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REVIEW article

Coral reef ecosystems under climate change and ocean acidification.

• 1 The Global Change Institute, The University of Queensland, St Lucia, QLD, Australia
• 2 ARC Centre for Excellence in Coral Reef Studies, The University of Queensland, St Lucia, QLD, Australia
• 3 School of Biological Sciences, The University of Queensland, St Lucia, QLD, Australia
• 4 CSIRO Oceans and Atmosphere, Queensland Biosciences Precinct, St Lucia, QLD, Australia
• 5 Coral Reef Watch, National Oceanic and Atmospheric Administration (NOAA), College Park, MD, United States

Coral reefs are found in a wide range of environments, where they provide food and habitat to a large range of organisms as well as providing many other ecological goods and services. Warm-water coral reefs, for example, occupy shallow sunlit, warm, and alkaline waters in order to grow and calcify at the high rates necessary to build and maintain their calcium carbonate structures. At deeper locations (40–150 m), “mesophotic” (low light) coral reefs accumulate calcium carbonate at much lower rates (if at all in some cases) yet remain important as habitat for a wide range of organisms, including those important for fisheries. Finally, even deeper, down to 2,000 m or more, the so-called “cold-water” coral reefs are found in the dark depths. Despite their importance, coral reefs are facing significant challenges from human activities including pollution, over-harvesting, physical destruction, and climate change. In the latter case, even lower greenhouse gas emission scenarios (such as Representative Concentration Pathway RCP 4.5) are likely drive the elimination of most warm-water coral reefs by 2040–2050. Cold-water corals are also threatened by warming temperatures and ocean acidification although evidence of the direct effect of climate change is less clear. Evidence that coral reefs can adapt at rates which are sufficient for them to keep up with rapid ocean warming and acidification is minimal, especially given that corals are long-lived and hence have slow rates of evolution. Conclusions that coral reefs will migrate to higher latitudes as they warm are equally unfounded, with the observations of tropical species appearing at high latitudes “necessary but not sufficient” evidence that entire coral reef ecosystems are shifting. On the contrary, coral reefs are likely to degrade rapidly over the next 20 years, presenting fundamental challenges for the 500 million people who derive food, income, coastal protection, and a range of other services from coral reefs. Unless rapid advances to the goals of the Paris Climate Change Agreement occur over the next decade, hundreds of millions of people are likely to face increasing amounts of poverty and social disruption, and, in some cases, regional insecurity.

Introduction

Both warm- and cold-water corals secrete calcium carbonate skeletons that build up over time to create a three-dimensional reef matrix that provides habitat for thousands of fish and other species. The production of limestone-like calcium carbonate is high enough in many warm-water coral reefs to establish carbonate structures. High rates of calcification are sufficient to overcome significant rates of bioerosion and wave driven physical erosion. These structures underpin the framework of barrier reefs and islands, which are critically important to tropical coastlines. Although they occupy less than 0.1% of the ocean floor, tropical coral reef ecosystems provide habitat for at least 25% of known marine species, with many reef species still to be discovered ( Fisher et al., 2015 ). The biological diversity of warm-water coral reefs has been estimated to include ~1–9 million species that live in and around coral reefs ( Reaka-Kudla, 1997 , Census of Marine Life, http://www.coml.org/census-coral-reef-ecosystems-creefs ). In deeper parts of these warm-water reef systems, the tendency toward carbonate dominated reef structures diminishes as light levels decrease ( Bongaerts et al., 2010a ). At low light levels, erosion and dissolution exceed calcium carbonate production, leading to coral communities that may be abundant yet with little or no three-dimensional calcium carbonate reef framework. Extending from 40 to 150 m, these “mesophotic” (low light) coral reefs also provide extensive habitat, with the rates of discovery of species remaining very high due to these reefs being difficult to visit ( Bongaerts et al., 2010a , 2011 ). Mesophotic reef systems probably cover a comparable area to shallow warm-water coral reefs ( Bongaerts et al., 2010a ; Slattery et al., 2011 ).

Both shallow or deeper mesophotic coral reefs are dominated by scleractinian corals that form symbiosis with dinoflagellate protists from the genus, Symbiodinium . On the basis of this symbiosis, their intracellular symbionts (i.e., living within the gastrodermal or digestive tissues of their coral hosts) are able to photosynthesize and provide the host coral with a rich source of sugars, glycerol, lipids, and other organic compounds ( Muscatine, 1990 ). This relationship enables corals to grow and calcify at high rates in the clear, warm, and shallow water conditions along tropical coastlines ( Muscatine and Porter, 1977 ). The abundance of Scleractinian corals hosting Symbiodinium decreases with depth beyond 20–40 m, depending on the clarity of the water column. The deepest Scleractinian corals that are symbiotic with Symbiodinium , are found 100 m or more below the surface of tropical waters ( Englebert et al., 2014 ). The productivity of this symbiosis is complemented by the ability of corals to capture and feed on waterborne particles and plankton (i.e., polytrophy). The combined ability to photosynthesise, as well as feed, underpins the success of the highly productive coral reef ecosystems that line many tropical coastlines. Evidence from isotope signatures within fossils reveal that Scleractinian corals have been symbiotic with Symbiodinium for over 230 million years ( Stanley and Fautin, 2001 ; Muscatine et al., 2005 ), most probably driving productive and diverse ecosystems that were not too different from those of today.

Cold-water coral reefs extend to depths of 3,000 m although some cold-water corals can be found growing in waters as shallow as 50 m (e.g., Norwegian shelf). Below 200 m depth there is so little light that photosynthesis is no longer possible. As a result, cold-water corals do not form a symbiosis with Symbiodinium and depend instead on particle feeding. Discoveries of the locations and extent of cold-water reefs has primarily been driven by advances in underwater technologies for surveying and mapping ( Turley et al., 2007 ; Ramirez-Llodra et al., 2010 ). For example, vast extents (~2,000 km 2 ) of cold-water coral reefs, some shown to be thousands of years old (>8,000 years), have been found in Norwegian waters in past decades ( Fosså et al., 2005 ). Cold-water coral reefs have now been discovered in every ocean, forming important assemblages within the deep ocean that provide critical habitat to thousands of other species, including many commercially important species.

Human communities derive many benefits from coral reefs including food, income, recreation, coastal protection, cultural settings, and many other ecological goods and services ( Cinner et al., 2009 ; Costanza et al., 2014 ). Despite their biological diversity, productivity and importance to humans, both warm and cold-water coral reefs are being heavily impacted by human activities due to both local and global influences ( Hall-Spencer et al., 2002 ; Burke et al., 2011 ). As a result, many coral reefs are rapidly declining across the world. While local factors can have significant impact on coral reefs (e.g., pollution, overfishing, and the physical destruction of reefs), changes in ocean temperature and chemistry due to anthropogenic activities are dramatically reducing the distribution, abundance, and survival of entire coral reef ecosystems ( Gattuso et al., 2014b ; Hoegh-Guldberg et al., 2014 ). Given these risks and the importance of coral reefs to humans and marine biodiversity, the present paper focuses on the challenges that warm and cold-water coral reef ecosystems and their human communities are facing, particularly those posed by rapidly warming and acidifying oceans.

Distribution, Abundance, and Importance of Coral Reef Ecosystems

Warm-water coral reefs are prominent ecosystems within coastal areas of the Pacific, Indian, and Atlantic oceans (Figures 1A,B ), where they are typically found in a broad band (30°S to 30°N) of warm, sunlit, alkaline, clear, and relatively nutrient deficient ocean waters ( Kleypas et al., 1999b ). Here, Scleractinian or reef-building corals proliferate, depositing copious amounts of calcium carbonate. As corals die, their dead skeletons build up over time and are “glued” together by the activities of other organisms such as encrusting red coralline algae ( Glynn and Manzello, 2015 ). Other organisms such as calcifying green algae, invertebrates, and phytoplankton also contribute to the overall carbonate budget of warm water coral reefs ( Hutchings and Hoegh-Guldberg, 2009 ), leading to three-dimensional calcium carbonate structures that build up over hundreds and thousands of years. In turn, the three-dimensional structures (Figure 1C ) within warm-water reef systems creates habitat for hundreds of thousands of species, many of which support coastal human populations with food, income, and other ecological goods and services such as coastal protection. Coral reefs are also important sources for bio-prospecting and the development of novel pharmaceuticals. The asset value of coral reefs has been estimated as close to $1 trillion ( Hoegh-Guldberg, 2015 ) with the economic value of goods and services from coral reefs exceeding $375 billion annually, with benefits flowing to over 500 million people in at least 90 countries worldwide ( Burke et al., 2011 ; Gattuso et al., 2014b ).

Figure 1. (A) Distribution of warm-water and cold-water coral reefs (credit: Hugo Ahlenius, 2008, UNEP/GRID-Arendal, http://www.grida.no/resources/7197 ). (B) Location of warm-water coral reef cells and provinces, from Hoegh-Guldberg et al. (2014) . (C) Warm-water carbonate coral reef from the Great Barrier Reef, Australia (credit: Ove Hoegh-Guldberg). (D) Mesophotic coral community of North Sulawesi, Indonesia. (Credit: Pim Bongaerts, University of Queensland). (E) Deep-water community of Lophelia pertusa from the Mississippi Canyon at ~450 m depth (Image from NOAA, licensed under the Creative Commons Attribution-Share Alike 2.0 Generic license).

As light levels decrease with depth, decalcification dominates and the overall carbonate balance of reef ecosystems shifts to negative ( Barnes and Chalker, 1990 ; Bongaerts et al., 2010a ). Under these conditions, Scleractinian corals and their symbionts persist with reefs being referred to as “mesophotic” ( Bongaerts et al., 2010a , 2011 ; Robinson C. et al., 2010 ). In these habitats, colonies of Scleractinian corals are often platelike in shape, orientating themselves to maximize light harvesting under these dim light conditions (Figure 1D ). Mesophotic reef systems are also primarily restricted to areas where water clarity, carbonate ion concentrations, and temperatures are relatively high. Like their counterparts in shallower regions, mesophotic reef systems play an important role in supporting fisheries and hence human livelihoods. Given the difficulty of working at depths of more than 30 m (beyond SCUBA-diving depth) many species remain to be discovered ( Bongaerts et al., 2010a ). Mesophotic reefs therefore have an unknown potential to be sources of novel pharmaceuticals and other potentially beneficial compounds ( Leal et al., 2012 ). As a result, their true value has probably been underestimated.

Cold-water corals generally form reefs at much greater depths from 200 to 2,000 m however in some regions they are found at shallower depths ( Fosså et al., 2002 ; Freiwald et al., 2004 ). Deep-water corals are not dependent on light levels as they are not symbiotic with Symbiodinium . Due to the colder and more CO 2 rich (and hence less alkaline) waters, deep-water corals grow slower than warm-water corals, forming aggregations that are variously termed patches, banks, thickets, bioherms, mounds, gardens, and massifs. In the absence of significant wave action, these fragile and slow growing reefs form aggregations that can cover vast tracks of the seabed (e.g., 2,000 km 2 in Norwegian waters http://www.lophelia.org/ ) ( Hall-Spencer et al., 2002 ) and involve near mono-specific stands of Scleractinian corals such as Lophelia pertusa and Oculina varicosa (Figure 1E ). In addition to Scleractinian corals, they often exhibit a wide variety of abundant coral-like organisms, including soft corals, gorgonians, and Alcyonaceans.

Recent Changes in the Extent of Anthropogenic Stresses on Coral Reef Ecosystems

Coral reefs are facing growing challenges from the local to global effects of human activities. Over the past 200 years, human activities have fundamentally changed coastlines, overexploited resources such as fish stocks, and polluted coastal waters, to a point where many coral reef ecosystems are degrading rapidly ( Jackson et al., 2001 ; Pandolfi et al., 2003 ; Hoegh-Guldberg, 2014b ). Warm-water coral reefs, for example, have declined by at least 50% over the past 30–50 years in large parts of the world's tropical regions ( Hughes, 1994 ; Gardner et al., 2003 ; Bruno and Selig, 2007 ; De'ath et al., 2012 ). Similar conclusions have been reached for cold-water reefs where human activities have put these systems under escalating pressure from the mid-1980s onwards. Key drivers of the destruction of cold-water reefs include commercial bottom trawling, hydrocarbon exploration and production, deep sea mining, cable and pipeline placement, pollution, waste disposal, coral exploitation, and trade, and destructive scientific sampling ( Hall-Spencer et al., 2002 ; Turley et al., 2007 ; Roberts and Cairns, 2014 ). The increase in impacts from human activities is a result of rapid advances in technologies for visualizing and exploiting the biological and mineral resources of deep water habitats ( Freiwald et al., 2004 ; Ramirez-Llodra et al., 2010 ). Many populations of deep-sea corals (Scleractinians, gorgonians) have very slow turn-over rates and may live for centuries, with some species such as black corals (Antipatharians) living for thousands of years. The longevity and slow growth rates of these taxa means that recovery from anthropogenic stressors will be very slow. The areas inhabited by the deep-sea reefs are also a “resource frontier” for hydrocarbon extraction and mining of high value and “high-tech” metals ( Roberts and Cairns, 2014 ). Hence, it is likely that anthropogenic impacts on these reefs will expand. These impacts are also likely to interact with ocean warming and acidification (Figure 2A ), which pose growing and serious risks to coral reef ecosystems on their own. The direct impact of these changes to coral reefs have been growing since the early 1980s ( Hoegh-Guldberg et al., 2007 , 2014 ; Eakin C. M. et al., 2010 ; Gattuso et al., 2014b ). The latter are the direct result of the burning of fossil fuels and have been driving growing impacts on warm water coral reefs since the early 1980s. Understanding and solving both local and global threats to coral reefs will be critically important if they are to survive some of the greatest rates of environmental change in the recent history of the Earth ( Hönisch et al., 2012 ; Pörtner et al., 2014 ).

Figure 2. (A) Linkages between the build-up of atmospheric CO 2 and the slowing of coral calcification due to ocean acidification. Approximately 30% of the atmospheric CO 2 emitted by humans has been taken up by the ocean (IPCC, 2013) where it combined with water to produce carbonic acid, which releases a proton that combines with a carbonate ion. This decreases the concentration of carbonate, making it unavailable to marine calcifiers such as corals. (B) Temperature, [CO 2 ] atm , and carbonate-ion concentrations reconstructed for the past 420,000 years. Carbonate concentrations were calculated ( Lewis et al., 1998 ) from [CO 2 ] atm and temperature deviations from conditions in the decade of the 2000s with the Vostok Ice Core data set ( Petit et al., 1999 ), assuming constant salinity (34 parts per trillion), mean sea temperature (25°C), and total alkalinity (2,300 mmol kg −1 ). Acidity of the ocean varies by ± 0.1 pH units over the past 420,000 years (individual values not shown). The thresholds for major changes to coral communities are indicated for thermal stress (+2°C) and carbonate-ion concentrations ([carbonate] = 200 μmol kg −1 , approximate aragonite saturation ~Ω aragonite = 3.3; [CO 2 ] atm = 480 ppm). Coral Reef Scenarios CRS-A, CRS-B, and CRS-C are indicated as A, B, and C, respectively, with analogs from extant reefs. Red arrows pointing progressively toward the right-hand top square indicate the pathway that is being followed toward [CO 2 ] atm of more than 500 ppm. From Hoegh-Guldberg et al. (2007) with permission of Science Magazine.

Warm-water coral reefs are largely dependent on the physical and chemical changes occurring in the surface of the ocean, whereas cold-water reef systems are tied relatively more to the broad scale conditions of the bulk ocean ( Freiwald et al., 2004 ; Eakin C. M. et al., 2010 ). In this respect, there are likely to be differences in terms of the rate and characteristics of the changes that are occurring. These differences also translate into different trajectories when it comes to near and long-term projections of planetary warming and ocean acidification.

Warm-water coral reef environments have experienced relatively small amounts of variability in terms of temperature and carbonate ion concentrations, even with the relatively substantial swings in average global temperature and atmospheric CO 2 concentration during the glacial cycle (Figure 2B ). Warm-water coral reefs contracted toward the equator during glacial periods, and re-expanded along the tropical and subtropical coastlines of the world during the intervening warm periods ( Hubbard, 2015 ). While these changes were rapid relative to geological time frames, they occurred over periods of 10,000 years or more and are slow when compared to climatic changes that have occurred since pre-industrial. While our understanding of how conditions have changed in terms of the habitat of deep-water coral reefs over geological time is limited, it is very likely that conditions varied even less over these long periods than those surrounding the warm-water coral reefs.

It is virtually certain that the upper ocean has warmed between 1971 and 2010 and likely that it has warmed between 1870s and 1971 ( IPCC, 2013 ). These changes are consistent with those expected from the associated rise in greenhouse gas concentrations in the atmosphere ( IPCC, 2013 ). The average sea surface temperatures (SST) of the Indian, Atlantic, and Pacific oceans have increased by 0.65, 0.41, and 0.31°C during 1950–2009 (Table 30-1 in Hoegh-Guldberg et al., 2014 ). The influence of long-term patterns of climate variability such as the Pacific Decadal Oscillation (PDO) contribute to variability at regional scales and confound efforts to detect and attribute regional changes to anthropogenic greenhouse gas emissions ( Hoegh-Guldberg et al., 2014 ). Nonetheless, examination of the Hadley Centre HadISST1.1 data ( Rayner et al., 2003 ) over 60 years (1950–2009) reveals significant warming trends in SST for many sub-regions of the ocean (Table 30-1 in Hoegh-Guldberg et al., 2014 ). Significant trends are clearly demonstrated within the six major warm-water coral reef regions, with the exception of the Gulf of Mexico/Caribbean Sea region (Table 1 ). Rates of increase in SST in the warm-water coral reef regions range from 0.07°C (west Pacific Ocean) to 0.13°C (Coral Triangle and southeast Asia) per decade, resulting in an overall increase in the regions of between 0.44 and 0.79°C during the period from 1950 to 2009.

Table 1. Changes in sea surface temperature (SST) in six major warm-water coral reef provinces (Figure 1B ) over the period 1950–2009 using 1 × 1 degree monthly SST data extracted from the Hadley Centre HadISST1.1 data set ( Rayner et al., 2003 ) .

In addition to the heat content and temperature of the upper layers of the ocean, the research community is virtually certain that ocean chemistry is also changing as a result of the increasing amounts of CO 2 entering the Ocean ( IPCC, 2013 ). Observed increases in salinity at tropical latitudes are consistent with the amplification of the global hydrological cycle ( Durack and Wijffels, 2010 ; Durack et al., 2012 ), including rainfall, which have significant implications for coastal ecosystems such as warm-water coral reefs. At regional levels, changes in storm and rainfall intensity also have the potential to influence coastal water quality, which is important to coral reefs, as a result of the interplay between droughts, coastal and catchment erosion, and sudden inundation (flood) events. The impact of climate change adds to those from other human activities that are already impacting water quality, coastal erosion and biological systems.

Average global sea levels are increasing by an average of 3.2 mm year −1 (over 1993–2010) as a result of warming of the ocean (thus increasing volume) and the melting of land ice ( IPCC, 2013 ). Sea level rise varies between regions as a result of differences in local oceanography and geology and the influence of long-term variation in regional climate. Some areas that have significant warm-water coral reefs, such as Southeast Asia and northern Australia, have reported rates of sea level rise of around 10 mm year −1 . While the direct attribution of changes in regional wind strength, storm intensity and frequency to global warming is challenging due to long-term variability, there is considerable evidence that the frequency and intensity of the strongest tropical storms in some regions (e.g., North Atlantic; IPCC, 2013 ) has increased since the 1970s. The combination of higher sea levels and more intense storm systems is likely to increase the amount of force exerted by wave action on coastal areas, which has implications for coastal infrastructure, as well as the state of ecosystems such as coral reefs, mangroves, and seagrass beds ( Hamylton et al., 2013 ; Saunders et al., 2014 ).

Changes have also occurred in the pH of ocean surface waters over the past 100 years, a phenomenon which is referred to as ocean acidification ( Kleypas et al., 1999a ; Caldeira and Wickett, 2003 ; Gattuso et al., 2014a ). As CO 2 enters the ocean, it reacts with water increasing hydrogen ion concentration (thus decreasing ocean pH) and decreasing the carbonate ion concentration. While the overall change in ocean pH appears small (0.1 pH units over the past 150 years), this is actually a 26% increase in the concentration of hydrogen ions. Experimental evidence shows a reduction in carbonate ions with ocean acidification is biologically significant, since it can affect the rate at which marine organisms, such as corals build their calcareous structures ( Kroeker et al., 2013 ). However, understanding of the mechanisms driving the sensitivity of coral calcification to ocean chemistry, such as the response of the pH of the internal calcifying fluid in which the coral skeleton forms to the concentration of dissolved organic carbon, are only being untangled ( Comeau et al., 2017 ). These changes in ocean chemistry are temperature dependent, with the CO 2 absorption and consequently acidification being highest when waters are cooler. The aragonite (one form of calcium carbonate) saturation state (Ω arag ) is essentially the ratio between the concentrations of calcium and carbonate ions ( Doney et al., 2009 ). The aragonite saturation state shows a similar distribution to sea surface temperature with Ω arag being highest in the warmest ocean regions and lowest in polar regions ( Jiang et al., 2015 ). Surface waters of the ocean are generally supersaturated with respect to aragonite (Ω arag > 1). However, in warmer waters where Ω arag is not projected to fall to <1 (thus undersaturated with respect to aragonite, Figure 3 ), substantial impacts are likely to still occur on calcifying organisms. There is substantial evidence that carbonate accretion on warm-water coral reefs approaches zero or becomes negative when Ω arag falls below 3.3 ( Hoegh-Guldberg et al., 2007 ; Chan and Connolly, 2013 ), a level likely to be reached in tropical surface waters within the next few decades at current rates of greenhouse gas emission ( Hoegh-Guldberg et al., 2007 ; Ricke et al., 2013 ).

Figure 3. Aragonite saturation state of the surface ocean simulated by the University of Victoria Earth System Model under different atmospheric concentrations of CO 2 . 280 ppm represents pre-industrial and 394 ppm levels in 2012 . Four hundred and fifty ppm is projected to be reached during 2030s under Representative Concentration Pathway (RCP) 4.5, 6.0, and 8.5, and to approach, but not reach 450 ppm, during 2040s under RCP 2.6 (IPCC 2013). Eight hundred ppm is projected to be reached during 2080s under RCP 8.5 only. Fields are calculated from the model output of dissolved inorganic carbon concentration, alkalinity concentration, temperature, and salinity, together with the chemistry routine from the OCMIP-3 project. Modified from Figure SM30-2 in Hoegh-Guldberg et al. (2014 ; reprinted with permission of IPCC AR5).

The global distribution of cold-water corals is at least partly limited by the depth of the aragonite saturation horizon, Ω arag = 1.0 ( Guinotte et al., 2006) . Aragonite saturation state diminishes with depth, due partly to hydrostatic pressure and lower temperature, with a distinct aragonite “saturation horizon” below which waters become under-saturated for aragonite (Ω arag <1) ( Jiang et al., 2015 ). The saturation horizon is a complex outcome of ocean circulation, temperature, CO 2 concentrations, salinity, metabolic activity, and the concentrations of organic compounds and occurs at depths between 200 and 3,500 m, depending on the latitude and the ocean ( Orr et al., 2005 ; Doney et al., 2009 ; Rhein et al., 2013 ; Jiang et al., 2015 ). Surface waters and waters at 50 m depth are mostly supersaturated throughout the global ocean ( Jiang et al., 2015 ), however in western Arctic waters, the area of under-saturated waters in the upper 250 m north of 70°N has increased from 5 to 31% between 1990s and 2010 ( Qi et al., 2017 ). At 500 m, large areas of undersaturated Ω arag water are found in the northern and equatorial Pacific ocean. At 1,000 m, Ω arag < 1.8 over all ocean basins and at 2.000 m, Ω arag < 1.0 across all the Pacific and Indian Ocean and parts of the Atlantic Ocean. Ocean acidification is proceeding at higher rates at high latitudes than at lower latitudes (Figure 3 ) resulting in a shoaling of the aragonite saturation horizon. There is now evidence to show that the aragonite saturation horizon has shoaled since the Preindustrial Period ( Turley et al., 2007 ). For example, in the north east Pacific (from 33.5 to 50.0°N) the aragonite saturation horizon has shoaled by 19.6 m in 11 years (2001–2012) and, at this rate, the entire water column in the northern section of this region is projected to become undersaturated within 50–90 years ( Chu et al., 2016 ).

Biological Responses to a Rapidly Warming and Acidifying Ocean

Not surprisingly, the scale and pace of the physical and chemical changes occurring in the ocean are driving a large range of fundamental responses in marine organisms, ecosystems, and regions ( Hoegh-Guldberg et al., 2014 ; Pörtner et al., 2014 ). Equally significant, is the observation that relatively small amounts of change have resulted in quite substantial biological impacts, with clear evidence of non-linear trends, tipping points, and otherwise complex responses. Coral responses to changes in ocean conditions, in particular mass coral bleaching, provide particularly compelling examples of the consequences of a rapidly changing ocean for organisms, ecosystems, and dependent societies.

The symbiosis between warm-water corals and Symbiodinium (Figures 4A,B ) is very sensitive to changes in the physical and chemical environment surrounding corals. Short periods of high or low temperature and/or light, or exposure to toxins like cyanide, can drive the breakdown of the symbiosis, resulting in the loss of the brown symbionts and a subsequent paling (hence “bleaching”) of the coral host ( Hoegh-Guldberg, 1999 ). Coral bleaching involves the breakdown of the symbiosis between Scleractinian corals and Symbiodinium , which may recover if conditions are not too anomalous for too long. While bleaching of coral tissues has been reported on the scale of colonies or groups of colonies for at least 100 years ( Yonge and Nichols, 1931 ), reports of bleaching at large geographic scales (Figures 4C,D , example of affected coral reefs in American Samoa from late 2015) was unknown to the scientific literature until 1979. Since the early 1980s, however, mass coral bleaching has affected entire reefs and regions, often resulting in significant mortality of reef-building corals. The absence of pre 1979 scientific reports in addition to the close relationship between bleaching and elevated sea temperature, plus considerable laboratory, and mesocosm studies, strongly support the conclusion that mass coral bleaching and mortality are novel and are caused by warm water coral reefs being exposed to rising sea temperatures ( Hoegh-Guldberg and Smith, 1989 ; Glynn, 1993 , 2012 ; Hoegh-Guldberg, 1999 ; Glynn et al., 2001 ; Hoegh-Guldberg et al., 2007 , 2014 ; Baker et al., 2008 ; Eakin C. M. et al., 2010 ; Strong et al., 2011 ; Gattuso et al., 2014b ). The latest cycle of mass coral bleaching in 2016 ( Hoegh-Guldberg and Ridgway, 2016 ) is reputedly the worst on record and accompanies the warmest years on record ( King and Hawkins, 2016 ; https://www.nasa.gov/press-release/nasa-noaa-data-show-2016-warmest-year-on-record-globally ).

Figure 4. (A) Scleractinian coral ( Turbinaria sp) and (B) Hydrozoan coral ( Millepora sp) showing respective Symbiodinium symbionts (each brown cell is about 10 μm in diameter) removed from coral tissues; Credit for (A,B) : Todd LaJeunesse, from Pennsylvania State University. ( https://www.flickr.com/photos/tags/linkflickrset72157631573740050 ). (C) The photo at left, taken in December 2014, shows coral reef near runway in American Samoa, without obvious bleaching of corals. (D) The photo at right shows the same coral reef, now heavily bleached, in February 2014 (Credit for C , D : Richard Vevers, The Ocean Agency).

Mass coral bleaching and mortality can be triggered by small (1–2°C) SST increases above the long-term summer maxima for a region ( Strong et al., 2011 ). If temperatures are higher for longer, the amount of coral bleaching will increase, driving increased mortality ( Hoegh-Guldberg, 1999 ; Hoegh-Guldberg et al., 2007 ; Eakin C. M. et al., 2010 ). There is a strong link between the size and length of temperature extremes and mass coral bleaching and mortality ( Hoegh-Guldberg, 1999 ; Strong et al., 2004 , 2011 ; Eakin C. M. et al., 2010 ). These relationships are used with satellite data to derive anomalies in SST to monitor the frequency and intensity of mass coral bleaching and mortality ( Strong et al., 2004 , 2011 ). For this reason, there is a high level of confidence that the increases in mass coral bleaching and mortality since the early 1980s are due to anthropogenic climate change in particular ocean warming ( Hoegh-Guldberg et al., 2014 ). The loss of symbionts from coral tissues can have immediate effects through the loss of photosynthetic energy, and lead to starvation, disease, reproductive failure, and a loss of competitive ability relative to other organisms on coral reefs ( Hoegh-Guldberg and Smith, 1989 ; Glynn, 1993 , 2012 ; Hoegh-Guldberg, 1999 ; Baker et al., 2008 ; Hoegh-Guldberg et al., 2014 ; Glynn and Manzello, 2015 ).

Understanding how the positions of ocean isotherms (lines of similar temperatures) are changing and how fast across the ocean surface (“velocity of climate change”, Burrows et al., 2011 , 2014 ) provides insight into whether or not coral populations will be able to move, adapt or acclimatize fast enough to changing sea temperatures ( Hoegh-Guldberg, 2012 ; Pörtner et al., 2014 ). Some of the highest rates of climate velocity (up to 200 km per decade) were observed in ocean tropical regions (over 1960–2010), driven by shallow spatial gradients in temperature ( Burrows et al., 2011 , 2014 ). Observed rates of distribution shifts for individual warm-water coral species linked to increases in sea surface temperatures range from 0 to 150 km per decade, with an average shift rate of 30 km per decade ( Yamano et al., 2011 ; Poloczanska et al., 2013 ), suggesting that corals and coral ecosystems may be unable to keep up with warming rates ( Hoegh-Guldberg, 2012 ; Burrows et al., 2014 ; García Molinos et al., 2015 ).

The possible reduced influence of extremes from climate change with depth has led to the speculation that deeper (>40 m) mesophotic coral reefs may offer a potential refuge against the otherwise rapid changes in temperature, storm intensity, and chemistry that are typical of shallow-water (0–30 m) coral reef environments ( Bongaerts et al., 2010a ). The “Deep Reef Refugia” hypothesis has been explored by a number of groups who are finding substantial differences in terms of the rate of warming and acidification with depth, as well as examples of species that may span the mesophotic zone to shallow reef areas. Recent work however, has revealed that mesophotic reefs may not be immune to the impacts of storms ( Bongaerts et al., 2013 ). Also, populations of what appear to be the same coral species appear to have considerable genetic structure as a function of depth. This is important given that it implies a high degree of specialization, local adaptation, and even speciation, by corals living at different depths, with the implication that mesophotic corals may not be able to survive in shallow-water environments and vice versa , reducing the potential for mesophotic environments to provide refugia for shallow water Scleractinian corals. This reduces the significance of deeper water populations as a source of recruits for regenerating damaged areas on shallow water coral reefs ( Bongaerts et al., 2010b , 2015 ). In addition to warming oceans, corals are also sensitive to changes to the pH and the carbonate chemistry of seawater as a result of ocean acidification ( Kleypas et al., 1999a ; Gattuso et al., 2014a ). These changes affect organisms in a variety of ways, including reducing calcification rates in a wide array of corals and other organisms in laboratory, mesocosm, and field studies ( Gattuso et al., 1998 ; Reynaud et al., 2003 ; Kleypas et al., 2006 ; Dove et al., 2013 ; Kroeker et al., 2013 ; Gattuso et al., 2014a ).

Long-lived corals from the field have provided an opportunity for retrospective analysis of how growth has varied over long periods of time ( De'ath et al., 2009 ; Lough, 2010 , 2011 ). Calcification measurements from coral cores from 328 colonies of the massive coral Porites growing on the Great Barrier Reef in Australia, for example, have revealed that calcification by these corals has declined by 14.2% since 1990. This appears to unprecedented on the Great Barrier Reef for at least the last 400 years ( De'ath et al., 2009 ) (but see D'Olivo et al., 2013 ; Figure 5 ). Given the complexity of the environmental changes occurring in places like the Great Barrier Reef, it is difficult to assign specific drivers of this decline. However, the combined effects of elevated warming and acidification from climate change, along with declining water quality, appear to be significant drivers of the changes observed ( D'Olivo et al., 2013 ). Declining growth and calcification rates have also been detected for Porites colonies in the Red Sea ( Cantin et al., 2010 ) and at several locations in Southeast Asia ( Tanzil et al., 2009 ).

Figure 5. (A–D) Partial-effects plots showing the variation of calcification (grams per square centimeter per year), linear extension (centimeters per year), and density (grams per cubic centimeter) in Porites from the Great Barrier Reef (GBR), Australia, over time. From De'ath et al. (2009) . Plots (A–C) are based on 1900–2005 data from 328 Porites colonies, and plot (D) on data for ten long cores. Light blue bands indicate 95% confidence intervals for comparison between years, and gray bands indicate 95% confidence intervals for the predicted value for any given year. Calcification declines by 14.2% from 1990 to 2005 (A) , primarily due to declining extension (B) . Density declines from 1900 onward (C) . The 1572–2001 data show that calcification increased weakly from ~1.62 before 1,700 to ~1.76 in ~1850, after which it remained relatively constant (D) before a weak decline since ~1960. (D–F) Decline coral cover of the GBR over 1985–2012. (E) Map of GBR with color shading indicating mean coral cover averaged over 1985–2014. Points show the location of 214 survey reefs in the northern, central, and southern regions, and their color indicates the direction of change in cover over time. (F) Box plots indicate the percentiles (25, 50, and 70%) of the coral cover distributions within each year and suggest a substantial decline in coral cover over the 27 years. Adapted from De'ath et al. (2012) and with the permission according to PNAS policy.

Studies of the influence of rapidly warming and acidifying conditions on mesophotic coral reefs are absent. Given that these reef systems cover roughly the equivalent area of shallow water coral reefs, understanding how environmental changes are likely to influence these important areas in terms of habitat the fisheries and biodiversity is important and should be a priority of future research ( Bongaerts et al., 2010a ). Linking the physiological and ecological response of mesophotic reefs to changes in pH and carbonate ion concentration will also be important in the context of understanding how mesophotic coral reef ecosystems will be affected by the shoaling of the saturation horizon in regions such as off Hawaii.

Our understanding of how deep ocean environments are likely to respond to changes in ocean temperature and chemistry are at an early stage. Like mesophotic coral reefs, little is known about the sensitivity of cold-water coral reefs to changes in temperature. As cold-water corals tend not to have a mutualistic symbiosis with Symbiodinium , their response is naturally different to that of symbiotic Scleractinian corals. As with mesophotic coral reefs, there is much more to be discovered with respect to how these critically important cold-water coral reefs are likely to respond to steadily warming and acidifying ocean. Coral reefs in the deep-sea have been identified as particularly vulnerable to the effects of ocean acidification; in part because of the numerical predominance of calcifying taxa, and in part because the pre-industrial carbonate levels at the depths and temperatures they inhabit were already low ( Freiwald et al., 2004 ). Experimental studies reveal that short-term exposures of important deep-water corals such as L. pertusa to a reduction in pH of around 0.15–0.3 units resulted in a decrease in calcification rates of between 30 and 56% ( Maier et al., 2009 ). However, subsequent work has shown that L. pertusa can acclimatize (i.e., maintain considerable calcification) to declining aragonite levels modifying skeletal structure and skeletal strength ( Form and Riebesell, 2012 ; Hennige et al., 2015 ). Observations of deep-sea corals in under-saturated waters from the SW Pacific also suggest some species-specific tolerance, however growth rates are extremely low and in under-saturated conditions dead coral skeletons dissolve rapidly ( Bostock et al., 2015 ; Thresher et al., 2015 ). Whether cold-water corals will be able to adjust to rapid warming and ocean acidification projected for the coming century is unknown. However, analyses of cold-water coral fossils suggest that a combination of declining aragonite and oxygen saturations will reduce the distribution of cold-water corals ( Thiagarajan et al., 2013 ).

While coral species and their symbionts have received a major amount of focus in terms of the effect of ocean warming and acidification on warm-water coral reef ecosystems, there is a growing number of studies that have revealed impacts on a broader range of reef organisms. Among the most affected are calcifying algae, calcareous phytoplankton, molluscs, and echinoderms, with the larval stages of some organisms being more sensitive than the adult phase ( Kroeker et al., 2013 ). Bioeroding organisms also respond to both warmer and more acidic conditions ( Dove et al., 2013 ; Fang et al., 2013 ; Reyes-Nivia et al., 2013 ). The sponge, Cliona orientalis , increased biomass and bioerosion capability when exposed to warmer and more acidic conditions, implicating a role of this sponge in helping tip the carbonate balance of reefs toward net erosion ( Dove et al., 2013 ; Fang et al., 2013 ). Similar observations have been made for bio-eroding endolithic algal communities, where small shifts in ocean temperature and acidity (i.e., CO 2 levels) enhanced skeletal dissolution and was associated with increased endolithic biomass and respiration under elevated temperatures and CO 2 levels ( Reyes-Nivia et al., 2013 ).

In addition to impacts on growth, calcification, and reproduction, there is growing evidence of impacts on a range of physiological systems of coral reef organisms. Ocean acidification, for example, impairs the homing ability and olfactory discrimination of some coral reef fish, with potential consequences for the ability of fish to detect and avoid predators ( Munday et al., 2009 ; Dixson et al., 2010 ). At present, there are few reports on the influence or not of ocean acidification on the metabolic performance of tropical fish species. In this regard, it will be important to explore whether or not tropical fish have the same challenges that temperate fish have when it comes to respiratory gas transport and acid–base balance ( Esbaugh et al., 2012 ; Pörtner et al., 2014 ). Physiological impacts combined with ecological impacts and habitat degradation, are likely to generate “surprises” for complex ecosystems such as those associated with both cold and warm water coral reefs.

Ecological Ramifications of Rapid Change

The impact of climate change on coral reef organisms has ramifications for ecosystems, some of which may be transformative in terms of their effects on primary productivity, food web dynamics, habitat forming species, disease ecology, and many other aspects ( Hoegh-Guldberg and Bruno, 2010 ). The recent decline in the abundance of warm-water coral reefs ( Hughes, 1994 ; Gardner et al., 2003 ; Bruno and Selig, 2007 ; De'ath et al., 2012 ), however, illustrates the complex yet fundamental ways that marine ecosystems are changing in response to rapid rates of ocean warming and acidification. The ecological ramifications of rapid global change for mesophotic coral reefs are less well-known or understood than those of warm-water shallow reef systems. Similarly, threats to cold-water coral reefs less well-understood and undoubtedly involve a different mix of local and global drivers ( Turley et al., 2007 ; Roberts and Cairns, 2014 ).

The major ecological responses of warm-water coral reefs to climate change have their origins in the response of reef-building corals to warming and acidification, and their role as framework builders within typical carbon reef systems ( Gattuso et al., 1998 ; Kleypas et al., 1999a ; Reynaud et al., 2003 ; Maier et al., 2009 ; Kroeker et al., 2013 ). As described above, corals are sensitive to small changes in temperature, light, and a number of other environmental variables, responding by disassociating from the dinoflagellate symbionts that populate their tissues (i.e., bleaching). Small changes in temperature are driving decreased growth and reproduction and increased mortality of corals in many parts of the world ( Hoegh-Guldberg and Smith, 1989 ; Hoegh-Guldberg, 1999 ; Hoegh-Guldberg et al., 2014 ). As corals lose their symbionts, they become vulnerable to death and disease, as well as being less able to compete with other benthic organisms. These changes have driven episodes of coral mortality associated with thermal stress, with the catastrophic loss of corals in particular regions over the past 30 years ( Hoegh-Guldberg, 1999 ; Baker et al., 2008 ; Eakin C. M. et al., 2010 ; Glynn, 2012 ). While some coral reefs have recovered over subsequent decades many others have not. Regional differences in the ability to recover are linked to the presence and absence of other factors affecting the resilience of reef building corals and other reef related organisms such as levels of herbivory, macroalgal cover, and coral recruitment rates ( Baker et al., 2008 ). The reduced resilience of reef building corals as a result of thermal stress is likely to be exacerbated by increasing ocean acidification, which has the potential to reduce the ability of corals to grow, calcify, and recover from disturbances. While teasing apart the effects of rising temperatures and increasing amounts of ocean acidification is difficult, both thermal stress, and acidification have the potential to reduce the ability of corals to recover from stresses ( Hughes et al., 2007 ). This may help explain why stressors such as cyclones, which do not appear to have increased in frequency over the past 30 years ( Callaghan and Power, 2011 ; IPCC, 2013 ), appear to be having longer-lasting impacts on coral communities on the Great Barrier Reef ( De'ath et al., 2012 ).

Mass coral bleaching reduces the energy available to corals, leading to physiological compromise. Warm-water corals, for example, exude mucus which is rich with the excess carbohydrates which provides food for a large number of molluscs, crustaceans, worms, ciliates, fish, and many other organisms ( Baker et al., 2008 ; Wild et al., 2011 ). It also appears to play an important role in preventing the settlement of fouling and disease organisms. Mucus secretion, however, is reduced in bleached corals, potentially leading to increased disease ( Harvell et al., 2007 ). Bleaching can also directly influence growth and reproduction of corals, as well is their tendency to succumb to a range of diseases ( Harvell et al., 1999 , 2007 ; Bruno and Selig, 2007 ; Baker et al., 2008 ). A reduction in reef-building corals raises the threat that a considerable proportion of the mega-diversity associated with coral reefs will face extirpation or, for some species, global extinction ( Glynn, 2012 ). A meta-analysis of 17 independent studies, undertaken by Wilson et al. (2006) , revealed that fish species reliant on live coral cover for food and shelter (some 62% of reef fish species) decreased in abundance within 3 years of disturbance events such bleaching, storms, and outbreaks of crown-of-thorns starfish that reduced coral cover by 10% or more.

The loss of calcifiers such as corals and calcareous algae due to warming and other stressors contributes to a reduced rate of community calcification, which is exacerbated by increases in dissolution and bioerosion as the water column becomes more acidified. Coral bleaching events driven by elevated temperatures has also been shown to shift carbonate budgets of coral reefs from net accretion to net erosion ( DeCarlo et al., 2017 ; Januchowski-Hartley et al., 2017 ). Sixteen years later, a third of reefs that were considered ecologically recovering ( Graham et al., 2015 ) did not show positive carbon budgets ( Januchowski-Hartley et al., 2017 ). Reefs remaining in negative carbonate budgets were those where massive coral loss was high and recovery of branched corals was low. The composition of reef benthic communities, which are sensitive to thermal stress, have an influence on the sensitivity of coral reef ecosystems to ocean acidification ( DeCarlo et al., 2017 ). In long-term studies done in mesocosms, carbonate balance of reefs tips toward overall dissolution under concentrations of CO 2 of more than 450 ppm ( Dove et al., 2013 ), which matches similar conclusions from previous experimental work ( Anthony et al., 2008 ; Wild et al., 2011 ; Andersson and Gledhill, 2013 ) and from the geographical distribution of coral reefs in relation to the aragonite saturation state of seawater ( Kleypas et al., 1999b ; Hoegh-Guldberg et al., 2007 ).

Evidence for Evolutionary Responses and the Relocation of Ecosystems

The strong relationship between short periods of elevated sea temperature in mass coral bleaching and mortality within warm-water coral reefs has been used to project how communities of reef building corals might change as ocean temperatures increase as a result of anthropogenic climate change ( Hoegh-Guldberg, 1999 ; Done et al., 2003 ; Donner et al., 2005 ; Frieler et al., 2012 ). Inherent to the conclusions of these studies, however, is the requirement that the thermal threshold of corals remain relatively constant over time. Evidence from the past 25 years, over which time satellite measurement programmes have used a simple algorithm based on sea surface temperature anomalies (relative to the average summer-time maxima 1985–1993) to predict when and where mass coral bleaching and mortality is likely to occur. This strongly suggests that little change has occurred in the sensitivity of reef building corals to thermal stress ( Eakin C. et al., 2010 ; Strong et al., 2011 ; Hoegh-Guldberg, 2012 ). Nonetheless, it is important to consider potential evolutionary responses of reef-building corals over the next 100 years as well as the potential for coral reef ecosystems to relocate, as conditions change. Due to the dearth of information available about mesophotic and cold-water corals, this discussion will be restricted to the evidence for warm-water coral reefs.

Other than dying, corals have the option of acclimatizing, evolving or relocating as conditions within a region become sub-optimal ( Hoegh-Guldberg, 2014a ). Reef -building corals, like all organisms, can adjust their phenotype or acclimate to match local conditions to some extent ( Gates and Edmunds, 1999 ; Middlebrook et al., 2010 , 2012 ). However, there is little or no evidence that acclimatization has resulted in an upward shift in the thermal tolerance of reef-building corals ( Eakin C. et al., 2010 ; Hoegh-Guldberg, 2012 ; Hoegh-Guldberg et al., 2014 ). Corals appear able to shift the relative ratio of different genetic clades or varieties of Symbiodinium within the one coral colony, which is correlated with tolerance to extreme temperatures ( Rowan et al., 1997 ; Berkelmans and van Oppen, 2006 ; Jones et al., 2008 ). Further investigation of these putatively more tolerant varieties reveals a physiological trade-off in terms of reduced growth and competitiveness ( Jones and Berkelmans, 2011 ).

A few studies ( Glynn et al., 2001 ; Maynard et al., 2008a , b ) have proposed that the thermal tolerance of reef building corals has increased over time, with less corals bleaching for similar amounts of thermal stress. The problem with these studies is several-fold ( Hoegh-Guldberg, 2009 ). For example, the assessment of stress levels was restricted to temperature alone despite the fact that variation in parameters such as the light intensity ( Hoegh-Guldberg, 1999 ; Mumby et al., 2001 ) and water flow rates over a reef ( Nakamura and Van Woesik, 2001 ) can significantly modify the overall stress levels arising from elevated temperature at small scales ( Hoegh-Guldberg, 2014a ). As well, studies like that of Maynard et al. (2008a) investigated community level responses, and hence are unable to distinguish the loss of fragile species as opposed to the specific acclimatization and/or adaptation of individual species. Evidence for acclimatization in other reef organisms has exposed some intriguing possibilities, such as transgenerational acclimatization, where organisms inherit improved tolerances from parents that have been previously exposed to high levels of stress. For example, some coral reef fishes being exposed to higher CO 2 levels prior to producing the next generation ( Donelson et al., 2012 ; Miller et al., 2012 ). Whether or not this mechanism operates within corals is unknown, although the very observation that the same satellite temperature threshold still works after more than 25 years is evidence that thresholds are not changing very rapidly. As observed by Donner et al. (2005) , the required rate of adaptation needs to match the rate of increase in sea temperatures or ~0.1–0.2°C per decade.

Genetic adaptation has also been suggested as a mechanism by which coral populations might be able to keep up with rapid changes in ocean temperature. Like all organisms, corals and their symbionts have adapted to local temperature conditions, a fact embodied by the fact that thresholds used by satellites for predicting mass coral bleaching and mortality are tied closely to local temperature conditions ( Strong et al., 2011 ). Adapting to local conditions, however, has probably taken hundreds if not thousands of years and is slowed by the fact that reef building corals have generation times from 5 to over 100 years ( Babcock, 1991 ). As a result, reef building corals do not have the population characteristics that would favor rates of evolution that would enable them to keep up with an environment that is changing faster than any time in the past 65 million years if not 300 million years ( Hönisch et al., 2012 ). Several researchers have suggested that corals might “evolve” by swapping their symbionts for more thermally adapted varieties ( Buddemeier and Fautin, 1993 ). Evidence of this, however, has not eventuated. These propositions also suffer from the problem that both the coral and the symbiont need to adapt to temperature change ( Hoegh-Guldberg et al., 2002 ; Stat et al., 2006 , 2009 ; Hoegh-Guldberg, 2012 , 2014a ). There are several observations of the shuffling of strains of Symbiodinium within the one host in response to warming ( Rowan et al., 1997 ). These changes, however, are examples of acclimatization as opposed to genetic adaptation ( Hoegh-Guldberg et al., 2002 ). In this regard, the advent of completely new symbiotic association between coral and a novel strain of Symbiodinium , hence a “new symbiotic genotype” has never been observed.

A third and final response by organisms facing rapidly changing conditions might be to relocate to new areas, which has been documented for a large number of marine plants and animals ( Poloczanska et al., 2013 , 2014 ). New records of several coral reef species have been reported at high latitude locations ( Precht and Aronson, 2004 ; Yamano et al., 2011 ) which is consistent with the proposition that corals might shift to higher latitudes. There is also ample evidence that small increases in ocean temperature in the past have resulted in the appearance of coral reefs at slightly higher latitudes than where they are found today ( Precht and Aronson, 2004 ; Greenstein and Pandolfi, 2008 ). While these reports are interesting, they are not sufficient to support the notion that whole coral reef ecosystems will shift successfully to higher latitudes as anthropogenic warming of the ocean continues which raises some important considerations. Firstly, how will ecosystem structure and function of coral reefs be affected if only a portion of species in the ecosystem shift, and which of these are the critical components for ecosystem services. Secondly, reduced light levels along with decreasing aragonite saturation are also critical factors in determining whether or not carbonate coral reef ecosystems will form successfully at higher latitudes. As recognized by Greenstein and Pandolfi (2008) , other factors (e.g., available shallow water shelf habitats) are crucially important for determining whether or not coral reef ecosystems will be able to move to higher latitudes. Thirdly, analogies to past shifts are limited given that current changes on coral reefs today include a multitude of other pressures in addition to temperature (e.g., pollution, ocean acidification). And finally, shifts in the past occurred over long periods of time during which conditions were relatively stable as compared to the extremely rapid changes typical of today. Current changes in ocean temperature and acidity will continue for centuries, if not millennia, under the current greenhouse gas emission pathway, thereby severely limiting the ability of populations and adaptive processes to keep up with a rapidly changing climate ( Hoegh-Guldberg, 2012 ).

Global Change: Projected Responses to Rapid Ocean Warming and Acidification

The close relationship between mass coral bleaching and mortality, and short periods of elevated sea temperature, provides an opportunity to explore how warm-water coral reefs are likely to be affected under different climate change scenarios ( Hoegh-Guldberg, 1999 ). Using projections of sea surface temperature (SST), future temperatures could be compared to established thermal thresholds for corals, and the frequency and intensity of future mass coral bleaching and estimated mortality. This led to the conclusion, which was somewhat controversial at the time, that coral reefs would experience mass coral bleaching and mortality on a yearly basis as early as 2030–2040. With field observations concluding that recovery from disturbances such as mass coral bleaching and mortality takes at least 10–20 years, the predictions of yearly mass coral bleaching and mortality events suggest strongly that coral dominated ecosystems would be unable to cope, and would start to disappear around this time. Subsequent studies revealed that these conclusions are not far-fetched and matched the expectations of a thermal threshold of corals was relatively fixed, as it appears to be ( Hoegh-Guldberg, 1999 ; Done et al., 2003 ; Donner et al., 2005 ; Eakin C. et al., 2010 ; Eakin C. M. et al., 2010 ; Frieler et al., 2012 ).

Hoegh-Guldberg et al. (2014) repeated the analysis of Hoegh-Guldberg (1999) using Coupled Model Intercomparison Project Phase 5 (CMIP5) data from an ensemble of 10–16 independent models. Historical and unforced temperature trends were compared with those from Representative Concentration Pathway (RCP) 2.6 (global temperature “about as likely as not” to exceed 1.5°C by 2100 relative to 1986–2005) and RCP 8.5 (global temperature “very likely” to exceed 3.0°C by 2100 relative to 1986–2005) in terms of model projections of the future for coral reef provinces. Model outputs were constrained to geographic areas (coral regions) known to contain warm-water coral reefs. The range in each case represents differences between models and model assumptions. Three things become apparent. First, the amount of SST warming that we have seen so far is very significant in each coral region given that average global temperature has warmed by 0.85°C over the period 1880–2012 (Table 1 , Figure 6 ). Second, differences between the two RCP scenarios do not become evident until mid to late century (Table 2 , Figure 6 ). Third, only conditions associated with the RCP 2.6 scenario stabilize, which is important if evolutionary processes are to be able to operate and re-establish coral reef ecosystems in these regions. In the context of the preceding discussion, this is the only scenario in which coral reefs have any chance of replenishing tropical coastal regions.

Table 2. Projected changes in sea surface temperature (SST °C) over the next 90 years for coral reef provinces (Figure 1B ) from AOGCM model simulations from the Coupled Model Intercomparison Project Phase 5 (CMIP5, http://cmip-pcmdi.llnl.gov/cmip5/ ) .

Figure 6. Past and future sea surface temperatures (SST) in six major coral reef provinces and locations (Figure 1B ) under historic, unforced (natural), and Representative Concentration Pathways (RCP)4.5 and 8.5 scenarios from Coupled Model Intercomparison Project Phase 5 (CMIP5) ensembles (see Table SM30-3 in Hoegh-Guldberg et al., 2014 ) . Observed and simulated variation in past and projected annual SST over various sites where coral reefs are prominent ecosystems. The black line shows estimates from the Hadley Centre Interpolated sea surface temperature 1.1 (HADISST1.1) data set ( Rayner et al., 2003 ) reconstructed historical SST dataset. Shading denotes the 5–95 percentile range of climate model simulations driven with “historical”changes in anthropogenic and natural drivers (62 simulations), historical changes in “natural” drivers only (25), the RCP4.5 emissions scenario (62), and the RCP8.5 (62). Data are anomalies from the 1986 to 2006 average of the HADISST1-1 data (for the HadISST1.q time series) or of the corresponding historical all-forcing simulations. Figure SM30-3 with the permission of IPCC AR5 ( Hoegh-Guldberg et al., 2014 ).

Hoegh-Guldberg et al. (2014) also looked at the annual incidence of bleaching and mortality events. The proportion of a coral grid cells with a reef province (Figures 1B , 7A ) that would have a particular stress level in any 1 year was calculated and the maximum for each decade then plotted. Two levels of stress were examined. The first being the amount of warming required to trigger mass bleaching, which is around one Degree Heating Months (DHM) ( Strong et al., 2011 ) and is shown in Figure 7B . The second was the amount of heat stress required to trigger mass mortality events like those that occurred in the Maldives, Okinawa, North-West Australia and Palau in 1998, and is calculated as five Degree Heating Months Hoegh-Guldberg, 1999 ; Figure 7C ). The conclusions from this analysis are very clear. Firstly, the risk of mass coral bleaching (DHM ≥ 1) increases steadily over the next few decades, affecting all regions of the Caribbean, Gulf of Mexico and eastern Pacific. By contrast, the Western Pacific Ocean, Coral Triangle, and Indian Ocean are likely to experience less stress and will still have large areas unaffected by annual mass coral bleaching by the end of the century. Secondly, conditions that drive mass mortality events today (DHM > 5) will spread across most regions by the end of the century under RCP 8.5. This risk decreases from RCP 8.5 to zero under RCP 2.6 with no regions experiencing annual conditions that would cause mass mortality event. Given the time that it takes for coral reefs to recover from mass mortality events (10–20 years), there is significant risk associated with high greenhouse gas emission scenarios given that the damage from these events, even in managed reef systems. Even 10% of grid cells being at risk of experiencing a mass mortality event would eventually add up to a very low number of unaffected areas by the end of century.

Figure 7. Annual maximum proportions of reef pixels with Degree Heating Months ( DHM, Donner et al., 2007 ) for each of the six coral regions (A) . (B) DHM ≥1 (used for projecting the incidence of coral bleaching; Strong et al., 1997 , 2011 ) and (C) DHM ≥5 (associated with bleaching followed by significant mortality; Eakin C. M. et al., 2010 ) for the period 1870–2009 using the Hadley Centre Interpolated sea surface temperature 1.1 (HadISST1.1) data set. The black line on each graph is the maximum annual area value for each decade over the period 1870–2009. This value is continued through 2010–2099 using Coupled Model Intercomparison Project Phase 5 (CMIP5) data and splits into the four Representative Concentration Pathways (RCP2.6, 4.5, 6.0, and 8.5). DHM were produced for each of the four RCPs using the ensembles of CMIP models. From these global maps of DHM, the annual percentage of grid cells with DHM ≥1 and DHM ≥5 were calculated for each coral region. These data were then grouped into decades from which the maximum annual proportions were derived. The plotted lines for 2010–2099 are the average of these maximum proportion values for each RCP. Monthly sea surface temperature anomalies were derived using a 1985–2000 maximum monthly mean climatology derived in the calculations for Figure 30-4 in Hoegh-Guldberg et al. (2014) . This was done separately for HadISST1.1, the CMIP5 models, and each of the four RCPs, at each grid cell for every region. DHMs were then derived by adding up the monthly anomalies using a 4-month rolling sum. Figure SM30-3 presents past and future sea temperatures for the six major coral reef provinces under historic, un-forced (no anthropogenic forcing), RCP4.5 and RCP8.5 scenarios. Reprinted with permission of the PCC AR5, Figure 30-10 ( Hoegh-Guldberg et al., 2014 ).

Considering cold-water coral ecosystems, at pre-industrial atmospheric CO 2 level, 9% of known cold-water coral ecosystems were in under-saturated water ( Cao et al., 2014 ). Under emission scenario IS92a (atmospheric CO 2 concentration 713 ppm and temperature increase of about 2.4°C by 2100), an estimated 70% of cold-water corals could be in under-saturated water by the end of the century with some ecosystems experiencing under-saturation by 2020s ( Guinotte et al., 2006 ; Turley et al., 2007 ). Even if mitigation efforts (e.g., through geoengineering) could reduce atmospheric CO 2 levels to pre-industrial by the end of the century, the lag in the recovery of deep ocean chemistry would result in longer-lasting threats to cold-water coral ecosystems ( Cao et al., 2014 ).

Living with Change: Implications for People and Livelihoods

Overall, the evidence presented above confirms earlier work ( Hoegh-Guldberg, 1999 ; Done et al., 2003 ; Donner et al., 2005 ; Frieler et al., 2012 ) and substantiates the serious concern regarding the vulnerability of carbonate coral reef systems to a rapidly changing world. Given the importance of coastal ecosystems such as warm-water coral reefs for hundreds of millions of humans ( Burke et al., 2011 ; Hoegh-Guldberg, 2015 ), these changes are likely to have implications for people and livelihoods, as well as regional security in some instances. It is also clear that we must increase our understanding of the effects of warming and acidifying oceans on mesophotic and cold-water coral reefs. These coral reefs represent important stores of biodiversity as well as habitat for fish, many of which are commercially important. As the deep ocean warms, the aragonite saturation horizon shoals and dissolved oxygen declines, it will be important to understand how these ecosystems will be affected. It will also be important to get a better understanding of how environmental conditions differ in the case of mesophotic reefs and whether or not they have the potential to act as refugia for coral reef species from the greater environmental extremes of shallow regions ( Bongaerts et al., 2010a , 2013 , 2015 ).

With regard to cold-water corals, management interventions are likely to be limited to regulating or banning fishing and mineral extraction in the locality of reefs ( Thresher et al., 2011 ). The highest priority for these sensitive ecosystems is to locate and protect sites that are likely to be refugia areas ( Thresher et al., 2015 ).

The recent consensus of the Intergovernmental Panel on Climate Change ( IPCC, 2014 ) identified a number of risks and vulnerabilities for coral reefs under rapid ocean warming and acidification, as well as exploring the ramifications and adaptation options (see Table 30.4 in Hoegh-Guldberg et al., 2014 ). Changes to the structure of ecosystems such as a loss of coral reefs, underpin a series of risks and vulnerabilities to fisheries production and consequently food and income security in tropical regions thus rates of unemployment and poverty. As coral reef ecosystems degrade or disappear, there is the risk that coastal fisheries production is reduced, decreasing food security and increasing unemployment. There is also a risk that the tourist appeal of tropical coastal assets may decrease as ecosystems take on less desirable states (i.e., from coral to seaweed domination), affecting the potential to attract tourist dollars. Reduced availability to food and income is likely to exacerbate coastal poverty in many equatorial countries. Strategies to reduce risks in both these cases involve strengthening integrated coastal zone management to reduce contributing stresses such as coastal pollution, overexploitation and physical damage to coastal resources.

As outlined by Poloczanska et al. (2013 , 2014 , 2016) , marine species are already redistributing toward higher latitudes. This has the potential to reorganize ecosystems including commercial fish stocks, drive changes to the distribution, and abundance of predators and prey, as well as increasing the risks of invasive species taking hold in new locations and ecosystems. Changes to the distribution of fish species in coral reef regions are expected as oceans warm ( Cheung et al., 2010 ; Pörtner et al., 2014 ; García Molinos et al., 2015 ). This has the potential to change national income in either direction, depending on location, as important fisheries stocks redistribute, increasing the likelihood of disputes over national ownership of fisheries resources ( Robinson J. et al., 2010 ). In this regard, key management actions to reduce risks include increased international cooperation over key fisheries as well as developing a better understanding of the linkages between ocean productivity, recruitment and fisheries stock levels ( Bell et al., 2011 ). International cooperation on measures which enable sustainable fishing of these valuable stocks that take into account the influence of climate change across international boundaries, are important adaptation measures that need to be implemented as soon as possible ( Robinson J. et al., 2010 ). Changing ecosystem structures in a warming and acidifying ocean is also likely to increase the risk of disease such as ciguatera and harmful algal blooms, with ramifications for human health and well-being. Strategies in this case involve increased monitoring and education surrounding key risks, plus the development of alternative fisheries and income for periods in which disease incident increases ( Bell et al., 2013 ).

Conclusion: The Key Role of Climate Stabilization and Non-Climate Change Factors

A recurrent theme within this review is the fact that we are already seeing major and fundamental change occurring in the world's ocean in response to climate change and that the rate of change is largely outstripping the ability for coral reefs to adapt genetically or relocate. If greenhouse gas emissions are not mitigated, it is very clear that the ocean will be a vastly different place by the mid to late century ( Gattuso et al., 2015 ). It is also clear that there are few or no adaptation strategies for humans to counter the risks of ocean warming and acidification at global scales. If they did exist, they would almost certainly be prohibitively expensive relative to the costs of developing solutions to the unprecedented rise of CO 2 in the earth's atmosphere.

This leaves us with two clear options with respect to preserving invaluable ecosystems such as coral reefs. The first is to stabilize planetary temperature and CO 2 concentrations as quickly as possible. Only then will biological responses such as acclimation and genetic adaptation have any chance of operating. The second is to dramatically reduce local stresses which are currently acting on coral reefs and which are reducing their resilience to climate change. By reducing these non-climate stresses, coral reefs will have the opportunity to develop greater robustness or resilience to the challenges of a changing planet. However, if this is not combined with stabilization of temperatures and acidification then it is likely to only temporarily put off the inevitable. If we do these two things, there is a chance that conditions will stabilize on planet earth by mid-to-late century, ensuring that some of the spectacular coral reef ecosystems will be able to flourish across the world's tropical regions.

Author Contributions

OH led the project and wrote 50% of this manuscript. EP contributed to core concepts in the manuscript and contributed 30% of text. SD and WS contributed 15 and 5% respectively to the writing.

Funded primarily by the Australian Research Council (Canberra; FL120100066, LP110200874, CE140100020), The University of Queensland, and the National Oceanic and Atmospheric Administration (Washington, DC).

Conflict of Interest Statement

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

Acknowledgments

The authors are grateful for support from the Australian Government, Australian Research Council, National Oceanic and Atmospheric Administration, and Queensland Government. OH was supported by an ARC Laureate Fellowship during the development and publications of this study, and was a member of the ARC Centre for Excellence in Coral Reef Studies through the University of Queensland. SD and OH would like to recognize the generous support of the Centre Scientifique de Monaco for the final analysis and writing stages of this study.

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Keywords: corals, climate change, ecosystems goods and services, decline, warming ocean, ocean acidification

Citation: Hoegh-Guldberg O, Poloczanska ES, Skirving W and Dove S (2017) Coral Reef Ecosystems under Climate Change and Ocean Acidification. Front. Mar. Sci . 4:158. doi: 10.3389/fmars.2017.00158

Received: 07 January 2017; Accepted: 10 May 2017; Published: 29 May 2017.

Reviewed by:

Copyright © 2017 Hoegh-Guldberg, Poloczanska, Skirving and Dove. This is an open-access article distributed under the terms of the Creative Commons Attribution License (CC BY) . The use, distribution or reproduction in other forums is permitted, provided the original author(s) or licensor are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.

*Correspondence: Ove Hoegh-Guldberg, [email protected]

† Present Address: Elvira Poloczanska, Alfred Wegener Institute for Polar and Marine Research, Integrative Ecophysiology, Bremerhaven, Germany

Disclaimer: All claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. Any product that may be evaluated in this article or claim that may be made by its manufacturer is not guaranteed or endorsed by the publisher.

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The Evolution of Coral Reef under Changing Climate: A Scientometric Review

Chandra segaran thirukanthan.

1 Institute of Marine Biotechnology (IMB), Universiti Malaysia Terengganu (UMT), Kuala Nerus 21030, Terengganu, Malaysia

Mohamad Nor Azra

2 Research Center for Marine and Land Bioindustry, Earth Sciences and Maritime Organization, National Research and Innovation Agency (BRIN), Pemenang 83352, Indonesia

Fathurrahman Lananan

3 East Coast Environmental Research Institute, Universiti Sultan Zainal Abidin (UniSZA), Gong Badak Campus, Kuala Nerus 21300, Terengganu, Malaysia

Gianluca Sara’

4 Laboratory of Ecology, Earth and Marine Sciences Department, University of Palermo, 90133 Palermo, Italy

Inga Grinfelde

5 Laboratory of Forest and Water Resources, Latvia University of Life Sciences and Technologies, LV-3001 Jelgava, Latvia

Vite Rudovica

6 Department of Analytical Chemistry, University of Latvia, LV-1004 Riga, Latvia

Zane Vincevica-Gaile

7 Department of Environmental Science, University of Latvia, LV-1004 Riga, Latvia

Juris Burlakovs

8 Mineral and Energy Economy Research Institute of the Polish Academy of Sciences, 31-261 Krakow, Poland

Associated Data

No data was generated from the study.

Simple Summary

Coral reefs are vital ecosystems with high biodiversity and ecological services for coastal communities. Climate change is accelerating, with detrimental consequences on coral reefs and related communities, but it is challenging to keep up with the literature given its current rapid expansion. The current review foresees three future trends in the area of coral reefs and climate change, including (i) incorporating future scenarios, (ii) climate-induced temperature changes, and (iii) adaptation strategies, which are expected to move society closer to the following Sustainable Development Goal: 13 Climate Action.

In this scientometric review, we employ the Web of Science Core Collection to assess current publications and research trends regarding coral reefs in relation to climate change. Thirty-seven keywords for climate change and seven keywords for coral reefs were used in the analysis of 7743 articles on coral reefs and climate change. The field entered an accelerated uptrend phase in 2016, and it is anticipated that this phase will last for the next 5 to 10 years of research publication and citation. The United States and Australia have produced the greatest number of publications in this field. A cluster (i.e., focused issue) analysis showed that coral bleaching dominated the literature from 2000 to 2010, ocean acidification from 2010 to 2020, and sea-level rise, as well as the central Red Sea (Africa/Asia), in 2021. Three different types of keywords appear in the analysis based on which are the (i) most recent (2021), (ii) most influential (highly cited), and (iii) mostly used (frequently used keywords in the article) in the field. The Great Barrier Reef, which is found in the waters of Australia, is thought to be the subject of current coral reef and climate change research. Interestingly, climate-induced temperature changes in “ocean warming” and “sea surface temperature” are the most recent significant and dominant keywords in the coral reef and climate change area.

1. Introduction

Scleractinians, or stony corals, emerged during the Cambrian period and constructed the earliest reefs, dating to approximately 410 million years ago [ 1 , 2 , 3 ]. Five major coral extinctions have occurred since then, all of which have been linked to rising temperatures and higher levels of carbon dioxide in the atmosphere [ 2 , 4 ].

While coral reef research commenced more than 100 years ago, concern over the state of coral reefs is relatively recent, occurring only in the last four decades. In 1981, at the 4th International Coral Reef Symposium, Edgardo Gomez initiated the conversation on threats to coral reefs by presenting his concerns to the scientific community [ 5 , 6 ]. Coral reefs are considered an important marine resource for coastal communities, and the conference participants were mainly focused on coral reef management and environmental impacts and related fisheries activities [ 5 , 7 , 8 ]. In addition to the natural stresses that have always existed on coral reefs, such as storms, freshwater inundation, and seismic and volcanic events, there is growing evidence of new emerging threats potentially causing global damage to coral reefs [ 7 , 8 , 9 , 10 ].

At current extinction rates, it is estimated that we are commencing the sixth mass extinction event [ 11 , 12 , 13 ], with individual extinctions occurring approximately 1000-fold faster than the expected background extinction rate. Theoretically, species extinctions occur at a rate proportional to the rate of speciation or the creation of new species [ 14 ]. Current extinction rates are much higher than speciation rates. This is largely due to the fact of anthropogenic factors [ 15 , 16 ], such as habitat destruction [ 17 ], deforestation [ 18 ], pollution [ 7 ], ocean acidification, climate change resulting from greenhouse gas emissions, and overexploitation of ecological resources [ 19 , 20 , 21 ]. It is estimated that 75% of species will go extinct unless human pressures on the environment are scaled back soon [ 20 , 22 , 23 , 24 , 25 ]. Further management efforts are required to reduce the impacts of climate change and human anthropogenic stress towards coral reef communities [ 7 ].

Anthropogenic pressures on reefs have been the dominant factor damaging coral reefs through a range of stresses ( Figure 1 ). Unsustainable land-based human activities, such as deforestation, poorly regulated agriculture, and urban/industrial development, are major contributors to the release of excessive sediments and nutrients into the environment [ 26 , 27 , 28 ]. Increases in human-caused greenhouse gas emissions are the primary factor in the current climatic shift around the world [ 29 , 30 ]. The ocean acts as a massive sink that absorbs carbon dioxide, resulting in the acidification of the oceans [ 31 , 32 ]. Coral bleaching and widespread damage to the coral reef ecosystem have become increasingly common because of thermal stress brought on by rising ocean temperatures [ 33 ]. Lack of food to sustain the coral reef ecosystem and disruption of larvae dispersal are both attributable to altered currents, upwelling, and/or vertical mixing brought on by changing currents and winds [ 34 ]. Storms and cyclones are “agents of mortality” on coral reefs and can have a direct impact on the structure and local distribution of coral reef assemblages, especially through the large waves they produce [ 35 ]. Reduced salinity caused by heavy rainfall and enhanced surface run-off onto nearshore reefs during cyclones, storms, and heavy precipitation rates can lead to algal blooms and other devastating results [ 36 , 37 ]. The rise in sea level, caused by thermal expansion and the melting of ice on land, has varied across different regions of the world over the past century, with an average increase of approximately 20 cm [ 38 ]. The sedimentary mechanisms triggered by the rising sea levels have the potential to intensify and jeopardize crucial physiological reef processes, such as photosynthesis, feeding, and recruitment, thereby posing a severe threat to coral reefs and related ecosystems, such as seagrass meadows and mangrove forests [ 39 , 40 ]. This threat, coupled with increasing carbon dioxide (CO 2 ) emissions, can negatively impact these vital ecosystems. Without effective local measures and a concerted effort to reduce carbon dioxide emissions, these effects are projected to intensify, leading to an unprecedented degradation of marine biodiversity and ecological balance [ 41 , 42 ].

Threats to coral reefs posed by factors related to climate change.

Given the length of time that scientists have been studying climate change [ 43 ], the sheer volume of published research in the field can make it difficult for scientists to develop an overview of the topic [ 44 ]. Bibliometric analysis can be used to provide an overview of voluminous scientific literature [ 45 ]. Research output in each field can be charted in terms of its characteristics and evolution through quantitative examination of publication data [ 45 , 46 ]. Performance and research patterns of authors, journals, countries, and institutions can all be evaluated with the help of bibliometric methods, and patterns of collaboration between these entities can be identified and quantified [ 47 ]. A research domain’s multidisciplinary nature and the variety of journals publishing on a given topic can be inferred from the subject categories assigned to publications and the number of journals publishing on a given topic. The most recent developments, research directions, and top-of-mind issues in a particular field can be gleaned from bibliometrics [ 48 ]. In addition, bibliometrics can be a useful tool for guiding scientific policy. Findings from bibliometric analyses not only inform researchers and policymakers but also aid in the distribution of funds for scientific investigation [ 49 ].

2. Scientometric Analysis

Software programs, such as VOSviewer (Centre for Science and Technology Studies (CWTS), Leiden University, Leiden, The Netherlands), Pajek (University of Ljubljana, Ljubljana, Slovenia), and CiteSpace (Drexel University, Philadelphia, PA, USA) can be used to create scientometric visualizations. CiteSpace is a scientometric-based analysis tool that provides two main outputs for researchers. Firstly, it includes three central concepts: burst detection, betweenness centrality, and heterogeneous networks. Secondly, CiteSpace addresses three practical issues: identifying the nature of a research front, labeling specialties, and detecting emerging trends and abrupt changes in a timely manner. Identifying these outputs involves six main procedures: time slicing, thresholding, modelling, pruning, merging, and mapping. CiteSpace’s functionalities, such as dual map overlay, burst detection, cluster explorer, and timeline view, are particularly helpful in identifying the current research trends of a specific field. Researchers can use this information to gain insights into the overall state of research in a given field, the pace of advancement, and the most prominent areas of interest [ 50 , 51 ].

Briefly, the dual map overlay graphically identifies both original documents and cocited networks and expresses their relationships by connecting them with lines. Burst detection refers to a frequency surge of a specific knowledge domain. A cluster is defined as the frequency of citations of cited references. A timeline arranges papers in chronological rows, with each cluster represented by one row and papers represented as nodes [ 52 ].

3. Objective of the Review

This review addresses the question, “How have coral reefs been affected by climate change and their interactions using bibliometrics?”. Specific objectives were to assess the literature in terms of (i) the annual number of articles published, (ii) countries/regions involved in the field, (iii) research topics, (iv) cocited networks (i.e., frequency of two different documents are cited together in other documents), (v) cluster networks, (vi) research topic (i.e., keywords) burstiness, (vii) dual map overlay, and (viii) the future trends of the knowledge domain (i.e., coral reef and climate change). The main article structure diagram is shared in Figure 2 .

Article structure diagram.

4. Systematic Data Collection

This study was analyzed based on the Web of Science database Core Collection (WOSCC) on 16 November 2022. Keywords used for data collection comprised terms related to climate change [ 53 ] and coral reefs [ 54 ]. CiteSpace, 6.1.4, version 64-bit for Windows, was used to visualize current trends. The search string are as follows: CLIMATE CHANGE: (“climat* chang*”) OR (“global warm*”) OR (“seasonal* variat*”) OR (“extrem* event*”) OR (“environment* variab*”) OR (“anthropogenic effect*”) OR (“greenhouse effect*”) OR (“sea level ris*”) OR (erosio*) OR (“agricult* run-off”) OR (“weather* variab*”) OR (“weather* extrem*”) OR (“extreme* climat*”) OR (“environment* impact*”) OR (“environment* chang*”) OR (“anthropogenic stres*”) OR (“temperature ris*”) OR (“temperature effect*”) OR (“warm* ocean”) OR (“sea surface* temperat*”) OR (heatwav*) OR (acidific*) OR (hurrican*) OR (“el nino”) OR (“el-nino”) OR (“la nina”) OR (la-nina) OR (drought*) OR (flood*) OR (“high precipit*”) OR (“heavy rainfall*”) OR (“CO 2 concentrat*”) OR (“melt* of the glacier*”) OR (“melt* ice*”) OR (“therm* stress*”) OR (“drought”) OR (“hypoxia”) AND CORAL REEF: (“coral reef*”) OR (“barrier reef*”) OR (“atolls”) OR (“fring* reef*”) OR (“coral island*”) OR (“atoll lagoon*”) OR (“biogenic deposit*”).

5. Evolution of the Literature

5.1. global publication.

The coral reef and climate change field showed an increase in published articles, indicating that the research is in rising momentum ( Figure 3 ). Since 2006, studies on coral reef and climate change have been published, amounting to approximately 1.5% to almost 10% of total articles on the WoS platform in 2021, especially in the WoS Core Collection (WOSCC) database. The field entered an accelerated uptrend phase in 2016, and it is anticipated that this phase will last for the next 5 to 10 years for research publications and citations. This could be because the WOSCC added a new edition of the Emerging Sources Citation Index (ESCI) database in 2015 [ 55 ]. The leading countries in coral reef and climate change research are shown in Figure 4 . The USA and Australia produced the most publications, with 2940 and 2933 articles, respectively, more than triple the output of third-ranked England, with approximately 719 articles. The USA and Australia contributed more than 75% of all publications. The study also found that there is a lack of studies conducted in the African regions. As expected, countries without coastal areas, such as Kazakhstan and Mongolia, showed no articles published in the coral reef and climate change fields. Additionally, most island nations, such as Japan, New Zealand, Australia, or Cuba, contributed to the knowledge of the coral reef and climate change.

Number of original research articles on the impact of climate change on coral reefs published annually from 1977 until 2021.

Nations publishing the most research on coral reefs and climate change generated from the freely editable map chart website ( https://www.mapchart.net/world.html (accessed on 10 January 2023)).

5.2. Leading Institutions, Funding, and Authorship Distribution

Ellegard and Wallin [ 46 ] opined that the distribution of research institutions is a useful indicator of academic support for a discipline. The network of institutions generated 948 nodes (i.e., group of entities) and 2666 collaborative links among institutions conducting research ( Figure 4 ). The 4159 documented affiliations reflect the importance of this field in academia and the intensity of the investigations. Institutions in this collaborative network (a collaborative affiliation appeared in the article) that have contributed the most to this field are shown in Figure 5 and Table 1 . An analysis of 7743 publications related to coral reefs and climate change identified a total of 29,890 affiliations. The top 20 institutions contributed to 7231 affiliations, which accounted for nearly 24% of all publications. Specifically, James Cook University was found to have contributed the highest number of publications (1119, 14% of the total), followed by the Australian Institute of Marine Science and the University of Queensland. These findings suggest that a small number of institutions have played a significant role in the research on the impact of climate change on coral reefs.

Collaborative network of institutions researching coral reefs and climate change from 1976 to 2021.

Top 20 institutions in the field of coral reef and climate change.

Table 2 lists the major funding agencies cited by publications in this field. The Australian Research Council funded 1001 publications, or nearly 13% of the total. Australian institutions have been at the forefront of climate change research partly because they have had greater access to funding. For instance, the Great Barrier Reef Foundation (GBRF) was awarded USD 443.3 million for the Reef Trust project between 2018 and 2019, the largest investment in reef protection to date. In 2019 and 2020, the budget allocated USD twenty-three million dollars for water quality projects, USD 4.33 million for Crown-of-Thorns (COT) control, USD 16.3 million for reef protection aimed at supporting traditional owners, USD 2.6 million for community reef protection, and USD 1.5 million for integrated monitoring and reporting. This funding allocation was based on the 2019 Great Barrier Reef Foundation’s budget plan (GBRF, 2019). In addition to these investments, the Australian Government also dedicated USD 6 million in 2018 to support the Reef Restoration and Adaptation Program’s concept feasibility phase. This program aimed to investigate the most effective science and technology options for restoring the reef, including methods for cooling and shading reef structures, coral reproduction and recruitment, biocontrol, field treatments, and coral plantation initiatives [ 56 ]. The ARC Centre of Excellence for Coral Reef Studies, located at James Cook University, conducts cutting-edge research on coral reefs. With strong collaborative ties to 24 other institutions in nine countries, it is Australia’s top contributor to coral reef sciences. Five of the ten most prominent scientists identified by CiteSpace analysis are associated with the ARC Centre of Excellence for Coral Reef Studies.

Top 20 funding agencies related to coral reef and climate-change-related studies.

The network of authors generated in CiteSpace is a valuable tool for understanding collaborations and identifying potential future collaborations within a given field. In this study, the network of authors in the field of coral reefs and climate change generated 1680 nodes and 6931 links, where nodes represent authors with the number of publications, and links between nodes represent collaborations between authors. As a general rule, a higher number of nodes generated from CiteSpace indicates a greater degree of collaboration between authors.

The top authors in Table 3 are crucial indicators of further scientific collaboration and advancement within the field of coral reefs and climate change. However, it is important to note that while first authorship is typically assigned to the individual who made the most significant contribution to the research, this does not mean that coauthors’ contributions are any less significant or valuable. Coauthors can provide specialized expertise, contribute to data analysis and interpretation, or support the project through funding, logistical support, or other means. The authorship order and the number of authors on a publication may vary based on the specific circumstances of the research project and the agreed-upon conventions of the discipline. Ultimately, understanding the network of authors and their collaborations can help identify potential areas of future research and scientific collaboration within the coral reefs and climate change.

Top 20 authors in the field of coral reef and climate change.

Emeritus Professor Terry Hughes of James Cook University, who served as the ARC Centre of Excellence for Coral Reef Studies Director from 2005 to 2020, topped the list. In 2016, Nature named Hughes one of the “10 people who mattered this year” for addressing the widespread coral bleaching event brought on by climate change. Hughes’s research has led to practical solutions to improve marine environmental management [ 57 ]. His work on the effects of climate change on coral reefs has been widely cited, especially his paper on the resistance of some coral reefs to climate change and anthropogenic factors [ 58 ]. Second on the list was Professor Ove Hoegh-Guldberg from the University of Queensland (UQ), Australia. He serves as Director of the Global Change Institute at UQ and also as a Chief Investigator at the ARC Centre of Excellence for Coral Reef Studies [ 59 ]. Dr. Katharina Fabricus is a coral reef ecologist and a Senior Principal Research Scientist at the Australian Institute of Marine Science [ 60 ]. Many of her highly cited publications are on topics related to ocean acidification [ 61 , 62 , 63 , 64 ], the impacts of water quality on coral reefs [ 65 , 66 ] and understanding the effects of terrestrial run-off on coral reefs [ 67 , 68 ].

Dr. Peter W. Glynn, from the National Center for Coral Reef Research, University of Miami, was among the pioneers in analyzing and reporting the impacts of the 1982–1983 El Niño warming event on Eastern Pacific coral reefs [ 69 ]. This was followed by Tim McClanahan, a senior conservation zoologist at the Wildlife Conservation Society and also an associate at the ARC Centre of Excellence for Coral Reef Studies [ 70 ]. His global study of more than 2500 reefs produced a Bayesian hierarchical model to predict how reef fish biomass is related to 18 socioeconomic drivers and environmental conditions [ 71 ]. Dr. Peter Mumby is a coral reef biologist from the University of Queensland and also a Chief Investigator at the ARC Centre of Excellence for Coral Reef Studies [ 72 ]. He collaborated with Professor Ove Hoegh-Guldberg to publish “Coral Reefs under Rapid Climate Change and Ocean Acidification”, which is one of the most cited papers in the field (see Table 4 ) [ 22 ]. Another study published in Nature reported the resilience of Caribbean coral reefs against moderate hurricanes [ 73 ]. Dr. David Bellwood, an Australian Laureate Fellow and Distinguished Professor at James Cook University [ 74 ], has reported on the effects of climate change on coral reef ecosystems, even though his primary research interests are in biology and the evolution of reef fish [ 58 , 75 , 76 ].

The most highly cited references about coral reefs and climate change.

* Burst index: the value generated from the CiteSpace indicates the level of importance of each article in the field. ** Centrality: the main focused article between cited references in the publications.

Dr. John Pandolfi from the University of Queensland is a paleoecologist and a Chief investigator at ARC. His research integrates long-term ecological and environmental time series data to discover past and future influences of natural variability, human impact, and climate change on coral reef resilience. Among his highly cited works is a projection of the future of coral reefs under global warming and ocean acidification [ 20 ]. Dr. Kenneth RN Anthony, an associate scientist at the Australian Institute of Marine Science and director of Environmental Strategies ES5, has published widely on ocean acidification [ 77 , 78 , 79 ]. Dr. John Bruno, from the University of North Carolina, is a marine ecologist focusing on the impacts of climate change on marine ecosystems, particularly coral reef ecology. His publication with Dr. Ove Hoegh-Guldberg, on the effects of climate change on global marine ecosystems is one of his most cited works [ 80 ].

Next on the list is Emeritus Professor Barbara E. Brown from Newcastle University, who conducted extensive research on coral bleaching, specifically on the role of zooxanthellae [ 81 ]. Next on the list is Professor Andrew C. Baker, a marine biologist at the University of Miami, who studies coral reefs and climate change. He leads the Coral Reef Futures Lab and focuses on developing and testing methods to increase coral reef resilience [ 82 ].

Glenn De’ath and Ray Berkelmans, both from The Australian Institute of Marine Science, are also highly cited for their research on coral reefs. De’ath’s work involves statistics and ecology, specifically on the Great Barrier Reef coral cover decline [ 83 ], while Berkelmans’ research focuses on thermal stress, adaptation to climate warming, the resilience of reef communities, and upwelling [ 84 ].

Joan Kleypas, a Senior Scientist from the National Center for Atmospheric Research, is also on the list, and her highly cited works revolve around the impact of ocean acidification on coral reefs [ 85 , 86 ]. Next is Professor Nick Graham from Lancaster University, who assesses the impacts of climate-induced coral bleaching on coral reef fish assemblages, fisheries, and ecosystem stability [ 87 ].

Emeritus Professor Michael Lesser from the University of New Hampshire is also highly cited for his work on climate change-related stressors’ biochemical and physiological impacts on coral reefs [ 88 ]. Professor Joseph Loya from Tel Aviv University quantifies changes in biodiversity and assesses reef health [ 89 , 90 ], while Professor Peter Edmunds focuses on the physiological ecology of tropical coral reefs [ 91 ].

Lastly, Toby Gardner is a Senior Research Fellow from Stockholm Environment Institute, known for his extensive work on Caribbean corals. He co-leads SEI’s Initiative on producer-to-consumer sustainability and the transparency for sustainable economies platform. His long-term observations revealed that the coral cover of the Caribbean basin declined by 80% in just thirty years [ 92 ].

5.3. Emerging Research Disciplines

CiteSpace’s “Category” node type was used to generate a visual map showing research disciplinary categories represented by papers addressing issues related to climate change’s impact on coral reefs. The centrality of a network (i.e., the center of collaborative activities) comprising 135 nodes and 336 links was computed after the data were simplified and merged (i.e., automatically generated from the CiteSpace algorithm and programming) ( Figure 6 ). The five disciplines with the most publications in descending order were marine and freshwater biology, environmental sciences, ecology, oceanography, and geosciences. The study of coral reefs is a multifaceted research topic that includes many fields of study, as demonstrated by the distribution map. Disciplines in related subjects such as biodiversity conservation, geography, physical sciences, biology, evolutionary biology, geology, paleontology, and water resources, show strong connections, represented by the sizes of the nodes. The number of published papers is comparably low in some research disciplines, such as toxicology, biotechnology and applied microbiology, green and sustainable science and technology, and biochemistry and molecular biology. However, the relatively high betweenness centrality values of these fields suggest their significant contribution to interdisciplinary research, signifying their pivotal position in the scientific network. This centrality may also hint at their potential for future development and advancement in the field.

Network of linked research disciplines. The sizes of the modes are proportional to the frequency of the subject category cooccurrence. The thickness of the lines between the two nodes is proportional to the strength of the linkages between the two research disciplines.

5.4. Research Cluster Analysis

Cluster analysis is a popular method of statistical data analysis and knowledge discovery because of its ability to uncover latent semantic themes in textual data [ 93 , 94 ]. Cluster analysis can divide a large body of research data into various units based on the relative degree of term correlation, making it easier to identify the research themes, trends, and connections within a given field of study [ 94 , 95 ]. A cluster’s homogeneity can be quantified using an index called the mean silhouette, with values ranging from −1 to 1. The average silhouette value for each cluster was determined using CiteSpace. The higher the value, the more similar the cluster’s members are to one another [ 96 ]. The network showed 24 clusters in the context of the scientometric analysis mapping the link between climate change and coral reefs ( Figure 7 ).

The reference co-citation research cluster network. Based on a one-year interval, a 24-cluster network of document co-citation with burst detection from 1976 to 2021. Node sizes are proportional to the frequency of the publications’ co-citations.

The largest cluster (#0) has 291 members (i.e., number of publications) and a silhouette value of 0.863 and is labeled as “coral reef.” The most cited article of this cluster is by Gilmour et al. [ 97 ]. They monitored and assessed the impacts of the 2016 heat stress event on Western Australian coral reefs. They found that mass bleaching in 2016 reduced coral cover by 70% at Scott Reef and caused widespread mortality (>30%) at Christmas Island, Ashmore Reef, and inshore reefs in southern Kimberley. A coral phase shift is characterized by a rapid decline in coral abundance or cover and an accompanying rise in non-reef-building organisms, like algae and soft corals [ 98 , 99 ]. The second largest cluster (#1) has 247 members and a silhouette value of 0.876, labeled as “phase shift.” This publication by Brodie et al. [ 100 ], entitled “Terrestrial pollutant run-off to the Great Barrier Reef: An update of issues, priorities and management responses”, is the most cited article in this cluster. They addressed findings from studies of problems caused by surface run-off of pollutants like nitrates from fertilizers, herbicides from crops, etc. Within the cluster of the study, there are three different types of management generated automatically and have mentioned (i) Reef Plant 2009, (ii) Reef Rescue, and the (iii) Reef Protection Package in the analysis. These topics are just some of the initiatives set up to continuously monitor and report on levels of discharges into the Great Barrier Reef. Multiple observations of specific facets of the topic have been published; Hughes [ 101 ]; McManus and Polsenberg [ 102 ]; Idjadi et al. [ 103 ]; Norström et al. [ 104 ]; Graham et al. [ 105 ]; Crisp et al. [ 106 ]. Many anthropogenic stressors have been linked to this phenomenon [ 106 , 107 , 108 ]. Nutrients play a pivotal role in conceptual models of how coral reef communities form. These studies show that corals have a competitive advantage over macroalgae in low nutrient conditions but that the advantage shifts to macroalgae in higher nutrient conditions [ 102 , 109 ]. Siltation, resulting in mud-bacterial complexes, collectively known as “marine snow,” is another factor that hinders coral growth. In addition, excess nutrients resulting in plankton blooms reduce light, thereby inhibiting coral growth [ 110 ].

The fourth largest cluster (#3), “ocean acidification”, has 233 members and a silhouette value of 0.959. The most cited article of this cluster is Bates [ 111 ], which reported twenty years (1996 to 2016) of marine carbon cycle observations at Devils Hole, Bermuda. Her findings shed light on the dynamic nature of biogeochemical processes like primary production, respiration, calcification, and CaCO 3 deposition in the Bermuda reef system. During this period, neither warming nor cooling of any significance was observed. However, increases in inorganic carbon in onshore waters were primarily due to increased salinity (45%), uptake of anthropogenic CO 2 (25%), and changes in Bermuda reef biogeochemical processes (30%). Increases in atmospheric carbon dioxide concentrations result in the absorption of more carbon dioxide by oceans, which in turn causes a decrease in pH [ 112 , 113 , 114 , 115 ]. The majority of research on ocean acidification has focused on the impact of changes in ocean chemistry towards suboptimal states of aragonite and calcite saturation on the calcification processes of pelagic and benthic organisms [ 77 , 116 , 117 , 118 , 119 ]. However, it is likely that ocean acidification also has an effect on other physiological processes, such as growth and reproduction in significant reef-building species [ 77 ].

The Indo-Pacific region’s Coral Triangle, the world’s epicenter of marine biodiversity [ 120 , 121 ], is predicted to become a “marginal” coral habitat between 2020 and 2050 unless CO 2 emissions are reduced [ 122 , 123 ]. In addition to reducing coral diversity, acidification also results in a decline in shellfish and fish species due to the loss of reef structure, which provides habitat for these other species and reduces the reefs’ capacity to mitigate the effects of storm waves and erosion [ 122 , 124 ]. Ocean acidification has a devastating impact on the economies of ocean-dependent sectors of the global economy. Previous studies have provided estimates of the economic impact of ocean acidification on marine mollusk and shellfish production, as well as the bioeconomic costs associated with coral reef damage [ 125 ]. These studies have shed light on the detrimental effects of ocean acidification on marine ecosystems, which in turn, can have severe economic implications. Estimating these costs can aid in developing policies aimed at reducing the negative effects of ocean acidification and promoting the sustainable use of marine resources. For instance, according to a study by Narita et al. [ 126 ], the global annual loss of mollusk production due to the fact of ocean acidification could amount to between USD 6 billion and USD 100 billion. Commercially valuable finfish populations will suffer as a result of global ocean changes that reduce coral reef coverage, resulting in a loss of habitat, reduced availability of prey, and increased predation [ 125 , 127 , 128 ]. The scientometric analysis has identified four prominent clusters, also referred to as topics, which represent distinct research areas based on their geographic location. These clusters include the “central red sea” (#3), the “eastern pacific” (#5), the “great barrier reef” (#8), and the “Dominican Republic” (#18). These geographic regions are frequently cited in scientific research as they represent the study location of many relevant studies.

The most cited article of the “Red Sea” cluster is by Osman et al. [ 129 ], which mapped coral microbiome composition along the northern Red Sea. The Red Sea is a distinctive body of water that is an evaporative basin with a high salinity above 38 ppt [ 130 , 131 ]. It is home to some of the world’s most thriving and productive coral reef ecosystems [ 132 ]. Osman et al. [ 129 ] research offered a fresh understanding of the coral microbiome’s exclusive and endemic characteristics along the northern Red Sea refugia. They looked into the surface mucus layer (SML) for bacterial communities from six dominant coral species and discovered five novel algal endosymbionts. Over the past four decades, the average annual sea surface temperature in most of the world’s tropics and subtropics has risen between 0.4 °C and 1 °C. However, in the central Red Sea, where reef growth and scleractinian coral diversity are abundant, warming is more extensive than the observed mean tropical temperature increase [ 133 ]. The 2010 “Thuwal bleaching” in the central Red Sea was caused by a temperature rise of 10–11 °C, the largest coral bleaching event ever recorded. Furby et al. [ 131 ] conducted a survey and found that the “Thuwal bleaching” event caused more severe bleaching of inshore reefs (74% of hard corals were bleached) than offshore reefs (14% of hard corals were bleached). One mechanism that can lead to higher tolerance is repeated exposure to thermal stress [ 134 , 135 ]. Based on current knowledge, it is hypothesized that the reefs in the Red Sea will be relatively resistant to bleaching as sea temperatures rise, as noted in a study by Grimsditch and Salm [ 136 ]. However, reports indicate that bleaching is beginning to occur in the Red Sea, as documented by Kleinhaus et al. [ 137 ]. For instance, Rich et al. [ 138 ] reported a winter bleaching event in the central Red Sea in January 2020 due to sea surface temperatures (SSTs) falling below 18 °C. Additionally, inshore bleaching events in the central Arabian Red Sea were observed during the “3rd global coral bleaching event” in 2015, as reported by Monroe et al. [ 139 ].

The Eastern Tropical Pacific (ETP) comprises the ocean basin extending from the Gulf of California in México to Peru and includes areas of the continental shelf and offshore islands (Coco Island, the Galápagos Islands, the Revillagigedo Archipelago and Clipperton Atoll). The most cited article of the cluster “Eastern Pacific” is Spencer [ 140 ], which discussed potentialities, uncertainties and complexities in the response of coral reefs to future sea-level rise of reef islands in the Pacific Ocean and the Caribbean Sea. Throughout the Holocene, sea levels rose without being stabilized, and reefs in the Caribbean grew in tandem with these elevation changes [ 141 ]. The once structurally complex coral reefs in the Caribbean have suffered a dramatic decline since the 1970s, with only a minority of reefs maintaining a mean live coral cover of 10% or more [ 142 ]. A strong hurricane season brought on by unusually warm waters in the tropical Atlantic, and the Caribbean in 2005 caused the worst bleaching event ever observed in the basin [ 143 ]. There was a 60% decline in coral cover on reefs in the US Virgin Islands due to the fact of a severe disease outbreak brought on by the 2005 bleaching events in the Caribbean region, as reported by Miller et al. [ 144 ].

The Great Barrier Reef is the largest coral reef ecosystem, with over 348,000 km 2 of coverage consisting of 2900 individual reefs and 900 islands stretching over 2300 kilometers [ 145 ]. Three major coral bleaching events within a span of five years (2016, 2017, and 2020) along with the effects of severe tropical cyclones, poor water quality from catchment run-off, population growth and urbanization, overexploitation of marine resources, and habitat loss have all been the factors towards the degradation of coral reefs in the Great Barrier Reefs [ 56 , 146 ]. Cluster #4, which is the fifth largest cluster, contains 176 publications and has a silhouette value of 0.9. This cluster is strongly associated with cluster #6, “symbiotic dinoflagellate,” and cluster #12, “coral disease”. The high silhouette value of 1.0 indicates that there is a focused field of study in the context of coral reefs and climate change. The most cited article in cluster #4 is by Reaser et al. [ 147 ] on scientific findings and policy recommendations for coral bleaching and global climate change. Coral bleaching, which is the ability of animals with a symbiotic relationship with Symbiodinium to turn white, is an important issue associated with climate change-based literature. According to Douglas [ 148 ], all animals that have a symbiotic association with the dinoflagellate algae of the Symbiodinium genus, which are also referred to as zooxanthellae, have the ability to undergo bleaching. Symbiodinium have been reported to form extracellular symbioses with giant clams and intracellular symbioses with various organisms, including corals, anemones, jellyfish, nudibranchs, ciliophora, foraminifera, zoanthids, and sponges [ 149 , 150 ].

Fujise et al. [ 151 ] reported that the expulsion mechanisms of Symbiodinium were temperature-dependent; however, under non-thermal stress conditions, the expulsions of this algae were part of a regulatory mechanism to maintain a constant Symbiodinium density. In response to moderate thermal stress, Symbiodinium becomes damaged, and corals either selectively digest or expel the damaged cells. During extended periods of thermal stress, damaged Symbiodinium may accumulate in coral tissues, resulting in coral bleaching. Multiple factors have been shown to cause bleaching, including high oxidative stress [ 152 ], intense light [ 153 ], high temperature [ 154 ], low salinity [ 155 ], sedimentation [ 156 ], pollutants [ 157 ], decreased seawater temperature [ 158 ], diseases [ 159 ], bacterial infection [ 160 , 161 ], and ENSO-related marine heatwave events [ 162 , 163 ]. Degree heating weeks (DHW), defined as 1°C above the long-term climate level for the warmest month at a given locality, have become a common global predictor of bleaching [ 164 ]. Severe bleaching is typical at 8 DHW and above [ 165 , 166 ]. A global analysis report of coral bleaching from 1998 to 2017 [ 166 ] found that coral bleaching was most prevalent in regions with high-intensity and high-frequency thermal-stress anomalies. In areas where sea-surface temperature (SST) anomalies varied greatly, such as the Gulf of Aqaba region [ 167 ], the Caribbean Sea [ 168 ], and the Indo-Pacific [ 169 ], coral communities were significantly less susceptible to coral bleaching [ 166 ].

Globally, coral reefs have been threatened by coral disease, which is now recognized as one of the biggest threats to these ecosystems [ 170 ]. Similar to bleaching, coral disease was not considered a severe threat to coral reefs until recently [ 170 ], despite its first documentation in 1965 [ 171 ]. Since their initial descriptions, both the variety of coral diseases and the number of reported cases have skyrocketed [ 172 , 173 ]. Approximately 76% of all coral diseases described worldwide are found within this relatively small basin, leading experts to label the Caribbean a “hot spot” for disease [ 174 ]. For example, two dominant Acropora species in the Caribbean have been replaced by low-encrusting Agaricia due to the fact of coral disease [ 175 , 176 ]. Common coral diseases include Black band disease, which is caused by increased seawater temperature and anthropogenic factors, ciliates cause the Brown band disease, Cyanobacteria cause the Red band disease, and the White plague is caused by a bacterial infection [ 177 ].

5.5. Timeline Co-citation Analysis

The timeline for the document co-citation analysis is an important indicator to explain the period when the study got the attention of the researcher worldwide ( Figure 8 ). From 2010 to 2021, there have been bursts in citations for research clusters on (#0) “coral reefs”, (#2) “ocean acidification”, (#3) “central sea”, (#11) “sea level rise”, and (#5) “eastern pacific”. When taken as a whole, these studies shed light on the growing interest in studying the effects of ocean acidification and sea level rise on coral reefs, with particular attention paid to the plight of these ecosystems within the eastern pacific area, such as in the central Red Sea and the Dominican Republic.

A timeline co-citation analysis. Nodes represent references, whereas lines represent connections between those references. Larger nodes indicate higher frequencies of citations. References with strong citation bursts are shown as red circles, whereas references with high centrality are shown as yellow circles. Longer line segments indicate longer time spans.

5.6. Highly Cited Articles in the Field

CiteSpace’s visualized analysis of 7743 publications yielded a co-citation network (frequency of two different documents are cited together in other documents) with 2525 cited documents (nodes) and 5440 links or connections indicating co-citations between nodes [ 178 ]. The larger the node, the more often a document is cited, demonstrating its impact on coral reef and climate change research. Document co-citation analysis locates essential literature. The given references (in the article) were the most cited among 7743 Web of Science references. Table 4 presents the twenty most cited references based on co-citation analysis along with their frequency, burstiness, and centrality indices. An increase in citations reflects increased interest in that topic. “Citation bursts” demonstrate correlations between publications and sudden increases in citations. When comparing clusters, the centrality index indicates how well they are connected (i.e., coral reef and climate change). An elevated centrality score indicates that the publication is located between two or more sizable subclusters [ 179 ]. Dr. Terry Hughes’s research was widely cited, with three of his Nature and one of his Science publications ranking among the top five most-cited references. The publication entitled Global warming and recurrent mass bleaching of corals topped the list with a frequency index of 546, a burst index of 157 and a centrality index of 0.7. The burst in citation period of this publication was from 2018 to 2021. The findings were based on the third global-scale pan-tropical coral bleaching episode that occurred between 2015 and 2016. The reef ecosystem of eastern and western Australia was studied using aerial and underwater surveys along with sea surface temperatures obtained from satellites. According to their findings, the devastating bleaching event in 2016 was only slightly impacted by water quality and fishing pressure, indicating that local reef protection offers little to no protection against extreme heat.

The second paper on the list, with a frequency index of 359 and burst index of 125, was a global study analyzing the bleaching records of 100 globally distributed reefs from 1980 to 2016 [ 28 ]. According to their findings, mass coral bleaching events happen every year regardless of the presence or absence of El Nino. They forecast that the intervals between recurrent events will eventually become too short to permit a complete recovery of mature coral assemblages, typically taking 10 to 15 years to reach the fastest-growing species. They warned that if temperatures rise by 1.5 or 2 degrees Celsius above preindustrial levels, it will exacerbate the already severe decline of coral reefs around the world. Similar findings were found in another study of his that also appeared on the list of the most-cited research. Research into the effects of climate change on coral reef ecosystems, with a special emphasis on the Great Barrier Reef, ranked fifth [ 28 ]. They found that the Great Barrier Reef’s 2016 record-breaking heatwave had caused widespread loss of functionally diverse corals across the reef’s most remote and pristine regions. Ranked third on the list was the study by Hoegh-Guldberg et al. [ 22 ], which investigated the effects of climate change and ocean acidification on coral reefs. This study was closely linked to the 3rd (#2) research cluster, also known as “ocean acidification”. The research review presented future scenarios for coral reefs, which suggested increasingly detrimental impacts on various sectors, including tourism, coastal protection, and the fisheries industry. These predictions were based on the assumption that global temperatures would rise by at least 2 °C between 2050 and 2100, coupled with atmospheric carbon dioxide concentrations exceeding 500 ppm. The findings of this study emphasize the urgent need for effective measures to mitigate climate change and ocean acidification to ensure the long-term survival and sustainability of coral reefs and the associated ecosystems. The article by LaJeunesse et al. [ 180 ] on coral endosymbionts has garnered significant attention, with a citation frequency of 134, a burst index of 68, and a burst period spanning from 2013 to 2017. This publication is associated with cluster (3), also known as “symbiotic dinoflagellate”. The article describes Symbiodinium clades and proposes that the divergent evolutionary Symbiodinium “clades” correspond to genera within the Symbiodiniaceae family. The study affirms that the long evolutionary history of the Symbiodiniaceae family is appropriately acknowledged within the suggested framework. The findings of this study provide valuable insights into the evolutionary relationships and ecological functions of these endosymbionts, highlighting the critical role they play in the health and survival of coral reefs.

The list of 11 to 20 top-cited articles on climate change on coral reefs cover a wide range of topics, including ocean acidification, declining coral cover, Symbiodinium diversity, and coral reef resilience. Several studies indicate that warming trends and bleaching stress are increasing, and coral bleaching protection mechanisms are becoming less effective, ultimately leading to significant declines in coral populations. Research on the impacts of ocean acidification and warming on marine organisms, as well as the interactions between these factors, has also shed light on the mechanisms underlying the sensitivity of coral reefs to climate change [ 79 , 181 , 182 ]. Studies on the diversity, distribution and stability of Symbiodinium [ 183 ] have provided insights into the potential for coral resilience, while research on the decline of coral cover in the Indo-Pacific region has highlighted the extent and timing of this phenomenon [ 183 , 184 ]. These studies demonstrate that collaborative research efforts are essential to understanding the impacts of climate change on coral reefs and developing more efficient conservation and management strategies.

5.7. Distribution of Keywords

Using the co-cited keyword analysis performed in CiteSpace, 963 unique keywords were generated. In order to better understand the connections between these terms, a clustering tool was used to categorize them into groups ( Figure 9 ). This generated seven major clusters consisting of “sea surface temperature”, “Symbiodinium”, “coral reef fish”, “marine protected area”, “water quality”, “ocean acidification”, and “hydrocorals”. Each of these clusters can be analyzed independently to determine which descriptors are most applicable. The major keywords used to discuss “sea surface temperature”—record, Indian ocean, and reef; “Symbiodinium”—scleractinian coral, diversity, zooxanthellae, population, nutrient enrichment, elevated pressure, and oxidative stress; “coral reef fish”—phase shift, ecosystem, disturbance, fish, dynamics, community, recruitment, abundance, thermal tolerance, Stylophora pistillata , and ecology; “marine protected area”—management, assemblage, biodiversity, susceptibility, degradation, adaptation, response, and recovery; “water quality”—sea level, Great barrier reef, rate, transport, coral bleaching, French Polynesia, and fringing reef; “ocean acidification”—climate, El Niño, impact temperature, coral reef, calcification, and carbon; and “hydrocorals”—seawater and carbon dioxide.

Distribution of co-cited keywords in the field of coral reef and climate change.

Table 5 displays the top 10 keywords with the strongest citation burst. With the exception of “ocean warming,” all the most frequently cited keywords emerged in the early 1990s and experienced a citation burst that extended until the late 2000s. The keywords identified were “French Polynesia,” “record,” “El Niño,” “Australia,” “Indian Ocean,” “Continental,” “Shelf,” “Sea level,” “Sea surface,” “temperature,” and “Island.” Notably, the keyword “ocean warming” only gained popularity in 2017, with citations peaking from 2018 to 2021. This demonstrates the significance of research on climate-related temperatures in the field of coral reefs and climate change. The keywords “Australia” and “Continental shelf” demonstrated citation bursts lasting over 15 years. In contrast, “French Polynesia” had the highest frequency of citations during a relatively shorter period, commencing in 1992 and concluding in 2006. French Polynesia, situated in the westernmost region of the South Pacific, comprises 118 islands and atolls, classified into five main clusters: the Marquesas, Society, Tuamotu, Gambier, and Austral islands [ 191 ]. These regions exhibit a north–south gradient for variables such as sea surface temperature (SST), solar insolation, evaporation, and humidity. The Millennium Coral Reef Mapping Project (MCRMP) has successfully mapped the Austral, Gambier, Society, and Tuamotu islands and atolls; however, significant research remains to be undertaken in this extensive region, which accounts for the enduring citation burst for this keyword. The findings of this study highlight the scientific interest and importance of French Polynesia as a unique and diverse region for further research and conservation efforts.

Top 10 keywords with the strongest citation bursts.

5.8. Dual Map Overlay

Figure 10 illustrates the dual-map overlay of the number of articles pertaining to the type or focus of the journal. The map labels represent the research subjects covered by the journals, with the citing journals displayed on the left side and the cited journals on the right. The trajectory of the citation links provides valuable insights into inter-specialty relationships. A shift in trajectory from one region to another would indicate the influence of articles from another discipline on a specific field. In the domain of coral reef and climate change interaction, the dominant fields were found to be “ecology, earth, and marine”. The most influential discipline was “plant, ecology, and zoology”, with a z-score of 7.66, followed by “earth, geology, and geophysics”, with a z-score of 4.99 and, lastly, “molecular, biology, and genetics”, with a z-score of 2.80. These findings provide a valuable understanding of the interdisciplinary relationships within the field of coral reef and climate change research, highlighting the influence of various disciplines in shaping the current research landscape.

Dual-map overlay on the impact of climate change on coral reefs research.

6. General Discussions

Coral reefs are a vital marine ecosystem service, providing high biodiversity and supporting the livelihoods of coastal communities. However, ocean warming and temperature are the largest threats to corals from anthropogenic climate change [ 192 ]. Between 1997 and 2018, the global average percentage of coral cover was approximately 32%, but by 2100, RCP 8.5 predicts a global decline in coral cover of 5 and 15%, equating to a relative global decline of more than 40% [ 115 , 193 ]. This decline is due to the fact that sea surface temperatures (SSTs) are projected to increase by more than 3 °C by the turn of the century [ 194 ]. These declines could have significant ecological and socioeconomic impacts, particularly in coastal communities that rely on coral reefs for food, tourism, and other ecosystem services.

For example, The Republic of Palau, a small Micronesian nation, has already experienced significant losses in coral reef cover [ 195 ]. Over 87% of Palau’s households are linked to coral reef-associated activities, which are critical to the country’s economic and social well-being. While tourism, particularly ecotourism, is a significant contributor to GDP growth, tax revenue, and employment, climate change-related stressors have caused a steady decline in coral reef cover. This decline has indirectly caused a major decline in tourism, threatening the country’s economic sustainability [ 196 ]. According to Barnett [ 197 ], climate change is a significant threat to food security for people in Pacific SIDS, primarily due to the decline in fisheries output resulting from the impact of climate change on total coral cover.

Apart from impacting the socioecological structure, the impact of climate change can have cascading effects on the entire reef ecosystem, affecting the abundance and diversity of other marine species that depend on corals for food and shelter. Up to 14% of species may be in imminent danger of extinction at a warming of 1.5 °C and up to 29% at a warming of 3 °C. This rise in ocean temperature will probably force coral to colonize higher latitudes that currently lack reefs [ 198 , 199 , 200 ]. However, various factors, including the need for a suitable substrate [ 201 ], connectivity to other reefs [ 202 ], ocean acidification [ 203 ], and light intensity [ 204 ], may outweigh the advantages of reefs as they expand to high latitudes [ 193 ].

Through the timeline co-citation analysis, we have observed a significant increase in research interest in the topic of climate change impacts on coral reefs between 2010 and 2022. The analysis identified several research clusters that gained traction in the scientific community, including those related to “coral reefs,” “ocean acidification,” “central red sea”, “Great Barrier Reef”, and “sea level rise”. These clusters have evolved to become research hotspots under the overarching topic of climate change impacts on coral reefs. For example, the research clusters related to the central red sea and the Great Barrier Reef have emerged as prominent research areas, given their unique characteristics and ecological importance. Similarly, the impact of ocean acidification and sea level rise on coral reefs has gained significant research interest, given their severe consequences on the health and survival of coral reefs. To better understand the impact of these research clusters, our overall discussions have been designed to incorporate the subtopics “climate change threats to coral reefs” and “adaptive strategies for coral resistance and resilience”.

6.1. The Threat of Climate Change to Coral Reefs: Investigating the Impacts of Temperature and Ocean Acidification

Climate-induced changes in temperature are a major threat to coral reef ecosystems, and extensive research has highlighted several key areas for investigation [ 53 , 205 ], with marine heatwaves, solar radiation, heat tolerance, and thermal thresholds representing the most promising areas for future research. Marine heat waves have become increasingly prevalent and intense as a result of climate change. These extreme events, characterized by prolonged periods of elevated water temperatures, significantly impact coral reef ecosystems. For instance, the mass global coral bleaching event of 2016–2017 was the most extensive and long lasting on record, as documented by Eakin et al. [ 206 ]. The event, which was associated with the El Niño Southern Oscillation (ENSO), had varying impacts on coral reefs worldwide [ 207 ], with some regions experiencing more severe bleaching than others, as reported by Kim et al. [ 208 ].

Corals are thermophilic, but their thermal tolerance is narrowly defined [ 169 , 209 ]. For instance, the rate of calcification increases with temperature up to a threshold level, beyond which it declines [ 210 , 211 , 212 ]. Tropical corals live close to their upper thermal limits and are, therefore, highly sensitive to periods of elevated sea surface temperatures and ocean warming [ 187 , 213 ]. Coral reefs in the Persian Gulf have been observed to have the highest upper-temperature thresholds of approximately 35–36 °C [ 214 ]. However, it has also been noted that these corals remain highly vulnerable to thermal stress when temperatures surpass their local maximum summer temperatures [ 215 ]. The escalating frequency and gravity of thermally induced mass bleaching events have sparked worldwide attention to the elevated temperature impacts on corals [ 28 ]. As a result, research endeavors have focused on establishing maximum thermal tolerance thresholds and variations in diverse coral species and regions and exploring potential coral refugia to brace for future ocean warming [ 216 ].

Corals rely on their symbiotic relationship with unicellular algae of the genus Symbiodinium for photosynthesis, and over 90% of their energy budget is needed for essential functions, such as calcification, tissue growth, and reproduction [ 212 ]. This critical association is threatened when corals experience thermal stress, such as elevated sea surface temperatures (SST), resulting in coral bleaching, where the algal endosymbionts are expelled. The resulting impairment and expulsion of the algal symbionts are linked to reactive oxygen species (ROS) generation from the host, the algal symbiont, or both, triggering a host immune response [ 217 ].

Protracted coral bleaching can lead to extensive coral mortality, severely affecting the ecosystem and associated reef fauna. Based on the timeline cocitation analysis, it was evident that the Red Sea (Cluster #3) and Great Barrier Reef (GBR) (cluster #8) are major research hotspots in terms of geographic regions. Although the Persian Gulf is a hot sea that supports coral reef ecosystems, the Red Sea harbors corals with greater thermal stress tolerance, with some coral genotypes capable of surviving temperatures over 5 °C above their summer maxima [ 216 , 218 ]. Corals in the southern end of the Red Sea are more heat resistant, surviving prolonged high temperatures, while the northern Red Sea benefits from heat-resistant genotypes that have migrated from the south [ 219 ]. The importance of broad latitudinal temperature gradients in promoting adaptation to high temperatures and exchanging heat-resistant genotypes across latitudes for genetic rescue in coral reefs is exemplified in the evolutionary history of coral reefs in the northern Red Sea [ 9 , 216 ]. On the other hand, the GBR, known as the world’s largest coral ecosystem, was severely impacted by the 2015–2016 climate change-amplified strong El Niño event that triggered the warmest temperatures on record. This resulted in a massive bleaching event affecting nearly 90% of reefs along the northern region, leading to a loss of approximately 30% of live coral cover in the following six months [ 28 , 220 , 221 ]. Research has increasingly linked climate change to a rise in coral diseases. Bruno et al. [ 222 ] used a high-resolution satellite dataset to investigate the relationship between temperature anomalies and coral disease on a large spatial scale of 1500 km in Australia’s Great Barrier Reef. Their findings showed a significant positive correlation between warm temperature anomalies and the incidence of the white syndrome, an emergent disease in Pacific reef-building corals. In a similar vein, Tignat-Perrier et al. [ 223 ] noted a decline in populations of two gorgonian species ( Paramuricea clavata and Eunicella cavolini ) found in the Mediterranean Sea due to the fact of microbial diseases during thermal stress events. These studies illustrate the growing concern that climate change is contributing to the increased incidence and severity of coral diseases, which could ultimately lead to a decline in the health of marine ecosystems.

In the past, studies on the impact of climate change on coral reefs primarily centered on the thermal tolerance of corals and the consequences of massive, abrupt coral loss on organisms associated with reefs [ 224 ]. However, research has recently shifted towards investigating the distinct and synergistic effects of ocean warming and ocean acidification resulting from increased atmospheric CO 2 levels. The timeline co-citation analysis reveals that these emerging research fields are highly significant with recent citation bursts, as evidenced by their identification as Cluster #2 (Ocean acidification) and Cluster #10 (Elevated CO 2 ), respectively.

The escalation of atmospheric carbon dioxide (CO 2 ) concentrations has resulted in ocean acidification, which is among the foremost threats to coral reef ecosystems. Forecasts for 2100 anticipate a rise in CO 2 concentrations to between 540 and 970 ppm, leading to a global decrease in seawater pH by 0.14 to 0.35 units [ 31 , 68 , 116 , 225 ]. As demonstrated by Fabricius et al. [ 68 ], ecological traits of coral reefs will gradually transform as seawater pH decreases to 7.8, and a decline below this level (at 750 ppm pCO 2 ) would be catastrophic for these ecosystems. Ocean acidification reduces the availability of carbonate ions that corals require to form their calcium carbonate skeletons, ultimately leading to a decrease in coral calcification rates [ 33 ]. Ocean acidification has also been shown to decrease the ability of coral larvae to settle and survive [ 226 ] and increase their susceptibility to disease [ 227 ]. Research has shown that even modest increases in ocean acidity can impact the physiological processes of corals. For example, exposure to high levels of CO 2 reduces coral growth and calcification rates [ 68 , 226 ]. In addition to the direct effects on coral physiology, ocean acidification can have cascading impacts on the entire coral reef ecosystem. For instance, reduced calcification by corals can reduce the complexity of the coral reef structure, potentially leading to the loss of important habitats for fish and other marine organisms [ 228 ]. Furthermore, ocean acidification can impact the symbiotic relationship between corals and their algal symbionts, potentially leading to a decline in the productivity of the reef ecosystem as a whole [ 229 ]. The combination of ocean warming and acidification is particularly concerning, as they act synergistically to exacerbate the negative impacts on coral reef ecosystems [ 22 ]. With continuing increases in atmospheric CO 2 levels, the effects of ocean acidification on coral reefs are expected to become even more pronounced, highlighting the need for urgent action to reduce greenhouse gas emissions and protect these valuable and vulnerable ecosystems.

The rate of atmospheric CO 2 increase continues to accelerate, with emission scenarios predicting CO 2 concentrations of 540–970 ppm and a decline in seawater pH by 0.14–0.35 units globally for 2100 [ 68 , 225 ]. Fabricius et al. [ 68 ] demonstrated that many ecological properties in coral reefs will gradually change as pH declines to 7.8 and that it would be catastrophic for coral reefs if seawater pH dropped below 7.8 (at 750 ppm pCO 2 ).

6.2. Adaptive Strategies for Enhancing Coral Resistance and Resilience in the Face of Climate Change

Coral resistance and resilience are scientific constructs that pertain to the capacity of coral reefs to withstand and recuperate from various stressors. Coral resistance is defined as the ability of corals to endure or tolerate perturbations and stressors, such as variations in water temperature, ocean acidification, pollution, and physical injury. Corals that possess a greater resistance to these stressors exhibit a greater ability to sustain their structure and function despite disturbances and are less prone to suffering from coral bleaching, disease, or mortality [ 229 , 230 ]. A myriad of studies has reported on the bleaching thresholds of corals inhabiting the Persian Gulf, despite conditions at least 2 °C higher than other coral reef ecosystems worldwide [ 231 ]. Additionally, corals from the Indo-Pacific and Caribbean regions have been found to maintain calcification rates even in low aragonite saturation states, present in naturally acidified locales [ 68 , 232 ]. The eastern Pacific region of Palau has revealed the thriving of reefs in waters with natural acidification, resulting from biological processes and reef system circulation patterns [ 232 , 233 ]. However, it is noteworthy that coral communities in Palau’s relatively acidic reef zones developed over thousands of years, fostering an inherent resistance that differs from coral communities in regions affected by higher anthropogenic interventions.

Coral resilience, in contrast, refers to the ability of coral reefs to recover from disturbances and stressors. Corals that exhibit higher resilience can reproduce, regenerate, and rebuild their structural complexity after experiencing bleaching [ 234 ]. These mechanisms are attributable to genetic diversity within coral populations and their symbiotic association with Symbiodinium algae, which are critical to their health and survival [ 235 , 236 ]. Genetic adaptation in corals is mediated through various factors, including the activation of heat-shock proteins, oxidoreductase enzymes, and microsporine-like amino acids. The coral surface micro-layer that absorbs UV radiation has also been identified as a significant mechanism for adaptation [ 180 , 237 , 238 ]. In-depth research on corals that thrive in the warm waters of the Persian Gulf has demonstrated their capacity for resilience, attributable to metabolic trade-offs, unique physiological characteristics, and specific genetic signatures, including a heat-specialist algal endosymbiont, Symbiodinium thermophilum [ 236 , 239 ]. S. thermophilum can thrive in high-temperature and high-salinity environments, allowing the coral to develop a temperature-stress-resistant phenotype [ 239 ].

Symbiodinium, a diverse group of dinoflagellates, is classified into nine clades (A–I) based on their phylogenetic characteristics [ 240 ]. Among these clades, Symbiodinium clade D has garnered attention for its exceptional thermal resilience ability, despite its relatively low representation (less than 10%) in the endosymbiotic community of coral hosts [ 241 ]. Various coral species, including fast-growing branching types, such as Acropora, Stylophora, and Pocillopora, as well as slow-growing massive, encrusting, and solitary corals, have been associated with Symbiodinium clade D [ 242 ]. The prevalence of clade D Symbiodinium in corals from the Persian Gulf has been linked to their higher thermal tolerance, particularly in comparison to corals associated with clade C, which is the dominant lineage in corals from the Great Barrier Reef and other Pacific coral reef ecosystems [ 243 ], and clade B in corals from the Atlantic [ 244 ]. These findings highlight the significance of Symbiodinium diversity in understanding the thermal resilience of coral reefs and the potential mechanisms underlying their adaptation to changing environmental conditions.

McCulloch et al. [ 234 ] explored the ability of coral species to withstand the adverse impacts of ocean acidification and global warming on coral reefs. Their study revealed that some coral species (i.e., Stylophora pistillata and Porites spp.) exhibit the capacity to increase pH levels within their calcifying fluid, crucial for the deposition of calcium carbonate and maintenance of the coral structure, even in the face of declining seawater pH levels. The study demonstrated the significance of acid-base regulation mechanisms for corals’ resilience to the effects of ocean acidification, allowing them to maintain or increase their calcification rates despite rising ocean acidification. Moreover, the study indicated that corals could acclimate to extended acidification, which enables them to maintain or increase their calcification rates by upregulating their internal pH levels, thus providing insight into potential strategies for mitigating the effects of climate change on coral reefs. A similar adaptation resilience strategy against ocean acidification was observed in cold-water scleractinian corals (i.e., Caryophyllia smithii , Desmophyllum dianthus , Enallopsammia rostrata , Lophelia pertusa , and Madrepora oculate ) [ 245 ].

Oceanographic processes, such as upwelling and tidal currents, also play a significant role in helping corals avoid bleaching. In areas where upwelling events mix deeper, cooler water with shallow warmer water, thermal stress is reduced [ 246 , 247 ]; for example, in northern Galapagos during the 2015/16 ENSO [ 248 ] and Nanwan Bay, southern Taiwan, during summer [ 249 ]. Similarly, a coral reef’s ability to resist bleaching is bolstered by the elimination of potentially damaging oxygen radicals due to the swift water flow associated with tidal currents [ 230 , 250 , 251 ].

Therefore, in summary, the scientific community has identified various adaptive strategies that could enhance the resilience and resistance of coral reefs to these challenges. Going forward, it is crucial to continue ongoing research efforts to better understand the mechanisms underlying coral resilience and resistance, identify research gaps, and develop new management strategies for protecting these vital ecosystems. This can be achieved through a multidisciplinary approach that combines laboratory-based experimentation, field research, and community engagement. In addition, collaborations between the scientific community and policymakers can facilitate the implementation of evidence-based management practices that promote the resilience and resistance of coral reefs to climate change and other stressors.

7. Conclusions

The scientometric analysis that is presented in this article demonstrates that research on coral reefs in relation to climate change has emerged as one of the potential fields, with interest in this topic has grown steadily since the 2000s. The increasing global temperatures are posing a significant threat to coral reefs, leading to widespread coral bleaching and mortality. Moreover, changes in ocean chemistry brought on by an increase in carbon dioxide levels lead to ocean acidification, which can worsen the effects of rising temperatures on corals [ 22 ]. In addition, sea level rise and coastal development are transforming the physical structure of coral reef ecosystems, exacerbating the negative effects of the other stressors [ 252 ]. These changes are harmful not only to the coral reefs but also to the plethora of species that rely on them for survival and the communities that rely on them for livelihoods and for protecting the coast. Future challenges for developing countries like those within the coral reef triangle initiative (i.e., Indonesia, Malaysia, the Philippines, Papua New Guinea, Timor Leste, and the Solomon Islands) will center on access to funding for conservation-restoration efforts and continued monitoring studies. There are several ways that ongoing research and coordinated action can help coral reefs cope with the effects of climate change:

• Monitoring: Regular coral reef monitoring can reveal vital details about the well-being and state of the reefs as well as the effects of climate change. These data can be used to pinpoint especially vulnerable regions and monitor long-term changes. Scientists and environmentalists can detect early warning signs of coral bleaching and other detrimental effects by monitoring coral reefs, which enables them to take action before it is too late;
• Research: Collaborative research efforts can contribute to a better understanding of the impacts of climate change on coral reefs and the mechanisms underlying these impacts and can also aid in developing and rigorously testing intervention and restoration techniques for coral reefs. Given that the preservation of coral reef ecosystems requires a range of interventions, including biological, ecological, and social strategies for mitigation and adaptation [ 19 ], such research can help create more efficient restoration, conservation and management strategies;
• Conservation and management: Collaborative conservation and management efforts can assist in mitigating the effects of climate change on coral reefs. Protected areas and marine reserves, for instance, can aid in mitigating the effects of overfishing and pollution, thereby making coral reefs more resilient to the effects of climate change;
• Mitigation: joint efforts can also aid in lowering atmospheric greenhouse gas concentrations, which are primarily responsible for climate change, for instance, by collaborating with regional organizations and authorities to advance sustainable development and lower carbon emissions;
• Public education and awareness: Raising public understanding of the effects of climate change on coral reefs can encourage support for management and conservation initiatives.

Funding Statement

The present study was supported by the Department of Higher Education, Ministry of Higher Education Malaysia, under the LRGS program (LRGS/1/2020/UMT/01/1; LRGS UMT Vot No. 56040) entitled “Ocean Climate Change: Potential Risk, Impact and Adaptation Towards Marine and Coastal Ecosystem Services in Malaysia’. The work was also supported by PASIFIC program GeoReco project funding from the European Union’s Horizon 2020 Research and Innovation programme, under the Marie Sklodowska-Curie grant agreement No. 847639, and from the Ministry of Education and Science.

Author Contributions

Conceptualization, C.S.T. and M.N.A.; methodology, M.N.A.; software, C.S.T.; validation, J.B., Z.V.-G. and V.R.; formal analysis, F.L.; investigation, G.S.; resources, I.G.; data curation, J.B.; writing—original draft preparation, C.S.T.; writing—review and editing, M.N.A.; visualization, F.L., G.S., I.G., V.R. and Z.V.-G.; supervision, M.N.A.; project administration, G.S.; funding acquisition, J.B., Z.V.-G., V.R. and I.G. All authors have read and agreed to the published version of the manuscript.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Data availability statement, conflicts of interest.

The authors declare no conflict of interest.

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Research Article

A systematic review of artificial reefs as platforms for coral reef research and conservation

Roles Conceptualization, Formal analysis, Investigation, Methodology, Validation, Visualization, Writing – original draft, Writing – review & editing

Affiliation Department of Biology, Dalhousie University, Halifax, Nova Scotia, Canada

Roles Conceptualization, Funding acquisition, Methodology, Project administration, Resources, Supervision, Validation, Writing – review & editing

* E-mail: [email protected]

Affiliation Department of Oceanography, Dalhousie University, Halifax, Nova Scotia, Canada

• Emily Higgins, 
• Anna Metaxas, 
• Robert E. Scheibling

• Published: January 21, 2022
• https://doi.org/10.1371/journal.pone.0261964
• Reader Comments

Artificial reefs (ARs) have been used on coral reefs for ecological research, conservation, and socio-cultural purposes since the 1980s. We examined spatio-temporal patterns in AR deployment in tropical and subtropical coral reefs (up to 35° latitude) and evaluated their efficacy in meeting conservation objectives, using a systematic review of the scientific literature. Most deployments (136 studies) were in the North Atlantic and Central Indo-Pacific in 1980s – 2000s, with a pronounced shift to the Western Indo-Pacific in 2010s. Use of ARs in reef restoration or stressor mitigation increased markedly in response to accelerating coral decline over the last 2 decades. Studies that evaluated success in meeting conservation objectives (n = 51) commonly reported increasing fish abundance (55%), enhancing habitat quantity (31%) or coral cover (27%), and conserving target species (24%). Other objectives included stressor mitigation (22%), provision of coral nursery habitat (14%) or source populations (2%) and addressing socio-cultural and economic values (16%). Fish (55% of studies) and coral (53%) were the most commonly monitored taxa. Success in achieving conservation objectives was reported in 33 studies. Success rates were highest for provision of nursery habitat and increasing coral cover (each 71%). Increasing fish abundance or habitat quantity, mitigating environmental impacts, and attaining socio-cultural objectives were moderately successful (60–64%); conservation of target species was the least successful (42%). Failure in achieving objectives commonly was attributed to poor AR design or disruption by large-scale bleaching events. The scale of ARs generally was too small (m 2 –10s m 2 ) to address regional losses in coral cover, and study duration too short (< 5 years) to adequately assess ecologically relevant trends in coral cover and community composition. ARs are mostly likely to aid in reef conservation and restoration by providing nursery habitat for target species or recruitment substrate for corals and other organisms. Promoting local socio-cultural values also has potential for regional or global impact by increasing awareness of coral reef decline, if prioritized and properly monitored.

Citation: Higgins E, Metaxas A, Scheibling RE (2022) A systematic review of artificial reefs as platforms for coral reef research and conservation. PLoS ONE 17(1): e0261964. https://doi.org/10.1371/journal.pone.0261964

Editor: Maura (Gee) Geraldine Chapman, University of Sydney, AUSTRALIA

Received: September 10, 2021; Accepted: December 14, 2021; Published: January 21, 2022

Copyright: © 2022 Higgins et al. This is an open access article distributed under the terms of the Creative Commons Attribution License , which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited.

Data Availability: All relevant data are within the paper and its Supporting Information files.

Funding: Funding for this study was provided by the Natural Sciences and Engineering Research Council of Canada Discovery Grants program (RGPIN-2016-04878 to AM and RGPIN-2016-04878 to RES). EH was partially supported by scholarships from the Natural Sciences and Engineering Research Council of Canada CGS Masters, Nova Scotia Research and Innovation Scholarship, and the Faculty of Graduate Studies at Dalhousie. The funders had no role in study design, data collection and analysis, decision to publish, or preparation of the manuscript.

Competing interests: The authors have declared that no competing interests exist.

Introduction

The global cover of scleractinian corals has declined dramatically since 1985 due to synergistic effects of increased ocean temperatures and acidification, predation, biological invasions, mechanical damage, and disease [ 1 , 2 ]. The increasing frequency and intensity of natural and anthropogenic stressors has altered coral reefs, contributing to large-scale phase shifts, in some regions, to alternative stable communities dominated by fleshy macroalgae [ 3 , 4 ], soft corals, corallimorpharia, or sponges [ 5 ]. It is estimated that more than 800 million people worldwide depend on coral reefs for food, coastal protection, and tourism [ 6 – 8 ], and that persistence of alternative stable states will cause a significant reduction in these ecosystem services [ 9 ].

Traditional conservation measures (e.g. no take-zones, reserves, and marine protected areas) have been used on coral reefs for decades [ 10 – 12 ], but attention has progressively shifted toward active restoration methods as a consequence of accelerating coral decline [ 13 , 14 ]. Ecological restoration is the process which assists the recovery of a degraded, damaged or destroyed ecosystem [ 15 ]. Since it may not be possible to remove the threat responsible for degradation or damage, the trajectory of recovery may allow adaptation to local and global changes [ 16 ]. The United Nations General Assembly recognized the pressing need to restore damaged ecosystems and proclaimed 2021–2030 to be the United Nations Decade on Ecosystem Restoration, with the primary goal being to prevent, halt and reverse the degradation of ecosystems worldwide. The United Nations Environment Assembly adopted a resolution that requested UNEP to specifically better define best practices for coral restoration [ 17 ]. Since the main threat to coral reefs is climate change [ 18 ], their restoration is likely most effective as a complementary tool in a larger management portfolio or as a temporary measure to minimize loss while global solutions are sought [ 17 , 19 ]. However, restoration of coral reefs has lagged behind and the spatial extent of restoration is the smallest compared to other major marine coastal ecosystems [ 20 ]. Thus, our knowledge on best practices for coral reef restoration is limited.

Motivations for coral reef restoration have ranged from ecological to cultural to legislative reasons, but experimental reasons appear to dominate [ 21 , 22 ]. Experimental approaches to active restoration include direct transplantation of corals, coral gardening, larval propagation, substrate manipulation, and substrate addition through the deployment of artificial reefs [ 17 , 19 ]. All approaches of active restoration have had certain shortcomings, such as short monitoring periods (average = 18 months) and small scales (< 100 m 2 ) and have often lacked objectives [ 19 ]. Of these, artificial reefs (ARs), although popular for fish enhancement, have not been used as extensively for coral restoration [ 19 ], possibly because of the logistics of deployment and, on average, an order of magnitude greater cost than other approaches [ 20 ]. ARs have been deployed in coral ecosystems globally to address various conservation objectives, including enhancing fish and invertebrate biomass [ 23 ], increasing habitat quantity and structural complexity of denuded reefs [ 24 , 25 ], conservation of target species [ 26 , 27 ], and as nursery habitat for transplantation initiatives [ 28 ]. Examining the objectives of artificial coral reefs, success in meeting these objectives, and assessing their potential benefits as a restoration strategy can inform management decisions in different regions and under projected climate scenarios. However, for management decisions to be effective, the benefits of AR must be quantified and the efficacy of the methodologies (e.g. AR type, size, distribution, deployment location and period) evaluated.

ARs deployed in different temperate and tropical ecosystems can provide benefits to both benthic and pelagic communities [ 29 ] by supplying additional hard substrate for settlement [ 30 ], reducing fishing and tourism pressure on natural reefs [ 31 ], increasing heterogeneity of natural substrata [ 32 , 33 ], and providing shelter from predators and human disturbances [ 34 , 35 ]. As with other active restoration approaches, clearly defined objectives for the deployment of ARs are not always provided [ 29 ], presenting challenges with monitoring their effectiveness. There is also concern that the scale of ARs is too small to have long-term impacts on conservation or restoration of target species and their functional relationships [ 36 ]. It has been argued that ARs can introduce alien materials onto reefs that may harm the recipient community by leaking toxic compounds [ 37 ] or by scouring natural reefs if detached during coastal storms [ 38 ]. Additionally, there is debate as to whether ARs act as a source or sink for fish and invertebrate populations [ 39 – 42 ].

To assess the functional importance of ARs, an understanding of the dynamics of established benthic communities and their relationship with demersal and pelagic species is imperative [ 35 ]. Deploying ARs for restoration of coral ecosystems specifically is a relatively new strategy, and most research to date has been largely descriptive [ 43 ], with few replicated comparisons to natural reefs [ 44 ]. For example, there is increasing evidence that fish and invertebrate assemblages on ARs deployed in coral ecosystems do not mimic those on natural reefs [ 45 – 47 ]; the role of ARs in colonization by reef invertebrates is unknown [ 35 ]. Long-term data on species’ residence time, growth and survival, and production patterns on adjacent natural coral reefs rarely are collected during studies of ARs [ 34 , 40 ].

Planning AR deployments in coral ecosystems with specific goals and objectives coupled with long-term monitoring plans can allow the assessment of conservation outcomes from these interventions [ 17 , 19 , 29 ]. Here, we present results of a systematic review of the scientific literature focussed specifically on the use of ARs as an active restoration strategy for coral ecosystems. In particular, we examine stated objectives of ARs over the past 100 years and across 8 marine realms, along with records of the spatial scale, monitored taxa, and study duration. For studies that recorded progress toward meeting conservation objectives, we evaluate and discuss the reported success, and identify factors that may limit the attainment of objectives. Based on our findings, we propose that among all prospective conservation objectives for artificial coral reefs, the provision of nursery habitats and additional hard substrate for colonization, and the promotion of local socio-cultural values are those most likely to achieve conservation success. However, given the limited evidence of setting conservation objectives specific to deployment, the large variation in size, spacing and monitoring effort, and the potential cost, much more research is needed to assess the use of ARs as a coral restoration strategy globally.

Literature search and data extraction

We conducted searches in ISI Web of Science Core Collection (1900–2020), Scopus ( https://www.scopus.com ), and Google Scholar ( https://scholar.google.ca ) for peer-reviewed publications that measured or monitored ecological and socio-cultural variables on ARs deployed in tropical and subtropical coral reef ecosystems (up to 35° latitude). In each database, we adapted the following general search terms to account for syntax differences: (TITLE-ABS-KEY ((artificial* OR “man-made” OR construct*) W/2 (coral* OR reef* OR habitat* OR nursery*)) AND TITLE-ABS-KEY (coral* OR tropic* OR subtropic*)). The first two sets of search terms were optimized to return studies that incorporate AR structures that both were designed purposefully and became de facto ARs. The last set narrowed the scope of the search to articles pertaining to ARs deployed in coral ecosystems. Studies on both vertebrate and invertebrate groups were included. Searches in all databases were completed on 31 December 2020. To ensure scientific rigour in the assessment of conservation objectives, we did not include the 1000s of studies from the grey literature, the validity of which had not been evaluated through peer review.

Over all databases, the search terms returned 4088 articles after duplicates were removed. All article citations and abstracts were imported into the web-based software review program Covidence ( https://www.covidence.org ), their titles and abstracts were screened, and 802 studies were extracted that included research on AR structures in coral reef ecosystems ( Fig 1 ). A full text review was conducted for 530 articles, and data were extracted from 136 that met one or more of the following secondary inclusion criteria: 1) included a date, precise location, and depth of deployment, 2) included the precise dimensions and number of ARs in the study, and 3) stated an objective of AR deployment.

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Articles were divided into two categories: 1) those that directly measured the success of meeting the objective(s) of ARs, and 2) those that were deployed for the purposes of scientific experiments or as de facto submergences (e.g. accidental ship groundings, dumping vehicles or building materials as waste). All 136 studies from both categories were surveyed for 1) duration of study, 2) clear description of AR dimensions, 3) targeted taxonomic groups, and 4) socio-cultural and ecological response variables used to assess whether the conservation objective(s) of the AR was being met. Latitude and longitude were extracted for each AR and then categorized into marine realms as defined in [ 48 ]. All 136 studies were used to examine spatio-temporal patterns of AR deployments as presented in the scientific literature. For the analyses of global AR abundance over time and to ensure representation of definitions used in the studies, we included all structures clearly defined as ARs by study authors and with a minimum area of ≥ 0.25 m 2 .

To ensure ecological relevance of conclusions about conservation success, studies reporting on progress towards attaining conservation objectives of deployed ARs fulfilled all secondary inclusion criteria listed above as well as two additional ones: (1) the monitored ARs were ≥ 1 m 2 to allow for comparison with natural reef formations and knolls; and (2) for studies reporting on multiple ARs, individual AR structures were defined as such if they were at least 2 m from the nearest adjacent AR. This spacing reflects what is considered an AR by study authors and the methods used to ensure connectivity of motile organisms and larvae between ARs. It has been shown that ARs > 2 m apart can form distinct benthic communities [ 49 ]. A total of 53 studies fulfilled all inclusion criteria and were used in this analysis of conservation objectives.

Classification of deployment objectives and response variables

Studies monitoring the success of an AR towards achieving one or more conservation objectives were further sub-classified into 8 categories of objectives: increase fish abundance, increase coral cover, conservation of target species (i.e. reef species of significant ecological or socio-cultural importance), socio-cultural value (e.g. economic evaluation, attractiveness to divers or tourists), serving as a source population for recruitment to the surrounding ecosystems, nursery or coral garden, increase habitat quantity, and stressor mitigation (i.e. deployment following catastrophic events, such as bleaching, severe tropical storms, and dredging). The ecological response variables used to assess success in meeting the conservation objective(s) of ARs were categorized according to the measurements (abundance, diversity, cover, recruitment, biomass, size distributions, survival/mortality, growth and reproduction rates, species turnover, connectivity/space use, and structural complexity) and by broad taxonomic groups (fish, coral, other invertebrates, and algae).

AR deployments on coral reefs

Definitions of ar.

There is little standardization or agreement about the definition of AR in the scientific literature. Definitions within the studies examined in this review were disparate or absent. Authors reported on a vast array of structures, from de facto or accidental deployments to purposefully designed and deployed ARs. De facto or accidentally deployed ARs are wide ranging. Most are wrecks (or pieces of wrecks) of various numbers (15 in one case [ 50 ]), sizes and types of vessels; retired oil rigs [ 51 ], breakwaters and coastal jetties [ 52 ], and ropes in a tuna farm [ 53 ] were also considered ARs. Purposefully deployed ARs ranged from piles of rocks [ 54 ] or tires on the seafloor [ 55 ] to specifically engineered structures optimized for recruitment of target species for conservation, such as casitas or Autonomous Reef Monitoring Structures [ 28 , 56 , 57 ]. This is a similar range in structures and materials as for all ARs and is not particular to tropical reefs [ 58 ]. We used a broad AR definition when examining spatio-temporal patterns of AR deployment to accurately characterize the wide variety of structures that are currently being categorized as ARs in the peer-reviewed literature.

There is also little consistency in AR area within the peer-reviewed literature. Deployments of ARs for conservation purposes were conducted on a larger scale than ARs deployed for scientific experimentation. Most ARs used in experimental studies (70%) were 1–5 m 2 ( Table 1 ), while more than a half (60%) of ARs with conservation objectives were > 150 m 2 ( Table 2 ). The small size in experimental studies likely reflects logistical constraints of monitoring large reef structures in scientific experiments or of experimentally controlling and disentangling confounding abiotic effects of reef development on larger ARs [ 36 ]. Spacing between individual ARs is not well reported in studies examining structures with conservation objectives, which often neglect to distinguish between ARs and AR modules. Nearly all studies that monitored communities on de facto reefs reported that the structures were > 150 m 2 ; only two studies monitored response variables on ARs of smaller area ( Table 1 ).

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Spatio-temporal patterns in deployments of ARs

There were only 4 reports of AR deployments in the scientific literature until the mid-twentieth century. More than 2200 ARs were deployed in the 1960s, most in Hawaii ( Fig 2 ). Comparatively few deployments were recorded from the 1970s to the 1990s, with a greater than 2-fold increase from the 1960s in the 2000s, followed by a similar increase between the 2000s and the 2010s ( Fig 2 ). The increase in the 2000s corresponds to the increased focus on effects of climate change on coral reefs in the late 1990s following the first major global bleaching event in 1998 [ 59 ]. In the 2010s, > 10000 deployments were reported during a single study on the Indian shelf [ 60 ], resulting in the highest recorded number of ARs in coral ecosystems globally. These temporal patterns parallel those for ARs in other coastal ecosystems, reflecting a general global transition in AR research [ 61 ].

Inset shows marine realms defined by [ 48 ].

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Following the large number of AR deployments in the Western Indo-Pacific, the Tropical Atlantic region has the next greatest number of AR deployments to date, with a coral restoration effort in Antigua contributing substantially to the region’s deployments (~ 3500 ARs deployed in 2004). However, the high abundance of ARs from the Tropical Atlantic is biased by a high intensity of study and frequency of publication from the southern United States (particularly Florida) from the 1960s onward [ 61 , 62 ]. Florida has a long history of AR deployment, with reefs often made from cheap waste materials (tires, metal construction materials, automotive parts) or de facto structures (sunken vessels, planes) [ 63 ].

The Central and Eastern Indo-Pacific exhibit similar numbers of AR deployments ( Fig 2 ). AR deployments in the Eastern Indo-Pacific are attributed mostly to a single location (Hawaii), while in the Central Indo-Pacific they are distributed across several countries (e.g. Australia, Indonesia, Malaysia, Taiwan, Thailand, Vietnam) but largely concentrated in Indonesia. The regional interest of ARs in the Central Indo-Pacific may be a consequence of increasing exploitation of marine habitats [ 64 ] and the reliance of Southeast Asian countries on the economic value of ecosystem services associated with coral reefs (e.g. fisheries, tourism, shoreline protection) [ 65 ].

Scientific experimentation and de facto AR deployments

Studies reporting on ARs that did not have a direct conservation-oriented objective were classified as either scientific experimentation or de facto submergences ( Table 1 ) and were not included in our exploration of AR conservation objectives ( Table 2 ). Over one third of the studies examined in this review (53 of 136) reported on scientific experiments conducted on ARs and 42% of these were conducted in the Tropical Atlantic realm. Overall, studies addressing only scientific objectives were marginally shorter than conservation-oriented projects, with mean durations of 1.7 y ( Table 1 ) and 2.0 y ( Table 2 ), respectively.

ARs recorded in peer-reviewed literature and deployed in the 1920s – 1950s were unplanned ship groundings that later were observed to have an AR effect by attracting fish and invertebrate colonizers [ 66 ]. Research efforts on de facto reefs (22 of 136 studies) reflect largely opportunistic monitoring, with data most often collected through digital imagery and a few manipulative experiments ( Table 1 ). De facto ARs are the most variable in terms of study duration, ranging from 4 months to 11 years.

The Tropical Atlantic and Central Indo-Pacific realms have the highest number of AR deployments for scientific objectives or de facto deployments, with 414 and 402, respectively. In the past two decades, more than half (57%) of all such AR deployments are from the Central Indo-Pacific, a region which has experienced significant coral mortality since 1998 [ 4 , 67 ].

Conservation-purposed AR deployments

Conservation objectives of ars..

The three most-commonly cited conservation objectives of ARs were increasing fish abundance (55%), increasing habitat quantity (31%), and increasing coral cover (27%) ( Table 2 ). These conservation objectives were most common in the Western and Central Indo-Pacific, Tropical Atlantic and Temperate Australasia ( Fig 3 ). Many of these ARs are in countries with substantial government funding for research and conservation, notably the USA (Florida) and Israel. In the Central Indo-Pacific, ARs with conservation objectives were predominantly deployed in countries with well-established national programs for AR development (Thailand, Malaysia), as well as those which received international funding in response to reef decimation by the Indian Ocean tsunami of 2004 (Thailand, Indonesia) [ 68 , 69 ].

See Fig 2 for map of bioregions.

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The most frequently cited conservation objectives reflect a concentration on enhancing the quantity of hermatypic coral habitat and its most economically valuable inhabitants, including commercial fish ( Fig 3 ). Fewer studies reported on ARs deployed for objectives related to the mitigation of natural and anthropogenic impacts on reef communities, such as conservation of target species (24%), mitigation of environmental stressors (22%), and provision of coral nurseries (14%), a relatively new restoration goal [ 70 ]. These AR conservation objectives are particularly common in the Central and Western Indo-Pacific ( Fig 3 ). Studies addressing socio-cultural value and economic analyses on ARs (16%) were most frequently conducted in the Western Indo-Pacific. More specifically, 8 out of 51 studies were from the Middle East, where sea surface temperatures (SST) have increased more than 3 times the global average since 1985 [ 1 ]. This region is a global hotspot for AR research, leading the publication output in many categories of conservation objectives ( Fig 3 ). Two studies (both from Malaysia) stated their conservation objective was to deter fishing trawlers and were not included in Table 2 .

Taxonomic groups monitored on Ars.

Globally, fish and coral (29 and 26 studies, respectively) were the most frequently monitored taxonomic groups in the 51 studies assessing progress towards achieving conservation objectives of an AR. Most studies on corals (79%) were conducted in the Central and Western Indo-Pacific, while studies addressing fish populations were more evenly distributed across realms ( Fig 4 ). Publications on corals were most frequent in the Central and Western Indo-Pacific, indicating biases in the AR conservation literature.

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ARs have been deployed on coral reefs to assess and increase abundance of fish populations since the 1980s, and fish taxa were monitored in 55% of studies evaluating the conservation success of ARs ( Fig 4 ). This is largely in response to declining fisheries on coral reefs due to overfishing and harmful fishing practices that have had catastrophic effects on coral reef fish since the 1980s, such as cyanide and dynamite fishing [ 4 , 71 ]. Many studies have focused on the population dynamics and behaviour of commercially or recreationally desirable fish species on and near ARs [ 72 , 73 ]. In the 1980s and 1990s, publications focused on protecting and increasing target fish species on reefs [ 74 – 76 ]. From the 1990s to 2010s, research effort on fish taxa has continually increased ( Fig 5 ). In the late 2010s and 2020, a few conservation-oriented publications from Temperate Australasia and the Eastern Indo-Pacific focused on space use or connectivity of fish populations on ARs and adjacent natural reefs, likely reflecting an increased focus on the importance of connectivity in the persistence of reef fish assemblages [ 77 – 79 ].

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Scleractinian corals were the other most frequently monitored (53%) taxonomic group on ARs in the peer reviewed literature. Similarly to fish population metrics, the number of studies monitoring coral communities increased every decade from the 1990s to 2010s ( Fig 5 ), reflecting the increasing scale and severity of anthropogenic impacts on coral reefs [ 80 , 81 ]. Due to the alarming decline in coral cover and associated biodiversity worldwide, objectives of ARs that focus on coral conservation (e.g. coral nurseries or transplantation initiatives) [ 10 , 82 ] will likely continue to increase into the 2020s and beyond. More studies on coral conservation were published in the first five months of 2020 than during an entire decade in the 1980s and 1990s ( Fig 5 ).

Benthic algae and invertebrates other than corals were the least monitored taxonomic groups on ARs ( Fig 4 ). Understanding the successional patterns of these organisms on different AR structures is important because they can attract or deter target species [ 35 ]. Monitoring frequency of these underrepresented groups has increased since the 1990s, but they were still only measured in 0.05% (algae) and 20% (other invertebrates) of conservation studies published in the 2010s ( Fig 5 ). However, despite increasing awareness of the importance of these groups for attaining conservation objectives of ARs, monitoring is still lacking in many regions [ 35 , 83 ]. Non-coral invertebrate groups were monitored in studies from the Indo-Pacific realms (16%), the Tropical Atlantic (17%), and Temperate Australasia (33%) ( Fig 4 ). Only 1% of conservation studies measured benthic algae, all in the Indo-Pacific and the Tropical Atlantic. Fouling invertebrates and macroalgae growing on ARs can attract fish and motile invertebrate grazers [ 84 – 86 ]. Structures designed to support the growth of these organisms on coral reefs can enhance reef complexity and the abundance of local consumer populations [ 87 , 88 ]. Alternatively, excessive fouling by toxic invertebrates (e.g. ascidians and sponges) and some species of macroalgae deter coral larvae from settling and increase post-recruitment mortality rates [ 89 – 91 ]. Therefore, it is unclear whether ARs designed to promote fouling communities for the attraction of target fish species are conducive to coral recruitment.

Potential of ARs as a conservation or restoration strategy on coral reefs

Reported success of achieving ar conservation objectives.

Deployment of ARs with specific conservation objectives has varied over time ( Fig 6 ) and geographic locations ( Fig 3 ). Of the 51 studies, 65% reported success or progress towards achieving the conservation objective of AR deployment. Objectives with the highest reported rates of success were provision of nursery habitat and increasing coral cover (each 71%), followed by increasing fish abundance and mitigating effects of environmental impacts (each 64%), and increasing habitat quantity and attaining socio-cultural objectives (each 63%) and ( Table 2 ). Conservation of target species was reported as successful in only 42% of studies. The most-commonly cited reasons for not achieving conservation objectives were poor AR design for target species and extensive bleaching during the study period ( Table 2 ). Effective AR design considerations can be integrated into management strategies and deployment plans; however, reducing the level of extensive bleaching on artificial and natural reefs will require global cooperation for reducing carbon emissions [ 92 ].

Numbers above bars indicate number of studies.

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Many studies reported multiple conservation objectives for each AR ( Table 2 ), and 35% did not draw conclusions on all stated objectives. For example, if an AR was deployed for both increasing fish abundance and mitigating an environmental stressor, researchers may have recorded progress towards attaining only one of the two objectives due to constraints of logistics or expertise. Deploying ARs with multiple conservation objectives may reduce the likelihood of evaluating success or measuring ecological function of the AR. Structural design, site, and monitoring should be tailored for specific conservation objectives to limit ambiguous conclusions about success.

Evaluation of reported success in achieving AR conservation objectives

While ARs have been deployed to increase fish abundance since the 1980s, many studies monitoring their success did not measure appropriate ecological response variables for detecting increased fish production on the reef ( Fig 7 ). For example, few studies examining the success of ARs in increasing fish abundance effectively monitored fish recruitment and movement between natural reefs and ARs. Therefore, authors were not able to distinguish whether ARs are attracting fish from adjacent habitats or enhancing abundance of resident populations. The three-dimensional structure and physical relief of the AR plays a significant role in attracting adult and juvenile fish from the water column [ 34 , 93 , 94 ]. Factors that contribute to the species composition of the colonizing fish community on ARs include distance from suitable substrate, distance from source populations, access by predators, access to food, and shelter for protection and egg-laying [ 34 , 63 ]. Disentangling whether ARs actually enhance production of fish or simply redistribute them within the ecosystem would enable researchers to evaluate whether ARs can be used to increase absolute fish abundance on coral reefs. This knowledge gap is well cited within the AR literature [ 34 , 40 , 42 ] and new approaches, such as modelling of biomass flux, may prove useful [ 95 ]. However, our results indicate that the gap remains poorly addressed in coral reef ecosystems specifically.

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Increasing coral cover has been a relatively successful AR conservation strategy ( Table 2 ). Overall, peer-reviewed studies used appropriate monitoring strategies for determining the success of this objective; however, there was regional variation in the measured response variables. Studies done in marine realms that encompassed ocean warming hotspots (Western and Central Indo-Pacific) concentrated on response variables pertaining to specific coral life history events (e.g. recruitment, survival/mortality, reproduction, and growth) ( Fig 7 ). However, the scale of ARs has been too small to address regional losses in coral cover and the study duration has been too short to adequately assess a sustained increase in coral cover ( Table 2 ), which can take decades to detect [ 96 , 97 ]. Small-scale rehabilitation projects using ARs to increase coral cover in denuded areas might be successful if proper design considerations and environmental stressors are taken into account [ 17 ]. For example, suspended ARs could be deployed on shallow water reefs and moved to deeper or cooler water during periods of peak SST to avoid bleaching [ 47 ].

Protecting select ecologically and socio-culturally important species was addressed through the objective of conserving target species. Authors reported limited success for this objective, with many studies citing inappropriate design for target species as the reason ( Table 2 ). One study reported that colonization of target fish species was interrupted by the presence of invasive lionfish [ 98 ]. Structural design and site selection must be considered using species-specific requirements to increase the overall success of this conservation objective [ 17 ]. ARs deployed for the purpose of restoring, rehabilitating, or mitigating reef degradation for conservation of selected species need to be specifically engineered to enhance settlement and survival of targeted species [ 94 ].

Stressor mitigation has been increasingly used as a conservation objective for ARs over the past two decades ( Fig 6 ). This is most likely a response to the increasing frequency and severity of coral bleaching events and concurrent climate change perturbations since the 1990s [ 1 , 80 , 99 ]. While this objective can be met on small spatial scales (e.g. preventing impacts of wave action and sedimentation) [ 60 ], our results suggest limited success when ARs are deployed to address ecosystem-wide stressors because ARs operate on a much smaller scale (m– 100s m) than natural reefs (10s – 100s km). Both scientific and conservation projects on ARs can be interrupted by large-scale bleaching events during the study period, making it difficult or impossible to assess the efficacy of ARs in mitigating stressors [ 26 , 68 ]. ARs do not directly alleviate underlying environmental stressors and may only be effective at remediating damages once the original perturbation has been substantially reduced or removed [ 17 , 100 ]. Coral reef restoration (including through the deployment of ARs) is most effective as an integrated component of wider management frameworks that include stressor mitigation [ 17 ]. The mismatch between the increasing spatial scale of stressors and the small scale of management interventions, such as ARs, reinforce the urgency for developing comprehensive management frameworks [ 101 ].

Arguably the most successful application of ARs is as nursery habitat for coral transplantation or source populations for which specific and appropriate ecological response variables (i.e. coral growth, reproduction, and survival) were used to determine success ( Table 2 ). As long as coral colonies or fragments of colonies experience low mortality, increased larval production and a high yield of functional adult colonies with low environmental impact are possible [ 102 , 103 ]. Native species predicted to respond well to anticipated climatic changes can be selectively bred as a biological bank to re-populate natural reefs after disturbances [ 104 ]. If ARs are suspended or designed to detach from the seafloor, they also can be moved horizontally or vertically to avoid unfavorable growing conditions [ 70 ]. While nurseries operate on a relatively small scale compared to natural reefs, the likelihood of an AR functioning as a small source population in the region can be maximized by seeding it with high densities of coral species [ 28 ]. As with many studies published on active coral restoration strategies, publications examining the success of ARs as coral nurseries were exclusively from the Western and Central Indo-Pacific ( Fig 3 ).

ARs deployed to increase habitat have been largely successful, likely because they add hard substrate to the benthic environment, making this a relatively attainable objective [ 39 , 105 , 106 ]. Measured response variables focused on benthic community development and fish presence at the AR ( Table 2 ). Study durations for this objective were too short (0.01–3.5 y) to characterize success beyond initial recruitment and colonization phases for fish and invertebrates [ 63 ]. However, increasing hard substrate is not considered a high priority in reef conservation compared to addressing large-scale tissue loss of scleractinian corals caused by ocean acidification and warming [ 36 ].

In studies where deployment of ARs for socio-cultural purposes was the primary goal, the ARs were monitored appropriately and can be considered successful. However, in studies that combined socio-cultural and ecological objectives, conclusions were only drawn about the latter. Studies that monitored AR success using socio-cultural objectives employed a variety of socio-cultural variables, which can be separated into those monitoring human behaviour and emotions relative to ARs and those concerned with economic valuation ( Table 2 ). In the Western Indo-Pacific, researchers surveyed the attractiveness of ARs to divers and diver behaviour on ARs [ 107 ]. Some studies examining the economic value of ARs lacked secondary inclusion criteria for this review but conducted a cost-benefit analysis [ 108 ] or estimated gross revenue generated from commercial fisheries as a consequence of ARs [ 109 ].

Limitations of ARs and current knowledge gaps

Overall, our results indicate that ARs have limited success in meeting regional-scale conservation objectives, such as increasing abundance of coral and fish species or stressor mitigation. Nonetheless, these objectives are being increasingly cited in studies examining AR success, likely because of the acceleration of coral decline globally and the increasing call for remediating losses with active restoration strategies [ 14 ]. Because ARs mostly operate on a much smaller scale than natural reefs (except possibly small patch reefs), their success in addressing large-scale objectives must be assessed. Reference or control sites can provide context for the observed outcomes on ARs [ 29 ]. For example, a meta-analysis of 39 studies documented no difference in fish community metrics between natural and artificial reefs [ 41 ]. While it has been suggested that larger ARs (> 150 m 2 ) support higher fish abundances [ 23 ], the extent to which ARs function as a source of fish production remains poorly understood [ 34 , 40 , 42 ]. Further, larger ARs are logistically difficult to fund, deploy, and monitor. The introduction of networks of ARs to regions with minimal environmental stressors may increase the success of abundance-oriented conservation objectives (i.e. increasing fish abundance and coral cover) by increasing colonizable reef area while fostering connectivity of fish and invertebrate species between degraded natural reefs [ 100 ]. Overall, small-scale objectives of ARs (e.g. increasing public education, selective coral breeding programs, training scientific and recreational divers) are far more achievable because they do not require additional intensive, long-term studies to determine their contribution to reef conservation and are generally successful when well defined and monitored.

Among all studies considered in this review, more than 73% spanned 3 years or less, which is too short a period for elucidating or predicting long-term shifts in coral reef populations. Studies that examined the success of ARs in meeting conservation objectives spanned 1 week to 5 years. This period matches the average for monitoring studies of several different coral restoration approaches [ 19 , 10 ] and may be adequate for addressing short-term goals of restoration at local scales [ 17 ]. For example, observation periods of months to years can allow monitoring colonization patterns in many short-lived organisms, such as reef-associated invertebrates (e.g. ascidians, bryozoans, and some sponges), that can settle, reproduce, and die on a substrate within months [ 47 , 110 , 111 ]. These durations also may be effective for monitoring fish populations on ARs, as many fish species have a life expectancy of under 5 years due to their inherent longevity or high rates of juvenile mortality [ 112 – 114 ]. Changes in coral community composition and dynamics, however, take much longer to detect [ 115 , 116 ]. For example, scleractinian coral communities require multidecadal monitoring to properly assess ecologically relevant trends in coral cover and species composition [ 96 , 97 ]. Longer monitoring periods also may be needed to capture effects of aperiodic or stochastic events, such as heatwaves or storms. Future studies examining the success of ARs in achieving coral-oriented conservation objectives must adjust study duration according to the relevant time scales of biotic and abiotic factors that govern the underlying ecological processes.

ARs can have some potentially negative impacts on the surrounding ecosystems. Often the materials used in ARs, such as rubber or plastics, are not biodegradable or may even leach toxic substances into the surrounding ecosystems [ 63 ]. Concrete, which is used for many ARs because of ease of production and low cost, in addition to leaching metals has a high alkalinity that may inhibit colonization [ 117 , 118 ]. To increase the ecological value of artificial structures, new materials using aggregate concrete with different chemistries are being developed [ 117 , 118 ]. ARs can also facilitate the introduction and spread of invasive species [ 119 , 120 ]; modification of the physical and chemical properties of the ARs and pre-seeding by native species may minimize colonization by non-native species [ 119 ]. Engineering solutions can provide potential mitigation strategies for the negative impacts of ARs.

Conclusions and recommendations

• To be an effective management tool, ARs deployed for conservation purposes must employ SMART (specific, measureable, achievable, realistic, time bound) objectives. Structural design, site, and monitoring should be tailored for specific conservation objectives to limit ambiguous conclusions about success. We showed that evaluation of success is less effective when ARs have multiple conservation objectives, either because some objectives are not evaluated or the measured ecological function is inappropriate for all objectives.
• To be useful, ARs deployed for conservation purposes must be clearly described and accurately contextualized within the recipient environments to facilitate comparisons across geographic locations, target species and conservation objectives. Only 136 of 530 studies on research on coral reef ARs included a date, precise location, and depth of deployment, precise dimensions and number of ARs and stated an objective of AR deployment. In the scientific literature alone, the array of materials, shapes, sizes and distribution of ARs on the seafloor varied widely from haphazardly deployed sunken ships or car tires to deliberately deployed cinder blocks or reef balls.
• The design of ARs must be suitable for the specific conservation objectives. Many of the studies we examined reported that ARs were unsuccessful in meeting their objective either because of an inappropriate design or because of loss of AR communities due to severe environmental perturbations, such as heatwaves. Future studies aiming to increase the efficacy of ARs for conservation purposes should choose structures and sites that are tailored for specific conservation objectives. The objective of using ARs to address ecosystem-wide restoration goals, such as increasing coral cover and stressor mitigation, has been met with limited success, because of the mismatch in scales between the AR and the recipient ecosystem. While nearly all AR projects are relatively small compared to adjacent natural reefs, they can address local, conservation-specific objectives, such as assisting the recovery of smaller patch reefs or local enhancement of coral cover if the stressor can be removed.
• Monitoring of the performance of ARs in meeting conservation objective(s) must be based on ecologically relevant variables measured over appropriate time scales. For example, although increases in fish abundance and habitat quantity were the most frequently documented conservation objectives of ARs, monitoring over short-term scales failed to capture recruitment and community succession. Overall, fewer than one quarter of the 136 studies measured the success of ARs in meeting conservation objectives.

Based on their reported success as active restoration tools for tropical coral reefs, ARs are most likely to achieve their conservation objectives by providing nursery habitat for rearing target reef species or by supplying additional hard substrate for settlement and recruitment of corals and other marine organisms. We suggest that promoting local socio-cultural values also has potential for success if it is prioritized as an objective and properly monitored. This objective can be effective also globally, by increasing awareness of coral reef decline among tourists who mostly originate from countries without corals. While the effectiveness of ARs per se in achieving regional-scale conservation objectives may be limited, their integration into a larger restoration program could prove beneficial if used conjunction with other conservation strategies. However, given their relatively high cost, the implementation of ARs into larger restoration programs would require the development of better practices in identifying objectives, selecting the appropriate designs, and monitoring the relevant ecological responses.

Supporting information

S1 checklist. prisma 2009 checklist..

https://doi.org/10.1371/journal.pone.0261964.s001

S1 File. Literature used in Tables 1 and 2 .

https://doi.org/10.1371/journal.pone.0261964.s002

Acknowledgments

We thank Claire Attridge for her help with the literature review and the preparation of the manuscript.

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New report on Great Barrier Reef shows coral cover increases before onset of serious bleaching, cyclones

by Australian Institute of Marine Science

Most of the underwater surveys contributing to these findings , published today, were conducted before and during the recent mass bleaching event, one of the most extensive and serious on record, and have not yet captured how many corals survived or died following the bleaching.

Surveys in the Central region were also completed before the passage of tropical Cyclone Jasper in December 2023.

AIMS' Long-Term Monitoring Program (LTMP) leader Dr. Mike Emslie said coral cover increases were a positive sign but did not reflect the potentially destructive consequences of the 2024 mass bleaching event.

"We saw evidence of early onset mortality, particularly in the Southern region, but the full picture of mortality was not yet apparent during this year's surveys," he said.

"While bleached corals are very stressed, they are still alive and are recorded as live coral on our surveys.

"Some types of corals can remain bleached for months, remaining on a knife edge between survival and death. This is why returning and repeating surveys of the reefs in this vast, complex and dynamic system is so important. This year's results serve as a very important reference against which to measure the impacts of the summer's events."

The next (LTMP) survey season recommences in September and will capture impacts on coral cover from this summer's mass bleaching event and the cyclones, with a full assessment complete by mid-2025.

"Climate change remains the greatest threat to the Reef because it drives these mass bleaching events. This most recent one was the fifth such event since 2016. These more frequent and extensive marine heat waves will lead to shortened 'windows' for coral recovery. Recent gains, while encouraging, can be lost in a short amount of time," Dr. Emslie said.

Surveys were conducted at 94 Reefs spread through the Northern, Central and Southern Great Barrier Reef between August 2023 and June 2024.

The Report recorded the following average hard coral coverage:

• Northern region (north of Cooktown)—39.5%, up from 35.8% last year;
• Central region (Cooktown to Proserpine)—34%, up from 30.7%;
• Southern region (south of Proserpine)—39.1%, up from 34%.

The AIMS report finds that small rises in coral cover this year brought the Northern and Central regions to their highest levels in 38 years of monitoring.

The surveys also found that crown-of-thorns starfish outbreaks have persisted on some reefs in the Southern region.

The long term monitoring team surveyed reefs off Townsville after the passage of tropical Cyclone Kirrily in late January, finding evidence of storm damage and declines in hard coral cover ranging from 6% to 10% at Kelso, John Brewer, Helix and Chicken Reefs. Other reefs appear to have escaped with little impact.

AIMS Research Program Director Dr. David Wachenfeld said the regional increases in coral cover are encouraging, showing the Reef's capacity for recovery after reaching their lowest levels within the last 15 years. However, climate change and other disturbances mean this recovery is fragile and Reef resilience is not limitless.

"In many ways the Reef has had some lucky escapes in recent years. The 2020 and 2022 mass bleaching events had levels of heat stress that were not as intense as the 2016 and 2017 events or the 2024 event. Coupled with very few other events causing widespread coral death, that has led to the levels of coral cover increase we have seen," he said.

"But the frequency and intensity of bleaching events is unprecedented, and that is only forecast to escalate under climate change, alongside the persistent threat of crown-of-thorns starfish outbreaks and tropical cyclones."

Aerial surveys undertaken by AIMS and the Great Barrier Reef Marine Park Authority in February and March found bleached corals in the shallows of 73% of reefs surveyed across all three regions.

In recent weeks, AIMS scientists in separate monitoring programs observed substantial mortality in reefs that were particularly hard hit by the 2024 event.

"We are only one large scale disturbance event away from a reversal of the recent recovery. The 2024 bleaching event could be that event—almost half of the 3000 or so reefs that make up the marine park experienced more heat stress than ever recorded," Dr. Wachenfeld said.

"We still don't know how much mortality this event has caused. Our monitoring over the next 12 months will help us to understand how this bleaching event stacks up against the others in the last decade."

AIMS CEO Professor Selina Stead said AIMS was prioritizing research to develop scientific solutions to boost reef resilience under a warming climate.

"Climate change is increasing pressure on reef systems around the world," she said. "The 2024 bleaching event was part of the fourth global bleaching event, announced in April.

"These vitally important ecosystems that millions rely upon need strong greenhouse gas emissions reduction, science-based management of local pressures, and input from multiple fields of research if they are to endure.

"At AIMS we are developing a toolbox of interventions to help reefs adapt to and recover from the effects of climate change ."

Provided by Australian Institute of Marine Science

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• Published: 07 August 2024

Highest ocean heat in four centuries places Great Barrier Reef in danger

• Benjamin J. Henley   ORCID: orcid.org/0000-0003-3940-1963 1 , 2 , 3 ,
• Helen V. McGregor   ORCID: orcid.org/0000-0002-4031-2282 1 , 2 ,
• Andrew D. King   ORCID: orcid.org/0000-0001-9006-5745 4 , 5 ,
• Ove Hoegh-Guldberg   ORCID: orcid.org/0000-0001-7510-6713 6 ,
• Ariella K. Arzey 1 , 2 ,
• David J. Karoly 4 ,
• Janice M. Lough 7 ,
• Thomas M. DeCarlo   ORCID: orcid.org/0000-0003-3269-1320 8 , 9 &
• Braddock K. Linsley   ORCID: orcid.org/0000-0003-2085-0662 10  

Nature volume  632 ,  pages 320–326 ( 2024 ) Cite this article

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• Climate change
• Environmental impact
• Palaeoclimate

Mass coral bleaching on the Great Barrier Reef (GBR) in Australia between 2016 and 2024 was driven by high sea surface temperatures (SST) 1 . The likelihood of temperature-induced bleaching is a key determinant for the future threat status of the GBR 2 , but the long-term context of recent temperatures in the region is unclear. Here we show that the January–March Coral Sea heat extremes in 2024, 2017 and 2020 (in order of descending mean SST anomalies) were the warmest in 400 years, exceeding the 95th-percentile uncertainty limit of our reconstructed pre-1900 maximum. The 2016, 2004 and 2022 events were the next warmest, exceeding the 90th-percentile limit. Climate model analysis confirms that human influence on the climate system is responsible for the rapid warming in recent decades. This attribution, together with the recent ocean temperature extremes, post-1900 warming trend and observed mass coral bleaching, shows that the existential threat to the GBR ecosystem from anthropogenic climate change is now realized. Without urgent intervention, the iconic GBR is at risk of experiencing temperatures conducive to near-annual coral bleaching 3 , with negative consequences for biodiversity and ecosystems services. A continuation on the current trajectory would further threaten the ecological function 4 and outstanding universal value 5 of one of Earth’s greatest natural wonders.

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Like many coral reefs globally, the World Heritage-listed GBR in Australia is under threat 4 , 6 . Mass coral bleaching, declining calcification rates 5 , 7 , outbreaks of crown-of-thorns starfish ( Acanthaster spp.) 8 , severe tropical cyclones 9 and overfishing 10 have placed compounding detrimental pressures on the reef ecosystem. Coral bleaching typically occurs when heat stress triggers the breakdown of the symbiosis between corals and their symbiotic dinoflagellates 11 . Although coral bleaching can occur locally as a result of low salinity, cold waters or pollution, regional and global mass bleaching events, in which the majority of corals in one or more regions bleach at once, are strongly associated with increasing SST linked to global warming 2 .

The first modern observations of mass coral bleaching on the GBR occurred in the 1980s, but these events were less widespread and generally less severe 3 than the bleaching events in the twenty-first century 4 . Stress bands in coral skeletal cores have provided potential evidence for pre-1980s bleaching in the GBR and Coral Sea, such as during the 1877–78 El Niño 12 . However, stress bands are evident in relatively few cores before 1980 (ref. 12 ),  suggesting that severe mass bleaching did not occur in the 1800s and most of the 1900s.

As the oceans have warmed, however, mass coral bleaching events have become increasingly lethal to corals 4 . Coral bleaching on the GBR 1 in 1998 coincided with a strong eastern-Pacific El Niño, and in 2002 with a weak El Niño. El Niño events can induce lower cloud cover and increased solar irradiance over the GBR 13 , increasing the risk of thermal stress and mass bleaching events 14 . In 2004, water temperatures were anomalously warm, and although bleaching occurred in the Coral Sea 15 , it was not widespread in the GBR, probably because there was reduced upwelling and an associated reduced influence of nutrients on symbiotic dinoflagellate expulsion 16 .

However, in the nine January–March periods from 2016 to 2024 (inclusive) there were five mass coral bleaching events on the GBR. Each was associated with high SSTs and affected large sections of the reef. GBR mass bleaching occurred in both 2016 and 2017, influenced by the presence of an El Niño event in 2016, and led to the death of at least 50% of shallow-water (depths of 5–10 m) reef-building corals 4 . Major bleaching events occurred again in quick succession in 2020 and 2022, with the accumulated heat stress for large sections of the GBR reaching levels conducive to widespread bleaching but lower levels of coral mortality 1 . The bleaching event in 2022 occurred, unusually, during a La Niña event, which is typically associated with cooler summer SSTs, higher than average rainfall and higher cloud cover on the GBR 1 . At the time of writing, researchers are assessing the impacts of the 2024 mass bleaching event.

The frequency of recent mass coral bleaching and mortality on the GBR is cause for concern. In 2021, the World Heritage Committee of the United Nations Educational, Scientific and Cultural Organization (UNESCO) drafted 17 a decision to inscribe the GBR on the List of World Heritage in Danger, stating that the reef is “facing ascertained danger”, citing recent mass coral bleaching events and insufficient progress by the State Party (Australia) in countering climate change, improving water quality and land management issues. The committee’s adopted decisions 18 have not included inscription of the ‘in danger’ status, but the draft inscription highlights the seriousness of the recent mass coral bleaching events. Authorities in Australia 5 have noted that climate change and coral bleaching have deteriorated the integrity of the outstanding universal value of the GBR, a defining feature of its World Heritage status.

Although rapidly rising SSTs are attributed to human activities with virtual certainty 19 , understanding the multi-century SST history of the GBR is critical to understanding the influence of SST on mass coral bleaching and mortality in recent decades. Putting aside a problematic attempt to do this 20 , which was discredited 21 , 22 , knowledge of the long-term context for GBR SSTs comes primarily from two multi-century reconstructions based on the geochemistry of coral cores collected from the inner shelf 23 and outer shelf 24 (Flinders Reef) in the central GBR. These reconstructions showed that SSTs in the early 2000s were not unusually high relative to levels in the past three centuries, with five-year mean SSTs (and salinities) estimated to be higher in the 1700s than in the 1900s. However, these records were limited by their relatively coarse five-year sampling resolution and their most recent data point being from the early 2000s. After these studies were published, SSTs in the GBR have continued to rise. Updated analysis of coral data from Flinders Reef provides valuable improved temporal resolution 25 , but interpretations of these records remain limited spatially.

Here, we investigate the recent high SST events in the GBR region in the context of the past four centuries. We combine a network of 22 coral Sr/Ca and δ 18 O palaeothermometer series (Supplementary Tables 1 and 2 ) located in and near to the Coral Sea region to infer spatial mean SST anomalies (SSTAs) for January–March, the months when maximum SST and thermal bleaching are most likely to occur in the Coral Sea 16 , 26 , each year from 1618 to 1995 ( Methods and Supplementary Information ). Anthropogenic climate change began and proceeded entirely within the multi-century lives of some of these massive coral colonies, offering a continuous multi-century record covering the industrial era. We use this 1618–1995 reconstruction and the available 1900–2024 instrumental data to contextualize the modern trend and rank four centuries of January–March SSTAs with greater precision than was previously possible. We then assess the degree of human influence on ocean temperatures in the region using climate model simulations run both with and without anthropogenic forcing.

The instrumental period (1900–present)

Mass coral bleaching on the GBR in 2016, 2017, 2020, 2022 and 2024 during January–March coincided with widespread warm SSTAs in the surrounding seas 1 , including the Coral Sea (Fig. 1a–e , using ERSSTv5 data 27 ). The Coral Sea and GBR have experienced a strong warming trend since 1900 (Fig. 1f ). January–March SSTAs averaged over the GBR are strongly correlated ( ρ  = 0.84, P   ≪  0.01) with those in the broader Coral Sea (Fig. 1f ), including when the long-term warming trend is removed from both time series ( ρ  = 0.69, P  < 0.01; Supplementary Fig. 4 ). Based on the strength of this correlation, we associate high January–March area-averaged Coral Sea SSTAs with increased thermal bleaching risk in the GBR.

a – e , SSTAs (using ERSSTv5 data) for January–March in the Australasian region relative to the 1961–90 average for the five recent GBR mass coral bleaching years: 2016, 2017, 2020, 2022 and 2024. The black box shows the Coral Sea region (4° S–26° S, 142° E–174° E). f , Coral Sea and GBR mean SSTAs for 1900–2024 in January–March relative to the 1961–90 average. The black vertical lines indicate the five recent GBR mass coral bleaching years.

Record temperatures were set in 2016 and 2017 in the Coral Sea, and in 2020 they peaked fractionally below the record high of 2017. The January–March of 2022 was another warm event, the fifth warmest on record at the time. Recent data (ERSSTv5) indicate that 2024 set a new record by a margin of more than 0.19 °C above the previous record for the region. The January–March mean SSTs averaged over the five mass bleaching years during the period 2016–2024 are 0.77 °C higher than the 1961–90 January–March averages in both the Coral Sea and the GBR. The multidecadal warming trend, extreme years and association between GBR and Coral Sea SSTs are similar for the HadISST 28 gridded SST dataset, with some notable differences in the 1900–40 period (Supplementary Fig. 3 ). Furthermore, analysis of modern temperature-sensitive Sr/Ca series from GBR corals for 1900–2017 provides coherent independent evidence of statistically significant multi-decadal warming trends in January–March SSTs in the central and southern GBR (Supplementary Information section  4.2 ).

A multi-century context (1618–present)

Reconstructing Coral Sea January–March SSTs from 1618 to 1995 extends the century-long instrumental record back in time by an additional three centuries (Fig. 2a and Methods ). The reconstruction (calibrated to ERSSTv5) shows that multi-decadal SST variability was a persistent feature in the past. At the centennial timescale, there is relative stability before 1900, with the exception that cooler temperatures prevailed in the 1600s. Warming during the industrial era has been evident since the early 1900s (Fig. 2a ). There is a warming trend for January–March of 0.09 °C per decade for 1900–2024 and 0.12 °C per decade for 1960–2024 (Fig. 1f ) using ERSSTv5 data. Calibrating our reconstruction to HadISST1.1 yields similar results, with some differences in the degree of pre-1900 variability at both multi-decadal and centennial timescales (Supplementary Information section  5.2.6 ).

a , Reconstructed and observed mean January–March SSTAs in the Coral Sea for 1618–2024 relative to 1961–90. Dark blue, highest skill (maximum coefficient of efficiency) reconstruction with the full proxy network; light blue, 5th–95th-percentile reconstruction uncertainty; black, observed (ERSSTv5) data. Red crosses indicate the five recent mass bleaching events. Dashed lines indicate the best estimate (highest skill, red) and 95th-percentile (pink) uncertainty bound for the maximum pre-1900 January–March SSTA. b , Central GBR SSTA for the inner shelf 23 in thick orange and outer shelf 25 (Flinders Reef) in thin orange lines; these series are aligned here (see Methods ) with modern observations of mean GBR SSTAs for January–March relative to 1961–90. Observed data are shown at annual (grey line) and five-year (black line with open circles, plotted at the centre of each five-year period and temporally aligned with the five-year coral series 23 ) resolution. Dashed lines indicate best-estimate pre-1900 January–March maxima for refs. 23 (red) and 25 (pink). Orange shading indicates 5th–95th-percentile uncertainty bounds. Red crosses indicate the five recent mass bleaching events. c , Evaluation metrics for the Coral Sea reconstruction (Supplementary Information section  3.1 ); RE, reduction of error; CE, coefficient of efficiency; Rsq-cal, R-squared in the calibration period; Rsq-ver, R-squared in the verification (evaluation) period. d , Coral data locations relative to source data region (orange box) and Coral Sea region (red box). Coral proxy metadata are given in Supplementary Tables 1 and 2 .

Our best-estimate (highest skill; Methods ) annual-resolution Coral Sea reconstruction (Fig. 2a ), using the full coral network calibrated to the ERSSTv5 instrumental data, indicates that the January–March mean SSTAs in 2016, 2017, 2020, 2022 and 2024 were, respectively, 1.50 °C, 1.54 °C, 1.53 °C, 1.46 °C and 1.73 °C above the 1618–1899 (hereafter ‘pre-1900’) reconstructed average. Using the same best-estimate reconstruction, Coral Sea January–March SSTs during these GBR mass bleaching years were five of the six warmest years the region has experienced in the past 400 years (Fig. 2a ).

By comparing the recent warm events to the reconstruction’s uncertainty range ( Methods ), we quantify, using likelihood terminology consistent with recent reports from the Intergovernmental Panel on Climate Change 19 , that the recent heat extremes in 2017, 2020 and 2024 are ‘extremely likely’ (>95th percentile; Fig. 2a ) to be higher than any January–March in the period 1618–1899. Furthermore, the heat extremes in 2016 and 2022 are (at least) ‘very likely’ (>90th percentile) to be above the pre-1900 maximum. We perform a series of tests that verify that our findings are not simply an artefact of the nature of the coral network itself (Supplementary Information section 5.2 ). In a network perturbation test, we generate 22 subsets of the reconstruction by adding proxy records incrementally in order from the highest to the lowest correlation with the target (Supplementary Information section  5.2.5 ). We confirm that 2017, 2020 and 2024 were ‘extremely likely’ (>95th percentile) to have been warmer than any year pre-1900 (using ERSSTv5 data) for all of these proxy subsets. Furthermore, in 20 of the 22 subsets, 2016 was also ‘extremely likely’ (>95th percentile), rather than ‘very likely’, to be warmer (2022 was ‘extremely likely’ in 14 of the 22 subsets). All our additional tests, including a reconstruction with only Sr/Ca coral data (thereby omitting the possibility of any non-temperature signal in δ 18 O coral on the reconstruction), achieve high reconstruction skill and confirm the extraordinary nature of recent extreme temperatures in the multi-century context (Supplementary Information section  5.2 ). Analyses using HadISST1.1 generally show lower correlations with the coral data and reconstructions with slightly warmer regional SSTs before 1900, along with more-muted centennial and multi-decadal variability in the pre-instrumental period. Nevertheless, the HadISST1.1-calibrated reconstructions show that the recent thermal extremes are well above the best estimate (highest skill) of the pre-1900 maximum of reconstructed January–March SSTAs (Supplementary Fig. 42 ). Furthermore, lower SSTAs (in the HadISST1.1 data) relative to the previous three centuries (as in our reconstructions calibrated to HadISST1.1), coupled with the recently observed mass coral bleaching events, could indicate that long-lived corals have a greater sensitivity to warming than is currently recognized.

Reconstructed regional GBR SSTAs based on a five-year-resolution, multi-century coral δ 18 O record from the central inshore GBR 23 (Fig. 2b ) show similarly strong warming since 1900 but more multi-decadal-to-centennial variability than the Coral Sea reconstruction. Recent five-year mean January–March GBR SSTAs narrowly exceed the best estimate of the maximum pre-1900 five-year mean since the early 1600s (Fig. 2b ). The averages for the five-year periods centred on 2018 and 2022 exceed the pre-1900 maximum by 0.11 °C and 0.06 °C, respectively. Results are similar using the five-year-resolution Flinders Reef (central outer shelf) 24 record (Supplementary Fig. 24 ), although its interpretation is limited by the lack of uncertainty estimates available for that record. Our Coral Sea reconstruction incorporates an updated (annual resolution) record from Flinders Reef 25 , which indicates similar centennial trends (thin orange line in Fig. 2b ) and shows that the recent high January–March SSTA events have approached the estimated local pre-1900 maximum SSTA. Although contiguous multi-century cores from within the GBR are limited in their spatial extent, twentieth-century warming is evident in these records.

The extraordinary nature of the recent Coral Sea January–March SSTs in the context of the past 400 years is further illustrated by comparing the ranked temperature anomalies (Fig. 3 ) for the combined reconstructed and instrumental period from 1618–2024, incorporating reconstruction uncertainty ( Methods ). The mass coral bleaching years of 2016, 2017, 2020, 2022 and 2024, and the heat event of 2004, stand out as the warmest events across the whole 407-year record. The warmest three years (2024, 2017 and 2020) exceed the upper uncertainty bound (95th percentile) of the warmest reconstructed January–March in the pre-1900 period (pink (upper) dashed line in Fig. 3 ); 2016, 2004 and 2022 exceed the 90th percentile bound (red (lower) dashed line in Fig. 3 ). The warming trend is clear in the association between the ascending rank of the temperature anomalies and the year (shown as the colour of the filled circles in Fig. 3 ). Despite high interannual variability, 78 of the warmest 100 January–March periods between 1618 and 2024 occurred after 1900, and the 23 warmest all occur after 1900. The warmest 20 January–March periods all occur after 1950, coinciding with accelerated global warming.

Ranked January–March SSTAs for 1618–2024 relative to 1961–90 (coloured circles) from the best-estimate (highest skill, full coral network) reconstruction (1618–1899) and instrumental (ERSSTv5) data (1900–2024). The year is indicated by the colour of the filled circles. The 5th–95th-percentile uncertainty bounds of the pre-1900 reconstructed SSTAs are shown by small grey dots. The year labels indicate the warmest six years on record, five of which were mass coral bleaching years on the GBR. The pink (upper) dashed line indicates the 95th-percentile uncertainty bound of the maximum pre-1900 reconstructed SSTA; the red (lower) dashed line indicates the 90th-percentile limit.

Assessing anthropogenic influence

Using climate model simulations from the most recent (sixth) phase of the Coupled Model Intercomparison Project 29 (CMIP6), we assess the human influence on January–March SSTAs in the Coral Sea. The model simulations are from two experiments in the Detection and Attribution Model Intercomparison Project (DAMIP) 30 . The first set of simulations represents historical climate conditions, including both the natural and human influences on the climate system over the 1850–2014 period (‘historical’; red in Fig. 4 ). The second experiment is a counterfactual climate that spans the same period and uses the same models but includes only natural influences on the climate, omitting all human influences (‘historical-natural’; blue in Fig. 4 ). The historical experiment includes anthropogenic emissions of greenhouse gases and aerosols, stratospheric ozone changes and anthropogenic land-use changes; the historical-natural experiment does not. Variations in natural climate forcings, such as from volcanic eruptions and solar variability, are incorporated in both experiments. We include models that have a transient climate response (the global mean surface-temperature anomaly at the time of a doubling of atmospheric CO 2 concentration) in the range 1.4–2.2 °C, which is deemed ‘likely’ by the science community 31 ( Methods and Supplementary Information ).

Climate-model simulations of Coral Sea January–March SSTAs relative to the 1850–1900 average for the period 1850–2014, for models within the ‘likely’ range for their transient climate response 31 . The blue line (median) and light blue shading (5th–95th-percentile limits) are from the ‘historical-natural’ climate model simulations (no anthropogenic climate forcing); the red line and light red shading are from the ‘historical’ simulations (anthropogenic influences on the climate included) using the same set of climate models. The climate-model-derived time of emergence of anthropogenic climate change, shown by the grey and black vertical lines (1976 and 1997), is when the ratio of the climate change signal to the standard deviation of noise/variability 32 across model ensemble members first rises above 1 and 2, respectively. All models are represented equally in the model ensemble.

It is only with the incorporation of anthropogenic influences on the climate that the model simulations capture the modern-era warming of the Coral Sea January–March SSTA (Fig. 4 ). The median of the historical simulations has statistically significant warming trends of 0.05 °C, 0.10 °C and 0.15 °C per decade for the periods from 1900, 1950 and 1970 to 2014, respectively; the equivalent historical-natural trends are smaller in magnitude than ±0.01 °C per decade. To further explore the centennial-scale trends, we use a bootstrap ensemble ( Methods ) of the two sets of 165-year simulations from 1850–2014. We found that 100% of the historical bootstrap ensemble has statistically significant positive trends ( Methods ) for 1900–2014, but this value is 0% for the historical-natural ensemble. The observed (ERSSTv5) mean SSTA for 2016–2024 of 0.60 °C relative to 1961–90 is warmer than any nine-year sequence in the 7,095 simulated years in the historical-natural experiments from models with transient climate responses in the ‘likely’ range 31 .

We also use the simulations to estimate the time of emergence of the anthropogenic influence on January–March Coral Sea SSTAs above the natural background variability. The anthropogenic warming signal 32 increases from near zero in 1900 to around 0.5 standard deviations of the variability (‘noise’) in 1960. The climate change signal-to-noise ratio then increases rapidly from 1960 to 2014, exceeding 1.0 in 1976, 2.0 in 1997 and around 2.8 by 2014, the end of these simulations (Fig. 4 , Methods and Supplementary Fig. 50 ). Anthropogenic impacts on the climate are virtually certain to be the primary driver of this long-term warming in the Coral Sea.

Previously, our knowledge of the SST history of the GBR and the Coral Sea region has been highly dependent on instrumental observations, with the exception of the five-year-resolution multi-century coral Sr/Ca and U/Ca SST reconstructions from the two point locations in the central GBR 23 , 24 , an update at one of these locations 25 , seasonal resolution ‘floating’ (in time) chronologies from the GBR in the Holocene 33 , 34 and point SST estimates further back in time 35 . Thus, the context of recent warming trends in the Coral Sea and GBR and their relation to natural variability on decadal to centennial timescales is largely unknown without reconstructions such as the one we developed here.

Our coral proxy network is located mostly beyond the GBR, in the Coral Sea, and some series are located outside the Coral Sea region (Fig. 2d ). The selection of the Coral Sea as a study region allowed for a larger sample of contributing coral proxy data than exists for the GBR. However, coral bleaching on the GBR can be influenced by factors other than large-scale SST, including local oceanic and atmospheric dynamics that can modulate the occurrence and severity of thermal bleaching and mortality events 13 . Nonetheless, warming of seasonal SSTs over the larger Coral Sea region is likely to prime the background state and increase the likelihood of smaller spatio-temporal-scale heat anomalies. Furthermore, where we use only the five-year resolution series directly from the GBR to reconstruct GBR SSTAs, we draw similar conclusions about the long-term trajectory of SSTAs as for our full coral network (Fig. 2b and Supplementary Fig. 24 ). Furthermore, short modern coral series from within the GBR, analysed in this study, document a multi-decadal warming signal that is coherent with instrumental data (Supplementary Figs. 29 and 30 ). Nonetheless, additional high-resolution, multi-century, temperature-sensitive coral geochemical series from within the GBR would help unravel the local and remote ocean–atmosphere contributions to past bleaching events and reduce uncertainties.

The focus on the larger Coral Sea study region also takes advantage of the global modelling efforts of CMIP6. The large number of ensemble members available for CMIP6 means that greater climate model diversity, and therefore greater certainty in our attribution analysis, is possible compared with most single model analyses. There is also a methodological benefit in having high replication of the same experiments run with multiple climate models. However, coarse-resolution global-scale models do not accurately simulate smaller-scale processes, such as inshore currents and mesoscale eddies in the Coral Sea or the Gulf of Carpentaria, which probably affect local surface temperatures and variations in nutrient upwelling in the GBR 36 , 37 . Upwelling on the GBR is linked to the strength of the East Australian Current 16 , the southward branch of the South Pacific subtropical gyre. The CMIP-scale models we use do capture these gyre dynamics. The models show that the East Australian Current is expected to increase in strength as the climate continues to warm through this century 38 , and this may lead to more nutrient inputs that can exacerbate coral sensitivity to rising heat stress 39 , 40 . As well as focusing our model analysis on the larger Coral Sea region, we use a three-month time step. In doing so, we minimize the impact of model spatio-temporal resolution on our inferences about the role of anthropogenic greenhouse-gas emissions on the SST conditions that give rise to GBR mass bleaching.

Remaining uncertainties

We present analyses and interpretations that are as robust as possible given currently available data and methods. However, several sources of remaining uncertainty mean that future reconstructions of past Coral Sea and GBR SSTs could differ from those presented here. Although bias corrections are applied to observational SST datasets such as ERSST and HadISST, these datasets probably retain biases, especially for the period during and before 1945 (ref. 41 ), and these may not be fully accounted for in the uncertainty estimates 42 . Because our reconstructions are calibrated directly to these datasets, future observational-bias corrections are likely to improve proxy-based reconstructions.

Reconstructions of SST that use coral δ 18 O records may be susceptible to the influence of changes in the coral δ 18 O–SST relationship on time periods longer than the instrumental training period, along with non-SST changes in the δ 18 O of seawater, which can covary with salinity. As such, new coral records of temperature-sensitive trace-element ratios such as Sr/Ca, Li/Mg or U/Ca may prove influential in future efforts to distinguish between changes in past temperature and hydroclimate. Owing to the limited availability of multi-century coral data from within the GBR itself, the reconstructed low-frequency variability of GBR SSTs in recent centuries is likely to change as more temperature proxy data become available. It is also likely that new sub-annual resolution records would aid in removing potential signal damping or bias from our use of some annual-resolution records to reconstruct seasonal SSTAs.

Ecological consequences

With global warming of 0.8–1.1 °C above pre-industrial levels 19 there has been a marked increase in mass coral bleaching globally 43 . Even limiting global warming to the Paris Agreement’s ambitious 1.5 °C level would be likely to lead to the loss of 70–90% of corals that are on reefs today 44 . If all current international mitigation commitments are implemented, global mean surface temperature is still estimated to increase in the coming decades, with estimates varying between 1.9 °C (ref. 45 ) and 3.2 °C (ref. 46 ) above pre-industrial levels by the end of this century. Global warming above 2 °C would have disastrous consequences for coral ecosystems 19 , 44 and the hundreds of millions of people who currently depend on them.

Coral reefs of the future, if they can persist, are likely to have a different community structure to those in the recent past, probably one with much less diversity in coral species 4 . This is because mass bleaching events have a differential impact on different coral species. For example, fast-growing branching and tabulate corals are affected more than slower-growing massive species because they have different thermal tolerance 4 . The simplification of reef structures will have adverse impacts on the many thousands of species that rely on the complex three-dimensional structure of reefs 4 . Therefore, even with an ambitious long-term international mitigation goal, the ecological function 4 of the GBR is likely to deteriorate further 5 before it stabilizes.

Coral adaptation and acclimatization may be the only realistic prospect for the conservation of some parts of the GBR this century. However, although adaptation opportunities may be plausible to some extent 47 , they are no panacea because evolutionary changes to fundamental variables such as temperature take decades, if not centuries, to occur, especially in long-lived species such as reef-building corals 48 . There is currently no clear evidence of the real-time evolution of thermally tolerant corals 48 . Most rapid changes depend on a history of exposure to key genetic types and extremes, and there are limitations to genetic adaptation that prevent species-level adaptation to environments outside of their ecological and evolutionary history 19 . Model projections also indicate that rates of coral adaptation are too slow to keep pace with global warming 49 . In a rapidly warming world, the temperature conditions that give rise to mass coral bleaching events are likely to soon become commonplace. So, although we may see some resilience of coral to future marine heat events through acclimatization, thermal refugia are likely to be overwhelmed 50 . Global warming of more than 1.5 °C above pre-industrial levels will probably be catastrophic for coral reefs 44 .

Our new multi-century reconstruction illustrates the exceptional nature of ocean surface warming in the Coral Sea today and the resulting existential risk for the reef-building corals that are the backbone of the GBR. The reconstruction shows that SSTs were relatively cool and stable for hundreds of years, and that recent January–March ocean surface heat in the Coral Sea is unprecedented in at least the past 400 years. The coral colonies and reefs that have lived through the past several centuries, and that yielded the valuable Sr/Ca and δ 18 O data on which our reconstruction is based, are themselves under serious threat. Our analysis of climate-model simulations confirms that human influence is the driver of recent January–March Coral Sea surface warming. Together, the evidence presented in our study indicates that the GBR is in danger. Given this, it is conceivable that UNESCO may in the future reconsider its determination that the iconic GBR is not in danger. In the absence of rapid, coordinated and ambitious global action to combat climate change, we will likely be witness to the demise of one of Earth’s great natural wonders.

Instrumental observations

The Coral Sea and GBR area-averaged monthly SSTAs relative to 1961–90 for January–March are obtained from version 5 of the Extended Reconstructed Sea Surface Temperature dataset (ERSSTv5) 27 . We compare our results using ERSSTv5 with those generated using the Hadley Centre Sea Ice and Sea Surface Temperature dataset (HadISST1.1) 28 . We use only post-1900 instrumental SST observations here. Although gridded datasets have some coverage before 1900, ship-derived temperature data in the region for that period are too sparse to be reliable for calibrating our reconstruction (Supplementary Information section  1.2 ). The regional mean for the GBR is computed using the seven grid-cell locations used by the Australian Bureau of Meteorology (Supplementary Information section  1.1 ). We define the Coral Sea region as the ocean areas inside 4° S–26° S, 142° E–174° E.

Coral-derived temperature proxy data

We use a network of 22 published and publicly available sub-annual and annual resolution temperature-sensitive coral geochemical series (proxies; Fig. 2d , Supplementary Tables 1 and 2 , and Supplementary Fig. 5a–v ) from the western tropical Pacific in our source data region (4° N–27° S, 134° E–184° E) that cover at least the period from 1900 to 1995. Of these 22 series, 16 are δ 18 O, which are in per mil (‰) notation relative to Vienna PeeDee Belemnite (VPDB) 51 ; the remaining six are Sr/Ca series. The coral data are used as predictors in the reconstruction of January–March mean SSTAs in the Coral Sea region. We apply the inverse Rosenblatt transformation 52 , 53 to the coral data to ensure that our reconstruction predictors are normally distributed. Sub-annually resolved series are converted to the annual time step by averaging across the November–April window. This maximizes the detection of the summer peak values, allowing for some inaccuracy in sub-annual dating and the timing of coral skeleton deposition 54 , 55 . A small fraction (less than 0.8%) of missing data is infilled using the regularized expectation maximization (RegEM) algorithm 56 (Supplementary Information section  2.3 ), after which the proxy series are standardized such that each has a mean of zero and a standard deviation of one over their common 1900–1995 period.

Reconstruction method

To produce our Coral Sea reconstruction, we use nested principal component regression 57 (PCR), in which the principal components of the network of 22 coral proxies are used as regressors against the target-region January–March SSTA relative to the 1961–90 average. We perform the reconstructions separately for each nest of proxies, where a nest is a set of proxies that cover the same time period. The longest nest dates back to 1618, when at least two series are available. The nests allow for the use of all coral proxies over the full time period of their coverage. The 96-year portion of the instrumental period (1900–1995) that overlaps with the reconstruction period is used for calibration and evaluation (or equivalently, verification) against observations. We reconstruct regional SSTAs from the principal components of the coral network of δ 18 O and Sr/Ca data, rather than their local SST calibrations, to minimize the number of computational steps and to aid in representing the full reconstruction uncertainty.

Principal component analysis (PCA) is used to reduce the dimensionality of the proxy matrix, as follows. Let P ( t , r ) denote the palaeoclimate-data matrix during the time period t  = 1,..., n at an annual time step for proxy series r  = 1,..., p . PCA is undertaken on this matrix during the calibration period, P cal . We obtain the principal component coefficients matrix P coeff ( r , e ) for principal components e  = 1,..., n PC and principal component scores P score ( t , e ), which are representations of the input matrix P cal in the principal component space. P score is truncated to include n PC,use principal components to form \({P}_{{\rm{score}}}^{{\prime} }\) such that the variance of the proxy network explained by the n PC,use principal components is greater than \({\sigma }_{{\rm{expl}}}^{2}\) (which we set to 95%). Reconstruction tests in which \({\sigma }_{{\rm{expl}}}^{2}\) is varied from 70% to 95% show that our results are not strongly sensitive to this choice, and tests based on lag-one autoregressive noise for \({\sigma }_{{\rm{expl}}}^{2}\) from 50% to 99% further support this choice (Supplementary Information section  3.2 ). These principal components are used as predictors against which the Coral Sea January–March instrumental SSTAs are regressed. We regress the standardized SSTA target data during the calibration period, I cal , against the retained principal components of the predictor data, \({P}_{{\rm{score}}}^{{\prime} }\) :

Thus, we obtain n PC,use estimates of the regression coefficients γ e with gaussian error term ε t  ~  N (0, \({\sigma }_{N}^{2}\) ). The principal components are extended back into the pre-instrumental period by multiplying the entire proxy matrix P ( t , p ) with the truncated principal component coefficient matrix \({P}_{{\rm{coeff}}}^{{\prime} }\) ( t , e ) to obtain \({Q}_{{\rm{coeff}}}^{{\prime} }\) :

The reconstruction proceeds with the fitted regression coefficients γ e and extended coefficient matrix \({Q}_{{\rm{coeff}}}^{{\prime} }\) to obtain a reconstruction time series R m ( t ) for a given nest of proxy series

The standardized reconstruction R m ( t ) is then calibrated to the instrumental data such that the standard deviation and mean of the reconstruction and target during the calibration interval are equal. As well as obtaining reconstructions for each nest of available proxies, we compute stitched reconstructions S c ( t ) for each calibration period c , which include at each time step the reconstructed data for the proxy nest with maximum coefficient of efficiency 58 , 59 (Supplementary Information section  3.1 ). This procedure is performed for contiguous calibration intervals between 60 and 80 years duration between 1900 and 1995, with interval width and location increments of two years, reserving the remaining data in the overlapping period for independent evaluation, and for all proxy nests. The reconstruction error is modelled with a lag-one autoregressive process fitted to the residuals. We evaluate the capacity of our reconstruction method to achieve spurious skill from overfitting by performing a test in which we replace the coral data with synthetic noise (Supplementary Information section  3.2i ). We find that reconstructions based on synthetic noise achieve extremely low or zero skill and as more noise principal components are included in the regression, the evaluation metrics indicate declining skill. Our reconstruction and evaluation methods therefore guard against the potential for spurious skill.

Pseudo-proxy reconstructions

Our reconstruction method is further evaluated by using a pseudo-proxy modelling approach based on the Community Earth System Model (CESM) Last Millennium Experiment (LME) 60 , for which there are 13 full-forcing ensemble members covering the period 850–2005. We use the pseudo-proxy reconstructions to evaluate our reconstruction method and coral network in a fully coupled climate-model environment. We form pseudo-proxies by extracting from each LME ensemble member the SST and sea surface salinity (SSS) from the 1.5° × 1.5° grid cell located nearest to our coral data. We then apply proxy system models in the form of linear regression models, basing δ 18 O on both SST and SSS, and Sr/Ca on SST only (Supplementary Information section  3.3 ). We set the spatial and temporal availability of the pseudo-coral network to match that of the coral network. We then apply our PCR reconstruction and evaluation procedure to the pseudo-proxy network, taking advantage of the availability of the modelled Coral Sea SSTA data across the multi-century period of 1618–2005, which allows for the evaluation of the pseudo-proxy reconstruction over this entire time period. We first test our method using a ‘perfect proxy’ approach (with no proxy measurement error) before superimposing synthetic noise on the pseudo-proxy time series, evaluating our methodology at two separate levels of measurement error, quantified by signal-to-noise ratios of 1.0 and 4.0. The evaluation metrics for these tests indicate that our coral network and reconstruction method obtain skilful reconstructions of Coral Sea SSTAs in the climate-model environment (Supplementary Figs. 17b , 18 , 20b , 21 , 22b and 23 ).

Comparison with independent coral datasets

We use two multi-century five-year-resolution coral series from the central GBR 23 , 24 (Fig. 2b and Supplementary Fig. 24 ) and a network of sub-annual and annual resolution modern coral series (dated from 1900 onwards but not covering the full 1900–1995 period) from 44 sites in the GBR (Supplementary Information section  4.2 ) for independent evaluation of coral-derived evidence for warming in the region. We estimate five-year GBR SSTAs (Fig. 2b ) by aligning the post-1900 mean and variance of the proxy and instrumental (ERSSTv5) data.

Reconstruction sensitivity to non-SST influences

Of the 22 available coral series, 16 are records of δ 18 O, a widely used measure of the ratio of the stable isotopes 18 O and 16 O. In the tropical Pacific Ocean, δ 18 O is significantly correlated with SST 61 , 62 , 63 , 64 . Coral δ 18 O is also sensitive to the δ 18 O of seawater 65 , which can reflect advection of different water masses and/or changes in freshwater input, such as from riverine sources or precipitation, which in turn co-vary with SSS. Thus, it is generally considered that the main non-SST contributions to coral δ 18 O are processes that co-vary with SSS 62 , 66 . Our methodology minimizes the influence of non-temperature impacts on the reconstruction by exploiting the contrast in spatial heterogeneity between SST and SSS in January–March (Supplementary Information section  5.1 ). SSS is spatially inhomogeneous in the tropical Pacific 66 , 67 , leading to low coherence in SSS signals across our coral network. By contrast, the strong and coherent SST signal across our coral network locations and the Coral Sea region leads to principal components that are strongly representative of SST variations. This produces a skilful reconstruction of SST, as determined by evaluation against independent observations, and low correlations with SSS across the Coral Sea region (Supplementary Fig. 31 ).

Although the likelihood of non-SST influences on our SST reconstruction is low, we nonetheless test the sensitivity of our reconstruction and its associated interpretations to the possibility of these influences on the coral data. The tests compute the correlations between our best-estimate SSTA reconstruction (highest coefficient of efficiency) and observations of SSS, along with a series of additional reconstructions based on subsets of our coral network. The correlations between our highest coefficient of efficiency January–March Coral Sea SSTA reconstruction and January–March SSS are mapped for the Coral Sea and its neighbouring domain using three instrumental SSS datasets (Supplementary Fig. 31 ). Correlations are not statistically significant over most of the domain. Noting differing spatial correlation patterns between the instrumental SSS datasets 68 , which also cover different time periods (Supplementary Information section  5.1 ), we undertake six sensitivity tests using subsets of the coral network (Supplementary Information section  5.2 ). We use the following combinations of coral series: (1) the full network of 22 δ 18 O and Sr/Ca series (Figs. 2a and 3 ); (2) a subset of the six available Sr/Ca series (Supplementary Figs. 32 – 33 ), to test how the reconstruction is influenced by the inclusion of coral δ 18 O records; (3) a fixed nest subset of the five longest coral series, extending back to at least 1700 (Supplementary Figs. 34 – 35 ), to test for the potential influence of combining series of differing lengths (from our splicing of portions of the best reconstructions from each nest); (4) a subset of the ten coral series that are most strongly correlated with the target (Supplementary Figs. 36 and 37 ), to test how our reconstruction is influenced by the inclusion of coral series that are less strongly correlated with our target; (5) a subset of coral series that excludes the six records that are reported to potentially include biological mediation or non-climatic effects, or have low correlation with the target (Supplementary Figs. 38 and 39 ), to test their influence on the reconstruction; and (6) a network perturbation test comprising 22 separate subsets of proxies, in which proxy records are added incrementally in order of highest to lowest correlation with the target, starting with a single coral series and increasing the number of included proxies to all 22 series in our network (Supplementary Information section  5.2.5 ), to systematically quantify the influence of gradually including more coral datasets on our reconstruction and its interpretations.

The evaluation metrics (Fig. 2c and Supplementary Figs. 32b , 34b , 36b and 38b ) indicate a skilful reconstruction back to 1618 for the reconstructions based on the Full, Sr/Ca only, Long, Best-10 and OmitBioMed networks. These reconstructions explain 82.7%, 80.6%, 77.6%, 79.8% and 80.4% (R-squared values) of the variance in January–March SSTAs, respectively, in the independent evaluation periods (using ERSSTv5b). All coral subsets in the network perturbation test produce skilful reconstructions (Supplementary Fig. 40 ). The highest-skill reconstructions for all subsets in the network perturbation test align with our key interpretations (Supplementary Figs. 41 and 42 ). Together, our sensitivity tests show that the coral network, observational data and reconstruction methodology are a sound basis for reconstructing Coral Sea January–March SSTAs in past centuries and contextualizing recent high-SST events ( Supplementary Information ).

Climate-model attribution ensembles and experiments

The multi-model attribution analysis used here is based on simulations from CMIP6. We analyse simulations from the historical experiment (including natural and anthropogenic influences for 1850–2014) and the historical-natural experiment (natural-only forcings for 1850–2014). We select climate models for which monthly surface temperature is available in at least three historical and historical-natural simulations (Supplementary Table 5 ). All model simulations are interpolated to a common regular 1.5° × 1.5° latitude–longitude grid. January–March SSTAs relative to 1961–90 are calculated for each simulation. The full historical all-forcings ensemble is composed of 14 models with 268 simulations for 1850–2014. The natural-only ensemble is composed of the same 14 models with 95 individual simulations. A subset of climate models in the CMIP6 ensemble are considered by the science community to be ‘too hot’, simulating warming in response to increased atmospheric carbon dioxide concentrations that is larger than that supported by independent evidence 31 . We omit these models from our analysis by including only models with a transient climate response in the ‘likely’ range 31 of 1.4–2.2 °C. Our results are not strongly sensitive to this selection (Supplementary Information section  6.3 ). The ten remaining models yield a total of 25,410 years from 154 historical ensemble members and 7,095 years from 43 historical-natural ensemble members. We weight the models equally in our analysis using bootstrap sampling. We report linear trends based on simple linear regression models fitted with ordinary least squares. The statistical significance of linear trends is assessed using the Spearman’s rank correlation test 69 .

Time of emergence of the anthropogenic impact

We assess the anthropogenic influence on SSTAs in the Coral Sea region by starting with the assumption that any anthropogenic influence on SSTAs in the Coral Sea is indistinguishable from natural variability at the commencement of the model experiments. We measure the impact of anthropogenic influence on the climate in the region using a signal-to-noise approach 32 , 70 . We calculate the anthropogenic ‘signal’ as the mean of the difference between the smoothed (using a 41-year Lowess filter) modelled historical Coral Sea SSTA and the mean smoothed modelled historical-natural SSTA. Our ‘noise’ is the standard deviation of the difference between the modelled historical SSTA and its smoothed time series (Supplementary Information section  6 ).

Methods additionally rely on Supplementary Information and refs. 71 , 72 , 73 , 74 , 75 , 76 , 77 , 78 , 79 , 80 , 81 , 82 , 83 , 84 , 85 , 86 , 87 , 88 , 89 , 90 , 91 , 92 , 93 , 94 , 95 , 96 , 97 , 98 , 99 , 100 , 101 , 102 , 103 , 104 .

Data availability

The ERSSTv5 instrumental SST data are available from the US National Oceanic and Atmospheric Administration at https://psl.noaa.gov/data/gridded/data.noaa.ersst.v5.html . The HadISST1.1 data are available from the UK Met Office at https://www.metoffice.gov.uk/hadobs/hadisst/ . The original coral palaeoclimate data are available at the links provided in Supplementary Table 2 . Land areas for maps are obtained from the Mapping Toolbox v.23.2 in Matlab v.2023b and the Global Self-consistent, Hierarchical, High-resolution Geography (GSHHS) Database at https://www.soest.hawaii.edu/pwessel/gshhg/ through the m_map toolbox by R. Pawlowicz, available at https://www.eoas.ubc.ca/%7Erich/map.html . Prepared data from the coral geochemical series, reconstructions and climate models that support the findings of this study are available at: https://doi.org/10.24433/CO.4883292.v1 .

Code availability

The code that supports the findings of this study is available and can be run at : https://doi.org/10.24433/CO.4883292.v1 .

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Acknowledgements

We acknowledge the originators of the coral data cited in Supplementary Tables 1 and 2 ; S. E. Perkins-Kirkpatrick and the deceased G. J. van Oldenborgh 105 for contributions to an earlier version of this manuscript; E. P. Dassié and J. Zinke for discussions and data; R. Neukom for advice on an earlier version of the reconstruction code; and B. Trewin and K. Braganza for advice about the Bureau of Meteorology GBR SST time series. B.J.H. and H.V.M. acknowledge support from an Australian Research Council (ARC) SRIEAS grant, Securing Antarctica’s Environmental Future (SR200100005), and ARC Discovery Project DP200100206. A.D.K. acknowledges support from an ARC DECRA (DE180100638) and the Australian government’s National Environmental Science Program. B.J.H. and A.D.K. acknowledge an affiliation with the ARC Centre of Excellence for Climate Extremes (CE170100023). H.V.M. acknowledges support from an ARC Future Fellowship (FT140100286). A.K.A. acknowledges support from an Australian government research training program scholarship and an AINSE postgraduate research award. Funding was provided to B.K.L. by the Vetlesen Foundation through a gift to the Lamont-Doherty Earth Observatory. Grants to B.K.L. enabled the generation of coral oxygen isotope and Sr/Ca data from Fiji that were used in our reconstruction (US National Science Foundation OCE-0318296 and ATM-9901649 and US National Oceanic and Atmospheric Administration NA96GP0406). We acknowledge the support of the NCI facility in Australia and the World Climate Research Programme’s working group on coupled modelling, which is responsible for CMIP. We thank the climate-modelling groups for producing and making available their model output. For CMIP, the US Department of Energy’s Program for Climate Model Diagnosis and Intercomparison provided coordinating support and led the development of software infrastructure in partnership with the Global Organisation for Earth System Science Portals.

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Environmental Futures, School of Earth, Atmospheric and Life Sciences, University of Wollongong, Wollongong, New South Wales, Australia

Benjamin J. Henley, Helen V. McGregor & Ariella K. Arzey

Securing Antarctica’s Environmental Future, University of Wollongong, Wollongong, New South Wales, Australia

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Benjamin J. Henley

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Andrew D. King & David J. Karoly

ARC Centre of Excellence for Climate Extremes, University of Melbourne, Parkville, Victoria, Australia

Andrew D. King

School of the Environment, The University of Queensland, Brisbane, Queensland, Australia

Ove Hoegh-Guldberg

Australian Institute of Marine Science, Townsville, Queensland, Australia

Janice M. Lough

ARC Centre of Excellence for Coral Reef Studies and School of Earth Sciences, University of Western Australia, Crawley, Western Australia, Australia

Thomas M. DeCarlo

Department of Earth and Environmental Sciences, Tulane University, New Orleans, LA, USA

Lamont-Doherty Earth Observatory of Columbia University, Palisades, NY, USA

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Contributions

B.J.H., H.V.M. and A.D.K. conceived the study and developed the methodology. B.J.H. did most of the analysis. A.K.A. contributed analysis of modern coral data (Supplementary Information section  4.2 ). T.M.D. contributed analysis of instrumental data coverage (Supplementary Information section  1.2 ). B.K.L. contributed sub-annual coral data. B.J.H. and H.V.M. led the preparation of the manuscript, with contributions from A.D.K., O.H.-G., A.K.A., D.J.K., J.M.L., T.M.D. and B.K.L. Generative artificial intelligence was not used in any aspect of this study or manuscript.

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Henley, B.J., McGregor, H.V., King, A.D. et al. Highest ocean heat in four centuries places Great Barrier Reef in danger. Nature 632 , 320–326 (2024). https://doi.org/10.1038/s41586-024-07672-x

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DOI : https://doi.org/10.1038/s41586-024-07672-x

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