THE YEAR IN ECOLOGY AND CONSERVATION BIOLOGY, 2009

Coral Reefs

Threats and Conservation in an Era of Global Change

Bernhard Riegl,a Andy Bruckner,b Steve L. Coles,c

Philip Renaud,b and Richard E. Dodgea

aNational Coral Reef Institute, Nova Southeastern University, Dania, Florida, USA

bKhaled Bin Sultan Living Oceans Foundation, Landover, Maryland, USA

c

Bishop Museum, Honolulu, Hawaii, USA

Coral reefs are iconic, threatened ecosystems that have been in existence for ∼500 million years, yet their continued ecological persistence seems doubtful at present. Anthropogenic modification of chemical and physical atmospheric dynamics that cause

coral death by bleaching and newly emergent diseases due to increased heat and irradiation, as well as decline in calcification caused by ocean acidification due to increased

CO2, are the most important large-scale threats. On more local scales, overfishing and

destructive fisheries, coastal construction, nutrient enrichment, increased runoff and

sedimentation, and the introduction of nonindigenous invasive species have caused

phase shifts away from corals. Already ∼20% of the world’s reefs are lost and ∼26%

are under imminent threat. Conservation science of coral reefs is well advanced, but its

practical application has often been lagging. Societal priorites, economic pressures, and

legal/administrative systems of many countries are more prone to destroy rather than

conserve coral-reef ecosystems. Nevertheless, many examples of successful conservation exist from the national level to community-enforced local action. When effectively

managed, protected areas have contributed to regeneration of coral reefs and stocks of

associated marine resources. Local communities often support coral-reef conservation

in order to raise income potential associated with tourism and/or improved resource

levels. Coral reefs create an annual income in S-Florida alone of over $4 billion. Thus,

no conflict between development, societal welfare, and coral-reef conservation needs to

exist. Despite growing threats, it is not too late for decisive action to protect and save

these economically and ecologically high-value ecosystems. Conservation science plays

a critical role in designing effective strategies.

Key words: coral reef; conservation; global climate change; phase shift; overfishering;

coral diseases; bleaching; ocean acidification; tourism; marine reserve

Introduction

Coral reefs have been identified as an endangered ecosystem because they are subject to multiple natural, man-made and manmediated stresses (Glynn 1996; Hughes et al.

2003; Hoegh-Guldberg et al. 2007). Most crucially, they are being considered one of the most

Address for correspondence: Professor Bernard Michael Riegl, Nova

Southeastern University—Oceanography, 8000 N. Ocean Drive, Dania,

Florida 33004. rieglb@nova.edu

sensitive ecosystems to global climate change

and are frequently likened to the proverbial canary in the coal mine. We, as the coal miners,

have reason to be a bit more concerned about

the canary’s health.

Statistics vary according to source, but estimates suggest that 20% of the world’s coral reefs

are already lost, 24% under imminent risk of

collapse, and another 26% in grave danger of

irreparable damage (Fig. 1; Wilkinson 2006).

Hardly any reef of the world is not overfished,

and few have escaped degradation of their

The Year in Ecology and Conservation Biology, 2009: Ann. N.Y. Acad. Sci. 1162: 136–186 (2009).

doi: 10.1111/j.1749-6632.2009.04493.x c 2009 New York Academy of Sciences.

136
Riegl et al.: Coral Reefs 137

Figure 1. Coral reefs around the world are threatened by a variety of natural and man-made factors. Threat is not contributed evenly, though. (A) from ReefBase online GIS (http://www.reefbase.org); (B)

ReefCheck data from Wilkinson (2006), the lower the index, the more degraded the reefs of the area. (Photo

by A. Hagan.)

biological components (Jackson et al. 2001; Sale

2008). Worldwide events, such as the 1998 El

Nino Southern Oscillation (ENSO) that caused  ̃

widespread coral bleaching and death (Baker

et al. 2008), have led to indiscriminate damage

in protected and unprotected systems (Fig. 1).

Global climate change is potentially threatening every single coral, and its associated fauna

worldwide (Hoegh-Guldberg et al. 2007). This

leads to the question whether it is even possible to conserve coral reefs. What are the odds?

Should we even bother?

Yet, despite this apparently gloomy outlook, coral reefs, similar to those we know today, have existed for approximately 215 million years (and, in another taxonomic guise,

for about 500 million years). They have survived the extinction of the dinosaurs and the

climate changes of the ice ages. This would

suggest remarkable evolutionary resilience and

would certainly suggest that there is scope

for ecological resilience as well. The ultimate

question is whether we need not worry about

the survival of coral reefs or whether the upheavals of the Anthropocene (= the present era

dominated by human activities; Crutzen 2002;

Crutzen and Steffen 2003) will turn out to be

more than these time-proven ecosystems can

sustain.

To understand actions needed to conserve

coral reefs and understand what will be lost

if stresses continue unabated, we require an
138 Annals of the New York Academy of Sciences

overview of their dynamics and the most important threats facing them. While many states

and societies have been more or less active in

coral-reef conservation, there have been mixed

results, and reef stresses keep rising (Sale 2008).

Is there more we should do? And if so, what?

On the following pages we will provide a

condensed overview of what we consider to be

the issues and we will show how they have been,

and can, be addressed. This review should thus

provide an easy entry point to the discussion

about the why, when, where, and how of coralreef conservation.

Coral Reefs in the Past: Crises and

Renewed Evolution

When decrying the “coral reef crisis,” losses

of biodiversity, and threats to the ecosystem,

we are well advised to read the pages of Earth

history in order to put what is happening today

into perspective. Crises and extinction are nothing new for coral reefs. They and analogous

sedimentary systems have a very long geological history and have persisted through all major Phanerozoic (i.e., much of the entire fossil

record, >600 Ma = million years) biotic crises

(Fig. 2). Over time, many reef crises and innumerable extinctions have occurred, but coral

reefs (in the widest sense) have persisted. Not

only have reefs survived or arisen repeatedly

after extirpation, they have been shown to be

evolutionary focal points, with more organisms

evolving within reefs and spreading to adjacent

habitats than the other way around (Kiessling

2005, 2008). However, each crisis brought major extinctions and faunal turnovers, and in

some cases it took evolution millions of years

to compensate for the damages. The question

is whether mankind has created the final crisis that will push these long-lived systems into

extinction.

In the immediate past, the corals themselves

record climate in their skeleton via variability in

geochemistry. While much of this information

is lost in older fossil corals that have been replaced by different types of calcium carbonate,

the skeletons of recent or subfossil corals provide an excellent monitoring tool for climate

variability, at least in the Holocene and Pleistocene (Eakin and Grottoli 2006) and perhaps

even further back in time. Thus, looking back

and observing patterns may indeed help us to

look forward in anticipation of what might be

forthcoming.

The earliest analogues to reefs were stromatolites, layered rocks formed by filamentous

blue-green algae (Riding 1999), arising at the

dawn of time in the Archaean Eon (∼3.5 Ga =

billion years ago; Walter et al. 1980). Stromatolites have persisted throughout the geological

record into the present day. The rise of biodiversity was the demise of stromatolite dominance.

As more grazers evolved, the algae making up

the stromatolites were consumed, suppressing

their formation (Copper 2001). Thus, we find

them primarily in extreme habitats (as today in

hypersaline western Australian lagoons, or in

tidal passes with extreme currents in the Bahamas, Fig. 3A) or during extreme times (biotic crises of all ages; Wood 1999). Although

many of today’s remaining stromatolites need

protection and could easily be destroyed, we

probably need not worry about their future—

evolutionary crises will recur, and stromatolites have demonstrated their ability to survive over billions of years. The first true reefs

were built by spongelike organisms, archaeocyaths, in the lower Cambrian (∼520 Ma).

These reefs were ecologically complex and

had zoned communities with niche separation (Zhuravlev 2001; Rowland and Shapiro

2002). They were wiped out by a mass mortality caused by a global transgression–regression

couplet (Zhuravlev 2001) and were immediately followed in the Ordovician (500–440 Ma)

by one of the most significant marine metazoan

radiations in Earth history (Sepkoski 1990).

This period saw the rise of corals with the

appearance of the Tabulata and Rugosa (also

called “Tetracorals,” since they always had a

multiple of four septa; modern “Hexacorals,”

which always have a multiple of six septa,
Riegl et al.: Coral Reefs 139

Figure 2. Reef abundances through time. (Modified from Fluegel 1997.) The curve shows the reconstructed number of reef sites in the geological record. Extinction events are plainly visible as strong downward

dips in the number of reefs recorded. The dominant framebuilding taxa, that is, those that built the reef rock,

are mentioned in the gray bar above the curves.

Figure 3. (A) Stromatolites, such as these at Lee Stocking Island in the Bahamas, were the first organisms that formed geological structures akin to reefs. (B) A Devonian coral reef in Austria. (C and D)

Pleistocene/Holocene ecological constancy. The coral Acropora palmata dominated shallow Caribbean for

the past few hundred-thousand years. (C) shows a Pleistocene A. palmata reef in Curacao, and (D) shows

recent A. palmata in Andros, Bahamas. (Part (B) courtesy of Bernhard Hubmann.)
140 Annals of the New York Academy of Sciences

arose in the Triassic, ∼240 Ma) and calcareous

sponges that dominated reefs throughout the

Paleozoic. Despite several crises (Ashgill and

Ludlow extinctions; Copper 1994), reefs flourished until the mid-Devonian (∼480 Ma), one

of the acmes of Phanerozoic reef building. Huge

barrier-reef systems extended over 2000 km in

Australia and Canada, and reefs were common

throughout the world (Copper 2002; Fig. 3B).

At about 470 Ma (Frasne/Famenne crisis),

these spectacular ecosystems were wiped out

by a complex series of sea-level rises, falls, and

oceanic anoxia. Interestingly, while we currently fear for the future of reefs due to increased atmospheric CO2 levels and anticipated effects on ocean chemistry (Guinotte and

Fabry 2008), the Frasne/Famenne extinctions

occurred during a time of the largest drop in

pCO2 in the Phanerozoic (Berner 1998, 1999),

which also shifted the oceans from a calcite to

an aragonite mode (Hardie 1996). This is in direct juxtaposition of scenarios for a near-future

reef crisis of increasing CO2 (Hoegh-Guldberg

et al. 2007). Paleozoic reefs never recovered to

their former glory. In the Mississippian, phylloid algae reefs existed and in the Permian,

reef complexes were built in North America

(e.g., the famous Permian reefs of Texas/New

Mexico)—only to be wiped out for good during

the greatest marine mass extinction on Earth at

the Permian/Triassic boundary (Newell 2001),

caused by a multitude of factors, worldwide

cooling being among them (Wood 1999).

After an ∼10 Ma interval with no reefal

record, the modern corals stormed onto

evolution’s stage in about the mid-Triassic

(∼230 Ma) and rapidly built major reef complexes. It is uncertain whether these corals

contained zooxanthellae or whether they functioned exactly like the modern ones, but some

authors suggest that this is likely (Stanley and

Swart 1995; Stanley and van de Schootbrugge

2009). The calcareous Alps are peppered with

well-developed reefal limestones of impressive

dimension built by scleractinian corals that

were closely related to today’s reef-builders.

In the Rhaetian (∼210 Ma) the reef period

collapsed during a brief ice-house (Fluegel

and Senowbari-Daryan 2001), but reefs subsequently diversified throughout the Jurassic

(∼203–135 Ma; Leinfelder et al. 2002). Many

more and different types of reef than today

(microbial, siliceous and calcareous sponges,

corals) occupied and built frameworks in more

environments. This rich reef age came to an

end due to sea-level regression, leaving reefs

high and dry. The following Cretaceous (∼135–

65 Ma) saw spectacular reef building during a

time of much higher average temperatures, far

higher atmospheric pCO2 (2–10 times today’s

levels; Berner 1994), and a more sluggish ocean

circulation than today. Counterintuitive to today’s CO2 discussion, the Cretaceous was a

time of prolific carbonate deposition and reef

building (Kiessling 2002). During this period,

corals moved more and more into the oligotrophic realm at the shelf edge—a trend that

had begun in the late Jurassic (Leinfelder et al.

2002)—with the inner shelf regions dominated

by giant reef-building bivalves called Rudists.

During the Creatceous calcite sea, which presumably made skeletal formation difficult for

scleractinia, some corals lost their skeleton altogether and became the Corallimorpharia

∼110–133 Ma (Medina et al. 2006). The great

end-Cretaceous cataclysm (∼65 Ma) caused by

a bolide impact exterminated this reef period.

The end-Cretaceous cataclysm wrought by

the famous Chicxulub bolide impact terminated the reign of the dinosaurs, rain forests,

and coral reefs. But within only a million years,

rain forests covered the Earth again (Johnson

and Ellis 2002) and coral reefs, of surviving Cretaceous species and new Cenozoic species reformed in the Danian (65–61 Ma; Perrin 2001).

The Paleocene (65–53 Ma) and the Eocene

53–33.7 Ma) were characterized by some of

the warmest temperatures ever, the Paleocene–

Eocene Temperature Maximum (PETM), during which coral reefs expanded and diversified

spectacularly. However, most of these reefs were

in shallow water, and few existed in deep water

(Perrin 2001). This is an interesting parallel to

what has been proposed as future scenarios for
Riegl et al.: Coral Reefs 141

Figure 4. Coral cover has decreased on many reefs around the globe, but rarely as spectacularly as

across the Caribbean basin. (A and B) From a meta-analysis by Gardner et al. (2003); a spectacular decline

is shown, but also, in (B) some regional recovery. Thus, there is reason for hope.

modern deepwater reefs. It is feared that higher

atmospheric pCO2 and concomitant acidification will make calcification impossible for about

70% of today’s deepwater reefs (Guinotte et al.

2006; Turley et al. 2007; Guinotte and Fabry

2008) and might lead to their demise. Maybe

the situation during the PETM is indeed a valid

analog—barring the fact that anthropogenic

CO2 input has no parallel during the PETM.

During the Oligocene (∼30 Ma), reefs expanded more, but at its end, the western

Atlantic–Caribbean area lost about half its genera (Edinger and Risk 1994). In the Micoene

(∼20 Ma), reefs started to look very much like

today, which is not surprising, since the configuration of the continents and ocean circulation were similar. The Mediterranean had

extensive reef systems, but it evaporated, ending its era of reef building. Increasing isolation of the Mediterranean basin had already

led to a decrease in species diversity until reefs

were almost entirely dominated by the genus

Porites (Braga et al. 1990; Pomar 1991; Perrin

2001). A similar situation, loss of biodiversity and a shift toward Porites dominance has

also been predicted to occur in the Caribbean

(Aronson et al. 2004), which is similarly isolated today as the Mediterranean was in the

Miocene.

After the closure of the Isthmus of Panama,

a new fauna emerged over the past ∼3 Ma

in the Caribbean via a gradual step-down of

the old Indo-Pacific fauna and a progressive

step-up of newly evolved purely Caribbean taxa

(Budd and Johnson 1999; Budd 2000). By the

Pleistocene (<1 Ma), a fauna had evolved that

was very similar to today’s and showed remarkable constancy even throughout the major upheavals of the ice ages (Fig. 3D, 3E).

Throughout much of the Caribbean, a community pattern similar to that observed until

the early 1980s (i.e., Acropora palmata dominance

in the shallows, Montastraea-group dominance

in the deeper areas) showed ecological persistence through time. Beginning in the late

1970s, in the matter of only a few decades,

the shallow Acropora zone was lost on most

reefs in the Caribbean (Hughes 1994; Gardner

et al. 2003). The near-total disappearance of

this well-established zonation pattern in such a

short time raised considerable alarm (Pandolfi

and Jackson 2006). While a temporary loss

of the A. palmata zone via storm activity is a

natural, and repetitive, occurrence (Blanchon

and Jones 1997), the wholesale loss of this

zone in many areas is a reason for concern

(Fig. 4). It is believed that newly emergent diseases are the primary culprits of the spectacularly fast dieback of Caribbean Acropora, and

beginning in the late 1990s similar phenomena affected Montastraea (Aronson et al. 1998;

Bruckner and Bruckner 2006b; Edmunds and
142 Annals of the New York Academy of Sciences

Elahi 2007). While the Indo-Pacific has not experienced similar extensive ravages at the hands

of coral diseases yet, severe outbreaks of corallivores, such as the crown-of-thorns starfish Acanthaster planci or the snail Drupella cornus, along

with unprecedented levels of bleaching-related

mortality (Baker et al. 2008) have taken their

toll. It appears that also in the Indo-Pacific,

well-established and apparently long-time stable zonation patterns might be at the tipping

point.

For over >600 Ma of reef evolution the

greatest threats to reefs were large-scale environmental perturbations. Obviously, when

ocean basins disappeared, reefs disappeared

with them, and nothing stands up against tectonic change. But climate change has always

played an important role, with probably the

greatest extinction of them all occurring at

the Permo/Triassic boundary. Among other

causes, it was due to cooling that made the

tropics disappear (other factors, like increased

volcanism that created huge epicontinental

basalt flows and deleterious effects on atmosphere and water chemistry also contributed;

Wood 1999). Almost every ice-house climate

caused a severe reef, crisis. Climate rarely got

too hot for reefs, and they persisted throughout the hot Cretaceous and PETM. However,

changes in ocean chemistry may have triggered the evolution of the corallimorpharia—

essentially naked corals without a skeleton

(Medina et al. 2006). Every dramatic change

in climate and/or ocean chemistry had some

evolutionary consequence—and herein lies the

true lesson. Since humans are manipulating

these very factors (Karl and Trenberth 2003),

we must expect major biotic upheaval. So what,

if any, are the signs indicating that we might be

steering toward a systemic modern reef crisis?

Modern Reef Crisis in the

Anthropocene: Global Threats

The Anthropocene (Crutzen and Stoermer

2002, 2003) is the present time, dominated

by the activities of Homo sapiens in all global

aspects of ecology, geomorphology, and evolution. Humans are as efficient geomorphic

change agents as riverine and glacial sediment transport (Haff 2003a,b). Evolutionary

change in the Anthropocene is largely due

to forcing extinctions (present extinction rates

may exceed by 1000–10,000 times those before

human intervention; Wilson 1988). There is

general debate concerning exactly when the

Anthropocene began (Crutzen and Steffen

2003; Ruddiman 2003), and for reefs we certainly do not know. Coral-reef science is relatively young, and therefore our documentation

of ecological trends in the past is sketchy at best.

While reefs were considered stable and well biologically accommodated ecosystems only three

decades ago (Endean 1977; Connell 1978), the

last two decades have revealed them to be very

dynamic (Mumby and Steneck 2008). In the final decade of the 20th century it appeared that

coral reefs had started to unravel ecologically

on a worldwide scale, mostly due to large-scale

changes in climate and environment brought

on by human activities (Bakeret al. 2008). While

ultimately all threats caused distally (by changing climate) or proximally (by direct, local impact) from human activity can be considered

man-made, we distinguish in the following between “global threats” by climatic or large-scale

environmental phenomena that reef management has no or little control over, and “local

threats” that refer to proximal, direct human

impacts that can be regulated and avoided.

Atmospheric Warming and Bleaching

A major and apparently very recent threat

to coral reefs, with the potential of negating

success to all conservation efforts, is bleaching and associated coral mortality (Baker et al.

2008). Dinoflagellate symbionts of the genus

Symbiodinium, referred to as “zooxanthellae,”

live within coral tissues. They exist in what

is an obligatory association for the host coral,

but not for the algae, which contribute photosynthates and aid calcification (Muscatine and
Riegl et al.: Coral Reefs 143

Figure 5. Bleaching is a discoloration of coral tissue due to loss of photosynthetic algae. (A) Partly

bleached Montastraea faveolata from Puerto Rico. (Photo by A. Bruckner.) (B) Fully bleached Acropora

cervicornis. (C) Nonphotosynthetic pigments give a bleached Siderastrea siderea an attractive pink color.

The coral has, however, lost the photosynthetic pigments needed for survival. In the background, bleached

Montastraea annularis and Acropora palmate. (B & C from Andros, Bahamas, 1998.)

Porter 1977). Stress caused by high temperature

or irradiance damages the symbionts’ photosynthetic system, leading to overproduction of

oxygen radicals that damage the symbionts and

their hosts (Goreau 1964). As a result, the symbionts can be expelled or die (Lesser 2006),

turning the coral white since the yellow-brown

pigmentation of the symbionts is lost—this phenomenon is referred to as bleaching. A variety of nonphotosynthetic pigments inside the

corals may not be diminished during bleaching

and corals can appear in a variety of attractive, mortality-masking pastel colors (Fig. 5).

Bleaching events, when they occur, are usually not confined to corals alone, but can also

affect numerous other organisms (gorgonians,

soft corals, anemones, foraminifera; Hallock

2001; Hallock et al. 2006; Rodolfo-Metalpa et al.

2006).

Since at least one of the primary culprits

of coral-reef bleaching appears to be elevated

temperature, it comes as little surprise that in a

rapidly warming world (IPCC 2007) the number of coral-reef bleaching events has risen dramatically since the early 1980s (Glynn 1993;

Hoegh-Guldberg 1999; Hughes et al. 2003;

Hoegh-Guldberg et al. 2007; Baker et al. 2008).

The frequency and scale of coral bleaching

over the past few decades have been unprecedented, with hundreds of reef areas bleaching at

some point, and occasionally even entire ocean

basins affected (Fig. 6). Bleaching is often variable and patchy over micro (mm to cm) to meso

(km) scales. This can be explained by fluctuations in environmental conditions, spatial heterogeneity of reef surfaces, genetic differences

in hosts or symbionts, and differences in environmental history. Bleaching has been reported
144 Annals of the New York Academy of Sciences

Figure 6. Documented bleaching events. ((A)from Baker et al. 2008) show that virtually

all reefs around the world have been affected. The events here are only such with noticeable

coral mortality. The frequency of bleaching events is predicted to increase (Sheppard 2003a).

from almost every coral-reef region and wherever corals occur (even if not reef building,

like the Mediterranean; Cerrano et al. 2000;

Rodolfo-Metalpa et al. 2006).

Corals and other reef organisms with zooxanthellae live very close to their upper thermal

tolerance limits, which makes them susceptible

to heat (∼1.0 to 1.5◦C above seasonal maximum mean temperatures). Interactions between temperature and light damage Photosystem II (Iglesias-Prieto et al. 1992; Fitt and

Warner 1995; Lesser 1996; Warner et al. 1996,

1999; Jones et al. 1999; Brown et al. 2000). At

high temperatures and light, the lipid composition of thylakoid membranes in the symbiont

changes and degrades (Tchernov et al. 2004).

Also increased nitric acid synthase accompanies bleaching (Trapido-Rosenthal et al. 2005).

In general, bleaching results from accumulated

oxidative stress on the thylakoid membranes

of symbiont chloroplasts (Lesser 1996, 1997;

Downs et al. 2002) as a result of damage to Photosystem II (see Lesser 2006 for review), which

causes degradation and expulsion of the symbionts from host tissue. Protective mechanisms

involve enzymatic antioxidants that degrade reactive oxygen species (Lesser et al. 1990), and

also the xanthophyll cycle can dissipate excess

absorbed energy (Brown et al. 1999). While

other stressors, like low temperatures (Coles

and Jokiel 1977; Glynn and D’Croz 1990; Coles

and Fadlallah 1991; Hoegh-Guldberg et al.

2005), can also cause bleaching, light/heat interactions cause the majority of events on tropical reefs.

Coral bleaching is patchy both on the scale

of reefs and individual corals. This is a result

of interaction between environmental stressors

and the patchy distribution and/or zonation of

different Symbiodinium within and among coral

species (Rowan and Knowlton 1995; Rowan

et al. 1997). Within the coral, different types of

zooxanthellae are found. Since these can respond differently to environmental stressors,

the distribution of symbiont diversity within

and among coral colonies and species can influence patterns of bleaching, and the proportion of the symbiont clades may change following a bleaching event. Symbiodinium in clade

D (particularly D1a) are resistant to elevated

temperature conditions (Rowan 2004) and can

remain much longer in coral-host tissues than

other clades (Baker 2001; Glynn et al. 2001;

Baker et al. 2004; Berkelmans and van Oppen 2006; Jones 2008). Thus, the heat resistance of corals may indeed be linked to the

type of zooxanthellae they harbor. Buddemeier

and Fautin (1993) suggested in their “adaptive bleaching hypothesis” that changes in algal symbiont communities following bleaching

might be a mechanism allowing coral adaptation to environmental change—a point still
Riegl et al.: Coral Reefs 145

very much in debate. Goulet and Coffroth

(2003) and Iglesias-Prieto et al. (2004) found no

change in symbionts after bleaching or transplantation, while Baker (2001) recorded shifts

in symbiont communities in several species of

Caribbean coral following bleaching due to

irradiance stress and showed that corals that

changed their symbiont communities experienced less mortality. Glynn et al. (2001) showed

that corals containing clade D did not bleach,

while those with clade C bleached severely.

Baker et al. (2004) and Berkelmans and van

Oppen (2006) observed increases in clade D after bleaching or after transplantation to hotter

sites. Clade D was found more commonly on

reefs recently affected by bleaching (e.g., Kenya)

and on reefs routinely exposed to high temperatures (e.g., Arabian Gulf), but rarely on reefs not

exposed to high temperatures (e.g., Red Sea),

or without a history of recent severe bleaching

(e.g., Mauritius). Also, Jones (2008) showed that

71% of colonies changed their symbiont communities to more heat-tolerant types following bleaching, with many corals shuffling preexisting symbiont communities at the colony

level. All this would suggest that some natural protection mechanisms to bleaching indeed

exist.

Bleaching events are predicted to recur

more rapidly due to global warming (Sheppard

2003a). Bleaching is episodic, with the most

severe events typically accompanying coupled

ocean–atmosphere phenomena, such as the

ENSO, which result in sustained regional elevations of ocean temperature (Glynn 1993,

1996). Bleaching episodes have resulted in

catastrophic loss of coral cover in some locations and have changed coral community

structure in many others, with a potentially

critical influence on the maintenance of biodiversity in the marine tropics (Fig. 7). This has

led many to develop models of coral-reef dynamics in future accelerated bleaching dynamics, none of which are particularly optimistic

(Done 1999, Hoegh-Guldberg 1999; Sheppard

2003a; McClanahan et al. 2007b). Bleaching

has also facilitated or initiated increases in coral

diseases, the breakdown of reef framework by

bioeroders, and the loss of critical habitat for

associated reef fish and other biota (Jones et al.

2004; Pratchett et al. 2008). Secondary ecological effects, such as the concentration of predators on remnant surviving coral populations,

have also accelerated the pace of decline in

some areas.

Baker et al. (2008) studied the regeneration

of coral reefs after bleaching events in a metaanalysis of published data and found variable

rates of recovery among sites. In some cases

it was high enough to be detected within only

2 years (Maldives), while no recovery was observed in other locations, even over 20 years

(Galapagos). The rate of recovery did not appear to be related to the severity of the bleaching disturbance, and the degree of recovery

was not related to the amount of coral cover

remaining after the disturbance. Many reefs

with high coral cover also continued to decline after a bleaching event (Cook Islands, U.S.

Virgin Islands). Other reefs with low cover

regenerated rapidly (Arabian Gulf recovered

from 0% to 42% in 9 years; American Samoa

recovered from 6% to 40% in 4 years). Numerical experiments (Fong and Glynn 2000;

Riegl and Purkis 2009) show that even with

repeated and severe bleaching mortality, at

least limited recovery is possible given enough

asexual regeneration or connected populations. However, changes in community structure must be expected at high bleaching recurrence. In particular, Acropora dominance may

be compromised—model predictions and empirical observations (McClanahan et al. 2007b)

seem to conform.

The species documented by Baker et al.

(2008) with most potential for successful regeneration were mostly broadcast spawners

(Harrison and Wallace 1990). This may be

due to a different life-history strategy, with larvae spending more time in the water column

than those of brooders and dispersing further

from the parent, thus reducing the extinction

debt (see text under heading “Potential for

Extinction”). While recruitment is important,
146 Annals of the New York Academy of Sciences

Figure 7. Conceptual model outlining the possible responses of coral communities to bleaching, in

particular when events recur with increasing frequency, as predicted by global change scenarios (Coles and

Brown 2003.)

the maintenance of reef framework is key for

the conservation of biodiversity associated with

corals. Loch et al. (2004) observed that collapsing Acropora tables, victims of the preceding bleaching, effectively negated the otherwise

high recruitment success in the Maldives. The

secondary effects of bioerosion continued to degrade potential settlement substrates, an observation also made by Sheppard et al. (2002) in

the Chagos Archipelago. Thus, not only settlement substratum and coral recruits are lost,

but also niche space for much associated fauna

(Pratchett et al. 2009).

Clearly, coral bleaching, largely caused

by global warming, is a major challenge for

the conservation of coral reefs. It is unclear

whether bleaching can be managed, but

emphasis is put on attempting to minimize

additional stressors, since bleaching is known

to facilitate the outbreak of diseases and to

weaken corals (Marshall and Schuttenberg

2006a, 2006b; Bruno et al. 2007). Most

strategies to manage bleaching by restoring

or maintaining ecosystem resilience search to

identify areas less prone to bleaching, thus

allowing conservation efforts to have the greatest opportunity for success. Proposed actions

are to (1) identify local physical or environmental conditions that naturally protect reefs

from bleaching, and (2) use climate models

to identify coral-reef areas or regions most

likely to escape the worst effects of warming

(Baker et al. 2008). Coupled ocean–atmosphere

climate models can help to forecast bleaching stress on reefs (Hoegh-Guldberg 1999;

Donner et al. 2005), but other approaches to

estimate bleaching susceptibility are also used

(McClanahan et al. 2007a; Kleypas et al. 2008),

including, for example, attempts to use the relative abundance of heat-tolerant Symbiodinium

in corals to help identify relatively bleachingresistant reefs. Other, more hands-on
Riegl et al.: Coral Reefs 147

suggestions include shading corals, sprinkling

reef surfaces with water to increase evaporative

cooling and reflection of UV (Baker et al.

2008), or even feeding corals since increased

heterotrophy seems to benefit bleached corals

(Grottoli et al. 2006).

However, the most efficient possible management action would be a more responsible

use of fossil fuels to slow the global greenhouse

effect.

Seawater Chemistry and Reef Building

A major concern in all scenarios regarding the future of coral reefs are changes in

seawater chemistry, most particularly acidification of ocean waters concomitant with rises

in atmospheric CO2 concentration. During

Earth history, seawater chemistry has changed

repeatedly and dramatically, and with it the

ability of marine skeletal organisms to calcify.

Hardie (1996) showed these large-scale, secular changes (Fig. 8) to be most strongly influenced by rates of seafloor spreading that

change the Mg/Ca ratio in seawater, which

strongly affects the type of calcium carbonate that can be precipitated by marine skeletal organisms (primarily aragonite at ratios >2,

and calcite at ratios <1; Stanley 2006). Conflicting information exists about evolutionary

consequences of changes in seawater chemistry for reef-builders. Some observed important consequences (Stanley and Hardie 1998)

and claim that major extinctions were related

to changes in ocean chemistry (Veron 2008a,b),

while others find no correlation (Kiessling et al.

2008). In the modern ocean, the most important cause of seawater chemistry change is the

anthropogenic enrichment of the atmosphere

with CO2, nitrogen, and sulfur (Doney et al.

2007). While mean atmospheric CO2 ranged

in the past 650.000 years between 200 and 300

parts per million (ppm), it now reached about

387 ppm and could reach 540–970 ppm by

2100—an increase about 100 times faster than

over the past 650 ky (Fabricius 2008). Only half

of the CO2 that reaches the atmosphere remains there—∼20% is absorbed by land and

∼30% by the sea—resulting in acidification

and shifts in the production of bi/carbonate

ions (Fig. 9). Increased availability of CO2 lowers ocean pH and shifts the balance toward

more formation of bicarbonate versus carbonate ions, making the precipitation of calcium

carbonate more difficult (Kleypas and Langdon

2006; Kleypas 2007). The topic has received

much recent review elsewhere (Stanley 2006;

Guinotte and Fabry 2008), and we will be brief

here. Present atmospheric CO2 concentrations

are the highest in the past 650,000 years and

are expected to increase rapidly (Guinotte and

Fabry 2008). This has the effect of reducing

the present aragonite supersaturation (already

down from 4.6 to 4.0 since preindustrial times;

Kleypaset al. 1999; Hoegh-Guldberg et al. 2007)

and thereby with the capability of many organisms to calcify efficiently (Kleypas 2007; Jokiel

et al. 2008). The present decline in aragonite supersaturation and the ease of its precipitation

in skeletons is being changed predominantly

by pH in the modern ocean, via uptake of

atmospheric CO2 in contrast to shifts in the

Mg/Ca balance in the fossil ocean. This illustrates the increasing disequilibrium in the

ocean that is being created by atmospheric

changes (Kleypas 2007). Unfortunately, among

the organisms of concern are the major reefbuilders and many of the important calcifiers

in the plankton. Furthermore, increased levels

of dissolved CO2 seem to lead to synergisms

with other damaging chemical factors such as

nutrient enrichment, leading to increased coral

mortality (Renegar and Riegl 2005). Some organisms can maintain, even increase, calcification in seawater; however, this can come at

a cost of reduced muscle mass (Wood et al.

2008).

Corals do not necessarily die when subjected to conditions unfavorable to the production of aragonite skeletons. When subjected to

seawater equivalent of calcite-ocean acidity

(and Mg/Ca ratio; i.e., conditions simulating a Cretaceous ocean), coral growth decreased and they produced a calcite, instead
148 Annals of the New York Academy of Sciences

Figure 8. Aragonite and calcite seas with their faunas throughout earth history. (A) Temporal variability

through geological time of the Mg/Ca molar ratio in seawater and the nucleation field for aragonite,

high-Mg calcite, and low-Mg calcite. (B) Skeletal mineralogy of some dominant reef-builders through time.

Aragonitic corals persist into the Cretaceous calcite sea, as do rudists, which have a part-aragonite skeleton.

(From Stanley 2006.) (C) Proportional abundance of stony corals in macrofaunal assemblages (proportions

are counts of genera with aragonite skeletons over the sum of aragonite and low-Mg calcite occurrences).

This shows that coral was calcitic in the Paleozoic, but aragonitic in the Meso- and Cenozoic. The major

evolutionary events do not coincide with changes in seawater chemistry, but with mass-extinction events

(vertical dotted lines). (From Kiessling et al. 2008.)

of an aragonite, skeleton (Ries et al. 2006) or

stopped making a skeleton altogether (Fine and

Tchernov 2007). Skeletal growth resumed

when acidity was again in an acceptable range.

It is assumed that double preindustrial atmospheric CO2 concentrations will lead to a

20–60% reduction in calcification (Langdon

et al. 2006; Kleypas and Langdon 2007), which

would then presumably lead to vastly increased

bioerosion rates (Guinotte and Fabry 2008).

Calcification by coralline red algae was found

to decrease dramatically in acidification experiments (Jokiel et al. 2008). In short, it appears

as if the most important modern reef-builders

will precipitate much less skeleton in more

acidic oceans. From the geologic record, it has

been demonstrated that reef building apparently progressed slower in calcite seas (Stanley

2006); however reef-builder diversity was not

affected by large-scale changes in ocean chemistry (Kiessling et al. 2008). In today’s ocean

and under the currently observed extremely

rapid rate of acidification, not only shallowwater coral reefs are at risk. Since the aragonite

saturation horizon is moving toward the surface

at a rate of 1–2 m per year (Guinotte and Fabry

2008), it is believed that deepwater coral reefs,

probably the most biodiverse system at greater

ocean depths, may be sooner at risk than their

brethren in the photic zone (Turley et al. 2007).
Riegl et al.: Coral Reefs 149

Figure 9. The carbonate system in seawater. (From Kleypas 2007) CO2 entering the

ocean forms (weak) carbonic acid, which rapidly dissociates to bicarbonate (HCO3

−) or carbonate (CO3

2−). Total CO2 is the sum of HCO3−, CO32−, and dissolved CO2 in seawater.

Total alkalinity is the excess positive charge in seawater that can be changed by shifts in the

availability of Mg2+ and Ca2+. The proportion of HCO3

− and CO32− adjusts to balance

this positive charge. Throughout Earth history, the major secular changes in seawater chemistry were primarily via changes in the Ca2+/Mg2+ ratio, which forced adjustment of the

bicarbonate/carbonate component. In the modern ocean, unusually strong CO2 input forces

the bicarbonate/carbonate system by favoring bicarbonate over carbonate, which leads to

decreases in calcification.

Projections from Orr et al. (2005) suggest that

by the end of the century, 70% of deepwater

reefs could be situated in a zone of aragonite

undersaturation, while today more than 95%

are situated in the supersaturated zone. This

trend is disconcerting and might severely compromise the functioning of these ecosystems.

Clearly, acidification is as big a challenge for

the conservation of coral reefs as any other factor. It appears that the only management action

possible is a more responsible use of fossil fuels

since we cannot otherwise regulate how much

CO2 enters the oceans. While the prospects

of ocean acidification are certainly a potential

threat to the future viability of corals, much remains to be learned about its exact mechanisms

with regards to coral health and long-term effects on reef building (Kleypas and Langdon

2006; Kleypas 2007). Nonetheless, concern is

big enough that some claim acidification could

cause a mass extinction of coral reefs (Veron

2008b).

Diseases Old and New

Coral diseases are one of the biggest threats

to the conservation of coral reefs and are predicted to become ever more prevalent with

global change (Harvell et al. 2007). Diseases

have primarily ravaged Caribbean reefs and

led to an interruption of long-established ecological patterns yet they are beginning to invade the Indo-Pacific as well (Fig. 10). Occurring globally in most coral-reef habitats

whether near human population centers or remotely offshore, their prevalence is generally

low. Localized outbreaks have produced significant mortalities of scleractinian corals, gorgonians, sea urchins, reef fish, sponges, algae,

and other coral reef organisms (Peters 1993;
150 Annals of the New York Academy of Sciences

Figure 10. Coral diseases are apparently an emergent problem, since their frequency

is seemingly increasing. (A) Black-band disease on the Caribbean Siderastrea siderea, is

the oldest known coral disease. (B) White-band disease on Acropora palmata has ravaged

Acropora populations throughout the Caribbean. (C) White plague on Caribbean Agaricia

lamarcki. (D) Several yellow-band-disease infections on Caribbean Montastraea faveolata. (E)

White syndrome on Indo-Pacific Diploastrea heliopora. (F) Corallivorous gastropod Drupella

cornus destroying an Indian Ocean Acropora sp. Outbreaks of these voracious predators

have reached plague proportions in Australia and the Red Sea. (All photos by A. Bruckner.)

Harvell et al. 1999; Williams and BunkleyWilliams 2000). In corals, additional to tissueloss, diseases can cause significant changes in

reproduction, growth, species diversity, abundance, and size structure.

Between 1972 and 2005 coral diseases were

reported on 39 coral genera and 148 species

worldwide, with observations in 63 countries.

Although Pacific reefs have a higher diversity

of reef-building corals than the Atlantic and

harbor 92% of the world’s coral reefs (Spalding

and Greenfell 1997), only 14% of the global observations of coral disease were from the IndoPacific (Green and Bruckner 2000; Sutherland

et al. 2004). The first Caribbean epizootics were

reported in 1978 and led to die-offs of two

dominant structure-forming corals, A. palmata

and A. cervicornis. Between 1982 and 1984, a
Riegl et al.: Coral Reefs 151

disease of unknown cause decimated populations of a keystone species, the herbivorous sea

urchin Diadema antillarum. This mortality triggered increases in fleshy macroalgae and concurrent losses of coral cover, biodiversity, and

habitat in many locations (Lessios et al. 1984;

Hughes 1994; Aronson and Precht 2001). Since

1998 there has been an emergence of new

and virulent diseases in the wider Caribbean,

where over 30 named diseases affect 45 species

of scleractinian corals, 3 hydrozoan corals,

10 octocorals, 2 zoanthids, 9 sponges, and

2 crustose coralline algae (Green and Bruckner 2000; Weil et al. 2006; Fig. 10). Thus the

Caribbean has been referred to as a “hotspot”

for disease, largely because of the rapid emergence, high prevalence, wide distribution, large

number of host species, and virulence of diseases in this region—and, not least, the large

body of research. Much less is known about

the Indo-Pacific, but surveys there revealed

new diseases, suggesting either a rapid emergence of diseases or a new realization of their

presence.

While first described in the western Atlantic

over 35 years ago (Antonius 1973), the realization of diseases as a significant factor accelerating the deterioration of coral reefs is relatively recent (Epstein et al. 1998; Harvell et al.

1999; Green and Bruckner 2000; Sutherland

et al. 2004; Weil 2004). Over half of all published coral disease records up to the year 2005

were for black-band disease (BBD, Fig. 10A),

white-band disease (WBD, Fig. 10B), and white

plague (WP, Fig. 10C) (Antonius 1973; Dustan

1977; Gladfelter 1982). Since the mid-1990s

new diseases have been reported with increasing frequency, including four [dark-spot disease (DSD), yellow-band disease (YBD), white

pox disease (WPX), and aspergillosis (ASP,

Table 1] that are widely distributed and four

(YBD, WPX, WP-II, and ASP) that are causing substantial coral mortality (Porter et al.

2001; Kim and Harvell 2002; Williams and

Miller 2005; Bruckner and Bruckner 2006a,

2006b). Close to 80% of all western Atlantic

coral species are now affected (45 scleractinian corals, 8 gorgonians, 2 hydrozoan corals),

with some corals (especially Montastraea annularis complex) susceptible to as many as eight

diseases and individual colonies showing signs

of two to three diseases at any one time (Weil

2004; Bruckner and Bruckner 2006a, 2006b;

Fig. 10D). Indo-Pacific corals had been known

to have tumors (growth anomalies, Squires

1965), and diseases similar to those in the

Caribbean (BBD and WBD; Antonius 1985,

1987, 1988), but recently many new diseases

were identified [Table 2; white syndrome, yellow band, skeletal eroding band (SEB), and

Porites ulcerative white spots; Littler and Littler 1996; Korrubel and Riegl 1998; Riegl and

Antonius 2003]. Currently, Acropora, Pocillopora,

and Porites appear to be most susceptible.

The recent emergence of diseases in the

wider Caribbean may be unprecedented over a

millennial time scale (i.e., >3800 yr) (Aronson

and Precht 2001). Unfortunately, characterization of the cause, prevalence, and consequences

of most disease outbreaks is limited or nonexistent. While hundreds of studies have been

published, the causative agent has been confirmed for only five diseases (Raymundo et al.

2008). Other infections may be caused by opportunistic, nonspecific pathogens that exploit

compromised health state of corals when exposed to environmental stressors (Lesser 2006).

Thermal anomalies and bleaching events seem

to be followed by outbreaks of disease (Harvell

et al. 2001; Bruno et al. 2007; Miller et al. 2006).

Changing environmental conditions could affect corals and lower their ability to fight infection and increase the virulence of potential

pathogens (Rosenberg and Ben-Haim 2002).

Pollution, nutrient loading, sedimentation, and

any other anthropogenic stressors could further

reduce coral health, alter the composition and

virulence of the microbial community found in

the surface mucopolysaccharide layer of corals,

and reduce their resistance to pathogenic organisms (Ritchie 2006).

Obviously, diseases pose a major challenge

for the conservation of coral reefs. Traditional

management tools for human and wildlife
152 Annals of the New York Academy of Sciences TABLE 1. Major Diseases Observed in Western Atlantic Scleractinian Corals and Gorgonians Syndrome Synonym Host range Prevalence/impact Rates of tissue loss White-band disease (WBD) White-line disease; white death; white plague, WBD type II Acropora palmata, A. cervicornis, A. prolifera 1–80% or more; most common in dense populations; up to 98% mortality over two decades 5 mm/day in 1970s; up to 2 cm/day today Plague Type I (WPX) White plague 13 species 1–4% average; up to 73% of individual species; maximum of 20–30% mortality 3.1 mm/day White-plague type II (WP II) Plague type III, white-band disease, white-line disease 40 species of plating and massive corals

<1–58%; losses of up to 38% of corals

from a single outbreak

2–10 cm/day

White-patch disease

(WPD)

White pox; patchy necrosis A.. palmata

<1–100%; caused 87% loss of A.

palmata in Florida from 1996 to 2002

2.5 cm2/day

Black-band disease

(BBD)

Black-line disease 24 scleractinian corals, 1 hydrozoan,

6 gorgonians

Mean

<0.2–3%; up to 50% in isolated

areas; increase during summer;typically causes partial mortality

3 mm–1 cm/day; various,

depending on species,depth, and season;

Yellow band disease

(YBD)

Yellow-blotch disease; ring

bleaching1, yellow-pox

disease2; yellow-band

syndrome3

Montastraea annularis complex, M.

cavernosa; possibly other faviidsand; A. agaricites

1–86%; infections may last 5–10 years

until coral dies; caused losses of40–80% of M. annularis over 10 yearsin Puerto Rico

Mean

= 0.5 cm/month in

winter and 1 cm/monthin summer; max

=

3 cm/month

Dark-spots disease

(DSD)

Dark-spot disease, dark-spot

syndrome; ring disease, DSDtype II; purple-bandsyndrome

11 species; mostly M. annularis

(complex), Siderastrea spp.;Stephanocoenia intersepta, and Agaricia

agaricites

Up to 94%; increases during summer. 0.5–1.33 cm2/month

Caribbean ciliate

infection (CCI)

Skeletal eroding band (SEB) 26 species, especially Dichocoenia,

Montastraea, Acropora

Up to 25%, increases in summer 1.2 cm/month in winter;

2.6 cm/month insummer

Growth anomaly Hyperplasia, calicoblastic

epithelioma, tumors.Gigantism, area ofaccelerated growth, chaoticpolyp development

Acropora, Diploria, Colpophyllia,

Montastraea, Agaricia, Porites,Dichocoenia, Madracis

Generally rare; can affect all corals in

restricted area

<1 cm/month

Aspergillosis (ASP) Sea-fan disease 7 species, 5 genera of gorgonians 8–60%; 31% mean in Florida in 1997,

declined to 6% by 2003:

>50% loss

of sea-fan tissue area over 6 years,due to loss of largest sea fans

Not reported
Riegl et al.: Coral Reefs 153

TABLE 2. Major Diseases of Indo-Pacific Scleractinian Corals

Syndrome Location Host species Prevalence ImpactBlack-band disease Australia, Egypt, Fiji, India, Jordan,

Papua New Guinea, Philippines,Saudi Arabia, Tonga, South Africa,CNMI, Palau

19 genera, 49 species; Pocillopora,

Acropora, and faviids mostfrequently affected

Generally <1%; outbreaks of up to

50% in restricted areas

Average spread of

3 mm/day; mostly causespartial mortality

White syndrome Egypt, Australia, Solitary Islands 38 species 1 Turbinaria, Acropora,

Goniastrea, Pocillopora, Stylophora,and Porites

Increased 20-fold in Australia

between 1998 and 2003

Up to 2–4 cm/day with

3–20% coral mortalityduring outbreaks

Ulcerative white-spot

disease

Philippines, East Africa, Indonesia,

Guam, Palau,

Porites: 7 species; Echinopora,

Goniastrea, Heliopora, Favia,Montipora

Up to 82% of susceptible corals;

increases in summer

Slow spread, lesions 2–5 mm

in diameter

Vibrionic bleaching Mediterranean, Israel, Tanzania Oculina pategonica; Pocillopora All corals may have Vibrios that

become pathogenic at hightemperatures

Can cause bleaching and

tissue sloughing iftemperatures elevated forlong periods

Atramentous necrosis Florence Bay and Bright Point,

Australia

Montipora Acropora Echinopora,

Turbinaria, Merulina

Unknown Unknown

Growth anomalies CNMI, Australia, Hawaii, Guam,

Palau, Enewetak, French Polynesia,New Caledonia, Maldives,Micronesia, Marshall Islands, Japan,Oman, China, Philippines

Acropora, Pocillopora, Pavona, Fungia,

Madrepora, Montipora, Platygyra

Generally <1%, but lesions increase

in number on affected corals

Unknown

Pink-line syndrome Papua New Guinea, Sri Lanka,

Kavaratti Island, Indian Ocean

Porites Increased from 20% in 1996 to 80%

in 1999 in one location

Unknown

Fungal syndrome East African Coast Astreopora, Montipora, Echinopora,

Acropora, Goniopora, Platygyra,massive Porites, Pocillopra,Goniastrea Hydnophora, Cyphastrea

Unknown; single event in 2002

caused extensive colony mortality

Unknown

Plagiosporus trematode

infection

Hawaii Porites compresssa, P. lobata Not reported to kill corals Does not advance; colony

may have multipleinfections

Skeleton eroding band

(SEB)

Australia 3 Egypt, Jordan 1 PNG,

Mauritius 2, Australia 2

21 genera scleractinian corals, 1

hydrozoan

In Australia increased 20× in

summer

1 mm/week to 1 mm/day

Yellow-band disease United Arab Emirates; Arabian Gulf;

Iran2

4 genera, 12 species Unknown? 9.4 mm/week in winter;

19.7 mm/week in summer

Brown-band disease Australia Acropora, Pocillopora, Echinopora, 16

spp.

Generally rare; 0.3% in Australia Unknown
154 Annals of the New York Academy of Sciences

diseases such as quarantine, culling, and vaccinating are not viable options. The oceans,

as well as the corals themselves, support vast

and diverse microbial communities, some of

which may be pathogenic or potentially so,

and we are just beginning to understand the

complex relationships between the coral and

its associated microflora. The coral holobiont

is a complex consortium involving beneficial

surface bacteria as well as potential pathogens,

and indiscriminate use of antibiotics may harm

this beneficial microbial community. Various

treatments, like removing infectious agents by

aspirator, covering the area with underwater

putty that forms an impermeable barrier preventing spread of the disease, or even phage

therapy (Efrony et al. 2007) have been tried. But

these methods are costly and time-consuming

and likely viable only to save certain high-value

colonies, such as particularly large, long-lived

corals (Raymundo et al. 2008).

Since disease outbreaks may be facilitated by

high nutrient levels and temperatures (Bruno

et al. 2003; Lesser 2004; Marshall and Schuttenberg 2006a), remedial actions that reduce

land-based pollution and other direct human

impacts may improve the coral’s chance of

resisting or recovering from infections. Some

mortality by diseases may have been facilitated

by bleaching events that weakened the corals

(Lesser 1997; Miller et al. 2006). Therefore, any

reduction of stresses would be to the coral’s

advantage. Improved watershed management

and better waste-water treatments have the potential to reduce nutrient levels, thereby reducing macroalgae, which stress a coral through

direct contact and also may provide refuge for

certain pathogens (Nugues et al. 2004). And

again, a more rational and responsible way

of using fossil fuels to slow temperature rise

would be at the core of successful management, since it might alleviate global warming, which not only stresses corals with heat

but also increases moisture content in the atmosphere, leading to more precipitation that

can deliver nutrients and pathogens to the

reefs.

Predator Outbreaks

Spectacular depredation has been visited

onto coral reefs by outbreaks of organisms

that prey on living coral. In the Indo-Pacific,

the most notable of these are the crownof-thorns starfish (COTS) Acanthaster planci and

the coral-eating snails Drupella cornus (Turner

1992; Schuhmacher 1992; Fig. 10F). In the

Caribbean, the primary predator is the coraleating snailCoralliophila abbreviata (Bruckneret al.

1997; Baums et al. 2003), but other corallivores, such as the fireworm (Hermodice carunculata), stoplight parrotfish (Sparisoma viride), and

three-spot damselfish (Stegastes planifrons), are

of growing concern, especially as coral populations are reduced and corallivores experience reduced predation pressure due to overfishing of their predators. Interestingly, while

Caribbean reefs have been racked by diseases,

predator outbreaks have been relatively mild,

with the opposite situation in the Indo-Pacific.

The best-studied and most devastating organism is the COTS, the outbreaks of which have

been widely studied in the literature. It is now

believed that a certain level of infestation will

always plague coral reefs, and it is assumed that

the clearly pulsed population outbreaks may be

natural. However, it is also possible that anthropogenic nutrification of coastal waters favors

the phytoplankton on which both starfish and

corallivorous snail larvae feed, thus enabling

a higher survival rate that then leads to outbreaks of adult populations (Birkeland 1982;

Ayukai et al. 1997). It has also been surmised

that removal of other predators, like the triton shell Charonia tritonis, a voracious predator

of starfish, allows coral predator populations to

go unchecked.

Bradbury et al. (1985a,b) investigated COTS

outbreaks on the Great Barrier Reef (GBR)

from a theoretical ecological viewpoint with

qualitative models harnessing graph theory

and were able to make a case for the existence of metastable points, cycles, and chaos

(Fig. 11). The observed dynamics conformed to

the ecological theory (May 1973, 1975, 1976)
Riegl et al.: Coral Reefs 155

Figure 11. Phase-space approximation of predator–prey dynamics as envisaged by

Bradbury et al. (1985a,b) for the crown-of-thorns (COTS)/coral system. (A) Shows the general phase-space representation of a stable equilibrium point. Wherever started on the phase

lattice, the system will converge toward the stable point. Such a pattern would be exhibited

by regular Lotka–Volterra type dynamics. (B) Metastable states observed in the Great Barrier

Reef (GBR) coral system. The loop returning to the starting point indicates that the system “rests

in itself” and will not transit into another state (in appreciable time). Some reefs seem to persist

indefinitely with high coral/low COTS or medium coral/medium COTS. (C) Expression of a

stable cycle, also as observed in the GBR. Some reefs go through cycles from high coral/high

COTS (outbreak) to low coral/high COTS, to regeneration of high coral/low COTS. (D) Some

reefs express chaotic “flip-flops” between high/low COTS and/or high/low coral.

of endogenously driven cyclicity. According to

Bradbury et al. (1985a,b), the most likely asymmetries driving qualitatively stable cycles would

be asymmetrical predator preferences or lags in

the responses of predator and prey to each other

due to the disparate nature of reefs in space. Recruitment, reproduction, and growth, would be

other logical factors (Birkeland 1982; Ayukai

et al. 1997). The distribution of these asymmetries over space would in itself be an interesting

indicator of how uniform large reef-complexes

are in their internal dynamics, that is, whether

these reef systems consist of a multiplicity

of separate bounded ecosystems or whether

general models can hold over the entire

system.

Whatever the causes, coral predator outbreaks have been intensely managed. From

early on, volunteers and paid workers have

responded to COTS outbreaks by gathering

up the starfish and bringing them onshore

for burial, or giving them lethal injections

(Yamaguchi 1986). Similarly, divers have collected corallivorous snails in an effort to reduce

their populations (Turner 1992). While some

outbreaks could be controlled, others were simply too big to allow the divers’ efforts to make

any dents into the vast numbers of predators.
156 Annals of the New York Academy of Sciences

Again it appears that a more responsible management of our watersheds might be the best

insurance for life in the sea. If nearshore nutrification is indeed a root cause for these outbreaks

via increased phytoplankton availability for

predator larvae, then better watershed management will be required. This will, however, require improved interagency cooperation, since

frequently the land and the sea are managed

by different government departments.

Alien Species and the Integrity of the

Local Ecosystem

Invasions by introduced (alien) species have

been ranked among the most potential perturbations of marine ecosystems (Carlton 1994),

and alteration of habitats and food webs by invasive species has been proposed as a major

factor in degradation of coral reefs (Birkeland

2004). Most information is available from harbors, the logical entry point for marine introductions, and much less from coral reefs (Coles

et al. 1999a,b, 2004; Paulay et al. 2002; Hewitt

2002; Buddo et al. 2003; Ray 2005). However,

in harbors most invasions originate and many

species escape onto the reefs. Of the 85 introduced marine species on Guam, 23% occurred

on reefs outside of Apra Harbor (Paulay et al.

2002), and from 28 introduced species in Pago

Pago Harbor on Tutuila, American Samoa, six

made their way onto coral reefs (Coles et al.

2003). Of ∼100 introduced and cryptogenic

(of uncertain origin) species in Hawaiian harbors and Kaneohe Bay, 54 were found on reefs  ̄

in the main Hawaiian Islands (max. 23 at a single site; Coles et al. 2006). None of these species

were considered invasive, that is, monopolizing

habitat or excluding native species, at any reef

site. Further afield from busy ports, only ten introduced or cryptogenic species were found at

Johnston Atoll (Coles et al. 2001), two at French

Frigate Shoals (DeFelice et al. 2002), and three

at Midway in the northwestern Hawaiian Islands (DeFelice et al. 1998).

Introduced or invasive species can become

a major threat and disrupt reef communities, especially when the latter are subject to

other disturbances. A striking example is Kaneohe Bay (Oahu, Hawaii), where reefs in the

south basin experience low circulation, have

restricted exchange with open ocean water, receive urban storm runoff, and had received

sewage for about 20 years that produced eutrophication and caused a shift from corals to

sponges and ascidians. After removal of the

sewage outfall in 1997, substantial coral recovery occurred, but subsequently invasive species

began to dominate shallow reefs. The invasives

are macroalgae (Gracilaria salicornia, Kappaphycus alvarezii, Kappaphycus striatum, and Eucheuma

denticulatum) that were intentionally introduced

into the bay for mariculture in the 1970s

(Russell 1992; Rodgers and Cox 1999; Woo

et al. 1999; Smith et al. 2002, 2004) and a sponge

(Mycale grandis, first noted in the mid-1990s;

Coles and Bolick 2007). Reef flats and back

reef slopes are now overgrown by dense algae

mats, providing habitat for other introduced

invertebrates and excluding native fauna. The

sponge now undermines and overgrows corals

(Fig. 12C). And these are only the most invasive

of 116 introduced or cryptogenic species (Coles

et al. 2002b), comprising a whopping 19% of

all species identified in Kaneohe Bay.  ̄ G. salicornia also dominates reef flats along Waik ̄ik ̄i

(Fig. 12B). Other problematical invaders are

Acanthophora spicifera, the earliest introduced algae (first noted in 1950; Doty 1961) and now

the most widespread in Hawaii (Smith et al.

2002) and Hypnea musciformis (Russell 1992), also

intentionally introduced into Kaneohe Bay in  ̄

1974 for mariculture. It spread and now occupies up to 80% of available substrata on some

reefs in Maui, where piles of rotting algae cause

net losses of over $20 million per year due to declining beachfront property values (Cesar et al.

2002; Smith et al. 2004).

Other notable Hawaii invasives are an alga

(Avrainvillea amedelpha), two invertebrates (Gonodactylaceus falcatus and Carijoa sp. aff. riisei), and

two reef fishes (Lutjanus kasmira and Cephalopholis

argus), all of which have proliferated and altered

coral-reef ecology and habitats (Fig. 12D). A.
Riegl et al.: Coral Reefs 157

Figure 12. Alien invaders on coral Kaneohe Bay reefs. (A) Carijoa sp. aff. reesei overgrowing black

corals in Hawaii. (B) A mass of Gracilaria covers what formerly was coral at Kaneohe Bay, Oahu. (C) The

sponge Mycale grandis busy killing a Montipora at Kaneohe Bay, Oahu, Hawaii. (D) A school of the common

introduced snapper Lutjanus kasmira on Oahu. (All images by S. Coles.)

amedelpha was first reported in 1981 and dominates soft or sandy bottom areas, displacing

sea grasses on Oahu (Smith et al. 2002). G. falcatus, a stomatopod common in coral rubble

that arrived from the Philippines at the end

of World War II (Kinzie 1968), now outnumbers the native species by a factor of 100 (Coles

and Eldredge 2002). The “snowflake” octocoral Carijoa sp. aff. riisei was first reported in 1972

in Pearl Harbor and has since spread to all the

Hawaiian Islands (Grigg 2003, 2004; Kahng

and Grigg 2005). It overgrows the deep blackcoral trees (Antipatharia sp.) that are a probable major source of larval replenishment for

the black corals that are harvested in shallow

depths for jewelry production (Fig. 12A).

In an attempt to increase the diversity and

abundance of reef game fish, federal and state

fisheries agencies imported 11 species of snappers and groupers from the Marquesas and Society Islands to Hawaii between 1951 and 1955

(Brock 1960; Randall 1987). Of these, three

have established breeding populations, and two

are considered invasive pests. The bluestripe

snapper Lutjanus kasmira is now abundant on

reefs. Its dispersal rate was 18–70 nm per year,

and it now has reached Laysan, 820 nm from

the introduction site (Grigg et al. 2008). It occurs

in large schools, is not considered a desired target fish, and is believed to compete with more

desired native reef fish for habitat and food. The

peacock grouper Cepholopholis argus is also not a

desired food fish because of reported Ciguatera

poisoning. It is an effective predator considered

to compete with native fishes for resources, although recent research has indicated that its
158 Annals of the New York Academy of Sciences

impact to the reef ecosystem may indeed be relatively low. It dispersed only 490 miles from the

introduction, suggesting a spread of 5–17 nm

per year (Grigg et al. 2008).

Two of only three known scleractinian coral

introductions are from the Caribbean (the third

being the ahermatypic Oculina patagonica into

the Mediterranean; Zibrowius 1974; Fine et al.

2001). The orange cup coral Tubastrea coccinea

was first reported at Curac ̧ao in 1943 (Cairns

2000) and has since spread to Brazil (Ferreira

2003; Creed 2006), Florida, and the Gulf of

Mexico (Fenner and Banks 2004). It may have

originated from the Cape Verde Islands or

the Gulf of Guinea (Fenner and Banks 2004),

since it would have been exposed to freshwater if transported as vessel fouling through the

Panama Canal. Or its planula larvae might

have been transported in ballast water. In Jamaica, T. F. Goreau introduced the solitary

mushroom coral Fungia scutaria at Discovery

Bay in 1966 for experimental purposes, and

despite hurricanes, ecological degradation of

the reef, and attempted eradication, they yet

persist (Bush et al. 2004), indicating a surviving

breeding population after more than 40 years.

Off southeast Brazil, the introduced soft coral

Stereonepthya aff. curvata competes with the native

octocoral Phyllogorgia dilatata (Ferreira 2003).

An incredibly efficient invader, the Pacific

lionfish Pterois volitans was apparently the first

successful introduction of a marine fish from

the western Pacific to the Atlantic (Whitfield

et al. 2002; Ruiz-Carus et al. 2006). First noted

off Palm Beach in October 1992 shortly after

an accidental release of six fish from a marine

aquarium during Hurricane Andrew in August

1992, it reached Long Island, New York, and

Bermuda a decade later. By 2008 it had occupied west Florida, other east Atlantic states,

the Bahamas, Cuba, the Cayman Islands, Jamaica, Turks and Caicos, the Dominican Republic, and Puerto Rico (Ray 2005). It is unknown if the entire population stems from the

originally documented accidental release, since

about 80,000 have been imported to Tampa

and Miami in the past 40 years (Ruiz-Carus

et al. 2006) and additional releases are probable. The lionfish is a highly effective predator

and could become a keystone species substantially altering food webs and energy availability

to native predators. It also carries a leach parasite that may infect native species with unknown

consequences.

Other introduced or invasive species on

Caribbean coral reefs are (1) the Mediterranean portunid crab Charybdis hellerii, first reported off Cuba in 1987 and now occurring

from Florida to Brazil (Tavares and De Mendonca 1996); (2) the invasive Pacific green mussel Perna viridis, first seen in Trinidad in 1990

and now in Jamaica and Tampa Bay since 1998

(Buddo et al. 2003); (3) the “world’s most common brittlestar” Ophicatis savignyi, which dispersed from the Pacific Ocean to the western

Atlantic (Roy and Spooner 2002); (4) the Pacific batfish Platax orbicularis and many species

of Pacific ornamental reef fish, all apparently

from aquarium releases (Semmens et al. 2004).

Although introduced species are usually a

minor component of the total biota of reef

ecosystems, alien invasions can have serious

negative impacts. Proliferation, competition

with native species, and alteration of the invaded habitat appears to be greater in reef areas

that are already disturbed by other negative environmental influences. Management implications are therefore relatively clear-cut and relate

to avoidance and early detection. Although the

vectors of introduction of alien species are often

unknown, some of the worst impacts occurred

from intentional introductions, for example,

macroalgae for mariculture and marine fishes

for “fisheries enhancement,” with unplanned

and unwanted results. Any such efforts must

be viewed skeptically and if alien species are

brought into proximity of any marine ecosystem, all effort and caution must be exerted

to isolate them from the natural environment.

There is often a lag time of about a decade between the time an alien species is introduced

and when it shows invasive properties. Early

detection and control are therefore important

in preventing serious environmental impacts.
Riegl et al.: Coral Reefs 159

Once an invasive alien species has become established, it is difficult to impossible to eradicate

it, and even reducing or controlling its population requires an intense and expensive effort. At

a cost of over Australian $2.2 million (Bax et al.

2002) the black-stripe mussel Mytilopsis (Congeria) sallei was successfully removed from Darwin

Harbor (Labowitch et al. 2005) and in Hawaii, a

vacuumcleaner-like device (the “Supersucker”;

Conklin et al. 2008) appears to be the only

device able to reduce the densities of introduced algae. Manual removal (Smith et al. 2004)

proved inefficient. To destroy Carijoa riisei in

Port Allen Harbor, Kauai, 738 pier pilings were

wrapped with plastic for 15 months and still did

not manage to remove the entire population (A.

Montgomery, personal communication).

Public awareness of the impacts of alien

species should be increased, discouraging the

release of imported maricultured or aquarium

species and intercepting introduced species that

may be transported by both commercial and

private vessels as hull fouling or ballast water.

Connectivity and the Threat Posed by

the Lack Thereof

While the propagation of nonnative fauna

is highly problematic, that of native fauna is

desired and absolutely necessary. The effectiveness and long-term success of a coral-reef

marine reserve will be compromised if its

propagules have no place to go, or even worse,

if the reserve itself is never reached by propagules from other areas (Gaines et al. 2003).

The concept of connectivity was early on

recognized as crucially important for coral-reef

conservation and is receiving much attention

(Cowen et al. 2006, 2007). Connectivity refers

to the linkage of populations in space through

the exchange of offspring (Roberts et al. 2006).

The existence of planktonic larval phases gives

the impression of easy connectivity across large

distances in the ocean. This would suggest that

habitat fragmentation should be less severe

than on land, where the loss of corridors

between intact habitat patches rapidly leads

to isolation and biodiversity loss (Turner 1996;

Lens et al. 2002). However, coral reefs may not

act all that much differently from terrestrial

systems, and while most organisms have larvae

with at least a short planktonic phase, withinreef retention of larvae emerges as apparently

very important (Cowen et al. 2007). Even if that

was not so, and long-range dispersal of species

was always readily feasible, local declines of

species in shrinking habitat patches reduce

the amount of larvae available, and thus the

chances of reaching a distant refuge dwindle.

Thus, to ensure survival, networks of protected reefs are needed that operate on all scales

of the landscape and the populations. Protected

reefs should link with unprotected reefs outside

the boundaries. But how much reef should be

protected? Theoretical considerations (Roberts

et al. 2006) suggest that many small reserves may

be as good as a few big reserves, since connectivity between many small and closely spaced reserves is higher than among few, widely spaced

ones. Gaines et al. (2003) have shown the overriding importance that local current patterns

can have. In the presence of strong currents,

due to connectivity between reefs, several small

reserves are more efficient than a single one

of the equivalent size. And well-connected reserves can outperform effort-based management strategies in terms of fisheries yield.

Management action in this case is soundly

based in good science prior to the declaration

of reserves (Gaines et al. 2003). Explicit information regarding currents and larval transport

in the areas targeted for new marine reserves

will be required. Thus investments in ocean

observation networks that give information on

water movements have an important payoff

for conservation planning. Salm et al. (2006)

also suggest the introduction of wide buffer

zones around existing reserves to account for

any shifts in habitat utilization due to climate

change and to avoid disruption of established

connectivity patterns.

When coral-reef species become rare, knowledge of connectivity patterns can help generate valuable management-related information.
160 Annals of the New York Academy of Sciences

Population genetic work on the two Atlantic A.

species, A. palmata and A. cervicornis has shown

spatial structure in their connectivity (Baums

et al. 2006; Vollmer and Palumbi 2007). A. palmata appears to have a largely open population

structure and seems to fall into two major populations (an eastern and a western Caribbean

population), with clear dispersal breaks caused

by oceanic phenomena (Baums et al. 2005;

2006a,b). Also, sexual recruitment appeared

more frequent in the species’ eastern range than

in the west. From this, one could deduce that

the eastern populations would have higher gene

flow with more likely adaptation to changed

environment (Baums et al. 2006a,b), making

them potentially more valuable for conservation. Zvuloni et al. (2008) showed in the Red

Sea that relatively rapid adaptation is indeed

possible, given frequent enough sexual recruitment. Both endangered species of Atlantic Acropora propagate only in relatively small regions

(Pineda et al. 2007; Vollmer and Palumbi 2007),

often within a few hundred meters of a spawning locus. This clearly shows that at least these

species need local source populations for their

recovery and that management must act on

local, not just large connectivity scales. Ridgway et al. (2008) using a similar genetic approach found larger-scale connectivity patterns

in southern Africa. Thus, population genetics

offers much insight into connectivity patterns

and can support the development of regional

and local management plans. Palumbi (2003)

showed how genetic patterns of isolation can

be useful to help estimate connectivity between

areas (e.g., reserve networks) where this may

be difficult to do by observation of recruitment

alone.

Potential for Extinction

Coral reefs are widely distributed, as are

many of the constituent species. Extinction of

widely distributed species is arguably less likely,

due to broad distribution and large populations.

However, it appears that many coral-reef organisms are indeed vulnerable and that, in fact,

many coral-reef species may be headed for extinction (Carlton et al. 1999; Dulvy et al. 2003;

Carpenter et al. 2008). Geologically speaking,

extinction of coral-reef taxa is not unusual;

however, the last faunal turnover took place

over 2 million years ago (Budd and Johnson

1999; Budd 2000), and even the upheavals of

the ice ages have not led to massive extinction

(Pandolfi and Jackson 2006). Rather, many new

species, in particular of the genera Acropora and

Montipora, have arisen only in the Qaternary,

and we are graced by several species in statu

nascendi (Veron 1995; Willis et al. 1997). It is

many of these young species that, according to

Carpenter et al. (2008), are headed for extinction. Why would their tenure on Earth potentially be such an unusually short one?

Many species accrue extinction debt once

their population numbers reach certain low levels. Besides extinction debt (Tilman et al. 1994),

the Allee effect, too low numbers of reproductive units to allow successful fertilization, can

come into play (Gascoigne and Lipcius 2004).

Many Caribbean corals have become locally

so rare that successful reproduction appears

more and more questionable (Edmunds 2007;

Edmunds and Elahi 2007). In the Caribbean

coral M. annularis fertilization success drops dramatically when the number of spawning corals

in an area becomes too low (Levitan et al. 2004).

But Allee effects in the ocean are difficult to

quantify and apparently do not always relegate

species to extinction, as is well demonstrated

by the case of the Caribbean sea urchin Diadema antillarum that is staging a slow comeback,

despite near extinction in the 1980s (Miller

et al. 2003). Extinction debt refers to species

extinction due to habitat shrinkage. Despite remaining locally abundant, the extinction-prone

species eventually are outcompeted by better

dispersers. The present decline of two Acropora in the Caribbean carries hallmarks of two

conflicting messages. It may well be the case

of species committed to extinction. They are

common where they exist but recruit rarely.

Thus, population crashes are dramatic and do

not reverse easily. However, also in this species
Riegl et al.: Coral Reefs 161

some interesting, at least temporary recovery

(Idjadi et al. 2006) and persistence (Keck et al.

2005; Riegl et al. 2009) have been observed.

But, asexual reproduction through fragmentation may help ensure persistence of these

branching corals, which occurs much less frequently in other endangered corals like Montastraea (Bruckner and Bruckner 2006b). Maybe

we ought not be too pessimistic. After all, no

extinctions of corals have yet been observed, although near extinction is on record (Glynn and

de Weerdt 1991; Glynn and Feingold 1992).

However, even if corals may not become fully

extinct, ecological extinction, that is, them becoming so rare that they can no longer fulfill

their role, would be worrying. Not only would

it mean an end to reef-framework production,

but a host of associated species that depend on

the corals for their existence would probably

suffer similar population restrictions, if not extirpation. Coral extinctions would only be the

tip of the iceberg for a much greater number of

extinctions of obligatorily associated species.

From a conservation standpoint the lessons

are clear. A multitude of protected reefs helps

keep extinction debt low, since sufficient habitat is available, even for the locally dominant

but badly dispersing species. A large number

of reserves can also reduce Allee effects by

maintaining sufficiently strong populations to

avoid Allee’s dip in the early phases of population rebuilding. Also, many reserves provide a

means of insurance in the face of large-scale disturbance (Halpern 2003; Palumbi 2003). The

more representative types of environments are

included in protected areas, and the more this

protection is replicated within reserves, the less

the likelihood that a single catastrophic event

will cause extinction (Allison et al. 2003; Salm

et al. 2006).

Alternative Stable or Quasi-stable States

What can be done to prevent a coral reef

from flipping into a state where structure and

function has deteriorated, goods and services

are reduced, and that we humans consider

as undesirable? The concept of phase shifts

on coral reefs, namely that such a complex

system can apparently come completely undone and fall into an alternative stable state,

has received much attention since it has been

demonstrated on Jamaican coral reefs (Hughes

1994). The idea is that, through a variety of

natural and anthropogenic stressors, corals declined dramatically and then, after a population crash of the dominant herbivore Diadema

antillarum, a formerly coral-dominated system

became dominated by fleshy algae (Fig. 13) due

to the absence of top-down control, that is, a

lack of grazers since the reefs were overfished

and the urchins had died. Absence of the natural predators, removed by overfishing, allowed

this situation to persist in a stable state and the

phase shift was accomplished (Hughes 1994;

Mumby et al. 2007; Fig. 13). Others have argued that such phase shifts are only quasi-stable

(Knowlton 2006), until one or the other organism returns to sufficient population density to

effect the return shifting of the phase. However, convincing case studies of reversals are

yet missing. While in some areas corals seem

to be recovering or persisting (Vargas-Angel

et al. 2003; Keck et al. 2005; Rowlands et al.

2008; Zubillaga et al. 2008; Riegl et al. 2009),

in other areas gains are undone by new disturbances. A phase-shift reversal observed by

Idjadi et al. (2006) and Precht and Aronson

(2006) in which A. cervicornis appeared to regain

dominance over macroalgae and coral cover

was on the general increase, unfortunately collapsed again in 2005 when yet another bleaching event depressed coral cover even further

than pre-reversal (Quinn and Kojis 2008). But

there seems to be some hope that even the badly

stressed Caribbean coral reefs still have some

resilience left (Fig. 14). There is also the possibility that Acropora and Montastraea may never

regain their dominance, but other branching

and massive corals may replace them, serving similar ecological and geological functions.

This trend appears to be underway in some remote locations in the Caribbean, such as Mona

Island in Puerto Rico (Bruckner et al. 2009).
162 Annals of the New York Academy of Sciences

Figure 13. The existence of stable and unstable equilibria in (A) coral reefs suggests that vistas of (B)

degraded coral reefs could remain permanent. Coral reefs, like many complex adaptive systems, exhibit

hysteresis. This means that stresses effect changes, but the removal of the stresses does not assure that

the system reverts to a stage it has previously been in. Stable (solid squares) and unstable (open squares)

equilibria are situated along model trajectories by Mumby et al. (2007). Once the system is pushed over a

threshold, it does not return to its previous state.

Figure 14. Areas in the Caribbean where the beleaguered Acropora species have been

reported to hold their own, or increase. Not all such increases have persisted, but they are

nevertheless a hopeful sign. (Modified from Precht and Aronson 2006; added points are

based on data in Keck et al. 2005; Macintyre and Toscano 2007; Rowlands et al. 2008;

Zubillaga et al. 2008; Riegl et al. 2009.)

Alternative stable states are not easily addressed in management action. With regard

to algae-dominated reefs, the best action is to

combat overfishing in order to maintain topdown control on proliferation of algae (Mumby

et al. 2006). Bottom-up control (i.e., algal proliferation caused by nutrient input) can be managed via better watershed and wastewater management that hopefully reduce the import of

nutrients into the system, a key factor in algal

proliferation. When corals are simply scarce,

but the habitat still suitable, coral nurseries and
Riegl et al.: Coral Reefs 163

transplantation (Jaap et al. 2006) or reseeding

reefs with mass-cultured larvae (Hayward et al.

2002) have proven successful in increasing coral

numbers.

Some key factors for management as outlined by Mangel (2006) are the realization that

thresholds exist, up to which ecosystems are

fairly resilient and changes are relatively small.

However, once these thresholds are exceeded,

changes happen fast and can be irreversible.

When radically altered, complex adaptive systems may never return to the original state, even

if the stress is removed. This process, called hysteresis, has been demonstrated to exist in coral

reefs (Mumby et al. 2007). Diversity is important, and even if productivity may not be decreased with the loss of some species or landscape components, the resilience of the system

may be affected. Also, since multiple spatial and

temporal scales interact in complicated ecosystems such as coral reefs, they cannot be understood from the perspective of a single time or

spatial scale. The strong linkages among components within reefs and among reefs, even over

wider geographical areas, need to be taken into

account. The impacts of disrupting these linkages are variable and need to be better understood. Coral reefs, like all ecosystems, change

with time, and the different components change

with different speed. It has to be borne in mind

that the human components change probably

more rapidly than any of the reef’s biological

components (Mangel 2006), and the danger of

outstripping the capacity of the ecosystem for

accommodating this change is real and its results plainly visible in the degradation of many

reefs around the world.

Local Threats

Many natural threats beleaguer coral-reef

ecosystems. Although ultimately caused by human action and the way we interfere with

climate and ecosystems, these are large-scale

threats that cannot be addressed by individuals, but rather require a concerted effort by

societies. However, there are many localized

threats, caused ultimately by our behavior as

individuals or small-scale economic units that

cause just as much destruction. If we could

learn to mitigate or eliminate those local disturbances, much could be won. To avoid phase

shifts, reefs must be managed to ensure the integrity of all its components. Reef fisheries must

be strictly controlled, since even the healthiest

corals will eventually deteriorate once the fish,

and other herbivores, are gone (Hughes 1994;

Mumby et al. 2007). Key ecological processes,

such as top-down control of algal growth by

herbivory, must be maintained to avoid knockon effects that will eventually disrupt all other

ecological processes as well (Hughes 1994).

Runoff, Sedimentation, and Nutrient

Enrichment

Runoff and associated siltation stress and pollutant loading frequently goes hand-in-hand

with development of coastal lands. Poor landuse practices in the headwaters of river basins

can have pronounced negative effects on coral

reefs. Australia’s Burdekin River, for example, has been putting ecological stress on the

GBR lagoon since colonists began altering

the land and raising cattle in the late 19th

century (McCulloch et al. 2003). In general,

significant increases in sediment yield to the

GBR have been documented with 14.4 million tons reaching the GBR lagoon, a threefold increase since precolonial times (Furnas

2003; Hopley et al. 2007). The situation is

similar in Kenya’s Sabaki River, which has

been transporting increasing amounts of sediment toward the sea ever since the uplands

came under the plough or were used as pastures (Fleitmann et al. 2007). This situation is

replicated thousandfold across the tropics, and

steep tropical islands or coastlines are particularly vulnerable to the effects of devegetation that increases runoff and sedimentation.

In much of the tropics, pulsed high-intensity

precipitation associated with tropical weather

systems (Rogers et al. 2008) exacerbates the
164 Annals of the New York Academy of Sciences

problem. Freshwater plumes and runoff can

travel dozens of kilometers into the sea

(Andrefou  ́ et ̈ et al. 2002). During such events,

coral mortality can be dramatic, especially

when preceded by other stresses like bleaching

(Riegl et al. 2009). Over geologic time, runoff

can shape coral communities more than even

temperature variability (Pandolfi 1999), clearly

disadvantaging the exposed corals (Ballantine

et al. 2008; Rogers et al. 2008). On Molokai,

rainfall and waves generated by storms can

generate more than 1000 times greater sediment deposition than normal (Jokiel 2008) due

to resuspension and deposition of fine, terrigenous sediment. Not only is sedimentation in

and of itself harmful to corals, the associated

increase in nutrient import from agriculture or

sewage poses a significant problem. Coral reefs

mostly grow in nutrient-poor environments (0.2

to 0.5 μM ammonium, 0.1 to 0.5 μM nitrate,

and <0.3 μM phosphate; Furnas 1991) and increases in nutrient loading not only favor algal

proliferation, that can lead to the displacement

of corals, but also directly interferes with the

corals’ calcification mechanism, and increases

mortality rates either directly (Renegar and

Riegl 2005) or by increasing susceptibility to

diseases (Bruno et al. 2003). There appears to be

interaction between nutrients and pCO2 that

increases mortality in some corals (Renegar and

Riegl 2003). Global climate-change scenarios

forecast more precipitation in a warmer tropical atmosphere, leading to more runoff, at least

during storms (Riegl 2007), as well as increased

pCO2.

In efforts to reduce grazing pressure and concomitant runoff on the Hawaiian Island Kahoolawe, over 20,000 goats are currently being

eliminated. This is demonstrating success, as

sediments are more rapidly winnowed from the

reefs than redeposited (Jokiel 2008). A classic

case study of sewage pollution and remediation

is that of Kaneohe Bay in Hawaii. From the

early 1960s, raw sewage was discharged into

the south basin of Kaneohe Bay, which led to

proliferations of phytoplankton and benthic algae that overgrew and killed corals. Removal

of the sewage outfall in 1979 led to dramatic

increase in water quality, decreases in nutrient

levels, and a dieback of algae. By 1983, coral

cover had begun recovering, going from 12%

to 26% (Maragos et al. 1985; Hunter and Evans

1993; Jokiel 2008).

Management actions to alleviate direct

sewage input are relatively straightforward, like

effectively managing watersheds so that soil

and vegetation naturally filter storm and waste

waters, thereby keeping rivers clean. However,

what sounds quite straightforward is in reality

often complicated, since different government

agencies with nonoverlapping jurisdictions and

aims must collaborate to succeed in this effort. In many countries that challenge has effectively negated efforts to relieve stress from

up-catchment onto the coastal reefs. In many

areas governance is so ineffective that only an

offshore location, which equates to a removal

from the stresses due to stronger dilution of pollutants, gives reefs a good chance of survival.

However, nitrification of coastal ocean areas

may lead to coral-predator outbreaks, therefore, indirect effects can be far-reaching.

Coastal Construction

A significant threat in many countries across

the reef belt is the proliferation of coastal construction directly affecting coral reefs. More

and more people live in the coastal zone, and

more and more infrastructure is needed to accommodate expanding populations along with

the increased trade that sustains their economic

system. It is assumed that coastal development

threatens 33% of the Caribbean region’s reefs

(Wilkinson 2006). In the Arabian Gulf, which is

undergoing an unprecedented building boom,

the entire coastline of some political entities

has been moved several kilometers into the

sea, burying coral reefs and other habitats under artificial islands. Virtually the entire southern shoreline of the Arabian Gulf is currently

subject to major reshaping. The Red Sea is

not much better off, with massive development projects taking priority over preserving
Riegl et al.: Coral Reefs 165

coral reefs. Florida has developed densely along

its southeastern continental reef tract (Banks

et al. 2008) and in the Florida Keys but has at

least avoided building on the reefs. Nonetheless,

coastal erosion that requires frequent beach

renourishment, pollutant runoff, sewage discharge and anchorage off major ports, and

ship groundings strain the system (Banks et al.

2008). Honolulu, Hawaii, and Tutuila, American Samoa, are of course not the only Pacific islands with an airport built on a reef flat.

Dredging for construction and maintenance of

coastal infrastructure near reefs often can cause

significant stress or mortality (Wolanski 1994).

Coastal construction is often a necessity for

local economic development and can be accomplished in relative harmony with the environment, when properly designed. But in many

countries the environmental impact assessment

(EIA) process, meant to avoid unnecessary environmental insults, is flawed or summarily ignored. Often, environmentally sound construction techniques, even where available and affordable, are not employed (Sheppard 2003b).

Thus, a strengthening of the EIA process

is required, in particular in smaller nations,

that should go hand-in-hand with feedback

monitoring, a form of adaptive management

that monitors impacts during the construction

phase for rapid correction of environmental

damage (Turner et al. 2006).

Management intervention would again

appear to be straightforward in this case. If

all development projects in the coastal zone,

in particular in coral-reef areas were properly

evaluated for their expected impacts, then those

could be mitigated. This, however, assumes an

efficient and functioning, noncorrupt EIA process, which may be surprisingly rare in countries with coral-reef resources. Unfortunately,

well-meaning laws or EIAs that identify potential problems are often discounted. In many

places, self-regulation of the construction industry is observed, which is often far more efficient than formal actions taken by the local

trustee of the marine resources. A worldwide

overhaul of the EIA process would be extremely

beneficial to coral-reef conservation (Sheppard

2003b).

Overfishing and Destructive Fishing

Overfishing has been well documented as a

root cause of coral-reef decline (Jackson 1997;

Jackson et al. 2001). Hughes (1994) saw overfishing of grazing fish (primarily parrot fishes)

as a reason for a phase shift away from corals

and toward fleshy macroalgae when the other

major grazer, Diadema antillarum, was wiped out

by a disease. It is unfortunate that coral reefs

are often by necessity a primary target for fisheries since they are in truth not very productive

fisheries systems (Hatcher 1997; McClanahan

2006). While algae production can be as high

as in agricultural systems, the high internal demand for this production reduces the amount

available for human consumption via fisheries

to a mere 1%. Oceanic upwelling systems produce more than 50 times the fish biomass per

unit algae as coral reefs do (McClanahan 2006).

Not surprisingly, coral reefs are rapidly overfished when a hungry populace depends on

them as protein source.

Overfishing is a primary threat to about 60%

of Caribbean coral reefs (Wilkinson 2006), but

this is probably a conservative number. Almost

all regions that are well studied report some

degree of overfishing (Ballantine et al. 2008;

Jokiel 2008; Rogers et al. 2008). The situation is

similar in the Pacific, with significant overfishing reported from American Samoa (Birkeland

et al. 2008) Guam and the Marianas (Richmond

et al. 2008), the main Hawaiian islands, and in

most other island and coastal nations. In U.S.

jurisdiction, only the northwestern Hawaiian

islands and the Pacific Remote Islands Area report near-pristine fish populations (Birkeland

et al. 2008; Grigg et al. 2008). However, even

there, some targeted species have collapsed and

not regenerated yet, like the overharvest of pearl

oysters on Pearl and Hermes Atoll in the early

1900s (Grigg et al. 2008). To combat overfishing, Australia declared 33% of its Great Barrier Reef Marine Park no-take areas (Day et al.
166 Annals of the New York Academy of Sciences

2003; Fernandez et al. 2005). This is the world’s

biggest no-take area, and it has already shown

success by increasing fish populations.

Overfishing is nothing new, since human–

coral reef interactions date back at least

35,000 years (Pollnac 2007). While some consider the effects of early, preindustrial human

exploitation mostly benign (Johannes 1978,

1981), others believe impacts to have been

deleterious (Kirch and Hunt 1997) and potentially even responsible for the long-term degradation that eventually led to the deteriorated

state of many of today’s coral reefs (Jackson

1997; Jackson et al. 2001; Diamond 2005).

Fitzpatrick and Donaldson (2007) provide evidence that coral-reef exploitation in Palau has

over the past few thousand years led to declines in the numbers of finfish and mollusks.

In the Palauan Rock Islands, as a clear sign

of fishing pressure over 1400 years, the proportion of snappers and emperors found in

middens declined, as did the overall size of

all consumed fish (Masse et al. 2006). Signs of

overharvesting in prehistoric times exist for giant clams (Tridacna sp. and Hippopus hippopus) as

well as the humped conch (Strombus gibberulus).

The list of overexploited species in prehistoric

times is long and very similar to those of today

(Fitzpatrick and Donaldson 2007). Also in the

Caribbean, clear evidence of overexpoitation of

coral-reef fishes exists at several Ceramic Age

sites (LeFebvre 2007). The archaeological signs

are a decrease in the size of exploited coral-reef

species and an increase in the use of inshore and

pelagic species over time (Wing 2001). Once

overfished, resources may take a long time to

regenerate, particularly if fishing pressure is

maintained. Overfishing in the Marianas dates

back to the Japanese period (1914–1944) and

may have been influential in molding the current nearshore coral-reef community structure

(Richmond et al. 2008). To compound the problem, Guam witnessed a further 70% reduction

in coastal fisheries catch from 1985 to 1996. In

response, no-take areas were established, that

originally met with strong local opposition, but

when finally enforced rapidly led to increases in

fish stocks inside the reserves (Richmond et al.

2008), demonstrating again the efficacy of protected areas for fisheries management (Fig. 15).

Destructive fisheries, like dynamiting, muroami, cyanide, and bleach fishing threaten many

coral reefs. Dynamite fishing is particularly

damaging and a major problem in parts of

Asia and Africa (Riegl and Luke 1998; Fox

and Caldwell 2006) and interesting cottage industries have developed around it. Sources of

explosives are often discarded munitions or leftover land mines from past conflicts. Recovery

of explosives is a specialized trade, and dynamiting teams consist of skilled bombers and

fish retrievers. Dynamiting is well planned and

in full knowledge of the damage that is done.

Parts of reefs are dynamited repeatedly, since

fish biomass (especially herbivores) can increase

briefly after corals are destroyed and fleshy algae take over. Only after herbivore biomass declines is a new section of the reef dynamited

(Riegl and Luke 1998). Many fishers are aware

of the unsustainable nature of such practices.

A frequent excuse was that the reefs were already lost to the fishermen, with the perspective

that tourism development and/or protection in

marine reserves would eventually make fishing

impossible. Thus, no need to protect the resource for future livelihood was perceived. In

the Caribbean, bleach fishing for lobster and

other shellfish is a major problem, degrading

primarily patch reefs and shallow reefs.

The successful management of reef fisheries

has proven to be an elusive goal. Fisheries regulations are regularly ignored either for reasons

of profit or because the reef fish are needed

as protein sources or shellfish can be sold as

curios (Fig. 16). Alcala et al. (2006) review the

highly successful case of the voluntary reserves

enacted by the local communities at Apo and

Selinog Islands in the Philippines, where the

reserve has indeed created a biomass source

for surrounding areas (Fig. 15). Other successful community-based conservation efforts

have been reported, for example, from Papua

New Guinea (Aswani et al. 2007). Obura et al.

(2006) report from Kenya that local community
Riegl et al.: Coral Reefs 167

Figure 15. (A and B) The efficacy of marine reserves in increasing fish stocks, here acanthurids (sturgeons), carangids (jacks), lutjanids (snappers), and lethrinids (emperors) show how vulnerable reef fisheries

are to overfishing. (From Alcala et al. 2006.) (C and D) Show the mismatch between economics of consumption and the ecological role of coral reef fishes. (C) Is the biomass of functional reef-fish groups in a Pacific

ecosystem, and (D) contains their economic value. (From Bellwood et al. 2004.) It is little surprise that certain

species are almost driven to annihilation.

Figure 16. Fisheries not only affects finfish but also corals and other components of coral reefs. Is this

really the best use for these organisms? (A) These corals could have built a reef and been home to fish,

and (B) this trumpet triton Charonia tritonis could have preyed on the corallivorous cown-of-thorns starfish

Acanthaster planci and helped control its populations.
168 Annals of the New York Academy of Sciences

efforts at coral farming to make up losses from

bleaching events also led to voluntary restrictions on fisheries, to avoid damaging the coral

farm with fishing gear. These examples clearly

show that, provided the local community has

enough sense of ownership in the resource,

successful fisheries management is indeed

possible.

Attempting the Balance: Managed

and Protected Coral Reefs

Ours is an increasingly crowded world with

vast numbers of people living in the coastal

zone. Some of the highest population densities

in the world are in tropical coastlines close to

coral reefs (Shi and Singh 2003). In the United

States, the 17% of land that defines the coastal

zone holds 53% of the nation’s total population (Bowen et al. 2006). Many coral reefs occur

in the world’s economically disadvantaged regions and thus have a special importance as a

source of scarce proteins, often the only one affordable, to a vast local populace. This dubious

privilege has led to widespread degradation,

and some claim that hardly any reef exists today that is not overfished (Jackson et al. 2001).

Thus, many perceive some sort of protection

from human exploitation as the only possibility to maintain even a semblance of healthy reef

ecosystems. Enter the marine reserve or marine

protected area (MPA). The world is dotted with

small to large marine reserves, a great many of

which attempt to preserve coral reefs (Fig. 17).

The only ingredient that is needed for coralreef survival is efficient management based on

good science.

Theoretical approaches to MPAs have

been elaborated in many publications (e.g.,

Lubchenco et al. 2003) and are outlined in

Mangel (2006), who also gives a good review

and pointers to pertinent literature. Both in theory and practice, it has been shown that MPAs

can work. When managed properly, they have

been shown to increase biomass and diversity

both within the reserve and the adjacent areas

via larval export and spillover of adults, which

reduces the extinction probability of resident

species. These beneficial effects are even observed in small reserves (Halpern 2003). Wellconnected networks of MPAs ensure against

catastrophes and provide for long-term persistence of marine communities (Lubchenco et al.

2003). Yet recent analyses of coral-reef MPAs

show that most are not very efficiently managed and many achieve little to no protection

of their entrusted resources (Alcala et al. 2006;

Wells 2006).

Very clearly, science has a strong case in supporting management decisions. Unfortunately,

too often the science is available, but implementation lags, and although the theory is

well known, the practical “how-to, where, and

when” of coral-reef conservation is hotly debated. Many traditional societies have quite

capably conserved their reefs, but it is easier

to conserve a reef when only a few hundred

villagers depend on it than if thousands or

even millions of people are the reef’s neighbors. Thus, even though unquestionably successful (Johannes 1978, 1981; Aswani et al.

2007; Richmond et al. 2007), traditional values and village-based approaches work only

where societal and political systems still accommodate them–and these traditional values can

also erode very quickly (Oles 2007). In many

coral reef areas, “Western-style” centralization

of governance and decision making results in a

top-down approach by legal fiat that frequently

does justice to the word of the law, but as often

does preciously little to conserve the resource

(Latin 1993). Thus, a place exists for a theory of

conservation that results in development of rational conservation principles upon which legislative action can be designed. The scientific

literature is replete with such studies (Mumby

and Steneck 2008).

MPAs (Fig. 17) are considered to be essential tools for conservation, perhaps most importantly for the conservation of biodiversity.

However, the declaration of a protected area,

and effective protection of the resources within,

are two quite separate issues, and probably
Riegl et al.: Coral Reefs 169

Figure 17. Many coral reefs are protected, at least on paper. Efficient marine reserves, however, appear

to be few and far between. The biggest marine reserves with coral reefs are the Phoenix Islands Protected

Area (PIPA, Kiribati), Papahanaumokuakea (northern Hawaiian Islands) National Marine Monument, the

Pacific Remote Islands Areas Wildlife Reserves, and the Great Barrier Reef Marine Park.

the majority of coral-reef reserves are at best

only marginally effective (Alcala et al. 2006;

Ricciardi et al. 2006; Wells 2006). To aid in the

development and assessment of protected areas, and to encourage standards for assessment

and reporting and use of appropriate indicators that focus on outputs and outcomes, clear

guidelines have been developed (Hockings et al.

2000; Pomeroy et al. 2004).

Conservation is traditionally considered to

be within the realm of governmental custodians

or nongovernmental organizations (NGOs),

and much has been written about the effectiveness of government- or community-based

efforts (Wells 2006). There also exist undeclared

but de facto marine reserves. Since the biggest

threat to coral-reef ecosystem integrity appears

to be unfettered access by man, any activity restricting access can end up benefiting the reefs.

For example, military use of coral reefs can

in instances create remarkably efficient nature

reserves. About two thirds of the Caribbean

island of Vieques was used for practice bombing and amphibious assault from the 1940s until 2004 (Hernandez-Cruz et al. 2006). When

the Navy left, unexploded ordinance made the

land difficult to use and enabled transfer into

the biggest National Wildlife Refuge in the

Caribbean, making Vieques one of the most

undisturbed Caribbean islands. Despite bombing, the health of the reefs was better than in

nearby St. Croix, where reefs had been protected for decades (Riegl et al. 2008).

The major impacts on reefs had been due to

hurricanes and coral diseases, and the absence

of stressors from land-based sources helped

Vieques reefs stay healthy. In the Pacific, Johnston Atoll, one of the world’s most isolated

atolls, was a military site for the storage and destruction of chemical weapons as well as some

nuclear testing (Lobel and Lobel 2008). Wake

atoll had been used as a Navy airfield since

1934. Military governance as “forbidden areas” isolated these reefs and despite modification of the islands, the reefs maintained excellent health (Lobel and Lobel 2008), and were

largely unfished. Johnston atoll is an important

stepping stone for larvae from the tropical Pacific to reach the Hawaiian Islands (Kobayashi

2006). Also the British Indian Ocean Territory

(BIOT) benefited from remoteness due to military activity, with Diego Garcia being a major

U.S. Naval base. The entire BIOT was depopulated in the 1950s to free the way for military installations that remained concentrated

on Diego Garcia. With the atolls being under British Naval surveillance, the BIOT had

effectively become one of the world’s largest
170 Annals of the New York Academy of Sciences

conservation areas, and one of the few uninhabited island chains. While the U.S. Navy

modified Diego Garcia island and lagoon, conservation measures on half of the atoll allowed

spectacular coral reefs to coexist with a major

military installation. The northern atolls are

home to spectacular and healthy coral reefs.

These reefs suffered like many others in the

1998 mass bleaching event but, possibly due to

the absence of any human stressors, but have

recovered well, which supports the concept of

resilience in areas subject to little direct anthropogenic stress (Sheppard et al. 2008). Nonformal and unintentional coral-reef conservation occurred at Bikini Atoll. Site of the test

explosion of two hydrogen bombs, the island

had been depopulated when chosen as a test

site and due to continued radiation risk, had

been kept population free. Early investigations

showed the extent of the damage when nuclear bombs explode on a reef, but recent studies reveal impressive regeneration of the corals

and how apparently resilient the corals were

(Richards et al. 2008). An equivalent situation

exists at France’s Mururoa Atoll (Guille et al.

1996; Planes et al. 2005).

MPAs appear to work best and contain

the healthiest reefs when distant from human

habitation, like Hawaii’s Papahanaumokuakea

(northewestern Hawaiian Island) National Marine Monument (Rooney et al. 2008; Grigg

et al. 2008), the U.S. Pacific Remote Islands

Area National Wildlife Refuges (Maragos et al.

2008a,b), parts of Australia’s GBR (Kelleher

1993) or Kiribati’s Phoenix Islands Protected

Area (PIPA). In these areas, management plans

can be—and are—primarily science-based.

Frequently, however, MPAs are sited within areas traditionally used by local communities and

can therefore directly impact their livelihoods

or social identity. The success or failure of many,

if not most MPAs in coral-reef areas (Wells

2006; Christie and White 2007) is, among a

multitude of capacity-related factors, linked to

design, declaration process, and the involvement of the local communities. Studies in the

Bahamas have shown that if an MPA is perceived as interfering with traditional activities

regarded as key social components, strong local

opposition can have deleterious impacts on the

declaration process (Stoffle and Minnis 2007)

or relegate an MPA to the already impressive

heap of dysfunctional paper parks (Ricciardi

et al. 2006; Wells 2006).

Market forces tied to emigration, transmigration, the free flow of goods and capital, and a

strengthening of cultural and material imports

lead to changes in the social fabric of communities (Oles 2007), with implications for conservation attitudes toward coral reefs. Traditional

values linked to coral reefs as well as ecological

traditions and knowledge erode rapidly. Fortunately, coral reefs provide economic opportunities firmly rooted in the wage labor system,

freeing people from the pressures of subsistence and providing access to more money

and the spoils of the free-market economy. The

potentially biggest opportunity arises through

tourism (Spurgeon 2006; Diedrich 2007). Coral

reefs are believed to have a worldwide income potential through tourism of at least US$

9.6 billion per year (Jobbins 2006). But despite

offering potentially huge economic gain, coralreef-based tourism can also lead to conflict

and degradation. Tourism, if well planned and

managed, should result in social, economic,

and environmental benefits to the host community. However, if benefits are not distributed

equitably, significant conflict can be created

and local communities can quickly begin to

resent conservation measures. The feeling of

loss of control and access to natural resources

that are considered property will be aggravated

if material gain from tourism activities cannot be realized. In Belize, touristic use of reefs

contributes ∼30% of the gross domestic product via extractive and nonextractive ecosystem services and a positive correlation between

tourism development, and coral-reef conservation awareness and support was observed

(Diedrich 2007). Support for conservation was

largely linked to realized or expected financial

gain, indicating that local communities expect

protected areas to attract business. Financial
Riegl et al.: Coral Reefs 171

gain from marketing coral reefs is certainly realized in four South Florida counties (Palm

Beach, Broward, and Miami-Dade) adjacent

to the Continental Florida Reef Tract (Banks

et al. 2008). Reef use creates an annual contribution to the local economy of US$8 billion in

sales and almost US$4 billion in income with

70,400 jobs (Johns et al. 2001). This number factors in direct and indirect spending in connection with the tourism, marine, infrastructure,

and transport industries in order to gain access

to the reefs. Many other countries also effectively market their reefs to build their coastal

tourism. Tourism overall generated receipts of

US$856 billion in 2007 (World Tourism Organization, http:/www.world-toursim.org). Such

numbers should convince that losing coral reefs

equates to losing money and that many people

can benefit from reef conservation, directly and

indirectly.

Which Reefs to Protect?

Do Refuges Exist?

Conservation strategies for coral reefs vary

from strengthening existing protection by better training (Wells 2006) and devising specific

management responses to crises such as bleaching (Marshall and Schuttenberg 2006a,b;

Obura et al. 2006), to searching specific attributes of reefs (such as expected resilience

in the face of global change) for declaration

of specially protected areas; and to scientifically tailor-making management approaches

that promise to maintain the highest possible ecological and biodiversity integrity (Salm

et al. 2006). In their efforts, coral-reef managers are aided by increasingly sophisticated

monitoring and warning systems, such as the

National Oceanic and Atmospheric Administration’s (NOAA)’s Coral Reef Watch program

(Strong et al. 2006).

Objectives for coral-reef conservation could

be defined as efforts (1) to maintain or rebuild

the biological and structural integrity of coral

reefs, and (2) to secure the sustained delivery

of ecosystem services to people (Roberts et al.

2006). Obviously, point 2 follows if, and only

if, point 1 is achieved or achievable. Debate

continues regarding which reefs should be selected for protection, that is, where best to invest

sparse resources for maximum benefit. Since

bleaching will likely continue to be a major

issue for coral conservation, characteristics of

areas have been sought that would lower the incidence of bleaching. These factors include low

light [due to depth, shading, turbidity, or cloud

cover (Mumby et al. 2001)], high flow (Craig

et al. 2001; Nakamura and van Woesik 2001;

Nakamura et al. 2003; Nakamura et al. 2005),

lower temperatures (Riegl 2003; McClanahan

2008), and natural higher nutrient settings

(Grottoli et al. 2006—higher nutrient availability can aid bleached corals, but increased nutrient levels by pollution can also damage—

see the preceding sections; Fig. 18). Often specific environments correlate with these factors:

deeper reefs (often near to shallow thermoclines), reefs in upwelling areas (although too

much cold can kill corals; Glynn and D’Croz

1990; Hoegh-Guldberg et al. 2005), coastal areas with high levels of suspended terrigenous

sediment, areas with strong currents, and shoreline and lagoonal reefs that are shaded by high

islands (Glynn 1996; Riegl and Piller 2003;

Salm et al. 2006; West and Salm 2003). Such

potential ‘‘refuge’’ habitats have been singled

out as preferred target sites for conservation

(Salm et al. 2006). Some argument of course exists whether refugia are the same everywhere.

McClanahan et al. (2005a,b) argued, for example, that high flow does not necessarily protect

corals. Rather, low-flow areas with dramatic

temperature fluctuations as a result of restricted

circulation would help corals acclimatize, and

those “spoilt” by the more uniform temperature

regime in high-flow areas would bleach more

easily during temperature anomalies. Thus, the

role of the environment in helping corals acclimatize to temperature stress may be more important than its role in alleviating the metabolic

effects of the stress itself. Such findings complicate identification of resilient areas and refugia
172 Annals of the New York Academy of Sciences

Figure 18. Conceptual diagram, based on information from or quantitative findings in

other studies (Marshall and Schuttenberg 2006a,b; Salm et al. 2006; Riegl et al. 2009) on

how the setting of coral reefs relates to the stress state of corals, likely degradation, and

conservation value.

by using physical habitat characteristics alone,

because previous ‘‘experience’’ (not necessarily

predictable from simple habitat characteristics)

is likely to play a critical role in determining

whether or not corals actually bleach (Brown

et al. 2000, 2002). This makes a strong case for

monitoring both environment and corals over

large scales and extended periods in order to

obtain more clarity.

Where the environment does not naturally

protect, legal protection may offer a solution. Recently, the corals Acropora cervicornis and
Riegl et al.: Coral Reefs 173

A. palmata, which have suffered spectacular

population restrictions (Bruckner 2003), were

listed as threatened under the U.S. Endangered

Species Act. It remains to be seen whether

this level of protection actually helps or hinders the conservation of the species. The strict

protection of stony corals and black corals under Appendix II of the Convention on International Trade in Endangered Species (CITES)

that only allows the export and import of corals

under permits issued by the exporting country seems to be more a hindrance to research

than to actually preserve the corals (Green and

Hendry 1999), although it has raised international awareness about the impacts of the curio

and aquarium trades (Bruckner 2001; Fig. 16).

Whatever their size, both small and large

reserves have positive effects (Halpern 2003).

In analogy to terrestrial-conservation science,

which has been able to produce exact-size

prescriptions based on theories of biogeography and biodiversity (McArthur and Wilson

1963; Diamond and May 1976; Wilson 1988;

Hubbell 2001), it is certainly not wrong to

also suggest the bigger, the better for coral-reef

reserves. Theoretic relationships between reef

sizes and contained biodiversity are less well

established than in terrestrial systems. Coral

reefs and habitats that contain corals follow

a power law in their frequency (Purkis et al.

2007). This means that many more small reefs

exist than big ones. Also, with regards to disturbances, there usually exist more small and

localized than severe and widespread disturbances (Riegl et al. 2009). The conservation

message is clear: If many small, localized disturbances recur randomly, any one of the many

small reefs is in danger of being compromised

by such a disturbance sooner or later. While

nothing can ensure against the few, widespread

disturbances, an MPA should be big enough

to contain many small reefs, so that enough

can escape the many small perturbances. Ideally, the biggest coherent reef structures, and

many of the small, should be contained, in order to give the system a reasonable chance to

contain refugia when the few large, widespread

disturbances strike. This is the principle of insurance against unforeseen events suggested by

Wilson and Willis (1975), Allison et al. (2003),

and Mangel (2006). Of the many coral-reef

MPAs, few fulfill such a size requirement (obvious examples being the Great Barrier Reef

Marine Park, the Papahanaumokuakea National Marine Monument in the northwestern

Hawaiian Islands, the Florida Keys National

Marine Sanctuary, among others). Small reserves lose species due to ecological truncation

(Wilson and Willis 1975), and ecological drift

(Hubbell 2001) can lead to different species

dominating in different small patches, requiring

more patches to maintain high biodiversity in a

refuge. If coral communities are neither nicheaccommodated nor neutral (Dornelas et al.

2006), but shaped by high-frequency stochastic

events, then the diversity of patterns and species

compositions on coral reefs is even higher and

less reproducible in small reserves. All this suggests that large and well-connected reserves are

desirable to avoid losses of coral-reef biodiversity, especially at a time where stress events, like

bleaching, will recur more frequently (HoeghGuldberg 1999; Sheppard 2003a).

If reefs cannot be protected, or if something

goes wrong, there is a place for restoration and

a ripe literature exists. Also artificial reefs can

be useful and are becoming very popular in

many parts of the world as a means to mitigate

for damaged or destroyed reef habitat (Clark

and Edwards 1994), to improve fisheries yields

(Pickering and Whitmarsh 1997), or to reduce

recreational dive pressure on natural reefs. Both

restoration and artificial reefs are often met

with skepticism since it appears better not to

damage an ecosystem in the first place. However, large-scale natural impacts as well as the

increasingly common insults visited upon reefs

by ship groundings, other industrial accidents,

and natural disasters make restoration capabilities desirable. Yet restoration remains a viable

alternative only if the original threats leading to

reef degradation in the first place are mitigated

and appropriate strategies based on the type

of species and environmental conditions are
174 Annals of the New York Academy of Sciences

applied (Bruckner et al. 2009). Larval culturing

techniques (Omori 2005; Miller and Szmant

2006; Omori et al. 2008), enhanced and directed recruitment (Hayward et al. 2002), asexual propagation (Epstein and Rinkevich 2001;

Soong and Chen 2003), coral transplantation

methods and coral gardening (Edwards and

Clark 1998; Rinkevich 2006) hold much hope

to artificially improve regeneration. It is in particular in the realm of reef rehabilitation and the

creation of artificial reef ecosystems that basic

and applied conservation research can generate synergy. It is a field where ecological theory can fruitfully be translated into informed

and directed efforts to recreate lost biodiversity. If natural coral reefs remain on the currently observed downward trajectory, we may

too soon require all the knowledge and techniques we can muster to giver nature a helping

hand. Before we come to that stage, however,

efficient conservation of what currently exists is

the preferable option.

Conclusion

Coral reefs, in one taxonomic and evolutionary guise or the other, have graced the Earth

for about 500 million years. Modern, scleractinian corals evolved about 215 Ma and have

survived several major extinction events. While

frequently of tectonic origins or caused by collision with extraterrestrial objects, most of these

mass extinctions had a climatic component.

The lesson is that rapid climatic changes have

always caused major extinctions. Thus, given

the currently observed rates of climatic change,

there is reason to worry about the future of

coral reefs.

The greatest global-scale threats currently

faced by coral reefs appear to be all linked

to man-made or man-mediated changes of

climate:

• Bleaching, a heat- and light-mediated loss

of symbiotic algae within the corals, has

increased markedly in impact and severity

over the past decades and affects virtually

every reef worldwide.

• Diseases have increased in incidence and

diversity and caused severe population declines of corals in the Caribbean and,

to a lesser extent, the Indian and Pacific

oceans.

• Predator outbreaks have recurred repeatedly, most particularly in the Indo-Pacific

and have caused severe degradation on affected reefs. These outbreaks seem to follow natural dynamics, enhanced by manmade changes in the coastal zone, and they

can be particularly severe when the host

population is diminished or weakened by

other threats.

• Losses in keystone predators and herbivores have created phase shifts away from

corals and to the establishment of stable states dominated by algae. This phenomenon is typical of the Caribbean.

• Ocean acidification is an emergent problem that may threaten deep as well as shallow coral reefs.

Smaller-scale, localized, and entirely manmade threats include:

• runoff, sedimentation, and nutrient enrichment;

• coastal construction leading to smothering

of habitat and creation of high turbidity

around coasts;

• overfishing and destructive fishing techniques.

Conservation is attempted in marine reserves

and by legal regulation of activities on coral

reefs. Conservation science has provided a rich

theoretical body with regard to the required

ingredients, size, and connectivity of coral-reef

reserves to be efficient. In general, larger protected areas with little extractive interference

(ideally no-take) have a higher likelihood of conserving coral reefs long term.

It is estimated that costs due to lost economic opportunities from destroyed coral reefs

will reach US$350–870 million per year by

2015 of the annualUS$3.1–4.6 billion of annual

benefits from coral-reef fisheries, dive tourism,
Riegl et al.: Coral Reefs 175

and shoreline protection services (Wilkinson

2006). These numbers are conservatively low

and relate only to direct benefits lost. Bearing in mind that three counties in Florida

alone benefit from annual revenues of circa

US$4 billion (Johns et al. 2001), the direct and

indirect financial losses associated with worldwide coral-reef degradation will have undesirable consequences to our global economy.

Economic impacts aside, we have a moral,

ethical, and cultural responsibility to prevent

one of our world’s most biodiverse and esthetically beautiful ecosystems from being wantonly

destroyed.

A study of Earth’s history teaches us that

coral reefs have periodically survived destructive natural perturbations. What presently unfolds is a consonance of coral-reef disturbances

imposed by both man and nature. If the current trajectory of coral-reef degradation continues unabated, we will remain on the path of

a mass coral extinction event on the scale of an

asteroid impact (after which corals took over a

million years to recover). Corals will not likely

go completely extinct, but the coral-reef ecosystems that currently harbor immense biodiversity, provide the necessities of life for millions of

people, and produce valuable global economic

services will disappear. We have demonstrated

capabilities to exert great changes on our planet

with negative consequences; now it is our obligation to harness our capabilities to arrest and

reverse the trajectory of coral-reef decline.

Acknowledgments

Two of the authors (BR) and (RED)

acknowledge funds from NOAA CSCOR

NA16OA1443 to the National Coral Reef Institute (NCRI). All of the authors are grateful for

support through the Khaled Bin Sultan Living

Oceans Foundation. NCRI publication #105.

Conflicts of Interest

The authors declare no conflicts of interest.

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