EVALUATION OF EMERALD ASH BORER POPULATIONS AND THE ASH RESOURCE AT THREE
STAGES OF THE INVASION WAVE
By
Stephen J. Burr

A THESIS
Submitted to
Michigan State University
in partial fulfillment of the requirements
for the degree of
MASTER OF SCIENCE
Entomology
2012

ABSTRACT
EVALUATION OF EMERALD ASH BORER POPULATIONS AND THE ASH RESOURCE AT THREE
STAGES OF THE INVASION WAVE
By
Stephen J. Burr
Questions consistently arise as to the status of Agrilus planipennis populations and the
persistence of ash in forested settings, particularly in the original core of the A. planipennis
infestation in southeast Michigan. I sampled A. planipennis populations in 24 green ash
(Fraxinus pennsylvanica) sites, each 1 ha in size. Eight sites were located in each of three areas
of southern Michigan representing (1) the original A. planipennis Core; (2) the invasion Crest, in
south central Michigan, where A. planipennis populations were currently peaking, and (3) the
invasion Cusp, in southwest Michigan where A. planipennis had recently become established.
Adult A. planipennis were captured in all 24 sites in both years of the study. No larvae
were present in the two most westerly Cusp sites in 2010, but larvae were present in all 24 sites
in 2011. Despite the depletion of overstory ash in southeast Michigan, A. planipennis continues
to infest and kill trees in Core sites. Beetle populations were highest in south central Michigan
and are rising in southwestern Michigan. Mortality of overstory ash decreased across an east
to west gradient across southern Michigan. Green ash was abundant in the understory of all
sites including Core sites. There was no evidence of current year ash germination in Core sites
in southeast Michigan. If A. planipennis persists in these areas, regenerating ash will continue
to become infested and killed, and potentially eliminating ash from these systems.

DEDICATION
I would like to dedicate this thesis to my grandmother, Margret Hansen (Gee), who
passed away while I was working on this project, as well as my other grandparents Ronald
Hansen (Bumpa), Frances Burr (Grandpa), and Ruth Burr (Grandma), and my great-grand
parents Lewis Bauser (Opa), and Ottomine Hansen (Grandma Minnie). They watched over me
as a child, and their strength of spirit is with me to this day.

iii

ACKNOWLEDGEMENTS

I thank, first and foremost my advisor, Dr. Deborah McCullough for her guidance and
leadership. Deb allowed me the freedom to find my own path through this adventure, stepping
in and putting me back on track only when I went astray. I would also like to thank the other
members of my research committee including Dr. Therese Poland, who employed me as an
undergraduate, and who sparked my first interest in the field of forest entomology, and Dr.
Richard Kobe, whose guidance in the forestry aspect of my research was invaluable.
I could not have successfully completed my program without the help of many people. I
would like to thank Paul Nelson, Rachel Posavetz, and Molly Robinett, my undergraduate
assistants, who worked beside me at all hours of the day, and sometimes in unpleasant site
conditions without complaint. I would like to thank my fellow graduate students Sara Tanis and
James Wieferich, and our lab technicians Andrea Anulewicz, Jacob Bournay, Nick Gooch, and
Andrew Tluczek, for allowing me to lean on their experience and insight. I would also like to
thank my other lab mates Elliott Berlin, Tara Dell, Russ Kibat, Rodrigo Mercader, Emily Pastula,
Kyle Redilla, and Nathan Siegert for their assistance in the field and laboratory.
This research could not have been conducted without the contribution of the land
managers who allowed me to work within their sites. I would like to thank the Campus Natural
Areas Committee at Michigan State University; Jim Curtis and Greg Kowalewski at MSU’s
Kellogg Forest; Mark Bishop, Randy Heinze, Dan Kennedy, Tim Machowicz, Nicole VanBloom,
and Denise Smith from the Michigan DNR; Paul Mullee from Huron-Metro Parks; Steve Alman
and Jim Rozonowski from Wayne County Parks; Richard Simek and Dr. David Susko from

iv

University of Michigan Dearborn; Chip Francke from Ottawa County Parks; Jackie Blanc, John
Greenslit, and Dan Patton from Eaton County Parks; and Jane Greenway and LuAnn Maisner
from Ingham County Parks.
I would also like to thank my family for their support through this endeavor. To my
parents Jill and James Burr, and to my sisters Rachel, Amy, and Maria, for always being there
when I needed you. I would also like to give a special thanks to my dog Bear, for every squirrel
and deer he chased, for every pool he swam in, and for every mud puddle he rolled in he
reminded me that though this is work, there is no need not to have fun along the way.
This research was funded by the USDA Forest Service, Northeastern Area, Forest Health
Protection.

v

PREFACE

Emerald ash borer (Agrilus planipennis) was discovered in southeast Michigan in June of
2002 (Cappaert et al. 2005), but is believed to have entered North America six to ten years prior
to its discovery. Since its arrival, A. planipennis has killed tens of millions of ash trees in
Michigan alone, and is currently established in 15 states and two provinces of Canada.
Initial research conducted in North America focused on A. planipennis biology, host
range, potential impact, and optimal trapping techniques. Researchers in Michigan, the
surrounding states, and Canada have worked to develop better methods for detection and
monitoring, and reducing the impact of this invasive pest.
This project focused on evaluating A. planipennis populations at three stages of the A.
planipennis invasion wave, and the impact A. planipennis on the overstory, and potential
regeneration of green ash (Fraxinus pennsylvanica). Previous literature has reported nearly
100% mortality of all Fraxinus spp. within infested stands (Gandhi et al. 2008). In Chapter 1, I
address adult trapping and larval surveys that were conducted to assess beetle populations
behind, within, and in front of the advancing A. planipennis invasion wave. Ash overstory trees
were assessed in fixed radius plots and belt-transects to determine ash basal area, mortality,
and canopy decline at all three stages of the invasion.
Studies reported in Chapter 2 focused on impacts of A. planipennis on overstory ash,
and three stages of ash regeneration. All species of overstory trees, recruits, saplings, and
seedlings were assessed in fixed radius plots and belt-transects to assess condition of overstory
trees, and understory species composition to determine what species are likely to replace ash

vi

in these systems. Readings were taken in the understory to measure available
photosynthetically active radiation (PAR), and sapling growth associated with gaps formed by
overstory ash mortality. My goal was to better understand A. planipennis population densities
across southern Michigan, and A. planipennis impacts on overstory green ash, potential
regeneration, and likely changes to ash stands in southern Michigan. Each chapter of the thesis
is intended to be prepared as a manuscript for publication.

vii

TABLE OF CONTENTS

LIST OF TABLES…………..………………………………………………………………………………………………………………x
LIST OF FIGURES……………………………………………………………………………………………………………….……….xi
CHAPTER 1
Emerald Ash Borer Populations and the Condition of the Green Ash Overstory
at Three Stages of the Invasion Wave………………………………………………………………………….….……..…1
Introduction……………………………………………………………………………………………………………..……1
Material and Methods…………………………………………………………………………………………..……...5
Study sites…………………………………………………………………………………………………..……..5
Adult A. planipennis populations…………………………………………..……………………………5
Larval A. planipennis densities……………………………………………………………………………8
Overstory…………………………………………………………………………………………………………..9
External signs of A. planipennis………………………………………………………………………..10
Statistical analysis…………………………………………………………………………………………….11
Results………………………………………………………………………………………………………………………….12
Adult A. planipennis captures……………………………………………………………………………12
Larval A. planipennis densities…………………………………………………………………….……14
Atanycolus cappaerti parasitism…..……………………………………………………………..……17
Overstory………………………………………………………………………………………………….………18
External signs of A. planipennis………………………………………………………………..………20
Discussion…………………………………………………………………………………………………………….………23
Tables…………………………………………………………………………………………………………………..………31
Figures………………………………………………………………………………………………………….………………33
Chapter 2
Green ash Overstory and Regeneration Ahead, Within, and Behind the Emerald
Ash Borer Invasion Wave...……………………………………………………………………………………………………..38
Introduction…………………………………………………………………………………………………………………38
Materials and Methods……………………………………………………………………………………………..…41
Study sites……………………………………………………………………………………………………..…41
Overstory…………………………………………………………………………………………………….……41
Recruits……………………………………………………………………………………………………….……42
Saplings………………………………………………………………………………………………………….…42
PAR and sapling growth……………………………………………………………………………….……43
Seedlings……………………………………………………………………………………………………..……44
Statistical analysis………………………………………………………………………………………….…44
Results……………………………………………………………………………………………………………………….…46
Overstory………………………………………………………………………………………………………….46
Canopy dieback………………………………………………………………………………………………..48
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Recruits……………………………………………………………………………………………………………48
Saplings……………………………………………………………………………………………………………49
PAR and sapling growth……………………………………………………………………………………51
Seedlings……………………………………………………………………………………………………….…51
Discussion………………………………………………………………………………………………………………….…53
Tables…………………………………………………………………………………………………………………..………58
Figures…………………………………………………………………………………………………………….……………60
APPENDIX A
Record of deposition of voucher specimens…………………………………………………………………………….67
APPENDIX B
List of study sites, Michigan county, and geographic coordinates……………………………….……………68
LITERATURE CITED………………………………………………………………………………………………………..………….69

ix

LIST OF TABLES

Table 1.1. Mean (± SE) trap tree DBH (diameter at breast height), total
captures of adult and larva A. planipennis, adults and larvae per
2

m , surface area of adult traps, and surface area examined
for larvae in 2010 and 2011 in 24 sites in southern
Michigan.…………………..…………………………………………………………………………………………………31
Table 1.2. Total number of ash trees and live ash trees, mean (± SE)
percent of ash mortality, ash canopy dieback and proportion
of ash with epicormic shoots, stump sprouts and woodpecker
predation in 2010 and 2011 in 24 sites in southern
Michigan. Canopy dieback and epicormic shoots were
recorded on live ash trees. Stump sprouts and holes left
by woodpeckers preying on A. planipennis larvae were
recorded on all ash …………..……………………………………………………………….…………………………32
Table 2.1. Total number of trees (N), mean (± SE) diameter at
breast height (DBH) and basal area, relative density, relative
frequency, relative dominance, and relative importance
values (RIV)of the most abundant tree species across 24 sites
representing three A. planipennis invasion stages.
Relative importance values rank species contribution to a stand's
overstory. All measurements were taken in summer 2010.…………………………………………58
Table 2.2. Total count (N) and mean (± SE) stems per ha of the five
most abundant species of recruits, saplings, and seedlings in
2010, sampled in 24 sites in southern Michigan representing the
three stages of the A. planipennis invasion wave…………………………..…………………………….59
Table A.1. Specimens submitted to the Michigan State
University (MSU) Albert J. Cook Arthropod Research Collection………………………..…………67
Table B.1. List of study sites by Michigan county and
geographic coordinates…………………………………………….………………………………………………..68

x

LIST OF FIGURES

Figure 1.1. Locations of 24 green ash sites representing three stages of
the A. planipennis invasion wave in 2010 and 2011. Core sites were
characterized by low densities of A. planipennis and high ash
mortality. Crest sites were characterized by high densities of
A. planipennis and moderate to high ash mortality. Cusp sites were
characterized by low densities of A. planipennis and low
ash mortality………………………………………………………………………………………………………………33
2

Figure 1.2. A. Mean (± SE) captures of A. planipennis adults per m of
trap area in Core, Crest, and Cusp sites in 2010 and 2011.
2

B. Mean (± SE) captures of A. planipennis larva per m of
exposed ash phloem in Core, Crest, and Cusp sites in 2010 and
2011. Means with the same letter are not significantly
different (Tukey’s protected HSD test; P < 0.05).
(a, b, and c for 2010; and y and z for 2011)…………………………………………………………………34
2

Figure 1.3. Linear regression of larval A. planipennis density per m of
2

exposed phloem and adult A. planipennis density per m on
traps in (A.) 2010 and (B.) 2011 at 24 sites in southern
Michigan (N = 24) (P < 0.05)………………….………..……………………………………………………………35
2

Figure 1.4. A. Mean (± SE) live basal area (m per ha) of green ash in
Core, Crest, and Cusp sites in 2010 and 2011. B.
2

Mean (± SE) dead basal area (m per ha) of green ash in
Core, Crest, and Cusp sites in 2010 and 2011. Means with
the same letter are not significantly different
(Tukey’s protected HSD test; P < 0.05)
(a, b, and c for 2010; x, y, and z for 2011)……………………………………………………………………36

xi

Figure 1.5. A flow diagram representing A. planipennis populations
and the negative effects on overstory ash. The positive effects
of overstory ash abundance on A. planipennis occurrences through
reproduction and immigration. The negative effects of
A. planipennis mortality factors such as parasitism and
woodpecker predation on larvae, and effects of the
reduction of ash overstory on A. planipennis
populations through emigration. Positive relationships are
represented by
and negative relationships are
represented by
………………………………………………………………………………….………………………37
Figure 2.1. Locations of 24 green ash sites representing three stages
of the A. planipennis invasion wave. Core sites were
characterized by low densities of A. planipennis and
high ash mortality. Crest sites were characterized by
high densities of A. planipennis and moderate to high ash mortality.
Cusp sites were characterized by low densities of
A. planipennis and low ash mortality……………………………………………………………………………60
Figure 2.2. Mean (± SE) density of green ash saplings per ha
in 2010 (N = 2133) and 2011 (N = 1905). Means with
the same letter are not significantly different among
the three invasion stages (a and b for 2010, y and z
for 2011).……………………………………………………………………………………………………………………..61
Figure 2.3. Mean (± SE) of percentage of full sun represented by
photosynthetically active radiation (PAR) measured after
dawn and before sunrise in 12 sites from 20 June to 15 July
2011 (N = 428 readings). Means with the same letter are
not significantly different (Tukey’s protected HSD test; P < 0.05).………………………..………62
Figure 2.4. Mean (± SE) growth (cm) of leaders of green ash saplings
measured in 12 sites between 12 June to 17 June at the three
A. planipennis invasion stages (N = 237). Means with the same
letter are not significantly different (Tukey’s protected HSD
test; P < 0.05).………………………………………………………………………………………………………………63
Figure 2.5. Linear regression of annual growth of terminal leaders of
green ash saplings and photosynthetically active radiation (PAR)
measured in 12 sites in southern Michigan (N=237 paired measurements).
Saplings were 1 to 2 m tall, diameter and exact height of saplings
were not measured (P < 0.05).……………………………………………………………….………………64
xii

Figure 2.6. A flow diagram representing effects of A. planipennis
populations on overstory ash and ash recruits,
and the positive relationship of overstory ash abundance on
seedlings, saplings, and recruits. Positive
relationships are represented by
and negative
relationships are represented by
……………………………………………………………………65

xiii

CHAPTER 1

Emerald Ash Borer Populations and the Condition of the Green Ash Overstory at Three Stages
of the Invasion Wave

INTRODUCTION
Emerald ash borer (Agrilus planipennis Fairmaire) (Coleoptera: Buprestidae), is a
phloem-feeding beetle native to Asia originally discovered in Detroit, MI and Windsor, Ontario
and identified in July 2002 (Cappaert et al. 2005). Dendrochronological studies indicate A.
planipennis entered North America at least six to ten years prior to its discovery (Siegert et al.
2007). Since its arrival, A. planipennis has attacked ash (Fraxinus spp.) trees in forest, rural, and
urban settings, with mortality rates approaching 100% (Gandhi et al. 2008). Unlike many native
phloem-feeding Agrilus spp. which predominantly infest stressed trees (Balch and Prebble
1940; Barter 1965; Dunn et al. 1986; Muzika et al. 2000), A. planipennis can attack healthy ash
trees (Poland and McCullough 2006). As ash resources are depleted in local areas, beetle
populations disperse and colonize new locations.
The resulting A. planipennis invasion wave is expanding in all directions from its origin.
In addition to natural dispersal, transportation of infested logs, nursery trees, and firewood has
resulted in satellite populations of A. planipennis established miles from the original infestation
sites (Cappaert et al. 2005), increasing the overall rate of expansion (Liebhold and Tobin 2010;
Mercader et al. 2011). Presently, A. planipennis is established in 16 states and two provinces of
Canada (EAB.info 2012).

1

Adult A. planipennis emerge in May and June from pupation cells in the outer sapwood
or outer bark (Cappaert et al. 2005). Adults feed on ash foliage for 5 to 7 days before mating,
and females feed for an additional 1 to 2 weeks before oviposition begins (Cappaert et al.
2005). Throughout July and August females deposit eggs beneath rough edges of bark or within
cracks in the outer bark of ash trees. Larvae hatch within two weeks, chew into the phloem and
begin feeding. Larvae overwinter in the thick outer bark or outer sapwood of trees, pupating
the following spring. In high density infestations, A. planipennis complete their life cycle in a
single year, but in healthy trees with low larval densities, most larvae require two years to
develop (Cappaert et al. 2005; Siegert et al. 2010; Tluczek et al. 2011). Prolonged development
of A. planipennis larvae may reflect host resistance or host quality (Poland and McCullough
2006; Tluczek et al. 2011).
As larval densities increase, feeding disrupts translocation of nutrients in trees, causing
canopy decline, and eventually death of the host. External signs of A. planipennis become
apparent, include D-shaped exit holes left by emerging A. planipennis adults, holes left by
woodpecker predation on late instar larvae, and frequent formation of epicormic shoots and
stump sprouts. As the ash component of a stand is depleted, A. planipennis disperse and
colonize new locations. Stands behind the invasion wave typically have few remaining live
overstory ash (Ghandi et al. 2008).
The long-term repercussions of the A. planipennis invasion on green ash in North
America are currently unknown. Agrilus planipennis is known to attack and kill all Fraxinus spp.
in North America, but preferentially infests green ash (Fraxinus pennsylvanica Marsh)
(Anulewicz et al. 2007, 2008; Rebek et al. 2008). Green ash is the most widely distributed
2

species of ash in North America, ranging from Nova Scotia to Saskatchewan in the north and
from Texas to Florida in the south (Wright 1959; Kennedy 2009). Green ash is found primarily
on alluvial soils in riparian zones, but is capable of surviving on upland sites (Wright 1959;
Kennedy 2009). It can grow in sites ranging from clay soils subject to frequent flooding to sandy
silt soils where water can be limited (Burns and Honkala, 1990). Green ash is a common species
in Michigan forests (Barnes and Wagner 1981). Millions of green ash have been killed in
Michigan by A. planipennis (Poland 2007; EAB.info 2012) and millions more are at risk (FIA
database 2012). Ecological impacts ensuing from widespread ash mortality could have
cascading effects throughout systems where ash is a dominant overstory species (MacFarlane
and Meyer 2005; Gandhi and Herms 2010).
In this study, I assessed A. planipennis population levels behind, within, and in front of
the expanding invasion wave in green ash sites located along an east-west gradient across
southern Michigan. The first invasion stage, classified as the “Core,” represents the original
infestation area where most ash trees have been killed, and A. planipennis density has
presumably dropped with the reduction of available resources. The second stage, designated
as the “Crest,” represents sites where A. planipennis populations are near peak densities, and
ash decline and mortality are moderate to high. The third stage of the invasion wave was
designated as the “Cusp.” Ash trees in Cusp sites appeared healthy with few external signs of
A. planipennis infestation, indicating A. planipennis populations had not yet built to damaging
levels capable of adversely impacting overstory ash trees. Adult A. planipennis populations
were assessed by capturing beetles on purple double-decker traps (McCullough and Poland
2009; McCullough et al. 2011; Poland et al. 2011) and sticky bands placed on girdled ash trees
3

and uninfested nursery ash trees planted in each site. Stressed ash trees attract A. planipennis
(McCullough et al. 2009a, 2009b; Siegert et al. 2010; Tluczek et al. 2011), at least partly due to
changes in volatiles emitted from foliage (Rodriguez-Soana et al. 2006; de Groot et al. 2008;
Grant et al. 2010) and bark (Crook et al. 2008). Sticky bands were also placed on ungirdled
“control” ash trees which were neither stressed nor baited, enabling me to characterize A.
planipennis activity uninfluenced by lures or visual stimuli. I assessed larval density by felling
and debarking control, girdled, and planted trees. I measured the condition of overstory ash to
evaluate how A. planipennis will affect stand with a substantial green ash component. This
study will provide insight into A. planipennis populations along the A. planipennis invasion wave
and the impact of A. planipennis on green ash stands. Results will be useful to A. planipennis
programs and resource managers as they attempt to manage this invasive pest.

4

MATERIALS & METHODS
Study sites: Twenty-four sites, each one ha, located on an east to west gradient across southern
Michigan, were selected in July and August 2009 (Fig. 1.1). Sites were on state, county, and city
property, and were second growth stands with trees between 60 and 90 years of age
(Michigan.gov 2012). The overstory at each site was comprised of a minimum of 20% green
ash, based upon available inventory data provided by managers, previous studies conducted in
these locations, and on site visits. Eight sites were selected to represent each of three invasion
stages designated as Core, Crest, and Cusp. Core sites were located in southeast Michigan (Fig
1.1,) in areas where the majority of ash trees were dead, and beetle populations were believed
to have dropped with the reduction of available resources. To ensure ash mortality was the
result of A. planipennis, trees in Core sites were examined for D-shaped exit holes in the bark
left by emerging beetles, and bark was removed from several trees to determine if A.
planipennis galleries were present. Crest sites in south central Michigan (Fig 1.1) were
characterized by ash trees in various stages of decline. Roughly half of the ash trees in Crest
sites were alive, and of those, most trees were obviously infested by A. planipennis. Exit holes
and serpentine galleries beneath the bark were found on sampled trees. Many trees in Crest
sites also had epicormic shoots, stump sprouts, and signs of woodpecker predation on larvae.
In Cusp sites in southwest Michigan (Fig 1.1), ash trees showed little evidence of infestation.
Exit holes and galleries were observed in some sites, but infrequently.
Adult A. planipennis populations: Adult A. planipennis populations were assessed in 2010 with
a variety of trapping methods, including double-decker traps and sticky bands on girdled,

5

planted, and control trap trees. Traps and sticky bands were deployed at sites beginning on 10
May and remained in place until 16 August.
Two purple double-decker panel traps were placed at each site, in full sun when
possible (McCullough and Poland 2009; McCullough et al. 2011; Poland et al. 2011). Traps
consisted of a 3-m-tall polyvinyl chloride (PVC) pipe (10 cm diameter), and two three-sided
prism panels constructed from purple corrugated plastic (Harbor Sales Inc., Sudlersville, MD).
Each panel was 4 mm thick, 60 cm in height and 40 cm in width. The first panel was attached to
the top of the PVC pipe using cable ties, and the second panel, similarly attached, was placed 60
cm below the top panel. Clear Pestick (Hummert international, Earth City, MO), was applied to
both panels. The top panel was baited with two bubble caps of cis-3-hexenol (combined
volatilized release rates of 7.4 mg/d determined in the laboratory at 20°C; Contech Enterprises,
Inc., formerly Phero Tech, Inc., Delta, BC, Canada). The bottom panel was baited with an 80:20
blend of Manuka oil and Phoebe oil (release rate of 50 mg/d determined in the laboratory at
20°C; Synergy Semiochemicals Corp., Burnaby, BC, Canada). Double-decker traps were placed
over 1.5 m tall T-posts set into the ground for support.
Three uninfested, green ash nursery trees, 3.8-6.4 cm DBH, (Bailey Nursery, Newport,
MN) were planted at each site, typically in full sun along edges or in gaps to optimize beetle
captures (Yu 1992; McCullough et al. 2009a, 2009b). I assumed transplant shock would be
adequate to elicit the chemical stress responses attractive to A. planipennis (McCullough et al.
2009a, 2009b; Tluczek et al. 2011). A band of clear plastic wrap, 30 cm wide and coated in
Tanglefoot (Contech Enterprises, Inc., formerly Phero Tech, Inc. Delta BC, Canada), was
wrapped around each planted tree, 1 m above the base to capture adult A. planipennis.
6

Two naturally regenerated ash trees in each site were girdled. Trees growing in gaps
with increased sun exposure were selected to optimize adult captures (Yu 1992; McCullough et
al. 2009a, 2009b), but were also selected to be representative of ash trees within each site. To
girdle trees, drawknives and handsaws were used to remove a 15 cm wide band of outer bark
and phloem, 1 m above the base of the tree. A band of clear plastic wrap, 30 cm wide, was
placed immediately above the girdle of each tree and coated with Tanglefoot.
Two additional ash trees growing on each site were selected using the same criteria
described for girdled trees. These trees were not stressed or baited, and were designated as
“control” trees. Sticky bands coated with Tanglefoot were applied 1 m above the base of the
tree.
Girdled and control trees, double-decker traps, and planted trees were a minimum of 10
m apart. Trap trees and traps of the same type (i.e. control and control, double-decker and
double-decker) were a minimum of 20 m apart.
Traps were checked biweekly. Adult A. planipennis were collected from each trap,
counted, and returned to the laboratory, where beetles were soaked in Histo-Clear II (National
Diagnostics, Atlanta GA) for two weeks to remove Pestick and Tanglefoot. Insects were then
examined to confirm identification. To estimate densities of A. planipennis adults on traps and
2

sticky bands, captures were standardized per m of trap surface area.
Adult A. planipennis trapping was repeated in 2011 with the following modifications.
Traps and sticky bands were placed in sites beginning on 9 May and remained in place until 12
August. Double-decker traps and planted trap trees were placed in roughly the same locations
as the previous year. Trees selected for girdling and controls were as near as possible to the
7

location of trees in 2010. Three planted trap trees were obtained from Law’s Nursery Inc. in
Hastings MN. Lures with cis-3-hexenol were again used on the top panel of double-decker traps
in 2011. Lures on the lower panel, however, were baited with manuka oil only, because Phoebe
oil was unavailable in 2011.
Larval A. planipennis densities: Densities of A. planipennis larvae were evaluated on trap trees
from October to December in 2010. All trees < 10 cm DBH, including planted trap trees and
small control and girdled trees, were debarked from the base to roughly 2 m above ground.
Stem diameters and tree height were measured prior to debarking. Because stem diameter
was different on the top and bottom of trap trees, the equation for a conical frustum was used
2

to determine m of exposed surface area. Trees ≥ 10 cm DBH were felled and bucked into 1 m
sections beginning just above the sticky bands. An area equal to 1/2 m in length and 1/2 the
circumference of the upper half of each log was debarked on alternate sections. Sections were
rolled to ensure all sides of the trees were sampled. Length and width of sections were
2

measured to calculate larvae per m of exposed phloem surface area.
Parasitoids found either as pupae in A. planipennis larval galleries, or as small larvae
attached to the larger A. planipennis larvae, were recorded. Logs with A. planipennis
parasitoids were returned to the laboratory and parasitoids were allowed to develop, and
emerge as adults. Representative adult parasitoids from each site where they were collected
were identified by Dr. John S. Strazanac from the University of West Virginia in Morgantown
WV, as Atanycolus cappaerti Marsh and Strazanac (Hymenoptera: Braconidae). The proportion

8

2

of A. planipennis larvae parasitized and parasitoid densities per m of exposed surface area
were calculated per tree.
Larval density was surveyed in 2011 using methods from 2010 but with two
modifications. Trap trees ≥ 10 cm were felled and cut into 1 m sections above the sticky bands,
and an area 1 m in length and by 1/2 the circumference of the log was debarked on alternate
sections. All trap trees were cut down and debarked in October and November.
Voucher specimens of ten larval A. planipennis, ten male and ten female adult A.
planipennis, and ten male and ten female adult A. cappaerti were submitted to the Albert J.
Cook Arthropod Research Collection at Michigan State University in East Lansing, Michigan.
Specimens were submitted in March 2012.
Overstory: Ash and non-ash trees (DBH ≥ 10 cm) were counted, recorded as alive or dead, and
DBH was measured in two belt-transects and four fixed radius plots. Belt-transects were 150 m
x 2 m, running diagonally across each site in an X-formation. Trees within 10 m of the belttransect intersections were marked to ensure individual trees were not measured more than
once. Belt-transects divided sites into four quadrats and one macro-plot (18 m radius) was
placed in the center of each quadrat. Basal area of all trees was calculated in cm and converted
2

2

-5

to m using the formula DBH X 7.854 X 10 (VanderSchaaf 2008). Basal area of all trees was
2

summed and divided by the area measured to calculate m per hectare. Trees were grouped
into five diameter classes (10 cm increments) for analysis. Basal area was calculated for all
species combined, and for live ash, dead ash, and total ash for each site. Dead ash trees were
assumed to be killed by A. planipennis if evidence such as holes left from woodpecker predation
9

of larvae, and A. planipennis exit holes were present. Sections of bark were removed from
dead ash trees if no external signs of A. planipennis infestation were apparent to confirm
presence of larval galleries. Dead ash trees without A. planipennis galleries were excluded from
A. planipennis mortality estimates.
Overstory assessments were repeated in 2011 with one modification. Basal area of nonash trees was assumed to remain relatively constant, so only ash trees were measured.
External signs of A. planipennis: Woodpeckers preying on late instar larvae and stump sprouts
growing from exposed roots at the base of trees were assessed on live and dead ash trees.
Woodpecker predation on ash trees was visually estimated as absent, low (1 to 6 woodpecker
attacks),and high (> 6 woodpecker attacks). Date of woodpecker predation cannot be
determined with visual surveys, so estimates represented cumulative woodpecker predation.
Stump sprouts were visually estimated as absent, low (1 to 4 stump sprouts), and high (> 4
stump sprouts).
On live trees, epicormic shoots and canopy dieback were also recorded. Epicormic
shoots growing on the trunk or branches were visually estimated as absent, low (1 to 4
epicormic shoots), and high (> 4 epicormic shoots). Canopy dieback was visually estimated in
10% increments, where 0% indicates a full canopy, and 90% indicates a nearly complete
absence of leaves (Zarnoch et al. 2004). Dieback was assessed from 21 June to 23 July, after
trees were fully flushed but before current-year larvae began feeding.
I estimated the external signs of A. planipennis again in 2011 with the following
modifications. Canopy dieback of ash was assessed from 18 June to 20 July. Canopy dieback of

10

non-ash trees was assumed to remain relatively constant, so non-ash trees were not assessed
in 2011.
Statistical analysis: Data were tested for normality using the Shapiro-Wilk test (Shapiro and
Wilk 1965) and residual plots. Adult A. planipennis captures, larval densities, and basal area
were normalized by log10(x + 1) transformations. Adult A. planipennis captures, larval
densities, and basal area were tested as unplanned comparisons to assess differences among
the three invasion stages. Tukey’s honestly significant difference (HSD) multiple comparison
procedure was applied if the overall analysis of variance (ANOVA) was significant (P < 0.05).
Two-way ANOVA was used to evaluate main effects of trap type, invasion stage and the
interaction of the two factors on adult captures and larval densities (PROC GLM; SAS Institute
2003) (Ott and Longnecker 2001). Estimates of trap surface area, trap tree DBH, the surface
area debarked, parasitoid densities, canopy dieback, and tree size class were not normalized by
transformations. Friedman’s two-way nonparametric test was therefore used to evaluate
differences among the different types of trap trees (i.e. control, girdled, and planted), the three
invasion stages, and the interaction between the two factors (PROC RANK; SAS Institute 2003)
(Friedman 1937). Friedman’s two-way nonparametric test was also used to evaluate effects of
invasion stage and diameter class on ash mortality, canopy dieback, and abundance of
epicormic shoots, stump sprouts, and woodpecker attacks (PROC RANK; SAS Institute 2003)
(Friedman 1937). When results for nonparametric tests were significant (P < 0.05),
nonparametric multiple comparisons were applied (Zar 1984). All analyses were conducted at P
< 0.05 level of significance using SAS statistical software (SAS Institute 2003).

11

RESULTS
Adult A. planipennis captures: In 2010, 2600 adult A. planipennis were captured on traps,
including 338, 1960, and 302 beetles in Core, Crest, and Cusp sites, respectively. The
proportions of total beetle captures were 13%, 75%, and 12% in Core, Crest, and Cusp sites,
respectively. Adults were captured in all 24 study sites. Beetle captures peaked from 21 June
to 2 July, when 1142 A. planipennis were captured, representing 44% of the years’ total.
2

Captures per m of A. planipennis adults in Crest sites were 5-fold higher than in Core sites, and
9-fold higher than in Cusp sites (F = 47.94; df = 2, 211; P < 0.001) (Fig. 1.2A). More A.
planipennis were captured in Core sites compared to Cusp sites, but the difference was not
significant.
Trap type affected adult A. planipennis captures in 2010. Double-decker traps
accounted for 75 to 80% of the beetles captured in all sites. Double-decker traps captured at
least one beetle on 100%, 100%, and 88% of traps in Core, Crest, and Cusp sites, respectively.
Sticky bands on girdled trees captured 12%, 17%, and 17% of the adults in Core, Crest, and Cusp
sites, respectively. At least one A. planipennis was captured on 69%, 94%, and 44% of sticky
bands in Core, Crest, and Cusp sites, respectively. Sticky bands on control and planted trees
captured < 10% of the adults. Sticky bands captured at least one beetle on 25%, 69%, and 27%
of control trees in Core, Crest, and Cusp sites, respectively, while sticky bands on 21%, 79%, and
6% of planted trees were positive in Core, Crest, and Cusp sites, respectively. The interaction
between trap surface area and invasion stage was not significant (P = 0.99). Surface area
differed among trap types (H = 28.38; df = 3, 216; P < 0.001) (Table 1.1). Area of double-decker
panels was 15-fold greater than the surface area of sticky bands on girdled and control trees,
12

and 25-fold greater than sticky bands on planted trees. The area of sticky bands on girdled and
control trees was nearly 2-fold greater than the area of sticky bands on planted trees. Surface
area of sticky bands on girdled and control trees were similar. Sticky bands on girdled trap
2

trees captured more A. planipennis per m than double-decker traps and sticky bands on
control and planted trees. (F = 12.69; df = 3, 210; P < 0.001) (Table 1.1). Differences in A.
2

planipennis captures per m between double-decker traps and the sticky bands on control and
planted trees were not significant. The interaction between the three invasion stages and trap
2

type on adult captures per m was not significant (P = 0.22).
In 2011, 2504 adults were captured on traps, including 319, 1498, 687 beetles in the
Core, Crest and Cusp sites, respectively. Beetle activity peaked later in 2011 than in 2010.
Beetle captures peaked from 5 July to 15 July, when 1116 A. planipennis were collected,
comprising 45% of the total captures. Captures of A. planipennis in Crest sites were 4-fold
higher than in Core and Cusp sites (F = 17.8; df, 2, 210; P < 0.001) (Fig. 1.2A), where captures
did not differ.
As in 2010, trap type affected adult A. planipennis captures in 2011. Double-decker
traps accounted for 62%, 44%, and 75% of the adults captured in Core, Crest, and Cusp sites,
respectively. At least one beetle was captured on 94%, 100%, and 94% of double-decker traps
in Core, Crest, and Cusp sites, respectively. Sticky bands on girdled trees captured 27%, 42%,
and 18% of adults in Core, Crest, and Cusp sites, respectively. Sticky bands captured at least one
A. planipennis on 81%, 100%, and 69% of girdled trees in Core, Crest, and Cusp sites,
respectively. Sticky bands on control and planted trees captured < 10% of the adults in all sites.
13

At least one beetle was captured on 31%, 62%, and 50% of control trees, and 33%, 71%, and
13% of planted trees in Core, Crest, and Cusp sites, respectively. Area of double-decker panels
was 15-fold greater than surface area of sticky bands on girdled and control trees, 30-fold
greater than sticky bands on planted trees, and the area of sticky bands on girdled and control
trees was 2-fold greater than the area of sticky bands on planted trees (H = 7.52; df = 3, 212; P <
0.001) (Table 1.1). Surface area of sticky bands on girdled and control trees was similar. The
interaction between trap type and invasion stage on trap surface area was not significant (P =
2

0.98) on an area basis. Sticky bands on girdled trees captured nearly 3-fold more adults per m
than sticky bands on planted trees, 4-fold more than sticky bands on control trees, and 6-fold
more than panels on double-decker traps. Sticky bands on planted trees captured twice as
2

many adults per m as panels on double-decker traps (F = 15.4; df = 3, 209; P < 0.001) (Table
2

1.1). Differences in A. planipennis captures per m among other trap types were not significant
The interaction between invasion stage and trap type did not affect captures of adults per m

2

(P = 0.06).
Larval A. planipennis densities: Rectangular areas of bark, hereafter referred to as “windows,”
were debarked on girdled, control, and planted trees in fall 2010 to assess A. planipennis larval
2

density. Area of exposed phloem per tree averaged 0.6 ± 0.04, 0.6 ± 0.04, and 0.9 ± 0.2 m in
Core, Crest, and Cusp sites, respectively, and did not differ among invasion stages (P = 0.93).
Overall, 1818 larvae were recorded in windows debarked on trap trees in 2010, including 441,
902, and 475 larvae on trees in Core, Crest and Cusp sites, respectively. Larvae were found on

14

trees in 22 of the 24 study sites. No larvae were found on trees in the two most westerly sites
in Ottawa County, MI. Larval densities differed among all three invasion stages (F = 17.03; df =
2, 163; P < 0.001) (Fig. 1.2B). Larval A. planipennis density was significantly related to adult A.
planipennis density on traps when all 24 sites were pooled (Figure 1.3A), but relationships
within each invasion stage were not significant [(P = 0.26), (P = 0.30), and (P = 0.18) in Core,
Crest and Cusp sites, respectively].
In 2010, girdled trap trees accounted for 68%, 69%, and 64% of all larvae recorded in
Core, Crest, and Cusp sites, respectively, and at least one A. planipennis larva was found in 94%,
100%, and 60% of the girdled trees in Core, Crest, and Cusp sites, respectively. Larvae in
control trees accounted for 22%, 18%, and 33% of all larvae recorded in Core, Crest, and Cusp
sites, respectively, and 44%, 88%, and 40% of control trees had at least one A. planipennis larva
in Core, Crest, and Cusp sites, respectively. Planted trap trees, accounted for only 10%, 13%,
and 1% of all larvae in Core, Crest, and Cusp sites, respectively. One or more A. planipennis
larvae were present on 54%, 63%, and 13% of trees in Core, Crest, and Cusp sites, respectively.
Average DBH of control and girdled trees was more than twice the DBH of planted trees
(H = 28.38; df = 2; 163; P < 0.001) (Table 1.1). Trap tree DBH did not differ between girdled and
control trees. Nor was the interaction between trap type and invasion stage significant (P =
0.99). Densities of larvae on girdled trees were twice as high as on control trees, and 11-fold
higher than on planted trees. Larval densities on control trees were 5-fold higher than on
planted trees (F = 43.04; df = 2, 163; P < 0.001) (Table 1.1). The interaction of trap type and
invasion stage did not affect larval density (P = 0.51).

15

Area of phloem exposed in bark windows in 2011 to assess larval density averaged 0.8 ±
2

0.1, 0.7 ± 0.04, and 0.8 ± 0.03 m per tree in Core, Crest, and Cusp sites, respectively, and was
not affected by invasion stage (P = 0.85). In 2011, 2895 larvae were recorded in all 24 sites,
including 584, 1245, and 1066 on trees, in Core, Crest and Cusp sites, respectively. Trap trees in
Crest sites had higher larval densities than those in Core and Crest sites (F = 7.2; df = 2, 162; P =
0.001) (Figure. 1.2B). Larval densities were higher in Cusp sites compared to Core sites, but the
difference was not significant. Larval A. planipennis density was significantly related to adult A.
planipennis density on traps when all sites were pooled (Figure 1.3B). Within each invasion
2

2

stage, larval density was significantly related to adult captures per m in Core sites (R = 0.71; P
= 0.05), but not in Crest (P = 0.13) or Cusp sites (P = 0.19).
Girdled trees accounted for 71% to 75% of all larvae in all sites, and at least one A.
planipennis larva was found in 88%, 100%, and 88% of girdled trees in Core, Crest, and Cusp
sites, respectively. Control trees accounted for 11%, 12%, and 31% of total larvae found in
Core, Crest, and Cusp sites, respectively, and 69%, 81%, and 63% of control trees had at least
one A. planipennis larva in Core, Crest, and Cusp sites, respectively. Larvae on planted trap
trees accounted for 16%, 17%, and 4% of total larvae found in Core, Crest, and Cusp sites,
respectively, and at least one A. planipennis larva was found on 57%, 75%, and 25% of trees in
Core, Crest, and Cusp sites, respectively. Similar to 2010, DBH of control and girdled trees in
2011 was more than twice that of planted trees (H = 7.52; df = 2; 162; P < 0.001) (Table 1.1), but
did not differ between girdled and control trees. Larval density on girdled trees was 4-fold
greater than on control trees and 10-fold greater than planted trees (F = 39.7; df = 2, 162; P <

16

0.001) (Table 1.1). More larvae were found on control trees than on planted trees, but
differences were not significant. The interaction between trap type and invasion stage on larval
density was not significant (P = 0.16).
Atanycolus cappaerti parasitism: In 2010, 283 A. planipennis larvae were parasitized by A.
cappaerti, including 57, 223, and 3 larvae in Core, Crest, and Cusp sites, respectively. Parasitism
rates by A. cappaerti averaged 12 ± 4.0%, 27 ± 6.0%, and 1.3 ± 1.2% in Core, Crest, and Cusp
sites, respectively. Parasitism rates in Crest sites were higher than in Cusp sites (H = 10.28; df =
2, 21; P = 0.005), but other differences were not significant. Densities of parasitoids averaged
2

1.4 ± 0.6, 7.2 ± 2.4, and 0.1 ± 0.1 parasitoids per m of exposed phloem in Core, Crest, and Cusp
sites, respectively. Average densities of parasitoids recorded on A. planipennis larvae were
higher in Crest sites than Core and Cusp sites (H = 20.42; df = 2, 163; P < 0.001), where densities
did not differ.
Parasitoid density was higher on A. planipennis larvae in girdled trees than on planted
and control trees, and higher on control trees compared to planted trees (H = 42.65; df = 2,
163; P < 0.001). Densities of A. cappaerti in girdled trees were lower in Cusp sites than girdled
trees in Core and Crest sites (Friedman’s F = 9.99; df = 4, 160; P < 0.001), where densities did
not differ. Parasitoids density on A. planipennis larvae averaged 4.7 ± 1.8, 22.2 ± 7.3, and 0.3 ±
0.2 parasitoids per tree on girdled trees in Core, Crest, and Cusp sites, respectively, compared
to 0.6 ± 0.4, 3.1 ± 1.3, and 0.2 ± 0.2 parasitoids in control trees in Core, Crest, and Cusp sites,
respectively. No parasitoids were observed on planted trap trees in 2010.
In 2011, 145 A. cappaerti parasitizing A. planipennis larvae were recorded which was
approximately half as many as had been recorded in 2010, this including 28, 84, and 33
17

parasitoids in Core, Crest, and Cusp sites, respectively. Parasitism rates by A. cappaerti
averaged 5.3 ± 1.2%, 7.7 ± 3.5%, and 2.3 ± 1.1% A. planipennis larvae in Core, Crest, and Cusp
sites, respectively, and did not differ significantly among stages (P = 0.21). Difference in
parasitoid densities on A. planipennis larvae among the three invasion stages were not
2

significant (P = 0.11), and averaged 0.7 ± 0.2, 2.4 ± 1, 0.5 ± 0.3 parasitoids per m in Core, Crest
and Cusp sites, respectively. Girdled trees had significantly higher average densities of
parasitoids on A. planipennis larvae then control and planted trees (H = 17.66; df = 2, 161; P <
0.001). Average densities of A. cappaerti were 1.3 ± 0.9, 2.3 ± 0.8, and 0.4 ± 0.2 parasitoids on
control trap trees, girdled trap trees, and planted trap trees, respectively. The interaction
between trap type and invasion stage on parasitoid densities was not significant (P = 0.89).
Overstory: In 2010, 3353 trees (DBH ≥ 10 cm) were measured, including 1015, 924, and 1414
trees representing 37 species in Core, Crest, and Cusp sites, respectively. Total basal area
2

averaged 17.3 ± 3.8, 11.1 ± 2.4, and 16.9 ± 1.3 m per ha in Core, Crest, and Cusp sites,
respectively, and did not differ among the invasion stages (P = 0.27).
In 2010, 1035 green ash trees (DBH ≥ 10 cm) were recorded, including 216, 376, and 443
trees in Core, Crest and Cusp sites, respectively. Most ash trees were < 30 cm in DBH (88%,
95%, and 87% in Core, Crest, and Cusp sites, respectively). Ash mortality varied among the
three invasion stages (H = 290.77; df = 2, 1032; P < 0.001) (Table 1.3), but not among diameter
classes (P = 0.53). In three of the eight Core sites, 100% of the overstory ash had succumbed to
A. planipennis. In contrast, 70% was the highest ash mortality recorded in any Crest site, and

18

34% was the highest mortality in Cusp sites. Ash killed by A. planipennis were present in all
eight Core sites, six Crest sites, and four Cusp sites.
In 2011, 1054 ash trees were recorded in transects and plots, including 207, 404, and
443 trees in Core, Crest, and Cusp sites, respectively. As in 2010, 85% to 97% were < 30 cm in
DBH. Ash mortality again varied among the three invasion stages (H = 259.55; df = 2, 1,052; P <
0.001) (Table 1.3), but not among diameter classes (P = 0.44). In three Core sites, all overstory
ash was dead, as in 2010. The highest overstory ash mortality recorded was 100%, 85%, and
50% in Core, Crest, and Cusp sites, respectively. Ash trees killed by A. planipennis occurred in
all Core sites, seven Crest sites, and seven Cusp sites.
Total ash basal area (dead and alive) in 2010 averaged 2.3 ± 0.5, 2.7 ± 0.6, and 5.2 ± 1
2

m per ha in Core, Crest, and Cusp sites, respectively. Ash basal area was more than twice as
high in Cusp sites compared to Core sites (F = 4.37; df = 2, 21; P = 0.026); other differences
among stages were not significant. Live ash accounted for 21%, 63%, and 97% of total ash basal
area in Core, Crest, and Cusp sites, respectively, and differed among invasion stages (F = 21.09;
df = 2, 21; P < 0.001) (Fig. 1.4A). Most live ash were < 30 cm in DBH (87%, 97%, and 84% in
Core, Crest, and Cusp sites, respectively). Basal area of dead ash in Core sites was more than 8fold higher than in Cusp sites (F = 4.64; df = 2, 21; P = 0.02) (Fig. 1.4B), but other differences
among stages were not significant. Most dead ash was < 30 cm in DBH (88%, 90%, and 98% in
Core, Crest, and Cusp sites, respectively).
2

Total ash basal area in 2011 averaged 2.2 ± 0.4, 2.7 ± 0.6, and 5.5 ± 0.9 m per ha, in
Core, Crest, and Cusp sites, respectively, and was higher in Cusp sites than in Crest or Core sites

19

(F = 6.69; df = 2, 21; P = 0.006), which did not differ significantly. The proportion of total ash
basal area still alive was 14%, 44%, and 85% in Core, Crest, and Cusp sites, respectively, and
differed among stages (F = 20.72; df = 2, 21; P < 0.001) (Fig. 1.4A). Most live ash (83 to 95%)
were < 30 cm in DBH. Dead ash basal area did not differ among Core, Crest and Cusp sites in
2011 (P = 0.16) (Fig. 1.4B). Most dead ash was < 30 cm in DBH (83%, 96%, and 88% in Core,
Crest, and Cusp sites, respectively).
External signs of A. planipennis: All sites had ash trees with some canopy dieback. Ash trees in
Cusp sites had healthier canopies than ash trees in Core and Crest sites (H = 112.3; df = 2, 713; P
< 0.001) (Table 1.2), where canopy dieback did not differ significantly. Ash canopy dieback did
not differ significantly among diameter classes (P = 0.31). There was no interaction of invasion
stage and diameter class on ash canopy dieback (P = 0.91).
Ash tree canopies in 2011 were again healthier in Cusp sites than Core and Crest sites (H
= 20.41; df = 2, 581; P < 0.001) (Table 1.2), where canopy dieback did not differ significantly.
The apparent reversal in ash canopy decline from 2010 to 2011 in Core and Crest sites reflected
increased ash mortality in these sites in 2011. A number of trees with relatively high canopy
dieback in 2010 did not survive, increasing the proportion of surviving trees with healthier
canopies. Trees measuring 10 to 20 cm DBH had healthier canopies than trees with a DBH of
30.1 to 40 cm DBH (H = 14.00; df = 2, 1,052; P = 0.007), while other differences among size
classes were not significant. There was no significant interaction between invasion stage and
tree size class on ash canopy dieback (P = 0.88).
Fewer ash trees had epicormic shoots in 2010 in Cusp sites than in Core and Crest sites
(H = 81.28; df = 2, 710; P < 0.001) (Table 1.2), which did not differ. Tree diameter class did not
20

affect epicormic shoot abundance (P = 0.88), and the interaction between invasion stage and
tree diameter class was not significant (P = 0.74). In Crest sites, 40% of trees had high levels of
epicormic shoots (> 4 shoots), compared to 29% and 11% of trees in Core and Cusp sites,
respectively. Few trees had low levels of epicormic shoots (1 to 4 shoots) 11%, 6%, and 4% of
trees in Core, Crest, and Cusp sites, respectively.
In 2011, the proportion of trees with epicormic shoots was higher in Core and Crest sites
than in Cusp sites (H = 24.23; df = 2, 581; P < 0.001) (Table 1.2). Tree diameter class did not
affect epicormic shoots (P = 0.30), and the interaction between invasion stage and tree
diameter class was not significant (P = 0.94). Trees with high levels of stump sprouts comprised
58%, 27%, and 22% in Core, Crest, and Cusp sites, respectively. Few trees had low levels of
epicormic shoots, including 5%, 9%, and 6% of trees in Core, Crest, and Cusp sites, respectively.
The proportion of live and dead ash trees in 2010 with stump sprouts was higher in Core
and Crest sites than in Cusp sites (H = 157.49; df = 2, 958; P < 0.001) (Table 1.2). Tree diameter
class did not affect stump sprout presence (P = 0.97), and the interaction between invasion
stage and tree diameter class was not significant (P = 0.62). Trees with high levels of stump
sprouts (> 4 sprouts) comprised 38%, 40%, and 6% of trees in Core, Crest, and Cusp sites,
respectively. Trees with low levels of stump sprouts (1 to 4 sprouts) comprised 6%, 7%, and 3%
of trees of trees in Core, Crest, and Cusp sites, respectively.
The proportion of live and dead ash trees in 2011 with stump sprouts was higher in Core
and Crest sites than in Cusp sites (H = 157.49; df = 2, 977; P < 0.001) (Table 1.2). Tree diameter
class did not affect stump sprout presence (P = 0.93), and the interaction between the three
stages and tree diameter class was not significant (P = 0.92). Trees with high levels of stump
21

sprouts comprised 39%, 53%, and 12% of trees in Core, Crest, and Cusp sites, respectively. Few
trees had low levels of stump sprouts (1 to 4 sprouts), including 9%, 8%, and 6% of trees in
Core, Crest, and Cusp sites, respectively.
Live and dead ash trees with holes left by woodpeckers preying on late instar larvae in
2010 were more abundant in Core sites compared to Cusp sites (H = 259.61; df = 2, 958; P <
0.001) (Table 1.2), but all other differences were not significant. Tree diameter class was not
related to woodpecker predation on late instar larvae (P = 0.11), and the interaction between
the invasion stage and tree diameter class was not significant (P = 0.10). Trees with > 6
woodpecker attacks comprised 73%, 56%, and 13% of trees in Core, Crest, and Cusp sites,
respectively. Trees with 1 to 6 woodpecker holes were rare, accounting for 4%, 6%, and 3% of
trees in Core, Crest, and Cusp sites respectively.
The proportion of ash trees with woodpecker predation increased in all three stages
from 2010 to 2011. Woodpecker attacks were higher in Core sites than Cusp sites (H = 249.28;
df = 2, 977; P < 0.001) (Table 1.2), but other differences were not significant. Tree diameter
class did not affect the abundance of woodpecker predation on late instar larvae (P = 0.42), and
the interaction invasion stage and tree diameter class was not significant (P = 0.99). Trees with
high levels of woodpecker predation comprised 88%, 75%, and 30% of trees in Core, Crest, and
Cusp sites, respectively. Few trees had low levels of woodpecker predation, including 3%, 6%,
and 7% of trees in Core, Crest, and Cusp sites, respectively.

22

DISCUSSION
Effects of A. planipennis in stands of green ash are similar to the effect of major forest
pathogens such as Cryphonectria parasitica (formerly Endothia parasitica, formerly Diaportha
parasitica) which caused chestnut blight in American chestnut (Castanea dentata). Similarly,
Ophiostoma ulmi (prior to 1940s) and O. novo-ulmi (post 1940s) vectored by both exotic and
native bark beetles caused Dutch elm disease in American elm (Ulmus americana) (Braier and
Buck 2001). Both chestnut blight and Dutch elm disease killed healthy trees, spread across the
natural ranges of their host trees in a matter of decades, and have caused near 100% mortality
of overstory American chestnut and American elm (Anagnostiakis 1987; Bey 1990;
Anagnostiakis and Hillman 1992; Brasier 2000). Like these fungal pathogens, A. planipennis has
become established in 16 states and two provinces of Canada (EAB.info 2012), attacks and kills
healthy ash trees (Poland and McCullough 2006), and ash mortality rates approach 100%
(Gandhi et al. 2008). By comparison, the invasive insect Anoplophora glabripennis
(Motschulsky), while capable of killing healthy trees (Haack et al. 2010), has not spread nearly
as quickly as A. planipennis. First discovered in New York in 1996, infestations of A. glabripennis
are currently established in four additional states (National Invasive Species Information Center
2012), and multiple A. glabripennis infestations have been successfully eradicated (National
Invasive Species Information Center 2012). Another invasive insect, Adelges tsugae (Annand),
introduced into Virginia in the 1950s (Souto et al. 1996), has become a major pest of eastern
hemlock trees (Tsuga canadensis) in the eastern United States. Mortality rates for eastern
hemlock ranges from 0 to 95% of trees within stands, and are related to stand size, aspect, and
local climate (Orwig and Foster 1998; Orwig et al. 2002). In contrast, mortality caused by A.
23

planipennis varies little among stands and approaches 100% of overstory trees (Gandhi et al.
2008; Knight et al. 2008).
My results indicate A. planipennis persists at low densities in Core sites. The number of live
ash stems and basal area was significantly lower in Core sites than in Crest and Cusp. Reduced
A. planipennis adult captures in Core sites compared to Crest sites presumably reflects reduced
numbers of live overstory ash trees, causing A. planipennis populations to expand outward
from the original infestation sites in southeast Michigan (Figure 1.5). Previous studies found A.
planipennis did not develop on non-ash hosts in North America (Anulewicz et al. 2006;
Anulewicz et al. 2008), so beetle populations will presumably decline as their host resource is
depleted (Figure 1.5). Low density populations can be susceptible to Allee effects, which can
lead to negative population growth rates (Courchamp et al. 1999) and eventual extinction
(Liebhold and Tobin 2008). Mortality factors such as parasitism by A. cappaerti, and perhaps
woodpecker predation, could further reduce local A. planipennis populations (Cappaert et al.
2005; Lindell et al. 2008; Tluczek 2009; Cappaert and McCullough 2009) (Figure 1.5). Ash trees
in Core sites, however, are still becoming infested. It is likely A. planipennis will persist as long
as live ash trees remain.
Despite the persistence of A. planipennis, and high levels of ash mortality, live overstory ash
trees were still present in Core sites. It is unknown why some ash trees have survived, but the
patchy distribution of A. planipennis colonization may reflect beetle dispersal patterns. Agrilus
planipennis do not disperse randomly. Ash phloem abundance (Mercader et al. 2009; Siegert et
al. 2010), and the presence of stressed ash trees (McCullough et al. 2009a, 2009b; Tluczek et al.
2011) affect A. planipennis dispersal in sites with low densities (Figure 1.5). Not every tree will
24

be colonized simultaneously. While A. planipennis populations are lower in Core sites
compared to Crest sites, beetles were captured in every Core site, and ash canopy dieback did
not differ between these two invasion stages. Ash canopy dieback does not become apparent
until trees are moderately to heavily infested, and often culminate in death of the host (Poland
and McCullough 2006). Future studies to monitor ash and A. planipennis populations over time
are needed.
My results indicate A. planipennis populations were peaking in Crest sites in 2010 and 2011.
Despite high densities of A. planipennis, on average > 50% of ash trees in Crest sites were still
alive, so it is likely A. planipennis populations will remain near peak densities until ash
populations are further depleted. Parasitism by A. cappaerti, and woodpecker predation on A.
planipennis larvae, do not appear to reduce A. planipennis populations sufficiently to prevent
new trees from being infested, so ash mortality is likely to rise in Crest sites. As ash mortality
escalates, populations of A. planipennis and green ash will likely be reduced to levels currently
observed in Core sites (Figure 1.5).
Given the abundance of available resources in Cusp sites, A. planipennis populations can be
expected to grow (McCullough and Siegert 2007) (Figure 1.5). Adult captures and larval
densities increased substantially in Cusp sites from 2010 to 2011. In five of the eight Cusp sites,
more adult A. planipennis were captured on a single trap over a two week period in 2011 than
were captured on all traps combined in those sites during the 2010 season. If no efforts are
made to reduce A. planipennis population growth, patterns observed in southeast Michigan are
likely to be repeated in Cusp sites, and across the range of green ash in the United States.

25

There are 56 known species of parasitoids In North America that attack Agrilus larvae
(Taylor et al. 2012), but Atanycolus cappaerti is emerging as an important mortality factor for A.
planipennis (Cappaert and McCullough 2009) (Figure 1.5). The mechanism(s) by which A.
cappaerti locates A. planipennis larvae are unknown. Many parasitoids are capable of detecting
plant volatiles produced as an induced response to insect attack (Stowe et al. 1995, Gols and
Harvey 2009). Trees stressed from A. planipennis feeding also emit volatiles from their foliage
(Rodriguez-Saona et al. 2006; de Groot et al. 2008; Grant et al. 2010) and bark (Crook et al.
2008) that are attractive to A. planipennis (McCullough et al. 2009a, 2009b; Siegert et al. 2010;
Tluczek et al. 2011). It seems likely that A. cappaerti, which evolved to prey on native phloemfeeding insects, is able to detect volatiles emitted from ash trees stressed by feeding A.
planipennis larvae. This is supported by my data, as the majority of parasitoids were found on
girdled trees. Whether A. cappaerti parasitism rates will increase and what impact they may
have on A. planipennis populations remains unknown, but I found no evidence in the literature
of a phloem-feeding insect controlled by a parasitoid.
The decline in parasitism between 2010 and 2011 may have been caused by the removal of
trap trees with parasitoids from the study sites. If A. cappaerti is attracted to girdled trees,
whether to plant volatiles or an increase of A. planipennis larval density, felling and debarking
trees in 2010 may have reduced their numbers. Future research on A. cappaerti is needed to
document the biology of the parasitoid, parasitism rates on A. planipennis larvae, prey finding
behavior of A. cappaerti, and options for possible augmentative biocontrol.
Woodpeckers are also continuing to cause mortality of A. planipennis larvae in southern
Michigan (Figure 1.5). In my research, holes left by woodpecker predation on larvae were
26

found on most ash trees in Core and Crest sites, and nearly half of the ash trees in Cusp sites by
2011. The accumulation of holes left by woodpecker attacks suggests A. planipennis
populations are being reduced by woodpeckers (Figure 1.5), but not every infested tree is being
targeted. Studies have shown there is a positive correlation between A. planipennis larval
density and woodpecker predation, but woodpecker predation can also be patchy (Cappaert et
al. 2005; Lindell et al. 2008; Tluczek 2009). This is consistent with my findings that woodpecker
attacks were more common on trees in Core sites where A. planipennis populations have been
present longer than in Crest and Cusp sites, but were not present on all trees. If native
woodpeckers are adapting to A. planipennis, and woodpecker populations increase due to
capitalization of the resource, larval mortality rates caused by woodpecker predation may
increase. However, to date, woodpecker predation on A. planipennis larvae has not stopped
the spread of the beetle, nor has it sufficiently reduced A. planipennis populations in Core sites
to prevent the infestation of remaining ash trees.
Effective methods to detect and monitor A. planipennis infestations are an important aspect
of managing this invasive pest. My study suggests trap trees and traps are a better detection
tool than visual surveys. Despite the increasing of A. planipennis populations in Cusp sites in
2011, most trees still appeared healthy with few sprouts or holes left by woodpecker predation.
This is consistent with previous studies reporting external signs of A. planipennis rarely appear
until a tree becomes heavily infested (Cappaert et al. 2005; Poland and McCullough 2006).
The question then arises as to what is the best trap for detecting A. planipennis? In my
study, adult A. planipennis were more attracted to girdled trap trees than other trap types, in
both 2010 and 2011. This is consistent with other studies conducted on A. planipennis in North
27

America which found girdled trees were more attractive than trees stressed by other injury or
baited with attractive volatiles (McCullough et al. 2009a, 2009b; McCullough et al. 2011;
Tluczek et al. 2011). Although higher densities of beetles were captured on sticky bands of
girdled trees, double-decker traps consistently were more successful at capturing at least one
beetle and had the highest total A. planipennis captures. These findings are consistent with a
study evaluating various trap types for A. planipennis (McCullough et al. 2011). Based on
surface area, girdled trees are more attractive, but girdling additional trees is destructive and
wrapping a larger proportion of the trunk is not practical. Furthermore, in Cusp sites in 2010,
when beetle populations were at their lowest density, at least one double-decker trap captured
an adult A. planipennis in all eight sites, but larvae in girdled trees were only found in six sites,
indicating that double-decker traps were better at detecting A. planipennis at low population
densities than girdling and debarking trees. As a detection tool, it is not density of beetle
captures, but rather the successful capture of at least one A. planipennis that is important.
Since double-decker traps are larger, more apparent, and captured the most A. planipennis of
all the traps used in this research, they have the highest probability of detection, and would be
the optimal trap for detecting A. planipennis in front of the invasion wave.
The proportion of total beetles captured on double-decker traps in Core and Crest sites
dropped substantially from 2010 to 2011. Reduced captures may be partially due to the change
in lures from a Manuka/Phoebe oil blend in 2010 to Manuka oil alone in 2011. A previous study
which traps baited with Manuka/Phoebe oil had significantly higher capture rates compared to
traps baited with just Manuka oil (Crook et al. 2008). This does not explain, however, why the
proportion of total beetle captures on double-decker traps in Cusp sites remained constant
28

between 2010 and 2011. The study comparing Manuka oil and Phoebe oil by Crook et al.
(2008) was conducted on the edges of infested ash stands. Rodriguez-Saona et al. (2006) found
that stressed ash emitted elevated levels of stress volatiles eliciting antennal responses and
were attractive to A. planipennis. Volatiles emitted from an increased number of stressed trees
may, however, mask lures on traps. In my study, few trees in Cusp sites were showing signs of
stress, and A. planipennis populations were low. My findings indicate that at low A. planipennis
densities, where few trees are emitting stress volatiles, traps baited with a Manuka/Phoebe oil
blend may not have a higher capture rate than traps baited with Manuka oil alone. Phoebe oil
contains the compound 7-epi-sesquithujene, which Manuka oil lacks (Crook et al. 2008). The
presence of 7-epi-sesquithujene may reduce the degree to which volatiles emitted from
stressed trees mask lures on artificial traps. Crest sites had the highest number of stressed
trees, which could explain the drop in effectiveness of double-decker traps in these locations.
Ash trees in Core sites are stressed, but few trees remain to emit stress volatiles compared to
Crest sites. Further studies to determine if lures become less effective as more trees become
stressed would be useful.
Natural enemies, such as woodpeckers and the braconid parasitoid A. cappaerti can
cause substantial mortality of A. planipennis (Cappaert et al. 2005; Lindell et al. 2008; Cappaert
and McCullough 2009; Tluczek 2009), but to date, have not prevented the spread and
successful colonization of new ash stands. Infested trees decline rapidly as A. planipennis
populations build; mortality rates, for example, nearly doubled in Cusp sites from 2010 to 2011.
Therefore, trapping to detect low density populations in front of the leading edge of the
invasion wave would be useful for timely implementation of control tactics to minimize ash
29

mortality. If A. planipennis populations are not reduced and prevented from colonizing new
locations, then trends in ash mortality observed in Core sites will be repeated in Crest and Cusp
sites, and in green ash stands across its range.

30

31

32

Figure 1.1. Locations of 24 green ash sites representing three stages of the A.
planipennis invasion wave in 2010 and 2011. Core sites were characterized by low
densities of A. planipennis and high ash mortality. Crest sites were characterized
by high densities of A. planipennis and moderate to high ash mortality. Cusp sites
were characterized by low densities of A. planipennis and low ash mortality.

33

2

Figure 1.2. A. Mean (± SE) captures of A. planipennis adults per m of trap
area in Core, Crest, and Cusp sites in 2010 and 2011. B. Mean (± SE) density
2

of A. planipennis larvae per m of exposed ash phloem in Core, Crest, and
Cusp sites in 2010 and 2011. Means with the same letter are not significantly
different (Tukey's protected HSD test; P < 0.05). (a, b, and c for 2010; and y
and z for 2011)

34

2

Figure 1.3 Linear regression of larval A. planipennis density per m of exposed phloem
2

and adult A. planipennis density per m on traps in (A.) 2010 and (B.) 2011 at 24 sites in
southern Michigan (N = 24).

35

A.

B.

2

Figure 1.4. A. Mean (± SE) live basal area (m per ha) of green ash in Core,
Crest, and Cusp sites in 2010 and 2011. B. Mean (± SE) dead basal area
2

(m per ha) of green ash in Core, Crest, and Cusp sites in 2010 and 2011.
Means with the same letter are not significantly different (Tukey’s
protected HSD test; P < 0.05) (a, b, and c for 2010; x, y, and z for 2011).

36

37

Chapter 2

Green ash Overstory and Regeneration Ahead, Within, and Behind the Emerald
Ash Borer Invasion Wave

INTRODUCTION
Since its introduction in the early to mid-1990s (Siegert et al. 2007), emerald ash borer,
(Agrilus planipennis Fairmaire) (Coleoptera: Buprestidae) has killed tens of millions of ash trees
(Fraxinus spp.) in the United States and Canada (EAB.info 2012). A. planipennis preferentially
feeds on green ash (Fraxinus pennsylvanica)(Anulewicz et al. 2007; Rebek et al. 2008) and has
caused widespread mortality of this species in eastern North America. Green ash, with the
widest distribution of all North American Fraxinus spp., is native to all states east of the Rocky
Mountains, as well as seven provinces of Canada (Wright 1959; Kennedy 1990). Green ash is
typically found in bottomlands, but can survive in upland sites (Wright 1959; Kennedy 1990). In
southern Michigan, green ash is widespread, particularly in lowlands and riparian zones (Barnes
and Wagner 1981), and is common in areas prone to frequent flooding (Kennedy 1990). With
the introduction and spread of A. planipennis, the future of green ash within these systems is
uncertain.
Ash comprises a substantial proportion of Michigan’s hardwood forest resource.
According to the USDA Forest Inventory and Analysis (FIA database 2012), there were 869
million Fraxinus spp. trees with diameter at breast height (DBH) ≥ 3 cm in Michigan as of 2011.

38

Female A. planipennis oviposit individual eggs in cracks in the outer bark, or beneath
rough edges of ash bark from mid to late summer. Eggs hatch within two weeks, and larvae
feed in the cambium and phloem, constructing serpentine galleries (Poland and McCullough
2006). Larval A. planipennis pass through four instars and typically complete feeding in late fall,
then excavate pupal chambers in the outer sapwood or bark (Cappaert et al. 2005). Over time,
as densities of larvae increase, feeding disrupts the translocation of nutrients, resulting in
canopy thinning or dieback. Epicormic shoots may appear on branches or trunks, and stump
sprouts may be produced by dying trees. As ash trees within stands die, beetles disperse and
colonize new locations. The demise of ash could cascade through the ecosystem, affecting
species in multiple trophic levels. To date, classical biocontrol, silvicultural techniques, and
predation by native predators and parasitoids have failed to stop A. planipennis from killing
tens of millions of ash trees.
Agrilus planipennis has the potential to drastically alter forest composition by virtually
eliminating ash as a major overstory component, and allowing shade intolerant species to
increase in the overstory. Classical theory on successional trajectories projects that in a system
with minimal disturbances, shade-tolerant species will grow to dominate stands (Kobe et al.
1995; Catovsky and Bazzaz 2000). Large canopy gaps formed by mortality of overstory ash tree,
however, result in increased light reaching the forest floor, providing an opportunity for shade
intolerant species to compete with more tolerant species (Messier et al. 1999).
Many questions arise regarding the long-term impact of A. planipennis in stands. How
does overstory and understory species composition vary with invasion stage? Does ash

39

continue to persist and regenerate in stands invaded by A. planipennis? What species will likely
replace ash in the overstory?
The goal of this study was to characterize ash and other species in the overstory and
understory in stands representing three stages of the A. planipennis invasion wave. The first
stage, classified as the “Core”, includes stands within the original infestation zone in southeast
Michigan. Most overstory ash trees in Core stands have succumbed, creating substantial
canopy gaps. The second stage of the invasion wave, designated as the “Crest,” represents
stands in south central Michigan where A. planipennis populations are near peak densities and
ash trees are in various stages of decline. The third stage of the invasion wave, termed “Cusp,”
includes stands in southwest Michigan where ash appear healthy with few external signs of A.
planipennis infestation. Results should provide insight into 1) current green ash populations at
three stages of the A. planipennis invasion wave; 2) the potential of green ash to persist in
Michigan forests; and 3) potential future stand composition.

40

MATERIALS & METHODS
Study sites: Twenty-four sites, each 1 ha in size, were selected in July and August 2009. Sites
were located on an east to west gradient across southern Michigan (Figure. 1.1) on state,
county, and municipal property, and were second growth stands with dominant trees between
60 and 90 years of age (Michigan.gov 2012). The overstory in each site (pre-A. planipennis) was
comprised of a minimum of 20% green ash. Sites were classified as Core, Crest, or Cusp sites
based on ash tree condition at the time of selection. Eight Core sites, primarily in southeast
Michigan were selected. In Core sites, most overstory ash were dead, creating large gaps.
Lateral growth from neighboring non-ash trees had often extended into these gaps. To ensure
A. planipennis was the cause of ash mortality, bark was removed from a subsample of dead ash
to identify the characteristic serpentine galleries left by A. planipennis larvae. Eight Crest sites
were selected in south central Michigan in areas where ash trees were in various stages of
decline. Large gaps were present but neighboring trees had not extended their branches into
the gaps. Dead ash trees in Crest sites were also examined to ensure mortality was caused by
A. planipennis. Eight Cusp sites, primarily in southwest Michigan, were selected in areas where
overstory ash had little or no signs of A. planipennis infestation, and no canopy gaps caused by
ash mortality were present.
Overstory: All trees (DBH ≥ 10 cm) were identified to species and DBH measured in June and
July 2010 along two 150 x 2 m belt-transects that ran diagonally across each site in an Xformation. Trees within 10 m of the intersection of the belt-transect were marked to prevent
multiple measurements of individual trees. The two belt-transects divided each site into four
quadrats. One macro-plot (18 m radius) was established in each quadrat. Trees within macro41

plots were identified to species and DBH recorded. Canopy dieback was recorded to assess A.
planipennis-related injury to ash trees, and to estimate the relative health of non-ash trees at
the site. Visual estimates of dieback were recorded in increments of 10%; 0% indicated a full
canopy and 100%, indicated the complete absence of live branches (Zarnoch et al. 2004).
Dieback was assessed in early to mid-summer of each year, after trees were fully flushed but
before current-year larval feeding could further affect ash overstories. Data from belt-transects
and macro-plots were used to calculate relative importance values (RIV) for overstory trees in
each site. The index is derived using relative frequency (representing the sum of occurrences
within plots of a species as a proportion of an individual species related to all species), plus
relative density (count of individuals of a species as a percentage of the total count of
individuals of all species), plus relative dominance (basal area of a given species as a percentage
of the total basal area of all species) (Kent and Coker 1992).
Overstory ash trees were assessed again in belt-transects and macro-plots in 2011. Nonash density and basal area presumably remained relatively constant between years, and were
not surveyed in 2011.
Recruits: Recruits included trees ≥ 3 cm and < 10 cm in diameter. Recruits were identified to
species in subplots with a 7 m radius located at the center of each macro-plot in 2010. Total
number of green ash recruit stems per ha were calculated. Green ash stem densities and
mortality attributed to A. planipennis did not differ among invasion stages, so recruits were not
resurveyed in 2011.
Saplings: Trees < 3 cm in diameter and ≥ 45 cm in height were classified as saplings. Sapling
density by species was recorded in the same subplots (7 m radius) used for recruits. Total
42

number of stems and green ash stems per ha were calculated. Presence of current-year deer
browse on green ash saplings was also recorded.
PAR and sapling growth: Ash saplings in subplots were counted again in 2011 and
photosynthetically active radiation (PAR) associated with individual ash saplings was measured.
Readings were recorded with a LI-COR 250A (LI-COR Inc. Lincoln, NE) in four randomly selected
Core, Crest, and Cusp sites (12 sites total). Light readings were taken after canopies were fully
flushed, but before injury caused by current-year larvae could further affect ash canopies.
Readings were conducted in one site per day, three sites per week, beginning 20 June and
concluding 15 July. A total of 37 readings over 1 to 2 m tall green ash saplings were collected
systematically around fixed radius plots and belt-transects across each site. Growth of
individual saplings was determined by measuring the distance between terminal bud scars of
the top segment on the terminal leader of each plant. Sapling growth was measured beginning
2

12 June and concluded 17 June. If an ash sapling was not present within a 2.5 m radius of a
selected location, a PAR reading was collected without an associated sapling.
Light readings began at dawn and were completed by sunrise to reduce variability from
random sunflecks penetrating the canopy. Adjustments for dawn and sunrise times by date
and location were determined by GAISMA.com (2011).Readings were 15 second averages.
While each reading was taken within the site, a paired reading was taken simultaneously in an
adjacent open field (synchronized by two-way radios). Readings within sites were divided by
the simultaneous open field reading to determine percentage of sun penetrating the canopy. If
PAR in the open field was zero, the reading within the stand was dropped from analysis. All

43

light readings, with and without an associated ash sapling, were used to calculate average PAR
levels in the understory of each site.
Seedlings: In 2010, seedlings < 45 cm in height were tallied by species in 1.8 m radius microplots, located at the center of each of the four subplots in each site. Total number of seedlings
and green ash seedling per ha were calculated. Ash seedlings were examined for cotyledons to
determine if they represented current-year germination. Ash seedlings, with and without
cotyledons, were counted again in 2011.
Statistical analysis: Variables recorded in plots, transects, and PAR readings were used to
assess species composition, basal areas, stems per ha, and sapling growth among the three
stages of invasion. Data were tested for normality using the Shapiro-Wilk test (Shapiro and
Wilk 1965) and residual plots. Several variables, including total basal area, ash basal area, PAR,
and green ash sapling growth were normalized by log10 (x + 1) transformations. Differences in
basal area, PAR, and green ash sapling growth among the three invasion stages were tested as
unplanned comparisons, and Tukey’s honestly significant difference (HSD) multiple comparison
test was applied if the overall analysis of variance (ANOVA) was significant (P < 0.05) (PROC
GLM; SAS Institute 2003) (Ott and Longnecker 2001). Regression analysis was used to evaluate
the linear relationship between PAR and green ash sapling growth (PROC CORR; SAS Institute
2003) (Ott and Longnecker 2001). Density of green ash recruits, saplings, seedlings, and ash
seedlings with cotyledons were not normalized by transformations. The nonparametric KruskalWallis statistic was applied to assess differences in these response variables among the three
invasion stages (Kruskal and Wallis 1952) (PROC NPAR1WAY; SAS Institute 2003). When results
were significant (P < 0.05), a nonparametric multiple comparison was applied (Zar 1984). The
44

five most abundant species of recruits, saplings, and seedlings based upon stems per ha were
ranked in each of the three invasion stages. All analyses were conducted at P < 0.05 level of
significance using SAS statistical software.

45

RESULTS
Overstory: In 2010, 3353 overstory trees (DBH ≥ 10 cm) (all species) were measured in the 24
sites including 1015, 924, and 1414 trees in Core, Crest, and Cusp sites, respectively. Most trees
were ≤ 30 cm in DBH (71%, 83%, and 87% in Core, Crest, and Cusp sites, respectively). Total
2

basal area averaged 17.3 ± 3.8, 11.1 ± 2.4, and 16.9 ± 1.3 m per ha in Core, Crest, and Cusp
sites, respectively. Live basal area of all species averaged 14.9 ± 3.4, 9.6 ± 2.4, and 16.2 ± 1.3
2

m per ha in Core, Crest, and Cusp sites, respectively. Differences among the three invasion
stages did not affect total basal area (P = 0.26), live basal area (P = 0.18), nor dead basal area (P
= 0.26). Of the 3353 trees recorded across all sites, 377 were dead. Of the 377 dead trees, 67%
were ash, 19% were American elm (Ulmus americana), and 14% were comprised of other
species. Galleries of the native elm bark beetle (Hylurgopinus rufipes (Eichhoff)) and the
smaller European elm bark beetle (Scolytus multistriatus (Marsham)) were found beneath the
bark of dead elm trees.
Overall, 1035 green ash trees with DBH ≥ 10 cm were recorded in 2010 including 216,
376, and 443 trees in Core, Crest, and Cusp sites, respectively, and 87% to 95% of ash in all sites
were ≤ 30 cm in DBH. Ash basal area (live plus dead ash) in 2010 averaged 2.3 ± 0.5, 2.7 ± 0.6,
2

and 5.2 ± 1 m per ha in Core, Crest, and Cusp sites, respectively, and differed among invasion
stages (F = 4.37; df = 2, 21; P = 0.03). Total ash basal area was twice as high in Cusp sites
compared to Core sites; other differences among stages were not significant. Live ash basal
2

area in 2010 averaged 0.5 ± 0.2, 1.7 ± 0.3 and 5.0 ± 1.1 m per ha, in Core, Crest and Cusp sites,
respectively. Live ash basal area in Cusp sites was 10-fold higher than in Core sites, and nearly
46

3-fold higher than in Crest sites, while live ash basal area in Crest sites was 3-fold higher than in
Core sites (F = 21.09; df = 2, 21; P < 0.001). Ash mortality attributed to A. planipennis in 2010
averaged 66.7 ± 11.0%, 26.7 ± 8.7%, and 11.1 ± 3.9% percent in Core, Crest, and Cusp sites,
respectively, and differed among invasion stages (H = 290.77; df = 2, 1032; P < 0.001). Dead ash
2

basal area averaged 1.7 ± 0.5, 1.1 ± 0.4 and 0.2 ± 0.1 m per ha, in Core, Crest and Cusp sites,
respectively, and also differed among invasion stages (F = 4.64; df = 2, 21; P = 0.02). Dead ash
basal area was 8-fold higher in Core sites compared to Cusp sites, but other differences among
stages were not significant.
In 2011, 1054 overstory green ash trees were tallied when sites were re-surveyed,
including 207, 404, and 443 trees in Core, Crest, and Cusp sites, respectively. As in 2010, 85%
to 97% of ash trees were ≤ 30 cm in DBH. Total ash basal area in 2011 averaged 2.2 ± 0.4, 2.7 ±
2

0.6, and 5.5 ± 0.9 m per ha, in Core, Crest, and Cusp sites, respectively. Total ash basal area
was more than twice as high in Cusp sites compared to Core and Crest sites (F = 6.69; df = 2, 21;
P = 0.006), which did not differ. Live ash basal area averaged 0.3 ± 0.1, 1.2 ± 0.3 and 4.7 ± 1 m

2

per ha, in Core, Crest and Cusp sites, respectively. Live ash basal area in Cusp sites was 15-fold
greater than in Core sites, nearly 4-fold greater than in Crest sites, while live ash basal area in
Crest sites was 4-fold greater than in Core sites (F = 20.72; df = 2, 21; P < 0.001). Dead ash basal
2

area averaged 2.0 ± 0.4, 1.6 ± 0.5 and 0.8 ± 0.4 m per ha, in Core, Crest and Cusp sites,
respectively, and did not differ among invasion stages (P = 0.161). In 2011, on average, 78.6 ±
10.1%, 44.8 ± 11.2%, and 19.8 ± 7.3% of overstory trees were killed by A. planipennis in Core,

47

Crest and Cusp sites, respectively, and ash mortality differed among all three invasion stages (H
= 259.55; df = 2, 1,051; P < 0.001).
Canopy dieback: Canopy dieback of non-ash trees in 2010 was low in all sites, averaging < 5% in
all three invasion stages. Ash trees, however, averaged 40.2 ± 4.7%, 34 ± 2.3%, and 11.0 ±
1.2% dieback in Core, Crest, and Cusp sites, respectively, and dieback differed among invasion
stages (H = 112.3; df = 2, 713; P < 0.001). Ash canopy dieback in Core sites was nearly 4-fold
greater than in Cusp sites, and 3-fold greater in Crest sites compared to Cusp sites Differences
in ash canopy dieback between Core and Crest sites were not significant.
Canopy dieback on ash trees in 2011 averaged 38.5 ± 6.0%, 28.4 ± 2.7%, and 19.9 ± 1.7%
in Core, Crest, and Cusp sites, respectively, and again differed among invasion stages (H =
20.41; df = 2, 581; P < 0.001). Canopy dieback was nearly twice as high in Core sites compared
to Cusp sites, and higher in Crest sites than in Cusp sites, but did not differ between Core and
Crest sites.
A total of 37 overstory species were recorded in the 24 sites in 2010. The five most
abundant species in the three invasion stages, based upon RIV, were similar in all sites (Table
2.1). Green ash, American elm, and black cherry (Prunus serotina) ranked in the top five in all
three invasion stages (Table 2.1). Red oak (Quercus rubra) and black walnut (Juglans nigra)
were relatively abundant in Cusp sites, and red maple (Acer rubrum) and sugar maple (Acer
saccharum) ranked highly in the Core and Crest sites.
Recruits: In 2010, 913 recruits representing 28 species were recorded, including 246, 321, and
346 recruits in Core, Crest, and Cusp sites, respectively. Densities of recruits averaged 499 ± 95,
655 ± 160, 702 ± 92 stems per ha in Core, Crest, and Cusp sites, respectively, and did not differ
48

among invasion stages (P = 0.06). The five most abundant recruit species based on stems per ha
were similar to overstory species composition (Table 2.2). Red oak recruits, however, were only
abundant in Core sites, were scarce in Crest sites, and were not observed in Cusp sites. Black
walnut recruits were represented in Core and Crest sites, but not in Cusp sites, and red maple
recruits were not observed in Crest sites.
Overall, 315 green ash recruits were tallied in 2010, including 63, 174, and 78 recruits
comprising 34%, 64%, and 29% of all recruits in Core, Crest, and Cusp sites, respectively.
Densities of total ash recruits averaged 120 ± 29, 353 ± 90, and 158 ± 38 stems per ha in Core,
Crest, and Cusp sites, respectively, and did not differ among invasion stages (P = 0.15). Ash
recruits were recorded in 23 of the 24 study sites, but were not encountered in one Cusp site
(Barry County). Of the 315 green ash recruits, however, 63 were dead. On average, 29.2 ±
16%, 21.9 ± 9.5%, and 30.3 ± 11.1% of ash recruits in Core, Crest, and Cusp sites, respectively,
were dead, and differences among invasion stages did not differ (P = 0.98). Only 31 of the 63
dead ash recruits had A. planipennis galleries beneath the bark. On average, 7.6 ± 5.6%, 12.9 ±
7.6%, and 9.7 ± 8.3% of ash recruits were killed by A. planipennis in Core, Crest, and Cusp sites,
respectively, and this mortality did not differ among invasion stages (P = 0.06).
Saplings: Overall, 2931 saplings representing 25 species were counted in 2010, including 1337,
1112, and 474 saplings in Core, Crest, and Cusp sites, respectively. Sapling densities averaged
2693 ± 354, 2273 ± 456 and 962 ± 235 stems per ha in Core, Crest, and Cusp sites, respectively,
and differed among stages (H = 19.37; df = 2, 93; P < 0.001). Sapling densities were twice as
high in Core and Crest sites compared to Cusp sites, but differences between Core and Crest
sites were not significant.
49

Saplings were recorded in all subplots in Core sites, while 6% and 16% of the 32 subplots
in each of the Crest and Cusp sites had no saplings, respectively. Species composition of
saplings was similar to that of overstory trees and recruits (Table 2.2), except that red oak and
black walnut saplings were only observed in Core sites. Red maple was abundant in Core and
Cusp sites, but was infrequent in Crest sites.
In 2010, 2133 green ash saplings were tallied, including 967, 911, and 255 saplings
accounting for 72%, 81%, and 54% of total sapling counts in Core, Crest, and Cusp sites,
respectively. Ash saplings were recorded in subplots at 22 of the 24 study sites; none were
recorded in two Cusp sites (Barry Co., Ottawa Co.). Densities of ash saplings were 3-fold greater
in Core and Crest sites compared to Cusp sites, but did not differ between Core and Crest sites
(H = 23.62; df = 2, 93; P < 0.001) (Figure 2.2). On average 14.4 ± 3.7%, 6.4 ± 2.5%, and 12.4 ±
4.3% of ash saplings had evidence of deer browse in Core, Crest, and Cusp sites, respectively,
but deer browse did not differ among invasion stages (P = 0.23).
In 2011, 1905 green ash saplings were recorded, including 892, 855, and 158 saplings in
Core, Crest, and Cusp sites, respectively. Ash saplings were again absent in the same two sites
as in 2010. Ash sapling densities were 6-fold higher in Core and Crest sites compared to Cusp
sites, but did not differ between Core and Crest sites (H = 27.90; df 2, 93; P < 0.001) (Figure 2.2).
Deer browsed on an average of 18.1 ± 3.8%, 13.2 ± 3.4%, of ash saplings in Core and Crest
respectively, which was higher than Cusp sites where 2.6 ± 1.5% of ash saplings were browsed
(H = 20.81; df = 2, 29; P < 0.001).

50

PAR and sapling growth: A total of 428 PAR readings were collected including 146, 144, and
138 readings in Core, Crest, and Cusp sites, respectively. In 2011, 237 green ash saplings were
paired with PAR readings (115, 91, and 31 saplings in Core, Crest, and Cusp sites, respectively).
Available PAR was higher in Crest sites than in Core and Cusp sites (F = 16.81; 2, 425; P <
0.001), where readings did not differ (Figure 2.3A). Ash sapling growth was also highest in
stands where A. planipennis had caused substantial ash mortality. Leader growth of ash
saplings in Crest sites was 5-fold greater than Cusp sites, and double that recorded in Core sites,
which was greater than Cusp sites (F = 49; df = 2, 234; P < 0.001) (Figure 2.3B). Annual growth
of ash saplings was significantly and positively related to PAR (Figure 2.4).
Seedlings: In 2010, seedlings were present in 94% of the 32 micro-plots in Core sites, 78% in
Crest sites, and 84% in Cusp sites. A total of 2249 seedlings, representing 22 species, were
tallied, including 733, 692, and 824 seedlings in Core, Crest, and Cusp sites, respectively.
Seedling densities averaged 25,219 ± 3552, 23,808 ± 5536, and 28,350 ± 6886 per ha in Core,
Crest, and Cusp sites, respectively, and did not differ among invasion stages (P = 0.46). Red oak
and black walnut seedlings were not observed in any site, while red maple seedlings were fairly
abundant in Crest sites (Table 2.2).
In 2010, 1704 green ash seedlings were counted, including 386, 656, and 662 seedlings
in Core, Crest, and Cusp sites, respectively. Green ash seedlings were recorded in every site.
Densities of green ash seedlings averaged 13,280 ± 2363; 22,560 ± 557; and 22,776 ± 5622
stems per ha in Core, Crest, and Cusp sites, respectively, and did not differ among invasion
stages (P = 0.95).

51

In 2011, 1,622 green ash seedlings were recorded in 23 of the 24 study sites, including
386, 467, and 769 seedlings in Core, Crest, and Cusp sites, respectively. Densities of ash
seedlings averaged 13,280 ± 2210; 16,067 ± 3519; and 26,457 ± 6930 stems per ha in Core,
Crest, and Cusp sites, respectively, and did not differ among invasion stages (P = 0.87).
In 2010, 102 ash seedlings had cotyledons, indicative of current-year germination. This
included 44 seedlings in six Cusp sites, and 58 seedlings in four Crest sites. No ash seedlings
with cotyledons occurred in micro-plots in Core sites. Extrapolated densities of ash seedlings
with cotyledons averaged 1995 ± 664 and 1513 ± 571 per ha in Crest, and Cusp sites,
respectively, and were significantly higher than Core sites, which had not ash seedlings with
cotyledons (H = 12.32; df = 2, 93; P = 0.002)/ Crest and Cusp sites did not differ.
In 2011, 99% of the ash seedlings with cotyledons occurred in plots in four Cusp sites.
Density of current-year ash seedlings averaged 34 ± 34 and 3371 ± 1610 seedlings per ha in
Crest and Cusp sites, respectively, but none were found in Core sites. Average densities of ash
seedlings with cotyledons were higher in Cusp sites (H = 16.45; df 2, 93; P < 0.001) than other
sites.

52

DISCUSSION
Millions of overstory ash trees have been killed by A. planipennis in forested sites in
Michigan and the surrounding states (Poland 2007; EAB.info.2012). Previous studies examining
ash stands in southeast Michigan reported overstory ash mortality rates approaching 100%
(Gandhi et al. 2008; Herms et al. 2009). In my study, 100% mortality of overstory ash occurred
in three of eight Core sites in southeast Michigan, but live overstory ash remained in the other
five sites. My objectives were to determine the potential for green ash to persist in Michigan
forests and potential future stand composition. With high levels of ash mortality, what are the
long term implications for green ash stands? Will overstory species composition be altered by
this invasive pest? My results indicate A. planipennis will largely remove green ash as a
functional component of stands in southern Michigan, but has yet to alter non-ash composition
of the overstory.
High levels of ash mortality created sizeable gaps in the overstories of Core sites in
southeast Michigan, but the percentage of PAR penetrating the canopy in Core sites was similar
to that in Cusp sites, where overstory ash remains relatively healthy. These data indicate
canopy gaps in Core sites have generally been closed by the lateral ingrowth of neighboring
non-ash overstory trees, and growth of recruits. Studies conducted in Appalachian hardwood
forests found crown expansion rates of overstory trees vary among and within species, but
lateral ingrowth of overstory trees (≥ 25 cm in DBH) averaged 18 cm per year (Trimble and
Tyron 1966; Runkle 1982; Runkle and Yetter 1987). In addition, recruits could have capitalized
on increased PAR within affected stands, and will are growing into gaps. Most trees in Core
sites were ≤ 30 DBH, indicating they were recruits at the time stands were originally infested by
53

A. planipennis. Similar size class distributions, however, were found in Crest and Cusp sites. If
gaps had been filled by recruits capitalizing on available PAR levels, then I would have expected
to see more live trees of the ≤ 30 cm DBH size class in Core sites compared to Crest and Cusp
sites, but that was not the case. Although my data indicate green ash is the most prominent
species at the recruit and sapling level in Core sites, ash has not reached the overstory. Based
on these data, I conclude most gap closures are likely due to lateral ingrowth of non-ash
overstory trees, rather than understory growth.
Ash seedlings with cotyledons were present in Crest and Cusp sites, but not in Core
sites. Seedling abundance is dependent on seed production by overstory trees (Figure 2.6), but
the surviving overstory ash trees in the Core sites are confronted with multiple barriers for
successful reproduction. Green ash is a dioecious species (Wright 1959; Barnes and Wagner
1981; Kennedy 1990), so not only does an individual ash need to survive, but an additional ash
of the opposite sex must also survive. Ash pollen is dispersed by wind over fairly short
distances (60 to 90 m), and surviving trees must be in range for pollination to occur (Kennedy
1990). Green ash of both sexes can flower annually (Wright 1959; Kennedy 1990), but this does
not guarantee reproduction will occur. If all other conditions are met (i.e. trees of both sexes
survive and are within range for pollination), surviving trees could fail to flower simultaneously.
The absence of newly germinated ash in Core sites is consistent with a previous study on ash
regeneration in southeast Michigan which documented a lack of ash in the seed bank of stands
where overstory ash were killed by A. planipennis (Herms et al. 2009). If ash is not reproducing
in Core sites, then future overstory ash is dependent on ash currently growing in the understory

54

(Figure 2.6). Trends in ash germination observed in Core sites are likely to repeat in Crest and
Cusp sites as overstory ash dies.
A study conducted in ash stands in southeast Michigan from 2007 to 2009 reported
green ash was still abundant in the understory (Kashian and Witter 2011). This is consistent
with my results; green ash was still the most prevalent species in the understory of Core and
Crest sites. Density of ash saplings was highest in sites where overstory trees were dead or
declining in the Core and Crest sites. In Core sites, the reduction of available PAR caused by
lateral ingrowth of non-ash overstory trees, and the lack of newly germinated ash seedlings will
likely reduce ash sapling densities over time (Figure 6.2). Current levels of PAR and vertical
growth of ash saplings were high in Crest sites, suggesting, saplings may yet grow into gaps,
however, there is no evidence that ash recruits and saplings in Core sites are successfully
reaching reproductive age, making it unlikely that ash saplings in Crest sites will either. Similar
patterns can be expected to occur in Cusp sites as A. planipennis populations continue to build.
Mortality of green ash recruits caused by A. planipennis was low in sites in all three
invasion stages. Trees at the recruit stage are the most likely to replace overstory ash as it
succumbs to A. planipennis (Figure 2.6). Green ash regeneration can survive in shaded
conditions up to 15 years, and can often outcompete other species when released (Kennedy
1990). Agrilus planipennis is more attracted to trees exposed to sunlight (Yu 1992; McCullough
et al. 2009a, 2009b), and recruits predominately grow in shaded conditions. Beetles are also
attracted to stressed ash trees (McCullough et al. 2009a, 2009b; Siegert et al. 2010; Tluczek et
al. 2011). Shade and stressed overstory trees may have allowed recruits to remain relatively
undetected, even in Core sites. As more overstory trees succumb, the beetles will likely infest
55

recruits growing in the understory, preventing them from reaching reproductive age (8 to 10
cm DBH) (Kennedy 1990) (Figure 2.6).
Species composition of recruits and saplings were generally consistent among all sites,
regardless of A. planipennis invasion stage. Large gaps in the overstory could provide an
opportunity for more shade intolerant species to grow into the overstory (Messier et al. 1999).
The prevalence of green ash in the understory at all stages of the invasion, however, may
prevent the growth of less tolerant species. If overstory ash is killed, and understory ash fails to
reach reproductive age, ash will eventually be lost from the understory. Based on this study, I
predict few ash currently in the understory of Core sites will reach the overstory, and that if
overstory trees survive, they will be too few and far between to successfully reproduce. If
efforts are not made to substantially slow A. planipennis population growth, green ash will be
functionally lost from these stands in the next 20 years. Furthermore, decimation of green ash
by A. planipennis will be repeated in Crest and Cusp sites, and throughout the range of green
ash.
Red maple and sugar maple are poised to succeed ash as dominant overstory species in
most sites I surveyed. Neither species of maple was abundant in the overstories of Core and
Crest sites, but sugar maple was common at all other stages of growth, and red maple was
abundant in the understory of Core and Cusp sites (Tables 2.1, 2.2). Both species of maple can
thrive on a variety of sites, within different communities, and have ecophysiological traits such
as shade tolerance, abundant seed production and seed dissemination that make them strong
competitors (Godman et al. 1990; Walters and Yawney 1990; Abrams 1998). American elm
also had a strong presence at all three invasion stages (Tables 2.1, 2.2), but over time, may
56

become less important because of Dutch elm disease (Barnes and Wagner 1981, Bey 1990).
Additionally, understory American elm often declines when suppressed by more shade tolerant
sugar maple (Barnes and Wagner 1981, Bey 1990). If sugar maple becomes more prevalent, the
ability of American elm to persist in the understory may decline. Black cherry is classified as
shade intolerant (Marquis 1990), but was abundant in all four stages of growth in Core sites
(Tables 2.1, 2.2,). In Crest and Cusp sites, black cherry was abundant in the overstory (Table
2.1), but sporadic and generally scarce in in the understory (Table 2.2). Black cherry may
become more prominent in Crest and Cusp sites due to the increased availability of PAR in gaps
formed by dead ash. Northern red oak and black walnut were common in the overstory in Core
and Crest sites, but were infrequently encountered in lower canopies (Tables 2.1, 2.2). Red oak
has poor seed survival, and typically requires the production of 500 acorns to produce one 1year-old seedling (Sander 1990). Black walnut is shade intolerant and must be dominant or
codominant in a stand to survive (Barnes and Wagner 1981; Williams 1990). If ash in the
understory does not reach the overstory, my data suggests American elm, black cherry, red
maple, and sugar maple will likely dominate these stands.

57

Table 2.1. Total number of trees (N), mean (± SE) diameter at breast height (DBH) and basal area,
relative density, relative frequency, relative dominance, and relative importance values (RIV) of the most
abundant tree species across 24 sites representing three A. planipennis invasion stages. Relative
importance values rank species contribution to a stand's overstory. All measurements were taken in
summer of 2010.
Mean ± SE Mean ± SE basal Relative Relative
Relative
1
Species
N
DBH (cm) area (m² per ha) density frequency dominance RIV

Core sites
Green Ash
Red Oak
Black Cherry
American Elm
Black Walnut
Crest sites
Green Ash
American Elm
Black Walnut
Black Cherry
Red Oak
Cusp sites
Green Ash
Red Maple
American Elm
Black Cherry
Sugar Maple
1

216
103
122
130
82

20 ± 0.6
32.1 ± 1.5
25 ± 1.2
17.6 ± 0.5
33.1 ± 1.2

2.3 ± 0.5
2.9 ± 1.4
2.2 ± 0.9
1.0 ± 0.4
2.2 ± 1.1

27%
13%
15%
16%
10%

19.5
17.1
19.5
14.6
7.3

18.4
23.7
18.4
8.3
18.5

65.9
53.8
52.9
39.0
35.8

376
88
84
39
24

16.8 ± 0.3
18.8 ± 0.8
26.5 ± 1.2
25.5 ± 1.8
22.9 ± 2.2

2.7 ± 0.6
0.8 ± 0.4
2.2 ± 1.1
0.7 ± 0.4
0.3 ± 0.2

61%
14%
14%
6%
4%

25.0
18.8
6.3
15.6
15.6

40.7
12.0
23.4
10.2
5.1

126.7
44.7
43.7
31.8
24.7

443
198
174
118
138

21 ± 0.4
17.8 ± 0.5
17.8 ± 0.6
19.5 ± 0.8
18.7 ± 0.8

5.2 ± 1.0
1.6 ± 1.1
1.5 ± 0.4
1.2 ± 0.4
1.3 ± 0.4

39%
17%
15%
10%
12%

15.1
13.2
15.1
15.1
11.3

37.1
11.4
10.4
8.5
9.5

91.2
41.6
40.5
33.6
32.8

Relative importance values = relative density + relative frequency + relative dominance.

58

Table 2.2. Total count (N) and mean (± SE) stems per ha., of the five most abudant species of recruits, saplings, and seedlings
in 2010, sampled at 24 sites in southern Michigan representing the three stages of the A. planipennis invasion wave.
Mean ± SE
Mean ± SE
Mean ± SE
1
2
3
N stems per ha
N
stems per ha
N
stems per ha
Recruits
Saplings
Seedlings
Core sites
Green Ash
Red Maple
Sugar Maple
Black Cherry
Red Oak
Crest sites
Green Ash
Basswood
American Elm
Sugar Maple
Shagbark Hickory
Cusp sites
Sugar Maple
Green Ash
American Elm
Red Maple
Black Cherry
1
2
3

63
56
28
24
16

120 ± 29
113 ± 56
56 ± 30
48 ± 20
32 ± 27

Green Ash
Black Cherry
Sugar Maple
Red Maple
American Elm

967
98
65
40
24

1962 ± 331
199 ± 71
132 ± 59
81 ± 37
49 ± 21

Green Ash
Red Maple
Black Cherry
Sugar Maple
American Elm

386
123
81
41
19

13,281 ± 2364
4232 ± 1895
3131 ± 963
1411 ± 249
654 ± 346

174
47
20
19
12

353 ± 90
95 ± 78
41 ± 20
39 ± 18
24 ± 24

Green Ash
Sugar Maple
Northern Hackberry
American Elm
Black Cherry

911
46
19
18
15

1849 ± 443
93 ± 51
39 ± 29
37 ± 26
30 ± 26

Green Ash
Northern Hackberry
Sugar Maple
Basswood
American Elm

656
18
4
3
3

22,570 ± 5570
619 ± 446
138 ± 82
103 ± 76
103 ± 103

79
78
50
46
13

160 ± 41
158 ± 38
102 ± 25
93 ± 43
26 ± 9

Green Ash
Sugar Maple
Black Cherry
American Elm
Red Maple

255
47
11
10
10

518 ± 212
95 ± 47
22 ± 10
20 ± 6
20 ± 10

Green Ash
Red Maple
Musclewood
Bitternut
Swamp White Oak

662
92
16
15
9

22,776 ± 5623
3166 ± 1926
551 ± 360
516 ± 320
310 ± 113

Recruits (≥ 3 cm and < 10 cm stem diameter).
Saplings (< 3 cm stem diameter and ≥ 45 cm).
Seedlings (< 45 cm in height).

59

Figure 2.1. Locations of 24 green ash sites representing three stages of the A.
planipennis invasion wave. Core sites were characterized by low densities of A.
planipennis and high ash mortality. Crest sites were characterized by high densities
of A. planipennis and moderate to high ash mortality. Cusp sites were
characterized by low densities of A. planipennis and low ash mortality.

60

Figure 2.2. Mean (± SE) density of green ash saplings per ha in 2010 (N =
2133) and 2011 (N = 1905). Means with the same letter are not
significantly different among the three invasion stages (a and b for 2010,
y and z for 2011).

61

Figure 2.3. Mean (± SE) percentage of full sun represented by
photosynthetically active radiation (PAR) measured after dawn and
before sunrise in 12 sites between 20 June to 15 July 2011 (N = 428
readings). Means with the same letter are not significantly different
(Tukey’s protected HSD test; P < 0.05).

62

Figure 2.4. Mean (± SE) growth (cm) of leaders of green ash saplings
measured in 12 sites between 12 June to 17 June 2011 at the three A.
planipennis invasion stages (N = 237 saplings ). Means with the same
letter are not significantly different (Tukey’s protected HSD test; P <
0.05).

63

Figure 2.5. Linear regression of annual growth of terminal leaders of green ash
saplings and percent of full sun represented by photosynthetically active radiation
(PAR) measured in 12 sites in southern Michigan (N=237 paired measurements).
Saplings were 1 to 2 m tall, but diameter and exact height of saplings were not
measured (P < 0.05).

64

65

APPENDICES

66

Appendix A

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

Voucher Number: 2012-03

Title of thesis:
Evaluation of Emerald Ash Borer Populations and the
Ash Resource at Three Stages of the Invasion Wave

Museum where deposited:
The Michigan State University (MSU) Albert J. Cook Arthropod Research Collection

Table A.1. Specimens submitted to the Michigan State University (MSU) Albert J. Cook
Arthropod Research Collection
Family
Life Stage
Quantity
Preservation
Buprestidae
Agrilus planipennis
Larva
10
Ethanol
Buprestidae
Agrilus planipennis
Adult, Male
10
Pinned
Buprestidae
Agrilus planipennis
Adult, Female
10
Pinned
Braconidae
Braconidae
Braconidae
Braconidae

Atanycolus cappaerti
Atanycolus cappaerti
Atanycolus cappaerti
Atanycolus cappaerti

Adult, Male
Adult, Female
Adult, Male
Adult, Female

67

5
5
5
5

Pinned
Pinned
Ethanol
Ethanol

Appendix B

Table B.1. List of study sites by Michigan county and geographic coordinates.

68

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69

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