHIS 2003). EAB has been responsible for killing millions of

ash trees in N. America (Cooperative emerald ash borer project 2009).

21

EAB larvae destroy ash trees (Fraxinus spp.) by feeding on phloem.
Adults feeding on ash foliage cause little defoliation. EAB develop in one or two
years under ash bark until emerging as adults beginning in May (Russell et al.
2003). Damage caused by EAB occurs when beetle density increases within a
tree and nutrient flow in the phloem is staunched. High densities of larvae create
feeding galleries that cover the majority of the cambium area girdling the tree
causing nutrient flow to cease, resulting in death of the tree (Russell et al. 2003).
EAB are able to develop in all ash trees greater than 2.5 cm, however, the
majority of EAB in a given area emerge from large trees (greater than 26 cm
diameter at breast height). Therefore, reducing the number of large ash trees
may be crucial in managing EAB in forested areas (McCullough and Siegert
2007).

Control Methods of EAB in North America

EAB eradication attempts have not been successful. This consisted of
cutting and disposing of infested trees, and removing all ash trees in 0.8 km zone
around outlier populations in the US. to slow down the natural spread of the
pest. Therefore, alternative methods of control are being evaluated, including
trapping (Poland et al. 2006; Crook et al. 2007; Poland and McCullough 2007),
chemical control (Cappaert et al. 2007; McCullough et al. 2007b; Rebek and
Smitley 2007; Tanis et al. 2007), and biological control (microbial and natural
enemies) (Liu et al. 2003; Bauer et al. 2004; Bauer et al. 2005; Marshall et al.
2005; Yang et al. 2005; Zhang et al. 2005; Yang et al. 2006; Liu et al. 2007; Liu

and Bauer 2008; Wang et al. 2008).

22

 

Detection of EAB

Several detection methods for EAB are being developed, including stress
induced girdling of trees, purple panel traps (used by USDA APHIS), and purple
panel traps baited with a leaf blend lure (Poland et al. 2006) and/or Manuka oil
(similar to ash bark volatiles) (Poland and McCullough 2007; Crook et al. 2007).
Purple traps placed in areas with high sun exposure and baited with Manuka oil
and leaf blend lures are very effective in capturing EAB (McCullough et al.
2007a). Traps placed along the edge of a tree line or in open fields, as well as
above ground (13 m) are most effective (Francese et al. 2008).

Cereceris fumipennis (Hymenoptera: Crabronidae), a known predator of
buprestid species, has been shown to effectively detect EAB in low and high-
density populations (Marshall et al. 2005). Females search for prey to provision
the nest; able to orientate to a new area and actively search for prey within 24
hours of transport. Since it attacks a variety of buprestid beetles and is not
typically found in great numbers, this predator will be most effective monitoring
new areas of infestation and not effective) at controlling EAB populations.
Chemical Control of EAB

Chemical control is being evaluated for control of EAB suppression and
protection of host trees. lmidacloprid applied by soil drench (Rebek and Smitley
2007; Cappaert et al. 2007) and trunk injections in the fall and spring (Tanis et al.
2007) were tested for effectiveness for tree protection. lmidacloprid is only mildly
effective in reducing EAB larval activity, resulting in only partial control and tree

mortality was still likely. Trunk injection with emamectin benzoate shows this

23

product is more effective with 100% adult mortality on foliage and over 99%
control of larvae within infested trees (McCullough et al. 2007b). This could
provide effective control of EAB for high-value urban ash trees. However, these
methods are expensive, labor intensive, and environmentally risky for the
protection of forests.
Biological Control

Invasive insect species may be successfully controlled with biological
control: the introduction and establishment of pathogens, parasite-s, or predators
from the pest’s home of origin. Although the presumed source of the infestation
is China (due to international trade), beetles may have originated from other
Asian countries. Even though the specific origin was unknown when EAB was
initially detected, a search for biological control agents across the native range
was initiated (Liu et al. 2003). Several parasitoids were identified that attack
EAB. Spathius agrili (Hymenoptera: Braconidae), a larval ectoparasitoid found in
China, develop up to 2 generations per year on EAB (Liu et al. 2003; Yang et al.
2005; Wang et al. 2008). Two generations of an egg parasitoid, Oobius agrilli
(Hymenoptera: Eulophidae) were found in high density attacking EAB in
northeastern China (Liu et al. 2003; Zhang et al. 2005; Liu et al. 2007). Another
biological control agent was located in the same location in northeastern China;
the larval endoparasitoid, Tetrastichus planipennisi (Hymenoptera: Eulophidae),
attacked EAB in four separate generations per year (Yang et al. 2006; Liu et al.
2007). O. agrili and T. planipennisi reduced populations by over 70% in 2005

(Liu et al. 2007). Balcha indica (Hymenoptera: Eupelmidae), an invasive

24

parasitoid from Southeast Asia, has also been recorded to attack EAB (Gibson
2005).

There have also been biological control agents found native to North
America that have attacked EAB. For example, Liu and Bauer (2008) evaluated
an entomopathogentic fungus, Beauveria bassiana strain GHA, sold as
BotaniGard ES, by foliar and trunk sprays on EAB colonized ash trees. Over
40% reduction in EAB population was achieved in treated trees with 47%
reduction of larval density. These authors concluded that B. bassiana might be
an effective method to slow the spread of EAB and possibly contain an
infestation in an outlier site.

Despite the potential for chemical control of EAB with ememectin
benzoate, and control with biological control agents identified from Asia and N.
America, the spread of EAB continues by natural dispersal and human transport.
Given the fact that natural enemies found from the same source population of
EAB will have a higher probability of success in control strategies, one goal of the
work described in this dissertation was to test EAB populations throughout its
native range in Asia to determine the specific origin of N. American EAB. This
knowledge will be useful in focusing effort on continued establishment of
biological control agents from the same region of origin to increase the likelihood

of success to control EAB in the introduced range.

25

Objectives

The overall objectives of this study were 1) to determine the geographic
origin of North American populations of EAB, 2) to determine if North American
EAB populations were the result of a single or multiple introductions, and 3) to
determine the spread of EAB from the source of its introduction in North America.
In an attempt to answer these questions, I used genetic methods to determine

the geographic origin and monitor the spread throughout the introduced range.

26

CHAPTER II
GEOGRAPHIC DISTRIBUTION OF EMERALD ASH BORER

Introduction

Emerald Ash Borer (EAB), Agrilus planipennis Fairmaire (Coleoptera:
Buprestidae), was first detected in Southeast Michigan near Detroit in May and
June 2002 (Haack et al. 2002). Collected adults were initially evaluated by Gary
Parsons (Department of Entomology, Michigan State University) and identified as
an Agrilus sp. unlike species known to the Great Lakes region. The beetles or
images were evaluated by five other US beetle experts to identify these
specimens to the species level. All five experts agreed on the identification of the
genus, suggesting it was most likely of Asian origin; however, species
identification remained unclear. Digital images, followed by specimens, were
then evaluated by an Asian Agrilus spp. expert in Slovakia, Eduard Jendek,
(State Forest Products Research Institute, Slovakia) for identification. He
positively identified the specimens as A. planipennis in July 2002 (Haack et al.
2002). Following the positive species identification, adult EAB were collected in
Windsor, Ontario and subsequently confirmed to be A. planipennis by Richard
Westcott (Oregon Department of Agriculture). Since the initial collection in 2002
in the Detroit area of Michigan and Windsor, Ontario, EAB has been discovered
in other areas of North America. In 2003 it had been discovered in Northern
Ohio; in 2004 Maryland, and Indiana; in 2006 throughout most of Michigan’s
Lower Peninsula and in areas of the Upper Peninsula Michigan. By the end of

2007, EAB range extended into Illinois, Pennsylvania, and West Virginia. Within

27

the next year (2008) populations were discovered in Missouri, Virginia,
Wisconsin, as well as Quebec, Canada with the current range in 2009 to include
locations in Iowa, Minnesota, New York, and Kentucky (Cooperative Emerald
Ash Borer Project 2009; Figure 2.1).

Information on EAB was limited before its discovery in N. America (Liu et
al. 2003; Wei et al. 2004). EAB is native to eastern Asia, including northeastern
China, Inner Mongolia, Japan, Korea, Mongolia, the Russian Far East, and
Taiwan (Figure 2.2) (Chinese Academy of Science 1986, Wei et al. 2004, Yu
1992). Several populations of EAB have been recognized as separate species in
the literature. In China, EAB is commonly referred to as A. marcopoli
Obenberger, in Korea and Japan A. marcopoli ulmi Kurosawa, and A. feretrius
Obenberger in Taiwan (Haack et al. 2002). These species were synonymized by
Jendek (1994) as A. planipennis. Although details were not provided for the
rationale for synonomyzing the species, it was stated the decision was based on
the overlap of morphological characteristics of type specimens for each species.
Adult EAB are normally 8.5-13.5mm in length in native China (Yu 1992) and
11.5-15mm in native Japan (Akiyama and Ohmomo 2000). The adults are
generally metallic green in color with purple on the dorsal surface of the
abdomen. Color morphs differing from the standard metallic green include
metallic bronze, gold, copper, or ruby color in the native range (Yu 1992;
Akiyama and Ohmomo 2000), however, morphs differing from green are rare in

the introduced range (personal observation).

28

 

Figure 2.1. Emerald Ash borer introduced distribution range in North America reported
as of July 2009. Dots are locations where EAB has been detected. Numbers correspond
to states or province: l-Minnesota, 2- Missouri, 3-Wisconson, 4-Illinois, 5-Michigan, 6-
Indiana, 7-Kentucky, 8-Ohio, 9-West Virginia, lO-Ontario, ll-New York, 12-
Pennsylvania, l3-Maryland, 14-Virginia. (Figure taken from Cooperative Emerald Ash
Borer Project 2009).

29

 

 

 

 

 

Figure 2.2. Emerald ash borer distribution in China, South Korea, and Japan
based on literature recordings and field surveys conducted from 2003-2006
in shaded areas

30

Host range

In Asia, EAB has been reported primarily on Fraxinus spp. (Oleaceae)
hosts with some variation depending on location. In China EAB has been
reported on F. chinensis subsp. chinensis Robx., F. chinensis subsp.
rhychophylla Hance, and F. mandshurica Ruprecht. but causes less damage to
these native species than to N. American spp. (Liu et al. 2003; Wei et al. 2004).
The distribution of native ash in China includes central and eastern China
extending from the southern to northern border (Wei et al. 2004; Figure 2.3).
EAB, however, aggressively attacks ash trees imported into China from North
America for use primarily as ornamental trees. F. velutina Torr (Velvet ash) was
first imported into China in the 1950’s, while F. americana Linnaeus (white ash),
F. pennsylvanica var. Ianceolata Marsh (green ash), and F. pennsylvanica Marsh
(red ash) were imported in the 1960’s. These imported species of ash must be
managed in China as EAB may kill entire plantings (Wei et al. 2004). The host
range of EAB in Korea includes F. mandshurica, F. rhynchophylla, F.
chiisanensis Nakai, and F. sieboldiana Blume (Haack et al. 2002). The host
range in Japan includes F. japonica Maximowicz (Japanese ash), F. spaethiana
Lingelsh (Spath’s ash), F. apertisquamifera Hara, F. Iongicuspis Siebold and
Zucc., F. platypoda Oliver and F. lanuginosa Koidzumi. In addition, EAB records
exist for collections from walnut and elm, including Jug/ans mandshurica var.
sieboldiana Maximowicz and var. sachalinensis Kitamura (Juglandaceae)
(Manchurian walnut), Pterocarya rhoifolia Siebold and Zucc. (Juglandaceae)

(Japanese Wingnut), and Ulmus davidiana var. japonica Nakai and U. propinqua

31

 

 

 

 

- p
g -

 

 

Figure 2.3. Fraxinus spp. host tree distribution of emerald ash borer in China
(shaded areas). Distribution based on Wei et al. 2004

32

Koidz (Ulmaceae) (Japanese elm) (Akiyama and Ohmomo 2000; Haack et al.
2002)

In the introduced range of N. America, EAB has developed successfully
on all ash species encountered to date (Anulewicz et al. 2008): F. Americana, F.
pennsylvanica, F. nigra Marshall (black ash), F. quadrangulata Michaux (blue
ash), and F. profunda Bush (pumpkin ash). EAB populations in N. America have
not been found to develop on nonash species, including walnut or elm species
(Anulewicz et al. 2008).
EAB Life History

Biological information on EAB is scarce (Haack et al. 2002); however the
following information is based on literature from the Chinese Academy of Science
(1986) and Yu (1992). Adult EAB typically emerge and are active from mid May
to August. Males and females live approx. two and three weeks, respectively
after emergence from the host tree. Adults feed on ash foliage (Fraxinus spp.)
during the daytime preferably when it is warm and sunny. Mating occurs after 7-
14 d of feeding with copulation lasting 20-90 min. Females oviposit eggs
individually in bark crevices 7-9 d after the first mating with a lifetime production
of 68-90 eggs. Eggs are found in early June to late July maturing for 7-21 d.
Larvae hatch from eggs and bore through bark to the cambium layer where they
feed in a serpentine pattern creating an S-shaped gallery while developing
through four instars. One generation per year or one generation every two years
occurs in both the native and introduced ranges. When univoltine, EAB

overwinter as mature larvae, however, when development occurs over two years

33

EAB overwinter as early instar larvae the first winter and mature larvae the
second winter. EAB pupate in spring lasting 14-21 d (at 24°C) and emerge as
adults after another 7-14 (I (Petrice and Haack 2006).
Methods for EAB collection in introduced and native range

EAB were sampled for this study from native populations in China, Japan,
and South Korea (Figure 2.4) and introduced populations in Illinois, Indiana,
Michigan, Ohio, Pennsylvania, West Virginia and Ontario (Figure 2.5). Samples
were also collected from an introduced location in Moscow, Russia.
EAB collection with collaborators

Due to the large geographic distribution of EAB, a network of collaborators
was established to assist in insect collection. Leah Bauer PhD (USDA Forest
Service), Houping Liu PhD, and Therese Poland PhD (USDA Forest Service)
throughout China and Michigan, Roger Fuester (USDA-ARS) in China, Paul
Schaefer (USDA-ARS) in Japan, Bob Haack PhD (USDA-FS) with Oleg Kulinich
PhD in Russia, and David Williams in South Korea coordinated collection of
Asian samples (Figure 2.4). Additional collaborators, Deb McCullough PhD and
David Smitley PhD, of the Department of Entomology at MSU and several
employees of the Michigan Department of Agriculture collected individual beetles
from populations throughout Michigan. Lucy Hunt (Ohio Department of
Agriculture), Mark Cinnamon (Illinois Department of Agriculture), Benjamin West
(Indiana Department of Natural Resources), and Dick Bean (Maryland
Department of Agriculture) assisted in beetle collection in Ohio, Illinois, Indiana,

and Maryland respectively. Ches Caister, Plant Protection officer Southwest

34

Figure 2.4. Collection locations for Emerald Ash Borer in the native range of Asia
(2002-2007). Province or provincial level cities with collection locations are as
followed: Tianjin City (1-Dagong, 2- Hangu), Liaonging (3 and 4- Shenyang, 5-
Benxi), Jilin (6-Laoniujuan, 7-Changchun City, 8- Jiutai, 9- Jilin City), Hebei (10-
Tangshan), Beijing (11-Chaoyang), Heilongjiang (12 — Harbin), Miyagi (13-
Shirioshi City), Chungchong-nam do (14- Daejeon), Gyeonggi-do (15- Suwon),
Gyeongsang-bukto (16- Sangju), Gangwon-do (17- Inge, 18- Samcheok)

35

 

 

 

 

 

36

 

 

 

 

 

 

 

 

 

Figure 2.5.a. Collection locations for emerald ash borer‘in North America (2002-2007). Boxed area is expanded in figure

2.5.b.

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region, collected specimens in Windsor, Ontario. R.M. Turcotte (USDA-FS)
collected specimens in West Virginia. Nathan Siegert PhD (Dept. of Entomology,
MSU) and Sven-Erik Spichiger (Pennsylvania Department of Agriculture)
collected specimens in Pennsylvania. North America collection locations are
shown in figure 2.5.

All collaborators were requested to follow the following field collection
technique. Visual inspections of Fraxinus spp. were conducted at breast height
for evidence of EAB activity, such as, the characteristic D-shaped exit hole or
external tree evidence of larval activity of bark cracks or woodpecker holes.
Adult beetles were collected in the field from mid May to late July with a sweep
net; samples were killed and stored in 90-100% ethanol or killed by freezing and
stored in ethanol and shipped to the lab in East Lansing, MI. To collect larvae,
bark was removed from trees with a drawknife to survey for larval serpentine
galleries. This was done at the field site on standing trees, or trees were cut into
logs (60 cm sections of cut tree) in the field and returned to East Lansing, MI for
bark removal with knives and chisels. When a gallery was located, the larvae or
pupae were removed with forceps and placed in 90% ethanol or on larval diet
(developed by Juli Gould, USDA-APHIS; Liu et al. 2007) for further development.
All dead insects were sent to East Lansing, MI for genetic analysis. Live insects
were shipped to the USDA-FS lab in East Lansing, MI for development on media
or in logs for natural enemy evaluation (project conducted by Dr. Bauer (USDA-

FS) and Dr. Liu (MSU). Live specimens that did not contain natural enemies

39

were preserved in 90% ethanol for genetic analysis.
Collection of North American EAB samples

In addition to samples collected by collaborators, samples were collected
directly. EAB were collected in the summer of 2005 and 2008 in Lansing, MI
from F. pennsylvanica. Four trees were felled in 2005 and three felled in 2008,
cut into two feet logs and transported to the laboratory. Bark was removed from
three logs from each tree. Larvae were removed and killed in liquid nitrogen, or
frozen at -20°C or -80°C and stored in 95% ethanol at -20°C. Six additional logs
from each tree (24 total) in 2005, held at 4°C to simulate winter conditions, were
used for EAB adult development. After 2-4 months in chill, the infested logs were
placed into heavy cardboard tubes capped with plastic ends at 24°C to stimulate
pupation and adult development. Emerged adults were collected on a daily
basis, killed and stored at -80°C.

Collection of EAB was also conducted in Shipshewana, Indiana in January
2005. Fraxinus spp. were located on a woodlot adjacent to farm land and
visually observed for EAB activity. Adult emergence holes (D-shaped) were
observed on 10 trees; therefore, bark was peeled at breast height to locate and
expose larval galleries. Larvae and pre-pupae were removed and placed
individually into wells of 24-well tissue plates and allowed to develop into adults.
Sampling was also conducted in Oregon, OH and N. Baltimore, OH in January
2005. Urban street trees were felled and visually observed for D-shaped adult
exit holes. Bark was peeled in the field and larvae and pre-pupae were removed

and placed into individual wells for further development. Six 60 cm logs were

40

also maintained at 5°C until used to rear adults as described above. In June
2005, a Fraxinus spp. tree farm in Windsor, Canada was inspected for adult
EAB. All trees were approx. 8-12 feet tall. Adult EAB were collected with an
aerial net from the leaf canopy and killed in 70% ethanol for transport to East
Lansing, MI in accordance with the Canadian Food Inspection Agency
requirements.

Collection of Asian EAB samples

One of the most challenging aspects of this research project was obtaining
samples from the native range of EAB in Asia. Because it was very difficult for
the foreign collaborators to take time to collect themselves, a sampling trip was
conducted in the native range of EAB in South Korea, China and Japan in the
summer of 2006. The focus for site selection was on locations EAB was
documented in the literature, collected in the past, and from personal
communication with collaborators. Collection occurred mostly in natural forests
and urban trees in South Korea, plantation and urban trees in China, and natural
forests in Japan (Table 2.1). All sites were visually inspected for D-shaped exit
holes and crown dieback. If symptoms were found, bark was removed at breast
height to locate larval galleries.

Several locations in South Korea were identified for survey with the
primary host tree sampled being F. rhynchophylla. Visual survey for host trees
were conducted at Mt. Juri near Jinju. One tree with significant dieback at the
Gyeongsang University Experimental Station was felled and sectioned into one

meter logs for transport to Gyeongsang National University for bark removal.

41

 

 

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Several more sites were visually surveyed for host trees and collection including
Mt Muju, Mt. Sangju, Mt. Wolak, and Suwon (Table 2.1.a).

In China, nine collection locations were surveyed and sampled from four
provinces: Liaoning, Jilin, Hebei, Heilongjiang, and two provincial-level cities:
Tianjin City and Beijing (Table 2.1.b). Collection locations were chosen, in part,
in accordance to sites previously successfully sampled within five years (Liu et al. .
2003) and personal communication with local foresters. Surveying concentrated
on native host species including F. mandshun'ca, F. rhynchophylla, F. chinensis,
and introduced host species F. velutina and F. pennsylvanica (Liu et al. 2003).

Finally, five geographic locations were surveyed and sampled in Japan
concentrating on native host trees known from the literature F. japonica, F.
spaethiana, F. apertisquamifera, F. Iongicuspis, F. platypoda, F. lanuginosa, J.
mandshun’ca var. sieboldiana and var. sachalinensis, P. rhoifolia, and U.
davidiana var. japonica (Akiyama and Ohmomo 1997; Table 2.1.c). Visual
surveys in Chichibu and Morioka were restricted to trees along roadsides for
safety precautions due to the presence of native poisonous snakes and bears.
Results of emerald ash borer collection

For this study, an extensive EAB collection was assembled. Because
EAB was first detected in Michigan (Haack et al. 2002), widespread surveys were
conducted for the presence of EAB in this state. For this project, I obtained
samples throughout the current range of EAB distribution to include EAB from 43
sites in 21 counties, including two in the upper peninsula for a total of 1652 EAB

individuals (Table 2.2). Samples were also obtained with collaborators or

49

 

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52

collected directly from Illinois (one site; 10 individuals), Indiana (1; 14), Maryland
(2; 13), Ohio (4; 55), Pennsylvania (2; 50), West Virginia (1; 5), and Windsor,
Canada (1; 114) (Table 2.2). Thus, the collection presented here represents
most of the current known range of EAB in N. America. Samples were not
obtained for this project from more recently detected (2008-2009) sites in
Kentucky, Minnesota, Missouri, New York, Virginia, Wisconsin, and Quebec,
Canada (Cooperative emerald ash borer project 2009), due to time limitations
and resource availability.

In the summer of 2006, field surveys for EAB in South Korea were carried
out at 11 sites in six provinces spanning the range of peninsula (Table 2.1.a).
Surveys included a range of locations to include urban street trees, plantation
forests and natural forests to include a range of possible habitats for EAB. EAB
adults were collected from six of the 11 sites in four provinces (Table 2.3). Host
tree associations could not be verified for these specimens since only adults
were obtained; however, all were collected in close vicinity of F. rhynchophylla.
Trees, mainly F. rhynchophylla, that exhibited signs of EAB infestation (crown
dieback, bark splitting, and D-shaped exit holes) were peeled of bark to locate
larval galleries. Several trees peeled were heavily infested with larvae; however,
identification of the collected samples indicated that they were in the family
Cerambycidae and not A. planipennis. Even though EAB larvae were not
detected at the time of collection, the presence of D-shaped exit holes on the
trees leads to the hypothesis that EAB (or a different Buprestid spp.) did infest

the tree in the past.

53

Table 2.3. Emerald ash borer specimens obtained/collected in Asia from 2003-

 

2008
Country Province Locality Collector #Adults #Larvae
(or equiv.)
China Heilongjiang Harbin Liu/Bauer/Zhao 4 3
Tianjin City Dagong Liu/Bray/Fuester 22 94
Hangu Liu/Bauer/Zhao O 1 1
Jilin Jilin City Liu/Bauer/Zhao O 4
SongHua
Lake Park Bray/Fuester O 13
Jiang Nan
Forestry
Center Bray/Fuester 0 8
Laoniujuan Liu/Bauer/Zhao 0 20
Jiutai Liu/Bauer/Zhao 0 20
Hebei Tangshan Liu/Bauer/Zhao 0 20
Zhuang-Zi Fuester O 21
Liaonging Benxi Liu/Bauer/Zhao 0 6
DongYingFang Fuester 2 O
Shenyang Fuester 1 5
Shenyang Liu/Bauer/Zhao 0 10
Beijing Chaoyang Liu/Bauer/Zhao 0 10
South
Korea Gangwon-do Hoengseong Williams 1 0
Samcheok Williams 1 O
Inge Williams 5 0
Chungchong-nam do
Daejeon Williams 9 O
Gyeongsang-bukto
Sangju Bray 1 0
Gyeonggi-do Suwon Bray 1 0
Japan Miyagi Shiroishi City Schaefer/Bray 1 0
Russia Moscow Kulinich 12 O

54

In China, field surveys for EAB in all Fraxinus spp. present were
conducted by collaborators or directly at 19 sites in four provinces and two
provincial-level cities (Table 2.1.b). Larvae were successfully collected from 15
of the sites, representing all provinces (Table 2.3). Evidence of past EAB
activity, crown dieback in edge trees, exit holes, and larval galleries, was found at
the remaining 4 sites although samples were not collected. Historical records of
EAB host range include Shandong province (CAS 1986, Yu 1992), however
surveys within the province have not been successful (Liu et al. 2003).

Surveys for EAB in Japan included all historical host species including,
ash, walnut and elm species. Extensive surveys for EAB were conducted at 11
sites in seven prefectures (Table 2.1.0). One location, Forest of Azalea near
Shiroishi City, was found to have an active population. One adult was collected
by collaborators Mr. Kaneko and Dr. Schaefer at this location in 2004 as well as
two larvae from F. Ianuginose by this project in 2006. The larvae collected were
relatively small, possibly 1st or 2nd instars. and difficult to identify to species.
There is another metallic woodboring beetle known in the same area to attack F.
lanuginosa, Agn'lus koyoi Ohmomo, which is difficult to distinguish from EAB in
early instars. Very little detail is known about the biology of A. koyoi since it was
recently described in 2002 by Ohmomo (Jendek 2007). It is much smaller than
A. planipennis being 5.1-6.2 mm adult length compared to 8.5-15mm, however, it
is similar in color being metallic green to bronze, have the same host tree, and

located in the same forest near Shiroishi City. Two adults of A. koyoi, were

55

collected in 2006 at this location confirming an active population of both species
in the area.
Discussion

Current efforts for emerald ash borer eradication and control have not
been successful in slowing the spread of this destructive pest. High quality basic
knowledge on the biology of an introduced pest and the speed at which action is
taken will facilitate efforts to slow the spread of any invasive pest (Simberloff
2005).

Given the large native region of EAB including eastern Russia, Korea,
China, Japan and Taiwan, there could be important biological differences
between these populations that could impact the management strategy of EAB in
North America. The collection efforts of this study have yielded an extensive
collection of EAB from the native and introduced ranges. Although surveys
throughout much of the native range included regions in China, South Korea, and
Japan, samples were rare in South Korea and Japan, reducing the probability the
origin of the introduced population was from these two countries. This

hypothesis is further evaluated by genetic methods in the following chapters.

56

CHAPTER III
AFLP and mtDNA COI analysis of EAB Populations in Asia and North
America
Introduction

Michigan has 693 million ash trees (Fraxinus spp.) throughout its forests
and urban environments with an estimated value > $18.92 billion (USDA-APHIS
2003). Ash has a variety of economic values including nursery sales, timber,
sports equipment, tool handles as well as recreational use in state and federal
forests. For example, ash trees were one of the primary replacement trees in
urban landscaping to replace the American elms lost during the Dutch Elm
Disease epidemic in the latter half of the 20th century (MacFarlane and Meyer
2005).

The United States and Canada are currently battling a destructive invasive
pest of ash, the emerald ash borer (EAB), Agrilus planipennis Fairmaire, which is
native to Eastern Asia. Larvae of EAB destroy northeastern species of North
American ash trees by feeding below the bark on phloem tissue (Anulewicz et al.
2008). EAB develop within one to two years inside the tree until emerging as an
adult beginning in May creating a characteristic D-shaped exit hole in the bark.
High densities of larvae create feeding galleries that girdle the tree causing
nutrient flow to cease, causing death of the tree within as little as two years
(Russell et al. 2003).

EAB was detected in North America in 2002 (Haack et al. 2002), and was

initially found in counties of southeast Michigan and southeastern Ontario (Scarr

57

et al. 2002) and one northern county of Ohio in 2003. It is suspected that EAB
was introduced in the mid-1990’s and potentially may have been introduced in
the early-1990’s (Siegert et al. 2007). The known current range of EAB in the

United States includes 15 US states as well as several counties in Ontario and
Quebec, Canada (Chapter 2 Figure 2.1).

Invasive species such as EAB are an increasing problem as global trade
and travel increase (Simberloff 2005). A species can become invasive when it is
introduced into a new environment free from natural enemies and competitors,
becoming released from population constraints (Orians 1986, Sakai et al. 2001).
Even though the incidences of species introduction continue to rise, the
obstacles they must overcome to become established and invasive remain the
same. Invaders must locate a suitable habitat, resources and potential mates to
continue the population apart from the native range (Elton 1958). Introduced
populations may also have to overcome decreased genetic variation due to small
founding populations that could lead to inbreeding depression (Allendorf and
Lundquist 2003, Allendorf and Luikart 2007). However, decreased genetic
variation has been found to be beneficial, such as in Argentine ants with the
genetic stock of the introduced population having a competitive advantage over
native ant populations (T sutsui et al. 2000; Tsutsui and Case 2001). in some
cases, there could also be an increase in genetic variation if the introduced
population is an admixture from multiple introductions or a single introduction of
multiple source populations that are genetically distinct from each other

(Allendorf and Luikart 2007). This has been observed in a variety of invasive

58

organisms including the invasive weed, Verbascum thapsus (Dlugosch and
Parker 2008) and the lizard, Anolis segrei (Kolbe et al. 2004). In both examples,
the authors detected increased genetic variation in the introduced range
compared to the native range presumably due to multiple introductions.
Knowledge of genetic variation has important implications for development of an
effective management strategy for invasive species (Allendorf and Lundquist
2003, Sakai et al. 2001). Understanding the population biology of an invasive
species could provide information on suitable habitat, potential location(s) of
effective biological control agents, or help predict the effectiveness of a control
agent (Kambhampati and Rai 1991; Lee 2002; Allendorf and Lundquist 2003;
Schefffer and Grissell 2003,).

Molecular markers have been used to elucidate the origin and number of
introductions (Grapputo et al 2005), invasion history (Elderkin et al. 2004),
genetic variation of an introduced species (Roderick 1996; Sakai et al. 2001),
and to develop and evaluate control strategies (Szalanski and Owens 2003;
Conord et al. 2006; Kim et al. 2006). For example, mitochondrial DNA gene
sequencing has been used to detect genetic constraints caused by an
introduction event (Graputto et al. 2005). Mitochondrial DNA is haploid and
maternally inherited (exceptions: Wallis 2000; Andolfatto et al. 2003); therefore it
may reveal haplotype confirmation of geographic origin of invasive species, as
well as events of strong genetic drift (Avise 1994; Villablanca et al. 1998).
However, it may not be able to detect genetic variation on a recent time scale

(Grapputo et al 2005; Shapiro et al. 2008). Cognato et al. (2005) successfully

59

used the mtDNA cytochrome oxidase subunit l to determine the geographic
origin of the red turpentine beetle, Dendroctonus valens LeConte, that was
introduced into China. Genetic variation was assessed from individuals from 32
populations across the native range in North America and in the introduced range
in China. Haplotypes found in China were most closely related to haplotypes
found in the Pacific North-West of the United States. High genetic variation was
detected in the introduced range measured by haplotype diversity (0.94),
indicating the founding population may have been large or there have been
multiple introductions.

Genetic variation may also be assessed using nuclear markers exhibiting
higher variation than mtDNA due to biparental inheritance (Villablanca et al.
1998). One such marker is DNA fingerprinting by amplified fragment length
polymorphisms (AF LP); it is able provide a high number of polymorphic loci
throughout the whole genome of an individual without prior genetic knowledge of
the organism (Vos et al. 1995). This allows for rapid detection of polymorphism
in a population that is not possible with gene or DNA fragment sequencing
(Mueller and Wolfenbarger 1999). This technique has successfully provided
information about population variation including, the genetic variability of insect
populations across large geographic areas (Conord et al 2006; Clark et al. 2007),
identification of strains within the same species (Dalirsefat and Mirhoseini 2007),
genetic structure of an invasive species in an introduced range (Elderkin et al.
2004), and identification of geographic origin of invasive species (Reineke et al

1999; Benavides et al. 2005).

60

Although the examples above describe studies using one molecular
marker to elucidate population structure and variation, the addition of markers
with different inheritance patterns can provide a more complete understanding of
the population variation. For example, Salvato et al. (2002) used AFLP and
mtDNA markers to assess variation and population structure in two sibling
species of the winter pine processionary moth (Thaumetopoea pityocampa and
Th. wilkinsonr). Strong genetic differentiation between the two species as well
as geographic structuring of Th. pityocampa was detected with both markers.
AFLP and mtDNA markers were also successfully used to assess
phylogeographical patterns of Adelges cooleyi in Western N. America (Ahern et
al. 2009). Paupy et al. (2004) compared AFLP’s, microsatellite and isoenzyme
markers in Aedes aegypti from Phnom Pehn, Cambodia and determined the
effectiveness of these markers for inference of population structure. All markers
revealed the same population structure; however, AF LP analysis revealed up to
five times greater genetic differentiation than the other two. AF LP and mtDNA
markers have been used to estimate the geographic origin and population
structure of an invasive species. Grapputo et al. (2005) used both to assess the
variation in Colorado potato beetles in the native range of N. America and
introduced range in Europe. The European populations only had a portion of the
AFLP native variability and only a single mtDNA haplotype. Hence, they
concluded that the introduced populations undenrvent a bottleneck. These

examples provide evidence on the importance of assessing genetic variation and

61

structure with different marker systems to obtain a clear understanding and
corroboration of hypotheses.

To provide further evidence of population origin of invasive species,
population assignment tests may be used with molecular data to assign an
individual to the most likely population it originated (first described by Paetkau et
al. 1995). This method has been used to answer a variety of questions including:
identifying dispersal patterns in a metapopulation (Berry et al. 2004), confirming
population structure of native and/or invasive populations (Elderkin et al. 2004;
Paupy et al. 2004), assessing population structure to design eradication
strategies (Abdelkrim et al. 2005), estimating the origin of invasive species
(Bonizzoni et al. 2001; Genton et al. 2005; Tsutsui et al. 2001), or to assess if
individuals are from an established population or recently migrated to an area
(Kim et al. 2006). These tests are often performed using microsatellite data,
however, AFLPs are useful alternative for species where microsatellite markers
are unavailable or logistically difficult to obtain (Campbell et al. 2003).

The purpose of this study was to characterize the genetic population
structure of EAB in its native range in Asia and introduced range in N. America.
A primary goal was to identify as specifically as possible the geographic location
of the source EAB population(s) that gave rise to North America’s EAB
infestation. Although the presumed source of the infestation is China (based on
specific trade records), beetles may have originated from other Asian countries
within its native range. Population structure and variability of EAB was assessed

using AFLP and mtDNA COl markers. Understanding the genetic structuring in

62

the native and introduced range could be helpful to determine movement
patterns to focus management efforts (Sakai et al. 2001) and estimating the
success of introduced biological control agents or developing chemical controls
(Bourguet et al. 2000).
Materials and methods
Sample collection

Larvae, pupae and adult EAB were collected from introduced populations
in Illinois, Indiana, Maryland, Michigan, Ohio, Pennsylvania, West Virginia and
Windsor, Canada and from native populations in China, Japan, and South Korea
(Table 3.1). A network of collaborators from several government agencies and
academic institutions was established to assist in insect collection. Field
collected EAB were killed in 70% EtOH and shipped to Michigan State University.
Upon arrival, samples were transferred to 90% ethanol and stored in a -20°C
freezer until needed. Some samples from Michigan were reared to adult and
killed in liquid nitrogen. To do this, ash bolts, 60 cm sections of trunk or branch,
were obtained from many locations in Michigan for insect rearing in the USDA
Forest Service laboratory in East Lansing, MI. Bolts were cut in the field,
transferred to the laboratory and stored at 4°C until rearing of larvae into adults.
Rearing adult EAB was accomplished by placing individual bolts in a circular
cardboard container closed at both ends with a plastic cap, allowing light to enter,
and maintained at room temperature (approx. 20-24°C). Emerged adult EAB

were killed with liquid nitrogen.

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The aim was to collect at least ten individual beetles from each location.
However, this was not possible for all locations (e.g., South Korea, Japan, and
West Virginia, Table 3.1).

DNA isolation

Total genomic DNA was extracted from larval tissue (with the GI tract
removed to avoid potential contamination from gut contents) or from adult femur
muscle tissue using a QlAamp DNA Mini Kit (QIAGEN cat. no. 51304) as
described by the manufacturer with the following modifications. Larval tissue,
two abdominal segments, was combined with 20 uL protease K and 180 pL
QIAGEN ATL buffer in a 1.5 mL microcentrifuge tube and homogenized with a
pestle (1mL pipet tip closed by flame). If an adult hind-leg was used, the femur
was removed from the thorax and cut in half perpendicularly with a sterile razor
blade to expose tissue and then combined with the protease K and QIAGEN ATL
buffer. Tissue homogenates were incubated overnight at 56°C on a rocking
platform. DNA was eluted from the spin columns by applying 80 uL QIAGEN AE
elution buffer to the spin column, incubation at room temperature for 5 min, and
centrifugation at 8000 rpm for 1 min. Elution step was repeated for a total
volume of 160 pL. Samples were stored at -20°C. Quantification of DNA was
carried out on a subset of samples using a Nanodop system (Thermo Scientific).
Mitochondrial DNA gene sequence

The partial sequence of cytochrome oxidase | (COI) was amplified for all
EAB individuals in the sample (Table 3.1). Primers C1-J-1751 (5’-GGA TCA
CCT GAT ATA GCA TTC CC- ‘3) and C1-N-2191 (5’-CCC GGT AAA A'lT AAA

68

ATA TAA ACT TC-‘3), desicribed by Simon et al. 1994, were used to amplify a
partial COI nucleotide sequence of approx. 450 bp. PCR reactions (25pL)
contained 9.3 pL ddH20, 2 uL 10X Taq polymerase buffer (lnvitrogen), 1.6 uL 50
mM MgClz (lnvitrogen), 0.4 uL Taq DNA polymerase (lnvitrogen 5 units/pL), 2 pL
10mM each deoxynucleotide triphosphates (dATP, d'l'l'P, dCTP, and dGTP), 1.5
uL of 10 pmol oligonucleotide fonivard and reverse primer, and 2.7 pL sample
DNA. Reactions were carried out on a PE9700 Thermal Cycler with the following
protocol: 5 min at 95°C, 35 cycles of 120 sec at 94°C, 90 sec at 52°C, and 120
sec at 72°C, followed by a final extension for 7 min at 72°C. PCR products were
purified using QIAGEN MinElute PCR Purification Kit (cat no. 28004) to remove
unincorporated primers and nucleotides. Purified product (5 0L) was combined
with 4 pL ddH20 and 3 pL of 10 umol primer and sent to the MSU Genomics
Technology Support Facility for forward and reverse sequencing. Alignment of
obtained sequences was performed using Squd Ver 1.0.3 (Applied
Biosystems).

Phylogenetic analysis of mitochondrial COI haplotypes was carried out by
parsimony analysis of sequence data using PAUP*4 (Swofford 2000). A heuristic
search was performed using tree-bisection-reconnection branch swapping, equal
weighting of characters, and random addition of replicates.

Haplotype relationships were also analyzed using statistical parsimony
with the computer program TCS program 1.21 (Clement et al. 2000). This
analysis allows the possibility the ancestral population is observed in the sample

set and potentially the most frequently observed. Haplotype relationships from

69

TCS were displayed as a network. Haplotype diversity was also determined
using the TCS program.

Analysis of molecular variance (AMOVA) using ARLEQUIN 3.0 (Excoffier
et al. 2005) was used to determine if population structure between the collection
locations in Asia and N. America was observed using the mtDNA haplotypes.
Within- and between- population variation was estimated considering the total
EAB samples as 11 populations (seven populations in China, two in South
Korea, one in Japan, one in Russia and 230 individuals pooled together in one
population in N. America). (bst in pairwise population comparisons were
considered to be significantly different at the p = 0.05 level.

Amplified Fragment Length Polymorphism (AFLP)

AF LP plant mapping protocol (Applied Biosystems Part #402083) was
used to generate fragments for fingerprinting of individual EAB. Restriction
fragments were generated by digesting genomic DNA with EcoRI and Msel
restriction endonucleases and ligating the fragments to specific EcoRI and Msel
adaptors. This created a large number of fragments used for subsequent
selective amplification. An enzyme master mix was prepared by combining 0.1
pL 10X T4 ligase buffer, 0.1 uL 0.5 M NaCl, 0.05 uL 1 mg/mL bovine serum
albumin (BSA), 1 unit Msel, 5 units EcoRl, 1 unit T4 DNA ligase, and brought to a
total volume of 1 uL with ddeo. The restriction/ligation master mix was
prepared by combining 1 pL 10X T4 ligase buffer, 1 pL 0.5 M NaCl, 0.5 pL 1
mg/mL BSA, 1 uL Msel adaptor, 1 uL EcoRl adaptor, and 1 pL enzyme master

mix. Reactions were carried out by combining 5.5 uL genomic DNA with 5.5 uL

70

master mix and incubating for 2.5 hr at 37°C. The product of this digestion was
diluted with 89 uL of TEo.1 buffer (20 mM Tris-HCI, 0.1 mM EDTA, pH 8.0).
Reduction of the number of fragments was accomplished by PCR amplification
with pre-selective primers; a Msel complementary primer (Msel adaptor with the
recognition site and 3'C) and an EcoRI complementary primer (EcoRl adaptor
with recognition site and 3’A). Solutions were prepared by adding 4 0L dilute
restriction/ligation product to 1 pL pre-selective primer pairs and 15 uL core mix
(Applied Biosystems). Pre-amplification solutions were placed in a Applied
Biosystems GeneAmp PCR System 9700 thermal cycler with the following
program: one cycle of 72°C for 2 min; 20 cycles of 94°C for 1 sec, 56°C for 30
sec, and 72°C for 2 min; with a final extension step of 60°C for 30 min . Pre-
amplification products were diluted with 89 pL of TEo,1 buffer for final selective
amplification. Four primer pair combinations used for selective amplification
were run on each EAB sample (EcoRl-AGG/Msel-CTT, EcoRl-ACT/Msel-CAG,
EcoRl-ACT/MseI-C‘l‘l', EcoRl-AGG/Msel-CAG. Selective amplification solutions
were prepared by combining 3.0 uL diluted pre-selective amplification product,
1.0 uL 5uM Msel primer, 1.0 uL 1pM EcoRl primer (fluorescent labeled), and
15.0 pL AF LP core mix and run in the above thermal cycler with the following
program: one cycle of 94°C for 2 min; 10 cycles of 94°C for 20 sec, 66°C for 30
sec, 72°C for 2 min decreasing by one degree at the second step each cycle; 20
cycles of 94°C for 20 sec, 56°C for 30 sec, and 72°C for 2 min; with a final
extension step of 60°C for 30 min. Selective amplification PCR products were

submitted to the MSU Genomics Technology Support Facility for electrophoretic

71

separation of DNA fragments on an Applied Biosystems ABI PRISM ® 3100
Genetic Analyzer, and analyzed using Applied Biosystems GeneScan Analysis
Software.

Profiles were observed with GeneScan 3.1 software (ABI Applied
Biosystems) for scoring of fragments between 50-500 bp in length. Scoring
fragments smaller than 50 bp increases the likelihood of homoplasy (Ahern et al.
2009) and therefore these were eliminated from the analysis. Fragments greater
than 500 bp were not observed in any profile. Fragments (peaks) were coded
manually as present (1) or absent (0) for each individual locus and compiled in a
Microsoft Excel spreadsheet for subsequent analysis. Loci were scored as
present if the peak height was greater than 100 (Ahern et al. 2009).

Descriptive statistics, including the percent polymorphic loci and average
heterozygosity for each population, as well as pairwise genetic differentiation (Fst)
were obtained using AFLP-SURV (Vekemans et al. 2002). A Bayesian method
of analysis with non-unifonn prior distribution of allele frequencies and assumed
Hardy-Weinberg equilibrium was used to determine allele frequencies. The
proportion of polymorphic loci and expected heterozygosities were calculated for
each population (Lynch and Milligan 1994). Higher values of polymorphism and
expected heterozygosity are expected in Asian populations than N. American
populations if the introduced population was the result of a single introduction or
several small introductions from the same geographic source. Overall genetic
differentiation, Fst, was calculated with 500 permutations to determine

differentiation among populations at the 1% level (Lynch and Milligan 1994). A

72

high value of overall Fst (higher than 0.15 (Wright 1951; Connor and Hartl 2004)
is considered to indicate strong differentiation among populations. Painivise
genetic differentiation (Fst) between each population was calculated to determine
which populations were significantly differentiated from each other. Low pain/vise
Fst values (close to zero) were interpreted to mean that two populations are either
not differentiated from each other or the loci used were unable to detect
differentiation.

Significance of pairwise (bpt values and population structure were analyzed
by AMOVA (999 permutations) using GenAlEx 6.1 (Peakall and Smouse 2006)
with presence/absence data in a Microsoft Excel file. The analysis was
performed for Asian and N. American collection locations with each collection
location defined as a population to determine proportion and significance of
variation within and between populations. Mantel tests were also performed to
determine if a pattern of isolation-by-distance model of population structure is
observed in the native and introduced range.

The number of distinct genetic groups (K) throughout the native range
given no a priori information was estimated using STRUCTURE 2.2 (Pritchard et
al. 2000). This program uses Bayesian algorithms to determine the proportion of
genetic variation in each sample from the estimated number of populations. The
admixture ancestry model was used for a different number of parent populations
(K = 1-8) and each run was conducted with a burn-in value of 50000 and 450000
iterations for data collection. The value of K with the lowest posterior probability

is considered the probable number of parent populations (K) for a given data set.

73

Genetic similarity among EAB populations were displayed via neighbor-
joining (NJ) analysis using pairwise <Dpt distances with MEGA 4 (T amura et al.
2007). Results from neighbor-joining analysis and structure analysis were
compared to determine if similar structuring was observed.

Population assignment tests were performed to determine the geographic
origin of N. American EAB samples from Asian populations using the AFLPOP
version 1.1 program (Duchesne and Bernatchez 2002). The program does not
permit missing data; therefore only individuals with all loci scored for all primer
pair combinations were entered for analysis.

Results
mtDNA variability

Partial COI gene sequencing of 326 EAB from 15 collection locations in
Asia and 37 locations in N. America (Table 3.1) resulted in nine different
haplotypes (Table 3.2; Appendix 1). Seven haplotypes differed from the common
haplotype (N=316) by a single bp change, one haplotype by two bp, with the final
haplotype differing by 21 bp changes (Shirioshi City, Japan). Twenty-nine
nucleotide changes occurred at 25 nucleotide sites along the 481 bp length.
Third codon position changes accounted for the majority of differences (96%)
with a single change at the first codon position. All changes were synonymous
with respect to amino acid translation.

Parsimony analysis of the nine haplotyes based on 1 parsimony

informative character, 24 variable uninformative characters, and 456 constant

74

 

 

 

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75

characters yielded 6 most parsimonious trees which had < 50% bootstrap values.
The strict consensus of these six trees was unresolved (Figure 3.1).

Statistical parsimony network analysis yielded two groups (Figure 3.2). All
samples in N. America and the majority in China had the same haplotype, H1
(Table 3.2, Figure 3.2). Three other haplotypes were observed in China; two
haplotypes in Benxi, H2 and H3, and the third haplotype in Shenyang, H4. The
common haplotype, H1, was observed in South Korea with five other haplotypes,
H4-H8. A single sample from Shirioshi City, Japan differed from the common
haplotype by 3.7% and separated on its own into the second group (Figure 3.2).

Genetic structuring of populations was assessed by AMOVA. Collection
locations in Asia were pooled into a single population it they were within 50 km of
each other (Tangshan and Zhenzhu zhenstulangai; Changchun City and Jiutai;
Shenyang and Yushutan; Samcheok and Inge). Significant population
structuring was detected between the six populations in China with 32% of the
variation explained between populations (Fst = 0.3207, P < 0.001). When the
analysis was expanded to the nine total populations in Asia (China, S. Korea,
and Japan), 77% of the variation was explained between populations (F5, =
0.7716, P < 0.001) (Table 3.3). Structuring the populations into 3 groups
according to country revealed significant structuring with 90% of the variation
explained between countries (Fct = 0.9045, P < 0.005). When the introduced
populations from N. America and Moscow, Russia were added to the into
AMOVA analysis for a total of five groups by country detected 74% of the

variation due to country grouping (Fct = 0.7424, P < 0.001) and 13% of the

76

Main Haplotype (H1)
‘—- Daejeon, South Korea (H5)

1 Daejeon, South Korea (H4)
Shenyang, China

 

 

2
Daejeon, South Korea
(H8)
1
—— Inge, South Korea (H6)
1
Inge, South Korea (H7)
21 1

—. Benxi, China (H2)

 

—— Benxi, China (H3)

 

 

Shirioshi City, Japan

Figure 3.1. Strict consensus of six parsimonious trees for 9 emerald ash borer
haplotypes. The main haplotype is a single individual representing 316 samples that
share this haplotype collected in China, South Korea, North America, and Moscow,
Russia.

77

 

94
418 _ _®
463

403 a
103

H1

Figure 3.2. Haplotype network for 326 Emerald Ash Borer individuals collected in Asia
and North America based on partial sequences (481 bp) of the mitochondrial cytochrome
oxidase I gene. Numbers on the lines between haplotypes represent single nucleotide
mutations

78

 

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79

variation between populations within groups. The subdivision was due to two
populations in China (Benxi and Shenyang), and both populations in S. Korea
and Japan. Pairwise Fst analysis did not detect significant differentiation between
the majority of populations (Table 3.4).
AFLP variability

AF LP analysis was conducted with four primer pair combinations yielding
273 polymorphic loci scored for 109 samples from 10 locations in Asia and 184
samples in 11 locations in N. America (Table 3.1). Two loci were unique to the
Japan samples and one locus was unique to the South Korea samples. Of the
eight collection locations in China, polymorphism level (% polymorphic loci) was
detected highest in Shenyang (64.8%) and lowest in Benxi (35.9%) while
estimated heterozygosity ranged from 0.136 in Benxi to 0.265 in Harbin (Table
3.5). In South Korea, polymorphism level and estimated heterozygosity was
47.6% and 0.169 respectively. Two individuals sampled in Japan recorded 100%
and 0.376 respectively. In the introduced range in N. America, polymorphism
was lowest in West Virginia (33.7%) to 53.5% in Petoskey, Michigan. Expected
heterozygosity ranged from 0.095 to 0.194 in West Virginia and Canada,
respectively (Table 3.5). Difference in estimated heterozygosity was marginally
significant between Asian and N. American locations (Mann-Whitney U test P=
0.043), however, when the value from Japan was removed they are not
significantly different (Mann-Whitney U test P= 0.080). Although the mean level
of polymorphism was slightly higher in Asia (55.92% with Japan; 51.02% without

Japan) than North America (46.34%), it was not significantly different (Mann-

80

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Table 3.5. Descriptive statistics for Asian and North American EAB collection locations
with (N) number of individuals using 273 total AF LP loci.

 

Population (N) % polymorphic loci (p) Expected heterozygosity (H) i SE

 

Asia
China
Dagong 25 64.1 0.19i0.01
Tangshan 22 61.9 0.18i0.01
Beijing 2 36.6 0.19:1:0.01
Benxi 6 35.9 0.14:1:001
Shenyang 13 64.8 0.17i0.01
Jilin 6 45.4 0.17zt0.01
Jiutai 20 40.3 0.15i0.01
Harbin 5 62.6 0.26i0.01
South Korea 8 47.6 0.17i0.01
Japan 2 100 0.38i0.01
North America
Canada 26 61.2 0.19:1:001
Michigan
Plymouth 13 40.3 0.141001
Livonia 2] 48.7 0.15i0.01
Lansing 29 45.1 0.152i:0.01
Petoskey 10 53.5 0.15i0.01
Moran 21 49.1 0.151001
Ohio 12 52.0 0.19:1:001
Indiana 16 39.6 0.14i0.01
Illinois 14 41.4 0.16:1:001
Pennsylvania 17 45 .1 0.16i0.01
West Virginia 5 33 .7 0.09i0.01

 

82

Whitney U test P= 0.251 with Japan; Mann-Whitney U test P= 0.412 without
Japan). Overall (Dpt was 0.094 (P=0.001), signifying slight differentiation
between collection locations.

AMOVA detected genetic differentiation between the populations from the
native and introduced ranges with 5% of the total variation attributed to variation
between country (P = 0.001) and 7% attributed to variation within country (P =
0.001). When the analysis was conducted without regional information to
determine the level of differentiation between the 10 Asian populations and 11 N.
American populations, 9% of the differentiation was attributed to among
population variation (P = 0.001). The amount of variation was higher among N.
America populations than among Asian populations when considered separately
((Dst = 0.083, P = 0.001 and Cbst = 0.046, P = 0.002, respectively) (Table 3.3).
Painlvise (Dpt among Asian populations did not detect any differentiation between
several populations ((Dpt = 0.000) with the highest differentiation between Japan
and Jilin, China ((Dpt = 0.461) (Table 3.6). Pairwise (Dpt among N. American
populations did not detect any differentiation between Cranberry, Pennsylvania
and Peru, Illinois (CDpt = 0.000) with the highest differentiation between Fayette
00., West Virginia and Petoskey, Michigan ((Dpt = 0.236) (Table 3.7). Painlvise
(Dpt between Asian and N. American populations ranged from no differentiation
detected between several population comparisons ((Dpt = 0.000) with the highest
differentiation between Fayette Go, West Virginia and Beijing, China ((Dpt =

0.624) (Table 3.8).

83

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86

Neighbor-joining analysis based on pairwise (Dpt of Asian populations did
not reveal clear differentiation between populations based on geographic
separation (Figure 3.3a). Analysis of N. American populations also did not
divulge clear differentiation between populations based on geographic separation
expected if introduced populations were isolated by distance (Figure 3.36).
Analysis of all populations revealed genetic similarity associated with geographic
proximity (Figure 3.4).

Structure analysis of Asian samples determined the most probable
hypothesis for the 10 collection locations is the genetic data developed from six
populations (Table 3.9, Figure 3.5). Structure was not realized because there
were no unique AFLP patterns to specific populations. Samples from Dagong,
China exhibit the most variation among the collection locations (Figure 3.5).
Clear separation of geographic regions was not possible given the level of
admixture from each of the collection locations (Figure 3.5).

A Mantel test conducted on Asian populations revealed a marginally
significant correlation between genetic and geographic distances (r = 0.772, P =
0.026). However, this relationship was not significant when the Japan population
was removed from the analysis (r = 0.234, P = 0.190). N. American populations
also showed a marginally significant correlation between genetic and geographic
distances (r= 0.471, P = 0.029).

Population assignment tests were performed to assess the most likely
hypothesis for the geographic origin of N. American individuals using the 10

Asian collection locations as potential sources. Individuals from N. American

87

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

I Jmn
‘ Tangshan
Benxi
Dagong
Hadfin
Bemng
Korea
~—-——Junm
Shenyang
Japan
W
B.
MHst
1;...
Fkioskey
C110
(Errata
hkxan
'——l —-—-——-———-lJv31a
r—-—————-——-kxiana
VNNfinjrna
Lansing
Fflynxxlh
"076"

Figure 3.3. Neighbor-joining analysis of EAB AFLP data set based on pairwise <1th
values for A) native range in Asia and B) introduced range in North America.

88

[— Indiana

 

 

 

 

 

 

 

 

     
  
 
     

 

  
 

 

 

 

 

 

VVVthfia
—-Lhonm
Momn
Lanflng
—l Plymouth
———-Ohk)
Canada
Petoskey
Penn
lmnoo
Jmn
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Bemng
Hanfin
Dagong
Korea
Juna
Shenyang
Japan
W

Figure 3.4. Neighbor-joining analysis of EAB AF LP data set based on pairwise (Dpt
values for the native range in Asia and introduced range in North America.

89

 

Figure 3.5. Structure analysis results representing the most probable number of
populations of EAB, K: 4 through K=7 in Asia. Shades are assigned
automatically by the program with a separate color for each inferred population
with the proportion of the samples variation on the y-axis. Collection locations,
denoted on the x-axis, 1-8 were obtained in China (Population 1-Dagong, 2-
Tangshan, 3-Beijing, 4-Benxi, 5-Shenyang, 6-Jilin, 7—Juitai, 8-Harbin), location 9
in South Korea and location 10 in Japan.

90

 

91

Table 3.9. Structure analysis results. AF LP samples collected from Asia were used to
test the hypotheses samples originated from varying number of populations (K) by
Bayesian analysis. Data presented in log likelihood with the lowest score representing
the highest posterior probability number of populations.

 

 

Population number (K) Log likelihood
2 48741 .5
3 -47080.0
4 -46457.3
5 -55289.2
6 —45888.1
7 -45894.6
8 -46872.6

 

92

populations were assigned to Tangshan, China with the most frequency ranging
from 12.5% in Petoskey, Mi to 80% assignment in W. Virginia with a mean
percentage of assignment of 45.51 % (T able 3.10). Individuals were also
assigned to Dagong, China with a frequency as high as 53.3% from Windsor,
Canada with a mean assignment of 22.26% (Table 3.10).

Discussion

There was very little COI variation detected in most locations in China as
a single haplotype was found in the majority of locations and only three
haplotypes detected in two locations that differed from the main haplotype
observed (Table 3.2). The majority of unique haplotypes were found in locations
in South Korea with these locations showing higher haplotype diversity compared
to China with a final haplotype detected in Japan. Since there were few samples
from locations where unique haplotypes were observed, ranging from one
individual in Japan to 13 in Shenyang, China, it is possible the haplotype diversity
in these locations are an underestimate of the true diversity of the location.

A reduction in mitochondrial variation is common in introduced insect
population studies (Kambhampati and Rai 1991; Villeblanca et al. 1998; Tsutsui
et al. 2000; Downie 2002; Scheffer and Grissell 2003; Corin et al. 2007). For
example, the Colorado potato beetle (Leptinotarsa decemlineata Say), had 20
haplotypes detected throughout the native range in N. America compared to a
single haplotype found in the introduced range in Europe (GrapUtto et al. 2005).
Another example of reduced mitochondrial variation due to a founder event was

found in a moth pest of cacao, Conopomorpha cramerella Snellen, in the Malay

93

 

 

 

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94

Archipelago. Shapiro et al. (2008) found only six COI haplotypes across
widespread islands concluding the reduction of variation compared to similar
studies (Juan et al. 1998; Mun et al. 2003) was very low and the result of a
bottleneck event. This pattern of reduction of COI variation is consistent with the
data found in this study for introduced EAB. All samples throughout the
introduced range of USA and Canada shared the single main haplotype found in
Asia. Since the main COI haplotype is found throughout EAB’s native range, this
could be explained by either a unique introduction event with an unknown
number of individuals into N. America or multiple introductions of the sample
haplotype (Sakai et al. 2001; Graputto et al. 2005).

Given the only COI haplotype found in N. America is found throughout
most of the native range in Asia, added genetic information was needed to
characterize the population structure of EAB. AF LP analysis has been a
powerful tool to estimate population genetic relationships of insects (Meng et al.
1996; Reineke et al. 1999; Ravel et al. 2001; Elderkin et al. 2004; Conord et al.
2006; Dalirsefat and Mirhoseini 2007; Clark et al. 2007). Reduced genetic
variation was weakly confirmed by AF LP analysis, however the reduction was
only marginally significant by expected heterozygosity and not statistically
significant by % polymorphic loci, demonstrating the introduced populations have
retained genetic variation with nuclear loci (Sakai et al. 2001). Genetic variation
retained by nuclear loci but reduced in mtDNA could have been caused by the
unique introduction of a large founding population or multiple introductions

(Allendorf and Lundquist 2003; Grapputo et al. 2005; Ahern et al. 2009). Similar

95

 

results have been found in other invasive pests with reduced genetic variation in
the introduced compared to native ranges. For example, the Argentine ant had
significant reduction of genetic variation was detected in introduced populations
in N. America, Australia, Chile, Italy, and South Africa compared to the native
range in Brazil and Argentina (T sutsui and Case 2001; Tsutsui et al. 2001). The
Eurasian spiny waterflea also demonstrated evidence of reduced genetic
variation in the introduced range in N. America compared to the historical range
in Europe (Colautti et al. 2005). However, reduction of genetic variation is not
always observed in all invasive pests. For example, zebra and quagga mussel
populations introduced into N. America showed similar genetic variability to
populations in the native range of Eurasia and did not show evidence of a
bottleneck event during introduction (Stepien et al. 2002).

Weak population structure was detected using AF LP in both native and
introduced populations of EAB. This structure was not consistent with an
isolation-by-distance hypothesis characteristic to natural spread after initial
introduction and more indicative of human-mediated transport (Sakai et al. 2001).
In addition, greater genetic differentiation was detected in the introduced
populations (03st = 0.083 vs. <I>st = 0.046) suggesting multiple introductions into
N. America (Sakai et al. 2001). Average pairwise (Dpt was lower in the native
range (0.089) compared to the introduced range (0.105) meaning the Asian
populations are more similar to each other than the N. American populations are
to each other (Mann-Whitney U test P= 0.023). This could have several

interpretations including 1) the introduced population(s) was the result of multiple

96

 

 

introductions (Sakai et al. 2001; Allendorf and Lundquist 2003; Kolbe et al. 2004),
2) the sampling in the native range does not represent the full variation and
differentiation between locations (Kim and Sappington 2005), 3) there is more
gene flow between populations in the native range than in the introduced range,
4) population comparisons in the introduced range are caused by incomplete
sorting of nuclear variation (Hartl and Clark 1997; Ahern et al. 2009), or 5)
introduced populations have developed local adaptations causing populations to
diverge (highly unlikely given EAB was introduced into N. America within the last
20 years). Incorporation of additional samples throughout the native and
introduced ranges and/or additional genetic markers are needed to differentiate
between these hypotheses.

Determining the geographic origin of an invasive insect can be challenging
(Roderick and Navajas 2003, Cognato et al. 2005, Shapiro et al. 2008). Both
mtDNA and AFLP analysis suggest the most likely origin of EAB is from a region
of China, more specifically near Dagong and Tangshan. Even with small sample
sizes, the genetic diversity detected in a couple populations in China and S.
Korea was not observed in the introduced range while the main haplotype was
fixed in higher sample sizes in Dagong and Tangshan. Although pairwise (1th
between locations in Asia and N. America indicated significant differentiation
between several locations in the introduced range to Dagong and Tangshan
(Table 3.8), the average pairwise (Dpt across all populations in N. America
revealed the lowest population differentiation between these two populations

((Dpt = 0.0877 and 0.0848, respectively) with all other average pairwise (Dpt

97

values ranging from 0.1044 against Shenyang, China to 0.3818 against Japan.

Assignment tests of individual beetles from N. America were assigned
over 67% of individuals were assigned to either of these two collection locations,
while there were no individuals assigned to the location in Shiroishi City, Japan
or any location from South Korea. More extensive sampling to include more
individuals from known source populations, incorporation of samples from
locations in the native range not included in this study (for example Far eastern
Russia), as well as including more samples across the current introduced range
of EAB may help strengthen'the assignment strength of individuals to the most
likely source population.

In conclusion, this is the first study to assess genetic variation of EAB
populations in its native and introduced range. Molecular analysis of an invasive
species is essential to understand the geographic origin, population structure,
and develop management strategies. Given eradication is most successful
before a species has become established and widespread (Ruesink et al. 1995;
Allendorf and Lundquist 2003; Perrings et al. 2005; Simberloff 2005), it is unlikely
EAB can be eradicated due to its widespread distribution in N. America.
Therefore, increased understanding of population structure, estimating direction
and mode of dispersal, and determining the origin can be important (Grapputo et
al. 2005). The use of mtDNA and AFLP markers was useful in assessing the
population structure in Asia and N. America, as well as estimating the geographic
origin of the introduced population. This study demonstrated weak population

structure in both Asia and N. America; however this structure was not

98

characteristic of population separated by geographic distance. In addition, the
majority of the AFLP genetic variability is within individual populations and not
among populations, possibly suggestive of minor gene flow (Clark et al. 2007), or
human-mediated transport. Stronger restrictions on long-distance transport of
ash material should be incorporated into any management strategy. The
reduction in mtDNA and AFLP variability and estimated geographic origin of
Dagong and Tangshan, China provides evidence that the introduced population
should be susceptible to the biological control agents from that region (Sakai et

al. 2001; Allendorf and Lundquist 2003).

99

CHAPTER IV
MICROSATELLITE ANALYSIS OF EMERALD ASH BORER

Introduction

Emerald ash borer (EAB), Agn'lus planipennis Fairmaire, is a serious
exotic pest of North American ash trees (Fraxinus spp.) that was first discovered
outside of its native range of Northeastern Asia in 2002 (Haack et al. 2002).
Since its initial discovery in Michigan and Ontario, it has spread to Illinois,
Indiana, Kentucky, Maryland, Minnesota, Missouri, New York, Ohio,
Pennsylvania, Virginia, West Virginia, Wisconsin, and Quebec (Cooperative
Emerald Ash Borer Project 2009). Greater than seven billion ash trees are in the
United States with > 1.13 billion located in Michigan, Ohio and Indiana (Liu et al.
2003). The economic value of these trees is considerable, with an estimated
value of $18.9 billion in Michigan alone (USDA-AHPIS 2003). Since the most
likely introduction into N. America was in the early to mid 1990’s based on
dendrochronology (Siegert et al. 2007) the spread of EAB has almost certainly
been expedited by unintentional human movement of infested ash material
(USDA APHIS 2006). Information on EAB was limited before the introduction
into N. America (Liu et al. 2003; Wei et al. 2004) because it was not considered a
pest in the native range of Northeastern China, Far eastern Russia, Japan,
Korea, Mongolia, and Taiwan (Chinese Academy of Science 1986; Yu 1992).
EAB host range in Asia includes several Fraxinus spp. (Chinese Academy of
Science 1986; Yu 1992) as well as Jug/ans mandshun'ca var. sieboldiana, J.

mandshun'ca var. sachalinensis, Pterocarya moifolia and Ulmus davidiana var.

100

 

 

 

japonica in Japan (Akiyama and Ohmomo 2000; Sugiura 1999). EAB is not
considered a pest species in its home range, however; it has caused serious
damage to imported N. American Fraxinus spp. plantations in China causing
many to be removed (Liu et al. 2003; Wei et al. 2004).

Molecular markers, in general, are useful in understanding the introduction
and expansion of invasive species (Roderick 1996; Sakai et al. 2001). More
specifically, microsatellite markers have been useful in studying invasive species,
notably in determining their geographic origins (T sutsui et al. 2001; Fonseca et
al. 2004; Scribner et al. 2003; Kim et al. 2006; Caldera et al. 2008); identifying
bottlenecks and gene flow (T arr et al. 1998; Colautti et al. 2005); estimating the
size and number of introductions (Bonizzoni et al. 2001; Walker et al. 2003; Miller
et al. 2005); and evaluating genetic population structure and dispersal (Vargo
2003a,b; Kim and Sappington 2005; Herborg et al 2007). Despite EAB’s
ecological and economic impact in N. America, relatively little is known about its
invasion history and geographic origin. In this chapter, I describe research to
isolate, identify, and characterize microsatellite markers to better understand the
demographics of EAB in its native Asian range and its introduced range in N.
America.

Materials and methods
Sample collection and DNA isolation

EAB were collected from native populations in China, South Korea and

Japan and from introduced populations in N. America (Table 4.1) with the

assistance of a network of collaborators from several government agencies and

101

Table 4.1. Locality Information for EAB Individuals Included in the

Microsatellite Analysis

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Country State/Province County Locality Pop # (*) # ind (N)
China Hebei Tangshan 1 19
China Jilin Chanchun 2 6
China Jilin Jiutai 2 15
China Liaoning Benxi 3 5
China Liaoning Shenyang 3 8
China Tianjin Daggg 4 8
China Beijing Beijirg 4 2
Russia Moscow 2
Chungchong-nam
South Korea do Daejeon 1
South Korea Gyeongg-do Suwon 1
USA Illinois Peru 5 13
USA Indiana LaGrange Shipshewana 6
USA Maryland Prince Georg's 3
USA Michigan Chippewa Brimley 4
USA Michigan Emmet Petoskey 3
USA Michigan lngham East Lansing 6 2
USA Michiggn lngham Lansing 6 12
USA Michigan Livingston Brighton 7 1
USA Michigan Livingston Howell 7 1
USA Michigan Livingston Pinckney 7 3
USA Michigan Oakland Bloomfield 7 1
USA Michigan Washtenaw Ann Arbor 7 6
USA Michigan Washtenaw Dexter 7 2
USA Michigan Wayne Belleville 7 1
USA Miclggan Wayne Westland 7 1
USA Michigan Mackinac Moran 8 10
Fulton and
USA Ohio Wood 5
USA Pennsylvania Butler Cranberry 9 15
Canada Ontario Essex Windsor 10 6
Total 162
Notes-

" Population number in Fstat Analysis; some localities pooled

102

 

 

academic institutions. Samples were collected in the field and then stored in 90-
100% ethanol and sent to the lab for analysis. Total genomic DNA was extracted
from larval tissue (with the GI tract removed to avoid potential contamination from
gut contents) or from adult femur muscle tissue using a DNA mini-kit (Qiagen cat.
No 51304) as described by the manufacturer with minor modifications.
Populations with few individuals and primers which did not amplify DNA in all

samples were excluded from statistical analysis in this report due to low

 

statistical power. However, these data are available for future studies.
Microsatellite marker development

Ninety-six plasmid clones were developed in collaboration with Dr. Travis
Glenn at the University of Georgia's Savannah River Ecology Laboratory (Glenn
and Schable 2005). In brief, extracted DNA from four larvae (Lansing, MI) was
used for enrichment of di-, tri-, and tetra nucleotide microsatellite motifs and
development of clones of DNA region. Sequences of clones were forwarded to
me for primer design and optimization. Primers were designed using the
program PRIMER (vers. 3.0; Rozen & Skaletsky 2000), with ideal primers being
18-26 base pairs long with 45-50% GC content. Test amplifications at each
locus were performed on a PE9700 Thermal Cycler (Applied Biosystems) using
unlabeled primers to amplify alleles from eight individuals from two collection
locations in Michigan (Lansing and Belleville). In each case, varying MgClz
concentrations and annealing temperatures were tested to determine optimal
PCR conditions. PCR products were visualized on 1.5% agarose gels. Primers

designed to amplify alleles at 41 microsatellite loci yielded successful PCR

103

 

amplification in the test samples. Twenty-five of these were selected as
candidates to detect variation in EAB populations; the remaining 16 loci were
excluded because the primer pair either yielded more than two PCR products or
the amplified product was not in the expected size range (the primer pair
amplified a different region of the genome). The forward primer for each of the
25 candidate loci was 5’ fluorescence-labeled and PCR was performed at
optimum conditions developed for each specific primer pair using DNA templates
from a test group of 16 individuals collected throughout the native range in China
to assess polymorphism among these loci. PCR products from these reactions
were analyzed using an ABI 3130 capillary DNA analyzer at the Research
Technology Support Facility (RTSF) at Michigan State University,

At the time this dissertation research, another researcher Dr. Jenny Cory
(Simon Fraser University, British Columbia, Canada) produced her own set of
candidate microsatellite loci for EAB. A collaboration was established with Dr.
Cory to develop and test each others candidate loci to assess for variation and
increase the likelihood of success in locating polymorphic loci for EAB. This
study screened four of these loci for variation within the test population of Asian
samples and found that primer pairs at two loci (C-CS and C-C8) were
polymorphic in test samples and used for further characterization of EAB genetic
variation in 162 individuals from Asia and N. America.

Data analysis
Microsatellite allele data were analyzed first using the program MSA

(Microsatellite Analyzer, vers. 4.05; Dieringer and Schlotterer 2003), which

104

 

 

allowed determination of basic population statistics, including observed number
of alleles and average observed and expected heterozygosity. FSTAT (Goudet
2001) was used to estimate EAB population differentiation between four regions
in China and six regions in N. America (Table 4.1). Population differentiation was
measured using the parameter, FST, which measures the reduction of
heterozygotes from the entire dataset due to differentiation among
subpopulations; population differentiation increases as FST increases from zero.
Results and Discussion

Forty-one microsatellite loci were successfully amplified in EAB, however,
none were polymorphic with respect to a test group of 16 Asian individuals. None
of the primer pairs at the 25 candidate loci yielded polymorphism in our test
sample that would allow the assessment of variation within and between
populations in Asia. Loci yielded a limited number of alleles (1-3) or did not
amplify in all Asian test samples. Therefore, variation in EAB was based on data
from the two loci, C-CS and C-C8 (developed at Dr. Cory’s laboratory), and a
study sample of 162 individuals drawn from 24 localities in Asia and N. America
(Table 4.1). The observed and effective number of alleles was eight and seven
alleles and 6.1678 and 4.6231 over all populations, respectively (C-CS and C-C8;
Table 4.2). The expected and observed heterozygosity, number of alleles and
allelic richness per locus per population are given in Table 4.3.

FST values informed structure among N. American and Chinese
populations. Most interesting, the population from Jilin Province (Chanchun and

Jiutai) appears most dissimilar to each of the six N. American populations (Table

105

 

00.0..0
0260.0 .0 .00E0: u m.z 00.0..0 .o .00E0: u 02 5.0003920: 00.0090 u mf ”3.0003290; 82008 u 0:

 

0<<0FF0<F0F00F<0<<0<00 ”m

FmNo... . 300000000 N0F 03-0.0 5.3.0.00... 0<0C0<FE0<<FF<00F00 H. 00-0
0F0F00EF00<E00F0 ”m.

0.80 0 483.023 N0F .0702 0.00.... 0F00<0<<<0<<00200F H. 00-0

 

02 oz 01.0... 9.038000. 00.00 ..0-.0.8F_m=820e_.n. 88..
0.0E0w 0.00 0.0.3

 

000:00
.0.0E:.oo 00....0 .>..0.0>_:D .000.n. :0E.m. boo >::0.. .0 >0 0000_0>00 0.02, .00.. 0:0..03000 :00_.0E< 0:02
0:0 :0_0< Eo.. 0.000.200. No. .0 .0.00 000 0_0.0E0 Eo.. 00.0.00. .00. 0....0.000.0.E N .0 02.000.00.000 .00 0.00 ..

106

Table 4.3. Genetic diversity of two microsatellite loci in emerald ash borer populations from Asia
and North America (Loci developed by Dr. Jenny Cory. Simon Fraser University, BC, Canada)

 

 

 

 

China
Tianjin Hebei Bemng Liaoning Jilin
Dagong Tangshan Beijing Benxi Shenyang Jiutai Chanchun
C-C5 Ho 0.5 0.578 1 0.6 0.857 0.6 0.5
HE 0.725 0.633 0.666 0.711 0.791 0.501 0.56
No 5 4 2 3 5 5 4
A 1.725 1.633 1.666 1.711 1.791 1.501 1.56
C-C8 HO 0.285 0.47 1 0.4 0.333 0.2 0.333 ’-
HE 0.527 0.513 1 0.355 0.484 0.336 0.439
No 2 2 2 2 2 2 3
A 1.527 1.513 2 1.355 1.484 1.336 1.439
North America
Ann Arbor, Livingston Co., Lansing, Moran, Windsor,
Ml Ml MI Ml Illinois Pennsylvania Canada
C-C5 Ho 0.833 0.75 0.5 0.444 0.615 0.533 1
HE 0.787 0.607 0.423 0.542 0.63 0.514 0.733
NA 4 3 4 4 4 3 3
A 1.787 1.607 1.423 1.542 1.63 1.514 1.733
C-CS Ho 0.166 0 0.583 0.625 0.636 0.769 0.166
HE 0.409 0.355 0.489 0.491 0.601 0.667 0.53
NA 2 2 2 3 3 4 2
A 1.409 1.355 1.489 1.491 1.601 1.667 1.53

 

H0 = observed heterozygosity; HE = expectted heterozygosity; N0 = number of alleles;

A = Allelic richness

107

 

4..

 

Coom ._0 .0 0000.0. :00.:0..2 0.:. :0..0:00..:. .0...:. 0:. .00: 00 0. 00.000000 0. 00.0 0.0...

.00..::00 0:>0>> 0:0 .30:0.000>> 0:00.00 000052.. 05020:. :00.00..2 :.0.0000.:00 Eo.. 00.00.60 0.03 0.000.205 u 00 t

.0:o00.0 0.... 30.00 E5000 E0. 5. 0:0..05000 :00..0E< 5.02 :00500 000:0.05 ..

0N.0 0_0E00 n z

 

 

 

o 000 000 _.\. 009 000

005.0 0 wow #00 5mm 0N0

Sum—20 NNmNd o 000 000 000

050.0 .5320 .0006 o 00.. com

034:. Nommd .3de 0000.... o 000

0500.0- 0.0.0- wmomd who... mmmmd o

wowed- mnwwd 0000.0 hNNodr wooed- moo—mo 0

006.0. 00070 000.20 NNmod 0.00.0 Nnmod onmod- 0

meed vamd 500.0 NMNNd 00.0—no nmmmd 00070 {0.0 0

005.0 PNFNd mum—mo 030.0 mood tum—mo 0000.0- hood- Nomad 0
000000 0.:0>_>0::0n. :0.0$.-=2 00-..). E0002 0.0:..._ Q00 0:40:00... 5.... .000:
082.2, - :2 .500:

8 u z. 000:00
..000:.>> .o.
$Fnz.
0.:0>.>0::0n. .0
SFuz.
:0.0_2 - =2 .0
aFuz.

Ciwm I —§ .5
€Fuz.
59.9.. - =2 .0
sFuz.
0.0:._._ .m
sFuz.
0:...00\:._:0... .0
aFuz.
0:.0:00... .m

:Nqu

cEaN
.anuz.
.000:.F

 

.00..0E< 0002 0:0 0:.00 :. 30003000 .0.00 000 0.0.0E0 .0 .0.... 00.5.00. :0..0..:0.0...0 0..0:0O .0... 0.00 .r

108

4.4). The Illinois population appears most similar to Liaonging Province.
However it is difficult to identify a single population that is most similar to the
other N. American populations. These microsatellite data do not provide enough
information to distinguish between a single EAB introduction or multiple EAB
introductions into N. America.

Neighbor-joining analysis based on pairwise FST of Asian and North
American populations did not reveal clear differentiation between populations

based on geographic separation (Figure 4.1). The topology of the genetic tree is

 

most likely due to several negative pairwise FST treated as zero’s by the program
MEGA and not likely due to biological similarity. Negative values are most likely
due to greater within population variation than among population variation
(Conner and Hartl 2004). Greater number of loci and individuals are needed to
reduce this error.

To assess the main mode of spread in N. America, pairwise FST values
were evaluated as a function of geographic distance between any two given N.
American populations. If the spread of EAB was due to natural movement,
genetic differentiation is expected to increase as geographic distance increases
(Sakai et al. 2001), however, little relationship between genetic and geographic
distance would provide evidence humans mediated transport is likely. If there is
any isolation by distance in this set of populations, it is very slight. There is a
slight relationship among Fst and distance, but this relationship is not significant

(Figure 4.2). This non-significant relationship suggests that humans have aided

 

109

Tianjin/Beijing

 

 

 

 

 

 

 

 

 

 

 

 

 

 

lngham, Ml
Hebei
Jilin
__ Liaoning
l Southeast MI
I | Moran, Ml
Windsor, Canada
Illinois
1 Pennsylvania
W

 

Figure 4.1. Neighbor-joining analysis of EAB microsatellite loci (two) based on
painrvise FST values for North American EAB Populations.

110

 

0.35.
. y = 0.0001x + 0.0712

R2 = 0.0518
0.3 .

0.25 4 o

0.15 ‘ , 0 °

Fst

0.1 4
0.05 ~ 0

f 7 Y I I T I

100 200 300 400 500 600 700 800 9 900 1 000

 

Kilometers

Figure 4.2. Test for Isolation by Distance in the North American EAB Populations
based on the analysis of microsatellite markers (2 loci).

111

 

the movement of EAB in N. America, however, it cannot be ruled out that natural
dispersal is a factor in the movement of EAB across some of these distances.
Analysis of genetic variation at polymorphic microsatellite loci holds
promise for answering questions about both the structure of EAB populations at
various spatial scales and about the basic biology of EAB. Even this dataset,
based on an analysis of 162 individuals at two genetic loci, has yielded some
interesting and thought-provoking results. Development of more microsatellite
loci and increased samples sizes will be needed to give a better understanding of
the invasion dynamics of EAB. EAB microsatellite markers also have
applicability with respect to answering some basic questions about the biology of
EAB, e.g., how many individuals are responsible for the eggs found on a single
tree? The outcome of further analysis thus has direct impacts on control
strategies since that could help concentrate efforts on the leading edge of
migration for control rather than on many long-range dispersal events caused by

humans.

112

 

 

 

APPENDIX

113

 

APPENDIX A

COI sequence haplotypes

114

 

 

- - - - - - - - - - - - - - - - - - - - - - - - - - 0...
- - - - - - - - - - - - - - - - - - - - - - - - - - 0:
- - - - - - - - - - - - - - - - - - - - - - - - - - v:

(EOOOEO._.._.<0<<._.<._.<<._.<<<._.<<OFI
N0 .0 00 av 0.0 0v 00 00300 NV SN 00 mm 00 50 00 mm 00 00 N0 .0 cm mm mm mm

.. - - - - - - - - - - - - - - - - - - - - - - .. - F 0..
- .. - - - - - - - - - - - - - - - - - - - - - - - F 0..
- - - - - - - - - - - - - - - - - - - - - - - - - F 0:
- - - - - - - - - - - - - - - - - - - - - - - - - 0 0..
0 0 0 0 F F < 0 0 < F < F < 0 F 0 0 < 0 F < 0 0 F 0:. F:
0N .N N NN FN ON 0F 0F 2 0F 0F 3 0F NF FF 9 0 0 N 0 0 v m N F c 0000.00:

 

80.80 02.8.82

 

00.02000 0000 .0 500000.. 00. u : 00.000. 0.0 0:0...000 00000.00: 0.0000) .2. 0030.000 :. 0000. 0. .00..:00. 0.0.00.
000000.00: 0020008 0.0000 00:... 00.00 000 0_0.0E0 .0 0000 F 000006 0Eo.000.>0 <20.E 00. :. :0..0..0> .0. 0.00 00.00.00... .F.< 0.00 F

115

no.

No.

.b..

00.

'I—<

- - - m:
- - - 0:
- - - E
- - - 0...
- - - 0...
- - - .0.
- - - 0:
- - .. N:
F < < F...

- - - m:
- - - 0:
- - - 0.
- - - 0:
- - - 0..
- - - 2.
- - - 0:

- - N:

0 F F:

:80-
3

002...:8 F... 0.00F

116

mvw

VNF

vvr

mm.

00.

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er

FNF

Fvw

ON.

00.

0F.

00.

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00.

new

000

or.

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mm.

0..

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00F

NF.

N0.

FF.

.0.

OFF

00F

00F

mm.

00.

0NF

no.

- - - 0:
- - - E
- - - 0:
- - - 0:
- - - .0..
- - - 0:

H. < .h VI
NNF 0N? va

- - - m...
- - - 0..
- - - 0.
- - - 0...
- - - 0..
- - - 0..
- - - 0..
- - - N:
< 0 0 F:

00F 00F 40F

82.....8 HF.< 0.00F

117

50F

00..

00.

00..

00.

00..

V0.

00..

00.

N0.

N0.

.0..

.0..

00..

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hm...

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F:

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03.

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00.

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00.

0...
0...
RI
01
m...
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0....
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F...

0....
01
2..
0....
m...
S...
m...
N...
Z...

118

 

- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - n:
- - - - - - - - - - - - - - - - - - - - - or
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - v:
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - N:
F <. <. h F (V <‘ o no my my my A< h o p » A< _< o .F F:

mmm mmm NNN omm mmm «Nu mum «Nu rum emu arm mrm firm ova arm v.~ mFN «Fm FFN arm mom

- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - “I
- - - - - - - - - - - - - - - - - - - - - o:
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - «I
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - N:
o o F o k < .< mu m. k “U my <‘ P F F .< o h F ”U .1

wow now mom mom vow new mom Pom com mm. mm. “a. mm. mm. «a. mm. mm. .a. om. mm. mm?

umzczcoo “Pi 2an

119

- - - - - - - - - - - - m:
- - - - - - - - - - - - h:
- - - - - - - - - - - - o:
- - - - - - - - - - - - m:
- - - - - - - - - - - - v:
- - - - - - - - - - - - m:

KN CNN m0N 00w NON 00m 00m v0N m0N NON PON o0N mmm 00m NON 00m 00m «0N mmm Nmm .‘0N

omm mvw 0cm D‘N 0VN 0vw 3w 9% NVN

- - - - - - - - - o - - a:

- - - - - - - - - - - - m:
- - - - - - - - - - - - h:
- - - - - - - - - - - - o:
- - - - - - - - - - - - m:
- - - - - - - - - - - - v:
- - - - - - - - - - - - m:
- - - - - - - - - - - - N:
.< o _< _< o ‘< H P 4< h P .P r:
va o¢~ mmu mmm 5mm mmm mmm «mm mmm umw qu emu

umacncoo Hwas. flow...

120

 

 

- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - n:
- - - - - - - - - - - - - - - - - - - - - o:
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - v:
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - N:
.< 0 (V H F _< P ”u m. h F _< F o p <. m. h o H mu .1

mFm «Fm .Fm arm mom mom “on mom mom «on new mom Fom com wow mam 5mm mmm mmm vow man

- - - - - - - - - - - - - - - - - - - - - a:
- - - - - - - - - - - - - - - - - - - - - w:
- - - - - - - - - - - - - - - - - - - - - n:
- - - - - - - - - - - - - - - - - - - - - o:
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - «I
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - N:
<_ P 0 (V p o F no .0 4< 0 (V <_ ‘< o o <_ m. <. H P PT.

wow Pam cam mmm mmm 5mm mmm mmm «mm mam «mm me omm mNN mum hum mum mNN «Nu mNN «Nu

8:52.50 “—34. 0.2m...

121

 

 

- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - oz
- - - - - - - - - - - - - - - - - - - - - 3..
- - - - - - < - - - - - - - - - - - - - - N:
< o o < o o o o 0 F F o F F o F o o < F F F:

mmm «mm mmm mmm me omm mvm mvm va mvm mvm «cm mwm Nvm va ova mmm mmm me mmm mmm

- - - - - o o - - - - - - - - - - - - - - m:
1 - - - - .. .. - - .. - - - - - .. .. - - - - w:
- - - - - - - - - - - - - - - - - - - - - o:
- - - - - - - - - - - - - - - - - - - - - «I
o o F < F F F F o F F o F F o F F o o o w E

«mm mmm mmm me omm mmm mmm me mmm mmm cum mmm mum me omm mFm mFm FFm oFm mFm VF”

8:528 F.< 0_an

122

 

- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - F:
- - - - - - - - - - - - - - - - - - - - - o:
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - v:
- - - - - - - - - - - - - - - - - - - - - m:
- - - - - - - - - - - - - - - - - - - - - «I
o F F 1< o F <2 o .< F _< .< 1< F F o 1< ,a ,\ m. 0 FT.

Fom mmm mmm «on mmm mmm me cam mmm mmm me mam mmm vmm mam «mm me own mFm me FFm

- - - - .. - - - - - - - - - - F - - o - - m:
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APPENDIX B

Deposition of voucher specimens

126

Appendix B
Record of Deposition of Voucher Specimens*
The specimens listed on the following sheet(s) have been deposited in the named museum(s) as
samples of those species or other taxa, which were used in this research. Voucher recognition
labels bearing the Voucher No. have been attached or included in fluid-preserved specimens.
Voucher No.: 2009-05
Title of thesis or dissertation (or other research projects):

AFLP, mtDNA, and Microsatellite Analysis of Emerald Ash Borer Population Structure from Asia
and North America

Museum(s) where deposited and abbreviations for table on following sheets:

Entomology Museum, Michigan State University (MSU)

Other Museums:

lnvestigator‘s Name(s) (typed)
Alicia Marie Bray

 

 

Date 18/Auq/2009

*Reference: Yoshimoto, C. M. 1978. Voucher Specimens for Entomology in North America.
Bull. Entomol. Soc. Amer. 24: 141-42.

Deposit as follows:
Original: Include as Appendix 1 in ribbon copy of thesis or dissertation.

Copies: Include as Appendix 1 in copies of thesis or dissertation.
Museum(s) files.
Research project files.

This form is available from and the Voucher No. is assigned by the Curator, Michigan State
University Entomology Museum.

127

 

 

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