The living cover of Caribbean reef-building Corals has declined by 50% since systematic reef monitoring began in the late 1970s. During this time, the majority of Caribbean reefs have been transformed from habitats dominated by reef-building Corals into habitats dominated by Macroalgae, Sponges, and/or non–reef-building Invertebrates. The decline in Corals has been attributed to fishing, land-based pollution, anthropogenic ocean warming, and outbreaks of Coral and Sea Urchin diseases. Modern ecological studies of Caribbean reefs began in the late 1960s, less than a decade before a series of acute events acted in synergy to rapidly transform Coral communities. Outbreaks of White Band Disease appeared on many reefs in the late 1970s and early 1980s, eventually killing over 80% of the populations of the Elkhorn Coral, Acropora palmata, which previously dominated reef crest zones, and the Staghorn Coral, Acropora cervicornis, which previously dominated midslope zones. Mass mortality of the Black Sea Urchin, Diadema antillarum, in 1983–1984 due to an unidentified pathogen removed this keystone herbivore from reefs that were already largely devoid of large herbivorous Fish because of overfishing. Diadema mortality exceeded 90%, precipitating an explosion of Macroalgae on reefs across the Caribbean. These events were followed by local outbreaks of Coral bleaching beginning in the late 1980s followed by regional outbreaks in the 1990s, leading to further increases in Coral disease and, in some instances, a further replacement of Corals by Macroalgae.
In a
paper published in the journal
Science Advances on 22 April 2020,
Katie Cramer of the
Julie Ann Wrigley Global Institute of Sustainability at
Arizona State University, and the
Center for Oceans at
Conservation International,
Jeremy Jackson of the
Center for Marine Biodiversity and Conservation at the
Scripps Institution of Oceanography, the
Center for Biodiversity and Conservation at the
American Museum of Natural History, the
Smithsonian Tropical Research Institute, and the
Department of Paleobiology at the
National Museum of Natural History,
Mary Donovan of the
Hawai‘i Institute of Marine Biology at the
University of Hawai‘i at Mānoa, and the
Marine Science Institute at the
University of California, Santa Barbara,
Benjamin Greenstein of the
School of Social and Natural Sciences at
Roger Williams University,
Chelsea Korpanty of the
MARUM Center for Marine Environmental Sciences at the
University of Bremen,
Geoffrey Cook of the Department of Biology and Health Science at
New England College, and
John Pandolfi of the
Centre for Marine Science,
School of Biological Sciences, and
ARC Centre of Excellence for Coral Reef Studies at the
University of Queensland, present the results of a study in which they used palaeoecological, historical, and survey data to track
Acropora presence and dominance throughout the Caribbean from the prehuman period to present.
Despite decades of research, the origin and transmission of White Band Disease in Caribbean Acropora are still poorly understood. However, multiple anthropogenic stressors appear to have played a role. Recent observations (1997–2004) of the presence of White Band Disease on Acropora Corals across the Caribbean show a link between contemporary white-band disease and elevated sea surface temperature from anthropogenic climate change. Although the coverage of early survey data is insufficient to investigate causes of the initial white-band disease epidemics of the late 1970s and early 1980s, they may also have been related to temperature stress: Anthropogenic warming of sea surface waters in the Caribbean first became pronounced in the 1970s. Initial and subsequent White Band Disease outbreaks may also have been caused by increased Macroalgal abundance related to overfishing of reef herbivores and/or reef eutrophication, as numerous experiments have found increased disease prevalence in other Caribbean Scleractinian Coral species associated with Macroalgal contact. Nutrient enrichment from land-based runoff has also likely exacerbated White Band Disease outbreaks by suppressing Coral immunity and encouraging growth of pathogenic microbes and allelopathic Algae (Algae which produce chemicals that influence the growth, survival, and reproduction of Corals).. Another hypothesis is that declines in Acroporid and other Corals in the Caribbean are related to an increase in hurricane frequency and intensity due to climate change, which could limit Acropora recovery on reefs already degraded by overfishing and nutrification. Although hurricanes have been a natural occurrence on low-latitude Caribbean Coral Reefs for millions of years, Corals in this region have increasingly failed to recover following major storms.
A Staghorn Coral, Acropora cervicornis, suffering from White Band Disease. Reefs.
The loss of Acroporids has profoundly altered the structure and functioning of Caribbean reef ecosystems, as this genus grows up to 5 to 10 times faster and taller than other Caribbean Coral species and disproportionately contributes to reef architectural complexity and carbonate production. The virtual absence of these species at the shallow zones of most reefs today represents an unnatural state for modern Caribbean reef ecosystems. Studies of uplifted Pleistocene and Holocene reefs revealed that Acropora dominance has persisted for at least the past 250 000 years despite marked fluctuations in temperature and sea level. Isolated surveys of altered reefs have tracked an increase in the relative abundance of disturbance-tolerant and low-relief 'weedy' species such as Porites and Agaricia over the past few decades following the loss of Acropora, indicating an unprecedented shift from dominance of superior competitors to that of stress-tolerant and weedy species.
'Weedy' Coral species, such as this Porites sp. colony in the Florida Keys, are becoming more predominant on Caribbean reeds following the loss of Acropora spp. Louis Wray/Wikimedia Commons.
The scarcity of quantitative ecological surveys before the 1970s leaves open the question of when the declines in Acroporids may have initially begun. Two isolated studies based on historical and palaeontological data suggest that Acroporids began to decline well before the 1970s, most likely due to increases in coastal runoff from land clearing for agriculture, but the geographic extent of these declines is unknown. In contrast, another isolated palaeocological study found that Acropora cervicornis dominance was continuous over the past thousands of years until its abrupt decline beginning in the 1980s. To resolve the initial timing of the widespread decline in Caribbean Acroporids, Cramer et al. compiled an extensive dataset of qualitative and quantitative observations of the presence and dominance of both Staghorn, Acropora cervicornis, and Elkhorn, Acropora palmata, Corals within reef crest and forereef zones at a number of reef sites throughout the Caribbean Sea, spanning the pre-Human Late Pleistocene epoch (roughly 125 000 years before present) to present (2011 AD). To explore the possible causes of declines in these corals, Cramer et al related Acropora dominance since the 1950s to available proxies of potential regional and local disturbances.
Occurrence data of Acropora Corals from the Pleistocene to 2011 were obtained in a variety of ways. Semiquantitative (number of colonies, species abundance rankings, and percent weight of Coral skeletons in reef matrix cores) and qualitative (observations of dominance/commonness/rarity or species’ presence/absence) data were compiled from the primary peer-reviewed scientific literature, government reports, and (less commonly) historical literature, including field notes from early explorers. Quantitative data (percent fossil abundance, percent living cover) were compiled from surveys of uplifted fossil reefs or underwater survey data of modern reefs that were received directly from contributors or gleaned from peer-reviewed literature to construct the Global Coral Reef Monitoring Network database that assessed trends in Caribbean reef benthic communities from 1970 to 2011. For data from the historical and modern periods (1500–2011 AD), presence and dominance values were reported from a single survey encompassing a single day or from a series of multiyear surveys spanning up to 6 years. For palaeoecological data from the Pleistocene and Holocene epochs, presence and dominance values were reported from samples of fossil Coral assemblages that represent one or more centuries of reef growth to entire geologic units that represent up to 60 000 years of reef growth. Fossil data were gathered from reef matrix cores collected below current sea level for the Holocene and from transect surveys of uplifted reefs for the Pleistocene, while modern data were primarily derived from underwater field surveys, although a small number were from boat-based observations and high-resolution aerial photographs.
To strike a balance between providing sufficient temporal resolution and ensuring adequate sample sizes and geographic coverage when assessing the original timing of declines in both Acropora species, data were grouped into 12 time bins: Pleistocene (roughly 125 000 to 12 000 years ago), Holocene (roughly 9100 years ago to 1500 AD), 1500 to 1949, 1950 to 1959, 1960 to 1969, 1970 to 1979, 1980 to 1984, 1985 to 1989, 1990 to 1994, 1995 to 1999, 2000 to 2004, and 2005 to 2011. Bins were reduced to 5-year increments after 1980 (except for a 6-year increment for the most recent bin) due to the large increase in reef survey effort following the mass die-off of the Black Sea Urchin Diadema antillarum.
Data from the literature were extracted from text, tables, figures, and maps. Qualitative data were included in the database if, in addition to presence/absence and/or dominance information for at least one Acropora species, the following set of associated information was also available: (i) age of fossil assemblage or year of observation of modern data, (ii) original source of data, (iii) country and island, coastline, or reef site, and (iv) water depth or reef zone. Data were recorded at survey level, with a survey constituting a unique combination of reef site, depth zone, and year/time period. Surveys constituted individual reef 'sites' in some cases and encompassed more extensive areas such as entire reef tracts, bays, or banks in other cases.
Data were compiled from 'reef crest' and 'midslope' reef zones, the zones where Acropora were noted to occur in early Caribbean reef surveys. Generally, reef crest data spanned 0 to 6 m water depth, and midslope data spanned between 6 and 20 m, as 6 m was the depth at which dominance typically transitioned from Acropora palmata to Acropora cervicornis in the semiquantitative and quantitative data. However, the reef crest/midslope zone delineation was made on a location-by-location basis by first considering water depth, and when available, considering additional environmental characteristics such as wave exposure and reef morphology. For some offshore reef locations with presumably higher water clarity, the boundary between reef zones was closer to 10 m. When a precise water depth was not available, Cramer et al. used Acropora species presence and/or dominance in addition to environmental characteristics to delineate between zones. When not reported in the papers from which data were extracted, palaeo-water depths were determined by (i) defining sea surface as the top of the Pleistocene fossil reef terrace sampled and (ii) constructing a composite Caribbean-wide sea level curve for the Holocene. Surveys from backreef habitats, reef flats, and reef pavements were excluded, as these reef zones are not the preferred environments for Acropora in the Caribbean.
Changes in Acropora presence and dominance were assessed from analysis of percent living Coral cover, abundance rankings, and presence/absence data. Species were ranked by percent Coral cover values, species rankings, and qualitative descriptions of relative abundance (e.g. 'principal reef-building Coral', 1; 'second most commonly found Coral', 2). Presence and ranking values were assigned for subordinate (nondominant) species only if a source contained abundance or relative abundance data for at least two species. For data sources that only listed the presence or dominance of one species, the remaining species were assumed to be nondominant (i.e., ranking 0), but their presence or absence could not be determined.
Temporal changes in the presence and dominance of Acropora across the Caribbean were tracked by computing the proportion of sites with each species present and dominant in each time bin. To account for the uneven geographic distribution of samples across time bins, Cramer et al. assigned each reef site to an island group, large reef tract, or country, and used binomial generalised linear mixed-effects models that included time bin as fixed effect and country as random effect. Cramer et al. chose the country level to represent spatial structure in the data because (i) the imprecision of the geographic locations noted in some of the historical records and sparser sample sizes in the historical and early survey periods did not permit a finer geographic partitioning, and (ii) the only reliable, multidecadal proxy for coastal pollution available, fertiliser consumption compiled by the United Nations Food and Agriculture Organization, is reported at this same spatial resolution. It was not possible to analyse temporal trends within individual countries, as data were too sparse for most countries. Separate models were run for both species in the reef crest and midslope zones. Cramer et al. determined the earliest timing of significant change in the presence and dominance of each Acropora Coral species compared with the prehuman baseline state (the Pleistocene epoch) from Wald’s Z tests (a statistical method intended to assess constraints on statistical parameters) of differences between Pleistocene values with those from other time bins. To account for the possibility that high Acropora presence and dominance values in Pleistocene and Holocene periods may be related to the relatively longer time spans these periods encompass, Cramer et al. excluded these periods and ran additional generalised linear mixed-effects models of the same form for Acropora palmata presence and dominance at the reef crest and Acropora cervicornis presence and dominance at the midslope zone. Estimates of uncertainty for percent presence and dominance were computed from binomial confidence intervals.
Distribution of dominance and presence/absence data for Acroporid Corals. Data from reef crest zones in magenta; data from midslope zones in blue. Size of dot proportional to total number of surveys across both reef zones and all bins combined (range, 1 to 541). Cramer et al. (2020).
To explore the possible causes of initial declines in Acropora, we assessed the influence of a suite of local and regional stressors on Acropora dominance from 1950 to 2011 using binomial generalised linear mixed-effects models. Cramer et al. compiled data for four potential drivers for which reliable data were available from at least the 1960s to the present: temperature stress represented by degree heating months, hurricane activity represented by average number of hurricanes per year, coastal agricultural pollution represented by total fertilizer consumption, and a general proxy of anthropogenic disturbance (including fishing) represented by human population density. Cramer et al. were not able to include an independent measure of fishing effort in their analysis due to the lack of reliable data over the broad geographic and temporal scale investigated in the study.
Temperature data were compiled from the Hadley Centre Global Sea Ice and Sea Surface Temperature dataset, a product that blends historical shipboard temperature records with satellite records. Monthly sea surface temperature averaged over 1 × 1 degree cells was extracted for each year from 1950 to 2011, and degree heating months were computed from latitude and longitude coordinates for each year and reef site by summing the positive monthly temperature anomalies relative to the maximum monthly temperature from the climatological base period from 1900 to 2011. Average and maximum degree heating months were then computed for each reef site and time bin. Geographic coordinates were obtained from Google Earth for reef sites with a specific identifiable location indicated, while coordinates for the center of a more general region or island were used for entries without a specific reef site name recorded. Although Cramer et al. calculated average and maximum degree heating months for each reef site, they chose to include the latter in their analyses because it often included values of over 2°C and, thus, was a more appropriate proxy of acute temperature stress.
Hurricane data were compiled by tallying the number of unique hurricanes (from categories 1 to 5) that were reported to cross within approximately 20 km of each reef site or island (an area that typically encompasses maximum wind speeds) during a time bin and then dividing by the number of years included in that bin. Cramer et al. chose to include hurricanes from all categories because Caribbean Acroporids are susceptible to fragmentation from lower-intensity storms and chose to focus on hurricane frequency because this variable greatly affects Acropora recovery potential. Data were obtained from the Caribbean Hurricane Network, whose analyses were based on 'best track' data taken from the National Oceanic and Atmospheric Administration’s National Hurricane Center’s North Atlantic hurricane database reanalysis project.
Fertiliser consumption data were computed from the quantity of fertiliser (in metric tons) of plant nutrient consumed in agriculture by a country annually from 1961 to 2002 (from the Food and Agriculture Organization of the United Nations’ Fertilizers archive). Annual estimates were averaged across each time bin. For island nations, the value reported for the entire country was used, while for large continental countries and islands (Colombia, Costa Rica, Cuba, Mexico, Panama, and Venezuela), the total country value was multiplied by the fraction of total country area comprised by the provinces, states, or departments in which the reef sites included in Cramer et al.'s database were located. Because the composition of reef sites varied across time bins, these calculations were performed separately for each bin. Reef sites claimed by a continental country but located on small islands well (over 70 km) offshore (San Andrés, Providencia, and Santa Catalina archipelago of Colombia, Corn Islands of Honduras, and Los Roques archipelago of Nicaragua) were assigned a fertiliser consumption value of zero. Because of the exceptionally high rate of agricultural fertiliser usage and geographic extent of the United States and their inability to locate state-level fertiliser consumption data for the United States that were reported in a comparable manner to the country-level Food and Agriculture Organization data, Cramer et al. excluded Florida reefs from the drivers analyses.
Human population density data (people/km²) were obtained by country for 5-year increments from 1950 to 2010 (e.g., 1950, 1955, 1960, 1965, 1970, 1975, 1980, 1985, 1990, 1995, 2000, 2005, and 2010) from the United Nations Population Division. Quinquennial or annual values were averaged to compute density values for each of the time bins considered in Cramer et al.'s analyses.
To ensure adequate sample size and comparable spatial extent of data on Acropora dominance through time and to separate the pre- and post-White Band Disease periods (late 1970s to early 1980s) and pre- and post-Diadema die-off periods (circa 1984), Cramer et al. computed coral dominance and driver values within four time bins: 1950–1969, 1970–1984, 1985–1994, and 1995–2011. (Acropora presence and dominance data were too sparse to conduct separate analyses for each of these countries or time bins or to partition data into finer-resolution time bins. Separate binomial generalised linear mixed-effects models were formulated to predict Acropora palmata dominance at the reef crest zone and Acropora cervicornis at the midslope zone as a function of (i) all four of the potential drivers, (ii) time bin, and (iii) country. While individual potential drivers were included as fixed effects, time bin and country were included as random effects to account for temporal autocorrelation and the uneven geographic distribution of samples across time bins, respectively.
To determine the models that best described patterns of change in Acropora palmata dominance at the reef crest zone and Acropora cervicornis dominance at the midslope zones since 1950, Cramer et al. (i) ran an initial 'full' model that included all four potential fixed effects and both random effects, (ii) inspected the significance of each fixed effect, and (iii) ran a 'final' model that included significant fixed effects and both random effects. In the case that a fixed effect was found to be 'nearly significant', R² values (in statistics the proportion of the variance in the dependent variable that is predictable from an independent variable) were compared across models that included and excluded this effect to determine the best-fit model. Cramer et al. inspected both the marginal R² (variance explained by fixed effects) and conditional R² (variance explained by the entire model, including fixed and random effects). Before running the generalised linear mixed-effects models analyses, the distributions of all four continuous predictor variables were log transformed to reduce the influence of extreme large values and improve model convergence. Linear correlations between potential drivers were assessed by conducting Spearman rank correlation tests. Because of the lack of discernable mechanistic causes for any correlative relationships between potential drivers and our interest in assessing the effects of each of these stressors, all four drivers were included as predictor variables in the initial models.
For the time series and drivers analyses, model performance was assessed via diagnostic plots of model residuals (quantile-quantile plots, pooled residuals versus predicted values, and residuals of random and all significant fixed effects versus predicted values) and via goodness-of-fit tests on pooled residuals (uniformity, outliers, and dispersion). Diagnostic plots and goodness-of-fit tests were produced for each mode, Tests were carried out via a simulation-based approach that transformed model residuals to a standardized scale. All statistical analyses were performed using the program R version 3.4.
Presence and dominance data for Acropora cervicornis and Acropora palmata were compiled from 2459 reef sites from 27 countries for the reef crest zone and 5185 reef sites from 30 countries for the midslope zone. Each of the 12 time bins contained data from a broad geographic area that spanned the greater Caribbean, including sites from the Lesser Antilles, Greater Antilles, Gulf of Mexico, Florida, and mainland coast of Central or South America. The number of reef sites exceeded 100 in both reef zones within each time bin except for the Pleistocene, 1500–1949, and 1950–1959 periods. Two time bins, 1500–1949 and 1950–1959, contained the fewest number of reef sites at both reef zones and accordingly contained the largest uncertainty values.
At the reef crest zone, the model-estimated percent of reef sites dominated by the Elkhorn Coral Acropora palmata declined from the Pleistocene to the present, from 78 to 6% of sites. Wald’s Z tests (a way to find out if explanatory variables in a model are significant) indicated that the first significant decline from baseline dominance values in the Pleistocene occurred in the 1950s, by which point, Acropora palmata was dominant at only 49% of reef sites. Dominance levels for Acropora palmata remained significantly lower than Pleistocene values from the 1950s to the present, although dominance increased to 60% in the 1970s. During the 1980–1984 time period, the percent of sites with Acropora palmata dominance shrunk to only 52% of the number of sites it dominated in the Pleistocene. The impact of initial White Band Disease outbreaks in the late 1970s/ early 1980s and the Diadema die-off in the early 1980s was reflected in the continued decline in dominance of Acropora palmata from 41 to 23% of sites between the 1980–1984 and 1985–1989 time periods. The Staghorn Coral Acropora cervicornis experienced a coincident but less marked decline in dominance at the reef crest zone. The percentage of sites with Acropora cervicornis dominance declined significantly from Pleistocene values beginning in the 1960s (from 18 to 7% of sites) and remained significantly lower than Pleistocene levels from that point on. By the time of the Diadema die-off, the percent of reef sites with Acropora cervicornis dominating the reef crest zone had already declined to only 3%.
Acropora cervicornis was the most abundant species at the midslope zone across the Caribbean during the prehuman and historical periods, dominating 63% of sites in the Pleistocene. Wald’s Z tests indicated that presence and dominance of Acropora cervicornis at the midslope zone declined significantly between the Pleistocene and Holocene periods. The second significant decline from the baseline dominance values in the Pleistocene occurred in the 1960s, by which time the species was dominant at only 12% of sites. By the time of the Diadema die-off in the early 1980s, Acropora cervicornis dominated only 4% of reefs at the midslope zone and did not significantly decline post-Diadema die-off. This species dominated the midslope zone at less than 1% of the 3293 reef sites included in the most recent time bin. Acropora palmata dominance at the midslope zone significantly declined from 4 to under 1% of sites from the Pleistocene to the early 1990s and has since remained significantly lower (less than 1% of sites) than Pleistocene values.
When Acropora presence and dominance values from the historical period (1500–1949 AD) were considered the baseline, results were similar to analyses that considered the Pleistocene period as the baseline. The first significant declines in Acropora palmata presence and dominance in the reef crest zone from the baseline period 1500–1949 occurred in the 1960s and 1950s, respectively (compared with the 1950s when the Pleistocene period was considered the baseline), and the first significant declines in Acropora cervicornis presence and dominance in the midslope zone from the baseline period 1500–1949 occurred in the 1950s (compared with the 1950s for Acropora cervicornis presence and the 1960s for Acropora cervicornis dominance when the Pleistocene period was considered the baseline).
Because of insufficient or incompatible data for one or more potential drivers, the analysis of potential drivers of change in Acropora dominance since 1950 included a subset of the countries and reef sites that were included in the time series. Relating declines in Acropora dominance since 1950 to a suite of potential disturbances suggested that local human stressors played a significant role in pre-White Band Disease declines of Acropora palmata. At the reef crest, Human population density had a significant negative effect on Acropora palmata dominance, and fertiliser consumption had a nearly significant positive effect on Acropora palmata dominance. While the generalised linear mixed-effects models (including country and time bin as random effects and population density and fertiliser consumption as fixed effects) explained 30% of the variance in Acropora palmata dominance, the fixed effects alone only explained 7% of the variance in Acropora palmata dominance. At the midslope zone, none of the fixed effects were found to have a significant effect on Acropora cervicornis dominance.
Comparisons of country-level temporal trends in potential drivers since 1950 revealed differing patterns across variables. In the datasets for Acropora palmata dominance at the reef crest and Acropora cervicornis dominance at the midslope zone, population density increased across all four time periods in each country, while values of the other three potential drivers fluctuated over time. Degree heating months wer highest within the 1970–1984 and 1995–2011 time periods for most countries, reflecting major El Niño–Southern Oscillation events in 1982–1983 and 1997–1998, respectively. Fertiliser consumption increased in most countries until the 1985–1994 or 1995–2011 time periods. Hurricanes per year fluctuated across the earliest three time periods, from 1950 to 1994, and reached peak values in most countries during the 1995–2011 period. The analysis of relationships among potential drivers revealed that reef locations are exposed to varying combinations of stressors and that reefs with high exposure to one or more stressors have a low degree of exposure to others. In both the reef crest and midslope zone datasets, all drivers were significantly correlated with one another although many of these correlations have no obvious ecological explanation.
The timeline of change in Acropora presence and dominance across the Caribbean from the prehuman period to the present revealed that the dominance of Acropora (Acropora palmata at reef crest and Acropora cervicornis at midslope) began to decline significantly by the 1950s and 1960s, predating the first recorded instance of White Band Disease by 10 to 30 years, the Diadema die-off by about 25 to 35 years, and large-scale Coral bleaching epidemics by 20 to 40 years. This pattern holds whether the Pleistocene period, Pleistocene/Holocene periods combined, or historical period (1500–1949 AD) are treated as the baseline time bin. Our long-term dataset shows that by the time the earliest widespread Coral bleaching occurred in the Caribbean in the late 1980s, Acropora corals were relatively rare in the Caribbean: The proportion of sites dominated by Acropora palmata at the reef crest had already declined from 78 to 22%, and the proportion of sites dominated by Acropora cervicornis at the midslope had already declined from 63 to 4% since the Pleistocene. Acropora presence and dominance values briefly increased in the 1970s and were particularly notable for Acropora palmata. This period also encompasses the first reported instances of White Band Disease and initial pronounced anthropogenic ocean warming in the Caribbean, precluding any obvious ecological explanation for Acropora recovery.
The analysis of potential drivers of Acropora loss for which reliable long-term data exist, temperature stress, fertiliser consumption for agriculture (as a proxy for eutrophication), hurricanes, and population density, suggests that local Human impacts may have played a role in the high levels of Acropora mortality that occurred decades before White Band Disease and widespread Coral bleaching. The strong negative effect of Human population density on Acropora palmata dominance at the reef crest zone since 1950 could implicate either land-based pollution or fishing effects, the importance of now-depleted herbivorous reef Fish to Coral health is well established, but there are no reliable proxies of fishing effort or reef Fish abundance at the broad temporal and spatial scales explored by Cramer et al. The weak positive effect of fertiliser consumption on Acropora palmata dominance at the reef crest was likely related to lower Human presence in agricultural areas, as values of fertiliser consumption were greatest at low levels of Human population density in the dataset. In contrast, neither population density nor fertiliser consumption was a significant predictor of Acropora cervicornis dominance in the midslope zone, suggesting that other factors played a role in the historical decline of this Coral.
Cramer et al. did not find evidence that hurricanes and ocean warming were responsible for the initial decline in Acropora dominance that occurred across the Caribbean in the mid-20th century. Earlier work has linked anthropogenic ocean warming to outbreaks of White Band Disease that caused massive die-offs of Acropora beginning in the late 1980s/early 1990s; however, the analyses upon which these conclusions were based could not identify earlier causes of initial decline, because they did not include pre-White Band Disease abundance or dominance data. Cramer et al.'s long record of Acropora dominance since the pre-Human period reveals that early Acropora declines predate region-wide Coral disease outbreaks, indicating that Coral populations across the Caribbean were substantially altered before catastrophic climate change impacts. The lack of association between declines in Acropora over the past half-century and hurricane exposure may be because storm events assist in the primary mode of Acropora reproduction, asexual propagation via colony fragmentation. It is possible that the recent lack of Acropora recovery following hurricanes is related to loss of herbivory on reefs rather than the hurricanes themselves, an analysis of total living Coral cover on reefs across the Caribbean since 1970 found that Coral cover was not related to hurricane exposure until after the die-off of the keystone herbivore Diadema Urchin.
Coral declines in the Caribbean have commonly been attributed, in part, to declining water quality due to inputs of land-based sediments and pollutants. Water turbidity has almost certainly increased across the Caribbean as agricultural and industrial activities have increased over the past century, but this trend has gone undetected due to the lack of established water quality monitoring programs for reef environments. The only long-term water clarity surveys reported for the Caribbean, from Belize and Puerto Rico, revealed a significant increase in turbidity from 1993 to 2012 attributed to land-based runoff. The earlier decline to close to 0% dominance of Acropora cervicornis at the midslope (early 1980s) compared with Acropora palmata at the reef crest (early 2000s) and the failure of
