TANDEM RESTORATION OF DIADEMA ANTILLARUM AND …

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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

2017

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© 2017 Kayla J. Ripple

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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.

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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 .......................................................................................22

2 HEALTH ASSESSMENTS OF EX SITU EXPERIMENTATION OF DIADEMA

ANTILLARUM AND TRIPNUESTES VENTRICOSUS .........................................................26

Introduction .............................................................................................................................26 Methods ..................................................................................................................................27 Results .....................................................................................................................................29 Discussion ...............................................................................................................................30

3 UNDERSTANDING THE RELATIONSHIP BETWEEN ACROPORA CERVICORNIS

CORAL DENSITY AND JUVENILE DIADEMA ANTILLARUM SURVIVORSHIP .........36


4 PREDICTING SUCCESSFUL DIADEMA ANITLLARUM RELOCATION SITES IN

THE FLORIDA KEYS ...........................................................................................................54


5 DISCUSSION .........................................................................................................................66

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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 2017-2020 .........................76

LIST OF REFERENCES ...............................................................................................................80

BIOGRAPHICAL SKETCH .........................................................................................................91

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LIST OF TABLES

Table page

2-1 Urchin Health Day 2 ..........................................................................................................35

2-2 Urchin Health Day 3 ..........................................................................................................35

2-3 Shelter Seeking Behavior in D. antillarum and T. Ventricosus .........................................35

3-1 Averages of urchins present within cages during experiment. ..........................................53

4-1 Proportion of Clusters Present for Each Category for Years 2017-2020...........................65

A-1 Clusters Present for Each Category of Cluster-Biomass at Restoration Sites in 2017. .....76




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LIST OF FIGURES

Figure page

1-1 Urchin grazing effects. .......................................................................................................24

1-2 The spread of Diadema antillarum mortality sightings. ....................................................25

2-1 Grid and design of experimental pools. .............................................................................33

2-3 Urchin abundances across times spent in pools. ................................................................34

3-1 Layout of coral density experimental grid. ........................................................................49

3-2 Cage Design. ......................................................................................................................50

3-3 Average urchins present for time across all treatments. ....................................................50

3-4 Urchins present over time for control cages. .....................................................................51

3-5 Urchins present over time for 10 corals. ............................................................................51




4-1 “Diadema Retention Model” for A. cervicornis restoration sites. .....................................62

4-2 A. cervicornis restoration sites identified in the Florida Keys for preliminary D.

antillarum relocations in 2017. ..........................................................................................63







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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

By

Kayla J. Ripple

May 2017

Chair: Mark Flint

Major: Fisheries and Aquatic Science

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

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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 2020.

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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),

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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).

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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,

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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 20th century due in part to the development of new commercial

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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 km2 (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.

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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/m2 (Bauer, 1980). In the Florida Keys, densities were lower than other areas of the

Caribbean around 4-5 individuals/m2 (Kier and Grant, 1965; Bauer, 1976, 1980), but higher than

present densities at maximums of 0.33 individuals/m2 (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).

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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,

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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 coral-

dominated 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,

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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 one-

third 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.

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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).

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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

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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

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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

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/0.25m2). 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 75-100% 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 2017-2020

so that preliminary urchin relocation trials may be tested in future work to achieve.

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).

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).

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.).

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 2015. 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

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

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

30

(2=1.011, p= 0.3148) (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= 0.0066) (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

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

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.

33

Figures

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

34

Figure 2-2: Temperature of urchin holding pools during experiment. Temperature of pools

fluctuated between 2-4ºC over a 24h period and declined overtime.

Figure 2-3: Urchin abundances across times spent in pools. 96-100% of total urchins remained in

pools for 3 days. Urchin abundances dropped at Day 4, and no urchins survived in

pools for 7 days.

0

5

10

15

20

25

30

1 2 3 4 5 6 7

Num

ber

of

Urc

hin

s

Days in Pool

Total Tripnuestes Diadema

35

Table 2-1: Urchin Health Day 2

Expected Observed Total

Normal 16 15 31

Very Poor 0 1 1

Total 16 16 32

Chi-squared table of observed urchin health versus expected urchin health (2=1.011, p=

0.3148).

Table 2-2: Urchin Health Day 3

Expected Observed Total

Normal 16 10 26

Poor 0 6 6

Total 16 16 32

Chi-squared table of observed urchin health versus expected urchin health (2=7.385, p=

0.0066).

Table 2-3: Shelter Seeking Behavior in D. antillarum and T. Ventricosus

Species Day Uncovered Under Coral Shaded

D. anti

llaru

m

Day 1 28.6 (22.10) 0 (0) 72.9 (22.68)

Day 2 31.0 (38.0) 11.9 (31.0) 55.0 (43.1)

Total 31.1 (30.2) 6.7 (22.1) 61.7 (34.8)

T. ve

ntr

icosu

s

Day 1 31.1 (23.5) 2.2 (8.6) 67.3 (26.6)

Day 2 35.6 (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.

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.

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 225.67 ± 26.06 100/m2 and 73.56 ± 19.11/100m2respectively, while at a nearby

reef off Utila, abundances ranged from 1.22 ± 0.43 and 0.33 ± 0.17/100 m2. 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

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 1983-

84, 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/m2 (Kier and Grant, 1965; Bauer, 1980; Forcucci, 1994; Kissling et al., 2014), but

these dropped and remain since 2011 <1 individual/m2 post-mortalities (Chiappone et al., 2011).

Furthermore, Chiappone et al. (2009) reported a predominance of smaller D. antillarum test sizes

between 1999-2007, 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

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 2000. CRF’s outplanting method involves outplanting ten or

more A. cervicornis corals over an area of approximately 1m2 (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

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/0.25m2) 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/0.25m2. 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 ~1m2 to form dense

continuous thickets of A. cervicornis over time.

Urchins were collected from a rubble patch inshore of Pickles Reef off Tavernier, Florida

(N2459.180’, W08025.126’) under permits designated by the Fish and Wildlife Conservation

Commission (SAL-16-1722-SCRP) and Florida Keys National Marine Sanctuary (FKNMS-

2015-028-A2). All urchins were collected and relocated to the experimental cages on the same

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 1-20.

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

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 3.2.2 ‘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

3.2.2 ‘stats’ package (2016) for spine breakage against treatment.

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 2.7 1.7 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 1.6

1.5 urchins were present within coral structures, while 2.2 1.5 urchins were found under

blocks, while the average urchins present within the plot throughout the experiments was 3.5

0.5 (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.

44

In treatments with 40 corals, an average of 2.4 1.3 urchins were found in coral clusters,

whereas an average of 1.51.25 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 2.9 1.5 urchins present in corals

versus an average of 0.9 1.3 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/0.25m2 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

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

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/0.25m2 retained 75% or more of urchins within

the plot, and clusters with <10 corals/0.25m2, 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/0.25m2.

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

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.

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/0.25m2, 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.

49

Figures

Figure 3-1: Layout of coral density experimental grid. Cages were constructed of chicken wire

with an area of 0.25m2 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.

50

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).

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.

0

1

2

3

4

5

1 2 3 7 17 20

Aver

age

Urc

hin

Count

Time (days)

55 Corals

40 Corals

25 Corals

10 Corals

0 Corals

75% Retainment

51

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.

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.

0

1

2

3

4

5

1 2 3 7 17 20

Aver

age

Urc

hin

Count

Time (Days)

Urchins in Coral

Urchins Under Blocks

Total Urchins

0

1

2

3

4

5

1 2 3 7 17 20

Aver

agde

Urc

hin

Count

Time (Days)

Urchins in Coral

Urchins Under Block

Total Urchins

52





within the plot. In 25 coral plots, stable urchin retention was maintained over the 20-

day period.






day period.

0

1

2

3

4

5

1 2 3 7 17 20

Aver

gae

Urc

hin

Count

Time (days)

Urchins in Coral


Total Urchins

0

1

2

3

4

5

1 2 3 7 17 20

Aver

age

Urc

hin

Count

Time (days)

Urchins in Coral


Total Urchins

53






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 0 0 2.2 1.7 2.3 1.7

10 corals 0 0.8 2.4 1.7 2.7 1.7

25 corals 1.6 1.5 2.2 1.5 3.8 0.5

40 corals 2.4 1.3 1.5 1.3 3.9 0.5

55 corals 2.9 1.5 0.9 1.3 3.9 0.5

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.

0

1

2

3

4

5

1 2 3 7 17 20

Aver

age

Urc

hin

Count

Time (days)

Urchins in Coral


Total Urchins

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

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 2017-2020 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

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 1m2 (cluster-

biomass TLE cm/m2). First, all coral density treatments in Chapter 3, were converted to units of

cluster-biomass using the “initial cluster biomass” equation:

𝐵 = 𝑆𝑖 × 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/m2) and their associated Day 20 urchin retention rates (.25, .25, .75, 1.0, 1.0) were

fed into R Version 3.2.2. 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 2017-2020, values were generated for cluster-biomass that would support urchin

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” (75-84.9%), “Better” (85-94.9%), and “Best” (>95%). These values were used to

suggest sites where necessary cluster-biomass currently existed from 2017-2020.

We calculated the yearly cluster-biomass (BY) for each cluster in relation to the amount of

time on the reef for years 2017-2020 using the following equation:

𝐵𝑌 = 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 ~1m2,

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 2017-2020.

Results

The Diadema Retention Model (DRM) was significantly correlated between cluster-biomass

and urchin retention rates (p=0.01604, F= 25.25, R2=0.8532) and is linear between 0-6000cm

TLE/m2 units of cluster-biomass where urchin retention is between 0-100% (Figure 4-1). Input

for desired urchin retention rates for “Good” (75-84.9%), “Better” (85-94.9%), and “Best”

(>95%) were specified as requiring 3782.6 - 4438.6cm TLE/m2, 4445.2 - 5101.2cm TLE/m2, and

>5107.8cm TLE/m2 respectively, to retain urchins for at least 20 days within A. cervicornis coral

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

2017-2020 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.

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

~1m2. 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).

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

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.

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

(3782.609cm TLE/m2).

63

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

antillarum relocations in 2017.



64





65

Table 4-1: Proportion of Clusters Present for Each Category for Years 2017-2020.

Year Good

(75-84.9%)

Better

(85-94.5%)

Best

(>95%)

Total

2017 246 (6.77) 153 (4.21) 565 (15.5) 964

(26.48)

2018 130 (3.58) 172 (4.73) 2,241 (61.68) 2,543

(69.99)

2019 8 (0.22) 51(1.40) 3,573 (98.27) 3,632

(98.89)

2020 0 (0) 0 (0) 3,636 (3636) 3,636

(100)

Table of total clusters present for each year in and the proportion they represent of total clusters

on the reef assigned for good, better, and best D. antillarum retention rates.

Proportions are recorded in parentheses.

66

CHAPTER 5

DISCUSSION

In the 1980s Diadema antillarum were decimated across the Florida Keys and wider

Caribbean reefs. This has been linked to macroalgae overgrowth and suffocation of scleractinian

corals resulting in loss of reef biodiversity (Bak et al., 1984; Lessios et al., 1984; Hughes et al.,

1987; Carpenter, 1990) creating a bleak future for the ecologically vital Florida Keys Reef Tract.

It has been postulated that restoration of both D. antillarum urchin species and Acropora

cervicornis coral species can create a symbiotic partnership between the two and assist with reef

recovery back to a coral dominated state. This study examined various strategies to assist this by

assessing six primary objectives using two experiments to generate a model predicting where

successful urchin relocations might take place to assist enhanced reef recovery. 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

restoration efforts that may facilitate the recovery of urchins by selecting suitable relocation to

67

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.

Diadema antillarum Health and Behavior Parameters Ex Situ: Applications for In Situ

Monitoring and Relocation Efforts

Current ex situ rearing efforts being conducted privately (Mr. Martin Moe, Islamorada,

FL) and in the aquarium industry (Mr. John Than, The Florida Aquarium, Tampa, FL) may soon

achieve standard protocols for mass production of these urchins allowing for a surplus of urchins

to be used for relocation efforts. On-land aquaculture provides many potential benefits for the

recovery of sea urchins such as mass production, control to produce uniform age class and size,

and potential annual restoration efforts versus reliance on seasonal factors to induce spawning

and settling in the wild.

Classification as an invertebrate renders certain ethical husbandry parameters required for

marine vertebrates, unnecessary to consider in design of experiments. D. antillarum is not

currently listed as a protected species under the U.S. Endangered Species Act and has not been

considered for listing on the IUCN Red List (IUCN 2016). In contradiction to the listing status of

D. antillarum, it is widely recognized that the species has been extensively affected from the

1980’s mass mortality event, where 93% of the Caribbean population was decimated (Lessios,

1988). Monitoring the health of urchins used in experiments can be beneficial for understanding

if current practices are detrimental to the health of the urchins, as well as contribute reasons for

good or poor urchin survivorship when relocated to coral restoration sites. Health and behavior is

also a concern for current efforts of D. antillarum propagation ex situ for translocation in situ.

Our first hypothesis- relocation of D. antillarum to coral restoration sites does not affect

health of the urchins- was addressed through the objectives: (1a) determine the parameters

necessary to promote D. antillarum health in an ex situ environment; (1b) understand potential

68

habitat usage and behavior of D. antillarum and T. ventricosus in a controlled environment when

compared to expected normal behavior; and (1c) understand proper health metrics for monitoring

D. antillarum health for in situ relocations and studies. These objectives specifically addressed

concerns with D. antillarum health in regards to moving them from the wild to another reef

habitat, as well as better understanding for behavior of ex situ urchins and parameters that could

potentially lead to poor health of urchins when relocated to reef sites.

Objective 1a: determine the parameters necessary to promote D. antillarum health in an ex

situ environment.

A protocol for parameters to maintain urchin health ex situ was not fully attained for this

objective. Urchins were adversely affected during trials in this experiment. By assessing health

of the urchins in ex situ trials, we highlighted the importance of controlling external abiotic

factors that may possibly cause stress in the urchins and decrease the overall accuracy of any

trials being conducted. Facilities conducting ex situ urchin breeding programs to increase

individuals for use in coral-urchin restoration efforts, should be fully equipped to provide

necessary temperature adjustments, supply of food for routine feeding regimes, and structure that

mimics in situ habitat for urchins to inhabit.

Objective 1b: understand potential habitat usage and behavior of D. antillarum and T.

ventricosus in a controlled environment when compared to expected normal behavior.

Concern also arises when considering altered behavior of ex situ raised urchins for urchin

relocation efforts. It has been observed that urchins raised in ex situ environments do not display

the same shelter-seeking or predator avoidance behaviors when compared to in situ collected D.

antillarum (Sharp and Delgado, pers. comm.). Urchins in ex situ trials showed abnormal shelter-

seeking behaviors before a decline in their health. It is uncertain whether these behaviors were

due to poor conditions or a result of the urchins being held ex situ, however it is clear their

behaviors were not similar to wild D. antillarum used for in situ experiments. In situ relocated

69

urchins immediately sought shelter under shaded habitat or within coral clusters. These urchins

appeared to possess innate knowledge to seek shelter and avoid predation. Selecting for natural

behaviors such as these or training ex situ raised urchins may increase the likelihood of urchin

relocation success where urchins seek shelter immediately, decreasing the likelihood of predation

and increasing their chances for survival.

Ex situ trials may provide necessary insight to urchin biology as we strive to fill in the

numerous gaps in our knowledge, but they need to closely approximate the target environment

where urchins will be relocated- the relatively climatically stable staghorn dominated restoration

sites of the Florida Keys. Further, future mass production through ex situ aquaculture of D.

antillarum as a restoration strategy needs to consider imprinting shelter seeking and predator

avoidance behavior if these behaviors are to be maintained throughout land-based holding prior

to translocation of D. antillarum to an environment that requires foraging and predator

avoidance.

Objective 1c: understand proper health metrics for monitoring D. antillarum health for in

situ relocations and studies.

Criteria used to evaluate the health of urchins in ex situ trials were outlined from a health

evaluation protocol developed by Francis-Floyd et al. (in press) as an evaluation tool for the

release of ex situ raised urchins in relocation efforts. These protocols not only proved to be

sufficient for evaluating the health of urchins ex situ, but also in situ studies for evaluating

stability in urchin health and effects of relocation practices. In contrast to the ex situ trials at

CFC, urchin health during in situ trials showed no significant declines in health throughout the

experiment where urchins exhibited the same healthy characteristics from collection to the end of

the experiment on Day 20. This suggests current relocation protocols from the collection,

70

transportation, and release onto coral clusters are not detrimental to the health of urchins, nor do

they promote predation when ample shelter is provided.

Understanding this provides practitioners with criteria to better interpret results of

relocation efforts. If efforts were unsuccessful and proper relocation protocols were followed,

failure may have resulted in areas outside of urchin health and relocation parameters, factors

including reef complexity, urchin homing behavior, or other unknown variables. Knowing this

can allow researchers to isolate other possibilities of failure, and explore further issues more in

depth. Experimental

Acropora cervicornis Densities to Promote Diadema antillarum Retention

Our second hypothesis- with increasing A. cervicornis density D. antillarum retention within

clusters increases- was addressed with the objective to (2a) determine the coverage of A.

cervicornis coral needed to promote urchin retention over time when urchins were relocated to

coral clusters. Current coral restoration efforts focus on propagation and restoration of A.

cervicornis fragments for outplanting at designated restoration sites (Young et al., 2012). While

not historically associated as D. antillarum habitat, A. cervicornis coral thickets as a result of

coral restoration efforts may provide an interim habitat for juvenile D. antillarum to grow

outside of predation prone test diameters and head start adult populations at restoration sites with

potential to reseed downstream reefs.

Objective 2a: determine the coverage of A. cervicornis coral needed to promote urchin

retention over time when urchins were relocated to coral clusters.

In situ trials identified units for cluster-biomass that may promote 75% or more retention of

juvenile D. antillarum when relocated to A. cervicornis clusters where appropriate coral

coverage was present. This study demonstrated a potential threshold for A. cervicornis coral

cluster density of 25 corals/0.25m2 or more needed for juvenile D. antillarum to have a retention

71

rate >75% of translocated urchins after 20 days. The implications of these findings are important

for managerial decisions, corroborate previous studies, and lay the ground work for additional

urchin relocation experiments on reef habitats. Providing habitat to maintain D. antillarum

populations could enhance both coral and urchin restoration success. Stocking restoration sites

with proper coral cover 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. In return, increased D. antillarum densities at the restoration site can decrease

overgrowth of macroalgae (Sammarco, 1982; Burdick, 2008; Hernandez et al., 2008) and

increase growth of the surrounding coral (Edmunds and Carpenter, 2001; Idjadi et al., 2010)

providing opportunity for reversal of the currently macroalgae dominated ecosystem.

Applications for Management

Our third hypothesis- With multiple restoration efforts taking place, suitable habitat

currently exists to support D. antillarum relocation- was supported with the objectives to: (3a)

create a tool for managers charged with implementing coral restoration programs in conjunction

with D. antillarum recovery programs, and (3b) predict and identify current restoration efforts

that may facilitate the recovery of urchins by selecting suitable translocation reefs with

appropriate coral infrastructure (coverage) Through these objectives, we were able to prove there

are sites that currently exist to begin coral-urchin restoration efforts and greater abundances of

sites that will become available over the next three years as A. cervicornis clusters continue to

grow.

Objective 3a: predict and identify current restoration efforts that may facilitate the

recovery of urchins by selecting suitable translocation reefs with appropriate coral

infrastructure (coverage).

Acropora cervicornis has been the focus of many restoration groups, however, a diverse

array of other coral species can greatly increase structural complexity on coral reefs (Alvarez-

72

Filip et al., 2011). Current efforts to propagate A. palmata, bouldering corals, and pillar corals

are underway, yet still require ample amount of time and replication to maximize success of

these programs (NOAA Workshop, 2016). In the meantime, the proven success of A. cervicornis

restoration efforts can provide an interim habitat for juvenile D. antillarum to be moved to

restoration sites either from in situ collection sites, or ex situ hatcheries.

A Diadema Retention Model (DRM) was created to predict urchin retention at restoration

sites when relocated to coral clusters that have properties of suitable coral density and ground

cover. The model allowed us to input desired urchin retention rates and output a specific cluster-

biomass needed to promote retention. Using this information, we identified sites where habitat

currently exists to support preliminary urchin relocations as well as which sites would become

available over the next 3 years as coral growth continues within clusters. As coral-urchin

restoration is scaled up, additional data can be incorporated into the DRM for enhanced

predictive results of a successful coral-urchin restoration site, however this tool currently

provides a baseline of sites where combined restoration efforts may begin.

Currently, less than 30% of coral clusters outplanted to reefs have sufficient A.

cervicornis cover to promote successful urchin retention based on our studies. However, in years

2018, 2019, and 2020, this greatly increases to 69.99, 99.89, and 100% respectively under

current outplanting and site restoration regimes. Clusters achieve optimal cluster-biomass in

about three years, however, unknown limitations such as coral disease, coral bleaching, and other

coral stressors may result in a diminished cluster-biomass over time. Recognizing that these

stressors may contribute to a decrease in number of clusters available, it is important to consider

increasing the abundance of corals in a shorter timeframe to incorporate urchin relocations as

soon as will allow. Coral cluster-biomass could be greatly increased if larger coral size and

73

greater coral thicket densities were outplanted to restoration sites in a shorter period of time.

CRF ECO strategies showed great promise in generating sufficient coral cluster-biomass needed

for successful urchin relocations, close to only one year after outplanting. These efforts to

maximize coral cluster density can be beneficial for not only increasing the thickets of corals

placed onto reefs, but also the amount of structure they provide for urchin relocations. If D.

antillarum retention within a restoration site can be maintained, urchins can restore natural

herbivorous benefits to the reef ecosystem, consuming macroalgae, and providing suitable

habitat for coral colonies to grow and thrive (Sammarco et al., 1982; Edmunds and Carpenter,

2001; Nedimyer and Moe, 2011). In return, the increased healthy coral structure can provide

habitat for relocated urchins to increase local populations, and possibly reseed downstream reefs.

Objective 3b: create a tool for managers charged with implementing coral restoration

programs in conjunction with D. antillarum recovery programs.

Resources for management and recovery of reef species can be limited with funding,

biologically available specimens, and resources for monitoring and maintenance of the targeted

objective in reef recovery. These resources are inextricably valuable and conserving resources

used, to generate a large impact, is of the utmost importance. Diadema antillarum populations

are in slow recovery at the majority of reef sites in the Florida Keys and the Caribbean, but have

been cited as essential for the recovery of coral reefs to a coral-dominated stage (Edmunds and

Carpenter, 2001). Utilizing relocation efforts to promote coral cover should be paired with the

best possible knowledge to maximize resources and promote protection of the urchins from

predators to ensure that their persistence at a reef site is met with the objectives of reducing

macroalgae cover and promoting survival into adult spawning sizes. As coral reefs have been

flattened (Alvarez-Filip et al., 2009), there may be less area for relocated D. antillarum to reside,

and reefs may hold carrying capacities for urchins based off of available complexity of the reef

74

(Diadema Workshop, 2017). Added structure to the reef through A. cervicornis restoration

efforts can increase the structural complexity of a reef short to long-term by adding coral cover

that was not previously existent. As shown above, a certain threshold of coral cluster-biomass

can increase the percentage of urchins retained when relocated to a coral cluster. Identifying sites

that currently exist to promote the greatest maximum urchin retention can provide managers and

practitioners with valuable knowledge to ensure that their efforts are successful and a wise use of

resources.

The DRM is a valuable tool for managers looking to incorporate urchins into their current

restoration programs, and holds important implications for restoration protocols at present. The

results of this study show that the goal of restoration programs should be to increase the amount

of coral on the reef, but also to increase the structural complexity of reefs that are restored.

Increasing complexity of the reefs can increase the number of dynamic interactions needed to

promote not only the recovery of D. antillarum, but also bring back dynamic reef interactions

that can promote overall normal function of reefs pre-1970.

Future Work

To effectively stimulate reversal back to coral dominated reefs, management strategies

are highly encouraged to include holistic species recovery together. Some sites currently exist to

begin preliminary D. antillarum relocation trials, however, to support greater urchin retention,

the potential reef sites identified as coming online in 2018-2020 should be the subject of greater

maintenance and care to ensure they can support D. antillarum relocation at this critical time in

the recovery of the Florida Keys Reef Tract.

In addition to future sites that were identified as potential successful urchin relocation

sites, it is important to consider the scaling up of coral restoration efforts so that greater areas of

habitat may return to structural complexity suitable to promote dynamic interactions of other reef

75

organisms. ECO methods to prompt greater coral cover in shorter amounts of time may be

favorable for increasing habitat necessary to promote D. antillarum urchin recovery and should

be tested in initial urchin relocation trials.

Coral density experiments should be replicated on reefs to evaluate urchin retention at

sites with greater fish abundance and variable substrate complexities. Underlying reef rugosity

should be consideration in urchin relocation sites and one that is investigated further, and

possibly added to the DRM. Sites with greater reef rugosity may promote greater urchin

retention rates where urchins can hide in crevices, and use complexity provided from A.

cervicornis to forage at night, where coral can provide greater protection from local predatory

fish and invertebrates.

These implications are important for successful local relocation and reef recovery efforts,

however, to generate an even larger impact, principles of large-scale reef connectivity should

also be considered. Future work to understand the connectivity of reefs and D. antillarum larval

dispersal may offer strategies for selecting reef sites where coral-urchin restoration efforts are

most beneficial across large-scale regions, ultimately contributing to a naturally functioning reef

habitat which may promote natural resiliency giving reef ecosystems a chance to overcome

effects of hysteresis and other global stressors including climate change, ocean acidification, or

increasing storm intensities. Coral reefs are complex and dynamic ecosystems with complex

factors that can affect their health and recovery. As such, greater attention is needed for the

development of recovery strategies that incorporate an equally complex and dynamic restoration

approach. The combination of programs that incorporate recovery of multiple niche groups, such

as coral restoration and urchin relocations, is one solution that may initiate the return of coral to

degraded reef communities and enhance reef recovery rates.

76

APPENDIX A: CORAL CLUSTERS AVAILABLE FOR URCHIN RELOCATIONS FROM

2017-2020

Table A-1: Clusters Present for Each Category of Cluster-Biomass at Restoration Sites in 2017.

CRF Restoration Site Good

(75-84.9%)

Better

(85-94.9%)

Best

(>95%) Total Clusters at Site

AC1 Patch Reef (ECO) 20 16 0 37


Alligator Reef 0 34 20 147

Bighead1 Patch (ECO) 0 0 0 32

Carysfort Reef 4 0 30 317

CNC2 0 0 0 125

Conch Reef 30 5 19 211

Crocker Reef 0 20 0 90

Davis Reef 0 0 20 100

DMPK1 Patch (ECO) 0 0 42 42

East Turtle Shoals 0 0 46 86

French Reef 0 0 20 20

Grecian Rocks Reef 0 0 30 63

Key Largo Dry Rocks 0 0 61 91

Little Conch Ledge 0 0 10 100

Little Conch Shallow 0 0 30 90

M1 Patch Reef 0 0 0 10






Marker 32 0 0 0 40

Molasses Reef 0 3 46 168

Neon Patch (ECO) 0 0 0 135

Nine Foot Stake 0 0 0 97

North Dry Rocks 0 0 30 133

Pickles 5 (ECO) 0 0 0 4

Pickles Reef 57 15 37 297

Pickles Inshore Site 1 (ECO) 0 0 0 105

Pipe Patch (ECO) 0 0 0 30

PK Spike (ECO) 0 0 2 2

Snapper Ledge 0 5 72 260

Sponge 2015 Patch (ECO) 0 0 0 100

U32 Patch Reef (ECO) 35 0 0 94



West Turtle Shoals 0 0 51 91

Western Sambo 0 0 0 40

White Bank Reef 0 0 61 101

Total 246 153 627 3636

77



(75-84.9%)

Better

(85-94.9%)

Best







CNC2 0 0 100 125

Conch Reef 26 2 177 211
















Marker 32 0 0 0 40


















Total 130 163 2425 3636

78



(75-84.9%)

Better

(85-94.9%)

Best




Alligator Reef 147 0 147



CNC2 125 0 0 125

















Marker 32 40 0 0 40











Sponge 2015 Patch (ECO) 100 0 100 100







Total 3499 66 712 3636

79



(75-84.9%)

Better

(85-94.9%)

Best







CNC2 0 0 125 125

















Marker 32 0 0 40 40


















Total 0 0 3636 3636

80

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BIOGRAPHICAL SKETCH

Kayla graduated spring of 2017 with a Master of Science in Fisheries and Aquatic

Sciences. Kayla was a distance student residing in Key Largo, Florida where she was employed

at Coral Restoration Foundation.

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