Global and Regional Threats to Coral Reef Ecosystems
Among all the marine and terrestrial ecosystems, the diversity of coral reefs is rivaled only by rainforests (NOAA). In terms of ecosystem benefits, coral reefs provide coastal protection, crucial protein and income from fisheries, diving-related tourism, bio-medical opportunities, and intrinsic cultural and aesthetic treasures (Wolanski et al. 2003). The value of coral reefs to the global community is estimated to be in the billions of dollars, with the net present value of tourism on the Great Barrier Reef alone estimated at US$18–40 billion. These ecosystem values, like the health of the ecosystems themselves, are in worldwide decline (Fernandes et al. 2005). Globally, coral reefs are being destroyed at an alarming rate with some countries already having lost over 50% of their coral reefs in the last two decades (Wolanski et al. 2003).
Awareness about major threats to coral reefs began in the 1970’s, and reached a climax in 1998 in the event of an extensive and global coral bleaching event (Bellwood et al. 2006). That year, the U.S. statute titled “Coral Reef Protection,” signed by ex-President Bill Clinton, summarized the principal contemporary threats to coral reef ecosystems. These are, “land based sources of water pollution, sedimentation, detrimental alteration of salinity or temperature, over fishing, over-use, collection of coral reef species, and direct destruction by activities such as recreational and commercial vessel traffic and treasure salvage (United States Executive Order 13089, 1998).” Notably, Clinton refers to “coral reef ecosystems,” acknowledging that coral reefs comprise of “a biological network in which the success of species is linked directly or indirectly, through various biological interactions, to the performance of other species in the community (Doney, 2012).” Unlike an exploited fish stock that has the potential to recover under a harvesting moratorium, coral reefs face challenges from a wide range of anthropogenic stressors.
Bellwood et al. 2006, describes coral reefs as “highly dynamic systems characterized by variable and stochastic recruitment and disturbance (Bellwood et al. 2006).” The apprehension concerning these ecosystems, that have persisted for as long as 50 million years, is that coral reefs may be surprisingly stenotopic, or possessing a narrow range of tolerances, in the event of unprecedented environmental change (Norse, 2005). These changes include the highest atmospheric carbon dioxide concentrations seen on Earth for at least the past 650,000 years (NOAA) (see fig. 1), and a projected increase in average sea temperatures surrounding coral reefs of 3–6°C by the end of the century (IPCC 2007). Warming of as little as 1oC can cause coral bleaching, reduced coral growth rates, diminished reproductive outputs, and increased vulnerability to disease (Hoegh-Guldberg, 2008). Thus, the context for endangered coral reef ecosystems is first global and then regional.
Fig. 1) Rise in atmospheric CO2 in industrial times.
The Synergistic Adversities of Climate Change
The IPCC’s 2007, Fourth Assessment Report, states that, “warming of the climate system is unequivocal, as is now evident from observations of increases in global average air and ocean temperatures, widespread melting of snow and ice, and rising global average sea level (Spillman et al. 2011).” Elevated ocean temperatures over prolonged periods of time are the primary cause for coral bleaching, which is the result of corals’ expelling their symbiotic algae from tissues during times of stress (Spillman et al. 2007).
Most IPCC scenarios predict a sea surface temperature rise of at least 2°C in the twenty-first century, a figure that is likely to surpass the majority of corals’ ability to recover from repetitive occasions of acute natural and anthropogenic stress. The intensity and scale of recorded bleaching events has increased since the 1960s, and major events in 1997-98, 2002, 2005, and 2010 have impacted entire coral reef systems (Spillman et al. 2007) (see Fig. 2).
Fig. 2) Major coral bleaching events since the1983-84 event with SST’s from Tahiti.
Fig.3) The above diagram shows where temperatures exceeded coral bleaching thresh-holds on August 27, 2005. Yellow to orange colors indicate areas above the bleaching threshold (NOAA), with red areas indicating in the Caribbean indicating certain bleaching events.
The effects of climate change are not completely uniform on a global scale, but they are working synergistically in opposition of global coral health (see Fig. 3). For example, sea level rise in conjunction with anthropogenic restrictions on water quality has the potential to doom any shallow water coral reef by limiting access to light. A warming climate is also likely to increase natural disturbances, that can have acute and/or chronic impacts on coral reefs, such as ENSO events, and cyclones enabled by warm sea surface temperatures (Spillman et. al. 2011).
Perhaps most threateningly, climate change is associated with carbonate ion concentrations that are beyond those previously experienced in at least the past 420,000 years (Hoegh-Guldberg et al. 2007) (see Fig. 4).
Fig. 4) The above diagram shows the chemical changes in ocean water resulting from increased atmospheric carbon dioxide.
These projected changes in ocean acidification will continue to affect growth, calcification, and a range of other physiological processes in corals and in other associated organisms (Fabry et al. 2007). Importantly, acidification will limit corals’ ability to overcome temperature thresholds, keep up with rising sea levels through accretion, and to meet potentially increased rates of bio-erosion resulting from changing coral reef community compositions (Hoegh-Guldberg, 2008).
Corals’ uniform dependence on endodermal symbiotic dinoflagellates for energy, meanwhile, can be both a disadvantage and a means of adaptation. Some scientists’ optimism is a result of the zooxanthellate genus Symbiodinium’s inherent genetic diversity.
Fig. 5) The above image shows actual zooxanthellae algal cells.
Although one algal type, or clade, is usually dominant in any given species and environment, many corals are flexible in their hosting of different clades. In other words, corals have the capacity to modify their symbiont communities in response to environmental changes, as witnessed in the Indo-Pacific and Western Indian oceans by Baker et al. 2004.
This study found that one out of three genetically unique algal clades was prevalent in affected areas after severe bleaching events. Clade D out competed other clades by more than 50% in areas where it was previously recessive in abundance. Bleaching can thus be seen as a facilitating process for coral resilience, and evidence to corals’ adaptive potential (Baker et al., 2004). Nevertheless, these observations are recent and the long-term resiliency of the coral-algal partnership with clade D has not yet been tested. Any adaptation, no matter how effective, is only temporary.
Moreover, different coral species’ inherent traits, such as the rate at which they generate individuals and replicate genetically identical populations through clonal propagation, will make them more or less competitive as temperatures rise. Community compositions are thus likely to change together with the climate, creating more homogeneous populations (Lasker & Coffroth, 1999). These groups of genets are at greater risk from stochastic events and further stress (Tkachenko, 2007). Community composition changes also have more complex consequences, such as alterations of chemical cues that corals and other species use to settle successfully (Lasker & Coffroth, 1999).
Corals’ relatively long generation times pose a further obstacle to their adaptive evolution. This is not the case for all coral species, but Hoegh-Guldberg, 2008, argue that most corals have long generation times; more than 3 years for the genus Acropora, between 33 and 37 years for a large number of other genera, and of over 100 years for Porites spp. (Hoegh-Guldberg, 2008). It is also probable that reproductive condition, and hence gamete output, is highly dependent on coral colony size, thus further impeding rapid evolutionary change in the event of patchy populations prone to the Allee Effect. Moreover, ten generations are apparently required to significantly shift the pattern of phenotypic expression of a single genotype across a range of environments (Skelly et al. 2007 via Hoegh-Guldberg, 2008). These traits, together with the high probability of enduring global warming, suggests that previous and contemporary geographic shifts of thermally tolerant coral genes cannot keep pace with the rapid rates of climate change (Hoegh-Guldberg, 2008).
In the event that certain Scelractinian, or reef building, coral species do survive temperature changes, these corals still require certain concentrations of CaCO3 compounds, like Aragonite, in order to build their skeletal frameworks (Bauman et al., 2010). The saturation level of Aragonite is decreasing as CO2 levels rise, and the acidity of the water increases, slowing coral accretion rates and enabling more rapid and widespread break-down by mechanical and chemical bio-eroding organisms, and by storm surges and human disturbances such as boat groundings. One prediction states that by 2030, fewer than half the world’s reefs are projected to be in areas where Aragonite levels are ideal for coral growth, and by 2050, only about fifteen percent (Bauman et al., 2010)(see Fig. 6).
Fig. 6) In the above images, Ken Caldeira, from Stanford University, modeled what would happen to the Aragonite saturation as you decreased or increased CO2, from what it is today at 450ppm. Parts of the GBR are currently below presumed viable levels for ideal growth.
These figures are supported by in-situ observations over recent decades in the Great Barrier Reef of Australia and in Thailand, where corals have shown an abrupt downturn in calcification, as well as in linear extension (Hoegh-Guldberg, 2008). Moreover, recent research shows that ocean acidification causes significant decreases in productivity (Hoegh-Guldberg, 2008). This research is crucial in gauging a coral reef’s energetic and evolutionary capacity to adapt, for example, to a minimum projected change in sea temperatures of 3°C over the next 100 years. Under this scenario, Hoegh-Guldberg, 2008, believe ￼that considerable and rapid movement of more temperature-tolerant genotypes, over distances of over 20 km, are needed annually, and that the likelihood of this occurring is highly improbable. The fact that higher latitudes, where surviving corals will presumably settle, are associated with decreases in carbonate ion concentrations due to upwards spirals of atmospheric carbon dioxide, casts further doubt on the likelihood of a successful pole-ward migration (Hoegh-Guldberg, 2008). The same authors believe it is unlikely that coral reefs will survive predicted atmospheric carbon dioxide levels of 600–1,000 ppm, making the danger of acidification perhaps the most dire of all global and regional threats (Hoegh-Guldberg, 2008).
The Coral Holobiont
When assessing the threat of temperature stress and acidification, we must consider that corals are “holobionts,” or whole entities composed of many synchronous, living parts. Some of these parts are indispensable, such as the calcifying coralline algae, which is among the most sensitive of marine organisms to the disabling effects of acidification (Diaz-Pulido, 2012). This functional group enables coral reef construction by depositing calcium carbonate in the form of calcite, and by binding adjacent substrata like biological cement. Coralline algae also induce larval settlement, not only for corals, but for keystone invertebrates such as sea urchins as well (Diaz-Pulido, 2012). Others of these parts are internal, and include marine Bacteria, Archaea, and Eukaryotes, locked into mutualisms with the invertebrate coral structure. Still others, such as bio-eroding fungi and pathogens, are in direct contest with coral growth and/or health (Rosenberg et al. 2007).
Although coral bleaching can be considered the most serious, global coral disease, more than twenty specific coral diseases have been described, and only six of them traced to an isolated causative agent (Rosenberg et al. 2007). On a regional basis, corals suffer greater damage from chronic disturbances and stresses such as disease outbreaks, than from one-off events such as storms (Aronson & Precht, 2001). Furthermore, Bythell et al. 2000, reported that, proceeding a mass disease event, sites in the same locality that were destroyed by disease showed little or no recovery compared with sites less than a kilometer away that were devastated by a hurricane and were recovering well (Bythell et al. 2000 via Tkachenko, 2007). Emerging coral diseases, likely enabled by climate change, may be primarily responsible for the approximately 30% global mortality of coral populations in recent times (Rosenberg et al. 2007).
At Carrie Bow Cay in the Belizean Barrier Reef, for example, white-band disease (WBD) was the principal cause of A. cervicornis mortality in the spur-and-groove zone in the late 1980s, and has since nearly eliminated A. cervicornis from the physically protected shelf lagoon (Aronson & Precht, 1997,2001). Regional mass mortality of Acropora spp. on this scale has not occurred for millennia, as has been deduced from core sampling and paleontological evidence, leading to the conclusion that there is an unprecedented enabler for WBD (Aronson & Precht, 1997,2001). Although poor water quality and nutrient pollution have previously been associated with coral diseases (Bruckner, 1997), the fact that in this case reefs both near and far from human population centers have been affected equally, suggests that climate change is the likely catalyst (Aronson & Precht, 2001). The well-studied bleaching events of O. patagonica in the Mediterranean Sea and of Pocillopora damicornis in the Indian Ocean and Red-Sea, support this conclusion. In these cases, microbial analyses proved that bacterial infections, Vibrio shiloi in the case of O. patagonica, and Vibrio coralliilyticus in the Pocillopora, occurred in the hottest time of year, post-coral bleaching (Rosenberg et al. 2007).
Altering Coral Reef Fish Communities
Much of the scientific focus on coral reefs has been directed at bottom-up processes (Norse, 2005). It is likely, however, that direct and cascading top-down effects pose an equal threat to coral reefs by removing functional groups from the food chain (Campbell & Pardede, 2005). In early studies of coral reef communities, an inverted food pyramid was apparent, with three to four times as many carnivorous species as herbivores and even less primary producers. These findings, even though they were observed on already over-fished reefs, suggest that cross habitat exchanges of energy are important, and support conservation of mangroves and sea grass beds that act as foraging grounds and nurseries for reef species (Campbell & Pardede, 2005).
Perhaps the biggest regional concern for the health of coral reefs occurs when this pyramid is over-turned, and changes such as an increase in macro-algal cover and a decline in predation of sea urchins are observed. This can occur as a result of relatively low and consistent fishing pressures as well as intense overfishing (Campbell & Pardede, 2005). In the recent history of coral reef “phase shifts” to alternate stable states from which reefs do not recover, we have learned that reefs may appear healthy long after fishable biomass has been reduced to the point where ecosystem function is compromised. The tipping point is likely to be different for every reef, depending on the functional redundancy of its faunal community and on the anthropogenic factors that undermine its resiliency (Valentine et Heck Jr., 2005). Nevertheless, regardless of the time frame and the displacing species, be it anemones, sponges, algae, urchins or homogenous and less three-dimensional corals, the outcome is never a more biologically diverse coral community. Since the establishment of the GBRMP in 1975, for example, the biomass of targeted reef fishes has been reduced by up to 60% outside no fishing areas, causing substantial changes in the abundance of their prey. For example, coral cover has significantly declined over the last forty years, largely due to three successive major outbreaks of the voracious echinoderm Acanthaster planci (Crown of Thorns) since the 1960’s, and two major bleaching events in 1998 and 2000. In 2003, more than half the reefs sampled by Bellwood et al. 2004, had less than 10% cover, reflecting
“marked demographic changes, reduced reproductive output of brood stocks, lower rates of recruitment, impaired connectivity, and species-level changes in coral composition, for example in favor of pioneering weedy taxa (Bellwood et al. 2004).”
A coral reefs’ inability to recover is in part due to positive re-enforcement, which can be physical, in terms of a compromised substrate such as a shifting rubble field unsuitable for larvae, biological, in the event of other species out-competing corals and their recruitments for space, food, or minerals, or anthropogenic, in collaboration with fishing pressures (Valentine et Heck Jr., 2005). McClanahan et al. 1995, revealed a biological example in the Caribbean, when intense harvesting of key-stone “inertivore” Trigger Fishes (Balistidae spp), directly caused a boom in their sea urchin (Diadema antillarum) prey. This raised the rates of mechanical bio-erosion by urchins, and promoted large-scale habitat changes as sea-grasses replaced the algae that the sea urchins ate, as well as the eroded corals (Valentine & Heck, 2005).
In 1983, it is believed that an extensive epidemic caused a mass-mortality of the Diadema antillarium urchin, which had already adopted the functionality of over-fished herbivores. This event caused a mass mortality of corals by sending many reefs into a macro-algae dominated phase shift. In a vicious cycle, spatial and/or temporal areas of high fishing risk are less likely to be grazed, leading to more competitive macro-algae dominating corals, while areas of lower risk are at elevated risk of disproportionately high grazing intensity, which raises the stakes of bio-erosion and prevents successful coral recruitment (Madin et al. 2010).
In the heart of the Coral Triangle, the most bio-diverse reef region in the world, home to one third of the world’s remaining reefs (NOAA), a variety of gears are used to fish on coral reefs. These include hook and line, spear guns, hand spears, traps, poisons, and a range of netting practices including gillnets, Ambai and the notoriously dangerous Muro-ami netting. Each method, aside from some of them being innately caustic, targets different trophic levels, as well as working synergistically with the others to reduce biodiversity (Campbell & Pardede, 2005).
Observed variation in the Western Pacific, in the numbers and efforts of fishermen over different spatial scales, fishing methods, and gear types, can be translated to a global scale. In other words, local reefs are often characterized by local and regional threats. Destructive fishing practices and trade in coral organisms, regardless of the current locality or scale of the threat, will be exacerbated by rapid human population growth, the growing demand for fishery resources, the use of more efficient and/or destructive fishery technologies, inadequate management and enforcement, and the afore mentioned affects of climate change (NOAA). Fishing pressure may thus be the most pervasive and immediate regional threat to coral communities, currently affecting more than 55% of the world’s reefs (Bauman et al., 2010).
Acute and Chronic Anthropogenic impacts to Reefs
Acute anthropogenic threats to coral reefs normally occur in underdeveloped countries, home to the majority of the world’s reefs, and include mining for limestone, dumping of mine tailings, destructive fishing practices, and land reclamation. Regional and global threats involve more chronic disturbances, largely in the form of run-off from adjacent river catchments containing harmful sediments, nutrients and pesticides. Unlike the more acute impacts, these pressures may not cause immediate mortality to reefs. Nevertheless, by jeopardizing a coral reefs’ key parameters, water and substratum quality, they limit recovery from acute impacts, recruitment from healthier reefs, and settlement, thus enabling phase shifts (Wolanski et al. 2003).
For example, tropical harmful algal blooms (HAB’s) are increasing in frequency and intensity as a result of climate change, and are having substantial impacts on coral reef communities (Bauman et al., 2010). One such large-scale HAB event in October-November of 2008, involving the dinoflagellate Cochlodinium polykrikoide, affected over 500 square km in the Gulf of Oman, causing the complete loss of the branching corals and substantial reductions in the abundance, richness and trophic diversity of local reef fish communities (Bauman et al., 2010). Biological detective work is sometimes needed to discover the proximate causes of these HAB’s, as was the case in a mysterious 1999 coral and fish die off linked to a period of cooler water, that affected 400km of reef in the Mentawai Islands in Indonesia.
By counting back coral growth bands, and matching the coral record of ocean temperature to instrumental records, the reef death was dated to December 1997, and was moreover established as an unprecedented event in the last 7,000 years. The year of greatest mortality, 1997, saw a robust drought in an over-populated area, resulting in massive wildfires and a simultaneous spike in marine productivity. This led scientists to believe that a massive ash-fallout fertilized the sea around Mentawai and triggered the lethal bloom. This example emphasizes how increased human populations in combination with anthropogenic climate changes can literally and metaphorically set fire to our most valuable marine resources (Nerilie, 2004).
Turbulent Times for our World’s Reefs
The water surrounding coral reefs naturally contains a matrix of suspended matter, or “marine snow,” that includes calcareous, fecal and organic detrital material as well as mucus secreted by plankton, algae, bacteria, and the corals themselves (Wolanski et al. 2003). These aggregates are almost neutrally buoyant, and remain in suspension for hours in turbulent reef waters, rapidly aggregating via mucus and flocculation. These marine snow “flocs,” become bigger and more prevalent in waters with excess anthropogenic nitrogen and phosphorous (Szmant 2002), and can exceed several centimeters in places like the Great Barrier Reef. Additional sediments weigh the flocs down and cause them to settle at rates of around five centimeters per minute, with high siltation levels of around four to five milligrams per square centimeter. Any flocs with sediment contents of over 0.5 milligrams per square centimeter will suffocate and kill coral polyps within an hour (Wolanski et al. 2003), with an increased effect on younger coral recruits (McKenna et al. 2001). Mortality of juvenile corals is ten times greater than that of adults, and in Australia’s nutrient-enriched coastal waters they will die within 43 hours of exposure to muddy marine snow (Wolanski et al. 2003).
Fortunately, settled muddy marine snow tends to be short-lived due to natural flushing (Wolanski et al. 2003), however, it can also be frequent with increasing floods, areas of large tidal ranges and in underdeveloped countries where sewage commonly flows directly into the sea (Praveena et al. 2012). In Port Dickson, western Malaysia, for example, sedimentation rates of 76.83 mg/cm2/day at 3 m depth were observed (Praveena et al. 2012). High turbidity also has other negative associations. These include the influx of pollutants from fields or from dredging, and an inverse relationship to the prevalence of herbivorous fish, (Wolanski et al. 2003) which have been shown to aid coral growth and recovery by grazing on filamentous algae (Papasian, 2007).
Meanwhile, pesticides, heavy metals and hydrocarbons further interfere with chemically sensitive processes that are still little understood, such as broadcast spawning synchrony, egg-sperm interactions, fertilization, embryological development, larval settlement and metamorphosis and acquisition of zooxanthellae by juvenile corals (Wolanski et al. 2003).
Coral reefs have all been affected by natural disturbances, such as lava flows, storms, and floods, for millions of years, much like terrestrial forests experience acute fires that help restore nutrients to the soil and re-instate biodiversity with pioneering species. It is the synergistic and cumulative effects of human disturbances, from tourists walking on reefs in Malaysia (Praveena et al. 2012), to bombing thirty square meters with homemade, kerosene-fertilizer soda bottle bombs in Palau Tiga (Praveena et al. 2012), to leveling entire deep water reefs with rock-hopper trawling gear in Norway (Fossa & Furevik, 2002), to the changing chemistry and temperature of the ocean, that overwhelm coral reefs.
In some cases, the final result will be a less desirable stable state that will last indefinitely, such as occurred on the reef at Discovery Bay, Jamaica in the 1980’s. Hurricane Allen principally ravished the dominant Acropora species, causing collateral mortality of corallivorous fish and invertebrates, which when combined with reduced herbivory from a long history of overfishing, and the Caribbean-wide mass mortality of the grazing echinoid Diadema antillarum, transformed the coral reef into a macro-algae dominated field of rubble (Aronson & Precht, 2001). This exemplary scenario, combining climate change, over-fishing, disease, and positive re-enforcement of a phase shift, may soon be similarly played out in many of the world’s coral reefs if management efforts fail to intervene. Moreover, if climate change is not curbed, even healthy and resilient reefs will dissolve and a new and poorer landscape will replace them.