University of Florida Thesis Or Dissertation

DOES CULTCH TYPE AND RESTORATION HISTORY INFLUENCE DECAPOD CRUSTACEAN COLONIZATION OF OYSTER REEFS?

MEGAN SCHERRER LAMB

A THESIS PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF SCIENCE

UNIVERSITY OF FLORIDA

2018

To the oysters

ACKNOWLEDGMENTS

I would like to thank my advisor, Dr. Behringer, for taking me on as a distance student and providing me with counsel and advice. Thank you to my committee members Dr. Andy Kane and Dr. Shirley Baker who provided valuable input and support. I would like to thank my lab mates for welcoming me to the lab and their willingness to help me, even though they saw me more over a computer screen than in person. Thank you to the Apalachicola National Estuarine Research Reserve which has provided support, water quality data, and equipment use to allow me to complete this project, especially J. Garwood and C. Snyder for technical assistance. Thank you to C. Jones and J. Shields at the Florida Department of Agriculture and Consumer

Services for answering questions and providing fossilized material used in collectors.

Additional shell material was provided by T. Ward and Paddy’s Raw Bar. I would like to thank D. Armentrout, T. Griffith, M. K. Davis, M. Davis, C. Snyder, E. Bourque, M.

Christopher, K. Peter, H. Heinke-Green, S. Simpson, and W. Annis for assistance with field sample collection. Thanks to J. Collee from UF-IFAS Consulting provided assistance with statistical methods.

I would like to thank my parents for all the support, encouragement, and education they have provided me with over my entire lifetime. Thank you to my sister, brother-in-law, and nieces who have also given me much encouragement, advice, and sent care packages. My friends have been loyal and understanding during this journey and I cannot express my gratitude enough to them. Finally, I could not have made it through graduate school without my husband Dave. I cannot thank you enough for all you have done to support our family while I have gone back to school, so I’ll just say thank you and I love you.

TABLE OF CONTENTS

page

ACKNOWLEDGMENTS ...... 4

LIST OF TABLES ...... 7

LIST OF FIGURES ...... 8

LIST OF ABBREVIATIONS ...... 9

ABSTRACT ...... 10

CHAPTER

1 INTRODUCTION ...... 12

Need for Better Understanding of Restoration Outcomes ...... 12 Oysters: Global Bivalves ...... 12 Oysters and Ecosystem Services ...... 14 Degradation of Reefs ...... 17 Interest in Restoration ...... 19 Oyster Reefs in the Apalachicola Bay: Fishery, Management, and Collapse ...... 21 Gaps in Knowledge ...... 22

2 PATTERNS OF COLONIZATION OF DECAPOD CRUSTACEANS TO NATURAL AND NEWLY RESTORED OYSTER REEFS IN APALACHICOLA BAY, FLORIDA ...... 26

Introduction: Ecosystem Services Provided by Oyster Reefs and Interest in Restoration Outcomes ...... 26 Methods ...... 29 Study Site Selection ...... 29 Decapod Crustacean Collection, Preservation, and Identification ...... 30 External Parasite Load of Xanthid Crabs ...... 31 Statistical Analyses ...... 31 Water Quality ...... 32 Results ...... 32 Decapod Crustacean Species Abundance ...... 32 Decapod Crustacean Species Richness ...... 33 Decapod Crustacean Species Biodiversity ...... 34 Size Structure of Most Abundant Crustacean Species ...... 34 Gender and Reproductive Status of Most Abundant Crustacean Species ...... 34 External Parasite Prevalence in Select Xanthid Crabs ...... 35 Water Quality Monitoring ...... 36 Discussion ...... 37

Abundance, Richness, and Biodiversity of Restored and Historic Oyster Reef Communities ...... 37 Restored Oyster Reef Decapod Crustacean Reproduction and Colonization .. 38 Influence of Nearby Communities Oyster Reef Colonization ...... 39 Parasite Prevalence amongst Xanthid Crabs ...... 41 Limitations of Present Methods and Future Directions ...... 42

3 CONCLUSIONS ...... 65

APPENDIX: SUPPLEMENTARY MATERIALS ...... 68

LIST OF REFERENCES ...... 72

BIOGRAPHICAL SKETCH ...... 81

LIST OF TABLES

Table page

2-1 Collector deployment and retrieval dates...... 60

2-2 Results for site age, substrate, and interactive effects between the two on species abundance, richness, and biodiversity by sampling period...... 61

2-3 Percent ovigerous E. depressus females by site and sampling period...... 62

2-4 Percent ovigerous P. armatus females by site and sampling period...... 62

2-5 Results for site age, substrate, and interactive effects between the two on parasite prevalence in E. depressus and P. herbstii by sampling period...... 63

2-6 Water quality parameter mean, variance, and standard deviation measured at all sites at each retrieval/deployment date...... 64

A-1 Complete list of decapod crustaceans found in collectors deployed in Apalachicola Bay sites over entire study period...... 68

A-2 Species abundance. The total abundance of all decapod crustaceans found in each collector during each sampling period...... 69

A-3 Species richness. The total number of individual species found in each collector by site during each sampling period...... 70

A-4 Shannon-Weiner Biodiversity Index Calculations by site, collector number, and shell type throughout the sampling period...... 71

LIST OF FIGURES

Figure page

1-1 The Apalachicola Bay. The Apalachicola Bay is in the Florida panhandle at the foot of the Apalachicola-Chattahoochee-Flint watershed...... 24

1-2 Oyster bar habitat in the Apalachicola Bay...... 25

2-1 Research sites. Map includes Apalachicola Bay oyster bars, research sites, and ANERR water quality stations...... 44

2-2 Green shell (left) compared to fossilized (right) oyster shell material ...... 45

2-3 Green shell collector (left) and fossilized shell collector (right) before deployment ...... 46

2-4 Collectors tied together prior to deployment...... 47

2-5 Collectors stored in individual bins after retrieval to keep organisms with each collector...... 48

2-6 Subadult P. herbstii with mature L. panopaei infection...... 49

2-7 Decapod crustacean abundance at sites by substrate type...... 50

2-8 Decapod crustacean species richness at sites by substrate type ...... 51

2-9 Decapod crustacean biodiversity, calculated by the Shannon-Weiner biodiversity index, at sites by substrate type ...... 52

2-10 E. depressus sizes during all sampling periods by site ...... 53

2-11 P. armatus sizes during all sampling periods by site...... 54

2-12 D. texanus sizes during all sampling periods by site ...... 55

2-13 M. mercenaria sizes during all sampling periods by site ...... 56

2-14 Percent E. depressus parasitized by L. panopaei by site and sampling period .. 57

2-15 Percent P. herbstii parasitized by L. panopaei by site and sampling period ...... 58

2-16 Salinity (psu) and temperature (°C) readings from Cat Point during the study period ...... 59

LIST OF ABBREVIATIONS

ACF Apalachicola Chattahoochee Flint

ANERR Apalachicola National Estuarine Research Reserve

ANOVA Analysis of Variance

C Celsius

CW Carapace width

DO Dissolved Oxygen

FDACS Florida Department of Agriculture and Consumer Services

FDEP Florida Department of Environmental Protection

FWC Florida Fish and Wildlife Conservation Commission

GOM Gulf of Mexico

NFWF National Fish and Wildlife Foundation

NOAA National Oceanic and Atmospheric Administration

NRDA Natural Resource Damage Assessment pH Potential hydrogen psu Practical salinity unit

UF University of Florida

Abstract of Thesis Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Master of Science

DOES CULTCH TYPE AND RESTORATION HISTORY INFLUENCE DECAPOD CRUSTACEAN COLONIZATION OF OYSTER REEFS?

Megan Scherrer Lamb

May 2018

Chair: Donald C. Behringer Major: Fisheries and Aquatic Sciences

Oyster reefs have historically been a significant ecological and economic component in temperate estuaries throughout the world. In addition to food and financial benefits from fisheries, in recent decades there has been an appreciation of these estuarine features for the unique ecosystems they support. The ecosystem services provided by oyster reefs include water quality improvement, contribution to nitrogen cycles, and shoreline stabilization. Oyster reefs have degraded due to overharvest, pollution, and disease. Up to 85% of reefs have been lost globally, and there is increased interest in restoration to recapture the fisheries, ecosystem services, and intrinsic ecological worth of these habitats. Restoration efforts largely concentrate on the addition of hard substrates to reef areas to increase habitat available for oyster settlement. To determine if restoration projects are considered successful, monitoring of restored sites must be conducted and compared to historic, non-degraded areas.

Little information is available about the success of restoration projects or viability of the techniques employed. Understanding how a restored site functions on an ecosystem level can contribute knowledge to future restoration efforts and inform how restoration efforts may be used most efficiently.

This study examined decapod crustacean richness, abundance, and biodiversity on newly created and historic oyster bars in the Apalachicola Bay, Florida. Wire mesh collectors containing fresh oyster shell (n = 3) and fossilized material (n = 3) (the latter being utilized at the restored areas), were deployed at two restored and two control sites. Population structure, reproductive status, and external parasite prevalence in select xanthid crabs were also examined. Results showed that during the majority of sampling periods, there was no statistically significant difference in richness, abundance, or biodiversity between sites of different age, different substrates, or interactive effects between the two. Ovigerous females of multiple species of xanthid crabs and shrimp were observed at all sites, and juveniles of many species of decapod crustaceans recruited to all sites during the year. These results indicate restored sites were quickly colonized by decapod crustaceans, which is an important component in restoring ecosystem function to degraded habitats.

CHAPTER 1 INTRODUCTION

Need for Better Understanding of Restoration Outcomes

Oysters are a significant habitat component of coasts around the world. They create biogenic reef structure that supports robust and unique biological communities and perform valuable ecosystem services. Degradation and poor health of reef structures has led to a burgeoning restoration industry, but restoration outcomes are not well understood. Restoration outcomes must be quantified, measured, and reported to provide guidance towards future project effectiveness and efficient use of funds.

This chapter provides background information on oyster biology, ecosystem services, and an overview of existing publications about the richness, biodiversity, and abundance of organisms living on oyster reefs in comparison to nearby habitats.

Following this ecological introduction, the causes of the degradation of oyster reefs, the restoration industry, and a case study of oyster fishery collapse in Apalachicola Bay,

Florida are reviewed. This background material informs the issue of how and why restoration outcomes need to be better monitored and reported. This information is needed to address gaps in understanding so that restoration efforts and funding can be utilized efficiently and economically.

Oysters: Global Bivalves

Oysters are a dominant group of bivalves that inhabit temperate estuaries around the world (Beck et al. 2009). True oysters belong to the order Ostreidae which consists of around one hundred and fifty species. The three largest genera within this order are

Crassostrea, Ostrea, and Saccostrea (Carriker & Gaffney 1996). Here, I focus on the

American or eastern oyster Crassostrea virginica, the dominant native oyster on the

12 western Atlantic coast, ranging from the Gulf of St. Lawrence, Canada, south to the

Brazilian coast (Buroker 1983).

Oysters can grow in a variety of different shapes and there are multiple terms for these groupings. Oysters grow together in reefs, bars, or beds, terms which are used interchangeably, and can grow horizontally and vertically, creating complex three- dimensional structures. They sometimes grow in groups called aggregates which are clumps of multiple oysters (Beck et al. 2009). Oyster reefs create habitat for future oyster generations, other organisms, and are considered ecosystem engineers (Jones et al. 1994, Grabowski & Peterson 2007).

To understand how oysters create biogenic reef formations, it is important to understand their life cycle. Crassostrea virginica are protandric broadcast spawners and fertilization occurs externally in the water column (Thompson et al. 1996).

Following a planktonic stage that can last days to weeks, pediveliger stage oyster larvae use chemical and auditory settlement cues to find appropriate, hard substrate on which to settle (Pawlik 1992, Tamburri et al. 1992, Lillis et al. 2013). They then metamorphose into oyster spat which will grow into adult oysters if environmental conditions are favorable (Kennedy 1996, Wallace 2001).

Crassostrea virginica are well adapted to survive in the temperate estuaries they inhabit, which experience wide ranging temperatures, salinities, oxygen levels, and water levels. Oysters are eurythermal and found in waters where yearly temperature can range from -2°C to 36°C (Shumway 1996). They are euryhaline, with an optimum salinity range of 14-28 psu. Though oysters can tolerate exposure to lower and higher salinities for some period of time, prolonged exposure may impact growth rates and other metabolic functions (Galtsoff 1964, Shumway 1996). Reproduction is temperature

13 dependent, and spawning of northern oysters occurs between 15-20°C while it occurs in southern groups at 20°C (Wallace 2001). Adult C. virginica can tolerate hypoxic and anoxic conditions by closing their valves, decreasing oxygen consumption and their metabolic rate (Shumway & Koehn 1982, Shumway 1996, Lombardi et al. 2013, Porter

& Breitburg 2016). However, hypoxic and anoxic conditions are very detrimental to larval survival, juvenile settlement, and recruitment (Baker & Mann 1992). Crassostrea virginica are able to regulate their metabolism under a wide variety of temperature and salinity combinations (Shumway & Koehn 1982). Larval and adult oysters are filter feeders, consuming particulate matter suspended in the water column (Newell &

Langdon 1996). These adaptive characteristics historically allowed oysters to thrive in high numbers, creating large reefs that served as significant habitat in estuaries around the globe.

Oysters and Ecosystem Services

Oyster bars provide a suite of ecosystem services associated with their ability to filter large volumes of water, buffer nearby land areas from wave energy, and provide habitat. Oysters are filter feeders and ingest a wide range of particles from the water column including living microorganisms, detritus, and inorganic particles (Newell &

Langdon 1996). Multiple studies have found less chlorophyll-a, an indicator of algal or phytoplankton abundance, in the water column associated with oyster beds (Dame et al.

1984, Cressman et al. 2003, Nelson et al. 2004), which may reduce the chance of a harmful algal bloom occurrence (Newell 2004). Oyster filtration can remove fecal coliform particles, and inorganic particles such as nitrogen, phosphorus, and carbon from the water column (Dame et al. 1989, Cressman et al. 2003). After these particles pass through oysters, digested material (feces) and undigested material (pseudofeces)

14 are deposited onto surrounding benthic surfaces (Newell & Jordan 1983). This process helps prevent eutrophication in estuarine systems by reducing high concentrations of ions and aids in benthic-pelagic coupling of nutrients (Coen et al. 2007, Piehler & Smyth

2011, Wall et al. 2011). Removal of suspended solids from the water column also improves water clarity, aiding seagrass growth (Newell 2004, Newell & Koch 2004).

Improvement in water quality may also boost the economy through activities such as tourism, fishing, and recreation (Lipton 2004).

Oyster reefs buffer coasts from wave action created by storm events, boat wakes, and more. The hard substrate reduces wave energy reaching nearby land areas, and it has been found that placing oyster material nearby can increase sediment accretion in marsh areas (Meyer et al. 1997, Piazza et al. 2005, Borsje et al. 2011,

Gedan et al. 2011). Cultch material (crushed, hard material used in restoration) has been found to significantly reduce shoreline retreat in low energy environments, though effectiveness may be limited in high energy environments (Piazza et al. 2005).

Oysters create large areas of complex, hard bottom habitat that supports a diverse and extensive community. A multitude of animals including fish, shrimp, crabs, molluscs, bivalves, polychaetes, birds, and many more rely on oyster bars for the food and refuge they provide. Reefs create hard substrate which provides a place of attachment in estuaries which may otherwise contain mostly soft sediment or vegetation. Many studies have looked at the diversity and abundance of communities on oyster bars compared to other estuarine habitats, though it is difficult to compare results because of differences in the characteristics of the oyster bars used (fished or unfished, inter- versus subtidal, natural or enhanced reefs, difference in sampling techniques, or lack of specification of these details). Despite differences, a review of findings shows nearly universal support that all oyster reef types support diverse and 15 abundant communities. Glancy et al. (2003) found that oyster habitat supported greater decapod biodiversity than marsh edges year-round and greater than seagrass habitats at certain times of year. They also considered decapod biomass with results varying with time. Some years biomass was similar to both marsh edge and seagrass while other years it was only similar to seagrass (Glancy et al. 2003). Gain et al. (2017) found intertidal oyster reefs supported higher macrofaunal densities and distinct communities when compared to neighboring seagrass and marsh edge habitats. Conversely Nevins et al. (2014) found that fish and crustacean densities were significantly higher in marsh edge than unfished oyster reefs, but the community supported by the oyster habitat was unique.

Several studies have shown oyster habitats support more abundant and diverse nektonic communities when compared to non-vegetated bottom habitats (Shervette &

Gelwick 2008, Stunz et al. 2010). Kingsley-Smith et al. (2012) found species abundance higher on oyster reefs, while diversity measures were higher on non- vegetated bottom habitats. The authors attributed this result to biodiversity calculations reflecting the overall much higher abundance of organisms on the oyster habitat; because oyster reefs had much higher abundance overall, diversity calculations were lower, conversely there was a low abundance of organisms found on non-vegetated habitats but high diversity amongst that smaller abundance. Beck (2012) compared communities on fished and unfished reefs and found similar species and species density on both types, though fished reefs had higher diversity.

Despite differences in methodology and types of habitat compared, the overall indication is that new and old oyster habitat support unique and abundant communities, and loss of oyster habitat leads to lower biodiversity (Lotze et al. 2006, Airoldi & Beck

2007). While this habitat has intrinsic value, many species associated with oyster reefs 16 also have significant economic value directly as fisheries and indirectly through tourism and recreation (Lipton 2004, Grabowski & Peterson 2007).

Degradation of Reefs

It is estimated that 85% of oyster reefs globally have been lost compared to historic abundance, and oysters are functionally extinct (classified as 99% loss of reefs) in 37% of estuaries and 28% of ecoregions (Beck et al. 2009, Beck et al. 2011). At present, most of the world’s wild harvest of native oysters occurs in just five ecoregions in North America, with the Gulf of Mexico (GOM) region yielding a harvest as large as the other four regions combined. Despite this distinction, GOM reefs are also threatened; 50-89% of Gulf reefs disappearing in the last 130 years (Beck et al. 2011).

Despite being recognized for their biogenic importance, legal protections that have been given to habitats with similar functions such as coral reefs and kelp forests have not been given to oyster habitat. This may be because their true value has not been recognized, or because of the value associated with oyster fisheries (Lenihan &

Peterson 1998).

Loss of oyster reefs is due to overharvest, disease, poor water quality, pollution, sedimentation, changes in environmental conditions, poor recruitment, non-native species, and loss of reef stability, which hampers recovery ability (MacKenzie Jr. 1996,

Kirby 2004, Beck et al. 2011, Seavey et al. 2011). Non-natives can introduce pathogens, prey on oysters and other estuarine organisms, potentially increasing mortality rates and altering food webs (Beck et al. 2011, Poirier et al. 2017). Though growth and mortality vary between size classes and region, the optimal temperature and salinity range for all sizes of wild Crassostrea virginica in the northern GOM are 20.0-

26.3°C and 10.7-16.1 psu (Lowe et al. 2017). Exposure to salinities outside of this range caused by environmental changes or other alterations in freshwater flows into 17 estuaries could cause reduced recruitment and increased mortality (Heilmayer et al.

2008, Petes et al. 2012, Rybovich et al. 2016, Lowe et al. 2017). Certain impacts may work synergistically to increase oyster mortality. Oysters are more likely to suffer mortality from naturally occurring diseases such as “Dermo” (caused by the protozoan parasite Perkinsus marinus) or carry the pathogen Vibrio vulnificus in high salinity, high temperature conditions such as those occurring in the summer months which coincide with seasons of high temperatures and salinities (Motes et al. 1998, Petes et al. 2012).

Once oyster reefs have begun to deteriorate, regardless of the reason, there is less hard substrate available for new oysters and reef stability is compromised (Beck et al.

2011).

Predation can have significant detrimental effects to oyster populations. Multiple predator species express preference for oysters of a certain size, for example, the stone crab Menippe mercenaria disproportionately consume medium sized oysters (25-70 mm) over other oyster size classes (Rindone & Eggleston 2011), and the oyster drill

Stramonita haemastoma preferentially consumes medium oysters (50-75 mm) over smaller and larger size classes (Pusack et al. 2018). When given a range of oyster spat between 1-29 mm, the flatback mud crab Eurypanopeus depressus and estuarine mud crab Rhithropanopeus harrisii exhibit preference for spat <8 mm in size (Kulp et al.

2011). The non-native green crab Carcinus maenas preferentially preyed on oysters up to 40 mm in size (Pickering et al. 2017, Poirier et al. 2017) and may contribute to overall increase in oyster predation in native beds (Poirier et al. 2017). Predation levels are dependent on both predator and prey size (Pickering et al. 2017) so population size structure should be taken into account when inferring impacts. An oyster population that is rebuilding may have smaller individuals and be subject to a wider range of predators than populations with a larger distribution of sizes. Saline conditions also 18 allow predators such as carnivorous conchs (multiple species including Stramonita haemastoma, Urosalpinx species, Melongena corona, and more), or stone crabs (M. mercenaria) that prefer high salinities access to previously estuarine area, increasing predation rates (Garland & Kimbro 2015, Pine III et al. 2015, Kimbro et al. 2017).

Interest in Restoration

Recognition of the ecosystem services, economic benefits, and intrinsic values provided by oyster reefs has led to interest amongst coastal managers in restoring this habitat. Ecologists and coastal planners also recognize the place that restored oyster habitat function can have in larger scale ecosystem restoration (Stokes et al. 2012).

The values provided by ecosystem services in the GOM alone are estimated to be in the billions of dollars, and shoreline protection in particular is estimated to be worth $23 billion dollars per year (Stokes et al. 2012).

Restoration of oyster habitat is an emerging industry in the Gulf Coast region.

Early efforts have focused on restoration aimed at improving fisheries and at its most simple level involves adding hard substrate to which oyster spat can attach and grow

(Stokes et al. 2012). However, interest is growing in restoration and management of oyster habitat to provide multiple ecosystem functions beyond the direct commodity of oyster harvest (Brumbaugh et al. 2006, Stokes et al. 2012). There are generally three types of oyster reef restoration techniques applied in the GOM: 1) application of loose cultch (crushed, hard) material, often applied in subtidal zones, 2) use of cultch material contained in a vessel, such as a mesh bag or affixed to mats, more often used in intertidal zones, or 3) precast concrete structures. Materials used during restoration include oyster shell, crushed limestone, granite, and concrete (Stokes et al. 2012). The restoration technique chosen depends on the physical characteristics of the area to be restored, desired goals of restoration, and funds available (Stokes et al. 2012). 19

After installation, monitoring is necessary to determine project success and inform future projects of the benefits and outcomes of different restoration methods.

With many projects, project goals and monitoring are still focused on oyster recovery alone. Monitoring after restoration should include ecosystem metrics such as community biodiversity and abundance measures (Hadley et al. 2010). There are few published reports of monitoring of oyster communities post-restoration. Brown et al.

(2013) compared historic oyster reefs to new (<5 years old) and old (>6 years old) created reefs and all artificial reef metrics of nekton and benthic macrofaunal assemblages were similar to natural reefs, except that benthic macroinvertebrate abundance was higher on old rock reefs. Humphries et al. (2011) found that unique and abundant benthic macroinvertebrates and fish quickly colonized newly constructed cages of oyster shell compared to mud bottom. Conversely, Gregalis et al. (2009) found that response of reef macroinvertebrates and fish to restoration was highly variable and linked to location-specific biophysical characteristics. Geraldi et al. (2009) found that crustacean communities showed no difference to the addition of oyster material, and demersal fish communities showed a weak positive difference. The results vary by project, and do not present a clear picture of effective or ineffective restoration techniques.

Restoration can be quite costly, for example, two restoration projects in Alabama cost $3.5 million and $750,000, respectively, in 2012 (Kroeger 2012). Because of this cost, maximizing restoration dollars by strategically directing them to where they will be most effective is of interest to community decision makers who are planning and funding restoration projects. Better understanding of how restored oyster communities function compared to natural ecosystems can inform planning for restoration projects.

Oyster Reefs in the Apalachicola Bay: Fishery, Management, and Collapse

Historically, the Apalachicola Bay was a relatively pristine and productive estuary located in the northern GOM along the Florida panhandle, encompassing approximately

550 km² (Edmiston 2008). The main source of fresh water to the system is the

Apalachicola River, an alluvial river that is the largest in Florida and ranks 21st in flow in the United States. It is formed at the confluence of the Chattahoochee and Flint Rivers, and flows 275 km south to the Apalachicola Bay. The Apalachicola-Chattahoochee-

Flint (ACF) Basin, which also includes a smaller contribution from the Chipola River, drains an area of 51,280 km² of Georgia, Alabama, and Florida. The upstream rainfall from these areas have a great influence on the Apalachicola River flows discharging into the bay (Fig. 1-1) (Livingston et al. 1975, Edmiston 2008).

Subtidal oyster bars populated by C. virginica are an important ecological and commercial component of this estuarine system, covering 10% of the bay (Fig. 1-2)

(Livingston 1984). Historically, approximately 90% of Florida’s oyster harvest and 10% of the nation’s oyster harvest came from Apalachicola Bay (Edmiston 2008). Franklin county oyster landings peaked in 1981 at >6 million pounds. Until recent years, oysters accounted for half of the county’s income and were a significant contributor to local and state economies (Edmiston 2008, Sempsrott et al. 2012).

The long history of the oyster fishery in the Apalachicola Bay has been accompanied by management and maintenance of the oyster habitat. Cultching, or preparing oyster beds with additional hard material to provide a point of attachment for spat, has occurred in Apalachicola Bay for over a century. Private lease holders and a variety of public entities including Franklin County, Florida Department of Health, Florida

Department of Natural Resources, Florida Department of Environmental Protection

(FDEP) and currently the Florida Department of Aquaculture and Consumer Services 21

(FDACS) have carried out cultching activities. Processed fresh shell, also called green shell, collected from oyster dealers has been the primary material used for this. Other materials, such as clam shells, have been used when green shell was not available. It was not until recently that an end of the FDACS cultching program, lack of enforcement of the state shell buyback program, and lack of processed shell availability due to lower harvest has led to reliance on alternate materials, namely mined, fossilized oyster shell material (J. Shields pers. communication).

Oyster landings in the Apalachicola Bay decreased drastically in 2012, due to the effects of high salinities in the bay caused by reduced freshwater flows from the

Apalachicola River (Havens et al. 2013, Kimbro et al. 2017). This led to a reduction in the number of all size classes of oysters in the Bay (Havens et al. 2013). On August 12,

2013, the Florida west coast oyster fishery was declared a fisheries disaster by NOAA and commercial oyster landings have continued to decline (NOAA 2013, FFWCC 2017).

This fisheries disaster has led to an increase in efforts to restore the oyster communities in the bay with managers, local officials, and seafood workers seeking restoration funds and project opportunities.

Gaps in Knowledge

Despite the establishment of the restoration industry in response to the degradation of oyster habitats, scientific understanding of community-wide response to restoration is lacking. Because oyster fisheries are so valuable, monitoring of projects often focuses on the oysters themselves and may or may not include additional components (Hadley et al. 2010, Stokes et al. 2012). While a few peer-reviewed publications looking at restoration effects exist, the literature reviewed in this chapter indicated varying results. Moreover, the reason for this lack of clarity is not well

22 understood. More study is needed to tease out the causes of discrepancies reported and identify factors of both successful and unsuccessful restoration projects.

Figure 1-1. The Apalachicola Bay. The Apalachicola Bay is in the Florida panhandle at the foot of the Apalachicola-Chattahoochee-Flint watershed. Map courtesy of C. Snyder, ANERR.

Figure 1-2. Oyster bar habitat in the Apalachicola Bay. Map courtesy of C. Snyder, ANERR.

CHAPTER 2 PATTERNS OF COLONIZATION OF DECAPOD CRUSTACEANS TO NATURAL AND NEWLY RESTORED OYSTER REEFS IN APALACHICOLA BAY, FLORIDA

Introduction: Ecosystem Services Provided by Oyster Reefs and Interest in Restoration Outcomes

Oyster reefs composed of the species Crassostrea virginica have historically been a significant ecological and economical component of temperate estuaries throughout the Atlantic coasts of North and South America (Beck et al. 2009). The biogenic structure oysters create provides complex physical habitat utilized by a large and unique assemblage of organisms, causing oysters to be included in an important group of habitat foundation species termed ecosystem engineers (Jones et al. 1994,

Grabowski & Peterson 2007). Oyster reefs are associated with a multitude of functions beyond direct oyster fisheries that are valuable to coastal areas. Oyster reefs are decreasing, with an estimated 85% of reefs declining globally and functionally extinct in

37% of estuaries (Beck et al. 2009, Beck et al. 2011). Causes of decline include overfishing, disease, poor water quality and pollution, poor recruitment, sedimentation, impacts from nonnative species, and loss of reef stability, which hampers recovery ability (MacKenzie Jr. et al. 1997, Kirby 2004, Beck et al. 2011, Seavey et al. 2011).

Multiple ecosystem services are associated with oyster reefs. During the process of filtration feeding, oysters consume phytoplankton, also taking nitrogen, phosphorus, and carbon from the water column (Dame et al. 1984, Dame et al. 1989,

Newell & Langdon 1996, Cressman et al. 2003). Removal of these components can reduce eutrophication and harmful algal blooms, and lead to increased water clarity, which can help nearby seagrass communities (Newell 2004, Newell & Koch 2004). The removal by oysters of particles from the water column and their subsequent deposition onto the benthos as feces and pseudofeces also aids in benthic-pelagic coupling of

26 nutrients (Coen et al. 2007, Piehler & Smyth 2011, Wall et al. 2011). Lastly, oysters buffer coastlines from wave action and increase sediment accretion in nearby marshes

(Meyer et al. 1997, Piazza et al. 2005, Borsje et al. 2011, Gedan et al. 2011).

Oyster habitat supports a high abundance of organisms which have both intrinsic and economic values. Multiple studies have found oyster habitats associated with unique communities, high abundance and biodiversity of organisms (see Glancy et al.

2003, Shervette & Gelwick 2008, Stunz et al. 2010, Kingsley-Smith et al. 2012, Gain et al. 2017). Not surprisingly, loss of oyster habitat has been linked to loss of biodiversity

(Lotze et al. 2006, Airoldi & Beck 2007). However, few studies have compared the communities living on restored oyster bars to those living in natural communities.

Brown et al. (2013) compared benthic macroinvertebrate and nekton assemblages in the northern GOM between natural oyster reefs, newly restored (<5 years old), old (>6 years old) restored reefs, and restored reefs made of either shell or rock. They found communities on created reefs similar to those on natural reefs, but benthic macroinvertebrate abundance higher on older created rock reefs than natural reefs.

Quan et al. (2012) looked at communities on restored C. ariakensis reefs in China, finding abundance, richness, and biomass of benthic macrofauna to be positively correlated with oyster development up to five years after site restoration, but associations between benthic macrofaunal and oyster population metrics were not always consistent suggesting other factors may influence this relationship. They also did not compare the restored site to any unrestored reefs or control sites. Hadley et al.

(2010) also found abundant crabs and mussels associated with restored oyster reefs, but again, did not compare these communities to natural sites. Luckenbach et al.

(2005) studied intertidal and subtidal reef systems, and found most metrics of oyster

27 reef communities were positively correlated with live oyster growth on restored reef sites, but again did not compare to other habitats or control areas.

Recognition of the multitude of ecosystem services provided by oyster reefs has heightened interest in restoration in recent decades. For example, in 2011 a large group of > 40 state, federal, and nonprofit entities launched the 100-1000 Restore

Coastal Alabama Partnership with the goal of restoring 100 linear miles of oyster reefs in Alabama waters (Stokes et al. 2012, Alabama Coastal Foundation 2017). The

Florida Oceanographic Society has restored nearly 5,575 m² of oyster habitat in the St.

Lucie Estuary, FL (Florida Oceanographic Society 2017). In 2014, the FDEP obtained over $5M from the 2010 Deepwater Horizon Oil Spill Natural Resource Damage

Assessment (NRDA) Trustees to add cultch to 85 hectares of oyster bars in three

Florida bays in the Northern GOM (Deepwater Horizon Project Tracker 2018). The list of projects and partners is long, but results of monitoring efforts after project construction are underreported. Without the results of clear, measured outcomes from multiple projects to inform them, development of best restoration practices is hindered.

The paucity of studies monitoring restoration effects, differences in methodology, time scales, types of oyster species and reefs (intertidal versus subtidal), and potential differences in other physical factors creates a lack of comparability amongst studies.

Lack of comparison with natural reefs may be impossible due to project restraints or lack of appropriate comparison sites, but still leaves a gap in understanding how restored sites compare to natural sites. Greater understanding of the function of restored oyster communities under various conditions is needed to better inform the most appropriate restoration efforts given an area and project goals. The purpose of this study is to monitor newly restored oyster reef communities in comparison to those

28 of surrounding natural oyster reefs and gain a better scientific understanding of how restored communities function.

The first objective of this chapter was to measure species richness, total abundance, and biodiversity of crustaceans on natural and restored oyster reefs in

Apalachicola Bay, Florida. I hypothesized that species richness, abundance, and biodiversity would all be lower on restored sites compared to natural sites. The second objective of this chapter was to use decapod crustaceans to determine if members of oyster reef communities show a preference for substrate type when recruiting to collectors. I hypothesized that substrate type would not make a difference in settlement patterns of crustaceans to collectors. The third objective of this chapter was to use select decapod crustacean species that are host to the parasitic barnacle Loxothylacus panopaei to determine if parasite prevalence differed between newly restored and natural study sites. Because L. panopaei castrates its host, effectively removing it from the reproducing population, and changes host behavior and life span, its presence has the potential to alter population dynamics (Hines et al. 1997, Kruse & Hare 2007, Eash-

Loucks et al. 2014). I hypothesized that parasite levels would be lower at newly restored sites compared to natural areas in the bay.

Methods

Study Site Selection

Study sites were selected from subtidal oyster bars within Apalachicola Bay,

Florida. Two areas where natural bars occur in Apalachicola Bay, Platform Bar and Cat

Point, were selected (Fig. 2-1). These sites undergo commercial harvest and periodic cultching by management agencies to replace shell material, which until recent years has consisted mainly of fresh processed or green oyster shell (a term used for fresh oyster shell material) (J. Shields, personal communication). Two newly constructed,

29 fossilized shell reefs located near historic oyster areas with little natural hard substrate remaining, Bulkhead Bar and Hotel Bar, were selected as study sites (Fig. 2-1). These sites were created in summer 2015 as part of a National Fish and Wildlife Foundation

(NFWF) funded project led by the Florida Fish and Wildlife Conservation Commission

(FWC), the University of Florida (UF), and FDACS to determine optimal restoration reef shelling densities. Each NFWF site includes five study plots at 0 (control), 153, 230,

305, and 382 cubic meters per 0.4-hectare shell plots. For this project, the 153-cubic meter per 0.4-hectare plot was used because this substrate density most closely resembled the substrate density at the natural sites. This similarity was desired to remove any confounding factors that may occur from comparison between differing reef densities between historic and fossilized sites. All sites were chosen in close proximity in the eastern area of the bay to lessen differences in water quality regimes that could create confounding factors when comparing communities between sites.

Decapod Crustacean Collection, Preservation, and Identification

Collectors designed to measure settlement by decapod crustacean communities were deployed at all research sites. Collectors measured 30 cm³ and were made of vinyl-coated 16-gauge welded wire mesh with 2.5 cm² openings. Each was filled with either fossilized or green oyster shell (Fig. 2-2, 2-3, and 2-4). Three fossilized shell collectors and three green shell collectors were daisy-chained together and deployed on the benthos at each site. The size and quantity of natural and fossilized oyster shells was similar in all collectors to provide a consistent amount of surface area and interstitial space for recruits.

Collectors were sampled every two months for one year. The first deployment began December 31, 2015, six months after site creation, and the final deployment ended December 16, 2016 (Table 2-1). This time frame included a complete range of 30 seasons. Collectors were retrieved mid-way between the full moon and new moon cycles. New collectors were deployed simultaneously with retrieval. Retrieved collectors were placed inside bins to retain any escaped organisms until they could be returned to the lab for processing (see Fig. 2-5). All material in each collector was carefully examined and organisms placed inside a sample jar and preserved with 10% buffered formalin. After 48 hours, samples were transferred to 70% ethanol. All decapod crustaceans were identified to species level and measured (mm, carapace length or width). Xanthid crabs were sexed and ovigerous females noted. One individual Hexapanopeus crab could not be identified to species level and is referred to as Hexapanopeus sp. One species of shrimp could not be identified and these individuals are listed as UID shrimp. Fifty-four shrimps from the genus Palaemonetes could only be identified to the genus level because of damaged or missing second legs necessary to differentiate these species. Due to logistical constraints, only decapod crustaceans are discussed in this document.

External Parasite Load of Xanthid Crabs

Xanthid crabs were examined externally for infection by L. panopaei according to

Hines et al. (1997). Crab abdomens were visually examined for a protruding sacculina externa to indicate parasite presence (Fig. 2-6). Because of the changes to the abdomen caused by L. panopaei, parasitized crabs could not be sexed.

Statistical Analyses

Taxonomic abundance was measured as the number of individuals of a given genus or species. Species richness was defined as the number of species present at a site, compared to the total number of species found at all sites over time. Biodiversity

(H) was calculated using the Shannon-Weiner Diversity Index, defined as follows:

퐻 = −푠푢푚[(푝푖) × ln(푝푖)] Eq. 2-1

Where sum = Summation, pi= Number of individuals of species i/total number of samples.

A split block design was applied to sampling sites, using each oyster bar as a plot and each substrate collector as a subplot, and time as a sub-sub plot. A linear mixed-effects model was used to test both main effects and interactive effects of site age (historic or new) and substrate type (green shell or fossilized material) on species abundance, richness, and biodiversity. An analysis of variance (ANOVA) was then used to test for significant differences between treatment groups. Each sampling period was tested individually for abundance, richness, and biodiversity, then all sampling periods were combined and tested. Data was tested for normality and found to be normally distributed. Models were analyzed using R version 3.3.3 and R Studio

1.1.383. Results were considered significant when α < 0.05.

Water Quality

The Apalachicola National Estuarine Research Reserve (ANERR) maintains a long-term water quality monitoring station at Cat Point in the Apalachicola Bay. A YSI

EXO II sonde is deployed approximately 15 cm off the benthos at this site and collects temperature, conductivity, salinity, pH, dissolved oxygen, and depth data year-round at

15 min intervals (Fig. 2-1). At each retrieval/deployment of collectors, water quality parameters including temperature, conductivity, salinity, pH, and dissolved oxygen were also taken at each site using a YSI Pro DSS instrument.

Results

Decapod Crustacean Species Abundance

Twenty-five decapod crustacean species were collected over the year-long sampling period (Table A-1). The most common decapod crustacean through all

32 sampling periods was Eurypanopeus depressus, the flat-backed mud crab, of which

28,988 individuals were collected, followed in abundance by the green porcelain crab,

Petrolisthes armatus (n=3,159), the Texas mud crab, Dyspanopeus texanus (n=994), and the Florida stone crab, Menippe mercenaria (n=829).

Abundance of all crustaceans followed seasonal trends, with abundance highest during late summer and early fall (Fig. 2-7, Table A-2). During sampling periods two

(Mar-Apr) and five (Sep-Oct), substrate had a significant effect on abundance (Mar-Apr:

F1,13 = 9.40, p = 0.009; Sep-Oct: F1,18 = 6.22, p = 0.023). No other significant differences were found by site age, substrate type, or interactive effects between the two at any sampling period (Table 2-2). With all sampling periods combined there was no statistically significant difference in abundance between sites of different age (F1,2 =

0.11, p = 0.78), collectors of different substrate type (F1,132 = 0.94, p = 0.33), or with interactive effects between the two (F1,132 = 0.02, p = 0.88).

Decapod Crustacean Species Richness

Species richness was determined by adding the total number of different species in each collector at each site and sampling time (Figure 2-8, Table A-3). During sampling period two (Mar-Apr), site age, substrate type, and interactive effects were all significant effects on species richness (Site age: F1,2 = 56.25, p = 0.017; Substrate: F1,18

= 12.00, p = 0.003; Interactive effects: F1,18 = 5.33, p = 0.033). Species richness was not linked to site age, substrate type, or interactive effects at any other sampling period

(Table 2-2). When all sampling periods were combined, there was no statistically significant difference in richness with site age (F1,2 = 1.35, p = 0.36), collectors of different substrate type (F1,132 = 0.03, p = 0.86), or interactive effects between site age and substrate type (F1,132 = 0.00, p = 0.98).

Decapod Crustacean Species Biodiversity

Biodiversity of decapod crustaceans was calculated using the Shannon-Weiner biodiversity index for each collector at each site and sampling time (Figure 2-9, Table A-

4). During sampling period six (Nov-Dec), substrate had a borderline significant effect on biodiversity (F1,18 = 4.31, p = 0.053). Biodiversity was not linked to site age, substrate type, or interactive effects at any other sampling period (Table 2-2). When all sampling periods were combined, there was no statistically significant difference in biodiversity between site age (F1,2 = 1.35, p = 0.37), substrate type (F1,54 = 0.03, p =

0.86), or with interactive effects between the two (F1,54 = 0.00, p = 0.98).

Size Structure of Most Abundant Crustacean Species

The size structures of the most abundant decapod crustacean populations were examined in more detail. Fig. 2-10 shows the size of E. depressus individuals by site and sampling period. The average size of E. depressus decreased through the year, but size increased in the final sampling period, December. Similar plots of the next most abundant species, P. armatus, D. texanus, and M. mercenaria did not show a similar trend in size over time as that apparent in E. depressus, but these species are not present at all sites during each time period, making trends difficult to detect (Fig. 2-

11, 2-12, and 2-13).

Gender and Reproductive Status of Most Abundant Crustacean Species

Sex ratios and percent ovigerous females were examined for three of the most abundant species, E. depressus, P. armatus, and D. texanus. Sex ratios and ovigerous females were not considered in M. mercenaria because the majority of this species collected were juveniles and sex was not easily distinguished below 10 mm CW.

The sex ratio of E. depressus females to males at all sites throughout the year was 1.6:1. E. depressus females were found ovigerous from sampling 1 (Jan-Feb)

34 through sampling 4 (Jul-Aug), but peaked during sampling 2 (Mar-Apr) (Table 2-3). The sex ratio of P. armatus females to males at all sites throughout the year was 0.8:1. P. armatus females were found ovigerous during sampling 4 (Jul-Aug) and sampling 5

(Sep-Oct) (Table 2-4). The sex ratio of D. texanus females to males at all sites throughout the year was 1.1:1. Only a single ovigerous D. texanus female was collected, during sampling 4 (Jul-Aug) at Platform Bar. It should be noted that P. armatus and D. texanus were not present at all sites throughout the year.

External Parasite Prevalence in Select Xanthid Crabs

The xanthid crabs E. depressus and Panopeus herbstii were hosts of the external barnacle parasite L. panopaei. In E. depressus, prevalence ranged from 0-

18.1% with an average 2.8 (±0.7) % per collector. During sampling period four (Jul-

Aug), the interactive effects between site age and substrate had significant effects on parasite prevalence (F1,18 = 4.49, p = 0.048). During sampling period five (Sep-Oct), substrate had significant effects on parasite prevalence (F1,18 = 5.36, p = 0.033). When all sampling periods were combined, there was no significant effect by site age, substrate, or interactive effects between the two (Site age: F1,132 = 0.00, p = 0.995;

Substrate type: F1,132 = 0.02, p = 0.875, Interactive effects: F1,132 = 0.11, p = 0.745). Fig.

2-14 shows the mean prevalence among collectors per substrate type by site during each sampling period, ANOVA results for all sampling periods are shown in Table 2-5.

Prevalence of the parasite in P. herbstii was 30.7 (±9.4) % and ranged from 0-

100% per collector. During sampling period two (Mar-Apr), substrate and interactive effects between substrate and site age were borderline significant (Substrate: F1,18 =

4.20, p = 0.055; Interactive: F1,18 = 4.20, p = 0.055). Site age, substrate, and interactive effects were not significant for P. herbstii during any other sampling period. When all sampling periods were combined, there was no significant effect by site age, substrate, 35

or interactive effects between the two (Site age: F1,132 = 0.30, p = 0.639; Substrate type:

F1,132 = 0.35, p = 0.557, Interactive effects: F1,132 = 0.63, p = 0.430). Fig. 2-15 shows the mean prevalence among collectors per substrate type by site and sampling period.

ANOVA results for all sampling periods are shown in Table 2-5. Abundance of P. herbstii individuals was much lower than E. depressus, so while prevalence may have been higher, overall number of parasitized individuals was much lower. Prevalence estimates in all crabs were conservative because they relied on external examination and presence of parasite externae. This examination did not detect early stage parasites that were not yet visible externally.

Water Quality Monitoring

Water Quality data recorded at Cat Point by ANERR’s data sonde followed seasonal trends and freshwater inputs. The mean dissolved oxygen (DO) throughout the sampling period was 7.10 (±0.01) mg L-1. Hypoxic events where DO fell below 3.0 mg L-1 were not common, and generally occurred with warmer water temperatures during the summer months and were less than an hour in duration. The mean pH value was 8.0 (±0.002), but ranged from 6.8-8.5 during the study period. Water temperature reflected seasonal trends, with a low of 11.8°C on 1/12/2016 and a high of 33.6°C on

7/12/2016. Mean salinity values were 20.18 (±0.09) psu, but varied greatly throughout the year with changing freshwater inputs and tidal changes. Fig. 2-16 shows the salinity and temperature recorded at Cat Point during the study period. Data was not recorded from 1/17/2016-2/17/2016 due to instrument power failure.

Water quality readings were taken at all four sites at each retrieval and deployment date. Sites showed similar water quality conditions at all sites at each time period. Dissolved oxygen, temperature, and pH were similar between sites during each retrieval/deployment date, while salinity and conductivity had a more variance between 36 sites. Table 2-6 shows the mean, standard deviation, and range of each water quality parameter during a given sampling period. An analysis conducted by the NWFW restoration team compared point water quality readings taken at NWFW project sites including Bulkhead Bar and Hotel Bar to long-term ANERR sites. The analysis showed no statistical difference between water quality at the restored sites and ANERR locations (A. Kane, pers. communication).

Discussion

Abundance, Richness, and Biodiversity of Restored and Historic Oyster Reef Communities

Our results suggest that newly created oyster reefs attract and support crustacean colonizers similar to nearby historic reefs, and differences in substrate type only resulted in small differences between species assemblages. There was little difference between 6-18-month-old restored reefs and natural reefs of indeterminate age or substrate makeup of green oyster shell compared to fossilized material when considering species abundance, richness, or biodiversity. We detected small differences during a minority of sampling periods between abundance, richness, and biodiversity in decapod crustacean communities colonizing collectors containing different substrate types. Site age was only significant to species richness during one sampling period. Overall, restored reefs and both substrate types attracted similar colonizers in terms of abundance, richness, and biodiversity.

These results are similar to the findings of the few other studies comparing restored and historic reefs. Brown et al. (2013), who compared natural reefs to new (<5 years old) and old (>6 years old) created shell and rock reefs, found similar assemblages at all site types except a greater macroinvertebrate abundance was present on old rock reefs. Likewise, Humphries et al. (2011) found that oyster reefs

37 were colonized very quickly and communities resembled nearby natural ones at four months post-construction. While they found no difference between community metrics at 4-16 months after creation, they did not look at these sites beyond 16 months of age.

Neither the present study, nor the Humphries et al. (2011) study, looked at reefs beyond

16 to 18 months of age to compare to the older reefs, as in Brown et al. (2013). These results are positive for the restoration methods used in this area and indicate that restored oyster reef ecosystems can closely mimic natural systems in a relatively short time frame.

Restored Oyster Reef Decapod Crustacean Reproduction and Colonization

In addition to community-scale metrics of richness, abundance, and biodiversity, additional measures of individual species including size structure, sex ratios, spawning periods, and parasite prevalence supported the idea that the ecological structure of newly constructed sites was similar to nearby, natural reefs. The flat-backed mud crab

E. depressus was present in great abundance at all sites during all sampling periods.

Females of this species have an extended spawning period lasting from February to

August. Abundance of this species increased in the spring and summer months at all stations, and the mean CW of individuals at all sites decreased through time. This indicates recruitment of many small juveniles to the population and possibly a yearly recruiting class. Because of its abundance and occurrence across sites and time, this species serves as an excellent indication that all sites are functioning similarly, regardless of type. Environmental conditions favored E. depressus populations with high salinities throughout the year. E. depressus fecundity and recruitment are inversely related to freshwater flow (Tolley et al. 2013), and the present study measured

Cat Point salinity values of over 20 psu for the study year.

While no other species was as prevalent as E. depressus, there was evidence of juvenile recruitment from multiple species at all sites regardless of age or substrate type. For example, juvenile M. mercenaria appeared in the late summer months. Adult

M. mercenaria including ovigerous females were not observed, but this is likely because they were excluded from collectors due to their carapace size. With the exception of a small number of individuals collected in the spring, D. texanus appeared only during the latter part of the year. We collected only a single ovigerous female and most individuals were under the size at maturation for this species (16 mm CW Strieb et al. 1995), indicating the individuals we saw may have represented recruitment from a year class.

In the present study, juveniles of multiple decapod crustacean species recruited to sites during the June, August, and October sampling periods. These species all utilize external reproduction and planktonic larval stages, and thus recruits are likely derived from larvae dispersed from nearby populations (Williams 1984, Anger 2001, Tolley et al.

2012).

Influence of Nearby Communities Oyster Reef Colonization

Despite the swiftness at which restored site collectors were colonized, other research indicates site selection is critical to restoration success. Gregalis et al. (2009) found that response to reef restoration by resident and transient reef species is extremely variable and likely related to local biophysical characteristics of neighboring restored areas. Geraldi et al. (2009) compared fish and crustacean communities before and after addition of oyster material in tidal creeks and find no difference in crustacean communities and only a weak positive difference in demersal fish communities. They attribute this lack of positive effect to redundancy in nearby habitats, drawing attention to the idea that close proximity to already healthy habitats may limit the effects of restoration. At the same time, donor communities are needed to provide recruits to new 39 restoration sites. Grabowski et al. (2005) find that ecosystem recovery is dependent on the characteristics of surrounding habitats when looking at crustacean and fish community response to oyster restoration. While larger nekton may be able to travel relatively easily to new sites, settlement of less motile organisms such as non-swimming decapod crustaceans may be limited by mechanisms of larval dispersal, pelagic larval duration, and variation in environmental factors such as water movement (Anger 2001).

The complex structure and refuge provided by oyster reefs is a significant factor explaining the organisms found there (Glancy et al. 2003, Brown et al. 2013). Structure provides organisms with valuable refuge from predators (Dittel et al. 1996, Wong 2013,

Hernández Cordero & Seitz 2014), and changes overall trophic dynamics within an oyster ecosystem by enhancing predator foraging efficiency (Grabowski & Powers

2004, Grabowski 2004). This study found no difference in the species richness, biodiversity, or abundance of organisms found on green oyster shells compared to fossilized material, which supports the theory that it is the available habitat created by reef materials that is important to organisms, not necessarily material composition.

Restored reefs serve the same functions as natural reefs, regardless of substrate composition.

The population of the Atlantic mud crab P. herbstii found in this study presents a different picture of life history than other xanthid crabs (E. depressus, M. mercenaria, and D. texanus) captured during the sampling period. P. herbstii can attain maximum sizes of 53 mm (female) to 62 mm (male) (Williams 1984), and most specimens collected in the present study fell within the subadult size range for this species. P. herbstii utilizes vertical migration to aid in larval retention within estuaries which is similar to other xanthids found here (Williams 1984, Rodriguez & Epifanio 2000). There are a few explanations for this observation. Few ovigerous females were observed, 40 which may be an artifact of adult exclusion from collectors due to opening size. Another explanation is the high prevalence of L. panopaei among P. herbstii, which could have suppressed reproduction in the species during the sampling period. The high prevalence is also notable because multiple reports looking at L. panopaei prevalence among multiple xanthid crab species in other estuaries report no infection in P. herbstii, even when the parasite is present in other species at that site, including E. depressus

(Hines et al. 1997, Freeman et al. 2013, Eash-Loucks et al. 2014). P. herbstii may also utilize the space available between oyster shells less than smaller xanthid crabs because of their larger size (McDonald 1982), another possible explanation of why abundance was less than E. depressus in substrate filled collectors.

Parasite Prevalence amongst Xanthid Crabs

The present study found parasitism by the castrating barnacle L. panopaei at low levels in the xanthid crab E. depressus (2.8 ± 0.7% per collector) and somewhat higher levels in P. herbstii (30.7 ±9.4% per collector). During most sampling periods for both xanthid species, there was no difference in parasite prevalence with regard to site age or substrate type within collectors. L. panopaei is native to the GOM, southeast Florida, and the Caribbean coasts. It has been introduced to the eastern Atlantic coast of North

American, now ranging from Long Island Sound in the north to northern Florida in the south (Kruse & Hare 2007, Freeman et al. 2013, Fofonoff et al. 2018). The

Apalachicola Bay falls within this native range of L. panopaei and multiple studies have published prevalence levels of the parasite in xanthid crabs within its introduced range, though fewer reports exist for its native range. Prevalence rates across all host xanthid species of 0-47%, 0-83%, and 0-91% have been reported in introduced ranges of the parasite North Carolina, coastal Virginia, and the Chesapeake Bay, respectively, and 0-

9% in its native range in the Indian River Lagoon in Florida, but large variations in 41 prevalence were noted (Hines et al. 1997). Interestingly, though L. panopaei is a reported parasite in P. herbstii, prevalence levels are not reported within its native range and studies within introduced ranges where the parasite occurs in other xanthid species find no infection in P. herbstii (Hines et al. 1997, Freeman et al. 2013, Eash-Loucks et al. 2014).

Nearby populations of benthic macrofauna and nekton serve as feeder populations to colonize restored sites for all oyster reef residents, including L. panopaei.

They have a relatively short larval stage of 48 hours (Walker et al. 1992), and dispersal seems to be limited in distance with infection rates much reduced at 1.0 and 10 km as compared to 0.1 km away from a source infected population (Grosholz & Ruiz 1995).

Dispersal is also likely limited in some environmental conditions because L. panopaei larvae are intolerant of low salinity (Reisser & Forward 1991, Walker et al. 1992), but in this study salinity remained high. Impacts of infection on mud crabs are not well understood. Xanthid mud crabs are not harvested commercially, but are known to feed on juvenile oysters and are serve as prey items for commercially harvested fish

(Fofonoff et al. 2018). Infection by L. panopaei results in castration of the crab host and suspends molting and growth, thus that crab is no longer a reproductive member of the population (O'Brien & Skinner 1990, Gould 1996, Fofonoff et al. 2018). Parasitized crabs may also have altered behavior (Hines et al. 1997, Kruse & Hare 2007), which has the potential to alter the population’s dynamics if prevalence is high. Changes in these population dynamics have the potential to alter estuarine food webs.

Limitations of Present Methods and Future Directions

Disturbance in the form of harvesting has the potential to impact oyster communities, and unfortunately could not be well quantified in the present study. Beck

(2012) found greater invertebrate diversity on harvested compared to unharvested 42 reefs, which he attributed to more diverse substrate availability or to the intermediate disturbance hypothesis attributed to Connell (1978), Roxburgh et al. (2004), and Shea et al. (2004). Unfortunately, disturbance could not be controlled for in this study.

Historic sites were open to public oyster harvest during part of the study period, but oyster fishermen are not required report the reef from which they harvest from, so harvest levels on specific reef are difficult to determine. Restored sites were closed to harvest, but anecdotal reports indicate illegal harvesting may have occurred; again, any levels are difficult to determine.

While the collectors used here were efficient at recruiting smaller decapod crustaceans and benthic nekton, this does not represent the entire reef community.

Collectors may also have excluded some larger forms of certain decapod crustacean species such as M. mercenaria, Callinectes sapidus, and more. Future studies may consider comparing additional groups of animals utilizing oyster reefs such as fish, which would need to be captured using alternate methods besides the collectors used here. Additional parameters such as turbidity levels, chlorophyll-a levels, and wave attenuation at nearby shorelines before and after restoration could add additional information about the efficacy of restoration depending on project goals. Decapod crustacean communities are only one of many groups using oyster habitats for food and refuge. The crustacean communities in this study indicate restored areas here function similarly to natural areas, a positive indication that restoration techniques used in the

Apalachicola Bay’s subtidal bars are effective at restoring oyster communities in this estuary.

Figure 2-1. Research sites. Map includes Apalachicola Bay oyster bars, research sites, and ANERR water quality stations. Map courtesy of C. Snyder, ANERR.

Figure 2-2. Green shell (left) compared to fossilized (right) oyster shell material. Photo courtesy of author.

Figure 2-3. Green shell collector (left) and fossilized shell collector (right) before deployment. Photo courtesy of author.

Figure 2-4. Collectors tied together prior to deployment. Photo courtesy of author.

Figure 2-5. Collectors stored in individual bins after retrieval to keep organisms with each collector. Photo courtesy of author.

Figure 2-6. Subadult P. herbstii with mature L. panopaei infection. Photo courtesy of author.

Figure 2-7. Decapod crustacean abundance at sites by substrate type. a) Period 1, Jan-Feb. b) Period 2, Mar-Apr; *During this sampling period, substrate had a significant effect on abundance (p = 0.009). c) Period 3, May-Jun. d) Period 4, Jul-Aug. e) Period 5, Sep-Oct; *During this sampling period, substrate had a significant effect on abundance (p = 0.023). f) Period 6, Nov-Dec. Error bars represent ± 1 SE. 50

Figure 2-8. Decapod crustacean species richness at sites by substrate type. a) Period 1, Jan-Feb. b) Period 2, Mar-Apr; *During this sampling period, site age, substrate type, and interactive effects had a significant effect on richness (p = 0.017, p = 0.003, and p = 0.033, respectively). c) Period 3, May-Jun. d) Period 4, Jul-Aug. e) Period 5, Sep-Oct. f) Period 6, Nov-Dec. Error bars represent ± 1 SE.

Figure 2-9. Decapod crustacean biodiversity, calculated by the Shannon-Weiner biodiversity index, at sites by substrate type. a) Period 1, Jan-Feb. b) Period 2, Mar-Apr. c) Period 3, May-Jun. d) Period 4, Jul-Aug. e) Period 5, Sep-Oct. f) Period 6, Nov-Dec; *During this sampling period, substrate had a borderline significant effect on biodiversity (p = 0.053). Error bars represent ± 1 SE.

Figure 2-10. E. depressus sizes during all sampling periods by site and substrate type. a) Period 1, Jan-Feb. b) Period 2, Mar-Apr. c) Period 3, May-Jun. d) Period 4, Jul-Aug. e) Period 5, Sep-Oct. f) Period 6, Nov-Dec. Boxplot summary: The dark band in the middle of the box represents the 50th percentile or median; the bottom and top of the box represent the 25th and 75th percentiles (lower and upper quartiles, respectively). The upper whisker = min(max(x), Q_3 + 1.5 * IQR), lower whisker = max(min(x), Q_1 – 1.5 * IQR), where IQR = Q_3 – Q_1, the box length. Circles represent outliers.

Figure 2-11. P. armatus sizes during all sampling periods by site and substrate type. a) Period 1, Jan-Feb*. b) Period 2, Mar-Apr. c) Period 3, May-Jun. d) Period 4, Jul-Aug. e) Period 5, Sep-Oct. f) Period 6, Nov-Dec. *No individuals were collected during this sampling period. Boxplot summary: The dark band in the middle of the box represents the 50th percentile or median; the bottom and top of the box represent the 25th and 75th percentiles (lower and upper quartiles, respectively). The upper whisker = min(max(x), Q_3 + 1.5 * IQR), lower whisker = max(min(x), Q_1 – 1.5 * IQR), where IQR = Q_3 – Q_1, the box length. Circles represent outliers. When the sample for a given period, site, and substrate type size was n = 1, only a single band representing the carapace width is shown.

Figure 2-12. D. texanus sizes during all sampling periods by site and substrate type. a) Period 1, Jan-Feb*. b) Period 2, Mar-Apr. c) Period 3, May-Jun. d) Period 4, Jul-Aug. e) Period 5, Sep-Oct. f) Period 6, Nov-Dec. *No individuals were collected during this sampling period. Boxplot summary: The dark band in the middle of the box represents the 50th percentile or median; the bottom and top of the box represent the 25th and 75th percentiles (lower and upper quartiles, respectively). The upper whisker = min(max(x), Q_3 + 1.5 * IQR), lower whisker = max(min(x), Q_1 – 1.5 * IQR), where IQR = Q_3 – Q_1, the box length. Circles represent outliers. When the sample for a given period, site, and substrate type size was n = 1, only a single band representing the carapace width is shown.

Figure 2-13. M. mercenaria sizes during all sampling periods by site and substrate type. a) Period 1, Jan-Feb. b) Period 2, Mar-Apr. c) Period 3, May-Jun. d) Period 4, Jul-Aug. e) Period 5, Sep-Oct. f) Period 6, Nov-Dec. Boxplot summary: The dark band in the middle of the box represents the 50th percentile or median; the bottom and top of the box represent the 25th and 75th percentiles (lower and upper quartiles, respectively). The upper whisker = min(max(x), Q_3 + 1.5 * IQR), lower whisker = max(min(x), Q_1 – 1.5 * IQR), where IQR = Q_3 – Q_1, the box length. Circles represent outliers. When the sample for a given period, site, and substrate type size was n = 1, only a single band representing the carapace width is shown.

Figure 2-14. Percent E. depressus parasitized by L. panopaei by site and sampling period. a) Period 1, Jan-Feb. b) Period 2, Mar-Apr. c) Period 3, May-Jun. d) Period 4, Jul-Aug; *During this sampling period the interactive effects between substrate and site age had a significant effect on parasite prevalence (p = 0.048). e) Period 5, Sep-Oct; *During this sampling period substrate had a significant effect on parasite prevalence (p = 0.033). f) Period 6, Nov-Dec. Error bars represent ± 1 SE.

Figure 2-15. Percent P. herbstii parasitized by L. panopaei by site and sampling period. a) Period 1, Jan-Feb. b) Period 2, Mar-Apr; *During this sampling period, substrate and interactive effects between substrate and site age had a borderline significant effect on parasite prevalence (Substrate: p = 0.055, Interactive: p = 0.055), c) Period 3, May-Jun. d) Period 4, Jul-Aug. e) Period 5, Sep-Oct. f) Period 6, Nov-Dec. Error bars represent ± 1 SE.

Figure 2-14: Cat Point Water Quality Readings Throughout Study Period, 2016 34 32 30

28 Celsius)

° 26 24 22 20 18 16 14 12 10 8 6 Salinity (psu) 4 2

Salinity (psu) and Temperature ( Temperature (psu) and Salinity Temperature (°C) 0

Figure 2-16. Salinity (psu) and temperature (°C) readings from Cat Point during the study period. Data from ANERR’s long-term water quality instrument deployed at that site.

Table 2-1. Collector deployment and retrieval dates.

Natural Sites Restored Sites Sampling Cat Point Platform Bar Bulkhead Bar Hotel Bar Period 1. Jan-Feb 12/31/15-2/28/16 12/31/15-2/28/16 12/31/15-2/27/16 12/31/15-2/27/16 2. Mar-Apr 2/28/16-5/7/16 2/28/16-5/7/16 2/27/16-5/8/16 2/27/16-5/8/16 3. May-Jun 5/7/16-6/25/16 5/7/16-6/25/16 5/8/16-6/25/16 Collectors Lost 4. Jul-Aug 6/25/16-8/28/16 6/25/16-8/28/16 6/25/16-8/28/16 6/25/16-8/28/16 5. Sep-Oct 8/28/16-10/20/16 8/28/16-10/20/16 8/28/16-10/20/16 8/28/16-10/20/16 10/20/16- 10/20/16- 10/20/16 - 10/20/16- 6. Nov-Dec 12/16/16 12/16/16 12/17/16 12/17/16

Table 2-2. Results for site age, substrate, and interactive effects between the two on species abundance, richness, and biodiversity by sampling period. Bold font indicates significance or borderline significance, with significance defined as an α < 0.05.

Abundance Richness Biodiversity Sampling period Testing for DF F Statistic P-value DF F Statistic P-value DF F Statistic P-value 1. Jan-Feb Site Age 2 0.29 0.645 2 0.74 0.480 2 0.95 0.434 Substrate 18 0.60 0.449 18 0.77 0.391 18 0.65 0.429 Interactive Effects 18 2.56 0.127 18 0.09 0.773 18 0.22 0.648 2. Mar-Apr Site Age 2 1.57 0.337 2 56.25 0.017 2 3.33 0.210 Substrate 18 2.20 0.156 18 12.00 0.003 18 1.34 0.263 Interactive Effects 18 3.59 0.074 18 5.33 0.033 18 0.01 0.944 3. May-Jun Site Age 1 0.10 0.803 1 0.59 0.582 1 0.09 0.810 Substrate 13 9.40 0.009 13 1.50 0.243 13 0.28 0.609 Interactive Effects 13 0.10 0.760 13 0.24 0.630 13 0.74 0.404 4. Jul-Aug Site Age 2 0.05 0.844 2 0.00 1.000 2 2.43 0.259 Substrate 18 0.42 0.526 18 0.38 0.546 18 1.66 0.214 Interactive Effects 18 0.32 0.577 18 0.38 0.546 18 0.06 0.808 5. Sep-Oct Site Age 2 0.00 0.980 2 0.32 0.629 2 0.90 0.443 Substrate 18 6.22 0.023 18 0.02 0.895 18 0.04 0.835 Interactive Effects 18 0.11 0.747 18 0.87 0.362 18 1.13 0.303 6. Nov-Dec Site Age 2 0.03 0.882 2 3.45 0.204 2 1.58 0.336 Substrate 18 0.16 0.691 18 0.91 0.353 18 4.31 0.053 Interactive Effects 18 0.31 0.585 18 0.46 0.505 18 0.01 0.921 All periods Site Age 2 0.11 0.775 2 1.35 0.366 2 1.35 0.366 combined Substrate 132 0.94 0.333 132 0.03 0.856 132 0.03 0.856 Interactive Effects 132 0.02 0.884 132 0.00 0.982 132 0.00 0.982

Table 2-3. Percent ovigerous E. depressus females by site and sampling period.

Sampling Period Platform Bar Cat Point Bulkhead Bar Hotel Bar 1. Jan-Feb 9.09 0.00 0.00 0.00 2. Mar-Apr 26.09 33.13 20.60 25.07 3. May-Jun 2.30 18.30 16.21 No samples 4. Jul-Aug 7.47 16.88 8.86 8.37 5. Sep-Oct 0.00 0.00 0.00 0.00 6. Nov-Dec 0.00 0.00 0.00 0.00

Table 2-4. Percent ovigerous P. armatus females by site and sampling period. NA indicates no P. armatus females were present at that site during that sampling period.

Sampling Period Platform Bar Cat Point Bulkhead Bar Hotel Bar 1. Jan-Feb NA NA NA NA 2. Mar-Apr 0.00 NA NA NA 3. May-Jun 0.00 NA NA No samples 4. Jul-Aug 82.35 0.00 0.00 NA 5. Sep-Oct 2.98 5.03 6.70 9.94 6. Nov-Dec 0.00 0.00 0.00 0.00

Table 2-5. Results for site age, substrate, and interactive effects between the two on parasite prevalence in E. depressus and P. herbstii by sampling period. Bold font indicates significance or borderline significance, with significance defined as an α < 0.05. NA indicates there was no parasitism found during the sampling period.

Eurypanopeus depressus Panopeus herbstii Sampling period Testing for DF F Statistic P-value DF F Statistic P-value 1. Jan-Feb Site Age NA 2 0.53 0.543 Substrate NA 18 0.53 0.476 Interactive Effects NA 18 1.47 0.241 2. Mar-Apr Site Age 2 0.06 0.823 2 1.00 0.423 Substrate 18 0.82 0.378 18 4.20 0.055 Interactive Effects 18 0.31 0.584 18 4.20 0.055 3. May-Jun Site Age 1 0.04 0.874 1 3.31 0.320 Substrate 13 0.96 0.345 13 1.29 0.277 Interactive Effects 13 1.63 0.224 13 0.78 0.393 4. Jul-Aug Site Age 2 2.43 0.260 2 0.68 0.497 Substrate 18 1.78 0.199 18 1.20 0.288 Interactive Effects 18 4.49 0.048 18 0.52 0.481 5. Sep-Oct Site Age 2 4.18 0.178 2 0.48 0.562 Substrate 18 5.36 0.033 18 3.00 0.100 Interactive Effects 18 3.53 0.077 18 0.24 0.630 6. Nov-Dec Site Age 2 0.59 0.522 2 0.04 0.859 Substrate 18 3.28 0.087 18 0.39 0.539 Interactive Effects 18 0.01 0.932 18 4.70 0.044 All periods Site Age 2 0.00 0.995 2 0.30 0.639 combined Substrate 132 0.02 0.875 132 0.35 0.557 Interactive Effects 132 0.11 0.745 132 0.63 0.430

Table 2-6. Water quality parameter mean, variance, and standard deviation measured at all sites at each retrieval/deployment date.

Date Parameter (units) Mean Range Standard Deviation 2/27- DO % 140.5 135-140 4.04 2/28/2016 DO mg/L 13.0 12.6-13.4 0.41 Salinity (psu) 15.4 10.2-23.4 5.63 Conductivity (mS/cm) 18.2 10.6-29.9 8.43 Temp °C 15.2 15.0-15.4 0.17 pH 8.5 8.48-8.57 0.04 5/7- DO % 104.6 98.9-107.3 3.87 5/8/2016 DO mg/L 8.5 8.1-8.7 0.27 Salinity (psu) 11.2 9.8-13.7 1.80 Conductivity (mS/cm) 18.8 16.6-22.7 2.78 Temp °C 22.5 22.3-22.6 0.14 pH 8.2 8.1-8.3 0.08 6/25/2016 DO % 107.2 100.7-114.6 5.72 DO mg/L 7.2 6.72-7.71 0.41 Salinity (psu) 23.4 21.0-26.3 2.22 Conductivity (mS/cm) 37.0 33.8-41.3 3.25 Temp °C 29.8 29.5-30.1 0.26 pH 8.5 8.4-8.56 0.07 8/28/2016 DO % 108.1 103.9-112.0 4.22 DO mg/L 6.8 6.6-7.1 0.24 Salinity (psu) 30.7 30.3-31.0 0.31 Conductivity (mS/cm) 47.5 47.5-47.9 0.43 Temp °C 30.8 30.6-30.9 0.13 pH 8.0 8.0-8.0 0.02 10/20/2016 DO % 99.5 97.2-101.7 1.84 DO mg/L 6.7 6.5-6.8 0.12 Salinity (psu) 31.3 30.2-31.9 0.76 Conductivity (mS/cm) 48.4 46.5-49.9 1.42 Temp °C 26.6 26.5-26.7 0.10 pH 8.1 8.1-8.1 0.02 12/16- DO % 108.9 106.1-112.3 2.99 12/17/2016 DO mg/L 9.0 8.8-9.4 0.30 Salinity (psu) 25.0 21.5-28.7 2.94 Conductivity (mS/cm) 39.2 34.4-44.3 4.07 Temp °C 17.1 16.5-17.6 0.56 pH 7.9 7.8-8.0 0.07

CHAPTER 3 CONCLUSIONS

Oyster reefs play a vital role in estuarine habitats, providing invaluable ecosystem services such as water filtration improving water clarity and quality, linking benthic-pelagic nutrient coupling, buffering coastlines from wave action and erosion, and supporting a highly abundant and diverse ecosystem with both intrinsic and economic values (Dame et al. 1989, Coen et al. 2007, Beck et al. 2011, Grabowski et al. 2012). Up to 85% of oyster reefs have been lost globally due to overharvest, pollution and reduced water quality, sedimentation, poor recruitment, loss of reef stability, and more (MacKenzie Jr. et al. 1997, Kirby 2004, Beck et al. 2009, Beck et al.

2011, Seavey et al. 2011). There is much interest in restoration of these degraded systems, however there is a lack of understanding of project outcomes. Measuring the ecological function of restoration projects is critical to informing future restoration efforts and funds efficiently.

The present study finds that the benthic macroinvertebrate and associated nekton communities of newly restored oyster sites are quickly colonized and decapod crustacean species richness, abundance, and biodiversity were comparable to nearby natural reefs in 6 - 18 months. The first hypothesis of this study was that species richness, abundance, and biodiversity would be lower on restored sites compared to natural sites during the post-restoration study period. There is little support for this hypothesis since richness, abundance, and biodiversity were not statistically different on newly restored and natural areas during all but one time periods sampled. The exception was a single sampling period when site age was significant when comparing species richness. The second hypothesis of this study, that decapod crustaceans will

recruit differently to collectors of different substrate types, has some support but results were not definitive. Substrate had a statistically significant effect on species abundance during two out of six sampling periods, a significant effect on species richness during one out of six sampling periods, and a borderline significant effect on biodiversity during one out of six sampling periods. During the majority of sampling periods, substrate does not make a different to richness, abundance, or biodiversity. While this warrants consideration in future studies, overall fossilized shell substrate recruited similarly abundant and diverse decapod crustacean assemblages relative to green shell in the present study. Finally, this study hypothesizes that parasite levels would be lower in xanthid crab populations at newly restored sites compared with natural areas. No support for this hypothesis was present and newly restored sites had levels of parasite prevalence similar to natural areas. However, substrate was significant to parasite prevalence during two of six sampling periods in E. depressus and one of six sampling periods in P. herbstii.

The global loss of oyster reef habitat has led to an appreciation of services provided by these ecosystem engineers and interest in restoring reefs to reestablish these important functions. Holistic management of oyster habitat will provide multiple ecosystem functions beyond direct oyster fisheries commodities (Brumbaugh et al.

2006, Stokes et al. 2012). While restoration techniques vary depending on the physical characteristics of the area to be restored, desired restoration, and funds available, monitoring is needed to identify both successful projects and causes of failure (Stokes et al. 2012). Many ongoing projects monitor oyster recovery but do not include additional ecosystem measures that better inform success of non-fisheries service

restoration (Hadley et al. 2010). The results of this study indicate that the restoration practices and materials utilized at these Apalachicola Bay study sites create oyster reef areas capable of quickly gaining ecosystem function similar to nearby natural areas.

Future studies may consider additional organism groups of the oyster reef food web, lengthier time spans to monitor long-term outcomes, and control for fisheries and disturbance impacts. Additional metrics examining water quality impacts such as chlorophyll-a in the water column, benthic pelagic coupling of restored sites, and wave attenuation on nearby shorelines would also be useful to inform future restoration efforts.

APPENDIX SUPPLEMENTARY MATERIALS

Table A-1. Complete list of decapod crustaceans found in collectors deployed in Apalachicola Bay sites over entire study period.

Species list Common Name Family Acetes americanus Sergestid shrimp Sergestidae Alpheus heterochaelis Bigclaw snapping shrimp Alpheidae Alpheus normanni Green snapping shrimp Alpheidae Callinectes sapidus Blue crab Portunidae Clibanarius vittatus Thinstripe hermit crab Diogenidae Dyspanopeus texanus Texas mud crab Panopeidae Eurypanopeus depressus Flatback mud crab Panopeidae Hexapanopeus angustifrons Smooth mud crab Panopeidae Hexapanopeus spp Panopeidae Latreutes fucorum Slender sargassum shrimp Hippolytidae Latreutes parvulus Sargassum shrimp Hippolytidae Leander tenuicornis Brown grass shrimp Palaemonidae Libinia dubia Longnose spider crab Epialtidae Lysmata wurdemanni Peppermint shrimp Lysmatidae Menippe mercenaria Florida stone crab Menippidae Pagurus longicarpus Longclawed hermit crab Paguroidae Palaemon floridanus Florida grass shrimp Palaemonidae Palaemonetes intermedius Brackish grass shrimp Palaemonidae Palaemonetes pugio Daggerblade grass shrimp Palaemonidae Palaemonetes vulgaris Marsh grass shrimp Palaemonidae Panopeus herbstii Atlantic mud crab Panopeidae Periclimenes americanus American grass shrimp Palaemonidae Periclimenes longicaudatus Longtail grass shrimp Palaemonidae Petrolisthes armatus Green porcelain crab Porcellanidae Pilumnus sayi Spineback hairy crab Pilumnidae

Table A-2. Species abundance. The total abundance of all decapod crustaceans found in each collector during each sampling period.

Collector Site Substrate No Time 1 Time 2 Time 3 Time 4 Time 5 Time 6 Platform Bar Green 1 20 128 221 610 642 120 Green 2 11 113 250 372 642 108 Green 3 30 109 182 693 879 153 Fossilized 1 17 123 221 393 566 219 Fossilized 2 8 87 176 686 629 121 Fossilized 3 7 101 187 522 560 105 Cat Point Green 1 45 188 267 383 488 167 Green 2 16 241 402 538 636 102 Green 3 28 267 359 552 412 136 Fossilized 1 30 196 240 458 297 136 Fossilized 2 13 195 248 349 532 165 Fossilized 3 29 164 241 419 439 116 Bulkhead Bar Green 1 18 93 231 672 776 100 Green 2 24 79 282 544 814 94 Green 3 22 82 237 367 384 103 Fossilized 1 22 67 177 502 504 59 Fossilized 2 21 77 230 610 608 102 Fossilized 3 30 75 196 420 435 115 Hotel Bar Green 1 1 102 471 521 186 Green 2 15 113 420 524 178 Green 3 12 88 590 766 133 Fossilized 1 4 89 455 325 164 Fossilized 2 15 147 530 471 210 Fossilized 3 16 124 526 561 132

Table A-3. Species richness. The total number of individual species found in each collector by site during each sampling period.

Collector Site Substrate No Time 1 Time 2 Time 3 Time 4 Time 5 Time 6 Platform Green 1 5 5 12 10 13 13 Bar Green 2 4 7 10 10 16 12 Green 3 4 8 6 6 14 15 Fossilized 1 3 8 8 11 12 15 Fossilized 2 4 9 8 13 12 12 Fossilized 3 3 8 9 10 13 11 Cat Point Green 1 3 5 8 10 14 13 Green 2 4 5 6 10 14 16 Green 3 4 6 8 9 13 11 Fossilized 1 2 7 6 8 16 10 Fossilized 2 4 6 7 6 17 12 Fossilized 3 3 7 4 8 12 10 Bulkhead Green 1 3 5 7 11 15 9 Bar Green 2 3 4 12 7 15 11 Green 3 5 8 8 10 14 13 Fossilized 1 4 4 8 10 10 7 Fossilized 2 4 4 8 7 14 11 Fossilized 3 7 4 8 7 15 10 Hotel Bar Green 1 2 4 10 15 11 Green 2 3 7 10 13 12 Green 3 2 5 7 13 12 Fossilized 1 2 7 13 13 11 Fossilized 2 1 3 11 13 13 Fossilized 3 1 5 7 15 10

Table A-4. Shannon-Weiner Biodiversity Index Calculations by site, collector number, and shell type throughout the sampling period.

Collector Site Substrate No Time 1 Time 2 Time 3 Time 4 Time 5 Time 6 Platform Green 1 1.36 0.67 0.75 0.65 1.61 1.76 Bar Green 2 1.12 0.81 0.61 0.66 1.79 1.97 Green 3 1.12 1.17 0.66 0.35 1.77 2.11 Fossilized 1 1.07 0.75 0.58 0.65 1.80 1.81 Fossilized 2 1.15 0.92 0.78 0.51 1.77 1.94 Fossilized 3 1.00 0.95 0.72 0.45 1.74 1.80 Cat Point Green 1 0.40 0.28 0.60 0.38 1.04 1.01 Green 2 0.63 0.39 0.39 0.36 0.86 1.59 Green 3 0.63 0.41 0.41 0.34 1.04 1.30 Fossilized 1 0.15 0.37 0.59 0.33 0.89 0.96 Fossilized 2 0.75 0.40 0.41 0.27 0.83 0.97 Fossilized 3 0.50 0.56 0.29 0.26 0.78 1.26 Bulkhead Green 1 0.73 0.35 0.60 0.28 0.79 1.11 Bar Green 2 0.60 0.25 0.78 0.13 0.91 1.02 Green 3 0.68 0.61 0.82 0.45 1.02 1.41 Fossilized 1 0.69 0.33 0.65 0.23 0.75 0.93 Fossilized 2 0.80 0.49 0.52 0.16 0.88 1.14 Fossilized 3 1.32 0.21 0.77 0.23 1.09 1.02 Hotel Bar Green 1 0.33 0.47 0.38 0.86 1.21 Green 2 0.82 0.62 0.33 1.13 1.11 Green 3 0.27 0.47 0.17 0.77 1.25 Fossilized 1 0.69 0.43 0.39 1.19 1.06 Fossilized 2 0.00 0.27 0.22 0.93 0.97 Fossilized 3 0.00 0.35 0.17 1.10 1.10

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BIOGRAPHICAL SKETCH

Megan grew up in Pennsburg, Pennsylvania as the daughter of two teachers who instilled her with a love of learning. For her undergraduate education, she attended

Wesleyan University where she majored in biology. During her undergraduate career she studied abroad in Australia at the University of Brisbane, where she was able to take field courses on the Great Barrier Reef. She also completed a semester at the

Williams-Mystic Maritime Studies Program where she continued her marine ecology studies and was introduced to the concept of non-native species and containerization in a global world.

After receiving her bachelor’s degree, Megan accepted a position in Eastpoint,

Florida as a water quality technician at the Apalachicola National Estuarine Research

Reserve. After she moved to Florida she learned to pilot boats, SCUBA dive, and coordinate field campaigns for coastal ecology studies. She then began working as the

ANERR site coordinator for the NOAA Environmental Cooperative Science Center, whose purpose was to promote coastal research and understanding with the goals of training the next generation of NOAA scientists and conduct research that could be used by coastal decision makers. As coordinator, Megan worked with graduate students and project PIs with their research around the Gulf Coast. Megan then became the nutrient and chlorophyll-a monitoring program lead the Reserve, before most recently becoming a biologist at ANERR. The combination of the crustacean work she did at Williams-Mystic and working on the Apalachicola Bay, the oyster capital of

Florida, led Megan to her interest in the crustacean communities that are the focus of her master’s research.