1 the Role of Biotic Resistance Through Predation on the Invasion Success

Home , Petrolisthes, Petrolisthes armatus, Resistance (ecology)

The role of biotic resistance through predation on the invasion success of the green

porcelain crab (Petrolisthes armatus) into nearshore oyster reef communities.

Thesis

Presented in Partial Fulfillment of the Requirements for the Degree Master of Science in

the Graduate School of The Ohio State University

Kaitlin Alyse Kinney

Graduate Program in Environment and Natural Resources

The Ohio State University

2017

Thesis Committee

Dr. Lauren M. Pintor, Advisor

Dr. James E. Byers

Dr. Stuart A. Ludsin

Dr. Christopher M. Tonra

Copyrighted by

Kaitlin Alyse Kinney

2017

Abstract

Studying the edge of a non-native species distribution can help us understand what limiting factors may be preventing its expansion. These factors may include physical tolerances, geographic barriers, competition with native species or predation by native predators and can even work in conjunction with each other to prevent invasions.

The alarmingly high abundances of some non-native species may even promote limiting factors like predation by increasing interaction rates and thus consumption by native predators. In this study, I examined whether biotic resistance through predation by native predators may limit the northward spread of the non-native, invasive filter feeding crab

Petrolisthes armatus into oyster reef communities along the Southeastern US. The only hypothesized limitation on the expansion of this species is the cold snaps associated with northern winters. However several native predators in oyster reefs have been shown to consume this abundant and profitable prey item, suggesting that biotic resistance through predation may be an additional factor limiting its northward spread. My main objectives were to 1) determine if the per capita predation risk exerted by native predators might be a factor that explains the current distribution of P. armatus, and 2) test whether the consumption and preference of P. armatus by the native predatory crab Panopeus herbstii varied across different relative abundances of native to invasive prey. I predicted that if predation limits the spread of P. armatus, then predation risk should be highest at

ii the northern edge of its range. Additionally, I predicted that if the relative abundance of native prey affects the consumption of P. armatus by a native predator, then consumption of P. armatus should be higher when P. armatus is proportionally more abundant than native prey. To test these hypotheses, I conducted a field study to quantify predation risk across 8 invaded estuary sites along the Southeastern US coast from St. Augustine, FL to

North Inlet, SC and conducted a controlled lab experiment to quantify the consumption and preference of P. armatus when in low to high abundance relative to alternative native prey. While predation rates were high across sites (68.2 – 98.2%), there was no significant relationship between predation and latitude across the 8 invaded estuaries.

Furthermore, while P. herbstii increased consumption of P. armatus in response to increased abundance in the tank, P. herbstii always showed a preference for native prey regardless of its relative abundance. Further analyses on environmental factors across the sites and a predator exclusion caging experiment suggest that habitat quality, density of

P. armatus, and trophic interactions may influence predation risk of an invasive prey.

Overall, I found no evidence that native predators are preventing the spread of P. armatus and this species is likely to continue its expansion into northern waters as sea temperatures increase with climate change.

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To my family,

Thank you for teaching me to chase my dreams

No matter where they take me

I will always love you

Acknowledgments

There are so many people and organizations I would like to thank for their support during my career without whom none of this would have been possible. I would like to start out by thanking my advisor, Dr. Lauren Pintor for this opportunity to pursue my degree and her unwavering confidence in me. It has been a wonderful experience being a part of such an inspiring female ecologist’s lab. From her willingness to have long, in depth meetings going over every detail of my project to encouraging text messages while

I was away at conferences or field work, she is always challenging me to do my best work while always having my best interest at heart. I would like to thank the OARDC

SEEDS organization for providing partial funding for this project. To my committee members and teachers at OSU, Dr. Stu Ludsin and Dr. Chris Tonra, thank you for your lessons and advice throughout this project.

I would like to thank my final committee member, Dr. Jeb Byers for his formative teaching, collaborative nature and guidance throughout my undergraduate and graduate career. I would not have met Lauren or had the opportunity to work on the Petrolisthes project without his guidance. Finally, I would like to thank Jeb for allowing us to use his lab space and field resources to complete this project. To the members of the Byers Lab:

Jenna Malek, Carrie Keogh, Alyssa Gehman, Rachel Smith, Daniel Harris and Linsey

Haram, thank you for the encouraging words, sharing your site selection wisdom, and

v teaching me how to science. I would especially like to thank Linsey Haram for taking me

“under her wing” all those years ago and teaching me to appreciate the small things, like

“snails not whales”.

I would like to thank all the former and current members of the Pintor Lab but especially Chris Johnson, Jenna Odegard, Chelsea Crosby, Liz Berg, and Alec Mell for always providing a safe space to share ideas and talk science. Particularly to Alec Mell, I would not have been able to do this project without his dedication to detail, readiness for anything and bravery in the face of gators. I would like to thank my fellow OSU grad students and “Aquatic Warrior” sister’s for their friendship. I would like to thank the many faculty, staff and friends at all the marine institutes I visited including Joel Fodrie’s team at the University of North Carolina-IMS, Paul Kenny at North Inlet NERR, the

College of Charleston Grice Marine Lab and GTM NERR. As for the Skidaway Institute of Oceanography, I would like to thank my friends, especially Bob Allen and Jessica

Pruett, for so many wonderful memories over the years and for making the island feel like my summer home.

Finally, I’d like to thank my family and friends. To my friend Lindsey Sherwood, thank you for always knowing how to make me laugh and for appreciating all things fluffy. To my family: Marilyn, Marc, Jacob and Samuel, thank you for always reminding me what is important in life, for loving me no matter what, for the many phone calls and for making anything an excuse to go to Disney World together. To Tyler Williams, thank you for reminding me to go outside, for enabling me to follow my dreams, for inspiring me to be my best self, and for defeating Gannon to save Hyrule.

Vita

May 2010 ...... Cartersville High School, GA

December 2013 ...... B.S. Biological Sciences,

University of Georgia, Athens

August 2015 to Present ...... Graduate Associate, School of Environment

...... and Natural Resources, The Ohio State

...... University

Fields of Study

Major Field: Environment and Natural Resources Minor Field: Fisheries and Wildlife

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Table of Contents

Abstract ...... ii Acknowledgments ...... v Vita ...... vii List of Tables ...... x List of Figures ...... xi Chapter 1 Does predation risk limit the spatial distribution of the non-native invasive green porcelain crab, Petrolisthes armatus?...... 1 INTRODUCTION ...... 1 METHODOLOGY ...... 5 Study Sites ...... 5 1 Pattern of abundance of P. armatus along the US Atlantic Coast ...... 6 2 Per capita predation risk on P. armatus along its invasive range ...... 8 3 Environmental factors that contribute to explaining risk of predation on P. armatus ...... 9 4 Identifying predators responsible for P. armatus predation ...... 12 RESULTS ...... 14 1 Pattern of abundance of P. armatus along the US Atlantic Coast ...... 14 2 Per capita predation risk on P. armatus along its invasive range ...... 16 3 Environmental factors that contribute to explaining risk of predation on P. armatus ...... 16 4 Identifying predators responsible for P. armatus predation ...... 16 DISCUSSION ...... 17 LITERATURE CITED ...... 24 TABLES AND FIGURES ...... 35 Chapter 2 The influence of relative prey abundance on consumption of the non- native, invasive green porcelain crab (Petrolisthes armatus) by a native predator ... 42 INTRODUCTION ...... 42 METHODOLOGY ...... 46 Data Analysis ...... 48 RESULTS ...... 50

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DISCUSSION ...... 52 LITERATURE CITED ...... 57 TABLES AND FIGURES ...... 62 Chapter 3 Applications to the study and management of invasive species...... 67 HOW ARE INVASIVE PREY ABLE TO PERSIST DESPITE BEING READILY CONSUMED BY NATIVE PREDATORS?...... 67 CONSERVATION MANAGEMENT EFFECTS ON BIOTIC RESISTANCE THROUGH PREDATION ...... 72 BIOTIC RESISTANCE THROUGH PREDATION IN THE FACE OF CLIMATE CHANGE ...... 74 LITERATURE CITED ...... 76 References ...... 82 Appendix A: Additional Figures for Chapter 1 ...... 98 Appendix B: Additional Figures for Chapter 2 ...... 101

List of Tables

Table 1.1 Average values of each predictive variable included in the GLMM model

examining which environmental factors affect predation risk on P. armatus

regardless of latitude...... 36

Table 1.2 Ranking of top models with a ΔAIC score less than five that predict the

predation events on P. armatus using the possible variables of the density of

native prey, the density of large oysters, the density of P. armatus, the abundance

of native predators and the mean summer temperature...... 40

Table 2.1 Sample size of each treatment combination in 2016, 2017, and the total of the

two years combined...... 63

Table 2.2 ANOVA table showing each of the 4 factors...... 63

Table 2.3 Estimates for population level preference shown for both native and non-native

prey as well as percent of individuals showing expected preference for P. armatus

across the six Bayesian analyses...... 63

Table A.1 The coordinates and habitat variables not used in analyses for the 25 reefs

across the 9 estuaries...... 98

Table B.1 Additional ANOVA analysis exploring interaction terms...... 101

List of Figures

Figure 1.1 Map of the 9 estuaries surveyed in this study...... 35

Figure 1.2 Mean density of P. armatus across the sampled range...... 37

Figure 1.3 Mean proportion of P. armatus within the prey base across the sampled

estuaries...... 38

Figure 1.4 The mean predation risk on P. armatus across the 8 estuaries invaded by P.

armatus...... 39

Figure 1.5 Proportion of P. armatus eaten by predators across different cage treatments.

...... 41

Figure 2.1 A-F Description of the 6 different relative abundance treatments in the prey

choice assays...... 62

Figure 2.2 A-D Bar plots of the mean logarithmic total number of P. armatus eaten + 1

per individual given A) the relative abundance treatment, B) the alternative native

prey species, C) the sex of the native predator and D) the year the trial was run. 64

Figure 2.3 A-F Population-level preferences (solid lines) and individual-level

preferences (dotted lines) of P. herbstii for both native and invasive prey items. 65

Figure A.1 Analysis of the predator exclusion experiment with the 6 compromised

blocks included...... 100

Figure B.1 A-F Population-level preferences (solid lines) and individual-level

preferences (dotted lines) of Female P. herbstii for both native and invasive prey

items...... 102

Figure B.2 A-F Population-level preferences (solid lines) and individual-level

preferences (dotted lines) of Male P. herbstii for both native and invasive prey

items...... 103

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Chapter 1 Does predation risk limit the spatial distribution of the non-native, invasive green porcelain crab, Petrolisthes armatus?

INTRODUCTION

The geographic range and abundance of non-native species following introduction often depends on the interaction of multiple factors within a native ecosystem (Sakai et al. 2001, Arim et al. 2006, Hayes & Barry 2008). Abiotic conditions, such as temperature, can often limit the spread of a non-native species depending upon its thermal tolerance (Ford 1996, Stachowicz 2002, Sorte et al. 2010). Yet if a non-native species can tolerate local environmental conditions, then biotic resistance can play a role in determining invasion success. Biotic resistance through predation by native predators has been frequently shown to limit the local abundance of an invader (Baltz & Moyle

1993, Reusch 1998, Byers 2002, DeRivera et al. 2005, Dumont et al. 2011, Yamanishi et al. 2012). For example, predation by native benthic predators has prevented the establishment of invasive ascidians (Ciona intestinalis) and restricted its invasion success to suspended artificial structures in marine systems (Dumont et al. 2011). Given that predation is a strong force that structures marine communities, it’s not surprising that this interaction frequently limits invasions in marine systems (Kimbro et al. 2013, Papacostas et al. 2017). Yet predation pressure may also very geographically (Hewitt 2002, Ruiz et al. 2009, Freestone et al. 2013) and can frequently depend on the specific predators with which a non-native species co-occurs (DeRivera et al. 2005, Jensen et al. 2007, Dumont 1 et al. 2009). Here, I examine whether biotic resistance through predation may limit the geographic distribution and local abundance of a non-native prey species hypothesized to be limited by physical tolerances.

Petrolisthes armatus, the green porcelain crab, is a non-native, invasive crab whose northern range is hypothesized to be limited by its susceptibility to cold temperatures (Knott et al. 2000, Stillman & Somero, 2000, Hadley et al. 2010, Canning-

Clode et al. 2011, Eash-Loucks et al. 2014). Specifically, P. armatus appears unable to withstand the severe winter temperatures or ‘cold snaps’ associated with northern sites along the eastern US coast (Canning-Clode et al. 2011). Furthermore, although summertime densities of P. armatus have been reported as high as several thousand per/m2 (Hollebone & Hay, 2007a), abundances decline in substantially colder months and at higher latitudinal sites (Hartman & Stancyk, 2001). The ephemeral nature of the species’ northern edge has made its northernmost distribution difficult to ascertain, but some evidence suggests it is slowly expanding (Wassick 2016). Though it seems likely temperature has a large influence on the leading edge, there may be other limiting factors affecting the range of this species.

Predation by native predators within invaded oyster reef communities suggests that biotic resistance from predation may interact with abiotic conditions to limit the geographic range of non-native P. armatus (Hollebone & Hay 2008, Pintor & Byers

2015). For example, Panopeus herbstii, the Atlantic Mud Crab, is a widespread, important generalist predator that has been shown to readily incorporate P. armatus within their diet (Hollebone & Hay 2008, Hostert 2014, Pintor & Byers 2015). Similarly,

2 predatory fish, such as the mummichog Fundulus heteroclitus, has been shown to consume P. armatus in a laboratory setting (Hollebone & Hay 2008). Additionally, predator species from the genus Callinectes that are common predatory crabs within these oyster reef communities and have been suggested to limit the spread of other invasive species (Harding 2003, DeRivera et al. 2005, Carlsson et al. 2011, Crosby & Pintor unpublished data). Although predation by these predators have only been directly observed under laboratory conditions and field tethering trials at a single invaded site

(Hollebone & Hay 2008), I hypothesize that biotic resistance may be a another process limiting the spread of this non-native, invasive species.

Although there is a diverse suite of predator species that have been shown to consume non-native P. armatus in the laboratory, the abundance of these predators can vary widely in natural ecosystems (Wenner & Wenner 1989, Gehman et al. 2017). Such variation can affect the strength of biotic resistance from predation, especially if it is the case that realized predation of P. armatus in the field is primarily due to only one or two predator species. For instance, every predatory species may not consume P. armatus at the same rate (Magoulick & Lewis 2002, Hollebone & Hay 2008, Dick et al. 2013).

Hollebone and Hay (2008) compared consumption rates of P. armatus within the lab by a suite of predator species thought to function as generalist predators on the reef. Although

P. herbstii, Callinectes similis and F. heteroclitus readily consumed P. armatus, the native fish, Leiostomus xanthurus (Spot) avoided P. armatus. Even within a species, there can be individual variation in consumption of a non-native prey (Réale et al. 2007, Pintor

& Byers 2015) that can alter the total strength of predation exerted on a non-native prey

3 population. For instance, female individuals within a single population of P. herbstii have been shown to incorporate more P. armatus within their diet in comparison to males

(Pintor & Byers, 2015). Consumption could also be structured by the size class of the native predator (Truemper & Lauer 2005, Toscano & Griffen 2012, Pintor & Byers

2015). For example, smaller P. herbstii that were less competitive for prey items regularly consumed P. armatus more than larger more dominant individuals (Pintor &

Byers 2015). Because of these known differences in consumption based on predator species and size, it is important to understand the predator composition, including the identity, abundance and demography of predator species.

Finally, additional factors such as habitat availability (Byers 2002, Dumont et al.

2011), the number of alternative prey items (Pyke 1984, Stephens 1986, Magoulick &

Lewis 2002) or abiotic conditions (Sanford 2002, Ferrari et al. 2015) can mediate the ability of native predators to consume P. armatus. In these intertidal communities invaded by P. armatus, Crassostrea virginica, the Eastern oyster, acts as an ecosystem engineer (Gutierrez 2003, Byers et al. 2015). Higher oyster abundance at a site increases habitat complexity (Gutierrez 2003) which can alter the interaction strength between predators and prey (Grabowski 2004). Yet, the consumption of a non-native prey species by native predators is also likely to be influenced by its density on a reef as well as the density of alternative native prey (Krebs & Davies 1981, Pyke 1984, Stephens & Krebs

1986). For example, in the Great Lakes, many native predators have switched to consuming zebra mussels, Dreissena polymorpha, as this species becomes very abundant in the freshwater communities it invades (French & Bur, 1993, Molloy et al., 1997).

Temperature is an abiotic factor that, in addition to directly affecting non-native species’ physiological tolerances, is also known to affect metabolic rates and influence consumption rates by predators (Sanford 2002, Ferrari et al. 2015). Studying the importance of temperature in predicting predation rates on a non-native prey species could elucidate how communities will respond to changing climates (Stachowicz et al.

2002, Sorte et al. 2010).

I set four main objectives to evaluate whether biotic resistance via predation may also limit the current distribution and abundance of P. armatus in its non-native range.

First I quantified the pattern of density of P. armatus along its invaded range. I predicted that the density of P. armatus would decline from southern to northern sites. Second I determined if the per capita predation risk changes with latitude and might be a factor that explains the current distribution of P. armatus. If predation by native predators helps to limit the spread of P. armatus northward, then predation rates by native predators should increase with increasing latitude. Third regardless of latitude, I examined whether other environmental factors contribute to explaining rates of predation on P. armatus.

Finally, I determined what predators might be responsible for consuming P. armatus in a field setting.

METHODOLOGY

Study Sites

In order to explore the latitudinal patterns of both abundance and predation of P. armatus, I systematically selected eight replicate estuaries equally spaced across the invasive range of P. armatus (Figure 1.1). I chose the lower limit of St. Augustine, FL

5 because south of this site estuarine habitat becomes mixed with both oyster and mangrove communities (Saintilan et al. 2014). I also included a control site that is outside the invasive range of P. armatus but is predicted to be invaded with changing climate

(Canning-Clode et al. 2011). These sites were visited during the summer season of 2016 from June 8 - July 27. For each of the nine estuaries (eight invaded, one control), I selected replicate oyster reefs that were of similar distance from the mouth of the estuary and were backed by Spartina alterniflora vegetation. I chose a minimum of two reefs, three when available, that were separated by at least 10 m, but similar in location and habitat within the estuary. To maximize the ability to detect differences in the effects of latitude and predation pressure, I chose reefs (25 total) across the nine estuaries that were similar in habitat quality and abiotic factors based on salinity, tidal range, slope of reef and height of reef (Table A.1).

1 Pattern of abundance of P. armatus along the US Atlantic Coast

To quantify the pattern of abundance of P. armatus across its invaded range I enumerated the density of P. armatus and the proportion of P. armatus in the prey base for each reef across the nine estuaries described above. P. armatus density and alternative native prey taxa was quantified at each reef by haphazardly placing a 0.25 m2 quadrat 1 m up from the bottom (waterward) edge of the reef. I rapidly excavated this area by hand, collecting all the oyster shell down to approximately 5 cm beneath the mud surface and rinsed the collected oyster reef material through a 2 mm grid sieve. To quantify the densities of P. armatus and the dominant native prey species on the reef, I counted the number of individuals of the following prey species: P. armatus, Geukensia demissa

(mussel), Crassostrea virginica (oyster), and Eurypanopeus depressus (crab). These are the most conspicuous prey items on the reef and have been shown to be the primary prey items that make up the diet of important generalist predators like P. herbstii (Lee and

Kneib 1994). Because they often have a size refuge from predation, G. demissa and C. virginica were separated into two size classes. I considered G. demissa and C. virginica that were less than 20 mm to be considered prey items. I calculated the relative proportion of P. armatus in the representative prey base as the number of P. armatus over the total number of all prey items per 0.25 m2. In total, I quantified density plots at 25 reefs across the nine estuaries.

I examined the effect of latitude on the density of P. armatus and on the proportion of P. armatus within the prey base using simple linear regressions in the statistical program R (R 3.2.2 Development Core Team 2010). I averaged the density of

P. armatus and the proportion of P. armatus in the prey base across the reefs for each estuary. Both variables were checked for normality using the Shapiro Wilkes test. Both the density of P. armatus and the proportion of P. armatus in the prey base were normally distributed. To account for the differences in the number of replicate reefs sampled at each estuary, I used a weighted simple linear regression to test for a relationship with latitude between both density of P. armatus and proportion of P. armatus. In addition to examining a linear relationship, I also checked for quadratic relationships between latitude and the density and proportion of P. armatus. Neither term was significant and were removed from the final models.

2 Per capita predation risk on P. armatus along its invasive range

To determine whether the per capita risk of predation varies with latitude, I conducted tethering experiments at each of the invaded estuaries (n = 8) (Heck &

Thoman, 1981). Tethered crabs are considered at higher risk and can result in higher predation than natural circumstances (Zimmer-Faust et al. 1994). Thus, this experiment aimed to quantify the relative predation risk across the estuaries invaded by P. armatus.

Individual crabs (6-11 mm carapace width) were collected by hand from oyster reefs at the Skidaway Institute of Oceanography (SkIO) in Savannah, GA. I attached an individual crab to 35 cm length of 6.8 kg strength fishing line on the back of the carapace using super glue. Because tethers were many times the length of a crab, tethered crabs had much freedom of movement including to potentially take shelter among the oyster reef habitat to avoid predation. I then attached the line to a roofing nail to be used as an anchor point for installment in the reef. Tethered crabs were kept in a flow through system for a minimum of 24 hours before deployment to ensure tether integrity and that no mortality occurred due to tethering. Tethered crabs were transported to each deployment site in 5 gallon buckets filled with filtered and aerated sea water. At each reef I placed approximately 20 P. armatus 1 m apart for 12 hours during a nocturnal high tide. I made sure that tethers were underwater during the night hours, so tethers were placed between 18:00-23:00 and picked up no later than 09:00 the next morning. I was limited to tidal cycles when low tide was in the evening and then again in the morning.

Predation was determined by counting the number of crabs that were visibly preyed upon

(e.g. a piece of carapace remained on the tether). Any replicates that were observed with

8 a cut line (presumably due to abrasion from the oyster shells) were not included in the estimated per capita predation risk and excluded from all further analyses. In total I placed 371 tethers across 22 reefs within the 8 invaded estuaries.

I examined the effect of latitude on the predation risk of P. armatus using simple linear regressions in the R statistical program (R 3.2.2 Development Core Team 2010).

First, the predation risk of P. armatus was calculated as the number of P. armatus eaten divided by the number of P. armatus placed per reef. I then averaged the proportion of P. armatus eaten across the reefs for each of the 8 estuaries. Because the proportion of P. armatus eaten was not normal, I used a Spearman’s Rank Coefficient to explore the pattern across latitude.

3 Environmental factors that contribute to explaining risk of predation on P. armatus

To determine other environmental factors, regardless of latitude that contribute to the risk of predation of P. armatus, I quantified multiple biotic and physical variables at the sites where the tethering experiments were conducted (n = 8). The factors I explored were the density of large oysters, the density of P. armatus, the density of alternative native prey, the mean summer temperature and the abundance of native predators. I used large C. virginica (> 20mm) from the 0.25 m2 density plots (discussed in 1 Pattern of abundance of P. armatus along its invasive range) as an indicator of habitat quality in my analyses. While I also measured reef slope and height as potential measures of habitat quality, because C. virginica is a foundational species that creates habitat for many species (Gutierrez 2003, Coen et al. 2007, Byers 2015), I felt that density of large oysters was the most direct measure of habitat quality. Furthermore, the density of large oysters

9 have been previously shown to effect predator-prey interactions through creation of complex habitats for prey to hide and evade predators (Grabowski 2004). Based on optimal foraging theory, predators should consume a prey item when it is more abundant

(Pyke 1984). I included the density of P. armatus from the 0.25m2 plots density plots described above to better understand if native predators forage optimally. Because foraging by predators can be a function of alternative prey abundance (Pyke 1984), I included the density of native prey items from the 0.25m2 plots described above (1

Pattern of abundance of P. armatus along its invasive range). This included the number of small C. virginica (< 20 mm), small G. demissa (< 20mm), and E. depressus. Because temperature can affect metabolic rates of predators and thus consumption (Sanford 2002,

Ferrari et al. 2015), I used the average water temperature for July 1-15 given on the

National Oceanographic and Atmospheric Administration National Oceanographic Data

Center for the closest available site to the chosen estuaries.

Finally, I characterized and quantified the invertivore predator community at each of the 8 sites using a combination of trapping and plot sampling. Specifically, I set out 1 crab trap (standard size 61 cm x 61 cm x 28 cm) and 1 minnow trap (standard 42 cm x

22.9 cm with 2.54 cm openings) at each reef within the sampled estuary. Traps were set at low tide, 1m up from the edge of the reef (waterward) and retrieved approximately 6 hours later at high tide. Each of the minnow traps were baited with approximately 132g of frozen shrimp. Each crab trap was baited with approximately 1500g of frozen chicken.

I ensured that the shrimp and chicken were from the same companies to ensure that differences were not based on bait type. Additionally, the common mud crab predator,

Panopeus herbstii was quantified using the 0.25m2 plot excavation method described above. Fish species that were identified as potential predators were identified as having

“decapods” in their diets on fishbase.org.

To determine whether per capita predation risk could be explained by these other environmental variables, regardless of latitude, I conducted a generalized linear mixed effects model (GLMM) using the lme4 package in the statistical program R 3.2.2

(Douglas et al. 2015). I coded the dependent variable, predation risk, as binary data (link type = logit) with predation as the success (“1”) and survival as a failure (“0”). Individual tethered P. armatus were nested within “estuary” and included in the model as a random factor. I fitted the density of alternative native prey (sum of small G. demissa, small C. virginica, and E. depressus), the density of P. armatus, the mean summer temperature for each estuary, the number of large oysters and the abundance of native predators. The density of native prey, the density of P. armatus, the abundance of native predators and the number of large oysters (i.e. measurement of habitat quality) per reef were averaged across reefs within each estuary. The mean summer temperature is already taken at the estuary level. I used the package MuMin in R to run an exhaustive search of all possible models (with five candidate independent variables: 25 = 32 possible models) and determine the best model using AIC criteria including calculating the delta AIC and

Akaike weight (w) (Burnham & Anderson 2002, Symonds & Moussalli 2011). I then determined the relative variable importance (RVI) of each of the five independent variables by adding together the w for each of the models that include that variable

(Burnham & Anderson 2002, Symonds & Moussalli 2011).

4 Identifying predators responsible for P. armatus predation

To more directly determine predator species or size classes might be consuming

P. armatus, I designed a caging experiment to differentially exclude predators based on size and feeding ecology and quantified predation on tethered P. armatus. I conducted this experiment at the Skidaway Institute of Oceanography (SKIO) during 6 night tides from June 28 –30 and July 12-14, 2017. The Skidaway Institute of Oceanography was 1 of the 8 invaded sites previously surveyed and is known to have all the predatory species quantified in the trapping regime (McFarlin & Alber 2005). Individual P. armatus were tethered using the same methodology as described above, but instead super glued to

17mm length of 6.8 kg strength fishing line. I then attached the line to a roofing nail for installment in the reef substrate, placing one crab every meter. Cages were made out of wire shelving grids that were 35.5cm x 35.5 cm x 35.5 cm. I then either left the grid as is with 40 mm grid openings (“Small/Medium Crabs and Small Fish”) or I covered the entire cage with 17 mm (“Small Crabs and Small Fish”) or 6 mm (“No Predator

Control”) birding mesh. Mesh was attached using 9.1 kg strength fishing line and looped along the edges in a sewing method to ensure no predators could enter the cage along the seams. To exclude predators based on feeding ecology, I also constructed a “roof” treatment that was 35.5 cm x 35.5 cm x 5cm, leaving the sides of the cage free of mesh and open for predators to enter from the sides but not the top which was covered in 6 mm bird mesh (“All Crabs and Small Fish”). This would prevent large fish, such as Red

Drum which are known decapod predators in this system, from being able to predate the tethered crabs, but allow other benthic predators to enter from the sides. Finally, we kept

12 one treatment a cage-less control allowing all predators’ access including crabs and fish of all sizes (“All Predators”). I named the five treatments to represent which predators were able to enter the cage.

To examine the effects of predator size on consumption of P. armatus, each experimental block contained one replicate of each of the five treatments in a random order on the reef. Blocks were established in areas with similar mud substrate type including shell hash for refuge but was muddy enough to allow the cages to be pushed ~

2.5 cm into the substrate to ensure no predators could enter through the bottom. I kept blocks at least 10 m apart. I placed the tethers during an evening tide and left them for approximately 12 hours until the morning tide. I then removed all cages and quantified predation events during the morning low tide. New blocks were established every time cages were installed to ensure that placement of cages was truly random. In total I had 32 blocks and 139 tethered P. armatus. The treatment “Small Crabs and Small Fish” was only replicated over 11 spatial blocks and three nights. There were 6 tethered crabs that were eaten in the “No Predator Control” treatment out of the 32 placed. Four of these replicates were run during a storm event that moved the cages allowing predators to enter and two cages were compromised and found to contain small (< 20 mm) P. herbstii. I ran analyses with and without these compromised blocks and found that results did not differ

(Figure A.1).

Similar to the models described above, I conducted GLMM models using the lme4 package in R (Douglas et al. 2015) to test whether the predation risk of P. armatus differed under the five predator exclusion treatments: 1) “No Predator Control”, 2)

“Small Crabs and Small Fish”, 3) “Small/Medium Crabs and Small Fish”, 4) “All Crabs and Small Fish”, and 5) “All-Predators”. As described above, I coded the response variable, predation risk, as binary data (link type = logit) with predation as the success

(“1”) and survival as a failure (“0”). Block (e.g. established plots) was included as a random factor to account for any spatial variation in microhabitat where cages were installed. Predator exclusion treatment was fitted as a fixed effect. Since “treatment” is a categorical variable, model output gives how each group differed from the “reference” group. In order to determine significant differences between all the groups, I ran separate models changing the reference group for the model to each of the five treatments to determine all 10 pairwise comparisons. I then used the estimates of beta and the associated p values to determine post hoc difference between treatments were significant.

Since no predation occurred in the “No Predator Control” treatment with those compromised blocks removed, I only report on the six pairwise comparisons between the other four treatments.

RESULTS

1 Pattern of abundance of P. armatus along the US Atlantic Coast

The mean density of P. armatus across the nine estuaries ranged between 0 – 21 individuals per 0.25 m2. The minimum of 0 individuals per 0.25 m2 was observed in

Morehead City, NC as well as North Inlet, SC, the two northernmost sites. However at

North Inlet, I observed P. armatus individuals at the reefs, but outside of the measurement plots. Additionally, P. armatus has been reported to be established and often at low abundances in this estuary (Hartman & Stancyk 2001). The maximum

14 density of P. armatus was observed at the Skidaway Institute of Oceanography with 21 individuals per 0.25 m2. Results of the field survey indicated that there was no relationship between the density of P. armatus and latitude (R2 = 0.25, p = 0.166, n = 9,

Figure 1.2).

The density of P. armatus at our most southern site, St. Augustine, FL (an average of 4.5 individuals per 0.25 m2) was lower than expected relative to how long P. armatus has been established at this site. Because habitats in St. Augustine, FL begin to shift from oyster/saltmarsh dominated communities to mangrove systems, I further examined whether habitat quality (measured as number of large oysters) also influenced the density of P. armatus. Although St. Augustine, FL site is within the range of other sites for quantities of habitat quality and abiotic variables (Table 1.1 and Table A.1), I ran an

ANCOVA including latitude and density of large oysters to examine whether these were significant predictors of the density of P. armatus. Results indicate that number of large oysters was not a significant covariate (F = 1.439, df = 1, p = 0.275) and this did not affect the relationship between latitude and the density of P. armatus (F = 1.715, df = 1, p

= 0.238).

The mean proportion of P. armatus in the prey base across the 9 estuaries ranged between 0 - 42% of the prey base. The minimum of 0% was observed at the 2 most northern sites, Morehead City, NC and North Inlet, SC. The maximum of 42% was observed at St. Catherine’s Island, GA. There was no relationship between the proportion of P. armatus in the prey base and latitude (R2 = 0.11, p = 0.39, n = 9, Figure 1.3).

2 Per capita predation risk on P. armatus along its invasive range

Predation risk across the 8 invaded estuaries was high across all sites and ranged from 68.2 – 98.2%. The lowest predation risk (68.2%) on P. armatus was measured at

Sapelo Island, GA. The maximum predation risk (98.2%) measured was at Skidaway

Island, GA. Results of the tethering experiment indicated that there was no relationship between predation risk and latitude (Spearman’s Rank Correlation: S = 96, p = 0.752, rho

= -0.14, n = 8, Figure 1.4).

3 Environmental factors that contribute to explaining risk of predation on P. armatus

Using AIC criteria, the model that best predicted per capita predation risk on P. armatus included the density of large oysters, the density of P. armatus, and the total abundance of native predators (AICc = 234.69, df = 5, R2 = 0.26, Table 1.2). Specifically, predation on tethered P. armatus was negatively associated with the number of large oysters (i.e., a proxy for habitat quality) and the total abundance of native predators but was positively associated with the natural densities of P. armatus. Looking across all models, the factor with the highest RVI was the density of large oysters (RVI = 0.95) followed by the density of P. armatus (RVI = 0.81) and then total abundance of native predators (0.55). The mean summer temperature and the density of native prey had far lower RVI (0.30 and 0.29, respectively).

4 Identifying predators responsible for P. armatus predation

Results of the caging experiment indicated that there was an overall treatment effect on the per capita predation on tethered crabs (Figure 1.5). Results reported here are with the compromised blocks removed however analyses with the compromised blocks 16 included yielded the same results (Figure A.1). There was no significant difference between the “All Predators” and the “Small Crabs and Small Fish” treatment, however all other treatments had lower predation risk compared to the “All Predators”

(“Small/Medium Crabs and Small Fish”: p = 0.041 and “All Crabs and Small Fish”: p =

0.013). There was no significant difference between the “Small/Medium Crabs and Small

Fish” and the “All Crabs and Small Fish” treatments.

DISCUSSION

Although P. armatus is readily consumed by native predators, my results suggest that native predators are not providing biotic resistance against the northward expansion of P. armatus along the southeastern coast of the U.S. Per capita predation risk of tethered crabs was relatively high across all sites and did not systematically vary with latitude. Variation in per capita predation risk was best explained by differences in habitat quality and the relative abundance of alternative prey. My results are somewhat ambiguous as to which predatory species are consuming P. armatus with predator abundances having negative relationships with the per capita predation risk of P. armatus. These results suggest that there is no one predator driving P. armatus consumption, a conclusion that is further supported by the caging experiments that demonstrated predators in several size ranges are capable of consuming P. armatus to a substantial degree.

Predation risk did not vary with latitude and neither did the ambient density of P. armatus. Rather, density of P. armatus varied substantially across the sampled range from St. Augustine, FL to Morehead City, NC, but did not systematically vary with

17 latitude. However, the density of P. armatus dropped dramatically at the northern edge of its range, with no individuals found in Morehead City, NC and North Inlet, SC. This drop off in density at its northernmost edge of the distribution of P. armatus where predation risk and predator abundance do not decline provides additional evidence that the northward spread of P. armatus is limited not by biotic resistance, but likely another factor such as temperature (Canning-Clode et al. 2011).

Predation risk was relatively high across all invaded sites and no pattern of predation with latitude indicates that biotic resistance by native predators is not limiting the spread or abundance of P. armatus where it has been established. Although prey tethering studies likely overestimate per capita predation (Zimmer-Faust et al. 1994), my experimental design compares the relative difference in predation risk across this invasive range. I also designed methods that allowed P. armatus to still access and use refuge within the reef. Despite this, on average P. armatus was consumed 89% of the time. P. armatus may be able to overcome this high predation risk through particularly high recruitment. High recruitment has been previously suggested to nullify or counteract any biotic resistance via competition that is faced by P. armatus when it invades a new system (Hollebone & Hay, 2007b). As reviewed by Rilov and Crooks (2009), little is known about how propagule pressure may interact with the native community to structure marine invasions. However it has long been known that invasive species are often characterized by high fecundity (Sakai et al. 2001). Despite these high abundances, other studies have suggested that the functional response of predators might indicate predatory pressure can only reach a certain strength, allowing invaders to persist at high densities

(Twardochleb et al. 2012). Taken together, current evidence suggests that biotic factors may not be limiting the distribution or abundance of P. armatus and adds additional support that temperature and/or other abiotic factors regulate P. armatus. This suggest that as temperatures continue to warm with changing climates, P. armatus is likely to continue its spread into northern latitudes (Canning-Clode et al. 2011).

Although per capita predation risk was high across invaded sites, habitat quality

(measured as the number of large oysters) reduced predation risk and was the most important factor mediating predator-prey interactions between P. armatus and native predators. This negative relationship between the predation risk on P. armatus and the number of large oysters may indicate that higher quality habitat allows P. armatus to evade native predators. Within the same system, other studies have shown that increased oyster abundance increases the survival of native prey items (Grabowski 2004). Even in other systems, habitat characteristics, like sediment type, can affect the ability of native predators to consume a non-native prey (Byers 2002). However, oysters are a commercial species, and many reefs are under peril for reduced habitat quality given historical and continued harvest (Beck et al. 2011). This reduction in the number of large oysters on a reef could affect the predator prey interactions occurring between P. armatus and native predators.

Predation on tethered P. armatus was higher at sites that had higher densities of P. armatus suggesting that predation on P. armatus is density-dependent. Although P. armatus is a relatively new prey, if a predator is foraging optimally, then it should increase their consumption of P. armatus as it becomes more common in the environment 19

(Krebs & Davies 1981, Pyke 1984, Stephens & Krebs 1986). Additionally, other studies have demonstrated that consumption of non-native prey is often a function of the density of the invader (e.g. functional response) (Twardochleb et al. 2012, Charbonnier et al.

2014). For example, the consumption of invasive New Zealand mud snails by native signal crayfish was dependent on prey density, with a consumption rate indicative of a type 3 functional response. This pattern of predation by native predators may be due to the predator switching to consuming the non-native prey species when it reaches a highly abundant, more profitable state (Magoulick & Lewis 2002). This relationship also points out an important issue when thinking about the progression of an invasion. If native predators are consuming P. armatus less often when it is in lower densities, there could be a crucial reprieve from predation when P. armatus are first arriving at a site and in low abundances. This reprieve would allow P. armatus to grow in abundance and establish a population before predation pressure is able to exert any influence on the success of this establishment (Twardochleb et al. 2012). Then once reaching a highly abundant state, predators may only be able to exert a constant level of predation, regardless of prey density, allowing P. armatus to grow in abundance (Twardochleb et al. 2012).

Predation risk on tethered crabs both across latitude in the field and in the caging exclusion experiment together suggest that there is likely a suite of predators consuming

P. armatus in the natural environment and that trophic interactions may have a strong influence on predation. Interestingly, predation on P. armatus was negatively associated with the index of abundance native predators. Many of these species (P. herbstii, F. heteroclitus, and Callinectes sp.) have been shown to readily consume P. armatus in the 20 lab and are relatively common within the community (Hollebone & Hay 2008, Hostert

2014, Pintor & Byers 2015, Crosby & Pintor in prep). Thus, the negative relationship between the density of these predators and predation risk on the tethered P. armatus may be associated with predation risk by top predators in the system. The cage exclusion experiment had similar patterns with smaller predators providing equivalent predation pressure on P. armatus as the entire predatory community but only when larger predators were excluded. These smaller predators could be exerting similar predation pressure on

P. armatus as the entire suite of native predators because they experience some reprieve from their own enemies within the smaller mesh size treatment. The native predators we quantified have complicated interactions, with many acting as top and mesopredators within a system. Callinectes often acts as a top predator consuming P. herbstii and other mesopredators on the reef (Seed 1980, Kneib 1982). Even within a species, larger

Callinectes and P. herbstii are known to exhibit cannibalism of smaller individuals

(Perkins et al. 1996, see methodology: Pintor & Byers 2015). Fewer and smaller individuals (like juvenile P. herbstii or Callinectes sp.) being able to exert equivalent or even greater predatory pressure on this invasive species could be indicative of mesopredator release or reduced competition experienced at sites with fewer predators or the smaller mesh exclusion treatments (see risk-reduction caused by predator-predator interactions: Sih et al. 1998, Finke & Denno 2004, Lavender et al. 2014). Similarly, when predators are in competition with one another, they will often change their diet breadth or behavior to relieve some of this intraguild competition (Clark et al. 1999, Pintor & Byers

2015). This has been shown for the native predator P. herbstii within this system, as

21 smaller less competitive individuals’ more readily consumed P. armatus in the laboratory setting. Trophic interactions similar to these have been shown in other systems to effect the invasion success of non-native species (Taniguchi et al. 2002, Needles et al. 2015) and could be an important interaction preventing native predators from controlling the P. armatus invasion. Overall these results suggests that many size classes of individuals may be driving consumption of P. armatus and consumption by smaller individuals could be mediated by the presence of larger more competitive native predators in the system.

Finally, although I did not find evidence of biotic resistance from predation, the predatory community in salt marshes varies seasonally which could lead to different patterns of predation throughout the year (Dahlberg & Odum 1970, McErlean et al. 1973,

Hines et al. 1990). This study and analyses are representative of native predator-invasive prey interactions during the summer season, when predatory pressure is most likely the highest and P. armatus is at their highest abundances (Dahlberg & Odum 1970,

McErlean et al. 1973, Hines et al. 1990, Hollebone & Hay 2007a). This means the annualized rate of loss due to predation is likely lower than the rates measured in our experiment. While predatory pressure could be similar across this latitudinal range during the summer, there could be differences in timing of migration movements of predators during the cooler months that result in different predation patterns along the coast. If predation is lower in the cooler months, these seasonal differences could create a reprieve for P. armatus during the cooler seasons when predators aren’t around or have different foraging patterns. For example, blue catfish in Lake Dardanelle, Arkansas showed a distinct prey preference based on season, with individuals consuming more invasive 22 zebra mussels during the summer months and native shad during the winter (Magoulick

& Lewis, 2002). This shift in diet was likely due to changes in prey profitability within the system, with predators switching to the more abundant prey (Magoulick & Lewis,

2002). P. armatus is known to decrease in abundance during the winter months which could influence predation risk if native predators switch to a more abundant native prey item. Future research is needed to understand whether these shifts in diet could be occurring for native predators within the P. armatus invaded sites and how this might affect the invasion dynamics of P. armatus.

Although the density of P. armatus did not vary systematically with latitude, the dramatic decline in their abundance at the northern edge continues to suggest that there is some limiting factor preventing P. armatus from spreading northward. My results suggest that biotic resistance through predation is not likely limiting the abundance of P. armatus.

Predation was high across sites and did not vary with latitude, i.e. it was equally high at the northern edge of its invaded range. The experiments to tease apart influential predators suggest that predators of many sizes are important and able to sufficiently consume P. armatus. This likely explains the low variation in predation across this range despite differences in the predatory community where a complementarity of predators are able to consume P. armatus. Given this evidence, we should expect P. armatus to continue to spread northward as sea water temperatures rise (Canning-Clode et al. 2011).

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TABLES AND FIGURES

Figure 1.1 Map of the 9 estuaries surveyed in this study. The site in red, Morehead City, NC, served as an uninvaded, control site. The 8 invaded sites include North Inlet, SC; Charleston, SC; Ashepoo, Combahee and Edisto (ACE) Basin, SC; Skidaway Institute of Oceanography (SkIO), GA; St. Catherine’s Island, GA; Sapelo Island, GA; Jacksonville, FL; and St. Augustine, FL. Latitude lines are shown.

Table 1.1 Average values of each predictive variable included in the GLMM model examining which environmental factors affect predation risk on P. armatus regardless of latitude. The number of replicate reefs ranged from 2-3 per estuary based on availability and access. The mean density of large oysters was the number of C. virginica > 20mm in the 0.25m2 excavated plots. The mean density of P. armatus is the number of P. armatus in the 0.25 m2 excavated plots. The mean density of native prey is the total number of C. virginica < 20mm, G. demissa < 20mm, and E. depressus in the 0.25m2 excavated plots. Values reported here represent the average value for each variable across reefs within an estuary. The mean summer temperature was taken at the estuary level from the average water temperature in Celsius for July 1-15 given on National Oceanographic and Atmospheric Administration National Oceanographic Data Center for the closest available site to the chosen estuaries.

Mean Density Mean Mean of Density Density Mean Number Large of P. of Native Summer Estuary of Reefs Oysters armatus Prey Temperature North Inlet, SC 3 91.67 0.00 174.33 27.22 Charleston, SC 3 133 14.67 151.00 27.78 ACE Basin, SC 3 36 3.33 8.33 27.78 SkIO, GA 3 94.33 21.00 38.33 28.89 St. Catherine's Island, GA 3 47 10.67 14.33 28.89 Sapelo Island, GA 2 153 11.00 56.50 28.33 Jacksonville, FL 3 119.67 17.33 120.67 27.78 St. Augustine, FL 2 66 4.50 123.50 28.33

35 2 30

25 per 0.25mper 20

15 P.armatus 10

5 Number Number of 0 29 30 31 32 33 34 35 Latitude of Estuary

Figure 1.2 Mean density of P. armatus across the sampled range. Each dot represents the average number of P. armatus per 0.25m2 across the reefs within each estuary with the standard error bars. There was no relationship between the density of P. armatus and latitude (R2 = 0.25, p = 0.166, n = 9).

1 0.9 0.8

2 0.7

0.6 / number of number / prey 0.5

0.4 P.armatus items items per 0.25m 0.3 0.2 0.1

Number Number of 0 29 30 31 32 33 34 35 Latitude of Estuary

Figure 1.3 Mean proportion of P. armatus within the prey base across the sampled estuaries. Each dot represents the average number of P. armatus over the number of total common prey items (C. virginica, G. demissa, E. depressus, and P. armatus) per 0.25 m2 plot across each reef within the estuary. The standard error bars are shown for each estuary. There was no relationship between the proportion of P. armatus of the prey base and latitude (R2 = 0.11, p = 0.39, n = 9).

1 0.9 0.8

0.7 eaten eaten / 0.6

0.5 tethered perreef

P. armatus P. 0.4 0.3

Number 0.2 Number Number 0.1 0 29 29.5 30 30.5 31 31.5 32 32.5 33 33.5 34 Latitude of Estuary

Figure 1.4 The mean predation risk on P. armatus across the 8 estuaries invaded by P. armatus. Predation was quantified over a nocturnal high tide, with tethered crabs placed on evening low tide (approximately 6-11pm) and picked up on the morning low tide (no later than 9 am) to ensure that tethers were underwater the entire trial. Each dot represents the mean number of P. armatus eaten out of the number placed per reef across each estuary with standard error bars. The number of replicate tethers per estuaries from northern to southern sites were as follows: North Inlet, SC: n = 49, Charleston, SC: n = 44, ACE Basin, SC: n = 44, SkIO, GA: n = 47, St. Catherine's, GA: n = 50, Sapelo, GA: n = 47, Jacksonville, FL: n = 51, and St. Augustine, FL: n = 39. There was no relationship between predation risk and latitude (Spearman’s Rank Correlation: S = 96, p = 0.752, rho = -0.14, n = 8).

Table 1.2 Ranking of top models with a ΔAIC score less than five that predict the predation events on P. armatus using the possible variables of the density of native prey, the density of large oysters, the density of P. armatus, the abundance of native predators and the mean summer temperature. Each row relates for each model estimates for the intercept and β’s for independent variables that were included, as well as model summary statistics: degrees of freedom, the log likelihood, corrected AIC value, ΔAIC, and Akaike weight of each model. The relative variable importance (or RVI) was calculated for each variable by adding up the model weight for each model including that variable. The RVI for each variable are as follows: Density of Native Prey: 0.29, Density of Large Oysters: 0.95, Density of P. armatus: 0.81, Mean Summer Temperature: 0.32 and Total Abundance of Native Predators: 0.55.

Total Density Density of Mean Abundance of Native Large Density of Summer of Native Rank Intercept Prey Oysters P. armatus Temperature Predators df logLik AICc ΔAIC w 1 2.50 -0.92 0.58 -0.36 5 -112.27 234.69 0.00 0.23 2 2.45 -1.11 0.49 4 -113.61 235.33 0.63 0.16 3 2.47 0.16 -0.90 0.65 -0.42 6 -112.00 236.23 1.54 0.10 4 2.48 -0.94 0.65 -0.14 -0.35 6 -112.09 236.41 1.72 0.10 5 2.42 -1.13 0.58 -0.17 5 -113.34 236.85 2.15 0.08 6 2.44 0.05 -1.11 0.50 5 -113.58 237.32 2.63 0.06 7 2.43 -0.87 3 -115.80 237.66 2.97 0.05 8 2.47 0.18 -0.89 0.64 0.03 -0.42 7 -112.00 238.31 3.61 0.04 9 2.46 -0.24 -1.15 0.61 -0.39 6 -113.14 238.51 3.82 0.03 10 2.42 -0.71 -0.24 4 -115.45 239.00 4.31 0.03 11 2.43 -0.86 0.12 4 -115.70 239.50 4.81 0.02 12 2.44 -0.11 -0.85 4 -115.72 239.54 4.85 0.02

Figure 1.5 Proportion of P. armatus eaten by predators across different cage treatments. Treatments are named to reflect which predators were able to access the tethered crab placed in the cage. Because the graph shows the proportion of P. armatus eaten across the entire trial period, there are no error bars. The letters above each treatment indicate significant differences between treatments at α = 0.05 determined based on the model output of GLMM showing estimates for direction of difference between each treatment (β) and the significance of those differences (p). The number of replicate blocks was n = 26 for all treatments except “Small Crabs and Small Fish” treatment which had n = 5 replicates.

Chapter 2 The influence of relative prey abundance on the consumption of the non- native prey, invasive green porcelain crab (Petrolisthes armatus) by a native predator.

INTRODUCTION

Although many predators forage adaptively in response to changes in the abundance of their prey, predators may be slow to respond to the introduction of a non- native prey. As a result many invasive species reach high abundances by escaping the control of their natural enemies (Mitchell & Power 2003, Torchin et al. 2003, Shwartz et al. 2009). Native predators may not readily consume an invasive prey if the prey is not profitable. Low profitability could stem from if the prey is poisonous (Phillips & Shine

2004, Schlaepfer et al. 2005), exhibits effective anti-predatory responses or if the native predators does not have the foraging skills to consume a novel prey item (Cox & Lima

2006, Sih et al. 2010). Alternatively, or additionally, a predator may not consume a non- native prey in the early stages of the invasion process because its abundance relative to native prey is low. If the predator forages optimally, then it should ignore the rare prey and only increase its consumption of the non-native prey as it become more abundant relative to alternative native prey (Krebs & Davies 1981, Pyke 1984, Stephens & Krebs

1986). Indeed, many studies have shown that predators will switch to consuming an invasive prey over time as it becomes abundant (French & Bur, 1993, Molloy et al. 1997,

Magoulick & Lewis 2002, Carlsson et al. 2009) and can even benefit from this consumption (Pintor & Byers 2015a). However, many of the studies reviewed in Pintor &

Byers 2015a quantified the consumption of a non-native prey in the absence of native prey. In the few observational studies they review, when the native predator had access to 42 non-native and native prey simultaneously, predator abundance increased significantly relative to pre-invasion abundance. In other words, taking into account the relative proportion of the invasive prey to alternative native prey in the environment may reveal a more accurate representation of consumption of a native predator, whereby predators could still largely ignore less preferred invasive prey but still use them as supplemental resource. Here, I aimed to understand whether the consumption of a non-native, invasive prey by a native predator is a function its abundance relative to that of alternative native prey.

The adaptive foraging of a native predator on a novel prey might also be a function of the similarity of the non-native prey to a native prey within the invaded community. For instance, because native mussels (Mytlius edulis) had decades of prior exposure to the invasive predatory crab (Carcinus maenas), upon introduction of a second invasive predatory crab (Hemigrapsus sanguineus) the mussels were able to rapidly evolve the ability to reduce predation by adapting and changing shell morphology— a presumed co-option of their existing response to C. maenas (Freeman &

Byers 2006). Although this scenario is dealing with an invasive predator, the response of native predators to a novel non-native prey may also be dependent on the similarity between the invasive and native prey (Carlsson et al. 2009). For instance, predators may specialize on certain prey types and ignore other prey items even if they become abundant within a system (Estes et al. 2003). This preference could allow predators to learn skills that make them proficient in capturing and consuming certain prey types which could transfer to consumption of a similar novel prey (Ellis 1965, Hughes &

O’Brien 2001). Still, native predators could largely ignore an invasive prey if it is different from native prey types, especially preferred prey types, in the system. Ignoring these novel prey types, perhaps even if the invasive prey is very abundant within a system, and choosing to consume their familiar native prey could explain how native predators are inadequate at providing biological resistance to an invasive prey.

Petrolisthes armatus, the green porcelain crab, is a filter feeding crab that has invaded oyster reef communities along the southeastern coast of the US and is commonly observed to reach densities of several thousand per meter square (Hollebone & Hay,

2007). Proportionally, P. armatus can make up on average anywhere from 35-42% of the most common prey items within oyster reef communities in Georgia (see Chapter 1

Figure 1.3). Therefore, if native predators are foraging adaptively, then we would expect that native predators should consume P. armatus when they are in higher abundance relative to alternative native prey. Furthermore, P. armatus has a caloric value that is similar to or higher than that of many native prey items on the reef (Hostert 2014).

Therefore, within oyster reef communities where P. armatus has become abundant, it should be a profitable prey item and readily consumed.

The Atlantic Mud Crab, Panopeus herbstii, has been shown to readily consume P. armatus in the lab (Hollebone & Hay 2008, Hostert 2014, Pintor & Byers 2015b), but the persistent high abundances of P. armatus in reefs along the Georgia coast suggests that P. herbstii are not yet consuming P. armatus in proportions that reflect their value relative to alternative native prey. P. herbstii is a generalist predator on the reef consuming common native prey species including Geukensia demissa, the Ribbed Mussel,

Crassostrea virginica, the Eastern Oyster, and Eurypanopeus depressus, the Depressed

Mud Crab (Lee & Kneib 1994). I would expect the prey preference shown by P. herbstii to be a function of caloric value, search and handling time of a prey item. Previous work has shown that of the prey species that are common within these reef communities, P. armatus is equivalent in caloric value to native oysters (C. virginica) and significantly higher in caloric value than another native crab prey (E. depressus; Hostert 2014).

However, P. armatus has fewer calories per gram of tissue than native mussels (G. demissa). Although these prey vary in their caloric value per gram, P. armatus should still be profitable, especially when in high relative abundance. The mobility of P. armatus is most similar in mobility to the native crab E. depressus and quite different from sessile organisms like the native oysters (C. virginica) and mussels (G. demissa). These sessile organisms are largely dependent on shell thickness as a means to avoid predation while

P. armatus is mobile on the reef, often using its abdominal flap to swim backwards through the water column (similar to crayfish). Therefore, even if P. armatus is in high abundance, sessile filter feeders like oysters or mussels may be easier to catch and consume by P. herbstii compared to a mobile prey (e.g. P. armatus), despite any differences based on caloric content. So while we might expect that the consumption of

P. armatus will increase as the abundance of P. armatus increases in the environment, consumption is likely affected by the caloric content and mobility of the alternative native prey in the environment.

Here, I used the invasion of P. armatus into oyster reef communities of the south eastern US to examine whether the consumption of non-native prey and preference by P.

45 herbstii is a function of the relative abundance of alternative native prey and morphological similarity of the native to non-native prey. If consumption of P. armatus is a function of its relative abundance, then I predicted that consumption of P. armatus will increase when its abundance relative to an alternative native prey is higher. If consumption of P. armatus is a function of its similarity to the alternative native prey, then I predict that consumption would be greater on P. armatus when given the option of a more similar mobility/less caloric content prey item (E. depressus). Finally, I predicted that P. herbstii would consume native and invasive prey options in proportion to their availability in the environment, showing no preferences across different relative abundances of native to invasive prey. I had two main objectives: 1) quantify the consumption of P. armatus and preference shown for P. armatus when it was in low, equal or high abundance relative to native prey, and 2) test whether the consumption of

P. armatus when in varying relative abundances to a native prey differed depending upon whether the native prey was ecologically similar (E. depressus) or dissimilar (G. demissa) to P. armatus.

METHODOLOGY

I used a simultaneous prey choice assay to determine whether predation levels and preference on P. armatus by P. herbstii was dependent on the relative abundance of alternative native prey and the similarity of P. armatus to a native prey. I conducted these trials at the Skidaway Island Institute of Oceanography in Savannah, GA between June 1

– July 30, 2016 and June 20 - July 10, 2017. P. armatus has been established at this site

46 since 1994 and previous work has shown that P. herbstii from this site readily consume

P. armatus (Hollebone & Hay 2008, Pintor & Byers 2015b).

I used a 3 x 2 factorial design to examine whether predation on P. armatus by P. herbstii was dependent on the relative abundance of prey and similarity of the invader to native prey. Specifically, an individual P. herbstii predator was randomly assigned to an aquarium that contained a ratio of 1:3 (abundant P. armatus), 1:1 (equal P. armatus to native prey), or 3:1 (rare P. armatus) native: invasive prey items (total prey abundance across all treatments was equal to 16). Additionally, the alternative native prey offered was one that was ecologically similar (E. depressus) or a dissimilar (G. demissa). This resulted in a total of 6 treatment combinations (Figure 2.1). I aimed to test an equal number of males and female P. herbstii across each of the treatment combinations to examine whether sex affected consumption of P. armatus.

I hand collected individual P. herbstii from oyster reefs and housed them in flow- through seawater tanks, and feed them frozen shrimp ad libitum. Male and female P. herbstii were housed separately. For prey items I collected P. armatus (7-10 mm), E. depressus (7-12 mm) and G. demissa (20-30mm) from the same oyster reefs and housed them in separate flow-through seawater tanks. The size ranges were chosen based on previous work looking at the consumption of these prey species by P. herbstii (Hostert

2014). Before the start of the trial individual P. herbstii were placed in an isolated holding tank (2.3 liter aquaria filled with 0.6 liters of aerated seawater) for 24 hrs. To standardize hunger levels of the predators, food was withheld for 24-hours. After 24- hours, the isolated individuals were randomly assigned to one of the six treatment tanks

47 and allowed to acclimate to the testing tank for 8-hours (23.4 liter aquaria filled with 15.9 liters of aerated seawater). Fifteen minutes before the start of a trial, the P. herbstii predator was isolated from the rest of the tank using an inverted and weighted, opaque cup. Sixteen total prey items were then added to the tank opposite of the focal predator in ratios of a native prey species to the non-native species of either 3:1 (rare P. armatus),

1:1 (equal P. armatus to native prey), or 1:3 (abundant P. armatus). The prey species were allowed to acclimate for three minutes before the P. herbstii predator was released from isolation and the experiment began. Each individual tank was checked every 30 minutes over a 4-hour period during the evening (19:00-23:00) for a total of 8 nighttime observations to quantify the number of each prey species eaten. Each prey item consumed was replaced to maintain prey densities and thus the corresponding relative abundance treatment. Any P. herbstii that did not consume any prey items (native or invasive) during the trial were removed from any further statistical analyses.

Data Analysis

I measured the predation rate on P. armatus of 98 individual P. herbstii predators.

I had unequal replication across the six treatments due to some individual predators not eating during trials. Thus in the end I had between 15 - 17 replicates per treatment (Table

2.1). I used an ANOVA to test for an effect of the relative abundance treatments and similarity of the alternative native prey. Sex of the focal predator and the year of the trial was run were included as a covariates in the model. We also checked for interaction terms between all variables but removed any non-significant interaction terms. I used a

Tukey's honest significant difference (HSD) post hoc test with 95% confidence interval

48 to determine the direction and magnitude of difference between all levels within each factor. Data was checked for normality and if necessary, transformed by taking the natural log + 1. I used the R 3.2.2 statistical program for all analyses and generate figures

(R Development Core Team 2010).

I also used a Bayesian approach to evaluate whether P. herbstii consumed prey in proportion to their relative abundance in treatment tanks, but also to determine whether

P. herbstii demonstrated a preference for one of the two prey types. Although these

ANOVA analysis is useful in understanding whether the mean consumption of P. armatus by native predators differed across treatments, it does not inform us whether native predators are consuming P. armatus in proportion to what was placed in the tank.

If P. herbstii are consuming P. armatus in proportion to what is available in the tank, then I would consider them to be foraging as expected. I used a Bayesian approach to estimate the preference of P. herbstii for both native and P. armatus prey items across the different treatments. There are many advantages of Bayesian analyses, one of which is allowing us to view not only the population level of the parameter but individual variation within the data (Fordyce et al. 2011). Along with the estimate of the parameter value, the analyses yield a 95% credible margin of error for the preference of P. herbstii at both the individual and population level.

I used the “BayesPref” package in R (Gompart & Fordyce 2011) to run a

Bayesian analysis of the count data: the number of invasive prey eaten and the number of native prey eaten per individual (Fordyce et al. 2011, Haram et al. 2017 in review). I estimated posterior densities from 5000 MCMC steps following a burnin of 1000

49 generations. For each of the ratio treatments, there is a known expected level of consumption for both P. armatus and the native prey items. This is simply the ratio of native to invasive prey items within the tank, either abundant P. armatus (1:3), equal P. armatus to native prey (1:1), or rare P. armatus (3:1). In order to determine whether the populations were consuming P. armatus and native species as expected or not, I compared the expected value (either 0.75, 0.5 or 0.25 depending on the treatment) to the generated 95% bootstrap population variance. If the population variance included the expected value, I stated that consumption was as expected and P. herbstii is consuming optimally. If the population variance did not include the expected value, I stated that consumption was either higher or lower than expected and that P. herbstii is either showing a preference or avoidance of a prey item.

RESULTS

Results of the ANOVA indicated that there was significant effect of ratio treatment (F2, 92 = 5.61, p = 0.005, n = 98, Figure 2.2A), but no effect of prey similarity on the consumption of P. armatus (F1, 92 = 0.157, p = 0.69, n = 98, Figure 2.2B).

Specifically, P. herbstii ate significantly fewer P. armatus when P. armatus was in low or equal abundance to native prey in contrast to when P. armatus was in high relative abundance (p = 0.008 and p = 0.023). There was no difference between the treatments where P. armatus was in equal or in low abundance compared to native prey (p = 0.92).

There was no effect of sex on the number of P. armatus eaten (F1, 92 = 1.30, p = 0.069, n

= 98, Figure 2.2C). There was no effect of the year the trial was run (F1, 92 = 2.06, p =

0.15, n = 98, Figure 2.2D). Analyses with interaction terms between all variables yielded the same results (Table B.1).

Although males and females did not differ in their consumption of P. armatus at a significance level of p < 0.05 there was a trend towards females consuming more P. armatus. Based on this and previous work (Pintor & Byers 2015, Hostert 2014), I decided to conduct the Bayesian analyses separately for males and females. However, I combined across the prey similarity treatments, e.g. dropping the distinction between individuals that received E. depressus versus G. demissa as their alternative native prey, but report analyses ran separately on both prey types in Appendix B (Figure B.1 and B.2). Thus, below I report six sets of Bayesian analyses on predator consumption data: Female with abundant P. armatus, Female with equal native and P. armatus, Female with rare P. armatus, Male with abundant P. armatus, Male with equal native and P. armatus, and

Male with rare P. armatus.

Results of Bayesian analyses indicated P. herbstii did not always consume P. armatus at the expected preference (Table 2.3). When the abundance of P. armatus is high relative to native prey, female P. herbstii consume native prey more than expected and consume P. armatus less than expected (Figure 2.3A, n = 18). Male P. herbstii consume native prey more than expected and consume P. armatus less than expected

(Figure 2.3D, n = 15) when P. armatus is in high relative abundance. When P. armatus is in equal abundance to native prey, female P. herbstii consume native prey as expected and consume P. armatus as expected (Figure 2.3B, n = 16). In contrast, male P. herbstii consume native prey more than expected and consume P. armatus less than expected

(Figure 2.3E, n = 17) when both prey species are in equal relative abundance. Finally, when P. armatus is in low abundance relative to native prey, female P. herbstii consume native prey as expected and consume P. armatus as expected (Figure 2.3C, n = 15).

Similarly, male P. herbstii consume native prey as expected and consume P. armatus as expected (Figure 2.3F, n = 17) when P. armatus is in low relative abundance. Individual variation in preference increased (i.e. a larger percentage of individuals were showing a preference other than what was expected) in the higher relative abundance of P. armatus treatments (Table 2.3).

DISCUSSION

Despite differences in the relative abundance of P. armatus and the similarity of alternative native prey, the native predator P. herbstii always preferred native prey.

Following optimal foraging theory, I predicted that if consumption of P. armatus is a function of its relative abundance to an alternative native prey, then P. herbstii should increase its consumption of P. armatus when its abundance increases relative to a native prey. In addition, if consumption of P. armatus is a function of its similarity to the alternative native prey, then I predicted that consumption would be greater on P. armatus when given the option of an alternative prey item that was more similar in mobility and of lower caloric content (e.g. E. depressus). The ANOVA results suggest that P. herbstii does increase its consumption of P. armatus as the relative abundance of the invader increases. However the Bayesian results suggest that P. herbstii predators always preferred native prey. This preference for native prey was the same regardless of the caloric content or ecological similarity of the prey to the invasive P. armatus. Finally I

52 found some evidence that sex may be an important factor mediating the consumption of this invasive prey.

P. armatus may escape its natural enemies, in part, because native predators, like

P. herbstii, do not forage optimally on this non-native prey. Although P. herbstii did increase its consumption of P. armatus as the relative abundance of P. armatus increased across treatments, its consumption was significantly lower than expected relative to the abundance of the two alternative prey. This was true regardless of proportion of P. armatus in the tank, i.e. P. herbstii always showed a higher preference for native prey items than P. armatus even when the native prey was rare. This mismatch in the expected and observed consumption of prey by P. herbstii could have implications for the ability of native predators to provide some biotic resistance to P. armatus. For example, when invasive species first arrive in an ecosystem, they are typically in low abundance and rare compared to other native prey options. Overtime, as they increase in abundance, native predators should switch to consuming this newly abundant and profitable prey item

(Krebs & Davies 1981, Pyke 1984, Stephens & Krebs 1986). These results suggest that P. herbstii have not switched to consuming the invasive prey P. armatus. Another study looking at the rate of consumption across different densities of prey found similar results to mine suggesting native signal crayfish were inefficient at consuming the invasive New

Zealand mud snail (Twardochleb et al. 2012). Native signal crayfish exhibited a type 3 functional response, having little effect on mud snails when they were present in low densities, but had an increasingly negative effect when present in medium densities.

However, when the snail was present at high densities, native signal crayfish were only

53 able to consume mud snails at a certain rate, regardless of their increased abundances in the environment (Twardochleb et al. 2012). While previous work has shown that predators will switch to consuming an invasive prey when it is abundant (Molloy et al.,

1997, French & Bur, 1993, Magoulick & Lewis 2002, Carlsson et al. 2009, Charbonnier et al. 2014), this work shows that this increase in consumption may still be lower than expected and needed to provide biological resistance of an invasive prey species.

The ecological and caloric similarity of native prey to P. armatus did not influence consumption by P. herbstii which suggest invasive species may be so ecologically novel in an environment that native predators still prefer native prey. The preference P. herbstii showed for the native mussel G. demissa could be largely based on the fact that this calorically dense sessile organism is unable to exhibit behavior to evade predation like P. armatus might within the tank. However, the preference P. herbstii showed for the similar mobility native crab E. depressus despite its lower caloric content compared to P. armatus, might indicate that P. herbstii are experiencing some difference in handling time between the prey items. Therefore, the native crab E. depressus may not be a truly ecologically similar species to P. armatus. Native P. herbstii predators may have skills that enable them to efficiently catch and consume E. depressus that do not transfer to consumption of P. armatus. Overall, given the fact that P. herbstii always showed a preference for the native prey species suggests that there must be some difference in profitability based on whether the prey is native or invasive. Prey naïveté has been suggested and often shown to drive the large negative impacts that an invasive predator has on a native prey (Case & Diamond 1986, Cox & Lima 2006, Freeman & Byers 2006,

Sih et al. 2010). Here, I suggest that naïveté of native predators explains this preference for native prey as they lack an evolutionary history with P. armatus which may be too big a barrier to overcome in the short term. When native predators do not efficiently consume invasive prey based on a lack of co-evolutionary history, this can allow invasive species to escape predatory pressure and persist in native environments (Mitchell & Power 2003,

Torchin et al. 2003, Shwartz et al. 2009).

In our Bayesian analyses, females tended to consume P. armatus at the expected preference more often compared to males which only consumed P. armatus at the expected preference in the rare P. armatus treatment. This suggests that demographic traits affect whether an individual consumes or avoids a novel invasive prey. Within this system, other studies have observed this same pattern of females consuming more P. armatus than males (Hostert 2014, Pintor & Byers 2015b). Differences in diet breadth and consumption rates based on sex is not a new idea as often the different sexes have varied energetic demands which lead to different prey selections (Williams and

McBrayer 2011, Hierlihy et al. 2013). This difference may even be indicative of behavioral differences based on sex, with females being forced to consume prey items avoided by more dominant male individuals (Pintor & Byers 2015b). Depending on the sex ratio observed in the wild, consumption of P. armatus could be limited to fewer individuals than anticipated, with only female individual’s regularly consuming P. armatus, while males largely avoid them. This reduction in the number of native predators consuming P. armatus could help explain why P. armatus is able to efficiently avoid predatory pressure, and persist and spread to new oyster reefs.

Although P. herbstii did not forage optimally on P. armatus, there are other native predator species within these communities that will consume P. armatus (Callinectes similis and Fundulus heteroclitus: Hollebone & Hay 2007), thus, resistance may come as a consequence of multiple predator species. Biotic resistance is the study of the entire predatory community and its ability to resist the invasion of a novel species (Elton 1958) and it’s not always the case that native predators prefer native prey (Li et al. 2011,

Alexander et al. 2015). So while P. herbstii may not forage optimally on P. armatus, little is known about how other predatory species consumption changes with abundance of P. armatus. Preliminary results on the consumption of blue crabs (Callinectes) suggest that there is less individual variation in consumption of P. armatus, with most individuals’ regularly consuming P. armatus (Crosby & Pintor unpublished data). All of these species shown to consume P. armatus (P. herbstii, Callinectes sp., and F. heteroclitus) exist throughout the invaded range of P. armatus and the northern sites where P. armatus is predicted to expand. Understanding their foraging habits on this invasive species may give insight into how the native community of predators will deal with this species expanding northward into warming waters. In addition to this drawback,

I purposefully designed the experiment without habitat structure to simplify species interactions and for accuracy when checking the tanks for signs of predation. While this allowed us to more directly examine the influence of relative abundance and similarity of prey, the design is a simplification of the complex and intricate habitat created by oyster reefs. For instance, habitat quality and complexity is known to mediate predator-prey interactions between invasive and native species (Byers 2002) as well as in oyster reefs

(Grabowski 2004). Depending on how native and invasive prey items compete for habitat space to evade native predators, we could see differences in the consumption rates and preference of native predators for the different prey types under various habitat quality.

The native predator Panopeus herbstii did not forage optimally on the invasive prey species Petrolisthes armatus, and that predators consistently preferred native prey over P. armatus regardless of its relative abundance. Similarity of the native prey to the invader did not affect the consumption of P. armatus by P. herbstii, suggesting that co- evolutionary history of P. herbstii with its native prey may still strongly influence its likelihood of switching to consume more P. armatus when it’s abundant. Overall, these results suggest that the native predator P. herbstii is not likely to provide biotic resistance through predation on this invasive prey.

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TABLES AND FIGURES

Figure 2.1 A-F Description of the 6 different relative abundance treatments in the prey choice assays. A) 1:3 native to invasive ratio with E. depressus as alternative prey where P. armatus is relatively abundant, B) 1:1 native to invasive ratio with E. depressus as alternative prey where P. armatus is equal to native prey, C) 3:1 native to invasive ratio with E. depressus as alternative prey where P. armatus is relatively rare, D) 1:3 native to invasive ratio with G. demissa as alternative prey where P. armatus is relatively abundant E) 1:1 native to invasive ratio with G. demissa as alternative prey where P. armatus is equal to native prey, F) 3:1 native to invasive prey with G. demissa as alternative prey where P. armatus is relatively rare.

Table 2.1 Sample size of each treatment combination in 2016, 2017, and the total of the two years combined.

Treatment 2016 2017 Total E. depressus x rare P. armatus 10 7 17 E. depressus x equal P. armatus to native 10 6 16 E. depressus x abundant P. armatus 9 7 16 G. demissa x rare P. armatus 9 6 15 G. demissa x equal P. armatus to native 10 7 17 G. demissa x abundant P. armatus 10 7 17 Total 98

Table 2.2 ANOVA table showing each of the 4 factors. “Relative Abundance” is the relative abundance treatment with 3 levels of either rare P. armatus, equal P. armatus to native prey, or abundant P. armatus. “Native prey” is the alternative native prey species in the tank (E. depressus or G. demissa). “Sex” is the sex of P. herbstii in the tank (male or female). “Year” is the year the trial was run (2016 or 2017). Tukey HSD host poc testing was used to determine relative differences in the levels. “.” indicants a significance at α = 0.10, “*” is significant at α = 0.05, and “**” is significant at α = 0.01.

Factor df SS MSS F p Relative Abundance 2 4.31 2.15 5.61 0.005 ** Native Prey 1 0.06 0.06 0.16 0.693 Sex 1 1.30 1.30 3.39 0.069 . Year 1 0.79 0.79 2.06 0.155 Residuals 92 35.34 0.38

Figure 2.2 A-D Bar plots of the mean logarithmic total number of P. armatus eaten + 1 per individual given A) the relative abundance treatment, B) the alternative native prey species, C) the sex of the native predator and D) the year the trial was run. Standard error bars are shown for each bar. Letters above each group indicant significant differences at α = 0.05.

Figure 2.3 A-F Population-level preferences (solid lines) and individual-level preferences (dotted lines) of P. herbstii for both native and invasive prey items. The green line represents the native prey preference and the grey line represents P. armatus prey preference. The expected preference for native prey is shown by the green arrow and the expected preference for invasive prey is shown by the grey arrow. Each panel is representative of each of the 6 treatments: A) female P. herbstii with abundant P. armatus (n = 18), B) female P. herbstii with equal native to P. armatus (n = 16), C) female P. herbstii with rare P. armatus (n = 15), D) male P. herbstii with abundant P. armatus (n = 15), E) male P. herbstii with equal native to P. armatus (n = 17), F) male P. herbstii with rare P. armatus (n = 17).

Table 2.3 Estimates for population level preference shown for both native and non- native prey as well as percent of individuals showing expected preference for P. armatus across the six Bayesian analyses. The expected preference for P. armatus and Native Prey are the ratios within the tank. The % of individuals showing expected preference for P. armatus was calculated by counting the number of individuals showing a range of preference that included the expected value out of the total number of individuals in that treatment.

% of individuals 95% estimate 95% estimate showing Expected for Population Expected for Population expected Preference for Preference of Preference for Preference of preference for Treatment P. armatus P. armatus Native Prey Native Prey P. armatus Female with Abundant P. armatus 0.75 0.38-0.61 0.25 0.39-0.62 33% Female with Equal P. armatus to Native Prey 0.50 0.24-0.55 0.50 0.45-0.76 88% Female with Rare P. armatus 0.25 0.17-0.44 0.75 0.56-0.83 100% Male with Abundant P. armatus 0.75 0.30-0.55 0.25 0.45-0.70 20% Male with Equal P. armatus to Native Prey 0.50 0.11-0.32 0.50 0.68-0.89 35% Male with Rare P. armatus 0.25 0.10-0.34 0.75 0.66-0.90 82%

Chapter 3 Applications to the study and management of invasive species

I used the range expansion of the “Caribbean Creep” species Petrolisthes armatus to examine the ability of native predators to consume and control this invasive prey. My results suggest that biotic resistance through predation does not appear to limit the spread of P. armatus. Rather cold temperatures seem to be the dominate factor limiting the distribution of P. armatus (Canning-Clode et al. 2011). Therefore, as climates warm we may be likely to see this species expand into northern waters (Canning-Clode et al. 2011).

Additional experiments and analyses emphasized factors like prey abundance, habitat quality, and trophic interactions that might be influencing the ability of native predators to consume an invasive prey. These findings build upon previous studies and have direct applications to conservation management. My study also points out the lack of understanding of how these native predator-invasive prey interactions will change given inevitable changes in climate.

HOW ARE INVASIVE PREY ABLE TO PERSIST DESPITE BEING READILY CONSUMED BY NATIVE PREDATORS?

Examining the biotic resistance hypothesis not only helps us better understand whether native species can control invasive species within the food web, but can also help identify some potential mechanisms that are allowing invasive species to escape their natural enemies. This study showed that even when native predators regularly consume invasive P. armatus both in the field and in the lab, predation pressure might not be enough to resist invasion. These results beg the question, how is P. armatus able to persist despite being readily consumed by native predators? Here I outline 4 mechanisms

67 by which P. armatus might be able to evade predation and continue its expansion. First,

P. armatus may be so highly fecund that propagule pressure allows them to persist

(Hollebone & Hay, 2007). Second, native predators may prefer native prey regardless of relative abundances within the environment. Third, habitat structure and quality may influence predation rates and even preference for both native and invasive prey items.

Finally, trophic interactions between native species may influence the consumption of an invasive prey species.

Other studies that have examined consumption and control of P. armatus suggest that P. armatus is so highly fecund that propagule pressure is able to overcome any predatory pressure (Hollebone & Hay, 2007). While this study had no measure of movement of individuals across reefs to address this hypothesis, it’s commonly found that invasive species are highly fecund which could be the reason they are able to persist in new environments (Colautti et al. 2006). Experimental evidence of high fecundity overcoming high predation rates is lacking but a similar study found evidence that increased fecundity or propagule pressure can overcome any biotic resistance through competition (Clark & Johnston 2009). In combination with disturbance treatments, increased dosages of free-swimming larvae of the invasive bryozoan Watersipora subtorquata was able to overcome any competition provided by the native fouling community (Clark & Johnston 2009). Understanding how many larval or adult P. armatus are moving between reefs or even estuaries would help us understand the importance of propagule pressure in overcoming these high predation rates.

While native predators may increase predation levels in response to increased invasive prey abundance, preference for native prey items may still persist regardless of abundances when given the choice of an alternative native prey. While it is not always the case that native predators prefer native prey (Li et al. 2011, Alexander et al. 2015), many studies do not incorporate the relative density of prey items in the environment to help explain whether this is a true preference or predators responding to increased relative abundance of prey. Because the effect of invasive prey on native predators is influenced by whether the invasive prey was offered in addition to native prey options or in isolation (Pintor & Byers 2015), studies exploring the effect of density on predation on invasive prey that are performed with alternative prey available or in isolation should be distinguished. Both the lab and field studies described included native prey options and found similar results, with predators increasing predation as the abundance of P. armatus increases. However when I included the Bayesian results which considers consumption of both native and invasive prey, I determined that consumption on invasive prey was lower than expected, and that predators were preferentially consuming native prey. When invasive prey were given in isolation, native crayfish predators (Pacifastacus leniusculus) predation rates on the invasive New Zealand mud snails (Potamopyrgus antipodarum) changed in response to different densities indicating some potential for biotic resistance

(Twardochleb et al. 2012). However, depending on the options of native prey in the community, biotic resistance could be insignificant if these native crayfish respond in a similar way to the native predators in my study and choose native prey regardless of the relative abundances. Field studies looking at consumption of invasive prey with other

69 native prey options available found evidence of predator switching to an abundant non- native prey (Magoulick & Lewis 2002, Charbonnier et al. 2014). Theoretically since these studies were performed in the field, alternative prey should be available for native predators to choose between. However, it is hard to know whether this switching is a true preference for the abundant invasive prey because most studies don’t include the abundance of native prey options. For instance, invasive prey often displace native prey species, which could force native predators to consume these invasive prey out of necessity and lack of alternative native prey options in the environment (Magoulick &

Lewis 2002).

Depending on how native and invasive prey are interacting, including habitat structure could also change the predation rates and preference for prey items of native predators. For instance, I found that the native crab Panopeus herbstii always preferred native prey items in the lab experiment without habitat structure. However, if habitat structure was introduced to this experiment, we might expect P. armatus to more easily evade predation (based on the results from the field study), strengthening this preference for native prey. While native predatory species readily consume invasive species, this predation can restrict invasive to “refuge” habitats where they are able to evade predators and establish populations (Byers 2002, Dumont et al. 2011). In these cases, the “refuge” habitats were different from the preferred habitat of native prey items, which can reduce competition for good habitat between native and invasive species (Byers 2002, Dumont et al. 2011). But native prey in oyster reef communities also benefit from increased habitat complexity or reefs with more large oysters (Grabowski 2004). Therefore,

70 consumption of invasive versus native prey in oyster reef communities likely depends on whichever prey are able to find and hold onto quality habitat that allows them to evade predation.

Although we might expect and have seen that multiple native predators consuming an invasive prey would lead to stronger biotic resistance (Harvey et al. 2004), however competition between predators may negatively affect the ability of native communities to resist invasive species or structure invasion success. I found evidence that the number or size of native predators may influence consumption of P. armatus. Given that both top and mesopredators may be consuming P. armatus, intraguild predation between these different predators may affect the ability of smaller or lower trophic level predators to consume this invasive prey. While some information exists on trophic interactions effects on invasive species establishment (Taniguchi et al. 2002, Needles et al. 2015), the literature base looking at native predators interfering with each other in consumption of invasive prey is lacking. But intraguild competition for prey between predators affecting consumption has long been studied (Sih 1998), with many theories as to why multiple predators do not always lead to higher predation rates. For instance, predators may diversify their diet or consume alternative prey types (Pintor & Byers 2015) in order to reduce risk of competition between each other for food (Sih 1998). While these theories can be applied to invasion ecology, experimental evidence of this influence could help us predict how diversity of the native predatory community affects the biotic resistance through predation of an invasive prey.

CONSERVATION MANAGEMENT EFFECTS ON BIOTIC RESISTANCE THROUGH PREDATION

These results indicating that relative abundance of prey, habitat quality, and trophic interactions affect consumption of a novel prey suggest that biotic resistance through predation could also be affected by management practices that alter these factors. In conservation management, one of the primary goals is to restore native species that are in decline. This goal can be met through many different conservation efforts such as captive breeding and re-releasing individuals to the wild, eradicating invasive species and/or restoring habitat (Seddon et al. 2007, Benayas et al. 2009, Veitch et al. 2002) however there is also some debate as to whether these practices truly meet conservation goals

(Zavaleta et al. 2001, Miller & Hobbs 2007, Fraser 2008). In oyster reef conservation, management efforts focus on either reintroduction of live oysters (Ravit et al. 2012) and creating habitat by deploying manmade structure near existing reefs, which is expected to be colonized by native species (Beck et al. 2011, La Peyre et al. 2014). However, there has been evidence that these manmade reefs and natural reefs can even differ in the abundance of invasive P. armatus and also in native predator abundance (Dunnigan

2015). The factors I found to be important in my analyses could be affected by introduction of these manmade reefs and result in different predator-prey interactions between native and invasive. Studies like these can help us understand and predict what the predatory pressure on invasive prey will be like if ecosystems are restored under different management practices.

Because so many trophic interactions depend on it, restoring or creating habitat to meet management goals could have complicated effects on native predator-invasive prey 72 interactions. When management creates new or restores existing habitat for native species to colonize, invasive species may also colonize this new habitat. For example, in the

Great Lake communities where management uses loose stone or riprap to increase habitat complexity, one study found evidence that the invasive round goby had increased abundances in these created habitat compared to sandy substrate (Jude & DeBoe 1996).

This was a cause for concern that management efforts may actually be increasing undesirable invasive species and failing at targeting efforts on native species. These increased abundances could be due to round gobies being able to avoid predation in these new habitats (Jude & DeBoe 1996). In regards to the P. armatus invasion, artificial oyster habitat designed to benefit native species had lower abundances of P. armatus compared to historical reefs which is hypothesized to be driven by higher abundances of the native predatory crab P. herbstii on the artificial reefs (Dunnigan 2015). However, these artificial oyster habitats were sampled relatively soon after implementation

(approximately 0.5-2 years after construction) and could change in community dynamics over time, ultimately ending up like the historical reefs sampled in the study. If this is the case, these artificial oyster habitats could become more reefs dominated by P. armatus like the historical reefs surveyed. In fact, I anecdotally observed more P. armatus in artificially made reefs at the Skidaway Institute of Oceanography site compared to the surrounding natural reefs (personal observation). Especially for any restoration actions that occur in northern waters, management should consider how trophic interactions will change if P. armatus dominate any artificial reefs created in the path of their predicted range expansion.

Current management doesn’t always take into account the complex interactions between invasive and native species and often methods are not used in conjunction despite their proven interactive effects (Didham, et al. 2007, Miller & Hobbs 2007,

Brudvig 2011). These results suggest predation is mediated by factors that could be altered by some conservation efforts. Therefore, management should consider a plan of restoration that uses a combination of efforts like captive reintroduction individuals to the wild or restoring/creating habitat in combination with eradicating invasive species if they are present. This way any negative impacts or indirect effects on native species may be minimized to achieve conservation goals. Understanding the complex interactions between invasive and native species will only further our understanding of how to effectively manage and maintain our native environments.

BIOTIC RESISTANCE THROUGH PREDATION IN THE FACE OF CLIMATE CHANGE

As climates continue to change, both native and invasive species will have to adapt to any abiotic changes to the environment, which could also affect the trophic interactions between them (Hellmann et al. 2008, Gilman et al. 2010, Hulme 2017). To reiterate, I found evidence that native predators in oyster reef communities are inadequate at preventing the spread of the invasive species Petrolisthes armatus and that as climates change this “Caribbean Creep” species is likely to spread into warming northern waters

(Canning-Clode et al. 2011). While we can somewhat predict how P. armatus will respond to these increased temperatures as we have already observed it’s successful range expansion in response to warming waters (Canning-Clode et al. 2011), how the native predators will respond is somewhat more ambiguous (Kennedy 1990, Hines et al. 2010, 74

Dodd et al. 2015). We have no experimental evidence for how trophic interactions between these native predators and invasive prey will change with the climate.

There are few studies that experimentally test predator-prey interactions between native and invasive species in the face of climate change (Cheng et al. 2017) and instead rely on comparisons between similar native and invasive species to predict how populations will respond (Chown et al. 2007, Sorte et al. 2013). For invasive predators and native prey adapting to climate change, there is evidence that predators were less tolerant of changes in the environment than their prey (see trophic sensitivity hypothesis

Cheng et al. 2017). If this scenario also applies to native predators and invasive prey, native predators may decrease predatory pressure on invasive prey thereby relieving any biotic resistance they were providing (Hellmann et al. 2008). This is a cause for concern for management using biotic resistance through predation to control invasive species

(reviewed in Hellmann et al. 2008). However, there are so few experimental studies on native predator-invasive prey interactions, that predicting the net effect of invasive prey in the face of climate change is impossible without some experimental evidence to ground predictions.

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