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Coral community decline at a remote Caribbean island: Marine no-take reserves are not enough

Authors:
  • Caribbean Institute of Science and Sustainability

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1.Coral reefs around the world have been deteriorating over decades owing to anthropogenic pressure. In the Caribbean recent rates of decline are alarming, particularly for coral reefs under high local human impact, many of which are severely degraded, although regions with lower direct anthropogenic influence seem less affected.2.Little Cayman is a relatively undeveloped island, with less than 150 permanent residents. About 20% of its reefs have been protected by no-take marine reserves since the mid-1980s. We analysed the dynamics of coral communities around the island from 1999 to 2004 in order to test the hypothesis that a lack of major local anthropogenic disturbances is enough to prevent decline of coral populations.3.Live hard coral coverage, coral diversity, abundance, mortality, size, and prevalence of disease and bleaching were measured using the Atlantic and Gulf Rapid Reef Assessment methodology (line transects) at nine sites. Despite the apparent undisturbed condition of the island, a 40% relative reduction of mean live coral coverage (from 26% to 16%, absolute change was 10%) was recorded in five years. Mean mortality varied from year to year from 23% to 27%. Overall mean diameter and height have decreased between 6% and 15% on average (from 47 to 40 cm for diameter, and from 31 to 29 cm for height).4.The relative abundance of large reef builders of the genus Montastraea decreased, while that of smaller corals of the genera Agaricia and Porites increased. Disease prevalence has increased over time, and at least one relatively large bleaching event (affecting 10% of the corals) took place in 2003.5.Mean live coral cover decline was similar inside (from 29% to 19%) and outside (from 24% to 14%) marine no-take reserves. No significant difference in disease prevalence or clear pattern in bleaching frequency was observed between protected and non-protected areas. It is concluded that more comprehensive management strategies are needed in order to effectively protect coral communities from degradation.Copyright © 2007 John Wiley & Sons, Ltd.
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AQUATIC CONSERVATION: MARINE AND FRESHWATER ECOSYSTEMS
Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
Published online in Wiley InterScience
(www.interscience.wiley.com) DOI: 10.1002/ aqc.822
Coral community decline at a remote Caribbean island:
Marine no-take reserves are not enough
VA
ˆNIA R. COELHO
a,c,
* and CARRIE MANFRINO
b,c
a
Department of Natural Sciences and Mathematics, Dominican University of California, San Rafael, California, USA
b
Department of Geology and Meteorology, Kean University, Union, New Jersey, USA
c
Central Caribbean Marine Institute, Princeton, New Jersey, USA
ABSTRACT
1. Coral reefs around the world have been deteriorating over decades owing to anthropogenic
pressure. In the Caribbean recent rates of decline are alarming, particularly for coral reefs under high
local human impact, many of which are severely degraded, although regions with lower direct
anthropogenic influence seem less affected.
2. Little Cayman is a relatively undeveloped island, with less than 150 permanent residents. About
20% of its reefs have been protected by no-take marine reserves since the mid-1980s. We analysed
the dynamics of coral communities around the island from 1999 to 2004 in order to test the
hypothesis that a lack of major local anthropogenic disturbances is enough to prevent decline of
coral populations.
3. Live hard coral coverage, coral diversity, abundance, mortality, size, and prevalence of disease
and bleaching were measured using the Atlantic and Gulf Rapid Reef Assessment methodology (line
transects) at nine sites. Despite the apparent undisturbed condition of the island, a 40% relative
reduction of mean live coral coverage (from 26% to 16%, absolute change was 10%) was recorded in
five years. Mean mortality varied from year to year from 23% to 27%. Overall mean diameter and
height have decreased between 6% and 15% on average (from 47 to 40 cm for diameter, and from 31
to 29 cm for height).
4. The relative abundance of large reef builders of the genus Montastraea decreased, while that of
smaller corals of the genera Agaricia and Porites increased. Disease prevalence has increased over
time, and at least one relatively large bleaching event (affecting 10% of the corals) took place in 2003.
5. Mean live coral cover decline was similar inside (from 29% to 19%) and outside (from 24% to
14%) marine no-take reserves. No significant difference in disease prevalence or clear pattern in
bleaching frequency was observed between protected and non-protected areas. It is concluded that
more comprehensive management strategies are needed in order to effectively protect coral
communities from degradation.
Copyright #2007 John Wiley & Sons, Ltd.
Received 12 May 2006; Revised 2 October 2006; Accepted 12 November 2006
*Correspondence to: V. R. Coelho, Department of Natural Sciences and Mathematics, Dominican University of California, 50 Acacia
Avenue, San Rafael, CA 94901, USA. E-mail: vcoelho@dominican.edu
Copyright #2007 John Wiley & Sons, Ltd.
KEY WORDS: coral community structure; coral community dynamics; marine reserves; coral disease; coral
bleaching; no-take zones; marine protected areas; marine parks; Caribbean reef; Cayman Islands; Little Cayman
INTRODUCTION
Coral reefs around the world have been deteriorating over decades owing to anthropogenic pressure, but
recent rates of decline in the Caribbean are particularly alarming. The main causes of deterioration have been
attributed to overfishing, pollution, increased sedimentation, climate change and disease (Harvell et al.,1999;
Jackson et al., 2001; Bruno et al., 2003; Gardner et al., 2003; Hughes et al., 2003; Pandolet al.,2003).
Hurricanes are natural disturbances that also negatively impact reefs, but affected areas usually recover with
time; however, if additional types of stress are present, such as high pollution levels that cause coral larvae
and juvenile mortality, recovery can be compromised (Hughes, 1994; Connell, 1997; Green et al.,1999;
Hughes and Connell, 1999; Bythell et al., 2000; Nystrom et al., 2000; Negri et al.,2002;Smithet al.,2003).
Coral reefs under strong local human influence typically are severely degraded, but some regions with
lower direct anthropogenic influence seem less affected (Pandolfi et al., 2003), however, Aronson and Precht
(2001) and Kramer (2003) have suggested that even reefs distant from main population centres are
susceptible to degradation. Presumably non-local sources of stress may still affect more remote reefs
(Allison et al., 1998; Harvell et al., 1999), since ocean currents can potentially transport pollutants and
pathogens to less developed areas (Dodge and Gilbert, 1984; Atwood et al., 1987; Guzma
´n and Garcı
´a,
2002; McCallum et al., 2003; Kim et al., 2005). Coral reefs are also vulnerable to climate change, which can
lead to more stronger hurricanes, thermal bleaching and increased coral mortality, as well as ocean
acidification and consequent decrease in calcification rates (Kleypas et al., 1999; Hughes et al., 2003;
Webster et al., 2005). The possible detrimental effects of airborne African dust on Caribbean coral reef
health is another suggested source of stress (Shinn et al., 2000; Griffin et al., 2003).
Currently, the reefs around the Cayman Islands are considered to be in average to slightly higher than
average condition for Caribbean reefs (Kramer, 2003; Pandolfi et al., 2003). Little Cayman, with an area of
approximately 18 km
2
, is the smallest of the three Cayman Islands, and the one with the best coral reef
conditions in terms of both benthic and fish communities (Manfrino et al., 2003; Pattengill-Semmens and
Semmens, 2003). It has fewer than 150 permanent residents and five small (10-room) to medium-size
(40-room) hotels. Furthermore, about 20% of its reefs have been within the boundaries of no-take marine
reserves since the mid-1980s, and even outside these protected areas fishing pressure is low (Manfrino et al.,
2003; Pattengill-Semmens and Semmens, 2003). For Caribbean reefs today, the conditions of this relatively
remote, undeveloped island may be considered exemplary of low local human impact on a coral reef system.
The dynamics of coral communities on the reefs around Little Cayman from 1999 to 2004 were analysed,
in order to test the hypothesis that a lack of major local anthropogenic disturbances is enough to prevent
decline of coral populations.
Overfishing, increased sedimentation due to urban development and point-source pollution are not major
problems in Little Cayman because of the low human population density and the presence of no-take
reserves. Thus, any signs of deterioration in reefs around the island should be linked to non-point-source
pollution, changes in global-scale processes, and/or natural impacts. Conversely, if the latter factors are
unimportant or minimal, little or no degradation should be recorded over time.
METHODS
The study sites were chosen according to the topography of the reefs, oceanographic setting, and degree of
protection from fishing. All sites were spur and groove formations, some were located on the south side of
V.R. COELHO AND C. MANFRINO
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
the island, which is more windward, while others were located on the north side which is considered leeward
of the prevailing easterly trades (Manfrino et al., 2003). For each oceanographic setting (windward or
leeward), sites were located inside and outside marine no-take reserves (Figure 1; Table 1). All sites were
between 9 and 13 m deep, and were sampled in summer (June 1999, July–August in 2002–2004).
The methodology was based on the Atlantic and Gulf Rapid Reef Assessment protocol (www.agrra.org).
Every year 8–15 randomly placed 10-m line transects at each site were used to measure live hard coral
cover, coral diversity, abundance, maximum diameter, maximum height, mortality, disease and bleaching.
Transects were placed parallel to the spurs, in order to avoid crossing the sandy groove areas. A 1-m pole,
divided into 10-cm intervals (the first 10-cm interval was subdivided into 1-cm intervals), was used to
measure live coral cover beneath the line of the 10-m transects. Each coral colony under the line transect
was identified to species, and its maximum diameter and height measured. Only colonies at least 10 cm in
diameter were included in the surveys. Although the number of colonies smaller than 10 cm was not
quantified, such small corals were rarely observed in the field (VRC and CM, pers. obs.). Coral mortality
extent per colony was estimated from a view from above perpendicular to the axis of growth. Presence and
type of disease were recorded after Sutherland et al. (2004). Bleached corals included corals that were
distinctively pale or had at least a portion of the colony bleached white. All researchers received a 3–4 day
training programme to standardize measurements and species identification after Humann and Deloach
(2002).
During the study period, 1999–2004, Little Cayman was within the ‘hurricane’s strike circle’ (62.5
nautical miles radius) of two large storms: Tropical Storm Isidore in September 2002, and Category 1
Hurricane Lili in October 2002. In November 2001, Category 4 Hurricane Michelle directly hit Cuba,
passing about 122 nautical miles off Grand Cayman and reportedly caused heavy surf, storm surge and
flooding in the Caymans (Caribbean Hurricane Network www.stormcarib.com, NOAA www.noaa.gov).
Two hurricanes affected the Cayman Islands in 2004 but the data of the current study was collected prior
to both of them: Category 2 Hurricane Charley passed in strike range of Little Cayman in August and
Figure 1. Map of Little Cayman showing sampling sites. Light grey area: no-take marine reserves (parks). Modified after Pattengill-
Semmens and Semmens (2003).
CORAL COMMUNITY DECLINE AT LITTLE CAYMAN
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
Category 5 Hurricane Ivan passed west of Grand Cayman in September (Caribbean Hurricane Network
www.stormcarib.com, NOAA www.noaa.gov).
Statistics
The statistical analyses were performed using InStat version 3.0b and Prism version 4.0a (GraphPad
Softwares, San Diego, California, USA; www.graphpad.com). Data on live coral cover failed normality
tests, and transformations either could not restore normality or revealed unequal variances after testing.
Therefore, Kruskal–Wallis tests were used to identify whether there were statistical differences over time.
Apost hoc test, Dunn’s Multiple Comparison, was also performed to determine pairwise differences. The
same type of analysis was used to investigate differences in coral mortality, diameter and height, as well as,
mean diameter, height and mortality of the most abundant coral genera over time (after normality testing
and transformation attempts failed).
Chi-squared (w
2
) tests for independence ðw2
iÞand trend ðw2
trÞwere used to analyse contingency tables in
order to assess, respectively, whether annual differences in prevalence of disease or bleaching and linear
trends occurred. The same types of w
2
tests were applied to examine relative abundance of the main genera
of corals over time, and which ones were most affected by disease and bleaching.
For the w2
tr test equally spaced scoring (1, 2, 3, 4) was used while it would have been better to use scoring
that reflected the gap in years (1, 4, 5, 6), but the software did not allow for these changes. Small difference
in scoring is not considered critical for this type of test (Altman, 1991).
Table 1. Sites sampled per year in Little Cayman
Site Lat (80N) Long (80W) D RE PL Y LT C
1 Coral City 19 40.845 80 01.398 10 Wi NP 1999 15 152
2002 08 112
2004 08 118
2 Grundy’s Garden 19 39.440 80 05.373 09 Wi P 1999 11 90
2002 12 151
2004 08 103
3 Richard’s Reef 19 39.396 80 05.819 11 Wi P 2003 10 145
2004 11 152
4 West Point 19 39.298 80 06.216 10 Wi NP 2003 09 119
2004 10 110
5 Jigsaw Puzzle 19 40.298 80 06.203 13 L NP 2002 10 133
2004 11 136
6 Mixing Bowl 19 41.096 80 04.673 12 L P 1999 15 196
2004 09 144
7 Paul’s Anchor 19 41.666 80 04.166 13 L P 1999 10 139
2003 08 112
2004 10 113
8 Snapshot 19 42.049 80 03.421 10 L NP 1999 12 142
2003 09 125
2004 11 126
9 Sailfin 19 42.410 80 00.732 09 L NP 1999 10 102
2002 11 120
2004 10 123
C: number of corals sampled per year in each site; D: depth (m); L: leeward; Lat: latitude; Long: longitude; LT: number of line
transects per year in each site; N: north; NP: site not protected from fishing; P: site protected from fishing, no-take reserve; PL:
protection level; RE: relative exposure; W: west; Wi: windward; Y: years sampled.
V.R. COELHO AND C. MANFRINO
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
A comparative analysis of disease and bleaching prevalence inside and outside marine protected areas per
year was carried out using Fisher’s Exact tests; when significant differences were observed the Odds Ratio
was also calculated.
RESULTS
Mean live coral cover, mortality and size
Mean live coral cover decreased over time, with a relative reduction of approximately 40% in 5 years (from
26.5% SE 1:3 to 16.3% SE 0:8;absolute change was 10.2%). The number of transects analysed
regarding live coral cover was: 73 in 1999, 41 in 2002, 36 in 2003 and 88 in 2004 (total ¼238). Mean
mortality was unequal over the years, peaking in 2003 at 27.4% SE 1:4;while during the other years it
fluctuated from 22.9% SE 0:8 to 24.5% SE 0:9:Mean coral size has decreased since 1999, i.e. about
15% for diameter (from 46.7 cm SE 1:3 to 39.9 cm SE 1:2) and 9% for height (from 31.4 cm SE 1:1
to 28.7 cm SE 1:0). All changes were statistically significant (Figure 2; Table 2). The total numbers of
colonies analysed regarding mortality, diameter and height were 821 in 1999, 516 in 2002, 501 in 2003, and
1125 in 2004.
0
Mean live coral cover or mortality (%)
Mean diameter or height (cm)
10
20
30
40
50
60
70
80
90
100
1999 2000/ 2001 2002 2003 2004
0
5
10
15
20
25
30
35
40
45
50
55
L
M
D
H
Figure 2. Mean live coral cover, coral mortality, diameter and height over time. Live coral coverage (L) and mortality (M) measured in
percentage (smooth lines), diameter (D) and height (H) measured in centimetres (dashed lines). Sample size in 1999, 2002, 2003, 2004:
number of transects for L: 73, 41, 36, 88; number of corals for M, D and H: 821, 516, 501, 1125. Bars represent standard error
of the mean.
Table 2. Multiple comparison analysis of mean live coral cover, mortality, diameter and height over time
KW DMC
1999–2002 1999–2003 1999–2004 2002–2003 2002–2004 2003–2004
L 47.217, p50:0001* p>0:05 p>0:05 p50:001* p>0:05 p50:001* p>0:05
M 25.11, p50:0001* p50:001* p50:001* p50:001* p50:001* p50:001* p50:001*
D 80.38, p50:0001* p50:001* p50:001* p50:001* p50:001* p50:001* p50:001*
H 39.9, p50:0001* p50:001* p50:001* p50:001* p50:001* p50:001* p50:001*
D: diameter; DMC: Dunn’s Multiple Comparison test; H: height; KW: Kruskal–Wallis test; L: live coral cover; M: mortality;
p:pvalue; *significant.
CORAL COMMUNITY DECLINE AT LITTLE CAYMAN
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
Disease and bleaching
Overall disease and bleaching prevalence differed annually. Disease frequency increased over time, and
bleaching peaked in 2003 but no upward trend was observed over the years (Figure 3(a); Tables 3 and 4).
Five different diseases were recorded, the most common ones being white plague and dark spots
(Figure 3(b)). These two disorders had different prevalence over the years; white plague was highest in 2002
(affecting 5.4% of the corals) but no linear trend was observed over time, while dark spots revealed an
upward trend (from 0% in 1999 to 3.9% in 2004). Black band was not common and showed a decreasing
trend over time. Red band and yellow blotch were extremely rare (Figure 3(b); Tables 3 and 4). (There is
some controversy as to whether red and black band diseases are the same or should be considered separate
syndromes, for a review see Sutherland et al. (2004)).
Coral community structure
Thirty-six species of hard corals belonging to 21 genera were recorded during the study (Table 5). Every
year more than 90% of all corals surveyed comprised five genera: Agaricia,Diploria,Montastraea,Porites
and Siderastrea.Montastraea was the most abundant genus but displayed a decreasing trend over time,
0
2
4
6
8
10
1999 2000/ 2001 2002 2003 2004
Diseased
Bleached
0
Diseased corals (%) Diseased or bleached corals (%)
2
4
6
1999 2000/ 2001 2002 2003 2004
BB
DS
RB
WP
YB
(a)
(b)
Figure 3. Frequency of all diseased or bleached corals (a) and frequency of different diseases (b) over time. BB: black band disease, DS:
dark spots disease, RB: red band disease, WP: white plague disease, YB: yellow blotch disease. Sample size: 821 corals in 1999, 516 in
2002, 501 in 2003 and 1125 in 2004.
V.R. COELHO AND C. MANFRINO
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
while the other two top genera, Agaricia and Porites, had increasing trends in relative abundance. Although
their relative abundance changed significantly annually, no linear trends were observed for Diploria,
Siderastrea and all corals of the other less abundant genera combined (Figure 4; Tables 3 and 6).
Agaricia was the only genus in which overall disease prevalence increased over time, and this was due
mainly to dark spots disease; Siderastrea also showed an upward trend for the latter disease over the years.
White plague prevalence differed significantly among the years in Montastraea (peak in 2002) and Agaricia
(peak in 2003), for all other corals no significant pattern was observed (Figures 5(a) and 6; Tables 3 and 6).
Bleaching frequency was unequal among the years, peaking in 2003 for Agaricia,Montastraea and
Siderastrea. The frequency of bleaching over time in the other genera was too low to be tested statistically
(Figure 5(b); Tables 3 and 6).
Table 3. Sample size of corals with bleaching, diseased, with different types of disease, and total number of corals overall and per most
abundant genera and all rare genera combined each year
Agaricia Diploria Montastraea Porites Siderastrea Other Total
1999 TG 180 53 381 101 57 49 821
B1 0 13 03 3 20
D0 0 18 0 3 1 22
BB 0 0 4 0 2 0 6
DS 0 0 0 0 0 0 0
RB 0 0 0 0 0 0 0
WP 0 0 14 0 1 1 16
YB 0 0 0 0 0 0 0
2002 TG 130 29 200 75 31 51 516
B3 1 6 01 0 11
D7 2 15 4 2 1 31
BB 0 0 0 0 0 0 0
DS 2 0 0 0 0 0 2
RB 0 0 0 0 1 0 1
WP 5 2 15 4 1 1 28
YB 0 0 0 0 0 0 0
2003 TG 118 56 165 79 62 21 501
B 8 4 16 5 14 0 47
D9 1 5 1 6 1 23
BB 0 0 0 0 1 0 1
DS 4 0 2 0 4 1 11
RB 0 0 0 0 0 0 0
WP 5 1 3 1 1 0 11
YB 0 0 0 0 0 0 0
2004 TG 297 88 350 222 84 84 1125
B0 2 4 00 0 6
D 40 3 16 5 9 3 76
BB 0 0 0 0 0 0 0
DS 35 1 1 0 7 0 44
RB 0 0 0 1 0 0 1
WP 5 2 13 4 2 3 29
YB 0 0 2 0 0 0 2
B: number of bleached corals; BB: number of corals with black-band disease; D: number of diseased corals; DS: number of corals with
dark spots disease; RB: number of corals with red band disease; TG: total number of corals per genera or all rare genera combined
(other); WP: number of corals with white plague; YB: number of corals with yellow blotch disease.
CORAL COMMUNITY DECLINE AT LITTLE CAYMAN
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
Coral size
Montastraea had the greatest mean diameter (ranging from ca 58 to 66 cm) and height (ca 43 to 52 cm) of all
main genera, followed by Diploria and Siderastrea respectively. Agaricia and Porites had the smallest
dimensions and were of overall similar size (Figures 7(a) and (b); Table 3).
Since 1999, the mean size of Agaricia and Porites has decreased significantly. There was also some
difference in the mean size of Siderastrea over time, but the multiple comparison analysis was unable to
provide significant pairwise differences. No other statistically significant changes in size were observed
(Figures 7(a) and (b); Tables 3 and 7).
Coral mortality
Porites was the genus with the least annual mean coral mortality overall, 15% on average, followed
by Siderastrea 16%, Agaricia 19%, Diploria 21% and Montastraea 37%. The lowest mean mortality
recorded was 11% for Siderastrea in 2002, and the highest was 41.7% for Montastraea in 2003 (Figure 7(c);
Table 3).
Mortality was unequal over the years for all main genera. It significantly decreased with time for Agaricia
and Porites, and increased for Diploria and Montastraea. Multiple comparison analysis failed to provide
significant pairwise differences in mortality over the years for Siderastrea (Figure 7(c); Tables 3 and 7).
Protected and non-protected areas
Mean live coral cover decreased significantly inside and outside marine no-take reserves over time
(Figure 8(a); Tables 8 and 9). Relative reduction in mean live coral cover was similar regardless of
protection level: 36% in protected areas (from 29.4% SE 2:2to18.8%SE1:4;absolute change was
10.6%) and 39% in non-protected areas (from 23.7% SE 1:2to14.4%SE0:8;absolute change
was 9.3%).
There was no significant difference in disease prevalence inside and outside protected areas each year
(Figure 8(b); Tables 8 and 10). Bleaching in protected areas was significantly lower in comparison to non-
protected areas in 1999, but the reverse was observed in 2002. No other significant differences were recorded
(Figure 8(c); Tables 8 and 10).
Table 4. Analysis of 2 4 contingency tables using w
2
tests for independence and trend regarding disease and bleaching over time for
all corals
Disease or bleaching w2
ipvalue w2
tr pvalue
All diseases combined 17.353 0.0006* 13.122 0.0003*
Bleaching 100.98 50.0001* 1.964 0.1611
Types of disease
Black band 12.435 0.0060* 8.494 0.0036*
Dark spots 46.304 50.0001* 44.448 50.0001*
Red band
a
}} } }
White plague 15.933 0.0012* 0.02339 0.8785
Yellow blotch
a
}} } }
w2
i: chi-squared test for independence, 3 degrees of freedom; w2
t: chi-squared test for trend, 1 degree of freedom; }no value; *significant;
a
not enough cases to calculate a valid w
2
test.
V.R. COELHO AND C. MANFRINO
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
DISCUSSION
Gardner et al. (2003) have described an overall 80% relative decline in live hard coral cover in Caribbean
reefs in 24 years, from 1977 to 2001. The current study shows a ca 40% relative reduction (absolute change
was 10%) in just 5 years, from 1999 to 2004, at Little Cayman. Although storms may have contributed to
the decline in Little Cayman their impact alone cannot explain the data, otherwise the reduction in live
coral cover should have happened only after their occurrence and not before or later on. The data show a
continuous decline from 2002 to 2004 in comparison to 1999 (Figure 2). The differences were statistically
significant when comparing 1999 to 2004, and 2002 to 2004, but not from 1999 to 2002, 1999 to 2003, or
2002 to 2003 (Table 2). If storms were the exclusive cause of the decline we would expect to find a significant
Table 5. List of hard coral genera and species found in Little Cayman
during this study. Except for Millepora and Stylaster, which are
hydrocorals (class Hydrozoa), all other coral genera are scleractinians
(class Anthozoa, order Scleractinia)
Genus Species
Acropora A. cervicornis
A. palmata
Agaricia A. agaricites
A. fragilis
A. tenuifolia
Colpophyllia C. natans
Dichocoenia D. stokesii
Diploria D. clivosa
D. labyrinthiformes
D. strigosa
Eusmilia E. fastigiata
Favia F. fragum
Helioceris (Leptoseris)H. cucullata
Isophyllia I. sinuosa
Madracis M. mirabilis
Manicina M. areolata
Meandrina M. meandrites
Millepora M. alcicornis
M. complanata
Montastraea M. annularis
M. cavernosa
M. faveolata
M. franksi
Mussa M. angulosa
Mycetophyllia M. ferox
M. lamarckiana
M. reesi
Porites P. astreoides
P. divaricata
P. furcata
P. porites
Siderastrea S. radians
S. siderea
Solenastrea S. bournoni
Stephanocoenia S. mechelinii
Stylaster S. roseus
CORAL COMMUNITY DECLINE AT LITTLE CAYMAN
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
difference in mean live coral cover from 1999 to 2002 (before and after Hurricane Michelle), or 2002 to 2003
(before and after Tropical Storm Isidore and Hurricane Lili), or 1999 to 2003 (before and after all storms),
which was not the case. Thus, other factors such as disease and bleaching must also be critically important
to the decline observed in Little Cayman.
Disease increased over time mainly owing to dark spots disease, and the presence of white plague was a
chronic condition of these reefs. Dark spots disease has been shown to cause coral tissue necrosis but the
long-term impacts of this disease on coral communities are not well-known (Cervino et al., 2001; Borger,
2005). Conversely, the devastating effects of white plague on Caribbean reefs are well documented
(Bruckner and Bruckner, 1997; Richardson, 1998; Richardson et al., 1998; Nugues, 2002).
Several coral diseases, including black band, dark spots and white plague, are usually more common in
summer owing to higher water temperatures (reviews in Kuta and Richardson (2002) and Sutherland et al.
(2004)). In the current study, sampling always occurred during the summer, thus seasonality cannot explain
the differences observed over the years. Eutrophication has also been linked to greater severity or
prevalence of diseases such as black band and yellow blotch (Kuta and Richardson, 2002; Bruno et al.,
2003). Both syndromes were uncommon at the study sites in Little Cayman and, considering the small
human population density, we would expect low nutrient levels in the reefs around the island. However,
further studies on physical and chemical water parameters are necessary to better understand the
relationship between environmental factors and disease prevalence and severity in Little Cayman.
The highest mean coral mortality and bleaching prevalence were recorded in the same year, suggesting
that this bleaching event had significant lethal effects in the community. Past bleaching occurrences around
the world have been linked to coral mortality, sometimes on a massive scale (Hoegh-Guldberg, 1999;
Aronson et al., 2000, 2002; Stimson et al., 2002; Hughes et al., 2003). However, it is also possible that this
high mean coral mortality could have been at least partially due to the storms’ impact in 2002.
Caribbean coral communities are undergoing profound changes in community structure. Areas where
frame-builders, such as Acropora and Montastraea, were the most abundant corals are now being
dominated by non-frame-builders of the genera Agaricia and Porites (Hughes, 1994; Aronson et al., 1998;
Miller et al., 2000; Aronson and Precht, 2001; Edmunds and Carpenter, 2001; Knowlton, 2001;
0
10
20
30
40
50
1999 2000/ 2001 2002 2003 2004
Agaricia
Diploria
Montastraea
Porites
Siderastrea
Other
Corals (%)
Figure 4. Frequency of coral colonies of the most abundant genera and of all rare genera combined over time. Sample size: 821 corals
in 1999, 516 in 2002, 501 in 2003 and 1125 in 2004.
V.R. COELHO AND C. MANFRINO
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
Gardner et al., 2003; Manfrino et al., 2003). The current study provides evidence of the structural shift in
the coral communities around Little Cayman; Montastraea is still the most abundant single genus, but
Agaricia and Porites are increasing in relative abundance and, when combined, form the dominant group of
corals. The latter two genera have the smallest size of all main coral genera found in Little Cayman, while
the former genus is the largest.
Despite their increase in relative abundance, Agaricia corals are being heavily affected by bleaching and
disease, including dark spots and white plague, while Porites colonies are the least affected. Both genera
comprise small, opportunistic corals that are more vulnerable during storms than large reef-builders
Table 6. Analysis of 2 4 contingency tables using w
2
tests for independence and trend regarding relative abundance, overall disease,
white plague, dark spots and bleaching over time for the most abundant genera and corals of the rare genera combined
Genus w2
ipvalue w2
tr pvalue
Relative abundance
Agaricia 5.520 0.1374 4.233 0.0396*
Diploria 13.574 0.0035* 3.142 0.0763
Montastraea 52.148 50.0001* 49.677 50.0001*
Porites 20.739 0.0001* 20.091 50.0001*
Siderastrea 17.660 0.0005* 1.276 0.2587
Others 14.506 0.0023* 0.1424 0.7059
All diseases combined
Agaricia 29.815 50.0001* 29.391 50.0001*
Diploria
a
}}
Montastraea 4.222 0.2385 0.2895 0.5905
Porites 6.283 0.0986 0.3251 0.5685
Siderastrea 1.567 0.6670 1.510 1.510
Others
a
}}
White plague
Agaricia 8.785 0.0323* 0.7183 0.3967
Diploria
a
}}
Montastraea 8.361 0.0391* 0.3481 0.5552
Porites 6.928 0.0742 0.08374 0.7723
Siderastrea
a
}}
Others
a
}}
Dark spots
Agaricia 36.967 50.0001* 32.991 50.0001*
Diploria
a
}}
Montastraea
a
}}
Porites
}
}}
Siderastrea 7.239 0.0647 6.645 0.0099*
Others
a
}}
Bleaching
Agaricia 25.712 50.0001* 0.1828 0.6690
Diploria
a
}}
Montastraea 24.274 50.0001* 0.6514 0.4196
Porites
a
}}
Siderastrea 27.699 50.0001* 0.2293 0.6321
Others
a
}}
w2
i: chi-squared test for independence, 3 degrees of freedom; w2
t: chi-squared test for trend, 1 degree of freedom; }no value; *significant;
a
not enough cases to calculate a valid w
2
test;
}
no dark spots disease recorded.
CORAL COMMUNITY DECLINE AT LITTLE CAYMAN
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
(Knowlton, 2001; Gardner et al., 2003), thus it is possible that this shift could decrease reef resilience to
disturbances such as hurricanes.
Mortality decreased over time in Agaricia and Porites. Conversely, in Montastraea and Diploria, the two
largest frame-builders, mortality increased. The latter genera were also the ones most affected by white
plague, and among the most affected by bleaching in 2003.
It appears that undeveloped areas may be more vulnerable to the decline of coral communities than
would be expected; which is consistent with the conclusion of a large-scale analysis of Caribbean reefs
(Kramer, 2003). Moreover, it is concluded that the presence of no-take marine reserves established 20 years
ago has proved inadequate to provide protection from degradation to the coral communities.
No-take marine reserves are not enough
No-take reserves help increase fish abundance and biomass, and may diminish competition for space
between corals and algae (McCook et al., 2001; Halpern, 2003; Mumby et al., 2006), but they cannot
prevent coral disease outbreaks or bleaching events from happening, as suggested by Allison et al. (1998)
and shown in the current study. Considering the rapid loss of live coral cover observed in Little Cayman,
it is clear that more comprehensive management programmes are necessary (Bellwood et al., 2004;
0
4
8
12
16
1999 2000/ 2001 2002 2003 2004
Agaricia
Diploria
Montastraea
Porites
Siderastrea
Other
0
5
10
15
20
25
1999 2000/ 2001 2002 2003 2004
(a)
(b)
Disease (%)Bleaching (%)
Figure 5. Frequency of disease (a) and bleaching (b) in the most abundant coral genera and in all rare coral genera combined over
time. Sample size can be found in Table 3.
V.R. COELHO AND C. MANFRINO
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
Birkeland, 2004; Grigg et al., 2005; Pandolfi et al., 2005). The creation of reserves should be seen as one step
and not as a final solution.
Coral communities are under multiple stresses and therefore require complex management strategies.
Several mechanisms to slow down or perhaps reverse decline rates can be implemented at local or national
levels such as: restoration initiatives (Heyward et al., 2002; Epstein et al., 2003; Rinkevich, 2005), disease
control programmes (Hudson, 2000; Bruckner, 2002), possible bleaching prevention or attenuation by
shading corals when water temperatures approach critical levels (West and Salm, 2003; Strong et al., 2004),
and methods to increase herbivory pressure (e.g. rearing Diadema in laboratory conditions (Idrisi et al.,
2003) and placing them on reefs, as well as the creation of no-take reserves (Mumby et al., 2006)).
Some critics may say that it is not possible to implement many of these measures, e.g. disease
management, on a large scale. Perhaps the solution lies in involving local stakeholders, such as diver
0
2
4
6
8
10
12
14
1999 2000/ 2001 2002 2003 2004
0
2
4
6
8
1999 2000/2001 2002 2003 2004
Agaricia
Diploria
Montastraea
Porites
Siderastrea
Other
(a)
(b)
White plague (%)Dark spots (%)
Figure 6. Frequency of the two most common diseases, white plague (a) and dark spots (b), in the most abundant coral genera and in
all rare coral genera combined over time. Sample size can be found in Table 3.
CORAL COMMUNITY DECLINE AT LITTLE CAYMAN
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
operators, in these initiatives by providing them with the necessary materials, training and expertise,
organizing a support network that would manage the data being gathered and offer guidance. Successful
examples of community involvement in protection of marine organisms and resources include Christie et al.
(1994), Pattengill-Semmens and Semmens (2001) and Goffredo et al. (2004).
0
20
40
60
80
1999 2000/ 2001 2002 2003 2004
Agaricia
Diploria
Montastraea
Porites
Siderastrea
0
20
40
60
1999 2000/ 2001 2002 2003 2004
0
10
20
30
40
50
1999 2000/ 2001 2002 2003 2004
(a)
(b)
(c)
Mean height (cm) Mean diameter (cm)Mean Mortality (%)
Figure 7. Mean diameter (a), height (b) and mortality (c) of the most abundant coral genera over time (sample size in Table 3). Bars
represent standard error of the mean.
V.R. COELHO AND C. MANFRINO
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
Identifying corals that are particularly resistant to bleaching (West and Salm, 2003; Obura, 2005) or
disease, harvesting their larvae and propagating recruits with these selected genotypes, and protecting areas
that have low bleaching historically because of natural hydrodynamic patterns or other factors (West and
Salm, 2003; Obura, 2005) are also important strategies to be considered. More and better managed no-take
marine reserves must be implemented, including international marine protected areas that take into account
migratory routes and breeding grounds (Roberts, 1997; Myers and Worm, 2005; Ormond and Gore, 2005;
Sale et al., 2005; Mora et al., 2006).
These initiatives do not address all causes of the decline but they would help reefs to survive over time.
This is important because, otherwise, by the time other measures may be effectively in place, the coral
populations could be beyond rescue owing to the Allee effect or inverse density dependence (Birkeland,
2004).
To address the other main causes of reef decline, global climate change and pollution, will require the
establishment and enforcement of legal measures on countrywide and worldwide scales. In order to slow
down global climate change, international agreements to curb CO
2
emissions and potentially to remove
CO
2
from the atmosphere are critical (Gru
¨bler et al., 1999; Anderson and Newell, 2004). Pollution sources
such as sewage, oil and heavy metal contamination etc. must be carefully monitored and minimized (Dodge
and Gilbert, 1984; Atwood et al., 1987; Guzma
´n and Garcı
´a, 2002; Negri et al., 2002; Bruno et al., 2003;
Smith et al., 2003).
The issues are difficult but can be resolved. It will take time and this is why delaying the decline of coral
communities is absolutely essential. Failing to do so will mean that these reefs will disappear in the next
decades. The cost of inaction will be the certain collapse of Caribbean reefs.
Table 7. Multiple comparison analysis of mean diameter, height and mortality of the most abundant coral genera over time
KW DMC
a
1999–2002 1999–2003 1999–2004 2002–2003 2002–2004 2003–2004
Diameter
A 64.404, p50:0001* p50:001* p50:001* p50:001* p>0:05 p>0:05 p>0:05
D 7.270, p¼0:0638 }}}}}}
M 3.250, p¼0:3547 }}}}}}
Po 30.547, p50:0001* p>0:05 p50:001* p50:001* p>0:05 p>0:05 p>0:05
S 8.350, p¼0:0393* p>0:05 p>0:05 p>0:05 p>0:05 p>0:05 p>0:05
Height
A 28.960, p50:0001* p>0:05 p50:001* p50:001* p>0:05 p>0:05 p>0:05
D 3.414, p¼0:3320 }}}}}}
M 5.605, p¼0:1325 }}}}}}
Po 15.913, p¼0:0012* p>0:05 p50:01* p>0:05 p50:05* p>0:05 p>0:05
S 8.067, p¼0:0446* p>0:05 p>0:05 p>0:05 p>0:05 p>0:05 p>0:05
Mortality
A 38.472, p50:0001* p50:001* p50:01* p50:001* p>0:05 p>0:05 p>0:05
D 17.060, p¼0:0007* p>0:05 p>0:05 p>0:05 p50:001* p>0:05 p>0:05
M 13.463, p¼0:0037* p>0:05 p50:05* p50:05* p>0:05 p>0:05 p>0:05
Po 13.645, p¼0:0034* p>0:05 p50:01* p>0:05 p>0:05 p>0:05 p>0:05
S 8.214, p¼0:0418* p>0:05 p>0:05 p>0:05 p>0:05 p>0:05 p>0:05
A: Agaricia;D:Diploria; DMC: Dunn’s Multiple Comparison test; KW: Kruskal–Wallis test; M: Montastraea; Po: Porites;
S: Siderastrea;p:pvalue; }no value;
a
post hoc test calculated only if p50:05 for Kruskal–Wallis test; *significant.
CORAL COMMUNITY DECLINE AT LITTLE CAYMAN
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DOI: 10.1002/aqc
0
2
4
6
8
1999 2000/ 2001 2002 2003 2004
0
2
4
6
8
10
12
1999 2000/ 2001 2002 2003 2004
0
5
10
15
20
25
30
35
1999 2000/ 2001 2002 2003 2004
P
NP
(a)
(b)
(c)
Disease (%) Mean live coral cover (%)Bleaching (%)
Figure 8. Mean live coral cover (a), disease prevalence (b) and bleaching frequency (c) in protected (P, no-take marine reserves) and
non-protected (NP) areas over time. Sample size can be found in Table 8. Bars represent standard error of the mean.
Table 8. Sample sizes for live coral cover, total number of corals, diseased and bleached corals in
protected (no-take reserves) and non-protected areas
Protected (no-take reserve) Not protected
LT C D B LT C D B
1999 36 425 8 5 37 396 14 15
2002 12 151 10 7 29 365 21 4
2003 18 257 12 23 18 244 12 24
2004 38 512 28 2 50 613 46 6
C: total number of corals; B: number of bleached corals; D: number of diseased corals; LT: number
of line transects.
V.R. COELHO AND C. MANFRINO
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
ACKNOWLEDGEMENTS
We would like to thank two anonymous reviewers and the editor, Dr John Baxter, for their comments on an earlier
version of the manuscript. Many thanks to all interns of the coral reef research internship programme at the Central
Caribbean Marine Institute, Jon Clamp manager of the Little Cayman Research Centre, and Marilyn Brandt, for their
help with fieldwork. This project was partially funded by the Central Caribbean Marine Institute and Cayman Airways.
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V.R. COELHO AND C. MANFRINO
Copyright #2007 John Wiley & Sons, Ltd. Aquatic Conserv: Mar. Freshw. Ecosyst. (in press)
DOI: 10.1002/aqc
... Fishers can reap the benefits of increased fish size and abundance by fishing near MPAs, their access regulated via catch share programs and fishing licenses (Gaines et al. 2010, Anderson andUchida 2014). However, poaching, lack of enforcement, limited spillover, and limited management resources often contribute to the failure of MPAs to achieve social and environmental objectives (Coelho 2007, Graham et al. 2008, Huntington et al. 2011, Edgar et al. 2014, Gill et al. 2017). ...
... Resources Bill in the coming years, which will require long-term buy-in by the fisherfolk and improved inclusion of them in management processes. However, poaching, lack of enforcement, inadequate fishing regulations, and limited spillover often contribute to the failure of MPAs to protect fish populations (Coelho 2007, Graham et al. 2008, Huntington et al. 2011, Gill et al. 2017, Bruno et al. 2019. ...
... Even the oldest and most protected sites (where fishing is fully restricted) experienced hard-coral loss and macroalgal increase. This conclusion is concordant with a majority of studies that have found MPAs to be ineffective at mitigating the decline of corals in response to large-scale disturbances including disease, bleaching, and storms(McClanahan et al. 2001, Coelho 2007, Graham et al. 2008, McClanahan 2008, Huntington et al. 2011, Toth et al. 2014, de Bakker et al. 2017). Our study contributes to the growing body of literature indicating that MPAs provide little protection to coral populations, even if they are successful in increasing fish populations. ...
Thesis
Full-text available
Commercial and subsistence fisheries provide livelihoods and fish protein to nearly three billion people annually. This demand has led to overfishing, which disrupts marine ecosystem functioning and threatens fisheries sustainability. Fisheries are some of the most challenging common-pool resource systems (CPRS) for which to develop effective management strategies because they are easily sub-tractable and non-excludable. Without effective institutions to regulate the extraction of the marine species, users are inclined to overharvest resources. Territorial User Rights for Fishing (TURFs) have recently emerged to encourage environmental stewardship in coastal communities by providing effective ownership of fish stocks, further incentivizing sustainable fishing practices. These such community-based fisheries management (CBFM) strategies grant fishers rights to fish in designated areas in exchange for reporting their catch. Belize became the first country in the Caribbean to implement a nationwide TURF system–known as Managed Access (MA)–in 2016, resulting from long-term collaborations between governmental and international fisheries agencies. In this dissertation, I applied Ostrom’s social-ecological systems (SES) framework to understanding and evaluating over forty years of marine resource management in Belize. Using mixed methods, I determined that marine resource management in Belize is institutionally robust (e.g. contains nested and decentralize enterprises), which could lead to the overcoming of collective action problems often found in CPR systems. Next, I described coral reef benthic community structure from 21 sites across the Belize Mesoamerican Barrier Reef (BMBR) following several major disturbances (bleaching, storms, and disease), and attributed them to ocean warming and local human impacts, from 1997-2016. I found two ecologically distinct assemblages between early and late sampling years, significant declines in mean coral cover, and significant increases in macroalgae cover over ~20 years. Lastly, I conducted quantitative interviews of fishers from 10 communities in southern Belize in 2019 and compared their knowledge, attitudes and perceptions to fishers from 2014. I discovered that respondents from both years understand the requirements for getting and renewing MA licenses, yet perceive lack of enforcement as an issue to success. The results of my dissertation provide holistic, science-based advice for sustaining fishers’ livelihoods while preserving coral reef ecosystems in a changing world.
... Within well-designed and enforced MPAs, fish abundance, biomass, and diversity often increase and in some cases spill over into adjacent, non-protected areas [73,[85][86][87][88]. MPAs can reduce other extractive activities that could directly or indirectly impact coral populations [89]. However, a large majority of studies have found that MPAs are not slowing or preventing the decline of reef-building corals [52,67,81,[90][91][92][93], particularly in response to large-scale disturbances. A recent meta-analysis of 18 studies, encompassing 66 MPAs, reported that MPAs did not affect coral loss or recovery in response to large-scale disturbances including disease, bleaching, and storms [39]. ...
Article
Full-text available
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.
... In contrast, a global analysis of coral cover inside and outside MPAs (Selig and Bruno, 2010) showed that from 1969 to 2006, surveyed coral reefs showed more stable coral cover inside MPAs compared to declining coral cover in fished grounds, suggesting that in the long term, the MPAs may offer more benefits to coral reefs than no protection at all. Given these mixed results and a growing body of literature showing empirically that many benthic communities and corals failed to recover inside MPAs due to local stressors and larger disturbances ongoing despite the removal of fishing (Jones et al., 2004;Coelho and Manfrino, 2007;Carassou et al., 2013;Anderson et al., 2014;McClure et al., 2020), there is, indeed, a need to tease apart hierarchical or interacting effects of these multiple disturbances as well as management efforts inside and outside the MPAs, to understand ways to improve benthic and coral recovery. ...
Article
Full-text available
Philippine coral reefs have been on the decline since the 1970s, and this degradation has posed a risk to biodiversity, food security, and livelihood in the country. In an effort to arrest this degradation, marine protected areas (MPAs) were established across the country. MPAs are known to improve fish biomass, but their effect on live coral cover and other benthos is not yet well documented and understood. In this study, 28 MPAs across the Philippines were surveyed comparing benthic cover and indices between protected reefs and adjacent unprotected reefs. No consistent differences were found between reefs inside and outside MPAs through all the benthic categories and reef health indices considered that are indicative of protection effects or recovery within MPAs. However, there were notable site-specific differences in benthic cover across the study MPAs-suggesting that factors other than protection play important roles in influencing benthic cover inside and outside of MPAs. Storm frequency and proximity to rivers, as a proxy for siltation, were the strongest negative correlates to live coral cover. Also, high coastal population, a proxy for pollution, and occurrence of blast and poison fishing positively correlated with high dead coral cover. The lack of significant difference in benthic cover between reefs inside and outside MPAs suggests that protection does not necessarily guarantee immediate improvement in benthic condition. Correlations between benthic condition and storm frequency, siltation, and pollution suggest that it is necessary to augment MPAs with other management strategies that will address the multiple stressors that are usually indiscriminate of MPA boundaries. Supplementing long-term and systematic monitoring of benthic cover and biodiversity inside and outside of MPAs with data on other important environmental and human impact variables will help improve understanding of benthic cover and biodiversity dynamics inside and outside of MPA boundaries.
... The success of an MPA in restoring populations of reef species sensitive to fishing and its long term stabilizing effect on coral cover fuelled the optimism that they could also be effective in managing coral disease. Albeit worldwide literature produced varied results on how an MPA influences coral disease prevalence (Coelho and Manfrino, 2007;McClanahan, 2008;Page, et al. 2009;) several analogies positively implicate an MPA to reduce coral disease progression and spread on the reef. For instance functionally diverse reef fish communities as operating within MPAs encourage the balance between corals and potential disease vectors that could have been ecologically released on fished areas ). ...
... Coral loss in Florida, however, was significantly greater than the Caribbean average throughout this period (Schutte et al., 2010), while the survival of coral recruits and reef recovery were limited (Toth et al., 2014;. Protecting fish stocks does not necessarily reduce the cover of macroalgae, increase coral populations, or preserve or increase the topographic complexity that is critical to maintaining and increasing those fish stocks (Alvarez-Filip et al., 2009Bates et al., 2019;Bood, 2006;Coelho and Manfrino, 2007;Cox et al., 2017;Huntington et al., 2011;Idjadi et al., 2006;Kramer and Heck, 2007;Ledlie et al., 2007;Lowe et al., 2011;McClanahan et al., 2011a,b;Mora, 2008;Myers and Ambrose, 2009;Reyes-Bonilla et al., 2014;Selig and Bruno, 2010;Stockwell et al., 2009;Toth et al., 2014;Ż ychaluk et al., 2012). ...
... Exclusion of activities that damage corals inside marine reserves that directly damage corals (e.g., Asoh et al., 2004;Yoshikawa and Asoh, 2004) and high-intensity tourism (e.g., Lamb and Willis, 2011;Lamb et al., 2014), is likely to mitigate disease by reducing entry points for opportunistic coral pathogens (Page and Willis, 2008;Nicolet et al., 2013;Katz et al., 2014;Lamb et al., 2014). Environmental influences that permeate reserve borders (e.g., Coelho and Manfrino, 2007;McClanahan et al., 2009;Page et al., 2009) have been shown to limit reserve effectiveness. It is also plausible that protected areas facilitate the spread of disease by increasing the number of susceptible coral hosts (McCallum et al., 2005;Bruno et al., 2007;Myers and Raymundo, 2009), or fishes that act as vectors for coral pathogens through feeding injuries (Aeby and Santavy, 2006;Raymundo et al., 2009). ...
Article
Full-text available
Diseases of tropical reef organisms is an intensive area of study, but despite significant advances in methodology and the global knowledge base, identifying the proximate causes of disease outbreaks remains difficult. The dynamics of infectious wildlife diseases are known to be influenced by shifting interactions among the host, pathogen, and other members of the microbiome, and a collective body of work clearly demonstrates that this is also the case for the main foundation species on reefs, corals. Yet, among wildlife, outbreaks of coral diseases stand out as being driven largely by a changing environment. These outbreaks contributed not only to significant losses of coral species but also to whole ecosystem regime shifts. Here we suggest that to better decipher the disease dynamics of corals, we must integrate more holistic and modern paradigms that consider multiple and variable interactions among the three major players in epizootics: the host, its associated microbiome, and the environment. In this perspective, we discuss how expanding the pathogen component of the classic host-pathogen-environment disease triad to incorporate shifts in the microbiome leading to dysbiosis provides a better model for understanding coral disease dynamics. We outline and discuss issues arising when evaluating each component of this trio and make suggestions for bridging gaps between them. We further suggest that to best tackle these challenges, researchers must adjust standard paradigms, like the classic one pathogen-one disease model, that, to date, have been ineffectual at uncovering many of the emergent properties of coral reef disease dynamics. Lastly, we make recommendations for ways forward in the fields of marine disease ecology and the future of coral reef conservation and restoration given these observations.
... Coral loss in Florida, however, was significantly greater than the Caribbean average throughout this period (Schutte et al., 2010), while the survival of coral recruits and reef recovery were limited (Toth et al., 2014;. Protecting fish stocks does not necessarily reduce the cover of macroalgae, increase coral populations, or preserve or increase the topographic complexity that is critical to maintaining and increasing those fish stocks (Alvarez-Filip et al., 2009Bates et al., 2019;Bood, 2006;Coelho and Manfrino, 2007;Cox et al., 2017;Huntington et al., 2011;Idjadi et al., 2006;Kramer and Heck, 2007;Ledlie et al., 2007;Lowe et al., 2011;McClanahan et al., 2011a,b;Mora, 2008;Myers and Ambrose, 2009;Reyes-Bonilla et al., 2014;Selig and Bruno, 2010;Stockwell et al., 2009;Toth et al., 2014;Ż ychaluk et al., 2012). ...
Chapter
Caribbean reefs have experienced unprecedented changes in the past four decades. Of great concern is the perceived widespread shift from coral to macroalgal dominance and the question of whether it represents a new, stable equilibrium for coral-reef communities. The primary causes of the shift—grazing pressure (top-down), nutrient loading (bottom-up) or direct coral mortality (side-in)—still remain somewhat controversial in the coral-reef literature. We have attempted to tease out the relative importance of each of these causes. Four insights emerge from our analysis of an early regional dataset of information on the benthic composition of Caribbean reefs spanning the years 1977–2001. First, although three-quarters of reef sites have experienced coral declines concomitant with macroalgal increases, fewer than 10% of the more than 200 sites studied were dominated by macroalgae in 2001, by even the most conservative definition of dominance. Using relative dominance as the threshold, a total of 49 coral-to-macroalgae shifts were detected. This total represents ~ 35% of all sites that were dominated by coral at the start of their monitoring periods. Four shifts (8.2%) occurred because of coral loss with no change in macroalgal cover, 15 (30.6%) occurred because of macroalgal gain without coral loss, and 30 (61.2%) occurred owing to concomitant coral decline and macroalgal increase. Second, the timing of shifts at the regional scale is most consistent with the side-in model of reef degradation, which invokes coral mortality as a precursor to macroalgal takeover, because more shifts occurred after regional coral-mortality events than expected by chance. Third, instantaneous observations taken at the start and end of the time-series for individual sites showed these reefs existed along a continuum of coral and macroalgal cover. The continuous, broadly negative relationship between coral and macroalgal cover suggests that in some cases coral-to-macroalgae phase shifts may be reversed by removing sources of perturbation or restoring critical components such as the herbivorous sea urchin Diadema antillarum to the system. The five instances in which macroalgal dominance was reversed corroborate the conclusion that macroalgal dominance is not a stable, alternative community state as has been commonly assumed. Fourth, the fact that the loss in regional coral cover and concomitant changes to the benthic community are related to punctuated, discrete events with known causes (i.e. coral disease and bleaching), lends credence to the hypothesis that coral reefs of the Caribbean have been under assault from climate-change-related maladies since the 1970s.
... Other studies have found little evidence that protected areas mitigate coral disease (Coelho and Manfrino 2007, Page et al. 2009), although authors cautioned that either poor compliance with fishing restrictions or the presence of environmental influences that permeate reserve borders could have negated reserve effectiveness in their studies. It is also plausible that protected areas might facilitate the spread of disease by increasing densities or cover of susceptible coral hosts (McCallum et al. 2005, Bruno et al. 2007, Myers and Raymundo 2009, or by increasing densities of fishes that are either vectors for coral patho gens or cause feeding injuries that increase coral susceptibility to opportunistic pathogens Santavy 2006, Raymundo et al. 2009). ...
Chapter
While disease is a part of all natural systems, emerging marine diseases are on the rise and many are exacerbated by anthropogenic stressors. Marine and terrestrial environments are fundamentally different, requiring a suite of new approaches to understanding and managing the host–pathogen–environment relationship. Promising strategies include establishing marine protected areas, developing forecasting tools, and using natural ecosystem filters to control pathogens. Aquaculture is one measurable avenue by which natural systems come into direct contact with managed systems, often with negative consequences. This chapter presents examples where pathogens, invasive species, and degraded water quality are associated with impacts on adjacent natural systems. While effective regulatory procedures exist, international transport presents a challenge to implementation and needs special attention. Ecological restoration, a growing management science, would benefit from consideration of disease processes, such as genotyping to determine differences in natural resistance that could be used to guide selective breeding efforts.
... It is not surprising that some papers have reported no difference for reefs inside and outside MPAs, given that these studies were not restricted to areas that were marine reserves with effective enforcement [73]. Some studies have shown significant coral declines even within no-take zones, attributable to bleaching [74]. The 2005 bleaching episode in the Caribbean, which was at least partly associated with climate change [47], was followed by a disease outbreak which led to substantial losses in coral cover within and outside of a marine protected area [26]. ...
Article
Coral reefs are one of an incredibly beneficial ecosystem on earth. [1] Coral reefs are important for our world for several reasons. Besides the fact that they are very scenic and attract tourists, thus they function as a very huge income in environmental and economic professional employment, such as tourism, and food, and coastal protection, and they have variant benefits for our marine environment and the world such as symbioses and a source to finding medicament. In this review, we point out the importance and possible extinction. Top-down controls of complex food webs maintain the balance among the critical groups of corals, herbivores, and algae, thus allowing the persistence of corals reefs as three-dimensional, biogenic structures with high biodiversity, resistance, resilience and connectivity, heterogeneity, and the delivery of essential goods and professional employment to community. With rapidly increasing rates of contemporary extinction, predicting extinction vulnerability and identifying, how multiple stressors drive non-random species loss have become key challenges in ecology. We developed a novel predictive framework of species extinction vulnerability and applied it to coral reef fishes. Although relatively few coral reef fishes are at risk of global extinction from climate disturbances, a negatory convex relationship between fish species locally vulnerable to climate change vs. fisheries exploitation indicates that the entire community is vulnerable to the many reefs. Techniques to restore the coral reefs impacted by human disturbance are; salvaging sponges and corals, removing loose debris from the reef, rebuilding three-dimensional (3-D) structures onto leveled-scarified reef surfaces, and transplanting sponges and corals back on the cleared reef surfaces. [2] A substantial proportion of the world's living species, including one-third of the reef-building corals, are threatened with extinction and in pressing the need for conservation action.
... However, only 3.6% of global ocean is reported as being protected within actively managed MPAs (Sala et al., 2018). MPAs, are regarded as management tools to conserve coral reefs globally; but many fail to achieve conservation objectives despite the tough management practices that have been implemented (Rinkevich, 2008;Parnell et al., 2005;Jameson et al., 2002;Epstein et al., 2005;Coelho and Manfrino, 2007). However, Russi et al. (2016), demonstrated the benefits from MPAs in Europe and Roberts et al. (2001), showed that MPAs can improve fishing. ...
Article
Tropical coral reefs render a large number of ecosystem services, although without sustainable use practices and conservation measures over the last couple of decades many tropical coral reef ecosystems have been damaged because of excessive use of reef resources. This study provides an estimation of economic benefits from both direct and indirect use of Saint Martin's Coral Island resources, one of the ecologically critical areas (ECAs) of Bangladesh. The coral reef and the associated habitats of St. Martin's Island contributes 33.6 million USD/year to the local economy from fishing, tourism, shoreline protection, seaweed culture, and gathering of intertidal shellfish. Tourism and fisheries are the major economic sectors, generating annual direct use values of 19.4 million USD and 13 million USD/year respectively. The indirect use value of shoreline protection is estimated to be about 180,000 USD/year. Economic benefits of around 1 million USD, could also be generated from an entrance fee collected from tourists visiting the island. The net present value (NPV) of benefits from all of the resources of Saint Martin's Island over a 25-year time frame, with a 6.5% discount rate, is about 545 million USD. A coupled socio-ecological-political, restoration and management framework is recommended. The government should come forward with a conservation, restoration and management plan so that the framework could be used for the management and restoration of a degraded coral reef ecosystem.
Article
Full-text available
Mass mortalities due to disease outbreaks have recently affected major taxa in the oceans. For closely monitored groups like corals and marine mammals, reports of the frequency of epidemics and the number of new diseases have increased recently. A dramatic global increase in the severity of coral bleaching in 1997–98 is coincident with high El Niño temperatures. Such climate-mediated, physiological stresses may compromise host resistance and increase frequency of opportunistic diseases. Where documented, new diseases typically have emerged through host or range shifts of known pathogens. Both climate and human activities may have also accelerated global transport of species, bringing together pathogens and previously unexposed host populations.
Chapter
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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.
Article
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A benthic assessment of the isolated Cayman Islands was completed at 42 sites. Major changes in the reef community structure were documented by comparison with earlier studies. Acropora palmata and A. cervicornis, once abundant as shallow framework builders, were uncommon. Diseased stony corals were seen in >90% of the study sites, with the highest averages in Little Cayman, especially at Bloody Bay which is one of the most highly regulated marine parks in the Cayman Islands. The Montastraea annularis species complex accounted for two-thirds of the diseased corals which, along with other massive species, were affected largely by white-plague disease. Recent partial-colony mortality was particularly high in Grand Cayman. However, small- to intermediate-sized (<1.5 m diameter) colonies and recruits of reef framework builders (including the M. annularis complex) suggest a strong potential for population regeneration. Algal competition generally did not appear to be a problem for stony corals, and bleaching was insignificant, yet more prevalent, in the deeper (>10 m) sites.
Article
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The prevalence and severity of marine diseases have increased over the last 20 years, significantly impacting a variety of foundation and keystone species. One explanation is that changes in the environment caused by human activities have impaired host resistance and/or have increased pathogen virulence. Here, we report evidence from field experiments that nutrient enrichment can significantly increase the severity of two important Caribbean coral epizootics: aspergillosis of the common gorgonian sea fan Gorgonia ventalina and yellow band disease of the reef-building corals Montastraea annularis and M. franksii. Experimentally increasing nutrient concentrations by 2–5· nearly doubled host tissue loss caused by yellow band disease. In a separate experiment, nutrient enrichment significantly increased two measures of sea fan aspergillosis severity. Our results may help explain the conspicuous patchiness of coral disease severity, besides suggesting that minimizing nutrient pollution could be an important management tool for controlling coral epizootics.
Chapter
Yellow band and dark spot syndromes have been frequently observed to affect coral species throughout the Caribbean within the last 10 years. These syndromes significantly impair at least three important reef-building species. Yellow band (also known as yellow blotch) appears as rings or blotches on Montastrea annularis throughout the Caribbean. The coral tissue necrosis occurs at a rate of approximately 0.6 cm/month. Transect measurements at various locations indicated that as many as 90% of M. annularis were affected by yellow band during 1997-98. Tissue samples reveal a 41-96.9% decrease in zooxanthellae/sample compared to healthy specimens, depending on distance from healthy tissue. Mitotic indices (MI) of zooxanthellae (symbiotic algae appearing as doublets) for M. annularis are 2.5%. MI in yellow band samples directly bordering healthy tissue are less than 0.9%, and zooxanthellae directly within the band bordering exposed skeleton had a mitotic index of 0.0%. This indicates impairment of zooxanthellae cell division in yellow band specimens. Zooxanthellae are not expelled and appear vacuolated and devoid of organelles. Dark spot, characterized by tissue necrosis as well as a depression of the colony surface, affects Stephanocoenia michelinii and Siderastrea siderea throughout the Caribbean. Transects showed that as many as 56% of S. michelinii and S. siderea showed signs of dark spot during 1997-98. Affected tissues of S. siderea died at a rate of 4.0 cm/month. In dark spot samples from S. siderea, the total number of zooxanthellae was 56% of that in healthy tissue; dark spot-affected specimens of S. michelinii showed a 14% decrease in the number of zooxanthellae compared to healthy tissue samples. Mitotic indices of zooxanthellae from healthy specimens of S. siderea were 1.20% compared to 0.40% in dark spot samples. Mitotic indices of healthy S. michelinii were 1.54% compared to 0.23% in dark spot samples, also indicating a decrease in algal cell division. Zooxanthellae from dark spot tissue are swollen and darker in pigment. Due to the changes that are evident in the symbiotic algae, we suggest that both syndromes act primarily on the zooxanthellae symbiont, and secondarily on the cnidarian host.
Article
Beginning in the late 1980s, white-band disease nearly eliminated the staghorn coral Acropora cervicornis from reefs in the central shelf lagoon of Belize. The lettuce coral Agaricia tenuifolia replaced Acropora cervicornis in the early 1990s, but anomalously high water temperatures in 1998 caused severe bleaching and catastrophic mortality of Agaricia tenuifolia. The short-lived transition in dominance from Acropora cervicornis to Agaricia tenuifolia left an unambiguous signature in the fossil record of these uncemented lagoonal reefs. Analysis of 38 cores, extracted from 22 sampling stations in a 375-km2 area of the central lagoon, showed that Acropora cervicornis dominated continuously for at least 3000 years prior to the recent events. Agaricia tenuifolia occasionally grew in small patches, but no coral-to-coral replacement sequence occurred over the entire area until the late 1980s. Within a decade, the scale of species turnover increased from tens of square meters or less to hundreds of square kilometers or more. This unprecedented increase in the scale of turnover events is rooted in the accelerating pace of ecological change on coral reefs at the regional level.
Article
The Atlantic and Gulf Rapid Reef Assessment (AGRRA) sampling strategy is designed to collect both descriptive and quantitative information for a large number of reef vitality indicators over large spatial scales. AGRRA assessments conducted between 1998 and 2000 across a spectrum of western Atlantic reefs with different histories of disturbance, environmental conditions, and fishing pressure were examined to reveal means and variances for 15 indicators. Twenty surveys were compiled into a database containing a total of 302 benthic sites (249 deep, 53 shallow), 2,337 benthic transects, 14,000 quadrats, 22,553 stony corals. Seventeen surveys contained comparable fish data for a total of 247 fish sites (206 deep, 41 shallow), 2,488 fish transects, and 71,102 fishes. Shallow (≤ 5 m) reefs were dominated by A. palmata, a good proportion of which was standing dead, while deep (>5m) reefs were nearly always dominated by the Montastraea annularis species complex. Fish communities were dominated by acanthurids and scarids with seranids making up less than 1% of the fish seen on shallow reefs and 4% on deep reefs. AGRRA benthic and fish indicators on deep reefs showed the highest variation at the smallest spatial scale (∼0.1 km), with recent mortality and macroalgal canopy height displaying the largest area and subregional scale (∼1-100 km) variation. A mean live coral cover of 26% for the 20 survey areas was determined for the deep sites. Significant bleaching and disease-induced mortality of stony corals associated with the 1998 (El Niño-Southern Oscillation) ENSO event were most apparent in the western Caribbean and Bahamas subregions and the Montastraea annularis complex was the most heavily impacted. The overall low number of sightings for larger-bodied groupers and snappers (∼< 1/100 m2 ) as a whole suggest that the entire region is overfished for many of these more heavily targeted species. More remote reefs showed as much evidence of reef degradation as reefs more proximal to human coastal development. Characterizing present-day reef condition across the region is a complex problem since there are likely multiple sources of stress operating over several spatial and temporal scales. Not withstanding the many limitations of this analysis, the value of making multiple observations across multiple spatial scales that can approximate the "normal" state for the region today is still very high.