Coral reefs throughout the world are facing the consequences of changing Earth’s climate. The ENSO phenomenon, which is a natural periodic fluctuation in sea surface temperature (El Niño) and air pressure of the overlying atmosphere (Southern Oscillation) across the equatorial Pacific Ocean (Dijkstra, 2006), has important consequences for the climate around the globe. Ocean warming worldwide and particularly intense ENSO episodes are leading to an increase in the severity, duration and frequency of coral bleaching events, which threatens the long-term stability of coral reefs, hampering their resilience to local human pressure (Hughes et al., 2018a). Reefs in the world have recently been affected by three major global bleaching events, in 1998, 2010, and 2016. The U.S. National Oceanic and Atmospheric Administration (NOAA) declared that the last bleaching event can be considered as the longest, the most widespread, and probably the most damaging on record (Hughes et al., 2017). More than 70% of coral reefs around the world bleached and experienced catastrophic levels of coral mortality (Eakin et al., 2016).

Corals suffer thermal stress when sea surface temperatures exceed the seasonal means and protract for extended periods to cause the disruption of the mutualistic symbiosis between corals and their dinoflagellate endosymbionts (zooxanthellae), resulting in the expulsion of zooxanthellae and consequent coral mortality (LaJeunesse et al., 2018). The relationship among heat exposure, bleaching, and consequent mortality of different taxa is still not well understood (Hughes et al., 2018b). However, a protracted value of 1°C above the maximum summer mean has been often described as high enough to cause coral bleaching (Kayanne, 2017), depending also upon sea and whether conditions. Current projections foresee the collapse of coral reefs by the end of this century due to the increased frequency and severity of coral bleaching events worldwide (Hoegh-Guldberg, 1999; Hoegh-Guldberg et al., 2007; Veron et al., 2009; Bay et al., 2017; Heron et al., 2017). However, these projections do not consider the potential of corals to survive a changing climate by adapting and/or developing responses apt to increase their resilience to climate-driven stressors. In this regard, two hypotheses have been proposed: (i) the adaptive bleaching hypothesis (hereafter ABH); and (ii) the deep refuge hypothesis (hereafter DRH).

Although coral reefs are confined to stable habitats and narrow environmental conditions, the ABH postulates that the loss of one or more zooxanthellae strains, rapidly replaced by different ones with the formation of a new symbiotic consortium, has the potential to create new and more tolerant organisms, capable of surviving a variety of extreme conditions (Buddemeier and Fautin, 1993). Capacity to adapt ensured coral reef communities to persist over geological time scales through significant climate and sea-level fluctuations (Buddemeier et al., 2004); however, oceans are presently warming at a faster rate with respect to the geological past and decadal evolutionary responses to warming may drive the selection of the most adequate species able to adapt rapidly to environmental change (Padfield et al., 2015). Genetic adaptation to sea warming has been predicted (Matz et al., 2018), and reduced bleaching and mortality by some corals exposed to previous bleaching events has been often reported (Coles et al., 2018, and references therein). Adaptive bleaching mechanisms are not common in all corals (McClanahan and Muthiga, 2014; McClanahan, 2017), and shifts in species composition and decline in coral diversity are predicted to occur as temperature increases (Bahr et al., 2016; Richardson et al., 2018).

According to the DRH, marine organisms in deeper habitats are buffered from high surface temperatures and increased temperature variability (Bongaerts et al., 2010; Morri et al., 2017; Bianchi et al., 2019, and references therein). Mesophotic coral ecosystems are light-dependent communities developing at tropical and subtropical latitudes that extend from 30 m depth to as deep as about 165 m (Kahng et al., 2010), at the lower limit of the photic zone (Hinderstein et al., 2010; Loya et al., 2016; Laverick et al., 2017). Although working at mesophotic depths is constrained by traditional scuba limits, investigations on mesophotic coral reefs have grown exponentially in recent years due to advances in technology and because of the increasing recognition of their potential role as refuge for shallow-water reefs (Hoegh-Guldberg and Bruno, 2010; Loya et al., 2016; Muir et al., 2017, 2018). Corals show bleaching and mortality especially in shallow waters (Jokiel, 2004) whilst the occurrence of bleaching events and consequent mortality is normally lower at greater depths (Baird et al., 2018). DRH therefore suggests that mesophotic coral ecosystems may provide refuge and a source of recruits to facilitate recovery of shallow-water reefs damaged after a disturbance (Glynn, 1996; Riegl and Piller, 2003; Holstein et al., 2015; Thomas et al., 2015; Prasetia et al., 2017; Baird et al., 2018). Mesophotic reefs may display an important degree of overlap in species composition, and maintain connectivity with shallow reefs (Lesser et al., 2018; Muir et al., 2018), thus contributing to the local persistence of heat sensitive species despite periods of elevated seawater temperatures during strong ENSO episodes (Gatti et al., 2017; Shlesinger et al., 2018). Most families of scleractinian corals have been shown to be represented in the mesophotic zone of the Great Barrier Reef, with many also extending to greater depths, thus giving each evolutionary lineage some potential for being safeguarded in the case of shallow-reef degradation (Muir et al., 2018).

Repeated heat waves in the Maldives during the last decades provided the opportunity to better understand the impact of bleaching and mass mortality on coral communities. The general trend of coral recovery after the severe bleaching event of 1998 has already been investigated in previous papers (Bianchi et al., 2003, 2006, 2017; Lasagna et al., 2008), while the bleaching event of 2010 was comparatively milder and did not cause widespread mortality (Morri et al., 2015). The consequences of the further severe bleaching event of 2016 have been analyzed by Montefalcone et al. (2018a), especially with respect to the constructional potential of Maldivian reefs. The availability of a twenty-three year series of data on coral cover in the Maldives, encompassing the two most important bleaching events of 1998 and 2016 allowed comparing their impact and exploring the following hypotheses: (1) coral mortality after the 2016 bleaching event should be lower than after the 1998 event, according to the ABH; (2) coral mortality in deep waters should be lower than in shallow waters, meaning that mesophotic reefs can offer refuge to thermal stress as postulated by the DRH. Clearly, we did not analyze the ecophysiological mechanisms of ABH and DRH but simply concentrated on their manifest consequence: increased coral survival.

This study aims at: (i) reconstructing the thermal regime in the Maldives to identify the heat waves that caused severe bleaching and mass coral mortality in the last two decades; (ii) reconstructing the trend of human pressure (tourism and resident population) in the Maldives to be compared with coral mortality rates in coincidence with the bleaching events; (iii) evaluating the trajectories of change and recovery patterns in shallow reef communities after the two mass mortality events; (iv) investigating geographical patterns of hard coral survival during the last bleaching event according to location (i.e., atolls with high or low human pressure) and reef exposure (i.e., lagoon- or ocean-facing reefs); (v) discussing the ABH using our long-term series of coral cover data collected on shallow reefs at about 5 m depth; (vi) discussing the DRH by analyzing coral cover reduction and mortality along depth profiles, from 10 to 50 m, surveyed in years with and without coral mortality.

Maldivian coral reefs experienced two severe bleaching events in 1998 and in 2016. After the 1998 bleaching event, more than 90% of hard corals died (Bianchi et al., 2003, 2006), and it took 16 years for reefs to recover the pre-bleaching values of live hard coral cover (Morri et al., 2015). Such a recovery was consistent with that observed in the eastern tropical Pacific Ocean (Romero-Torres et al., 2020), but was surprisingly slow as compared to the fast coral cover recovery observed in the neighboring Chagos Archipelago (Sheppard et al., 2008) and in other remote Indian Ocean locations with similar oceanographic conditions (Gilmour et al., 2013). All these latter studies concerned nearly uninhabited islands; on the contrary, the Maldives are experiencing a still moderate but nevertheless continuously increasing level of human pressure because of population increase, coastal development, and tourism intensification (Jaleel, 2013; Nepote et al., 2016). In addition, starting from 2016, the atolls of North and South Malé have been subject to important land reclamation engineering works with consequent increase of sediments in the lagoons (Pancrazi et al., 2020).

Visual estimation of substratum cover is among the most universally used metric to quantify sessile benthic organisms (Bianchi et al., 2004), but might be an insufficient descriptor of recovery when considered alone (Johns et al., 2014). In the Maldives, coral species richness was recovered after 4 years (Benzoni and Pichon, 2007) and Acropora colonies density and size after 6 years (Lasagna et al., 2010b). However, recover of all descriptors at ecosystem level (community structure, seascape, trophic organization, and structural complexity) took 14 to more than 16 years and was rarely complete (Bianchi et al., 2017). While previous papers already investigated the general trend of coral recovery in the Maldives after the mass bleaching event in 1998 (Lasagna et al., 2008; Morri et al., 2015; Bianchi et al., 2017), this study showed that during the 2016 bleaching event, 60% of hard corals in the Maldives bleached and died when the accumulated heat exposure exceeded the critical regional bleaching threshold of degree heating weeks (DHW = 8.0) for 2 months. Hard corals bleached in almost all reefs surveyed during 2016, although differences in coral mortality were observed among the reefs investigated in the subsequent years. Local ecological and environmental differences influenced the severity of coral bleaching in the various areas (McClanahan et al., 2018). Reefs in remote areas or affected by a comparatively lower degree of human pressure, such as in Felidhoo and Ari atolls, kept a higher value of live hard coral cover than reefs in more developed and urbanized areas, such as in North and South Malé atolls (Dhunya et al., 2017; Stevens and Froman, 2019). Reefs affected by local human pressure have already been shown to be more vulnerable to climate disturbance (Montefalcone et al., 2011; Nepote et al., 2016).

Corals in ocean reefs proved more resistant to the 2016 bleaching event than those in lagoon reefs. Ocean reefs are close to deep cooler waters, have higher water movement and waves that may attenuate light intensity and ensure mixing of water, and are generally unaffected by nutrients coming from land: all factors that can potentially reduce bleaching and mortality (Muir et al., 2017). Differences between ocean and lagoon reefs are expressed not only in morphology, topography and exposure (Lasagna et al., 2008, 2010a; Rovere et al., 2018), but also in the species composition of the coral communities. Maldivian lagoon reefs are typically dominated by tabular and branching Acropora corals (Bianchi et al., 1997; Lasagna et al., 2010b). Susceptibility to thermal stress of Acropora corals is known to be widely variable (Muir et al., 2018), but in the Maldives they suffered a catastrophic die-off in both 1998 (Tkachenko, 2014) and 2016 (Pisapia et al., 2019), as already reported by studies in other localities (Loya et al., 2001; Pratchett et al., 2013). On the contrary, ocean reefs are dominated by massive corals (Bianchi et al., 1997), which experienced only partial colony mortality after the two bleaching events (Bianchi et al., 2006; Perry and Morgan, 2017; Pisapia et al., 2019). Higher resistance of ocean reefs is also likely linked to higher propagule exchanges due to stronger currents and water movements in comparison to lagoon reefs (Kinlan and Gaines, 2003).

Although the 2016 heat wave has been declared as the most damaging on record (Eakin et al., 2016; Hughes et al., 2017), corals in the Maldives survived to a larger extent with respect to 1998. According to our calculation for the Maldives, the thermal anomaly of 1998 was distinctly higher than that of 2016 in terms of DHW, but the two were comparable according to the NOAA coral reef watch database (which, however, considered an area wider than the Maldives). In any case, the 1998 heat wave was a relatively longer event, whilst the 2016 heat wave was less prolonged. The lower coral mortality in 2016 might give some support for the ABH in some species of shallow Maldivian coral reefs, as already hypothesized by McClanahan and Muthiga (2014). However, differences in duration and intensity between the two heat waves hamper a definite answer. Coral adaptability might be not the only explanation for a lower coral mortality in 2016: processes of acclimatization, ecological reorganization, and an effective coral heterotrophy (Houlbrèque and Ferrier-Pagès, 2009; Grottoli et al., 2014; McClanahan, 2017; Coles et al., 2018) have also been evoked.

In the Maldives (as elsewhere in the world) shallow reefs in the upper photic zone, within 10 m depth, are usually the target of most coral-reef monitoring programs (Jimenez et al., 2012; Tkachenko, 2012). Due to scuba diving constrains, comparatively fewer explorations have been conducted in the lower photic zone, below 10 m depth, where coral reefs can still be constructional (Morri et al., 1995; Bianchi et al., 1997) and even exhibit high coral cover in some cases (Sheppard, 1980). Although mesophotic reefs are rarely included in reef assessments (Pyle et al., 2016; Garavelli et al., 2018; Studivan and Voss, 2018a), there is a growing interest to investigate the upper mesophotic coral-reef ecosystems, 30 m to 50 m depth, for their potential to serve as thermal refuge (Semmler et al., 2016). Nevertheless, ecological information on mesophotic reefs in the Maldives is poor compared to that on shallow reefs (Morri et al., 1995, 2010; Bianchi et al., 1997).

Our results showed significantly reduced bleaching in the upper mesophotic reefs and a negligible coral mortality after the two bleaching events at depths between 30 and 50 m, thus providing some support to DRH. In many coral reefs of the world, the upper mesophotic reef community represents an extension of the shallow-water community (Pyle et al., 2016; Lesser et al., 2018), as many coral species are shared across these depths (Kahng et al., 2017; Studivan and Voss, 2018b; Zlatarski, 2018). However, our results also indicated that coral cover reduced drastically with depth from shallow to deep waters, with cover values always <20% in the latter. These values are lower than that of the survivors on shallow reefs. Even assuming a great fecundity for mesophotic coral species, there is no reason to think that their potential for contributing to reef recovery after severe events surpasses that of the survivors in photic reefs.

Although widely used as indicator of reef health (Montefalcone et al., 2018a), coral cover does not account for differences in susceptibility between species. Susceptibility of the different species to bleaching over depth has never been widely quantified, and available information on deep reefs during severe bleaching events is limited (Muir et al., 2017; Morais and Santos, 2018). Except for the ultrasensitive fire corals of the genus Millepora (Smith et al., 2014; Morri et al., 2017), some recent studies suggest that the applicability of the DRH may be only site- and species-specific (Semmler et al., 2016), and should not be considered as a general phenomenon (Smith et al., 2016; Bongaerts et al., 2017). Some controversy occurs regarding the hypothesis that deep reefs can provide a refuge from bleaching. An important role of mesophotic habitats has been demonstrated in the Indo-Pacific Ocean due to the high species richness of scleractinian corals, particularly at 30–45 m depths, able to preserve evolutionary lineages (Muir et al., 2018). During the 2016 bleaching event, 73% of the Maldivian coral species at 24–30 m depth have been not affected by bleaching (Muir et al., 2017). Deep areas of the ocean-side reefs on the Great Barrier Reef provided a refuge from the 2016 bleaching event for some colonies of the most abundant and ecologically important coral genera, e.g., Acropora, Pocillopora, and Porites (Baird et al., 2018). The survivors at depth may well be species that have been severely depleted in shallow reefs, thus making their contribution potentially critical. Shallow-reef assemblages in the Maldives share species with reefs at 30 m depth, but there are species that are not depth generalists (Benzoni and Pichon, 2007; Bigot and Amir, 2012): contrasting results might reflect local variations in both species composition and depth distribution of individual species down the reef slope (Laverick et al., 2018; Rocha et al., 2018). More data on recovery through time and at different depths would be necessary to further test this hypothesis and to understand to which extent mesophotic Maldivian reefs may act as refuge to safeguard shallow reefs.

Similarly, to investigate the ABH, estimates on coral cover can only provide a first picture on the effects of thermal stress. Further investigations on community taxonomic composition and on the relative sensitivity of species to thermal stress should be envisaged to prove both the ABH and the DRH in the Maldivian coral reefs. Ecophysiology, molecular biology, and genetics would be fundamental approaches to prove resistance and adaptability of corals to ocean warming.

Periods of unusually high sea surface temperatures have already become long-lasting and frequent (Frölicher et al., 2018), and this trend is predicted to accelerate under future global warming scenarios (Hughes et al., 2018a). Coral tolerance limits are expected to be frequently exceeded due to the incessant increase in sea surface temperature (Heron et al., 2017) and predictions indicate that the majority of coral reefs will not survive the most pessimistic scenarios of global warming (Perry et al., 2018). Even trusting in coral adaptation to ocean warming, this process would probably be slower than the rate at which sea temperatures are currently rising - therefore it will not prevent the extinction of many stenothermal coral species (Frieler et al., 2013; Bay et al., 2017). Up to 16 years had been necessary for Maldivian coral reefs to recover from the severe bleaching event of 1998 (Morri et al., 2015; Pisapia et al., 2016; Perry and Morgan, 2017), and the predicted frequency of two severe bleaching events per decade (Hughes et al., 2018a) would thus jeopardize future recovery of Maldivian reefs (Montefalcone et al., 2018a). In addition to climatic effects, the increase in local human pressure in the last decades is cause of concern (Nepote et al., 2016; Pancrazi et al., 2020). Resident population in the Maldives, and especially in the highly anthropized atolls (such as those of Malé), is continuously growing, but this is not the case for the most remote and less developed atolls (such as Ari and Felidhoo). Tourism, in particular, has become the main component of Maldivian economy (Bertaud, 2002) and may be expected to keep growing exponentially in the near future. The only flexion in tourism growth has been in 2005, as a consequence of the tsunami of December 2004 (Morri et al., 2015), whilst in the atolls of Malé tourism has been showing a huge rise since 2013. Our study showed that the impact of 2016 bleaching and mortality was significantly greater on coral communities in atolls with higher human pressure. The observed 16 years required to Maldivian coral reefs to recovery from a severe bleaching event are likely to be not enough for recovering from the last mass bleaching event of 2016, especially for those reefs where the disturbance regime has growth exponentially. While containing climate stress requires international actions, regional management practices may prove successful in reducing local human pressure (Lasagna et al., 2014; Dhunya et al., 2017), thus making Maldivian coral reef ecosystems more resilient to climate change (Brown et al., 2013; Shaver et al., 2018). Keeping local impacts under control may represent a better conservation strategy than relying upon coral adaptive responses and depth refuge efficacy.

All datasets generated for this study are included in the article/Supplementary Material.

All authors listed have made a substantial, direct and intellectual contribution to the work, and approved it for publication.

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.