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02 agosto 2020
Il primo articolo sullo stato dei coralli maldiviani con dati raccolti oltre i 50 mt

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Influence of Local Pressures on Maldivian Coral Reef Resilience Following Repeated BleachingEvents, and Recovery Perspectives
 

Monica Montefalcone*, Carla Morri and Carlo Nike Bianchi

DiSTAV, Department of Earth, Environment and Life Sciences, University of Genoa, Genoa, Italy



Two severe heat waves triggered coral bleaching and mass mortality in the Maldives in

1998 and 2016. Analysis of live coral cover data from 1997 to 2019 in shallow (5 m

depth) reefs of the Maldives showed that the 1998 heat wave caused more than 90%

of coral mortality leaving only 6.8  0.3% of survived corals in all the shallow reefs

investigated. No significant difference in coral mortality was observed among atolls with

different levels of human pressure. Maldivian reefs needed 16 years to recover to the prebleaching

hard coral cover values. The 2016 heat wave affected all reefs investigated,

but reefs in atolls with higher human pressure showed greater coral mortality than reefs

in atolls with lower human pressure. Additionally, exposed (ocean) reefs showed lower

coral mortality than those in sheltered (lagoon) reefs. The reduced coral mortality in 2016

as compared to 1998 may provide some support to the Adaptive Bleaching Hypothesis

(ABH) in shallow Maldivian reefs, but intensity and duration of the two heat waves were

different. Analysis of coral cover data collected along depth profiles on the ocean sides

of atolls, from 10 to 50 m, allowed the comparison of coral mortality at different depths

to discuss the Deep Refuge Hypothesis (DRH). In the upper mesophotic zone (i.e.,

between 30 and 50 m), coral mortality after bleaching was negligible. However, live

coral cover did not exceed 15%, a value lower than coral survival in shallow reefs.

Low cover values of corals surviving in the mesophotic reefs suggest that their role as

refuge or seed banks for the future recovery of some species in shallow-water reefs of

the Maldives may be small. The repeatedly high coral mortality after bleaching events

and the long recovery period, especially in sites with human pressure, suggest that

the foreseen increased frequency of bleaching events would jeopardize the future of

Maldivian reefs, and ask for reducing local pressures to improve their resilience.

Keywords: coral reefs, local human pressure, sea water warming, adaptive bleaching hypothesis, mesophotic

reefs, deep refuge hypothesis, Maldives

INTRODUCTION

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 

and

Frontiers in Marine Science | www.frontiersin.org 2 July 2020 | Volume 7 | Article 587

Montefalcone et al. Maldivian Coral Reef Resilience

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 oceanfacing

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.

MATERIALS AND METHODS

Study Area and Field Activities

TheMaldives Archipelago consists of 27 atolls with ca. 1190 small

coral islands stretched over an area of 860 km long from about

7070 N to 0400 S in latitude and 72330 E to 73450 E in

longitude, with more than 99% of its territory covered by water.

Coral reefs of the Maldives are considered as the seventh largest

coral reef system on Earth, representing 3.14% of the World’s reef

area (Dhunya et al., 2017).

Scientific cruises took place annually in late April – early May

between 1997 and 2019. Every year, eight to eleven sites were

chosen randomly and surveyed across the atolls of Ari, Felidhoo,

Gaafu Alifu (Suvadiva), North Malé, South Malé, Rasdhoo, and

Thoddoo (Supplementary Table S1). Based on the occurrence

of inhabited islands and infrastructures, the atolls of North Malé

and SouthMalé were considered to be subjected to higher human

pressure with respect to the remaining atolls (Godfrey, 2006).

Data (see below for detail) were collected by scuba diving at reef

sites located either on the ocean-exposed reefs or in lagoon sites

(lagoon-facing sides of the atoll rim or patch reefs).

To evaluate the overall reef status of theMaldives (with a main

focus on the central atolls because of logistic constraints), our

sampling design implied a completely randomly selection of the

study sites each year, to assure data independency. Monitoring

the history of a specific reef site was indeed not within the scope

of this long series of cruises. Each year an equal number of ocean

reefs and lagoon reefs has been sampled, always distinguishing

between atolls with high or low human pressure. Although the

cruise route differed from year to year, some reefs (31%) were

casually revisited in different years (Supplementary Table S1).

In total, 168 sites were surveyed; their geographical position was

recorded using a portable GPS.

Climate and Local Human Pressure

Regimes

Sea surface temperature data was used as a proxy for

climate change (Montefalcone et al., 2018b). A twenty-three

year trend (1997–2019) of monthly maximum sea surface

temperature (SST) was plotted from data provided by the U.S.

National Oceanic and Atmosphere Administration (NOAA)

(available at http://coralreefwatch.noaa.gov/vs/gauges/maldives.

php) for the area of the Maldives Archipelago. Satellite data

were calibrated by the usual process of linear regression with

discontinuous field data on sea surface temperature from our

own archives, collected contemporaneously with the biological

data. Maximum SSTs were compared to the two regional

bleaching thresholds defined for the Maldives, corresponding

to: (i) 30.9C for severe bleaching events that are likely to

cause widespread coral mortality and live coral cover reduction

(NOAA, 2016); and (ii) 30.5C for moderate bleaching events

that usually have no wide-scale impacts on Maldivian coral reefs

(Montefalcone et al., 2018a).

Since both the intensity and the duration of thermal stress

are key factors in bleaching response, the Coral Bleaching

Degree Heating Week (DHW) has been developed by NOAA

(Kayanne, 2017) as an index of thermal stress, which measures

the amount of heat stress accumulated in an area over the

past 12 weeks. Intensity and duration of thermal stress are

combined into the DHW single number, which is thus expressed

as “degree C-weeks”. DHW values >4.0 are likely to induce

some bleaching, whilst DHW values >8.0 result in widespread

bleaching and mass coral mortality (Kayanne, 2017). To estimate

DHW for theMaldives we counted the number of weeks per year

when the monthly mean SST was higher than the temperatures

triggering moderate (30.5C) or severe (30.9C) bleaching events

in the Maldives, and we added up all temperatures exceeding

the regional bleaching thresholds during that time period.

The estimated values of DHW for the Maldives were then

compared with DHW values available from the NOAA coral reef

watch database1 considering the region of the “northern-eastern

hemisphere centre” (i.e., the central Indian Ocean) over an area

of 60 degrees  40 degrees. A linear regression was performed on

a total of 19 observations to test relationships between estimated

values of DHW for the Maldives and DHW values available from

NOAA coral reef watch.

As regards human pressure, the twenty-three year trends

(1997–2019) of the resident population (number of inhabitants)

and of the number of tourist arrivals in the Maldives were

plotted from data provided by the United Nations (Department

of Economic and Social Affairs Population Division), by the

National Bureau of Statistics in the Maldives (Department

of National Registration), and by the Ministry of Tourism

data (compiled from annual reports, available at https://www.

tourism.gov.mv/downloads/stats). For the four main central

atolls investigated in this study (North and SouthMalé, Felidhoo,

and Ari), the resident population (number of inhabitants)

obtained from the national census carried out in 1995, 2000, 2006,

and 2014 (available at http://statisticsmaldives.gov.mv/yearbook/

statisticalarchive) and the twenty-three year trends (1997–2019)

of the number of beds in the resorts were also provided.

Data Collection

Shallow reefs at 51m depth were surveyed each year.

Composition and status of reef communities were described

using live hard coral, bleached (but still alive) coral, and recently

dead coral. The latter has been defined as coral deprived of

living polyps but with the whole colony still in place and with 

a well-defined shape of the corallites (meaning that the main

erosive processes have not started yet, i.e., usually within 1 year

from coral death). The percent substratum cover for each of

the three categories was visually estimated by the plain view

technique of Wilson et al. (2007). Divers hovered 1–2 m above

the bottom observing an area of 20 m2, in three replicate spots

(tens of meters apart) at each reef site and at each sampling

time. To reduce bias in visual estimations, the same observers

collected data during the surveys from 1997 to 2013, while new

observers trained by the former estimated cover values from

2014 to 2019.

In a limited number of years (i.e., 1997, 1998, 1999, 2007,

2012, 2013, 2015, 2016, 2017, 2018, and 2019) the same surveying

method (making sure to maintain a constant depth during each

data collection) has also been used to visually estimate the percent

substratum cover of live hard coral, bleached coral, and recently

dead coral down depth profiles in each of the surveyed reefs

(Supplementary Table S1), at depths corresponding to the lower

photic reef (i.e., 10 and 20 m) and at depths corresponding to the

upper mesophotic reef (i.e., 30, 40, and 50 m) (Loya et al., 2016,

and references therein).

Data Management and Analysis

Twenty-three year (1997–2019) trends of the mean ( standard

error) percent cover of live hard coral and recently dead coral

in shallow reefs were obtained. While the severe bleaching event

of 1998 had been already described in detail in previous papers

(Bianchi et al., 2003, 2006, 2017; Lasagna et al., 2008; Morri

et al., 2015), here we analyzed with greater detail the bleaching

event of 2016, exploring patterns of live hard coral cover in

the pre-bleaching year (2015), during the event (2016), in the

post-bleaching year (2017), and in the following early recovery

years (i.e., 2018 and 2019), contrasting atolls with high or

low human pressure.

To compare the two bleaching events of 1998 and 2016, two

subsets were extrapolated from the whole dataset (Montefalcone

et al., 2018a): the pre-bleaching years (1997 vs. 2015), the years

of the bleaching (1998 vs. 2016), and the 2 years after (1999 and

2000 vs. 2017 and 2018).

The mean ( standard error) coral mortality (in %) was

computed as the ratio of recently dead coral to live coral in

the two subsets (i.e., 1997–2000 and 2015–2018), using the

following formula:

coral mortality D TRDC=.HC C BC C RDC/U  100

where RDC is the cover of recently dead coral, HC the cover of

live hard coral, and BC the cover of bleached coral.

The mean ( standard error) of the coral mortality and of

the percent cover of HC and RDC was computed down the

depth profile (10 m to 50 m) in two periods: i) years without

coral mortality (i.e., 1997, 1998, 2012, 2013, 2015, and 2019),

and ii) years with coral mortality during or short after either

the moderate or severe bleaching events (i.e., 1999, 2007, 2016,

2017, and 2018).

Non-parametric permutational analysis of variance (one-way

PERMANOVA, PRIMER6 C PERMANOVA; Anderson, 2001),

run on untransformed data using the Euclidean distance, was

used to test for differences in:

(i) live hard coral cover in shallow reefs (5 m depth) between

atolls with high human pressure (North Malé and South Malé)

and atolls with low local human pressure (2 levels) after the

bleaching event of 2016;

(ii) live hard coral cover in shallow reefs (5 m depth) between

lagoon reefs and ocean reefs (2 levels) after the bleaching

event of 2016;

(iii) coral mortality in shallow reefs (5 m depth) among the

years when coral mortality occurred (4 levels: 1999, 2016, 2017,

and 2018);

(iv) coral mortality across depths (5 levels: 10, 20, 30, 40, and

50 m) when coral mortality occurred (pooling data from 1999,

2007, 2016, 2017, and 2018).

To test for differences in the percent cover of live hard

corals, a two-way PERMANOVA was also performed with the

factor “Mortality occurrence” (2 fixed levels: years without coral

mortality and years with coral mortality during or short after

bleaching events) and the factor Depth (5 levels: 10, 20, 30, 40,

and 50 m) fixed and orthogonal.

The pair-wise test was used to discriminate among levels of

significant factors.

RESULTS

Climate and Human Pressure Regimes in

the Maldives

Peaks in maximum SST exceeded the moderate regional

bleaching threshold in various years of the investigated period,

but only the heat waves of 1998 and 2016 surpassed the severe

regional bleaching threshold of 30.9C (Figure 1). The heat wave

of 1998 lasted from April to June and reached an estimated

DHW value of 11.2, distinctly higher than the value of 8.2

reported in the NOAA database (Table 1). In 2016, we estimated

a DHW value of 8.8, comparable to that of 8.9 provided by

NOAA (Table 1), which was reached between mid-April and

mid-June 2016. Notwithstanding occasional discrepancies due

to the different size of the cells investigated, the DHW values

we estimated for the Maldives proper were positively correlated

(r = 0.96, n = 19) with those contained in the NOAA database for

the whole central Indian Ocean.

Resident population in the Maldives increased slowly but

steadily in the last decades (Figure 2A), passing from 244,814

people in the census of the year 1995 to 298,968 in 2006 and

to 344,023 in 2014; the trend has been linear, leading to an

estimated growth rate of 1.8% yearly. The same constant increase

of the resident population was observed for the atolls of North

and South Malé (Figure 2B), which have seen their population

doubled in the last 20 years with an estimated growth rate of

4.2% yearly. On the contrary, the atoll of Ari passed from 11,955

people in the census of the year 1995 to 14,050 in 2014 and that of

Felidhoo reduced from 1,678 people to 1,601 in the same period.

Tourist arrivals in the country grew exponentially, passing

from around 366 thousands in 1997 to one million and half

in 2019, with an increase rate of over 400%; a reduction in 

2005 - possibly a consequence of the December 2004 tsunami –

was rapidly recovered in 2006 (Figure 2C). The number of

beds in the resorts increased exponentially in the atolls of Ari 

and Felidhoo, and more than exponentially in North and South

Malé, where an impressive increase occurred especially after 2013

(Figure 2D); the number of beds in North and South Malé

was twice than in Ari and one order of magnitude higher than

that in Felidhoo.

Impact of Bleaching on Shallow Reefs

Live hard coral cover on shallow reefs of the Maldives dropped

from values around 70% to a value lower than 8% after the 1998

bleaching event, and gradually returned to a value comparable to

the pre-bleaching one only by 2014 (Figure 3). In 2015, recovery

was apparently complete, with live hard coral cover reaching

over 70% in several reefs. However, in reefs of North Malé atoll

subjected to high human pressure live hard coral cover hardly

exceeded 40% (Figure 4 and Supplementary Table S2). In 2016,

many hard corals died following the new mass bleaching and

dropped to a mean live cover value of around 20% (Figure 3).

Bleaching affected indistinctly all sites (Supplementary Table S2)

in both lagoon and ocean reefs. However, coral mortality after

the bleaching event showed high spatial variability (Figure 4

and Supplementary Table S2), with reefs in atolls affected by

comparatively lower human pressure exhibiting higher values

of live hard coral cover compared to reefs in the two atolls of

North Malé and South Malé, where human pressure is high on

average (Table 2). Reefs started to recover already in 2017, and

reached a mean live hard coral cover value of over 30% (Figure 3).

Recovery, however, was distinctly lower in the atolls of North

and South Malé (Figure 4) than in atolls with lower human

pressure. Pattern of recovery was also different according to reef

type: ocean reefs had suffered lower mortality with respect to

lagoon reefs (Figure 5), and were therefore able to show live

hard coral cover values higher than lagoon reefs in the immediate

aftermath of the bleaching event (Table 3). Between 2018 and

2019 live hard coral in reefs subjected to low human pressure

exhibited cover values up to 50–70%, whilst most reefs in atolls

with high human pressure hardly reached 40% (Figure 4 and

Supplementary Table S2).

After the two bleaching events of 1998 and 2016, at 5 m depth,

cover of recently dead corals was higher in 1999 than in 2017

(Figure 3). Accordingly, coral mortality was significantly higher 

in 1999 than in the remaining years with mortality (Figure 6

and Table 4). However, differently from the 1998 bleaching

event, bleached and recently dead corals were observed also in 

the 3 years following the 2016 bleaching event (Figure 3 and

Supplementary Table S2). Mortality was always higher in high

human pressure atolls than in low pressure atolls (Figure 6).

Impact of Bleaching With Depth

Hard corals bleached from shallow water down to about 30 m

depth. Consistently, bleaching-induced coral mortality was high

at 10 m and 20 m, low at 30 m and negligible at 40 m and 50 m

depth (Figure 7A and Table 5). The cover of live hard corals

decreased gradually with depth (Figure 7B). At the lower photic

depths (10 and 20 m), live coral cover decreased after the mass

mortality episodes, whilst at the upper mesophotic depths (i.e.,

30, 40, and 50 m) live hard coral cover showed no significant

decline (Figure 7B). Only at lower photic depths (10 and 20 m)

differences in live hard coral cover between years with or without

mortality were significant (Table 6). 

DISCUSSION

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 byMcClanahan 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

Frontiers in Marine Science | www.frontiersin.org 10 July 2020 | Volume 7 | Article 587

Montefalcone et al. Maldivian Coral Reef Resilience

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.

DATA AVAILABILITY STATEMENT

All datasets generated for this study are included in the

article/SupplementaryMaterial.

AUTHOR CONTRIBUTIONS

All authors listed have made a substantial, direct and intellectual

contribution to the work, and approved it for publication.

FUNDING

Part of this work received economic support from University of

Genoa internal funds (FRA, Fondi Ricerca d’Ateneo).

ACKNOWLEDGMENTS

Albatros Top Boat (Verbania, Milan and Malé) organized

our scientific cruises in the Maldives: we especially thank

Donatella ‘Dodi’ Telli, Massimo Sandrini, and Herbert Fontana

for their support, and all the staff of Conte Max and Duke

of York boats for assistance during field work. We also thank

Save the Beach Maldives, especially Beybe (Assan Hamed),

Sara Montagnani (University of Genoa), and all participants

in field activities, who helped collecting data. The advice of

Luigi Piazzi (University of Sassari) on statistical analyses is

greatly acknowledged.
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