Experimental biology of coral reef ecosystems - CiteSeerX

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Journal of Experimental Marine Biology and Ecology

300 (2004) 217–252

Experimental biology of coral reef ecosystems

Michael P. Lesser*

Department of Zoology and Center for Marine Biology, University of New Hampshire, Durham, NH 03824, USA

Received 23 November 2003; received in revised form 18 December 2003; accepted 28 December 2003

Abstract

Coral reef ecosystems are at the crossroads. While significant gaps still exist in our understanding

of how ‘‘normal’’ reefs work, unprecedented changes in coral reef systems have forced the research

community to change its focus from basic research to understand how one of the most diverse

ecosystems in the world works to basic research with strong applied implications to alleviate

damage, save, or restore coral reef ecosystems. A wide range of stressors on local, regional, and

global spatial scales including over fishing, diseases, large-scale disturbance events, global climate

change (e.g., ozone depletion, global warming), and over population have all contributed to declines

in coral cover or phase shifts in community structure on time scales never observed before. Many of

these changes are directly or indirectly related to anthropogenically induced changes in the global

support network that affects all ecosystems. This review focuses on some recent advances in the

experimental biology of coral reef ecosystems, and in particular scleractinian corals, at all levels of

biological organization. Many of the areas of interest and techniques discussed reflect a progression

of technological advances in biology and ecology but have found unique and timely application in

the field of experimental coral reef biology. The review, by nature, will not be exhaustive and reflects

the author’s interests to a large degree. Because of the voluminous literature available, an attempt has

been made to capture the essential elements and references for each topic discussed.

D 2004 Elsevier B.V. All rights reserved.

Keywords: Coral reef ecosystems; Experimental biology; Global climate change

Scleractinian, or reef-building corals, are a central component to coral reef ecosystems

worldwide between 30jN and 30jS latitude and contribute to thousands of square

kilometers of critical marine habitat. The prolific growth rates (3–15 cm year�1) of

reef-building corals in optically clear, oligotrophic tropical seas are responsible for the

three-dimensional framework of coral reef systems (Fig. 1). While other organisms serve

0022-0981/$ - see front matter D 2004 Elsevier B.V. All rights reserved.

doi:10.1016/j.jembe.2003.12.027

* Tel.: +1-603-862-3442; fax: +1-603-862-3784.

E-mail address: [email protected] (M.P. Lesser).

Fig. 1. Underwater photograph of coral reef in Indonesia with almost 100% cover of Acropora sp. (Photograph by

M. Lesser).

M.P. Lesser / J. Exp. Mar. Biol. Ecol. 300 (2004) 217–252218

to consolidate the framework of the reef structure together (e.g. calcareous algae) and use

it as essential habitat (e.g. fish, algae, invertebrates and bacteria), corals are the functional

group that has contributed significantly to coral reef ecosystems for at least 200 million

years (Veron, 1995) and have built the primary structure of entire reefs, islands and such

massive oceanic barriers as the barrier reefs of Mesoamerica and Australia. Coral reefs are

a source of food and livelihood for at least 100 million people worldwide, support major

industries (fishing and tourism), play a key role in stabilizing coastlines, and their high

species and genetic diversity rivals that of tropical rainforests (Connell, 1978; Hoegh-

Guldberg, 1999). This biodiversity is now just beginning to be exploited in the search for

bioactive compounds that could benefit humankind (Quinn et al., 2002).

Unfortunately, coral reefs are also experiencing unparalleled levels of anthropogenically

induced stress. Current estimates on the rate of decline in the health of coral reefs and the

loss or change in community structure of reefs are of worldwide concern (Wilkinson, 2000).

It is estimated that a combination of physical, chemical and biological stresses will cause the

decline of between 40% to 60% of the world’s coral reefs over the next 50 years unless

appropriate steps are taken (Wilkinson, 2000). Until recently, global climate change was

seen as just one of many factors (e.g., eutrophication, coastal development, sedimentation,

over-fishing) responsible for the decline in the health of coral reefs (Wilkinson, 1999) while

the time scales of change due to global climate effects were believed to be slow and other

anthropogenic causes a higher priority for study. In 1998, however, an estimated 16% of the

world’s living corals were eliminated in a single warming event related to El Nino

(Wilkinson, 2000). During this event, sea temperatures warmed to 2–3 jC above long-

M.P. Lesser / J. Exp. Mar. Biol. Ecol. 300 (2004) 217–252 219

term average summer temperatures and resulted in a catastrophic ‘‘bleaching’’ event that

caused significant mortality of several species of coral (e.g., both the expulsion of

zooxanthallae and host tissue death occurred). The impact of this thermal event on the

percent cover of shallow coral reefs worldwide and the projection of continued rising sea

temperatures under greenhouse warming (Hoegh-Guldberg, 1999) has radically changed the

focus of a large proportion of the research community towards understanding the potential

impact of greenhouse-driven climate change on the world’s coral reefs. Bleaching as a result

of thermal stress is not the only threat from global climate change and coral reef biologists

from around the world have had to use new experimental tools at all levels of biological

organization in their efforts to understand how reefs work, determine which corals will

survive anthropogenically driven change, and predict what reefs will look like at the end of

the next century. In essence, who will be the winners and the losers (Loya et al., 2001)?

1. The coral–algal symbiosis

Coral reef communities contain a wide variety of mutualistic associations none more

important than the relationship between corals and their symbiotic dinoflagellates of the

genus Symbiodinium sp., commonly referred to as zooxanthellae. Scleractinian corals first

appeared in the Triassic (Veron, 1995), and it is widely accepted that their rapid ecological

success was directly related to the acquisition of dinoflagellate endosymbionts that enabled

the symbiosis to survive in oligotrophic and high solar irradiance habitats. Corals acquire

the majority of their energetic and nutrient requirements by two mechanisms: photosyn-

thesis by their zooxanthellae and heterotrophy, or the direct ingestion of zooplankton and

other organic particles in the water column by the cnidarian host. The zooxanthellae reside

within vacuoles in the cells of the host gastrodermis (Fig. 2a and b; Trench, 1979, 1987)

where they serve as primary producers and supply their coral host with up to 95% of their

photosynthetic products, such as sugars, amino acids, carbohydrates and small peptides

(Trench, 1979; Muscatine, 1990) making corals autotrophic with respect to carbon in most

habitats. These compounds provide the coral with energy for respiration, growth, and the

deposition of its CaCO3 skeleton (Muscatine, 1990).

Supplying translocated photosynthate to the host contributes significantly to the fitness

of the symbiosis (Muscatine, 1990; Mueller-Parker and D’Elia, 1997) while in return the

zooxanthellae receive essential nutrients such as ammonia, phosphate, and carbon dioxide

from the metabolic wastes of the coral (Trench, 1979; Mueller-Parker and D’Elia, 1997).

Additionally, photoautotrophy is not the only source of nutrition for corals. An increasing

amount of experimental evidence continues to document that heterotrophy in corals is

essential for providing nitrogen, phosphorus, and other nutrients which make it possible

for the coral host to use the available carbon skeletons for protein synthesis and other

essential metabolic requirements. Initially, the degree of heterotrophy appeared to be

positively correlated with coral polyp size (Porter, 1976). Porter (1976) described a

bathymetric gradient from autotrophy in shallow waters to heterotrophy in deeper waters

that was correlated with polyp size in the Caribbean. Species with small polyps that were

more dependent on autotrophy were found in shallow waters while more heterotrophic

large polyp species of coral were found in deep waters (Porter, 1976). Clearly, heterotro-

Fig. 2. (a) Electron micrograph of zooxanthellae in hospite. (b) Phase-contrast micrograph of zooxanthellae in

tentacle squash preparation (Photographs by M. Lesser and T. LaJeuness).


phy in corals is important. Glynn (1973) described plankton depletion on a coral reef as

water flowed past and Wellington (1982) provided experimental, multifactorial, evidence

that supported Porter’s autotrophy to heterotrophy gradient, but also showed that

heterotrophy did not compensate for the decrease in solar irradiance with depth when

growth rates were measured. Recently, Sebens and colleagues (Sebens and Johnson, 1991;

Johnson and Sebens, 1993; Sebens et al., 1996, 1998) have shown quite convincingly that

both small and large polyped corals are successful at capturing certain size classes of

zooplankton and that any differences in the efficiency of capture were due largely to the

escape ability of the zooplankton.

Whether from autotrophy or heterotrophy, the tight recycling of nutrients within the

coral symbiosis and the close coupling between trophic levels at reasonably high


efficiencies contribute to the very high productivity of corals (Muscatine and Porter, 1977;

Falkowski et al., 1984; Muscatine, 1990; Mueller-Parker and D’Elia, 1997). Nutrient

limitation imposed by the host on the algal symbionts is also believed to be part of a highly

regulated control mechanism on the growth of zooxanthellae that would otherwise out-

divide their host cells at rates approaching those of free-living phytoplankton (Muscatine

and Porter, 1977). From an organismal and experimental perspective, it would appear that

the role of autotrophy and heterotrophy in the energetics and nutrient metabolism of corals

should be vigorously revisited. This will require simultaneous and interdisciplinary studies

by groups of collaborators on a range of coral species in different habitats using a range of

tools (e.g., fluorescence measurements, feeding studies, stable isotopes) to fill in what

appear to be large gaps in our understanding. By definition mutualistic associations incur

both benefits and costs for the partnered species. For any mutualistic symbiosis to develop

and persist, a constant evaluation of the costs and benefits must be occurring such that the

selective pressure favors those associations where the benefit to both partners outweighs

the costs (Cushman and Beattie, 1991). Under the continuing scenario of rapid change on

coral reefs, it is important to understand, at an organismal level, which species will survive

in the broad range of trophic strategies that span the dependence on autotrophy versus

heterotrophy.

2. Hurricanes, overfishing, eutrophication, bleaching, and community phase shifts

Both the growth forms and species of corals show typical and well-described zonational

patterns on reefs worldwide (Loya, 1972; Huston, 1985; Done, 1995). While heterogeneity

exists, species diversity along a bathymetric gradient is predictable to a certain degree and

reflects both biotic and abiotic processes. Much of the recent ecological work on coral reefs

has been framed around the concept that reefs are non-equilibrium systems whose

community structure and diversity are largely determined by the intensity and rate of

disturbance as described in the intermediate disturbance hypothesis (Connell, 1978, 1997;

Connell et al., 1997). Additionally, strong latitudinal and bathymetric gradients in abiotic

factors such as solar irradiance, water flow, and calcium carbonate saturation state

significantly influence the community structure, growth forms, and state of photoacclima-

tization over both small and large spatial scales (Falkowski et al., 1990; Done, 1995;

Wilkinson, 1999; Lesser et al., 2000). The scale-dependent variability in coral reef

community structure continues to be an important area of study for understanding not

only the range of scales at which different patterns occur but also what processes at different

scales may be driving that variability (Murdoch and Aronson, 1999).

The current concern by coral reef biologists is that the periodicity and intensity of

disturbance events, which now include a suite of anthropogenic factors over large (e.g.,

kilometer) spatial scales, is rapidly changing coral reefs and threatening their existence

which is in juxtaposition to the long-term persistence of coral reefs over geological time

scales (Pandolphi, 1999). Most coral reef biologists do agree that coral reefs are changing

and will exist in the near future but they will not be the ‘‘coral reefs’’ we have come to know

inmany parts of the world (Knowlton, 2001). The outcome on each reef systemwill likely be

determined by a combination of the number and severity of insults, but also which set of the


unique and varied life-history traits will be able to cope with these stressors on ecological

and evolutionary time scales (Hughes et al., 1992; Done et al., 1996).

Jackson et al. (2001) demonstrate from several sources of historical data that a range of

disturbances including overfishing and coastal development have consistently led to major

changes in coral reefs ecosystem structure and health. The most poignant example of the

effects of anthropogenic influences is the state of reefs in the Caribbean. A recent meta-

analysis of coral cover throughout the Caribbean has shown an 80% decline in percent

coral cover that has been both long-term (e.g., decadal) and region-wide (Gardner et al.,

2003). Though many reefs worldwide have suffered similar reductions in coral cover

(McClanahan, 2002), most Caribbean reefs have undergone a shift from being coral-

dominated to algal-dominated in this time period (Hughes, 1994). The causes of this shift

vary from reef to reef but are the result of several types of disturbance that include

hurricane damage (Hughes, 1994; Hughes and Connell, 1999), eutrophication (Lapointe,

1997), thermal stress resulting in coral bleaching (Hoegh-Guldberg, 1999; Aronson et al.,

2000, 2002; Ostrander et al., 2000), coral diseases (Harvell et al., 2002; Richardson, 1998;

Rosenberg and Ben-Haim, 2002), the transport and deposition of sand and dust from the

Sahara in the Caribbean, which may be a factor that partially explains the increase in coral

diseases (Shinn et al., 2000), and reduced herbivory from over-fishing compounded by an

epizootic of unknown etiology that decimated Diadema populations in the 1980s

(Carpenter, 1988; Hughes, 1994).

Hughes (1994) described the rapid and significant ecological changes that occurred on

coral reefs in Jamaica when herbivores were removed by fishing, to the point where reef

resilience (i.e. ability to recover from a disturbance) was lost and a permanent phase shift to

algal-dominated communities began. Additionally, natural factors conspired with anthro-

pogenic stresses to produce this outcome. First, Hurricane Allen, a category five hurricane

struck Jamaica after almost 40 years without any significant storm damage to coral reefs.

While most of the damage occurred in shallow waters (<10 m) a period of recovery began

but was short-lived due to the loss of the sea urchin, Diadema antillarum, between 1982

and 1984 from an epizootic disease of unknown etiology. This Caribbean-wide loss of a

critical herbivore that controlled algal growth led to significant changes on these coral reefs.

Without urchin and fish herbivores on coral reefs around the Caribbean, large populations

of foliose macrophytes formed and prevented coral settlement and growth with the net

result that coral cover dramatically declined. Significant bleaching events beginning in the

late 1980s also caused coral mortality and further shifted the change to algal-dominated

reefs. This scenario has been repeated in many parts of the world as key elements of coral

reef communities, such as grazing fishes and invertebrates, have been removed resulting in

dramatic changes in community structure (Lewis, 1986; Jackson et al., 2001). Recently,

however, there is evidence that Diadema populations may be recovering. Edmunds and

Carpenter (2001) reported that at several sites along the North coast of Jamaica, urchin

populations have increased significantly by 10-fold and is correlated with an 11-fold

increase in the density of juvenile corals. Recovery of urchins may initiate beginning of a

shift from one alternative stable state to another throughout Jamaica and the rest of the

Caribbean.

While the work of Edmunds and Carpenter (2001) appears to clearly indicate that top-

down control of macroalgal growth is essential to ‘‘reef health’’ or maintenance of coral reef


community structure as we know it, many coral reef biologists worldwide are still

discussing what controls macroalgal abundance on coral reefs. The two sides of this debate

have fallen along the ‘‘bottom-up’’ versus ‘‘top-down’’ dichotomy. The controversy is

centered around whether algal blooms are kept in check by herbivory as discussed in a

critique of Lapointe (1997) by Hughes et al. (1999), or responding to nutrient availability at

critical ‘‘threshold concentrations’’ that determine the balance between algal bloom and

non-bloom conditions Lapointe (1997,1999). In reading these papers, one can appreciate

the fact that each group recognizes the potential role of the ecological process discussed by

the other, but neither group has incorporated the opposing ecological process in their

experimental evaluation of this issue. Complicating the debate had been the lack of rigorous

multifactorial field experiments examining both herbivory and nutrient concentrations in

the field. Additionally, many of the sites examined (e.g., Jamaica and the Florida Keys)

have also experienced repeated disturbances, both natural and anthropogenic (see above),

that complicate the assessment of whether ‘‘bottom-up’’ or ‘‘top-down’’ processes from

either of these sites can be universally applied to all reefs.

Miller et al. (1999), however, provide us with experimental insight into what may be

occurring along the Florida reef tract. First, their design was multifactorial (herbivory and

nutrients), well replicated (i.e., without pseudoreplication), and appropriately analyzed

despite the complications associated with the field work that occurred during their study. An

aspect of their paper that will be of interest to experimental ecologists is the unique method

of nutrient addition amongst treatments. Fertilizer spikes commonly used in gardens were

partially covered in paraffin, and entombed in cinder blocks with holes to allow for the

diffusion of nutrients into the environment. The results of the field experiments by Miller et

al. were interpreted in the context of the Littler and Littler (1991) ‘‘relative dominance

model’’ for corals, turf algae, crustose coralline algae, and foliose macrophytes and the

‘‘threshold nutrient’’ model of Lapointe (1997). The Miller et al. paper provides a test of the

relative dominance model, that supports a strong role for herbivore control of algae on coral

reefs while not supporting the ‘‘threshold nutrient’’ hypothesis of Lapointe (1997). Another

paper by Jompa and McCook (2002) experimentally demonstrated the subtle affects of

herbivory when nutrients were also in abundance. Their experimental design clearly

revealed that herbivores are the single most important factor affecting coral growth and

mortality but that nutrients can also have an effect, but only if herbivory is weak and allows

nutrients to control algal growth and subsequently competitive interactions with corals.

Finally, a recent paper by Aronson and Precht (2000) examined the effects of herbivory on

Jamaican reefs over time in a nonmanipulative manner, essentially a ‘‘natural experiment’’.

Despite the shortfalls of a natural experiment, the Aronson and Precht (2000)results support

the experimental work of Miller et al. (1999) and Jompa and McCook (2002) which both

show that herbivory, or top-down control of macroalgae on coral reefs, has a greater impact

than nutrients on the community structure of coral reefs.

3. Global climate change

There can be little doubt at this point in time that global climate change, principally the

emission of greenhouse gases (e.g., CO2, CH4), and its subsequent effects on seawater


temperature, calcium carbonate saturation point, large-scale changes in atmospheric/

oceanic coupling [e.g., El Nino-Southern Oscillation (ENSO)], and changes in sea level,

is occurring (Smith and Buddemeier, 1992; Huppert and Stone, 1998; Hoegh-Guldberg,

1999; Kleypas et al., 1999; Wilkinson, 1999; Crowley, 2000; Stott et al., 2000; Urban et

al., 2000; Wellington et al., 2001). There is concern that within the framework of

evolutionary adaptation, scleractinian corals will not be able to physiologically adapt at

the current rates of environmental changes (Gates and Edmunds, 1999). In particular,

increases in seawater temperatures are thought to be the primary cause of the unprece-

dented number of ‘‘coral bleaching’’ events since the early 1980s (Brown, 1997; Glynn,

1991, 1993), with predictions for continued increases of seawater temperature in the future

(Hoegh-Guldberg, 1999). In 1998, coral reefs experienced the largest and most widespread

thermally induced mass bleaching of corals ever recorded with an estimated 16% mortality

of the world’s living corals as a result of an ENSO event (Wilkinson, 2000).

Even within tropical and sub-tropical environments, temperature is a pervasive abiotic

factor controlling the distribution and abundance of corals. Many species exhibit

population specific ranges of temperature where growth, reproduction, and survival will

occur, and is also related to the mean temperature of the warmest month in that

geographical location. It is generally believed that corals are living very close to their

upper thermal limits (Jokiel and Coles, 1990) where temperature effects on respiration are

much greater than on photosynthesis, resulting in a decrease in the P/R ratio and reduction

in net photosynthesis during exposure to elevated temperatures. When seawater temper-

atures warm anywhere from 2 to 3 jC above long-term average summer temperatures,

corals exhibit the stress response known as bleaching. Several field and laboratory studies

on bleaching in corals and other symbiotic cnidarians have established a causal link

between temperature stress and bleaching (Hoegh-Guldberg and Smith, 1989; Jokiel and

Coles, 1990; Lesser et al., 1990; Glynn, 1991; Fitt et al., 1993; Lesser, 1997; Hoegh-

Guldberg, 1999; Coles and Brown, 2003), and the extent of bleaching, subsequent

mortality, and the underlying mechanism (s) are related to the magnitude of temperature

elevation and the duration of exposure for any individual event. As is typical, and

consistent with experimental results, the severity for coral bleaching events varies in space

and time. For example, during the 1998 ENSO event 48% of corals in the Indian Ocean

died while only 3–5% of corals died on the Great Barrier Reef. The 1998 bleaching event,

unlike any previously observed, began a new appreciation for the projection of rapidly

rising sea temperatures under greenhouse warming (Fig. 3) and has increased the concern

of scientists, governments, and the general population about the potential impact of

greenhouse-driven climate change on the world’s coral reefs (Hoegh-Guldberg, 1999;

Wilkinson et al., 1999; McClanahan et al., 2002).

The number and severity of coral bleaching events are believed to be a ‘‘biological

signal’’ (sensu Hughes, 2000) for the consequences of global climate change on coral reefs

that is occurring worldwide, and is predicted to continue if current trends persist (Hoegh-

Guldberg, 1999; Sheppard, 2003). Coral bleaching is defined here as a response to

environmental stress that leads to a series of cellular responses that culminates in the

expulsion of the symbiotic zooxanthellae from the coral host tissues causing a paling or

whitening of the affected coral (Fig. 4a and b). Defining bleaching is not a trivial issue. It

is important to put bleaching in the context of the seasonal cycling of zooxanthellae

Fig. 3. Sea surface temperature data generated by the global-ocean-ice-model forced by greenhouse emissions that

conform to the IPCC scenario IS92a (Tahiti, 17.5jS, 149.5jW; Phuket, 7.5jN, 98.5jE; Jamaica, 17.5jN, 76.5jW).

With permission from Ove Hoegh-Guldberg and Marine and Freshwater Research (SO: 839–866, 1999).


densities in reef corals, which has recently been documented (Fagoonee et al., 1999; Fitt et

al., 2000).The annual high in sea surface temperatures coincides with yearly lows in

zooxanthellae densities in all studies made to date (Stimson, 1997; Fagoonee et al., 1999;

Fitt et al., 2000) suggesting the phenomenon is universal amongst symboitic corals.

Seasonal cycles in the quantum yields of chlorophyll fluorescence of corals have also been

observed (Warner et al., 2002), revealing seasonal acclimatization in solar irradiance and

seawater temperature.

While thermal stress is seen as the principal cause of coral bleaching, other environ-

mental factors, including those that are affected by anthropogenic influences, can act

synergistically by effectively lowering the threshold temperature at which coral bleaching

occurs. The principal abiotic factor that has significant influence on the severity of

thermally induced coral bleaching is solar radiation, both its visible (photosynthetically

active radiation, PAR: 400–700 nm; Hoegh-Guldberg and Smith, 1989; Dunne and

Brown, 2001) and ultraviolet (UVR: 290–400 nm, UVB: 290–320 nm, UVA: 320–400

nm; Shick et al., 1996) components.

The global decrease of stratospheric ozone from anthropogenic inputs of chlorinated

fluorocarbons has resulted in an increase in the amount of harmful UVB radiation reaching

Fig. 4. (a) Underwater photo of bleached M. faveolata in the vicinity of Lee Stocking Island (Rainbow Gardens),

Bahamas (Photograph by M. Lesser). (b) Underwater photograph of bleached Acropora sp. on the Great Barrier

Reef (Photograph by R. Berkelmans).


the sea surface (Madronich et al., 1998). Although earlier concerns were centered on the

Antarctic, tropical ecosystems, with their smaller solar zenith angle and thinner layer of

ozone (Cutchis, 1982) have exposed tropical ecosystems over evolutionary time to higher

irradiances of UVR, and UVB in particular (Green et al., 1974; Frederick et al., 1989). In

absolute terms, even a small percentage decrease in stratospheric ozone over the tropics

would be important because the UVB irradiance there is already high. The optical

properties of tropical waters also result in low attenuation coefficients and allow UVR

to penetrate to depths of 15 m or more (Fig. 5, Smith and Baker, 1979; Gleason and

Wellington, 1993; Shick et al., 1996; Lesser, 2000; Lesser and Gorbunov, 2001). Although

tropical waters are generally more transparent to UVR than temperate waters, the water

Fig. 5. Depth profile of spectral irradiance data (300–700 nm) collected from the outer fore reef at Carrie Bow

Cay, Belize (17jN) in Spring 1996 using a LiCor 1800 UW scanning spectroradiometer. For details of instrument,

see Lesser (2000). Lesser, unpublished data.


column overlying coral reefs in coastal areas is susceptible to terrigenous inputs,

upwelling, and variations in dissolved organic matter that can affect its optical properties

(absorption and scattering) and increase the attenuation of UVR (Kirk, 1994).

Ultraviolet radiation is known to have a detrimental effect on photosynthesis and

growth in zooxanthellae (Shick et al., 1996) with the harmful effects of UVR involving

damage to DNA, proteins, and lipids. This damage may be the result of both the direct and

indirect effects of UVR on many cellular targets. For sessile corals, exposure to solar UVR

in shallow tropical waters is unavoidable and exposure to UVR is particularly important

during hyperoxic conditions (Dykens and Shick, 1982; Kuhl et al., 1995) that occur

intracellularly in corals during photosynthesis and leads to the photodynamic production

of reactive oxygen species (ROS) (Valenzeno and Pooler, 1987).

An important response of corals during exposure to UVR includes the synthesis of UVR-

absorbing compounds and enzymes involved in the protection of both the host and symbiont

from oxidative stress (Dykens, 1984; Dykens and Shick, 1984; Lesser and Shick, 1989;

Dykens et al., 1992; Shick et al., 1996; Lesser, 1996; Shick and Dunlap, 2000; Brown et al.,

2002; Lesser and Farrell, in press). UVR-absorbing compounds are believed to provide

protection from the high-energy wavelengths within the UVR part of the spectrum by

providing a broad-band filter as these compounds have absorption maxima in the UVR


portion of the spectrum from f310 to 360 nm. The concentration of UVR-absorbing

compounds [mycosporine-like amino acids (MAAs)] in corals shows an exponential

decrease with depth (Dunlap et al., 1986; Banaszak et al., 1998; Lesser, 2000) and

experimental evidence has shown that MAAs are produced by the symbiotic zooxanthellae

(Shick et al., 1999). MAA concentrations in corals, zoanthids and cultured zooxanthellae

have been shown to decrease upon exposure to elevated seawater temperatures (Lesser et al.,

1990; Glynn et al., 1992; Lesser, 1996; Shick and Dunlap, 2000), potentially leaving both

host tissues and symbiotic zooxanthellae more susceptible to biological damage caused by

exposure to UVR. Lastly, the MAA mycosporine-glycine is now know to also have

antioxidant activity (Dunlap and Yamamoto, 1995; Kim et al., 2001; Suh et al., 2003).

Other than Shick et al. (1999), there is no published work on the shikimic acid pathway in

zooxanthellae which synthesizes MAAs, and other environmental factors that could

influence carbon flux through this pathway. Jokiel et al. (1997) showed that differences in

the rates of water flow could influence MAA concentration under identical solar irradiances,

suggesting that photosynthetic rates andMAA synthesis are also under the control of carbon

delivery due to differences in the thickness of the diffusional boundary layer (Lesser et al.,

1994; Jokiel et al., 1997). In addtion to studies that show an increase in MAA concentration

upon exposure to UVR, Jokiel et al. (1997) suggested that the syntheses of MAAs may also

be controlled by the flux of carbon through the shikimic acid pathway which is in turn

affected by rates of photosynthesis. This MAAs follows carbon flux scenario has not

received general acceptance because Jokiel et al. (1997) did not describe a specific

stoichiometry between carbon flux and MAA concentration (Shick et al., 2000). Shick et

al. (2000), however, argues against MAAs synthesis being affected by carbon flux using the

calculations of photosynthesis and MAA concentration from laboratory experiments on a

free-living dinoflagellate (Neale et al., 1998). In the Neale et al. (1998) study the

concentration of MAAs was normalized to chlorophyll content, a co-varying factor that

overstates any differences inMAA concentration when cultures photoacclimated to different

irradiance regimes are compared. In any case the effects of changes in carbon flux on MAA

synthesis through the shikimic acid pathway has not received the appropriate experimental

attention to clearly show whether there is any relationship between carbon flux and MAA

concentrations and what other environmental factors affect the shikimic acid pathway and

subsequent biosynthesis of MAAs.

Exposure to elevated temperatures alone (Iglesias-Prieto et al., 1992), UVR alone

(Lesser and Shick, 1989), or in combination (Lesser, 1996, 1997) can result in photo-

inhibition of photosynthesis in zooxanthellae. Photoinhibition occurs as a result of the

reduction in photosynthetic electron transport, combined with the continued high absorp-

tion of excitation energy (Osmond, 1981). One consequence of reducing electron transport

is the production of ROS such as singlet oxygen [1O2] superoxide radicals [O2�], hydrogen

peroxide [H2O2], and hydroxyl radicals [OH], for which there are many cellular targets

including photosystem II and the primary carboxylating enzyme, Rubisco in zooxanthellae

(Lesser, 1996). The enzymes superoxide dismutase, catalase, and ascorbate peroxidase act

in concert to inactivate superoxide radicals and hydrogen peroxide, thereby preventing the

formation of the most reactive form of ROS, the hydroxyl radical, and subsequent cellular

damage (Fridovich, 1986). Enzymic defenses in the animal host occur in proportion to the

potential for photooxidative damage in symbiotic cnidarians (Dykens and Shick, 1982;


Dykens et al., 1992). However, high fluxes of ROS in the host (Dykens et al., 1992; Nii

and Muscatine, 1997) or zooxanthellae (Lesser, 1996) can overwhelm the protective

enzymatic response and result in hydroxyl radical production via the Fenton reaction

(Asada and Takahashi, 1987). Oxidative stress has been proposed as a unifying mecha-

nism for several environmental insults that cause bleaching (Lesser, 1996). Oxidative

stress can lead to bleaching of zooxanthellae via exocytosis from coral host cells (Lesser,

1996, 1997) or apoptosis (Gates et al., 1992; Dunn et al., 2002; Lesser and Farrell, in

press). A cellular model of bleaching in symbiotic cnidarians has been developed and

includes oxidative stress, PSII damage, DNA damage, and apoptosis as underlying

processes (Lesser et al., 1990; Gates et al., 1992; Lesser, 1996, 1997; Warner et al.,

1999; Lesser and Farrell, in press). This model is consistent with a variety of biomarker

proteins expressed in corals during thermal stress (Brown et al., 2002; Downs et al., 2000,

2002).

Damage to photosystem II (PSII) reaction centers in the zooxanthellae, specifically at

the D1 protein of PSII, following exposure to elevated temperatures and solar radiation, is

believed to be an important factor leading to the bleaching of corals (Iglesias-Prieto et al.,

1992; Lesser, 1996; Warner et al., 1996, 1999), and caused by ROS (Lesser, 1997; Richter

et al., 1990). Damage or impairment of PSII function is easily detected using nondestruc-

tive active chlorophyll fluorescence techniques (Brown et al., 1999; Jones et al., 1998;

Gorbunov et al., 2001; Warner et al., 1999; Winters et al., 2003). Instruments have been

developed that incorporate protocols to measure the multiple photochemical turnover

[pulse amplitude-modulated (PAM)], and single photochemical turnover of PSII [fast

repetition rate (FRR)] in the laboratory and in the field (Schreiber et al., 1986; Gorbunov et

al., 2000). These instruments measure, nondestructively, fluorescent transients that provide

information on the efficiency of PSII and can discern chronic photoinhibition from

dynamic photoinhibition, the former representing damage to PSII and the latter a

protective regulatory response of the photosynthetic apparatus. The underwater FRR has

been used to examine diel cycling and dynamic versus chronic photoinhibition of corals in

shallow and deep waters (Gorbunov et al., 2001; Lesser and Gorbunov, 2001). One

advantage of the FRR versus the PAM instrument is that because of the protocol used to

measure fluorescent transients, a series of flashlets that saturate PSII in microseconds (Fig.

6a), this instrument can also measure the optical cross section of PSII which is a valuable

parameter for discerning dynamic versus chronic photoinhibition (Gorbunov et al., 2000).

An underwater version of the PAM instrument is commercially available and is capable of

examining the photoacclimatization state of corals by measuring the relationship between

electron transport rates and irradiance (Fig. 6b) which can be interpreted, with caution, to

traditional photosynthesis versus irradiance (P– I) curves. Using these data and non-linear

fitting techniques the user can then can fit curves to the data and calculate photosynthetic

parameters. The instrument has also been widely used to study diel changes in the

quantum yield of PSII fluorescence and its relationship to differences between photo-

chemical and non-photochemical quenching, or dynamic photoinhibition (Brown et al.,

1999; Hoegh-Guldberg and Jones, 1999). Brown et al. (1999) suggested that the diurnal

patterns in quantum yields of PSII and xanthophyll cycling they observed were suggestive

of photoinhibition followed by photoprotection. Hoegh-Guldberg and Jones (1999) and

Jones et al. (1998) observed similar patterns and suggested that sink limitations were also


important in regulating the quantum yields of PSII. Warner et al. (1999) and Lesser and

Farrell (in press) have correlated changes in PSII fluorescence with changes in the

concentration of D1 protein during exposure to thermal stress and/or solar radiation.

Other models of thermally induced bleaching have suggested that the dark reactions of

photosynthesis are affected initially, leading to sink limitation, overreduction of photo-

synthetic electron transport, oxidative stress, and damage to PSII (Jones et al., 1998).

From the available data, Lesser and Farrell (in press) have proposed a model of

bleaching induced by damage to PSII that incorporates simultaneous damage to both

photochemistry and carbon fixation in a feedback loop that greatly enhances the damage to

PSII. There is evidence that PSII is already affected directly by high solar radiation without

any thermal stress (Gorbunov et al., 2001) and that thermal stress without high solar

irradiances affects PSII (Warner et al., 1999). Additionally, high levels of ROS are a

consistent feature of coral physiology, especially in the presence of thermal stress and UVR

in both the symbiont and host (Lesser, 1996, 1997; Lesser and Farrell, in press). Reactive

oxygen species, especially hydrogen peroxide, are well-known inhibitors of Rubisco

(Asada and Takahashi, 1987) and cause damage to PSII (Asada and Takahashi, 1987;

Richter et al., 1990). The work by Jones et al. (1998) illustrated the importance of carbon

sink limitation in exacerbating this damage. This observation is significant because carbon

limitation has been observed in shallow water corals (Muscatine et al., 1989), and can be

significantly effected by water flow (Lesser et al., 1994). Damage to PSII and sink

limitation under these conditions are likely to be occurring simultaneously with thermal

stress with the resulting formation of ROS which overwhelms all of the host and algal

antioxidant defense systems. The cascade of events that ultimately induces the expulsion of

zooxanthellae from their host could include the accumulation of ROS and damage to PSII

or a decrease in the amount of translocated photosynthate, or both (Lesser et al., 1990).

The host also responds to thermal stress. In particular, heat shock proteins (HSPs) are

up-regulated in response to thermal stress (Black et al., 1995; Fang et al., 1997; Sharp et

al., 1997). Heat shock proteins are inducible by a number of environmental factors and

appear to be a generalized stress response that is evolutionarily conserved. Under stressful

conditions, HSPs interact with proteins to maintain their conformation and function or in

targeting damaged proteins for degradation. This function is also consistent with patterns

of expression in markers of protein degradation observed in corals (Downs et al., 2000,

2002). Studies on the effect of UVR and thermal stress on corals have also shown

significant DNA damage in host tissues upon exposure to UVR (Anderson et al., 2001)

and thermal stress combined with exposure to solar radiation (Lesser and Farrell, in press).

DNA damage can lead to apoptosis or programmed cell death if not repaired. One of the

key cell cycle genes activated after DNA damage is p53. If DNA repair is not possible,

then p53-mediated apoptosis may be initiated. The expression pattern of a putative p53

protein in Montastraea faveolata after exposure to thermal stress and high irradiances of

Fig. 6. (a) Fast repetition rate fluorometer (FRRF) measurements on the coral M. faveolata from 1–2 m at Lee

Stocking Island, Bahamas. Note difference between dark adapted (Night) and steady state (Day) quantum yield

measurements of PSII fluorescence. Lesser, unpublished data. (b) Pulse amplitude-modulated (PAM) rapid light

curves on M. faveolata from different depths at Carrie Bow, Cay, Belize. Note the sensitivity of electron transport

rates (ETR) and depression of maximum rates at higher irradiances with depth as observed with photosynthesis–

irradiance ( P versus I ) curves. Lesser, unpublished data.


solar radiation was consistent with the observed pattern of DNA damage (Lesser and

Farrell, in press).

Recent work has shown that both apoptosis and cell necrosis are occurring in host and

algal cells of thermally stressed symbiotic sea anemones (Dunn et al., 2002). Based on the

ultrastructural evidence that apoptosis and necrosis both occur in thermally stressed

symbiotic cnidarians, and that a putative p53 protein is up-regulated in response to

DNA damage, the data are supportive for the occurrence of apoptosis and possibly cell

necrosis mediated by ROS in thermally stressed symbiotic cnidarians. Apoptosis and cell

necrosis are the extreme case in a range of likely cellular responses to thermal stress in

corals (Gates et al., 1992, Fig. 7). As in the case of community-wide responses to thermal

stress, the cellular mechanism of bleaching is a function of both susceptibility to and

severity of the environmental stress.

Coral bleaching results in the breakdown of a mutualistic symbiosis that is essential for

the survival of corals. There is growing evidence that the range of responses of corals to

environmental stress (Fitt et al., 2001) is also a function of the genotype(s) of zooxan-

thellae within the host. The availability of molecular genetic data on zooxanthellae

Fig. 7. Range of mechanisms for zooxanthellae expulsion with increasing severity of environmental stresses.

ZX=zooxanthellae (adapted from Gates et al., 1992). With permission of Tracey Saxby.


genotypes (see below) and their micro- and macroscale distributions will very likely play a

significant role in who are the winners and losers under any continuing scenario of global

climate change.

4. Coral symbiont systematics

Until recently most symbiotic dinoflagellates were considered to be members of a

single pandemic species, Symbiodinium microadriaticum. Early work by Schoenberg and

Trench (1980a,b) showed clear physiological and biochemical differences between many

cultured zooxanthellae from different hosts which were characteristics of different species,

and culminating in work by Blank and Trench (1985) and Trench and Blank (1987) which

used many independent lines of evidence to show that several different species of

zooxanthellae do exist and that there were probably more undescribed species in this

genus. Rowan and Powers (1991a,b) then used molecular genetic tools, restriction

fragment length polymorphisms (RFLPs) of the small ribosomal subunit (ssRNA) and

sequencing of ssRNA, to show that the zooxanthellae of reef-building corals and other

symbiotic invertebrates are a highly diverse group of organisms organized at that time into

three major ‘‘clades’’; A, B, and C. Additionally, many corals contain as many as two or

three species per host that appear to be ecologically segregated based on small-scale

gradients in their physical environment (e.g., light) (Rowan and Knowlton, 1995; Rowan

et al., 1997). Additional work using RFLP genotypes using ssRNA and large subunit

ribosomal RNA (lsRNA) (Baker and Rowan, 1997), chloroplast 23S-rDNA sequencing

(Santos et al., 2002), and sequencing of the internal transcribed spacer regions (ITS)

(LaJeunesse, 2001) have all provided a wealth of information on the diversity of this genus

(Baker, 2003) and its seven clades (A–G) along with host diversity (Fig. 8a) to the point

where we can now begin to compare within and between regional diversity over large

spatial scales (LaJeuness, 2002; LaJeuness et al., 2003).

More interesting experimental work remains to be done on whether all of this genetic

diversity translates into physiological diversity and the ability to tolerate various types of

environmental stress. Different clades have already been assigned different functional

groupings (e.g., stress-tolerant generalists and narrowly adapted specialists) based on

latitudinal or bathymetric gradients (Fig. 8b) in clade type within a single host species

(Rodriguez-Lanetty et al., 2001; LaJeuness, 2002), on small-scale spatial patterns of

multiple genotypes within a single coral (Rowan and Knowlton, 1995; Toller et al.,

2001a), or on repopulation studies of bleached corals (Rowan et al., 1997; Toller et al.,

2001b). Much of the significance of these studies has been attributed to putative

physiological differences between clades. While simultaneous works on genetics and

temperature tolerances have revealed some differences (Kinzie et al., 2001; LaJeuness et

al., 2003), other works on general photosynthetic capabilities have shown less difference

between clades (Savage et al., 2002).

This brings us to the vigorously debated issue of the Adaptive Bleaching Hypothesis

(ABH) originally formulated by Buddemeier and Fautin (1993). While environmental

stress can reduce fitness, it is believed that under certain conditions the same stress can

result in positive, directional, selection for phenotypes that arose from the generation and

Fig. 8. (a) Cladogram of currently known groups of zooxanthellae (Symbiodinium sp.) based on ITS 2 sequence

data and associated hosts. Numbers at branch nodes are boot strap values. (b) Community structure of holobionts

and distribution of Symbiodinium sp. Types as determined by ITS 2 sequencing at Puerto Moreles, Mexico

(20j50VN, 86j52VW). Symbiont types are compared to the number of host species associated with that type at

each depth. With permission from T. LaJeuness and Marine Biology (141: 387–400, 2002).



maintenance of variability stimulated by that stress (Hoffmann and Hercus, 2000). Other

studies have shown generally that significant constraints to adaptive evolution in response

to global warming exist (Etterson and Shaw, 2001). The ABH contends that after bleaching

occurs, the ‘‘shuffling’’ of zooxanthellae genotypes, from the large number currently

available and presumably representing a range of thermal tolerances, is possible and

adaptive in an ecological and evolutionary sense for the holobiont (Buddemeier and Fautin,

1993; Ware et al., 1996). It is believed that corals bleach and can survive future episodes of

thermal stress by allowing zooxanthellae genotypes with greater thermal tolerances that

were not initially expelled to become the new dominant genotype residing in the host

tissues (Baker, 2001, 2003). Several of the assumptions of the ABH have been experi-

mentally tested on a limited number of zooxanthellae cultures in terms of thermal tolerance

and shown to be related to their placement in specific clades as discussed above and would

therefore appear to be consistent with the ABH (Kinzie et al., 2001). Recently, Toller et al.

(2001b) showed that Montastraea sp. corals experimentally bleached and placed back in

the field were repopulated with zooxanthellae from different clades. However, the same

occurred for field populations that had not bleached, suggesting that thermal stress, and

subsequent bleaching, is not the only mechanism for symbiont ‘‘shuffling’’.

As the number and severity of bleaching events increases, the percent mortality of

corals has increased significantly (Hoegh-Guldberg, 1999), and the change in temperature

required for bleaching in many species of coral, which is already small (Jokiel and Coles,

1990), will be exceeded for most species of corals. Under these conditions, any

catastrophic mortality that would occur might be perceived as a limitation for the

opportunity of zooxanthellae ‘‘shuffling’’ consistent with the ABH. While most critiques

of the ABH have suggested that there is little evidence in the last 20 years that bleached

corals (i.e., the holobiont) have adapted to thermal stress, there is some evidence that

suggests otherwise. A prediction of the ABH is that corals that have recovered from prior

bleaching should be more resistant to subsequent thermal stress because of a change in

their symbiont composition compared to corals that have not experienced bleaching. In the

tropical far eastern Pacific, severe coral bleaching during the 1982–1983 El Nino resulted

in coral mortality of 50–90% (Glynn, 1988). The survivors of this event avoided

significant bleaching or mortality during the 1997–1998 El Nino even though the

temperature anomalies were as high in 1997–1998 as they were in 1982–1983,

presumably as a result of changes in zooxanthellae genotype (Glynn et al., 2001). The

corals in the eastern Pacific are now reverting back to their original, temperature-sensitive,

genotype (Baker, personal communication). This potentially adds another complication

(i.e., time dependence of genotype stability) to the ABH. It may be that only corals

experiencing chronically high temperatures (e.g., Arabian Gulf) are able to maintain

zooxanthellae genotypes that exhibit greater tolerances to thermal stress.

The ABH provides an experimental framework to examine interesting questions about

the stress response of corals and their symbionts. A question consistently raised is whether

we are now dealing with time scales and rates of environmental change where the results

of testing the ABH would be useful in the context of large-scale changes on coral reefs.

There is no agreement on this currently; however, producing new experimental work on

the thermal tolerances, genetic diversity, and infection capabilities of zooxanthellae is

essential at this time. We now have the molecular genetic tools to identify and follow


zooxanthellae populations over space and time, but little knowledge about the range of

environmental tolerances of those symbionts.

5. Coral host systematics

Presently there are approximately 1300 species of scleractinian corals (Veron, 1995,

2000) that have been identified primarily by morphological characters (e.g., corallite

structure). Molecular data are now being applied routinely to answer questions regarding

taxonomic and evolutionary questions on corals. Molecular markers such as ribosomal

RNA, the internal transcribed region (ITS), or amplified fragment-length polymorphisms

(AFLP) have been used to answer interesting questions on systematics, presence of

sibling species, and hybridization. Recent molecular and morphological studies have

clearly placed the Anthozoa as the basal class within the phylum Cnidaria (Bridge et al.,

1995) with the Scleractinia embedded within the Class Anthozoa and probably evolved

from the Corallimorpharia (Fautin and Lowenstein, 1992; Stanley and Fautin, 2001).

Zooxanthellate scleractinian corals have been present since the mid-Triassic (Veron,

1995) and based on mitochondrial 16S rDNA, two distinct clades of scleractinian

families appear to have diverged before the appearance of modern scleractinian taxa in

the fossil record (Romano and Palumbi, 1996; Romano and Cairns, 2000). These results

do not support traditional morphological systematics, but instead a polyphyletic origin of

the scleractinian skeleton. While understanding the evolution of these higher taxonomic

levels is problematic, the identification of coral ‘‘species’’ is no less troublesome. For

several prominent and ecologically important species there are now several different lines

of evidence that support the presence of cryptic or sibling species. The most prominent

example is from the coral Montastrea annularis, which exhibits a high degree of

morphological variation that was ascribed to phenotypic plasticity in response to solar

irradiance or other abiotic factors. These sympatric morphotypes are now recognized as

at least three closely related species (Knowlton et al., 1992; Weil and Knowlton, 1994)

known as Montastraea annularis, M. faveolata, and M. franksi. While character analysis

and assessments of reproductive isolation have generally supported the existence of these

sibling species (Knowlton et al., 1997), it is also known that hybrid larvae can be

produced and that intermediate coral morphologies exist in the field (Szmant et al.,

1997). Molecular analyses have been equivocal with an analysis using AFLPs and

microsattelites supporting reproductive isolation and therefore some degree of divergence

(Lopez et al., 1999), and sequencing of the ITS region and mitochondrial cytochrome

oxidase (COI) showing no support for distinct species (Medina et al., 1999). Other

groups of closely related morphotypes present similar questions regarding plasticity

versus sibling species. One explanation for the speciose nature and morphological

variability of certain groups is ‘‘reticulate evolution,’’ where repeated episodes of

hybridization and fusion of lineages occur over evolutionary time leading to variants

whose species boundaries become ‘‘fuzzy’’ (Veron, 1995). Recent molecular analyses

(i.e., ITS sequences) on several species in the genus Madracis from the Caribbean have

shown high levels of intraspecific and intra-individual variability (Diekmann et al.,

2001). Based on these data, and paleontological data, the most parsimonious interpre-


tation for these results was that reticulate evolution was the dominant mechanism

responsible for the observed variability. Probably the best example of reticulate evolution

in corals comes from the genus Acropora. In the Caribbean, three species of Acropora

exist; Acropora cervicornis, Acropora palmata, and Acropora prolifera. A. prolifera has

been described as an intermediate morphology and a hybrid between A. cervicornis and

A. palmata. Recent nuclear, mitochondrial, and ribosomal sequence data have supported

the hypothesis that A. prolifera is a hybrid and that the hybrids are long-lived and

propagate principally by asexual reproduction (Van Oppen et al., 2000; Vollmer and

Palumbi, 2002).

In the coming years we will see the next wave of molecular influence on coral biology

with the initiation of several coral genome projects. The case for a cnidarian genomics

database has been made (Ryan and Finnerty, 2003), and progress has been made on an

EST library for Acropora millepora (Kortschak, 2003) which has already revealed a

surprising number of conserved genes previously believed to have arisen during the

evolution of vertebrates. Which coral(s) to use as subjects for genomic sequencing, and

proteonomics as well, is under debate but the benefits for evolutionary, physiological, and

ecological studies on corals will be significant. The development of coral-specific micro-

and macroarrays for stress related markers, the continued identification of developmental

and cell cycle genes homologous with higher taxa in these tissue-grade diploblastic

animals, and DNA profiling to characterize genetic differences between individuals and

populations will benefit tremendously from the completion of one or more coral genome

projects (Ball et al., 2002; Gibson, 2002).

6. Fluorescent proteins

Recently, there has been a flurry of activity surrounding the identification of host

fluorescence (Kawaguti, 1969) in corals and other cnidarians as a group of homologous

fluorescent proteins related to green fluorescent protein (GFP; Fig. 9a and b; Matz et al.,

1999; Salih et al., 2000; Dove et al., 2001; Mazel et al., 2003) originally isolated and

described from the hydromedusae, Aequorea victoria (Tsein, 1998). The 238-amino-acid

protein, within which three residues at positions 65–67 form the active chromophore, is

extremely resistant to extremes in pH and temperature and requires the presence of oxygen

as it is translated and folds into its native configuration (Tsein, 1998). It requires no

cofactors for fluorescence, is not coupled to a bioluminescent (e.g., luciferin-luciferase)

system in corals, and in corals these fluorescent proteins are located principally in the

epithelial cells of the cnidarian host (Salih et al., 2000; Mazel et al., 2003), although they

can be observed in gastrodermal tissue as well (Salih et al., 2000). While considerable

interest in the utility of coral fluorescent proteins as transcription reporter genes has

dominated the literature, little is known about the function of this protein in corals. It has

been proposed that these fluorescent proteins provide photoprotection under high-light

conditions (Kawaguti, 1969; Salih et al., 2000), enhance photosynthesis under low-light

conditions (Salih et al., 2000), or both depending on the position of the fluorescent

pigment relative to the zooxanthellae (Salih et al., 2000; Dove et al., 2001). The

fluorescence can also contribute significantly to the spectral signature (e.g., reflectance)

Fig. 9. (a) Underwater photograph of Montastraea cavernosa taken under white light at Lee Stocking Island,

Bahamas. (b) Same coral taken under blue light excitation showing the dominant green light emission of a green

fluorescent protein in the epithelial cells of this coral (Mazel et al., 2003). Photographs by C. Mazel.


of corals under daylight illumination (Dove et al., 2001; Mazel and Fuchs, 2003) and

therefore the perceived color of corals by the human eye.

Many fluorescent proteins, based on fluorescence emission spectra and molecular data,

have been identified in corals (Mazel, 1995; Matz et al., 1999; Dove et al., 2001; Labas et

al., 2002; Kelmanson and Matz, 2003; Mazel et al., 2003), and recent data suggests they

may have arisen through gene duplications at several loci (Kelmanson and Matz, 2003).

Of the proposed functions described above, recent data have clearly shown that there is no

role for fluorescent proteins in enhancing photosynthesis under low-light conditions

(Gilmore et al., 2003; Mazel et al., 2003). Fluorescence resonance energy transfer (FRET)

between fluorescent proteins occurs but transfer from fluorescent proteins to chlorophyll


does not occur in corals. Additionally, for the Caribbean coral M. faveolata, there is no

evidence that fluorescent proteins protect the holobiont from the deleterious effects of

UVR (Mazel et al., 2003). While new fluorescent proteins continue to be discovered,

understanding their role in coral physiology and ecology remains elusive.

7. Coral diseases

One of the most significant changes on coral reefs along the Florida Keys Reef tract and

in the Caribbean generally has been the emergence of diseases and the potential

relationship to global climate change (Richardson, 1998; Rosenberg and Ben-Haim,

2002; Harvell et al., 2002). While coral bleaching is most commonly associated with

thermal stress and its physiological consequences, bleaching in at least one species of

coral, Oculina patagonica, is caused by the bacterium, Vibrio shiloi, subsequent to thermal

stress (Kushmaro et al., 1996). Additionally, the mortality of gorgonian corals, caused by a

fungal pathogen, Aspergillus sydowii, has been linked to transatlantic dust transport from

Africa that has increased in the last 25 years due to desertification (Shinn et al., 2000).

For many years the dominant, and ecologically significant, disease of corals was black

band disease (Rutzler et al., 1983; Kuta and Richardson, 1996), which is now known to be

a microbial consortium of cyanobacteria and sulfide-oxidizing bacteria that exploit

gradients of oxygen and sulfide while simultaneously producing anoxic zones that kill

the coral tissue (Richardson, 1998). Acropora cervicornis and A. palmata were once the

dominant species across the Caribbean including the Florida reef tract. In the 1980s,

however, disease (white band disease) resulted in almost the complete mortality of these

species changing the community structure of these reefs for the foreseeable future. White

band disease is now recognized as two variants of the same disease, type I and type II, with

the etiological agent believed to be a gram-negative bacterium (Richardson, 1998). Plague

or ‘‘white plague’’ is another disease of corals that also has two variants with plague type

II emerging in the mid-1990s and affecting mostly a single species of Caribbean coral,

Dichocoenia stokesi. The etiological agent for this disease has been identified as a new

species of Sphingomonas, a gram-negative bacterium (Richardson, 1998; Richardson et

al., 1998).

In the 1990s it appears that a suite of new coral diseases has emerged. Some of these

diseases are associated with elevated nutrients, either from agricultural runoff or from

human sewage. Recent evidence has shown that an enteric bacterium, Serratia marces-

cens, is responsible for the ‘‘white pox’’ disease of A. palmata (Patterson et al., 2002).

Continued development along coastal waterways paired with preexisting dwellings has

resulted in an enormous discharge of human sewage. While in many instances sewage is

first treated prior to discharge, evidence has demonstrated that not all of the infectious

organisms present in sewage are inactivated prior to their discharge. Several investigators

have noted alarming amounts of pathogenic organisms in and around the Florida canals

and around the Florida Keys (Lipp et al., 2002). Two issues have frustrated managers and

scientists interested in the study of coral diseases; nomenclature and a common set of

symptoms associated with a particular disease name, and isolation, identification, and

reinfection studies on the putative causative agents of these diseases. Fulfilling Koch’s


postulates has been a central tenant of studies with pathogenic microorganisms for over a

century and only the bacteria associated with black-band, white plague II, and white pox

coral diseases have fulfilled the requirements to be identified as the causative agent with a

high degree of confidence. An additional problem for these studies is culturability. Many

marine prokaryotes remain non-culturable unlike the majority of pathogens in clinical,

wildlife, or veterinarian settings. Molecular techniques such as quantitative PCR and

fluorescent in situ hybridization (FISH; Bythell et al., 2002) and others should provide

excellent proxies for the culturing usually required for fulfilling Koch’s postulates.

8. Remote sensing

Our understanding of coral reef ecology is still hampered by the inability to map and

monitor large expanses of reef area over any reasonable temporal scale. One way to assess

changes in the aerial coverage of coral reefs on large spatial and temporal scales is using

remote sensing imagery taken from airplanes or satellite (Fig. 10) platforms (Green et al.,

1996; Mumby et al., 1997). Several practical and analytical hurdles remain to be solved

before this approach can be widely used. There is a practical reason to examine reefs as

they relate to shallow water habitats and remote sensing. Coral reefs are generally found in

optically clear, Case I waters. Although chlorophyll concentrations of the waters over reefs

are typically 0.1 to 0.2 mg Chl a m�3, typical of open ocean waters, there is significant

absorption in the blue wavelengths because of benthic-derived CDOM (from corals and

Fig. 10. 1998 SeaWifs image of Florida, Florida Straights, and the Bahamas including the Bahama Banks and the

Tongue of the Ocean. (NASA archives).


sea grasses), and there is often extra scattering due to mineral particles derived from water

column precipitation of aragonite or resuspension of sediments. An optical approach to

monitoring coral reefs should include an understanding of the underlying reasons for

changes in the optical signal(s) of choice and whether one can generalize these

mechanisms from reef to reef. One goal of this approach is to obtain optical closure,

assessing which photons are absorbed, reflected, or re-emitted as fluorescence, essentially

a photon budget. Areas of high absorbance reflect the presence of primary producers

containing photosynthetic and accessory pigments whereas areas of low absorbance or

high reflectance indicate areas of low coral or algal cover. Another product of an optical

approach is to utilize reflectance or fluorescence signatures of the benthic community to

establish a reef classification scheme that can then be used with remote sensing imagery

(Mazel, 1995; Hochberg and Atkinson, 2000). Coral reefs therefore provide an excellent

test of our abilities to extract both water column optical and benthic properties of reefs

from remote sensing reflectance data.

Up to this point, when we compare traditional classification schemes (e.g., SCUBA

transects) to remote sensing classifications, the range of agreement between the two

methods is anywhere from 30% to 85%. This is not very satisfying at the moment but new

sensors and analytical approaches continue to be developed, as this is a high priority area

for managers and ecologists. Most remote sensing approaches still require a significant

amount of sea truthing over large spatial and temporal scales to validate the use of the

imagery and analytical routines. Several investigators have constructed irradiance reflec-

tance libraries of various functional groups (e.g., corals, macrophytes, seagrasses,

microbial mats) that span the possible pigment compositions observed on a reef and

can be used to develop algorithms (Hochberg and Atkinson, 2000; Hedley and Mumby,

2003b). Additionally, not all images are optimal for use in classification schemes. One

consistent problem is sea surface glint. Because of consistent wind patterns, sea surface

glint from wavelets often introduces errors in the water leaving radiances or remote

sensing reflectances. Routines to remove sea surface glint from remote sensing imagery

have recently been developed (Hochberg et al., 2003) and should contribute to more

accurate benthic classifications.

There has been significant interest in coral reef mapping using multispectral platforms

such as Landsat 7 and Ikonos imagery. The advantage of these satellite platforms is that

they provide global synoptic coverage of coral reefs anywhere from once to several times

annually and there is already a large database that has recently, and continues to be,

collected as part of the Long Term Acquisition Program (LTAP). There are presently

significant limitations in using these data sets for detailed bottom classifications but

progress has been made using multispectral data from coral reef environments. A

comparative study using the compact airborne spectrographic imager (CASI) showed that

CASI consistently outperformed satellite sensors (e.g. Landsat, SPOT) and aerial photog-

raphy in classifying bottom features (e.g., live corals, sand, seagrass) (Mumby et al., 1997,

2000). Recently, a radiative transfer approach and multispectral imagery was used on reefs

in Hawaii down to a depth of 25 m with accuracy exceeding 85% (Isoun et al., 2003).

Hyperspectral remote sensing imagery from airborne (e.g., CASI) and satellite (e.g.,

Hyperion) platforms holds the promise of providing the detailed information required for

remotely mapping coral reefs at 1-m2 resolution and consistent capabilities to discriminate


between functional end-members such as macroalgae and corals. Additionally, hyper-

spectral imaging spectrometers are viewed as a potentially important tool for the

assessment and management of tropical coastal resources (Green et al., 1996). New and

novel analytical approaches will be needed to take advantage of the additional information

contained within hyperspectral imagery. One of the disadvantages of current algorithms is

that because of the spectral mixing that occurs on certain types of reefs (e.g., Caribbean),

spectral, and therefore classification, information is lost during conventional supervised

routines. This translates into pixel to pixel errors because current algorithms are con-

strained to one pixel-one end-member classifications. Recognizing this problem has

resulted in new approaches of analysis and classification. Hochberg and Atkinson

(2000) and Andrefouet et al. (2003) have used derivative analysis coupled with linear

discriminant function analysis to analyze hyperspectral data for coral reefs in Hawaii and

French Polynesia. Hedley and Mumby (2003a,b) have developed a mathematical approach

that resolves both depth and subpixel spectral composition which is suitable for hyper-

spectral imagery. Another method of spectral classification uses a ‘‘lookup table’’ (LUT)

approach where a library of end-member remote sensing reflectances was generated from

radiative transfer computations (Louchard et al., 2003). An initial use of this approach was

applied to shallow waters (<5 m) in the Bahamas with 60–80% accuracy in classification

and 83% accuracy in bathymetry (Lochard et al., 2003). Lastly, multiple spectral

signatures in a single pixel can be resolved using a blend of different algorithms in a

‘‘Fuzzy’’ logic classification scheme. Fuzzy logic classifications allow for multiple end-

members to be recognized in an individual pixel. This type of classification scheme should

reduce errors associated with one pixel-one end-member algorithms and better reflect the

heterogeneous nature of coral reef habitats on small scales. Fuzzy classification schemes

have been used for ocean color data (Moore et al., 2001) and on coral reefs (Andrefouet et

al., 2000; Andrefouet et al., 2003). In the future, both multispectral and hyperspectral

imagery will require the development of algorithms capable of consistently extracting

accurate bottom classifications from remote sensing reflectance (Rrs) or extracting enough

optical information from the original image to effectively ‘‘remove the water’’ to be able to

use reflectance libraries of different end-members.

9. Future exploration and experimentation

Despite the voluminous amount of data on the biology and ecology of shallow water

reef systems, there are significant gaps in our knowledge about coral reef communities

beyond the 30-m depth of most studies. Deep fore reef communities have largely escaped

the effects of global climate change as it relates to ultraviolet radiation and global

warming. Because of the close proximity of near-shore fringing reefs to human popula-

tions and their relatively shallow depths (5 to 30 m), these reefs are most susceptible to

harmful human activities (sedimentation, nutrient enrichment, physical damage, over-

fishing) but also to the effects of natural disturbances (storm wave damage, high sea

surface temperatures, high irradiance). The mid-shelf coral reefs, either fringing reefs

associated with the offshore cays or nonemergent linear reefs, are less susceptible to

human induced stresses as described above, but are exposed to similar natural impacts due


to their comparable depths (5 to 30 m). The deep reefs (>30 m) are largely free from

human-induced stresses (excluding fishing and anchoring) and natural impacts due to their

greater distance from human populations and their greater depth, respectively. There is a

major gap in our overall understanding of coral reef communities and our lack of

understanding about deep reef environments, and in particular sponge and coral popula-

tions, has been keeping us from realizing an untapped resource of bioactive compounds,

understanding the ecology and biodiversity of deep reef communities, and assessing

whether deep reef communities are a potential source of larvae for shallow reef

communities. Most studies on coral reefs have focused on shallow reef (<30 m) systems

because of the technical limitations of conducting studies deeper than 30 m and the

expense of using submersible technology in relatively shallow depths (<150 m). The

technical limitations are slowly being overcome by the introduction of mixed gas technical

diving, both open-circuit SCUBA and closed-circuit rebreathers, to the scientific diving

community. A better understanding of deep reef ecology and biodiversity is critical if we

are to establish criteria that characterize the ‘‘health’’ of coral reefs and formulate

management plans in response to anthropogenic and natural disturbances.

Additionally, many cold, deep coral reefs composed of azooxanthellate taxa have

recently been discovered in deep continental shelf habitats. These communities appear to

be extremely fragile and may provide important three-dimensional habitat for juvenile fish

of commercial importance. In several locations these reefs have been severely damaged by

fish trawling activities, prompting fishing closures and proposals for marine protected

areas which include these reefs. These two unique communities, deep fore reef and deep

and cold reefs, will require experimental studies to understand their trophic and

reproductive biology, two critical life history traits which will help guide managers as

they attempt to formulate strategies for protecting these unique environments.

10. Conclusions

In closing, I want to restate that this review is not exhaustive and reflects the authors

bias of what represents the latest developments in the area of experimental biology of coral

reef ecosystems. It is essential for the coral reef community to work together for the

common good of the ecosystem. Time is short, as are finances, to conduct the integrative

studies required to understand the range of acclimative capabilities that the holobiont has

in the face of continued environmental change, and to potentially predict what reefs will

remain, and what will they look like in the future.

Acknowledgements

The author wishes to thank the editors of JEMBE for the invitation to write this review.

Support for the author’s work has been provided by NSF (Biological Oceanography),

ONR (Environmental Optics), NOAA (National Undersea Research Program and Ocean

Exploration), UNESCO, and the Smithsonian Institution (Caribbean Coral Reef

Ecosystems). Photographs or figures were generously provided by Charles Mazel, Todd


LaJeuness, NASA, Tracey Saxby, Ray Berkelmans, and Ove Hoegh-Huldberg. A special

thank you to Bill Fitt whose comments greatly improved this manuscript. [SS]

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