Coral reef bleaching: ecological perspectives. P.W. Glynn. Division of Marine Biology and Fisheries, Rosenstiel School of Marine and Atmospheric Science, ...
Coral Reefs (1993) 12:1-17
Coral Reefs 9 Springer-Verlag 1993
~%~,0~
Reports Coral reef bleaching: ecological perspectives P.W. Glynn Division of Marine Biology and Fisheries, Rosenstiel School of Marine and Atmospheric Science, University of Miami, 4600 Rickenbacker Causeway, Miami, Florida 33149, USA Accepted 19 June 1992
Abstract. Coral reef bleaching, the whitening of diverse invertebrate taxa, results from the loss of symbiotic zooxanthellae and/or a reduction in photosynthetic pigment concentrations in zooxanthellae residing within the gastrodermal tissues of host animals. Of particular concern are the consequences of bleaching of large numbers of reef-building scleractinian corals and hydrocorals. Published records of coral reef bleaching events from 1870 to the present suggest that the frequency (60 major events from 1979 to 1990), scale (co-occurrence in many coral reef regions and often over the bathymetric depth range of corals) and severity ( > 95% mortality in some areas) of recent bleaching disturbances are unprecedented in the scientific literature. The causes of small scale, isolated bleaching events can often be explained by particular stressors (e.g., temperature, salinity, light, sedimentation, aerial exposure and pollutants), but attempts to explain large scale bleaching events in terms of possible global change (e.g., greenhouse warming, increased UV radiation flux, deteriorating ecosystem health, or some combination of the above) have not been convincing. Attempts to relate the severity and extent of large scale coral reef bleaching events to particular causes have been hampered by a lack of (a) standardized methods to assess bleaching and (b) continuous, long-term data bases of environmental conditions over the periods of interest. An effort must be made to understand the impact of bleaching on the remainder of the reef community and the longterm effects on competition, predation, symbioses, bioerosion and substrate condition, all factors that can influence coral recruitment and reef recovery. If projected rates of sea warming are realized by mid to late AD 2000, i.e. a 2 ~ C increase in high latitude coral seas, the upper thermal tolerance limits of many reef-building corals could be exceeded. Present evidence suggests that many corals would be unable to adapt physiologically or genetically to such marked and rapid temperature increases.
Introduction Mass coral mortalities in contemporary coral reef ecosystems have been reported in all major reef provinces since the 1870s (Stoddart 1969; Johannes 1975; Endean 1976; Pearson 1981; Brown 1987; Coffroth et al. 1990). Why, then, should the coral reef bleaching and mortality events of the 1980s command great concern? Probably, in large part, because the frequency and scale of bleaching disturbances are unprecedented in the scientific literature. For example, no less than 60 major "coral reef bleaching events" (Fig. 1 a) were reported over the 12 year period, 1979-1990 (Coffroth et al. 1990; Williams and BunkleyWilliams 1990; Glynn 1991), compared with 45 "mass coral mortalities" (Fig. 1 b, c) caused by various other disturbances. In contrast, only three bleaching events were reported among 63 mass coral mortality records during the preceding 103 years (1876--1979; Coffroth et al. 1990; see Fig. 1 caption for additional references). An alternative explanation for the increased frequency of disturbances to coral reefs can be attributed to more observers and a greater interest in reporting in recent years. The incidences of mass coral mortalities, with 12 reports before 1940 and 96 from 1940 to 1990, suggest this possibility. [The lack of recent (1988-1990) mass coral mortality events may be due to a lag in reporting.] However, only 9 coral reef bleaching events were reported in the 1960s and 1970s, during a period of active coral reef research. The increased incidences of mass coral mortalities caused by the crown-of-thorns predator Acanthaster planci Linnaeus in southern Japan since the 1930s are also probably valid and largely independent of observer frequency and interest in ecological implications (see Fig. 1 in Birkeland and Lucas 1990). Considering the broad geographic areas affected and the abruptness with which coral reef bleaching became a concern, this disturbance is reminiscent of the severe coral mortalities caused by Acanthaster during the past two decades. A minimum of 41 major Aeanthaster outbreaks have been documented on Indo-Pacific coral reefs during the period 1968-1988 (Birkeland and Lucas 1990).
DTL 18'70
18180
18190
GBRE PTBS SCUBA
19'00
,o? 2Luu i i u u U u u u u u u zm4 22
a
20 18
19'20
IdlO
19130
u u U uu' uu
19f40
DBML 19150
~u""
19'60
uu U
19'70
19~80
19~90
uU u U'-'u' 'LI u ud
Coral Reef Bleaching Events N= 6:5
16 14 12 I0 o3
(D
0 tO ,.13 .t.m
8 6 4
~5 o
.Q
, ,I
2 I
,4
b
I
II d o c u m e n t e d
I0
I~ u n d o c u m e n t e d
8
Z
6
I
I
I
I
Major Acanthaster Outbreaks1
12
E
I
N=48
I
N= 26
I!
4
,,i
2 I
I
I
I
I
I
I
I
I
I
I
.
84
12 I0
..
C I MaSSN=Coral108M~
~
I
8
6 54
4 2
25
29 m
8 I
2
3
4
61
5
9 I0
II
12
13 30
1870
1880
1890
1900
1910
1920
Fig. 1 a-e. Reported natural disturbances to coral reefs worldwide, 1870-1990. a Coral reef bleaching events are from Williams and Bunkley-Williams (1990) and Glynn (1991). The 1973 mass mortality event in American Samoa was preceded by extensive coral bleaching (Lamberts 1983 personal communication). The following milestones and respective sources denoted in the above plot are: DTL, Dry Tortugas Laboratory (Colin 1980); GBRE, Great Barrier Reef Expedition (Yonge 1930); PTBS, Palao Tropical Biological Station (Hatai 1937); SCUBA (Somers 1972); and DBML, Discovery Bay Marine Laboratory (UNEP/FAO 1985). E1 Nifio events, and their respective relative intensities and durations, are from Quinn et al. (1987) and Quinn (pers. comm.), b Major Aeanthaster outbreaks are tallied as documented (solid bars) and undocumented (cross-hatched bars) cases (after Moran 1986; Birkeland and Lucas 1990). The 1903 record in Thailand is after Dawydoff(1952). e Mass
19:50
1940
1950
1960
90
1970
1980
105
1990
coral mortalities resulted largely from physical stressors such as violent storms, heavy rains, seismic activity and sudden sea level drops, but also include biotic disturbances such as epizootics, Terpios sponge proliferation, and non-Acanthaster corallivore outbreaks. All mass coral mortality events are numbered chronologically by year. The numbers above the bars indicate the cumulative numbers of mass coral mortality events that occurred to the end of a given year, and also refer to the following sources: Coffroth et al. 1990 (1, 2, 5-8, 12-15, 20-23, 26-28, 30-33, 36, 37, 40, 43-45, 55, 73, 88, 91, 92), Brongersma-Sanders 1957 (3), UNEP/IUCN 1988 (4, 9, 18, 24, 25, 29, 34, 35, 38, 39, 42, 46, 47, 50-54, 58, 60-72, 74-86, 89, 90, 93-108), Stoddart 1969 (10, 19), Yonge and Nicholls 1931 (11), Goreau 1959 (16), Endean 1976 (17, 41, 48), Bryan 1973 (49), Brown 1987 (56), Loya 1976 (57), Moyer et al. 1982 (59), M. Yamaguchi, pers. comm. 1991 (87)
Fig. 2 a, b. Coral reef bleaching in the eastern Pacific Ocean and Caribbean Sea. a Bleached (white) and dead (green) pocilloporid corals on the seaward slope, 7 m depth, Uva Island reef, Panama, 28 April 1983. The King Angelfish total lengths are 20-25 cm. b Bleached Montastrea and other coral species, 6 m depth, Enrique reef, La Parguera, Puerto Rico, 5 December 1987
Fig. 4. Dead Millepora platyphylla Ehrenberg (arrow) surrounded by bleached and dead pocilloporid corals, Uva reef, 6 m depth, 28 April 1983, Gulf of Chiriqui, Panama. Since no live colonies of this hydrocoral species have been found following the 1982-83 ENSO, it is now likely locally extinct in the eastern Pacific region (Glynn and de Weerdt 1991)
Fig. 3 a, b. Bleaching of non-scleractinian coelenterates, a Clownfish sea anemone Heteractis crispa (Ehrenberg) bleached on a Sesoko Island fringing reef, 4 m depth, 21 September 1991, Okinawa, Ryukyus Islands, Japan. b Partially bleached Palythoa caribbea Duchassaing (right side) and Erythropodium caribaeorurn (Duchassaing and Michelotti) (left side), the latter surrounding the non-bleached scleractinian coral Mycetophyllia lamarckiana Milne-Edwards and Haime, Triumph Reef, 8 m depth, 1 November 1991, Florida Keys
Fig.5. Acapulco Damselfish, Eupornacentrus acapulcoensis (Fowler), territory on a dead patch of Gardineroseris planulata (Dana) resulting from the /982-83 E1 Nifio warming event, 4 m depth, Uva Island reef, Gulf of Chiriqui, Panama, 13 August 1985. A newly expanded patch of algal lawn replaces live coral tissue on the right side
In addition, these disturbances to coral reefs are occurring during a time of unusual, although not necessarily related, disturbances to other marine species, e.g. mass mortalities of sea urchins (Scheibling and Stephenson 1984; Lessios et al. 1984) and ulcerative syndrome of fishes (Tonguthai 1985; FAO 1986; Sindermann 1988), and see Williams and Bunkley-Williams (1990) for additional examples. Moreover, major terrestrial taxa, such as amphibians (Blaustein and Wake 1990) and fungi (Jaenike 1991), have revealed marked population declines in recent years. Finally, some of these large-scale disturbances (e.g. ulcerations and amphibian declines) have coincided with stresses from marginal or extreme natural environmental factors and with the concerns of global climate change (Singer 1989; Houghton et al. 1990; MacCracken 1990). Coral bleaching and mortality are caused by a variety of stressors (external factors or stimuli causing organismic responses, see Brown and Howard 1985) and the bleaching events reported before the 1980s were generally confined to relatively small areas or to certain reef zones (Coffroth et al. 1990). For example, localized bleaching was seen after torrential downpours and river discharges lowered near-shore salinities (e.g. Goreau 1964). Extreme low tidal exposures and sudden sea level lowerings caused coral bleaching and mortality on shallow coral reef flats (Fishelson 1973; Yamaguchi 1975; Glynn 1976; Loya 1976). And sudden temperature drops of reef waters due to atmospheric chilling or intense upwelling have been observed to induce coral bleaching and death (Glynn and Stewart 1973; Hudson 1981). Prior to the 1980s, most mass coral mortalities were related to non-thermal disturbances such as storms, aerial exposures during extreme low tides, and Acanthaster outbreaks (see Coffroth et al. 1990). Coral bleaching accompanied some of the mortality events prior to the 1980s during periods of elevated sea water temperature, but these disturbances were geographically isolated and restricted to particular reef zones (e.g., Yonge and Nicholls 1931). In contrast, many of the coral bleaching events observed in the 1980s occurred over large geographic regions and at all depths (Coffroth et al. 1990; Williams and Bunkley-Williams 1990; Glynn 1990, 1991). For example, in 1983 and 1987 coral reef bleaching affected respectively the entire equatorial eastern Pacific region to a maximum depth of 20 m and parts of most of the Caribbean Sea and adjacent waters up to 90 m depth (Fig. 2a, b). Some bleaching events (e.g. 8 of 13 during 1982-83, Coffroth et al. 1990) coincided with periods of elevated sea surface temperature (SST), when global atmospheric temperatures were abnormally high (e.g., Kerr 1988; Folland et al. 1990; Parker and Folland 1991). Since it is possible that two classes of disturbance are involved, we will distinguish between (a) small scale, generally isolated and acute bleaching events often with known causes, and (b) large scale, usually widespread and possibly chronic bleaching events with causes that are still unknown. The environmental conditions potentially responsible for the latter class of disturbances are considered below under "Environmental Correlates".
Because several reef-dwelling taxa (e.g. foraminifers, sponges, hydrocorals, sea anemones, alcyonarians, soft corals, zoanthids and tridacnid clams) besides scleractinian corals contain zooxanthellae or other kinds of symbiotic algae, and often lose these algae during bleaching events, it is appropriate to employ the generic term "coral reef bleaching" (sensu Williams and Bunkley-Williams 1990) to describe this kind of disturbance (Fig. 3a, b). Moreover, bleaching disturbances are not confined solely to symbiotic algae, but also affect some corals and sponges without endosymbionts (Williams and BunkleyWilliams 1990) and numerous other species that depend on live coral for shelter, food and other requisites (Glynn 1990). Indeed, if coral bleaching is caused by physicochemical state changes of the sea and atmosphere, then it is likely that such changes could also modify the biological and ecological activities (e.g., feeding, growth, reproduction, larval development and recruitment success) of many other reef taxa. Recent coral reef bleaching events Global distribution
Before I examine the geographic extent and effects of recent bleaching events it is important to underscore the nature of existing bleaching records. Certainly some bleaching events in remote regions must have gone unnoticed, and it is likely that some corals were reported to be bleached, but were in fact recently killed by predators, disease or some other agent. Also, some corals that appeared to have bleached may have been light-adapted, i.e. they contained a normal complement of zooxanthellae, but with a reduced concentration of photosynthetic pigments per algal cell (Jokiel 1988; Falkowski et al. 1990). Observations on bleaching are accumulating rapidly, yet no standardized method has been adopted to assess the degree of bleaching in the affected coral reef organisms. Some of the difficulties with standardizing the degree of bleaching are (1) the presence of genetically different types of zooxanthellae with possibly different pigment concentrations and environmental tolerances in different coral hosts (e.g., Blank and Trench 1985; Blank 1987; Rowan and Powers 1991 a, b), (2) the varying physiological and biochemical responses of identical types of zooxanthellae in different coral hosts and of different types of zooxanthellae in identical hosts (Trench 1971; Schoenberg and Trench 1976, 1980), and (c) the confusion arising from sibling species with different patterns of tissue coloration (Knowlton et al. 1992). Another problem concerns the areal extent of coral reef bleaching and the boundaries of the disturbance. Often these are not reported, making it difficult to compare the extent of different bleaching events. Nearly all of the world's major coral reef regions (Caribbean/western Atlantic, eastern Pacific, central and western Pacific, Indian Ocean, Arabian Gulf, Red Sea) experienced some degree of coral bleaching and mortality during the 1980s (Glynn 1984, 1991; Brown 1987; Coffroth et al. 1990; Williams and Bunkley-Williams 1990).
Even corals normally exposed to high summer temperatures (30-33 ~ C) also are susceptible to anomalous warming events. Early to mid-summer SSTs in Oman during 1990 reached 34-35~ and 39~ respectively along stretches of the southern Arabian (Persian) Gulf and Gulf of Oman, coinciding with widespread coral bleaching and mortality (IUCN 1990; Salm 1990). Some areas near the Strait of Hormuz experienced 90-95% coral mortality where high temperatures (32-33 ~ C) persisted to 10 m depth. No coral mortality was observed in the Strait of Hormuz (Arabian Gulf) where strong currents caused mixing and lowered SSTs. In only two relatively well studied coral reef systems has bleaching not been reported: Gulf of Eilat, a northern arm of the Red Sea (Y. Loya personal communication) and the barrier reef system in Belize (I.G. Macintyre personal communication). The extent of coral reef bleaching also is unknown in several other areas, including the Brazilian coral reef subprovince, New Guinea, the equatorial western Pacific warm pool (usually centered west of 180 ~ and northeast of Papua New Guinea, Philander 1990), the Philippines and western Australia, but it is uncertain if these areas were unaffected or incompletely observed. Most of the coral reef bleaching events of the 1980s occurred during years of large-scale ENSO activity (Glynn 1988 a; Jaap 1988). Four bleaching events were reported in the non-ENSO year of 1988: two occurred first in 1987 and continued into 1988 and two were confined to 1988. ENSO conditions known to cause coral bleaching and mortality include (a) sudden sea level drops resulting in reef exposures and reduced circulation, (b) low cloud cover, increased irradiance and warming of shoal reef waters, (c) high rainfall and lowered salinities, (d) largescale sea warming, and (e) calm seas with doldrum-like conditions. During ENSO events, conditions "a" and "b" often occur in the western Pacific, "c" and "d" in the central and eastern Pacific, and "e" in the western Atlantic. The strong association of such ENSO-generated stressors with bleaching is indicated in Fig. 1, which depicts ENSO events and their respective durations and intensities. A Fisher exact probability test demonstrated a significant relationship between ENSO years and coral bleaching during 1960-1990 (P=0.012)and 1980-1990 (P=0.030). Before 1979-80, coral reef bleaching was reported only during the 1963-64 and 1972-73 ENSO events even though ENSO events of moderate or stronger intensity occurred every 3.9 years on average from /8701990 (Quinn et al. 1987).
Range of responses to disturbances The well substantiated immediate effects of bleaching on coral hosts are decline in zooxanthellae density, loss of chlorophyll pigments, increase in respiration rate and declines in coral protein, lipid and carbohydrate (Glynn et al. 1985 a; Hoegh-Guldberg and Smith 1989; Kleppel et al. 1989; Porter et al. 1989; Ghiold and Smith 1990; Glynn and D'Croz 1990; Goreau and Macfarlane 1990; Jokiel and Cotes 1990; Szmant and Gassman 1990). In addition, three classes of non-lethal responses that might have longer-term, and possibly important effects, are (a)
diminished rates of coral growth and calcification, (b) impairment of reproduction, and (c) tissue necrosis. Reduced growth could decrease the capacity of corals to compete favorably for space with other reef benthos such as algal turf, coralline algae, macroalgae, sponges, bryozoans and tunicates. Benthic algae rapidly overgrew moribund and dead corals on reefs disturbed by the E1 Nifio warming event in Costa Rica (Cortes et al. 1984), Panama and the Galapagos Islands (Robinson 1985; Glynn 1990), and Indonesia (Brown and Suharsono 1990). The ratio of benthic algae/live coral cover has increased on some coral reefs for reasons apparently unrelated to bleaching, such as increased sedimentation (Cortes and Risk 1984; Rogers 1990), eutrophication (D'Elia et al. 1981; Birkeland 1988), and reduced grazing by herbivores (Hay 1984; Lewis 1986; Hughes et al. 1987). Any reduction in the growth rate of bleached corals would probably accelerate detrimental changes already in progress for other reasons. The diminished capacity of bleached corals to reproduce (Glynn and D'Croz 1990; Jokiel and Coles 1990; Szmant and Gassman 1990) could be expected to have a negative impact on coral recruitment, and this might be especially critical in annually reproducing species where the result would be missing an entire reproductive season. Even if reproduction is successful the settlement of coral planulae may be depressed by slightly elevated sea temperatures that are below the bleaching threshold of adult corals (Jokiel and Guinther 1978). Some bleaching events were coincident, or nearly so, with coral disease epizootics (Williams and Bunkley-Williams 1990). It is possible that diseased corals may be more prone to bleaching or that corals stressed by bleaching become more susceptible to various coral diseases. Regardless, diseased corals lose live tissue and when this results in colony fragmentation (partial mortality) remnant patches become more susceptible to continued mortality (Hughes and Jackson 1985). Further, if the size of surviving remnants is reduced below a critical minimum area then the coral's ability to reproduce sexually is lost until additional growth occurs (Kojis and Quinn 1985; Szmant-Froelich 1985). Since the different coral reef bleaching events of the 1980s were of varying severity, it is not surprising that coral mortality and recovery also varied greatly. If the intensity of a disturbance is not great, many corals will recover a few weeks after bleaching. In contrast, an intense disturbance can cause massive bleaching and the death of corals and other reef taxa (Glynn 1985 a, 1990; Glynn et al. 1985 a; Robinson 1985). Species differences in bleaching (see below) and the variable species composition of coral communities further complicate comparisons of coral bleaching reports. Several of the bleaching events of the 1980s were relatively minor with virtually complete coral recovery. In such cases coral bleaching may best be regarded as a physiological response that often is associated with seasonally high sea temperatures, UV-radiation or other conditions (Oliver 1985; Coffroth et al. 1990; Gates 1990; Hayes and Bush 1990; Williams and Bunkley-WiUiams 1990). Nonetheless, some bleaching events had catastrophic consequences. As a result of the 1982
83 ENSO (El Nifio-Southern Oscillation), eastern Pacific coral mortality ranged from 50% to 99% over the bathymetric range of all reef-building species (Glynn 1990), and coral mortality on some Indonesian reefs ranged from 80-90% in shallow reef flat habitats (Brown and Suharsono 1990). Numerous non-symbiotic species associated with eastern Pacific coral reefs, such as crustaceans and molluscs, also were affected in 1982-83 (Glynn 1985 a, 1990; Scott et al. 1988), particularly in the Galapagos Islands (Robinson 1985). Several major bleaching events were reported from all over the world during 1986-88, but few quantitative data on the mortalities of corals or other reef species have been published (Williams and Bunkley-Williams 1990). In many instances fast-growing, branching coral species with high metabolic rates were among the first to bleach and die (Brown and Suharsono 1990; Glynn 1990; Jokiel and Coles 1990; Williams and Bunkley-Williams 1990). However, some colonies of a fast-growing Caribbean coral [Aeropora palmata (Lamarck)] did not bleach at six locations in 1987 when other slower-growing species were bleaching (Williams and Bunkley-Williams 1990). Some massive coral species did not bleach, bleached just partially or bleached only near the end of a particular warming event (Glynn 1990; Williams and Bunkley-Williams 1990). Zoanthids (Palythoa spp.) and reef-building hydrocorals (Millepora spp.) often bleached first in both the Caribbean and Indo-Pacific regions, and alcyonaceans [Cladiella pachyelados (Klunzinger), Lobophytum crassum Marenzeller, Sinularia ?polydactyla (Ehrenberg)] showed early bleaching responses in the Indo-Pacific region (Williams and Bunkley-Williams 1990; personal observation in Okinawa). A few especially sensitive scleractinian corals and a hydrocoral (Fig. 4) suffered local extinctions from coral reefs in the eastern Pacific (Glynn 1988b, 1990) and one reef-building hydrocoral (Millepora boschmai de Weerdt and Glynn) apparently became globally extinct following the 1982-83 warming event (Glynn and de Weerdt 1991; de Weerdt and Glynn 1991). Not surprisingly, intraspecific differences in susceptibilities also were apparent. On Panamanian reefs, bouts of bleaching were observed with a patchy distribution among monospecific stands of pocilloporid corals (Glynn 1984). That is, multiple colonies of single species inhabiting the same habitat were affected differently (Fig.2a). Hoeksema (1991) reported notable intraspecific differences in extent of bleaching among mushroom corals during the 1983 sea warming event in Indonesia. In Hawaii in the summer of 1988 colonies of Porites compressa Dana that bleached were apparently clone-mates of a genotype that may be extremely sensitive to sea warming (C. Hunter and R.A. Kinzie III, cited in Jokiel and Coles 1990). It is reasonable to assume that various taxa that are nutritionally dependent upon reef corals or other symbiotic hosts must suffer soon after their hosts are affected. In only a few cases have such collateral disturbances been demonstrated. The gastropod predators and crustacean symbionts of pocilloporid corals died shortly after their prey or hosts, respectively, had perished
(Glynn 1985 a, b; Glynn et al. 1985 a), and flatworm parasitic biomass production declined in a montiporid coral host exposed to high temperature stress (Jokiel and Coles 1990). Strong ENSO events can cause diverse disturbances resulting in the mortality of both symbiotic and non-symbiotic reef organisms. Examples of such impacts ranged from prolonged reef-flat exposures (Yamaguchi 1975), to reduced circulation in reef lagoons (Perez in Glynn 1984), physical storm damage (Robinson 1985; HarmelinVivien and Laboute 1986), and reduced food supplies for suspension feeding populations due to nutricline depression (Robinson 1985; Glynn 1990). Intense upwelling following ENSO activity also was associated with coral bleaching and mortality in the eastern Pacific in 1985. Corals in Panama died during strong upwelling pulses (Glynn and D'Croz 1990) and in Costa Rica during massive dinoflagellate blooms (Guzman et al. 1990). Delayed or long-term effects resulting from the 198283 E1 Nifio warming event in the eastern Pacific have continued to the present (1991). These have involved (a) altered corallivore foraging patterns and the disruption of spatial refugia, (b) increased predator concentration relative to declining prey populations, (c) a shift in live cover dominance to algal turf, macroalgae and other non-reef building taxa, (d) increases in sea urchin abundances on dead reef frameworks, leading to severe bioerosion (Glynn 1988 c, 1990), and (e) the establishment and enlargement of damselfish lawns on massive corals that experienced partial mortality (Fig. 5). From preliminary findings, Scott et al. (1988) claimed that the recruitment of infaunal bioeroders (boring mussels and sponges) also has increased dramatically following E1 Nifio-induced coral mortality. Bioerosion now exceeds coral framework production on several eastern Pacific coral reefs (Glynn 1988 c; Eakin 1991). Since this process can probably critically influence reef growth turn-on, i.e. the initiation of framework development (Buddemeier and Hopley 1988), it will be re-examined below under "Coral Reef Recovery". Environmental correlates
Several potential stressors, e.g. salinity, light, temperature, sedimentation, aerial exposure, xenobiotics and epizootics, have been implicated convincingly in causing coral bleaching in particular instances (Brown and Howard 1985; Glynn 1990; Williams and Bunkley-Williams 1990). However, the apparent widespread coral reef bleaching episodes of the 1980s require causes of greater spatial magnitude (Fisk and Done 1985; Harriott 1985; Oliver 1985; Hoegh-Guldberg and Smith 1989; Glynn and D'Croz 1990; Jokiel and Coles 1990; Lesser et al. 1990). Increased sea temperatures and solar radiation (especially UV radiation), either separately or in combination, have received consideration as plausible largescale stressors. In most instances, wherever coral reef bleaching was reported, it occurred during the summer season or near the end of a protracted warming period (Williams and Bunkley-Williams 1990; Glynn 1991).
Many workers report that coral bleaching occurred during periods of low wind velocity, clear skies, calm seas and low turbidity (e.g., Causey 1988; Jaap 1988; Lang 1988), when conditions favor localized heating and high penetration of short wave length (UV) radiation (Jerlov 1968; Smith and Calkins 1976; Fleischmann 1989). Also, less oxygen is held by water at higher temperatures. Potentially stressful high sea temperatures and UV radiation flux could conceivably cause coral reef bleaching on a global scale with suspected greenhouse warming (Hansen and Lebedeff 1987; Tsonis and Elsner 1989) and the thinning of the ozone layer (Fredrick and Lubin 1988; Blumthaler and Ambach 1990). Many tropical marine organisms, including reefbuilding corals and associated species, live near their upper thermal tolerance limits (Moore 1972; Vernberg and Vernberg 1972; Johannes 1975; Coles et al. 1976). Small increases in sea temperature (0.5-1.5 ~ C) over several weeks or large increases ( 3 4 ~ C) over a few days will lead to coral dysfunction and death (Glynn and D'Croz 1990; Jokie! and Coles 1990). Because anomalously high sea temperatures often were reported in the Caribbean-wide series of coral reef bleaching events that occurred during 1986-88, some workers hypothesized that global warming was having an effect on the coral reefs in this region (Williams and Bunkley-Williams 1990; Goreau et al. in press). Ancient coral reef biotas also existed during periods of high, near-lethal sea temperatures (e.g. Fagerstrom 1987) and it has been suggested that widespread extinctions of early reef species occurred during maximum global warm intervals (Fischer and Arthur 1977; Thompson and Newton 1988). Superficially such hypothesized, ancient disturbance events bear some resemblance to contemporary coral reef bleaching disturbances. However, critical comparisons must consider not only the occurrence and magnitude of past warm intervals, but also the rates of temperature change. A further complication is the potential interaction of temperature change with other factors, such as anoxic waters, nutrient levels, and solar radiation flux (Fischer and Arthur 1977; Emiliani et al. 1981). The frequency and kinds of disturbances affecting contemporary coral reefs seem to be changing over decadal intervals; as a result, comparisons with disturbances of ancient reef assemblages are highly speculative. A large-scale analysis of the relationship between coral reef bleaching and positive SST anomalies in the tropical western Atlantic, based on observations from ships and buoys, failed to disclose significant sea warming during major coral bleaching in 1987 (Atwood et al. 1988). In contrast, blended high resolution satellite SST data have shown that coral reef bleaching occurred during significant temperature increases at 5 of 7 Caribbean sites in the 1980s (Goreau et al. in press). However, the latter study has not been substantiated by other satellitederived data sets (Atwood and I-Iendee in press) or by long-term SST data obtained from ships offshore (Bottomley et al. 1990; Lang et al. in press; J.D. Elms, personal communication). Indeed, for the period 1950-1989 a significant cooling trend, which varied in magnitude from -0.1 to - 0 . 9 ~ C per 100 years, occurred at 13 of
15 areas analyzed from ship data obtained in the Caribbean Sea and Gulf of Mexico. Perhaps contributing to this discrepancy are the different instruments employed, the spatial and temporal differences in sampling protocols, the time intervals selected for time series analyses, and the quality of the research. Finally, organismic responses are complex and are influenced not only by prevailing temperatures, but also by thermal history (i.e., preceding temperature regimes, including rates of change, and duration) and likely other interactive physical and biological factors (Coles and Jokiel 1978; Glynn et al. 1988; Glynn and D'Croz 1990). Solar ultraviolet radiation is potentially harmful to numerous shallow-water, tropical marine taxa (Jokiel 1980; Worrest 1982), including reef corals and their symbiotic zooxanthellae (Jokiel and York 1982; Siebeck 1981, 1988; Lesser and Shick 1989, 1990). UV-B (280320 rim) and UV-A (320400 nm) radiation can readily penetrate clear sea water (Jerlov 1968; Smith and Calkins 1976). UV (300400 nm) radiation has been measured to 25 m depth on a Jamaican coral reef, and was found to range between 20-25% of surface values to 10 m depth (Fleischmann 1989). Relatively high levels of UV-B and UV-A radiation, about 50% of surface values, were measured between 3 4 m depth on an inshore coral reef in Okinawa (Fig. 6). Reef-building corals contain UV-absorbing compounds ("S-320", Shibata 1969), a family of mycosporine-like amino acids (MAA) that are capable of blocking potentially damaging UV radiation (Dunlap and Chalker 1986). These compounds are produced in response to ambient UV levels (Jokiel and York 1982) and the concentration in corals is usually an inverse function of depth (Dunlap et al. 1986). Because natural UV flux and the concentration of MAAs were not measured at any one site during the coral reef bleaching events of the 1980s, it is not known if bleaching responses were related to variations in UV flux that could have exceeded the protective capacity of UV-absorbing compounds. Adding to the complexity of the problem is a possible interactive effect between temperature and UV radiation. Increases in temperature significantly reduced zooxanthellae densities and also the concentration of UV absorbing compounds in a reefzoanthid, thus potentially increasing the exposure of the symbionts to the direct effects of UV radiation (Lesser et al. 1990). Finally, the perceived deteriorating condition of the world's coastal marine environment (Rogers 1985; Brown 1987; Birkeland 1988; Sindermann 1988; UNEP/ IUCN 1988; Williams and Bunkley-Williams 1990) could have a global impact on coral reefs and interact with high SST and UV radiation to negatively affect coral reef organisms. Increased activities by man, such as dear-cut logging, land clearing, coastal development, agricultural and landscape fertilization, fishing, sewage disposal, pesticide use, and accidental chemical spills, have all caused unequivocal damage to coral reefs on local scales. These activities commonly affect corals directly through siltation, eutrophication and direct toxic effects (Loya and Rinkevich 1980; Risk et al. 1980; Smith et al. 1981; Pastorok and Bilyard I985; Tomascik and Sander 1985,
Mean S u r f a c e Irradiance (W crn -2] 0--&-"--
PAR UV A UV B
sea surface
i
4.67.10 -2 9.34-10 -4 7 . 9 8 - 10- 6
i
I
u
I
I
I
I
Ag
0.2
I 1 i.lai/ (3)
I.I /II
E /
c-
i
T:
/ /"
(3)
9 /" 9
/i j i
i
i i9
/ 1"0
i I
IO
20
I 30
I
40
[
t 60
I
80
I
[00
Log Percent Irradiance at 0.2 Meters Fig.6. Logarithmicdeclineof photosyntheticactiveradiation(PAR, 417-761 nm, half power points in spectral response plot), UV-A (356-367 nm) and UV-B (290-301 nm) radiation on a nearshore coral reef, Motobu-cho, Okinawa. Measurementswere made with an IL 1700 Research Radiometer(InternationalLight, Inc.) on 24 July 1990, 1105-1147 under clear sky and lightlyrippled sea surface conditions. Each regression line was calculated from 5 measurements made at each of 6 depths (n= 30) with the respectivemean values indicated. Mean percent and absolute irradiance values immediatelyabovethe sea surfaceare indicatedat the top of the graph. UV-B absorptionwas significantlygreaterthan PAR and UV-A (T' method for comparisons of regression coefficients, e = 0.05, Sokal and Rohlf 1981) 1987; Jackson et al. 1989; Cortes 1990). Also, indirect effects favoring non reef-building taxa (e.g. benthic algae, suspension feeders, bioeroders and predators) will suppress reef growth over the long term. A most difficult task is to demonstrate that these kinds of local impacts are having large-scale and possibly global effects on coral reefs. The complexity of the problem, and very likely the seriousness of the threat, would be much greater in the face of global temperature rise and increased UV radiation flux.
Coral reefreeovery Measuring coral reef recovery (Pearson 1981; Colgan 19 87) and understanding its meaning in the context of incessant community change (Grigg and Maragos 1974; Connell 1978; Grigg 1983; Hughes 1989; Brown and Su-
harsono 1990; Done et al. 1991) are not straightforward exercises. Evidence from major natural disturbances, such as cyclones, volcanic activity, extreme low tidal exposures, low temperature stress, and outbreaks, suggests that the rate of coral reef recovery often is related to the severity and scale of the disturbance, but varies greatly with location. Rapid recovery, i.e. the restoration of live coral surface cover over decadal periods, has been observed in many areas where the presence of coral survivors nearby could serve as seed populations facilitating recruitment (Endean and Stablum 1973; Loya 1976; Pearson 1981; Colgan 1987). Low rates of recovery may occur in coral communities that experience unusually frequent disturbances or continued predation on remnant coral prey (Loya 1976; Rinkevich and Loya 1977; Pearson 1981; Endean and Cameron 1990; Knowlton et al. 1990; Bunkley-Williams et al. 1991). Coral reef recovery also can be influenced by the reproductive mode of surviving corals. The presence of brooding species with high reproductive activity has lead to rapid recruitment to disturbed areas (Grigg and Maragos 1974; Hughes 1985; Smith 1991). Coral recovery may also occur rapidly via asexual fragmentation, but recovery will be slow if recruitment is dependent upon sexual reproduction, especially from distant populations (Highsmith 1982; Harrison and Wallace 1990). Reductions in population size, range restrictions and local extinctions also can affect coral species diversity and coral reef recovery (Pearson 1981). For example, if disturbances destroy large corals that contribute to the topographic complexity of reefs, then their loss would eliminate habitats for reef-associated biota with resultant declines in species diversity. In addition, since coral recruitment is favored by the presence of firm, stable substrata, disturbances that eliminate reef frameworks or convert them to rubble must interfere with recruitment and recovery. Most studies of coral reef bleaching have concentrated on the recovery of individual coral colonies in terms of their biomass and metabolic states (e.g., Kleppel et al. 1989; Porter et al. 1989; Glynn and D'Croz 1990; Jokiel and Coles 1990; Szmant and Gassman 1990). Only a few workers have reported on the longer term changes in coral communities following partial or complete colony mortality. One of the first such studies, carried out on the Great Barrier Reef complex, Australia, found notable recovery within 2 years (Fisk and Done 1985). The continued growth of acroporid and pocilloporid corals, and hydrocorals that survived the 1982 bleaching event quickly increased coral cover, by about 36% overall at 8 m depth. In the Java Sea, Brown and Suharsono (1990) reported lower rates of recovery on reefs that experienced high mortality (80-90%) during the 1982-83 ENSO warming event. Five years after the disturbance, coral cover had increased to about 50% of its former level, but with marked differences in rates and patterns in two initially similar reef flat communities separated by only 2.5 km. At one site, recovery was rapid and species diversity increased, while at another site, recovery was delayed and diversity did not change. The recovery pattern at the latter site was probably due to a predominance
Acanthaster
(Millepora)
of Acropora species, which suffered disproportionately high mortalities, and the survival of a Montipora species, which grew rapidly and increased in abundance following the disturbance. Further, the death and breakage of branching Aeropora corals produced an unstable rubble surface that was not suitable for coral recruitment. Almost a decade after the 1982-83 ENSO sea warming event caused extensive coral reef bleaching and mortality in the tropical eastern Pacific, ongoing studies reveal relatively slow recovery in severely impacted regions. The recruitment of corals onto reefs that experienced 5070% coral mortality in continental Costa Rica and Panama has been erratic and slow. Coral reefs that experienced 95-99% mortality in the Galapagos Islands and at Cocos Island (Costa Rica) are not experiencing significant coral recruitment (Glynn 1990; Guzman and Cortes in press). Moreover, many of the most severely impacted reefs are undergoing rapid bioerosion from the destructive grazing activities of sea urchins (Glynn 1988 c; Eakin 1991). If recovery is defined as the replacement of 100 to 300 year old reef frameworks and massive coral colonies, then full community restoration will probably not occur for several hundred years (Glynn 1990).
The future
Sea temperature rise Heuristically, it is worth considering two scenarios regarding coral reef bleaching and its possible links to sea warming: (I) the sea warming events of the 1980s were the result of an episodic warming trend that will soon end or (2) recent sea warming is a result of global warming and will increase 1-2~ on average by AD 2030-2050 (Mitchell 1988; Manabe et al. 1991). If incidents of tropical high sea temperatures decrease in frequency during the 1990s, then the frequency of coral reef bleaching events also should decrease. Most of the world's coral reefs that experienced bleaching with minor mortality would be expected to sustain their growth. However, some coral colonies that suffered partial mortality (tissue loss and size reduction) will probably be more susceptible to future disturbances and the fecundity of some corals may be reduced at least temporarily (Szmant and Gassman 1990). Coral reef bleaching from causes other than increased sea warming (e.g., increased sedimentation, nutrients, toxic pollutants and UV radiation) may well continue to increase with the accelerating impacts of man on our planet's atmosphere and waters generally and on tropical coastal environments in particular. Tropical marine organisms seem to be especially vulnerable because of their narrow environmental tolerances (Moore 1972; Vernberg and Vernberg 1972; Johannes 1975; but see Brown and Howard 1985, and Grigg and Dollar 1990 for opposing views) and the large-scale landscape alterations that have occurred at low latitudes in recent years (Salvat 1987; U N E P / I U C N 1988; Yamazato 1988). If a global warming trend impacts on shallow tropical and subtropical seas, we could expect an increase in the frequency, severity and scale of coral reef bleaching.
Coral mortality could exceed 95 % regionally with species extirpations and extinctions, as occurred in the tropical eastern Pacific during the 2-4 ~ C positive thermal anomalies that accompanied the 1982-83 ENSO event (Glynn 1988b, 1990; Glynn et al. 1988; Scott et at. 1988; Glynn and de Weerdt 1991). Secondary, long-term disturbances, such as corallivore concentration on surviving corals, bioerosion, destabilization of reef substrata arid loss of coral frameworks, would also be expected to follow. Disturbances of such scale and magnitude would greatly slow recovery processes. Other types of benthic communities, perhaps dominated by opportunistic, temperature-resistant algae and suspension feeders, would replace contemporary coral reef communities. Coral reefs subject to high nutrient inputs, such as in the eastern Pacific (Highsmith 1980) and Arabian Sea (Sheppard and Salm 1988), could be at greater risk than reefs in oligotrophic waters. Modelling results of global scale sea temperature change provide some indication of possible warming trends in relation to particular coral reef regions (Fig. 7). Responses of a general circulation model of the coupled atmosphere-ocean-land surface system to a gradual increase in atmospheric CO2 of + 1% per year (compounded) indicate that the upper mixed layer (0-50 meters) mean ocean temperature, between 30~ and 30~ could increase by 1.5 ~ C around AD 2030 (Manabe et al. 1991). A marked interhemispheric asymmetry is predicted with the greatest sea temperature increases in the temperate/subpolar northern hemisphere (2-3 ~ C), and relatively lower increases in tropical and temperate/ subpolar southern hemisphere latitudes (1 2 ~ C). The coupled ocean-atmospheric model indicates that all coral reef regions between approximately 25~ and S latitudes would experience temperature increases of between 1 and 2 ~ C. In addition to the model runs with gradual change in atmospheric CO2, Manabe et al. (199 i) also show the results from a steady state computation, one in which the coupled ocean-atmosphere system has an infinite amount of time to respond to the doubled concentration of CO2 in the model's atmosphere. This simplified model assumes no changes in the horizontal or vertical heat transport by the oceans. The comparison between the steady state and transient results allows a rough estimation of the lag effects inherent in the transient computation due to the incorporation of the realistic ocean model and slow increase in the atmospheric concentration of CO2. Considering the more conservative temperature increases of 1-2 ~ C, several coral reef regions between 2030~ would experience sustained warming that falls within the lethal limits of most reef-building coral species. Especially at risk would be mid- to high latitude coral reefs in the western Pacific (Bonin Islands, Volcano Islands, Daito Islands, Ryukyu Islands, Taiwan, Pescadores Islands), central Pacific (northwestern Hawaiian Islands), and western Atlantic (Bermuda, Bahama Islands, Florida Keys). Some Indian Ocean coral reefs at 15-30~ latitude (Madagascar, Mascarene Islands) would also experience relatively high sea temperatures. The relatively low resolution of the G F D L model does
10 - -
CENTRAL PACIFIC WESTERN PACIFIC 151.9 E
163.1
I
2
155.6 W
148.1
I
2
-
Coupled
Ocean-Atmospheric
Model
........
Coupled
Atmospheric-Mixed
Layer
Ocean
Model
3~
3~
WESTERN 80.9 W
73.1
ATLANTIC I
2
3~
50 40
I 40
I I 30
I INDIAN
30
OCEAN
65.6 E
76.9
2
3"C
20
EASTERN
PACIFIC
t ! L#
.2
20
95.7 W
88,1
I
2
3~
#:
i{ Ii
I0
I ! I I t i 1 I I
0
! I0
20
20 -
-
20
2O
Austral Iss 30
30
3 0 ' - 30
3O 40
4 0 - -
40
50
50
Kergue[en
5 0 - -
.~ill Is$
5O
Fig. 7. Meridional sections through selected coral reef regions illustrating projected sea temperature differences from 1958 in the upper ocean layer (mean depth = 25 m) computed for a doubling of atmospheric CO2 content with two different adjustment scenarios9 The dashed curves indicate the response of a coupled ocean-atmosphere model to a gradual increase of 1% (compounded) CO2 per year, and represent a twenty year average centered on the doubling time, AD
not exclude the possibility of regional differences of larger amplitude. Finally, it is cautioned that other modelling results forecast lower rates of global warming than that considered here, but still very high rates that would be five times those that have occurred during the past century (Houghton et al. 1990; Wigley and Raper 1992). These synoptic predictions fail to account for various physical oceanographic processes that m a y counteract or
20309 The dotted curves show the equilibrium warming for doubled CO 2 concentration produced by a simplified version of the coupled model that assumes no changes in the horizontal and vertical heat transports by oceanic currents. From the NOAA Geophysical Fluid Dynamics Laboratory (GFDL) coupled ocean-atmospheric climate model (Manabe et al. 1991)
depress sea warming on local scales (Andrews and Pickard 1990). One such meso-scale process is the "island mass effect", which has been shown to cause turbulence and vertical mixing on the leeward sides of islands subject to strong current flow. This effect has been observed in the Hawaiian Islands (Gilmartin and Revelante 1974), Marquesas Islands (Sournia 1976), and the Seychelles (Sorokin 1990).
11
Sea level rise Some workers have suggested that sea level rise that would accompany global warming might initially favor vertical reef flat accretion when coral growth could keep pace with the flooding (Buddemeier and Smith 1988; Hopley and Kinsey 1988). This prediction is based on a globally averaged most probable sea level rise of 15+_3 ram/year (Hoffman et al. i983). Predictions of coral reef growth responses to sea level rise are complicated, however, by the high susceptibility of important reef-building coral species to the sea warming events observed in the 1980s (Williams and Bunkley-Williams 1990; Glynn 1991). It is probable that sustained elevated sea temperatures that would accompany sea level rise would suppress coral growth or kill many reef flat corals before they could respond to reef flooding. If coral growth is retarded, it may be more susceptible to the destructive effects of corallivores and bioeroders that would probably not be affected by higher temperatures. Compared with the mass coral mortalities in Panama caused by the 1982-83 ENSO warming event, most corallivores, herbivores, and bioeroding sea urchin populations remained at pre-1983 abundances or increased in size after that disturbance (Glynn 1985 a, 1988 c, 1990). Coral reefs at critical high latitudinal (Grigg 1982) and deep water locations (Grigg and Epp 1989) might be the first to drown. Rising sea level may also interfere indirectly with a pole-ward expansion of coral reefs in response to more favorable thermal conditions. For example, increased sedimentation, nutrient loading and light attenuation, caused by the flooding and erosion of continental areas, could potentially inhibit reef growth over large areas (Neumann and Macintyre 1985; Hallock and Schlager 1986). However, coral reefs removed from continental influences, such as atolls and those present on oceanic islands, might be spared from such impacts. Finally, since projected sea level rise is globally averaged, we can expect varying effects under different geologic settings. In some tectonically active areas that experience uplift, such as southeast Asia, the Ryukyu Islands, and in the south Pacific (Pirazzoli 1991), reef building would probably continue during rapid sea level rise, at least in the absence of other disturbing effects.
Coral adaptation In spite of present limited understanding of the genetic structure of coral populations, and the mechanisms and rates of speciation of reef corals (Potts 1985; Potts and Garthwaite 1991), some available information may provide clues concerning the ability of corals to adapt to sudden increases in sea temperature, solar radiation and other impacts related to global climate change. This brief discussion will be limited to projected sudden increases in sea temperature, of the order of 1-2 ~ C in the upper mixed layer tropical ocean over the next 40-50 years (Stouffer et al. 1991). As reviewed by Jokiel and Coles (1990), individual coral colonies living in high temperature environments can survive and photosynthesize at temperatures a few
degrees higher than their congeners in lower temperature environments. Depending upon the area, sustained temperatures of 30~ can be tolerated for several weeks (Coles et al. 1976; Coles and Jokiel 1977) or 32-34 ~ C for several days to a few weeks (Coles 1988). Moreover, some individual corals have the capacity to acclimate physiologically to higher temperatures (Clausen and Roth 1975). However, as noted earlier in this review, even warm-adapted corals will succumb to excessive and prolonged periods of warming, as occurred in the Arabian Gulf and Gulf of Oman in 1990. If sea warming suddenly occurs, then corals (or their symbiotic zooxanthellae or both) must evolve rapidly to cope with this environmental change. The magnitude of such evolutionary responses could range from major changes, including permanent differentiation of populations to form new species, down to local physiological, morphological and/or ecological adaptations without taxonomic significance (Potts and Garthwaite 1991). Paleontological evidence suggests that many coral species may be incapable of rapid evolution. The fossil record demonstrates that numerous coral species have persisted since Pliocene or early Pleistocene times. The average age of Indo-Pacific and Caribbean species is approximately 20 my (Frost 1977; Veron and Kelley 1988). Some coral species are known only from Holocene (_
