Rob Ruzicka, Kate Semon, Michael Colella, Vanessa Brinkhuis, Jim Kidney, Jennifer Morrison, 

Kevin Macaulay, James W. Porter, Meredith Meyers, Mary Christman, and James Colee

 The United States Environmental Protection 

National Oceanic and Atmospheric Administration/National Ocean Service

NOAA/NOS

2009 Annual Report 

Rob Ruzicka, Kate Semon, Michael Colella, Vanessa Brinkhuis, Jim Kidney, Jennifer Morrison, 

James W. Porter, Meredith Meyers, Mary Christman, and James Colee

Funded by 

United States Environmental Protection Agency 

ceanic and Atmospheric Administration/National Ocean Service

USEPA grant award X7-97468002-8 

NOAA/NOS MOA-2001-683 (Amendment No. 004)/7477

Rob Ruzicka, Kate Semon, Michael Colella, Vanessa Brinkhuis, Jim Kidney, Jennifer Morrison, 

James W. Porter, Meredith Meyers, Mary Christman, and James Colee

ceanic and Atmospheric Administration/National Ocean Service

683 (Amendment No. 004)/7477 
2009 CREMP Final Report 2 F2345-02-10-F & F2635-06-09-F

2009 CREMP Annual Report 

Principal Investigators 

Rob Ruzicka, Fish and Wildlife Research Institute, 100 8th Avenue S.E., St. Petersburg, FL, 

33701 Telephone: 727-896-8626; Email: Rob.Ruzicka@MyFWC.com

Dr. Kathleen Semon, Fish and Wildlife Research Institute, 100 8th Avenue S.E., St. Petersburg, 

FL, 33701, Telephone: 727-896-8626; Email: Kathleen.Semon@ MyFWC.com

CREMP Team and Collaborators 

Michael Colella, Fish and Wildlife Research Institute, 100 8th Avenue S.E., St. Petersburg, FL, 

33701, Telephone: 727-896-8626; Email: Mike.Colella@ MyFWC.com

Vanessa Brinkhuis, Fish and Wildlife Research Institute, 100 8th Avenue S.E., St. Petersburg, 

FL, 33701, Telephone: 727-896-8626; Email: Vanessa.Brinkhuis@ MyFWC.com

Jennifer Morrison, Fish and Wildlife Research Institute, 100 8th Avenue S.E., St. Petersburg, FL, 

33701, Telephone: 727-896-8626; Email: Jennifer.Morrison@MyFWC.com

Kevin Macaulay, Fish and Wildlife Research Institute, 100 8th Avenue S.E., St. Petersburg, FL, 

33701, Telephone: 727-896-8626; Email: Kevin.Macaulay@MyFWC.com

Jim Kidney, South Florida Regional Laboratory, 58117 Overseas Highway, Marathon, FL 

33050, Telephone: 305-289-2332; Email: Jim.Kidney@ MyFWC.com

Dr. Jim Porter, University of Georgia, 104 East Green Street, Athens, GA 30602, 

Telephone: 706-542-3410; Email: Jporter@uga.edu

Meredith Meyers, University of Georgia, 104 East Green Street, Athens, GA 30602, 

Telephone: 706-542-3410; Email: Mkmeyers@uga.edu

Dr. Mary Christman, University of Florida, 406 McCarty Hall C, P.O. Box 110339, Gainesville, 

FL 32611, Telephone: 352-392-3724; Email: Mcxman@ufl.edu

James Colee, University of Florida, 406 McCarty Hall C, P.O. Box 110339, Gainesville, FL 

32611, Telephone: 352-392-3724; Email: colee@ufl.edu

Suggested Citation 

Ruzicka R, Colella M, Semon K, Brinkhuis V, Morrison J, Kidney J, Porter J, Meyers M, 

Christman M, and Colee J. 2010. CREMP 2009 Final Report. Fish & Wildlife Research 

Institute/Florida Fish & Wildlife Conservation Commission. Saint Petersburg, FL. 110 pp 

Acknowledgements 

CREMP monitoring is conducted under the Florida Keys National Marine Sanctuary Permit 

FKNMS-2007-121-A1. 
2009 CREMP Final Report 3 F2345-02-10-F & F2635-06-09-F

TABLE OF CONTENTS 

EXECUTIVE SUMMARY ................................................................................................................... 5

INTRODUCTION ................................................................................................................................. 7

METHODS .......................................................................................................................................... 10

SITE SELECTION AND STATION SETUP ................................................................................................ 10 

STATION SPECIES INVENTORY (SSI)................................................................................................... 13 

VIDEO TRANSECTS ............................................................................................................................ 13 

BIOERODING (CLIONAID) SPONGE SURVEY ........................................................................................ 14 

STATISTICAL & DESCRIPTIVE ANALYSES ........................................................................................... 14 

RESULTS & DISCUSSION ................................................................................................................ 16

2007 VS. 2008 .................................................................................................................................... 16 

All Benthic Cover Variables ........................................................................................................... 16

Macroalgal cover .......................................................................................................................... 17

Octocoral Cover ............................................................................................................................ 18

Sponge Cover ................................................................................................................................ 18

Stony Coral Cover ......................................................................................................................... 18

STONY CORAL SPECIES RICHNESS ..................................................................................................... 19 

STONY CORAL CONDITION ................................................................................................................ 23 

LONG-TERM TRENDS ......................................................................................................................... 25 

All Benthic Cover Variables ........................................................................................................... 25

Stony Coral Cover ......................................................................................................................... 28

Macroalgal cover .......................................................................................................................... 32

Octocoral Cover ............................................................................................................................ 34

Sponge Cover ................................................................................................................................ 36

Sentinel Stony Coral Species .......................................................................................................... 38

Montastraea annularis complex ..................................................................................................... 40

Montastraea cavernosa .................................................................................................................. 42

Colpophyllia natans ....................................................................................................................... 43

Porites astreoides .......................................................................................................................... 44

Clionaid Cover .............................................................................................................................. 47

CONCLUSIONS.................................................................................................................................. 51

LITERATURE CITED ....................................................................................................................... 53

APPENDICES ..................................................................................................................................... 63

APPENDIX 1. PERCENT COVER VALUES FOR MAJOR BENTHIC TAXA CATEGORIES FOR ALL YEARS AT ALL 

REGION*HABITAT GROUPINGS. ........................................................................................................... 63 

APPENDIX 2. PERCENT COVER VALUES FOR MAJOR BENTHIC TAXA CATEGORIES FOR ALL YEARS AT ALL 

CREMP SITES. ................................................................................................................................... 67 

APPENDIX 3. PERCENT COVER VALUES FOR FIVE SENTINEL CORAL SPECIES FOR ALL YEARS AT ALL 

REGION*HABITAT GROUPINGS. ........................................................................................................... 78 
2009 CREMP Final Report 4 F2345-02-10-F & F2635-06-09-F

APPENDIX 4. PERCENT COVER VALUES FOR FIVE SENTINEL CORAL SPECIES FOR ALL YEARS AT ALL 

CREMP SITES. ................................................................................................................................... 82 

APPENDIX 5. LONG-TERM TRENDS IN STONY CORAL COVER 1996 TO 2008. ......................................... 92 

APPENDIX 6. LONG-TERM TRENDS IN STONY CORAL COVER FROM 1999 TO 2008. ............................... 93 

APPENDIX 7. LONG-TERM TRENDS IN MACROALGAL COVER FROM 1996 TO 2008. ............................... 94 

APPENDIX 8. LONG-TERM TRENDS IN MACROALGAL COVER FROM 1999 TO 2008. ............................... 95 

APPENDIX 9. LONG-TERM TRENDS IN OCTOCORAL COVER FROM 1996 TO 2008. .................................. 96 

APPENDIX 10. LONG-TERM TRENDS IN OCTOCORAL COVER FROM 1999 TO 2008. ................................ 97 

APPENDIX 11. LONG-TERM TRENDS IN SPONGE COVER FROM 1996 TO 2008. ....................................... 98 

APPENDIX 12. LONG-TERM TRENDS IN SPONGE COVER FROM 1999 TO 2008. ....................................... 99 

APPENDIX 13. LONG-TERM TRENDS IN MONTASTRAEA ANNULARIS COMPLEX COVER FROM 1996 TO 2008.

 ........................................................................................................................................................ 100 

APPENDIX 14. LONG-TERM TRENDS IN MONTASTRAEA ANNULARIS COMPLEX COVER FROM 1999 TO 2008.

 ........................................................................................................................................................ 101 

APPENDIX 15. LONG-TERM TRENDS IN MONTASTRAEA CAVERNOSA COVER FROM 1996 TO 2008. ........ 102 

APPENDIX 16. LONG-TERM TRENDS IN MONTASTRAEA CAVERNOSA COVER FROM 1999 TO 2008. ........ 103 

APPENDIX 17. LONG-TERM TRENDS IN COLPOPHYLLIA NATANS COVER FROM 1996 TO 2008. .............. 104 

APPENDIX 18. LONG-TERM TRENDS IN COLPOPHYLLIA NATANS COVER FROM 1999 TO 2008. .............. 105 

APPENDIX 19. LONG-TERM TRENDS IN PORITES ASTREOIDES COVER FROM 1996 TO 2008. ................. 106 

APPENDIX 20. LONG-TERM TRENDS IN PORITES ASTREOIDES COVER FROM 1999 TO 2008. ................. 107 

APPENDIX 21. LONG-TERM TRENDS IN SIDERASTREA SIDEREA COVER FROM 1996 TO 2008. ................ 108 

APPENDIX 22. LONG-TERM TRENDS IN SIDERASTREA SIDEREA COVER FROM 1999 TO 2008. ................ 109 

APPENDIX 23. LONG-TERM TRENDS IN CLIONA DELITRIX COVER FROM 2001 TO 2008. ....................... 110
2009 CREMP Final Report 5 F2345-02-10-F & F2635-06-09-F

EXECUTIVE SUMMARY 

The purpose of the Coral Reef Evaluation and Monitoring Project (CREMP) is to monitor the 

status and trends of selected reefs in the Florida Keys National Marine Sanctuary (FKNMS). 

CREMP assessments have been conducted annually at fixed sites since 1996 and data collected 

provides information on the temporal changes in benthic cover and diversity of stony corals and 

associated marine flora and fauna. The core field methods continue to be underwater 

videography and timed coral species inventories. Findings presented in this report include data 

from 109 stations at 37 sites sampled from 1996 through 2008 in the Florida Keys and 1999 

through 2008 in the Dry Tortugas. The report describes the annual differences (between 2007 

and 2008) in the percent cover of major benthic taxa (stony corals, octocorals, sponges, and 

macroalgae), mean coral species richness and the incidence of stony coral conditions. 

Additionally, it examines the long-term trends of the major benthic taxa, five coral species that 

are the most spatially abundant (Montastraea annularis complex, Montastraea cavernosa,

Colpophyllia natans, Siderastrea siderea, and Porites astreoides) and the clionaid sponge, 

Cliona delitrix. 

In 2008, mean benthic cover values in the Florida Keys (N=97 stations) were 13.6% for 

octocorals, 12.6% for macroalgae, 6.6% for stony corals and 2.2% for sponges. In the Dry 

Tortugas (N=12 stations) cover was 8.7% for octocorals, 12.3% for macroalgae, 10.3% for stony 

corals and 1.4% for sponges. From 2007 to 2008 the cover of octocorals and macroalgae 

increased in the Florida Keys while cover remained similar for stony corals and sponges. Cover 

for all taxa remained similar between years in the Dry Tortugas. No significant differences in 

mean coral species richness (number of species per station) were observed between 2007 and 

2008 in the Florida Keys and Dry Tortugas. The long-term trends of the four major benthic taxa 

varied. Throughout the Florida Keys, stony coral and sponge cover has significantly declined 

from 1996 to 2008. During this time octocoral cover has significantly increased, and no trend 

has been observed for macroalgae. The trends within the Dry Tortugas mostly mirror those 

occurring in the Florida Keys; the lone divergence being octocoral cover has decreased there. 

The demise of stony corals is reflected by the declines in cover of four of the five most spatially 

prominent species. Overall trends in cover for M. annularis complex, M. cavernosa, C. natans, 

and P. astreoides were all negative. Only the trend for S. siderea indicated cover for this species 

remained similar. To date, the mass bleaching event during the 1997/1998 El Nino Southern 

Oscillation (ENSO) resulted in the most substantial reductions in cover. However, during the 

last decade, after effects of the 1997/1998 ENSO had dissipated, the overall trends remain 

negative for these four species. This is mostly attributed to a regional effect of the Lower Keys 

and the Dry Tortugas, whereas the sites in the Middle or Upper Keys do not exhibit this trend. 

It has been widely reported that following the disappearance of stony corals on reefs in the 

Caribbean and Western Atlantic large-scale shifts to macroalgae or sponge dominance have 

occurred. While CREMP has recorded single year spikes or ephemeral blooms after major 

disturbances (e.g. after the 1997/1998 ENSO and 2005 hurricane season) a prolonged shift 

towards increased macroalgal cover has not occurred. Likewise, the slow but steady decreases in 

sponge cover as well as decreases in Cliona delitrix suggest that sponges may be vulnerable to 

the same stressors that affect stony corals. Instead, CREMP findings support a transition to 

octocorals at many sites in the Florida Keys. The transition is most apparent in the shallow 

forereef where Elkhorn coral, Acropora palmata, and the blade fire coral, Millepora complanata, 
2009 CREMP Final Report 6 F2345-02-10-F & F2635-06-09-F

were previously abundant. Although the demise of A. palmata mostly predates the 

implementation of CREMP, the mortality of the few remnant A. palmata colonies on shallow 

forereefs during or shortly after the 1997/1998 ENSO has been followed by significant increases 

in octocoral cover in this habitat. Considering that octocoral cover has rebounded twice 

following major disturbances while little or no recovery of stony coral cover has occurred (even 

during intervals lacking major perturbations), reefs in the Florida Keys have likely entered into a 

new alternate state where octocorals are replacing stony corals as the dominant taxa. 

 
2009 CREMP Final Report 7 F2345-02-10-F & F2635-06-09-F

INTRODUCTION 

The purpose of the Coral Reef Evaluation and Monitoring Project (CREMP) is to monitor the 

status and trends of selected coral reefs, patch reefs, and hardbottom areas in the Florida Keys 

National Marine Sanctuary (FKNMS). CREMP was initiated in 1995 as part of the Water 

Quality Protection Program (WQPP) which mandated a comprehensive monitoring program be 

established for corals, seagrasses and water quality in the FKNMS. CREMP assessments have 

been conducted annually at 34 fixed sites since 1996 and data collected provide information on 

the temporal changes in spatial cover and diversity of stony corals and associated marine flora 

and fauna. CREMP is one of the longest tenured monitoring programs in the State of Florida and 

has been instrumental in documenting how a variety of perturbations have resulted in both 

widespread and localized changes to benthic community composition in the FKNMS. Some 

changes are due to direct disturbances that have occurred since CREMP was initiated, but many 

are likely the result of centuries’ old anthropogenic modifications to the South Florida 

environment well before the designation of the Sanctuary in 1990 or the inception of CREMP in 

1995. 

The Florida Keys contains a small year-round resident population of about 70,000 people with a 

present day economy that is heavily reliant on tourism and fisheries. The Florida Reef Tract is 

considered the most widely used and heavily exploited coral reef ecosystem on the planet 

(Kruczynski 1999; Johns et al. 2001; U.S. Census 2010) with human exploitation dating back to 

the 1500’s. As early as the mid 1800’s, Key West had already become one of the most populated 

cities in Florida due to a prolific commercial fishing industry for conch, sponges, and turtles. 

Green turtle and sponge fisheries were major marine exports during the late 19th and early 20th

century but overexploitation of both fisheries led to their collapse by the 1920’s. Since then 

other commercial and recreational fisheries targeting a wide range of fin- and shellfish (e.g. 

grouper, snapper, lobster, and stone crab) have expanded. Through the 1900’s intense fishing 

pressure on many of these resources has forced management agencies to establish catch limits 

and enact fishery closures (e.g. Goliath grouper, Nassau grouper and Queen Conch). Today 

fishing continues for many species, however there is a general consensus that by the 1970’s and 

80’s many of the Florida Keys fisheries had reached their peak (Tilmant et al. 1989; Bohnsack 

and Smith 1994; Ault et al. 1998; Ault et al. 2005; McClenachan 2008). 

Major population increases and onshore development began around the turn of the 20th century 

with the completion of the Florida East Coast Railway and, in 1912, the completion of the 

Overseas Railroad that connected Key West to the mainland. Later, in 1938, the Overseas 

Highway and a water pipeline from the mainland were constructed, setting the stage for intensive 

coastal development projects in the Keys during the following decades. To meet the demands of 

a rapidly growing population and the expanding agricultural economy of South Florida, 

hydrographic features around the region were permanently changed via the building of dams, 

canals, and miles of impervious roadways. During this time, the removal of large areas of 

mangroves, wetlands, and seagrasses severely reduced the volume and altered the chemistry of 

water flowing out of the Florida Everglades into Florida Bay and eventually across the Florida 

Reef Tract (Light and Dineen 1994; McIvor 1994; Fourqurean and Robblee 1999). As Key West 

became a center for tourism in the 1980’s large cruise ships began visiting the island and now 

bring over 350,000 visitors annually. 
2009 CREMP Final Report 8 F2345-02-10-F & F2635-06-09-F

Deteriorating water quality in South Florida and the Florida Keys was first recognized in the 

1970’s. In addition to the altered flow and chemistry of water leaving the Everglades, sewage 

pollution from local development is often emphasized as the leading contributor to declining 

water quality in the Florida Keys (Lapointe et al. 2004). In the 1980’s and 1990’s nutrients and 

bacteria associated with storm water run-off and waste water treatment were recognized as 

threats to the ecosystem (Lapointe 1990; Lapointe and Clark 1992; Lapointe and Matzie 1996; 

Szmant and Forrester 1996; Porter et al. 1999; Paul et al. 2000; Lipp et al. 2002; Griffin et al. 

2003). Wastewater entering nearshore waters caused algal blooms (Lapointe and Clark 1992; 

Lapointe et al. 1994; Lapointe 1997) and led to physiological problems in several species, most 

notably the reproductive failure of nearshore Queen Conch, an iconic and commercially 

important species (Glazer and Quintero 1998). The presence of fecal coliform bacteria has been 

detected in nearshore waters and corals (Paul et al. 1997; Lipp et al. 2002) and has been linked to 

coral diseases (Patterson et al. 2002; Sutherland and Ritchie 2004). Although coral diseases 

were first documented in Florida in the 1970’s, the higher prevalence of some diseases in recent 

decades may be due to increased nutrient availability in conjunction with a greater frequency of 

warm sea surface temperature anomalies (Richardson 1998; Bruno et al. 2003; Lesser et al. 

2007). In addition, it is believed that eutrophication increases the severity and duration of coral 

disease outbreaks and lowers the resilience of coral reefs to recover after disturbance events 

(McCook 2001; Bruno et al. 2004). 

Historically, the shallow forereefs of the Florida Keys have caused numerous shipwrecks. Many 

of the most prominent reefs of the FKNMS are named after vessels that collided with them. As 

recently as the 1980’s, a number of groundings occurred that destroyed large sections of reef. 

Most notably, within an 18-day period in 1989, three large ship groundings occurred in iconic 

areas of the Florida Keys. The M/V Alec Owen Maitland, a 130 meter freighter, crushed over 

1.5 km2

 of coral habitat near Carysfort Reef, the M/V Elpis, a 143 meter freighter, completely 

destroyed 2.5 km2 of coral habitat near Elbow Reef, and the M/V Mavro Ventranic, another large 

freighter, decimated the reef environment near Pulaski Shoal. The cumulative impacts of these 

groundings, in conjunction with reports of deteriorating water quality throughout the region, 

were the primary factors leading to the establishment of the FKNMS. 

Florida’s coral reefs are also subject to naturally occurring disturbances. Florida is influenced by 

the El Nino Southern Oscillation (ENSO) in the Pacific Ocean, which brings periods of low 

winds, little rainfall and extremely high water temperatures to the Florida Keys. Extensive 

bleaching of Florida’s coral reefs occurred at least twice, in 1983 and 1987, prior to CREMP 

monitoring, although minor incidences of bleaching were observed during other years (Aronson 

and Miller 2005). CREMP also documented a severe bleaching event in 1997/1998 that resulted 

in substantial mortality (Somerfield et al. 2008). While coral communities can recover from 

mild bleaching events, the frequency and intensity of these events has increased during the last 

two decades (Hoegh-Guldberg 1999) and is expected to intensify even further (Baker et al. 

2008). While most bleaching events are associated with thermal stress during ENSO years, other 

factors such as ultraviolet radiation, turbidity, and eutrophication can trigger a bleaching 

response (Glynn and Colgan 1993; Brown 1997; Fitt et al. 2001; Manzello et al. 2007; Wagner et 

al. 2010). Conversely, prolonged periods of low water temperatures have been observed due to 

Florida’s subtropical location. Occasional influxes of extremely cold arctic air migrates 
2009 CREMP Final Report 9 F2345-02-10-F & F2635-06-09-F

southward and can result in widespread coral mortality when water temperatures become 

depressed below the 16°C threshold for extended periods of time (Roberts et al. 1982). 

Tropical cyclone activity has always influenced South Florida, but historical records only 

reliably date back to the late 1800’s. In the early 1900’s several major hurricanes impacted the 

state of Florida including the Labor Day Hurricane of 1935, which made landfall in the Florida 

Keys as a category 5 hurricane and is still the strongest hurricane on record at the time it made 

landfall in the United States (National Hurricane Center 2010). Approximately 150 tropical 

storms or hurricanes affected the state of Florida from 1950 to 1995. Six of these storms made 

landfall as category 3 hurricane or higher in South Florida between 1950 and 1970. Only one 

storm of such intensity made landfall in the region (Hurricane Andrew, 1992) between 1970 and 

1995 (National Hurricane Center 2010). However, the 2004 and 2005 hurricane seasons were 

particularly volatile, especially in the Dry Tortugas, with five hurricanes and one tropical storm 

directly impacting the region (FWRI and DTNP 2010). Hurricanes are naturally occurring events 

from which coral reef organisms are adapted to survive and recover. Hurricanes can propagate 

new individuals via asexual reproduction of storm-induced fragments (Gilmore and Hall 1976; 

Shinn 1976; Fong and Lirman 1995; Fong and Lirman 1997; Lirman 2000) and under the right 

conditions, a rapid recovery can occur within several years after the passage of major storms 

(Shinn 1976). However, combined with other chronic stressors such as periodic disease 

outbreaks, deteriorating water quality, or elevated sea temperatures, hurricanes can have severe, 

long lasting effects on coral reef habitats. 

The interconnected oceanography of the Caribbean basin and Western Atlantic threatens the 

Florida Reef Tract with natural borne marine pandemics. The Caribbean-wide mass mortality of 

the long-spined sea urchin, Diadema antillarum, affected the Florida Reef Tract in the summer 

of 1984 after the disease was first observed at Panama in January 1983 (Lessios et al. 1984). 

Prevailing surface currents (e.g. Yucatan & Gulf streams) transported the disease rapidly 

throughout much of the Caribbean and the mortality was estimated at 95% of the population in 

nearly all locations (Lessios et al. 1984). Similarly, white-band disease (WBD) has been a 

primary cause of Acropora mortality in Florida (Precht and Aronson 1997; Precht and Miller 

2006), as well as in the Caribbean and Western Atlantic (Aronson and Precht 2001). It was first 

documented in the U.S. Virgin Islands (Robinson 1973) but had spread to the Florida Reef Tract 

by the late 1970’s and early 1980’s. Mass mortalities of spatially-dominant acroporid species (A. 

palmata and A. cervicornis) due to WBD were observed at several reefs, and consequently, cover 

of these species in Florida was estimated to have declined >90% by the late 1990’s (Miller et al. 

2002; Precht et al. 2002) prior to CREMP. 

Coral reef resources in the Florida Keys were first publicly recognized as degraded as early as 

1960 when the John Pennekamp Coral Reef State Park was established. Additional management 

efforts were initiated when the public began to call for more protection. The Key Largo National 

Marine Sanctuary, adjacent to Pennekamp Park, was established in 1975, and the Looe Key 

National Marine Sanctuary was established in 1981. Although these two marine sanctuaries 

encompassed only a small fraction of the Florida Keys’ marine environment, they laid the 

foundation for broader protection of the resource. On November 16, 1990, after a century of 

heavy exploitation, coastal development, and numerous catastrophic ship groundings, Congress 

passed a bill that established the Florida Keys National Marine Sanctuary and Protection Act. 
2009 CREMP Final Report 10 F2345-02-10-F & F2635-06-09-F

The Act designated approximately 2,800 square nautical miles of state and federal waters in the 

Keys as the Florida Keys National Marine Sanctuary. The FKNMS included the previously 

established marine protected areas, state parks, and wildlife refuges but also helped designate 

another 24 special protection areas (SPAs) and four ecological reserves as no-take areas. In 

conjunction with the EPA WQPP, the FKNMS has also built advanced wastewater treatment 

facilities to remediate deteriorating water quality. 

CREMP was established in 1996 with the primary goal of examining temporal changes in coral 

benthic communities in the FKNMS and to provide information to management organizations on 

the status of coral reef resources. Reports provided by CREMP have assisted in evaluating the 

efficacy of the SPAs and have aided in the permit review process for coastal construction 

projects and collection of corals in the FKNMS. Since monitoring was initiated, CREMP has 

documented changes in benthic community composition due to both large-scale and localized 

disturbances in the environment. These records include widespread coral mortality that occurred 

in association with the 1997/1998 El Nino and following multiple hurricanes that impacted the 

region in 2004 and 2005. More localized events include the 2001 harmful algal bloom in the 

Gulf of Mexico/Florida Bay area that dramatically reduced coral cover at two CREMP sites as 

well as a disease outbreak which reduced the cover of stony corals at one site in the Dry Tortugas 

between 2001 and 2003. While it’s difficult to distinguish what proportion of the changes 

observed by CREMP is directly due to present day stressors as opposed to the anthropogenic 

disturbances propagated over the last century, CREMP has predominately observed a decline in 

the cover and diversity of corals in the FKNMS. CREMP has observed multiple years where 

coral cover remained similar, but little recovery was observed during these intervals. This report 

describes the long-term trends observed by CREMP and provides insight as to how benthic 

community structure in the Keys has changed since 1996 and may be shaped in the future. 

Additionally, the report outlines the changes in benthic cover and coral diversity between 2007 

and 2008. The results presented in this report include data from 34 sites sampled from 1996 

through 2008 in the Florida Keys, and from three sites sampled from 1999 through 2008 in the 

Dry Tortugas. 

METHODS 

Site Selection and Station Setup 

Sampling locations were initially chosen in 1994 using a stratified random sampling procedure 

(US EPA EMAP). Forty reef sites were selected within the FKNMS and four permanent stations 

were installed at each site in 1995. Each station is approximately 22m long x 2m wide and is 

demarcated by two permanent station markers. Site selection was stratified across four habitat 

types defined as nearshore hardbottom, patch reefs, shallow spur and groove reefs (~3 to 6m 

depth), and deep spur and groove reefs (~10 to 20m depth). Although sites were randomly 

selected, station placement was designed to monitor specific stony coral populations within the 

selected habitats (e.g. stands of Acropora palmata on shallow forereef habitats). In 1999, an 

additional 12 stations at three sites were installed in the Dry Tortugas. As of 1999, CREMP 

monitoring included a total of 172 stations at 43 sites. 

A statistical analysis of CREMP’s experimental design assisted in streamlining the sampling 

effort by decreasing the number of stations sampled beginning in 2001. As a result of the 
2009 CREMP Final Report 11 F2345-02-10-F & F2635-06-09-F

reduction, all CREMP sites presently consist of two to four monitoring stations. Also starting in 

2001, hardbottom sites were removed from the effort due to consistently low coral cover. 

Monitoring of additional hardbottom locations was discontinued in subsequent years and only 

one site has been retained (Content Keys). Content Keys is located in an area commonly 

referred to as the backcountry, the gulf-side area of the Lower Florida Keys region, and has been 

more appropriately re-designated as a backcountry patch reef. The current CREMP sampling 

effort now includes 109 stations at 37 sites. Stratified by habitat, CREMP monitors six stations 

at two backcountry patch reef sites, 10 patch reef sites with 30 stations, 12 shallow reef sites with 

39 stations and 13 deep reef sites with 34 stations (Table 1). 

The core field methods for CREMP continue to be the underwater videography of three transects 

and a timed station species inventory at each sampling station (Figure 1). Image analysis from 

video provides relative estimates of benthic community composition and traditional station 

species inventories provide information on coral species richness, presence/absence of disease 

and bleaching, and Diadema abundance. 

Figure 1: CREMP sites consist of two to four monitoring stations delineated by permanent markers. Stations are 

approximately 2m x 22m and are generally perpendicular to the shoreline. Three video (benthic survey) transects, a 

station species inventory (SSI), and a clioniad sponge survey are conducted annually at each station. 
2009 CREMP Final Report 12 F2345-02-10-F & F2635-06-09-F

Table 1. Location, reef type, and number of stations sampled at 34 CREMP sites monitored from 1996 through 2008 

in the Florida Keys and three sites from 1999 through 2008 in the Dry Tortugas. 

Region Habitat/Reef Site No. of Stations 

Dry Tortugas Deep Bird Key Reef 4 

Dry Tortugas Deep Black Coral Rock 4 

Dry Tortugas Patch White Shoal 4 

Lower Keys Backcountry Patch Content Keys 3 

Lower Keys Backcountry Patch Smith Shoal 3 

Lower Keys Deep Eastern Sambo Deep 3 

Lower Keys Deep Looe Key Deep 3 

Lower Keys Deep Rock Key Deep 2 

Lower Keys Deep Sand Key Deep 3 

Lower Keys Deep Western Sambo Deep 3 

Lower Keys Shallow Eastern Sambo Shallow 3 

Lower Keys Shallow Looe Key Shallow 3 

Lower Keys Shallow Rock Key Shallow 4 

Lower Keys Shallow Sand Key Shallow 3 

Lower Keys Shallow Western Sambo Shallow 3 

Lower Keys Patch Cliff Green 2 

Lower Keys Patch Jaap Reef 3 

Lower Keys Patch West Washer Women 2 

Lower Keys Patch Western Head 3 

Middle Keys Deep Alligator Deep 2 

Middle Keys Deep Sombrero Deep 2 

Middle Keys Deep Tennessee Deep 2 

Middle Keys Shallow Alligator Shallow 3 

Middle Keys Shallow Sombrero Shallow 4 

Middle Keys Shallow Tennessee Shallow 3 

Middle Keys Patch Dustan Rocks 3 

Middle Keys Patch West Turtle Shoal 4 

Upper Keys Deep Carysfort Deep 2 

Upper Keys Deep Conch Deep 2 

Upper Keys Deep Molasses Deep 3 

Upper Keys Shallow Carysfort Shallow 3 

Upper Keys Shallow Conch Shallow 3 

Upper Keys Shallow Grecian Rocks 4 

Upper Keys Shallow Molasses Shallow 3 

Upper Keys Patch Admiral 4 

Upper Keys Patch Porter Patch 3 

Upper Keys Patch Turtle 2 
2009 CREMP Final Report 13 F2345-02-10-F & F2635-06-09-F

Figure 2. Location of 34 CREMP sites sampled from 1996 through 2008 in the Florida Keys and three sites sampled 

from 1999 through 2008 in the Dry Tortugas. 

Station Species Inventory (SSI) 

CREMP has used an established protocol since 1996. Station species inventories comprise a 

census of all stony corals (Milleporina and Scleractinia), presence/absence of coral disease or 

bleaching, and counts of Diadema antillarum within a station (44m2). Two divers conduct 

simultaneous, 20 minute inventories within the survey station and enter the data on underwater 

data sheets. Diseases are categorized as black band, white complex (including white plague, 

white band, white pox), and other (dark spot, yellow band, etc). Upon completion of the survey, 

observers compare data underwater (maximum of five minutes) to confirm the stony coral 

species richness. In the lab, data forms are reviewed, edited, and entered into the FWRI coral 

database. Data are validated using established coral monitoring quality assurance/quality control 

protocols. 

Video Transects 

Video is captured using a SONY TRV 900 to film three 40cm x 22m transects at each station 

(Figure 1). All transects are filmed at a distance of 40cm above the reef to yield a ~40cm wide 

image. A convergent laser light system aids in maintaining the camera at the appropriate 40cm 

distance above the reef surface. Prior to filming each transect, the videographer films a 
2009 CREMP Final Report 14 F2345-02-10-F & F2635-06-09-F

clapperboard that provides information on the date and location of each transect. Filming is 

conducted perpendicular to the substrate and follows a chain which marks the center of each 

transect. The videographer maintains a steady swim speed that roughly equals four meters per 

minute (~4-5min to complete one 22m transect). 

In the lab, video is captured digitally and each filmed transect is separated into about 9,000 

frames. RavenViewTM, an image processing program, is used to rejoin a small subset of the 

images to create a mosaic. Overlap between the images is set to 8% or less, resulting in 60-75 

nearly abutting images per transect. The images are then formatted for PointCount ’99 image 

analysis software. Fifteen random points are placed on each image and under each point 

observers identify benthic taxa (e.g. stony coral to species, octocoral, zoanthid, sponge, seagrass 

and macroalgae) and substrate. To ensure identification of benthic fauna is consistent across 

observers, an inter-observer file is completed by all counters prior to image analysis. After all 

images are analyzed, the data are checked for quality assurance and entered into the Microsoft 

Access database. 

Bioeroding (Clionaid) Sponge Survey 

Data on the abundance and cover of three clionaid sponges (Cliona delitrix, C. lampa, and C. 

caribbaea) are collected. Clionaid sponge data are acquired using a 1m belt transect centered on 

the video transects. A PVC pole is held perpendicular to the survey tape and the area of each 

clionaid sponge is enumerated with a 40cm x 40cm quadrat divided into 25cm2

 (5cm x 5cm) 

grids. The total number of 25m2

 grids occupied by the sponge is used to estimate the cover for 

each sponge enumerated. Data are entered into the FWRI coral database and are used to quantify 

the total clionaid spatial coverage at each station. Clionaid area (cm2

/m2) is calculated for each 

station by summing the total area of 25cm2 grids enumerated and dividing this value by total size 

of the station. 

Statistical & Descriptive Analyses 

The results presented in this report include data from 109 stations at 37 sites sampled from 1996 

through 2008 in the Florida Keys and from 1999 through 2008 in the Dry Tortugas (Table 1). 

Where appropriate, the results are presented FKNMS-wide (data pooled for all stations), by 

habitat, region, site, and for specific region*habitat groupings (e.g. Middle Keys patch reefs). 

The regions are defined as Upper Keys or “UK” (north Key Largo to Conch Reef), Middle Keys 

or “MK” (Alligator Reef to Moser Channel), Lower Keys or “LK” (Looe Key to Smith Shoal), 

and the Dry Tortugas or “DT” (Dry Tortugas to Tortugas Banks), and habitats defined as 

offshore deep reefs (OD), offshore shallow reefs (OS), patch reefs (P), and backcountry patch 

reefs (BCP). The results presented here may vary slightly from those reported in previous 

CREMP reports. Slight differences for values such as percent cover of benthic taxa or mean 

species richness are due to the differences in the total number of stations used for the analyses in 

this report compared to those of previous years. Values for all years have been recalculated using 

the current number of stations. Additionally, two sites, Content Keys and Smith Shoal, have had 

their habitat grouping re-designated to backcountry patch reef. These sites were formerly 

classified as a hardbottom and a patch reef site, respectively, and were included in the analyses 

for these habitats in previous CREMP reports. 
2009 CREMP Final Report 15 F2345-02-10-F & F2635-06-09-F

Mean species richness was calculated as the number of species per station. To determine if mean 

species richness was significantly different across all years in the Florida Keys (N=97 stations) 

and Dry Tortugas (N=12 stations), repeated measures ANOVA were used. Post hoc Tukey tests 

(Tukey-Kramer or Holm-Sidak) were used to identify significant differences between 2007 and 

2008 and across all years of the project. Long-term trends in mean species richness at the habitat 

and regional level (N=number of stations for each habitat or region) were examined using a 

generalized mixed model regression in SAS v9.2. Richness data for each station was logtransformed and trends were identified as increasing or decreasing if the slope of the curve was 

significantly differed from zero. 

Changes in benthic cover variables between 2007 and 2008 were examined for macroalgae, 

octocorals, sponges, and stony corals in SAS v9.2. To test for single year differences of mean 

percent cover within region*habitat groupings, a Kenward-Roger mixed model ANOVA was 

used with year and region*habitat groupings (and their interaction) as fixed effects and sites 

within region*habitat groupings treated as a random effect. Point count data from video 

transects were pooled for individual stations and square root-transformed. F-tests were used to 

identify differences in benthic cover means between years for each region*habitat grouping, and 

to identify differences among region*habitat groupings within years. Only data from the Florida 

Keys were analyzed because having only three Dry Tortugas sites limit the power of the analysis. 

For the Florida Keys, the analysis included 10 region*habitat groupings from 34 sites 

(N=number of stations for each region*habitat grouping). Differences in percent cover for the 

Dry Tortugas are described at the site level. 

Long-term trends in benthic cover variables (stony coral, macroalgae, octocoral, sponge), five 

sentinel stony coral species (Colpophyllia natans, Montastraea annularis, Montastraea 

cavernosa, Porites astreoides, and Siderastrea siderea), and clionaid area were examined using a 

generalized mixed model regression in SAS v9.2. Percent cover data for the benthic variables 

and sentinel coral species for each station were pooled and square root-transformed. Data for 

clionaid area (cm2

/m2) were log-transformed. Stations were nested within sites and sites nested 

within region*habitat groupings to provide long-term trend information at the site and 

region*habitat level. Regression lines were calculated from 1996-2008 to understand how the 

benthic variables and sentinel coral species cover have changed throughout the history of the 

project. Additionally, regressions were calculated from 1999-2008 to specifically assess how the 

benthic variables and sentinel coral species have responded in the aftermath of the 1997/1998 

ENSO. Long-term trends were calculated for clionaid cover beginning in 2001 when CREMP 

initiated clionaid monitoring. For all datasets, a regression analysis for each site or 

region*habitat grouping was calculated from annual percent cover values and the slope was 

identified as increasing or decreasing by t-tests demonstrating that the slope was significantly 

different from zero. The analysis included a total of 12 region*habitat groupings from 34 sites 

in the Florida Keys and three sites in the Dry Tortugas (N=number of stations for each 

region*habitat grouping). Region*habitat groupings and the corresponding number of stations 

and sites can be found in Table 1. To reduce the possibility of Type I errors due to repeating the 

same test on multiple region*habitat groupings or sites, a Bonferroni correction was used to 

adjust the p-value for identifying a trend as significantly increasing or decreasing. At the 

region*habitat level the p value was adjusted to p≤0.004 and at site level to p<0.002. 
2009 CREMP Final Report 16 F2345-02-10-F & F2635-06-09-F

RESULTS & DISCUSSION 

2007 vs. 2008 

All Benthic Cover Variables

When pooled for all stations in the Florida Keys (N=97) mean percent cover values were greatest 

for octocorals and macroalgae (Figure 3). Macroalgal cover did not significantly change 

between 2007 and 2008 (p=0.52; df=1, 90; F=0.41), remaining similar at 11.3% and 12.7%, 

respectively. Octocoral cover did significantly increase (p<0.001, df=1, 90; F=15.58), rising 

from 12.1% in 2007 to 13.6% in 2008. There was no significant change in stony coral cover 

(p=0.78; df=1, 90; F=0.07) between years with values of 6.7% in 2007 and 6.6% in 2008. 

Sponges, which make the smallest contribution to benthic cover, significantly increased between 

years (p=0.015; df=1, 90; F=6.09) rising from 1.9% in 2007 to 2.2% in 2008 (Figure 3). 

When combined for both years and pooled for all stations within each habitat type, back country 

patch reefs averaged the highest macroalgal cover (49.9%). Deep forereefs were next (14.7%) 

followed by patch reefs (8.6%) and then shallow forereef sites (6.6%). In contrast, octocoral 

cover was most abundant at patch reefs (20.3%) and shallow forereefs (12.0%). The greatest 

stony coral cover was found at patch reefs (16.0%), then shallow (3.9%) and deep forereef 

habitats (2.4%), followed by backcountry patch reefs (2.0%). Sponge cover was highest at patch 

reefs (3.6%) and offshore deep reefs (3.0%) but contributed <1.0% cover to shallow sites (0.6%) 

and backcountry patch reefs (0.8%). 

Macroalgae Octocoral Sponge Stony Coral

Mean % Cover

0

5

10

15

20

2007

2008 *

*

Figure 3. Mean (± SE) of macroalgae, octocoral, sponge and stony coral cover for 2007 and 2008 pooled for all 

stations (N = 97) in the Florida Keys. Error bars represent the standard error of the mean. Asterisks (*) denote a 

significant difference in percent cover between 2007 and 2008. 

When combined for both years and pooled for all stations within each region, macroalgae was 

the most spatially abundant benthic taxa in the Lower Keys (15.4%), while octocorals provided 
2009 CREMP Final Report 17 F2345-02-10-F & F2635-06-09-F

the greatest cover in the Middle (18.6%) and Upper Keys (17.5%). Stony coral cover was 

similar across the three regions, ranging from 5.7% in the Middle Keys to 7.0% in the Upper 

Keys. Lower Keys stony coral cover was 6.9%. Likewise, sponge cover was similar across all 

three regions ranging from 1.6% in the Upper Keys to 3.0% in the Middle Keys. 

When combined for both years and pooled for all stations within each region*habitat grouping, 

macroalgae was highest at Lower Keys backcountry patch reefs (49.9%), which had the lowest 

mean cover values for the other three benthic fauna groups. Octocoral and sponge cover were 

highest at Middle Keys patch reefs at 31.0% and 5.1%, respectively. The highest stony coral 

cover values were found on the Lower Keys patch reefs (19.7%). For a complete list of benthic 

cover values for all sites and region*habitat groupings, please see Appendices 1 and 2. 

For all stations in the Dry Tortugas (N=12), macroalgae was the most prevalent benthic taxa in 

both years (Figure 4). Macroalgal cover was 14.8% in 2007 and decreased slightly to 12.5% in 

2008. Stony coral was the next highest contributor to benthic cover, averaging 9.8% in 2007 and 

10.3% in 2008, followed by octocorals with 8.2% cover in 2007 and 8.7% in 2008. Similar to 

the Keys, sponges made the smallest contribution to benthic cover and averaged <1.5% for both 

years. 

Macroalgae Octocoral Sponge Stony Coral

Mean % Cover

0

5

10

15

20

2007

2008

Figure 4. Mean (± SE) of macroalgae, octocoral, sponge and stony coral cover for 2007 and 2008 pooled for all 

stations (N = 12) in the Dry Tortugas. Comparisons across years were not statistically examined. 

Macroalgal cover

Although there was no significant change in the overall cover of macroalgae in the Keys between 

2007 and 2008, the mixed model ANOVA yielded a significant interaction effect between year 

and region*habitat (p<0.001; df=10, 90; F=3.51). This result suggests changes in macroalgae 

were localized or restricted to certain habitat*region groupings. For example, in the Lower 

Keys, macroalgal cover significantly increased on backcountry patch reefs from 27.2% in 2007 
2009 CREMP Final Report 18 F2345-02-10-F & F2635-06-09-F

to 72.6% in 2008 (p<0.001; df=1, 90; F=14.39), but significantly decreased in deep (19.8% to 

12.2%, p=0.007; df=1, 90; F=7.46) and shallow forereef habitats (7.3% to 2.9%, p=0.009; df=1, 

90; F=7.05). All other region*habitat groupings showed no difference across years. 

 

Differences in percent cover were not examined statistically for the Dry Tortugas. Only 

descriptive analyses are presented for Bird Key Reef, Black Coral Rock (deep sites) and White 

Shoal (patch reef). Two sites, White Shoal and Black Coral Rock, decreased in macroalgae 

percent cover from 2007 to 2008 (17.2% to 13.6% and 10.2% to 1.2%, respectively). Bird Key 

Reef increased from 17.0% to 22.8%. 

Octocoral Cover

There was a significant interaction effect between year and region*habitat, indicating that 

octocoral cover did not increase at all region*habitat groupings. Mean octocoral percent cover 

was significantly higher in 2008 than in 2007 at three region*habitat groupings (Lower Keys 

shallow and deep sites, and Middle Keys shallow sites). The remaining region*habitat groupings 

demonstrated no change. Octocoral cover increased from 5.4% to 7.1% (p=0.001; df=10, 90; 

F=10.54) in the Lower Keys shallow forereef, and from 8.3% to 9.9% (p=0.007; df=10, 90; 

F=7.58) in the Lower Keys deep forereef. The shallow forereef in the Middle Keys increased 

from 13.8% to 18.1% (p<0.001; df=10, 90; F=21.39). 

Differences in percent cover were not examined statistically for the Dry Tortugas. For the three 

Dry Tortugas sites, octocoral percent cover increased slightly from 2007 to 2008 at White Shoal 

(6.1% to 8.8%) and at Black Coral Rock (6.4% to 7.4%). Bird Key Reef had a decrease in cover, 

from 12.2% in 2007 to 9.7% in 2008. 

Sponge Cover

Similar to octocorals, there was a significant interaction effect between year and region*habitat, 

indicating that sponge cover did not increase at all region*habitat groupings. Mean sponge cover 

significantly increased from 2007 to 2008 in two region*habitat groupings: Lower Keys deep 

sites from 1.5% to 2.7% (p<0.001; df=10, 90; F=22.44) and Middle Keys deep reefs from 2.3% 

to 3.4% (p=0.01; df=10, 90; F=5.85). 

Differences in percent cover were not examined statistically for the Dry Tortugas. At White 

Shoal and Bird Key Reef, sponge cover changed very little: <0.1% from 2007 to 2008. 

However, sponge cover at Black Coral Rock increased from 1.6% in 2007 to 2.7% in 2008. 

Stony Coral Cover

Although there was no significant change in the overall cover of stony corals in the Keys, a 

significant interaction effect suggests that one or more region*habitat groupings significantly 

changed from 2007 to 2008 (p=0.04; df=10, 90; F=2.02). A significant increase in coral cover 

was found at the Middle Keys shallow forereefs where cover changed from 1.2% to 1.7% 

(p=0.01; df=10, 90; F=6.88). 
2009 CREMP Final Report 19 F2345-02-10-F & F2635-06-09-F

Differences in percent cover were not examined statistically for the Dry Tortugas. In the Dry 

Tortugas, stony coral percent cover increased at Black Coral Rock (from 17.5% to 18.5%). 

Stony coral is the largest contributor to benthic fauna at Black Coral Rock. Stony coral percent 

cover remained similar between 2007 and 2008 at White Shoal (~1.5%) and Bird Key Reef 

(~10.5%). 

Stony Coral Species Richness 

In the Florida Keys, mean coral species richness was 13.9±0.45 (SE) and 13.6±0.44 species per 

station in 2007 and 2008 respectively (N=97). Results from the post hoc Tukey-Kramer 

pairwise comparisons demonstrated there was no significant difference between 2007 and 2008 

(Tukey-Kramer q=0.795; p>0.05); however, richness in both years was significantly higher than 

2006 (Tukey-Kramer; q=4.8 and 4.1; p<0.05). In 2006, mean species richness was 12.7±0.42 

species per station, which is the lowest value recorded since the inception of CREMP in 1996. 

This low occurred after a record number of tropical systems impacted the Keys during a 16 

month period between 2004 and 2005. Mean species richness in 2008 was lower than in 1996 

(Friedman’s ANOVA; x2

=300.35, df=12, p<0.001). Mean richness values can be clustered into 

four groups (Figure 5). Values reported for 2007 and 2008 are similar to those reported for 

2003-2005 and 1999 (Figure 5). Richness values during these years, however, are significantly 

lower than those from1996-1998 and 2000-2002 (Figure 5). The period from 1996-1998 

represents the peak in mean species richness observed by CREMP. Although the overall trend is 

negative, an encouraging sign is that mean species richness has rebounded twice after 1999 and 

2006. 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Mean Coral Spp. Richness

0

5

10

15

20

25

30

A A A C B B B CC C C D C

Figure 5. Boxplot of annual mean species richness in the Florida Keys (N = 97). Lower point, error bar and bound 

of the box show the 5th, 10th, and 25th percentiles of the data respectively. The solid black bar within each box 

represents the median. The solid red bar within each box is the mean. Upper bound of the box, error bar and upper 

point show the 75th, 90th, and 95th percentiles of the data, respectively. Letters above each box indicate summarized 

results of Tukey-Kramer pairwise comparisons. Letters denote significant differences in mean species richness 

across year(s) with group A>B>C>D. 
2009 CREMP Final Report 20 F2345-02-10-F & F2635-06-09-F

 

In the Dry Tortugas, mean coral species richness was 18.9±0.82 (SE) and 19.2±0.84 species per 

station in 2007 and 2008, respectively (N=12). There was no significant difference between the 

two years (Holm-Sidak, t=0.44; p>0.05), but as in the Florida Keys, richness in both years was 

significantly higher than in 2006 (Holm-Sidak, t=3.53; p<0.05). Also similar to the Keys, overall 

mean species richness has declined since monitoring was initiated in the Dry Tortugas 

(Friedman’s ANOVA; x2=48.15, df=9, p<0.001). Although not as high as in 1999 and 2001, 

species richness values for 2007 and 2008 are comparable to those observed in 2002 and 2003 

(Figure 6). Mean species richness in the Dry Tortugas reached its lowest level between 2004 

and 2006. This may be in part due to a coral disease outbreak that occurred at Bird Key Reef in 

2001, but the successive and violent hurricane seasons of 2004 and 2005 likely played a role in 

temporarily removing some species from the sites between 2004 and 2006. After three years in 

which no major disturbances affected the region, coral diversity appears to be recovering. 

1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Mean Coral Spp. Richness

0

5

10

15

20

25

30

A B A B B C C C B B

Figure 6. Boxplot of annual mean species richness Dry Tortugas (N = 12). Lower point, error bar and bound of the 

box show the 5th, 10th, and 25th percentiles of the data respectively. The solid black bar within each box represents 

the median. The solid red bar within each box is the mean. Upper bound of the box, error bar and upper point show 

the 75th, 90th, and 95th percentiles of the data, respectively. Letters above each box indicate summarized results of 

Tukey-Kramer pairwise comparisons. Letters denote significant differences in mean species richness across year(s) 

with group A>B>C. 

The long-term trends show mean species richness has declined across all regions and in all 

habitat types (Figures 7 and 8). By region, the largest decreases have occurred in the Lower 

Keys with an average decrease of ~2.6 species per station from 1996 to 2008. By habitat, the 

largest decreases have occurred at the backcountry patch reefs with an average decrease of ~5.0 

species per station from 1996 to 2008. The Dry Tortugas (19.4±0.85) maintains the highest 

diversity of any region and the deep sites have the highest diversity of any habitat (16.7±0.6 SE). 

When pooled for all stations in the Florida Keys (N=97), mean species richness has declined by 
2009 CREMP Final Report 21 F2345-02-10-F & F2635-06-09-F

~2.3 species per station since 1996 and by ~1.8 species per station (N=12) in the Dry Tortugas 

since 1999. 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Mean Spp. Richness (no./station)

0

5

10

15

20

25 Dry Tortugas

Lower Keys

Middle Keys

Upper Keys

Figure 7. Mean species richness by region from 1996-2008 in the Florida Keys and 1999- 2008 in the Dry Tortugas. 

The total number of stations sampled annually for each region was 12 in the Dry Tortugas, 46 in the Lower Keys, 23 

in the Middle Keys, and 28 in the Upper Keys. A mixed model regression indicated a declining trend in all habitats 

(p<0.001). 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Mean Spp. Richness (no./station)

0

5

10

15

20

25 Patch 

Shallow 

Deep

Backcountry 

Figure 8. Mean species richness by habitat from 1996-2008 in the Florida Keys. The total number of stations 

sampled annually for each habitat was 26 at patch reefs, 39 at shallow forereefs, 26 at deep forereefs, and 6 at 

backcountry patch reefs. A mixed model regression indicated a declining trend in all habitats (p<0.001). 

 
2009 CREMP Final Report 22 F2345-02-10-F & F2635-06-09-F

Siderastrea siderea was the most widely distributed species, occurring at all 109 stations in 

2008. The next four most recorded species were Porites astreoides (107 stations), Millepora 

alcicornis (107), Agaricia agaricites (99) and Montastraea cavernosa (93). A majority of the 

species observed by CREMP have decreased in presence at stations in the Florida Keys. The 

species that have had the greatest decrease in presence since 1996 are Favia fragum (33 stations), 

Mycetophyllia lamarckiana (21), Acropora cervicornis (17), Colpophyllia natans (15) and 

Porites porites (15). In contrast, only a few species have increased in distribution, including 

Agaricia fragilis (13 stations), Stephanocoenia intersepta (9), Siderastrea siderea (6), and

Siderastrea radians (4). The absolute change in the presence/absence for each species, pooled 

for all stations from 1996 to 2008 is demonstrated in Figure 9. 

Number of Stations

-40 -30 -20 -10 0 10 20

Favia fragum

Mycetophyllia lamarckiana

Acropora cervicornis

Colpophyllia natans

Porites porites

Mycetophyllia ferox

Diploria labyrinthiformis

Leptoseris cucullata

Diploria strigosa

Eusmilia fastigiata

Millepora complanata

Montastraea annularis

Acropora palmata

Meandrina meandrites

Scolymia cubensis

Madracis mirabilis

Mycetophyllia aliciae

Mycetophyllia danaana

Agaricia agaricites

Dichocoenia stokesi

Mussa angulosa

Diploria clivosa

Oculina diffusa

Agaricia lamarcki

Isophyllia sinuosa

Manicina areolata

Phyllangia americana

Dendrogyra cylindrus

Solenastrea bournoni

Cladocora arbuscula

Millepora alcicornis

Montastraea cavernosa

Madracis decactis

Porites astreoides

Siderastrea radians

Siderastrea siderea

Stephanocoenia michelinii

Agaricia fragilis

Figure 9. Change in the presence/absence (number of stations) of individual coral species from 1996 to 2008 in the 

Florida Keys (N=97 stations). 
2009 CREMP Final Report 23 F2345-02-10-F & F2635-06-09-F

Stony Coral Condition 

CREMP records the incidence of four coral conditions, including bleaching and three stony coral 

disease categories: black band disease (BBD), white diseases (e.g. white plague, white band, and 

white pox), and other diseases (e.g. dark spot, yellow band, and red blotch). In 2007, when 

pooled for all stations in the Florida Keys and Dry Tortugas (N=109), BBD occurred at nine 

stations, bleaching at 63, other diseases at 77, and white diseases at 46. The incidence of all four 

conditions declined from 2007 to 2008: BBD to seven stations, bleaching to 31, other diseases to 

67, and white diseases to 32. The incidence of disease and bleaching varied among coral 

species. Of the four conditions recorded, BBD affected the fewest number of coral species in 

both 2007 and 2008 (four species each year; Figure 10). BBD primarily attacks large boulder 

and brain corals, with the highest incidence occurring on Montastraea annularis complex and M. 

cavernosa. Bleaching was the most common condition recorded during the 2007 survey, 

affecting 80% (32 of the 40) of the observed coral species (Figure 10). The incidence of 

bleaching was much lower in 2008, affecting 17 species of the 41 observed (41.5%). Other 

diseases were observed on 13 species in 2007 and eight species in 2008 (Figure 10). The species 

most commonly affected by other diseases were Siderastrea siderea, Stephanocoenia intersepta, 

and M. annularis complex. The number of species affected by white diseases, as well as the 

number of stations where the disease was recorded for each species, remained relatively 

consistent between 2007 and 2008 (Figure 10). The species most affected by white diseases 

were Agaricia agaricites, M. annularis-complex, M. cavernosa, Porites astreoides, S. intersepta, 

and S. siderea. 

Black Band Disease 2007

Acer Cnat Mann Mcav Ssid

No. of colonies

0

20

40

60

80

100

120

Black Band Disease 2008

Acer Cnat Mann Mcav Ssid

No. of colonies

0

20

40

60

80

100

120
2009 CREMP Final Report 24 F2345-02-10-F & F2635-06-09-F

Bleaching 2007

Aaga

Acer

Afra

Alam

Carb

Cnat

Dcli

Dlab

Dsto

Dstr

Efas

Ffrag

Isin

Lcuc

Mang

Mann

Mcav

Mmea

Mpalc

Mpco

Myali

Mydan

Myfer

Mylam

Odif

Past

Ppor

Sbour

Scub

Sint

Srad

Ssid

No. of colonies

0

20

40

60

80

100

120

Bleaching 2008

Aaga

Acer

Afra

Alam

Carb

Cnat

Dcli

Dlab

Dsto

Dstr

Efas

Ffrag

Isin

Lcuc

Mang

Mann

Mcav

Mmea

Mpalc

Mpco

Myali

Mydan

Myfer

Mylam

Odif

Past

Ppor

Sbour

Scub

Sint

Srad

Ssid

No. of stations

0

20

40

60

80

100

120

Other Diseases 2007

Aaga

Dcli

Dcyl

Dlab

Dsto

Efas

Mann

Mcav

Mmea

Myali

Mylam

Past

Sbour

Sint

Srad

Ssid

No. of stations

0

20

40

60

80

100

120

Other Diseases 2008

Aaga

Dcli

Dcyl

Dlab

Dsto

Efas

Mann

Mcav

Mmea

Myali

Mylam

Past

Sbour

Sint

Srad

Ssid

No. of stations

0

20

40

60

80

100

120

White Diseases 2007

Aaga

Afrag

Alam

Apal

Cnat

Dcyl

Dlab

Dsto

Dstr

Lcuc

Mann

Mcav

Mdec

Myfer

Odif

Past

Ppor

Sbou

Sint

Ssid

No. of stations

0

20

40

60

80

100

120

White Diseases 2008

Aaga

Afrag

Alam

Apal

Cnat

Dcyl

Dlab

Dsto

Dstr

Lcuc

Mann

Mcav

Mdec

Myfer

Odif

Past

Ppor

Sbou

Sint

Ssid

No. of stations

0

20

40

60

80

100

120

Figure 10. Coral species affected by black band disease, bleaching, other diseases, and white diseases in CREMP 

stations in 2007 and 2008. Red bars show the number of stations where the species was present. Dark bars indicate 

the number of stations where the species was observed with the disease/condition. N=109 stations in the Florida 

Keys and Dry Tortugas. 

Throughout the project, BBD has been the least common of the four conditions (Figure 11). 

Although BBD peaked at 19 stations in 1998, it has not been recorded at more than 11 stations 
2009 CREMP Final Report 25 F2345-02-10-F & F2635-06-09-F

within a single year since (Figure 11). Peak incidence of both white diseases and other diseases 

occurred in 2002 (92 and 99 stations, respectively). Since then the incidence of both disease 

categories has declined. The decreasing incidence of white syndromes may be partially linked to 

the decline of Acropora cervicornis and A. palmata across CREMP stations (see species richness 

section). Both species of Acroporid corals are commonly targeted by white diseases and have 

dramatically reduced their populations (Precht et al. 2002, Miller et al. 2002). Bleaching 

incidence has followed a similar trend to white and other diseases, peaking in 2002 at 90 stations, 

with decreasing incidence since then (Figure 11). It is important to note that observations of 

bleaching and disease by CREMP occur throughout the entire field season (May through August) 

rather than strictly during the peak heating periods of August and September, when bleaching is 

reportedly most prevalent. Because CREMP records the incidence of disease and bleaching at 

the station level, but does not estimate the prevalence or severity of the condition of the corals 

within a station, the results presented here provide a general reference to stony coral condition 

during a particular year, but do not necessarily reflect peak bleaching years or disease outbreaks. 

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

No. of stations

0

20

40

60

80

100 Black Band

Bleaching

Other Diseases

White Diseases

Figure 11. Incidence of black band disease, white diseases (e.g. white plague, white band, and white pox), other 

diseases (e.g. dark spot, yellow band, and red blotch), and bleaching. N=109 stations in the Florida Keys and Dry 

Tortugas. 

Long-term Trends 

All Benthic Cover Variables

In the Florida Keys (N=97 stations), the long-term trends in the mean percent cover of the four 

major benthic groups (macroalgae, octocorals, stony corals, sponges) differ from 1996 to 2008. 

Stony corals and sponges have significantly declined, octocorals have significantly increased, 

and no significant change in macroalgae has been observed (Figure 12). Over the project’s 

duration, coral cover has declined ~48%, from 12.7% in 1996 to 6.6% in 2008 (Figure 12). 

During this period coral cover reached its lowest level, 6.4%, in 2006. Sponges have always 

provided the smallest contribution to biotic cover, but like stony corals, demonstrate a negative 
2009 CREMP Final Report 26 F2345-02-10-F & F2635-06-09-F

trend since the project’s inception. Sponge cover has decreased from 3.3% in 1999 to 2.2% in 

2008, with the lowest value for sponge cover attained in 2006 (1.4%). Although octocoral cover 

initially declined during the first four years of the project, it has steadily increased every year 

since except in 2006. Octocoral cover reached a minimum value of 8.9% in 2000, and a 

maximum value of 13.7% in 2005. In 2008, mean octocoral cover throughout the Florida Keys 

was 13.6%, a value only slightly below the 2005 peak. Macroalgae has been the most variable 

component of biotic cover. When combined across all years, mean macroalgal cover is ~11%. 

The highest cover was attained in 1998 (18.9%) while the lowest cover was found in 2005 

(7.4%). 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

5

10

15

20

Stony Coral

Macroalgae

Octocoral

Sponge

Figure 12. Mean annual percent cover for the four major benthic taxa recorded in CREMP image analysis. Mean 

percent cover is pooled from 97 stations in the Florida Keys. A mixed model regression indicates a decreasing trend 

for stony corals and sponges (p<0.001), an increasing trend for octocorals (p<0.001), and no trend for macroalgae 

(p>0.05). 

In the Dry Tortugas (N=12 stations), the long-term trends in mean percent cover from 1999 to 

2008 are decreasing for stony corals, sponges and octocorals. There was no significant change in 

macroalgal cover (Figure 13). From 1999 to 2008, coral cover declined ~45% from 18.9% to 

10.3% (Figure 13). As in the Keys, coral cover reached its lowest level, 8.8%, in 2006. Sponge 

cover decreased from 3.6% in 1999 to 1.4% in 2008. Macroalgae has been extremely variable 

annually, but mean macroalgal cover is ~11% when averaged across all years. The minimum 

value of 5.7% was observed in 2005 while a maximum value of 14.7% was recorded in 2007. In 

contrast to the Keys, octocoral cover has not increased. Octocoral cover remained relatively 

unchanged from 1999 to 2003 (~12.2%), then decreased from 2003 to 2006 to a low of 7.6%, 

and has subsequently increased to a value of 8.7% in 2008. 
2009 CREMP Final Report 27 F2345-02-10-F & F2635-06-09-F

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

5

10

15

20

Stony Coral 

Macroalgae

Octocoral

Sponge

Figure 13. Mean annual percent cover for the four major benthic taxa recorded in CREMP image analysis. Mean 

percent cover is pooled from 12 stations in the Dry Tortugas. A mixed model regression indicates a decreasing 

trend for stony corals, octocorals, and sponges (p<0.001) and no trend for macroalgae (p>0.05). 

Multiple factors have likely contributed to the observed changes in benthic cover during the last 

13 years. Most notable was the protracted ENSO of 1997/1998. It is well documented that this 

event caused extensive bleaching and mortality in the wider Caribbean and Western Atlantic and 

was responsible for dramatically reducing coral cover in the Keys (Somerfield et al. 2008). It is 

also evident that the impact of this ENSO reverberated across multiple taxonomic groups. In 

addition to the declines in stony coral cover, CREMP recorded consecutive years with decreases 

in octocoral and sponge cover between 1997 and 1999. The cause of mortality most likely 

varied across the different taxonomic groups. Declines in stony corals may have been directly 

due to the inability to recover from bleaching or may have been caused by a subsequent disease 

infection following bleaching. Although not documented during the 1997/1998 ENSO, studies 

since have demonstrated that disease outbreaks occur following mass bleaching events because 

elevated water temperatures increase pathogen virulence and the physiological stress endured by 

corals during bleaching weakens their ability to fight off infection (Harvell et al. 2002; Bruno et 

al. 2007; Miller et al. 2009). The expulsion of sponge symbionts leading to their bleaching has 

been concomitant during times of coral bleaching (Vicente 1990; Cowart et al. 2006) and may 

have been responsible for the drop in sponge cover. It is more difficult to ascertain the 

mechanisms that facilitated the decrease in octocoral cover. Massive bleaching events for 

octocorals have not been documented like those for stony corals and sponges; however, they are 

highly susceptible to Aspergillosis caused by the fungus Aspergillus sydowii (Nagelkerken 1974; 

Smith et al. 1996; Nagelkerken 1997; Nagelkerken et al. 1997; Geiser et al. 1998) and red band 

disease caused by cyanobacteria of the genus Oscillatoria spp. (Santavy and Peters 1997; 

Richardson 1998). It is possible that elevated sea temperatures facilitated an increase in these 

conditions. Confounding the identification of the causes of mortality is the fact that two 

hurricanes struck the Keys during this time. Hurricane Georges was especially destructive 
2009 CREMP Final Report 28 F2345-02-10-F & F2635-06-09-F

because it followed on the heels of the 1997/1998 ENSO, making landfall in Key West in 

September of 1998. Hurricane Irene struck a year later in 1999. Benthic organisms that escaped 

direct physical damage may have eventually succumbed to sedimentation or scouring because 

they were already surviving in a sub-optimal condition. It is impossible to identify what 

proportion of the mortality was due to bleaching, disease, or storm damage during these years, 

but the most substantial loss of stony coral, octocoral and sponge cover the project has recorded 

followed this period of combined disturbances. In isolation, hurricanes have resulted in 

reductions to benthic cover but not to the extent witnessed between 1997 and 1999. For 

example, following the volatile hurricane season of 2005, CREMP observed a reduction in 

benthic cover for all three taxonomic groups, but these declines were minor compared to the 

declines following the combined impacts of the 1997/1998 ENSO and Hurricane Georges. 

Recurring patterns of how benthic communities respond to periods of major disturbances are 

evident. As discussed above, all three major taxa groups are extremely vulnerable to acute 

disturbances. Declines in the spatial coverage of these taxa create a favorable environment for 

macroalgae to proliferate and rapidly colonize newly available substrate. CREMP has recorded 

several instances where the decrease in coral and sponge cover was concomitant with an increase 

in macroalgal cover. This shift was observed in 1998 when macroalgal cover doubled from 

9.2% to 18.8% and to a lesser extent in 2000. This pattern also repeated itself after the 2005 

hurricane season. Conversely, even though hurricanes reduce coral cover by breaking or 

crushing stony corals or ripping octocoral holdfasts from the substrate, the heavy scouring and 

wave action also removes macroalgae. Two of the lowest values in macroalgal cover were 

observed during or immediately after periods when hurricanes directly impacted the region. 

CREMP data also suggests that sponges and octocorals have the ability to recover after major 

disturbances while stony corals are less resilient. Octocoral and sponge cover began to rebound 

only two years after the 1997/1998 ENSO and the passages of Hurricanes Georges and Irene. 

This trend was evident again after 2006. 

Stony Coral Cover

In addition to evaluating the overall trends, trends in benthic cover were examined systematically 

by region*habitat grouping and by site to determine if the observed trends were localized or 

widespread (N varied and was dependent upon the number of stations within a region*habitat 

grouping or site). Of the 12 region*habitat designations, all but one has significantly decreased 

in coral cover since 1996 (Bonferroni adjusted, p<0.004). The lone exception was the Middle 

Keys patch reefs (Figure 14; Table 2). Coral cover for this region*habitat grouping was 15.8% 

in 1996 and 14.3% in 2008. It reached its highest point in 2002 (16.3%). This trend is 

remarkable considering how severe the declines were in all other areas during the first four years 

of monitoring. The Lower Keys backcountry patch reefs also fared well until cover collapsed at 

these two sites following a diatom algal bloom in 2001 (Hu et al. 2003). Between 2001 and 

2002, coral cover in this region*habitat grouping dropped from 10.0% to 2.7% and has never 

recovered. The most notable losses in coral cover between 1996 and 1999 occurred at the 

shallow sites. A majority of CREMP shallow sites are located on shallow spur and groove 

forereefs where Acropora palmata was once abundant. Both A. palmata and Millepora 

complanata suffered mass mortalities due to bleaching, disease and hurricanes that occurred 

during this time period (Miller et al. 2002; Patterson et al. 2002; Precht and Miller 2006; 

Somerfield et al. 2008). Although much of the Acropora palmata cover on shallow forereefs had 
2009 CREMP Final Report 29 F2345-02-10-F & F2635-06-09-F

already disappeared before CREMP sampling began, the remaining stands were reduced from 

3.0% to 0.4% during these four years (Somerfield et al. 2008, Miller et al. 2002). The deep 

forereefs also had substantial reductions in coral cover, from 6.7% to 3.7%, during the first four 

years of monitoring. Losses in this habitat were driven by the mortality of large boulder and 

brain corals like Montastraea spp., Colpophyllia natans, and Diploria spp. Stony coral cover at 

Patch reefs in the Lower Keys dropped from 27.5% in 1996 to 19.7% in 1999, and patch reef 

cover in the Upper Keys dropped from 16.3% in 1996 to 11.9% in 1999. Like deep forereef 

sites, patch reefs primarily consist of large reef-building corals (Montastraea spp., Colpophyllia 

natans, and Diploria spp.) but contain the highest coral cover values of any habitat in the Keys. 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

5

10

15

20

25

30

LK BCP

DT OD

LK OD

MK OD

UK OD

LK OS

MK OS

UK OS

DT P

LK P

MK P

UK P

Figure 14. Mean annual percent stony coral cover for 12 region*habitat groupings in the Florida Keys and Dry 

Tortugas. N for each region*habitat grouping varies (see methods). The trend for each region*habitat grouping was 

determined from a mixed model regression. The direction all slopes is reported in Tables 2 & 3. 

At the site level, long-term trends reflect those at the region*habitat level. Twenty-four of 34 

sites (70%) in the Florida Keys show a decreasing trend in coral cover since 1996 (Table 4, 

Appendix 5). Five of eight sites that did not have a significant change were patch reefs while the 

other three were Upper Keys deep sites. All shallow forereef sites show a negative trend which 

is indicative of the mass mortalities and lack of recovery of A. palmata, and in some instances, 

M. complanata in this habitat. No sites show an increasing trend in coral cover. 

Because of the substantial coral mortality incurred during the first three years of the project, the 

inclusion of these data will always produce a negative trajectory on the overall trends in coral 

cover. In order to evaluate how the benthic community responded after the 1997/1998 ENSO, 

we created a new baseline using the values from 1999 as a starting point. Most of the acute 

mortality associated with 1997/1998 ENSO and Hurricane Georges had abated by 1999. 

Although the incidence of coral diseases may have remained high, coral cover did increase 

slightly between 1999 and 2000, so it is reasonable to assume that the major declines in cover 
2009 CREMP Final Report 30 F2345-02-10-F & F2635-06-09-F

from these events had ceased. If a recovery in coral cover was to begin, 1999 appears to be an 

appropriate starting point. 

Table 2. Long-term trends of the four major benthic taxa recorded in CREMP image analysis. Trends were 

determined for 10 region*habitat groupings from a mixed model regression between 1996 and 2008. Interpretation 

of trends for each region*habitat grouping are based on Bonferroni corrected p values for repeated testing (adjusted 

p≤0.004). 

Reg*Hab Stony Corals Macroalgae Octocorals Sponges 

LK BCP decreasing increasing no change no change 

LK OD decreasing decreasing no change no change 

LK OS decreasing no change increasing decreasing 

LK P decreasing no change increasing no change 

MK OD decreasing no change no change no change 

MK OS decreasing no change increasing decreasing 

MK P no change no change no change decreasing 

UK OD decreasing no change increasing no change 

UK OS decreasing no change increasing decreasing 

UK P decreasing no change no change no change 

Table 3. Long-term trends of the four major benthic taxa recorded in CREMP image analysis. Trends were 

determined for 12 region*habitat groupings from a mixed model regression between 1999 and 2008. Interpretation 

of trends for each region*habitat grouping are based on a Bonferroni corrected p value adjusted for repeated testing 

(adjusted p≤0.004). 

Reg*Hab Stony Corals Macroalgae Octocorals Sponges 

DT OD decreasing no change decreasing decreasing 

DT P decreasing no change decreasing decreasing 

LK BCP decreasing increasing no change decreasing 

LK OD decreasing decreasing increasing decreasing 

LK OS decreasing no change increasing decreasing 

LK P no change no change increasing no change 

MK OD decreasing no change increasing no change 

MK OS decreasing no change increasing decreasing 

MK P no change no change increasing decreasing 

UK OD no change no change increasing no change 

UK OS decreasing no change increasing decreasing 

UK P no change no change no change no change 

Table 4. Long-term trends of the four major benthic taxa recorded in CREMP image analysis for all sites from 

1996-2008 and 1999-2008. Thirty-four sites were sampled from 1996 and 37 sites since 1999. Trends were 

determined from mixed model regression. Interpretation of trends for each site is based on a Bonferroni corrected p 

value adjusted for repeated testing (adjusted p<0.002). 

Stony Corals Macroalgae Octocorals Sponges 

96-08 99-08 96-08 99-08 96-08 99-08 96-08 99-08 

Decreasing 24 15 1 1 1 2 4 4 

No Change 10 21 31 34 23 19 28 33 

Increasing 0 1 2 2 10 16 0 0 
2009 CREMP Final Report 31 F2345-02-10-F & F2635-06-09-F

Unfortunately, the timeline adjustment improves the results very little. Since 1999 coral cover at 

eight of the twelve region*habitat groupings continued to decline (Table 3). This result includes 

the shallow and deep reef habitats in all regions except the Upper Keys deep sites (Table 3). 

Patch reefs in all regions, except in the Dry Tortugas, are the only areas that have not had 

significant coral declines during the last decade. At some locations, substantial losses of coral 

cover have continued to occur since 1999. As described previously, the 2001 diatom algal bloom 

coincided with a decline in cover at the backcountry patch reef sites from 10.0% to 2.7% within 

a single year. Since 1999 coral cover at the two Dry Tortugas deep reef sites has dropped from 

23.6% to 14.5%. This result is due to a disease outbreak on Montastraea annularis complex and 

Colpophyllia natans that occurred between 2001 and 2003, and to the physical destruction of six 

storms passing within 100 miles of the Dry Tortugas between August 2004 and October 2005 

(FWRI & DRTO, 2010). The region*habitat grouping for patch reefs in the Dry Tortugas 

consists of only one site (White Shoal) and cover there was diminished when the remaining 

Acropora cervicornis was struck by disease in 2003 and then decimated by the 2004 hurricanes 

(FWRI & DRTO, 2010). In all, of the 37 sites monitored by CREMP since 1999, 15 (40%) still 

continue a declining trend in coral cover, 21 (57%) show no change, while one site (3%) did 

increase in coral cover (Table 4, Appendix 6). Although some sites may demonstrate no 

significant change in coral cover over the last decade, significant decreases may be less 

detectable because coral cover values are so low. For example, the Middle Keys shallow sites 

(Sombrero and Tennessee) had a coral cover value of 2.0% after 1999, but had dropped to 1.7% 

by 2008. 

The most notable instances of coral cover declines since 1999 have occurred in the Dry Tortugas 

and Lower Keys regions. Only two sites show a declining trend in the Middle and Upper Keys 

(Appendix 7). Sixteen of 19 (84%) sites in the Dry Tortugas and Lower Keys show a negative 

trend in cover. While linkages between disturbance events and coral decline are easier to 

propose in the Dry Tortugas (FWRI & DRTO, 2010), it is more complicated to distinguish why 

the Lower Keys sites have not fared as well as those in the Middle and Upper Keys. Fourteen of 

the 16 sites in the Florida Keys with a decreasing trend in coral cover from 1999 to 2008 were 

based in the Lower Keys (Appendix 7). One reason for this decline could be due to forereef sites 

in the Lower Keys possessing higher cover values beginning in 1999 as compared to those in the 

Middle or Upper Keys. If the variance around the interannual mean cover values is equal for 

most sites (e.g. ±1% cover), significant changes in cover would be easier to detect at sites with 

higher coral cover values than at those with lower ones. Although the proportional change in 

cover at two sites could be similar (e.g. a 25% decrease in cover), the absolute change in cover 

would be greater at sites with higher cover. This scenario would yield a significant trend at a site 

starting with 8% cover but not necessarily for one with only 2% cover. Although the 

discrepancies in coral cover across sites should be considered, anthropogenic activities related to 

the thriving tourism industry of Key West need to be considered and may be responsible for the 

trends observed in the Lower Keys. Key West is the cruise ship port for the entire Keys region 

and the most intensively developed island. 

The one site that showed a trend of increasing stony coral cover was Jaap Reef. Another site, 

Molasses Shallow, did show a significant increase in cover since 1999, however it was slightly 

over the Bonferroni corrected p value of p<0.002 (p=0.01). Both sites are worth discussing 
2009 CREMP Final Report 32 F2345-02-10-F & F2635-06-09-F

because they are dissimilar in reef geology and community structure, yet both show trends that 

may be indicative of a slow, gradual recovery of the spatially dominant species that exist within 

these habitats. Jaap Reef is a high relief nearshore hardbottom composed of a monotypic stand 

of Montastraea annularis complex. Molasses Shallow represents a typical shallow spur and 

groove reef formation of the Florida Keys that was previously dominated by Acropora palmata. 

The cover of the sentinel species at each of these sites had significantly declined by 2000. 

Montastraea annularis cover decreased from 28% to 14% at Jaap Reef and A. palmata decreased 

from 6% to <0.6% at Molasses Shallow. Since that time, minor increases in cover have been 

detected and the cover of these two species has averaged 17% and 1.5%, respectively, over the 

last three years (2006 through 2008). Why corals at these two particular sites have responded 

positively since 2000 is difficult to explain, but this pattern demonstrates why sites need to be 

evaluated independently of one another. It should also be pointed out that while the trend over 

the last decade is positive, cover at these sites is still far below the amount that was recorded in 

1996. If anything, these trends suggest the time required for M. annularis complex and A. 

palmata to return to levels of 1996 is on the order of decades, given no further substantial 

bleaching or other large scale disturbances occur. 

 

Macroalgal cover

First, it is important to consider that CREMP image analysis does not identify individual 

macroalgae species and lumps calcareous, fleshy, and filamentous forms into a single 

macroalgae category. Annual observations of macroalgae can fluctuate because of variability in 

the timing of annual sampling, the potential of sampling during a periodic algal bloom, and the 

ephemeral nature of many algal species that occur on coral reefs. These factors can confound 

interannual observations of macroalgae, but interpretation of long-term patterns do appear to be 

consistent. In terms of the overall trend for the region, CREMP results indicate that macroalgal 

cover has remained relatively similar. Contrary to reports for other Caribbean and Western 

Atlantic regions (Hughes 1994; Aronson 2000; Rogers and Miller 2006), reefs in the Florida 

Keys have not undergone a phase-shift towards macroalgal dominance. Although some 

increases in macroalgal cover have persisted in specific region*habitat groupings, these instances 

are localized and do not reflect the broader trend for the entire region. 

Seven of the ten region*habitat groupings show no long-term trend in macroalgal cover since 

1996 (Table 2, Figure 15). One grouping, the Lower Keys backcountry patch reefs, indicates an 

increasing trend while the Lower Keys deep forereef demonstrates a decreasing trend (Table 2). 

Even when the first three years of data are excluded, the trends remain the same for all 

region*habitat groupings. Examining the results at the site level supports the region*habitat 

grouping trends, where >90% of all sites indicate no trend in macroalgal cover dating back to 

1996 or to 1999 (Table 4; Appendices 7 and 8). 

The results of both the region*habitat and site levels time–series analyses suggest that the entire 

region has avoided a prolonged phase-shift towards increased macroalgal cover. While single 

year spikes or short term blooms have been observed after major disturbances (e.g. after 1997 

and 2005), they have not been sustained. Whether due to periodic scouring by hurricanes, 

herbivorous fish abundance, or other top-down controls, macroalgal cover values remain 

between 10% and 15% in most areas. Whether these cover values represent an ecologically 

balanced level of macroalgae on Keys reefs can be debated, but the target level for fleshy 
2009 CREMP Final Report 33 F2345-02-10-F & F2635-06-09-F

macroalgae on Mesoamerican reefs (Belize, Mexico, and Honduras) is set at 15% 

(HealthyReefs.Org 2010). Most areas in the Keys repeatedly fall under this target for macroalgal 

cover. In addition, recent evidence demonstrates that mean Caribbean macroalgal cover 

calculated from 1996-2005 was ~21%, a value nearly double that of the Florida Keys (Schutte et 

al. 2010). Although it has been previously proposed that Keys reefs have undergone a phase 

shift towards macroalgae (Maliao et al. 2008), the long-term trends beyond 2000 do not 

corroborate this finding. Maliao et al. (2008) only used the first five years of CREMP data 

(1996-2000) and while the brief increase in macroalgal cover following the mass bleaching event 

lead to this conclusion, elevated levels of macroalgal cover (like those of 1998) have not been 

observed since. 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

20

40

60

80

LK BCP

DT OD

LK OD

MK OD

UK OD

LK OS

MK OS

UK OS

DT P

LK P

MK P

UK P

Figure 15. Mean annual percent macroalgal cover for 12 region*habitat groupings in the Florida Keys and Dry 

Tortugas. N for each region*habitat grouping varies (see methods). The trend for each region*habitat grouping was 

determined from a mixed model regression. The direction for all slopes is reported in Tables 2 and 3. 

Deep forereef sites consistently have the highest macroalgal cover and are the most susceptible 

to periodic blooms (Figure 15). Upwelling is known to bring excess nutrients to the Florida 

continental shelf and can deliver up to 40 times the amount of nitrogen and phosphorus as 

compared to terrestrial/nearshore inputs (Leichter et al. 2003). Upwelling varies in space and 

time so periodic blooms may not uniformly affect all regions of the Keys. This variability also 

makes assigning causality to periodic blooms challenging. Pulses in macroalgal cover can also 

be attributed to temperature (Lirman and Biber 2000). It may be no coincidence that the highest 

macroalgal values observed by CREMP occurred after the extremely warm water temperatures 

of the 1997/1998 ENSO. Both upwelling and temperature can work simultaneously to foster 

conditions characteristic of macroalgal blooms, albeit in spatially distinct locales. Since 

upwelling is usually accompanied by cooler water reaching the continental shelf, CREMP should 

be able to better distinguish between warm water temperature anomalies and upwelling in the 

future now that subsurface temperature loggers have been deployed at a subset of CREMP sites 

throughout the FKNMS. 
2009 CREMP Final Report 34 F2345-02-10-F & F2635-06-09-F

Octocoral Cover

Similar to macroalgal cover, CREMP image analysis does not differentiate across octocoral 

species and pools encrusting, branching, and flabellate forms into a single octocoral category. It 

should also be mentioned that unlike corals, macroalgae, and sponges, quantification of octocoral 

cover provides an estimation of both canopy and benthic (e.g. encrusting forms) cover. This is 

important to remember when interpreting the long-term trends for octocorals for at least two 

reasons. First, canopy cover can increase without requiring additional substrate or the 

recruitment of new colonies. Thus, octocoral cover can increase without a net gain of 

individuals. Second, octocoral community structure varies greatly by habitat and by site in the 

Florida Keys. Because all octocorals are grouped into a single category, a variety of species 

and/or growth forms may be responsible for apparently similar trends at different sites. Because 

CREMP pools the results for all phenotypes, interpretation and discussion of trends are only 

summarized at the gross taxa level. 

Five of the ten region*habitat groupings show an increasing trend in octocoral cover since 1996 

(Table 2, Figure 16). This trend includes all shallow forereef groupings and the Upper Keys 

deep and Lower Keys patch reef designations. All other region*habitat groupings demonstrated 

no change in cover. Considering the decline in octocorals during the first several years of 

monitoring, the long-term trends demonstrate a resiliency other benthic fauna groups (e.g. stony 

corals and sponges) have not shared. For most region*habitat groupings, octocoral cover 

reached low points in 1999 or 2000 (Figure 16). Octocoral cover has been trending positively for 

all region*habitat groupings except those in the Dry Tortugas and the Lower Keys backcountry 

patch reefs (Table 2). When using 1999 as a baseline to evaluate recovery after the 1997/1998 

ENSO in the Florida Keys, it is clearly evident that a majority of sites are gaining in octocoral 

cover. This result is most apparent at shallow forereef locations. Of the 12 shallow sites 

CREMP monitors, ten show an increasing trend (p<0.002) in octocoral cover from 1999 to 2008. 

The two other shallow sites, Carysfort Reef and Rock Key, also had considerable gains in 

octocoral cover but the regression yielded values slightly above the Bonferroni adjusted p value. 

In all, 16 of 37 sites (43%) have been increasing in octocoral cover since 1999, while 19 (51%) 

have not changed, and only 2 (6%) have decreased (Appendix 10). The two sites that decreased 

in octocoral cover are the deep sites in the Dry Tortugas. Dating back to 1996, when the trend 

includes the mortality associated with the first five years, only one site shows a decrease in cover 

suggesting that octocoral cover at nearly all sites has fully recovered (Appendix 9). 

Octocorals are susceptible to physical damage from storms and disease (Yoshioka 1994; Cerrano 

et al. 2000). Mean octocoral cover (pooled for all stations) declined gradually between 1996 and 

2000 but there was a substantial decrease after 1998. Another large decline also occurred 

between 2005 and 2006. During 1998 and 2005, powerful storms (e.g. Hurricane Georges and 

Wilma) impacted the Keys. Physical abrasion during storms causes partial mortality when 

branches are severed and whole mortality when holdfasts are detached from the substrate (Coma 

et al 2004). The amount of partial or whole mortality inflicted varies greatly by depth (because 

storm surge is usually strongest at shallow forereef locations) and predominately causes 

complete mortality at these sites due to holdfast detachment (Yoshioka and Yoshioka 1991). 

The decline in octocoral cover after 1998 may also be linked to thermal stress following the 

1997/1998 ENSO. Although CREMP has no direct evidence to support this hypothesis, the 1999 
2009 CREMP Final Report 35 F2345-02-10-F & F2635-06-09-F

mass mortality of octocorals in the Mediterranean was attributed to extremely high water 

temperature anomalies which facilitated the onset of a disease outbreak (Cerrano et al. 2000). It 

would not be unreasonable to assume that disease also contributed, in some part, to the mortality 

observed because the relationship between elevated sea temperatures and increased disease 

virulence has been substantiated (Harvell et al. 1999; Bruno et al. 2007). 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

10

20

30

40

LK BCP

DT OD

LK OD

MK OD

UK OD

LK OS

MK OS

UK OS

DT P

LK P

MK P

UK P

Figure 16. Mean annual percent octocoral cover for 12 region*habitat groupings in the Florida Keys and Dry 

Tortugas. N for each region*habitat grouping varies (see methods). The trend for each region*habitat grouping was 

determined from a mixed model regression. The direction for all slopes is reported in Tables 2 and 3. 

The time required for octocoral cover to be restored after a disturbance is highly dependent upon 

the type of mortality inflicted. Regeneration can be much faster if it begins from already 

established colonies that only suffered partial mortality. The onset of recovery is delayed when 

the settlement and growth of new recruits is required. The slow and gradual increase in 

octocoral cover at many reefs may reflect the latter recovery process. It is likely that hurricanes, 

in part, were responsible for reductions in octocoral cover after 1998 and 2005 by stripping 

whole colonies from the substrate. It would be expected that re-establishing octocoral cover to 

previous levels would require several years of recruitment and growth. In addition, CREMP 

image analysis favors large, spatially dominant organisms because the probability of a random 

point landing on smaller taxa is reduced. Even if recruitment of octocorals was initially high 

following a hurricane it would only contribute a minor fraction to overall benthic cover because 

of the small size of recruits. An increase in octocoral cover would likely require several years of 

uninterrupted growth before CREMP image analysis would begin to detect the increase in cover. 

This lapse explains why octocoral cover values after a disturbance remain stagnant for a year or 

two before gradually increasing in many region*habitat groupings. 

The most dramatic increases in octocoral cover have occurred at the shallow forereef sites. In 

2008, cover at most shallow forereef sites had not only recovered to 1996 levels but exceeded 

them. When combined for all sites in this habitat designation, octocoral cover has increased from 
2009 CREMP Final Report 36 F2345-02-10-F & F2635-06-09-F

8.7% in 1996 to 12.9% in 2008. The increase is even more impressive considering octocoral 

cover had declined to 5.8% by 2000. Prior to 2000, Acropora palmata and Millepora 

complanata were still the two predominant stony coral species at many shallow forereef 

locations. As discussed above, both of these species suffered mass mortalities due to bleaching 

and disease during the early years of CREMP monitoring and have shown little or no recovery 

since. The loss of these species opened up large areas of substrate for other organisms to 

colonize. Octocorals appear readily adapted to persist in the harsh shallow forereef environment 

and can tolerate wave action and intense ultraviolet radiation at the shallow depth. The transition 

from stony coral to octocoral dominated communities has been observed before, however all 

examples are exclusive to the Pacific (Endean et al. 1988; Fox et al. 2003; Stobart et al. 2005). 

Phase shifts from stony coral to macroalgae or sponge dominated assemblages have been 

reported for Belize, Jamaica and other Caribbean/Western Atlantic locations (Aronson 2005; 

Rogers and Miller 2006; Norstrom et al. 2009), but the long-term trends reported here suggest a 

shift to octocoral dominated communities. Although this trend is readily apparent on shallow 

forereefs, and is concurrent with the demise of stony corals, increasing octocoral cover at some 

deep forereef and patch reef sites suggests the transition is starting within other habitats in the 

Florida Keys. 

Sponge Cover

Like macroalgae and octocoral categories, all growth forms are pooled into a single sponge 

category during CREMP image analysis. Of the four major taxonomic groups CREMP 

identifies, sponges generally are the smallest contributor to biotic cover at most sites. Six of the 

ten region*habitat groupings show a negative trend in sponge cover since 1996 (Table 2). 

Middle and Upper Keys deep forereefs and Lower and Upper Keys patch reefs did not change in 

sponge cover (Figure 17, Table 3). None of the region*habitat groupings show a positive trend 

in cover since 1996. The outlook does not change if 1999 is used as a starting point. All but 

four of the 12 region*habitat groupings show a decreasing trend in sponge cover (Table 3) with 

the same four region*habitat groupings exhibiting no change in cover. Both region*habitat 

groupings in the Dry Tortugas show a declining trend since 1999. When examined by site, only 

four of the 34 sites show a negative trend in cover since 1996 (Appendix 11). The same four sites 

are the only sites to show a decreasing trend when analyzed from 1999 to 2008 (Appendix 12). 

This result suggests that sponge cover is only marginally decreasing at many sites, but when 

these sites are clustered within their specific region*habitat grouping, a stronger negative trend is 

revealed for the larger spatial area. 

Sponges, similar to stony corals and octocorals, are vulnerable to disease, bleaching, 

sedimentation, physical damage from storms, and harmful algal blooms. Many of these factors 

have contributed to the decline of sponge cover at localized and regional spatial scales. 

Widespread declines were recorded after the 1997/1998 ENSO. Between 1997 and 1999 sponge 

cover in the Florida Keys was reduced from 2.7% to 2.1%. Most notable were the declines at the 

deep forereef locations. Xestospongia muta provide a large portion of the sponge cover in these 

habitats and are susceptible to both disease and bleaching (Cowart et al. 2006; Lopez-Legentil et 

al. 2008). Bleaching of X. muta has been observed on other reefs during periods of severe 

thermal stress and lack of recovery may have been a cause of the decline in sponge cover 

observed from 1997 to 1999. Another regional pattern of decline was observed following the 

2005 hurricane season. Several large storms like Wilma directly impacted the Florida Keys. 

Hurricane Wilma alone inundated parts of the Keys with >10’ storm surge. Wave action and 
2009 CREMP Final Report 37 F2345-02-10-F & F2635-06-09-F

heavy storm surge increase sedimentation which smother and clog sponge oscules, resulting in 

suffocation (Schmahl 1990). Localized events have also influenced sponge cover at specific 

areas. The harmful algal bloom that spread across the West Florida shelf and Florida Bay in 

2001 decimated sponge cover on Lower Keys backcountry patch reefs. This diatom bloom 

significantly reduced the cover of sponges (much like stony corals) from 3.3% in 2001 to 0.6% 

in 2002 in this region*habitat grouping (Figure 17). Algal and cyanobacteria blooms occur 

periodically in Florida Bay and also have virulent effects on sponge populations and seagrass 

communities (Butler IV 1995). 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

2

4

6

8

10

LK BCP

DT OD

LK OD

MK OD

UK OD

LK OS

MK OS

UK OS

DT P

LK P

MK P

UK P

Figure 17. Mean annual percent sponge cover for 12 region*habitat groupings in the Florida Keys and Dry Tortugas. 

N for each region*habitat grouping varies (see methods). The trend for each region*habitat grouping was 

determined from a mixed model regression. The direction for all slopes is reported in Tables 2 and 3. 

It is noteworthy that none of the region*habitat groupings nor individual sites have shown an 

increase in sponge cover since 1996 or 1999. Sponges on other Caribbean reefs have rapidly 

increased in cover or abundance in areas previously occupied by stony corals following a major 

disturbance (Williams et al. 1999; Aronson et al. 2002; Rutzler 2002; Weil et al. 2002). Although 

this transition was reported for the Florida Keys, it was mostly drawn from nearshore hardbottom 

environments undergoing eutrophication, rather than from patch and forereef habitats (WardPaige et al. 2005). It may be the frequency of major disturbances affecting the Florida Keys that 

limits the potential transition from stony coral to sponge dominated communities. It also 

appears, at least in the interim, that sponges along the Florida reef tract are at competitive 

disadvantage in colonizing newly available substrate compared to the ability and success of 

octocorals. 
2009 CREMP Final Report 38 F2345-02-10-F & F2635-06-09-F

Sentinel Stony Coral Species

To further evaluate the decline of stony corals, long-term trends were analyzed from the percent 

cover data for five sentinel coral species. Following the decline of Acropora palmata in the late 

1990s, Montastraea annularis complex (includes all annularis subspecies), Montastraea 

cavernosa, Colpophyllia natans, Siderastrea siderea, and Porites astreoides represent the five 

most spatially-dominant scleractinians in the Florida Keys and Dry Tortugas. Of these five 

species, M. annularis complex has historically maintained the highest spatial coverage of any 

coral species. At project inception, M. annularis complex cover was greater than 5% in the 

Florida Keys and more than 8% in the Dry Tortugas (Figure 18). The next most spatially 

abundant species is M. cavernosa, with cover ranging between 2% and 4%, depending upon the 

region. For both Montastraea spp., coverage is higher in the Dry Tortugas than in the Florida 

Keys. Colpophyllia natans, S. siderea, and P. astreoides round out the top five, comprising 

between 0.5% and 2% cover and represent the third through fifth most spatially abundant corals.

When percent cover data are pooled for all stations in the Florida Keys (N=97) from 1996 to 

2008, and in the Dry Tortugas (N=12) from 1999 to 2008, M. annularis complex, M. cavernosa, 

C. natans, and P. astreoides exhibit decreasing trends in spatial cover (Figures 18 and 19). 

Siderastrea siderea is the only species to buck the negative trend as cover for this species has 

remained relatively similar throughout the project. Although the trends in this report present an 

overall decline for M. annularis complex, M. cavernosa, and C. natans, the decrease in cover for 

these species is often driven by the death of a few, large individuals. The size-specific mortality 

of large boulder and brain corals colonies is not a new phenomenon and has been described by 

Porter and Meier (1992) in Florida and by Hughes et al. (2000) in Jamaica. These three species 

are all broadcast spawners, and poor recruitment rates have not been able to mediate the loss of 

cover for these massive species (Hughes and Tanner 2000; Miller et al. 2000). Surprisingly, 

even the brooding species, P. astreoides, shows an overall decreasing trend. This is not a 

massive, reef building coral but a “weedier” species that is adept at quickly colonizing available 

substrate after disturbances. In some parts of the Caribbean, Porites astreoides has increased in 

relative abundance and mediated the loss of coral cover (Green et al. 2008) but this scenario is 

not reflected by the CREMP trends observed in the Florida Keys and Dry Tortugas. Siderastrea 

siderea appears to be the most tolerant of these five sentinels, and has shown no response to the 

acute and chronic stressors that have reduced the cover of the other four species. 
2009 CREMP Final Report 39 F2345-02-10-F & F2635-06-09-F

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

1

2

3

4

5

Mann

Mcav

Cnat

Past

Ssid

Figure 18. Mean annual percent cover of Montastraea annularis complex (Mann), M. cavernosa (Mcav), 

Colpophyllia natans (Cnat), Porites astreoides (Past), and Siderastrea siderea (Ssid) in the Florida Keys (N=97). 

Mean percent cover is pooled from 97 stations in the Florida Keys. A mixed model regression indicates trends for 

all species were decreasing (p<0.002) except Siderastrea siderea. 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

2

4

6

8

10

Mann

Mcav

Cnat

Past

Ssid

Figure 19. Mean annual percent cover Colpophyllia natans (Cnat), Montastraea annularis complex (Mann), M. 

cavernosa (Mcav), Porites astreoides (Past), and Siderastrea siderea (Ssid) in the Dry Tortugas (N=12). Mean 

percent cover is pooled from 12 stations in the Dry Tortugas. A mixed model regression indicates trends for all 

species were decreasing (p<0.002) except Siderastrea siderea. 
2009 CREMP Final Report 40 F2345-02-10-F & F2635-06-09-F

Montastraea annularis complex

Montastraea annularis complex has suffered the largest loss of spatial cover during the CREMP 

monitoring timeframe, with widespread declines occurring in most regions and habitats. Since 

1996, M. annularis complex cover has had a negative trend at seven of ten region*habitat 

groupings (Table 5). These trends are upheld across all habitat types in the Lower Keys region, 

and at deep forereefs in the Middle and Upper Keys. Even at patch reefs in the Middle and 

Upper Keys, the p values (p=0.008 for both regions) were only slightly above the adjusted 

Bonferroni corrected value (p≤0.004) and is supportive of a decreasing trend in both 

region*habitat groupings. Of all the decreases observed during the initial years, the most 

dramatic occurred at the Lower Keys patch reefs when cover was reduced from ~11.6% in 1997 

to ~6.0% in 2000. When the analyses are conducted starting in 1999 (after the immediate effects 

of the 1997/1998 ENSO are removed), the number of declining region*habitat groupings 

decreases from seven to five (Table 6). Since 1999, all habitat types in the Lower Keys show a 

continued decline except for the patch reefs. The widespread decline in M. annularis is also 

clearly evident when data are analyzed at the site level: 15 of 34 (44%) sites have had a 

significant reduction in cover since 1996 (Appendix 13). Site-level declines were less extreme 

after 1999, yet a third of all sites (12 of 37 sites) continued to lose M. annularis complex cover at 

a significant rate (Appendix 14). The most alarming rate of loss since 1999 has occurred in the 

Lower Keys and Dry Tortugas deep sites. When combined with the results of the Lower Keys 

shallow forereefs, the trends suggest a strong regionalized effect that may be tied to the intense 

development of the Lower Keys. Montastraea annularis is highly susceptible to disease 

outbreaks, bleaching, and hurricane damage, and the synergistic effect of these multiple stressors 

may be exacerbated by additional stressors related to intensive coastal development in the Lower 

Keys. The decline of this species has been widely reported throughout the Caribbean even on 

remote, isolated reefs (Hughes 1994; Gardner et al. 2003; Edmunds 2007). 

Table 5. Long-term trends for: Montastraea annularis complex (Mann), M. cavernosa (Mcav), Colpophyllia natans 

(Cnat), Porites astreoides (Past), and Siderastrea siderea (Ssid). Trends were determined for 10 region*habitat 

groupings from a mixed model regression from 1996-2008. Interpretation of trends for each region*habitat 

grouping are based on Bonferroni corrected p values for repeated testing (adjusted p≤0.004). 

Reg*Hab Mann Mcav Cnat Past Ssid 

LK BCP decreasing no change no change decreasing no change 

LK OD decreasing decreasing decreasing decreasing no change 

LK OS decreasing decreasing decreasing decreasing no change 

LK P decreasing decreasing no change no change no change 

MK OD decreasing decreasing no change no change no change 

MK OS no change decreasing no change no change no change 

MK P no change no change no change no change no change 

UK OD decreasing decreasing no change no change no change 

UK OS decreasing decreasing no change no change no change 

UK P no change no change no change no change no change 
2009 CREMP Final Report 41 F2345-02-10-F & F2635-06-09-F

Table 6. Long-term trends for Montastraea annularis complex (Mann), M. cavernosa (Mcav), Colpophyllia natans 

(Cnat), Porites astreoides (Past), and Siderastrea siderea (Ssid). Trends were determined for 12 region*habitat 

groupings from a mixed model regression from 1999-2008. Interpretation of trends for each region*habitat 

grouping are based on Bonferroni corrected p values for repeated sampling (adjusted p≤0.004). 

Reg*Hab Mann Mcav Cnat Past Ssid 

DT OD decreasing decreasing decreasing decreasing no change 

DT P no change no change decreasing no change no change 

LK BCP decreasing no change no change decreasing no change 

LK OD decreasing decreasing decreasing no change no change 

LK OS decreasing no change no change decreasing no change 

LK P no change no change no change no change no change 

MK OD decreasing no change no change no change no change 

MK OS no change no change no change no change no change 

MK P no change no change no change no change no change 

UK OD no change no change no change no change no change 

UK OS no change decreasing no change no change no change 

UK P no change no change no change no change no change 

Table 7. Long-term trends for Montastraea annularis complex (Mann), M. cavernosa (Mcav), Colpophyllia natans 

(Cnat),Porites astreoides (Past), and Siderastrea siderea (Ssid) at CREMP sites from 1996-2008 and 1999-2008. 

Trends were determined for 34 sites sampled since 1996 and 37 sites since 1999 from a mixed model regression. 

Interpretation of trends for each site are based on Bonferroni corrected p values for repeated sampling (adjusted 

p<0.002). The trends at each site are shown in Appendices 13-22. 

MANN MCAV CNAT PAST SSID 

96-08 99-08 96-08 99-08 96-08 99-08 96-08 99-08 96-08 99-08 

Decreasing 15 12 6 5 5 3 5 5 0 0 

No Change 19 25 28 32 29 33 28 32 33 37 

Increasing 0 0 0 0 0 1 1 0 1 0 
2009 CREMP Final Report 42 F2345-02-10-F & F2635-06-09-F

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

2

4

6

8

10

12

14

LK BCP 

DT OD 

LK OD 

MK OD 

UK OD 

LK OS 

MK OS 

UK OS 

DT P 

LK P 

MK P 

UK P 

Figure 20. Mean annual percent cover of Montastraea annularis complex at 12 region*habitat groupings in the 

Florida Keys and Dry Tortugas. N varies for each region*habitat grouping (see methods). Slope direction for each 

region*habitat grouping is reported in Tables 5 and 6.

Montastraea cavernosa

The long-term trends for Montastraea cavernosa are similar to those for M. annularis complex. 

The overall trajectory (pooled for all stations) in the Florida Keys or the Dry Tortugas is negative 

whether the analysis starts in 1996 or 1999. Since 1996, seven of ten region*habitat groupings in 

the Florida Keys showed a decline in M. cavernosa cover (Table 5). Only three region*habitat 

groupings have historically had cover values >1.0%: Middle Keys patch reefs, Lower Keys patch 

reefs and Dry Tortugas deep sites (Figure 21). Of the seven region*habitat groupings showing a 

decreasing trend since 1996, only the Lower Keys patch reefs started with cover values >1.0%. 

When data from 1996-1998 are excluded from the analysis, only three of the 12 region*habitat 

groupings show a decline in percent cover (Table 6). Of the three region*habitat groupings 

showing a decreasing trend since 1999, only the Dry Tortugas deep sites started with cover 

values above 1.0%. This area has experienced a steep decline in cover, decreasing from 5.2% in 

1999 to 2.9% in 2008. The difference in trends if the baseline begins prior to or after 1999 

suggests the stress associated with 1997/1998 ENSO was primarily responsible for the reduction 

of M. cavernosa cover at many locations. For example, the number of sites where the cover of 

M. cavernosa was declining in the Florida Keys was reduced by 50% after 1999 (Appendices 15 

and 16). Although this result may be partially due to diminished cover values for M. cavernosa

at some sites, the cover of this species does not show a decreasing trend since 1999 in the 

region*habitat groupings where it is most abundant: Middle and Lower Keys patch reefs. Thus, 

the trends for M. cavernosa may be directed by a species-specific vulnerability to large-scale 

bleaching events, but the coral may be less sensitive to other stressors that continue to reduce the 

cover Montastraea annularis complex. A consistency across both Montastraea spp. is that both 

corals appear to have fared better on patch reefs than at deeper sites since 1999 (Appendices 13 

through 16). 
2009 CREMP Final Report 43 F2345-02-10-F & F2635-06-09-F

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

2

4

6

8

10

LK BCP 

DT OD 

LK OD 

MK OD 

UK OD 

LK OS 

MK OS 

UK OS 

DT P 

LK P 

MK P 

UK P 

Figure 21. Mean annual percent cover of Montastraea cavernosa at 12 region*habitat groupings in the Florida Keys 

and Dry Tortugas. N varies for each region*habitat grouping (see methods). Slope direction for each region*habitat 

grouping is reported in Tables 5 and 6. 

Colpophyllia natans

Similar to Montastraea cavernosa, Colpophyllia natans cover is highest on patch reefs in the 

Middle and Lower Keys and the deep sites in the Dry Tortugas, ranging between 1.0% and 2.8%, 

while cover values for all other are region*habitat groupings are <1.0% (Figure 22). When 

pooled for all stations, long-term trends are negative in the Florida Keys and Dry Tortugas 

whether the analysis starts in 1996 or 1999. The declines at the region*habitat level are not as 

widespread as for Montastraea annularis and M. cavernosa. At the region*habitat level, only 

deep and shallow forereefs in the Lower Keys indicate a decrease in cover since 1996 (Table 5). 

This trend continues from 1999-2008 on the Lower Keys deep reefs and is evident in both Dry 

Tortugas habitats. The noticeable decline in cover between 2001 and 2003 in the Dry Tortugas 

was due to a disease outbreak at Bird Key Reef that reduced C. natans cover from 3.5% in 2001 

to 0.7% in 2003 at this site (Morrison et al. 2008). Colpophyllia natans is vulnerable to a variety 

of diseases and disease outbreaks have occurred following major disturbance events (Whelan et 

al. 2007). Also similar to M. cavernosa, C. natans cover was strongly influenced by the 

1997/1998 ENSO. For example, when examined by site in the Florida Keys from 1996-2008, 

five sites show a negative trend (Appendix 17). After 1999, the number of sites showing a 

decreasing trend is reduced to three but two of the sites were in the Dry Tortugas. Only one site 

continued to decline in the Florida Keys after 1999 (Appendix 18). On a positive note, cover has 

increased at one site, Western Head, a Lower Keys patch reef. Like M. cavernosa, this species 

appears more vulnerable to large-scale bleaching events and may be more tolerant of chronic 

stressors. 
2009 CREMP Final Report 44 F2345-02-10-F & F2635-06-09-F

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0.0

0.5

1.0

1.5

2.0

2.5

3.0

LK BCP 

DT OD 

LKOD 

MK OD 

UK OD 

LK OS 

MK OS 

UK OS 

DT P 

LK P 

MK P 

UK P 

Figure 22. Mean annual percent cover of Colpophyllia natans at 12 region*habitat groupings in the Florida Keys 

and Dry Tortugas. N varies for each region*habitat grouping (see methods). Slope direction for each region*habitat 

grouping is reported in Tables 5 and 6. 

Porites astreoides

Porites astreoides is found in a variety of habitats. The overall trend for P. astreoides (pooled 

for all stations) in the Florida Keys or for the Dry Tortugas is negative whether the timeline starts 

in 1996 or 1999. The trend in the Florida Keys is principally driven by three region*habitat 

groupings that show a significant decline when analyzed from either 1996 or 1999: Lower Keys 

deep and shallow forereef sites and backcountry patch reefs. When analyzed from 1999 to 2008, 

three region*habitat groupings show a significant decline: the Lower Keys shallow forereef and 

backcountry patch reefs and the Dry Tortugas deep sites. Although the 1997/1998 ENSO event 

reduced the cover of the large boulder and brain corals discussed above, the cover of P. 

astreoides did not significantly change early in the project except at the Lower Keys deep reefs. 

Of the three region*habitat groupings that had negative trends from 1996 to 2008, this was the 

only grouping in which cover did not show a significant trend when analyzed from 1999 to 2008. 

The other two groupings in the Florida Keys, as well as Dry Tortugas deep sites, had negative 

trends from1999 to 2008, demonstrating a large proportion of P. astreoides mortality has been 

recent. At both the Dry Tortugas deep sites and Lower Keys patch reefs, precipitous and 

consecutive annual declines in P. astreoides cover were observed starting in 2003 and continued 

through 2006 (Figure 23). At the backcountry patch reefs, the collapse in cover coincided with 

the diatom algal bloom in 2001. Site level results reflect the patterns of loss observed at the 

region*habitat level, with five sites indicating a negative trend in both time series from 1996 to 

2008 or 1999 to 2008 (Appendices 19 and 20). This pattern indicates that the most severe losses 

of P. astreoides have occurred following the 1997/1998 ENSO event. No consistent pattern was 

found among the sites with a decreasing trend, as at least one site in nearly all regions and 

habitats had a negative trend (Appendix 20). 
2009 CREMP Final Report 45 F2345-02-10-F & F2635-06-09-F

The results presented here are somewhat surprising based upon the findings reported elsewhere 

in the Caribbean. Given both the acute and chronic nature of the stressors influencing the Florida 

Reef Tract, corals with r-selected life histories would be expected to thrive. Porites astreoides is 

considered a faster growing, “weedier” species that has high recruitment rates because of its 

brooding reproductive strategy (McGuire 1998; Knowlton 2001). This is an ideal solution for 

surviving in periodically disturbed habitats. An increase in the relative abundance and cover of 

P. astreoides has been documented on multiple reefs in the Caribbean during the last several 

decades (Green et al. 2008), and would be expected in the Florida Keys due to the widespread 

declines of acroporid corals and Montastraea annularis complex. Recruitment rates and juvenile 

abundance appear to be high for P. astreoides on Florida reefs (Miller et al. 2000; Keller and 

Donahue 2006)) but survival of larger adult colonies must be low since cover has remained 

similar or declined at many locations. Although Porites astreoides is reported to be resistant to 

acute stressors such as warm water anomalies (Gates et al. 1990) and some chronic stressors such 

as sedimentation (Gleason 1998), the synergistic effect of multiple chronic stressors such as 

nutrient enrichment, over-fishing, and sedimentation do inhibit the recruitment and survival of 

smaller colonies for many species (Bak and Meesters 1999; Mumby et al. 2007). It can only be 

suggested that the level of chronic stressors influencing the Florida Keys and Dry Tortugas are 

strong enough to have deleterious impacts on P. astreoides. Another possibility is that the 

increasing shift to octocoral dominance may be limiting the recruitment and survival of P. 

astreoides at some locations, especially on shallow forereefs. Direct spatial competition and 

allelopathic interactions between octocorals and stony corals have been shown to inhibit the 

recruitment of stony coral larvae (Maida et al. 1995; Maida et al. 2001). 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0.0

0.5

1.0

1.5

2.0

2.5

3.0

LK BCP 

DT OD 

LK OD 

MK OD 

UK OD 

LK OS 

MK OS 

UK OS 

DT P 

LK P 

MK P 

UK P 

Figure 23. Mean annual percent cover of Porites astreoides at 12 region*habitat groupings in the Florida Keys and 

Dry Tortugas. N varies for each region*habitat grouping (see methods). Slope direction for each region*habitat 

grouping is reported in Tables 5 and 6. 
2009 CREMP Final Report 46 F2345-02-10-F & F2635-06-09-F

Siderastrea siderea

Siderastrea siderea is the only species of the five sentinel stony corals that does not show an 

overall decreasing trend in cover whether pooled for all stations in the Florida Keys or the Dry 

Tortugas, regardless of analyses started in 1996 or 1999. Mean annual percent cover for S. 

siderea has remained relatively unchanged since 1996 (Figures 18 and 19). All ten 

region*habitat groupings indicate no trend in cover in the Florida Keys since 1996 and the trend 

in cover for all 12 region*habitat groupings has remained flat in the Florida Keys and Dry 

Tortugas since 1999. Results at the site level mimic those of the region*habitat groupings. 

Thirty-three of 34 sites demonstrate no change in cover from 1996 to 2008 and all 37 sites show 

no change in cover from 1999 to 2008 (Table 7; Appendices 21 and 22). The only site to show a 

change in cover was Cliff Green, a Lower Keys patch reef, and the trend at this site was positive 

(Appendix 21). Siderastrea siderea mean cover was highest on patch reefs in the Florida Keys, 

especially in the Middle and Lower Keys (between 2.1% and 4.0%), followed by deep sites 

where cover ranges between 0.5% and 1.5%. 

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

Percent Cover (%)

0

1

2

3

4

5

LK BCP 

DT OD 

LK OD 

MK OD 

UK OD 

LK OS 

MK OS 

UK OS 

DT P 

LK P 

MK P 

UK P 

Figure 24. Mean annual percent cover of Siderastrea siderea at 12 region*habitat groupings in the Florida Keys and 

Dry Tortugas. N varies for each region*habitat grouping (see methods). Slope direction for each region*habitat 

grouping is reported in Tables 5 and 6. 

Of the five sentinel coral species examined, Siderastrea siderea is the most resilient in the 

Florida Keys and Dry Tortugas. Unlike the other large, framework building corals, Montastraea 

spp. and Colpophyllia natans, dramatic reductions in the cover of S. siderea did not occur 

following the 1997/1998 ENSO. Siderastrea siderea appears tolerant of adverse conditions 

during these extreme disturbances and the fact that the long-term trend in cover has remained 

unchanged at nearly all locations suggests the species is also resistant to chronic stressors. 

Siderastrea siderea is widely distributed around Florida, is common on shallow, nearshore 

carbonate Gulf of Mexico platform reefs on the Florida West coast, and is also found at the very 

northern extension of the Florida Reef Tract in Martin County (Colella et al. 2008; Gilliam 
2009 CREMP Final Report 47 F2345-02-10-F & F2635-06-09-F

2009). Reefs in both of these localities frequently experience turbid conditions, wide 

fluctuations in sea temperature, and are subject to terrestrial inputs and runoff from adjacent 

estuaries (Riegl and Dodge 2008). The relatively high abundance of S. siderea in both stable and 

harsh environments lends evidence to the highly adaptive nature of this species. It may be 

unexpected that S. siderea would be the most suited to conditions in the Florida Keys and Dry 

Tortugas, but changes in coral species composition can be regionally and locally specific. 

Although P. astreoides has partially mediated the decrease in coral cover on many reefs in the 

Caribbean (Green et al. 2008), a transition from Acropora cervicornis to Agaricia tenuifolia has 

been observed in Belize (Aronson and Precht 2001) and from A. cervicornis to Porites porites in 

the Bahamas (Curran et al. 1994). If a shift in coral community structure is to occur in the 

Florida Keys, it certainly appears that S. siderea would be the viable candidate to become the 

most spatially abundant coral. 

Clionaid Cover

Colony area (cm2

/m2) and abundance (number of colonies per station) were recorded for three 

species of clionaid sponges (Cliona delitrix, C. lampa, and C. caribbaea). Since 2001, the mean 

clionaid area, when pooled from all stations (N=109) for all three species, has ranged between 

four and 8cm2

/m2 annually. Mean clionaid area was highest in 2001 (7.9±1.9cm2/m2) and 

reached its lowest point in 2003 (4.8±1.4 cm2/m2). After 2003, clionaid area steadily climbed 

until 2006 and dipped slightly in 2007 to 5.4±1.5cm2/m2 before slightly increasing to 5.9±1.7 

cm

2

/m2 in 2008. 

At the species level, when pooled for all stations (N=109), both mean area and abundance for C. 

delitrix was higher than those of C. caribbaea and C. lampa. Cliona delitrix mean area has 

fluctuated between 2-5cm2

/m2 annually, C. caribbaea between 1-3cm2/m2, and C. lampa from 0-

3cm2/m2. Cliona lampa has been the most variable of all species, decreasing to nearly zero in 

2002 while increasing to >2cm2/m2 in 2006. Cliona lampa is only present at a few CREMP sites 

in the Florida Keys, and is predominately found at one of the Lower Keys backcountry patch 

reefs, Content Keys. In 2001, this site was directly impacted by the harmful algal bloom that 

killed much of the benthic fauna at these sites, and in 2002, C. lampa was not observed at 

Content Keys. In general, C. delitrix area and abundance has been decreasing since 2001. 

Although the area of C. caribbaea and C. lampa fluctuated from 2001 through 2008, the relative 

abundance of the two species has remained similar (Figure 25). The decline in spatial cover of 

C. delitrix is most likely concomitant with decreasing abundance. Between 2001 and 2008, 

Cliona delitrix abundance nearly dropped in half. 
2009 CREMP Final Report 48 F2345-02-10-F & F2635-06-09-F

2001 2002 2003 2004 2005 2006 2007 2008

Mean Clionid Area (cm2/m2

)

0

2

4

6

8

10 C. delitrix 

C. carribbeae

C. lampa 

All Cliona spp. 

Figure 25. Mean area for Cliona delitrix, C. caribbaea and C. lampa for all CREMP stations surveyed in the Florida 

Keys and Dry Tortugas from 2001 to 2008 (N=109 stations). 

2001 2002 2003 2004 2005 2006 2007 2008

Mean Clionid Abundance (sponges/m2

)

0

1

2

3

4

5

C. delitrix 

C. carribbeae 

C. lampa 

Figure 26. Mean abundance for Cliona delitrix, C. caribbaea and C. lampa for all CREMP stations surveyed in the 

Florida Keys and Dry Tortugas from 2001 to 2008 (N=109 stations). 

To evaluate the long-term trends of clionaid cover, Cliona delitrix was selected as a proxy 

because it was the most widely distributed and spatially dominant clionaid species within 

CREMP stations. When the data is pooled for all stations in the Florida Keys and Dry Tortugas 

(N=109), there is an overall decreasing trend from 2001 to 2008 (p<0.001). Three of the 12 
2009 CREMP Final Report 49 F2345-02-10-F & F2635-06-09-F

region*habitat groupings had a significant negative trend in mean C. delitrix area between 2001 

and 2008, Middle and Upper Keys offshore deep reefs and Lower Keys patch reefs (Table 8). 

All other region*habitat groupings did not indicate a trend, but the results for Lower Keys and 

Dry Tortugas deep sites suggest that C. delitrix cover is decreasing in these areas as well (Table 

8). The highest cover of C. delitrix was observed at the patch reefs in the Lower Keys (Figure 

27), while shallow forereef sites generally had the lowest mean area. By site, five of the 37 sites 

showed a decrease in C. delitrix cover: two Lower Keys sites, Cliff Green patch reef and Smith 

Shoal backcountry patch reef, and three deep forereef sites, Bird Key Reef in the Dry Tortugas, 

Sombrero Reef in the Middle Keys, and Carysfort Reef in the Upper Keys (Table 9; Appendix 

23). No change in mean area was detected at the other 32 sites. 

2001 2002 2003 2004 2005 2006 2007 2008

Mean C. delitrix Area (cm2/m2

)

0

5

10

15

20

25

LK BCP 

DT OD 

LK OD 

MK OD 

UK OD 

LK OS 

MK OS 

UK OS 

DT P 

LK P 

MK P 

UK P 

Figure 27. Mean Cliona delitrix area (cm2/m2) for 12 region*habitat groupings in the Florida Keys and Dry 

Tortugas from 2001 to 2008 (N=109 stations). Trends in C. delitrix area were calculated from a mixed model 

regression. Slope direction for each region*habitat grouping is reported in Table 8. 
2009 CREMP Final Report 50 F2345-02-10-F & F2635-06-09-F

Table 8. Long-term trends of mean Cliona delitrix area (cm2/m2) from 2001-2008. Trends for 12 region*habitat 

groupings in the Florida Keys and Dry Tortugas were determined with a mixed model regression. Interpretation of 

trends for each region*habitat grouping based on Bonferroni corrected p values for repeated testing (adjusted 

p≤0.004). 

Reg*Hab P-value Trend 

LK BCP 0.013 no change 

DT OD 0.008 no change 

LK OD 0.006 no change 

MK OD <0.001 Decreasing 

UK OD <0.001 Decreasing 

LK OS 0.338 no change 

MK OS 0.591 no change 

UK OS 0.144 no change 

DT P 0.917 no change 

LK P <0.001 Decreasing 

MK P 0.422 no change 

UK P 0.576 no Change 

Table 9. Long-term trends of mean Cliona delitrix area (cm2/m2) by site (N=37) in the Florida Keys and Dry 

Tortugas from 2001 to 2008. Trends for each site were determined with a mixed model regression. Interpretation of 

trends for each site based on Bonferroni corrected p values for repeated testing (adjusted p<0.002). 

Trend Number of Sites 

Decreasing 5 

No Change 32 

Increasing 0 

The trends for mean Cliona delitrix area mirror those for overall sponge cover in the Florida 

Keys and Dry Tortugas. A majority of the region*habitat groupings show mean C. delitrix area 

remaining similar while a few indicate a declining trend in cover. Previous reports in the 

literature have suggested that clionaid cover was increasing in the Florida Keys due to nutrient 

enrichment of the nearshore waters (Ward-Paige et al. 2005). While CREMP data confirms that 

the greatest abundance and cover of clionaids exist on reefs in closest proximity to highly 

urbanized cities like Key West (e.g. Lower Keys patch reefs), CREMP results do not support that 

clionaid cover is increasing in the Florida Keys. The hypothesis that clionaid cover increases 

with deteriorating water quality has been used elsewhere because sponges are filter feeders and 

can store excess nutrients in their tissues (Ward-Paige et al. 2005; Chaves-Fonnegra et al. 2007). 

While it can only be speculated at this point, CREMP findings, using C. delitrix as a proxy, may 

suggest the level of nutrients entering nearshore waters has been reduced as a result of advanced 

wastewater and sewage treatments facilities built in the Florida Keys during the last two decades. 

Alternatively, the decrease of clionaid cover at some locations may be related to the decline of 

massive star and boulder corals, such as M. cavernosa and the M. annularis complex. The 

region*habitat groupings that contain the highest C. delitrix area also possess many of the 

highest coral cover values for these species (see Sentinel coral species section). Montastraea 

spp. are frequently targeted by C. delitrix, which slowly erodes their skeletons. It is probably no 
2009 CREMP Final Report 51 F2345-02-10-F & F2635-06-09-F

coincidence that sites that have experienced significant declines in C. delitrix area (e.g. Bird Key 

Reef, Carysfort Deep, Cliff Green, Smith Shoal) have also had dramatic reductions M. annularis

complex cover (see Appendices). Although it is difficult to identify specific drivers of the 

observed trends in C. delitrix cover, it is most likely some combination of changing water quality 

and decreasing coral cover. 

CONCLUSIONS 

Throughout the Caribbean, coral reefs have been in a state of decline for the last 40 years with 

reports of many reefs losing more than half of their stony coral cover (Gardner et al. 2003). 

Starting in the late 1970s and continuing through the 1980s, the loss of stony corals in the Florida 

Keys was due in large part to major stressors such as the white band disease outbreak on 

acroporid corals and periodic episodes of bleaching (Causey 2001; Precht and Miller 2006). 

Since the project’s inception in 1996, CREMP has continued to document how these acute 

events have been correlated with negative impacts to Florida’s coral reefs. Without question, the 

greatest regional declines observed by CREMP were associated with the 1997/1998 ENSO and 

the 2005 Hurricane season. However, more localized events have contributed to reduced coral 

cover at a much smaller spatial scale. For example, a diatom bloom decimated stony coral cover 

on backcountry patch reefs in 2001, and a coral disease outbreak between 2001 and 2003 at Bird 

Key Reef in the Dry Tortugas resulted in substantial mortality to Montastraea annularis complex 

and Colpophyllia natans at that site. Although some of these acute stressors are naturally 

occurring phenomena (e.g. hurricanes), the long history of anthropogenic activities in South 

Florida has likely played a role in coral decline. Even during periods that lack major 

disturbances, coral cover has remained similar or continued to decline with only a few examples 

of recovery. Chronic stressors related to the development and extraction of South Florida’s 

natural resources (e.g. overfishing, sedimentation, nutrient enrichment) may be preventing 

widespread stony coral recovery and may contribute to the ongoing decline in cover, yet at a less 

pronounced rate when compared to that of major acute disturbances. 

This report summarizes how coral reef community structure has changed at CREMP sites over 

the last 13 years. Assuming that current environmental conditions surrounding the Florida Keys 

remain similar and major periodic disturbances continue, the synthesis provided here provides 

insight into how reefs in Florida Keys may be transformed in the years ahead. In contrast to 

many areas in the Caribbean, CREMP has not observed a shift from stony coral to macroalgal 

dominance (Hughes 1994; Rogers and Miller 2006). Although previous research has led to the 

conclusion that coral reefs in the Florida Keys were undergoing a shift to macroalgal or sponge 

dominated states (Ward-Paige et al. 2005; Maliao et al. 2008), the results from 13 years of data 

indicate that octocorals are becoming the most abundant taxa. This transition is most apparent at 

shallow forereef habitats and succeeds the demise of Acropora palmata. This trend has also 

begun at sites where large Montastraea spp. corals have declined (e.g. patch reefs and deep 

forereefs) and suggests the transition may be widespread. Identifying the reasons behind the 

transition to octocorals and why they have recovered faster than other benthic fauna after major 

disturbances is difficult to determine. Twice during the tenure of CREMP, octocoral cover has 

declined and rebounded. Although there is little information regarding the status and trends of 

octocoral cover prior to the inception of CREMP, octocorals were reported to exceed stony coral 

species richness and colony density at some localities in the Florida Keys prior to 1996 

(Wheaton and Jaap 1988; Chiappone and Sullivan 1997). It is possible that the shift toward 
2009 CREMP Final Report 52 F2345-02-10-F & F2635-06-09-F

octocorals commenced many decades ago in the Florida Keys starting with the decline in 

acroporid corals in the 1970’s and 1980’s, and CREMP is simply documenting the manifestation 

of this gradually developing transition. It is also uncertain what consequences this transition will 

bring to Florida Keys coral reefs. Octocorals are direct spatial competitors with stony corals and 

may limit their recovery by reducing the settlement of coral planulae (Maida et al. 1995; Maida 

et al. 2001). Alternatively, it is feasible that octocoral dominated benthic habitats will continue 

to support diverse and abundant fish and invertebrate assemblages, albeit unconventional, as they 

provide critical habitat and serve as an important food source for a variety of predators (Wolff et 

al. 1999; Gratwicke et al. 2006). 

While the general consensus is that the future of coral reefs is one of doom and gloom there are 

some reasons for optimism in the Florida Keys. It is important to understand that changes in 

stony coral communities take place over long temporal scales. The current state of reefs in the 

Florida Keys is the culmination of at least 100 years of intense fishing pressure and 50 years of 

physical changes to the coastal environment. Conversely, actions taken to remediate 

environmental perturbations such as fishing regulations (e.g. catch limits, closures, marine 

reserves, etc.) and improving water quality (e.g. advanced wastewater treatment in the Florida 

Keys and the restoration of the Florida Everglades) have been enacted only recently. Even 

though global climate change continues to threaten coral reefs, the alleviation of local stressors 

may temper impacts from severe acute stressors (e.g. mass bleaching events) and aid in the 

recovery following the disturbance. Since 1999, a few sites have shown a positive trend in coral 

cover (e.g. Jaap Reef and Molasses Shallow) indicating that some recovery may be occurring at 

isolated locations. Although the increases in stony coral cover at these sites are minor, corrective 

actions to improve water quality and restore fish populations may prove beneficial over longer 

temporal scales. While results like these are encouraging, continued monitoring by CREMP will 

provide the necessary data to evaluate the efficacy of these management actions and if they are 

reversing the long history of decline of scleractinian corals in the Florida Keys. 
2009 CREMP Final Report 53 F2345-02-10-F & F2635-06-09-F

LITERATURE CITED 

Aronson A, Precht W, Toscano M, and Koltes K. 2002. The 1998 bleaching event and its 

aftermath on a coral reef in Belize. Marine Biology 141(3):435-447. 

Aronson RB, Macintyre, I. G., Wapnick, C. M., O'Neill, M. W. 2005. Phase Shifts, Alternative 

States, and the Unprecedented Convergence of Two Reef Systems. Ecology [Ecology] 

85(7):1876-1891. 

Aronson RB, and Precht WF. 2001. White-band disease and the changing face of Caribbean 

coral reefs. Hydrobiologia 460(1-3):25-38. 

Aronson RB, Precht, W. F. 2000. Herbivory and algal dynamics on the coral reef at Discovery 

Bay, Jamaica. Limnology and Oceanography 45(1):251-255. 

Ault JS, Bohnsack JA, and Meester G. 1998. A retrospective (1979-1996) multispecies 

assessment of coral reef fish stocks in the Florida Keys Fishery Bulletin 96(3):19. 

Ault JSB, J.A., Smith SG, and Luo J. 2005. Towards Sustainable Multispecies Fisheries in the 

Florida, USA, Coral Reef Ecosystem Bulletin of Marine Science 76(2). 

Bak RPM, and Meesters EH. 1999. Population Structure as a Response of Coral Communities to 

Global Change. American Zoologist: From the Symposium Coral Reefs and Environmental 

Changes?Adaptation, Acclimation, or Extinction presented at the annual Meeting of the Society 

for Comparative and Integrative Biology, January 3-7, 1998, at Boston, Massachusetts 39(1):56-

65. 

Baker AC, Glynn PW, and Riegl B. 2008. Climate change and coral reef bleaching: An 

ecological assessment of long-term impacts, recovery trends and future outlook. Estuarine, 

Coastal and Shelf Science 80(4):435-471. 

Bohnsack JA, and Smith WFG. 1994. The impact of fishing on coral reefs. Global aspects of 

coral reefs: Health, hazards and history pp C8 C12-Proceedings of the colloquium on global 

aspects of coral reefs, Miami, 1993 76(2):196-200. 

Brown BE. 1997. Coral bleaching: causes and consequences. Coral Reef 16(1):129-138. 

Bruno JF, Aronson RB, Precht WF, Glynn PW, Harvell CD, Kaufman L, Rogers CS, Shinn EA, 

and Valentine JF. 2004. Causes of Coral Reef Degradation Science 302(5650):2. 

Bruno JF, Petes LE, Harvell DC, and Hettinger A. 2003. Nutrient enrichment can increase the 

severity of coral diseases. Ecology Letters 6(12):1056-1061. 

Bruno JF, Selig ER, Casey KS, Page CA, Willis BL, Harvell CD, Sweatman H, and Melendy 

AM. 2007. Thermal stress and coral cover as drivers of coral disease outbreaks. PLoS Biol 

5(6):1-8. 
2009 CREMP Final Report 54 F2345-02-10-F & F2635-06-09-F

Butler IV MJ. 1995. Cascading disturbances in Florida Bay, USA: cyanobacteria blooms, sponge 

mortality, and implications for juvenile spiny lobsters Panulirus argus. Marine Ecology Progress 

Series 129:119-125. 

Causey B. 2001. Lessons learned from the intensification of coral bleaching from 1980–2000 in 

the Florida Keys, USA. Salm RV, Coles SL (eds) Coral bleaching and marine protected areas: 

Proc Workshop ‘Mitigating coral bleaching impact through MPA design’ (Bishop Museum, 

Honolulu, HI) Asia Pacific Coastal Marine Program Report 0102. 

Cerrano C, Bavestrello G, Bianchi CN, Cattaneo-vietti R, Bava S, Morganti C, Morri C, Picco P, 

Sara G, Schiaparelli S et al. . 2000. A catastrophic mass-mortality episode of gorgonians and 

other organisms in the Ligurian Sea (North-western Mediterranean), summer 1999. Ecology 

Letters 3(4):284-293. 

Chaves-Fonnegra A, Zea S, and Gomez ML. 2007. Abundance of the excavating sponge Cliona 

delitrix in relation to sewage discharge at San Andres Island, sw Caribbean, Colombia. Bol 

Invest Mar Cost 36:63-78. 

Chiappone M, and Sullivan JR. 1997. Rapid assessment of reefs in the Florida Keys: Results 

from a synoptic survey. Proceedings from the 8th International Coral Reef Sypmosium 2:1509-

1514. 

Colella M, Richardson B, Johnson D, Ruzicka RR, Callahan MK, Bertin M, Wheaton JW, 

Schmitt S, and Sheridan N. 2008. Assessment of Population and Community Structure of Sessile 

Macro Invertebrates following a Benthic Mortality Event in the Eastern Gulf of Mexico. Fish 

and Wildlife Research Institute. St. Petersburg, FL. p 49. 

Cowart JD, Henkel TP, McMurray SE, and Pawlik JR. 2006. Sponge orange band (SOB): a 

pathogenic-like condition of the giant barrel sponge, Xestospongia muta. Coral Reef 24(4):513-

514. 

Curran HA, Smith DP, Meigs LC, Pufall AE, and Greer ML. 1994. The health and short-term 

change of two coral patch reefs, Fernandez Bay, San Salvador Island, Bahamas. In: Ginsburg RN 

(ed) Proc Colloquium on Global Aspects of Coral Reefs: Health, Hazards, and History, 1993 

Rosenstiel School of Marine and Atmospheric Science, University of Miami, FL,. p 147–153. 

Edmunds PJ. 2007. Evidence for a decadal-scale decline in the growth rates of juvenile 

scleractinian corals. Marine Ecology Progress Series 341:1-13. 

Endean R, Cameron AM, and DeVantier LM. 1988. Acanthaster planci predation on massive 

corals: the myth of rapid recovery of devasted reefs. Proceedings of the 6th International Coral 

Reef Symposium, Australia 2:143-148. 
2009 CREMP Final Report 55 F2345-02-10-F & F2635-06-09-F

Fitt WK, Barbara BE, Warner ME, and Dunne RP. 2001. Coral bleaching: interpretation of 

thermal tolerance limits and thermal thresholds in tropical corals. Coral Reefs 20(1):51-65. 

Fong P, and Lirman D. 1995. Hurricanes Cause Population Expansion of the Branching Coral 

Acropora palmata (Scleractinia): Wound Healing and Growth Patterns of Asexual Recruits. 

Marine Ecology 16(4):317-335. 

Fong P, and Lirman D. 1997. Patterns of Damage to the Branching Coral Acropora palmata 

Following Hurricane Andrew: Damage and Survivorship of Hurricane-Generated Asexual 

Recruits. Journal of Coastal Research 13(1):67-72. 

Fourqurean JW, and Robblee MB. 1999. Florida Bay: A history of recent ecological changes. 

Estuaries and Coasts 22(2B):345-357. 

Fox HE, Pet JS, Dahuri R, and Caldwell RL. 2003. Recovery in rubble fields: long-term impacts 

of blast fishing. Marine Pollution Bulletin 46:1024-1031. 

FWRI and DTNP. 2010. Long Term Monitoring & Assessment of Selected Coral Reef Sites In 

the Dry Tortugas National Park. 2006 Annual Report. Fish and Wildlife Research Institute. St. 

Petersburg, FL. p 35. 

Gardner TA, Cote IM, Gill JA, Grant A, and Watkinson AR. 2003. Long-tern region-wide 

declines in Caribbean corals. Science 301:958-960. 

Gates RD, Baghdasarian G, and Muscatine L. 1990. Temperature stress causes host cell 

detachment in symbiotic cnidarians: Implications for coral bleaching. Biological Bulletin, 

Marine Biological Laboratory, Woods Hole 182(3):324-332. 

Geiser DM, Taylor JW, Ritchie KB, and Smith GW. 1998. Cause of sea fan death in the West 

Indies. Nature 394:137-138. 

Gilliam DS. 2009. 2008 Year 6 Final Report. Southeast Florida Coral Reef Evaluation and 

Monitoring Project Florida DEP report #G0099:39. 

Gilmore MD, and Hall BR. 1976. Life history, growth habits, and constructional roles of 

Acropora cervicornis in the patch reef environment. Journal of Sedimentary Research 46(3):519-

522. 

Glazer R, and Quintero I. 1998. Observations on the sensitivity of queen conch to water quality: 

Implications for coastal development. Proceedings of the Gulf and Caribbean Fisheries 

Institute(50):78-93. 

Gleason DF. 1998. Sedimentation and distributions of green and brown morphs of the Caribbean 

coral Porites astreoides. Journal of Experimental Marine Biology and Ecology 230(1):73-89 
2009 CREMP Final Report 56 F2345-02-10-F & F2635-06-09-F

Glynn PW, and Colgan MW. 1993. Sporadic disturbances in fluctuating coral reef environments: 

El Nino and coral reef development in the eastern Pacific. American Zoologist 32(6):707-718. 

Gratwicke B, Petrovic C, and Speight MR. 2006. Fish distribution and ontogenetic habitat 

preferences in non-estuarine lagoons and adjacent reefs. Environmental Biology of Fishes 

76:191-210. 

Green DH, Edmunds PJ, and Carpenter RC. 2008. Increasing relative abundance of Porites 

astreoides on Caribbean reefs mediated by an overall decline in coral cover. Marine Ecology 

Progress Series 359:1-10. 

Griffin DW, Donaldson KA, Paul JH, and Rose JB. 2003. Pathogenic human viruses in coastal 

waters. Clinical Microbiology Reviews 16(1):129-143. 

Harvell CD, Burkholder JM, Colwell RR, Epstein PR, Grimes DJ, Hofmann EE, Lipp EK, 

Osterhaus ADME, Overstreet RM, Porter JW et al. . 1999. Emerging Marine Disease-Climate 

Links and Anthropogenic Factors. Science 285(3):1505-1510. 

Harvell DC, Mitchell CE, Ward JR, Altizer S, Dobson AP, Ostfeld RS, and Samuel MD. 2002. 

Climate Warming and Disease Risks for Terrestrial and Marine Biota. Science 296(5576):2158-

2162. 

HealthyReefs.Org. 2010. Report card for the Mesoamerican reef. An evaluation of ecosystem 

health. Healthy Reefs for Healthy People. p 44. 

Hoegh-Guldberg O. 1999. Climate change, coral bleaching and the future of the world's coral 

reefs. Marine & Freshwater Research 50(8):839-866. 

Hu C, Hackett KE, Callahan MK, Andrefouet S, Wheaton JL, Porter JW, and Muller-Karger FE. 

2003. The 2002 ocean color anomaly in the Florida Bight: A cause of local coral reef decline? 

Geophysical Research Letters 30(3):1151-1154. 

Hughes TP. 1994. Catastrophes, phase shifts, and large-scale degradation of a Caribbean coral 

reef. Science 265(5178):1547-1551. 

Hughes TP, and Tanner JE. 2000. Recruitment failure, life histories, and long-term decline of 

Caribbean corals. Ecology 81(8):2250-2263. 

Johns GM, Leeworthy VR, Bell FW, and Bonn MA. 2001. Socioeconomic Study of Reefs in 

Southeast Florida. Hazen and Sawyer Environmental Engineers & Scientists. p 348. 

Keller B, and Donahue S. 2006. 2002 - 03 Florida Keys National Marine Sanctuary Science 

Report: An Ecosystem Report Card After Five Years of Marine Zoning. Marine Sanctuaries 

Conservation Series NMSP-06-12 Marine Sanctuaries Conservation Series NMSP-06-12 US 
2009 CREMP Final Report 57 F2345-02-10-F & F2635-06-09-F

Department of Commerce, National Oceanic and Atmospheric Administration, National Marine 

Sanctuary Program, Silver Spring, MD. p 358. 

Knowlton N. 2001. The future of coral reefs. Proceedings from the National Academy of Science 

USA 98:5419-5425. 

Kruczynski B. 1999. Water quality concerns in the Florida Keys: Sources, Effects, and Solutions. 

Florida Keys National Marine Sanctuary Water Quality Protection Program p68. 

Lapointe BE. 1990. Caribbean coral reefs: Are they becoming algal reefs? Sea Frontiers(MarchApril):83-91. 

Lapointe BE. 1997. Nutrient thresholds for bottom-up control of macroalgal blooms on coral 

reefs in Jamaica and southeast Florida. Limnology and Oceanography 45:1119-1131 

Lapointe BE, Barile PJ, and Matzie WR. 2004. Anthropogenic nutrient enrichment of seagrass 

and coral reef communities in the Lower Florida Keys: discrimination of local versus regional 

nitrogen cycles. Journal of Experimental Marine Biology and Ecology 308:23-58. 

Lapointe BE, and Clark MW. 1992. Nutrient Inputs from the Watershed and Coastal 

Eutrophication in the Florida Keys. Estuaries 15(4):465-476. 

Lapointe BE, and Matzie WR. 1996. Effects of stormwater nutrient discharges on eutrophication 

processes in nearshore waters of the Florida keys. Estuaries and coasts 19(2):422-435. 

Lapointe BE, Tomasko DA, and Matzie WR. 1994. Eutrophication and the trophic state 

classification of seagrass communities in the Florida Keys. Bulletin of Marine Science 54(3):21. 

Leichter JJ, Stewart HL, and Miller SL. 2003. Episodic Nutrient Transport to Florida Coral 

Reefs. Limnology and Oceanography 48(4):1394-1407. 

Lesser MP, Bythell JC, Gates RD, Johnstone RW, and Hoegh-Guldberg O. 2007. Are infectious 

diseases really killing corals? Alternative interpretations of the experimental and ecological 

data. Journal of Experimental Marine Biology and Ecology 346:36-44. 

Lessios HA, Cubit JD, Robertson DR, Shulman MJ, Parker MR, Garrity SD, and Levings SC. 

1984. Mass mortality of Diadema antillarum on the Caribbean coast of Panama. Coral reefs 

Heidelberg etc 3(4):173-182. 

Light SS, and Dineen WJ. 1994. Water Control in the Everglades: A historical perspective. In: 

Everglades: the ecosystem and its restoration 47-84. 

Lipp EK, Jarrell JL, Griffin DW, Lukasik J, Jacukiewicz J, and Rose JB. 2002. Preliminary 

evidence for human fecal contamination in corals of the Florida Keys, USA. Marine Pollution 

Bulletin 44(7):666-670. 
2009 CREMP Final Report 58 F2345-02-10-F & F2635-06-09-F

Lirman D. 2000. Fragmentation in the branching coral Acropora palmata (Lamarck): growth, 

survivorship, and reproduction of colonies and fragments. Journal of Experimental Marine 

Biology and Ecology 251(1):41-57. 

Lirman D, and Biber P. 2000. Seasonal Dynamics of Macroalgal Communities of the Northern 

Florida Reef Tract. Botanica marina 43(4):305-314. 

Lopez-Legentil S, Song B, McMurray SE, and Pawlik JR. 2008. Bleaching and stress in coral 

reef ecosystems: hsp70 expression by the giant barrel sponge Xestospongia muta. Molecular 

Ecology 17(7):1840-1849. 

Maida M, Sammarco PW, and Coll JC. 1995. Effects of soft corals on scleractinian coral 

recruitment. I: Directional allelopathy and inhibition of settlement. Marine Ecology Progress 

Series 121:191-202. 

Maida M, Sammarco PW, and Coll JC. 2001. Effects of soft corals on Scleractinian coral 

recruitment. 2: Allelopathy, spat survivorship and reef community structure. Marine Ecology 

22(4):397-414. 

Maliao RJ, Turingan RG, and Lin J. 2008. Phase-shift in coral reef communities in the Florida 

Keys National Marine Sanctuary (FKNMS), USA. Marine biology 154(5):841-853. 

Manzello DP, Brandt M, Smith TB, Lirman D, Hendee JC, and Nemeth RS. 2007. Hurricanes 

benefit bleached corals. Proceedings of the National Academy of Sciences 104(29):12035-

12039. 

McClenachan L. 2008. Documenting Loss of Large Trophy Fish from the Florida Keys with 

Historical Photographs. Conservation Biology 23(3). 

McCook LJ, Wolanski, E., Spagnol, S. 2001. Modelling and visualizing interactions between 

natural disturbances and eutrophication as causes of coral reef degradation. Monographs Aquatic 

Ecology and Autecology 113-125. 

McGuire MP. 1998. Timing of larval release by Porites astreiodes in the northern Florida Keys. 

Coral Reefs 17:369-375. 

McIvor CC. 1994. Food Habits of mangrove fishes: A comparison along estuarine gradients in 

Northeasetern Florida Bay. Bulletin of Marine Science 54(3):881-899. 

Miller J, Muller E, Rogers C, Waara R, Atkinson A, Whelan KRT, Patterson M, and Witcher B. 

2009. Coral disease following massive bleaching in 2005 causes 60% decline in coral cover on 

reefs in the US Virgin Islands. Coral Reefs 28(4):925-937. 

Miller MW, Bourque AS, and Bohnsack JA. 2002. An analysis of the loss of acroporid corals at 

Looe Key, Florida, USA: 1983-2000. Coral Reefs 21(2):179-182. 
2009 CREMP Final Report 59 F2345-02-10-F & F2635-06-09-F

Miller MW, Weil E, and Szmant AM. 2000. Coral recruitment and juvenile mortality as 

structuring factors for reef benthic communities in Biscayne National Park, USA. Coral Reefs 

19:115-123. 

Mumby PJ, Harborne AR, Williams J, Kappel CV, Brumbaugh DR, Micheli F, Holmes KE, 

Dahlgren CP, Paris CB, and Blackwell PG. 2007. Trophic cascade facilitates coral recruitment in 

a marine reserve. Proceedings of the National Academy of Science 104:8362-8367. 

Nagelkerken I, Buchan K, Smith GW, Bonair K, Bush P, Garzon-Ferreira, J. B, L., Gayle P, 

Harvell CD, Heberer C et al. . 1997. Widespread disease in Caribbean sea fans I. Spreading and 

general characteristics. Proc 8th Int Coral Reef Sym I:679-682. 

Nagelkerken I, Buchan, K., Smith, G.W., Bonair, K., Bush, P., Garzon-Ferreira, J., Botero, L., 

Gayle, P., Harvell, C.D., Heberer, C., Kim, K., Petrovic, C., Pots, L., Yoshioka, P. 1997. 

Widespread disease in Caribbean sea fans II. Pattern of infection, and tissue loss. Mar Ecol Prog 

Ser 160:255-263. 

Nagelkerken WP. 1974. Studies on the fauna of Curacao and other Caribbean Islands: No. 147. 

Onthe occurrence of fishes in relation to corals in Curacao. Uitg Natuurwet Stud Suriname Ned 

Antillen no 77:118-141. 

National Hurricane Center. 2010. http://www.nhc.noaa.gov/. 

Norstrom AV, Nystrom M, Lokrantz J, and Folke C. 2009. Alternative states on coral reefs: 

beyondcoral–macroalgal phase shifts. Marine Ecology Progress Series 363:295–306. 

Patterson KL, Porter JW, Ritchie KB, Polson SW, Mueller E, Peters EC, Santavy DL, and Smith 

GW. 2002. The etiology of white pox, a lethal disease of the Caribbean elkhorn coral, Acropora 

palmata. Proceedings of the National Academy of Sciences, USA 99(13):8725-8730. 

Paul JH, Rose JB, Jiang S, Kellogg, C., and Shinn EA. 1997. Occurrence of fecal indicator 

bacteria in surface waters and the subsurface aquifer in Key Largo, Florida. Applied and 

Environmental Microbiology 61(6):2235-2241. 

Paul JH, Rose JB, Jianga SC, Zhoua X, Cochrana P, Kellogga C, Kanga JB, Griffina D, Farrah S, 

and Lukasik J. 2000. Evidence for groundwater and surface marine water contamination by 

waste disposal wells in the Florida Keys. Pergamon 31(6):1448-1454. 

Porter JW, Harvell CD, Kim K, Burkholder JM, Colwell RR, Epstein PR, Grimes DJ, Hofmann 

EE, Lipp EK, Osterhaus ADME et al. . 1999. Emerging Marine Diseases--Climate Links and 

Anthropogenic Factors. Science 285(5433):1505-1510. 

Precht W, Bruckner A, Aronson R, and Bruckner R. 2002. Endangered acroporid corals of the 

Caribbean. Coral Reefs 21(1):41-43. 
2009 CREMP Final Report 60 F2345-02-10-F & F2635-06-09-F

Precht WF, and Aronson RB. 1997. Stasis, Biological Disturbance, and Community Structure of 

a Holocene Coral Reef,. Paleobiology 23(3):326-346. 

Precht WF, and Miller SL. 2006. Ecological Shifts along the Florida Reef Tract: The Past as a 

Key to the Future. In: Geological Approaches to Coral Reef Ecology. Aronson RB, editor. 

Springer, NY. p 237-312. 

Richardson LL. 1998. Coral diseases: what is really known? Trends in Ecology & Evolution 

13(11):438-443. 

Riegl BM, and Dodge RE. 2008. Coral Reefs of the World 1. Coral Reefs of the USA. 2008: 

Springer. 

Roberts HH, Rouse LJ, Walker ND, and Hudson JH. 1982. Cold-water stress in Florida Bay and 

northern Bahamas; a product of winter cold-air outbreaks Journal of Sedimentary Research 

52(1):10. 

Robinson A. 1973. Natural vs. Visitor-related Damage to Shallow Water Corals: 

Recommendations for Visitor Management and the Design of Underwater Nature. NPS 

Symposium Series 1. 

Rogers CS, and Miller J. 2006. Permanent 'phase shifts' or reversible declines in coral reef 

cover? Lack of recovery of two reefs in St. John, US Virgin Islands Marine Ecology Progress 

Series 306:103-114. 

Rutzler K. 2002. Impact of crustose clionid sponges on Caribbean reef corals. Acta Geology 

Hispanica 37:61-72. 

Santavy DL, and Peters EC. 1997. Microbial pests: coral disease in the Western Atlantic. Proc 

8th Int Coral Reef Sym 1:607-612. 

Schmahl GP. 1990. Community structure and ecology of sponges associated with four southern 

Florida coral reefs. in: K Rützler, ed New perspectives in sponge biology Smithson Inst 

Press:376–383. 

Schutte VGW, Selig ER, and Bruno JF. 2010. Regional spatio-temporal trends in Caribbean 

coral reef benthic communities. Marine Ecology Progress Series 402:115-122. 

Shinn EA. 1976. Coral reef recovery in Florida and the Persian Gulf. Environmental Geology 

1(4):241-254. 

Smith GW, Harvell CD, and Kim K. 1996. Observations on the pathogenesis of sea fans infected 

with Aspergillus sp. Revista de biologia tropical [Rev Biol Trop] 46:205-208. 
2009 CREMP Final Report 61 F2345-02-10-F & F2635-06-09-F

Somerfield PJ, Jaap WC, Clarke KR, Callahan M, Hackett K, Porter J, Lybolt M, Tsokos C, and 

Yanev G. 2008. Changes in coral reef communities among the Florida Keys, 1996–2003. Coral 

Reefs 27(4):951-965. 

Stobart B, Teleki K, Buckley R, Downing N, and Callow M. 2005. Coral recovery at Aldabra 

Atoll, Seychelles:five years after the 1998 bleaching event. Philosophical Transaction of the 

Royal Society A 363:251-255. 

Sutherland KP, and Ritchie KM. 2004. White Pox Disease of the Carribean Elkhorn coral, 

Acropora palmata. In: Coral Health and Disease Springer-Verlag (Berlin) E Rosenberg and Y 

Loya (Eds) Chapter: 16:289-297. 

Szmant AM, and Forrester A. 1996. Water column and sediment nitrogen and phosphorus 

distribution patterns in the Florida Keys, USA. Coral Reef 15(1):21-41. 

Tilmant JT, Rutherford ES, and Thue EB. 1989. Fishery harvest and population dynamics of red 

drum (Sciaenops oscellatus) from Florida Bay and adjacent waters. Bulletin of Marine Science 

44(1):126-138. 

U.S. Census. 2010. United States Census Bureau. http://wwwcensusgov/. 

Vicente VP. 1990. Response of sponges with autotrophic endosymbionts during the coralbleaching episode in Puerto Rico (West Indies). Coral Reefs 8(4):199-202. 

Wagner DE, Kramer P, and Woesik RV. 2010. Species composition, habitat, and water quality 

influence coral bleaching in southern Florida Marine Ecology Progress Series 408:65-78. 

Ward-Paige CA, Risk MJ, Sherwood OA, and Jaap WC. 2005. Clionid sponge surveys on the 

Florida Reef Tract suggest land-based nutrient inputs. Marine Pollution Bulletion 51(5-7):570-

579. 

Weil E, Hernandez-Delgado EA, Bruckner AW, Ortiz AL, and Nemeth MRH. 2002. Distribution 

and status of Acroporid coral (Scleractinia) populations in Puerto Rico. In: Bruckner, AW (ed) 

NOAA Tech Memo NMFS-OPR-24(Silver Spring, MD):71-98. 

Wheaton JL, and Jaap WC. 1988. Corals and other prominent benthic Cnidaria of Looe Key 

National Marine Sanctuary, Florida. Florida Marine Research Publications 43:1-25. 

Williams EHJ, Bartels PJ, and Bunkley-Williams L. 1999. Predicted disappearance of coral-reef 

ramparts: a direct result of major ecological disturbances. global Change Biology 5(8):839-844. 

Wolff N, Grober-Dunsmore R, Rogers CS, and Beets J. 1999. Management implications of fish 

trap effectiveness in adjacent coral reef and gorgonian habitats Environmental Biology of Fishes 

55:81-90. 
2009 CREMP Final Report 62 F2345-02-10-F & F2635-06-09-F

Yoshioka PM. 1994. Size-specific life history pattern of a shallow-water gorgonian. Journal of 

Experimental Marine Biology and Ecology 184(1):111-122. 

Yoshioka PM, and Yoshioka BB. 1991. A comparison of the survivorship and growth of 

shallow-water gorgonian species of Puerto Rico. Marine Ecology Progress Series 69:253-260. 
2009 CREMP Final Report 63 F2345-02-10-F & F2635-06-09-F

APPENDICES 

Appendix 1. Percent cover values for major benthic taxa categories for all years at all 

region*habitat groupings. 

Region Habitat 

Sample 

Year Macroalgae 

Stony 

Corals Octocorals Sponges Zoanthids 

DT OD 1999 9.1 23.6 14.8 4.0 0.1 

2000 12.6 22.1 13.0 1.4 0.0 

2001 12.5 23.3 15.2 2.1 0.0 

2002 12.7 18.3 14.1 2.2 0.0 

2003 12.5 18.0 14.2 1.8 0.1 

2004 9.2 16.6 12.2 1.6 0.0 

2005 3.8 16.3 10.5 1.3 0.0 

2006 9.6 12.6 9.0 1.1 0.0 

2007 13.6 14.0 9.3 1.1 0.0 

2008 12.0 14.5 8.6 1.7 0.0 

DT P 1999 6.1 9.5 8.4 2.9 0.0 

2000 8.9 8.0 7.6 4.0 0.0 

2001 11.4 5.8 8.5 1.9 0.0 

2002 10.3 5.6 6.9 1.7 0.0 

2003 14.9 4.2 8.7 1.6 0.0 

2004 6.7 1.9 4.1 0.4 0.0 

2005 9.6 1.6 3.4 0.3 0.0 

2006 14.8 1.1 5.1 0.6 0.0 

2007 17.2 1.2 6.1 0.9 0.0 

2008 13.6 1.8 8.8 0.8 0.0 

LK BCP 1996 19.3 8.3 0.6 2.6 0.0 

1997 24.2 7.8 0.4 3.7 0.0 

1998 2.7 8.4 0.1 1.4 0.0 

1999 1.0 9.4 0.2 2.3 0.1 

2000 3.8 10.7 0.2 2.7 0.0 

2001 11.6 10.0 0.0 3.3 0.0 

2002 12.6 2.7 0.0 0.6 0.0 

2003 13.2 2.3 0.0 1.1 0.0 

2004 19.7 2.9 0.0 0.9 0.0 

2005 28.3 3.9 0.0 1.5 0.0 

2006 2.9 2.6 0.0 2.7 0.0 

2007 27.2 2.2 0.1 0.8 0.0 

2008 72.6 1.8 0.0 0.7 0.0 

LK OD 1996 15.5 7.6 10.9 3.7 0.3 

1997 19.8 7.9 9.7 3.4 0.3 

1998 43.0 5.3 11.6 2.5 0.1 

1999 15.6 4.2 8.5 1.8 0.2 

2000 27.7 4.3 8.4 2.5 0.2 

2001 23.8 4.4 7.5 2.5 0.1 

2002 15.5 4.2 8.7 4.0 0.2 

2003 11.0 3.9 9.6 2.3 0.1 

2004 20.4 3.1 11.3 2.1 0.2 

2005 4.5 3.2 10.6 2.5 0.2 

2006 14.2 1.9 8.8 1.0 0.1 
2009 CREMP Final Report 64 F2345-02-10-F & F2635-06-09-F

2007 19.8 2.3 8.3 1.6 0.2 

2008 12.2 2.4 9.9 3.1 0.2 

LK OS 1996 4.6 16.9 3.7 0.9 4.5 

1997 3.8 16.4 4.0 0.8 4.8 

1998 7.6 13.2 3.9 0.3 4.5 

1999 3.5 7.3 3.2 0.7 4.2 

2000 4.3 7.4 2.8 0.7 4.4 

2001 4.1 6.9 3.4 0.4 4.5 

2002 7.2 7.4 3.8 0.4 5.3 

2003 7.1 7.3 5.0 0.3 5.1 

2004 8.9 7.0 6.7 0.5 4.4 

2005 2.2 5.5 5.1 0.3 3.7 

2006 11.9 4.9 4.9 0.2 3.8 

2007 7.3 4.9 5.4 0.2 4.0 

2008 2.9 4.7 7.1 0.2 3.9 

LK P 1996 6.2 27.6 8.4 4.0 0.0 

1997 5.0 25.4 9.8 2.9 0.0 

1998 8.2 24.5 8.7 3.8 0.1 

1999 6.5 19.7 7.8 3.5 0.0 

2000 3.9 20.1 7.7 5.0 0.1 

2001 11.1 19.4 9.6 4.4 0.0 

2002 8.1 21.4 10.3 4.6 0.0 

2003 8.6 21.5 11.9 4.8 0.0 

2004 10.5 19.2 8.6 4.8 0.0 

2005 4.3 21.1 11.3 2.8 0.0 

2006 8.7 20.9 8.2 2.1 0.1 

2007 10.2 20.0 9.6 4.6 0.1 

2008 10.8 19.3 11.2 4.8 0.0 

MK OD 1996 4.0 4.1 12.5 4.9 0.5 

1997 6.2 4.8 10.7 4.9 0.1 

1998 41.7 3.4 10.3 3.2 0.0 

1999 15.3 3.0 5.3 2.2 0.0 

2000 20.3 2.8 4.9 3.4 0.0 

2001 16.6 3.0 6.1 4.6 0.0 

2002 4.1 2.7 6.4 4.0 0.0 

2003 5.9 2.9 6.7 4.5 0.0 

2004 13.5 2.5 10.4 4.4 0.0 

2005 8.1 2.7 10.4 3.8 0.0 

2006 22.7 2.3 7.7 2.3 0.0 

2007 13.3 2.0 8.4 2.7 0.0 

2008 16.2 2.2 8.7 3.6 0.0 

MK OS 1996 11.5 4.6 16.0 3.7 6.9 

1997 6.5 4.2 14.8 2.1 7.5 

1998 20.7 3.7 13.8 1.6 6.4 

1999 11.4 2.0 8.5 2.2 6.6 

2000 11.1 2.0 8.3 2.5 7.3 

2001 12.4 1.9 8.4 1.7 7.0 

2002 5.5 2.1 12.2 1.7 8.2 

2003 18.2 1.8 14.7 1.7 6.5 

2004 12.0 1.7 15.9 1.4 6.7 
2009 CREMP Final Report 65 F2345-02-10-F & F2635-06-09-F

2005 10.6 1.8 16.3 1.6 6.0 

2006 11.9 1.4 13.7 0.8 5.5 

2007 9.6 1.2 13.8 1.2 5.2 

2008 8.2 1.7 18.1 1.3 4.9 

MK P 1996 4.6 15.8 32.5 5.5 2.0 

1997 4.6 15.6 31.2 5.4 2.3 

1998 11.4 14.4 27.0 5.5 1.8 

1999 5.9 13.2 26.2 5.5 2.2 

2000 5.3 14.2 24.4 6.3 2.3 

2001 6.4 14.1 32.4 6.2 1.8 

2002 3.8 16.3 33.1 7.1 2.2 

2003 3.3 14.7 34.4 5.2 2.2 

2004 6.2 13.6 36.1 3.8 1.6 

2005 4.2 15.1 34.0 3.9 1.8 

2006 3.1 14.2 28.7 3.3 2.0 

2007 5.9 15.3 29.5 5.2 2.4 

2008 4.4 14.3 32.5 4.9 1.7 

UK OD 1996 13.9 7.4 12.6 6.2 0.0 

1997 16.6 6.1 11.2 5.4 0.2 

1998 26.4 3.5 10.4 4.2 0.0 

1999 7.9 3.4 8.3 4.2 0.2 

2000 20.4 3.6 9.5 4.4 0.0 

2001 15.0 2.9 9.7 3.9 0.0 

2002 11.7 3.0 12.3 5.1 0.1 

2003 15.5 3.1 12.1 4.5 0.0 

2004 19.9 2.7 14.4 4.2 0.1 

2005 8.2 2.7 14.1 4.1 0.1 

2006 11.1 2.3 14.4 3.2 0.0 

2007 10.2 3.2 13.3 4.2 0.1 

2008 12.7 3.0 11.8 4.7 0.0 

UK OS 1996 8.8 12.1 9.5 1.7 2.3 

1997 5.6 11.2 7.8 1.2 3.2 

1998 11.9 8.3 8.7 0.8 2.7 

1999 4.8 5.9 7.3 0.5 2.6 

2000 6.3 5.9 7.4 1.0 2.9 

2001 7.2 5.6 9.1 0.8 2.5 

2002 7.9 5.6 9.9 0.6 3.2 

2003 12.6 5.2 13.1 0.7 3.1 

2004 11.0 4.9 14.0 0.4 2.6 

2005 8.3 4.8 15.9 0.7 2.7 

2006 3.1 4.1 15.6 0.4 2.7 

2007 5.7 4.5 15.8 0.5 2.8 

2008 7.9 4.8 16.1 0.5 3.4 

UK P 1996 7.4 16.3 24.0 3.9 0.3 

1997 5.6 15.2 23.4 1.7 0.4 

1998 17.9 14.0 22.8 1.4 0.3 

1999 8.1 11.9 22.2 1.4 0.4 

2000 7.5 11.3 20.8 1.8 0.4 

2001 5.7 12.2 23.9 1.9 0.4 

2002 12.2 12.2 24.4 1.9 0.4 
2009 CREMP Final Report 66 F2345-02-10-F & F2635-06-09-F

2003 8.6 12.6 25.6 1.7 0.3 

2004 7.9 12.5 26.1 1.4 0.3 

2005 7.5 11.8 26.9 1.5 0.4 

2006 6.2 11.0 22.1 1.4 0.3 

2007 9.7 13.2 22.2 1.2 0.3 

2008 8.8 12.7 23.7 1.1 0.3 
2009 CREMP Final Report 67 F2345-02-10-F & F2635-06-09-F

Appendix 2. Percent cover values for major benthic taxa categories for all years at all CREMP 

sites. 

Site Name and 

Location 

Sample 

Year Macroalgae 

Stony 

Corals Octocorals Sponges Zoanthids 

Admiral 1996 3.4 30.0 25.7 0.2 0.1 

25.0447 oN 1997 3.7 27.6 23.0 0.4 0.1 

80.3948 oW 1998 7.4 26.6 26.8 0.3 0.1 

1999 5.0 22.5 23.8 0.4 0.1 

2000 7.2 20.0 23.3 0.9 0.1 

2001 2.7 21.2 25.0 0.4 0.1 

2002 6.6 21.6 24.4 0.4 0.2 

2003 6.6 23.1 28.0 0.6 0.2 

2004 3.5 23.1 27.7 0.5 0.1 

2005 6.1 21.7 30.3 0.5 0.1 

2006 2.7 19.9 23.7 0.6 0.1 

2007 8.9 24.6 23.0 0.5 0.1 

2008 5.6 23.3 26.9 0.3 0.1 

Alligator Deep 1996 0.9 1.9 16.4 2.9 0.0 

24.8452 oN 1997 11.3 2.3 14.6 3.3 0.1 

80.6209 oW 1998 25.3 1.2 15.5 2.9 0.0 

1999 1.1 1.3 9.5 2.8 0.0 

2000 12.4 1.1 7.0 3.2 0.0 

2001 2.6 1.3 7.5 3.9 0.0 

2002 7.4 1.3 9.2 2.9 0.0 

2003 3.4 1.1 9.7 4.3 0.0 

2004 1.8 0.9 14.9 4.5 0.1 

2005 3.6 1.1 14.8 3.3 0.0 

2006 8.0 0.8 10.4 2.0 0.0 

2007 2.5 1.0 13.2 1.6 0.0 

2008 1.5 1.3 13.0 2.5 0.0 

Alligator Shallow 1996 26.2 2.0 15.2 7.7 1.3 

24.8461 oN 1997 14.4 2.1 14.0 0.9 0.6 

80.6236 oW 1998 25.0 1.4 15.3 0.8 0.4 

1999 28.4 0.9 8.2 1.9 0.9 

2000 19.7 1.0 6.2 1.2 0.7 

2001 38.1 0.6 6.4 0.4 0.6 

2002 13.9 0.5 10.5 0.7 0.9 

2003 37.6 0.3 11.2 0.4 0.7 

2004 22.3 0.6 14.0 0.4 0.6 

2005 15.8 0.9 17.8 0.7 0.7 

2006 10.1 0.4 15.9 0.7 1.4 

2007 14.3 0.3 17.0 0.5 1.3 

2008 8.0 0.5 20.1 0.6 1.2 

Bird Key Reef 1999 11.0 20.6 19.6 2.0 0.0 

24.6117 oN 2000 14.0 21.5 17.1 0.7 0.0 

82.8702 oW 2001 10.6 22.3 17.5 0.5 0.0 

2002 21.0 15.0 18.8 1.0 0.0 
2009 CREMP Final Report 68 F2345-02-10-F & F2635-06-09-F

2003 1.8 15.0 17.4 0.6 0.0 

2004 0.6 12.6 14.3 0.5 0.0 

2005 6.5 10.8 13.1 0.4 0.0 

2006 2.6 9.5 12.3 0.5 0.0 

2007 17.0 10.6 12.2 0.6 0.0 

2008 22.8 10.6 9.7 0.7 0.0 

Black Coral Rock 1999 7.1 26.5 10.0 6.0 0.2 

24.6993 oN 2000 11.3 22.6 9.0 2.1 0.0 

83.0022 oW 2001 14.4 24.3 13.0 3.8 0.0 

2002 4.5 21.6 9.4 3.3 0.0 

2003 23.3 21.1 11.1 3.0 0.1 

2004 17.7 20.6 10.0 2.8 0.1 

2005 1.2 21.8 7.9 2.3 0.0 

2006 16.7 15.8 5.7 1.7 0.0 

2007 10.2 17.5 6.4 1.6 0.0 

2008 1.2 18.5 7.4 2.7 0.0 

Carysfort Deep 1996 24.7 14.3 8.4 4.0 0.0 

25.2208 oN 1997 31.0 11.0 10.2 2.1 0.3 

80.2099 oW 1998 20.1 5.4 9.2 0.6 0.0 

1999 12.8 5.2 6.6 0.8 0.6 

2000 27.3 6.4 9.3 0.8 0.0 

2001 17.7 4.8 9.8 0.9 0.0 

2002 9.2 5.0 14.4 2.2 0.0 

2003 31.1 5.0 10.1 0.6 0.0 

2004 44.5 3.6 10.6 1.2 0.0 

2005 6.6 3.7 10.6 1.0 0.0 

2006 9.3 3.6 9.5 0.8 0.0 

2007 9.3 4.9 12.1 1.6 0.0 

2008 10.0 4.1 10.4 1.3 0.0 

Carysfort Shallow 1996 6.0 10.9 8.3 0.4 2.7 

25.2210 oN 1997 3.3 10.9 7.5 0.2 4.6 

80.2103 oW 1998 4.5 7.5 8.0 0.2 4.0 

1999 1.1 3.6 6.0 0.0 2.9 

2000 2.9 3.5 6.7 0.2 3.1 

2001 2.8 3.7 9.0 0.1 3.2 

2002 2.4 4.1 8.9 0.1 4.9 

2003 1.3 3.3 11.3 0.0 5.3 

2004 1.1 3.1 10.1 0.0 4.1 

2005 0.7 3.5 9.4 0.1 4.2 

2006 0.9 2.5 9.5 0.1 4.1 

2007 1.1 2.5 9.1 0.1 3.2 

2008 3.1 2.9 9.7 0.1 5.2 

Cliff Green 1996 0.7 19.6 22.4 7.8 0.0 

24.5036 oN 1997 1.7 16.3 21.8 4.9 0.0 

81.7677 oW 1998 1.7 15.6 25.1 4.1 0.1 

1999 0.4 16.4 22.3 5.9 0.0 

2000 1.0 14.9 26.1 4.7 0.0 
2009 CREMP Final Report 69 F2345-02-10-F & F2635-06-09-F

2001 1.0 14.8 25.8 6.0 0.1 

2002 0.1 16.8 33.6 5.9 0.1 

2003 2.3 15.0 34.2 5.0 0.1 

2004 2.8 11.3 24.6 5.0 0.1 

2005 0.2 13.7 29.7 2.8 0.0 

2006 0.7 12.4 22.0 2.9 0.3 

2007 1.2 13.3 23.3 7.4 0.3 

2008 0.8 12.3 30.8 6.8 0.0 

Conch Deep 1996 6.4 5.0 9.4 5.8 0.0 

24.9519 oN 1997 13.2 3.5 8.6 4.3 0.0 

80.4513 oW 1998 33.3 3.2 8.3 6.4 0.0 

1999 10.0 2.6 7.8 3.4 0.0 

2000 30.0 1.7 6.2 3.4 0.0 

2001 22.1 2.4 6.9 4.2 0.0 

2002 19.7 2.4 9.9 5.0 0.0 

2003 13.5 2.1 9.8 5.2 0.0 

2004 10.9 2.7 13.4 5.3 0.0 

2005 15.0 2.6 10.8 4.6 0.0 

2006 8.9 1.8 13.8 3.0 0.0 

2007 11.2 2.9 13.2 4.9 0.0 

2008 16.2 3.3 11.6 5.6 0.0 

Conch Shallow 1996 18.8 5.9 8.6 3.2 0.9 

24.9559 oN 1997 8.3 5.9 5.7 2.9 1.2 

80.4574 oW 1998 24.1 4.1 6.4 0.6 1.5 

1999 10.9 3.0 4.2 0.5 1.1 

2000 8.3 3.5 3.0 1.3 1.3 

2001 11.7 3.3 3.4 1.0 0.9 

2002 13.0 2.4 6.1 0.3 1.3 

2003 32.9 2.2 7.8 0.5 1.0 

2004 17.4 1.6 10.2 0.3 1.1 

2005 25.0 1.1 15.8 0.6 1.2 

2006 4.5 1.1 13.9 0.4 0.8 

2007 9.2 1.4 14.2 0.3 1.0 

2008 14.1 1.6 14.6 0.3 1.2 

Content Keys 1996 31.3 1.2 0.0 1.9 0.0 

24.8221 oN 1997 33.4 1.1 0.1 5.7 0.0 

81.4889 oW 1998 0.8 1.4 0.0 1.7 0.0 

1999 1.1 1.4 0.0 2.6 0.0 

2000 7.3 1.2 0.0 2.5 0.0 

2001 20.7 2.3 0.0 3.5 0.0 

2002 0.3 0.5 0.0 0.2 0.0 

2003 2.6 0.6 0.0 0.6 0.0 

2004 31.3 0.4 0.0 0.5 0.0 

2005 55.6 0.6 0.0 0.8 0.0 

2006 5.1 0.3 0.0 3.3 0.0 

2007 35.2 0.4 0.2 0.4 0.0 

2008 84.6 0.2 0.0 1.0 0.0 
2009 CREMP Final Report 70 F2345-02-10-F & F2635-06-09-F

Dustan Rocks 1996 8.3 17.0 34.8 3.9 1.2 

24.6895 oN 1997 4.3 16.5 37.9 3.1 0.9 

81.0302 oW 1998 22.3 15.5 26.0 1.9 0.7 

1999 9.3 14.8 28.1 2.9 0.8 

2000 9.6 16.2 27.4 3.0 0.8 

2001 11.9 14.6 38.7 3.3 0.7 

2002 7.6 17.1 40.7 3.2 0.7 

2003 7.0 16.1 36.7 2.3 0.7 

2004 12.6 15.2 37.7 1.5 0.8 

2005 7.0 16.1 37.4 1.8 0.9 

2006 4.0 16.9 28.2 1.3 0.7 

2007 7.1 18.0 32.0 2.5 0.7 

2008 9.4 17.2 34.8 2.5 0.6 

Eastern Sambo Deep 1996 6.0 8.5 9.0 6.0 0.1 

24.4884 oN 1997 11.5 8.0 9.7 4.4 0.1 

81.6659 oW 1998 49.1 6.2 9.5 2.7 0.0 

1999 7.8 5.9 9.3 2.1 0.1 

2000 17.5 6.4 8.0 2.6 0.0 

2001 17.6 5.9 8.4 3.0 0.0 

2002 15.1 5.9 10.1 6.1 0.0 

2003 13.6 5.2 7.9 1.7 0.0 

2004 26.6 4.3 10.0 1.7 0.0 

2005 10.9 4.6 11.2 2.9 0.0 

2006 5.6 2.9 11.3 1.2 0.0 

2007 11.1 3.1 7.0 2.2 0.0 

2008 8.4 2.9 9.4 6.2 0.0 

Eastern Sambo 

Shallow 

1996 8.9 18.5 1.1 1.0 10.0 

1997 7.4 16.2 0.8 1.6 9.8 

24.4917 oN 1998 4.1 12.1 0.9 0.9 8.6 

81.6636 oW 1999 1.4 6.0 0.5 0.4 9.0 

2000 1.9 5.8 0.5 1.2 9.0 

2001 3.3 5.2 0.9 0.3 11.0 

2002 5.3 5.7 1.6 0.5 11.9 

2003 9.6 5.2 2.1 0.5 10.4 

2004 15.3 4.5 2.3 0.7 8.9 

2005 2.4 3.8 1.8 0.2 7.1 

2006 3.3 3.4 2.6 0.4 8.6 

2007 9.5 3.2 3.6 0.2 8.9 

2008 6.8 3.4 5.1 0.2 9.0 

Grecian Rocks 1996 4.0 17.8 11.2 1.8 2.3 

25.1075 oN 1997 7.4 16.3 9.7 1.0 2.7 

80.3069 oW 1998 13.7 13.6 10.6 1.3 2.4 

1999 6.0 11.2 11.0 1.0 2.9 

2000 11.2 11.2 10.7 1.3 3.0 

2001 11.7 10.3 12.4 0.8 2.8 

2002 12.8 10.2 13.3 1.1 2.7 

2003 11.1 9.9 15.9 1.4 2.6 
2009 CREMP Final Report 71 F2345-02-10-F & F2635-06-09-F

2004 15.1 8.9 17.5 0.9 1.9 

2005 5.2 8.6 16.8 1.1 2.0 

2006 3.4 7.8 18.3 0.6 1.9 

2007 5.6 8.0 18.6 0.9 2.5 

2008 7.0 8.3 20.7 0.9 2.5 

Jaap Reef 1996 16.0 30.7 1.3 0.5 0.1 

24.5857 oN 1997 9.8 33.5 1.9 1.7 0.0 

81.5826 oW 1998 18.1 27.3 1.2 3.5 0.0 

1999 15.9 18.1 0.7 2.7 0.0 

2000 12.1 15.9 0.4 4.1 0.2 

2001 26.1 16.0 1.1 1.4 0.0 

2002 25.9 17.3 0.7 1.7 0.0 

2003 26.3 21.4 1.1 1.4 0.0 

2004 31.4 18.6 0.9 1.5 0.0 

2005 13.7 20.4 2.0 1.4 0.1 

2006 23.8 21.2 0.9 0.9 0.0 

2007 27.3 18.9 1.0 1.2 0.0 

2008 33.1 20.8 1.2 1.0 0.0 

Looe Key Deep 1996 17.3 8.9 15.3 4.8 0.3 

24.5423 oN 1997 8.0 8.5 12.5 5.7 0.4 

81.4145 oW 1998 46.3 5.7 15.8 3.1 0.1 

1999 13.3 5.5 10.4 2.0 0.3 

2000 33.8 5.5 9.6 4.2 0.2 

2001 23.7 6.6 10.4 3.3 0.2 

2002 6.3 5.3 9.5 6.6 0.2 

2003 7.7 4.8 12.9 4.2 0.0 

2004 18.0 3.9 14.1 4.1 0.2 

2005 4.8 4.5 14.6 3.9 0.2 

2006 36.5 2.9 7.5 1.4 0.2 

2007 24.0 2.9 9.7 2.2 0.1 

2008 34.4 2.6 10.9 3.5 0.2 

Looe Key Shallow 1996 7.1 22.1 9.1 1.6 4.2 

24.5453 oN 1997 0.2 21.6 9.5 0.8 4.7 

81.4066 oW 1998 10.8 22.0 8.0 0.4 4.8 

1999 4.1 16.6 8.8 0.9 4.3 

2000 9.3 16.9 6.7 0.6 5.4 

2001 4.4 17.6 7.0 0.7 4.2 

2002 10.4 17.1 8.3 0.8 5.3 

2003 13.8 17.2 10.7 0.3 5.5 

2004 7.8 16.3 14.2 0.5 4.5 

2005 6.4 14.3 12.5 0.7 4.2 

2006 3.0 14.6 9.8 0.2 4.7 

2007 8.6 14.5 9.2 0.4 3.9 

2008 2.9 12.6 14.4 0.4 4.3 

Molasses Deep 1996 10.7 3.0 20.0 8.6 0.1 

25.0072 oN 1997 5.6 3.7 14.8 9.7 0.4 

80.3756 oW 1998 25.7 2.0 13.6 5.5 0.1 
2009 CREMP Final Report 72 F2345-02-10-F & F2635-06-09-F

1999 0.8 2.3 10.5 8.3 0.2 

2000 3.8 2.6 13.0 9.0 0.0 

2001 5.3 1.4 12.3 6.5 0.1 

2002 6.1 1.7 12.7 8.3 0.2 

2003 1.9 2.3 16.5 7.8 0.1 

2004 4.4 2.0 19.2 6.2 0.2 

2005 3.1 1.9 20.9 6.7 0.3 

2006 15.3 1.6 19.9 5.7 0.1 

2007 10.2 1.8 14.6 6.3 0.3 

2008 11.8 1.6 13.3 7.1 0.1 

Molasses Shallow 1996 8.0 11.7 9.2 1.2 3.3 

25.0089 oN 1997 2.8 10.1 7.7 0.6 4.5 

80.3753 oW 1998 4.9 6.3 9.0 0.7 3.0 

1999 0.8 3.9 6.9 0.5 3.3 

2000 1.3 3.5 8.0 1.2 4.1 

2001 1.3 3.6 10.6 1.3 2.8 

2002 1.8 4.4 10.2 0.7 4.1 

2003 5.8 3.9 16.5 0.5 3.8 

2004 9.1 4.9 17.1 0.3 3.6 

2005 3.3 4.6 21.1 0.9 3.7 

2006 3.4 3.8 19.6 0.4 4.2 

2007 6.9 4.9 20.2 0.5 4.5 

2008 7.9 5.1 17.9 0.6 4.8 

Porter Patch 1996 9.0 3.4 14.3 10.6 0.4 

25.1032 oN 1997 9.3 3.3 14.2 3.3 0.7 

80.3243 oW 1998 32.6 3.1 13.9 2.4 0.4 

1999 13.1 2.7 14.6 2.7 0.5 

2000 6.2 3.6 12.4 3.0 0.4 

2001 12.4 3.9 17.5 3.7 0.6 

2002 15.9 3.3 17.3 3.7 0.6 

2003 14.7 3.1 14.5 2.9 0.2 

2004 16.4 2.5 18.4 2.5 0.3 

2005 10.6 2.7 15.2 2.7 0.4 

2006 12.1 2.4 11.6 2.2 0.3 

2007 9.2 3.3 12.2 2.3 0.2 

2008 16.8 2.9 12.8 2.0 0.3 

Rock Key Deep 1996 1.2 4.8 13.7 4.4 0.5 

24.4532 oN 1997 15.9 8.0 10.2 3.4 0.4 

81.8568 oW 1998 30.7 4.4 12.9 2.9 0.4 

1999 9.6 3.9 10.0 2.4 0.3 

2000 14.8 3.1 9.3 2.4 0.3 

2001 21.8 3.2 6.6 3.3 0.3 

2002 3.8 4.3 7.8 3.0 0.4 

2003 5.5 3.3 13.6 2.7 0.3 

2004 7.1 2.7 13.4 2.0 0.2 

2005 1.6 2.7 12.3 2.6 0.2 

2006 5.5 1.7 7.9 0.8 0.3 
2009 CREMP Final Report 73 F2345-02-10-F & F2635-06-09-F

2007 11.6 2.6 9.6 1.5 0.3 

2008 12.8 2.5 8.3 2.4 0.3 

Rock Key Shallow 1996 1.7 11.5 3.1 0.9 4.3 

24.4546 oN 1997 2.5 10.5 4.4 0.6 4.8 

81.8584 oW 1998 9.1 7.4 5.1 0.2 4.6 

1999 5.2 3.3 3.1 0.4 4.2 

2000 4.0 3.0 3.8 0.5 4.2 

2001 5.7 2.8 4.2 0.2 4.7 

2002 10.5 4.0 3.7 0.2 5.0 

2003 2.5 4.0 5.5 0.3 5.3 

2004 3.9 3.7 6.6 0.3 5.1 

2005 0.5 2.3 4.4 0.2 4.4 

2006 15.4 1.7 4.2 0.1 3.9 

2007 6.6 1.7 4.9 0.2 4.4 

2008 1.8 1.8 5.3 0.1 4.0 

Sand Key Deep 1996 2.1 4.5 11.6 1.9 0.9 

24.4517 oN 1997 34.7 5.8 11.0 1.8 0.4 

81.8798 oW 1998 32.3 4.8 14.9 3.1 0.3 

1999 12.7 2.5 11.2 2.4 0.2 

2000 29.5 2.8 13.3 2.7 0.3 

2001 26.6 2.6 9.7 2.3 0.2 

2002 7.4 2.5 11.6 2.7 0.5 

2003 11.5 2.7 10.3 2.2 0.4 

2004 7.1 2.3 12.4 1.8 0.5 

2005 2.3 1.9 10.7 2.3 0.5 

2006 17.6 1.0 12.9 1.1 0.3 

2007 19.9 1.6 11.4 1.6 0.6 

2008 2.6 2.2 12.9 2.6 0.6 

Sand Key Shallow 1996 1.1 12.9 2.8 0.6 0.7 

24.4520 oN 1997 4.4 13.2 3.3 0.4 1.2 

81.8775 oW 1998 2.5 10.1 2.4 0.1 1.1 

1999 0.4 5.7 2.1 1.1 1.1 

2000 1.4 6.1 1.5 0.4 1.1 

2001 3.2 4.9 2.4 0.1 0.9 

2002 1.8 5.6 2.1 0.4 1.6 

2003 0.8 4.7 2.4 0.3 1.9 

2004 1.7 5.1 4.5 0.5 1.3 

2005 0.2 4.2 3.3 0.0 0.9 

2006 27.7 3.2 4.3 0.2 0.9 

2007 5.2 3.4 3.8 0.1 1.4 

2008 0.2 4.3 5.4 0.1 1.0 

Smith Shoal 1996 7.2 15.4 1.2 3.4 0.0 

24.7197 oN 1997 14.9 14.6 0.7 1.7 0.1 

81.9195 oW 1998 4.6 15.4 0.2 1.2 0.0 

1999 1.0 17.4 0.3 2.0 0.2 

2000 0.3 20.2 0.4 2.9 0.0 

2001 2.5 17.7 0.0 3.0 0.0 
2009 CREMP Final Report 74 F2345-02-10-F & F2635-06-09-F

2002 24.9 4.8 0.1 1.0 0.0 

2003 23.8 4.0 0.1 1.5 0.0 

2004 8.0 5.3 0.0 1.3 0.0 

2005 1.0 7.2 0.0 2.2 0.0 

2006 0.8 4.9 0.0 2.1 0.0 

2007 19.3 4.1 0.0 1.2 0.0 

2008 60.6 3.4 0.0 0.5 0.0 

Sombrero Deep 1996 7.6 3.7 10.8 3.9 1.3 

24.6231 oN 1997 4.0 4.6 7.1 3.3 0.1 

81.1105 oW 1998 54.0 3.8 6.7 2.2 0.0 

1999 25.2 2.8 2.1 1.7 0.0 

2000 31.2 2.0 2.0 3.1 0.0 

2001 25.7 3.2 4.2 4.9 0.0 

2002 1.9 2.2 2.7 2.4 0.0 

2003 11.1 3.0 3.3 3.0 0.0 

2004 10.2 3.0 6.4 2.0 0.0 

2005 9.6 2.5 5.5 3.1 0.0 

2006 40.6 1.6 3.9 1.4 0.0 

2007 23.9 1.3 3.3 1.5 0.0 

2008 29.3 1.5 3.6 2.6 0.0 

Sombrero Shallow 1996 5.0 7.1 10.1 1.3 15.6 

24.6259 oN 1997 4.7 6.2 10.6 2.9 17.5 

81.1101 oW 1998 20.8 5.5 9.5 1.7 15.1 

1999 4.8 3.1 7.1 2.3 15.0 

2000 9.0 2.6 8.6 2.7 17.0 

2001 2.1 2.7 8.2 1.7 15.9 

2002 2.8 3.2 12.6 1.5 18.6 

2003 12.9 2.8 15.6 1.1 15.3 

2004 10.6 2.5 15.2 1.3 15.5 

2005 5.6 2.3 13.6 1.7 13.5 

2006 14.4 2.1 10.8 0.9 11.7 

2007 6.5 2.0 11.6 1.2 11.0 

2008 8.0 2.5 16.2 1.4 10.3 

Tennessee Deep 1996 3.6 6.6 10.3 7.9 0.3 

24.7527 oN 1997 3.5 7.5 10.4 8.1 0.0 

80.7578 oW 1998 45.6 5.1 8.6 4.6 0.0 

1999 19.6 4.7 4.2 2.2 0.0 

2000 17.2 5.2 5.8 3.9 0.1 

2001 21.4 4.4 6.7 5.1 0.0 

2002 2.8 4.5 7.4 6.7 0.0 

2003 3.2 4.7 7.3 6.1 0.0 

2004 28.4 3.5 10.0 6.7 0.0 

2005 11.1 4.6 10.7 4.8 0.0 

2006 19.6 4.7 8.7 3.6 0.0 

2007 13.3 3.6 8.6 5.1 0.0 

2008 17.8 3.7 9.5 5.8 0.0 

Tennessee Shallow 1996 5.6 3.8 24.7 2.9 0.9 
2009 CREMP Final Report 75 F2345-02-10-F & F2635-06-09-F

24.7450 oN 1997 1.1 3.8 21.1 2.1 1.0 

80.7812 oW 1998 16.4 3.5 18.1 2.5 0.9 

1999 3.0 1.8 10.8 2.2 1.0 

2000 5.3 2.3 10.1 3.5 1.1 

2001 0.5 2.3 10.7 3.0 1.5 

2002 0.8 2.2 13.3 3.0 1.6 

2003 5.7 1.9 17.1 3.7 0.7 

2004 3.7 1.8 18.8 2.4 1.1 

2005 12.1 2.1 18.5 2.5 1.3 

2006 10.5 1.4 15.5 0.9 1.2 

2007 8.9 1.1 13.6 1.9 1.4 

2008 8.7 1.9 18.5 2.0 1.3 

Turtle 1996 12.8 8.3 35.0 1.2 0.7 

25.2947 oN 1997 3.7 8.3 38.2 2.0 0.5 

80.2191 oW 1998 17.1 5.4 28.1 2.2 0.7 

1999 6.8 4.4 30.1 1.4 0.8 

2000 10.0 5.5 28.3 1.9 0.9 

2001 1.6 6.6 31.4 2.1 0.7 

2002 17.9 6.4 35.3 2.2 0.7 

2003 3.3 5.9 37.3 2.2 0.8 

2004 3.9 6.1 34.6 1.3 0.7 

2005 5.7 5.5 37.6 1.6 0.9 

2006 4.1 6.1 34.8 1.8 0.8 

2007 12.1 5.5 35.6 1.0 0.7 

2008 3.3 6.3 33.7 1.3 0.9 

West Turtle Shoal 1996 1.8 14.9 30.7 6.6 2.7 

24.6993 oN 1997 4.8 14.9 26.2 7.1 3.4 

80.9669 oW 1998 3.3 13.5 27.7 8.1 2.6 

1999 3.4 12.0 24.8 7.4 3.3 

2000 2.0 12.7 22.2 8.8 3.4 

2001 2.3 13.7 27.6 8.4 2.6 

2002 0.9 15.7 27.4 10.0 3.4 

2003 0.5 13.8 32.7 7.4 3.2 

2004 1.4 12.4 35.0 5.6 2.3 

2005 2.1 14.4 31.5 5.5 2.5 

2006 2.4 12.2 29.0 4.8 2.9 

2007 4.9 13.4 27.7 7.3 3.6 

2008 0.7 12.1 30.8 6.7 2.5 

West Washer 

Women 1996 3.1 31.9 17.5 2.2 0.0 

24.5475 oN 1997 1.9 25.2 24.2 1.0 0.0 

81.5866 oW 1998 8.3 27.0 16.2 2.8 0.0 

1999 4.9 23.6 15.4 1.3 0.0 

2000 0.4 25.4 11.8 3.2 0.0 

2001 2.5 23.4 19.2 3.1 0.0 

2002 0.8 29.1 16.3 2.8 0.0 

2003 0.6 23.9 21.3 2.2 0.1 

2004 1.5 24.6 16.5 2.8 0.1 
2009 CREMP Final Report 76 F2345-02-10-F & F2635-06-09-F

2005 0.5 25.3 21.3 2.7 0.0 

2006 2.7 26.8 17.4 2.0 0.1 

2007 5.8 25.7 18.9 3.0 0.0 

2008 1.6 21.3 21.9 3.0 0.0 

Western Head 1996 2.2 26.9 0.2 6.0 0.0 

24.4977 oN 1997 4.4 23.5 0.1 4.0 0.0 

81.8056 oW 1998 2.7 26.0 0.4 4.7 0.4 

1999 2.3 21.1 0.1 4.1 0.0 

2000 0.1 24.3 0.0 7.1 0.0 

2001 8.5 23.3 0.8 7.3 0.0 

2002 0.3 23.5 0.4 7.9 0.1 

2003 0.4 24.4 1.6 9.7 0.0 

2004 0.6 21.3 0.3 9.4 0.0 

2005 0.2 24.0 1.6 4.2 0.0 

2006 2.8 22.3 0.3 2.7 0.0 

2007 2.0 21.7 2.8 7.3 0.0 

2008 1.3 21.4 1.1 8.4 0.0 

Western Sambo Deep 1996 45.9 10.2 5.9 1.5 0.0 

24.4785 oN 1997 27.4 9.4 5.2 2.0 0.0 

81.7161 oW 1998 52.3 5.2 5.3 0.7 0.0 

1999 32.5 3.3 2.3 0.3 0.0 

2000 38.7 3.1 2.2 0.8 0.0 

2001 28.7 3.3 2.3 0.9 0.0 

2002 41.3 3.2 4.0 1.1 0.0 

2003 14.8 3.4 4.5 0.7 0.0 

2004 39.0 2.3 7.5 0.8 0.0 

2005 1.8 2.3 4.6 0.9 0.0 

2006 3.0 1.0 4.2 0.5 0.0 

2007 29.8 1.2 4.3 0.6 0.0 

2008 2.9 1.7 7.6 0.8 0.0 

Western Sambo 

Shallow 

1996 5.0 21.1 2.7 0.6 3.5 

1997 5.0 22.4 2.0 0.7 3.7 

24.4797 oN 1998 11.0 16.5 2.6 0.1 3.4 

81.7170 oW 1999 5.8 6.2 1.6 0.8 2.3 

2000 4.9 6.5 1.4 0.7 2.4 

2001 3.4 5.3 2.0 0.9 1.7 

2002 6.8 5.7 3.3 0.3 2.5 

2003 10.1 6.2 4.2 0.2 2.2 

2004 17.6 6.6 5.8 0.4 2.0 

2005 1.9 3.8 3.9 0.3 1.7 

2006 9.0 2.9 3.8 0.4 1.2 

2007 6.9 2.7 5.6 0.1 1.1 

2008 3.0 2.6 5.9 0.2 1.0 

White Shoal 1999 6.1 9.5 8.4 2.9 0.0 

24.6416 oN 2000 8.9 8.0 7.6 4.0 0.0 

82.8964 oW 2001 11.4 5.8 8.5 1.9 0.0 

2002 10.3 5.6 6.9 1.7 0.0 
2009 CREMP Final Report 77 F2345-02-10-F & F2635-06-09-F

2003 14.9 4.2 8.7 1.6 0.0 

2004 6.7 1.9 4.1 0.4 0.0 

2005 9.6 1.6 3.4 0.3 0.0 

2006 14.8 1.1 5.1 0.6 0.0 

2007 17.2 1.2 6.1 0.9 0.0 

2008 13.6 1.8 8.8 0.8 0.0 
2009 CREMP Final Report 78 F2345-02-10-F & F2635-06-09-F

Appendix 3. Percent cover values for five sentinel coral species for all years at all region*habitat 

groupings. 

Region Habitat 

Sample 

Year 

Montastraea 

annularis 

Montastraea 

cavernosa 

Colpophyllia 

natans 

Siderastrea 

siderea 

Porites 

astreoides 

DT OD 1999 11.1 5.2 2.6 1.1 0.8 

2000 11.3 4.3 2.5 1.0 0.7 

2001 12.0 4.7 2.7 1.1 0.8 

2002 8.5 4.6 2.0 0.9 0.9 

2003 8.2 4.1 1.5 1.0 0.9 

2004 7.9 3.7 1.4 1.3 0.7 

2005 7.9 3.7 1.5 1.2 0.5 

2006 5.8 3.3 1.3 1.0 0.3 

2007 7.2 2.9 1.4 1.0 0.5 

2008 7.2 2.9 1.5 1.4 0.4 

DT P 1999 0.2 0.1 0.3 0.4 0.1 

2000 0.3 0.1 0.2 0.3 0.2 

2001 0.4 0.1 0.5 0.4 0.0 

2002 0.3 0.1 0.3 0.1 0.0 

2003 0.4 0.1 0.2 0.2 0.1 

2004 0.2 0.1 0.2 0.3 0.0 

2005 0.3 0.1 0.1 0.2 0.1 

2006 0.2 0.0 0.1 0.2 0.0 

2007 0.1 0.1 0.1 0.2 0.1 

2008 0.2 0.0 0.3 0.2 0.1 

LK BCP 1996 4.1 0.7 0.0 0.1 1.3 

1997 4.7 0.5 0.0 0.1 1.4 

1998 3.8 0.8 0.0 0.1 2.0 

1999 6.1 0.6 0.0 0.2 0.9 

2000 5.0 0.7 0.0 0.3 2.4 

2001 4.2 0.8 0.0 0.3 2.2 

2002 0.0 0.6 0.0 0.3 0.3 

2003 0.0 0.5 0.0 0.2 0.4 

2004 0.0 0.6 0.0 0.3 0.3 

2005 0.0 0.6 0.0 0.2 0.6 

2006 0.0 0.7 0.0 0.2 0.4 

2007 0.0 0.6 0.0 0.2 0.2 

2008 0.0 0.8 0.0 0.0 0.2 

LK OD 1996 2.2 0.9 0.3 0.8 0.2 

1997 2.4 0.6 0.4 1.0 0.3 

1998 2.1 0.5 0.3 0.6 0.3 

1999 2.0 0.5 0.2 0.8 0.1 

2000 2.0 0.6 0.2 0.6 0.1 

2001 1.9 0.6 0.2 0.6 0.2 

2002 1.7 0.7 0.2 0.6 0.2 

2003 1.4 0.5 0.2 0.6 0.2 

2004 1.1 0.4 0.1 0.5 0.1 

2005 1.3 0.4 0.1 0.6 0.1 
2009 CREMP Final Report 79 F2345-02-10-F & F2635-06-09-F

2006 0.6 0.3 0.0 0.5 0.1 

2007 0.6 0.2 0.1 0.7 0.1 

2008 0.5 0.3 0.1 0.7 0.1 

LK OS 1996 3.9 0.4 0.2 0.3 1.6 

1997 3.8 0.4 0.3 0.3 1.6 

1998 4.1 0.4 0.4 0.3 1.6 

1999 3.4 0.2 0.1 0.2 1.9 

2000 3.4 0.3 0.1 0.3 1.9 

2001 3.6 0.2 0.2 0.2 1.6 

2002 3.3 0.3 0.1 0.3 2.1 

2003 3.5 0.4 0.1 0.3 1.6 

2004 3.5 0.2 0.1 0.2 1.5 

2005 2.7 0.2 0.1 0.3 1.4 

2006 2.8 0.2 0.1 0.2 1.2 

2007 2.7 0.3 0.1 0.2 1.1 

2008 2.5 0.1 0.1 0.2 1.2 

LK P 1996 10.2 8.6 2.4 2.9 0.5 

1997 11.6 6.0 2.3 2.4 0.7 

1998 9.6 7.5 2.6 2.8 0.5 

1999 6.8 6.6 2.1 2.3 0.4 

2000 6.0 7.4 2.1 2.6 0.5 

2001 6.0 7.3 1.9 2.4 0.4 

2002 6.3 7.4 2.6 2.9 0.8 

2003 7.2 7.2 2.5 2.2 0.6 

2004 5.9 6.4 2.5 2.4 0.5 

2005 7.1 7.1 2.4 2.7 0.5 

2006 7.5 6.8 2.2 2.5 0.5 

2007 6.4 6.8 2.7 2.5 0.4 

2008 6.7 6.4 1.5 2.5 0.5 

MK OD 1996 1.0 0.5 0.2 0.7 0.1 

1997 1.3 0.2 0.4 0.6 0.2 

1998 1.1 0.1 0.2 0.5 0.2 

1999 1.2 0.1 0.0 0.7 0.2 

2000 1.1 0.3 0.0 0.7 0.2 

2001 0.9 0.2 0.0 0.7 0.2 

2002 0.8 0.1 0.0 0.6 0.2 

2003 1.0 0.2 0.0 0.7 0.2 

2004 0.7 0.2 0.1 0.6 0.1 

2005 0.7 0.2 0.0 0.6 0.2 

2006 0.7 0.2 0.0 0.7 0.1 

2007 0.7 0.1 0.0 0.5 0.1 

2008 0.5 0.1 0.0 0.6 0.1 

MK OS 1996 0.7 0.3 0.2 0.4 0.4 

1997 0.5 0.2 0.1 0.2 0.6 

1998 0.6 0.2 0.3 0.3 0.4 

1999 0.4 0.2 0.1 0.4 0.2 

2000 0.4 0.2 0.0 0.3 0.3 
2009 CREMP Final Report 80 F2345-02-10-F & F2635-06-09-F

2001 0.3 0.2 0.1 0.3 0.3 

2002 0.4 0.2 0.0 0.5 0.2 

2003 0.3 0.1 0.1 0.3 0.2 

2004 0.2 0.2 0.1 0.4 0.2 

2005 0.2 0.1 0.0 0.4 0.2 

2006 0.1 0.1 0.0 0.4 0.2 

2007 0.1 0.1 0.1 0.4 0.1 

2008 0.1 0.1 0.0 0.4 0.2 

MK P 1996 2.6 4.1 1.3 3.6 1.0 

1997 3.7 3.6 1.1 3.6 1.1 

1998 2.4 4.0 1.8 3.4 1.2 

1999 2.3 4.6 1.3 2.9 0.9 

2000 2.1 4.4 1.3 4.0 1.1 

2001 2.2 4.0 1.5 3.8 1.2 

2002 2.5 4.7 1.9 3.9 1.1 

2003 2.1 4.5 1.8 3.3 1.0 

2004 2.3 3.5 1.6 3.4 0.9 

2005 2.9 3.9 1.9 3.2 0.9 

2006 2.4 4.0 1.6 3.4 1.1 

2007 2.3 4.4 1.9 3.8 1.0 

2008 2.3 3.8 1.5 3.7 1.0 

UK OD 1996 2.9 0.4 0.2 0.8 0.2 

1997 2.1 0.3 0.0 0.8 0.2 

1998 0.9 0.3 0.0 0.8 0.2 

1999 0.8 0.2 0.1 0.7 0.1 

2000 0.9 0.3 0.0 0.8 0.1 

2001 0.6 0.3 0.1 0.7 0.2 

2002 0.7 0.2 0.0 0.6 0.1 

2003 0.7 0.2 0.1 0.5 0.3 

2004 0.6 0.2 0.0 0.6 0.2 

2005 0.6 0.2 0.0 0.7 0.1 

2006 0.5 0.1 0.1 0.6 0.3 

2007 0.6 0.2 0.1 0.7 0.4 

2008 0.4 0.2 0.1 0.6 0.2 

UK OS 1996 4.1 0.2 0.1 0.5 0.4 

1997 3.6 0.1 0.1 0.3 0.5 

1998 2.8 0.1 0.1 0.3 0.5 

1999 2.8 0.1 0.1 0.3 0.5 

2000 2.7 0.1 0.0 0.6 0.5 

2001 2.7 0.1 0.1 0.5 0.5 

2002 2.4 0.2 0.1 0.5 0.6 

2003 2.5 0.1 0.0 0.4 0.6 

2004 2.4 0.1 0.0 0.3 0.5 

2005 2.2 0.1 0.1 0.3 0.6 

2006 2.1 0.0 0.0 0.4 0.5 

2007 2.1 0.1 0.1 0.4 0.5 

2008 2.0 0.0 0.0 0.5 0.6 
2009 CREMP Final Report 81 F2345-02-10-F & F2635-06-09-F

UK P 1996 12.4 0.1 0.0 1.2 0.4 

1997 11.4 0.2 0.1 1.1 0.3 

1998 11.1 0.2 0.0 1.2 0.5 

1999 9.7 0.2 0.0 0.9 0.3 

2000 8.5 0.2 0.1 1.1 0.4 

2001 9.2 0.1 0.0 1.4 0.4 

2002 9.2 0.2 0.0 1.2 0.4 

2003 10.0 0.1 0.1 1.0 0.6 

2004 9.7 0.0 0.0 1.1 0.5 

2005 9.1 0.1 0.1 1.2 0.4 

2006 8.3 0.1 0.0 1.2 0.5 

2007 10.3 0.1 0.0 1.2 0.5 

2008 9.8 0.1 0.0 1.0 0.5 
2009 CREMP Final Report 82 F2345-02-10-F & F2635-06-09-F

Appendix 4. Percent cover values for five sentinel coral species for all years at all CREMP sites.

Site Name and 

Location 

Sample 

Year 

Montastraea 

annularis 

Montastraea 

cavernosa 

Colpophyllia 

natans 

Siderastrea 

siderea 

Porites 

astreoides 

Admiral 1996 27.0 0.0 0.1 1.5 0.2 

25.0447 oN 1997 24.9 0.0 0.1 1.2 0.2 

80.3948 oW 1998 24.1 0.0 0.0 1.4 0.5 

1999 20.8 0.0 0.0 1.0 0.3 

2000 18.2 0.1 0.0 0.8 0.3 

2001 19.3 0.1 0.0 1.3 0.2 

2002 19.5 0.1 0.0 1.4 0.3 

2003 21.4 0.0 0.0 1.1 0.3 

2004 21.1 0.0 0.0 1.2 0.4 

2005 19.6 0.0 0.0 1.4 0.4 

2006 17.9 0.1 0.0 1.1 0.5 

2007 22.4 0.0 0.0 1.3 0.5 

2008 21.2 0.0 0.0 1.2 0.5 

Alligator Deep 1996 0.1 0.1 0.0 0.1 0.0 

24.8452 oN 1997 0.4 0.1 0.0 0.2 0.0 

80.6209 oW 1998 0.2 0.0 0.0 0.1 0.0 

1999 0.1 0.0 0.0 0.2 0.0 

2000 0.1 0.1 0.0 0.1 0.1 

2001 0.1 0.0 0.0 0.2 0.0 

2002 0.0 0.0 0.0 0.2 0.0 

2003 0.0 0.0 0.0 0.1 0.0 

2004 0.0 0.0 0.0 0.1 0.0 

2005 0.0 0.0 0.0 0.2 0.0 

2006 0.0 0.0 0.0 0.2 0.0 

2007 0.0 0.0 0.0 0.2 0.0 

2008 0.0 0.0 0.0 0.2 0.0 

Alligator Shallow 1996 0.0 0.1 0.0 0.1 0.3 

24.8461 oN 1997 0.0 0.1 0.0 0.0 0.3 

80.6236 oW 1998 0.0 0.0 0.0 0.0 0.2 

1999 0.0 0.0 0.0 0.1 0.1 

2000 0.0 0.0 0.0 0.1 0.0 

2001 0.0 0.0 0.1 0.1 0.0 

2002 0.0 0.0 0.0 0.1 0.0 

2003 0.0 0.0 0.0 0.1 0.0 

2004 0.0 0.0 0.0 0.1 0.0 

2005 0.0 0.0 0.0 0.1 0.0 

2006 0.0 0.0 0.0 0.1 0.0 

2007 0.0 0.0 0.0 0.1 0.0 

2008 0.0 0.0 0.0 0.0 0.0 

Bird Key Reef 1999 10.5 1.9 3.0 1.9 0.9 

24.6117 oN 2000 11.5 2.5 2.9 1.6 1.0 

82.8702 oW 2001 11.6 2.2 3.5 1.7 1.0 

2002 7.8 2.5 1.1 1.2 1.3 

2003 7.1 1.9 0.7 1.7 1.1 

2004 6.7 1.4 0.6 2.0 0.9 

2005 5.5 1.5 0.4 1.7 0.5 
2009 CREMP Final Report 83 F2345-02-10-F & F2635-06-09-F

2006 4.8 1.8 0.6 1.4 0.2 

2007 6.1 1.0 0.7 1.6 0.4 

2008 5.9 0.8 0.6 2.1 0.4 

Black Coral Rock 1999 11.8 8.4 2.3 0.4 0.6 

24.6993 oN 2000 11.1 6.0 2.1 0.4 0.5 

83.0022 oW 2001 12.3 7.2 2.0 0.5 0.5 

2002 9.2 6.7 3.0 0.7 0.4 

2003 9.2 6.2 2.4 0.2 0.7 

2004 9.0 6.0 2.2 0.7 0.6 

2005 10.3 5.8 2.7 0.7 0.5 

2006 6.8 4.9 2.0 0.5 0.5 

2007 8.3 4.9 2.1 0.4 0.6 

2008 8.6 4.9 2.3 0.7 0.4 

Carysfort Deep 1996 7.4 1.0 0.4 1.3 0.2 

25.2208 oN 1997 5.4 0.6 0.1 1.4 0.4 

80.2099 oW 1998 1.7 0.6 0.1 1.9 0.4 

1999 1.9 0.3 0.4 1.6 0.1 

2000 2.3 0.9 0.1 1.7 0.3 

2001 1.3 0.6 0.2 1.3 0.4 

2002 1.8 0.6 0.1 1.0 0.4 

2003 1.8 0.4 0.3 0.9 0.8 

2004 1.5 0.4 0.1 0.8 0.3 

2005 1.3 0.5 0.1 1.4 0.1 

2006 1.3 0.2 0.3 1.0 0.3 

2007 1.7 0.5 0.1 1.0 0.9 

2008 0.8 0.4 0.3 1.0 0.3 

Carysfort Shallow 1996 0.8 0.2 0.0 0.2 0.9 

25.2210 oN 1997 0.1 0.0 0.0 0.2 1.3 

80.2103 oW 1998 0.2 0.0 0.0 0.1 1.2 

1999 0.2 0.0 0.0 0.1 1.3 

2000 0.1 0.1 0.0 0.4 1.7 

2001 0.2 0.0 0.0 0.7 1.7 

2002 0.3 0.2 0.0 0.4 2.0 

2003 0.1 0.0 0.0 0.2 1.8 

2004 0.1 0.0 0.0 0.2 1.5 

2005 0.2 0.0 0.0 0.2 2.0 

2006 0.1 0.0 0.0 0.4 1.6 

2007 0.1 0.0 0.0 0.3 1.4 

2008 0.2 0.0 0.0 0.3 1.7 

Cliff Green 1996 2.8 9.9 0.8 1.2 0.2 

24.5036 oN 1997 4.7 5.7 2.3 1.3 0.6 

81.7677 oW 1998 3.1 6.7 2.7 1.4 0.3 

1999 2.7 8.0 2.0 1.4 0.2 

2000 2.0 7.7 2.0 2.0 0.1 

2001 2.1 7.1 2.2 1.6 0.2 

2002 2.0 7.5 2.9 2.2 0.6 

2003 1.6 7.3 2.5 1.4 0.2 

2004 0.8 5.3 1.6 1.3 0.1 

2005 0.9 6.7 1.6 2.5 0.2 
2009 CREMP Final Report 84 F2345-02-10-F & F2635-06-09-F

2006 1.0 6.8 1.4 2.2 0.0 

2007 0.3 7.3 2.4 2.1 0.2 

2008 0.9 5.5 1.2 2.5 0.2 

Conch Deep 1996 0.9 0.1 0.0 0.9 0.3 

24.9519 oN 1997 1.0 0.2 0.0 0.5 0.2 

80.4513 oW 1998 0.8 0.2 0.0 0.4 0.2 

1999 0.6 0.3 0.0 0.5 0.2 

2000 0.5 0.1 0.0 0.4 0.1 

2001 0.6 0.1 0.0 0.9 0.1 

2002 0.2 0.1 0.0 0.7 0.0 

2003 0.2 0.0 0.0 0.5 0.0 

2004 0.4 0.1 0.0 0.7 0.3 

2005 0.6 0.0 0.0 0.6 0.1 

2006 0.2 0.1 0.0 0.7 0.3 

2007 0.2 0.0 0.0 0.7 0.2 

2008 0.4 0.0 0.0 0.8 0.3 

Conch Shallow 1996 0.0 0.0 0.0 0.3 0.0 

24.9559 oN 1997 0.0 0.0 0.0 0.2 0.0 

80.4574 oW 1998 0.0 0.0 0.0 0.3 0.0 

1999 0.0 0.0 0.0 0.2 0.0 

2000 0.0 0.0 0.0 0.3 0.0 

2001 0.0 0.0 0.0 0.3 0.0 

2002 0.0 0.0 0.0 0.3 0.0 

2003 0.0 0.0 0.0 0.2 0.0 

2004 0.0 0.0 0.0 0.2 0.0 

2005 0.0 0.0 0.0 0.2 0.0 

2006 0.0 0.0 0.0 0.3 0.0 

2007 0.0 0.0 0.0 0.2 0.0 

2008 0.0 0.0 0.0 0.2 0.0 

Content Keys 1996 0.0 0.0 0.0 0.1 0.5 

24.8221 oN 1997 0.0 0.0 0.0 0.0 0.4 

81.4889 oW 1998 0.0 0.0 0.0 0.0 0.5 

1999 0.0 0.0 0.0 0.3 0.7 

2000 0.0 0.0 0.0 0.0 0.6 

2001 0.0 0.0 0.0 0.4 1.2 

2002 0.0 0.0 0.0 0.2 0.2 

2003 0.0 0.0 0.0 0.1 0.2 

2004 0.0 0.0 0.0 0.1 0.1 

2005 0.0 0.0 0.0 0.1 0.2 

2006 0.0 0.0 0.0 0.1 0.1 

2007 0.0 0.0 0.0 0.1 0.1 

2008 0.0 0.0 0.0 0.0 0.1 

Dustan Rocks 1996 5.1 3.6 0.6 4.4 2.1 

24.6895 oN 1997 5.5 3.2 0.4 3.8 1.4 

81.0302 oW 1998 4.6 3.8 0.6 3.7 2.1 

1999 4.4 3.7 0.8 3.7 1.2 

2000 4.3 4.6 0.7 4.0 1.4 

2001 3.8 3.8 0.6 3.6 1.5 

2002 4.7 4.4 0.9 4.1 1.1 
2009 CREMP Final Report 85 F2345-02-10-F & F2635-06-09-F

2003 4.3 4.6 0.5 3.6 1.5 

2004 4.5 3.4 0.8 3.8 1.3 

2005 5.4 3.5 0.7 3.5 1.2 

2006 4.8 4.3 0.6 3.9 1.5 

2007 4.8 5.0 0.8 4.5 1.4 

2008 4.7 4.3 0.8 4.2 1.6 

Eastern Sambo Deep 1996 3.4 1.5 0.3 1.2 0.4 

24.4884 oN 1997 2.8 1.4 0.7 1.7 0.3 

81.6659 oW 1998 3.0 1.0 0.5 0.9 0.4 

1999 2.4 1.3 0.3 1.3 0.1 

2000 3.5 1.4 0.0 0.8 0.3 

2001 2.4 1.3 0.2 0.9 0.2 

2002 2.8 1.6 0.1 0.7 0.3 

2003 2.1 1.3 0.3 0.7 0.1 

2004 1.8 1.0 0.2 0.8 0.1 

2005 1.5 1.0 0.2 1.1 0.3 

2006 0.7 0.7 0.0 1.1 0.1 

2007 0.7 0.6 0.2 1.0 0.2 

2008 0.4 0.5 0.2 0.9 0.2 

Eastern Sambo 

Shallow 

1996 0.3 0.0 0.0 0.0 3.7 

1997 0.4 0.0 0.0 0.1 3.6 

24.4917 oN 1998 0.2 0.1 0.0 0.2 4.2 

81.6636 oW 1999 0.3 0.0 0.0 0.2 5.0 

2000 0.4 0.0 0.0 0.0 5.1 

2001 0.5 0.1 0.0 0.2 4.3 

2002 0.1 0.0 0.0 0.2 5.1 

2003 0.7 0.1 0.0 0.1 4.0 

2004 0.5 0.1 0.0 0.1 3.6 

2005 0.2 0.0 0.0 0.1 3.2 

2006 0.2 0.0 0.0 0.2 2.8 

2007 0.2 0.0 0.0 0.2 2.6 

2008 0.1 0.0 0.0 0.1 2.8 

Grecian Rocks 1996 10.9 0.4 0.3 1.0 0.5 

25.1075 oN 1997 10.2 0.1 0.3 0.6 0.6 

80.3069 oW 1998 8.1 0.1 0.3 0.6 0.5 

1999 7.6 0.2 0.4 0.7 0.5 

2000 8.3 0.2 0.0 1.2 0.3 

2001 8.0 0.3 0.2 0.6 0.2 

2002 7.0 0.3 0.1 1.1 0.4 

2003 7.3 0.2 0.0 1.0 0.4 

2004 6.8 0.2 0.1 0.5 0.3 

2005 6.4 0.2 0.1 0.6 0.4 

2006 6.2 0.1 0.0 0.7 0.4 

2007 5.9 0.2 0.2 0.6 0.4 

2008 5.8 0.1 0.1 1.1 0.4 

Jaap Reef 1996 26.8 0.8 0.7 0.4 1.0 

24.5857 oN 1997 29.7 0.2 0.6 0.0 1.3 

81.5826 oW 1998 25.1 0.6 0.4 0.0 0.7 

1999 16.5 0.3 0.3 0.0 0.8 
2009 CREMP Final Report 86 F2345-02-10-F & F2635-06-09-F

2000 14.0 0.1 0.4 0.2 0.9 

2001 14.4 0.0 0.5 0.0 0.7 

2002 15.3 0.2 0.6 0.1 0.8 

2003 19.2 0.2 0.6 0.2 0.8 

2004 16.4 0.3 0.6 0.1 0.8 

2005 18.6 0.1 0.5 0.0 1.0 

2006 19.8 0.0 0.4 0.0 0.7 

2007 17.3 0.2 0.6 0.1 0.6 

2008 18.5 0.1 0.3 0.1 0.9 

Looe Key Deep 1996 2.4 1.5 0.6 1.3 0.3 

24.5423 oN 1997 3.4 1.0 0.5 1.3 0.3 

81.4145 oW 1998 2.7 0.8 0.5 0.6 0.3 

1999 2.9 0.6 0.2 1.1 0.1 

2000 2.3 0.6 0.3 1.1 0.1 

2001 3.0 0.8 0.2 0.9 0.2 

2002 2.1 0.6 0.3 0.9 0.2 

2003 1.6 0.5 0.2 0.9 0.1 

2004 1.5 0.3 0.1 0.6 0.2 

2005 2.1 0.5 0.1 0.8 0.1 

2006 1.3 0.3 0.0 0.6 0.2 

2007 0.8 0.3 0.0 1.0 0.1 

2008 0.7 0.4 0.0 0.8 0.0 

Looe Key Shallow 1996 16.0 1.3 0.7 0.4 0.5 

24.5453 oN 1997 16.1 1.3 1.5 0.5 0.5 

81.4066 oW 1998 16.6 1.7 1.4 0.5 0.4 

1999 13.6 0.8 0.4 0.1 0.7 

2000 13.7 1.0 0.6 0.2 0.4 

2001 14.5 0.7 0.8 0.4 0.3 

2002 13.5 1.0 0.5 0.3 0.6 

2003 14.0 1.2 0.4 0.4 0.4 

2004 13.8 0.7 0.4 0.3 0.4 

2005 11.4 0.7 0.5 0.4 0.3 

2006 11.9 0.8 0.4 0.3 0.4 

2007 11.5 1.1 0.4 0.3 0.4 

2008 9.8 0.5 0.4 0.4 0.5 

Molasses Deep 1996 0.4 0.1 0.1 0.1 0.1 

25.0072 oN 1997 0.0 0.2 0.0 0.3 0.0 

80.3756 oW 1998 0.2 0.1 0.0 0.1 0.1 

1999 0.1 0.1 0.0 0.2 0.1 

2000 0.0 0.0 0.0 0.2 0.0 

2001 0.0 0.1 0.0 0.1 0.0 

2002 0.0 0.0 0.0 0.2 0.1 

2003 0.0 0.1 0.0 0.2 0.1 

2004 0.0 0.0 0.0 0.2 0.1 

2005 0.0 0.1 0.0 0.0 0.1 

2006 0.0 0.0 0.0 0.1 0.1 

2007 0.0 0.1 0.1 0.2 0.1 

2008 0.0 0.0 0.0 0.1 0.1 

Molasses Shallow 1996 2.2 0.3 0.0 0.1 0.2 
2009 CREMP Final Report 87 F2345-02-10-F & F2635-06-09-F

25.0089 oN 1997 2.1 0.2 0.3 0.1 0.3 

80.3753 oW 1998 1.3 0.1 0.0 0.2 0.2 

1999 1.7 0.1 0.0 0.1 0.2 

2000 0.7 0.1 0.1 0.1 0.2 

2001 0.9 0.1 0.1 0.3 0.3 

2002 0.8 0.2 0.1 0.1 0.2 

2003 1.1 0.1 0.0 0.1 0.3 

2004 1.3 0.1 0.0 0.1 0.1 

2005 1.0 0.1 0.2 0.1 0.2 

2006 0.7 0.0 0.0 0.2 0.2 

2007 1.1 0.1 0.1 0.2 0.2 

2008 0.9 0.1 0.0 0.2 0.2 

Porter Patch 1996 0.6 0.0 0.0 1.0 0.1 

25.1032 oN 1997 0.3 0.1 0.1 1.1 0.0 

80.3243 oW 1998 0.6 0.1 0.0 0.7 0.1 

1999 0.7 0.2 0.0 0.7 0.0 

2000 0.5 0.2 0.0 1.5 0.0 

2001 0.7 0.1 0.0 1.4 0.1 

2002 0.8 0.2 0.0 0.9 0.1 

2003 0.5 0.1 0.0 1.0 0.0 

2004 0.4 0.1 0.0 0.7 0.1 

2005 0.5 0.0 0.0 1.1 0.1 

2006 0.3 0.0 0.0 1.0 0.0 

2007 0.5 0.0 0.0 1.2 0.1 

2008 0.4 0.0 0.0 1.0 0.0 

Rock Key Deep 1996 1.3 0.9 0.6 0.5 0.0 

24.4532 oN 1997 2.2 0.9 0.8 0.9 0.6 

81.8568 oW 1998 1.2 1.0 0.4 0.5 0.4 

1999 1.4 0.5 0.4 0.3 0.4 

2000 0.7 1.1 0.1 0.3 0.2 

2001 1.0 0.7 0.2 0.3 0.3 

2002 1.0 1.6 0.3 0.6 0.3 

2003 0.8 0.6 0.0 0.3 0.5 

2004 0.5 0.4 0.0 0.5 0.3 

2005 0.6 0.5 0.3 0.3 0.2 

2006 0.2 0.5 0.0 0.5 0.2 

2007 0.6 0.4 0.0 0.7 0.2 

2008 0.3 0.4 0.1 0.6 0.2 

Rock Key Shallow 1996 0.4 0.1 0.0 0.2 1.3 

24.4546 oN 1997 0.3 0.2 0.0 0.2 1.2 

81.8584 oW 1998 0.4 0.0 0.0 0.1 1.1 

1999 0.4 0.0 0.0 0.1 1.2 

2000 0.3 0.1 0.0 0.2 0.9 

2001 0.3 0.1 0.0 0.1 1.1 

2002 0.3 0.0 0.0 0.1 1.4 

2003 0.8 0.3 0.0 0.1 0.6 

2004 0.4 0.1 0.0 0.1 1.1 

2005 0.2 0.0 0.0 0.1 0.8 

2006 0.1 0.1 0.0 0.1 0.8 
2009 CREMP Final Report 88 F2345-02-10-F & F2635-06-09-F

2007 0.2 0.0 0.0 0.1 0.7 

2008 0.0 0.0 0.0 0.1 0.7 

Sand Key Deep 1996 1.2 0.3 0.3 0.5 0.0 

24.4517 oN 1997 1.2 0.0 0.2 0.4 0.1 

81.8798 oW 1998 1.2 0.0 0.1 0.4 0.1 

1999 0.8 0.0 0.2 0.5 0.0 

2000 1.0 0.0 0.3 0.5 0.0 

2001 0.8 0.0 0.2 0.5 0.0 

2002 0.9 0.0 0.2 0.4 0.0 

2003 0.8 0.0 0.1 0.4 0.0 

2004 0.6 0.0 0.1 0.3 0.0 

2005 0.5 0.0 0.0 0.5 0.0 

2006 0.0 0.0 0.1 0.3 0.1 

2007 0.2 0.0 0.1 0.6 0.0 

2008 0.2 0.0 0.1 0.7 0.0 

Sand Key Shallow 1996 2.3 0.5 0.1 0.6 0.7 

24.4520 oN 1997 1.8 0.5 0.0 0.4 0.6 

81.8775 oW 1998 3.3 0.2 0.1 0.5 0.4 

1999 2.6 0.3 0.0 0.3 0.6 

2000 2.6 0.5 0.0 0.9 0.7 

2001 2.7 0.3 0.0 0.4 0.7 

2002 2.4 0.4 0.0 0.7 0.8 

2003 1.6 0.5 0.0 0.6 0.7 

2004 2.4 0.3 0.0 0.3 0.8 

2005 1.5 0.5 0.0 0.8 0.7 

2006 1.7 0.2 0.0 0.4 0.7 

2007 1.8 0.6 0.0 0.3 0.4 

2008 2.9 0.2 0.0 0.3 0.6 

Smith Shoal 1996 8.1 1.3 0.0 0.0 2.0 

24.7197 oN 1997 9.4 1.0 0.0 0.1 2.4 

81.9195 oW 1998 7.6 1.6 0.0 0.1 3.4 

1999 12.2 1.2 0.0 0.0 1.2 

2000 9.9 1.3 0.0 0.5 4.2 

2001 8.4 1.6 0.0 0.3 3.2 

2002 0.0 1.3 0.0 0.4 0.4 

2003 0.0 1.0 0.0 0.3 0.6 

2004 0.0 1.1 0.0 0.6 0.4 

2005 0.0 1.1 0.0 0.4 1.0 

2006 0.0 1.4 0.0 0.3 0.7 

2007 0.1 1.2 0.0 0.3 0.3 

2008 0.0 1.6 0.0 0.0 0.3 

Sombrero Deep 1996 0.6 0.5 0.5 1.2 0.0 

24.6231 oN 1997 0.9 0.4 1.0 1.0 0.1 

81.1105 oW 1998 0.5 0.0 0.6 1.0 0.2 

1999 0.5 0.2 0.0 1.3 0.2 

2000 0.3 0.2 0.0 1.1 0.0 

2001 0.6 0.2 0.0 1.1 0.1 

2002 0.3 0.2 0.1 1.0 0.0 

2003 0.4 0.3 0.0 1.4 0.1 
2009 CREMP Final Report 89 F2345-02-10-F & F2635-06-09-F

2004 0.4 0.4 0.2 1.3 0.1 

2005 0.3 0.1 0.1 0.8 0.2 

2006 0.1 0.2 0.1 0.7 0.1 

2007 0.1 0.2 0.0 0.8 0.0 

2008 0.1 0.1 0.0 0.8 0.1 

Sombrero Shallow 1996 1.8 0.5 0.4 0.9 0.7 

24.6259 oN 1997 1.3 0.4 0.2 0.6 0.9 

81.1101 oW 1998 1.3 0.4 0.6 0.6 0.8 

1999 0.9 0.3 0.1 0.8 0.5 

2000 0.8 0.3 0.0 0.4 0.6 

2001 0.7 0.3 0.0 0.6 0.6 

2002 0.9 0.3 0.0 1.0 0.5 

2003 0.8 0.2 0.1 0.6 0.5 

2004 0.3 0.3 0.1 0.7 0.5 

2005 0.5 0.2 0.1 0.7 0.5 

2006 0.3 0.2 0.1 0.8 0.3 

2007 0.2 0.2 0.0 0.7 0.3 

2008 0.2 0.2 0.0 0.9 0.4 

Tennessee Deep 1996 2.4 0.9 0.1 0.8 0.3 

24.7527 oN 1997 2.6 0.2 0.1 0.6 0.4 

80.7578 oW 1998 2.6 0.4 0.0 0.4 0.5 

1999 2.9 0.2 0.0 0.6 0.3 

2000 2.9 0.5 0.0 0.7 0.4 

2001 2.1 0.4 0.0 0.7 0.5 

2002 2.1 0.2 0.0 0.7 0.5 

2003 2.5 0.2 0.0 0.6 0.3 

2004 1.6 0.2 0.1 0.4 0.2 

2005 1.9 0.4 0.0 0.7 0.4 

2006 1.9 0.3 0.0 1.2 0.3 

2007 2.0 0.0 0.1 0.5 0.3 

2008 1.5 0.1 0.0 0.8 0.3 

Tennessee Shallow 1996 0.1 0.2 0.1 0.1 0.2 

24.7450 oN 1997 0.1 0.2 0.0 0.1 0.4 

80.7812 oW 1998 0.2 0.3 0.0 0.0 0.2 

1999 0.0 0.2 0.1 0.1 0.1 

2000 0.1 0.2 0.1 0.1 0.2 

2001 0.1 0.3 0.1 0.1 0.1 

2002 0.1 0.2 0.1 0.2 0.1 

2003 0.0 0.1 0.1 0.1 0.0 

2004 0.1 0.2 0.1 0.2 0.0 

2005 0.0 0.1 0.0 0.2 0.1 

2006 0.1 0.0 0.0 0.2 0.1 

2007 0.0 0.1 0.1 0.1 0.1 

2008 0.1 0.1 0.0 0.1 0.1 

Turtle 1996 0.9 0.5 0.0 1.1 1.3 

25.2947 oN 1997 1.0 0.8 0.1 1.1 0.8 

80.2191 oW 1998 0.9 0.5 0.0 1.4 1.0 

1999 1.0 0.5 0.1 0.9 0.7 

2000 1.0 0.3 0.4 1.1 1.0 
2009 CREMP Final Report 90 F2345-02-10-F & F2635-06-09-F

2001 1.6 0.4 0.0 1.3 1.2 

2002 1.3 0.6 0.1 1.2 1.2 

2003 1.2 0.2 0.3 0.6 1.8 

2004 1.1 0.1 0.1 1.3 1.2 

2005 0.9 0.5 0.4 1.0 1.2 

2006 1.0 0.1 0.0 1.6 1.0 

2007 0.7 0.3 0.0 1.2 1.2 

2008 0.9 0.5 0.1 0.9 1.3 

West Turtle Shoal 1996 0.7 4.4 1.8 3.1 0.2 

24.6993 oN 1997 2.3 3.9 1.6 3.4 0.8 

80.9669 oW 1998 0.8 4.2 2.7 3.2 0.4 

1999 0.8 5.3 1.6 2.3 0.6 

2000 0.4 4.3 1.7 3.9 0.9 

2001 0.9 4.2 2.2 3.9 0.9 

2002 0.8 4.9 2.7 3.8 1.1 

2003 0.5 4.4 2.7 3.1 0.6 

2004 0.6 3.6 2.1 3.1 0.6 

2005 1.0 4.3 2.8 3.0 0.7 

2006 0.6 3.7 2.3 3.1 0.8 

2007 0.4 4.0 2.7 3.3 0.8 

2008 0.5 3.5 2.1 3.3 0.6 

West Washer 

Women 1996 3.4 10.6 6.8 7.9 0.6 

24.5475 oN 1997 5.9 5.0 2.7 6.2 0.4 

81.5866 oW 1998 3.2 8.2 6.0 7.1 0.8 

1999 4.0 6.4 4.7 6.8 0.5 

2000 3.9 7.0 5.1 6.6 1.0 

2001 4.1 7.2 3.5 6.1 0.6 

2002 4.4 8.7 6.1 7.4 1.0 

2003 2.9 6.8 5.2 4.9 1.1 

2004 2.8 6.6 5.7 6.7 0.9 

2005 3.4 7.1 5.0 6.9 0.6 

2006 4.0 7.9 4.4 7.5 0.9 

2007 4.5 7.4 5.1 6.7 0.8 

2008 3.0 7.4 3.0 5.3 0.6 

Western Head 1996 3.1 14.1 2.4 3.3 0.2 

24.4977 oN 1997 1.8 12.5 3.5 2.9 0.2 

81.8056 oW 1998 2.7 14.3 2.6 3.6 0.3 

1999 1.7 12.2 2.3 2.4 0.0 

2000 2.2 14.8 1.8 2.7 0.2 

2001 1.3 14.7 2.0 2.9 0.1 

2002 1.3 13.8 2.2 3.1 0.7 

2003 1.8 14.4 2.4 2.9 0.2 

2004 0.8 13.0 2.8 2.7 0.3 

2005 2.2 14.2 2.9 2.6 0.2 

2006 1.9 12.8 3.2 1.8 0.3 

2007 0.9 12.8 3.5 2.5 0.3 

2008 1.3 12.7 2.0 3.0 0.2 

Western Sambo 

Deep 1996 2.4 0.0 0.0 0.5 0.4 
2009 CREMP Final Report 91 F2345-02-10-F & F2635-06-09-F

24.4785 oN 1997 2.4 0.0 0.1 0.5 0.3 

81.7161 oW 1998 2.1 0.0 0.1 0.5 0.3 

1999 2.1 0.1 0.1 0.5 0.2 

2000 2.1 0.0 0.1 0.4 0.1 

2001 1.9 0.0 0.1 0.5 0.3 

2002 1.7 0.2 0.0 0.3 0.3 

2003 1.6 0.0 0.2 0.4 0.3 

2004 1.0 0.0 0.1 0.3 0.2 

2005 1.5 0.1 0.1 0.3 0.1 

2006 0.7 0.0 0.0 0.1 0.0 

2007 0.7 0.0 0.0 0.3 0.1 

2008 0.9 0.0 0.0 0.4 0.1 

Western Sambo 

Shallow 

1996 1.7 0.1 0.3 0.4 1.8 

1997 1.7 0.0 0.1 0.2 2.1 

24.4797 oN 1998 1.5 0.0 0.4 0.5 2.1 

81.7170 oW 1999 1.0 0.1 0.0 0.4 2.2 

2000 1.2 0.0 0.1 0.3 2.5 

2001 1.0 0.1 0.2 0.2 2.1 

2002 1.1 0.0 0.1 0.2 2.6 

2003 1.1 0.1 0.1 0.3 2.7 

2004 1.6 0.0 0.1 0.3 1.9 

2005 1.2 0.0 0.1 0.2 2.0 

2006 0.8 0.0 0.0 0.3 1.6 

2007 0.6 0.1 0.0 0.3 1.5 

2008 0.5 0.0 0.0 0.4 1.3 

White Shoal 1999 0.2 0.1 0.3 0.4 0.1 

24.6416 oN 2000 0.3 0.1 0.2 0.3 0.2 

82.8964 oW 2001 0.4 0.1 0.5 0.4 0.0 

2002 0.3 0.1 0.3 0.1 0.0 

2003 0.4 0.1 0.2 0.2 0.1 

2004 0.2 0.1 0.2 0.3 0.0 

2005 0.3 0.1 0.1 0.2 0.1 

2006 0.2 0.0 0.1 0.2 0.0 

2007 0.1 0.1 0.1 0.2 0.1 

2008 0.2 0.0 0.3 0.2 0.1 
2009 CREMP Final Report 92 F2345-02-10-F & F2635-06-09-F

Appendix 5. Long-term trends in stony coral cover 1996 to 2008. 
2009 CREMP Final Report 93 F2345-02-10-F & F2635-06-09-F

Appendix 6. Long-term trends in stony coral cover from 1999 to 2008. 
2009 CREMP Final Report 94 F2345-02-10-F & F2635-06-09-F

Appendix 7. Long-term trends in macroalgal cover from 1996 to 2008. 
2009 CREMP Final Report 95 F2345-02-10-F & F2635-06-09-F

Appendix 8. Long-term trends in macroalgal cover from 1999 to 2008. 
2009 CREMP Final Report 96 F2345-02-10-F & F2635-06-09-F

Appendix 9. Long-term trends in octocoral cover from 1996 to 2008. 
2009 CREMP Final Report 97 F2345-02-10-F & F2635-06-09-F

Appendix 10. Long-term trends in octocoral cover from 1999 to 2008. 
2009 CREMP Final Report 98 F2345-02-10-F & F2635-06-09-F

Appendix 11. Long-term trends in sponge cover from 1996 to 2008. 
2009 CREMP Final Report 99 F2345-02-10-F & F2635-06-09-F

Appendix 12. Long-term trends in sponge cover from 1999 to 2008. 
2009 CREMP Final Report 100 F2345-02-10-F & F2635-06-09-F

Appendix 13. Long-term trends in Montastraea annularis complex cover from 1996 to 2008. 
2009 CREMP Final Report 101 F2345-02-10-F & F2635-06-09-F

Appendix 14. Long-term trends in Montastraea annularis complex cover from 1999 to 2008. 
2009 CREMP Final Report 102 F2345-02-10-F & F2635-06-09-F

Appendix 15. Long-term trends in Montastraea cavernosa cover from 1996 to 2008. 
2009 CREMP Final Report 103 F2345-02-10-F & F2635-06-09-F

Appendix 16. Long-term trends in Montastraea cavernosa cover from 1999 to 2008. 
2009 CREMP Final Report 104 F2345-02-10-F & F2635-06-09-F

Appendix 17. Long-term trends in Colpophyllia natans cover from 1996 to 2008. 
2009 CREMP Final Report 105 F2345-02-10-F & F2635-06-09-F

Appendix 18. Long-term trends in Colpophyllia natans cover from 1999 to 2008. 
2009 CREMP Final Report 106 F2345-02-10-F & F2635-06-09-F

Appendix 19. Long-term trends in Porites astreoides cover from 1996 to 2008. 
2009 CREMP Final Report 107 F2345-02-10-F & F2635-06-09-F

Appendix 20. Long-term trends in Porites astreoides cover from 1999 to 2008. 
2009 CREMP Final Report 108 F2345-02-10-F & F2635-06-09-F

Appendix 21. Long-term trends in Siderastrea siderea cover from 1996 to 2008. 
2009 CREMP Final Report 109 F2345-02-10-F & F2635-06-09-F

Appendix 22. Long-term trends in Siderastrea siderea cover from 1999 to 2008. 
2009 CREMP Final Report 110 F2345-02-10-F & F2635-06-09-F

Appendix 23. Long-term trends in Cliona delitrix cover from 2001 to 2008.