In Tandem Reef Coral and Cryptic Metazoan Declines and Extinctions

BULLETIN OF MARINE SCIENCE. 87(4):767–794. 2011 CORAL REEF PAPER http://dx.doi.org/10.5343/bms.2010.1025

Peter W Glynn

ABSTRACT Coral reef degradation and loss have been extensively documented worldwide during the last few decades. While much attention has been directed toward the mortality of reef-building corals vis-à-vis various observed disturbances (e.g., bleaching, diseases, overfishing, nutrification), the fate of other reef-associated metazoans, especially invertebrates, has not received sufficient attention. Living and dead corals, reef frameworks, and carbonate sediments provide essential habitat niches for a multitude of symbiotic and cryptic species. Thirty-one animal phyla contain species that inhabit coral reefs with known global species richness estimated at 93,000. Pos- sibly as many as 1,000,000 reef-associated metazoans occur globally. Many of these species are undiscovered because of their cryptic or sibling nature. Metazoan reef associates have important functional roles on reefs, e.g., increasing survivorship of coral hosts, aiding in reef framework construction (calcification, consolidation), providing trophic resources, affecting coral mortality (corallivores) and erosion (bioerosion). Despite widespread bleaching and mortality, no reef-building corals (Scleractinia) have yet to become globally extinct. Three populations of Millepora spp. (Hydroida) were severely impacted in Pacific Panama during the 1982–83 El Niño–Southern Oscillation event. Present status indicates recovery of Millepora intricata Milne-Edwards and Haime, 1860 to shallow reef zones from relatively deep (10–15 m) refugia. Furthermore, two hydrocoral species have suffered regional extinctions in the eastern Pacific with populations still present in the Indo-Pacific (Millepora platyphylla Hemprich and Ehrenberg, 1834) and eastern Indian Ocean (Millepora boschmai de Weerdt and Glynn, 1991). Considering the large numbers of obligate symbionts and other coral reef metazoan associates, there is a strong likelihood of large-scale extinctions following the loss of reef-building corals.

Modern coral reefs were perhaps the first reported marine ecosystems to experience region-wide impacts from present-day climate change, principally from sea warming episodes (Glynn 1984, Brown 1987). Several other kinds of disturbances, acting alone or in combination with elevated temperature, have resulted in unprecedented global declines in coral reefs during the past few decades. Documented examples of local to regional-scale damage have been reported for cyclones and related disturbances (Woodley et al. 1981, Knowlton et al. 1990, Hughes and Connell 1999), diseases, including coral and sea urchin epizootics (Lessios et al. 1984, Hughes 1994, Aronson and Precht 1997, Harvell et al. 1999), coastal development and nutrient loading (Wilkinson 2008), and overfishing and the loss of herbivores (Jackson et al. 2001, Pandolfi et al. 2003). Additionally, recent studies have shown that calcification rates in corals and coralline algae are declining due to increasing CO2 levels and associated ocean acidification L( angdon and Atkinson 2005, Kuffner et al. 2008). Critically, most reef-building or zooxanthellate corals, i.e., those engaged in an obligate symbiotic relationship with photoautotrophic dinoflagellates Symbiodinium( spp.), live perilously close to their upper thermal (Glynn and D’Croz 1990, Brown

1997, Coles and Brown 2003) and irradiance (Shick et al. 1996, Wellington and Fitt 2003) tolerance limits. If the holobiont (coral host and algal endosymbionts or zooxanthellae) is unduly stressed, the symbiotic association is disrupted, leading to a loss of algae with visible coral bleaching and possible coral death. In the context of contemporary climate change, my aim here is three-fold, namely, to (a) compile evidence of the vulnerability of coral reefs to changing environmental conditions, (b) examine the habitat niches and roles of coral reef invertebrates with special reference to assessing their species richness and potentially destructive activities, and (c) offer some projections on the fate of coral reefs and how declines in coral hosts and reef habitat structures could impact associated biotas. In contrast to the loss of rain forests, which has resulted mainly from direct over-exploitation by humankind, coral reef loss is largely an indirect result of climate change, i.e., an unprecedented rapid warming of the world’s oceans due to the retention of anthropo- genically-produced greenhouse gases. However, in certain areas (Caribbean), coral mortality has been due primarily to diseases (e.g., Aronson and Precht 2001a, Weil et al. 2006), although elevated sea temperatures may also lower some coral’s resistance to disease (Lesser et al. 2007). Of the various kinds of disturbances, I will focus on increasing seawater temperature because of its demonstrated severe global impact to coral reef ecosystems. When first reported in the 1980s, region-wide coral bleaching/mortality events were regarded as acute disturbances, but because of their subsequent frequent occurrences, they are now considered to be chronic (Hoegh-Guldberg 1999, Baker et al. 2008,

Fenner and Heron 2009). Even if CO2 emissions can be stopped, modeling results indicate that residual levels of greenhouse gases will allow for continued atmospheric warming for at least 1000 yrs into the future (Solomon et al. 2009).

Vulnerability of Reef-Building Corals

Reef-building corals live close to their upper temperature tolerance limits, thus contributing to their vulnerability during periods of elevated thermal stress (Hoegh- Guldberg 1999, Baker et al. 2008). This susceptibility to thermal stress was evident in the Florida Keys during the warmest summer months of 1997, 1998, and 2005 when sea surface temperature anomalies reached 0.5–1.0 °C and resulted in widespread coral reef bleaching and mortality (Fig. 1). Recent laboratory and field studies also suggest an interactive effect of elevated pCO2 and temperature that together depress coral calcification (Reynaud et al. 2003, De’ath et al. 2009, Tanzil et al. 2009). Thus, with increasing temperature stress and a declining aragonitic saturation state, coral health and skeletal growth could become rapidly compromised. Anomalously elevated temperatures also interact synergistically with high irradiance levels, including ultraviolet radiation, that cause coral bleaching (Gleason and Wellington 1993, Dunne and Brown 1996). Reduced salinity, generally more influ- ential locally, also has been identified as a stressor causing coral bleaching (Brown 1997). Finally, disease prevalence has been shown to increase at higher temperatures as well (Brandt and McManus 2009). Whether bacterial infections during periods of high temperature stress can initiate mass bleaching events (Kushmaro et al. 1998, Ben-Haim Rozenblat and Rosenberg 2003) or are the result of weakened corals offer- ing an increased opportunity for microbial invasions (Mydlarz et al. 2006, Lesser et al. 2007, Ainsworth et al. 2008) remains to be determined. glynn: coral reef metazoan extinctions 769

Figure 1. Time series plots of monthly mean sea surface temperature anomalies at five Florida Keys coral reef sites during the warmest seasonal period (July–September). Each bar represents 1 mo, white bars identify bleaching years. Modified after Manzello et al. (2007).

According to Wilkinson’s (2008) assessment of the condition of global coral reefs during the past few decades, 19% have been effectively destroyed, and 35% are under threat of loss in 20–40 yrs. Jackson’s (2008) assessment of reef coral decline is even higher, with an overall 61% loss of coral cover globally. Coral cover losses at 263 monitored Caribbean reef sites demonstrated a mean decline of from ~54% to 9% over a three decade period (Gardner et al. 2003). In 2005, the northern hemisphere experienced the highest mean temperature since reliable records began in 1880. At that time, much of the eastern Caribbean experienced record high coral mortality. Coral reefs in the US Virgin Islands, the heaviest impacted area, suffered 51.5% coral mortality overall (Wilkinson and Souter 2008). In terms of species’ vulnerabilities, a comprehensive analysis concluded that one-third of all known reef-building corals (~231 of 704 species that could be assigned conservation status) are at an elevated risk of extinction from climate change and local impacts (Carpenter et al. 2008). These recent assessments indicate that coral losses and reef degradation have been substantial on a global scale. Because in recent years some coral reefs have shown significant recovery of live coral cover (e.g., Connell 1997, Coles and Brown 2003, Guzman and Cortés 2007, Wellington and Glynn 2007), it could be argued that perhaps these ecosystems pos- sess an inherent resilience with a capacity for continued existence, albeit in some altered state. An analysis of recovery after major bleaching events has shown that the majority of disturbed reefs in the Indian and Central/West Pacific Oceans have 770 BULLETIN OF MARINE SCIENCE. VOL 87, NO 4. 2011 revealed increases in live coral cover (Baker et al. 2008). Whether this recovery will continue at the present increasing pace of ocean warming, disease outbreaks, and acidification is considered unlikely by many (e.g., Hughes et al. 2003, Pandolfi et al. 2005, Aronson and Precht 2006, Hoegh-Guldberg et al. 2007). Following coral mortality, some ecological processes can complicate the recovery potential of reefs. For example, critical species population sizes are necessary for successful spawning, the so-called Allee effect (Knowlton 2001). If a coral population is severely reduced in size, then coral gametes spawned in the water column will have a diminished chance of encounter and successful fertilization. Another potentially serious problem is the continuing erosion of coral skeletons and framework loss after death. Bioerosion, the transformation of coral skeletal calcium carbonate to sediments, is caused by the feeding, boring, and etching activities of a variety of reef organisms. These destructive effects typically accelerate after coral death (Hutchings 2008). There is a delicate balance between coral reef growth and the destructive biotic agents causing reef degradation (Glynn 1988, Scott et al. 1988, Eakin 1996, 2001, Reaka-Kudla et al. 1996). If coral reef recovery—the regeneration of surviving colonies, and the recruitment and growth of newly settled corals—is greatly delayed or fails, then coral reef structures can quickly degrade. Reef frameworks of 1–3 m vertical relief were bioeroded to rubble in 10 yrs in the Galápagos Islands (Glynn 1994), and lost 1.5 m of vertical relief in 3 yrs in the Chagos Archipelago, Indian Ocean (Sheppard et al. 2002). Phase shifts, sudden changes from one dominant community type to another, can also occur on disturbed coral reefs (Done 1992, Hughes 1994). One of the more common phase shifts is a change from a coral-dominated community to one dominated by macroalgae (McCook 1999, McClanahan et al. 2001, Bellwood et al. 2004, Graham et al. 2006, Hughes et al. 2007, Ledlie et al. 2007). Phase shifts can interrupt coral reef recovery: once a coral reef has become dominated by macroalgae, reversal to the pre-disturbance community state is often difficult because (a) macroalgae are relatively unpalatable and not readily consumed by many herbivores, and (b) macroalgae interfere with the settlement of coral larvae. Macroalgal dominance can persist on reefs under any of the following conditions (or combinations thereof): (1) reduced herbivore abundances, (2) low abundances of piscine sea urchin predators, (3) reduced topographic relief of reef structures (which eliminates shelter for herbivores), and (4) elevated nutrient concentration. Reversals are possible, however, and may be mediated in unexpected ways. For example, Bellwood et al. (2006) demonstrated experimentally that the batfish Platax pinnatus (Linnaeus, 1758), by removing and ingesting the tough thalli of the dominant alga Sargassum at Orpheus Island, Great Barrier Reef, can induce a reversal from macroalgae to coral [the commonly cited state change from coral to macroalgal dominance has been challenged recently and may not be as common as assumed (Bruno et al. 2009)].

Coral Reef Biodiversity

Tropical coral reefs have long been celebrated as supporting one of the highest concentrations of biodiversity of all global ecosystems. Of 34 metazoan phyla, no fewer than 31 contain species that occur on coral reefs (Paulay 1997, Reaka-Kudla 1997). Enormous numbers of metazoan taxa inhabit and contribute to the structure of coral glynn: coral reef metazoan extinctions 771 communities, but estimates of their species richness differ widely due in large part to difficulties inherent in sampling complex reef habitats. Based on an analysis of known species-area relationships, an estimated 93,000 species could inhabit global coral reef communities (Reaka-Kudla 1997). Extrapolating from a conservative estimate of global rain forest species (2 million), Reaka-Kudla further concluded that the true number of species on global coral reefs is at least 950,000. If ~71% of mac- roscopic animals (including fishes) make up this total, then it might be reasonable to expect a predicted actual number of ~675,000 metazoan species. This does not include minute metazoans of the cryptic fauna, for example most members of the meiofauna, epizoic symbionts, or the reef zooplankton community. Intensive sampling of a Caribbean coral reef microcosm, in combination with calculations based on an area/diversity relationship, prompted Small et al. (1998) to suggest that Reaka- Kudla’s (1997) estimate should be at least three times higher. It has now been over 10 yrs since Reaka-Kudla’s (1997) estimate of coral reef species richness. To what extent has this changed? On the one hand, newly discovered species are being described yearly as more investigators (with new sampling methods) participate in reef research. Bouchet (2006) estimated that 1300–1500 valid species are being added yearly to the inventory of marine life. Adding to this increase is the recognition of large numbers of sibling species, aided by the development of molecular genetic analyses (Knowlton 1993). Species complexes have been revealed, for example, in the Placozoa (Pearse and Voigt 2007), among snapping shrimps (Mathews 2006), and Acanthaster (Vogler et al. 2008). On the other hand, the discovery of syn- onymies have reduced the number of valid species, there has been a significant at- trition of systematists over the past few decades, and probably, some species have suffered extinctions due to ongoing losses of habitat space (Bellwood and Hughes 2001). In Bouchet’s (2006) recent comprehensive review, he concludes that the best current estimate of the magnitude of global marine biodiversity is within the range of 230,000–275,000 species, which encompasses Reaka-Kudla’s estimate of 274,000 species. Whatever the actual total species richness of global coral reefs, the associated micro- and macro-invertebrate fauna is assuredly highly diverse. Enigmatic Taxa.—The death, erosion, and loss of corals and reef framework structures has a direct impact on the myriad of reef-associated animals, and hence on coral reef biodiversity. The habitat niches of numerous species in the majority of coral reef higher taxa, particularly those occupying cryptic habitats (Table 1), would render them particularly vulnerable to extinction following the disintegration of reef structures. For example, an experimental study in Panama comparing metazoan communities associated with dead or live corals found the following alterations to dead compared with live colonies: (1) disappearance of obligate invertebrate symbionts, (2) significant reductions in symbiont biomass, and (3) changes in multidi- mensional species assemblages (Enochs and Hockensmith 2008). Additionally, over the 1-yr course of the Panamanian study, the difference in habitat space (volume of branches plus inter-branch spaces) between dead and live corals increased mark- edly due to continued growth of live colonies and the bioerosion of dead colonies. Live colonies demonstrated a mean annual increase in total volume of 57.4% where- as dead colonies lost 34.1% of their volume (IC Enochs, University of Miami, pers comm). This difference was highly significant (paired t-test: P < 0.0001). 772 BULLETIN OF MARINE SCIENCE. VOL 87, NO 4. 2011

In a survey of the higher invertebrate taxa associated with coral communities, no fewer than 94% (44 of 47 taxa in 31 phyla) contain members belonging to the cryptos (i.e., species that reside within a variety of different kinds of cavities; Table 1). This class of organisms is phyletically rich and speciose, contributing significantly to the biodiversity of coral reefs. Ginsburg (1983) observed that this hidden reef biota has an important trophic link to all reef organisms because it equals or possibly exceeds in biomass that of surface-dwelling animals. Many, perhaps the majority, of these species are undescribed and/or unstudied. Whether or not some of them could or do survive in other types of non-calcareous habitat niches is unknown (Paulay 1997). A variety of potential challenges could occur in moving from a coral to non-coral environment (e.g., quality of habitat niche, parity of nutritional benefits, and/or exposure to novel predators and competitors, as well as different parasites and diseases). Numerous reef species also display a variety of symbiotic relationships in coral communities, and many of these partnerships are obligate in nature (Table 1). Exam- ples of some of the better-known symbioses involve ecto- (Cnidaria, Acoelomorpha, Platyhelminthes, Annelida, Copepoda, Decapoda) and endoparasites (Rhombozoa, Acanthocephala, Annelida), commensals (Annelida, Cirripedia) and diverse mutual- isms (Porifera, Cnidaria, Crustacea, Mollusca, Echinodermata). Kinne (1980) noted that all metazoan phyla have members that are either parasites or hosts. Hundreds of copepod (Humes 1985, 1994) and decapod crustacean species (Bruce 1998, Ng et al. 2008) have been described living in association with reef-building corals. And some taxa believed to be species poor, e.g., Placozoa and benthic Ctenophora, are probably highly diverse according to recent molecular genetic studies (Rudman 1999, Pearse and Voigt 2007). In many of these examples, should host species become rare or go extinct, so too would their obligate symbionts. Destructive Taxa.—Once considered rare to non-existent, numerous invertebrate corallivores (macro- and micro-predators) are now known to feed on reef-building corals. At least eight higher order taxa contain species that prey on coral tissues (Table 1). These include species belonging to Platyhelminthes, Annelida, Decapoda, Copepoda, Cirripedia, Gastropoda, Asteroidea, and Echinoidea (e.g., Robertson 1970, Glynn 1990, 2004, Rotjan and Lewis 2008). Several species of decapods, copepods, gastropods, and asteroids are obligate corallivores. Some of these corallivores, e.g., species of molluscs, asteroids and echinoids, are more resistant to elevated temperature stress than their prey (Glynn l985, 1988, Baird 1999, Adjeroud et al. 2002, Craig et al. 2005). This can result in a concentration of corallivores preying on reduced abundances of surviving coral prey, thereby leading to delayed mortality and continuing decline after a mortality event. Corallivore concentration after a bleaching event has been observed in the gastropods Coralliophila in the Caribbean (Knowlton et al. 1981) and Drupella in the western Pacific (Baird 1999), and in the seastar Acanthaster in the eastern Pacific (Glynn 1985, Guzman and Cortés 2001) and Indo-Pacific M( cClanahan et al. 2000, van Woesik et al. 2004). The preferred prey of these corallivores are also coral species that are among the most sensitive to bleaching (Baker et al. 2008). Bioerosion can reduce reef biodiversity by degrading habitat structure directly, and indirectly, by weakening corals and reef frameworks, making them more sus- ceptible to storm-generated waves and projectiles. Nine major reef-associated taxa in six phyla contain bioeroding species that are ubiquitous on global reefs (Table 1). glynn: coral reef metazoan extinctions 773

Table 1. Marine invertebrate phyla, denoting global numbers of described species, habitat occurrences, and principal ecological roles for species associated with coral reef communities. Species numbers are from Bouchet (2006) unless otherwise noted. “Cryptic” refers to organisms that live in concealed spaces protected from full or direct exposure to major physical factors (Kobluk 1988). “Symbiosis” (sensu lato) indicates a close interspecific relationship involving, for example, commensals, mutualists, or parasites (Castro 1988). Key: crp, cryptic; epb, epibenthic; par, parasitic; wcl, water column (including planktonic and nektonic species or life history stages); sym, symbiotic; str, structural; con, species aiding in reef consolidation; cal, calcifiers; cor, corallivores, species consuming coral tissues; bio, bioeroders, species eroding reef framework structures. Dashes under ecological roles indicate further study necessary. General sources: Paulay 1997, Brusca and Brusca 2003, Glynn and Enochs 2011. Specific sources are referenced in the text and footnotes to this table.

Phylum Number Habitat Ecological role Porifera 5,500 crp, epb sym, strg, con, calg, bio Placozoa 11a crp, epb − Rhombozoa 82 par symh Cnidaria 9,795 crpc, epb, wcl sym, str, con, cal Ctenophora 166 crp, epbd, wcl − Acoelomorpha > 340b crp, epb sym, cord Platyhelminthes 15,000 crp, epb sym, cor Nemertea 1,180 crp, epb − Rotifera 50 crp, epb − Gastrotricha 390 crpe − Kinorhyncha 130 crpe − Nemata 12,000 crpe − Nematomorpha 5 wcld − Acanthocephala 600 par symi Entoprocta 165 crp, epb − Gnathostomulida 97 crpe − Priapula 8 crpe − Loricifera 18 crpe − Sipuncula 144 crp bio Echiura 176 crp, epb − Annelida 12,148 crp, epb sym, cal, cor, bio Tardigrada 212 crpe − Arthropoda 47,217 crpe, epb, wcl symj, cor, bio Mollusca 52,525 crp, epb, wclf sym, strk,l, cal, cor, bio Phoronida 10 crp − Ectoprocta 5,700 crp, epb cal Brachiopoda 550 crp, epb cal Echinodermata 7,000 crp, epb sym, cal, cor, bio Chaetognatha 121 crp, epb, wcl − Hemichordata 106 crp − Chordata 4,932 crp, epb sym, con, cor, bio a—Based on known haplotypes of Placozoa (Pearse and Voigt 2007); b—From Barneah et al. (2007) and Tyler et al. (2005); c—Benthic hydroids and diverse octocorals and hexacorals are members of the reef cryptos. d— Questionable assignments, more study necessary. e—Commonly members of meiofauna in reefal sediments. The developmental stages of numerous other taxa also are temporary members of the meiofauna and plankton. f—Squids and cuttlefishes are nektonic. g—Many minute copepod species are epizoic and consume coral tissue (Humes 1985, Stock 1988); these could be regarded as micro-predators or ectoparasites. h—Habitat is host dependent; exclusively parasitic in nephridia of cephalopod hosts. i—Habitat is host dependent; parasitic in some coral reef fishes. j—Sclerosponges inhabiting reef cavities and deep fore-reef slope environments. k—Vermetid gastropods can coalesce, forming rim-like features on some reef flats and algal ridges. l—Larger species of tridacnid clams can contribute to the structural components of reef frameworks. 774 BULLETIN OF MARINE SCIENCE. VOL 87, NO 4. 2011

The skeletons and shells of all calcifying reef organisms are subject to the erosive activities of sponges, annelid worms, sipunculids, decapod crustaceans, barnacles, chitons, bivalve and gastropod molluscs, and echinoids. Internal borers such as clionid sponges, sipunculids, annelids, and bivalve molluscs can seriously weaken −2 −1 coral skeletons, in aggregate causing losses of up to 10–25 kg CaCO3 m yr . Certain mollusc and echinoderm species cause significant external erosion on reefs. Echi- noids, especially species of Diadema and Eucidaris, are among the most effective reef bioeroders. Maximum erosion rates for Diadema mexicanum A. Agassiz, 1863 −2 −1 of 3–10 kg CaCO3 m yr and for Eucidaris galapagensis Döderlein, 1887 of 3–22 −2 −1 kg CaCO3 m yr have been reported on eastern Pacific coral reefs (Glynn 1988). Reef parrotfishes, particularly Bolbometopon muricatum (Valenciennes, 1840), contribute greatly to bioerosion on outer reef crests in the western Pacific. The foraging activities of B. muricatum schools can cause losses of structural reef carbonates of −2 −1 up to 28 kg CaCO3 m yr (Bellwood et al. 2003).

Types of Extinctions

Five distinct types of extinctions, based on spatial and functional scales, have been recognized in the marine environment (Carlton et al. 1999). Three of these relate to the size of the area over which species losses occur: local, regional, and global. A local extinction occurs when a species population is displaced from a small area or habitat (e.g., a particular coral reef). When a species population disappears from a part of its normal or fundamental range and then occupies a smaller realized range it has undergone a regional extinction. The disappearance of a wide-ranging Indo-Pacific species from the eastern Pacific or Arabian Gulf regions would be classified as a regional extinction. A global extinction occurs when a species has been lost from its entire range (i.e., worldwide) and is no longer extant. Functional extinctions refer to reductions in species populations that lead to important changes in their ecological roles, with cascading effects on community structure and/or function. Severe reductions in fish and sea urchin herbivore populations have been implicated in phase shifts in many areas of the Caribbean (Hughes 1994, Jackson et al. 2001, Pandolfi et al. 2003). For example, overfishing in Jamai- ca has severely reduced the abundances of herbivorous fishes, and disease-related mass mortalities have decimated populations of the sea urchin Diadema antillarum Phillipi, 1845, resulting in the effective ecological removal of these important algal herbivores, thus allowing for elevated spatial competition between macroalgae and corals. Macroalgae, no longer under herbivore control, grow more rapidly than corals, blocking sunlight, increasing sedimentation, and interfering with coral larval settlement and recruitment. When a commercially important species is over-exploited (over-fished or over- hunted) and no longer provides profitable yields, exploitation ceases leading to commercial extinction. In a review of extinction vulnerability in marine populations, mainly with reference to vertebrate species (mammals, birds, and fishes), Dulvy et al. (2003) reported that the key threats were exploitation (55%) and habitat loss/degradation (37%). Commercially extinct species may recover if allowed adequate protection. Examples of recovery of over-exploited reef species are tridacnid giant clams at several sites in the south Pacific and a sea urchin species at Bolinao in the Philippines (see “Observed Species Declines”). glynn: coral reef metazoan extinctions 775

Species extinctions may not immediately follow a disturbance event, but may be subject to lengthy time delays (i.e., an extinction debt). Ecological adjustments to habitat loss and fragmentation may not allow for the long-term persistence of a species (e.g., Tilman et al. 1994, Hanski and Ovaskainen 2002). Metapopulation persistence depends on the number, size, and spatial configuration of habitat fragments necessary to satisfy a species’ threshold condition. If the threshold condition or metapopulation capacity is not met for a given species, then the species is expected to go extinct. Since this extinction can occur several years following habitat loss, it incurs an extinction debt in the fragmented community. An early modeling study with reference to terrestrial extinctions (Tilman et al. 1994) concluded that species at greatest risk were abundant dominant competitors that were poor colonizers. Similar to the results of Tilman et al. (1994), in a modeling study of coral reef habitat destruction at Eilat (Red Sea), Stone et al. (1996) found exceptionally high extinction rates, particularly in coral species (dominant competitors for space) with poor recoloniza- tion ability, however, these species were present at low rather than high abundances.

Observed Species Declines

In comparison with documented terrestrial and marine vertebrate extinctions, relatively few reef-associated invertebrate species have been listed as extinct (e.g., Carlton et al. 1999, Dulvy et al. 2003). Aside from the various difficulties in documenting extinctions in terrestrial ecosystems, validating marine extinctions on coral reefs has been hampered by a meager research effort directed toward reefs until only the past few decades, and the difficulty in locating and monitoring subtidal populations. Here, examples of severe declines and purported extinctions of reef-building coral species are examined first, followed by reef-associated invertebates. The present status of these at-risk or endangered species is examined where relevant information is available. The only two western Atlantic acroporid species, Acropora palmata (Lamarck, 1816) and Acropora cervicornis (Lamarck, 1816), have undergone major region-wide declines since the early 1980s. These losses are primarily a result of disease-related mortality, elevated temperature-induced bleaching, and physical damage from hurricanes. Based on observed high rates of population decline through their ranges, the National Marine Fisheries Service of the National Oceanic and Atmospheric Administration (USA) listed these acroporids as “threatened” species in 2006 under the Endangered Species Act (Acropora Biological Review Team 2005). More recently, since early 2000, both Acropora species have shown signs of recovering at several Caribbean localities, in The Bahamas, and at the Flower Gardens Banks in the Gulf of Mexico (Precht and Aronson 2006). Whether these will serve as source populations to ensure survival of these species and eventually renewed reef framework construction over the wider Caribbean remains uncertain. Four scleractinian coral species experienced extreme reductions in population size in the eastern Pacific as a consequence of the 1982–1983 El Niño bleaching event (Glynn 1997). These species and their respective localities were Pocillopora capitata Verrill, 1864, Costa Rica; Porites panamensis Verrill, 1864, Costa Rica; Psammocora stellata Verril, 1866, Panama and Galápagos Islands; Gardineroseris planulata (Dana, 1846), Costa Rica and Galápagos Islands. In addition, two species [Acropora valida (Dana, 1846) and Porites rus (Forsskål, 1775)] disappeared from the eastern Pacific 776 BULLETIN OF MARINE SCIENCE. VOL 87, NO 4. 2011 during the same period or slightly later, but the cause(s) of their extirpations is less certain. Since these species were only discovered in 1983, it is unclear what their status was prior to El Niño or why they disappeared. Acropora valida was discovered in 1983 at Gorgona Island, Colombia (Zapata and Vargas-Ángel 2003), and P. rus in 1983 at Samaná, Costa Rica (Cortés and Jiménez 2003). Because these species range widely throughout the Indo-Pacific, their losses from the eastern Pacific represented regional extinctions. To illustrate some of the difficulties in documenting regional and global-scale extinctions, the following encapsulates the step-by-step investigation of Millepora hydrocoral disappearances in the eastern Pacific in the aftermath of the severe 1982–1983 El Niño mortality event. Before the unprecedented El Niño warming event, three species of milleporid hydrocorals—all discovered in 1970—were observed on coral reefs in Panama (Glynn 1972, Glynn et al. 1972, Porter 1972). These zooxanthellate species were Millepora intricata Milne-Edwards and Haime, 1860 (Fig. 2A), Millepora platyphylla Hemprich and Ehrenberg, 1834 (Fig. 2B), and Millepora boschmai de Weerdt and Glynn, 1991 (Fig. 2C), all narrowly restricted to the Gulf of Chiriquí in the tropical Panamic Pacific Province. Millepora intricata and M. platyphylla are well known Indo-Pacific species, occurring on coral reefs across the Indian Ocean and western, central, and southern Pacific Ocean (Boschma 1948, Randall and Cheng 1984, Razak and Hoeksema 2003, Fenner 2006, 2007). Relative to the eastern PacificP rovince, the nearest reported Indo-Pacific populations of M. intricata and M. platyphylla occur in Fiji and the Marquesas Islands, 10,800 km and 6600 km west of Panama, respectively. When first observed in Panama in 1970, M. boschmai was considered a possible new species; this was verified from material sent to the hydrocoral systematist H Boschma, who died before describing the new species. In order to avoid a “centinelan extinction,” or the loss of a species before it is known to science (Wilson 1992), a concerted effort was undertaken to search for Millepora spp. over their former ranges. Extensive surveys in previously occupied habitats over a 7-yr period (1983–1990), with a total search effort of 204 diver hrs, failed to locate live colonies of M. platyphylla or M. boschmai, thus it was concluded that these two species experienced regional and global extinctions, respectively. Only 7 mo after the publication of the extinction of M. boschmai (Glynn and de Weerdt 1991), five live colonies were discovered at a cove on the north shore of Uva Island, Gulf of Chiriquí (Glynn and Feingold 1992). Based on the sizes of these colonies, and their estimated growth rates, it is probable that they recruited to this site after 1983. A total of eight live colonies of M. boschmai were observed at Uva and Coiba Islands from 1992 to 2001 (Brenes et al. 1993, Maté 2003). However, despite intensive surveys since, no living colonies have been encountered through May 2003, at Coiba Island (Guzman et al. 2004) and through March 2010, at Uva Island (P Glynn, pers obs). Millepora platyphylla has not been seen alive in the eastern Pacific since 1983 (i.e., for 27 yrs). Millepora intricata was again abundant in shallow reef zones at Uva Island and other reefs 14 yrs after the 1982–1983 mortality event. Living populations were present in deeper water (12–25 m) at several reef sites after 1983, and it is likely these served as source populations for re-establishment in shallow reef zones (Glynn et al. 2001). All of the shallow reef populations of M. intricata that had recovered after 1983, again bleached and died during the 1997–1998 El Niño event. As in 1983, deeper populations at 12–20 m depth did not bleach or experience any marked increase in mortality. The lower temperatures and reduced light levels glynn: coral reef metazoan extinctions 777

Figure 2. Hydrocoral species severely impacted during the 1982–1983 El Niño–Southern Oscillation bleaching event in Panama, Gulf of Chiriquí, eastern tropical Pacific. (A) Millepora intricata experienced a temporary (3–4 yr) decline, disappearing entirely from shallow reef zones. Secas Islands reef, 4 m depth, 18 March, 1990. (B) Millepora platyphylla, all known colonies on the Uva Island reef bleached and died, and none has been sighted since in the eastern Pacific. Kingman reef, northern Line Islands, 4 m depth, 2000 (colony diam 2–2.5 m, height 2 m). (C) Millepora boschmai, now likely regionally extinct. Lazarus Cove, Uva Island, 6 m depth, 22 February, 1992. Photographs A and C courtesy of JS Feingold, photograph B by JE Maragos. at depth likely contributed to the survival of these corals (Fig. 3). Shallow reef areas again were colonized by M. intricata (~2 cm high colonies), at two sites as early as 2000, and by 2002, several colonies (one 21 cm in height), were present at 2–3 m depth. Surprisingly, Razak and Hoeksema (2003) recognized M. boschmai based on skeletal morphological characters in collections from Indonesia. The occurrence of five colonies at south Sulawesi and Sumba, ~17,600 km west of Panama, indicates that M. boschmai can no longer be considered an eastern Pacific endemic. In summary, M. boschmai and M. platyphylla are best considered regionally extinct species since they both are now known to occur elsewhere in the Indo-Pacific region. Millepora intricata is also a wide-ranging Indo-Pacific species that is confined to a single gulf 778 BULLETIN OF MARINE SCIENCE. VOL 87, NO 4. 2011

(Gulf of Chiriquí, Panama) in the eastern Pacific. Since it is capable of re-populating shallow reefs from deep-water refuges following El Niño disturbances, it is not at present considered regionally endangered. Among reef-associated invertebrates, the three mollusc [Stiliger vossi Ev. Marcus and Er. Marcus, 1960, Hippopus hippopus (Linnaeus, 1758), and Tridacna gigas (Linnaeus, 1758)] and single echinoid [Tripneustes gratilla (Linnaeus, 1758)] species listed by Dulvy et al. (2003) represent severe population reductions that could result in local extinctions. It is difficult to judge the status of S. vossi, a minute (ca. 1 mm in length) ascoglossan mollusc described in 1960. Clark (1994) noted that it is a very rare species, but long-term monitoring at its type locality has not been performed. The two tridacnid giant clams, H. hippopus and T. gigas, were overfished and may best be regarded as local, commercial extinctions. Tripneustes gratilla, a sea urchin harvested for its gonads at Bolinao in the Philippines, experienced severe declines in the 1990s, but has since demonstrated a degree of recovery from grow-out culture efforts and management intervention (Juinio-Meñez et al. 2008). Diadema antillarum population declines during the early 1980s were region-wide in the Caribbean, Bahamas, and Gulf of Mexico, but populations have recovered in Jamaica (Edmunds and Carpenter 2001) and some other areas of the western Atlantic during the past decade (Jaap et al. 2008, Rogers et al. 2008, Furman and Heck 2009, Vermeij et al. 2010).

Projected Trends, Outlook

Temperature increases projected for the end of this century range from 1.8 to 3.4 °C. Temperature increases in the tropics during past interglacial periods are thought to have been within this range although probably there is more evidence for the lower end of this range (IPCC 2007). While corals survived the warming associated with past glacial to interglacial transitions there is concern that both the magnitude and certainly the rate of increase will be greater in this century and that this will exceed the tolerance limits of certain reef-building coral species that do not have the capacity to acclimate or adapt to such changes (Hoegh-Guldberg et al. 2007, Baker et al. 2008).

In addition to ocean warming, the projected values of atmospheric CO2 concentrations of 555–825 ppmV would lead to ocean acidification and reduced calcification in coral reef ecosystems. In experimental studies doubling present-day pCO2 concentration (to simulate expected mid-21st century values), calcifying organisms have invariably demonstrated a diminished growth capacity. For example, based on experimental control of pCO2 in a coral mesocosm, Langdon et al. (2000) predicted a 40% decline in coral reef calcification rates. This reduction in calcification could transform some reefs from net accreters (capable of keeping up with bioerosion and sea level rise), to reefs that begin to slowly lose this capacity. Juvenile corals also have lowered skeletal growth rates, reductions of 50%–78%, at 560 and 720 pCO2 concentrations, respectively (Albright et al. 2008). Slower colony growth of juvenile corals will lengthen their time in this vulnerable stage in their life history and likely result in greater mortality of these new recruits. Crustose coralline algae, important in binding and consolidating reef structures, showed severe reductions in recruitment and growth rates in elevated pCO2 treatments (Kuffner et al. 2008). Finally, increased acidification has also been demonstrated to have various negative effects glynn: coral reef metazoan extinctions 779

Figure 3. Reef bottom temperature traces at three depth zones, Uva Island coral reef, Gulf of Chiriquí, Panama (March 2003–March 2004). Data entries = 48 per day (every 30 min) at each depth. Optic StowAway recorder, accuracy: ± 0.2 °C @ 21.1 °C, Onset Computer Corp. on invertebrate reproduction and early life history stages (Shirayama and Thornton 2005, Fabry et al. 2008). I consider here the survival of reef-building corals, the essential architects of coral reef structures. Their survival will determine the continued presence or demise of the multitude of reef metazoans that live as obligate symbionts on live coral hosts (Jones et al. 2004, Glynn and Enochs 2011). Living corals, of course, generate a variety of dead-coral substrates that provide shelter for motile cryptic metazoans that, in the eventuality of wide-spread coral mortality, may exceed the species richness associated with live corals (Mikkelsen and Cracraft 2001; IC Enochs, University of Miami, pers comm 2010). Reef-Building Corals.—Certain coral genera, especially those with species with non-branching colony morphologies, have been observed to survive severe bleaching events. These generally resistant and resilient genera, which often contribute importantly to reef building include: Porites (Hoegh-Guldberg and Salvat 1995, Gates and Edmunds 1999, Hueerkamp et al. 2001, Loya et al. 2001), Diploastrea (Schuhmacher et al. 2005), Astreopora (McClanahan 2000), Pavona (Hueerkamp et al. 2001, McClanahan et al. 2004), Siderastrea (Gates and Edmunds 1999), Goniastrea (Loya et al. 2001, Brown et al. 2002), Platygyra, Leptastrea, Favites, Favia (Loya et al. 2001, LaJeunesse et al. 2003), Cyphastrea, Goniopora, and Galaxea (McClanahan et al. 2004). Coles and Brown (2003) and Baird et al. (2009) have offered reviews of the known and potentially effective physiological mechanisms that could play a role in the acclimatization and adaptation of coral animals to elevated temperature and irradiance stress. Some coral hosts can minimize stressful solar radiation flux to their symbionts by producing fluorescent pigments (FPs) and sequestering mycosporine-like amino acids (MAAs). FPs have been shown to absorb, scatter, and dissipate high-energy 780 BULLETIN OF MARINE SCIENCE. VOL 87, NO 4. 2011 radiation, thus reducing photodamage. MAAs also absorb and dissipate UV energy, thus limiting the formation of toxic intermediate byproducts of photosynthesis. Two additional mechanisms that could reduce bleaching damage are antioxidant systems and heat shock proteins, both observed in corals. Also, coral species that can supple- ment their energy needs by heterotrophic feeding during stressful periods have been shown to survive experimental bleaching better than non-feeding corals (Grottoli et al. 2006). Finally, Loya et al. (2001) hypothesized that the survival of non-branching corals during a major bleaching event on Okinawa was related to the possession of thick tissues, and a colony morphology that facilitates a high mass transfer of reac- tive oxygen species produced during photosynthesis. The survival of brooding coral species (in contrast to broadcast spawning corals) in the Caribbean during the 1980s and 1990s, have prompted some workers to consider the brooding reproductive mode to be advantageous in a global warming scenario (Knowlton 2001). In the western Atlantic, brooding poritid, agariciid, and siderastreid corals have demonstrated relatively high survivorship during disturbance events (Hughes 1994, Aronson and Precht 2001b, Kikuchi et al. 2003) and increasing relative abundances of poritids (Green et al. 2008), but broadcast spawning corals in the Montastraea complex seem to have fared better than brooders in the US Virgin Islands (Rogers et al. 2008). The broadcast spawning eastern Pacific scleractinian coral fauna has survived recent El Niño–Southern Oscillation bleaching events (Glynn and Colley 2008). It is premature to conclude which if any of these reproductive traits might be advantageous. Since the pioneering study of Rowan et al. (1997), who demonstrated the existence of diverse, multi-cladal communities of Symbiodinium in the Caribbean Montastraea species complex and their differing sensitivities to increasing temperature and solar radiation, similar responses have been observed in several additional coral species. Evidence is mounting that the potential for different Symbiodinium variants (e.g., based on the internal transcribed spacers 1 and 2 of a multi-copy ribosomal gene family, ITS-1 and ITS-2 rDNA) to enhance the physiological tolerance of at least some reef coral species to climate change stressors may be effective into the next century, providing greenhouse gas emissions are moderated (Baskett et al. 2009). A differential adaptability of coral holobionts could result in certain coral species dominating future reef community structure(s). Recently identified resistant or resilient host species and the Symbiodinium variants that contribute to their thermotolerance include: Pacific—Acropora millepora (Ehrenberg, 1834), ITS-1 types C1 and D (Berkelmans and van Oppen 2006, Abrego et al. 2008, Jones et al. 2008); Acropora tenuis (Dana, 1846), ITS-1 type C1 (Abrego et al. 2008); Montipora digitata (Dana, 1846) and Porites spp. (LaJeunesse et al. 2003); Pocillopora spp., symbionts in clade D (Rowan 2004); eastern Pacific—Pocillopora spp., clade D (Baker et al. 2004); Caribbean—Montastraea annularis (Ellis and Solander, 1786), Montastraea faveolata (Ellis and Solander, 1786), and Montastrea franksi Gregory, 1895, symbionts in clades A and D (Rowan et al. 1997, Thornhill et al. 2006). This list is likely more reflective of the experiments and observations performed to date than of the complete diversity of coral-Symbiodinium associations that are relatively thermo-tolerant. In response to thermal stress in some coral species, symbiont community shifts have been observed to enhance a coral host’s ability to resist and/or recover from bleaching. In the case of Acropora millepora on the Great Barrier Reef, Berkelmans and van Oppen (2006) reported an increase in temperature tolerance of ~1–1.5 °C for glynn: coral reef metazoan extinctions 781 colonies hosting Symbiodinium in clade D. It is important to note that heat and light tolerant Symbiodinium types are not universally resistant across different host coral taxa but involve species-specific interactions between hosts and endosymbionts that may modify the physiological response of the holobiont. Refugia.—With the likelihood that various coral species could survive moderate thermal bleaching (low-emission scenario B1, 1.1–2.9 °C increase; IPCC 2007) through the 21st century, the availability of refugia becomes of critical importance. Some habitats with relatively benign environmental conditions have already provid- ed a degree of protection to corals during major bleaching events. Observations of coral survival and recovery in the Caribbean, northern Red Sea, and off southeastern Africa during the 1998 bleaching event demonstrated a strong correlation with local upwelling and medium depth sites where conditions did not exceed bleaching threshold limits (Riegl and Piller 2003). Based on an analysis of sea surface temperature (SST) time series off Madagascar, McClanahan et al. (2009) found a close relationship between the condition of coral communities and long-term trends in site-specific thermal environments. Ostensibly undisturbed coral communities in northern Madagascar with high coral cover and relatively high numbers of coral genera occur in thermally benign environments, characterized by moderate, relatively constant SSTs, and weak interannual periodicities. However, some evidence suggests that several shallow-occurring coral species subject to severe daily temperature variations (plus marked fluctuations in irradiance, pH, and salinity) are relatively resistant to thermally-induced bleaching. Such communities, observed in the Gulf of Oman (Coles 1997), American Samoa (Craig et al. 2001), and off the east African coast (McClanahan et al. 2007), may also serve as refugia. High coral survivorship on offshore bank reefs (compared with insular fringing reefs) was also monitored and modeled in the western Caribbean during the 1998 bleaching event (Riegl et al. 2008). The high survivorship of A. cervicornis at offshore locations was attributed to vigorous flushing of reef waters and reduced runoff. In the eastern Pacific P( anama), no coral bleaching was observed in an upwelling center during 1998, but corals were severely impacted in non-upwelling areas during the height of the warming event (Glynn et al. 2001). As noted in “Observed Species Declines,” deep-living zooxanthellate hydrocorals that experienced tidally-forced pulses of cool water also survived the 1998 bleaching event. Another example comes from a recently discovered, deep-coral community on the insular shelf off the SU Virgin Islands (Armstrong et al. 2006). This mesophotic community, occurring between 33–47 m depth, is dominated by corals in the M. annularis complex (> 90% dominance by cover, Smith et al. 2010). Mean live coral cover is 43%, greatly exceeding that of Montastraea cover (< 10%) on nearshore reefs at shallow depth (Smith et al. 2008). Broadcast spawning of Montastraea spp. between 33–45 m depth at the Flower Garden Banks in the Gulf of Mexico (Vize 2006) dem- onstrates the reproductive viability of some mesophotic populations and the pos- sibility that they could figure effectively in repopulating depleted shallow reef zones. The role that deeper-living corals will play in ameliorating the risk of extinction will depend on their connectivity with shallow populations and their ability to maintain relative resistance to mortality vis-à-vis regional disturbances of increasing severity and frequency. 782 BULLETIN OF MARINE SCIENCE. VOL 87, NO 4. 2011

In some regions where structural coral reef development is limited (Coiba Island), coral communities, i.e., non-reefal coral assemblages, may serve as sources for the repopulation of degraded reefs (Guzman et al. 2004). Coral cover and species richness were higher in coral communities than on reefs, and also contained rare and endangered coral species. Additionally, the surface cover of coral reefs in the Coiba study area was < 10% of the coral community cover (1560 ha). Therefore, it is possible that corals and associated species that are widely distributed over a diverse range of environments would have a greater likelihood of survival during bleaching disturbances. Coral Community Structure and Reef-Building.—With the extirpation of thermally-sensitive corals, coral community diversity would decline both locally and regionally. It is also likely that some of the 200+ species at an elevated risk of extinction could become globally extinct (Carpenter et al. 2008). From theoretical estimations of species extinctions, based on species number and area relationships (Reaka-Kudla 1997), Carlton et al. (1999) have argued that numerous reef species could already have become extinct. If global coral reefs support 3–4 million species total (known and unknown), and if 5% of the world’s reef area has been irreparably degraded, it is possible that a loss of 50,000–60,000 species has already occurred. Species losses would amount to 300,000–400,000 if 30% of global reefs were destroyed. As noted earlier, Wilkinson (2008) estimated that 19% of the world’s coral reefs have already been effectively destroyed and predicted that 35% are under risk of collapse during the next 20–40 yrs. These are daunting predictions and are not without uncertainties. Firstly, not a single coral reef invertebrate species has been documented to have become globally extinct in recent years. Although a coral-inhabiting goby of restricted distribution is under extreme threat of extinction (Munday 2004), no reef fishes are known to have disappeared during the past few decades. In the Galápagos Islands, a seastar [Heliaster solaris (A. H. Clark, 1920)] and damselfish (Azurina eupalama Heller and Snodgrass, 1903) have not been seen for 30 yrs and are quite possibly extinct (Edgar et al. 2009), but these species were not commonly associated with coral reefs. Sec- ondly, coral recovery has occurred during the past decade in several regions, particularly on reefs in the Indian and Pacific Oceans (Wilkinson 2008, Baker et al. 2008). With coral recovery in progress, what proportion of reef-associated species would be expected to re-occupy coral hosts and former coral communities? Like corals that have survived severe bleaching events in refugia, it is likely that many reef associates can also persist in such environments. Since the majority of reef associates have planktonic larvae, and given sufficient connectivity between disturbed sites and refugia, there exists a potential for the repopulation of severely degraded coral reefs. Even if coral reef structure is greatly altered, many reef associates may be able to survive, at least temporarily, in dispersed coral colonies, dead coral skeletons, loose reef debris, or relatively fine-grained limestone sediments. A recurrent theme in coral reef degradation centers on phase shifts or alternate stable states in community structure (e.g., Knowlton 2001, Hughes et al. 2003, Hoegh-Guldberg et al. 2007). Marked declines in coral cover may cause coral communities to shift to other community types dominated by turf algae, macroalgae, or sea urchin barrens. If grazing by reef herbivores (fishes and sea urchins) is severely reduced for example, beyond a critical threshold, hysteresis could occur with a change to a stable macroalgal-dominated community (Hughes et al. 2003, Mumby glynn: coral reef metazoan extinctions 783 et al. 2007). With algal communities dominating reef substrates, a different suite of associated organisms would be expected to develop. Algal turf communities main- tained by damselfishes favor the proliferation of internal bioeroders, and this could accelerate reef framework erosion (Risk and Sammarco 1982). Evidence from some studies suggests changes in the trophic structure of diverse coastal marine ecosystems due to the functional extinction of large consumers and the introduction of non-native species at lower trophic levels (e.g., Duffy 2003, Byrnes et al. 2007). Examples of large fish and other vertebrate consumer losses on coral reefs have been documented in several regions (e.g., Jackson et al. 2001, Jokiel 2008, Paddack et al. 2009). These opposing changes alter food web structure such that higher consumer species are reduced in numbers, and lower trophic levels, composed dominantly of filter feeders and scavengers, increase in abundance. Some examples of exotic invertebrate introductions have been reported, but these are generally rare compared with the loss of large vertebrate consumers. Two examples of probable introductions to coral reefs in Hawaii are the filter feeding sponge Mycale (Mycale) armata Gray, 1867 (Coles and Bolick 2007) and the octocoral Carijoa riisei (Duchassaing and Michelotti, 1860) (see Jokiel 2008). Both of these species are highly aggressive competitors, overgrowing reef corals and possibly leading to increased bioerosion. Whether or not trophic skew occurs on coral reefs needs further study because disproportionate attention has been given to conspicuous fishes and little information is available on changing abundances of lower level invertebrates. With respect to the loss of habitat space, including live coral hosts, several cryptic invertebrate taxa would likely become severely depleted, particularly acoelomorph worms and obligate crustacean symbionts such as decapods, barnacles, and copepods (Table 1). Compared with zooxanthellate corals, most non-symbiotic epibenthic species are not thermally sensitive. As free space becomes available, groups that are already abundant in some reef zones, such as sponges, molluscs (vermetid gastropods and bivalves), echinoderms (sea urchins), and colonial tunicates, would likely contribute more importantly to benthic cover. Reef plankton species that shelter in reef structures during the day and emerge at night (Porter and Porter 1977, Carleton and Hamner 1989) would likely decline with the erosion and loss of their structural habitat. This would be expected to cause a delayed effect, an extinction debt, with the loss of calcifying species that generate reef framework and dead coral substrates. In terms of ecological roles, those symbionts dependent on reef organisms for nu- triment or habitat niches during development, would be impacted. The loss of cryptic species would diminish trophic resources to fish communities, slow down decomposition processes, nutrient cycling, and filtering capacity (Rothans and Miller 1991, Hatcher 1996, Richter and Wunsch 1999, Yahel et al. 2006). Other facultative symbionts may well be able to survive in altered reef environments. Non-symbiotic reef organisms contributing to the structural integrity and consolidation of reefs (e.g., certain sponge, cnidarian and mollusc species), would be expected to survive. With lowered carbonate saturation states, calcification would become depressed in adult species with calcareous skeletons as well as in larval stages with supporting and pro- tective skeletal elements (Kurihara and Shirayama 2004). Reduced skeletal protection in mollusc and echinoderm larvae could significantly increase their vulnerability to planktonic predators. Compared with their heat-sensitive prey, facultative corallivores would survive and concentrate their feeding on greatly depleted coral populations. Obligate corallivores, such as certain decapods, gastropods, and asteroids, 784 BULLETIN OF MARINE SCIENCE. VOL 87, NO 4. 2011 may not be capable of switching their diets to alternative prey. Since the majority of bioeroders more effectively erode dead coral surfaces (Highsmith 1981, Hutchings 1986), the pace of reef framework breakdown would be expected to accelerate. In summation, if coral reef-bleaching disturbances become more severe and remain in a chronic state it is probable that several coral species will become extinct by the end of the century. With the demise of many corals, we could then expect the loss of tens of thousands of known and unknown coral associates. Remaining invertebrate corallivores and bioeroders would continue to damage depleted coral populations and deteriorating reef structures. Extinctions of cryptic invertebrates that utilize reefs for shelter and other resources would continue for decades as carbonate structures disintegrate and disappear. These are not encouraging prospects, but unfortunately remain squarely within the realm of possibilities.

Acknowledgments

I thank TB Smith and DP Manzello for permission to reproduce temperature and coral bleaching graphics, and D Holstein for technical assistance in producing depth-related temperature traces. Additionally I thank JS Feingold and J Maragos for allowing me to reproduce their Millepora photographs. Acquisition of pertinent literature was greatly facilitated by L McManus, C Hurt, A Campbell, and D Holstein. Discussions with VW Brandtneris, SD Cairns, IC Enochs, C Langdon, DP Manzello, AMS Correa, TB Smith, and C Wilkinson greatly improved the focus and quality of this contribution. I am also grateful for the constructive comments offered by reviewers and RJ Araújo. Research results in the eastern Pacific were funded by NSF grant OCE-0526361 and earlier awards.

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