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Convergent mortality responses of Caribbean coral species to seawater warming T. B. SMITH,1, M. E. BRANDT,1 J. M. CALNAN,1 R. S. NEMETH,1 J. BLONDEAU,1,3 E. KADISON,1 M. TAYLOR,1 AND P. ROTHENBERGER2 1Center for Marine and Environmental Studies, University of the Virgin Islands, 2 John Brewers Bay, St. Thomas, US Virgin Islands 00802 USA 2East End Marine Park Office, Division of Coastal Zone Management, Department of Planning and Natural Resources, 45 Estate Mars Hill, Frederiksted, St. Croix, US Virgin Islands 00840 USA Citation: Smith, T. B., M. E. Brandt, J. M. Calnan, R. S. Nemeth, J. Blondeau, E. Kadison, M. Taylor, and P. Rothenberger. 2013. Convergent mortality responses of Caribbean coral species to seawater warming. Ecosphere 4(7):87. http://dx.doi. org/10.1890/ES13-00107.1 Abstract. Species-specific responses to disturbance are a central consideration for predicting the composition, dynamics, and function of future communities. These responses may be predictable based on species traits that can be analyzed systematically to understand those characteristics important in determining susceptibility and potential for recovery. Scleractinian coral communities of the Western Atlantic are experiencing increased frequency and severity of extreme thermal disturbance, coral bleaching, and mortality. A conceptual thermal bleaching response model developed in this study suggests multiple susceptibility pathways that can lead corals to partial mortality and the loss of biomass, or complete mortality and the loss of genotypes, with implications for species-specific persistence and recovery. Coral assessments from annual to semi-annual surveys at 18 sites in the U.S. Virgin Islands, northeastern Caribbean Sea, before, during, and after the catastrophic 2005 coral bleaching event and during the mild 2010 bleaching event were used to evaluate bleaching, disease, and mortality responses. Three convergent groupings of species emerged based predominantly on their responses to the 2005 event: Type I—high bleaching and initial mortality, no subsequent white disease, and severe losses of cover (exhibited by Agaricia agaricites and branching Porites species); Type II—moderate bleaching and initial mortality, high subsequent white disease prevalence, and severe losses of cover (exhibited by Colpophyllia natans, Montastraea annularis species complex, and M. annularis sensu stricto); Type III—moderate to low bleaching and paling, low to no subsequent white disease, and low to no loss of cover (exhibited by Diploria strigosa, Montastraea cavernosa, Porites astreoides, and Siderastrea siderea). Whole colony mortality was uncommon, even in the most susceptible species, suggesting a potential for recovery among the majority (19 of 27) of scleractinian corals studied. Type II species performed worse than predicted by species traits because of their susceptibility to disease, a factor that needs to be incorporated more fully in models of thermal stress response. Responses of all species to the milder 2010 event were less severe, with limited bleaching and no detectible mortality. Future community composition of Caribbean coral reefs under seawater warming will likely be increasingly dominated by resistant Type III species. Key words: Caribbean bleaching event 2005; coral bleaching; coral disease; partial mortality; recovery; resistance; species traits; thermal stress; tolerance; United States Virgin Islands. Received 26 March 2013; revised 31 May 2013; accepted 3 June 2013; published 26 July 2013. Corresponding Editor: D. P. C. Peters. Copyright:  2013 Smith et al. This is an open-access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited. http://creativecommons.org/licenses/by/3.0/ 3 Present address: Southeast Fisheries Science Center, National Oceanic and Atmospheric Administration, 75 Virginia Beach Drive, Miami, Florida 33149 USA.  E-mail: tsmith@uvi.edu v www.esajournals.org 1 July 2013 v Volume 4(7) v Article 87 INTRODUCTION Disturbance affects communities by changing the relative abundance of constituent species, the strength of their interactions, and the feedbacks between organisms and the environment. There- fore, predicting the responses of individual species will be central to predicting the emergent properties of reorganized communities. This is particularly true for foundational species that, by virtue of their structure and function, have an inordinate influence on community and ecosys- tem processes and stability (Dayton 1972, Ellison et al. 2005). Light-dependent scleractinian corals are the foundational species of coral reefs, providing physical structure that influences most community processes. Thus, understanding the individual responses of coral species and their patterns of recovery during and after disturbance will increase our understanding of how coral reef community and ecosystem properties will also change with disturbance. Thermally induced mass coral bleaching has become one of the most critical disturbances to coral reefs (Hoegh-Guldberg 1999, Baker et al. 2008) and may stimulate increasing global degradation of coral reefs even under optimistic global carbon emission and coral adaptation scenarios (Frieler et al. 2012). The major effects of mass coral bleaching are shifts in community structure of corals and their symbionts and, therefore, the ecological function of coral ecosys- tems (Loya et al. 2001, Brandt 2009). The bleaching response defined here is related not only to the visual manifestation of loss of symbionts, loss of chlorophyll in symbionts, or acute effects of mortality in bleached corals, but also the delayed mortality that occurs in corals during recovery, long after the apparent visual evidence of bleaching has disappeared (Glynn 1996). Short-term shifts in community structure will be guided by species-specific resistance to disturbance, as seen in patterns of mortality during the bleaching response (Marshall and Baird 2000, McClanahan 2000, Loya et al. 2001, McClanahan et al. 2001, Baird and Marshall 2002), while long-term shifts in community structure will be controlled by both species- specific resistance during disturbance and recov- ery patterns after disturbance has passed (McCla- nahan 2008, van Woesik et al. 2011). In any given environment, species or community persistence over multiple disturbances may be due to endogenous processes where resistance is en- hanced by partial colony survival that favors regrowth and both sexual and asexual reproduc- tion. In addition, persistence may be favored by exogenous processes, such as the supply of larvae from connected populations. Within a community, patterns of thermal stress response among species are governed by biolog- ical characteristics of the coral holobiont and secondary interactions between corals and other organisms or the environment. These aspects of the thermal stress response vary among species and can result in the species-specific patterns of bleaching, mortality, persistence, and recovery. The thermal susceptibility of the coral animal, apart from its relationship with Symbiodinium, is not well understood. Most non-symbiotic organ- isms have much higher thermal tolerance limits than the coral symbioses, suggesting that thermal susceptibility is strongly related to the partial or complete breakdown of symbiosis (Baird et al. 2009). However, although Indo-Pacific coral species do demonstrate a pattern whereby increased bleaching severity and subsequent mortality have been shown to be related (Baird and Marshall 2002), the relationship can be nuanced, with some species bleaching and showing low mortality and other species not bleaching but showing high mortality (McClana- han 2004). The latter species may be directly susceptible because of host vulnerability to thermal stress. In practice, host and holobiont responses are hard to separate in the context of species-specific responses to high thermal stress and in most cases the thermal response can be considered integrated among host and symbiont. In contrast to host susceptibility to thermal stress, the susceptibility of coral symbiosis to bleaching is highly related to numerous biolog- ical traits of corals (see Gates and Edmunds 1999, Baker et al. 2008, van Woesik et al. 2012). Since bleaching is typically a light mediated response to thermal stress (Lesser 1997), factors that reduce light or heat stress, or increase resistance to light and heat stress can buffer the effects of v www.esajournals.org 2 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. thermal stress. Both Baker et al. (2008) and van Woesik et al. (2012) summarized known colony traits that imparted higher tolerance in a ther- mally fluctuating environment, including massive morphology (versus branching mor- phology), thick tissues, large inter-corallite spac- ing, low growth rates, large colony size, large corallite size, association with Symbiodinium Clade D, and a porous skeletal structure. Each of these characteristics can impart resistance to bleaching and also resistance to thermally in- duced mortality. This may include limiting mortality on colonies to partial, rather than whole colony (genet). For example, perforate corals (e.g., Porites) have intra-skeletal refuges of tissue that may survive over severe thermal stress and regrow after stress has subsided (Baker et al. 2008). Also, heterogeneous symbiont associations have been shown to protect coral species that can support thermally resistant symbiont types (e.g., Symbiodinium Clade D) as long as these symbionts are in high abundance within the coral colony at the time of thermal stress (Baker 2001, Baker et al. 2004). In the Western Atlantic, large faviids have been shown to host a diversity of symbiont clades (Rowan et al. 1997), including Clade D (LaJeunesse 2002). Colonies possessing unequal distributions of symbionts with differing susceptibilities to ther- mal stress may show intra-colony variation of bleaching, favoring partial mortality over total colony mortality (Baker et al. 2008). Although there is strong evidence to support traits that impart thermal resistance, when colony traits conflict in their potential for resistance to thermal stress (e.g., branching Porites species with perfo- rate skeletons), it is unclear a priori whether a species will show high or low resistance. Thermally stressed and/or bleached corals are in a precarious position for maintaining their energy balance. Lowered densities of zooxan- thellae lead to decreasing photosynthesis to respiration ratios and the possibility of declining energy reserves to maintain metabolism (Porter et al. 1989, Anthony et al. 2007, Rodrigues and Grottoli 2007). For some Western Atlantic coral species there is a positive relationship between colony energy and resistance to mortality among conspecifics (Thornhill et al. 2011), and this may apply more broadly between coral species (van Woesik et al. 2012). If the latter is true, then species that can maintain higher energy stores may be more successful at resisting mortality if bleached, while those species that tend to fall into negative energy states may be at an increased risk of starvation (Grottoli et al. 2004). Positive, or less negative, energy states could be favored through greater species-specific tissue mass ratios or increased capability for heterotrophic subsidy during thermal stress (Grottoli et al. 2006). A stressed and starved coral may also show increased susceptibility to secondary agents of stress. Abiotic stressors such as physical damage, excess nutrients, and toxins, which may normally be tolerated, can have greater impacts on corals already thermally stressed. For example, dis- solved inorganic nutrients may cause cascading effects that increase the severity of bleaching (Wooldridge 2009a) by decreasing the availability of dissolved inorganic carbon to symbionts (Wooldridge 2009b) and by increasing symbiont densities prior to bleaching, which may increase reactive oxygen production at the initiation of thermal stress (Cunning and Baker 2012). Sec- ondary biological stressors, such as predation, competition, parasitism, and disease, may also have greater effects on corals that are already thermally stressed, particularly if the effects of thermal stress cause a weakening of coral defense mechanisms. Recently there have been an increasing number of studies demonstrating a link between thermal stress and coral disease. For example, thermal stress was related to increased white diseases prevalence on the Great Barrier Reef when coral cover was high (Bruno et al. 2007). In addition, Muller et al. (2008) showed that Caribbean acroporids had increased disease prevalence when bleached or thermally stressed, and had increased disease severity when corals were bleached. Furthermore, Brandt and McManus (2009) showed that Caribbean Montastraea faveo- lata colonies with higher bleaching extent had higher incidence of white plague disease after thermal stress, Siderastrea siderea with dark spots disease later had a greater extent of bleaching, and Colpophyllia natans showed an association between the incidence of black band disease and bleaching. Thus, thermal stress can increase the prevalence and extent of disease in corals; however, because corals show species-specific v www.esajournals.org 3 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. patterns of susceptibility to various diseases (Calnan et al. 2008), increased susceptibility to disease with thermal stress may also have species-specific patterns. Recovery from disturbance may also show species-to-species variability. Recovery after ther- mal disturbance, bleaching, and mortality is facilitated by resistance to the disturbance and rapid reestablishment within the environment. Biological processes that might favor recovery include broadcast spawning (Glynn and Colley 2008) and high recruit density, strong potential for regrowth after partial mortality, and high colony growth rates (Darling et al. 2012, van Woesik et al. 2012). In the Indo-Pacific, rapid recovery has been seen in many species with these characteristics. For example, recovery from 4% to 47% coral cover ten years after thermal disturbance on a reef flat in Sesoko Island, Japan was led by resistant massive and encrusting species, resistant small (,4 cm), fast-growing acroporids, and the nonresistant but fast asexu- ally recruiting species Montipora digitata (van Woesik et al. 2011). In the Western Atlantic, recovery after mortality has tended to be slower, even in coral colonies that persisted through disturbance (Rogers and Miller 2006). This may relate to a higher influence of chronic secondary stressors (e.g., land-based sources of pollution and overfishing) that interact with acute distur- bance in the Western Atlantic (Connell 1997), and the changing demographics of dominant reef building species, such as Montastraea annularis (Edmunds and Elahi 2007). However, some species (e.g., sub-massive Porites astreoides and massive Diploria strigosa) have shown high resistance to thermal and storm disturbances and in some locations have been steadily increasing in abundance over the last few decades (Green et al. 2008, Edmunds 2010). The growing body of literature on coral’s response to thermal stress makes it clear that thermal stress and bleaching can impact multiple aspects of coral physiology and survival. We propose a conceptual thermal bleaching stress response model that can be separated into four physiological stress components encompassing host susceptibility, holobiont susceptibility (bleaching), susceptibility to starvation, and susceptibility to secondary stressors (Fig. 1). In the model, the components of physiological stress can increase susceptibility to other compo- nents within the same vertical column. For example, increased thermal susceptibility leading to increased bleaching can increase susceptibility to disease, and vice versa (Brandt and McManus 2009). The severity of the thermal response in any physiological component can also lead to partial mortality (the loss of coral biomass) or genet mortality (the loss of biomass and genotype), with consequences for recovery. This model provides a conceptual framework for under- standing how coral species with their respective traits respond to thermal stress and highlights different pathways that can lead to different outcomes of physiological stress, partial mortal- ity, and genet mortality. High thermal stress and coral bleaching events affected the northeastern Caribbean in 2005 and 2010 and provided the opportunity to examine species-specific responses of corals to thermal stress. These events had contrasting signatures in the United States Virgin Islands. The year 2005 was the most severe high sea surface temperature (SST) event on record for the northeastern Caribbean (Eakin et al. 2010). In the Virgin Islands, satellite SST records registered a peak of 10.25 degree heating weeks (DHW) and a period of approximately 59 days above the local bleaching threshold of 29.58C (Aug. 20–Oct. 18); a level of thermal stress accumulation associated with severe coral bleaching and some mortality (NOAA 2012). The warm season of 2010 started as warm or warmer than 2005, with the bleaching threshold surpassed for 21 days between August 12 and September 2. In a clear example of ameliorative storm cooling (Manzello et al. 2007), the passing of the storm center of Hurricane Earl on August 30th, approximately 100 km to the northeast of the St. Thomas-St. John, caused a rapid decline in SST’s below the bleaching threshold to 29.38C. Then from October 5–8, the passage of Hurricane Otto caused windy and cloudy weather that further reduced SST below 29.18C (see Fig. 1 in Brandt et al. 2013). Total DHW accumulated in 2010 began to decrease after the beginning of October, when it had reached 5.1 DHW; a level associated with some bleaching and limited mortality. This study presents six years of bleaching response and recovery data for 27 Western Atlantic coral species over catastrophic 2005 v www.esajournals.org 4 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. and mild 2010 thermal stress events. The responses of Western Atlantic coral species to large magnitude thermal stress events, particu- larly those as strong as 2005, are not well documented. In addition, this study investigated multiple aspects of colony response, including partial mortality and disease; at a resolution able to capture the major proximate drivers of coral cover change and recovery. Our specific aim was to examine the empirical response of coral species across these thermal disturbances to determine if species traits predicted response pathways that predisposed taxa to greater susceptibility, greater resistance, or greater re- covery. We hypothesized that in 2005 all species would show severe bleaching (.50% preva- lence), but that branching and foliose species would be more highly affected. We also hypoth- esized that species-specific patterns of coral bleaching, disease, and partial mortality could be characterized by the thermal bleaching stress response model and would be strong predictors of cover loss within species. Those species with higher physiological susceptibility at any level of the model or with interactions between compo- nents of physiological susceptibility should be more predisposed to partial or total mortality. The potential for recovery among Western Atlantic corals was also investigated using both empirical observations and mortality character- istics that may promote future resilience. Lastly, we examined species groupings or trait group- ings that can be used to predict the response of these or similar communities to future distur- bance and used this information to assess which Caribbean species may be ‘‘winners’’ or ‘‘losers’’ under the influence of increasing thermal stress events. METHODS The U.S. Virgin consists of three major islands, St. Croix, St. John, and St. Thomas, surrounded by shelves with widths varying from 0.2–25 km (Fig. 2). St. Thomas and St. John lie adjacent on the wide Puerto Rican shelf, whereas St. Croix Fig. 1. A conceptual model of the thermal bleaching stress response in any given environment with respect to corals, coral species, and community change. The negative impacts at the community level increase from left to right. Thermal stress is a function of susceptibility at four levels, including thermal stress on the coral animal, the bleaching response, starvation, and secondary agents that affect survival. Stress can lead to partial mortality and the loss of cover, but retention of extant genotypes. Partial mortality can include whole colony mortality if the colony is part of a larger genet that is not linked by tissue. In the most severe case, genet mortality can cause the loss of genotypes. Recovery is governed by the speed and completeness of return in each stress and mortality component, and becomes increasingly difficult with loss of biomass and genotypes. v www.esajournals.org 5 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Fig. 2. Study sites around St. Croix, St. Thomas, and St. John, U.S. Virgin Islands. Numbers in parentheses are the mean site depths in meters. Sites marked with an asterisk were not sampled during the 2010 thermal stress event and were not used in statistical comparisons for 2010 bleaching prevalence and extent. Yellow box in inset indicates the U.S. Virgin Islands in relation to the eastern Caribbean. Gray shaded bathymetry indicates seafloor deeper than 65 m. v www.esajournals.org 6 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. lies on its own narrow platform about 60 km to the south and separated by the Anegada Passage with depths greater than 4000 m. A total of 18 sites on all three major islands of the U.S. Virgin Islands were surveyed annually or semi-annually between the years 2004 to 2010 as part of the Territorial Coral Reef Monitoring Program (TCRMP; Fig. 2). Here we consider sites less than 25 m in depth in an effort to understand the response of species within the ocean surface mixed layer of the northeastern Caribbean. In addition, sites dominated by acroporid corals were not assessed, but their response to the 2005 bleaching event in the US Virgin Islands has been reported elsewhere (Muller et al. 2008). Each site consisted of six permanent 10 m long transects marked by steel rods and separated by 3–5 m distance. However, prior to 2007 transects at seven sites in St. Thomas were laid in random directions with the same spacing and in the same area of reef as where permanent transects were established in 2007. These sites were Black Point, Buck Island (St. Thomas), Flat Cay, Savana Island, Seahorse Cottage Shoal, and South Capella. It is likely that this had the effect of increasing the year-to-year variability in coral cover estimates from prior to bleaching through 2006, and decreasing variability after 2007. However, there were no significant differences in species-specific coral cover from 2006 to 2007, with the exception of Colpophyllia natans. This suggested that increased variability from random transects caused no spurious differences in year- to-year coral cover that was unrelated to the bleaching event; however, C. natans cover pat- terns between these periods should be treated with some caution. Sampling consisted of direct visual observation of coral health and video- based recording of benthic cover (methods described below). Sites were assessed at least once as a baseline prior to the 2005 mass bleaching event (2004– May 2005), during the mass bleaching event (September–December 2005), in early recovery (health sampling only; January–May 2006), after bleaching (June–December 2006), annually be- tween 2007 and 2009, and then during the 2010 bleaching event (subset of 12 sites, Fig. 2). The 2010 bleaching event was relatively minor in severity; thus, only data from the height of the event are reported to investigate the differences in the bleaching and paling response among species between 2005 and 2010. Coral cover Video sampling was conducted along 10 m transects following standard methodologies (Rogers and Miller 2001, Smith et al. 2008). A SCUBA diver swam at a uniform speed (3–5 min. per transect), pointing a video camera perpen- dicular to the substrata and following the vertical contour of the reef at the approximate height of a guide wand, 0.4 m in length. Non-overlapping captured images represented a planar area of reef approximately 0.31 m2 (0.64 m 3 0.48 m). Ten randomly located points were superimposed on each captured image and the benthic cover under each point was then identified to the lowest identifiable taxonomic level and used in calcula- tion of percent cover. For coral species in the Montastraea annularis species complex (M. annu- laris, M. faveolata, and M. franksi) there was difficulty in assigning species, particularly after the 2005 bleaching event where partial mortality often led to small tissue patches without discern- able taxonomic characteristics. This was not usually the case with M. annularis, whose macro-morphology tended to remain obvious. Therefore, two taxonomic categories of the M. annularis species complex are considered in this study: M. annularis and the Montastraea annularis species complex (MX), which hereafter refers to a combination of M. faveolata and M. franksi, with the likelihood of small numbers of amorphous M. annularis. This lumping of M. faveolata, M. franksi, and small M. annularis could mask some of the variation in individual species responses, such as susceptibility to disease. In addition, branching poritids, including Porites divaricata, Porites furca- ta, and Porites porites were combined under the species designation branching Porites spp. to avoid confusion in species assignment. However, from in situ observations on health transects, 92.4% of branching poritids were identified as P. porites. Coral health All scleractinian coral and hydrocoral colonies greater than 10 cm in largest diameter and intersected by the transect line were assessed in situ. Prior to 2008, only colonies greater than 10 cm in largest diameter were evaluated. Starting v www.esajournals.org 7 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. in 2008, all colonies, regardless of size were recorded, although the analyses were con- strained to colonies greater than 10 cm to match the pre-2007 data set. This resulted in a sample size of 9,989 coral colonies. Colonies were assessed for signs of bleaching, disease, and mortality following a modified Atlantic and Gulf Rapid Reef Assessment protocol (Kramer et al. 2005, Calnan et al. 2008, Smith et al. 2008). Variables measured were colony longest hori- zontal length (including dead colony portions), bleaching (prevalence and extent), white diseases (prevalence), recent partial and total mortality (prevalence), and old partial mortality (preva- lence and extent). Prevalence represents the number of individuals in the population affected, while extent represents the proportion of the colony affected and takes into account only affected individuals. Abnormal lightening of the colony color was assessed by experienced observers and catego- rized as bleaching (colony areas with a stark white appearance), or paling. Visual estimates of bleaching have been used extensively in the literature, with good correlations to symbiont chlorophyll concentration, and order of magni- tude reductions in symbiont density from normal conditions to those indicated by stark or skeletal white coloration (McClanahan et al. 2007). When bleaching or paling was present on a colony, the extent (proportion) of white or paling tissue on the colony affected was estimated. For colonies that had both bleached and pale tissue, the colony was scored as bleached for prevalence calculations and only the extent of the bleached tissues was used in comparisons, giving greater weight to the most severe signs of poor coral health. Some level of low prevalence and low extent bleaching and paling (combined preva- lence ,20%, or 40% for S. siderea) was common in the focal species before the bleaching event in the period prior to bleaching and in periods that preceded the study (T. Smith, unpublished data). Although this was not specifically controlled for, bleaching and paling associated with thermal stress in 2005 and 2010 was typically many times greater than background bleaching. Diseases were conservatively categorized into recognized Western Atlantic scleractinian diseas- es and syndromes (following Bruckner 2007). Only the diseases characterized as white disease, including white plague, were examined in this study as these had disproportionate impacts on coral tissue loss after bleaching (Miller et al. 2009). We use the more general term of white disease (Bythell et al. 2004) as it is not clear that the etiologies were specifically related to charac- terized white plague types in all cases. In general, we defined white disease lesion signs as multi- focal to coalescing areas of recent tissue loss that appeared to have originated basally or peripher- ally on colonies and proceeded as a solid to discontinuous band (Appendix B). Background white disease levels were not specifically con- trolled for, as levels of white disease in the focal species at the study sites from 2002–2004 (0.5%, N ¼ 2018 colonies) were many times lower than seen in our study (T. Smith, unpublished data). Recent mortality was defined as areas of skeleton recently denuded of living tissue with very little to no filamentous algae colonization on non-eroded coral skeleton. Old mortality was identified as areas of the colony with no living tissue with thick turf or macroalgae growth on top of eroded coral skeleton. Complete colony mortality was scored where there was evidence of recent mortality and no indication of any living tissue anywhere on the colony. Data analysis The prevalence of bleaching and paling, and the change of coral cover were assessed for all coral species sampled and presented as summary statistics. Nine focal scleractinian coral species with the highest abundances on the reefs sampled were chosen for comparative analysis because their sample size was amenable to parametric statistical comparisons. These species were Agaricia agaricites, Colpophyllia natans, Dip- loria strigosa, Montastraea annularis, Montastraea annularis species complex, Montastraea cavernosa, Porites astreoides, branching Porites spp., and Siderastrea siderea. In addition, these species represented a range of traits, facilitating evalua- tion of the performance of different biological characteristics during and after bleaching. For variables that represented chronic responses to the bleaching event and were therefore linked between time periods, such as old partial mortality and coral cover, a statistical model was used that incorporated the nine focal species and time. For variables that were acute and not v www.esajournals.org 8 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. necessarily linked across time periods, such as bleaching/paling, white disease, and recent mor- tality, the pattern among the nine focal species was assessed within a specific time period or using the mean values at the height of the response. Comparisons among species for the bleaching response were carried out at time periods that represented the maximum response during height of the bleaching event and, separately, for the first time period of recovery after the 2005 thermal stress event (Jan.–May 2006). Bleaching prevalence and extent was compared among the nine focal species for surveys conducted at the height of the 2005 thermal stress event. Bleaching among species was also tested at the height of the 2010 thermal stress event at a subset of 12 sites surveyed between August and November. Paling was tested within the 2005 and 2010 thermal stress events, when they had the potential to co- vary with the bleaching response, and an additional comparison was made among species in the 2005 bleaching recovery period (Jan.–May 2006). When significant differences were found among species, the difference between species was compared with a Tukey’s HSD post-hoc test. With the exception of principal components analysis, all statistical comparisons were con- ducted in JMP v9 (SAS Institute). The prevalence of white disease and recent mortality among coral species was tested in the two sampling periods just subsequent to bleach- ing in 2005. The purpose was to determine the peak prevalence among species of white disease or recent mortality stimulated by catastrophic bleaching in 2005. For each species the peak prevalence of white disease and recent mortality was taken from the early 2006 sampling period (Jan.–May) or the late 2006 period (Jun.–Dec.), whichever value was higher for the species. Species were compared with a Kruskal-Wallis test due to violations of parametric analysis and pair wise differences between species were then tested with a Tukey’s HSD post-hoc test. For white disease only the six focal species that exhibited these disease signs were used to compare peak prevalence (C. natans, M. annularis, M. annularis species complex, M. cavernosa, P. astreoides, and S. siderea). Within each species the temporal pattern of recent partial mortality was also examined with individual Kruskal-Wallis tests followed by Tukey’s HSD post-hoc tests. White diseases were not tested over time because of frequent periods with zero values, and the comparisons were made with visual inspection of the temporal pattern. Old partial mortality was tested among the nine focal species and time periods using a repeated-measures analysis of variance (RM- ANOVA). Pairwise post-hoc comparisons be- tween species and time periods were performed using a Tukey’s HSD test. Coral cover was tested among the nine focal species and time periods using a RM-ANOVA after an arcsine transformation. The pattern of individual species over time was then tested with a separate RM-ANOVA with a Bonferroni ad- justment procedure or a Kruskal-Wallis test when data did not meet assumptions for parametric analysis. Post-hoc comparisons among times were performed using a Tukey’s HSD test. The change in coral colony abundance caused by the 2005 bleaching event was tested for each species. The abundance of colonies greater than 10 cm maximum length was calculated from coral health intercept transects prior to bleaching in 2005 and after the cessation of the response to bleaching in 2007. The relative change was calculated for each species on each transect and used to test the hypothesis that the change in abundance was significantly different than zero using a Wilcoxon test on ranks. Live tissue size frequencies of the focal coral species were compared among length size categories; small (10–15 cm), medium (16–30 cm), and large (.30 cm) and the years 2005, 2007, and 2010. Data was not available on multiple size dimensions (length, width, height) prior to the 2005 bleaching event and a reliable estimate of the actual total and living area of a colony could not be calculated. Instead, live tissue lengths were calculated as the total colony length, inclusive of dead colony portions, multiplied by the percent of colony living (1 minus the extent of partial mortality). Since the extent of partial mortality is an areal measure and live length is linear measure, the live length calculated in this study provided a mutually consistent, but not exact, estimate of longest length of living tissue. As a caveat, this estimate of live tissue length underestimates the true live length for colonies with a high degree of partial mortality, since, all v www.esajournals.org 9 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. other things equal, the percent area decreases by the square while the percent length decreases linearly. Thus, there may have been an overesti- mation of the downward size class transition frequencies. As a check against undue bias, the direct change of maximum length (living and dead colony portions) was tested between the same time periods and results were robust for all species except MX (v ¼ 4.2, p ¼ 0.376). MX have very distinct individual skeletal structure and very dense skeletons that were readily identifi- able even after a high degree of partial mortality (see Appendix B), thus, they only showed size frequency changes when an estimate of live tissue length was used. Frequencies of live lengths in size classes were tested among time periods with a contingency table and v2 analyses. The species A. agaricites, C. natans, and P. astreoides had low expected frequency in some cells that should lead to cautious interpretation of the v2 analysis. For example, A. agaricites and C. natans both had no representative colonies in some size classes over some years. Relationships between the response variables for the 2005 and 2010 bleaching events among the focal coral species were explored with a correlation matrix. Principal component analysis (PCA) was then used to analyze the pattern of response to the 2005 bleaching event among the nine focal species. For each species the following variables were included in the analysis: (1) mean colony abundance change between 2005 and 2007, (2) mean coral cover change between 2005 and 2007, (3) peak of mean bleaching prevalence during the 2005–2006 event, (4) mean paling prevalence between January and May 2006 (recovery), (5) peak of mean disease prevalence following the 2005–2006 event, (6) peak of mean recent mortality prevalence in 2006, and (7) peak of mean old mortality prevalence following the 2005–2006 event. All data were normalized and the correlation matrix was used as the basis for the analysis using the PCA routine in the statistical software package PRIMER (v. 6). The ranking of coral species responses the 2005 bleaching event recorded in this study were compared to a trait-based ranking of coral genera responses to thermal stress (van Woesik et al. 2012). Thermal responses in van Woesik et al. (2012) were based on the consensus opinion of 10 coral reef scientists, who produced a list of eight coral traits associated with tolerance to thermal stress (Appendix A). Coral genera responses were evaluated by the same scientists against a hypothetical þ38C thermal stress anomaly during the solar insolation maximum. From the present study, coral cover change was chosen as the response metric as it is an integrated measure of coral response that is the most commonly used in studies of coral community disturbance. Scores for each trait and overall trait rankings of van Woesik et al. (2012) were used as published but evaluated against published quantitative infor- mation (Appendix A). Ties in rank within traits or cover loss were assigned the mean value of the rank slots that would have been taken by the tied ranks if arbitrarily assigned higher or lower (e.g., two species with a tie in ranks over slots 3 and 4 would have each been assigned a value of 3.5). The differences between hypothetical rankings from van Woesik et al. (2012) and empirical rankings of coral cover change from this study were used as the comparative metric. RESULTS The extreme thermal disturbance event in 2005 initiated a distinct sequence of coral responses across the shallow and intermediate depth coral reefs monitored from 2004 to 2010 (Fig. 3). Overall, the coral stress responses showed a lagged pattern; with a high mean prevalence of bleaching from September through December 2005 preceding the highest recorded mean prevalence of white diseases by three months to a year. The response of coral cover was closely associated with the onset of white diseases, suggesting that disease lesions and their associ- ated recent partial mortality were responsible for cover loss. Recent partial mortality spiked within three months to a year after bleaching, concur- rent with a 48% loss of relative coral cover, and the rapid and pervasive accumulation of signs of old partial mortality. These responses are evalu- ated below for the 27 individual species recorded and the nine focal species. Bleaching All of the nine focal coral species were affected by bleaching during the 2005 thermal stress event. The combined prevalence of stark white bleaching and paling on these species was v www.esajournals.org 10 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. greater than 70% during the bleaching period (Fig. 4). However, the prevalence of coral bleaching and paling during the height of the bleaching event and during the recovery period was significantly different among coral species (Table 1). During the height of the bleaching event, Agaricia agaricites had the highest preva- lence of stark white bleaching (90%) and the highest extent of bleaching (92%). Paling preva- lence was relatively low in A. agaricites largely because most colonies were almost entirely bleached. With paling included, only 3% of sampled A. agaricites colonies showed no bleach- ing or paling. Branching Porites spp., which had a high prevalence of bleached colonies (78%) at a high extent (87%), with 11% of prevalence of colonies showing no bleaching or paling. Non- focal species with fewer sampled colonies that also showed a strong response to the 2005 thermal stress event included Agaricia grahamae, Isophyllia spp., Madracis decactis, Madracis mirablis, Millepora complanata, Mycetophyllia spp. and Stephanoceonia intercepta (Table 2). Species showing an intermediate bleaching response during the 2005 thermal stress event included Colpophyllia natans, Siderastrea siderea, and the two taxonomic divisions of the Montas- traea annularis species complex (Fig. 4). Bleaching Fig. 3. Summary of total community response for all scleractinian coral species to bleaching across sites and time periods. Upper panel, the prevalence of the coral stress responses stark white bleaching (left-axis) and white diseases (right-axis). Lower panel, the prevalence of recent and old partial mortality (left-axis) and total hard coral cover (right-axis). Error bars are 6SE. v www.esajournals.org 11 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. prevalence for these species was between 55% and 70%, but when bleaching was present it affected the majority of the colony (bleaching extent .75%). Few colonies of C. natans, M. annularis, M. annularis species complex, and S. siderea escaped impact, as the number of colonies showing no bleaching or paling was 14%, 18%, 8%, and 14%, respectively. A non-focal species Fig. 4. The prevalence and extent of coral bleaching and paling (6SE) for nine focal coral species across eight periods spanning 2004–2010. MX ¼ Montastraea annularis species complex. v www.esajournals.org 12 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. that also showed an intermediate response was Diploria labyrinthiformis (Table 2). The focal species with the lowest levels of bleaching in 2005 were D. strigosa, M. cavernosa, and P. astreoides (Fig. 4). The species D. strigosa and P. astreoides had a similar response, with about half of colonies bleached to an extent of 66% of tissue surface, and with a relatively large proportion of colonies that showed no bleaching or paling (;30%). M. cavernosa had fewer colonies that showed no bleaching or paling during the height of bleaching (18%), and among Table 1. Statistical comparison of the prevalence and extent of bleaching among nine focal species during height of bleaching in 2005. Test F/v2 df Prob. AA CN DS MA MX MC PA BP SS Prevalence BL 2005 38.8 8 ,0.0001 A ABC BC ABC AB C BC AB BC PA 2005 30.3 8 0.0002 B AB AB AB AB A B B AB BL 2006 32.8 8 ,0.0001 A B B B B B B B B PA 2006 87.1 8 ,0.0001 C BC ABC AB A C C C C BL 2010 13.0 8/337 ,0.0001 AB AB CD A BC D D BC BC PA 2010 5.9 8/337 ,0.0001 C BC ABC ABC AB ABC C C A Extent BL 2005 18.4 8/476 ,0.0001 A AB BC AB A C C AB AB PA 2005 5.6 8/256 ,0.0001 AB AB AB AB B A AB AB A BL 2006 4.0 8/57 0.0011 A AB AB AB AB AB AB B B PA 2006 2.8 8/283 0.0050 AB AB AB A A AB AB AB B BL 2010 2.1 8/273 0.0358 A A A A A A A A AB PA 2010 6.6 8/274 ,0.0001 ABC ABC ABC AB BC BC C AB A Notes: Test code abbreviations are: BL ¼ bleached, PA ¼ pale. Species code abbreviations are: AA ¼ Agaricia agaricites, CN ¼ Colpophyllia natans, DS ¼ Diploria strigosa, MA ¼ Montastraea annularis, MX ¼ Montastraea annularis species complex, MC ¼ Montastraea cavernosa, PA ¼ Porites astreoides, BP ¼ branching Porites species, SS ¼ Siderastrea siderea.  Wilcoxon v2 statistic for non-parametric comparisons. Table 2. Response summaries of 27 stony coral species for mean prevalence of bleaching and paling during the height of the 2005 bleaching event (Sep.–Dec. 2005) and the 2010 bleaching event (Sep.–Oct. 2010). Species 2005 BL % (SE, N) 2005 PA % (SE) 2010 BL % (SE, N) 2010 PA % (SE) Agaricia agaricites 89.8 (4.5, 36) 7.0 (3.5) 66.6 (10.6, 18) 18 (8.5) Agaricia grahamae 78.5 (14.8, 7) 7.1 (7.1) 0 (0, 2) 100 (0) Agaricia lamarcki ... ... 59 (14.7, 11) 9 (9) Colpophyllia natans 57.1 (13.7, 14) 28.5 (12.5) 50 (15, 11) 9 (9) Dendrogyra cylindrus 33.3 (33.3, 3) 66.6 (33.3) 0 (0, 3) 0 (0) Dichocoenia stokesii 0 (0, 3) 66.6 (33.3) ... ... Diploria clivosa 11.1 (11.1, 3) 66.6 (33.3) 0 (0, 0) 0 (0) Diploria labyrinthiformis 61.1 (11.8, 18) 27.7 (10.8) 16.6 (16.6, 6) 33.3 (21) Diploria strigosa 52.5 (11.1, 20) 17.5 (8.3) 0 (0, 13) 30.7 (12) Eusmilia fastigiata 33.3 (33.3, 3) 66.6 (33.3) 16.6 (16.6, 6) 33.3 (21) Helioseris cucullata 0 (0, 1) 100 (0) 0 (0, 1) 100 (0) I. sinuosa/I. rigida 100 (0, 3) 0 (0) ... ... Madracis decactis 100 (0, 5) 0 (0) 0 (0, 9) 22.2 (12.1) Madracis mirablis 50 (50, 2) 0 (0) 0 (0, 1) 100 (0) Meandrina meandrites 0 (0, 13) 25.6 (12) 0 (0, 9) 5.5 (5.5) Millepora alcicornis 63 (12.3, 14) 9.5 (7.3) 9 (6.2, 22) 9 (6.2) Millepora complanata 53.3 (22.6, 5) 6.6 (6.6) ... ... Montastraea annularis 65.7 (9.8, 21) 16.4 (6.5) 58 (6.5, 35) 27.8 (6.1) Montastraea annularis spp. cpx. 65.6 (4.6, 72) 26.6 (3.8) 34.7 (4.2, 56) 46.5 (4.7) Montastraea cavernosa 38.1 (6.1, 53) 43.3 (6.2) 3.8 (2.3, 48) 31 (6.2) Mycetophyllia spp. 100 (0, 1) 0 (0) 8.3 (8.3, 6) 0 (0) Porites astreoides 50.7 (6, 52) 18.9 (4.8) 10.7 (3.1, 61) 21.3 (4.1) Branching Porites spp. 78.3 (6.2, 32) 10.1 (4.2) 35.2 (6.4, 40) 12.6 (4.1) Scolymia spp. 0 (0, 1) 100 (0) ... ... Siderastrea siderea 58.9 (7, 45) 26.6 (6.4) 36.3 (5.5, 51) 51.8 (5.5) Stephanocoenia intercepta 100 (0, 3) 0 (0) 5.8 (5.8, 17) 52.9 (12.4) Notes: Abbreviations are: BL ¼ bleached; PA ¼ paling; I. sinuosa ¼ Isophyllia sinuosa; I. rigida ¼ Isophyllastrea rigida; Cpx. ¼ complex. An ellipsis (...) indicates no data. v www.esajournals.org 13 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. the focal species it exhibited the lowest preva- lence of bleaching (38%), but the highest preva- lence of paling (43%). Some non-focal species that were also less responsive to bleaching included Dendrogyra cylindrus, Diploria clivosa, Dichocoenia stokesii, Eusmilia fastigiata, Helioseris cuculata, Meandrina meandrites, and Scolymia spp. (Table 2). In the recovery period from January to May 2006 there were also species-specific differences in prevalence of bleaching and paling that corresponded to prolonged signs of stress. The phenomena of ‘‘apparent recovery’’ may have affected species that showed low prevalence of bleaching or paling in the recovery period. In this case tissues or colonies that were severely bleached may have been more prone to mortality during and just after the bleaching event, leading to fewer bleached colonies and tissues in the recovery period. This is a caveat to interpreting the data for species that appeared to recover more rapidly but had a high prevalence of recent partial mortality, such as branching Porites species. In contrast, the highly bleached species Agaricia agaricites maintained a significantly higher prevalence of bleaching into the recovery period, indicating a failure of colonies to recover their symbionts (Fig. 4, Table 1). In addition, the two large faviids M. annularis and the M. annularis species complex both exhibited a prolonged recovery period with significantly higher prevalence of paling into early 2006. Species-specific bleaching and paling respons- es to the 2010 thermal stress event showed some differences and similarities to the 2005 event. In 2010 all species had a prevalence of bleaching or paling that was elevated above background bleaching levels in non-bleaching periods (prior to 2005, 2007–2009). However, the extent of bleaching and paling on colonies was low and similar among species (Table 1), indicating spotty and partial bleaching, in contrast to 2005 when the extent of bleaching was severe and affected the majority of the colony. In 2010, the most severely affected species were A. agaricites, C. natans, and M. annularis, in contrast to 2005, when A. agaricites and branching Porites spp. were the most severely affected. Also in contrast to 2005, the M. annularis species complex and S. siderea were lightly affected in 2010, with a moderate prevalence of bleaching and paling at a low extent. The most bleaching resistant species in 2005 were also the most resistant in 2010; D strigosa, M. cavernosa, and P. astreoides showed a moderate prevalence of paling at a low extent, but no apparent increase in bleaching prevalence. White disease Identifiable white diseases after the 2005 bleaching event only affected six of the nine focal species within transects, including C. natans, M. annularis, M. annularis species complex, M. cavernosa, P. astreoides and S. siderea (Fig. 5). White disease also affected an additional three of 18 non-focal species (Table 3), although sample sizes were low. The prevalence of white diseases on affected species in 2006 was the highest recorded for any period at these sites between 2001 and 2012 (Calnan et al. 2008, Smith et al. 2008). The massive faviids, C. natans and M. annularis species complex tended to have the greatest prevalence of white disease in 2006 (Fig. 5, Table 4). However, the occurrence of bleaching or paling that was common for these species in 2006 (delayed recovery) was not related to the occurrence of white disease on a colony basis (contingency analysis, n ¼ 912, v2 ¼ 2.2, p ¼ 0.323), suggesting that recovery of algal pigmen- tation did not prevent the later development of disease signs. Although not recorded as suscep- tible to white disease in sampled transects, marked D. strigosa colonies at a subset of five of the sites assessed in a separate project showed signs of white disease between Dec. 2005 and Apr. 2006 (N ¼ 25, prevalence ¼ 8.0%). Of note is the fact that no cases of white disease were recorded during the bleaching period in 2005 on any species of coral. Partial mortality Recent and old partial colony mortality showed a strong response to the 2005 bleaching event with a pattern that differed among species. Recent partial mortality varied across time among and within species (Table 4, Table 5), lagging behind bleaching and peaking in 2006 for some species (Fig. 6). In general, recent partial mortality patterns followed patterns of bleaching severity across species, with the highest preva- lence in A. agaricites and branching Porites spp. Recent partial mortality also occurred earlier in A. agaricites, starting in the 2005 bleaching period and peaking in Feb.–Apr. 2006 (Fig. 6). However, v www.esajournals.org 14 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. few A. agaricites were sampled in the later part of 2006 (N ¼ 13), as the abundance of colonies greater than 10 cm in diameter had already declined by 81%. Other species that had signif- icantly elevated recent partial mortality after the peak of bleaching in 2005 included; C. natans, M. annularis species complex, M. cavernosa, and S. siderea, whereas there was no significant eleva- Fig. 5. The prevalence of white disease (6SE) for nine focal coral species across eight periods spanning 2004– 2010. Species with no data indicate that disease was not seen in those species. MX ¼ Montastraea annularis species complex. v www.esajournals.org 15 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. tion of recent partial mortality in the species D. strigosa, M. annularis, and P. astreoides (Fig. 6). The prevalence of old partial mortality had the same general pattern among all coral species. The pattern of old partial mortality was not signifi- cantly different among species, as indicated by a non-significant species 3 time interaction term (Table 6). Therefore, post hoc comparisons are shown separately among species and time in Table 6. Old partial mortality increased signifi- cantly and rapidly with time starting in early 2006, following the height of bleaching in 2005 and the onset of white diseases and recent mortality (Fig. 7). The prevalence of old partial mortality continued to build into 2007 followed by a leveling of prevalence in 2007 and 2008. By 2009 the prevalence of old partial mortality on all species began to decline; however, prevalence values were still much higher than pre-distur- bance levels. Although the pattern of old partial mortality was similar among species over time, species such as M. annularis and the M. annularis species complex had the highest overall preva- lence, while species A. agaricites, D. strigosa, and P. astreoides had lower overall prevalence (Table 6). Complete colony mortality that occurred dur- ing and after bleaching was moderate to low in Table 3. Response summaries of 27 stony coral species for mean prevalence of peak white disease (2006), percentage of total colony mortality (bleaching period to 2007), coral cover before bleaching in 2005, and the change in coral cover after the 2005 bleaching event (bleaching period to 2007). Species M N WD % (SE, N) Mortality % 2005 cover % Cover change % Agaricia agaricites F 434 0 (0, 22) 15.9 0.7 91.8 Agaricia grahamae F 35 0 (0, 22) 8.0 0.0 100.0 Agaricia lamarcki F 31 0 (0, 2) 0.0 0.004 87.0 Colpophyllia natans M 154 17.5 (8.3, 20) 6.7 0.2 81.1 Dendrogyra cylindrus C 26 0 (0, 5) 7.1 0.04 22.0 Dichocoenia stokesii M 14 100 (0, 2) 0.0 0.02 25.8 Diploria clivosa M 34 0 (0, 2) 0.0 0.1 55.3 Diploria labyrinthiformis M 119 5.8 (5.8, 17) 5.3 0.3 73.5 Diploria strigosa M 284 0 (0, 20) 2.0 0.4 35.2 Eusmilia fastigiata B 44 0 (0, 5) 0.0 0.02 5.9 Helioseris cucullata P/S 5 0 (0, 3) 0.0 0.0 100.0 I. sinuosa /I. rigida M 4 0 (0, 1) 0.0 0.0 14.6 Madracis decactis N 92 0 (0, 4) 0.0 0.01 14.2 Madracis mirablis B 43 12.5 (12.5, 8) 3.7 0.3 96.1 Meandrina meandrites Sm 141 0 (0, 11) 0.0 0.1 21.9 Millepora alcicornis B 269 0 (0, 20) 0.0 0.3 16.8 Millepora complanata VP 90 0 (0, 2) 3.7 0.2 76.3 Montastraea annularis M 1074 4.9 (3, 37) 0.0 2.6 60.4 Montastraea annularis spp. complex M 2704 16.6 (3.2, 82) 0.8 7.2 58.7 Montastraea cavernosa M 1140 2.5 (1.7, 65) 0.4 1.2 2.8 Mycetophyllia spp. P/S 15 0 (0, 1) ... 0.1 60.0 Porites astreoides M 1428 4.3 (2.5, 57) 0.7 1.5 14.6 Branching Porites spp. B 610 0 (0, 36) 3.9 1.2 63.3 Scolymia spp. S 9 ... 0.0 0.0 ... Siderastrea siderea M 900 3.7 (2.5, 54) 0.0 0.7 24.4 Stephanocoenia intercepta M 139 0 (0, 8) 0.0 0.03 34.2 Notes: Abbreviations are: M ¼morphology; N¼number of colonies; Mortality¼total colony mortality; Cover¼coral cover; I. sinuosa¼Isophyllia sinuosa; I. rigida¼Isophyllastrea rigida; B¼branching; C¼columnar; F¼Foliose; M¼massive; N¼Nodular; P/ S ¼ plating/sheeting; S ¼ solitary; Sm ¼ Sub-massive; and VP ¼ vertically plating. An ellipsis (...) indicates no data. Table 4. Statistical comparison of the peak prevalence of white disease and recent partial mortality among coral species in 2006 during recovery from bleaching in 2005. For white disease only data for affected species were tested. Test v2 df Prob. AA CN DS MA MX MC PA BP SS White disease 38.3 5 ,0.0001 ... AB ... AB A B B ... B Recent partial mortality 123.0 8 ,0.0001 A BCD E CDE C DE E AB DE Notes: Species codes as in Table 1. An ellipsis (...) indicates white disease did not occur for these species on transects. v www.esajournals.org 16 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. frequency, despite the severity of the bleaching response for some species, the strong increase in partial mortality, and the frequent health moni- toring after bleaching in 2005 (Table 3). Total colony mortality ranged from 0 to 15.9% for all species, with the highest loss in A. agaricites followed by Millepora complanata. Branching Porites spp. had low total colony mortality (3.9%) although they were highly susceptible to bleaching and partial mortality. Coral cover, colony abundance, and size frequencies Coral cover declined after the 2005 coral bleaching event in most of the focal species but remained unchanged in D. strigosa, P. astreoides, and S. siderea (Fig. 8, Table 7). In addition, M. cavernosa had a declining pattern of coral cover that was not significantly different after Bonfer- roni adjustment of p-values and appeared to be unrelated to the 2005 bleaching event. For those species that lost coral cover there was little indication of recovery in the five years following the mortality related to 2005 bleaching, with the possible exception of a slight increase in cover for the M. annularis species complex. The majority of non-focal species also showed losses of coral cover over the 2005 bleaching event (Table 3); however, these species had very low coral cover initially and were represented by a small number of transects, therefore change values should be treated with caution. From observations in 2011, the mild 2010 bleaching event did not have any appreciable effect on coral cover at these study reefs ( T. Smith, unpublished data). The change in abundance for colonies greater than 10 cm among the focal species was variable (Fig. 9, Table 8). The only species to have a significant loss in abundance was A. agaricites, while four species showed a significant increase in abundance (M. annularis species complex, M. cavernosa, P. astreoides, and S. siderea). The remainder of the focal species had no detectible change in abundance, including species that showed large decreases in coral cover (C. natans, M. annularis, M. annularis species complex, and branching Porites spp.). The live length size frequencies of most focal species declined between 2005, 2007, and 2010 (Table 9, Fig. 10). For most species, colonies in the 10–15 cm size class increased while corals greater than 30 cm decreased. While some of the change may have resulted from recruitment into the smallest size classes, a more parsimonious explanation is that living tissue on large colonies regressed in size due to partial mortality. As with other indications of response to the 2005 bleach- ing event, A. agaricites was the most negatively affected species, with a loss of all colonies in the largest size class; however, the M. annularis species complex, M. cavernosa, and S. siderea also lost a large proportion of colonies in the largest size class. Only branching Porites spp. had live length size frequencies that did not significantly change among years. Comparison of the coral species response The species to species correlations among the major bleaching response variables showed strong relationships between many variables (Table 10). The among species 2005 bleaching prevalence was highly negatively correlated with the 2005 paling prevalence and the change in abundance (higher bleaching was related to a decrease in abundance), and highly positively correlated with the peak of recent mortality in 2006. Coral cover change and abundance change were highly positively correlated, such that species that lost the most cover also lost the greatest number of colonies. Cover and abun- dance change were also highly negatively corre- lated with peak of recent mortality in 2006, indicating high recent mortality was associated with loss of tissue and colonies. The 2010 bleaching prevalence was positively correlated with 2005 bleaching prevalence as well as the Table 5. Statistical comparison of recent partial mortality prevalence across sampling time periods for individual species. Source v 2 df Prob. Total 222.9 7 ,0.0001 AA 84.0 7 0.0001 CN 27.4 7 0.0003 DS 6.4 7 0.490 MA 9.0 7 0.256 MX 134.3 7 ,0.0001 MC 21.3 7 0.0034 PA 12.1 7 0.100 BP 66.5 7 ,0.0001 SS 24.1 7 0.0011 Notes: Post hoc comparisons shown in Fig. 6. Species codes as in Table 1. v www.esajournals.org 17 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. peak prevalence of old mortality after 2005. Additionally, 2010 bleaching prevalence was negatively correlated with the abundance change after 2005 bleaching; thus, species that responded negatively to the 2005 bleaching event tended to be more susceptible to bleaching in 2010. The exception was S. siderea, which had relatively high bleaching in 2010 but showed no cover loss Fig. 6. The prevalence of recent partial mortality (6SE) for nine focal coral species across eight periods spanning 2004–2010. Post-hoc comparisons are Tukey’s HSD across time periods. MX ¼ Montastraea annularis species complex. v www.esajournals.org 18 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. and an increase in colony abundance after 2005. The first two axes of the PCA were the only axes with eigenvalues greater than 1 and together these axes comprised 81% of the variation in the multivariate dataset, while axes 3, 4 and 5 comprised the remaining variation (Table 11). Therefore, axes 1 and 2 adequately characterized the multivariate space, justifying the presentation of just these two axes (Fig. 11). Along the first axis, the eigenvector values for coral cover change and coral abundance change were of the largest magnitudes and were comparable but opposite to bleaching prevalence and recent mortality (e.g., in those species where coral cover change was negative, indicating a loss of coral cover, bleaching prevalence was higher in 2010; Table 10). When plotted along this first axis, there is a clear separation between those species that experienced high bleaching and high levels of recent mortality along with larger declines in coral cover (M. annularis, M. annularis species complex, C. natans, branching Porites spp. and A. agaricites) and those species less affected by bleaching and recent mortality which experi- enced little to no change in densities and cover (P. astreoides, M. cavernosa, S. siderea, and D. strigosa) (Fig. 11). In addition, A. agaricites stood farther out to the right on axis 1 because this species also had a large decline in abundance. In axis 2, the eigenvector values were far greater for disease prevalence and paling than for any other factor (Table 11). Along this axis, three distinct groups of species can be discerned: (1) A. agaricites and branching Porites spp. which experienced no disease and low post-event paling, (2) P. astreoides, M. cavernosa, S. siderea, and D. strigosa which experienced some disease and some post-event paling, and (3) C. natans, M. annularis, and M. annularis species complex which experienced high levels of disease and high levels of paling after the bleaching event. Thus, we can produce roughly three groupings of species corresponding to the response vari- ables: Type I species had large initial response during bleaching that included high bleaching prevalence, high recent mortality, and a large decline in coral cover (A. agaricites and branching Porites species); Type II species had a high to moderate initial response of bleaching preva- lence, followed by delayed recovery of normal coloration, high disease prevalence, high recent mortality, a large decline in coral cover, and stable or increasing colony abundance (C. natans, M. annularis, and the M. annularis species complex); and Type III species had a moderate initial response of bleaching prevalence, low to no disease increase after bleaching, low to no increase in recent mortality, low to no decline in coral cover, and stable or increasing colony abundance (D. strigosa, M. cavernosa, P. astreoides, and S. siderea). Comparison of hypothetical responses based on species traits (Appendix A) showed that many of the focal species had different empirical response (cover change) ranks (Table 12). Over performing species, where coral cover loss rank was less than the trait grouping rank, included the Type III species P. astreoides and S. siderea. Underperforming species included the Type II species C. natans, M. annularis, and M. annularis spp. complex. DISCUSSION Species performance over the 2005 bleaching event As predicted, the responses of corals to the 2005 high thermal stress and mass coral reef bleaching event were highly species-specific, ranging from high tissue mortality and decreas- ing adult abundance to undetectable loss of cover and colonies. Bleaching of corals during 2010 was Table 6. Statistical comparison of the prevalence of old partial mortality between the nine focal coral species and eight time periods. Source F df Prob. JP/MA BL/MX 2006a/BP 2006b/MC 2007/CM 2008/SS 2009/AA 2010/DS PA Time 20.8 7/45 ,0.0001 D D C BC A A B BC Species 5.14 4/51 0.0015 A AB ABC BC BC C D D D Species 3 Time 1.26 28/164 0.1861 Notes: Abbreviations are: JP ¼ just prior to bleaching in 2004 and 2005; BL ¼ 2005 bleaching event; 2006a ¼ Jan.–May 2006; 2006b ¼ Jun.–Dec. 2006. Tukey’s HSD post hoc comparisons are shown between species (combined periods) and between time periods (combined species). Species codes as in Table 1. v www.esajournals.org 19 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. much less severe than in 2005, but also showed species-specific rankings of susceptibility. These results underscore the species-specific variability seen in the thermal stress and bleaching respons- es in studies from the Indo-Pacific (Marshall and Baird 2000, Loya et al. 2001, Baird and Marshall 2002, McClanahan 2004, van Woesik et al. 2011), eastern Pacific (Glynn et al. 2001), and Western Atlantic (Lasker et al. 1984, Gates 1990, Lang et al. 1992, McField 1999, Oxenford et al. 2008, Villamizar et al. 2008, Brandt 2009). However, this study showed that species characteristics favored different outcomes to thermal stress reflected in patterns of bleaching, disease, and Fig. 7. The prevalence of old partial mortality (6SE) for nine focal coral species across eight periods spanning 2004–2010. MX ¼ Montastraea annularis species complex. v www.esajournals.org 20 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. partial to total mortality. These patterns con- verged to the three types of response: Type I species, with high bleaching and initial mortality, no subsequent white disease, and severe losses of cover; Type II species, with moderate bleaching and initial mortality, high subsequent white disease prevalence, and severe losses of cover; and Type III species, with moderate to low bleaching and paling, low to no subsequent white disease, and low to no loss of cover. Non-resistant Type I and Type II species whose cover decreased by half or more arrived at the loss of cover by two different routes. Type I species most susceptible to the visual manifesta- Fig. 8. Coral cover (6SE) for nine focal coral species across seven periods spanning 2004–2010. Post-hoc comparisons are Tukey’s HSD across time periods. Note the changing y-axis scale among the different rows. MX ¼ Montastraea annularis species complex. v www.esajournals.org 21 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. tion of coral bleaching, A. agaricites and branch- ing Porites spp., were also the most susceptible to early and high mortality during the peak of bleaching. This pattern has been noted previous- ly for Agaricia spp. in the Western Atlantic (Lasker et al. 1984). In contrast to Type II species, this mortality was not associated with white disease. Within the thermal bleaching response model, Type I species have high thermal and bleaching susceptibility that leads directly to partial or genet mortality (Fig. 1). As predicted by species traits, Type I A. agaricites and branching Porites spp. were the least tolerant to thermal stress. These species showed the lowest a priori tolerance scores based on traits and ranked among the species with the greatest cover loss in this study. Type I species were small foliose and branching species, mor- phological characteristics that are predicted to predispose corals to high mortality under ther- mal stress (Gates and Edmunds 1999, Baird and Marshall 2002, van Woesik et al. 2012). Branching and foliose morphologies may be a good trait predictor of susceptible Type I species, whilst other traits need further scrutiny. There is little background trait information available for A. agaricites; however, this species does seem to be slow growing (linear extension) relative to other Atlantic coral species (Huston 1985), and this may impart tolerance to thermal stress. It is likely that branching Porites spp., like Porites porites (Edmunds and Davies 1986), shows fast calcifi- cation rates relative to other species (Edmunds et al. 2011), potentially making them less tolerant. On the other hand, branching Porites spp. have traits associated with tolerance, including a perforate skeleton and deep skeletal tissue reservoirs, high biomass per area (Edmunds and Davies 1986, Edmunds et al. 2011), and the ability to feed heterotrophically to offset loss of Table 7. Statistical results of a repeated-measures ANOVA of coral cover among the nine focal coral species and seven time periods and the individual tests of coral cover within species over time. Source F/v2 df Prob. Among species Species 65.9 8/963 ,0.0001 Time 18.1 6/958 ,0.0001 Species 3 Time 6.1 48/4718 ,0.0001 Within species AA 113.1 6/748 ,0.0001 CN 21.3 6/748 0.0016 DS 1.7 6/748 0.115 MA 36.6 6/748 ,0.0001 MX 22.8 6/748 0.0009 MC 2.3 6/748 0.035 PA 1.8 6/748 0.104 BP 2.5 6/748 0.017 SS 0.7 6/748 0.664 Note: Species codes as in Table 1.  Indicates a non-parametric Kruskal-Wallis test. Fig. 9. The change in abundance (6SE) of focal species for colonies greater than 10 cm maximum length over the 2005 bleaching event. Periods were prior to bleaching (2004–2005) and after thermal stress mortality in 2007. Species codes in Table 1. v www.esajournals.org 22 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. energy (Seemann et al. 2012). These traits may have contributed to the low relative levels of total colony mortality and maintenance of colony abundance, even though the species experienced the third highest loss of cover. Thus, there may be different trait combinations in Type I species that lead to different degrees of tolerance to the more immediate effects of thermal stress. Type II species possess traits that may have made them more resistant to thermally induced pigment loss (van Woesik et al. 2012); however, they lost far more cover than was predicted a priori by species traits. Colpophyllia natans, M. annularis, and the M. annularis spp. complex are all large faviids, possessing the traits of tolerance: massive morphologies, thick tissues, large colony sizes, and low calcification rates. Furthermore, M. annularis and M. annularis spp. complex host Symbiodinium Clade D in higher abundances after thermal disturbance (Thornhill et al. 2006), whereas this has not been demonstrated for C. natans. Hosting Clade D may have contributed to greater heterogeneity in bleaching across colony surfaces (Rowan et al. 1997, Baker et al. 2008) (Appendix B), and greater levels of partial versus whole colony mortality. Yet resistance to direct bleaching induced mortality was negated to some degree by susceptibility to disease. During recovery from the 2005 bleaching event white disease was also noted on both C. natans and M. annularis in shallow marine protected areas of the U.S. Virgin Islands (see Miller et al. 2006, Whelan et al. 2007, Miller et al. 2009), but not in Barbados (Oxenford et al. 2008). In this study, the occur- rence of disease led to the relatively high losses of tissue as coral cover did not decrease until after the bleaching event and incidence of disease was not related to bleaching, paling, or normal pigmentation in the recovery period of early 2006. This suggests that disease incidence was not related to delayed recovery from bleaching, but was related to some aspect of the colony or community response during the height of the bleaching. The bleaching response pattern for Type II species suggested a path in the bleaching response model of progression from bleaching susceptibility to starvation susceptibility directly to susceptibility at the level of secondary agents (disease). We hypothesize that susceptibility to disease was increased due to negative energy states in the colonies as the result of the reduction of photosynthate transfer from Symbiodinium, followed by a depletion of colony tissue energy reserves. Montastraea annularis can lose half its energy content during bleaching stress (Porter et al. 1989) and, although tissue energy was not measured in this study, we can assume that severely bleached colonies in this study were similarly affected. Energy reserves can take many months to years to recover after bleaching (Rodrigues and Grottoli 2007), thus, even nor- mally pigmented colonies may have been sus- ceptible. The post-bleaching white disease outbreak has been suggested as appearing similar to White Plague Type II (Whelan et al. 2007, Miller et al. 2009); however, it was not clear if white disease signs indicated an infectious disease (Richardson et al. 1998), an environmen- Table 9. Size frequency analysis of focal coral species, with sample sizes for each period indicated. Species N 2005 N 2007 N 2010 v2 Prob. AA 61 14 46 30.9 ,0.0001 CN 14 15 17 15.4 0.0039 DS 37 43 31 12.8 0.012 MA 124 180 173 42.4 ,0.0001 MX 381 339 360 236.8 ,0.0001 MC 125 167 129 32.7 ,0.0001 PA 145 185 246 18.3 0.0011 BP 49 69 102 2.9 0.576 SS 70 113 146 55.1 ,0.0001 Notes: Contingency table analysis was conducted between years 2005, 2007, 2010 and the colony live tissue size classes 10–15 cm (small), 16–30 (medium), and .30 (large). Species codes as in Table 1.  Indicates size bins without corals and expected frequen- cies were ,5 in more than 20% of cells, suggesting a cautious interpretation of v2 results. Table 8. Statistical results of a Wilcoxon signed-ranks test for the hypothesis of relative colony abundance change greater than zero for the nine focal coral species between time periods prior to bleaching in 2005 and after the response to bleaching 2007. Source Signed-rank value df Prob. AA 286.5 42 ,0.0001 CN 19.5 26 0.559 DS 96.5 40 0123 MA 205.0 67 0.152 MX 389.5 89 0.044 MC 539.5 79 ,0.0001 PA 786.5 95 ,0.0001 BP 115.5 59 0.229 SS 433.0 69 0.0004 Note: Species codes as in Table 1. v www.esajournals.org 23 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Table 10. Correlations among focal coral species for major bleaching response variables. Measure 2005 BL 2006 PA Peak WD Peak recent Peak old Abundance change Cover change 2010 BL 2005 BL 1.000 2006 PA 0.212 1.000 Peak WD 0.146 0.446 1.000 Peak recent 0.900 0.022 0.044 1.000 Peak old 0.477 0.589 0.413 0.519 1.000 Abundance change 0.813 0.097 0.024 0.841 0.488 1.000 Cover change 0.689 0.337 0.268 0.751 0.606 0.817 1.000 2010 BL 0.704 0.418 0.322 0.628 0.758 0.774 0.679 1.000 Notes: Bleaching (BL) values represent mean prevalence for each species during the 2005 or the 2010 bleaching event, or paling (PA) prevalence between January and May 2006. Peak white disease (WD), peak recent mortality, peak old mortality represent the maximum average mean prevalence in 2006, after 2005 bleaching event. Abundance change and cover change represent change between prior to the 2005 bleaching event to 2007. Significant correlations are bolded (a ¼ 0.05). Fig. 10. Size frequencies of live tissue length for small (10-15 cm), medium (16-30 cm) and large (.30 cm) colonies of the focal coral species across years 2005, 2007, 2010. MX ¼ Montastraea annularis species complex. v www.esajournals.org 24 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. tal response (Lesser et al. 2007), or an epidemi- ological response that combined environmentally induced susceptibility and infectious transmis- sion (Bruno et al. 2007, Brandt et al. 2013). The susceptibility to disease increased the empirical response above that predicted by species traits, making Type II species that were apparently somewhat resistant to bleaching less tolerant to thermal stress overall. Models of species traits used to predict response to thermal disturbance should also investigate the possibility of traits that decrease tolerance to thermally linked diseases. Type III species were moderately resistant to bleaching, but tended to be very tolerant of thermal stress. This tolerance included less partial mortality, smaller reductions in size and cover, and low genet mortality relative to the other focal species. The low loss of cover indicated low susceptibility to thermal stress, bleaching, starvation, or secondary agent com- ponents of the thermal bleaching response model. Primary tolerance of thermal stress may also have prevented interactions with bleaching, starvation, and disease that would have led to greater partial or total mortality, as seen in Type II species. Type III species all share some characteristics that fit concepts of thermal toler- ance, including a massive to sub-massive mor- phology and high calcification rates. In addition, all of these species have some capacity for heterotrophy (Mills et al. 2004, Mills and Sebens 2004). However, other traits are not consistent among species. M. cavernosa and D. strigosa had the most traits associated with tolerance, with the exception that both possess an imperforate skeleton and D. strigosa has not been associated with Symbiodinium Clade D at a level perceived to confer tolerance. The empirical response of these two species was very close to that which was predicted by their trait rankings. On the other hand, P. astreoides and S. siderea over performed relative to predictions based on Fig. 11. Multivariate comparison of focal species’ response to 2005 bleaching event. Represented are the first two principal components of the nine species using the indices maximum 2005 bleaching prevalence, paling prevalence during the recovery period (Jan.– May 2006), coral cover change (prior to 2005 bleaching to 2007), coral abundance change (prior to 2005 bleaching to 2007), and the peak prevalence of white disease, recent partial mortality and old partial mortality after the 2005 bleaching period. Species codes as in Table 1. Table 11. Eigenvalues and percentage variation explained by the five principal components axes using coral condition and status indices for each of the nine species. Eigenvectors for each of the five measurement indices are also presented. Measure Principal component 1 2 3 4 5 Eigenvalues 3.91 1.74 0.635 0.339 0.22 Variation explained (%) 55.9 24.9 9.1 4.8 3.1 Cumulative variance explained (%) 55.9 80.8 89.9 94.7 97.9 Eigenvectors Cover change (2005–2007) 0.457 0.044 0.209 0.398 0.515 Abundance 0.452 0.201 0.17 0.241 0.215 Peak disease 0.099 0.61 0.655 0.063 0.417 Paling 2006 (recovery) 0.189 0.574 0.612 0.359 0.09 Peak old 0.376 0.35 0.151 0.759 0.304 Peak bleaching 2005 0.444 0.238 0.287 0.084 0.564 Peak recent 0.449 0.277 0.14 0.261 0.309 v www.esajournals.org 25 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. species traits, as they each possessed only half the traits considered to confer tolerance, but were still very tolerant of thermal stress. This was particularly true for P. astreoides, which had many traits that may make them more susceptible to thermal stress (e.g., small distance between corallites, small colonies, and a lack of Symbiodi- nium Clade D; van Woesik et al. 2012). Large distance between corallites, large colonies, and possession of Clade D are perceived to allow partial, versus total, mortality and increased energy storage. However, the perforate skeleton of P. astreoides and S. siderea may outweigh the lack of other traits that provide these mecha- nisms of tolerance by facilitating partial mortality and energy storage in deep skeletal tissue reservoirs. The influence of traits such as a perforate skeleton that cover a range of possible mechanisms for thermal tolerance may need additional weighting in trait based coral suscep- tibility models. Impacts of partial mortality Although the 2005 coral bleaching event was severe by multiple measures and unprecedented for the Caribbean (Eakin et al. 2010), the majority of cover loss among even the most severely affected coral species was the result of partial mortality rather than genet mortality. With the exception of A. agaricites and C. natans, seven of the nine focal species showed little change in abundance of colonies greater than 10 cm length and had few cases of complete colony mortality. Also, with the exception of branching Porites spp., all species after 2005 exhibited significant transitions of live length size frequencies away from large colonies to small colonies, suggesting a high-degree of partial mortality rather than total colony mortality (for examples see Appen- dix B). It also seems unlikely that in the five-year time frame of recovery in this study that there would have been any substantial sexual recruit- ment and growth of colonies into the .10cm size classes that would have been detectible in our study design. Increases in colony abundance in this study were therefore the result of fission, not sexual recruitment (Hughes and Jackson 1980, Loya et al. 2001). The impact of partial mortality on the resil- ience of Western Atlantic corals depends on how size class transitions affect overall demographic rates. Demographic feedbacks within adult mor- tality, fission, and downward size transitions that are not well represented in abundance changes can trump recruitment and lead to losses of colonies and cover (Hughes and Tanner 2000, Edmunds and Elahi 2007). For populations of M. annularis where demographic models have been applied, no reasonable amount of recruitment restored the largest size classes of corals and prevented ultimate population extinction when positive size transition probabilities were low, including a model parameterized with data from a study site within the US Virgin Islands (Edmunds and Elahi 2007). In addition, a demographic model parameterized with M. annularis population data from eastern Puerto Rico over the 2005 thermal stress event showed very low cover or extirpation with thermal stress events of 2005 level severity recurring at a frequency of less than every 17 years (Herna´n- dez-Pacheco et al. 2011). Consequently, partial mortality in large framework building corals may be a short-term step on a downward trend Table 12. Comparison on species performance rank predicted based on species traits (van Woesik et al. 2012) and empirical performance from this study based on cover loss. Species Trait score Trait rank Cover loss Cover loss rank Trait-cover rank difference Agaricia agaricites 4 8.5 92 9 0.5 Colpophyllia natans 2 5.5 81 8 2.5 Diploria strigosa 4 3 35 4 1 Montastraea annularis 4 3 60 5.5 2.5 Montastraea annularis spp. complex 4 3 60 5.5 2.5 Montastraea cavernosa 5 1 3 1 0 Porites astreoides 0 7 15 2 5 Branching Porites spp. 4 8.5 63 7 1.5 Siderastrea siderea 2 5.5 25 3 2.5 Note: Ties in rank were assigned the mean value of the rank slots that would have been taken by the tied ranks if arbitrarily assigned higher or lower. v www.esajournals.org 26 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. ending in low abundance or extirpation. This susceptibility may be increased if low cover inhibits positive feedbacks that facilitate coral recruitment and growth (Bozec et al. 2013). However, none of these models parameterized with empirical data included the ability for M. annularis populations to adapt to recurring stresses, which might occur with selection for resistant coral genotypes or adaptive bleaching (Buddemeier and Fautin 1993, Baker 2001). The latter involves the ability of corals to associate with different clades of algal endosymbionts (e.g., Clade D) with increased tolerance of thermal stress and resistance to bleaching. These considerations may be important in assessing the long-term potential for large faviids to endure repeated thermal stress. Regrowth of extant genotypes may speed the pace and scale of recovery (van Woesik et al. 2011). There has been only slight restoration of total coral cover in the five years after the thermal stress in 2005; however, by 2010 the M. annularis species complex, P. astreoides, and branching Porites spp. have shown evidence of increasing cover. Genotypes that survived the disturbance should buffer the genetic diversity within spe- cies, including the genotypes of corals that are centuries old. In addition, an increase in the relative abundance of resistant algal endosymbi- ont genotypes that may occur with adaptive bleaching might be preserved within the system for a window of time (Thornhill et al. 2006). In this case, surviving tissues possess a ‘‘memory’’ of high thermal disturbance that could lead to greater tolerance over future events, although the long-term costs of reversion to previous host- symbiont combinations or maladaptive combi- nations need to be considered (Jones 2008). Some of the species-specific differences in response to the 2005 and 2010 thermal stress events could be consistent with a shift to greater thermal toler- ance through selection of resistant coral geno- types and adaptive bleaching. For example, relative to other species and their response in 2005, the M. annularis species complex and branching Porites spp. had a lower bleaching response in 2010, and this could be taken as evidence of adaptation to thermal stress after 2005. However, colony-specific monitoring is necessary to determine the potential for adapta- tion when comparing two very different stress events, particularly without consideration of spatial variation in the stress (e.g., environmental variation around M. annularis versus M. annularis species complex dominated reefs). The susceptible Type I species A. agaricites provides a good example of the potential for partial mortality to influence long-term recovery. As the most thermally sensitive species, A. agaricites, had a 66% decrease of colonies greater than 10 cm length and the greatest prevalence of complete colony mortality (15.9%). Thus, 50% of colonies either transitioned to size classes smaller than 10 cm or suffered complete mortality. While some whole colony mortality was likely to have been missed in the sampling that occurred within the time frame that recent mortality transitioned to old mortality, approximately three months (Smith et al. 2008), much of the abundance decrease may have represented partial mortality and a transition to size classes less than 10 cm. Between 2008 and 2010, when all sizes of colonies were assessed, 80% of A. agaricites colonies were less than 10 cm in maximum diameter (N ¼ 499), indicating there was no shortage of small colonies that could represent colonies that sur- vived through 2005 and regressed in size. Despite the susceptibility of A. agaricites to the direct effects of thermal and bleaching stress, surviving colonies often had intra-colony refuges of tissue within the cryptic and shaded environ- ments where small tissue patches (,10 cm) often survived (Appendix B). The sheeting to foliose morphology of A. agaricites may create heteroge- neous light environments across the colony surface that act as intra-colony refuges from thermal-light induced bleaching. Observations suggest that these tissue patches are beginning to regrow out of these intra-colony refuges (Appendix B). A similar pattern of susceptibility and recovery can be seen in branching Porites spp. Within the expert ranking of physical and biological traits that favor tolerance and recovery conducted by van Woesik et al. (2012), A. agaricites and branching Porites spp. may be extremely intolerant of thermal stress (i.e., foliose morphology, high calcification rate, small size, low symbiont diversity, or non-porous skeleton), but resilient in the long-term because of high recovery potential (i.e., high recruit density, colony regrowth from surviving tissue, and high colony growth rates). Thus, these taxa may not v www.esajournals.org 27 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. be stress-resistant but could be classified as weedy due to a high capacity for recovery favored by brooding reproduction and high growth rates (Darling et al. 2012). Conclusions The results of this study support the concept that there will be winners and losers among coral species in response to a future of increasing frequency and intensity of thermal stress events. Many species traits predicted to influence in- creased tolerance to thermal stress performed well for Caribbean coral species over the 2005 stress event, with perforate skeleton seeming to have a relatively large positive influence on some Type III species. On the other hand, in Type II species some combinations of traits predicted for tolerance (massive morphology, thick tissues, large distance between corallites, large colony size, and the ability to host populations of Symbiodinium Clade D) were superseded by the susceptibility to disease. Large faviids seem very susceptible to long-term population declines because they fare poorly over the whole stress response, when bleaching, disease, and mortality are considered. In addition, M. annularis and the M. annularis spp. complex are likely less equipped for the recovery phase because they tend to grow slowly and have lower fecundity and greater susceptibility to mortality when small (Edmunds and Elahi 2007). While faviids may be considered ‘‘losers’’, Type III species are likely the clear ‘‘winners’’ because they are resistant to thermal stress and may be capable of recovery on short time scales when disturbed. Type I species are losers in the short-term, but may show long-term persistence due to their high potential for recovery. A key component of persistence, versus continued decline, for Type I species may be the return frequency of thermal disturbance. Continued monitoring of these communities will determine if initial recovery trajectories result in a restoration of community components prior to the next disturbance. We predict that future thermal stress events are likely to generate a community of diminutive, fast- growing colonies with higher proportions of Type III species. In summary, we found that there were species- specific differences in the bleaching response that corresponded to different traits among species and, as predicted, the most highly affected species possess branching and foliose colony morphologies. There was also a significant correlation between species-specific bleaching prevalence during the 2005 thermal stress event and subsequent recent partial mortality and cover change, indicating that bleaching responses and mortality were linked within species. Those species that had a higher prevalence of bleaching in 2005 also tended to have a higher prevalence of bleaching in 2010, with the exception of the M. annularis species complex and branching Porites spp., suggesting that for most species tolerance is similar over catastrophic and mild bleaching events, but that for some species there may have been limited adaption to thermal stress, warrant- ing further investigation. Finally, because of the preponderance of partial mortality among coral species, the ultimate resilience and future trajec- tory of reefs in the Caribbean will be intimately tied to the demographics of surviving, smaller colonies and the recurrence of future disturbance. ACKNOWLEDGMENTS We thank G. Bosire, L. Carr, G. Gentius, S. Herzleib, C. Loeffler, A. Paul, L. Allen-Requa for field assistance, S. Prosterman for diving support, and M. Henderson and C. Joseph for administrative support. We thank K. Baltzer, P. Edmunds, P. Glynn, L. Henderson, and J. Lang for helpful comments on the manuscript. This manuscript benefitted through interactions among the Future Reefs workgroup of the NSF National Center for Ecological Analysis and Synthesis, grant #EF- 0553768 to P. Edmunds and R. Gates. The Virgin Islands Department of Planning and Natural Resourc- es, Division of Coastal Zone Management and the NOAA Coral Reef Conservation Program funded this work. The Virgin Islands Experimental Program to Stimulate Competitive Research (NSF award # 346483 & 0814417) and the Lana Vento Charitable Trust provided equipment support. Views and opinions expressed in this manuscript do not reflect those of the funding agencies. This is contribution #89 from the Center for Marine and Environmental Studies, Uni- versity of the Virgin Islands. LITERATURE CITED Anthony, K. R. N., S. R. Connolly, and O. Hoegh- Guldberg. 2007. Bleaching, energetics, and coral mortality risk: Effects of temperature, light, and sediment regime. Limnology and Oceanography 52:716–726. v www.esajournals.org 28 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Baird, A. H., R. Bhagooli, P. J. Ralph, and S. Takahashi. 2009. Coral bleaching: the role of the host. Trends in Ecology & Evolution 24:16–20. Baird, A. H. and P. A. Marshall. 2002. Mortality, growth and reproduction in scleractinian corals following bleaching on the Great Barrier Reef. Marine Ecology Progress Series 237:133–141. Baker, A. 2001. Reef corals bleach to survive change. Nature 411:765–766. Baker, A. C., P. W. Glynn, and B. Riegl. 2008. Climate change and coral reef bleaching: an ecological assessment of long-term impacts, recovery trends and future outlook. Estuarine, Coastal and Shelf Science 80:435–471. Baker, A. C., C. J. Starger, T. R. Mcclanahan, and P. W. Glynn. 2004. Corals’ adaptive response to climate change. Nature 430:741. Bozec, Y.-M., L. Yakob, S. Bejarano, and P. J. Mumby. 2013. Reciprocal facilitation and non-linearity maintain habitat engineering on coral reefs. Oikos 122:428–440. Brandt, M. E. 2009. The effect of species and colony size on the bleaching response of reef-building corals in the Florida Keys during the 2005 mass bleaching event. Coral Reefs 28:911–924. Brandt, M. E. and J. W. McManus. 2009. Disease incidence is related to bleaching extent in reef- building corals. Ecology 90:2859–2867. Brandt, M. E., T. B. Smith, A. M. S. Correa, and R. Vega-Thurber. 2013. Disturbance driven colony fragmentation as a driver of a coral disease outbreak. PLoS ONE 8:e57164. Bruckner, A. W. 2007. Field guide to Western Atlantic coral diseases and other causes of coral mortality. NOAA, UNEP-WCMC, PADI. Bruno, J. F., E. R. Selig, K. S. Casey, C. A. Page, B. L. Willis, C. D. Harvell, H. Sweatman, and A. M. Melendy. 2007. Thermal stress and coral cover as drivers of coral disease outbreaks. PLoS Biology 5:e124. Buddemeier, R. W. and D. G. Fautin. 1993. Coral bleaching as an adaptive mechanism. BioScience 43:320–326. Bythell, J., O. Pantos, and L. L. Richardson. 2004. White plague, white band, and other ‘‘white’’ diseases. Page 488 in E. Rosenberg and Y. Loya, editors. Coral health and disease. Springer-Verlag, Berlin, Germany. Calnan, J., T. Smith, R. Nemeth, E. Kadison, and J. Blondeau. 2008. Coral disease prevalence and host susceptibility on mid-depth and deep reefs in the US Virgin Islands. Revista Biologia Tropical 56 (Supplement 1):223–224. Connell, J. H. 1997. Disturbance and recovery of coral assemblages. Coral Reefs 16 (Supplement):S101– S113. Cunning, R. and A. C. Baker. 2012. Excess algal symbionts increase the susceptibility of reef corals to bleaching. Nature Climate Change 3:259–262. Darling, E. S., L. Alvarez-Filip, T. A. Oliver, T. R. McClanahan, and I. M. Cˆote´. 2012. Evaluating life- history strategies of reef corals from species traits. Ecology Letters 15:1378–1386. Dayton, P. K. 1972. Toward an understanding of community resilience and the potential effects of enrichments to the benthos at McMurdo Sound, Antarctica. Pages 81–95 in Proceedings of the colloquium on conservation problems in Antarcti- ca. Allen Press, Lawrence, Kansas, USA. Eakin, C. M., et al. 2010. Caribbean corals in crisis: record thermal stress, bleaching, and mortality in 2005. PLoS ONE 5:e13969. Edmunds, P. 2010. Population biology of Porites astreoides and Diploria strigosa on a shallow Carib- bean reef. Marine Ecology Progress Series 418:87– 104. Edmunds, P., H. Putnam, R. Nisbet, and E. Muller. 2011. Benchmarks in organism performance and their use in comparative analyses. Oecologia 1–12. Edmunds, P. J. and P. S. Davies. 1986. An energy budget for Porites porites (Scleractinia). Marine Biology 92:339–347. Edmunds, P. J. and R. Elahi. 2007. The demographics of a 15-year decline in cover of the Caribbean reef coral Montastraea annularis. Ecological Monographs 77:3–18. Ellison, A. M., et al. 2005. Loss of foundation species: consequences for the structure and dynamics of forested ecosystems. Frontiers in Ecology and the Environment 3:479–486. Frieler, K., M. Meinshausen, A. Golly, M. Mengel, K. Lebek, S. D. Donner, and O. Hoegh-Guldberg. 2012. Limiting global warming to 28C is unlikely to save most coral reefs. Nature Climate Change 3:165–170. Gates, R. 1990. Seawater temperature and sublethal coral bleaching in Jamaica. Coral Reefs 8:193–197. Gates, R. and P. J. Edmunds. 1999. The physiological mechanisms of acclimitization in tropical reef corals. American Zoologist 39:30–43. Glynn, P. W. 1996. Coral reef bleaching: Facts, hypotheses and implications. Global Change Biol- ogy 2:495–509. Glynn, P. W. and S. B. Colley. 2008. Survival of brooding and broadcasting reef corals following large scale disturbances: is there any hope for broadcasting species during global warming? Pages 361–365 in 11th International Coral Reef Symposium, Ft. Lauderdale, FL. Glynn, P. W., J. L. Mate, A. C. Baker, and M. O. Calderon. 2001. Coral bleaching and mortality in Panama and Ecuador during the 1997-1998 El Nino-Southern Oscillation event: Spatial/temporal patterns and comparisons with the 1982-1983 v www.esajournals.org 29 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. event. Bulletin of Marine Science 69:79–109. Green, D. H., P. J. Edmunds, and R. C. Carpenter. 2008. Increasing relative abundance of Porites astreoides on Caribbean reefs mediated by an overall decline in coral cover. Marine Ecology Progress Series 359:1–10. Grottoli, A. G., L. J. Rodrigues, and J. E. Palardy. 2006. Heterotrophic plasticity and resilience in bleached corals. Nature 440:1186–1189. Grottoli, A. G., M. J. Rodrigues, and C. Juarez. 2004. Lipids and stable carbon isotopes in two species of Hawaiian corals, Porites compressa and Montipora verrucosa, following a bleaching event. Marine Biology 145:621–631. Herna´ndez-Pacheco, R., E. A. Herna´ndez-Delgado, and A. M. Sabat. 2011. Demographics of bleaching in a major Caribbean reef-building coral: Montas- traea annularis. Ecosphere 2:art9. Hoegh-Guldberg, O. 1999. Climate change, coral bleaching and the future of the world’s coral reefs. Marine & Freshwater Research 50:839–866. Hughes, T. P. and J. B. C. Jackson. 1980. Do corals lie about their age? Some demographic consequences of partial mortality, fission, and fusion. Science 209:713–715. Hughes, T. P. and J. E. Tanner. 2000. Recruitment failure, life histories, and long-term decline of Caribbean corals. Ecology 81:2250–2263. Huston, M. 1985. Variation in coral growth rates with depth at Discovery Bay, Jamaica. Coral Reefs 4:19– 25. Jones, R. 2008. Coral bleaching, bleaching-induced mortality, and the adaptive significance of the bleaching response. Marine Biology 154:65–80. Kramer, P., J. Lang, K. Marks, R. Garza-Perez, and R. Ginsburg. 2005. AGRRA methodology. Version 4.0, June 2005. University of Miami, Miami, Florida, USA. LaJeunesse, T. 2002. Diversity and community struc- ture of symbiotic dinoflagellates from Caribbean coral reefs. Marine Biology 141:387–400. Lang, J., H. R. Lasker, E. H. Gladfelter, P. Hallock, W. C. Jaap, F. J. Losada, and R. G. Muller. 1992. Spatial and temporal variability during periods of ‘‘recovery’’ after mass bleaching on western Atlan- tic coral reefs. American Zoologist 32:696–706. Lasker, H. R., E. C. Peters, and M. A. Coffroth. 1984. Bleaching of reef coelenterates in the San Blas Islands, Panama. Coral Reefs 3:183–190. Lesser, M. P. 1997. Oxidative stress causes coral bleaching during exposure to elevated tempera- tures. Coral Reefs 16:187–192. Lesser, M. P., J. C. Bythell, R. Gates, R. Johnstone, and O. Hoegh-Guldberg. 2007. Are infectious diseases really killing corals? Alternative interpretations of the experimental and ecological data. Journal of Exper- imental Marine Biology and Ecology 346:36–44. Loya, Y., K. Sakai, K. Yamazato, Y. Nakano, H. Sambali, and R. van Woesik. 2001. Coral bleaching: the winners and the losers. Ecology Letters 4:122– 131. Manzello, D. P., M. E. Brandt, T. B. Smith, D. Lirman, J. C. Hendee, and R. S. Nemeth. 2007. Hurricanes benefit bleached corals. Proceedings of the Nation- al Academy of Sciences 104:12035–12039. Marshall, P. A. and A. H. Baird. 2000. Bleaching of corals on the Great Barrier Reef: differential susceptibilities among taxa. Coral Reefs 19:155– 163. McClanahan, T. 2008. Response of the coral reef benthos and herbivory to fishery closure manage- ment and the 1998 ENSO disturbance. Oecologia 155:169–177. McClanahan, T. R. 2000. Bleaching damage and recovery potential of Maldivian coral reefs. Marine Pollution Bulletin 40:587–597. McClanahan, T. R. 2004. The relationship between bleaching and mortality of common corals. Marine Biology 144:1239–1245. McClanahan, T. R., M. Ateweberhan, C. A. Muhando, J. Maina, and M. S. Mohammed. 2007. Effects of climate and seawater temperature variation on coral bleaching and mortality. Ecological Mono- graphs 77:503–525. McClanahan, T. R., N. A. Muthiga, and S. Mangi. 2001. Coral and algal changes after the 1998 coral bleaching: interaction with reef management and herbivores on Kenyan reefs. Coral Reefs 19:380– 391. McField, M. 1999. Coral response during and after mass bleaching in Belize. Bulletin of Marine Science 64:155–172. Miller, J., E. Muller, C. Rogers, R. Waara, A. Atkinson, K. Whelan, M. Patterson, and B. Witcher. 2009. Coral disease following massive bleaching in 2005 causes 60% decline in coral cover on reefs in the US Virgin Islands. Coral Reefs 28:925–937. Miller, J., R. Waara, E. Muller, and C. Rogers. 2006. Coral bleaching and disease combine to cause extensive mortality on reefs in US Virgin Islands. Coral Reefs 25:418. Mills, M., F. Lipschultz, and K. Sebens. 2004. Particu- late matter ingestion and associated nitrogen uptake by four species of scleractinian corals. Coral Reefs 23:311–323. Mills, M. and K. Sebens. 2004. Ingestion and assimi- lation of nitrogen from benthic sediments by three species of coral. Marine Biology 145:1097–1106. Muller, E. M., C. S. Rogers, A. S. Spitzack, and R. van Woesik. 2008. Bleaching increases likelihood of disease on Acropora palmata (Lamarck) in Hawksn- est Bay, St John, US Virgin Islands. Coral Reefs 27:191–195. NOAA. 2012. US Virgin Island’s degree heating weeks v www.esajournals.org 30 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. plots. http://www.osdpd.noaa.gov/data/cb/time_ series/all_USVirgin.txt Oxenford, H. A., R. Roach, and A. Brathwaite. 2008. Large scale mortality in Barbados: a delayed response to the 2005 bleaching episode. Pages 505–509 in Proceedings of the 11th International Coral Reef Symposium, Ft. Lauderdale, FL. Porter, J. W., W. K. Fitt, H. J. Spero, C. S. Rogers, and M. W. White. 1989. Bleaching in reef corals: Physiological and stable isotopic responses. Pro- ceedings of the National Academy of Sciences USA 86:9342–9346. Richardson, L. L., W. M. Goldberg, K. G. Kuta, R. B. Aronson, G. W. Smith, K. B. Ritchie, J. C. Halas, J. S. Feingold, and S. L. Miller. 1998. Florida’s mystery coral-killer identified. Nature 392:557–558. Rodrigues, L. J. and A. G. Grottoli. 2007. Energy reserves and metabolism as indicators of coral recovery from bleaching. Limnology and Oceanog- raphy 52:1874–1882. Rogers, C. and J. Miller. 2001. Coral bleaching, hurricane damage, and benthic cover on coral reefs in St. John, U.S. Virgin Islands: a comparison of surveys with the chain transect method and videography. Bulletin of Marine Science 69:459– 470. Rogers, C. S. and J. Miller. 2006. Permanent ‘phase shifts’ or reversible declines in coral cover? Lack of recovery of two coral reefs in St. John, US Virgin Islands. Marine Ecology Progress Series 306:103– 114. Rowan, R., N. Knowlton, A. Baker, and J. Jara. 1997. Landscape ecology of algal symbiont communities explains variation in episodes of coral bleaching. Nature 388:265–269. Seemann, J., R. Carballo-Bola˜nos, K. L. Berry, C. T. Gonza´lez, C. Richter, and R. R. Leinfelder. 2012. Importance of heterotrophic adaptations of corals to maintain energy reserves. In Proceedings of the 12th International Coral Reef Symposium, Cairns, Australia. Smith, T. B., R. S. Nemeth, J. Blondeau, J. M. Calnan, E. Kadison, and S. Herzlieb. 2008. Assessing coral reef health across onshore to offshore stress gradients in the US Virgin Islands. Marine Pollution Bulletin 56:1983–1991. Thornhill, D. J., T. C. LaJeunesse, D. W. Kemp, W. K. Fitt, and G. W. Schmidt. 2006. Multi-year, seasonal genotypic surveys of coral-algal symbioses reveal prevalent stability or post-bleaching reversion. Marine Biology 148:711–722. Thornhill, D. J., R. D. Rotjan, B. D. Todd, G. C. Chilcoat, R. Iglesias-Prieto, D. W. Kemp, T. C. LaJeunesse, J. M. Reynolds, G. W. Schmidt, T. Shannon, M. E. Warner, and W. K. Fitt. 2011. A connection between colony biomass and death in Caribbean reef-building corals. PLoS ONE 6:e29535. van Woesik, R., E. C. Franklin, J. O’Leary, T. R. McClanahan, J. S. Klaus, and A. F. Budd. 2012. Hosts of the Plio-Pleistocene past reflect modern- day coral vulnerability. Proceedings of the Royal Society B 279:2448–2456. van Woesik, R., K. Sakai, A. Ganase, and Y. Loya. 2011. Revisiting the winners and the losers a decade after coral bleaching. Marine Ecology Progress Series 434:67–76. Villamizar, E., H. Camisotti, B. Rodrı´guez, J. Pe´rez, and M. Romero. 2008. Impacts of the 2005 Caribbean bleaching event at Archipe´lago de Los Roques National Park, Venezuala. Revista Biologia Tropical 56:255–270. Whelan, K., J. Miller, O. Sanchez, and M. Patterson. 2007. Impact of the 2005 coral bleaching event on Porites porites and Colpophyllia natans at Tektite Reef, US Virgin Islands. Coral Reefs 26:689–693. Wooldridge, S. A. 2009a. Water quality and coral bleaching thresholds: Formalising the linkage for the inshore reefs of the Great Barrier Reef, Australia. Marine Pollution Bulletin 58:745–751. Wooldridge, S. A. 2009b. A new conceptual model for the warm-water breakdown of the coral-algae endosymbiosis. Marine and Freshwater Research 60:483–496. v www.esajournals.org 31 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. SUPPLEMENTAL MATERIAL APPENDIX A Table A1. Traits of thermal tolerance for the nine focal species taken from van Woesik et al. (2012). To corroborate expert consensus, quantitative assessments of species traits from data published in Edmunds et al. (2011) were used to generate a median value among species and compared against the value for the species, when available. Values for A. agaricites were taken from the congener A. tenuifolia, as this was the closest available species with quantitative trait information. Quantitative trait information from Edmunds et al. (2011) was available for tissue thickness, distance between corallites, and calcification rate. Other traits were evaluated from the literature as referenced in the article text, or taken as given from expert consensus. Cover loss taken from this study (Table 3). Ties in rank were assigned the mean value of the rank slots that would have been taken by the tied ranks if arbitrarily assigned higher or lower. Morphology abbreviations as in Table 3. Trait Morph. Tissue thick./ mass Distance between corallites Calcification rate Colony size Corallite size Clade D Skeletal structure Trait score Agaricia agaricites Description F thin (4.8) small (11.5) fast medium small yes non-porous Score 0 1 1 1 0 1 1 1 4 Colpophyllia natans Description M thick (4.8) large Slow large large no non-porous Score 1 1 1 1 1 1 1 1 2 Diploria strigosa Description M thick large (2) Slow large large no non-porous Score 1 1 1 1 1 1 1 1 4 Montastraea annularis Description M thick (8.5) large (6.4) low (0.48) large small yes non-porous Score 1 1 1 1 1 1 1 1 4 Montastraea annularis spp. complex Description M thin (6.5) large low large small yes non-porous Score 1 1 1 1 1 1 1 1 4 Montastraea cavernosa Description M thick large (1.25) low medium large yes non-porous Score 1 1 1 1 0 1 1 1 5 Porites astreoides Description M, E thick small (18) low small small no porous Score 1 1 1 1 1 1 1 1 0 Branching Porites spp. Description B thick (17.1) small (23.8) low (0.46) small small no porous Score 1 1 1 1 1 1 1 1 4 Siderastrea siderea Description M thin small (9) low medium medium yes porous Score 1 1 1 1 0 0 1 1 2 Note: Tissue thick./mass refers to tissue thickness and biomass. Values from Agaricia tenuifolia. v www.esajournals.org 32 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. APPENDIX B Fig. B1. Photographic time course of two of the nine focal species, with examples of different phases of the response to the 2005 or 2010 high thermal stress events affecting the U.S. Virgin Islands. Agaricia agaricites: These corals showed severe bleaching and mortality early in the 2005 bleaching event. Surviving tissue remnants had begun to grow out of refuges and recolonize skeletons by 2009, with more robust recovery by 2012. Note that these are not the same colony. Colpophyllia natans: The response of bleached C. natans to the 2005 and 2010 event was typified by this marked colony followed over 2010 and 2011. Severe bleaching persisted during the height of the thermal disturbance. During restoration of colony pigmentation, white disease lesions (orange arrows) started from colony bases or margins and caused a high degree of partial mortality. v www.esajournals.org 33 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Fig. B2. Photographic time course of Diploria strigosa, with examples of different phases of the response to the 2005 high thermal stress event affecting the U.S. Virgin Islands. Colonies marked with an asterisk represent the same marked colony that was followed over time. Unmarked colonies are ad hoc examples of thermal stress response. The majortity of D. strigosa bleached in 2005 recovered with relatively little loss of tissue. However, a subset of colonies marked and followed between 2005 and 2006 showed the development of white disease during recovery and high loss of tissue. Mortality was partial, with sparse patches of living tissue remaining (orange arrows). v www.esajournals.org 34 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Fig. B3. Photographic time course of Montastraea annularis with examples of different phases of the response to the 2005 high thermal stress event affecting the U.S. Virgin Islands. Colonies are ad hoc examples of thermal stress response. Colonies of M. annularis bleached at moderate levels in 2005, with high intra-colony heterogeneity that ranged from unbleached to severly bleached tissue. Many colonies developed white disease in the recovery period, with high prevalence of partial mortality. Evidence of paling on surviving tissues often lasted for a year or more. v www.esajournals.org 35 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Fig. B4. Photographic time course of Montastraea annularis species complex with examples of different phases of the response to the 2005 high thermal stress event affecting the U.S. Virgin Islands. Colonies marked with an asterisk represent the same marked colony that was followed over time. Unmarked colonies are ad hoc examples of thermal stress response. The response of the Montastraea annularis species complex was similar to M. annularis, except that white diseases tended to be more severe on M. annularis species complex. Active or recently active lesion are denote in middle photos with orange arrows. A marked colony of M. faveolata that was impacted by severe bleaching (top left) had partial mortality that covered approximately 90% of the colony surface by 2006. This colony had shown relatively little recovery by 2011 (pink areas denote living tissue in bottom right image). v www.esajournals.org 36 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Fig. B5. Photographic time course of Montastraea cavernosa with examples of different phases of the response to the 2005 high thermal stress event affecting the U.S. Virgin Islands. Colonies are ad hoc examples of thermal stress response. Colonies of M. cavernosa showed a very weak bleaching response during thermal stress in 2005, with mostly colonies presening no bleaching or partial paling. Nonetheless, some colonies lost tissue in patches that rarely led to whole colony mortality. v www.esajournals.org 37 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Fig. B6. Photographic time course of Porites astreoides with examples of different phases of the response to the 2005 high thermal stress event affecting the U.S. Virgin Islands. Colonies marked with an asterisk represent the same marked colony that was followed over time. Unmarked colonies are ad hoc examples of thermal stress response. Some colonies of P. astreoides showed severe bleaching in 2005. Many colonies that were bleached became susceptible to algal overgrowth, although that was not directly measured in this study. One marked colony recovered pigment by 2006, but showed small losses of tissue (,25% of colony surface). v www.esajournals.org 38 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Fig. B7. Photographic time course of branching Porites spp. with examples of different phases of the response to the 2005 high thermal stress event affecting the U.S. Virgin Islands. Colonies of branching Porites spp. were severly bleached in 2005, with the majority of partial mortality occuring by January 2006. Algal overgrowth was common on bleached and recovering colonies. v www.esajournals.org 39 July 2013 v Volume 4(7) v Article 87 SMITH ET AL. Fig. B8. Photographic time course of Siderastrea siderea with examples of different phases of the response to the 2005 high thermal stress event affecting the U.S. Virgin Islands. Colonies marked with an asterisk represent the same marked colony that was followed over time. Unmarked colonies are ad hoc examples of thermal stress response. Colonies of S. siderea were moderately bleached in the 2005 bleaching event. A marked colony showed a typical pattern of pigment recovery with little loss of living tissue. v www.esajournals.org 40 July 2013 v Volume 4(7) v Article 87 SMITH ET AL.