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Regional spatio-temporal trends in Caribbean coral reef benthic communities


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Coral cover has declined on reefs worldwide with particularly acute losses in the Caribbean. Despite our awareness of the broad-scale patterns and timing of Caribbean coral loss, there is little published information on: (1) finer-scale, subregional patterns over the last 35 yr, (2) regional-scale trends since 2001, and (3) macroalgal cover changes. We analyzed the spatiotemporal trends of benthic coral reef communities in the Caribbean using quantitative data from 3777 coral cover surveys of 1962 reefs from 1971 to 2006 and 2247 macroalgal cover surveys of 875 reefs from 1977 to 2006. A subset of 376 reefs was surveyed more than once (monitored). The largest 1 yr decline in coral cover occurred from 1980 to 1981, corresponding with the beginning of the Caribbean-wide Acropora spp. white band disease outbreak. Our results suggest that, regionally, coral cover has been relatively stable since this event (i.e. from 1982 to 2006). The largest increase in macroalgal cover was in 1986, 3 yr after the regional die-off of the urchin grazer Diadema antillarum began. Subsequently, macroalgal cover declined in 1987 and has been stable since then. Regional mean (±1 SE) macroalgal cover from 2001 to 2005 was 15.3 ± 0.4% (n = 1821 surveys). Caribbeanwide mean (±1 SE) coral cover was 16.0 ± 0.4% (n = 1547) for this same time period. Both macroalgal and coral cover varied significantly among subregions from 2001 to 2005, with the lowest coral cover in the Florida Keys and the highest coral cover in the Gulf of Mexico. Spatio-temporal patterns from the subset of monitored reefs are concordant with the conclusions drawn from the entire database. Our results suggest that coral and macroalgal cover on Caribbean reef benthic communities has changed relatively little since the mid-1980s.
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Mar Ecol Prog Ser
Vol. 402: 115122, 2010
doi: 10.3354/meps08438
Published March 8
Based on a recently published global map of the
magnitude and geographic extent of 17 anthropogenic
threats, Halpern et al. (2008) argued that coral reefs
are one of the world’s most heavily impacted marine
ecosystems. This finding is consistent with a vast body
of literature documenting the global degradation of
tropical reefs over the last several decades (Wilkinson
1992, Grigg 1994, Gardner et al. 2003, Bruno & Selig
2007, Edmunds & Elahi 2007). Coral loss can lead to
reductions in fish abundance and diversity (Jones et al.
2004, Pratchett et al. 2008) and declines in topographi-
cal complexity (Alvarez-Filip et al. 2009). Coral loss is
also associated with compensatory increases in the
abundance of several other taxa, including sponges,
© Inter-Research 2010 ·*Email:
Regional spatio-temporal trends in Caribbean
coral reef benthic communities
Virginia G. W. Schutte
1, 3,
, Elizabeth R. Selig
1, 2, 4
, John F. Bruno
Department of Marine Sciences, The University of North Carolina at Chapel Hill, 340 Chapman Hall CB# 3300, Chapel Hill,
North Carolina 27599-3300, USA
Curriculum for the Environment and Ecology, 207 Coastes Building CB# 3275, The University of North Carolina at Chapel
Hill, Chapel Hill, North Carolina 27599-3275, USA
Present address: Odum School of Ecology, The University of Georgia, 140 E. Green St., Athens, Georgia 30602-2202, USA
Present address: Center for Applied Biodiversity Science, Conservation International, 2011 Crystal Drive, Suite 500,
Arlington, Virginia 22202, USA
ABSTRACT: Coral cover has declined on reefs worldwide with particularly acute losses in the
Caribbean. Despite our awareness of the broad-scale patterns and timing of Caribbean coral loss,
there is little published information on: (1) finer-scale, subregional patterns over the last 35 yr,
(2) regional-scale trends since 2001, and (3) macroalgal cover changes. We analyzed the spatio-
temporal trends of benthic coral reef communities in the Caribbean using quantitative data from 3777
coral cover surveys of 1962 reefs from 1971 to 2006 and 2247 macroalgal cover surveys of 875 reefs
from 1977 to 2006. A subset of 376 reefs was surveyed more than once (monitored). The largest 1 yr
decline in coral cover occurred from 1980 to 1981, corresponding with the beginning of the
Caribbean-wide Acropora spp. white band disease outbreak. Our results suggest that, regionally,
coral cover has been relatively stable since this event (i.e. from 1982 to 2006). The largest increase in
macroalgal cover was in 1986, 3 yr after the regional die-off of the urchin grazer Diadema antillarum
began. Subsequently, macroalgal cover declined in 1987 and has been stable since then. Regional
mean (±1 SE) macroalgal cover from 2001 to 2005 was 15.3 ± 0.4% (n = 1821 surveys). Caribbean-
wide mean (±1 SE) coral cover was 16.0 ± 0.4% (n = 1547) for this same time period. Both macroalgal
and coral cover varied significantly among subregions from 2001 to 2005, with the lowest coral cover
in the Florida Keys and the highest coral cover in the Gulf of Mexico. Spatio-temporal patterns from
the subset of monitored reefs are concordant with the conclusions drawn from the entire database.
Our results suggest that coral and macroalgal cover on Caribbean reef benthic communities has
changed relatively little since the mid-1980s.
KEY WORDS: Coral cover · Macroalgae · Coral disease · Coral bleaching
Resale or republication not permitted without written consent of the publisher
Mar Ecol Prog Ser 402: 115122, 2010
gorgonians, and macroalgae (Hughes 1994, Aronson et
al. 2002, Norström et al. 2009). A wide variety of causes
and human activities have been implicated in driving
these changes, including overfishing,
increasing ocean temperatures, coral
disease and predator outbreaks, and
poor land-use practices that lead to sed-
imentation and nutrient pollution
(Endean 1977, Jackson et al. 2001,
Hughes et al. 2003, Fabricius 2005,
Hoegh-Guldberg et al. 2007).
Although the key drivers of coral loss
have been identified, there is no con-
sensus on their relative importance and
how this varies in space and time
(Grigg & Dollar 2005, Precht et al.
2005). Quantifying the spatio-temporal
changes in coral reef benthic communi-
ties at regional and decadal scales can
lead to a broader understanding of the
patterns and causes of reef degradation
and provide information which will
result in more effective management
actions (Côté et al. 2005).
The purpose of the present study was
to quantify the regional-scale trends in
coral and macroalgal cover on Carib-
bean coral reefs over the last 35 yr in
each of 7 subregions (Fig. 1A). Our
study builds on previous analyses of
benthic changes on particular reefs
(Edmunds & Elahi 2007), whole islands
(Hughes 1994), and a meta-analysis of
the entire region (Gardner et al. 2003).
Although the broad-scale patterns and
timing of Caribbean coral loss were dis-
cussed in Gardner et al. (2003), there is
no published information on regional-
scale trends for macroalgal cover or for
coral cover since 2001. Our analysis
was based on data from 3777 surveys of
1962 reefs performed between 1971
and 2006 the largest compilation of
quantitative Caribbean reef surveys to
All analyses were based on quantita-
tive in situ surveys of subtidal coral
reefs in the Caribbean basin that mea-
sured the percentage of the bottom cov-
ered by living scleractinian coral tissue.
We grouped these survey sites into 7
subregions, which were similar to previous delineation
schemes and were based on areas of similar biodiver-
sity and biogeography, major bathymetric changes, or
Fig. 1. (A) Locations of survey sites (purple dots; n = 1667) and delineations of the 7
subregions used in our analyses. The 295 sites for which we could not obtain ex-
act coordinates are not shown. (B) Number of thermal stress anomalies during
2005. Sea surface temperature data were derived from the Advanced Very High
Resolution Radiometer sensor and were processed to a resolution of approxi-
mately 4.6 km at the equator. An anomaly is defined as 1°C more than the typical
maximum climatological week defined for each grid cell (see also Text S3 in the
supplement, pdf)
Schutte et al.: Caribbean reef trends
management regimes (Fig. 1A; Spalding et al. 2007).
The coral and macroalgal cover survey database was
compiled by searching the scientific literature using a
variety of search engines (e.g. ISI Web of Science and
Google Scholar) using terms including ‘Caribbean’
and ‘coral cover’. We also examined every issue of
Coral Reefs, Atoll Research Bulletin, the Proceedings
of the Colloquium on Global Aspects of Coral Reefs
(Ginsburg 1993), and the Proceedings of the Interna-
tional Coral Reef Symposiums. When cover data were
presented in graphical form, we used ImageJ (Rasband
2006) to extract the raw data (the percent cover). In
addition, we obtained survey data directly from sev-
eral monitoring programs (Text S1, Fig. S1 & Table S1
in the supplement,
m402p115_app.pdf), including the Florida Coral Reef
Evaluation and Monitoring Project (CREMP; data com-
prise 35.6% of surveys and 13.1% of sites), the Atlantic
and Gulf Rapid Reef Assessment (AGRRA) Program
(17.1% of surveys and 30.3% of sites), and Reef Check
(16.5% of surveys and 22.8% of sites).
The database included 3777 coral cover surveys from
1962 sites surveyed between 1971 and 2006 (Text S1 &
Table S2 in the supplement,
suppl/m402p115_app.pdf). We pooled observations
across depths (1 to 30 m depth, 8.2 ± 0.08 m, mean ±
1 SE; Fig. S2 in the supplement,
articles/suppl/m402p115_app.pdf) and reef zones for
our analyses because there was little variation in sur-
vey depth among subregions or years and there was
no relationship between depth and cover (Text S2 &
Fig. S3 in the supplement,
suppl/m402p115_app.pdf). The cover of Millepora spp.
(hydrozoans) was included in the total live coral cover
value for 181 surveys (4.1% of all surveys). Macroalgal
cover, including both fleshy and calcifying macroalgae
(sensu Steneck 1988), was measured in 2247 surveys
on 875 reefs (58.4 and 44.6% of the total, respectively).
We did not include cover from turf or coralline encrust-
ing algae. In 64.6% of surveys, benthic cover was mea-
sured in situ using related techniques such as the
AGRRA Program methodology (Lang 2003) and point
intercept cover estimates (Hodgson et al. 2006, Lam et
al. 2006). For the other 35.4%, the substrate was re-
corded with video or still photographs (Edmunds &
Bruno 1996), which were analyzed in the laboratory
using image analysis software or the point count tech-
nique (Aronson et al. 2002, Idjadi & Edmunds 2006,
Lam et al. 2006, Rogers & Miller 2006).
Annual sample sizes for the whole region (~200 to
400 surveys; Fig. 2) and for many subregions (Table S1)
over the last decade are quite large, and most survey
sites were more or less randomly or haphazardly se-
lected (particularly with respect to the hypotheses being
tested in this analysis). In other words, we used survey
data regardless of the original research purpose. There-
fore, our analyses are likely representative of the broad-
scale trends on benthic reef communities, particularly
since the mid-1990s. Yet meta-analyses of field surveys
such as this have a number of limitations, and several
caveats need to be considered when interpreting the re-
sults (Bruno & Selig 2007). First, year-to-year variations
in which sites are sampled can cause apparent short-
term temporal trends that are likely sampling artifacts
and not real phenomena. As in any long-term analysis,
short-term and temporary fluctuations are expected,
may not be real, and should not be over-interpreted.
Second, for a variety of reasons, some sites and sub-
regions have been sampled far more intensively than
others. We used a repeated measures linear regression
analysis to test the null hypothesis that there was
no relationship between percent coral or macroalgal
cover and time (using Stata Version 9.1, STATA). This
test was based on the individual trajectories of the 7
subregions rather than pooled data for each year. Per-
forming this analysis on yearly subregional averages
equalizes the influence of each subregion. However, it
does not remove the influence of highly sampled sites
on subregional values. Additionally, some important
locations were sampled comparatively sparsely rela-
tive to their size, e.g. Belize. Therefore, the regional
and subregional trends may not be representative of
all parts of the Caribbean. Our general trends also
might differ from the true, but unknowable, population
values. The population-level trends we quantified are
not meant to represent the trajectory of any individual
reef. Additionally, trends from the early years of the
database should be interpreted with great caution
given their relatively small annual sample sizes. We
still know very little about the natural baselines of
Caribbean reefs and even less about their state even in
the near-past, e.g. the 1960s. Thus, putting our results
into a broader historical context is very difficult. It is
possible that Caribbean reefs began to degrade long
before we began surveying them (Pandolfi et al. 2003),
in which case the actual degree of degradation is much
greater than we can currently quantify.
The database largely consisted of reefs surveyed only
once, but included 376 reefs surveyed in 2 or more
years. We analyzed this subset of monitored reefs sepa-
rately from the entire database, restricting the analysis
to surveys performed between 1996 and 2006 (n = 331
reefs; Table S3 in the supplement,
journals/articles/m402p115_app.pdf), because few
reefs were monitored before this time period. Re-
peated-measures regression analysis was used to de-
termine whether coral or macroalgal cover changed
significantly from 1996 to 2006. We performed separate
analyses for each subregion and also an analysis of all
the monitoring data combined.
Mar Ecol Prog Ser 402: 115122, 2010
Regional temporal trends
Numerous studies indicate that Caribbean reef ben-
thic communities changed dramatically in the 1980s
(Hughes et al. 1985, Carpenter 1990, Liddell & Ohl-
horst 1992, Hughes 1994, Gardner et al. 2003). Despite
the general perception that Caribbean reefs have con-
tinued to degrade, the broad-scale cover of hard corals
and macroalgae appears to have changed very little
since at least the mid-1980s. Our analysis suggests that
the regional coral cover average has not changed sig-
nificantly since the 1981 Acropora spp. mass mortality
(i.e. from 1982 to 2006, p = 0.99, based on repeated-
measures regression of annual subregional means;
Fig. 2A). This pattern of regional stasis does not contra-
dict studies that have documented recent, i.e. post-
1980s, coral loss on individual reefs (e.g. Aronson et
al. 2002, Edmunds 2007) and in some subregions such
as the Florida Keys (Fig. S4 in the supplement, www.; Porter &
Meier 1992, Maliao et al. 2008). Instead it suggests that
local losses on some reefs have apparently been
roughly balanced by local coral recovery on other
reefs. In other words, the observed pattern of regional
stasis is probably a dynamic equilibrium, masking
greater spatio-temporal variance in benthic commu-
nity structure at finer spatial scales. Additionally,
because we were only able to analyze broad-scale pat-
terns in total coral and macroalgal cover, we could not
evaluate other reef characteristics such as trophic com-
plexity (Paddack et al. 2009), reef rugosity (Alvarez-
Filip et al. 2009), or changes in species composition
(e.g. Aronson et al. 2004), which may indeed have
changed during the study period and are also impor-
tant indicators of coral reef health.
Disturbances before the mid-1980s caused substan-
tial regional losses in coral cover. Caribbean-wide
coral cover averages were highest from 1971 to 1980
(Fig. 2A). Annual regional means during this period
ranged from ~25 to 40% (n = 43 surveys), which is
somewhat lower than generally assumed (Gardner et
al. 2003). Our results suggest that regional absolute
coral cover declined by ~18% between 1972, the year
with the highest yearly coral cover mean (38.3%, n = 1
1970 1975 1980 1985 1990 1995 2000 2005 1970 1975 1980 1985 1990 1995 2000 2005
A) Coral B) Macroalgae
D) FLK coralC) Coral-no FLK
Cover (%)
No. of studies
Fig. 2. Annual cover values (±1 SE, closed circles, left y-axis) and site sample sizes (open circles, right y-axis) for (A) mean coral
cover for all sites in the Caribbean basin (n = 1962; star: 1980, the year in which Hurricane Allen struck and white band disease
outbreaks began); (B) mean macroalgal cover for all sites for which data were available (n = 875; star: 1983, the year in which the
Diadema antillarum die-off began); (C) mean coral cover for all sites in the greater Caribbean except those in the Florida Keys
(FLK; n = 1515); and (D) mean coral cover for all sites in the FLK subregion (n = 447)
Schutte et al.: Caribbean reef trends
survey), and 1982 (20.8 ± 4.2%, n = 7 surveys), when
the current period of regional stasis began (Fig. 2A).
However, the annual coral cover means during this
period are based on a small number of surveys
(Table S2) and it is possible that the historical regional
baseline was higher. The largest 1 yr decline in coral
cover (24.9 ± 3.2% absolute coverage) took place from
1980 to 1981, coincident with and presumably due to
the regional Acropora spp. die-off from the white band
disease epizootic (Aronson & Precht 2001, 2006) and
the passage of Hurricane Allen through Jamaica
(Woodley et al. 1981), where 9 of the 21 surveys con-
ducted in 1980 and 1981 were performed.
Just over half of the coral cover surveys had accom-
panying macroalgal cover values; our database inclu-
ded 2247 macroalgal surveys conducted between 1977
and 2006. Macroalgal cover did not change between
1987 and 2006 (p = 0.13; Fig. 2B), although the small
number of quantitative surveys in the 1980s and early
1990s makes estimates of subregion-specific macroal-
gal trends during this period less reliable (Table S2). It
is possible that some other rarely measured attribute of
macroalgae changed during the period of stasis, e.g.
biomass, height, or composition, although macroalgal
cover is generally a good predictor of macroalgal bio-
mass (Miller et al. 2003).
The regional mean coverage of macroalgae increa-
sed dramatically in 1986 (26.0 ± 13.3%; Fig. 2B) follow-
ing the Caribbean-wide die-off of the sea urchin Dia-
dema antillarum in 1983 and 1984 (Hughes et al. 1985,
Lessios 1988). Mean macroalgal cover before the D.
antillarum die-off (based on 37 surveys performed
from 1977 to 1983) was 8.0 ± 1.6%. In 1986, macroalgal
cover increased to 38.1 ± 13.3% (n = 8 surveys). How-
ever, it declined again to 14.5% in 1987 (± 5.7, n = 13
surveys) and the regional mean remained below 20%
for most years between 1987 and 2006. The drop in
macroalgal cover following the spike in 1986 could be
due to compensatory population increases or behav-
ioral responses by other fish and urchin grazers to the
loss of the once dominant herbivore, D. antillarum
(Aronson et al. 2000, Haley & Solandt 2001). Popula-
tions of D. antillarum have since recovered on some
Caribbean reefs (e.g. Carpenter & Edmunds 2006,
Myhre & Acevedo-Gutiérrez 2007), which could ex-
plain the general absence of macroalgal cover changes
since 1987. Conversely, the 1986 spike in macroalgal
cover could be an artifact of non-random site selection;
many macroalgal cover studies conducted in the mid-
1980s focused on reefs that experienced significant
losses in coral cover or were designed to document the
indirect effects of the D. antillarum die-off (e.g.
Hughes 1994).
The combined multi-decade, regional patterns of
changes in coral and macroalgal cover indicate that,
although the region has experienced substantial coral
losses, there has not been a concomitant increase in
macroalgal cover (Fig. 3). Almost half (48.9%) of the
2247 macroalgal surveys documented a higher percen-
tage of macroalgal cover than coral cover, but macroal-
gal cover has rarely exceeded 50% (just 5.2% of sur-
veys). The observed regional increase in macroalgal
cover in 1986 occurred 5 yr after the collapse of coral
cover in 1981, supporting the argument that coral loss
in the 1980s was not caused by an increase in macro-
algal cover (Aronson & Precht 2006, Bruno et al. 2009).
The relationship between coral and macroalgal
cover on reefs over time can be divided into 3 temporal
categories (Fig. 3): (1) the late 1970s, the baseline for
the present study; (2) the 1980s and early 1990s, after
the Acropora spp. and Diadema antillarum disease
outbreaks; and (3) from 1993 to 2006, a period of post-
disease stability. We pooled the annual means within
each temporal period in order to compare the 3 peri-
ods, and the greatest coral cover loss occurred from
Time Period 1 to Time Period 2. There was not a long-
term increase in macroalgal cover proportional to the
coral cover loss in the 1980s.
Recent spatial patterns
Recent (2001 to 2005) macroalgal cover varied signi-
ficantly among subregions (ANOVA, p < 0.0001; Fig. 4)
and ranged from 6.2 ± 4.0% (n = 16 surveys) in the Gulf
of Mexico to 22.8 ± 1.6% (n = 100 surveys) in the south-
Fig. 3. Average annual coral and macroalgal cover values
from 1977 to 2006. Horizontal and vertical lines represent
1 SE. The 3 points from the 1981 to 1992 group that are
clumped with the 1993 to 2006 values are from 1981, 1987,
and 1988
Mar Ecol Prog Ser 402: 115122, 2010
western Caribbean. The region-wide mean cover of
macroalgae from 2001 to 2005 was 15.3 ± 0.4% (n =
1821 surveys), while mean coral cover was 16.0 ± 0.4%
(n = 1547 surveys). Recent coral cover also varied sig-
nificantly among subregions (ANOVA, p < 0.0001;
Fig. 4; see also Fig. S4). Subregional coral cover dif-
ferences are concordant with a previous analysis of the
Caribbean (Gardner et al. 2003). Assuming the histori-
cal coral cover baseline was similar across the region,
this pattern of current spatial variability could be inter-
preted as evidence of variable rates of coral loss.
Recent coral cover was highest in the Gulf of Mexico
(58.1 ± 3.5% from 2001 to 2005, n = 10 surveys). The
high cover in this subregion was likely due to the
absence of Acropora spp. host populations in the
Flower Garden Banks (until recently; Precht & Aron-
son 2004), where most of the surveys from the Gulf of
Mexico were conducted, which precluded white band
disease from reducing Acropora spp. coral cover there
(Aronson et al. 2005). The Flower Garden Banks reefs
are also atypical sites because survey depths there
were from 20 to 30 m (e.g. Dokken et al. 2003), more
than twice the mean depth of most surveys in the
Recent coral cover was lowest in the Florida Keys
(FLK; 8.6 ± 0.4% from 2001 to 2005, n = 747 surveys).
This subregion was extensively sampled (50.6% of all
surveys performed from 1996 to 2006 were conducted
in the FLK; Table S2) and this overrepresentation
could have unduly influenced the Caribbean-wide
analyses. Therefore, we also analyzed the trends from
1996 to 2005 in regional coral and macroalgal cover
without the FLK data. This resulted in annual regional
coral cover means as much as 13.1% higher than when
the FLK data were included (Fig. 2C,D), but did not
noticeably influence macroalgal cover values. Mean
regional Caribbean coral cover without the FLK data
from 1996 to 2005 was 21.8 ± 0.3%.
The southeastern Caribbean experienced a severe
warming event in late 2005, with levels of thermal
stress exceeding standard bleaching thresholds
(Fig. 1B; Donner et al. 2007, Wilkinson & Souter 2008).
This led to widespread coral bleaching, subsequent
coral disease outbreaks, and moderate to severe local
coral mortality and loss in some locations, including
the US Virgin Islands and the Lesser Antilles (Miller
et al. 2006, Wilkinson & Souter 2008). There was a mi-
nor regional reduction in coral cover in 2006 (Fig. 2C),
which could have been caused in part by coral mortal-
ity in several subregions, particularly those that expe-
rienced the most severe temperature anomalies (Fig. 1;
see also Fig. S4), e.g. the Northern Caribbean (–6.0%
absolute cover from 2005 to 2006) and the Lesser An-
tilles (–3.8% absolute cover). However, we did not
have enough post-bleaching event data to reliably es-
timate subregional declines since 2005 (Table S2).
Monitored sites
Our analysis was primarily based on reefs that were
randomly selected and surveyed only once. The results
of such randomized population sampling can be gener-
alized to a greater degree than those from longitudinal
monitoring studies. However, randomized sampling
has some drawbacks, including sensitivity to sample
composition. Minor, short-term fluctuations in coral
and macroalgal cover (Fig. 2; see also Fig S4) could be
due to the subpopulation of reefs that were surveyed in
a given year, rather than to real year-to-year changes
in community state. Although non-random initial site
selection can cause similar biases, monitoring studies
are generally less sensitive to sample composi-
tion and have several advantages, including
greater power to detect small changes in com-
munity state. Unfortunately, there are still rela-
tively few quantitative reef monitoring pro-
grams and most are focused on well-studied
reefs within Marine Protected Areas.
Spatio-temporal patterns from the subset of
376 monitored reefs are very similar to those
from the entire database (Fig. S5 in the sup-
p115_app.pdf) and are concordant with our
general conclusions. The regional trends from
the sites monitored from 1996 to 2006 indicate
that, since 1996, there has been very little tem-
poral variation in both coral (n = 331) and
macroalgal cover (n = 215) and the trends rein-
force the finding that reefs in the Caribbean
entered a period of general regional stasis in
Fig. 4. Recent (2001 to 2005) macroalgal and coral cover in the 7
Caribbean subregions (Fig. 1). Values are means (+1 SE) and sample
sizes (no. of sites) are shown to the right of the bars. Mesoam Reef:
Mesoamerican Reef; N Caribbean: Northern Caribbean; SW Carib-
bean: Southwestern Caribbean
Schutte et al.: Caribbean reef trends
the mid-1990s, at least in terms of coral and macroalgal
cover. Linear repeated-measures regression analyses
indicate that there was no significant change in coral
cover (p = 0.32, n = 331 reefs) or macroalgal cover
(p = 0.109, n = 215 reefs) from 1996 to 2006 across the
region (Table S4 in the supplement,
articles/suppl/m402p115_app.pdf). We also analyzed
monitoring data from each subregion independently
from 1996 to 2006; in most cases, there was no sig-
nificant change in cover (Tables S5 & S6 in the supple-
pdf. There was a statistically, although perhaps not
ecologically, significant decrease in both coral and
macroalgal cover in the intensively monitored Florida
Keys. Coral cover slightly increased in the Northern
Caribbean and decreased in the Lesser Antilles and
the Southwestern Caribbean. Otherwise, our regres-
sion analyses of the monitoring data suggest that there
have been few other signs of temporal change within
subregions since 1996.
The first meta-analysis of Caribbean hard coral cover
documented a reduction from ~50% in 1977 to ~10%
in 2001 (Gardner et al. 2003). Our study built on that
work by adding surveys from an additional 1699 sites
and macroalgal data and by expanding the analysis to
2006. The timing of coral loss we documented is consis-
tent with the hypothesis that the acroporid white band
epizootic was the primary cause of hard coral cover
loss in the Caribbean (Aronson & Precht 2006). Like-
wise, the regional macroalgal bloom that occurred sev-
eral years after the disease-induced Diadema antil-
larum die-off supports the argument that a reduction
in herbivory, rather than increased nutrient availabil-
ity, caused the observed increases in macroalgal cover
(Hughes et al. 1999). Therefore, disease, whether nat-
ural or exacerbated by human activities, appears to
have been the primary driver of regional-scale
changes in Caribbean reef benthic communities over
the last 35 yr. Our results indicate that, since these 2
major disturbances, regional coral and macroalgal
benthic coverage has been relatively stable. However,
the future of Caribbean reefs is uncertain. There are
many factors besides disease that could cause fur-
ther changes in reef benthic community structure,
including climate change, ocean acidification, and
direct anthropogenic stressors like overfishing and
nutrient pollution (Hughes et al. 2003). Although our
results could be interpreted as relatively good news,
the observed regional pattern could also be a tempo-
rary plateau preceding a potential collapse in coral
Acknowledgements. We thank K. France, R. Katz, L. Ladwig,
S. C. Lee, M. I. O’Connor, C. Shields, G. Smelick, I. Vu, and
A. M. Melendy for assistance with this project, and J. C. Lang,
G. Hodgson, W. F. Precht, and W. K. Fitt for very helpful com-
ments on this paper. We are especially grateful to everyone
who shared data with us, including Reef Check, AGRRA,
Florida CREMP, the scientists and data managers who com-
municated with us personally, and all the volunteers and
researchers who collected the data. This project was funded
in part by the National Science Foundation, an Environmen-
tal Protection Agency STAR fellowship to E.R.S., and the Uni-
versity of North Carolina at Chapel Hill.
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Editorial responsibility: Tim McClanahan,
Mombasa, Kenya
Submitted: May 18, 2009; Accepted: November 23, 2009
Proofs received from author(s): February 18, 2010
... Introduction Coral reefs worldwide have experienced remarkable changes over the past 40-50 years, particularly the widespread declines of reef-building corals and large, predatory fishes [1][2][3][4][5][6][7]. These changes have caused a reduction in or effective loss of essential ecological functions, including the provisioning of habitat for fisheries production and the maintenance of reef structure for shoreline protection [8,9]. ...
... Corals declined in the first few years of our study (1997)(1998)(1999), then remained relatively constant at both protected (~11.8 ± 1.5%, fishing prohibited) and unprotected sites (~11.9 ± 0.9%, fishing allowed; Fig 2). In fact, this general stasis in coral cover has been apparent across the region, especially on low-coral-cover reefs, for the last several decades [2,75]. Given the frequent disturbances on the BRR during this period (Table 1), stability in coral cover is technically evidence of "resilience" [76]. ...
... Twenty years of change in benthic communities across the Belizean Barrier Reef bleaching events [2,3,[78][79][80]. Across seven subregions in the Caribbean, Schutte et al. [2] found significant declines in hard-coral cover and increases in macroalgal cover from 1970-2005. ...
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Disease, storms, ocean warming, and pollution have caused the mass mortality of reef-building corals across the Caribbean over the last four decades. Subsequently, stony corals have been replaced by macroalgae, bacterial mats, and invertebrates including soft corals and sponges, causing changes to the functioning of Caribbean reef ecosystems. Here we describe changes in the absolute cover of benthic reef taxa, including corals, gorgonians, sponges, and algae, at 15 fore-reef sites (12–15m depth) across the Belizean Barrier Reef (BBR) from 1997 to 2016. We also tested whether Marine Protected Areas (MPAs), in which fishing was prohibited but likely still occurred, mitigated these changes. Additionally, we determined whether ocean-temperature anomalies (measured via satellite) or local human impacts (estimated using the Human Influence Index, HII) were related to changes in benthic community structure. We observed a reduction in the cover of reef-building corals, including the long-lived, massive corals Orbicella spp. (from 13 to 2%), and an increase in fleshy and corticated macroalgae across most sites. These and other changes to the benthic communities were unaffected by local protection. The covers of hard-coral taxa, including Acropora spp., Montastraea cavernosa , Orbicella spp., and Porites spp., were negatively related to the frequency of ocean-temperature anomalies. Only gorgonian cover was related, negatively, to our metric of the magnitude of local impacts (HII). Our results suggest that benthic communities along the BBR have experienced disturbances that are beyond the capacity of the current management structure to mitigate. We recommend that managers devote greater resources and capacity to enforcing and expanding existing marine protected areas and to mitigating local stressors, and most importantly, that government, industry, and the public act immediately to reduce global carbon emissions.
... Los arrecifes del Caribe se vienen deteriorando reconocidamente desde hace más de tres décadas (Gardner et al. 2003, Schutte et al. 2010. Estas afectaciones se han manifestado de diferentes maneras en toda la región (Jackson et al., 2014). ...
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Los ecosistemas marinos y costeros del Caribe garantizan la mayor parte de las actividades socioeconómicas que sostienen a más de 43 millones de personas. Esta región es altamente vulnerable a los impactos potenciales del cambio climático, principalmente los producidos por el incremento del nivel del mar y los efectos de eventos meteorológicos extremos como huracanes, fuertes lluvias o sequías intensas, y se encuentra también afectada por los cambios de hábitat, invasiones biológicas, sobreexplotación de los recursos marinos y costeros y la contaminación, que son presiones que ejerce el desarrollo descontrolado en las zonas marinas y costeras, cuyos efectos continuarán amplificándose por los impactos del cambio climático en la región. Para evitar la magnificación de los impactos asociados al cambio climático sobre los ecosistemas marinos y costeros y por tanto sobre el beneficio que recibe la sociedad caribeña de los servicios ecosistémicos que estos proveen, resulta imprescindible disminuir las presiones que el hombre ejerce sobre los ecosistemas fomentando para ello su resiliencia. La integración en planes de adaptación de las estrategias de conservación, rehabilitación ecológica y gestión sostenible a nivel local, nacional y regional deberá promoverse destacando el papel de los ecosistemas para la adaptación y mitigación al cambio climático, encaminando en una sola vía los vínculos entre diversidad biológica, cambio climático, reducción de desastres y desarrollo sostenible, lo que ha sido ampliamente reconocido como una necesidad a nivel mundial. La Adaptación basada en Ecosistemas (AbE) es una propuesta para construir resiliencia y reducir la vulnerabilidad de las comunidades al cambio climático, integrando justamente el uso sostenible de la biodiversidad y de los servicios ecosistémicos en una estrategia para ayudar a las personas a adaptarse al cambio climático considerando como puntos de partida tanto el conocimiento científico como el conocimiento comunitario local. La AbE propone que los ecosistemas pueden ser manejados para limitar los impactos del cambio implementando enfoques basados en el ecosistema para la adaptación que incluyan la gestión sostenible, la conservación y la rehabilitación de ecosistemas teniendo en cuenta los múltiples beneficios sociales, económicos y culturales para la sociedad. Para la implementación de la AbE en el Caribe resulta esencial prestar atención a temas como el incremento de la resiliencia a partir de la rehabilitación ecológica de arrecifes coralinos, manglares y playas, entre otros ecosistemas relevantes en esta región por su función como protectores de la costa, de la degradación de los suelos agrícolas y de la calidad del agua por la intrusión salina; al estudio, control y manejo de las invasiones biológicas y de nuevas y crecientes amenazas a estos ecosistemas posiblemente asociadas al cambio climático como las arribazones de sargazos a las costas caribeñas; y a la definición de mejoras en las herramientas esenciales para el manejo y gestión de la zona marina y costera como el planeamiento espacial marino, la elaboración de estrategias locales de adaptación, la evaluación de la vulnerabilidad ecológica y la evaluación de la salud de los ecosistemas, compartiendo experiencias a través de redes de intercambio como las redes CYTED. La profundización en estos temas contribuirá a la implementación de la AbE con una mirada hacia la naturaleza incluyendo al hombre, como una especie componente esencial del ecosistema, capaz de revertir la situación ambiental actual promoviendo la integración de voluntades, como una alternativa de convivencia entre el cambio climático y el bienestar socioeconómico para el desarrollo sostenible en la región caribeña
... Overview. Given the extent to which Caribbean coral reefs have changed in the last half-century 10,56,57 , as well as changes to the physical (e.g., temperature) and chemical (e.g., pH) conditions of the marine environment 58,59 , it is likely that the current success of octocorals is based on multiple aspects of their organismal biology. Presumably, the mortality of scleractinians 10,56 , the reduction in topographic complexity of reef surfaces 6 , and the provision of vacant space, which often has been exploited by macroalgae 13 , has also modified the coral reef habitat to the benefit of octocorals. ...
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Declines in abundance of scleractinian corals on shallow Caribbean reefs have left many reefs dominated by forests of arborescent octocorals. The ecological mechanisms favoring their persistence require exploration. We quantified octocoral communities from 2014 to 2019 at two sites in St. John, US Virgin Islands, and evaluated their dynamics to assess whether portfolio effects might contribute to their resilience. Octocorals were identified to species, or species complexes, and their abundances and heights were measured, with height ² serving as a biomass proxy. Annual variation in abundance was asynchronous among species, except when they responded in similar ways to hurricanes in September 2017. Multivariate changes in octocoral communities, viewed in 2-dimensional ordinations, were similar between sites, but analyses based on density differed from those based on the biomass proxy. On the density scale, variation in the community composed of all octocoral species was indistinguishable from that quantified with subsets of 6–10 of the octocoral species at one of the two sites, identifying structural redundancy in the response of the community. Conservation of the relative colony size-frequency structure, combined with temporal changes in the species represented by the tallest colonies, suggests that portfolio effects and functional redundancy stabilize the vertical structure and canopy in these tropical octocoral forests.
... Bleaching-induced coral mortality has led to drastic changes in coral biodiversity, productivity, structure, and functioning (Cornwall et al., 2021;Darling et al., 2017;Graham et al., 2015;McWilliam et al., 2018). Important hotpots, such as the Caribbean and Indo-pacific, are decreasing in coral cover by an average of 80% and 50% respectively over the last 40 years De'ath et al., 2012;Gardner, 2003;Schutte et al., 2010). There are studies showing that the transmission and prevalence of coral diseases are strongly associated with ocean warming (Aeby et al., 2020;Howells et al., 2020;Jones et al., 2004;Wall et al., 2018), high ultra violet radiation (Boyett et al., 2007;Coles and Brown, 2003), pollution, and low water quality (Nalley et al., 2021;Redding et al., 2013;Zhao et al., 2021). ...
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Although knowledge on the diseases affecting corals has been accumulating exponentially since the 2000s, even more effort is required to summarize and guide further investigation. Here, we used the Web of Science database to review 226 studies published, between 2000 and 2020, to identify the major geographic and taxonomic gaps in the literature, and propose future directions for the study of coral diseases. We classified the studies according to the ocean, ecoregion, coral species, disease types, approach (e.g., observational or experimental), and depth. In total, 22 types of diseases were reported for 165 coral species. Acropora spp. was the most studied taxa with 12 types of diseases and 8.2% of the records. Black band, white plague, white syndromes, skeletal eroding, dark spot, and yellow band were the six most common diseases, accounting together for 76.8% of the records. As expected, most studies were conducted in the Caribbean and Indo-Pacific (34.0% and 28.7%, respectively), but only in 44 of the 141 global ecoregions that harbour corals. Observational approaches were the most frequent (75.6% of the records), while experimental approaches accounted for 19.9% and were mainly done on Acropora. The vast majority of studies (∼98%) were performed in shallow waters (<30 m depth). We conclude that over the past two decades, coral diseases have been assessed on a very small fraction of coral species, in very few locations around the globe, and at a limited range of their depth distribution. While monitoring bleaching is mandatory for reef ecology and conservation, the ecoepidemiology of coral diseases deserves more space in the research agenda of reef ecosystems.
... Coral reefs are among the most biodiverse and productive ecosystems in the world (Hoegh-Guldberg 1999). However, their prevalence has dramatically declined over recent decades (e.g., Pandolfi et al. 2003;Burke and Maidens 2004;Schutte et al. 2010;Burke et al. 2011;Eddy et al. 2021). This decline has been largely attributed to the effects of various stressors, both natural and anthropogenic (Gardner et al. 2003;Carpenter et al. 2008;Vega Thurber et al. 2014). ...
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Coral reefs ecosystems are facing an unprecedented decline due to the action of natural and anthropogenic stressors. The Caribbean Sea is considered to be one of the most impacted areas, as the average estimated scleractinian coral cover in this region decreased from approximately 50% to 10% over the last 30 years. In this study, a ten-year biodiversity survey was used to examine changes in abundance and percentage cover of benthic invertebrates on permanent transects located at four shallow coral reefs around Roatán, Honduras. This study represents the first long-term investigation of the coral ecosystem of Roatán and reports a decrease in scleractinian coral cover from 37.45 [± 5.37]% to 28.95 [± 3.62]% and a concomitant increase in macroalgal (7.02 [± 3.59]% to 13.94 [± 2.69]%) and turf (5.11[ ± 0.84]% to 7.23 [± 1.00]%) cover although no significant differences in the abundance of scleractinian corals, soft corals, or sponges were observed on the transects. While the four reef sites supported more variable benthic communities at the onset of the study, an overall homogenization of the benthic community composition occurred during the study period. Although our study sites were limited to a small region of Roatán’s southern coral reef system, these observations add to results from other Caribbean locations and provide insights into how Mesoamerican coral reefs have changed over the last decade.
... Coral reefs in the twenty-first century differ from those described by ecologists in the twentieth century (e.g., Goreau 1959;Loya 1972), notably through large reductions in the population sizes of scleractinian corals (Bruno and Selig 2007;Schutte et al. 2010;Hughes et al. 2019). Even though corals may have constructed the framework upon which many reefs are built (Stoddart 1969;Allemand et al. 2011), extant communities often are dominated by taxa other than scleractinians (Norström et al. 2009;Pawlik and McMurray 2020). ...
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Most tropical coral reefs are experiencing declining coral cover, yet interpretation of this generality is tempered by spatial variation in coral cover among reefs separated over 20–200 km. This study addresses such landscape scale variation in coral reefs at 12 sites (7–10 m depth) around St. John (18° 18´ 37.04 N, 63° 43´ 23.17 W) and St. Thomas (18° 20´ 43.57 N, 65° 55´ 13.88 W), US Virgin Islands. Surveys completed from 2011 to 2021 were used to test for spatial variation in community dynamics among islands, shores, and sites. Community synchrony (φ) was used to evaluate portfolio effects in mediating changes in coral communities. From 2011 to 2021, changes in benthic communities differed among sites and times (i.e., there were site × year interactions), and, while coral cover declined at 11 sites, the decline was 2.6-fold faster around St. Thomas than St. John. The loss of coral cover was driven by multiple taxa that differed among sites, thus revealing asynchronous responses to prevailing conditions. Asynchrony suggests that coral communities at some of the sites have the capacity to exploit portfolio effects to modulate stability, yet these effects did not prevent declines in coral cover that reflect island-scale phenomena. In the US Virgin Islands, coral death is overwhelming the capacity for resilience of coral communities that historically may have benefitted from portfolio effects. Until coral assemblages are depleted through taxonomic extirpation, maintenance of the assemblage composition retains the possibility that increased resilience might emerge if environmental challenges can be alleviated.
... These losses from disease and destruction were exacerbated by overfishing and nutrient pollution resulting from widespread and uncontrolled coastal development (Grigg & Dollar, 1990;Rogers & Beets, 2001;Hughes et al., 2003). By the turn of the century, surveys were reporting a regional decline in coral cover of more than 50% across the Western-Atlantic Reef Province (Gardner et al., 2003;Schutte, Selig & Bruno, 2010;Jackson et al., 2014;Cramer et al., 2020), producing a functional homogenization of coral species and a flattening of reef structure (e.g., Álvarez Filip et al., 2009;González-Barrios, Cabral-Tena & Alvarez-Filip, 2021). Understanding the extent of this decline on a regional scale, however, has relied exclusively on the data pooling from limited local observation, and this has led to significant uncertainty regarding its principal cause (Murdoch & Aronson, 1999;Wood, 2007;Jackson et al., 2014). ...
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The development of coral reefs results from the interaction between ecological and geological processes in space and time. Their difference in scale, however, makes it difficult to detect the impact of ecological changes on geological reef development. The decline of coral cover over the last 50 years, for example, has dramatically impaired the function of ecological processes on reefs. Yet given the limited-resolution of their Holocene record, it is uncertain how this will impact accretion and structural integrity over longer timescales. In addition, reports of this ecological decline have focused on intrinsic parameters such as coral cover and colony size at the expense of extrinsic ones such as geomorphic and environmental variables. Despite these problems, several attempts have been made to predict the long-term accretion status of reefs based entirely on the contemporary health status of benthic communities. Here we explore how this ecological decline is represented within the reef geomorphic structure, which represents the long-term expression of reef development. Using a detailed geomorphic zonation scheme, we analyze the distribution and biodiversity of reef-building corals in fringing-reef systems of the Mesoamerican Reef tract. We find a depth-related pattern in community structure which shows that the relative species distribution between geomorphic zones is statistically different. Despite these differences, contemporary coral assemblages in all zones are dominated by the same group of pioneer generalist species. These findings imply that first, coral species distribution is still controlled by extrinsic processes that generate the geomorphic zonation; second, that coral biodiversity still reflects species zonation patterns reported by early studies; and third that dominance of pioneer species implies that modern coral assemblages are in a prolonged post-disturbance adjustment stage. In conclusion, any accurate assessment of the future viability of reefs requires a consideration of the geomorphic context or risks miscalculating the impact of ecological changes on long-term reef development.
... Stony corals are the architects of coral reefs, and as Bruno and Valdivia (2016) have concisely stated, "Coral loss is thus a direct measure of habitat degradation, " and "Macroalgal cover is an indirect measure of reef degradation." In general, deterioration of coral reefs from excessive sedimentation or other causes, is reflected in declining amounts of coral and increasing amounts of algae (e.g., Rogers and Miller, 2006;Schutte et al., 2010;Jackson et al., 2014;Arias-González et al., 2017), although these changes are not always tightly linked. Not all shifts in relative abundance between benthic components involve corals and algae. ...
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Coral reefs in the western Atlantic and Caribbean are deteriorating primarily from disease outbreaks, increasing seawater temperatures, and stress due to land-based sources of pollutants including sediments associated with land use and dredging. Sediments affect corals in numerous ways including smothering, abrasion, shading, and inhibition of coral recruitment. Sediment delivery resulting in deposition and water quality deterioration can cause degradation at the spatial scale of corals or entire reefs. We still lack rigorous long-term studies of coral cover and community composition before, during and after major sediment stress, and evidence of recovery after watershed management actions. Here we present an overview of the effects of terrestrial sediments on corals and coral reefs, with recent advances in approaches to watershed assessment relevant to the delivery of sediments to these ecosystems. We present case studies of northeastern Caribbean watersheds to illustrate challenges and possible solutions and to draw conclusions about the current state of knowledge of sediment effects on coral reefs. With a better understanding of erosion and the pathways of sediment discharge to nearshore reefs, there is the increased potential for management interventions.
... Over the last 50 years, the state of FCR has changed drastically. In the 1970s and 1980s, many reefs were documented with 30-50% living coral cover, high fish diversity and significant structural integrity [10]. However, recent surveys indicate living coral cover has declined to only 5% of the benthic substrate [11], and further declines have occurred as a result of the most recent disease outbreak [12,13]. ...
Knowledge of multi-stressor interactions and the potential for tradeoffs among tolerance traits is essential for developing intervention strategies for the conservation and restoration of reef ecosystems in a changing climate. Thermal extremes and acidification are two major co-occurring stresses predicted to limit the recovery of vital Caribbean reef-building corals. Here, we conducted an aquarium-based experiment to quantify the effects of increased water temperatures and pCO2 individually and in concert on 12 genotypes of the endangered branching coral Acropora cervicornis, currently being reared and outplanted for large-scale coral restoration. Quantification of 12 host, symbiont and holobiont traits throughout the two-month-long experiment showed several synergistic negative effects, where the combined stress treatment often caused a greater reduction in physiological function than the individual stressors alone. However, we found significant genetic variation for most traits and positive trait correlations among treatments indicating an apparent lack of tradeoffs, suggesting that adaptive evolution will not be constrained. Our results suggest that it may be possible to incorporate climate-resistant coral genotypes into restoration and selective breeding programmes, potentially accelerating adaptation.
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The world’s oceans are warming at an unprecedented rate, causing dramatic changes to coastal marine systems, especially coral reefs. We used three complementary ocean temperature databases (HadISST, Pathfinder, and OISST) to quantify change in thermal characteristics of Caribbean coral reefs over the last 150 years (1871–2020). These sea surface temperature (SST) databases included in situ and satellite-derived measurements at multiple spatial resolutions. We also compiled a Caribbean coral reef database identifying 5,326 unique reefs across the region. We found that Caribbean reefs have been warming for at least a century. Regionally reef warming began in 1915, and for four of the eight Caribbean ecoregions we assessed, significant warming was detected for the latter half of the nineteenth century. Following the global mid-twentieth century stasis, warming resumed on Caribbean reefs in the early 1980s in some ecoregions and in the 1990s for others. On average, Caribbean reefs warmed by 0.18°C per decade during this period, ranging from 0.17°C per decade on Bahamian reefs (since 1988) to 0.26°C per decade on reefs within the Southern and Eastern Caribbean ecoregions (since 1981 and 1984, respectively). If this linear rate of warming continues, these already threatened ecosystems would warm by an additional ~1.5°C on average by 2100. We also found that marine heatwave (MHW) events are increasing in both frequency and duration across the Caribbean. Caribbean coral reefs now experience on average 5 MHW events annually, compared to 1 per year in the early 1980s, with recent events lasting on average 14 days. These changes in the thermal environment, in addition to other stressors including fishing and pollution, have caused a dramatic shift in the composition and functioning of Caribbean coral reef ecosystems.
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In recent decades, the cover of fleshy macroalgae has increased and coral cover has decreased on most Caribbean reefs. Coral mortality precipitated this transition, and the accumulation of macroalgal biomass has been enhanced by decreased herbivory and increased nutrient input. Populations of Acropora palmata (elkhorn coral) and A. cervicornis (staghorn coral), two of the most important framework-building species, have died throughout the Caribbean, substantially reducing coral cover and providing substratum for algal growth. Hurricanes have devastated local populations of Acropora spp. over the past 20–25 years, but white-band disease, a putative bacterial syndrome specific to the genus Acropora, has been a more significant source of mortality over large areas of the Caribbean region. Paleontological data suggest that the regional Acropora kill is without precedent in the late Holocene. In Belize, A. cervicornis was the primary ecological and geological constituent of reefs in the central shelf lagoon until the mid-1980s. After constructing reef framework for thousands of years, A. cervicornis was virtually eliminated from the area over a ten-year period. Evidence from other parts of the Caribbean supports the hypothesis of continuous Holocene accumulation and recent mass mortality of Acropora spp. Prospects are poor for the rapid recovery of A. cervicornis, because its reproductive strategy emphasizes asexual fragmentation at the expense of dispersive sexual reproduction. A. palmata also relies on fragmentation, but this species has a higher rate of sexual recruitment than A. cervicornis If the Acropora spp. do not recover, macroalgae will continue to dominate Caribbean reefs, accompanied by increased abundances of brooding corals, particularly Agaricia spp. and Porites spp. The outbreak of white-band disease has been coincident with increased human activity, and the possibility of a causal connection should be further investigated.
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The coral reefs of the Flower Garden Banks (FGB) are among the most sensitive biological communities in U.S. Federal waters of the Gulf of Mexico. In 1973, the Minerals Management Service (MMS) established a program of protective activities at those reefs. The MMS and the National Oceanic and Atmospheric Administration (NOAA) have been monitoring coral populations on a long-term basis to detect incipient changes caused by oil and gas activities. The results also help in explaining the widespread degradation of reef ecosystems observed in the Caribbean region over the past few decades. Two sites, each 100 × 100 m and 17-26 m deep, have been monitored since 1988: one on the East FGB and the other on the West FGB. The mean coverage of living hard corals exceeded 50% at the two banks in 2002-2003, consistent with estimates of coral cover in previous years. We compared our results from 2002-2003 with data collected during the same period on protected reefs within the Florida Keys National Marine Sanctuary (FKNMS). Low values of coral cover on the reefs in the FKNMS exemplify how catastrophic mortality of the formerly dominant Acropora spp. led to degradation of coral assemblages throughout the Caribbean. The FGB remained in exceptionally good condition, largely for reasons of geography; their northern location excluded the cold-sensitive acroporids, so the regional-scale loss of acroporids did not reduce coral cover. The continuing multidecadal baseline of reef condition generated by the monitoring program at the FGB will enable managers to make informed decisions in the event of future changes to their biota. © 2005 by the Marine Environmental Sciences Consortium of Alabama.
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Populations of Diadema antillarum and Tripneustes ventricosus were monitored on the forereef and backreef at Discovery Bay, Jamaica, from 1995 to 2000. On the primarily coralline substrate of the forereef, T. ventricosus densities were normally low but showed a ten-fold increase in 1998, followed by a decline to normal levels. On the backreef, with both coral and seagrass, T. ventricosus densities were normally higher but a similar peak occurred in 1999. Diadema antillarum densities remained relatively constant on the backreef but declined on the forereef in 1998, and then increased rapidly. Both urchins remained on their respective coral and seagrass habitats on the backreef, but occurred together in similar habitat on the forereef. We propose that the appearance and subsequent decline of T. ventricosus enabled D. antillarum to increase after T. ventricosus had mechanically cropped macroalgae to levels easier for D. antillarum to manage. This explanation suggests that the ephemeral appearance of T. ventricosus on coral covered with macroalgae can act as a successional stage for the reestablishment of D. antillarum.
Six coral reef locations between Miami and Key West were marked with stainless steel stakes and rephotographed periodically between 1984-1991. The monitored areas included Looe Key National Marine Sanctuary, Key Largo National Marine Sanctuary, and Biscayne National Park. All six areas lost coral species between the initial survey year and 1991. Survey areas lost between one and four species; those losses constituted 13-29% of their species richness. Five of the six areas lost live coral cover. Net losses ranged from 7.3-43.9%. In the one station showing an increase in coral cover, the increase was only for the canopy branches of Acropora palmata; understory branches of this same species lost surface area at the same rate as canopy branches gained area. For most of the common species, there was a reduction in the total number of living colonies in the community, and a diminution in the number of large, mature colonies. There was no recruitment by any of the massive frame building coral species. Sources of mortality identifiable in the photographs include: 1) black band disease and 2) "bleaching'. Loss rate of this magnitude cannot be sustained for protracted periods if the coral community is to persist in a configuration resembling historical coral reef community structure in the Florida Keys. -from Authors