TANDEM RESTORATION OF DIADEMA ANTILLARUM AND ACROPORA CERVICORNIS FOR ENHANCED REEF RECOVERY

Transcription

1 TANDEM RESTORATION OF DIADEMA ANTILLARUM AND ACROPORA CERVICORNIS FOR ENHANCED REEF RECOVERY By KAYLA J. RIPPLE 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

2 2017 Kayla J. Ripple 2

3 ACKNOWLEDGMENTS I would like to thank all of the organizations and agencies who supported this project. Without funding and resources provided by the Florida Fish and Wildlife Conservation Commission, Coral Restoration Foundation, University of Florida, and staff at the Florida Aquarium s Center for Conservation, this project would not have been possible. I would like to thank the members of my committee, Dr. Mark Flint, Dr. Don Behringer, and Dr. Scott Winters. They have provided me with the resources and support necessary to carry out this project. Their diverse backgrounds and joint caliber taught me the process and inner workings of the scientific process. I would like to thank most of all my Committee Chair, Dr. Mark Flint, for his patience, continuous check-ins, and calming tactics to get me through every moment of this degree. Mark has taught me the fun side of science, how to tell a story through my research, and imparted on me his care and kindness with a side of wit. I would also like to thank my family and close friends for their unwavering support. To my mother and father, who have supported my dreams and ambitions no matter how crazy. To my brother and sister who always remind me life should be taken a little less seriously. And to Jessica and Jana, who kept me sane, listened to my rants, and rejoiced in my eureka moments and happy times. This was not just a degree or a research project, it was a life lesson and transforming journey. 3

4 TABLE OF CONTENTS page ACKNOWLEDGMENTS...3 LIST OF TABLES...6 LIST OF FIGURES...7 ABSTRACT...8 CHAPTER 1 A REVIEW OF THE BIOLOGY, ECOLOGY, AND JUSTIFICATION FOR CORAL- URCHIN RESTORATION...10 Introduction...10 Biology and Ecology of the Long-Spined Sea Urchin...11 The Diadema Die-Off...14 Rationale for Coral-Urchin Restoration...20 General Hypotheses and Objectives HEALTH ASSESSMENTS OF EX SITU EXPERIMENTATION OF DIADEMA ANTILLARUM AND TRIPNUESTES VENTRICOSUS...26 Introduction...26 Methods...27 Results...29 Discussion UNDERSTANDING THE RELATIONSHIP BETWEEN ACROPORA CERVICORNIS CORAL DENSITY AND JUVENILE DIADEMA ANTILLARUM SURVIVORSHIP...36 Introduction...36 Methods...39 Results...43 Discussion PREDICTING SUCCESSFUL DIADEMA ANITLLARUM RELOCATION SITES IN THE FLORIDA KEYS...54 Introduction...54 Methods...55 Results...57 Discussion DISCUSSION

5 Diadema antillarum Health and Behavior Parameters Ex Situ: Applications for In Situ Monitoring and Relocation Efforts...67 Acropora cervicornis Densities to Promote Diadema antillarum Retention...70 Applications for Management...71 Future Work...74 APPENDIX A: Coral Clusters Present for Urchin Relocations from LIST OF REFERENCES...80 BIOGRAPHICAL SKETCH

6 LIST OF TABLES Table page 2-1 Urchin Health Day Urchin Health Day Shelter Seeking Behavior in D. antillarum and T. Ventricosus Averages of urchins present within cages during experiment Proportion of Clusters Present for Each Category for Years A-1 Clusters Present for Each Category of Cluster-Biomass at Restoration Sites in A-2 Clusters Present for Each Category of Cluster-Biomass at Restoration Sites in A-3 Clusters Present for Each Category of Cluster-Biomass at Restoration Sites in A-4 Clusters Present for Each Category of Cluster-Biomass at Restoration Sites in

7 LIST OF FIGURES Figure page 1-1 Urchin grazing effects The spread of Diadema antillarum mortality sightings Grid and design of experimental pools Urchin abundances across times spent in pools Layout of coral density experimental grid Cage Design Average urchins present for time across all treatments Urchins present over time for control cages Urchins present over time for 10 corals Urchins present over time for 25 corals Urchins present over time for 40 corals Urchins present over time for 55 corals Diadema Retention Model for A. cervicornis restoration sites A. cervicornis restoration sites identified in the Florida Keys for preliminary D. antillarum relocations in A. cervicornis restoration sites identified in the Florida Keys for preliminary D. antillarum relocations in A. cervicornis restoration sites identified in the Florida Keys for preliminary D. antillarum relocations in A. cervicornis restoration sites identified in the Florida Keys for preliminary D. antillarum relocations in

8 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 TANDEM RESTORATION OF DIADEMA ANTILLARUM AND ACROPORA CERVICORNIS FOR ENHANCED REEF RECOVERY Chair: Mark Flint Major: Fisheries and Aquatic Science By Kayla J. Ripple May 2017 The long-spined sea urchin, Diadema antillarum, is often considered a keystone species on coral reefs throughout the Caribbean. In 1983, a disease epidemic resulted in the mass mortality of D. antillarum, correlating with a phase shift of coral to macroalgae dominated reefs. Since the epidemic, few reefs have shown signs of urchin recovery. Efforts to reverse reefs to coral dominance have taken place in the form of coral population enhancement programs and urchin relocation efforts, but the two have yet to be combined in earnest. This study aimed to understand multi-species reef-restoration techniques, specifically that of coral-urchin restoration, to promote recovery of both urchin and coral populations in tandem, and promote reef health. To achieve this, experiments were conducted to assess the general hypotheses that: (1) relocation of D. antillarum to coral restoration sites does not affect health of the urchins; (2) with increasing Acropora cervicornis density, D. antillarum retention within A. cervicornis clusters increases; and (3) with multiple restoration efforts taking place, suitable habitat currently exists to support D. antillarum relocation. Urchin health was addressed through the adoption of protocols outlined by Francis-Floyd et al. (in press) for studies in situ, and showed there were no changes to urchin health when relocated from one site to another. Coral density manipulation experiments showed that there 8

9 appears to be an apparent threshold for coral density where certain cluster-biomass is needed to support >75% urchin retention within A. cervicornis coral clusters. A Diadema Retention Model was developed to predict cluster-biomass needed to promote urchin retention, and sites were identified that to begin preliminary coral-urchin restoration efforts in as soon as 2017 and incrementally until

10 CHAPTER 1 A REVIEW OF THE BIOLOGY, ECOLOGY, AND JUSTIFICATION FOR CORAL-URCHIN RESTORATION Introduction Coral reefs are an ecosystem of great importance both biologically and economically. Ecosystem services coral reefs provide globally, have been estimated at $9.9 trillion (Costanza et al., 2014). Often referred to as the rainforests of the sea, the system fosters great biodiversity with a multitude of species filling various niches. Populations of different species act in various roles within the community to promote checks and balances within the system that contribute to overall health and resiliency of the biome (Tuya et al., 2004; Bellwood 2004; Graham and Nash, 2013). Large reef-building corals are essential to provide continuous habitat and shelter over time for the diverse group of organisms. In the Florida Keys and Caribbean, the branching corals Acropora cervicornis and A. palmata, were once the most abundant of reef-builders, providing three-dimensional habitat and shelter for reef fishes and invertebrates. The intricate morphology of the thickets they form, complimented with the morphology of massive mounds of boulder and pillar corals, provides opportunity for a wide range of biodiversity (Gratwicke and Speight, 2005). Current reef decline debates target climate change as the present critical stressor affecting reefs today. However, fishing pressures are considered to be the source of early reef decline (Hay, 1984; Jackson, 2001). In the 1970s, hard coral cover, specifically A. cervicornis and A. palmata, drastically decreased due to an outbreak of coral disease (Aronson and Precht, 2001; Miller et al., 2002). Periods of coral bleaching and consistently poor water quality compounded the issues, as greater coral numbers began to decline. The long-spined sea urchin, Diadema antillarum, is often considered a keystone species on coral reefs throughout the Caribbean (Lessios, 1988, Knowlton, 2001; Tuya et al., 2004), 10

11 contributing to the control of macroalgae abundances and allowing the growth of Caribbean scleractinian coral species. A mass mortality event of D. antillarum in the 1980 s is thought to have contributed to the ultimate demise of Caribbean coral reefs (Carpenter, 1990; Knowlton, 1992; Hughes, 1994). The following review provides greater detail for the biology of D. antillarum, factors contributing to reef decline, management tools for recovery, and rationale to promote a holistic approach to reef recovery by combining the restoration of both coral and urchins. Biology and Ecology of the Long-Spined Sea Urchin D. antillarum are distinguished from other herbivorous urchins by their dark bodies, long, ubiquitous spines, and iridescent receptors that allow them to detect differences in light and dark (Randall et al., 1964). D. antillarum are light sensitive (Millot, 1953, 1954) and exhibit diel activity patterns preferring dark, sheltered habitats and exhibiting nocturnal behavior to lessen predation pressures (Thornton, 1956; Randall et al., 1964; Ogden et al., 1973). A number of anecdotal observations describe the urchins spines as exhibiting a diverse array of color patterns, with juveniles often possessing black and white variegated spines and adults with all black, or gray spines. Spines are used as protection, and respond by vigorous movement when exposed to higher light intensities and potential threats such as predators (Millot, 1954). D. antillarum can be found in multiple tropical habitats including Thalassia testudinum beds, mangrove propagation roots, sand flats, and, most notably, on coral reefs and inshore rubble habitats (Randall et al., 1964). In years leading up to the catastrophic die-off of D. antillarum, urchin populations were considered to be the most abundant herbivorous species on Caribbean reefs, often deleterious to recreational activities for tourists on reefs (Randall et al., 1964). At the Virgin Islands Marine Park in St. Thomas, a management plan was even enacted to remove urchins to improve the experience of visiting sightseers (Kumpf and Randall, 1961). 11

12 The density of D. antillarum has been shown to be positively correlated with reef complexity (Lee, 2006; Clemente and Hernandez, 2008; Dame, 2008; Bodmer et al., 2015). Historically, high D. antillarum densities were correlated with shallow reef areas represented by high reef complexity comprised of multiple coral species including A. cervicornis, A. palmata, Orbicella annularis, and Millepora complanta (Weil et al., 1984). Urchins even have the ability to evaluate crevice quality (Carpenter, 1984) relative to their test size to avoid predator attacks (Carpenter, 1988), where individuals often exhibit homing behavior to their selected reef crevice (Tuya at el., 2003). Consequently, high structural complexity is an important characteristic to maintain urchin populations at a reef site. D. antillarum are observed as an aggregative species (Randall et al., 1964; Bauer, 1976; Levitan, 1988), where they can exist individually, or in aggregations of up to 100 individuals (Randall et al., 1964). They are broadcast spawners expelling their gametes for fertilization within the water column (Randall et al., 1964; Bauer, 1976). Increased numbers in urchin aggregations are thought to be correlated with reproductive purposes where urchins form denser aggregations around the full moon and new moon (Bauer, 1975). Gonadal index has been positively correlated for females with the lunar cycle where egg development is highest around the new moon and decreases after the new moon (Bauer, 1975), however, spawning tendencies appear to be asynchronous where males and females have been documented as spawning across different time periods (Levitan, 1988). Population structure for multiple organisms is often determined by food availability and predation pressure (Holt, 1977; Anholt and Werner, 1995). Like many other species, the local population of D. antillarum and their habitat preferences are heavily influenced by predator abundance. With increasing predator abundance, D. antillarum populations tend to decrease, 12

13 resulting in smaller populations of urchins (Ogden, 1973; Carpenter, 1984). High predation pressure within reef sites can result in D. antillarum exhibiting high crevice fidelity where urchins remain in the same crevice during the day, leave the crevice at night to graze, and return to the same crevice before dusk (Carpenter, 1984; Tuya et al., 2004). This homing behavior may be an indirect result of the individual to avoid predation pressure and increase their chances of survival (Carpenter, 1984). When predation pressure is reduced in an area, D. antillarum are less likely to exhibit nocturnal behavior and crevice fidelity and increase total grazing time (Frike, 1974). This type of behavior has been observed across many Caribbean reefs including reefs of Jamaica and St. Croix (Miller et al., 2003; Carpenter and Edmunds, 2006), where declines in D. antillarum predators have been recorded (Hughes, 1994). D. antillarum, in concert with other herbivorous fish species, consume macroalgae and expose rock suitable for the formation of crustose coralline algae (CCA) needed to promote recruitment of reef fish and invertebrate species whose larval metamorphosis and settlement are queued by CCA (Macintyre et al., 2005; Carpenter and Edmunds, 2006). Their intense grazing effects can be seen by the clear delineation of clean reef substrate where urchins have grazed and adjacent macroalgae covered substrate where urchins have not grazed (Figure 1.1, Idjadi et al., 2010). D. antillarum grazing effects can also be seen from satellite images as halo s around patch reefs (Ogden et al., 1973). D. antillarum were once the most abundant reef herbivore in the Caribbean (Lessios et al., 2001). Their large populations were thought to be attributed to an already declining reef fish trajectory of both predatory fish which decreased numbers of urchins, and herbivorous fishes which competed for macroalgae resources on the reef (Hay, 1984). Reef fish populations were reported as declining before the 20 th century due in part to the development of new commercial 13

14 and recreational fishing techniques (Jackson, 2001). While sportfish were the main target, multiple herbivorous reef species also dwindled (Paddock et al., 2009). As fish populations began falling, D. antillarum faced less predation pressure and increased food resources, resulting in a Type II predator-prey functional response where the urchins took over the role as predominant herbivore on the reef (Hay, 1984; Carpenter, 1984; McClanahan and Muthiga, 1988; McClanahan et al., 1996). Large populations of D. antillarum kept macroalgal abundances low, allowing corals to thrive even in the absence of normal herbivorous fish abundances (Ogden 1973; Sammarco, 1980; Carpenter, 1981; Sammarco, 1982). This changed with the mass die-off in 1983, and the decline has not yet been corrected along the Caribbean reef tract. The Diadema Die-Off In 1983, a plague swept through D. antillarum populations resulting in the largest documented near mass extinction event recorded for a marine species (Figure 1.2; Lessios, 1988). In one year, the plague spread from Panama throughout the Caribbean over an area of approximately 3.5 million km 2 (Lessios, 1988). Up to 93% of the long-spined urchins perished in most areas of the Caribbean (Lessios, 1988), and there appeared to be no populations untouched by the plague (Lessios, 2016). The cause of the mortality remains unknown, but is suspected to be a species-specific water-borne pathogen, whose transmission was facilitated by currents throughout the Caribbean (Lessios, 1988). In 1987, Bauer and Agerter (1987) received infected D. antillarum where they were able to isolate and culture Clostridium spp. bacteria. When this bacterium was injected into healthy D. antillarum, results were very similar to the disease symptoms that plagued wild populations, possibly implying the bacteria as an agent for the 1983 mass mortality. However, Bauer and Agerter caution highly that these findings did not fulfill Koch s Postulates thus do not indisputably demonstrate cause of the mortality event. 14

15 Before the D. antillarum epidemic, urchin densities were recorded in high abundance across various areas of the Caribbean. Urchin densities varied across regions from 3-64 individuals/m 2 (Bauer, 1980). In the Florida Keys, densities were lower than other areas of the Caribbean around 4-5 individuals/m 2 (Kier and Grant, 1965; Bauer, 1976, 1980), but higher than present densities at maximums of 0.33 individuals/m 2 (Chiappone et al., 2009). It is possible these high densities were a result of overfishing of predatory and competitive fish species (Jackson, 2001), however, analysis of the mtdna region of D. antillarum in the Caribbean, Atlantic, and of D. mexicanum in the Pacific, suggests that pre-mortality populations were likely high in abundance even 100,000 years ago (Lessios et al., 2001). It is unclear from the literature, whether these high population densities contributed to the mass dissemination of the fatal pathogen in D. antillarum populations, or if similar epidemics occurred in evolutionary history of the population (Lessios, 1988). Regardless of cause, the devastating reduction of D. antillarum populations facilitated a spike in macroalgae abundance across Caribbean reefs, which is strongly correlated with mass mortality of already failing Caribbean Acroporid populations (Hughes, 1994; Jackson, 2001). Today, D. antillarum populations are 25 times less dense than populations on reefs before the mass mortality event (Hughes et al., 2010). This drastic decline in herbivore abundance, coupled with an increase in macroalgae abundance, signified a possible phase shift on Caribbean reefs from a coral dominated to macroalgae dominated state (Carpenter, 1990; Knowlton, 1992; Hughes, 1994). There were no populations left untouched by the die-off (Lessios, 2016), but recent improvements for populations in Jamaica and Honduras show a decrease in macroalgae and increase in scleractinian coral recruits (Edmunds and Carpenter, 2010, Bodmer et al., 2015). 15

16 This evidence suggests that D. antillarum are key in promoting health of the reef as no other species has been able to fill the niche of D. antillarum for removing macroalgae. It is increasingly accepted that coral reefs throughout the Caribbean have undergone phase shifts from a coral-dominated state to a macroalgae-dominated state (Knowlton, 1992; Hughes, 1994; Jackson, 2001; Bellwood et al., 2004; Maliao et al., 2008). Environmental conditions can vary overtime, and stochastic events such as hurricanes or disease outbreaks can cause direct effects to a system, often affecting the populations of an organism (Scheffer et al., 2001). Factors causing these shifts can allow for simple or difficult reversal based on the type of perturbation or event that caused the change (Beisner et al., 2003). If the system only has one stable state it exists in, it is expected to settle back to original parameters after the event, however, if the system has multiple stable states, it may shift to an alternative stable state (Scheffer et al., 2001). The prolonged period of a macroalgae dominated state on most Caribbean reefs has led some scientists to believe reefs now exist in an alternative stable state, where effects of hysteresis may require drastic management measures to overcome its effects and revert back to coral dominant states (Mumby et al., 2007; Hughes et al., 2010; Fung et al., 2011; Graham et al., 2013). As coral cover declines and macroalgae grows, positive feedback loops keep reefs in an algae-dominant state (Scheffer et al., 2001; Norstrom et al., 2009; Hoey and Bellwood, 2011). Macroalgae causes direct interference with coral growth, often times overshadowing slower growing corals and secreting allelopathic chemicals (Lirman, 2001; McCook et al., 2001). Some algae species competing for space on substrate create unsuitable habitat for coral recruitment, smothering new coral recruits, or providing no space for fragments to settle and cement themselves to the substrate (Lirman, 2001). Further, significantly smaller coral populations, 16

17 suffer from allee effects, decreasing the opportunity for coral species to successfully spawn and to expand populations (Knowlton, 1992; Williams et al., 2008). Although declines in coral cover have slowed since the 1980 s, recovery back to a coraldominated state is not recorded for most regions of the Caribbean and in the Florida Keys (Gardner et al., 2003). It is thought that the resilience of these ecosystems has been weakened by the multiple stressors they face, making them more susceptible to pulse perturbations and resulting in difficulty reversing back to coral-dominated states (Bellwood et al., 2004). Models and empirical studies provide considerable evidence that the Florida Keys and areas of the Caribbean have currently shifted to this alternate state as a direct result of the Diadema die-off (Mumby et al., 2007; Lessios, 2016). A decrease in reef-building corals and increase in reef erosion, lessens reef rugosity and complexity (Alvarez-Filip et al., 2009). This flattening of reefs, along with other anthropogenic stressors such as overfishing, has been linked to a decrease in economically important fish species and overall deterioration of the biodiverse food web that once existed in coral-dominant ecosystems (Alvarez-Filip, 2009; Dixson et al., 2014). This current state of macroalgae dominated substrate offers little to no refuge for D. antillarum individuals, inhibiting the recovery of populations (Bodmer et al., 2015; Roger and Lorenzen, 2016). These observations may not encourage optimism for reversion of reefs back to coral dominance. Demonstrated shifts in other ecosystems however, prove it can be done. Urchin barrens of kelp forests in California provide a prime example. Sea otters control populations of the sea urchin Strongylocentrotus droebachiensis which graze on kelp holdfasts releasing floating kelp into the water column where it then drifts away (Estes and Palmisano, 1974). In the 1800 s, sea otter populations were decimated by fisherman harvesting sea otters for fur-trade, 17

18 and an explosion of the sea urchin population occurred, resulting in near decimation of kelp forests and a shift from kelp-dominated state to a barren urchin dominated state, or urchin barrens (1974). Regulations put in place to assist recovery of sea otter populations, and also increase farming pressure on sea urchin populations has resulted in an increase in the kelp beds to begin to come back where the distribution of kelp forest habitats has now expanded to onethird of the habitat it once was (Estes and Palmisano, 1974; Filbee-Dexter and Scheibling, 2014). Evidence from this shift, show that with proper regulation and conservation of natural resources, ecosystems are able to revert to a more biodiverse state to support original productivity and function of the ecosystem. Overall, D. antillarum populations have shown little recovery since their ecological demise in the 1980 s with the exception of a few reefs (Hughes, 2010; Lessios, 2016). Their inability to recover to pre-mortality densities is likely attributed to a multitude of compounding factors. Five main theories exist for the poor recovery of D. antillarum (Bodmer et al., 2015) including: (1) increased competition from vertebrate reef herbivores, (2) suppressed recruitment resulting from natural asynchronous spawning and allee effects, (3) predation pressure driving high mortality, and (4) loss of reef structural complexity removing microhabitat provision (5) ecological interactions with heterospecifc echinoids. D. antillarum have historically interacted with other reef echinoid species either by complementing grazing activities or competing with urchins for macroalgae resources. The specifically includes the East-Indian sea egg, Tripnuestes ventricosus (Haley and Solandt, 2001), and the rock-boring urchin, Echinometra viridis (Shulman, 1990). The sea urchin, T. ventricosus, feeds on larger macroalgae on reefs, that allows for juvenile macroalgae to recruit to reef substrate (Haley and Solandt, 2001). This type of algae is the preferred food source for D. 18

19 antillarum (Haley and Solandt, 2001; Bechtel et al., 2006). Because of great increases in macroalgae, it is possible T. ventricosus cannot keep up with current levels of macroalgae, leaving an inadequate food sources for D. antillarum (Liddel and Ohlhorst, 1986; Carpenter, 2005). E. viridis and D. antillarum have been observed as aggressive towards one another, and can compete for space on a reef (Shulman, 1990). However, there have been no population spikes in E. viridis since the Diadema die-off, and this is not expected to impact local D. antillarum population recovery (McClanahan, 1999). D. antillarum and herbivorous reef fishes are essential for the removal of macroalgae from reefs to maintain hard coral cover and recruitment. It is suspected that recovery may also be inhibited by the competition between herbivorous fishes and D. antillarum for food (Robertson, 1991). However, the two can exist harmoniously, and grazing pressure exerted by both are needed together for maximum macroalgae reduction (Carpenter, 1986). Moreover, studies have even shown that large populations of D. antillarum can drive down local herbivorous fish populations (Hay and Taylor, 1985), suggesting urchin populations are able to stabilize and can be unaffected by normal abundances of grazers in the area. Despite high fecundity in females who can produce up to one million eggs in a single spawning event (Levitan, 1989), recruitment of D. antillarum still appears low (Lessios, 2010) suggesting possible recruitment limitation may also be an issue (Karlson and Levitan, 1989). Sparse adult populations may contribute to allee effects, resulting in low fertilization success and the inability to produce enough larvae to overcome typically high mortality in the larval stage (Levitan, 1995). In the Florida Keys, current D. antillarum distributions are thought to have a <1% fertilization success rates as opposed to >96% before 1983 (Feehan et al., 2016). 19

20 D. antillarum have >15 known predators (Randall et al., 1964). Juvenile urchins (20-30mm maximum test diameter) are prone to increased predation pressure at reef sites, but after reaching sizes >40mm, D. antillarum are still vulnerable, but less prone to predator attacks (Clemente et al., 2007). Harborne et al. (2008) found that increased fish biomass of species known to feed on D. antillarum is linked with smaller urchin populations inside marine protected areas on Bahamian reefs, demonstrating how predation pressure can impact urchin abundances. Greatest recovery of D. antillarum populations have been documented in sheltered lagoonal areas and back reef habitat where these populations were also once abundant (Miller et al., 2003, Debrot and Nagelkerken, 2006; Steiner and Williams, 2006; Vermeij et al., 2010). This is likely a result of the increased habitat complexity offered in these areas (Rogers & Lorenzen, 2016). Greater habitat complexity is linked to larger D. antillarum populations (Clemente and Hernandez, 2008). Increasing habitat structure at a reef site can also increase retention of relocated D. antillarum (Dame, 2008). With decreased hard coral populations, there is a documented decrease in reef structural complexity throughout Florida and the Caribbean (Alvarez-Filip et al., 2009). It is possible, predation pressure, coupled with lack of structural complexity is a major rate-limiting step in recovery of D. antillarum populations (Bodmer et al., 2015). Rationale for Coral-Urchin Restoration The benefits of healthy D. antillarum populations are abundant. Their ability to reduce macroalgae abundance at reef sites is important for promoting the dominance of coral cover, which in turn, promotes reef biodiversity. Recovery of D. antillarum is potentially the largest driving factor in coral reef recovery (Edmunds & Carpenter, 2001). Scarids and other herbivorous fishes are responsible for only a proportion of grazing pressure on reefs (Mumby, 2006). Increased herbivory from Scarids may even cause harmful effects to existing coral 20

21 populations by inhibiting the healing of hard corals from parrotfish scrapes in areas with poor water quality and eutrophication on reefs from agricultural runoff (Zaneveld et al., 2016). D. antillarum proves to be beneficial for removing fast-growing macroalgae enhancing coral growth on reefs (Hernandez et al., 2008). Increasing abundance of D. antillarum at reef sites may break the positive feedback loop by decreasing the abundance of macroalgae at reef sites (Chiappone et al., 2001; Burdick, 2008; Nedimyer and Moe, 2011) increasing coral cover (Edmunds and Carpenter, 2001; Idjadi et al., 2010), which in turn can increase reef complexity necessary to increase D. antillarum abundances (Sammarco, 1982; Clemente and Hernandez, 2008; Bodmer et al., 2015). Since their die-off in the 1980 s, D. antillarum population enhancement efforts have taken place in the form of preliminary relocation trials at reef sites (Burdick, 2008; Nedimyer and Moe, 2011), development of comprehensive strategies for relocation (Hunt and Sharp, 2014), and workshops to promote ex situ production for relocation efforts (Diadema Workshop, 2017). Preliminary relocation efforts were successful in reducing macroalgae cover at reef sites, but these population numbers ultimately disappeared (Burdick, 2008; Nedimyer and Moe, 2011). Low urchin retention rates at relocation sites were attributed to the low complexity and predation pressure. Coral restoration efforts may provide a solution for low reef complexity. In 2012, over 60 programs were in place to restore degraded reef sites at 14 countries (Young et al., 2012). These programs focus on several threatened species of corals, including Acropora species which, grow quickly in nursery programs. As coral restoration programs become well established, best practices are better understood, as well as management strategies for reef recovery. The NOAA Acropora Recovery Plan provides an outline of criteria needed for the species to recover over the 21

22 long-term (2015). It is clear that without addressing environmental concerns such as climate change, overfishing, poor water quality, and increased macroalgae abundance, coral reefs cannot recover to their full potential. While not historically associated with thickets of A. cervicornis, in the absence of more favorable coral species like A. palmata, Orbicella annularis, and Millepora complanta (Weil et al., 1984), A. cervicornis, when configured in the right densities, may act as an alternate habitat for D. antillarum. Coupling the relocation of D. antillarum individuals with already existing A. cervicornis restoration sites could possibly lead to enhanced reef recovery, where A. cervicornis thickets can provide shelter to urchins and urchins can reduce macroalgae to increase coral growth and health at the restoration site. General Hypotheses and Objectives This study aims to support the necessary steps to generate an effective coral-urchin restoration strategy for coral reefs in the Florida Keys. To achieve this, experiments were conducted to assess the following general hypotheses and objectives for each: Hypothesis 1: Relocation of D. antillarum to coral restoration sites does not affect health of the urchins and its objectives (a) determine the parameters necessary to promote D. antillarum health in an ex situ environment; (b) understand potential habitat usage and behavior of D. antillarum and T. ventricosus in a controlled environment when compared to expected normal behavior; (c) understand proper health metrics for monitoring D. antillarum health for in situ relocations and studies. Hypothesis 2: With increasing A. cervicornis density, D. antillarum retention within coral clusters increases and its objective (a) determine the coverage of A. cervicornis coral needed to promote urchin retention over time when urchins were relocated to coral clusters. Hypothesis 3: With multiple restoration efforts taking place, suitable habitat currently exists to support D. antillarum relocation and its objectives (a) predict and identify current 22

23 restoration efforts that may facilitate the recovery of urchins by selecting suitable relocation to reefs with appropriate coral infrastructure (coverage); (b) create a tool for managers charged with implementing coral restoration programs in conjunction with D. antillarum recovery programs. For the first hypothesis, experiments were conducted at The Florida Aquarium s Center for Conservation in Apollo Beach, FL. Habitat preference and health were assessed during these experiments. Objectives 1a and 1b were partially achieved where we observed parameters that may potentially lead to demised urchin health ex situ, and observed abnormal shelter-seeking behavior. Results from urchin behavior were inconclusive however, as abnormal behavior may have been attributed to urchin health. Objective 1c was achieved through the adoption of health assessment protocols outlined by Francis-Floyd et al. (in press), for aquacultured urchin release onto reef sites, and demonstrated applicability in the field during in situ trials. Caging experiments in the Coral Restoration Foundation coral nursery offshore Tavernier, FL tested the second hypothesis. Objective 2a was achieved by manipulating coral densities from low to high (0-55 corals/0.25m 2 ). A trend was discovered for apparent thresholds of A. cervicornis cluster-biomass that can promote stable urchin retention rates >75%, and provide a baseline for coral-urchin restoration projects to test in future work. Coral densities and their associated urchin retention rates gathered from coral density manipulations were used to generate a Diadema Retention Model to achieve Objective 3b, which was used to predict coral densities needed to promote % urchin retention when urchins were relocated to A. cervicornis clusters to achieve Objective 3a. The model allowed the identification of coral clusters that currently exist on reefs of the Florida Keys from so that preliminary urchin relocation trials may be tested in future work to achieve. 23

24 Figures Figure 1-1: Urchin grazing effects. The delineation between urchin grazing zones and their boundary limits on a reef in Discovery Bay, Jamaica (After Idjadi et al., 2010). 24

25 Figure 1-2: The spread of Diadema antillarum mortality sightings. The spread of Diadema antillarum mass mortality through the Caribbean in relation to surface currents. The direction of spread was deduced from the timing of outbreaks at each locality. 1: Panama, 2: Curacao, 3: Tobago, 4: Barbados, 5: Jamaica, 6: Flower Garden Banks, 7: St. Croix, 8: St. Thomas. 9: Bermuda (After Lessios, 1988). 25

26 CHAPTER 2 HEALTH ASSESSMENTS OF EX SITU EXPERIMENTATION OF DIADEMA ANTILLARUM AND TRIPNUESTES VENTRICOSUS Introduction Current plans to contribute to the recovery of Diadema antillarum are widespread and include the collection and relocation of natural D. antillarum recruits on reefs, and aquaculture for restocking efforts. While the ex situ aquaculture provides a potential source of tens of thousands of larvae each batch, the health risk of introducing these animals to natural reefs and the effect artificial environments have on natural feeding and predator avoidance behaviors of the urchins is unknown. To evaluate viability of aquaculture raised urchins, a health assessment was developed by Francis-Floyd et al. (in press) to provide some degree of confidence that cultured sea urchins are not introducing pathogens to already compromised reef ecosystems. This work serves as an assessment tool for management agencies as efforts move forward in on-land mass production. Breeding programs that exist for population enhancement of terrestrial species have demonstrated that animals raised in captivity must undergo an imprinting stage where they learn behaviors needed for survival when released into the wild. Breeding, imprinting, and release techniques are well documented for avian species like the Andean Condor (Wallace and Temple, 1987). Sport fish restoration programs also demonstrate the need to initiate wild instincts and homing behavior for fish bred in aquaculture programs and released for re-stocking efforts (Hendricks et al., 2002). The effect aquaculture environments have on the feeding and predator avoidance behavior of sea urchins for use in population enhancement programs has not yet been comprehensively assessed. It has been anecdotally reported that long-spined sea urchins held in captivity for prolonged periods of time become habituated to their environment and actively seek food from handlers and do not seek shelter to avoid threats (Sharp and Delgado, pers. comm.). 26

27 As such, there is a concern that breeding sea urchins in artificial environments will remove their basic instincts to forage and seek shelter. Equally important, there is a risk that land-based experimentation to determine home range, feeding behavior, sheltering patterns, and threat response are not reliable with habituated subjects. Objectives of this study were to determine the parameters necessary to promote D. antillarum health in an ex situ environment, understand habitat usage and behavior of D. antillarum in a controlled environment when compared to expected wild behavior, and understand proper health metrics for monitoring D. antillarum health for in situ relocations and studies. Methods Experiments took place at The Florida Aquarium s Center for Conservation (CFC) greenhouse in Apollo Beach, FL over the winter of Tripnuestes ventricosus were cultured by Mr. Martin Moe in Islamorada, FL, and D. antillarum were caught on patch reefs off Marathon, FL by the Florida Fish and Wildlife Research Institute and obtained under the authorization of Florida Keys National Marine Sanctuary. Sixteen pools were established in a greenhouse, filled with clean seawater imported from the Gulf of Mexico, and aerated using individual bubbler systems. All pools were independent of one another to ensure separation of replicates. Water quality was assessed by The Florida Aquarium to ensure water was approximating natural sea water parameters and was free of pathogens. In twelve pools, one staghorn structure (~50cm in diameter), and one planter pot structure were placed on one end of the pool each with a food source (Gracilaria tikvahiae and Acanthophora spicifera) (Figure 2-1). Four pools were left empty as controls. Twenty adult T. ventricosus and six adult D. antillarum were available at CFC for study. Ten adult T. ventricosus and six adult D. antillarum were added separately to each pool at the beginning of the study (one 27

28 urchin per pool). Urchins were assigned randomly to a pool and placed at the opposite end of the coral and pot structures. As T. ventricosus began exhibiting signs of ill health, they were replaced with a new urchin for a total of 24 urchins used throughout the study (18 T. ventricosus and 6 D. antillarum). Routine water changes were performed daily, and urchin fecal pellets were removed through siphoning. Food was provided to urchins on the second day of experiments and secured in the middle of the pools by use of a clothes peg tethered to a dive weight. Food was replaced every day regardless of grazing or no grazing to ensure stable water quality and an equal fresh supply of food across pools and treatments. Temperatures were recorded daily using a point and shoot laser thermometer and a data-logger (LabJack, Digit-TL) submerged in one pool. Shelter seeking behavior was recorded as uncovered if urchins were found in open space of the pool with no surrounding cover, under coral if urchins were under the coral structure, and shaded if urchins were within the flower pot structure, or on shaded edges of the pool. The average percent of time and standard deviation for time spent in each location for Days 1, 2, and combined averages were compared in a table. Urchin health criteria was adopted from the protocol established by Francis-Floyd et al. (in press). Health categories were recorded as spine loss, test lesions, and urchin appearance (alert and active vs. lethargic and minimally responsive). Urchins were removed as signs of poor health became apparent. Appearance was classified as normal, poor, or very poor. Normal appearance was defined as normal spine position, no spine loss, or normal body position. Poor appearance was identified through spine loss or drooping spines, and very poor appearance was assigned when the urchin exhibited multiple spine loss (>30%), test lesions, or had abnormal body position (upside-down, or on their side). Observed urchin health 28

29 and expected urchin health was analyzed using a chi-square test of contingency for Day 1, Day 2, and Day 3 of the experiment. Urchins were observed the first hour that they were introduced to the tank and observations recorded. On Days 1-2, urchins were observed at 800, 1200, and 1600 for their selected habitat in the pool (uncovered, under coral, or shaded). Urchin location was recorded again on Days 5-6, and on Day 7, the trials were ended due to poor urchin health. Results Temperatures of the pools fluctuated between 2-4ºC every 24-hours throughout the experiment and steadily declined over the seven days of the trial (Figure 2-2). Urchin replacements took place as of Day 3 of the experiment (Figure 2-2), and very few urchins remained in pools more than 3 days. The count of urchins remaining in pools at 24h time intervals decreased as the number of days increased (Figure 2-3). D. antillarum numbers remained stable at six urchins until Day 6 when all urchins were removed and 0 remained. T. ventricosus urchins were replaced on Day 3, and abundances dropped by 61% for urchins held in the pools for 4 or more days. No urchins survived 7 days in the pools, and only 11% of total T. ventricosus survived to Day 6 (Figure 2-3). Urchin shelter seeking behavior was variable across all treatments where mean percent of time spent overlapped across treatments (Table 2-3). D. antillarum spent an average of 31.1% 30.2 of time uncovered, 6.7% 22.1 under coral, and 61.7% 34.8 of time in shaded habitat (Table 2-1). T. ventricosus spent 32.3% 30.4 of time uncovered, 8. 6% 24.3 under coral, and 58.4% 36.3 in shaded habitat (Table 2-3). On Day 1 all urchins appeared in normal health. On Day 2, 1 urchin exhibited signs of very poor health, but there was not a significant difference in health amongst all urchins 29

30 ( 2 =1.011, p= ) (Table 2-1). The first urchin removal event took place on Day 3 where 6 urchins exhibited signs of poor health. On Day 3, urchins exhibited significant differences in expected normal health and those observed with normal health, where 6 urchins appeared to be in poor health, and 10 urchins remained healthy ( 2 = 7.385, p= ) (Table 2-2). Discussion We were able to partially fulfill objectives to determine the parameters necessary to promote D. antillarum health in an ex situ environment, understand habitat usage of D. antillarum in a controlled environment when compared to expected normal behavior, and understand proper health metrics for monitoring D. antillarum health for in situ relocations and studies. Due to a sudden decline in urchin health on Day 3 of the study, conclusions drawn from analysis are speculative, and require further investigation to draw substantive conclusions. Techniques for monitoring urchin health were adopted from Francis-Floyd et al. (in press) and were used in later field experiments demonstrated in the following chapter. This validated that these adopted protocols for ex situ assessment can serve as a standardized data collection tool for coral-urchin restoration programs monitoring urchin health in field studies. D. antillarum shelter-seeking behavior is typically associated with shaded habitats (Randall et al., 1964; Carpenter, 1984). This shelter-seeking behavior can be altered in the absence of predators, where urchins remain in the open taking advantage of grazing time (Carpenter, 1984). Upon movement to a new environment in situ, D. antillarum aggressively seek shaded shelter (pers. obs.). Urchins in this study showed differences in percent of time spent in habitats, which were highly variable, suggesting expected natural behaviors are not being exhibited in an artificial environment. Urchin behavior in artificial environments has been observed as dissimilar to wild urchin behavior (Sharp and Delgado, pers. comm.). All urchins 30

31 used in this study had been long-term residents of artificial environments. Their shelter-seeking behavior was not consistent nor showed patterns of preferred habitat throughout the trial. It is possible they did not seek shaded shelter in the absence of predators, but evidence to support this is minimal, and therefore cannot be conclusive. Observed abnormal behavior may also be a product of compromised urchin health, so these conclusions are suggestions for further studies. These experiments subjected urchins to multiple parameters that did not align with natural environment measurements. Urchins are sensitive to harmful effects of metals and other toxicants in the environment (Kobayashi, 1980; Bielmyer et al., 2005). Pools were washed thoroughly before use in experiments, however, it is possible that chemicals from the lining of the pool may have contributed to poor urchin health on Day 3 of the experiment as well. Experimentation outside of normal temperature ranges and stark artificial environments did not produce normal behaviors for both T. ventricosus and D. antillarum and may induce health issues not expected in a natural environment. It is suggested that great care must be taken when designing artificial holding tanks for either D. antillarum experimentation aiming to predict animal response out on the reef, and when developing mass production aquaculture plans. Effects of temperature on urchin health ex situ are well documented for the development of gonads within many urchin species (McBride et al., 1997; Spirlet et al., 2000; Lawrence et al., 2009), however very little peer-reviewed literature exists to document culture of D. antillarum and effects of sharp temperature changes in the environment. Lawrence et al. (2009) found that temperature has insignificant effects on the survival and metabolic functions of the urchin species Strongylocentrotus intermedius, in culture. These experiments were conducted using gradually adjusted temperatures however, dissimilar to the drastic lulls in temperature experienced during the cold snap during this experiment. Average water temperature of tropical 31

32 coral reef ecosystems ranges from 23-29ºC annually and does not typically fluctuate more than 1-2ºC in a 24h period, rather fluctuates over longer periods of time in weeks or months (Lee and Williams, 1999). While temperatures in the pool remained within the normal tropical reef temperature ranges, they did undergo greater than natural diurnal fluctuations. Despite attempts to warm the greenhouse with external heaters to counteract a cold snap experienced during the trial, pool temperatures fluctuated between 2-4ºC in less than a 24h period each day of the experiment. It is suspected that this great range in temperature fluctuations may have contributed to possible health deterioration; for example- hypothermic shock. Attempts to stabilize water temperatures of the ex situ environment be a priority in maintaining urchin health. Multiple variables exist on reefs where urchin relocations may take place, and a comprehensive understanding of factors affecting the survivorship of urchins is essential when designing land based experiments and aquaculture facilities for these species. This experiment highlighted the need to closely approximate the target natural environment of the Florida Keys when considering ex situ rearing efforts to enhance D. antillarum stock for use in coral-urchin restoration programs. Further, it highlighted the health assessment protocols established for release of D. antillarum onto restoration sites (Francis- Floyd et al., in press) are a viable source for monitoring health of experimental animals. These monitoring protocols can also serve as a baseline for monitoring urchin health at restoration sites. Outcomes of this experiment demonstrate the vital nature of documenting health in any scenario where urchin relocations are tested. 32

33 Figures Figure 2-1: Grid and design of experimental pools. 33

34 Number of Urchins Figure 2-2: Temperature of urchin holding pools during experiment. Temperature of pools fluctuated between 2-4ºC over a 24h period and declined overtime Days in Pool Total Tripnuestes Diadema Figure 2-3: Urchin abundances across times spent in pools % of total urchins remained in pools for 3 days. Urchin abundances dropped at Day 4, and no urchins survived in pools for 7 days. 34

35 T. ventricosus D. antillarum Table 2-1: Urchin Health Day 2 Expected Observed Total Normal Very Poor Total Chi-squared table of observed urchin health versus expected urchin health ( 2 =1.011, p= ). Table 2-2: Urchin Health Day 3 Expected Observed Total Normal Poor Total Chi-squared table of observed urchin health versus expected urchin health ( 2 =7.385, p= ). Table 2-3: Shelter Seeking Behavior in D. antillarum and T. Ventricosus Species Day Uncovered Under Coral Shaded Day (22.10) 0 (0) 72.9 (22.68) Day (38.0) 11.9 (31.0) 55.0 (43.1) Total 31.1 (30.2) 6.7 (22.1) 61.7 (34.8) Day (23.5) 2.2 (8.6) 67.3 (26.6) Day (36.7) 8.9 (23.5) 53.3 (42.0) Total 32.3 (30.4) 8.6 (24.3) 58.4 (36.3) This table presents the average percent and standard deviation in parentheses of time urchins spent across three habitat options within pools. 35

36 CHAPTER 3 UNDERSTANDING THE RELATIONSHIP BETWEEN ACROPORA CERVICORNIS CORAL DENSITY AND JUVENILE DIADEMA ANTILLARUM SURVIVORSHIP Introduction The 1980s mass mortality event of Diadema antillarum coincided with a stark decline in scleractinian coral cover and a sharp increase in macroalgae on reefs around the Caribbean (Jackson, 2001). As hard coral cover declined and macroalgae increased, bare reef structure was eroded away, resulting in a shift from high structural complexity reefs to more homogenous, low complexity reefs dominated by macroalgae. Many reefs have been flattened through decrease in structural complexity (Alvarez-Fillip et al., 2009) leaving poor habitat for fish and invertebrate populations (Dixson et al., 2014). Over three decades after the mass mortality of D. antillarum, many countries have reported slow or no recovery of the populations to pre-mortality densities. Amongst other theories, reduced habitat complexity and opportunistic predation pressure may be linked to slow recovery (Bodmer et al., 2015). D. antillarum require enough reef rugosity to provide crevices or ledges to escape from predators (Levitan and Genovese, 1989). D. antillarum are preyed upon by many fish and invertebrate species including triggerfish, parrotfish, wrasses, and lobsters (Randall et al., 1964). Numerous studies suggest that sea urchin populations are heavily structured by predation pressures within the system, where increased predation pressure can decrease urchin populations and vice versa (McClanahan and Muthiga 1989; Hereu et al., 2005, Harborne et al., 2009). Decline in fish populations, due to overfishing, have shown to be correlated with an increase in D. antillarum abundance and an increase in grazing at reef sites (Carpenter, 1984). It is possible that predation pressure coupled with lack of reef habitat is attributed as one of the bottlenecks in D. antillarum recovery. 36

37 While recovery is minimal in many areas of the Caribbean, some countries have documented recent increases of D. antillarum individuals recolonizing their reefs. Jamaica and the Bay Islands of Honduras have reported increased densities of D. antillarum in areas such as Discovery Bay, Jamaica and Banco Capiro, Honduras (Edmunds and Carpenter, 2001; Bodmer et al., 2015). Lee (2006) found that habitat complexity and D. antillarum density were positively correlated on reefs in Jamaica and suggested that low habitat complexity may be one explanation for the failure of D. antillarum populations to recover to pre-mortality densities. However, in Discovery Bay, Jamaica, densities of D. antillarum have increased 10-fold with an inversely correlated decrease in macroalgae, and resultant correlation of an increase in juvenile corals up to 11-fold (Carpenter, 2001). Recovery of D. antillarum populations in this region are likely attributed to depauperate fish populations from overfishing, releasing urchins of the need to seek shelter and avoid predation allowing for exploitation of resources and increased urchin densities (Hughes et al., 1994). Bodmer et al. (2015) used the isolated population boom of D. antillarum individuals off Banco Capiro, Honduras to study factors that may have contributed to an increase in D. antillarum individuals at depths of 10m and 15m. On Banco Capiro, D. antillarum individuals were found at ± /m 2 and ± 19.11/100m 2 respectively, while at a nearby reef off Utila, abundances ranged from 1.22 ± 0.43 and 0.33 ± 0.17/100 m 2. Banco Capiro was more rugose, suggesting that amongst other factors, structural complexity may be a driving factor in D. antillarum adult population recovery. Increased reef complexity may protect smaller D. antillarum from local predation pressures and increase their likelihood of surviving to less predator prone sizes. D. antillarum with a test diameter <20mm are most prone to predation while larger urchins with a test diameter 37

38 >40mm seem to be less affected by predation (Clemente et al., 2007). Bodmer et al. (2015) recorded large abundances of juvenile D. antillarum individuals across two Honduran reefs, but no resultant significant increase in adult populations the following years. This suggests that recruitment is occurring and is not a significant bottleneck for population recovery, but rather smaller urchins, not surviving into adulthood because poor complexity and high predation pressure is a major factor in slow recovery of populations. In the Florida Keys, D. antillarum populations experienced mass mortalities from , and another disease outbreak in 1991 (Forcucci, 1994). These recurring outbreaks contributed to slow recovery rates (1994). Little data was collected before 1983 in the Florida Keys, but what was collected, was able to be reconstructed and analyzed to show that urchins abundances were high across reefs in the Florida Keys, similar to the rest of the Caribbean (Kissling et al., 2014). In some cases, D. antillarum densities were greater than 7.9 individuals/m 2 (Kier and Grant, 1965; Bauer, 1980; Forcucci, 1994; Kissling et al., 2014), but these dropped and remain since 2011 <1 individual/m 2 post-mortalities (Chiappone et al., 2011). Furthermore, Chiappone et al. (2009) reported a predominance of smaller D. antillarum test sizes between , which are similar to the findings of Bodmer et al. (2015) for Honduran reefs, supporting the theory there is poor survivorship of juveniles into adulthood for D. antillarum populations in the Florida Keys. Increasing habitat complexity may increase the survivorship of juveniles into adulthood, helping to re-establish adult population sizes. While decreased habitat complexity is common throughout the Caribbean (Alvarez-Filip et al., 2009), coral restoration efforts may provide opportunity to improve habitat complexity by adding coral to sites where it has diminished. When structural complexity was enhanced at reef sites, urchins relocated to the site showed 38

39 greater retention rates and decreased the amount of macroalgae in the area (Macia et al., 2007). Establishing a targeted threshold of habitat complexity would be desirable for reef managers when planning A. cervicornis coral restoration efforts may help enhance D. antillarum populations when relocated to these sites. Coral Restoration Foundation (CRF) has outplanted over 40,000 A. cervicornis clusters to >20 reef sites in the Florida since CRF s outplanting method involves outplanting ten or more A. cervicornis corals over an area of approximately 1m 2 (CRF, unpublished data). Multiple clusters are outplanted to the same general area of a restoration site to increase coral cover of the area. Anecdotal results reported by CRF suggest this method of outplanting provides opportunity for multiple coral thickets to develop and persist over time, increasing the structural complexity of the area, and possibly allowing re-establishment of habitat suitable for D. antillarum relocation. The objective of this experiment was to determine the coverage of A. cervicornis coral needed to promote urchin retention over time when urchins were relocated to coral clusters. Methods To assess the relationship between coral density and the retention of juvenile D. antillarum (20-30mm test diameter) over time when relocated to coral plots, experiments were performed in a moderately controlled environment at the Coral Restoration Foundation coral nursery offshore Tavernier, Florida (N24 58'55.60, W080 26'12.11) approximately 30ft in depth. Preliminary trials were conducted to understand optimal cage design that would allow urchins to inhabit the cages and reduce risk of escape. Cages were installed in a 3 x 5 grid and spaced 2m apart on each side (Figure 3-1) and constructed of double-sided chicken wire panels to decrease hole size. Each cage was 50cm x 50cm, constructed on land and taken to the nursery, cages were zip-tied to 1m rebar rods that had been installed in a sandy bottom area just outside the CRF nursery approximately 30ft deep (Figure 3-2). Cages were constructed of chicken wire on four sides and 39

40 left open at the top to allow predator access to urchins, but limit urchins from dispersal away from the coral cluster. In preliminary trials, corals experienced high mortality, and urchins immediately sought to escape cages in the sandy bottom habitat. To elevate corals and urchins off the sand, cement blocks, previously used for coral propagation, were added to each cage (Figure 3-2A). The blocks were originally created for coral propagation and growth, but were since retired. The platform of the blocks was made of cement with twelve 15cm PVC pipes cemented into the concrete. The blocks were placed with the PVC side down creating a table-like platform for corals. Coral density treatments were randomly assigned a cage number, and the assigned coral density was added to each cage using medium sized A. cervicornis corals (~30cm TLE) from the CRF coral nursery. Four coral density treatments (10, 25, 40, and 55 corals/0.25m 2 ) were replicated three times. These densities are visually assessed at approximately 30, 50, 80 and 95% ground coverage. An additional three control treatments were included in the experiment with a coral density of 0 corals/0.25m 2. Coral densities selected for this experiment began at 10 corals and were scaled up to test for the optimal A. cervicornis cluster density needed to enhance retention of small (2-3cm test diameter) D. antillarum when relocated to A. cervicornis coral clusters at established restoration sites. Minimum densities were based on current restoration practices of planting 10 A. cervicornis fragments of 30cm TLE, clustered approximately over an area ~1m 2 to form dense continuous thickets of A. cervicornis over time. Urchins were collected from a rubble patch inshore of Pickles Reef off Tavernier, Florida (N , W ) under permits designated by the Fish and Wildlife Conservation Commission (SAL SCRP) and Florida Keys National Marine Sanctuary (FKNMS A2). All urchins were collected and relocated to the experimental cages on the same 40

41 day. General characteristics of the collection site were taken including site substrate (rubble, sand, hard bottom), depth, tidal flow (incoming, outgoing, high tide, low tide), and surface and water temperature as outlined in the health assessment protocol for release of D. antillarum onto reefs (Francis-Floyd et al., in press). All urchins were assessed for tissue loss, spine loss, normal spine movement and normal body position before they were collected to ensure that urchins were healthy upon relocation to the nursery (Francis-Floyd et al., in press). Each urchin was measured using a plastic ruler, and only those that were 2-3cm in test diameter were included in the study. Four urchins were added to each cage and monitored over 20 days when weather was appropriate, for health condition, behavior, movement and urchin counts within coral, under blocks, and total present within the cage on Days Health of the urchins was assessed at the time of collection and throughout the trial to understand how urchin health was affected by collection and relocation methods. Categories of health assessment criteria were selected from the D. antillarum Health Assessment developed for FWC by Francis-Floyd et al. (in press). The following criteria were visually assessed, any urchins exhibiting abnormal behavior or appearance were not selected for the trials: 1) Spine position: healthy urchins should have spines erect and extended, drooping spines were considered abnormal, as they are indicative of a disease. 2) Spine movement: spine movement simply assessed by the movement or non-movement of spines when approached for capture. Abnormal spine movement was considered slow spine movement or no spine movement. 3) Spine loss: spine loss was assessed by observing the surrounding area the urchin inhabited for whole spines and a visual assessment of the test. Spines that were dropped from the test and spines that were broken, were distinguishable by the type of lesion left on the test. If spines are dropping from the test, skeleton is exposed at the band and socket joint where the spine is 41

42 attached to the test. 4) Spine breakage: spine breakage does not leave visual lesions to the test, but can be seen by shortened spines on the urchin and indicated by broken spines in the surrounding area. Urchins with 5% or more spine breakage were considered abnormal, and not selected for the trials. 5) Test lesions: a healthy urchin test should exhibit no signs of tissue exposure. If test was exposed, urchins were not selected for trials. The criteria outlined here, were used in visual assessments of the urchins while in coral plots at the CRF nursery. The number of urchins exhibiting abnormal signs within each plot was counted during the experiment and analyzed for differences between collection site, and the end of the experiment, as well as differences amongst treatments. A one-way analysis of variance was used to compare the average of urchins present in cages across treatments and a Tukey HSD test was used to determine differences amongst treatments in R version stats package (2016). We used descriptive statistics (averages and standard deviations) to demonstrate trends in the average number of urchins for treatments at each monitoring point to evaluate trends in urchin abundance in plots and within coral overtime. For each treatment, the average number of urchins present within the plot, within coral, and under blocks was plotted on a graph where the relationship between time and urchin abundance could be evaluated for trends. Frequencies of abnormal health parameters were recorded for each plot by counting the number of urchins exhibiting abnormal signs. The health of urchins did not change over time, with the exception of one event of spine breakage for one urchin on Day 2. Variance in spine loss across treatments for Day 2 was evaluated using one-way analysis of variance in R version stats package (2016) for spine breakage against treatment. 42

43 Results The average abundance of urchins across treatments was significantly different (p<0.01). further analysis with a Tukey HSD test, cages with 0 and 10 corals were not significantly different from one another (p=0.95), and cages >25 corals were significantly different from cages with 0 corals (p<0.05). Over time, there was a decreasing abundance of urchins within cages <25 corals per cage (Figure 3-3). In cages with >25 corals, there was an inverse relationship between urchins in corals and urchins in clusters- as more urchins sheltered under blocks, less were found in corals, and vice versa. Within increasing coral cover above 25 corals per structure, urchin numbers remained relatively stable. Overall, counts of urchins found in each cluster per treatment showed that after Day 7, there was a decrease in urchin abundance in cages with less than 25 corals. Plots with >25 corals showed stability in urchin abundance throughout the experiment where 75% of urchins added to plots were retained in treatments (Figure 3-3). An average of urchins persisted over time within plots of 10 corals. Plotted averages of urchins within 10 coral plots decreased over the 20 days of the experiment, notably on day 7 where averages dropped from 4 urchins to 2 urchins and ended with an average of 1 urchin per plot (Figure 3-5). Treatments with 25 corals showed a variable response in urchins in corals and urchins under blocks compared to control and 10 coral treatments (Figure 3-6). Overall, an average urchins were present within coral structures, while urchins were found under blocks, while the average urchins present within the plot throughout the experiments was (Table 3-1). Urchins present within corals and under blocks with 25 corals per unit appeared to have an inverse relationship (Figure 3-6). As urchin abundances rose in corals, urchin abundances decreased under blocks, and vice versa. 43

44 In treatments with 40 corals, an average of urchins were found in coral clusters, whereas an average of urchins were found under blocks (Table 3-1). These plots showed a stable trend in overall urchin presence over time. For plots with 40 corals, urchin abundances were inversely related, with a trend of greater urchin presence within coral plots (Figure 3-7). Treatments with 50 corals per plot had an average of urchins present in corals versus an average of urchins found under blocks (Table 3-1). Similar to the 40-coral treatment, 50 coral plots showed a stable trend in urchin abundance over the duration of the experiment (Figure 3-8). Recruitment occurred in two of the examined plots (one 40 coral and one 55 coral treatment), with 5 urchins present on Day 17. Urchin health parameters were recorded on each monitoring trip. Urchins did not exhibit signs of abnormalities or poor health throughout the experiment. There were no significant signs of ill health for urchins relocated to plots within the nursery or across coral density treatments. Urchins were either present or absent from plots and there were no signs of predation or poor health (broken spines, test lesions, or loss of spines) with one exception where one urchin showed signs of broken spines. All urchins exhibited normal spine movement and normal spine position. Discussion This study demonstrated a potential threshold for A. cervicornis coral density of 25 corals/0.25m 2 needed for small D. antillarum to have retention rates greater than 75% of translocated urchins after 20 days. Urchin abundance was significantly different across treatments, where cages with <10 corals had smaller urchin abundances overtime and cages >25 corals had greater urchin abundances. This provides an important piece of information for 44

45 managers involved in coral restoration activities and D. antillarum translocation with the purpose of restoring reef biodiversity and habitat complexity. Cages were used in this experiment to keep urchins from wandering away and specifically address the relationship of coral density and the shelter it provided urchins as protection from predation. Caging artifacts are a concern in any caging experiment, where cage and mesh size can affect water flow, sedimentation, light attenuation amongst many other parameters (Stocker, 1986; Hall et al. 1990; Como et al., 2006; Miller and Gaylord, 2007). Cages can also increase the abundance of other biofauna that may be attracted to increased structure at the experiment site and lead to unexpected artifacts of predation on the organism of study (Stocker, 1986). Because the aim of this study was to understand which A. cervicornis coral densities could promote urchin retention over time, cages were left open at the top to keep urchins within the cage, but still allow predator access to the urchins. It is possible that these cages attracted predator abundance to the area as it is well established that reef fishes and other invertebrates are attracted to increased artificial or natural complex structures in the marine environment. Since one of the goals was to understand how corals provided protection from predation, this was not a concern for results. D. antillarum exhibit photosensitive behaviors, showing preference for habitats with lower light levels (Millot, 1953, 1954; Woodley, 1982). Cement blocks at the bottom of cages were somewhat elevated from sand where urchins retreated under blocks when added to plots. Over time, urchins inhabited corals or remained underneath blocks. As strong winds brought greater wave action and turbidity, sand began to build up under blocks. Plots with less than 25 corals experienced a loss of urchins, some plots ending the experiment with no urchins present. It appears as habitat became unavailable under the blocks, coral cover became a major contributing 45

46 factor for the ability of the urchin to stay within the plot. Multiple factors contributing to a declining trend in D. antillarum abundances within cages may have occurred including predation from surrounding fish populations or urchins leaving to seek better shelter (Carpenter, 1984). With unsuitable coral cover, urchin abundances could not be sustained. This is important for relocation site considerations. Sites with greater reef rugosity and more coral cover may increase the urchin retention rate. If reef rugosity is insufficient, greater coral densities may be required to maintain urchin retention rates. There was a general trend that as coral density increased, urchin retention also increased, indicating a positive linear correlation between coral density and urchin retention rates. However, there was an apparent threshold of coral density that promoted stable urchin retention rates >75% where coral clusters with >25 corals/0.25m 2 retained 75% or more of urchins within the plot, and clusters with <10 corals/0.25m 2, were similar in retention rates (<25%). To further understand if this is a true linear relationship, future studies should increase coral density replicates and the number of urchins relocated to coral clusters. These findings of a threshold however, do indicate there is a tipping point where urchin retention is promoted therefore, it is recommended that urchins are only relocated to clusters with enough structural complexity that will promote high retention rates, which is possibly coral densities >25 corals/0.25m 2. While cautiously interpreted, two events of urchin recruitment occurred during the experiment, showing that with greater coral density, there is potential for recruitment of urchins to coral clusters. On day 17, one plot each with 40 and 55 corals, had five urchins present (one more than initially added). It is unclear if these urchins were from another cage or from a wild source. Because of the great distance to a wild population, it is speculated that the increase in urchins was from urchins migrating from one cage to another. Unfortunately, on Day 20, these 46

47 cages had been dismantled by strong wind and waves, and urchin abundances could not be recorded a second time. While this only occurred twice during the experiment, it may be important for future studies. As coral cover increases, there may be the potential to promote and foster aggregates of urchins within clusters that can chemically attract other urchins. Aggregates of urchins may have greater impact on surrounding macroalgae cover (Hernandez et al., 2007) and contribute to natural re-population of urchins to reseed downstream reefs (Bauer, 1976; Karlson and Levitan, 1989). This idea should be developed and further explored. Throughout the experiment, there were no signs of poor health. All urchins retained normal spine position and movement and did not show signs of spine loss, spine breakage, or test lesions, indicating current practices for urchin relocation were not detrimental to the health of the urchin where ample coral cover was provided. Therefore, relocation is not expected to have an impact on relocated populations through this management strategy if enough coral cover is provided. Because of limitations to the study, it is important to address some of the findings with further experiments. Greater replication and investigation is warranted. Thresholds found in this study should be identified on reef sites and preliminary D. antillarum relocations should be conducted and monitored to understand translation of nursery retention to reef site retention rates. Urchin densities have been shown to affect the abundance of urchins within a cluster or artificial habitat (Sharp and Delgado, pers. comm.). To understand retention rates when above four urchins per plot, studies should be repeated with various coral and urchin densities. This type of information could be used in management plans to enhance success of coral-urchin restoration projects. 47

48 The implications of these findings are important to consider when combining coral and urchin restoration projects into reef recovery programs. They corroborate previous studies that suggest increased structural complexity can increase D. antillarum abundances, and lay the ground work for additional urchin relocation experiments on reef habitats. With coral cover of 25 corals/0.25m 2, urchin retention can be stably maintained and possibly >75%. Providing habitat to maintain D. antillarum populations could enhance both coral and urchin restoration success, providing opportunity for reversal of the currently dominated macroalgae ecosystem. Stocking restoration sites with multiple juvenile individuals could provide D. antillarum with the necessary habitat to outgrow predation vulnerability and form breeding populations to increase recruitment of urchins to the reef site. 48

49 Figures Figure 3-1: Layout of coral density experimental grid. Cages were constructed of chicken wire with an area of 0.25m 2 and installed in a sandy bottom area on the edge of the CRF coral nursery. Cages were 2m apart from each other. Coral density treatments were randomly assigned to each cage for densities of 0, 10, 25, 40, and 55 corals shown here. 49

50 Average Urchin Count Figure 3-2: Cage Design. Coral densities began at 10 corals per cluster and scaled up by increments of 15 corals (A= control, B= 10 corals, C= 25 corals, D= 40 corals, and E= 55 corals) Time (days) 55 Corals 40 Corals 25 Corals 10 Corals 0 Corals 75% Retainment Figure 3-3: Average urchins present for time across all treatments. Average urchin counts with standard error bars for all treatments at each monitoring point throughout the experiment. Plots with >25 corals retained a stable abundance of urchins over the 20- day period, where 75% or more of the urchins relocated to plots were retained. The red line depicts the threshold for 75% urchin retention within the plots. Standard error bars depict the deviation from the population s true value. 50

51 Averagde Urchin Count Average Urchin Count Urchins in Coral Urchins Under Blocks Total Urchins Time (Days) Figure 3-4: Urchins present over time for control cages. Average counts with standard error bars for urchins present at each monitoring point during the experiment for control plots across all potential locations within the plot, which includes in coral and under blocks. The grey line represents overall average of total urchins present within the plot. In control plots, overall urchin retention dropped quickly over a short-term period Urchins in Coral Urchins Under Block Total Urchins Time (Days) Figure 3-5: Urchins present over time for 10 corals. Average counts with standard error bars for urchins present at each monitoring point during the experiment for plots with 10 corals across all potential locations within the plot, which includes in coral and under blocks. The grey line represents overall average of total urchins present within the plot. In 10 coral plots, overall urchin retention dropped quickly over a short-term period. 51

52 Average Urchin Count Avergae Urchin Count Urchins in Coral Urchins Under Blocks Total Urchins Time (days) Figure 3-6: Urchins present over time for 25 corals. Average counts with standard error bars for urchins present at each monitoring point during the experiment for plots with 25 corals across all potential locations within the plot, which includes in coral and under blocks. The grey line represents overall average of total urchins present within the plot. In 25 coral plots, stable urchin retention was maintained over the 20- day period Urchins in Coral Urchins Under Blocks Total Urchins Time (days) Figure 3-7: Urchins present over time for 40 corals. Average counts with standard error bars for urchins present at each monitoring point during the experiment for plots with 40 corals across all potential locations within the plot, which includes in coral and under blocks. The grey line represents overall average of total urchins present within the plot. In 40 coral plots, stable urchin retention was maintained over the 20- day period. 52

53 Average Urchin Count Urchins in Coral Urchins Under Blocks Total Urchins Time (days) Figure 3-8: Urchins present over time for 55 corals. Average counts with standard error bars for urchins present at each monitoring point during the experiment for plots with 55 corals across all potential locations within the plot, which includes in coral and under blocks. The grey line represents overall average of total urchins present within the plot. In 55 coral plots, stable urchin retention was maintained over the 20- day period. Table 3-1: Averages of urchins present within cages during experiment. Treatment Urchins in Coral (avg std dev) Urchins Under Blocks (avg std dev) Urchins Present (avg std dev) 0 corals corals corals corals corals Table of the average ( standard deviations) abundance of urchins present within experimental plots for the number of urchins in coral, under blocks, and the total urchins present with standard deviation values. 53

54 CHAPTER 4 PREDICTING SUCCESSFUL DIADEMA ANITLLARUM RELOCATION SITES IN THE FLORIDA KEYS Introduction When done on a large scale, coral restoration efforts can create a positive trajectory for threatened Acroporid populations (Miller et al., 2016). Despite positive trajectories, corals outplanted to reefs are still subject to various local and global stressors including climate change, rising sea surface temperatures, and the persistence of macroalgae which makes habitat unsuitable for coral recruitment and recovery (Jackson et al., 2001; Knowlton, 2001; Bellwood et al., 2004; Brown et al., 2014; Dixson et al., 2014). The National Oceanic and Atmospheric Administration (NOAA) Acropora Recovery Plan (2015), notes that for coral restoration efforts to be successful, populations must be present across approximately 5% of consolidated reef habitat and sustained for 20 or more years. Abating local stressors, such as poor water quality, eutrophication, and decreasing surrounding macroalgae may help recovering coral populations become resilient in the face of global climatic pressures (Mumby et al., 2007; Anthony et al., 2015) and decrease requirements for continued restoration. Diadema antillarum are critical to the removal of macroalgae on Caribbean coral reefs. In the absence of D. antillarum, as has been the case since their epidemic mortality in 1983, coral reefs have become dominated by fast-growing macroalgae which outcompetes and smothers the scleractinian corals (Lirman, 2001; McCook et al, 2001). Recovery of D. antillarum has the potential to reverse reef states, currently existing as macroalgae dominated, back to coral dominated states (Edmunds and Carpenter, 2001). Many studies have demonstrated the significant positive impacts reintroduction of D. antillarum can have to surrounding macroalgae (Nedimyer & Moe, 2003; Burdick, 2008; Dame, 2008). Natural recovery of D. antillarum on reefs in Jamaica have corresponded with a remarkable increase in scleractinian coral cover 54

55 (Edmunds and Carpenter, 2001; Carpenter and Edmunds, 2006; Idjadi et al., 2010). D. antillarum are in need of enhanced coral cover for protection from predators (Levitan and Genovese, 1989), as such, the recovery of coral may promote the recovery urchins, and the recovery of urchins may in turn promote the recovery of corals. The objectives of this study were to predict and identify current restoration efforts that may facilitate the recovery of urchins by selecting suitable translocation reefs with appropriate A. cervicornis coral coverage, and create a tool for managers charged with implementing coral restoration programs in conjunction with D. antillarum recovery programs, so that preliminary coral-urchin restoration trials may be tested using the information presented in this study. Methods Datasets used to generate information for this study were obtained from previous coral density experiments in the CRF nursery (Chapter 3), annual A. cervicornis coral density and growth (Lirman et al., 2014), and outplanting data with permission from Coral Restoration Foundation (CRF, unpublished data). Information regarding restoration site name, coordinates, number of clusters at the restoration site, number of corals outplanted within the cluster, and time on reef were extracted from the CRF outplanting database. The average size of each coral fragment outplanted to the reef was ~30cm total linear extension (TLE) (CRF, unpublished data) with the exception of ECO sites, where average coral size was ~60cm TLE. The average annual growth rate of the corals was set at 4cm per initial size of the coral according to Lirman et al. (2014). Using the datasets above, we generated a total cluster-biomass for each coral cluster and predicted from which sites would be suitable for coral-urchin restoration projects based off the cluster-biomass of outplanted A. cervicornis to reef sites in the Florida Keys. A. cervicornis were outplanted to reefs in one of two ways. The first method, which is the protocol used for the majority of outplant sites, involved clearing the substrate of algae and 55

56 calcareous algae to the clean limestone rock underneath with a hammer. Corals were attached to the reef using a two-part marine epoxy at three points of attachment. Once attached, the epoxy hardened and the coral grew over the epoxy on to the substrate, and eventually fused with other surrounding A. cervicornis colonies within the cluster. Corals outplanted with this first method were on average 30cm TLE. The second method, which is still being refined, is an expedited coral outplanting (ECO) strategy where large corals (avg. 60cm TLE) were taken to patch reefs, wedged into crevices, and then overlaid on one another to create a coral thicket. The latter method was first deployed in September 2015, and has been used experimentally to increase the amount of coral at a reef site in a short amount of time. Diadema Retention Model. The calculated cluster-biomass of coral clusters used in this study was defined as the cumulative TLE (cm) of all corals outplanted in an area of 1m 2 (clusterbiomass TLE cm/m 2 ). First, all coral density treatments in Chapter 3, were converted to units of cluster-biomass using the initial cluster biomass equation: B = S i n A Where Si is the initial coral size, n is the number of coral fragments within the designated cluster, and A is the area over which the corals were placed. The converted cluster-biomass for coral density treatments (0, 1200, 3000, 4800, and 6600cm TLE/m 2 ) and their associated Day 20 urchin retention rates (.25,.25,.75, 1.0, 1.0) were fed into R Version using the stats Package (2016) to generate a Diadema Retention Model (DRM) with cluster-biomass as a predictor of urchin retention. The model was analyzed using regression analysis. Identifying Suitable Sites. To identify sites where coral-urchin relocations could take place from , values were generated for cluster-biomass that would support urchin 56

57 retention >75% for at least 20 days. Outputs from the DRM for desired thresholds were assigned one of three categories based on the percent of urchin retention each cluster could maintain: Good ( %), Better ( %), and Best (>95%). These values were used to suggest sites where necessary cluster-biomass currently existed from We calculated the yearly cluster-biomass (BY) for each cluster in relation to the amount of time on the reef for years using the following equation: B Y = B (AP) t (AS) t Where B is the beginning cluster-biomass (calculated above), t is the time the cluster has been on the reef in years, AP is the annual productivity of the corals (4cm/year; Lirman et al., 2014), and AS is the annual survivorship of outplanted corals (0.80; CRF, unpublished data). This calculation assumes the following parameters: (i) corals outplanted to reef sites have an annual survivorship rate of 0.80, (ii) corals >30cm TLE have an annual productivity rate of 4cm annually per initial size of the coral, (iii) coral clusters are outplanted over an area ~1m 2, and (iv) coral growth is exponential. Coordinates of clusters identified as good, better, and best for urchin relocations were imported into ESRI ArcGIS. Sites where specific ranges of corals categorized as good, better, or best were plotted using yellow, blue, and green dots respectively for years Results The Diadema Retention Model (DRM) was significantly correlated between cluster-biomass and urchin retention rates (p= , F= 25.25, R 2 =0.8532) and is linear between cm TLE/m 2 units of cluster-biomass where urchin retention is between 0-100% (Figure 4-1). Input for desired urchin retention rates for Good ( %), Better ( %), and Best (>95%) were specified as requiring cm TLE/m 2, cm TLE/m 2, and >5107.8cm TLE/m 2 respectively, to retain urchins for at least 20 days within A. cervicornis coral 57

58 clusters outplanted to restoration sites. A table of restoration sites with the corresponding number of clusters was generated as a reference for optimal coral densities for urchin relocation trials to be used by managers for easy identification of reef sites suitable for coral-urchin restoration (Appendix A). Over 3,000 A. cervicornis coral clusters have been outplanted across 40 restoration sites along the Florida Keys Reef Tract since 2014 and were incorporated into this model. In 2017, very few clusters suited necessary parameters needed for successful urchin retention. Currently in 2017, 6.77%, 4.21%, and 15.5% of A. cervicornis clusters are considered good, better, and best respectively for urchin relocations (Figure 4-5). Based on growth rates in 2018, 3.58%, 4.73%, and 61.68% of coral clusters will be considered, good, better, or best for urchin retention respectively (Table 4-5). In 2019, 0.22%, 1.40%, and 98.27% will be considered good, better, and best (Table 4-5). In 2020, 100% of coral clusters used in this study are expected to be suitable for urchin relocation efforts (Table 4-5). Restoration sites clustered around areas of the Upper Keys and spread out overtime as coral-biomass increased along the Florida reef tract (Figure 4-2 to 4-5). Discussion The DRM predicted which reefs sites are suitable for urchin relocation now and will become so over the next three years. It serves as a baseline for predicting the amount of A. cervicornis cluster-biomass needed to support urchin retention at restoration sites and may offer explanations for past D. antillarum relocation efforts that were unsuccessful as a result of poor reef complexity (Nedimyer & Moe, 2003; Burdick, 2008). As coral-urchin restoration is scaled up, additional data can be incorporated into the DRM for enhanced results to predict restoration sites that are suitable for urchin relocation efforts. 58

59 Based on the applied information from the DRM, <30% of A. cervicornis clusters outplanted to reef sites are currently able to support a target minimum of 75% urchin retention for 20 days. Relocating urchins at present to A. cervicornis clusters may result in mediocre retention rates, but provide opportunity to begin preliminary relocation trials to test the results of these studies further. Corals outplanted in clusters of 10 with an average size of 30cm TLE appear to reach cluster-biomass necessary to promote >95% urchin retention around 3 years of growth. In 2018, the total proportion of clusters suitable for potential urchin relocation trials increases to 69.99%, then 99.89% and finally 100% in years 2019 and 2020, where all corals have been on the reef for 3 or more years. Scaling up current coral restoration programs in the Florida Keys and Caribbean is of great emphasis for many reef managers (NOAA Workshop, Nov. 2016). Increasing the abundance of coral on degraded reef sites and promoting their long-term recovery is a priority outlined in the NOAA Acropora Recovery Plan (2015) for recovery of the species. Unprecedented amounts of corals have been outplanted across the Florida Keys since 2014, but may take a few years to reach optimal sizes for urchin relocations. Recently, an effort to scale-up restoration projects has taken place using ECO methods. These methods have been deployed at targeted patch reef restoration sites, where more than 50 large corals can be placed in an area ~1m 2. In combination with current methods, this ECO strategy produced 7 restoration sites in less than one year that are suitable to promote >75% urchin retention if relocated to clusters at these restoration sites. As more reef area becomes available overtime, it is timely to begin consideration of alternative strategies to increase D. antillarum production and collections. Ex situ D. antillarum rearing may increase numbers of urchins available for relocation trials and speed recovery of the species (Chandler et al., in press). 59

60 Although promising, the DRM has limitations in design and available data that warrant further investigation. The DRM was targeted towards urchins relocated to a caged coral cluster. It is assumed that with a greater number of suitable coral clusters, you may add more urchins to a restoration site. However, it is important to understand the relationship between urchin retention and coral cover over a greater area of the reef. Carrying capacity of urchins on reefs is currently an unknown factor and one recently identified as a potential critical factor in future management decisions (Diadema Workshop, Feb. 2017). Data used to create the DRM was based on caged experiments in a semi-controlled environment. It is imperative to replicate experiments on restoration sites and compare against nursery experiments to better understand the relationship of D. antillarum retention more practically. The sites suggested for preliminary urchin relocations can serve as a viable starting point for these experiments. It is known that with increasing reef complexity, D. antillarum densities can increase (Clemente and Hernandez, 2008; Dame 2008; Rogers and Lorenzen, 2016). As A. cervicornis clusters become larger and more complex suitable area for urchin relocations can increase. The DRM currently predicts the sites where cluster-biomass will become suitable over the period of a few years using an A. cervicornis annual productivity rate (Lirman et al., 2014) and average A. cervicornis survivorship across multiple reef sites in the Florida Keys (CRF, unpublished data). The equations used to predict this, assume exponential growth, which is impractical for a biotic organism that can experience multiple stressors that can cause mortality. Multiple variables including differences in reef habitat (e.g. Dizon and Yap, 2005; Mieog et al., 2009), coral genotype (e.g. Baums, 2008; Williams et al., 2014), and bleaching or disease events (e.g. Aronson and Precht, 2001; Eakin et al., 2010) have the ability to alter survivorship of coral 60

61 clusters. If known, specific survivorship percentages for each reef site or region should be incorporated in to the DRM for greater predictive accuracy of cluster-biomass over sequential years. Consequently, it is important to consider in recovery plans which sites may have seen greater mortalities of restored corals in the sequential years that are presented in tables for reef site identification so that urchins are not relocated to sites where mass coral mortalities have taken place. Finally, the DRM only used A. cervicornis to model cluster-biomass and urchin retention. A. cervicornis can rapidly form densely branching thickets that can provide juvenile D. antillarum with natural habitat as they grow into adults and work to re-establish local populations. While other coral species and growth forms, such as Orbicella spp., A. palmata, and Millepora complanata, may provide more suitable habitat for larger urchins (Weil et al., 1984), past declines in coral cover coupled with their slower growth rates may not provide viable options for timely restoration of corals and urchins. As A. palmata or other coral species become more abundant in coral restoration programs, perhaps they too can begin to provide increased habitat complexity for urchins. A holistic approach to reef restoration combining coral and urchin recovery may lead to enhanced results, where corals provide greater habitat complexity for urchins, and urchins reduce the amount of macroalgae in the area increasing coral survivorship by reducing competition. Some sites currently exist to begin preliminary D. antillarum relocation trials, however, to support greater urchin retention, efforts to scale up restoration and configure present coral outplantings so that optimum cluster-biomass is achieved in a shorter amount of time should be the next steps taken to increase efficacy of coral-urchin restoration programs for enhanced reef recovery. 61

62 Figures Figure 4-1: Diadema Retention Model for A. cervicornis restoration sites. The relationship of cluster-biomass and urchin retention is represented by the black line and the 95% confidence intervals are represented by the red-dotted lines. The blue dashed line represents the point at which 75% urchin retention is achieved for cluster-biomass ( cm TLE/m 2 ). 62

63 Figure 4-2: A. cervicornis restoration sites identified in the Florida Keys for preliminary D. antillarum relocations in Figure 4-3: A. cervicornis restoration sites identified in the Florida Keys for preliminary D. antillarum relocations in

64 Figure 4-4: A. cervicornis restoration sites identified in the Florida Keys for preliminary D. antillarum relocations in Figure 4-5: A. cervicornis restoration sites identified in the Florida Keys for preliminary D. antillarum relocations in