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Assessing the Effects of “Stress” on Reef Corals

B. E. Brown and L. S. Howard

Department of Zoology, University of Newcastle upon Tyne Newcastle upon Tyne, England

I. Introduction . . . . . . . . . . . . . . . . . . . . 1 11. Natural Fluctuations and Man-made Influences . . . . . . . . . . 3

A. Assessing changes on coral reefs . . . . . . . . . . . . 3 B. Interpreting temporal changes on coral reefs . . . . . . . . 5 C. 9 D. Predicting recovery of reefs . . . . . . . . . . . . . . 17

111. Experimental Studies on Effects of Pollutants on Corals . . . . . . 20 A. Growth rate . . . . . . . . . . . . . . . . . . 20 B. Metabolism . . . . . . . . . . . . . . . . . . 27 C. Loss of zooxanthellae . . . . . . . . . . . . . . . . 29 D. Behavioural responses . . . . . . . . . . . . . . . . 35 E. Reproductive biology . . . . . . . . . . . . . . . . 46 F. Histopathology . . . . . . . . . . . . . . . . . . 48 G. Biochemical and cytochemical indexes . . . . . . . . . . 50

IV. Discussion and Future Research Needs . . . . . . . . . . . . 51 References . . . . . . . . . . . . . . . . . . . . 55

Effects and apparent lack of effects of pollution on coral reefs . . . .

1. Introduction

Some years ago Johannes (1975) published the first major literature review on the effects of marine pollutants on coral reefs. At that time he

1

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2 B . E. BROWN A N D L. S. HOWARD

highlighted the paucity of knowledge in many areas of the subjects. Al- though research efforts in the field have increased, particularly with respect to potential effects of pollution by oil (Loya and Rinkevich, 1980) and drilling muds (Dodge and Szmant-Froelich, 1984) there are still enor- mous gaps in our knowledge and serious contradictions in the existing literature. Much of this lack of information may be attributed to our limited understanding of the physiology of corals, although recent papers have contributed valuable data on growth (Highsmith, 1979), reproduction (Highsmith, 1982; Kojis and Quinn, 1981, 1982; van Moorsel, 1983), behavior patterns (Lasker, 1979), carbon turnover (Crossland ef ul., 1980b), calcification (Barnes and Crossland, 1978, 1982; Gladfelter 1982a), mucus production (Crossland et al., 1980a), and associated bacterial populations on living corals (Ducklow and Mitchell, 1979).

The aims of the present article are to consider long-term ecological studies in the light of known effects of disturbances on coral reefs and to ask whether the effects of disturbances can be distinguished from long- term fluctuations on the reef and also where other difficulties lie in assessment of pollution in the field. In addition, in an attempt to improve understanding of the overall susceptibility of reef corals to marine pollution, an assessment is made of the responses of corals to stress and methods by which these responses have been monitored to date.

The definition of “stress” has been much discussed in the literature (Grime, 1979; Pickering, 1981; Stebbing, 1981; Rosen, 1982), particularly with reference to problems involved in identifying and quantifying such a condition. Working with plants, Grime (1979) defined stress as the external constraints limiting dry matter production by all or part of the vegetation, while Rosen (l982), considering corals, described stressful conditions as those resulting in restricted growth and reproduction. Such specific definitions as these have not always been applied in many of the publications referred to in this article. For our purposes Rosen’s broader view of stress as a gradient between ideal conditions and the ultimate limits of survival will be adopted. As noted by Stebbing (1981), the term stress may be used as a cause or as an effect. In agreement with Stebbing and most other authors, we choose to view stress as an external force or stimulus.

The article is divided into three sections, the first section dealing with observations in the field, the second with laboratory assessment of pollutant effects, and the third section incorporating a general discussion of the validity of generalizations made to date on the overall vulnerability of coral reefs to man-made disturbance.

EFFECTS O F STRESS ON REEF CORALS 3

II. Natural Fluctuations and Man-made Influences

A. Assessing Changes on Coral Reefs

Assessing change necessarily implies that reefs are monitored regularly by standard, repeatable methods. It is only in recent years, however, that quantitative methods have become routinely employed on permanent transects over time intervals. The first review of field methods applied on coral reefs was published by Stoddart in 1972. Subsequently other workers (Loya, 1972, 1978; Done, 1977; Bouchon, 1983; Dodge er al., 1982) have successfully tested various quantitative and semiquantitative methods on the reef. Strictly quantitative techniques vary from plotless (Loya, 1978) to quadrat methods (Bak and Engel, 1979), and more recently workers have compared both approaches at the same sites in an attempt to gauge their relative efficiencies (Dodge et al., 1982; Bouchon, 1983). Al- though there appear to be no major differences in the results obtained by either methods, there are variations in the quantity and type of information generated and the time required for use; the line transect or “inter- sected length” method generally makes the most efficient use of the time spent underwater.

Measurements made on coral reefs using these techniques include coral cover, diversity, evenness (Loya, 1972, 1976a; Brown and Holley, 1982; Dodge et af . , 1982; Bouchon, 1983), colony number and colony size (Loya, 1972; Fishelson, 1977), and more recently spatial complexity (Rog- ers er af . , 1982, 1983) and spatial arrangement of both living and dead substratum components (Bak and Luckhurst, 1980).

Although monitoring of coral cover and diversity may yield fundamental information about coral assemblages, few studies incorporate measurements of the cover and diversity of other components of the coral community, such as soft corals, zoanthids, algae, sponges, and ascidians. Recently, the importance of monitoring these groups has been highlighted by the publications of Benayahu and Loya (1977), Bak er af. (1981), and Tursch and Tursch (1982). Invasion and/or overgrowth of scleractinian corals by many species of soft corals (Nishihira, 1981; Tursch and Tursch, 1982) and ascidians (Bak er al., 1981; Sammarco er al., 1983; Sya’rani, 1983) has frequently been observed in both the Indo-Pacific and the Carib- bean provinces.

In studies where dominant components of the coral community have been considered (Rogers er af . , 1982), the effect of disturbance on coral reef diversity may be complex. Diversity of scleractinian corals as a result of hurricane damage in St. Croix was shown to decrease in shallow reef

4 B . E. BROWN A N D L. S. HOWARD

zones, whereas diversity of the community as a whole actually increased because of colonization of new substrata by a wide variety of reef organisms, e.g., algae, sponges, tunicates, bryozoa, and hydroids. Clearly quantitative measurements on coral reefs affected by disturbance should include some account of all major components of the reef community.

Such measurements are also improved by an appreciation of the structural complexity of the coral reef environment. Rogers et al. (1982) used a modified transect method, with a linked chain following the contours of the reef, to obtain an index of reef topography or structural complexity. Done (1981) has applied the use of stereophotography to permanent transects on the Great Barrier Reef. A stereo pair of photographs provides a great resolution of detail, a means of determining the three-dimensional coordinates of colonies and substratum, and a means of determining true dimensions and shapes of benthic organisms at any depth in the photo- graph. With automated stereoanalysis it should be possible to accurately map three-dimensional growth patterns of living corals and/or surface area of other substratum components. So complex are the interactions on coral reefs (Bak et al., 1982; Porter et al., 1982) that standard measurements of areal coverage, diversity, and abundance may not always be sensitive to changes in interactions which would be detected in a three- dimensional approach to community analysis.

Such an analysis should also consider the nonliving components of the substratum. Bak and Luckhurst (1980) have highlighted the importance of monitoring not only living cover but also nonliving substrata such as rock and sediments. Their study showed that alteration of spatial arrangement through dislodgement and collapse of substrata and changes in sediment flow were of paramount importance in describing the community, particularly in shallow-water (10- and 20-m) quadrats. As the authors note, a continuous change in the cover of nonliving components must have serious implications for the settlement and survival of juvenile ben- thos.

One further factor should be considered when assessing changes on coral reefs, and that is the measurement of colony size. This parameter has been used by various workers (Loya, 1972, 1967a; Fishelson, 1977) involved in monitoring the effects of disturbance on coral reefs. It has recently been recognized (Hughes and Jackson, 1980) that partial colony mortality, colony fission, and colony fusion may affect any simple relationship between the size and age of reef corals. Following known corals in photographs for successive years demonstrated that, in foliose Carib- bean corals, size and age are seldom related. Measurements such as those of Fishelson (1977) on age groups of faviids from polluted and nonpolluted sites, estimated from size dimensions, may require reinterpretation in the

EFFECTS OF STRESS ON REEF CORALS 5

light of these more recent studies. Loya (1976a), however, recognizes the regenerative ability of corals when interpreting the effects of low tides at Eilat, and stresses that in the few cases where corals did not fully regener- ate the separate parts were considered as one individual.colony. His results suggested that whereas before the low tide coral colonies on the control reef fell into relatively large size categories, in 1973 after the catastrophic low tide, colonies fell into small size categories. Corals with marked regenerative ability included Cyphastrea microphthalma (La- marck), Pauona decussata (Dana), Millepora dichotoma (Forskal), and Porites lutea (Milne-Edwards and Haime). Generally the recovery of the control reef was mainly due to recolonization by coral planulae rather than regeneration of survivors. Nevertheless, regeneration of corals after partial mortality is an important process on all reefs, and it may be very difficult to decide if a small coral has recently settled or whether it is actually part of a much larger colony which has suffered partial colony mortality or colony fission. Such difficulties may be compounded in polluted areas, particularly those suffering from a high sediment load (personal observation) (Fig. 1). Clearly this aspect requires further study on reefs affected by sedimentation where the growth form of massive species such as P. lutea and Goniastrea retiformis (Lamarck) appears nodular and where partial colony mortality is high.

Many long-term monitoring programmes incorporating techniques described earlier have now been initiated on coral reefs in both the Carib- bean and the Indo-Pacific, and much interesting information should gradually become available over the next decade-to quote Lewis (1976), considering long-term ecological surveillance on temperate rocky shores, “to record ‘change’ is no problem. There is much and it would be a remarkable investigation that showed none. The major need is to ensure that the change recorded is real and relevant.”

B. Interpreting Temporal Changes on Coral Reex7

Table I demonstrates major long-term changes observed as a result of mainly natural disturbances, while Table I1 records instances of man- made damage on coral reefs. It is clear from these tables that recent regular monitoring of fixed stations and transects in CuraGao (Bak and Luckhurst, 1980) and Eilat (Loya, 1976a) have produced interesting data on coral distributions and their spatial distributions with time. In addition, surveys before and after damaging natural events such as hurricanes (Stoddart, 1974; Shinn, 1976; Rogers et al., 1982), low temperatures (Shinn, 1976), and low tides (Loya, 1972) provide some insights into reef recovery and development.

TABLE I. LONG-TERM SURVEILLANCE OF NATURAL DISTURBANCES ON REEFS

Time Environmental Site span history Major changes observed Reference

British Honduras 1964- 1966

Heron Island, Aus- 1963-1970

Key Largo, Florida 1950-1965 tralia

Key Largo, Florida 1965-1967

St. Croix, U.S. 1978-1979 Virgin Islands

Gulf of Eilat 1970

Humcane damage

Hurricane damage

Hurricane damage

(1961)

(1966)

(1960)

Repeated hurricane damage (1965)

Humcane damage ( 1979)

Catastrophic low tide

Branching corals more susceptible than massive species

No marked change in coral abundance

Although colonies broken and much destruction within 1 year, difficult to recognize damage; by 1965 damage completely healed

Damage not noticeable by 1967

Effect of humcanes complex-may result in reduction in coral diversity but increase in community diversity due to provision of more light for slower growing corals and new substrate for algae and other invertebrates

Change in the community structure with rare species affected

Stoddart (1974)

Connell (1973)

Shinn (1976)

Shinn (1976)

Rogers er a / . (1982)

Loya (1972, 1976a)

Qatar, Persian Gulf

Dry Tortugas, Flor-

St. Croix, U.S. ida

Virgin Islands

Carysfort Reef, Key Largo, Florida

CuraGao

John Brewer Reef,

Discovery Bay, Australia

Jamaica

1965-1967 Low temperatures

1881-1976 Thermal shock

1976-1979 Bacterial infection (1976-1977)

March- No obvious natural November disturbances 1975

1973-1978 No obvious natural disturbances

1976-1980 No obvious natural

1976-1980 Humcane Allen disturbances

(1979)

Regeneration of Acroporu sp. after chill; 2 years later colonies 2-20 cm high

Little change in area occupied by hermatypic corals; major changes were in coral species distributions

Death of Acropora palmara as result of “white band disease” caused decrease in structural complexity of reef surface, decrease in living coral tissue and a reduction in CaC03 deposition on reef

suggested decline over 14-month study period Estimates of net recruitment and mortality of reef corals

Cover of living and nonliving components relatively constant throughout study; major differences lay in spatial arrangement of substrate components

Net increase in colony abundance with a peak in recruitment in 1979

A trend of reduction in number of rarer coral species on the reef was reversed by Humcane Allen with storm-induced mortality being greatest in the most abundant species (Acroporu spp.)

Shinn (1976)

Davis (1982)

Gladfelter (1982b)

Dustan (1977)

Bak and Luckhurst (1980)

Done (1981)

Porter et a/ . (1981)

1 0.1 m i FIG. I . Portion of transect on intertidal reef affected by sedimentation where new recruits and colonies affected by partial mortality

and subsequent regeneration are difficult to distinguish (Brown, unpublished).


Studies of strictly long-term changes on coral reefs, however, are limited to those of Dahl and Lamberts (1978) and Davis (1982), who reas- sessed transects established 56 and 85 years before these recent surveys, respectively. In Dahl and Lamberts’ study in American Samoa, where results of dredging and cannery effluents were suspected of exerting influences on Arua Reef between 1917 and 1973, the total number of coral heads decreased by 28% during this 56-year period. Although the same genera were dominant in 1917 and 1973, the relative proportions of each differed in 1973 from that recorded in 1917. These authors conclude that the status of the reef in 1973 may reflect gradual recovery, or alterna- tively, a reef subject to intermittent “stress/sfresses.”

Although little change (6%) in the overall area occupied by corals was noted in Dry Tortugas between 1881 and 1976 (Davis, 1982) and in Cura- qao between 1973 and 1978 (Bak and Luckhurst, 1980), a major difference in the distribution and spatial arrangement of major coral species was observed in both studies during these time intervals. In the Dry Tortugas (Davis, 1982) in 1976 a lush Acropora ceruicornis (Lamarck) reef occupied what had been an octocoral-dominated region in 1881, while a considerable area of Acropora palmata (Lamarck) on the reef crest in 1881 was reduced to 600 m in 1976. During the relatively short time span of 5 years at Curaqao (Bak and Luckhurst, 1980) the combined effects of settlement, growth, dislodgement, and death of corals, coupled with variations in sedimentation, resulted in considerable temporal instability of both living and nonliving components.

Generally shallow reefs show less short-term stability and lower pre- dictability than deep reefs (Loya, 1976a; Bak and Luckhurst, 1980), the latter study demonstrating considerable change in spatial arrangement of substratum components and less stability at depths of 10-20 m when compared with deeper sites at 30-40 m. During the 5-year period of their study, Bak and Luckhurst comment on the constancy of coral cover in the reef habitats studied as compared with the shallow reef at Heron Island, Australia, where the area covered by living coral varied by a factor of 2-3 over 7 years of study (Connell, 1973). Similar variability was recorded by Glynn (1976) in Panama, where coral cover was observed to decrease by a factor of 2.5 in 15 months.

C. Effects and Apparent Lack of Effects of Pollution on Coral Reefs

Table I1 documents selected examples of recent studies on the effects of man-made disturbances on coral reefs. While studies on the effects of chronic oil and mineral pollution (Fishelson, 1973; Loya, 1975, 1976a), thermal pollution (Jokiel and Coles, 1974), sewage (Walker and Ormond,

TABLE 11. EFFECTS, OR LACK OF EFFECTS, OBSERVED I N THE FIELD FROM “SUSPECTED” POLLUTION

Time Pollutant Site scale Species Reference

“Suspected” man-made influences (possibly dredging and cannery effluent)

Sedimentation from change in ag- ricultural practice

Dredging and increased sedimenta. tion

Dredging

Dredging Kaolin clay spill

Offshore oil drilling

Arua Reef, Samoa

Low Isles, Austra- lia

Castle Harbour Bermuda

Piscadera Bay, Cu- raGao

Diego Garcia Hawaii

N.W. Palawan Is- land, Philippines

191 7- 1973

1928-1978

1974

1972

1980

1981

Acroporu continued to be important throughout; Psummucoru reduced 213. Pocilloporu increased 5 x , Poritc7.s trndrcwsii reduced (three genera missing in 1973-Merrtlinrr. Goniopor~r , Cyplirist r-eu )

rians very abundant Reef flat with few surviving corals but holothu-

Diploria strigoscr more susceptible to sedimentation than Diplorio lubyrintirifortnis; most susceptible would appear to be Steplian- coeniu tnichilini

result of inability to reject sediment; calcification rates of Modrueis mirubilis and Agori- ciu oguricites decreased by 33% over 4-week period at least

Porites ustreoides (plating form) died as a

Coral diversity unaffected by dredging Corals survived discharge, although some Po-

cilloporrr meondrinu were temporarily bleached: coral cover in area dominated by P . meundrinu and Porites lobatu

Massive species (e.g., P . Irrteu) appear to have survived in preference to branching species (Poci l /oporo , A croporu) which showed est i- mated 70-9096 reduction in nonstained area around wellheads

Dahl and Lamberts (1978)

C. M. Yonge, personal communication

Dodge and Vaisnys ( 1977)

Bak (1978)

Sheppard (1980) Dollar and Grigg

(1981)

Hudson ct 01. (1982)

Experimental shading San Cristobal Reef, S. W. Puerto Rico

Chronic oil pollution (and mineral Gulf of Eilat dust)

Chronic oil pollution

Thermal pollution

Sewage pollution

Gulf of Eilat (Na- ture Reserve)

Kahe Point, Oahu, Hawaii

Gulf of Aqaba

1979

1966-1972

1966- I972

1971- 1972

A . ceruicornis most susceptible to partial shading; A . uguricires, Montustreu unnuluris, D. labyrinthiformis, Siderustreu sidereo, and Colpophylliu nutuns show bleaching and variable recovery; no visible response in Eitsmiliu fustigiutu, Montustreo cuuernosa, or Mussu ungulosa

AcroPorU Seriutoporu

Stylophora Milleporu dichotomu: unchanged Nature reserve colonized after low tide by:

These genera were all reduced I P . luteu (16 colonies); M. dichotomu (8 CO~O-

nies); C . microphrhulmu (7 colonies); Fu- via favites ( 5 colonies); Acanthastreu echinata (2 colonies); Stylophoru pistillatu and C . microphtholmu suffered

Order of resistance to high T" from field observations: strongly resistant-~epra.~rreu purpurea, Porites compressu, P . lobatu, Monti- poru pufula, and Montiporu uerrucosu; least resistant-P. meandrinu

Only surviving coral species was S . pistillutu at polluted site, while control site displayed Fu- via spp., Fuvites spp., Seriutoporu hystrix, Pocilloporu dunue, and S . pistillutu

Rogers (1979)

Fishelson (1973)

Loya ( 1976a)

Jokiel and Coles ( 1974)

Walker and Or- mond (1982)

(continued)

TABLE 11. (CONTINUED)

Pollutant Time

Site scale Species Reference

Recreational activities

Rotenone derivative (fish-collect- E. Sambo Reef, 1973-1974 Octocorals apparently less susceptible than Jaap and Wheaton ing chemical) Florida scleractinians; for the rotenone derivative, A . (1975)

ceruicornis was more resistant than A. pulrnura, S. siderecc, D. strigo~ci. or Dicho- coeniu stokesii; for quinaldine (generally less toxic to all scleractinia) A. qyiricites proved to be the most susceptible to the chemical

cies A . ceruicornis, A . pcclnzutrc, and P. pori- fes: Millepora highly susceptible; generally soft corals suffered more than scleractinians

Biscayne National 1977-1980 Scleractinian damage greatest in branching spe- Tilmant and Park, Florida Schmahl (1983)

Heavy metal pollution and sedi- Intertidal reef Rats, 1979 No apparent effect upon coral diversity and Brown and Holley mentat ion Phuket, Thailand coral cover at site affected by heavy metal 1982; Brown (un-

pollution, although increased incidence of partial mortality of faviids suspected at this site

published)

EFFECTS OF STRESS O N REEF CORALS 13

1982), fish-collecting chemicals (Jaap and Wheaton, 1975), dredging (Dodge and Vaisnys, 1977; Bak, 1978), and recreational activities (Til- mant and Schmahl, 1983) clearly show the impact of these disturbances on corals in the field, there are also a number of studies which demonstrate apparent lack of serious damage as a result of man-made interference. Such studies include the limited effects of a major spill of 2200 tons of kaolin clay on a reef in the Hawaiian Islands (Dollar and Grigg, 1981), of elevated metal levels from tin smelting and tin dredging activities on intertidal reef flats at Phuket, Thailand (Brown and Holley, 1982), of dredging activities in Diego Garcia Lagoon (Sheppard, 1980), and of drilling muds in the Palawan Islands, Philippines (Hudson et al., 1982). It may be argued that application of more detailed and longer term survey techniques may yet reveal subtle changes in the community structure at these sites. Nevertheless, no major deterioration in reef structure was evident in any of these examples.

The possible reasons for this apparent lack of effect have been documented by the authors concerned. In the case of the kaolin spill, factors contributing to the lack of extensive damage were cited as the nontoxic nature of the kaolin, the small particle size of the clay, the presence of a wetting agent, and the rapid dispersal of the kaolin plume. In addition, rapid removal of sediments by coral cleansing aided the recovery process in the majority of coral colonies which were lobate and branching and hence less likely than platelike varieties to suffer heavy mortality (Dollar and Grigg, 1981).

Apparent lack of effect of tin smelting and tin dredging processes at Phuket were ascribed to the possible reduced “biological availability” of toxic metals to corals, the general tolerance of intertidal reef species to stresses, and the possible acquisition of specific metal tolerance mechanisms by the corals themselves (Brown and Holley, 1982).

At Diego Garcia dredging during the last decade was probably short term and limited in extent, any resulting damage being overcome by rapid recovery (Sheppard, 1980).

Limited damage to branching corals only (an area 115 X 85 m2) was recorded in the vicinity of wellheads around Palawan Island, but the authors (Hudson et al., 1982) conclude that drilling mud probably consti- tutes a minor threat to coral growth under the conditions described in the study.

In terms of overall tolerance of reef corals to disturbance in the field, the literature contains several references inferring the likely ability of intertidal and shallow-water corals to withstand physical stresses (Ed- mondson, 1928; Loya, 1972; Kojis and Quinn, 1981). Some authors also suggest that reef flat corals differ from deeper water species not only in

14 B. E. BROWN A N D L. S . HOWARD

increased physical tolerances, such as exposure to elevated temperatures (Jones and Randall, 1973), but also in reproductive strategies which they have evolved to reduce planktonic life to a minimum and to retain larvae on the reef (Stimson, 1978). Not surprisingly, then, it would appear that corals from shallow-water environments are more likely to be tolerant of environmental extremes, a finding reflected by Hudson (1981) on transplanting Montastrea annularis (Ellis and Solander) from a deep-water to a shallow-water location in the Florida Keys. The transfer resulted in severely reduced growth rates and mortality in one case. In contrast, transplantation of inshore M. annularis to offshore sites produced only slightly reduced growth rates compared with resident colonies. Should such increased environmental tolerances exist in shallow water corals, then there may be parallels with estuarine species in temperate ecosystems which have been described as “preadapted” to pollution stresses (Jones, 1975; Reeve et al., 1976). Indeed, Jokiel and Coles (1974) describe the Caribbean reef coral Siderastrea siderea (Ellis and Solander) as a “hardy estuarine coral” capable of establishing itself within a zone of maximum thermal effect around a thermal power plant in Florida during the winter and spring months of 1971.

Although there may be some overall pattern in increased tolerances to stress shown by corals from different reef habitats and even varying geographical locations (Coles and Jokiel, 1977, having demonstrated lethal temperatures for Enewetak corals to be 2-5°C higher than for Hawaiian corals), Table I1 highlights the variability in response of different corals to the same stress at any one individual site.

Considering, for example, temperature effects, Mayor (1914, 1918) was the first worker to note the ability of coral species to resist high temperatures in laboratory tests was inversely related to their metabolic rate. Since this date, Jokiel and Coles (1974) have camed out field observations in Hawaii, and their results confirm that this generalization also holds in the field-the most temperature-resistant coral being the large polyped species Leptasirea purpurea (Dana) with a low metabolic rate, and the most sensitive, Pocillopora meandrina (Dana) with a high metabolic rate. Recent observations by Neudecker (1981) in Guam showed Pocillopora damicornis (Linnaeus) to be more sensitive to thermal stress than Porites andrewsii (Vaughan), the most sensitive species being Acropora formosa (Dana). In the latter study, differences in temperature tolerance were considerable, colonies of P . andrewsii surviving up to 77 days in elevated (4-6°C above ambient) temperatures, whereas A . formosa generally died within 2 days.

Whereas Pocillopora is cited in many studies as being relatively sensi-

EFFECTS OF STRESS ON REEF CORALS I5

tive to stress in the form of increased temperatures, some authors (Dahl and Lamberts, 1978) describe P . dumicornis in American Samoa as “most tolerant of adverse conditions, being found near shore where there is silt and fluctuating water temperatures.” Since Mayor’s (1924) original transect at this site in 1917, P . dumicornis had increased fivefold, while P. andrewsii, once dominant in the midzone at Arua transect, was considerably reduced. Factors suspected of affecting the reef included dredging activities and discharge of cannery waste. Increased sedimentation resulting from drilling processes at Palawan Island, Philippines (Hudson et al., 1982), caused branching corals (including Pocillopora sp.) considerable mortality when compared with head corals such as P . futeu ; Dollar and Grigg (1981) also describe short-term effects of a kaolin spill affecting P . meandrinu but cite no short-term damage in P . lobata as a result of sedimentation.

Clearly, then, tolerance of Pocilloporu to either increased temperature or sedimentation, as described in Dahl and Lamberts’ study (1978), would appear to be at variance with observations of the above authors. It is, of course, acknowledged that there are probably few cases where a single factor is responsible for damaging effects observed in the field. Increased sedimentation, for example, will present at least three problems to exposed corals. These are decreased light values, increased energy-consum- ing processes such as sediment cleansing, and possibly reduction in planktonic food (Bak, 1978). One feature which does appear to be consistent throughout most studies involving sedimentation and/or shading is the particular susceptibility of branching corals to these stresses when compared with massive species. Considering shading alone, such observations would be in line with Porter’s (1976) conclusions that branching corals with small polyps may depend more upon light than upon planktonic capture and so are less able to withstand reduced light intensities than massive corals (Rogers, 1979), though evidence for such resource partitioning is now questioned (Rosen, 1982). With respect to sedimentation, branching corals are very effective in passive rejection of sediment because of their colony morphology (Hubbard and Pocock, 1972; Bak and Elgershuizen, 1976), and as Rogers (1979) showed, the branching Carib- bean coral A . ceruicornis was unaffected by daily exposure to sediments. Plating colonies of Porites ustreoides (Lamark), however, which are reported as inefficient sediment rejectors (Bak, 1978), were unable to reject sediments resulting from dredging activities at Curasao and either wholly or partially died.

Damage due to sedimentation summarized in Table I1 may therefore be variously interpreted. Where branching corals are observed to have suf-

16 B. E. BROWN AND L . S. HOWARD

fered (Dollar and Grigg, 1981; Hudson et al., 1982) it may be speculated that light levels were sufficiently reduced to induce damage. Where plating varieties, e.g., P . astreoides, suffered more than branching species, e.g., Madracis mirabilis (Duchassaing and Michelotti) (Bak, 1978), it is likely that lack of sediment rejection capabilities proved to be more lethal than reduction in light intensity and subsequent zooxanthellae loss. As Jokiel and Coles (1977) point out, zooxanthellae loss may be temporary and reversible, bleached corals recovering within 2 months of return to normal conditions.

In considering the response of a branching coral such as Pocillopora to temperature increase or sedimentation, it becomes apparent that generalizations must be applied with care when attempting to predict the response of an individual species to a pollutant or indeed the effect of a pollutant on a reef community.

One further example will serve to demonstrate the need for more critical data on responses of corals to pollution. Stylophora pistillata (Esper) has been described as an opportunistic or “weedy” species, a typical colonizer of unpredictable environments (Loya, 1976a,b) and polluted habitats (Walker and Ormond, 1982). The latter study demonstrated that S. pistillata was the only surviving coral species on a reef flat in Aqaba affected by sewage and phosphate pollution. Fishelson (1973) also showed that Stylophora was apparently relatively resistant to oil pollution and phosphate dust at Eilat, the coral representing 47% of all branching species in 1966 and 73.7% of the sample in 1968. The combined effects of a low tide and chronic oil pollution, however, severely reduced recolonization by S. pistillata at the nature reserve at Eilat, the coral previously having been dominant at this site (Loya, 1975). In subsequent surveys, P . lutea proved to show more successful recolonization than Stylophora, with a relatively large number of small-sized colonies being recorded. P . lutea, however, could hardly be described as an opportunistic species; according to Highsmith (1982), the life history characteristics of P . lutea include a high growth rate, large adult size, a long life expectancy, but no apparent release of larvae. So, in contrast to Endean’s speculation that opportunistic species should be well represented among early colonizers of polluted habitats (Endean, 1973), in this instance-the only detailed study of its kind to date-such a hypothesis does not appear to hold. The reproductive biology of the adult and the settling behaviour of the larvae of Stylophora are detrimentally affected by exposure to oil (Rinkevich and Loya, 1977, 1979a), but no similar information is available for P . lutea, so it is impossible at this stage to say where the observed tolerance in the latter species lies.


D. Predicting Recouevy of Reefs

Short-term phenomena on reefs are highlighted in Davis’ study (1982) on Dry Tortugas, with the destruction of 90% of an extensive A. ceruicornis stand in 1976-1977 as a result of lowered seawater temperatures in Janu- ary, 1977. A . cervicornis, however, is known to show rapid recovery rates after partial destruction (Shinn, 1976). A growth rate of 10 cm/year, combined with a geometrical progression of branch formation, has been attributed to the short-term recovery (5 years) of A. ceruicornis reefs after storm damage in the Florida Keys (Shinn, 1976). Speedy partial recovery of Acropora sp. in Qatar, Persian Gulf, 2 years after death due to lowered seawater temperatures has also been reported by Shinn (1976). Highsmith (1982) concludes that S ~ W recovery occurs when disturbance is so severe that hardly any fragments of reef-building corals survive and when survival depends upon sexual reproduction, rapid recovery ensuing when asexual reproduction and regeneration are possible. Such conclusions are supported by the long-term recovery (10-20 years) of A. cervicornis in Belize after suffering high mortality as a result of Hurricane Hattie (Stod- dart, 1974) and the relatively short-term recovery of A. cervicornis in studies of limited reef damage in Florida cited earlier (Shinn, 1976).

In other instances, however, high rates of recruitment have been attributed to rapid recovery of reefs (Loya, 1975, 1976a), but care must be taken in noting the time scale of such “rapid” recovery. Initial recolonization of reef flats at Eilat between 1969-1973 was shown to be 23X greater on a control reef when compared to a reef affected by chronic oil pollution (Loya, 1976a), although by 1973 there was still a significantly lower coral cover on the control reef compared with the initial survey in 1969. It is interesting to note that the dominant corals on reef flats at Eilat, e.g., Stylophora and Cyphastrea, are species cited as showing life histories in which sexual reproduction is predominant over asexual reproduction, and therefore possibly a slower recovery when compared with corals reproducing primarily by asexual methods (Highsmith, 1982). Corals with high rates of recruitment have been observed to be among the most common species on submerged lava flows in Hawaii, where 20-50 years were required for recovery (Grigg and Maragos, 1974), and at Heron Island, Australia (Connell, 1973). Such a correlation was absent in CuraGao and Bonaire (Bak and Engel, 1979), where common species such as S . siderea, M . annularis, and Montastrea cauernosa (Linnaeus) had very few recruits. The authors suggest that such a finding may indicate a higher level of environmental disturbance on the examined shallow reefs in the Pacific when compared with those studied in the Caribbean.

18 B . E. BROWN AND L . S. HOWARD

The importance of considering recruitment rates as only part of the overall life history strategy of a coral when attempting to explain distribution and abundance patterns has been shown in recent studies (Bak and Engel, 1979; Bak and Luckhurst, 1980). Two of the most common corals on the reef slopes of Curaqao are Agaricia agaricites (Linnaeus) and M . annularis. Agaricia displays a high recruitment rate, a low rate of survival, and a high mortality, whereas Montastrea combines a low recruitment rate, moderately good survival, and low mortality (Bak and Engel, 1979).

Clearly, then, life history strategies of individual corals are all important in determining the rate of recovery of a reef after disturbance. Inter- pretation of data is further complicated by species exhibiting different life history features in different geographical locations. Highsmith (1982) describes P . darnicornis in Hawaii as a fugitive species, competitively sub- ordinate, with a low growth rate, small adult size, and noted for production of planulae larvae. However, in Panama it is the major reef builder, competitively dominant, with a high growth rate and described as rarely showing recruitment of planulae. Contrasting characteristics are similarly shown in geographically isolated Porites haddoni (= P . lutea), which is reported as planulating from January to June at Low Isles, Australia, by Marshall and Stephenson (1933); no planulation, however, has been observed in P . futea at Enewetak (J. S. Stimson, personal communication, in Highsmith, 1982).

As mentioned earlier, the depth of the reef may be important in assessing damage and also in predicting recovery. Hurricane damage has been reported to depths of 20 m (Highsmith et al., 1980; Luckhurst in Bak and Luckhurst, 1980), although damage is generally considered to be greatest in shallow waters (see Endean, 1971, 1973, for reviews). Shallow-water habitats, however, may not always show the greatest effects of hurricanes, as recent studies demonstrate. Rogers et al. (1982) describe the number of broken branches of A . palmata per metre as decreasing with depth (0.6-6.1 m) as a result of hurricane damage at St. Croix, U.S. Virgin Islands, but whereas shallow branches broke only at their distal ends, exposing relatively little surface area for healing, deeper branches broke at their bases and consequently exposed much greater surface areas for healing and recolonization. Generally the larger branches took longer to heal, the healing process being more effective in fractured small branches which predominated in shallow water.

Theoretical aspects of recovery of coral reef communities devastated by catastrophic events have previously been reviewed by Endean (1973). At that time Endean speculated that in cases of extreme disturbance, opportunistic species with a high fecundity might be expected to be well


represented among the early colonizers; that substrata for colonization might be successfully invaded by benthic organisms other than corals such as algae and alcyonarians; that mortality of juvenile corals might be high; and that coral growth could be retarded, so increasing the time taken for the newly established coral to grow to maturity. Considering that Endean published his review 10 years ago, it is interesting to reflect how his speculations have been borne out by recent work. Earlier discussions in this article would suggest that while opportunistic species with a high fecundity might be found in polluted environments (Loya, 1976a; Walker and Ormond, 1982), they are not necessarily well represented among early colonizers in certain polluted conditions (see Section II,C of this article).

Colonization of substrata by algae rather than corals has been observed in areas receiving sewage discharge such as Kaneohe Bay (Smith, 1977) and Aqaba (Walker and Ormond, 1982). Although algal growth was greatly stimulated in polluted areas at Aqaba, the authors maintained this factor was not the direct cause of coral death. They suggested that en- hanced algal growth, stimulated by increased nutrient concentrations, may have acted as a sediment trap, thus exposing corals to a considerable sediment load. Other workers (Benayahu and Loya, 1977), looking at reef flats affected by periodic low tides at Eilat, have shown that resulting mass mortality of benthic communities opens up new spaces for settlement. Such unpredictable disturbances are believed to prevent potential dominant competitors from monopolizing the available space, the observed coexistence of stony corals, soft corals, and algae being due to different environmental tolerances and competitive abilities of each group. More recently, useful papers on space monopolization by some of the less well-known groups such as the soft corals have appeared in the literature (Samrnarco et af. , 1983; Tursch and Tursch, 1982).

Studies on the mortality ofjuvenile corals are still limited. The work of Connell(l973) and Bak and Engel (1979) suggests that on reefs unaffected by human disturbance, approximately 36 and 32% of the juvenile corals died or disappeared during the 1 I-month and 6-month study periods, respectively. Bak and Engel cited causes of mortality as sedimentation and competition from coralline algae on shallow reefs and possibly random grazing and/or direct predation by parrot fishes on the reef slope. Another third of the juvenile population in this study were described as limited in growth by factors such as spatial competition, which was similar at all depths. Mortality of juvenile corals in polluted environments is unknown, apart from related work by Rinkevich and Loya (1977) on the effects of crude oil on planulae and juveniles of S. pistillata. Effects observed in the laboratory included a decrease in the viability and successful settlement


of the planulae, which was manifest in the field by a limited recolonization by this coral at the polluted Nature Reserve at Eilat.

It is clear that we are still lacking much fundamental information on aspects of recovery and recolonization of reefs (Pearson, 1981). From the limited knowledge we have gained during the last 10 years, it would seem that both generalizations and predictions are dangerous and that until more evidence is forthcoming we should consider each case individually. Recent work by Grigg (1983) suggests that disturbance is a primary mechanism governing diversity, community structure, and succession of coral reefs in Hawaii. Furthermore, Grigg depicts the effects of disturbance occurring at different stages of successional processes on coral reefs and concludes that in Hawaii, reef community structure is primarily a function of the interaction between disturbance and recovery time.

111. Experimental Studies on Effects of Pollutants on Corals

The tolerance of scleractinian corals to factors such as increased temperature and sedimentation was first studied over 50 years ago by Mayor (1914, 1918), Edmondson (1928), and Marshall and Orr (1931). These early workers established the broad tolerances of a variety of coral species to physical disturbances which might be encountered in the field. Experiments were largely performed in the laboratory, where mortality was used as a measure of tolerance, though Mayor and others (Vaughan, 1915) saw the value of actually transplanting corals to a variety of environments and measuring growth rate and survival in siru as a reflection of environmental quality.

Despite the short-term nature of laboratory experiments and their shortcomings, a variety of responses have been monitored, both in the laboratory and experimentally in the field, by exposure of corals to a wide selection of chemical and physical parameters (Tables 111-VIII). These responses are discussed below.

A. Growth Rate

Growth rate of corals has been cited as one of the best quantitative measures of testing stress due to a disturbance since this parameter integrates a variety of physiological processes (Birkeland et al., 1976; Neudecker, 1983). It is also widely accepted, however, that coral growth rates may be inherently variable (Buddemeier and Kinzie, 1976; Barnes and Crossland, 1982) for a single species within reef zones (Gladfelter et al., 1978) and


even within individual colonies (Rogers, 1979; Brown et al., 1983). Glad- felter et al. (1978) have described some species as “conservative” in their growth whereas others are not. They cite M . annularis as showing relatively little response in growth rate to varying environmental conditions, while A . ceruicarnis shows marked variations under similar circumstances.

Methods employed to measure growth rates of corals have been reviewed by Buddemeier and Kinzie (1976) and Gladfelter et al. (1978). Table I11 illustrates that the majority of studies, in which growth rate has been used as a parameter to measure the effect of disturbance, have involved either x-radiography (Dodge and Vaisnys, 1977; Hudson, 1981, 1983; Hudson and Robbin, 1980; Hudson et al . , 1982), reference marking by alizarin red S stain or a fixed base line (Shinn, 1976; Rogers, 1979; Dodge, 1982; Bak and Criens, 1983; Neudecker, 1983), measurement of an increase in skeletal weight (Jokiel and Coles, 1977; Bak, 1978), or 4SCa deposition rate in the skeleton (Neff and Anderson, 1981).

The use of x-radiography, in the above context, has been applied solely to massive corals such as P . lutea (Indo-Pacific) and M . annularis, Diplo- ria strigosa (Dana), and Diploria labyrinthiformis (Linnaeus) (Caribbean). Significant suppression of coral growth as a result of disturbance has been observed using this method during short-term exposure of M . annularis to “extremely high” concentrations of drilling mud (Hudson and Robbin, 1980). Inhibition of coral growth was also obtained on transferring M . annularis from an offshore location to a more stressful inshore site (Hud- son, 1983) and in D . strigosa and D . labyrinthiformis as a result of dredging in Bermuda (Dodge and Vaisnys, 1977). No suppression of growth was observed in M . annularis and P . lutea as a result of bombing activities at Vieques, Puerto Rico (Dodge, 1983) and drilling processes off N.W. Palawan, Philippines (Hudson et al . , 1982), respectively. Results of the latter study, however, showed a 70-90% reduction in area coverage of branching coral species around the wellhead. Death of these low-profile corals was believed to be due to smothering by a prolonged and localized buildup of cuttings, surviving corals being primarily massive head corals in an elevated position above the bottom.

Such differences between apparent tolerances of branching and massive species of coral emphasize the need for sensitive methods of assay.

In long-term growth studies of M . annularis from the East Flower Gardens, Texas (Hudson et al., 1982), where exploratory drilling sites have been set up in recent years, and also at sites within the Key Largo coral reef marine sanctuary, Florida (Hudson, 1981), a decline in growth rates has been observed. However, in both studies the authors cannot directly attribute apparent growth suppression to any single environmen-

TABLE 111. THE USE OF GROWTH RATE TO ASSESS THE EFFECTS OF DISTURBANCE ON CORALS I N THE LABORATORY A N D EXPERIMENTALLY I N THE FIELD

Criterion: Nature of study growth rate and location Species Disturbance Results Reference

As determined Analysis after field collec- by x-ray tion, N.W. Palawan, analysis Philippines

Analysis after application of drilling mud by divers to transplanted specimens, Florida Keys, USA, and after field collection, Texas

Analysis after transplanting specimens, Florida Keys, USA

Analysis after field collection, Castle Harbour, Bermuda

Analysis after field collection, Vieques, Puerto Rico

Analysis after field collection, Florida Keys, USA

P. lutea

M . annularis

M . annularis

D . strigosa and D . labyrinrhifiormis

M . annularis

M . annularis

Drilling activities

Drilling muds

Change of habitat from offshore to inshore site

Dredging activities

Military bombing activities

Increased dredge and refill operations (?)

Little apparent suppression of growth due to drilling

growth shown (though other factors may be responsible)

Possible decrease in

Reduced growth rate and deposition of dense skeleton

death Decline in growth prior to

Apparent lack of effect

Decline in coral growth (1953-1968) at some midshore and inshore reef sites

Hudson et a / . (1982)

Hudson and Robbin (1980)

Hudson (1983)

Dodge and Vaisnys (1977)

Dodge (1983)

Hudson (1981)

As determined by alizarin S stain

As determined by skeletal growth from a baseline

As determined by weight of skeletal growth deposited

Analysis after transplanting corals to site of thermal enrichment, Guam

Experimental quadrats in the field, CuraCao

Laboratory experiment

Analysis after shading and application of sediments to corals in experimental channels and on the reef, San Cristobal, Puerto Rico

Laboratory experiment using corals from a Hawaiian reef

P . andrewsii, P . damicornis, A. formosa

M . mirabilis, A . palmata, A. ceruicornis

M . annularis

A. ceruicornis

P . damicornis, Montipora uerrucosa, Fungia scutaria

Thermal pollution 4-6°C above ambient

Fragmentation

1, 10, 100 ppm drilling mud doses for 6 weeks

Experimental shading and sedimentation (receiving up to 800 mglcm2 and 200 mgicm2 once a day, once a week, and once a month during 40- day period)

Temperature increase

Coral growth impeded by higher temperature; suggests slower growing species more tolerant of high temperature than faster growing species

M . mirabilis grew significantly more slowly after fragmentation

Skeletal extension de- clined significantly in 100 pprn treatment

Shading significantly affected growth rate but no observed effect on growth rate as a result of exposure to sedimentation

Exposure of corals to temperatures of 30°C reduced calcification

Neudecker (1983)

Bak and Criens (1981)

Dodge (1983)

Rogers (1979)

Jokiel and Coles (1977)

(continued)

TABLE 111. (CONTINUED)

Criterion: Nature of study growth rate and location Species Disturbance Results Reference

Observations in the field, Selection of Carib- Dredging Exposure to increased Bak (1978) CuraGao bean corals turbidity and sedimen-

tation caused a decrease in calcification rates of M . rnirabilis and A . agaricites

Laboratory experiment M . annularis Drilling mud 1, 10, Calcification rates de- Szmant-Froelich et 100 ppm doses creased at 100 ppm al. (1981) for 6 weeks dose after 4 weeks'

exposure As determined Laboratory experiment Millepora sp., Water-soluble Variable results with no Neff and Anderson

by "Ca incor- Madracis decac- fractions of fuel indications of a signifi- (1983) poration tis, M . annularis, oil and Louisiana cant effect of hydrocar-

Oculina diffusu, crude oil bons on 4SCa incorpora- Favia fragitrn tion


tal disturbance, although reduced growth rates in the Florida Keys coin- cide with a period of dredge and fill operations. Dodge and Lang (in Dodge and Szmant-Froelich, 1984) suggest that the decline in coral growth at the Flower Gardens may be due to water temperature fluctuations and increasing river discharge in the area.

Another feature revealed by x-radiography is the presence of high- density skeletal deposits or “stress” bands which have been observed in sections of M . annularis during periods of rapid chilling and mixing of shallow inshore waters (Hudson et af . , 1976; Hudson, 1977, 1981) (Fig. 2) . In addition, Highsmith (1979) has noted that, in M . annulavis from Belize, the high-density bands appear to be deposited for only short periods of

2.2 cm FIG. 2. “Stress bands” revealed in sections of M . annularis exposed to periods of chilling

and mixing of shallow inshore waters. B and C indicate the boring activities of sponges (Hudson, 1977).

26 B . E . BROWN A N D L. S. HOWARD

time while the low-density band is produced for a greater part of the year when compared with M . cavernosa and P . astreoides from the same locality. Highsmith attributes this difference to the contrasting distribution pattern of the corals ( M . cauernosa and P . astreoides being relatively restricted with respect to the broader tolerances of M. annularis), which may be reflected in the density banding pattern. It would seem, then, that massive corals living under similar environmental conditions are likely to reflect environmental variables to different degrees and that, while M . annularis is widely used in sclerochronological techniques because of the clarity of its banding pattern and wide distribution, other massive corals may be more sensitive indicators of changing environmental conditions. One alternative suggested by Hudson (1981) is S . sidereu which, although environmentally tolerant, does have a very close banding pattern enabling test cores to record longer time periods when compared with M . annularis. Whether S . siderea reflects lesser or greater sensitivity to environmental change than M . annularis remains to be seen.

Reduction in growth rate of branching corals as a result of thermal discharge (Neudecker, 1983) and fragmentation (Bak and Criens, 1983) has been observed using alizarin staining. Neudecker concluded that slower growing coral species were more tolerant of high temperatures than faster growing species, the fastest growing coral in his study being A . formosa, extending at a rate of 4.9 k 0.3 mm, while the slower growing coral P . andrewsii grew at a rate of 4.2 2 0.2 mm over the same 63-day period. No significance values were attributed to this comparison, and since there were no measurements of the weight of calcium carbonate deposited or the density of the skeleton laid down, care must be taken in interpreting the data as suggesting slower growing corals are more tolerant of thermal enrichment than faster growing species. Measurements of skeletal growth (as weight of CaC03 deposited) in Hawaiian corals exposed to increased temperatures (4-5°C) did not indicate similar results (Jokiel and Coles, 1977). The order of increasing thermal tolerance was P . damicornis < Montipora verrucosa (Lamarck) and Fungia scutaria (La- marck), while M . verrucosa calcified most rapidly and F . scutaria least rapidly of the three corals tested.

Rogers (1979), in her estimation of the effects of sedimentation on growth rate in A . cervicornis (determined by measuring skeletal growth from a base line), stresses the importance of making adequate measurements of branch extension on a large number of branches from different colonies of the same species before reliable data can be obtained. Results of this study indicated that even daily sediment doses of 200 mg/cm2 for 45 days did not affect growth rates of treated corals when compared with controls.


In contrast, Bak (1978) demonstrated an acute decrease in growth rate (measured as a 33% decrease in calcification) of M. mirabilis and A . agaricites as a result of increased sedimentation from dredging activites in CuraGao. Depressed calcification rates were noted for more than 1 month after reduction in light levels and suggest that the decrease in growth is not just the result of reduced light but also of metabolic shock that ex- ceeded the period of environmental disturbance.

The effects of water-soluble fractions of a fuel oil and Louisiana crude oil on the rate of calcium deposition (measured as 45Ca incorporation into the skeleton) in Millepora sp., Madracis decactis (Lyman), M . annuluris, Oculina diffusa (Lamarck), and Favia fragum (Esper) were quite variable, with sample variability being greater in hydrocarbon-exposed animals than in controls (Neff and Anderson, 1981). Such variability was attributed to the individual variation between colonies or parts of colonies in their sensitivity to oil, an explanation also favoured by Birkeland et ul. (1976) using coral growth as a parameter in assessing the effects of bunker oils on corals.

Despite the variability encountered in growth rate data, it would appear that this parameter has considerable value in many field observations, particularly since both branching and massive species can be transplanted into different reef sites.

B. Metabolism

A criticism of the use of metabolism as an indicator of stress in short-term experiments carried out in temperate waters has been the environmentally unrealistic levels of pollutants required to produce any effect (R. C. Newell, personal communication). In the limited number of experiments carried out with tropical scleractinians (Table IV) efforts have been made in many cases to carry out experimental manipulations in the field (Rog- ers, 1979; Dallmeyer et al . , 1982), and in all examples cited some impair- ment of an aspect of metabolism has been noted as a result of experimental disturbance (Rogers, 1979; Dallmeyer et al . , 1982; Szmant-Froelich et al., 1983). It is difficult to say, however, how experimental conditions in each case correspond to those observed in the field. Exposure of M . annularis to gradually increasing suspended peat levels of 175, 350, and 525 mg/litre may reduce photosynthesis during the day and oxygen levels during the night (Dallmeyer et al . , 1982), but no indication is given in this article of the levels of peat in natural waters at Negril, even in the brown plume reported in the field (Dallmeyer et al . , 1982). A need for improved information on levels of pollutants in the field and experimental designs that more accurately approach water quality conditions in situ has been

TABLE 1v. THE USE OF METABOLISM TO ASSESS THE EFFECTS OF DISTURBANCE ON CORALS I N T H E LABORATORY A N D EXPERIMENTALLY I N THE FIELD

Nature of study Reference Results Criterion and location Species Disturbance

Metabolism Primary pro-

ductivity and respiration

Respiration, gross photosynthesis. NO1 UP- take, NH, uptake

Respiration and net photosynthetic production

Photosynthesis and respiration

Experimental channels in the field, San Cristobal

Laboratory experiment in Row-through seawater system

I n situ measurement of oxygen metabolism, Negril, Jamaica

Laboratory experiment, Hawaii and Enewetak

Caribbean corals

M . unnuluris

M . unnuluris

P . darnicornis, M . uerrucosa, P . compressa, Fun- gin scutariu

Experimental shading

Exposure to 1. 10, and 100 ppm drilling mud for 6 weeks

Exposure to suspended peat

Thermal increase

Primary productivity and respiration decreased as a result of shading

Respiration and photosynthesis, NO3 and NH4 uptake all decreased as a result of exposure to 100 ppm drilling mud

Reduced net oxygen production as a result of exposure to suspended peat

Coral metabolism closely adapted to ambient T o conditions; results suggest lethal temperatures for Enewetak specimens to be 2-5°C higher than for Hawai- ian corals

Rogers (1979)

Szmant-Froelich c r l . (1983)

Dallmeyer et 01. (1982)

Cotes and Jokiel (1977)


highlighted by Hudson et al. (1982) and Dodge and Szmant-Froelich (1984), respectively.

In all examples cited, a key factor leading to reduced production as a result of shading (Rogers, 1979), exposure to drilling mud (Szmant- Froelich et al., 1983), and exposure to suspended peat (Dallmeyer et al., 1982) was expulsion of zooxanthellae-a response which will be discussed in detail in Section II1,C. Short-term ( 1 - to 2-h intervals) exposure of M. annularis to suspended peat concentrations of a maximum of 525 mg/litre resulted in a 50% fall in production and respiration rates when pre- and postexposure rates were compared over a 24-h period. Longer term exposure of M. annuluris to 100 mg/litre drilling mud in a flow- through system for 4 weeks resulted in a 25% fall in respiration rate and a decline in gross photosynthesis of 75% when compared with controls after 5 weeks (Szmant-Froelich et al., 1983). Shading alone of a selection of Caribbean corals (including M. annularis) produced a fall in production of approximately 50% after cover of 4 weeks (Rogers, 1979).

Recent work by Barnes (1983) may have some application to the study of stressed environments. Using a buoy equipped with pH and oxygen electrodes and a sensitive thermistor, he obtained measurements of changes in oxygen concentration, pH, and temperature of water across the reef flat, from which he deduced values for reef productivity and calcification. Barnes cites Kinsey (1979), who suggested that reef flats operate within narrow metabolic limits, any departures from these limits possibly reflecting perturbation. Once the respiratory and metabolic characteristics of reef communities are better understood, such a method as that described above may have a place in pollution studies. The technique has one distinct advantage over other “metabolism” studies in that it could be carried out in the field and potentially could give a direct measurement of the “health” of similar reef areas.

C. Loss oJ’ Zooxunthellur

The loss of zooxanthellae from coral tissue has been described by several authors (Franzicket, 1970; Jokiel and Coles, 1974; Jaap and Wheaton, 1975; Neff and Anderson, 1981) as a useful indicator of stress. Discolour- ation of corals as a consequence of zooxanthellae release may result from natural factors such as elevated temperatures and low tides (Vaughan, 1916; Yonge and Nicholls, 1931; Jaap, 1979), decreased temperatures (Wells, personal communication, in Jaap, 1979), salinity changes due to storms (Goreau, 1964), and also laboratory-induced elevated temperatures, darkness, and starvation (Yonge and Nicholls, 1931). In addition, as illustrated in Table V, loss of zooxanthellae may result from man-made

TABLE v. LOSS OF ZOOXANTHELLAE AS A RESULT OF MAN-MADE A N D NATURAL DISTURBANCES

Nature of study Criterion and location Species Disturbance Results Reference

Expulsion of Field and laboratory ob- zooxanthellae servations. Hawaii

Application of pollutant by divers in the field at Western and Eastern Sambo. Florida

Field observations at CuraGao

Field observations at Middle Sambo, Florida

Experimental chambers in the field, San Cristobal, Puerto Rico

Hawaiian reef corals

Caribbean reef corals




Elevated temperatures of 2-4°C

Exposure to quinaldine and rotenone deriva- tives

Dredging activities

Combined high temperatures and low midday tides

Shading for 4 weeks

Loss of zooxanthellar pigment

A . ceruicmwis. A . pulmcttu. S . sidereu. D . strigosu. and Dicho- coeniu stokesii showed bleaching as a result of application of chemicals

P. astrvoidrs lost zooxanthellae and subsequently died

Millrporcr complunutu displayed greatest discolouration; A. pulmutu, M . unnuloris, and Polvtlioo sp. were mildly discoloured

les, Millepora ulci- cornis, M . unnuluris. D . lubyrinthiformis, S .

A . c<jruic,ornis. A . t ~ g ~ r i c ' i -

Coles (1975); Jokiel and Coles (1974); Jokiel and Coles (1977)

Jaap and Wheaton (1975)

Bak (1978)

Jaap ( 1979)

Rogers (1979)

Experimental chambers in the field. Carysfort Reef, Florida

Field observations, Ha. waii

Transplant of corals to varying depths, Discov- ery Bay. Jamaica

Experimental studies in situ. Negril, Jamaica

Transplant studies at sites of thermal enrichment, Guam

A . ceruicornis

Seven Caribbean reef species

Hawaiian reef corals

M . einnirltrri~

M . cinnirloris

Sedimentation

Drilling mud

Kaolin spill

Transplanting of coral to 10 and 30 m depth

Addition of suspended peat to environmental chambers. during night and day

Elevated temperatures 4-6°C above ambient

sidereit. and C . nc i ions all show some zooxanthellae expulsion; large polyped species least affected

Both controls and test colonies showed small areas of bleached tissue

A . ceruicornis exposed to “mud 3” lost all zooxanthellae after 41 h: all tissue disintegrated within 52 h

Expulsion of zooxanthellae in P. mcwndrino and Poc~illlporci evdomi

Corals transplanted from 30 to 15 m showed loss of zooxanthellae

Loss of zooxanthellae by corals exposed to peat

A . fornio.sii and P. dcinii- cornis showed loss of zooxant hellae

Thompson 1’1 i t / .

(1980)

Dollar and Grigg (1981)

Dustan ( 1979)

Dallmeyer et ril. (1982)

Neudecker (1983)

32 B. E. BROWN AND L. S. HOWARD

disturbances such as thermal discharge (Coles, 1975; Jokiel and Coles, 1974, 1977; Neudecker, 1983), dredging (Bak, 1978), exposure to drilling mud (Thompson et al., 1980), kaolin (Dollar and Grigg, 1981), peat (Dallmeyer et al., 1982), quinaldine (Jaap and Wheaton, 1975), oil (Peters et al., 1981), shading (Rogers, 1979), and transplantations from deep- to shallow-water reefs (Lang, 1973; Dustan, 1979) and from offshore to inshore sites (Shinn, 1976).

The loss of zooxanthellae in different coral species subject to similar stresses under the same environmental conditions is extremely variable; the response within an individual coral may even differ over the colony (Rogers, 1979; Neff and Anderson, 1981). Triggering mechanisms for zooxanthellae expulsion have been cited as reduced supplies of nutrients available to the algae from the stressed coral host (Yonge and Nicholls, 1931; Muscatine, 1971), decrease in space available to the algae caused by atrophied host tissues (Muscatine, 1971), and secretion of substances by the coral host which produce a hostile environment around the algae (Jaap, 1979). Such mechanisms do not seem a wholly adequate explanation of the phenomenon for at least two reasons. First, the response may take place very rapidly, loss of zooxanthellae being reported within hours by Jokiel and Coles (1974), Thompson et al. (1980), Dallmeyer et al. (1982), and Neudecker (1981), and within 5 min by Jaap and Wheaton (1975). Second, although space may decrease in atrophied host tissues (Muscatine, 1971), Peters et al. (1981) have shown that under such conditions many zooxanthellae actually degenerate in situ. Clearly this aspect requires further clarification; it may well be that the speed and intensity of polyp retraction following disturbance is sufficient to expel zooxanthellae via the coelenteron, differing responses observed in corals simply reflecting the intensity of the reaction. In two studies where other aspects of coral behaviour were noted, in addition to tissue colouration, bleaching generally followed a period of polyp retraction which continued for longer than 5 min posttreatment (Jaap and Wheaton, 1975) or for the whole 96-h period of the experiment (Thompson et al., 1980). The route taken during expulsion is via the absorptive zone of the mesenteries, just below the filaments (Fig. 3), although the actual stimulus that causes the algae to be conveyed is unknown.

Some authors have described connections between zooxanthellae loss, polyp size, and other stressful environmental parameters (Rogers, 1979), but results obtained in other studies do not appear to show any consistent patterns (Goreau, 1964). In Goreau’s study (1964) of the effect of salinity changes, following Hurricane Flora, on 16 Caribbean coral species, M . annularis was observed to be the species most susceptible to zooxanthellae loss. In this study no correlation between tendency to lose zooxanthel-


FIG. 3. Transverse section through a portion of mesenterial filament of Goniu.s~rm sp. after exposure to elevated temperatures, showing zooxanthella in process of ejection. Ab- breviations: az, absorptive zone; z, zooxanthella; ze, zooxanthella being ejected; m, mes- ogloea; gm, glandular margin (after Yonge and Nicholls, 1931).

lae and polyp size was observed. Other studies indicate rather variable results, with Millepora complanata (Lamarck) displaying a greater tendency to discolour over A. palmata and M . annularis as a result of tern- perature stress at low tide; the latter two species, with small and medium- sized polyps, respectively, both showing medium discolouration (Jaap, 1979). P. astreoides displayed a marked loss of zooxanthellae as a result of dredging (Bak, 1978) when compared to other small polyped corals such as A . agaricites, and in other small polyped varieties A . formosa lost greater numbers of zooxanthellae than P. damicornis when exposed to a 5 4 ° C temperature increase (Neudecker, 1983).

The two studies where a correlation between polyp size and zooxanthellae loss has been described are those of Jokiel and Coles (1974) and Rogers (1979). Jokiel and Coles (1974) noted that the larger polyped coral L . purpurea was more tolerant of thermal increases than the smaller polyped species P. rrteandrina and P. lobata, which lost their zooxanthellae more rapidly on response to temperature increases 4-5°C above ambient. Rogers (1979) described a fairly close connection between increased susceptibility to shading, as demonstrated by bleaching, with decreased polyp size, e.g., A. cervicornis, with small polyps being the first species to bleach, followed by M. annularis, D. labyrinthiformis, and D. strigosa,

34 B. E . BROWN A N D L. S. H O W A R D

with medium-sized polyps and large polyped species such as Eusrnalia fastigiata (Pallas), Mussa angulosa (Pallas), and M . cauernosa appearing relatively unaffected by shading. The author acknowledges, however, that the latter species were growing at the edge of experimental channels where light may not have been totally excluded.

The apparent lack of a relationship between polyp size and the tendency for zooxanthellae loss in some studies, and yet the existence of an apparent relationship between the two parameters in others, probably reflects the likely complexity of the response. Apart from the obvious variability in responses from different corals at a given depth, the same species collected from a variety of depths may also exhibit widely differing responses. M . annularis from different depths contains zooxanthellae which are photoadapted to specific habitats (Dustan, 1979). In transplant experiments, zooxanthellae adapted to high light intensities function poorly in deeper water habitats, while zooxanthellae adapted to low light intensities are actually damaged by high light intensities in shallow waters. At deep water sites, corals transplanted from shallower depths show some reduction in algal content; in contrast, corals transplanted from deep water to shallow water display considerable bleaching and high mortality (Dustan, 1982). Hence, the tendency to lose zooxanthellae in response to changing light intensities is very much dependent upon the original position of the coral within the reef environment.

Considering, then, the limited literature to date, it would seem that loss of zooxanthellae in response to a particular stress does give some indication of the relative tolerance of a coral species to parameters such as temperature increase, salinity change, etc. There are problems, however, in interpreting such responses since they may be reversible (Jokiel and Coles, 1974, 1977; Jaap, 1979); they may be very localized, the undersur- face of an affected coral displaying little apparent loss of zooxanthellae (Jokiel and Coles, 1977); or they may reflect not only one stress but a combination of two or more, such as combined temperature and strong light intensity (Jokiel and Coles, 1977). In addition, loss of zooxanthellae may occur unperceived by the observer but calculated through some other measurable parameter such as reduced calcification (Bak, 1978) or production (Dallmeyer et al., 1982). Quantification of the response is possible by extraction of photosynthetic pigment from coral tissues, and this has been successfully carried out by a number of workers (Jokiel and Coles, 1974; Dallmeyer et al., 1982) in response to thermal stress and addition of peat, respectively. The value of the response of loss of zooxanthellae by coral tissues as an indicator of stress could be considerably improved (1) by quantification of the magnitude of algal loss using pigment extraction techniques, (2) by better understanding of the mechanism


of expulsion of the algae, and (3) by relating this response to other physiological parameters.

D. Behaviourul Responsvs

1 . Feeding and mesenterial filament extrusion

Mesenterial filament extrusion by scleractinian corals has been widely reported by early workers (Duerden, 1902; Carpenter, 1910; Vaughan, 1912; Matthai, 1918; Yonge, 1930) in connection with feeding activities. These early studies revealed the ability of many scleractinians to engage in extracoelenteric digestion of food material by means of filaments extruded either through the mouth or through temporary openings on the colony surface (Fig. 4). Lewis and Price (1975) failed to observe corals feeding routinely by this method and commented that in many cases where extracoelenteric digestion was described in the literature this was connected with the presence of large particles of food on the oral disc. Nevertheless, numerous other workers have noted mesenterial filament extrusion as a result of exposure of scleractinians to feeding inducers such as the amino acids glutathione and proline (Brown and Phillips, unpublished; Mariscal and Lenhoff, 1968; Goreau et al., 1971), where tactile stimuli were not involved.

FIG. 4. The extrusion of mesenterial filaments through the mouth and body wall of a coral polyp. Abbreviations: t , tentacle; m, mesentery; f, filament; c, calyx; s, skeleton (after Goreau et at., 1979).


The extrusion of mesenterial filaments in interspecific aggression responses by corals has been reported both in the Caribbean (Lang, 1970, 1971; Bak et al., 1982) and in the Indo-Pacific (C. R. C. Sheppard, personal communication). Sheppard (1979) implies that the response is trig- gered as a result of chemoreception properties of the aggressor, a sugges- tion first put forward by Muscatine (1973) when considering the nutrition of corals.

Filament extrusion has also been cited in recent years as a response of corals to adverse conditions (Table VI). Goreau et al. (1971) report severe starvation resulting in extrusion of mesenterial filaments by M . angulosa; Lewis (1971) describes the same response in Madracis asperula (Milne, Edwards, and Haime) when exposed to crude oil and oil dispersant; Bak and Elgershuizen (1976) recorded filament extrusion in M. decactis as the result of the presence of oil droplets in the gastrovascular cavity, while M . annularis was observed to extrude mesenterial filaments in response to exposure to drilling muds (Thompson et a/., 1980). Similar behaviour has also been recently reported by S. Wyers (personal communication) in D . strigosa exposed to fuel oil. All these observations apply to laboratory experiments, although J. H. Thompson (personal communication, in Dodge and Szmant-Froelich, (1984) observed mesenterial filament extrusion on four heads of M . annularis in the field after application of 5 ml drilling mud slurry to each. The same behaviour was observed on four heads of M . annularis treated with a similar amount of carbonate sand.

The majority of examples of filament extrusion cited above have been the result of exposure of animals to oil. Both Loya and Rinkevich (1980) and Ormond and Caldwell (1982), reviewing the literature concerning anthozoan feeding behaviour and oil exposure, highlight the work of Blu- mer et al. (1971), which discussed the possible interference of crude oil products with chemoreception in marine invertebrates. The mechanism suggested involved the mimicking of natural stimuli by the oil product, which in turn elicited feeding behaviour. Tentacle responses and/or mouth-opening reactions also associated with feeding (Table VI) have been observed in anthozoans in response to crude oil pollution (Reimer, 1975a,b; Loya and Rinkevich, 1980; Ormond and Caldwell, 1982). Experi- ments with the temperate anthozoan Actinia equina (Linnaeus) suggest that crude oil or some component of crude oil does act as a feeding inducer, whereas the pure hydrocarbons tested did not. It was also shown that crude oil presented on filter paper to the anemones either interfered with or diluted the action of natural feeding inducers present in fish mus- cle extract (Ormond and Caldwell, 1982).

Clearly the mechanism(s) involved in the appearance of mesenterial filaments and other feeding responses such as mouth opening as a result of exposure to pollutants are unclear. The use of filament extrusion as an

TABLE VI. THE USE OF MESENTERIAL FILAMENT EXTRUSION TO ASSESS THE EFFECTS OF DISTURBANCE ON CORALS IN THE LABORATORY AND

EXPERIMENTALLY IN THE FIELD


Mesenterial Laboratory experiment, Porites porites, Exposure to crude M . asperula exhibited Lewis (1971) filament Barbados, West Indies Madracis asper- oil and oil spill greatest effects on extrusion ula, F. fragum, dispersant addition of 50 ppm oil,

A . agaricites involving an increase of extrusion of mesenterial filaments; all species were more affected by the dispersant than by the crude oil

Laboratory experiment at Nineteen Caribbean Exposure to crude Extrusion of mesenterial Bak and Caribbean Marine coral species oil and sediments filaments in response to Elgershuizen Biological Institute, introduction of oil ( 1976) CuraGao drops into the gastro-

vascular cavity Experimental chambers Seven Caribbean Exposure to drilling Some polyps of M . annu- Thompson et a / .

sited on a sand flat at coral species mud laris extruded mesenter- (1980) Carysfort Reef, Key Largo. Florida

ial filaments; other corals tested did not show this response

Field experiment, Key M . annularis Exposure to drill Although sediments were Thompson, in Largo, Florida mud slurry or cleared within 2 h, all Dodge and Sz-

carbonate sand corals showed mesenter- mant-Froelich ial filament extrusion (1984) later

Flow-through laboratory D . strigosa Exposure to crude Extrusion of mesenterial S. Wyers, personal experiment, Bermuda oil filaments communication

38 B . E . BROWN A N D L . S . HOWARD

indicator of stress, often cited by workers (Lewis, 1971; Bak and Elgershuizen, 1976; Thompson, et al., 1980), would appear to be complicated by the variability in response of individual corals (S. Wyers, personal communication), the often temporary nature of the response (Thompson et al., 1980), its regular appearance in certain coral species and not in others (Goreau et al., 1971), and a lack of understanding of natural stimuli which might elicit such behaviour.

Other behavioural responses normally associated with feeding and also sediment shedding, such as coenosarc distension, have been observed on exposure of certain coral species to pollutants (Thompson et al., 1980; Dallmeyer et al., 1982). Bak and Elgershuizen (1979) describe the se- quence of events which follows a sediment rain on an expanded coral surface as initial contraction of the polyps followed by a marked expansion of tissues including the coenosarc. Such behaviour was also observed in M . annularis at night after exposure to drilling mud, although other corals, e.g., Porites diuaricata (Lesueur), Porites frrrcata (La- marck), P . astreoides, A . cervicornis, A . agaricites, and Dichocoeniu stokesii (Edwards and Haime), under similar conditions did not display coenosarc expansion (Thompson et al., 1980). Addition of peat to M . annularis (Dallmeyer et al., 1982) also resulted in conspicuous coenosarc distension at night.

A reverse phenomenon, i.e., polyp retraction, has been used in some studies to assess the toxicity of pollutants (Jaap and Wheaton, 1975; Thompson et al., 1980). Contraction of polyps as a reaction to stress in the form of electrical stimulation was described by Horridge (1957), and to extreme water currents by Hubbard (1974), while Bak and Elgershuizen (1979) observed contraction as a reaction to contact with nonfood particles. Polyp retraction was a common reaction of all 12 Car- ibbean corals exposed to fish-collecting chemicals (Jaap and Wheaton, 1975), while in experiments with drilling muds five species ( M . annularis, A . agaricites, A . cervicornis, P . furcata, and P . astreoides) demonstrated significant polyp retraction to 100 ppm and higher concentrations of drilling mud. D . stokesii, however, showed no reaction to drilling muds at any of the tested concentrations, while P . diuaricata displayed significant retraction at 316 ppm. Exposure of corals to water-soluble fractions of fuel oil resulted in polyps of M . decactis and M . annularis remaining partially or totally retracted while polyps of control corals were fully expanded during all or most of the exposure period (Neff and Anderson, 1981). These authors describe retraction of polyps as a sign of the effects of severe stress in corals.

While acknowledging that individual polyps can alternate between an expanded and a retracted state (Sweeney, 1976; Sebens and de Riemer,


1977), with some species remaining expanded for long periods of time while others engage in die1 cycles of expansion, Dodge and Szmant- Froelich (in press) believe coral polyp retraction may be useful as a bioassay technique for pollutants. Lasker (1979, 1981), however, has shown that in at least one example ( M . cauernosa) expansion cycles may vary within a species. On the Caribbean coast of Panama, colonies of M . cauernosa can be divided into two morphs based on their activity cycles and polyp morphology (Lehman and Porter, 1973; Lasker, 1976, 1977, 1979), with polyps of the diurnal morph of M . cauernosa expanded both day and night, while those of the nocturnal morph expand only at night. Furthermore, in diurnal morph colonies which exhibited zooxanthellae loss and bleaching, only the “normal” part of the colony expanded during the day and night, while the “bleached” area expanded solely at night. The loss of daytime expansion appeared to be linked to the absence of zooxanthellae. Since zooxanthellae loss has been cited as a response of many scleractinians to exposure by pollutants (see Section lll,C), careful consideration should be given to behavioural responses involving polyp expansion and retraction which may be influenced by the presence or absence of these symbionts.

2. Mucus production

Mucus production by corals has been described during the course of feeding (Lewis and Price, 1975; Lewis, 1977), sediment rejection (Yonge, 1930; Bak, 1978), and shading (Rogers, 1979) and also as a result of exposure to pollutants such as crude oil (Mitchell and Chet, 1975; Neff and Anderson, 1981), copper sulphate (Mitchell and Chet, 1973, fish- collecting chemicals (Jaap and Wheaton, 1973, drilling muds (Thompson et al., 1980), peat (Dallmeyer et al., 1982), and increased temperatures (Neudecker, 1983) (see Table V11). In histological studies an increase in the number and size of mucous secretory cells was observed as a result of exposure of Manicina areolata (Linnaeus) to chronic oil pollution (Peters, et al., 1981).

Although excess mucus production as a result of man-made stress has often been cited in the literature, Bak and Elgershuizen (1976) showed that the response of 19 hermatypic Caribbean corals to oiled sediments was not an obvious increase in mucus secretion compared with secretion resulting from exposure to clean sediments. Mucus production may, in some cases, actually delay cleaning of the coral surface, particularly when very small particles (e.g., carborundum powder) become trapped in the mucus. In Bak and Elgershuizen’s study such a process caused death in P . astreoides and A . agaricites.

TABLE VII. THE USE OF Mucus PRODUCTION TO ASSESS THE EFFECTS OF DISTURBANCE ON CORALS IN THE LABORATORY AND



Mucus produc- Fish-collecting chemicals tion applied in the field

Caribbean reef Exposure to corals quinaldine and a

rotenone derivative

Laboratory experiment

Laboratory experiment at Caribbean Marine Biological Institute, CuraCao

Field experiment, San Cnstobal Reef, S.W. Puerto Rico

Experimental chambers sited on a sand flat at Carysfort Reef, Key Largo, Florida

Platygyra spp. Exposure to crude oil, copper sulphate, potassium phosphate or dextrase

Nineteen Caribbean Exposure to crude coral species oil and sediments

Ten Caribbean Shading coral species

Seven Caribbean Exposure to drilling coral species mud

~ ~

All corals secreted increased amounts of mucus, although no long-term damage was evident

produce large quantities of mucus

All corals stimulated to

Jaap and Wheaton (1975)

Mitchell and Chet (1975)

Certain species (e.g., A . palmata, A. cervicornis, P . porites, and P . astreoides) proved to be copious mucus secreters

Some areas on collines of M. annularis secreting mucus after 8 weeks’ shading

All corals exhibited an increased mucus production, mucus being produced either as a sheath or as strands

Bak and Elgerhuizen ( 1976)

Rogers (1979)

Thompson et al. ( 1980)

Flow-through experiment M . annularis, A . cervicornis, A . on corals from Car-

ysfort Reef, Florida palmata

Experimental transplanta- P . andrewsii, P . tion in the field, Guam damicornis, A .

for m o s a

Laboratory experiment Manicina areolata

Laboratory experiment, M . annularis Jamaica

Exposure to surface oil slick of south Louisiana crude oil

Exposure to elevated temperatures (4-6°C above ambient)

Exposure to 0.15 ppm oil hydrocarbons for 3 months

Exposure to particulate peat

Mucus production particularly stimulated in M . annuluris

A . formosa appeared most sensitive to elevated temperatures and produced large quantities of mucus after a few hours of exposure; mucus production was also noted during alizarin staining in P . andrewsii, although these colonies were not used in subsequent experiments

Histological studies revealed an increase in the number of mucus- secreting cells in tissues of exposed corals

Clumps of mucus produced on the surface of the coral trapped peat, but were subsequently removed. After 15 h the coral showed a normal appearance

Neff and Anderson (1981)

Neudecker (1983)

Peters er al. (1981)

Dallmeyer et al. (1982)

42 B. E. BROWN AND L. S . HOWARD

Under normal circumstances mucus production may result in up to 40% net carbon fixation being lost from a coral such as Acropora acuminata (Verrill) (Crossland, Barnes and Borowitzka, 1980), so mucus production in corals under stress may constitute a considerable energy loss, as noted by other authors (Loya and Rinkevich, 1980).

Workers also suggest that mucus may bind or absorb pollutants such as aromatic hydrocarbons (Neff and Anderson, 1981) or heavy metals (Ho- ward and Brown, 1984) and so confer some protection to the underlying coral tissues either by physically protecting them or by acting as an ave- nue for hydrocarbon release from contaminated corals (Neff and Ander- son, 1981). However, it is also well known that mucus strands and nets may be ingested by corals (Lewis and Price, 1975); should oil particles be trapped in the mucus, then these may also be ingested during the feeding process (Bak and Elgershuizen, 1976). An increase in mucus production by a coral under stress may also result in significant increases in the bacterial population of the mucus (Ducklow and Mitchell, 1979), and these effects will be discussed later in this article (see Section 111,F). The existing literature on effects of drilling muds and increased temperature exposure on selected corals suggests that Acropora spp. readily produce mucus when compared with other coral species exposed to the same stress. In experiments with drilling muds (Thompson et al., 1980), A . cervicornis produced mucus strands after 30 min exposure, whereas mucus production in other species (P. divaricata, P. furcata, P. astreoides, and M . annularis) was not observed until at least 24 h after mud application. (It should be noted that these responses were not shown by all colonies of the test species.) Exposure of A . formosa, P . andrewsii, and P. damicornis to 5-6°C temperature increase at the Cabros Power Plant, Guam, resulted in considerable mucus production in A. formosa after a few hours; all test colonies in one experiment were dead within 48 h of exposure, whereas P . damicornis died within 30 days and P. andrewsii survived the entire test period of 77 days. No note of excessive mucus production in Pocillopora or Porites was reported (Neudecker, 1983).

If mucus production were to be used as a measure of stress, then some attempts would have to be made to assess the amount of mucus produced during exposure to a pollutant. Using techniques adopted by Crossland et al. (1980a), it would be possible to quantify the amount of mucus produced and also the rate of production in a stressed coral. Quantification of mucus production rates is limited to A . acuminata, where the rate of mucus output does not appear to vary diurnally. Extension of quantitative methods of mucus production to corals under stress would necessarily demand some base line information of the type already obtained for A . acuminata.


3. Sediment shedding

Sediment-shedding behaviour has been observed in a variety of corals by numerous authors (Hubbard and Pocock, 1972; Hubbard, 1974; Schuma- cher, 1979; Lasker, 1980; Fisk, 1983). Hubbard and Pocock described a ranking in the ability of Caribbean coral species to reject sediments which was related to the diameter of coral colonies and their abundance on the colony surface, but subsequent work by Bak and Elgershuizen (1976) failed to demonstrate such a relationship in Caribbean corals.

Other workers, however, have used sediment-shedding behaviour to assess the effects of drilling mud (Thompson, no date, in Dodge and Szmant-Froelich, 1984; Thompson and Bright, 1980), oiled sediments (Bak and Elgershuizen, 1976), and heavy metals (Brown and Holley, unpublished) on reef corals (Table VIII).

Results of experiments with drilling muds showed that while clearing rates varied from species to species, all three species tested (D. strigosa, M. annufaris, and M . cauernosa) could clear barite, bentonite, and CaC03 but no species was able to remove the used drilling mud. The authors suggested that dissolved, toxic components of the mud were responsible for this dramatic difference in response. Similar results were obtained with M . annularis and P . astreoides exposed to drilling muds and carbonate sands. Montastrea failed to move the drilling mud and died within 15 h, while treated P. astreoides died within 10 days. Those specimens covered by carbonate sand recovered completely.

In the 19 hermatypic coral species tested by Bak and Elgershuizen (1976), the efficiency of removal of oil sediment particles was the same and performed by an identical rejection mechanism as when they were covered with clean particles of the same size or quality. These authors highlighted the complexity of mechanisms involved in sediment rejection by corals, a feature noted recently by Fisk (1983) working with sediment shedding in fungiids. Bak and Elgershuizen state that the relationship between rejection efficiency and density of sediment particles and their size may involve a very rapid rejection, a rejection over time, or no rejection at all; e.g., M. cauernosa removes large oil-sand particles (3-mm diameter) by movements of the tentacles and polyp expansion, whereas smaller oil-sand particles (1-mm diameter) fall between the polyps and are rejected by ciliary activity; much smaller particles (0. I-mm diameter) are transported very rapidly by the cilia. Bak and Elgershuizen also noted that in some species rejection patterns varied considerably depending on the degree of expansion of the living tissue, a finding supported by Brown and Holley (unpublished) working with Fungia fungites (Linnaeus) exposed to metal-laden sediments.

TABLE VIII. THE USE OF SEDIMENT SHEDDING TO ASSESS THE EFFECTS OF DISTURBANCE ON


Nature of study Criterion and location Species Disturbance

\

Sediment shed- Laboratory experiment at Nineteen Caribbean Exposure to crude Patterns of

Biological Institute to ding 1 Caribbean Marine coral species oil and sediments rejection

of clean rates of ding are size and particles

D. strigosa, M. Exposure to drilling Cleaning annularis, M. mud, barite, from cauernosa bentonite, and all

calcium carbon- barite, ate calcium

species remove mud

Laboratory study

TABLE VIII. THE USE OF SEDIMENT SHEDDING TO ASSESS THE EFFECTS OF DISTURBANCE ON CORALS IN THE LABORATORY AND



\

Sediment shed- Laboratory experiment at Nineteen Caribbean Exposure to crude Patterns of oil sediment Bak and

Biological Institute to patterns of rejection (1976) ding 1 Caribbean Marine coral species oil and sediments rejection are identical Elgers huizen

of clean sediments; rates of sediment shedding are dependent on size and density of particles

D . strigosa, M . Exposure to drilling Cleaning rates varied Thompson and annularis, M . mud, barite, from species to species; Bright (1980) cauernosa bentonite, and all corals could remove

calcium carbon- barite, bentonite, and ate calcium carbonate; no

species was able to remove all the drilling mud

Laboratory study

Laboratory study M . annularis

Field experiment San M . annularis, D.

cliuosa, A. palmata, A. cervicornis

Cristobal, Puerto Rico srrigosa, Diploria

Exposure to 10-ml aliquots of different (A and B) drilling muds and natural fine- grained muds

Exposure to calcar- eous sediments in different frequen- cies and doses

Corals were able to clear natural sediments and one type of drilling mud (A); only one specimen was able to clear the more concentrated drilling mud (B)

A. palmata was the least tolerant of all species tested; although A . cervicornis and D . srrigosa were not significantly affected, single applications of 800 mglcm* to M . annularis and 200 mgl cm2 to A . palmata colonies caused death of underlying coral tissue

Thompson, in Dodge and Szmant- Froelic h (1984)

Rogers (1984)

46 B. E. BROWN AND L . S. HOWARD

Overall then, sediment-shedding efficiencies in corals exposed to stress may have some value as a bioassay, particularly when the environmental influence under consideration involves a combination of increased sedimentation and a pollutant. However, it must be borne in mind that patterns and efficiency of sediment rejection are often specific to particular coral species, that feeding status may affect ciliary rates (Holley, unpublished), and that observations in the laboratory must be extrapolated to the field with care since rates of sediment rejection in the natural environment may be accelerated by water movements (Bak and Elgershuizen, 1976; Dodge and Szmant-Froelich, 1984).

E. Reproductive Biology

Information on sexual (Rinkevich and Loya, 1979b,c; Kojis and Quinn, 1981, 1982; van Moorsel, 1983) and asexual (see Highsmith, 1982, for review) reproduction in corals has increased considerably in recent years. Several authors (Loya and Rinkevich, 1980; Dodge and Szmant-Froelich, 1984) reviewing the effects of pollutants on corals, recognize the value of considering reproductive biology and larval recruitment as top priorities in any toxicity assessment.

Not surprisingly in view of our recently acquired knowledge of reproductive strategies in corals, studies on the effects of pollutants on reproductive biology are limited, being restrkted to those of Rinkevich and Loya (1977, 1979a). Observations in thk field and laboratory suggested that populations of S . pistillata at a chronically oil-polluted reef at Eilat showed a smaller number of breeding colonies, a decrease in the average number of ovaria per polyp, a smaller number of planulae produced per coral head, and a lower settlement rate of planulae on artificial objects when compared with control colonies. The authors attributed the significant decrease in the number of ovaria in the oldest colonies from the oil- polluted site to the expulsion of premature and mature planulae, an observation recorded also by Cohen (1973) in the alcyonarian Hetoroxemia fuscescens (Hemprich and Ehrenberg) and Ormond and Caldwell (1982) in the anemone A . equina exposed to oil pollutants.

In further laboratory experiments, Rinkevich and Loya (I979a) cut large and mature colonies of S . pistillata into two halves, exposing one half to pollutant and the other to a control solution. Such a procedure was devised to reduce expected variatjbn between colonies, and after 2 months a significant decrease in t& number of female gonads per polyp were recorded in 75% of the polluted halves as compared with control halves of the colonies in clean sea water.

Similar detrimental effects of oil have been demonstrated by Peters et


al. (1981) in the Caribbean coral M . areolata and by Ormond and Caldwell (1982) in the temperate anemone A . equina. In these studies the anthozoans showed either degenerating ova and abnormal gonad development (Peters et al., 1981) or reduced size of ova (Ormond and Caldwell, 1982).

The settlement of planulae on artificial substrata in chronic exposures to pollutants in the laboratory would appear to be an attractive bioassay, but care must be taken in selection of planulae for use in such experiments. Rinkevich and Loya (1979a) do not state how planulae were collected for settlement experiments, but experience has shown that planulae released by P . dumicornis in the laboratory display variable development, which may subsequently be reflected in their settlement behaviour during experiments (Brown and Holley, unpublished). In much earlier studies, Marshall and Stephenson (1933) speculated that coral collection in the field produced sufficient disturbance to trigger release of planulae in certain corals. Premature ejection of juveniles as a result of field collection of A . equina has also been described by Ormond and Caldwell (1982). It may be, then, that planulae which are to be used in settlement experiments in the laboratory should be collected in situ by methods adopted by Rinkevich and Loya (1977) when investigating the number of planulae expelled by individual colonies on the reef. In these experiments plankton nets (mesh size 125-pn diameter) were secured to cover individual coral colonies in the late afternoon, the nets being removed 2 h after sunset when the planulae were collected. Although the assessment of oil (Loya and Rinkevich, 1980) and heavy metal (Brown and Holley, unpublished) pollution using criteria based on sexual aspects of reproductive biology of corals may have yielded interesting results, the importance of asexual reproduction in reef coral life histories should not be neglected. Highsmith (1982) has already demonstrated that asexual reproduction by fragmentation is an important mode of reproduction in major reef-building corals. Separation of single colonies into independent units may also occur as a result of fission on partial mortality. Based on field data, Brown and Holley (1982, and unpublished) believe that the incidence of partial mortality in massive species may be increased on intertidal reefs affected by heavy metals and sedimentation in Thailand. Such observations could not be demonstrated easily in the laboratory, but nevertheless they do highlight the need to consider all life history characteristics when assessing the effects of pollutants, especially when many corals reproduce predominately by asexu$l methods (Highsmith, 1982). Coral fragments exposed to stress in thedorm of coral disease proved to be more vulnerable than undisturbed colonies in studies on survival after fragmentation in Curaqao (Bak and Criens, 1983).

An unusual regeneration phenomenon, described as “polyp bail-out, ”

48 B. E. BROWN A N D L . S. HOWARD

has recently been demonstrated in Seriatopora hystrix (Dana) (Sam- marco, 1982). This response, in which polyps detach from the skeleton and subsequently settle and calcify, has been described as an escape response to environmental stress. Polyp bail-out was induced in the laboratory by maintaining coral colonies in nonaerated, noncirculating sea water. However, the response was also noted in the field where no obvious adverse environmental conditions were apparent. Clearly, this mode of “reproduction” warrants further investigation in both the field and the laboratory.

F. Histopathology

In a recent review on the pathology of reef corals (Antonius, 1983a) the author summarizes the four known coral diseases as bacterial infections, algal infections, white band disease, and shutdown reaction. According to Antonius, bacterial infections are often the result of corals protecting themselves against outside stresses by mucus secretion. Mitchell and Chett (1975) provide evidence of the involvement of predatory bacteria, Desulfouibrio, and Beggiatoa in the destruction of living tissue of Platy- gyra exposed to chemical pollutants. Excessive mucus production resulting from natural and man-made influences (e.g., increased sedimentation, toxic chemicals) may also enhance the numbers of the blue-green alga Oscillatoria submembranacea (Ardissone and Strafforella) thought to be responsible for black band disease. This disease may prove to be more damaging to some corals than to others, and Antonius (1983b) cites a tentative list of susceptibility for Caribbean corals, with D. strigosa and M . annularis as most prone to the disease and A . palmata, A . ceruicornis, and Acropora prolifera (Lamarck) as most resistant to black band infections. No pathogen is yet known to be responsible for white band disease (Antonius, 1983b), although work by E. C. Peters (personal communication, in Gladfelter, 1982b) suggests that bacteria may be responsible. The disease is currently widespread in the northeast Caribbean, Panama, and South Florida, and although Antonius cites A, pafmata as the most affected coral in certain parts of the Caribbean, Bak and Criens (1983) report that A . ceruicornis was more affected by the disease than A . palmata in Curasao.

Shutdown reaction has been obydrved as the result of exposure of corals to continuous sedimentation or excessive temperature plus a slight additional stress (e.g., a scratch on the coral surface) which would normally not cause any effect in a healthy coral. The speed of this response, which involves tissue regression, may be spectacular (10 c d h ) , and it now appears that shutdown reactions are highly contagious (Antonius,


1983a). Antonius suggests that this response may be a purely physiological reaction, although the mechanism is unknown.

Despite the apparent lack of understanding of causative agents involved in coral diseases, tumours which incorporate filamentous algae in Carib- bean gorgonids have been cited as potential indicators of environmental stress (Morse rt a f . , 1977, 1981). According to these authors, tumour-like growths have previously been observed in scleractinia from the Pacific (Squires, 1965; Soule, 1965; Cheney, 1975). Although the etiology of these growths is unclear in most cases, Kaufman (1980) has described gall-like nodules in Acropora in response to predation by damselfish. The incidence of algal tumours in populations of the Caribbean gorgonid Gorgoniu uentulina (Linnaeus) is highly localized at Bonaire, and Morse et al. (1981) infer that the tumours may be produced in response to high levels of chronic or intermittent hydrocarbon pollution from nearby petroleum tanker lanes and loading depots. Similar tumours, however, have been described in the gorgonian coral Pseudoplexaura spp. in the Florida Keys (Goldman and Makemson, 1983), where the authors make no environmental implications in detailing the widespread occurrence of the condition.

Tumours have been previously reported in the gonad tissues of clams collected from an oil-polluted area (Barry and Yevich, 1975), but there is no direct evidence that such pollutants result in malignant tumours. Simi- lar histopathological studies in corals are restricted to observations made by Peters et al. (1981) on the effects of No. 2 fuel oil on M. arraluta after 3 months’ chronic exposure to 0. I ppm and 0.5 ppm concentrations. Cellu- lar degeneration and atrophy of coral tissues were noted in both high and low concentrations of oil, a finding corroborated by S. Wyers (personal communication) working on the effects of oil on the Caribbean coral D . strigosa. Peters et a f . (1981) describe histopathological examinations as providing an early indication of tissue damage which might only be real- ized at a much later date in benthic ecological studies (i.e., mortality of adults and species recruitment). Despite this statement, it would seem that the general histology of coral tissues is poorly known, and the possible effects of laboratory entrainment and starvation (no mention of feeding of corals is given in the above experiment) on coral tissues remain to be documented. Effects observed by Peters et ul. may not have been the result of only one stress, b u t d m b i n a t i o n of pollutant and starvation. In addition, histopathological techniques generally involve the use of a limited number of specimens (three per assay). Considering the variability highlighted by other workers in measurement of physiological parameters between and even within different coral colonies, the potential pitfalls in histopathological examination of coral tissues are clear.

50 B. E. B R O W N A N D L. S. H O W A R D

G . Biochemical and Cytochemical Indexes

Interest in biochemical indexes of effects of stress in reef corals has been shown by those workers studying the effects of drilling muds on Carib- bean species (Szmant-Froelich et al., 1983; Dodge and Szmant-Froelich, 1984). Szmant-Froelich and Pilson (1980) had earlier shown that the ratio of lipid to protein reflected the nutritional status of an ahermatypic coral, Astrangia danae (Agassiz). Similar analyses were applied to M . annularis, and although there was no significant difference in nitrogen content (a measure of protein content) between control corals and corals exposed to 100 ppm drilling mud, changes were evident in the lipid content and composition of exposed corals (Dodge and Szmant-Froelich, 1984). These authors also describe several unpublished studies in which biochemical indexes have been used in assessing effects of drilling muds. These include a study by Powell on A . ceruicornis in which exposure to 100 and 500 ppm drilling mud for 24 h produced dramatic increases in the protein content and in the total ninhydrin-positive substances (mostly free amino acids) of coral tissues, and work by Prassad in which he found that treatment of M . decactis with drilling mud at 10 and 100 ppm concentrations actually inhibited protein synthesis. Clearly such measurements need to be interpreted in the light of further data on coral material. Nevertheless, the use of biochemical indexes represents an interesting development in the assessment of effects of stress in reef corals.

Such indexes have been used with some success by temperate workers (Jeffries, 1972; Bayne et al . , 1976; Moore, 1976; Moore and Stebbing, 1976). Measurements of the taurine/glycine ratio appear to offer some promise as an empirical indicator of stress, at least when the stressor is temperature.

Another index of stress used by these workers has been the cytochemical demonstration of latency of lysosomal hydrolases in mussels exposed to temperature stress (Moore, 1976), in hydroids exposed to increased concentrations of Cu, Cd, and Hg (Moore and Stebbing, 1976), and in homogenates of the sterile septa of the anemone Cerianthus lloydii (Gosse) (Tiffon, 1971). The exposure of Mytilus edulis (Linnaeus) to temperatures of 25-28°C over a period of 4 days induced a significant decrease in the latency of lysosomal glucosaminidase. In experiments with the hydroid CampanulaQexuosa (Hincks), cytochemical threshold concentrations were comparable to known environmental levels and were about one order of magnitude lower than those obtained by measuring inhibitory effects in colony growth rates (Moore and Stebbing, 1976). As such, then, these techniques appear to offer a very sensitive way of assessing effects of stress in invertebrate tissues.


IV. Discussion and Future Research Needs

Statements such as those of Johannes (1972) and Rogers et af. (1982) that coral reefs are adapted to natural catastrophes but susceptible to unnatu- ral stresses, and that of Loya (1976a) who concluded that human-per- turbed reefs may not return to their former configurations, while naturally denuded areas of reef may recover, given time, are obviously very broad generalizations. In the light of recent data, these conclusions are not necessarily supported, for lack of recovery as a result of natural damage has been noted in several cases (see Endean, 1973, for review), and both chronic and acute man-made pollution have been shown to have limited impact in a significant number of examples (see Section 11,C of this article). Furthermore, stresses (e.g., salinity and temperature) in the tropics have been cited as greater than those in temperate waters (Moore, 1972). Moore concluded that tolerances shown by tropical animals would tend to decrease under stress, and that as a result the effect of any additional stress would be exaggerated. This, in turn, led Johannes (1975) to consider that tropical marine communities would appear to be less tolerant to pollution than their temperate counterparts, which has further contributed to the view that cord reefs are fragile communities, highly sensitive to a wide range of man-induced pollutants (a view discussed by Dollar and Grigg, 1981).

Perhaps we should question the basic premise in the above statement. Considering temperature stress, then, tropical and temperate organisms have similar metabolic rates (measured as O2 consumption) when determined at their respective habitat temperatures (Vernberg, 1962; Vern- berg and Vernberg, 1972), although tropical animals may have a limited ability to alter their metabolic response to different temperatures (Vernberg, 1981) when compared with homeostatic mechanisms shown by temperate organisms. Certainly data on one metabolic parameter- primary production-show no clear pattern when compared between tropical and temperate systems, higher values being obtained in tropical benthic communities but lower values in tropical phytoplankton (Johannes, 1975).

It is our view, in accor ance with that of Dollar and Grigg (1981), that

lead us to believe. We believe that greater attention should be paid in the tropics to interpretation of individual examples of natural and man-made disturbances. Given the variability in responses of corals both in the field and in the laboratory to stress, as highlighted in this article, how should we approach each problem? Here a generalized approach to the assessment of natural and man-made disturbances may have some value in

reef ecosystems may ,;d be as fragile as previous generalizations would

52 B . E . BROWN A N D L . S. HOWARD

providing comparative data on tolerances both within and between different coral species and also between different provinces.

Conclusions from the foregoing article may be summarized as follows:

1 . Of primary importance is the assessment of the disturbance (man- made or natural) in the field. A lesson learned from temperate studies is that relatively little can be gained from isolated laboratory experiments executed without any field context. Obviously, methods employed to characterize man-made pollution in the field will depend on the form of pollution. It is important that the pollutant should be both physically and chemically characterized in the environment (i.e., waters, sediments, bi- ota, etc.) as completely as possible. Such analyses would ideally be carried out over time, particularly in a region which displayed a marked seasonal regime, i.e., reversing monsoon, where heavy rains and increased sediment loads may alter the biological availability of the pollutant at different times of the year.

Measurement of the response of the reef community in the field should involve estimates of abundance and diversity of all dominant mem- bers of the benthic community, as described in earlier sections of this article. Estimates of diversity, however, should be interpreted with care, since they may result in misleading conclusions (Hedgepeth, 1973; Jo- hannes, 1975). High diversities may be variously interpreted (Rogers et al., 1982; Brown and Holley, 1982). The latter study showed a high diversity of scleractinians at a polluted site which could be directly attributable to a high diversity of faviids when compared with a control reef. Branch- ing corals, however, were restricted in both number of species and abundance at the polluted location, a feature masked by the high diversity value obtained overall.

Estimates of colony size of scleractinians, recruitment, mortality of juveniles, and spatial distribution of living and dead cover in permanent quadrats over time enable man-made disturbances to be evaluated with respect to natural fluctuations at both polluted and unpolluted sites (Bak and Engel, 1979; Bak and Luckhurst, 1980; Bak and Criens, 1983). Such observations constitute a vitally important part of any monitoring pro- gram.

3. Experimentahfanipulations in the field offer considerable scope in assessing the effects of both man-made (Hudson, 1981; Neudecker, 1983) and natural (Bak and Criens, 1983) disturbances. Where appropriate, sediment-shifting capabilities of scleractinians measured in the field (Rogers, 1979, 1984) provide comparative data on a variety of species in different habitats. In addition, contaminated sediments may be used to assess the

2.


relative tolerances of species from polluted and unpolluted sites under natural conditions. Growth rate measurements also offer a useful and sensitive technique for assessing stress in the field, especially when both branching and massive species are analyzed at selected sites. Unfortu- nately, very few studies to date have incorporated measurements of both types of coral, so we have no comparative data on tolerances of these species to stress. Techniques such as x-radiography and alizarin staining lend themselves well to transplantation of corals to different habitats, which again enables intra- and interspecific differences between sites to be established. Assessment of the reproductive potential of corals transplanted to different environments may also highlight the effects of stress at polluted sites. Measurement of the number of planulae produced per colony and the fecundity of individual species will ultimately reflect the structure of the whole community. Such techniques (Rinkevich and Loya, 1977) have already clearly demonstrated that pollutants such as oil, previously thought to be relatively harmless to adult corals (for review see Johannes, 1975), are in fact capable of exerting a significant effect on reproductive processes. Asexual reproduction, recognized as an important process on coral reefs (Highsmith, 1982; Tunnicliffe, 1983), may also be assessed in unpolluted and polluted sites by experimental manipulation (Bak and Criens, 1983). Using alizarin red S staining, skeletal extension, weight of calcium carbonate deposited and calcification rates, and growth of fragments may be compared with intact colony measurements in different habitats.

Experimental biossays in the laboratory should be closely related to field measurements and if possible be used to extend or confirm observations made in the field. There is little convincing evidence in the literature reviewed so w o f a sensitive laboratory bioassay to any pollutant tested; neither is there any indication of how many of the parameters measured relate to the ultimate survival of the coral. Work in progress in Bermuda (S. Wyers, personal communication) attempts to combine a series of laboratory assays on the effects of oil on scleractinians with a closely monitored recovery period in the field. Overall, however, results of laboratory exposure of corals to pollutants tend to be very variable, particularly with respect to behavioural responses, and it may well be argued that a new approach is urgently required here. Such an approach has already been adopted with temperate marine organisms in the use of a “scope for growth” model which measures the response of individuals to environmental stress and pollution (Bayne and Widdows, 1978). These authors argue that it is unlikely that analyses of physiological parameters such as behaviour, growth, and reproduction will provide information on

4.

54 B . E. BROWN AND L. S. HOWARD

specific environmental stresses since changes in the biochemical targets of particular pollutants will be integrated into more generalized physiological changes as part of a response syndrome.

The equation for the scope for growth is as follows:

P = A - (R + U)

where P = production or scope for growth A = product of consumption and effi-

ciency of absorption of energy from food

R = respiratory heat loss U = energy lost as excreta

Using the mussel M. edulis from natural environments, the authors were able to show a reduced scope for growth in animals from polluted habitats. Such an equation may not be entirely applicable to corals, for it appears that less than 1% of assimilated input appears as growth (P. Spencer Davies, personal communication). In addition, it would be extremely difficult to measure the energy intake from carnivorous sources in many corals. However, energy budgets have been constructed for several species, including Pocillopora eydouxi (Milne, Edwards, and Haime) (P. Spencer Davies, personal communication). A study of the overall energy budget of the coral, together with measurements of growth rate, may therefore be more appropriate. Combining this approach in in situ studies with biochemical and cytochemical assays of coral tissues from field and laboratory exposures would permit a reasonable interpretation of the condition of corals in the environment.

The disadvantageLo3 these methods are clear-all require sophisti- cated equipment and techniques which are not always readily available at remote tropical laboratories. Although it is possible that any meaningful laboratory measurement of effects of stress in reef corals will involve a much more subtle approach than has been applied to date, the value of critical methods of evaluation of pollution in the field, such as those described here, should not be underestimated. On the basis of these field methods alone, considerable advances in our understanding of the tolerances of reef corals to effects of stress-both man-made and natural- could be made during the next decade.

Acknowledgments

We would like to acknowledge authors who allowed us to see copies of papers in press and also the staff of the library at the Marine Biological Association, Plymouth, especially David Moulder, for their cooperation.


We should also like to acknowledge the helpful comments of Sir Maurice Yonge, Dr. Peter Spencer Davies, Dr. Tony Stebbing, and Mr. Martin Le Tissier on the manuscript draft, and the useful discussion with Dr. B. Widdows and Dr. Moore of the Institute of Marine Envi- ronmental Research, Plymouth.

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