A TRIAL OF FIRE AND ICE: ASSESSING THE ABILITY OF INVASIVE TREE
PYRUS CALLERYANA TO RESIST DISTURBANCE DURING GRASSLAND
INVASION IN THE AMERICAN MIDWEST
Thesis
Submitted to
The College of Arts and Sciences of the
UNIVERSITY OF DAYTON
In Partial Fulfillment of the Requirements for
The Degree of
Master of Science in Biology
By
Margaret E. Maloney
Dayton, Ohio
May 2021
A TRIAL OF FIRE AND ICE: ASSESSING THE ABILITY OF THE INVASIVE TREE
PYRUS CALLERYANA TO RESIST DISTURBANCE DURING GRASSLAND
INVASION IN THE AMERICAN MIDWEST
NAME: Maloney, Margaret E.
APPROVED BY:
______
Ryan W. McEwan, Ph.D. Faculty Advisor Professor, University of Dayton
______
Chelse M. Prather, Ph.D. Committee Member Associate Professor, University of Dayton
______
Albert J. Burky, Ph.D. Committee Member Professor, University of Dayton
ii ABSTRACT
A TRIAL OF FIRE AND ICE: ASSESSING THE ABILITY OF THE INVASIVE TREE
PYRUS CALLERYANA TO RESIST DISTURBANCE DURING GRASSLAND
INVASION IN THE AMERICAN MIDWEST
Name: Maloney, Margaret E. University of Dayton
Advisor: Dr. Ryan McEwan
Ecological invasion is one of the most important processes of global change influencing ecosystems in practically every biome on Earth. Invasive species shift ecosystem dynamics, community structures, nutrient cycling, and ecosystem function.
Invasive species pose a significant challenge to land managers who are charged with maintaining biodiversity and managing long-term ecosystem structure. Pyrus calleryana is an ornamental tree species that escaped cultivation and grows rapidly in highly disturbed soils with high light intensities. Former agricultural fields are highly susceptible to invasion from Pyrus calleryana. Observational evidence suggests that two features of this species’ biology may be particularly important to invasion success: (a) an extended leaf phenology and (b) the ability to persist in grasslands following disturbance via epicormic sprouting. While this tree is one of the most problematic invasive species within the Eastern and Central parts of the United States, it is relatively understudied.
This thesis is divided into two distinct studies that address the invasion biology of Pyrus calleryana. The first study, detailed in Chapter One, is an analysis of the timing of leaf development (phenology) in relationship to co-occurring native trees. In the second
iii study, detailed in Chapter Two, I conducted a replicated field experiment to assess P. calleryana sprouting as a mechanism of persistence in response to experimental disturbance treatments including prescribed fire and experimental freezing using a treatment of liquid nitrogen. The experiments we conducted in this project are in cooperation with the conservation staff at the Five Rivers Metroparks and all activities took place on their properties. In the first study, I discovered that P. calleryana has a longer leaf duration than native trees through earlier leaf out, and delayed abscission. In addition, a late season frost event allowed for a serendipitous study that revealed much greater frost tolerance in P. calleryana than native trees. Extended phenology and frost tolerance indicate strong potential for growing season carbon acquisition for P. calleryana compared to native trees. In the field experiment (Chapter Two), I discovered that common practices such as mowing and prescribed fire may be facilitating the invasion of Pyrus calleryana due to its aggressive sprout response. Data indicate that experimental freezing was marginally more useful than cutting or herbicide in sprout control; however, herbicide application was the only reliable method for preventing resprouting. The results of these studies illuminate the invasion biology and may help inform more effective management of this problematic invasive species across the region, leading to more sustainable management of biodiverse habitats.
iv ACKNOWLEDGEMENTS
I would like to thank my advisor Dr. Ryan McEwan, for supporting me throughout my budding scientific career and providing me with the best guidance and sage wisdom. I would like to thank many of the current and former members of the
McEwan lab for shaping me into a better scientist, mentor, advisee, and friend. A special thank you to Julia Chapman, Charlotte Shade, Celia Montemurri, Taylor Sparbanie,
Corey Kuminecz, Abby Hay, Grace Attea, and many others who helped with field work, data analysis, and lab work. I would also like to thank my committee members for their support and guidance throughout this process. This research would not have been completed without the kind support from the University of Dayton Biology Department and Sigma Xi, The Scientific Research Honor Society. Most of all, thank you to Eric
Borth for essentially doing all field work with me (especially due to COVID) and helping tremendously with data analysis. Your support added far more depth to this project. I would also like to thank SARS-COV-2 for making this thesis far more complicated than it needed to be but also for pushing my creativity and innovation to continue my work despite the chaos in the world. You will not be missed but prove that science is far superior and needed in times of crisis.
v TABLE OF CONTENTS
ABSTRACT ...... iii
ACKNOWLEDGEMENTS ...... v
LIST OF FIGURES ...... viii
LIST OF TABLES ...... ix
CHAPTER 1 LEAF PHENOLOGY AND FREEZE TOLERANCE OF INVASIVE
TREE PYRUS CALLERYANA (ROSEACEAE) AND POTENTIAL NATIVE
COMPETITORS ...... 1
INTRODUCTION ...... 1
METHODS ...... 8
Site Description ...... 8
Experimental Design ...... 8
Statistical Analysis ...... 9
RESULTS ...... 10
DISCUSSION ...... 12
CHAPTER 1 TABLE AND FIGURES ...... 16
LITERATURE CITED ...... 20
vi CHAPTER 2 A TRIAL OF FIRE AND ICE: EXPERIMENTAL ASSESSMENT OF
NOVEL ECOLOGICAL RESTORATION TECHNIQUES IN MIDWESTERN
PRAIRIES INVADED BY PYRUS CALLERYANA ...... 25
INTRODUCTION ...... 25
METHODS ...... 32
RESULTS ...... 35
DISCUSSION ...... 38
CHAPTER 2 FIGURES ...... 44
LITERATURE CITED ...... 51
vii LIST OF FIGURES
Figure 1.1 Spring leaf development from 1=bud dormant to 5=leaf at summer green of two native tree species…………….………….………..………………………17
Figure 1.2 Percent leaf mortality for two native tree species (Populus deltoides and Platanus occidentalis)…………………………………………………..………...…18
Figure 1.3 Fall leaf development from 5=leaf at summer green to 8=leaf abscised of two native tree species……………………………….………………………………..19
Figure 2.1 Examples of experimental treatments for Pyrus calleryana on recently restored grasslands……………………….……………………………………..44
Figure 2.2 Total number of resprouts on previously disturbed Pyrus calleryana in an old field…………..……………………………..…………………………………45
Figure 2.3 Total dry mass (g) on previously disturbed Pyrus calleryana in an old field in Dayton………………………………………………………………….……46
Figure 2.4 Total number of resprouts on undisturbed Pyrus calleryana in an
old field in Dayton…….……………………..……………………………………….…47
Figure 2.5 Total dry mass (g) on undisturbed Pyrus calleryana in an old field in Dayton, OH…………………………………………………………………48
Supplemental Figure 1 The pre and post- treatment basal diameters on previously disturbed Pyrus calleryana…………………………………………..……… 49
Supplemental Figure 2 The pre and post- treatment basal diameters on
Undisturbed Pyrus calleryana in old fields………………………………………………50
viii LIST OF TABLES
Table 1:1 Observation categories for the vegetative changes of the invasive and native…………………...…………………………………………………..16
ix CHAPTER 1
LEAF PHENOLOGY AND FREEZE TOLERANCE OF INVASIVE TREE PYRUS
CALLERYANA (ROSEACEAE) AND POTENTIAL NATIVE COMPETITORS
INTRODUCTION
Invasive plant species are a significant threat to biodiversity in a variety of habitats and understanding the mechanisms these species employ to gain dominance is a pressing scientific concern (Sakai et. al 2001; Van Kleunen et. al 2010; Wolfe, 2002).
There are thousands of introduced species within the United States, though many never reach the designation of an invasive species (Congress, 1993; Pimentel et. al 2000). The mechanism from which an invasive species moves past its lag phase and into a rapid increase in distribution with concurrent negative effects on native habitats is still relatively understudied (Crooks, 2005). A variety of factors could potentially lead to invasion in some species, including opportunities arising in empty niche space, demographic changes through human mediated activities, genetic change, or changing climate envelopes (Mack et. al 2000; Callaway and Ridenour, 2004). Many invasive species exhibit a complex profile of traits that contribute to their invasive success such as
Lonicera maackii, Ranunculus ficaria, and Pueraria montana (McNeish and McEwan,
2016; Axtell et. al 2010; Forseth and Innis, 2010).
Within the eastern and central part of the United States, many agricultural fields have been left fallow leaving introduced plants the opportunity to invade due to empty niche space and naturalized establishment of introduced species from nearby ecological
1 spaces (Inouye et. al 1987). Abandoned farmland, otherwise known as old fields, are usually diverse, forb dense ecosystems that transition through succession into woody dominated ecosystems (Cramer et. al 2008). In these forb dense systems, it is common to see native species competing against non-native species that have also entered the system
(Gross and Emery, 2007). Gross and Emery (2007) found an increase in species richness in both native and non-native species over a 15-year period of time since agricultural abandonment. Both native and non-native species present in old fields usually have a predictable pattern of secondary succession as noted first by Catherine Keever in her
1950 groundbreaking paper (Keever, 1950). She noted that life history strategies, seed dispersal, and allelopathy all contribute to the interaction of species and mathematical models such as Tilman’s resource ratio hypothesis can be used to help better predict successional changes (Keever, 1950; Tilman, 1985). However, a combination of nutrient and moisture limitation in the soil of old fields from years of unsustainable agricultural practices leads to a competitive environment that creates advantages for species better adapted to disturbance (Cramer et. al 2008). This is especially true for invasive species, who usually thrive in highly disturbed soil and empty niche space. With more introduced invasive species within these systems, we do not see traditional successional trajectories
(Flory and Clay, 2010). The new combination of invasive species mixed with some native species is leading to new novel ecosystems being formed. In many ways, this is leading to drastically altered successional trajectories during old fields succession.
Problems arise as non-native species begin to alter successional trajectories within a system, leading to the formation of new, novel ecosystems (Flory and Clay, 2010).
Often, these non-native species are introduced into the system from human mediated
2 activities such as escaping cultivation in nearby suburban communities, arriving through shipping containers, exotic plant trade, or humans releasing species in native areas (Sakai et. al 2001). Once present in the ecosystem, a variety of factors can lead to the non-native species dominating the system. From escaping natural predation, releasing harmful chemicals into the soil that suppress native species, to utilizing other non-native species for pollination or seed dispersal, invasive species can quickly dominate old fields and change the trajectory of succession (Wolfe, 2002; Culley and Hardiman, 2007). With all these factors at play, the novel ecosystems we are seeing today across much of the eastern and central parts of the United States have left ecologists baffled on how to mediate succession. An especially complicated factor in restoration ecology is that many land managers and ecologists are attempting to speed up the natural cycle of succession that usually takes hundreds of years (Young et al. 2005). For example, in Dayton, Ohio, land managers are trying to transition old fields into forests in as little as ten to twenty years.
This type of rapid succession is relatively understudied and there are a variety of approaches that can aid in the suppression of invasive species and the growth of natives.
Beyond the challenges of aiding rapid succession of species, large climatic changes are also complicating the natural patterns of succession and opening opportunities for invasive to thrive in new novel habitats (Hellman et. al 2008).
As CO2 continues to rise rapidly, climate zones are shifting which is exacerbating the spread of species into new habitats. Climate change is altering seasonal weather patterns such as temperature and precipitation, in addition to increasing extreme weather events. Since 1981, the average increase in global temperature is 0.18O C according to
NOAA’s 2019 Global Climate Summary (NOAA, 2019). 2019 was also the 2nd warmest
3 year on record and 2020 was one of the worst years for natural disasters within the United
States (NOAA, 2019). As climate change shifts traditional weather patterns, more species will be driven to adapt to new conditions as well as compete with new species present in the ecosystem (Schwartz, 1992). As resources become scarcer, and competition increases, the stress and disturbance will increase the chance of invasion from non-native species
(Hellman et. al 2008). For delicate species that have only adapted to very niche climate patterns within the system, expanding the range to adjust to new conditions may be difficult and could result in species loss (McLaughlin et. al 2002). Numerous studies have shown that species compositions are changing and shifting with climate change, and one study estimated that species are moving at approximately 6.1km per decade, though it may vary from species to species (Parmesan and Yohe, 2003). As species continue to evolve due to climatic pressure and climate envelope changing, we are also seeing significant changes to the behavior of plants and animals due to this shift.
Change in plant phenology is one the earliest and most studied responses to climate change (McEwan et. al 2011). Changes in plant phenology give researchers a better understanding of how species may adapt to current climatic pressure and insight to understand which species are more at risk for extinction. In recent studies, it was found that species are blooming earlier in the growing season in comparison to studies from decades earlier (McEwan et. al 2011). This was especially prevalent in species that bloom early in the spring, such as spring ephemerals like Crocus and Galanthus (McEwan et. al
2011). This is alarming, as the timing of blooming, flowering, or seed dispersal changes can have severe consequences on pollination, light competition, and predation (Ghazoul,
2004; Miller-Rushing et. al 2010). This in turn can lead some species to extinction while
4 other species populations increase in richness (Miller-Rushing et. al 2010). Climate change and changing climate envelopes introduces phenological mismatch, which is the idea that species that depend on one another for survival are no longer in sync (Miller-
Rushing et. al 2010). All ecosystem relationships on this planet, such as competitive interactions, nutrient cycling, and seed dispersal, have a temporal component. If the temporal component is disrupted, it threatens the entirety of the system. While we understand that this disruption will hurt sensitive or endangered species, how will this mismatch phenology aid non-native species within a system? A relatively understudied but important mechanism that non-native species may be utilizing to shift to becoming invasive is extended leaf phenology. In previous studies, invasive species often have earlier leafing out periods and longer leaf duration during the growing season than native species in the American Midwest. For example, the deciduous shrub Lonicera maackii is successful due to extended leaf duration when compared to common native shrubs, allowing for greater access to carbon (McEwan et. al 2009). The same was seen when studying the leaf phenology of Berberis thunbergii, a shrub that invades deciduous forests in the northeastern United States (Xu, 2007). When compared two native understory shrubs, its leaf expansion initiated several weeks before the native shrubs (Xu,
2007). As more invasive species proliferate within an ecosystem, especially old fields, this mechanism of prolonged phenology may be utilized as a competitive advantage over native species.
Pyrus calleryana (Callery pear) is an ornamental invasive species that thrive in highly disturbed areas, such as old fields (Culley and Hardiman, 2007). P. calleryana originated in Asia but was introduced to the United States to combat the fire blight
5 (Erwinia amylovora) that was decimating Pyrus communis (Culley and Hardiman, 2007).
Originally thought to be sterile, P. calleryana was viewed horticulturally as extremely valuable and soon became the most widely planted boulevard tree within the United
States (Culley and Hardiman, 2007). It quickly escaped cultivation and has had a prolific impact on invading any system with disturbed soils and ample light, spreading rampantly throughout the United States (Culley et. al 2011; Vincent, 2005). Most often, we see P. calleryana growing along highway corridors, prairies, and old fields (Culley et. al 2011).
Pyrus calleryana has aggressive sprouting and establishment behavior. Its seeds are easily dispersed by birds and individuals replicate quickly (Culley and Hardiman, 2007).
While P. calleryana has been a problematic species for land managers over the past decade, there are few studies to understand the mechanisms through which it invades. As a whole, its phenological behavior has never been studied and we still do not understand how it was able to proliferate so quickly in old fields. In this study, we aim to understand the phenology of P. calleryana and see if mismatched phenology is used as a mechanism to outcompete native species.
We observed Pyrus calleryana leaf phenology at three sites in Dayton, OH. We were comparing the phenology to two native woody species, Populus deltoides (Eastern
Cottonwood) and Plantus occidentalis (American Sycamore), which are often seen as one of the first successional species in old fields. The overall objective of this study is to (a) understand the phenological traits of P. calleryana and (b) compare phenological differences to native early successional woody species. A late spring freeze also allowed us to examine differences in freeze tolerance for these species. We hypothesize that: (H1 )
P calleryana will leaf out significantly before native woody species; (H2 ) have an
6 extended phenology in the summer as well as keep its leaves longer in the fall; and (H3 ) and will be more resistant to frost.
7 METHODS
Site Description
Observations for this study took place at three study sites, two of which were located in the Medlar Conservation Area (MCA) in Miamisburg, OH (39°36'09.4"N
84°16'25.2"W) and the other in the Shiloh Conservation Area (SCA) near Dayton, Ohio,
USA (39°50'24.8"N 84°14'17.9"W). These sites are managed by the Five Rivers
Metroparks of Dayton, Ohio and are within a 50-kilometer radius of one another. All three sites were previously agricultural fields and have been managed as grasslands for conservation purposes by the Five Rivers Metroparks.
Experimental Design
At each of the three sites, 15 trees were selected for the experiment of which 5 were invasive Pyrus calleryana, 5 were the native tree Platanus occidentalis, and 5 were the native tree Populus deltoides. The native species were chosen due to their pattern of early establishment in regional old fields and prairies, making them potential competitors with Pyrus calleryana. In regional grassland habitats, both P. occidentalis and P. deltoides regularly co-occur with P. calleryana and while the size at maturity is much larger for both native species, competition in open habitats during the sapling stage may be highly relevant for longer-term competitive outcomes. Only two P. deltoides were identified at SCA, so there were only 12 total trees within the study. In August of 2020, these two trees were accidentally cut down by Five Rivers Metroparks staff leaving 10 P. deltoides specimens at the end of the study. All trees were ≤ 3 meters tall and were saplings that had not yet reached reproductive maturity.
8 From November 2019 to November 2020, we monitored the phenology of the trees and marked their vegetative features as they progressed throughout the year (Table
1.1). To do so, a single branch was selected and tagged on each tree. Starting at the outermost bud and moving inwards, the first ten buds of the branch were counted, and the phenotypic vegetative features of each leaf were recorded. If multiple leaves grew out of the same bud, we selected the leaf closest to the stem. In the winter of 2019- 2020, the leaves were checked once a month. Beginning in February 2020, we monitored leaves every 3 days to document leaf development until all trees had leaves that were fully developed and had reached summer green color. During the summer, trees were observed once a month until autumnal color change began at which point leaves were checked once a week until they abscised.
Statistical Analysis
The phenology data were separated into two periods, fall and spring, for the purpose of analysis. A mixed model ANOVA (one-way) was used to test if leaf duration for Pyrus calleryana differed significantly in the spring or fall in comparison with native woody species. Post-hoc comparisons were made using a pairwise t-test with a
Bonferroni correction. When analyzing leaf mortality after a frost event, a Kruskal-Wallis with a pairwise-wilcox test was used. This was used because, upon testing for normality, the data showed non-normal distribution. All analyses were conducted using R v. 3.6.2.
9 RESULTS
Leaf development of P. calleryana began significantly earlier in the Spring 2020 growing season than the native woody species (Figure 1.1; P<0.001). Populus deltoides and P. occidentalis were statistically indistinguishable during the spring growing season until April 22nd through April 28th and the very end of May after a significant frost event. Pyrus calleryana was significantly different (P< 0.001) from both P. deltoides and
P. occidentalis from March 22nd, 2020 until May 23rd, 2020. In the beginning of March,
P. calleryana leafed out relatively quickly (vegetative feature 2) and had its leaves completely exposed (vegetative feature 3) before P. deltoides and P. occidentalis even exposed their leaf blades. This pattern continued until mid-April, when both P. deltoides and P. occidentalis began expanding their leaves. For a brief period from April 22nd
2020 to April 28th 2020, P. calleryana, P. occidentalis, and P. deltoides were all significantly different from one another (P< 0.001).
A late frost event in spring of 2020 led to a serendipitous opportunity to assess freeze tolerance in the study species. On the evening of May 11th, 2020, the frost event occurred that resulted in leaf mortality on the majority of P. occidentalis and P. deltoides trees. This late frost led to the phenology of P. deltoides and P occidentalis becoming significantly different from one another (Figure 1.1; P <0.001). The frost event caused substantial damage to the native tree species, killing all the emerging leaves on the P. occidentalis trees across all three sites aside from one tree and 68% of P. deltoides leaves
(Figure 1.2). P. calleryana, however, was left generally unaffected, averaging only 6% leaf mortality from the frost (Figure 1.2). Starting May 23rd 2020, P. calleryana and P.
10 deltoides were not significantly different from one another, but P. occidentalis was significantly different from P. deltoides and P. calleryana until June 3rd of 2020.
In the fall, P. calleryana retained its leaves for a longer duration than both P. deltoides and P. occidentalis (Figure 1.3; P < 0.001). From August 22nd 2020 until
November 6th 2020, P. calleryana was significantly different from both P. deltoides and
P. occidentalis whereas P. occidentalis and P. deltoides were not significantly different from one another except for September 18th 2020 where all three species were significantly different from one another. Most notably, the leaves of P. occidentalis and
P. deltoides began showing rust spots and browning in early August (vegetative feature
6) and quickly moved to full color change by the end of September (vegetative feature 7).
Pyrus calleryana had some leaf discoloration in late September but did not start changing colors until mid-October (vegetative feature 7). Most P. calleryana trees did not reach full leaf abscission until November with one tree maintaining leaves into December
(Figure 1.3).
11 DISCUSSION
Invasive plant species pose a significant threat to biodiversity and ecosystem function of grassland systems and have the potential to alter old-field to forest succession
(Flory and Clay, 2010). With invasive species changing successional trajectories and potentially creating novel ecosystems, native species are highly susceptible to being outcompeted and lost from the ecosystem (Tognetti and Chaneton, 2012). Multiple mechanisms exist that invasive species can utilize to proliferate in ecosystems. In this study, we evaluated how P. calleryana utilizes the mechanism of prolonged phenology to outcompete other early successional native woody species. Pyrus calleryana proved all of our hypothesis correct as it leafed out before two co-occurring native species (H1), had a significantly longer duration for photosynthesis in the summer and fall (H2) and a greater frost tolerance (H3) than two co-occurring native species. It can utilize these mechanisms to facilitate the invasion in early successional habitats of the central and eastern United
States.
Pyrus calleryana has traits that amplify its ability to spread quickly and dominate ecosystems. Pyrus calleryana thrives in disturbed soils and high light environments, giving it the ability to grow in a multitude of habitat conditions (Culley and Hardiman,
2007). It mostly favors early and mid-successional ecosystems, especially old fields. It is less commonly found as an invasive plant in forest ecosystems, potentially due to the limited light availability from the overstory; however, recent observations (R.W.
McEwan, University of Dayton, personal observation) suggest this species may ultimately invade more shady habitats. It has been found that P. calleryana can grow in
12 poor soil conditions, including soils that have been deprived of important nutrients which is why it does so well in old fields (Warrix and Marshall, 2018).
The longer leaf phenology of P. calleryana may provide a significant competitive advantage and pose a threat to biodiversity in old fields and prairies. Previous studies have shown that woody invaders can benefit from extended phenology in prairie and deciduous forest ecosystems (Schuster and Dukes, 2017; McEwan et. al 2009), as it can aid in the establishment of woody encroachment within the system. Fridley (2012) found that most non-native woody invasive species capitalized on an extended autumn phenology, a behavior that was absent in native species across multiple phylogenetic groups. Xu et. al (2007) discovered a slightly different behavior in the invasive understory shrub Berberis thunbergii, which was comparable to native shrubs but had a carbon gain during the spring which might be aiding its success. On the contrary, a recent study from O’Connell and Savage (2020) found that while woody invasive plants retain leaves later than most native plants, they did not necessarily gain more carbon than native species. In the case of P. calleryana, it leafed out almost one month before the native woody species and retained leaves until almost a month after the other native species.
This extended phenology is likely conferring a competitive advantage in comparison to other native woody species. Further research would be needed to understand at what point P. calleryana reached its highest photosynthetic capacity and carbon gain.
The frost tolerance of P. calleryana is relatively understudied but an important mechanism of its continual dominance in early successional systems. The late frost in our study resulted in the leaves of both native species to be decimated. While the native species had a quick regrowth in their leaves, P. calleryana continued successful
13 photosynthesis throughout the frost event. The morphology of P. calleryana summer green leaves is glossy and there may be epidermal morphology that is aiding P. calleryana in its ability to withstand cold temperatures. This ability to continue growing despite a frost event gives P. calleryana a superb advantage as it can continue photosynthesizing and does not need to expend energy on regrowing leaves. The ability to withstand frost events also contradicts some early studies that noted that P. calleryana distribution was restricted by cold winters (Culley and Hardiman, 2007). Our findings opens the possibility that with warmer temperatures due to climate change, P. calleryana will continue to expand its range. O’Connell and Savage (2020) noted that withstand freezing temperatures may not limit the range of invasive species, but may indicate that at the most northern edge of their range, they lose the competitive advantage of an extended phenology. In Dayton, Ohio, we are not very close to the most northern edge of their range (as wild P. calleryana has been detected as far north as Wisconsin and Michigan) which means that the mild winters in Dayton will pose no threat to its continual spread throughout the state.
Prolonged phenology of P. calleryana may play an important role in this species invasion biology and ultimately influence native early successional species and alter regional successional trajectories. Old fields and young prairies have ample light availability and accessibility to resources (Gross and Emery, 2007). Since P. calleryana spends most of its energy elongating their shoots vertically and attaining a height over other species, P. calleryana can shade out native prairie species (Kikuzawa, 1995). We noted that during our study, specifically at MCA, that most P. calleryana were growing in groups right next to one another, leaving little room for other species to grow
14 underneath. Since they were limiting the light availability below and taking up a substantial amount of space, they are leaving minimal resources for native species.
Moreover, we noted that native woody species did not grow as closely with one another and did not produce nearly as many leaves as P. calleryana. If P. calleryana continues to minimize the light availability early on, traditional successional trajectories for old fields will be diminished.
Our research adds to the growing understanding of the mechanisms and the behavior of how P. calleryana continues to become a prolific species across much of the eastern and central parts of the United States. Through this study, we were able to understand the phenological behavior of P. calleryana and scratch the surface of its ability to tolerate frost. However, much is still unknown about this species. Further research is needed to understand how it is allocating its nutrients and resources and understanding its photosynthetic capacities during its growing season. Our study did shed light on some important considerations for both land managers and researchers. With an earlier start to its growing season, this may be the best time to treat P. calleryana without harming other native species. Moreover, further research in broadening our understanding of P. calleryana’s freeze tolerance may give us more insight into how it is able to sustain this competitive advantage.
15 CHAPTER 1 TABLE AND FIGURES
Table 1:1 Observation categories for the vegetative changes of the invasive and native shrub phenology. All shrubs were observed in Dayton, OH grasslands.
Vegetative features
1) Bud dormant
2) Leaf blade visible
3) Entire leaf exposed
4) Entire leaf exposed and flat
5) Leaf at summer green color
6) Leaf different than summer color
7) Leaf at fall color
8) Leaf abscission
16
Figure 1.1 Spring leaf development from 1=bud dormant to 5=leaf at summer green (Table 1) of two native tree species (Populus deltoides and Platanus occidentalis) and an exotic invasive tree (Pyrus calleryana) at three sites in Southwest Ohio, USA. Presence of stars indicate statistical significance (p<0.001) between median values for P. calleryana and native species.
17
Figure 1. 2 Percent leaf mortality for two native tree species (Populus deltoides and Platanus occidentalis) and an exotic invasive tree (Pyrus calleryana) resulting from a last frost event occurring on May 11th, 2020 at three sites in Southwest Ohio, USA There was a statistical significance (p<0.001) between median values for P. calleryana and native species.
18
Figure 1. 3 Fall leaf development from 5=leaf at summer green to 8=leaf abscised (Table 1) of two native tree species (Populus deltoides and Platanus occidentalis) and an exotic invasive tree (Pyrus calleryana) at three sites in Southwest Ohio, USA. Presence of stars indicate statistical significance (p<0.001) between median values for P. calleryana and native species.
19 LITERATURE CITED
Axtell, A. E., DiTommaso, A., & Post, A. R. (2010). Lesser celandine (Ranunculus
ficaria): A threat to woodland habitats in the northern United States and southern
Canada. Invasive plant science and management, 3(2), 190-196.
Callaway, R. M., & W. M. Ridenour. (2004). Novel weapons: invasive success and the
evolution of increased competitive ability. Frontiers in Ecology and the
Environment, 2(8), 436-443.
Congress, U. S. (1993). Office of Technology Assessment. (1993). Harmful non-
indigenous species in the United States. US Government Printing Office,
Washington.
Cramer, V. A., Hobbs, R. J., & Standish, R. J. (2008). What's new about old fields? Land
abandonment and ecosystem assembly. Trends in ecology & evolution, 23(2),
104-112.
Crooks, J. A. (2005). Lag times and exotic species: The ecology and management of
biological invasions in slow-motion. Ecoscience, 12(3), 316-329.
Culley, T. M., & N. A. Hardiman. (2007). The beginning of a new invasive plant: a
history of the ornamental Callery pear in the United States. BioScience, 57(11),
956-964.
Culley, T. M., N. A. Hardiman, & J. Hawks. (2011). The role of horticulture in plant
invasions: how grafting in cultivars of Callery pear (Pyrus calleryana) can
facilitate spread into natural areas. Biological invasions, 13(3), 739-746.
20 Flory, L. S. & K. Clay. (2010). Non-native grass invasion suppresses forest succession.
Oecologia, 164(4), 1029-1038.
Forseth, I. N., & Innis, A. F. (2004). Kudzu (Pueraria montana): history, physiology, and
ecology combine to make a major ecosystem threat. Critical reviews in plant
sciences, 23(5), 401-413.
Fridley, J. D. (2012). Extended leaf phenology and the autumn niche in deciduous forest
invasions. Nature, 485(7398), 359-362.
Ghazoul, J. (2004). Alien abduction: disruption of native plant‐pollinator interactions by
invasive species. Biotropica, 36(2), 156-164.
Gross, K.L. & S.M. Emery. (2007). Succession and Restoration in Michigan Old Field
Communities. Old Fields: Dynamics and Restoration of Abandoned Farmland, 9,
162-176.
Inouye, R. S., N. J. Huntly, D. Tilman, J.R. Tester, M. Stillwell, & K. C. Zinnel. (1987).
Old‐field succession on a Minnesota sand plain. Ecology, 68(1), 12
Hellman, J. J., J. E. Byers, B.G. Bierwagen, & J. S. Dukes. (2008). Five potential
consequences of climate change for invasive species. Conservation Biology,
22(3), 534-543.
Keever, C. (1950). Causes of succession on old fields of the Piedmont, North Carolina.
Ecological Monographs, 20(3), 229-250.
Kikuzawa, K. (1995). Leaf phenology as an optimal strategy for carbon gain in plants.
Canadian Journal of Botany, 73(2), 158-163.
21 Mack, R. N., Simberloff, D., Mark Lonsdale, W., Evans, H., Clout, M., & Bazzaz, F. A.
(2000). Biotic invasions: causes, epidemiology, global consequences, and control.
Ecological applications, 10(3), 689-710.
McEwan, R. W., M.K. Birchfield, A. Schoergendorfer, & M. A. Arthur. (2009). Leaf
phenology and freeze tolerance of the invasive shrub Amur honeysuckle and
potential native competitors. The Journal of the Torrey Botanical Society, 136(2),
212-221.
McEwan, R. W., R. J. Brecha, D. R. Geiger, & G. P. John. (2011). Flowering phenology
change and climate warming in southwestern Ohio. Plant Ecology, 212(1) 55-61.
McLaughlin, J. F., J. J. Hellman, C. L. Boggs, & P. R. Ehrlich. (2002). Climate change
hastens population extinctions. Proceedings of the National Academy of Sciences
of the United States of America, 99(9) 6070-6074.
McNeish, R. E., & McEwan, R. W. (2016). A review on the invasion ecology of Amur
honeysuckle (Lonicera maackii, Caprifoliaceae) a case study of ecological
impacts at multiple scales1. The Journal of the Torrey Botanical Society, 143(4),
367-385.
Miller-Rushing, A. J., Høye, T. T., Inouye, D. W., & Post, E. (2010). The effects of
phenological mismatches on demography. Philosophical Transactions of the
Royal Society B: Biological Sciences, 365(1555), 3177-3186.
NOAA, National Oceanic and Atmospheric Administration. (2019). National Climate
Data
Center (NCDC), Climate Normals US. 2019. Retrieved 2021.
22 O’Connell E. & J. Savage. (2020). Extended phenology has limited benefits for invasive species growing at northern latitudes. Biological Invasions, 22(10), 2957-2974. Parmesan, C., & Yohe, G. (2003). A globally coherent fingerprint of climate change impacts across natural systems. Nature, 421(6918), 37-42. Pimentel, D., Lach, L., Zuniga, R., & Morrison, D. (2000). Environmental and economic costs of nonindigenous species in the United States. BioScience, 50(1), 53-65. R Development Core Team. R: a language and environment for statistical computing. Vienna, Austria: R Foundation for Statistical Computing; 2014. Sakai, A. K., F. W. Allendorf, J. S. Holt, D. M. Lodge, J. Molofsky, K. A. With, S. Baughman, R. J. Cabin, J. E. Cohen, N. C. Ellstrand, D. E. McCauley, P. O’Neil, I. M. Parker, J. N. Thompson, and S. G. Weller. (2001). Annual Review of Ecology and Systematics, 32, 305-332. Schwartz, M. W. (1992). Potential effects of global climate change on the biodiversity of plants. The Forestry Chronicle. Schuster, M. J. & J. S. Dukes. (2017). Rainfall variability counteracts N addition by promoting invasive Lonicera maackii and extending phenology in prairie. Ecological Applications, 27(5), 1555-1563. Tilman, D. (1985). The resource-ratio hypothesis of plant succession. The American Naturalist, 125(6), 827-852. Tognetti, P. M., & Chaneton, E. J. (2012). Invasive exotic grasses and seed arrival limit native species establishment in an old-field grassland succession. Biological Invasions, 14(12), 2531-2544 23 Van Kleunen, M., Weber, E., & Fischer, M. (2010). A meta‐analysis of trait differences between invasive and non‐invasive plant species. Ecology letters, 13(2), 235-245. Vincent, M. A. (2005). On the spread and current distribution of Pyrus calleryana in the United States. Castanea, 70(1), 20-32. Warrix, A., & Marshall, J. (2018). Callery pear (Pyrus calleryana) Response to Fire in a Managed Prairie Ecosystem. Invasive Plant Science and Management, 11(1), 27- 32. doi:10.1017/inp.2018.4 Wolfe, L. M. (2002). Why alien invaders succeed: support for the escape-from-enemy hypothesis. The American Naturalist, 160(6), 705-711. Xu, C. Y., K. L. Griffin, & Schuster, W. S. F. (2007). Leaf phenology and seasonal variation of photosynthesis of invasive Berberis thunbergii (Japanese barberry) and two co-occurring native understory shrubs in a northeastern United States deciduous forest. Oecologia, 154(1), 11-21. Young, T.P., D.A, Petersen, & J.J. Clary. (2005). The ecology of restoration: historical links, emerging issues, and unexplored realms. Ecology Letters, 8(6), 662-673. 24 CHAPTER 2 A TRIAL OF FIRE AND ICE: EXPERIMENTAL ASSESSMENT OF NOVEL ECOLOGICAL RESTORATION TECHNIQUES IN MIDWESTERN PRAIRIES INVADED BY PYRUS CALLERYANA INTRODUCTION Ecological invasion presents a threat to a wide variety of ecosystems on Earth and understanding the factors that drive invasion success may be crucial to developing control and restoration strategies. The threat of invasive species may be accelerating due to climate change (NOAA, 2019) as recent evidence suggests that species richness and diversity are threatened by invasive species who benefit from the warming climate (Hejda et al. 2009; Chapman et al. 2012). Invasive species management is a pressing concern in a wide variety of habitats as these species may negatively influence biodiversity, interrupt nutrient cycling, alter patterns of succession, and drive shifts in the composition of native plant communities (Ruesink et al. 1995; Vilá et al. 2011; Flory and Clay, 2010). A variety of factors contribute to the success of invasive species such as escaping natural predation, genetic advantages, taking advantage of empty niche space, and allocating more resources towards reproduction instead of a defense against enemies (Wolfe, 2002; Lavergne and Molofsky, 2007). Some invasive plants release chemicals that may suppress the germination or growth of potential competitors and alter soil conditions (Batten et al. 2004; Callaway and Ridenous, 2004; McEwan et al. 2010). The potential longer-term legacy effects of invasive plants are relatively understudied; however, some 25 work suggests that allelopathic chemicals released by invasive species could remain in the soil for decades (Corbin and Antonio, 2012). Invasive species often have accelerated growth and reproduction relative to native species and can proliferate rapidly in disturbed habitats via the acquisition of available light, nutrients, and space (Lukens and Goessling, 1995; Baruch and Goldstein, 1999) creating a unique physiognomy in some instances (Rowekamp et al. 2020). Many invasive species also have prolonged phenology, giving them a longer growing season and a competitive advantage (McEwan et al. 2009; Xu et al. 2007). Disturbance response is an important factor in invasion success and is strongly enhanced in woody plants that can respond to disturbance by sprouting (Bond and Midgley, 2001; Herrero et al. 2016). The ability to persist and outcompete other species through the mechanism of resprouting can create a significant competitive advantage for some invasive species. Bond and Midgley (2001) describe the concept of a “persistence niche” related to resprouting in which a species that deploys a set of ramets in response to disturbance can establish competitive primacy in a space much more rapidly than potential competitors. When light is abundant, these species will create long shoots (sprouts) and rapidly produce leaves to attain a greater plant height and avoid being shaded (Kikuzawa, 1995). In highly disturbed, early successional environments, some invasive plants channelize most of their energy into resprouting as a way to outcompete other early successional species (Ditomaso et. al, 2006). This poses a significant challenge for land managers because tactics such as fire and mowing can initiate resprouting in some woody invasive species, and rather than impede the success of these species, the management activity may facilitate the invasion process. For example, 26 Herrero et al (2016) found that after a prescribed fire in a fire adapted ecosystem, the invasive species did as well as the native species, and the resprouts from the invasive species were more viable than the native species. In early successional prairies in the US, most native species are adapted to disturbances such as fire and drought. However, woody invasive species, which often invade early successional prairies, are relatively understudied in the literature. Some studies have examined woody early successional invasive species such as Lonicera maackii, Pyrus communis and Ailanthus altissima. These species tend to resprout after a prescribed burn and continue to invade because of their ability to resprout aggressively from their base (Ditomaso et. al, 2006; Grace et. al, 2000). Although resprouting and the persistence niche are important processes during some woody plant invasions, relatively little is known about the biology of this process. Climatic conditions regulate the abundance and distribution of invasive plants and climate change may advance ecological invasion. Previous studies have shown that specific climate envelopes could prevent further invasion by certain invasive species, but with current temperature increases, invasive species are continuing to advance into new ecosystems (Hellman et. al 2008). Increasing climate warming may create conditions in which some cold sensitive invasive species become more widespread due to milder winter temperatures. For example, Ikeda et. al (2014) found that Tamarix invasion will increase by 62% by 2080 due to warmer environments and the lack of frost tolerance needed. In addition, McEwan et al. (2009) found a similar pattern as the invasive shrub L. maackii exhibited greater frost tolerance than co-occurring native shrubs. Lastly, Bradley et al. (2010) predicted that some of the worst invaders within the Eastern United States, including kudzu (Pueraria lobata), privet (Ligustrum sinense; L. vulgare), and 27 cogongrass (Imperata cylindrica) will continue to expand their range due to the warming climate. Invasive species that were once held back by certain climate barriers will now be able to adapt to new environments leaving land managers responsible for trying to manage these new invasive species. However, there are many invasive species that are currently range limited due to their inability to adapt to current climate conditions. For example, common Eastern woody invasive plants are currently range limited such as Amur Honeysuckle (Lonicera macckii), Callery pear (Pyrus calleryana), and Japanese barberry (Berberis thunbergii) but have the potential to continue spreading with climate change. While some woody invasive species such as Amur Honeysuckle are well studied, certain woody species like Pyrus calleryana are relatively understudied. With climate change, little is known about whether this species will expand into new ranges or is limited by other climatic pressures. Pyrus calleryana is an invasive tree that has become an ecological threat across much of the eastern and central United States (Culley and Hardiman, 2007; Culley et. al, 2017). This species was intentionally introduced to the United States in the early 1900s from its native range that spans China, Japan, and Korea, due to horticultural interest related to its resistance to fire blight (Erwinia amylovora) (Meyer, 1918; Culley and Hardiman, 2007). During the 1950’s, many nurseries and horticulturalists recognized its promising ornamental characteristics including visually attractive spring flowering and fall color, along with physiological characteristics such as tolerance to heat, pollution, and drought, resulting in this species becoming a popular ornamental tree in residential areas and in cities (Culley et al. 2011). The original cultivars had minimal reproductive capacity due to an inability to self-pollinate; however, flowering from root sprouts and 28 the introduction of new cultivars released this species from this limitation leading to highly successful reproduction (Culley and Hardiman, 2009; Culley et al. 2011). The original invasion of this species was into habitats in areas adjacent to residential plantings that have ample light and disturbed soils (Vincent, 2005). Pyrus calleryana is bird dispersed and exhibits rapid growth and establishment within habitats that experience regular disturbance (Vincent, 2005, Culley and Hardiman, 2007). Invasion by P. calleryana may be slowed by shady conditions and cold temperatures (Bell et al. 2004) as the spread of this species has been limited in regions that experience winters during which temperature drop below -24°F (Culley and Hardiman, 2007; Phillips, 2004). This aspect of its biology may indicate a vulnerability that could be utilized to slow the spread of Pyrus calleryana. Current treatment for invasive plants varies by species and ecosystem and is currently costing millions of dollars in annual costs across the United States (US Congress, 1993; Kovacs et al. 2010). Management techniques for removal of invasive plants can be labor intensive and common techniques include prescribed fire, herbicide treatment, and tree cutting or girdling (Hartman and McCarthy, 2004; Loh and Daehler, 2007). A variety of techniques have been used to control P. calleryana. For example, some land managers recommend cutting and spraying the tree with glyphosate (Missouri Department of Conservation, 2018; Vogt et al. 2020). However, resprouting has been observed in some herbicide-treated trees (personal communication, Five Rivers Metropark; personal observation, Maloney 2020). In addition, because Pyrus calleryana often invades prairies, prescribed fire and mowing are often techniques that have been used to limit invasion. Recent studies have found that the P. calleryana will re-sprout 29 vigorously after burning and mowing (Warrix and Marshall, 2018; Just et al. 2017). One particularly important component of the invasion biology of P. calleryana is its ability to resprout following disturbance. Pyrus calleryana can begin resprouting as quickly as two weeks after damage (personal observation) and Warrix and Marshall (2018) found an average of 3.1 sprouts emerged from stems following disturbance (Warrix and Marshall, 2018). Advancing scientific understanding of epicormic sprout biology in P. calleryana is a pressing concern for land management in the eastern United States and has strong potential for yielding insights into the role of the persistence niche during woody plant invasion. In this project we aim to (a) understand resprouting as a mechanism of persistence in response to disturbance of P. calleryana (b) understand if sprouting is the dominant mechanism that P. calleryana utilizes for invasion and (c) test various treatments to understand how current management practices impact P. calleryana. We will be testing common restoration methods used in prairies such as fire, mowing, and herbicide, in addition to a new method, freezing with liquid nitrogen. We tested these common treatment methods on previously disturbed P. calleryana(which had been mowed and regrown within the past year), and undisturbed P. calleryana, which was approximately 5-8 years old and has never been disturbed. Our experiments took place in two grasslands in Dayton, OH that are managed by our local partner at Five Rivers Metroparks. We hypothesize that (H1) post-treatment basal diameter will be significantly lower in freezing and herbicide treatments, (H2) post-treatment number of sprouts will be significantly lower in freezing and herbicide treatments, (H3) post-treatment basal diameter will be significantly smaller in undisturbed pears in comparison to disturbed 30 pears and lastly (H4) post-treatment number of sprouts will be significantly less in undisturbed pears in comparison to resprouting pears. 31 METHODS Two experiments were conducted in grasslands near Dayton, Ohio, USA. The first study took place at Shiloh Conservation Area (SCA) near Dayton, Ohio, USA (39°50'24.8"N 84°14'17.9"W) which is owned and managed by the Five Rivers Metroparks. This site is a grassland that is undergoing restoration following > 50 years of use for row crop agriculture. Grassland restoration began approximately 10 years prior to the beginning of this experiment and the principal management technique used to keep the site open has been fall mowing. Vegetation in the site is dominated by agricultural weeds and woody invasive species including Amur honeysuckle (Lonicera maackii) Pyrus calleryana, and multiflora rose (Rosa multiflora). For pretreatment data, we marked and measured the basal diameter for 100 P. calleryana that were cut during mowing approximately one year prior to the start of the experiment and had resprouted. Treatment application began with cutting of all trees using loppers as close to the soil level as possible. Then we randomly assigned n = 20 of the resprouting P. calleryana to one of five treatments: no treatment (negative control), cut only (control), cut and experimental burning, cut and freezing (0.5L of liquid nitrogen application), and cut and 50% glyphosate herbicide. To burn each tree, we mimicked a prairie burn by igniting a 1m circular plot with a drip torch (Figure 2.1). Experimental freezing consisted of placing a piece of metal piping around the base of where each tree was cut to ensure all the liquid nitrogen was applied directly to the tree base (Figure 2.1). The temperature of each tree was taken afterward with an infrared thermometer aimed at the stump after each pour to ensure it reached below -20OF. Resprout success was recorded 6 weeks after treatment in 32 November of 2019. In addition, sprouts were recorded in the spring and fall of 2020. In the Fall of 2020, each of the surviving seedlings was cut at ground level, the basal diameter and number of sprouts were measured in the field. All cut materials were brought into the lab where wet and dried biomass was measured. Experiment 2 of this study tested similar treatments to Experiment 1; however, these treatments were applied to previously uncut P. calleryana at Medlar Conservation Area (MCA) in Miamisburg, OH (39°36'09.4"N 84°16'25.2"W). This site was an agricultural field for >50 years and had been converted to grassland for conservation purposes in 2015. Prior to the beginning of our experiment, the site had not been mowed nor treated, so the trees in this experiment were large, single-stemmed, trees that had not been subjected to management activity following their initial establishment. Treatment application began with cutting of all trees and then each was randomly assigned (n = 10) to one of four treatments: cut only (control), cut and experimental burning, cut and freezing, and cut and 50% glyphosate herbicide. Due to conditions at the site, it was not possible to create a series of small, prescribed fires as was done in the previous experiment; therefore, to mimic a prairie fire, we used a blow torch to heat each stump to 500OF. Resprout success was recorded 6 weeks after treatment in November of 2019. In addition, resprouts were recorded in the spring. In the Fall of 2020, each of the surviving seedlings was cut at ground level, the basal diameter and number of sprouts were measured in the field, and the wet and dried biomass were measured in the laboratory. After all of the resprouting trees were collected, they were placed into brown paper bags and all biomass was weighed within 24 hours of collection to get a wet biomass. Each bag was then placed in an oven for 48 hours at 700C. After the bag was 33 removed from the oven, the bag was reweighed. We then removed all material from the bags, separated the leaf and woody material to get foliar and woody biomass and also measured the bag itself to get a tare weight. For statistical analysis, we ran mixed model ANOVAs and Kruskal-Wallis with a pairwise-wilcox test using R statistical software. To compare pre- and post-treatment, a mixed model ANOVA was completed. To compare biomass and sprouts post treatment, Kruskal-Wallis with a pairwise-wilcox test was conducted. 34 RESULTS In Experiment 1, resprouting P. calleryana stems that had been cut one year earlier exhibited wide variation in response to experimental treatment. For example, the median number of sprouts in the negative control at the conclusion of the experiment was statistically indistinguishable from the cut only treatment and the cut then freeze treatments; however, prescribed burning of stumps resulted in a significant increase in the number of sprouts at the conclusion of the experiment (P = 0.001; Figure 2.2). All trees treated with herbicide died throughout the course of the experiment and therefore herbicide was statistically different from all other treatments. Freezing stumps following cutting resulted in a lower median number of resprouts (8 ± 1.63) than cutting alone (13 ± 1.6); however, these values were statistically indistinguishable (Figure 2.2). The final biomass of resprouting P. calleryana varied in response to experimental treatments (P = 0.001; Figure 2.3). The negative control, which had not been cut or treated, was statistically separated from all treatments and the median biomass (1077 ± 357g) was nearly double the biomass of any of the other treatments. All trees that were treated with herbicide failed to resprout resulting in no accumulation of biomass (Figure 2.3). The cut-only, cut-fire and cut-freeze treatments resulted in final sprout biomass values that were not statistically separable. Even so, we did find that the freeze treatment resulted in a median dry biomass of 250 ± 62.2 g whereas cut and freeze had relatively similar biomass at 385±55.7 g and 319±77.3 g respectively. Lastly, we evaluated how the basal diameter shifted in trees treated before and after treatment. All treatments had a significant effect on the size of their basal diameter 35 pre and post treatment (P<0.001; Supplemental Figure 1) indicating that the treatments did work in reducing the size of Pyrus calleryana. In terms of significance between treatments, the negative control was significantly different from herbicide, cut, freeze and fire due to its increase in the basal diameter. Herbicide was significantly different from all other treatments because all the trees died, and its basal diameter was zero. The basal diameter was reduced in the cut, fire and freeze but they were not significantly different from one another and had approximately the same range of sum basal diameters post- treatment. In Experiment 2, we cut and applied treatments to larger P. calleryana that had not been previously subjected to cutting or other management activity and were not resprouting. There was a significant overall effect of treatments on resprouting (P< 0.001) that was largely driven by the effectiveness of the herbicide treatment relative to the other methods (Figure 2.4). In fact, all the trees that were treated with herbicide died whereas all trees treated with any of the other treatments resprouted (Figure 2.4). The median values for the number of sprouts in the cut-none, cut-fire and cut-freeze treatments were relatively similar (ca. 10 sprouts/tree) (Figure 2.4). These results were generally reflected in the final biomass of resprouts, although the trees that were cut and treated with herbicide did not resprout, so no biomass accumulated whereas all other stems resprouted and therefore had greater biomass (Figure 2.5; P = 0.001). The cut- only, cut-burn and cut-freeze treatments were statistically indistinguishable; however, there was an apparent pattern where the fire treatment had lower biomass than the cut- only stems and the freeze treatments resulted in substantially lower biomass (Figure 2.5). Similar to Experiment 1, all the basal diameters decreased in Experiment 2 between pre 36 and post-treatment (P<0.001; Supplemental Figure 2). While there was significance in basal diameter from pre to post-treatment, there was no significant difference between treatments. 37 DISCUSSION In this study, our goal was to (a) understand resprouting as a mechanism of persistence in response to disturbance of P. calleryana (b) understand if sprouting is the dominant mechanism that P. calleryana utilizes for invasion and (c) test various treatments to understand how current management practices impact P. calleryana. Through this study, we found that P. calleryana has an aggressive resprouting behavior and uses its sprouting to persist within an ecosystem. We believe that this is one of the dominant mechanisms that P. calleryana utilizes for invasion success. Through analyzing various restoration treatment strategies, we found that current restoration techniques may be aiding in the resprouting of P. calleryana, especially in systems that are treated by fire or mowing. In Experiment 1, our hypotheses were confirmed. Post-treatment basal diameter was lower in freezing and herbicide treatments (H1), though only herbicide was able to eradicate P. calleryana. The post-treatment number of sprouts was significantly lower in herbicide treatment (H2), and freezing did produce fewer sprouts but not significantly more than cut fire or cut only. We also saw a decrease in biomass post- treatment in cut only and cut-fire treatment rather than freezing. In Experiment 2, (H3) post- treatment basal diameter was significantly smaller in undisturbed pears in comparison to disturbed pears and (H4) post-treatment number of sprouts were also significantly smaller in undisturbed pears in comparison to resprouting pears. The difference in responses from untreated P. calleryana and previously disturbed P. calleryana is extremely interesting and could be a key in understanding the persistence of 38 P. calleryana within grassland systems and can aid in our understanding of how to approach restoration and conservation of systems where P. calleryana is present. Persistence niche is relatively understudied (Bond and Midgley, 2001) but a mechanism that species like Pyrus calleryana use to proliferate within an ecosystem. Woody species that have the ability to resprout often spend more time sending up shoots when they are younger and less time putting energy into fruits (Kikuzawa, 1995). This is also true in Pyrus calleryana as none of our resprouting trees produced fruits, and only some of our untreated pears were producing fruits before they were treated. The age of the untreated pears was roughly 5-8 years old, meaning that Pyrus calleryana does not usually produce fruit until around this time. There is a potential they can start reproducing earlier (perhaps at 3 years old) but we did not have any trees younger than 5 within our study. This corresponds with other studies that found that the ideal fruiting age of Pyrus calleryana is correlated more with the root collar diameter (Warrix et. al 2017). We also concur with this theory as only the trees with the largest basal diameter in our study were producing fruits. The larger basal diameter of the tree may be a significant factor within our study as we saw older trees respond differently to freezing. This may be due to the fact they use a different mechanism to invade and spend more time putting energy into fruits than resprouting. While all the undisturbed pears still sprouted, their biomass was significantly less than the resprouting pears and they had far fewer sprouts. To understand the age question and also how repeated freezing may affect the growth of Pyrus calleryana, more studies will have to be done. While not statistically significant, we did see fewer sprouts on the pears treated with liquid nitrogen. A study had suggested that Pyrus calleryana cannot grow in 39 climates that are cold as they store more water within their cells and freezing temperature would lead to significant disruptions within the plants (Culley and Hardiman, 2007; Phillips, 2004). Current ongoing studies suggest that Pyrus calleryana is more resilient to freezing than previously thought and can grow in a variety of climates or disturbances (Warrix and Marshall, 2018). With further studies, exploring the relationship on how repeated treatments of liquid nitrogen could potentially aid in a natural remedy to eradicating Pyrus calleryana. The ability to tolerate liquid nitrogen also begs the question about its ability to thrive in more climates than previously thought. With climate change, more habitats will become suitable for Pyrus calleryana and its range will continue to expand. Due to its ability to persist within an ecosystem, we can assume (though it has not been tested) that Pyrus calleryana can store carbohydrates despite disturbance which aid in its ability to resprout. However, we do not understand how repeated disturbance could potentially harm or aid Pyrus calleryana. We noted in one study that the only trees to root sprout at MCA were those that were treated with freezing. While all the trees that were treated with liquid nitrogen still had epicormic sprouting, would repeated freezing cause the tree to choose only root sprouting over epicormic sprouting? There have been no documented cases of Pyrus calleryana escaping cultivations in areas with hard freezes constantly all winter, which means that potentially repeated freezing could aid as a natural remedy to treat Pyrus calleryana. The resprouting ability in Pyrus calleryana is aggressive and previous studies have shown that fire and cutting only contribute to its ability to resprout, which might aid the plant in proliferating within a system (Warrix and Marshall, 2018). Part of the reason 40 that Pyrus calleryana is able to sprout so aggressively is because early successional conditions have ample resources, where plants are able to replace leaves with higher photosynthetic ability in a short amount of time (Kikuzawa, 1995). Usually, plants that are efficient at utilizing light in high intensity environments do poorly in low light environments, hence why we do not see Pyrus calleryana succeeding into forest canopy where there is competition for light availability. In addition to putting out leaves, Pyrus calleryana aggressively sends out new shoots through epicormic sprouting replacing leaves and quickly building a productive system with many leaves. Because of the ample space available in old fields, Pyrus calleryana also grows in height relatively quickly. After 1-year post-treatment, almost all trees grew back to over 1 meter tall. This result likely occurred because plants with ample space will elongate their shoot vertically to avoid shading and create a structure that is best suited for high light environments (Kikuzawa, 1995). One of the relatively surprising results of our experiment was the shoots on the larger non-resprouting Pyrus calleryana trees at Medlar did not start resprouting until late June. On the other hand, the trees at Shiloh began resprouting as early as March. We believe this phenomenon is related more to the fact that the trees at Shiloh had already been resprouting whereas the ones at Medlar had not put resources into resprouting behavior. Further research into the phenological timing of resprouts will need to be done in order to understand this phenomenon. Woody plant encroachment into grasslands is a global concern that is increasingly augmented by the ever-expanding panoply of woody invasive species (Grace et. al 2000; Ratajczak et. al 2012). Woody invasive species change traditional successional trajectories by limiting the growth of native species and forming new novel ecosystems 41 (Flory and Clay, 2010). In this study, we saw that Pyrus calleryana can readily persist in systems with high disturbance. With Pyrus calleryana outcompeting native species, the successional trajectories will shift, leading the system to a new novel system and deterring the system from restoration or conservation targets for land managers. The largest area of concern for land managers should be that Pyrus calleryana can persist within a system without needing aid from pollinators or seed dispersal. Once established, it can clone itself within the system and continue spreading (Culley et. al 2011). As we have seen, it feeds off systems that are highly disturbed as disturbance will aid in its proliferation. From a land management perspective, treating Pyrus calleryana with herbicide through a cut spray method may be necessary prior to burning or mowing a prairie. Through conversations with land managers within Dayton Ohio, it is also critically important to think about when you are applying your herbicide. As summer progresses, leaves on Pyrus calleryana will become waxier, potentially making it more resistant to foliar spray of herbicide. This is not proven in this study but has been brought up as a reason why some land managers believed that Pyrus calleryana was resistant to herbicide. However, Vogt et. al (2020) concluded that all herbicides work well in killing Pyrus calleryana, and a foliar glyphosate solution was the most consistent. The best method to minimize harming native species would be to cut and spray the stump of the tree to ensure the tree does not resprout. It is also important to ensure the whole stump was treated, as we found some Pyrus calleryana that appeared to be growing out of dead stumps, but on further investigation, the stump was not completely coated with herbicide. With the resprouting pear trees at SCA, we saw that disturbances aids significantly in the pear’s ability to resprout. While the basal diameter was not as large after disturbance, 42 more resprouts mean more leaves that will aid in significant plant regeneration. This means that current land management practices are aiding in the growth of P. calleryana. We saw a similar response in older P. calleryana trees at MCA. Therefore, land managers need to think carefully about how they are going to manage systems with heavy Pyrus calleryana invasion. As we discovered, Pyrus calleryana will die with herbicide, but large broadcast spraying of fields will severely damage native species that are living within the system. Pyrus calleryana is quickly becoming one of the most problematic invasive species for land managers within the eastern and central parts of the United States (Culley and Hardiman, 2007; Warrix and Marshall, 2018). In this study, we found that resprouting is one of the main mechanisms of persistence in response to disturbance of P. calleryana and identified that current land management practices such as fire and mowing are aiding to the resprout and growth of P. calleryana. Herbicide application was the best method to control P. calleryana resprout response and aid in stopping the spread. It is recommended that in grassland and prairies that fire is used as a management practice, that herbicide is used before burn season. In addition, more research will be needed to understand if multiple freeze treatments may be a viable option for P. calleryana control and why P. calleryana chooses root sprouts over epicormic sprouting. Lastly, more public outreach is needed to educate the public about removing or preventing plantings of P. calleryana to ensure that the species does not enter more grassland ecosystems. 43 CHAPTER 2 FIGURES Figure 2. 1 Examples of experimental treatments for Pyrus calleryana on recently restored grasslands near Dayton, Ohio, USA. The first treatment (left) consisted of a prescribed fire using a drip torch to ignite a 1m plot containing a single P. calleryana stump. The second treatment (right) consisted of pouring 0.5L of liquid nitrogen on a P. calleryana stump using a small dewar. 44 Figure 2. 2 Total number of resprouts on previously disturbed Pyrus calleryana in an old field in Dayton, OH. The number of sprouts were recorded after 1 year of growth post- treatment. This metric was calculated by counting all the resprouts (sprouts that grew post- treatment) on each tree within each treatment type. Letters represent statistically significant differences (P<0.001). The centerline of the box plot represents the median number of sprouts for each treatment and each dot represents the number of sprouts for a tree within each treatment. 45 Figure 2. 3 Total dry mass (g) on previously disturbed Pyrus calleryana in an old field in Dayton, OH. The total dry mass was recorded after 1 year of growth post treatment. This metric was calculated by harvesting all above-ground biomass, drying all the biomass in an oven at 70OC for 48 hours and weighing it. Letters represent statistically significant differences (P<0.001). The centerline of the box plot represents the median dry mass for each treatment and each dot represents the total dry mass for a tree within each treatment. 46 Figure 2. 4 Total number of resprouts on undisturbed Pyrus calleryana in an old field in Dayton, OH. The number of sprouts were recorded after 1 year of growth post -treatment. This metric was calculated by counting all the resprouts (sprouts that grew post - treatment) on each tree within each treatment type. Letters represent statistically significant differences (P<0.001). The centerline of the box plot represents the median number of sprouts for each treatment and each dot represents the number of sprouts for a tree within each treatment. 47 Figure 2. 5 Total dry mass (g) on undisturbed Pyrus calleryana in an old field in Dayton, OH. The total dry mass was recorded after 1 year of growth post- treatment. This metric was calculated by harvesting all above-ground biomass, drying all the biomass in an oven at 70OC for 48 hours and weighing it. Letters represent statistically significant differences (P<0.001). The centerline of the box plot represents the median dry mass for each treatment and each dot represents the total dry mass for a tree within each treatment. 48 Supplemental Figure 1 The pre and post- treatment basal diameters on previously disturbed Pyrus calleryana in old fields within Dayton, OH. Pre-treated basal diameter was taken by calipers before cutting each tree. Post-treatment basal diameter was taken after 1 year of growth following the application of a treatment with calipers. All treatments saw a significant decrease in basal diameter from pre and post- treatment (red line; P<0.001). There was a significant difference between treatments- herbicide and negative control were significantly different (P<0.001) from the cut, fire and freeze treatment. The centerline of the box plot represents the median basal diameter for each treatment. 49 Supplemental Figure 2 The pre and post- treatment basal diameters on undisturbed Pyrus calleryana in old fields within Dayton, OH. Pre-treated basal diameter was taken by calipers before cutting each tree. Post- treatment basal diameter was taken after 1 year of growth following the application of a treatment with calipers. All treatments saw a significant decrease in basal diameter from pre and post- treatment (red line; P<0.001). There was no significant difference between treatments. The centerline of the box plot represents the median basal diameter for each treatment. 50 LITERATURE CITED Baruch, Z., & Goldstein, G. (1999). Leaf construction cost, nutrient concentration, and net CO2 assimilation of native and invasive species in Hawaii. Oecologia, 121(2), 183-192. Batten, K. M., Scow, K. M., Davies, K. F., & Harrison, S. P. (2006). Two invasive plants alter soil microbial community composition in serpentine grasslands. Biological Invasions, 8(2), 217-230. Bell, A. C., Ranney, T. G., Eaker, T. A., & Sutton, T. B. (2005). Resistance to fire blight among flowering pears and quince. HortScience, 40(2), 413-415. Bond, W. J., & Midgley, J. J. (2001). Ecology of sprouting in woody plants: the persistence niche. Trends in ecology & evolution, 16(1), 45-51. Bradley, B. A., Wilcove, D. S., & Oppenheimer, M. (2010). Climate change increases risk of plant invasion in the Eastern United States. Biological Invasions, 12(6), 1855-1872. Callaway, R. M., & Ridenour, W. M. (2004). Novel weapons: Invasive success and the evolution of increased competitive ability. Frontiers in Ecology and the Environment, 2(8), 436–443. Chapman, J. I., Perry, K. L., & McEwan, R. W. (2012). Changing flora of an old-growth mesophytic forest: Previously undetected taxa and first appearance of non-native invasive species. The Journal of the Torrey Botanical Society, 139(2), 206-210. Congress, U. S. Office of Technology Assessment. (1993). Harmful non-indigenous species in the United States. US Government Printing Office, Washington. 51 Corbin, J. D., & D'Antonio, C. M. (2012). Gone but not forgotten? Invasive plants' legacies on community and ecosystem properties. Invasive Plant Science and Management, 5(1), 117-124. Culley, T. M., & Hardiman, N. A. (2007). The beginning of a new invasive plant: a history of the ornamental Callery pear in the United States. BioScience, 57(11), 956-964. Culley, T. M., & Hardiman, N. A. (2009). The role of intraspecific hybridization in the evolution of invasiveness: a case study of the ornamental pear tree Pyrus calleryana. Biological Invasions, 11(5), 1107-1119. Culley, T. M., Hardiman, N. A., & Hawks, J. (2011). The role of horticulture in plant invasions: how grafting in cultivars of Callery pear (Pyrus calleryana) can facilitate spread into natural areas. Biological invasions, 13(3), 739-746. Culley, T. M. (2017). The rise and fall of the ornamental Callery pear tree. Arnoldia, 74(3), 2-11. DiTomaso, J. M., Brooks, M. L., Allen, E. B., Minnich, R., Rice, P. M., & Kyser, G. B. (2006). Control of invasive weeds with prescribed burning. Weed technology, 20(2), 535-548. Flory, S. L., & K. Clay. (2010). "Non-native grass invasion suppresses forest succession." Oecologia 164(4): 1029-1038. Grace, J. B., Smith, M. D., Grace, S. L., Collins, S. L., & Stohlgren, T. J. (2000). Interactions between fire and invasive plants in temperate grasslands of North America. In Proceedings of the invasive species workshop: the role of fire in the control and spread of invasive species. Fire conference (pp. 40-65). 52 Hejda, M., Pyšek, P., & Jarošík, V. (2009). Impact of invasive plants on the species richness, diversity and composition of invaded communities. Journal of Ecology, 97(3), 393-403. Hellman, J. J., J. E. Byers, B.G. Bierwagen, & J. S. Dukes. (2008). Five potential consequences of climate change for invasive species. Conservation Biology, 22(3), 534-543. Ikeda, D. H., Grady, K. C., Shuster, S. M., & Whitham, T. G. (2014). Incorporating climate change and exotic species into forecasts of riparian forest distribution. PLoS One, 9(9), e107037. Just, M. G., Hohmann, M. G., & Hoffmann, W. A. (2017). Invasibility of a fire- maintained savanna–wetland gradient by non-native, woody plant species. Forest ecology and management, 405, 229-237. Kikuzawa, K. (1995). Leaf phenology as an optimal strategy for carbon gain in plants. Canadian Journal of Botany, 73(2), 158-163. Luken, J. O., & Goessling, N. (1995). Seedling distribution and potential persistence of the exotic shrub Lonicera maackii in fragmented forests. American Midland Naturalist, 124-130. Hartman, K. M., & McCarthy, B. C. (2004). Restoration of a forest understory after the removal of an invasive shrub, Amur honeysuckle (Lonicera maackii). Restoration Ecology, 12(2), 154-165. Herrero, M. L., Torres, R. C., & Renison, D. (2016). Do wildfires promote woody species invasion in a fire-adapted ecosystem? Post-fire resprouting of native and non- 53 native woody plants in central Argentina. Environmental management, 57(2), 308-317. Kovacs, K. F., Haight, R. G., McCullough, D. G., Mercader, R. J., Siegert, N. W., & Liebhold, A. M. (2010). Cost of potential emerald ash borer damage in US communities, 2009–2019. Ecological Economics, 69(3), 569-578. Lavergne, S., & Molofsky, J. (2007). Increased genetic variation and evolutionary potential drive the success of an invasive grass. Proceedings of the National Academy of Sciences, 104(10), 3883-3888. Loh, R. K., & Daehler, C. C. (2008). Influence of woody invader control methods and seed availability on native and invasive species establishment in a Hawaiian forest. Biological Invasions, 10(6), 805-819. McEwan, R. W., Birchfield, M. K., Schoergendorfer, A., & Arthur, M. A. (2009). Leaf phenology and freeze tolerance of the invasive shrub Amur honeysuckle and potential native competitors. The Journal of the Torrey Botanical Society, 136(2), 212-221. McEwan, R. W., Arthur-Paratley, L. G., Rieske, L. K., & Arthur, M. A. (2010). A multi- assay comparison of seed germination inhibition by Lonicera maackii and co- occurring native shrubs. Flora-Morphology, Distribution, Functional Ecology of Plants, 205(7), 475-483. Meyer, F. N. (1918). Typescript of South China Explorations. Special Collections of the National Agricultural Library. 54 Missouri Department of Conservation. (2018). Invasive Species: Callery Pear [PDF File]. Retrieved from https://mdc.mo.gov/sites/default/files/downloads/callerypearinvasive.pdf NOAA, National Oceanic and Atmospheric Administration. (2019). National Climate Data Center (NCDC), Climate Normals US. 2019. Retrieved 2021. Phillips, L. (2004). The 2005 urban tree of the year. City Trees, 40, 34-38. Ratajczak, Z., Nippert, J. B., & Collins, S. L. (2012). Woody encroachment decreases diversity across North American grasslands and savannas. Ecology, 93(4), 697- 703. R Development Core Team. R: a language and environment for statistical computing. Vienna, Austria: R Foundation for Statistical Computing; 2014. Rowekamp, E. C., Chapman, J. I., & McEwan, R. W. (2020). Assessing the influence of riparian invasion by the shrub Lonicera maackii on terrestrial subsidies to headwater streams. Acta Oecologica, 105, 103580 Ruesink, J. L., Parker, I. M., Groom, M. J., & Kareiva, P. M. (1995). Reducing the risks of nonindigenous species introductions. BioScience, 45(7), 465-477. Vilà, M., Espinar, J. L., Hejda, M., Hulme, P. E., Jarošík, V., Maron, J. L., ... & Pyšek, P. (2011). Ecological impacts of invasive alien plants: a meta‐analysis of their effects on species, communities and ecosystems. Ecology letters, 14(7), 702-708. Vincent, M. A. (2005). On the spread and current distribution of Pyrus calleryana in the United States. Castanea, 70(1), 20-32. 55 Vogt, J. T., Coyle, D. R., Jenkins, D., Barnes, C., Crowe, C., Horn, S., ... & Roesch, F. A. (2020). Efficacy of five herbicide treatments for control of Pyrus calleryana. Invasive Plant Science and Management, 13(4), 252-257. Warrix, A. R., Myers, A. L., & Marshall, J. M. (2017). Estimating invading Callery pear (Pyrus calleryana) age and flowering probability in an Indiana managed prairie. In Proc Indiana Acad Sci, 126. Warrix, A., & Marshall, J. (2018). Callery pear (Pyrus calleryana) Response to Fire in a Managed Prairie Ecosystem. Invasive Plant Science and Management, 11(1), 27- 32. doi:10.1017/inp.2018.4 Wolfe, L. M. (2002). Why alien invaders succeed: support for the escape-from-enemy hypothesis. The American Naturalist, 160(6), 705-711. Xu, C. Y., Griffin, K. L., & Schuster, W. S. F. (2007). Leaf phenology and seasonal variation of photosynthesis of invasive Berberis thunbergii (Japanese barberry) and two co-occurring native understory shrubs in a northeastern United States deciduous forest. Oecologia, 154(1), 11-21. 56