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Bleaching drives collapse in reef carbonate budgets and reef growth potential on southern Maldives reefs

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Sea-surface temperature (SST) warming events, which are projected to increase in frequency and intensity with climate change, represent major threats to coral reefs. How these events impact reef carbonate budgets, and thus the capacity of reefs to sustain vertical growth under rising sea levels, remains poorly quantified. Here we quantify the magnitude of changes that followed the ENSO-induced SST warming that affected the Indian Ocean region in mid-2016. Resultant coral bleaching caused an average 75% reduction in coral cover (present mean 6.2%). Most critically we report major declines in shallow fore-reef carbonate budgets, these shifting from strongly net positive (mean 5.92 G, where G = kg CaCO3 m⁻² yr⁻¹) to strongly net negative (mean −2.96 G). These changes have driven major reductions in reef growth potential, which have declined from an average 4.2 to −0.4 mm yr⁻¹. Thus these shallow fore-reef habitats are now in a phase of net erosion. Based on past bleaching recovery trajectories, and predicted increases in bleaching frequency, we predict a prolonged period of suppressed budget and reef growth states. This will limit reef capacity to track IPCC projections of sea-level rise, thus limiting the natural breakwater capacity of these reefs and threatening reef island stability.
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Scientific RepoRts | 7:40581 | DOI: 10.1038/srep40581
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Bleaching drives collapse in reef
carbonate budgets and reef growth
potential on southern Maldives
reefs
C. T. Perry & K. M. Morgan
Sea-surface temperature (SST) warming events, which are projected to increase in frequency and
intensity with climate change, represent major threats to coral reefs. How these events impact reef
carbonate budgets, and thus the capacity of reefs to sustain vertical growth under rising sea levels,
remains poorly quantied. Here we quantify the magnitude of changes that followed the ENSO-
induced SST warming that aected the Indian Ocean region in mid-2016. Resultant coral bleaching
caused an average 75% reduction in coral cover (present mean 6.2%). Most critically we report major
declines in shallow fore-reef carbonate budgets, these shifting from strongly net positive (mean 5.92
G, where G = kg CaCO3 m2 yr1) to strongly net negative (mean 2.96 G). These changes have driven
major reductions in reef growth potential, which have declined from an average 4.2 to 0.4 mm yr1.
Thus these shallow fore-reef habitats are now in a phase of net erosion. Based on past bleaching
recovery trajectories, and predicted increases in bleaching frequency, we predict a prolonged period
of suppressed budget and reef growth states. This will limit reef capacity to track IPCC projections of
sea-level rise, thus limiting the natural breakwater capacity of these reefs and threatening reef island
stability.
Coral reefs support a wealth of ecosystem goods and services that extend across the provisioning of food
resources, through cultural and tourism benefits, to those associated with shoreline protection1–3. Climate
change poses a major threat to the capacity of reefs to sustain these functional roles. Warming sea waters, ocean
acidication and rising sea levels all have the potential to impact reefs across a range of spatial and temporal
scales, regardless of local management or degree of geographic isolation4–7. Critical in the context of ecosys-
tem service sustainability will be the extent to which these various climate change related stressors, acting
either independently or in tandem with other direct human-induced disturbances (such as long-term resource
over-extraction), will modify reefs both ecologically and in terms of their carbonate budgets8–10. A reefs carbonate
budget represents the balance between the rate at which carbonate is produced by corals, coralline algal and other
carbonate producing processes, less the rate at which carbonate is removed by either biological erosion (‘bioero-
sion’), physical processes, or chemical dissolution8. Such budget state measures have recently been identied as a
key metric for assessing reef “health11, but in terms of ecosystem service provisioning the budget state of a reef is
of particular relevance because of the direct inuence exerted on the capacity of reefs to maintain both their phys-
ical 3-dimensional structures and their vertical growth potential8. Shis towards net negative budgets may thus
lead to reef structural collapse, loss of reef growth potential, and diminished ecosystem service provisioning9,10,12.
Negative budget trajectories have already become evident across regions such as the Caribbean, where the
eects of prolonged and systematic ecological decline, caused both by direct human disturbances and a suite
of coral diseases, have resulted in steady transitions to states of low coral carbonate production13,14, reduced
carbonate budgets and diminished reef growth potential12. us the capacity of most Caribbean reefs to keep
pace with future sea-level rise is likely to be extremely limited12. What is more poorly understood, however,
are how individual events or disturbances may modify reef carbonate budgets. In this context, and in relation
to predictions of ever more frequent and severe sea-surface temperature warming15–17, the impacts of coral
bleaching-driven mortality events are especially important. is is because the ecological impacts of bleaching are
Department of Geography, College of Life and Environmental Sciences, University of Exeter, Exeter EX4 4RJ, United
Kingdom. Correspondence and requests for materials should be addressed to C.T.P. (email: c.perry@exeter.ac.uk)
Received: 08 November 2016
Accepted: 07 December 2016
Published: 13 January 2017
OPEN
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near-instantaneous and can be severe. Such events thus have the capacity to also drive very rapid, and potentially
severe, declines in carbonate budgets18 and in resultant reef growth potential. Here we quantify the impacts of the
most recent large-scale sea-surface temperature (SST) anomaly event in the central Indian Ocean, driven by the
strong ENSO-induced warming of 2016. Specically, we quantify the impacts on atoll interior reefs in the south-
ern Maldivian atoll of Gaafu Dhaalu (Fig.1A), a location where other common drivers of major reef degradation
exert minimal inuence (shing pressure on these reefs is relatively low, and immediate point sources of nutri-
ent input absent). Satellite SST anomaly data indicates that strong warming in this region started in late March
2016 (Fig.1B) and persisted at levels above the regional bleaching threshold of ~30.9°C through until mid-May
2016 (Fig.1B–D), the bleaching threshold being dened as the point where SST is 1 °C warmer than the highest
monthly mean temperature (NOAA, 2016). Here we use pre- (January 2016) and post-warming (September
2016) measured rates of both gross carbonate production and bioerosion from shallow fore-reef habitats (2 m
depth) on ve reefs (Mahutigala, Kandahalagala, Kagahlaa, Kodehutigalaa and Kadumaigala) to determine their
net biological carbonate budgets (G, where G = kg CaCO3 m2 yr1). We use these data to address two specic
questions: 1) what impact did the 2016 warming event have on the ecological composition and carbonate budgets
of the shallow fore-reef (2 m depth) habitats, and how consistent were the responses across these reefs?; and 2)
what have been the resultant impacts on reef growth potential?
Results
e extensive coral mortality that occurred between January 2016 and September 2016 at all our study sites can
be linked with a high degree of certainty to the strong SST warming that occurred across the central Indian Ocean
region in mid-2016 (Fig.1B,C). Coral cover (measured as cover of the 3-dimensional surface of the reef) declined
signicantly between January 2016 (across site mean: 25.6 ± 5.8%; range: 20.7 to 34.9%; SupplementaryTable1)
and September 2016 (mean: 6.3 ± 1.9%; range: 5.2 to 9.7%) (p < 0.01; SupplementaryTable2), and at each site
Figure 1. Study location and SST anomaly data. (A) Location of Gaafu Dhaalu atoll (inset map) in the
southern Maldives, and the location of the study sites in the southern part of the atoll (white boxed area). Inset
map (traced using Adobe Illustrator Version CS5) and satellite image from Google Earth Imagery (Map data:
Google, Landsat 2016); (B) Time-series data showing satellite-derived sea-surface temperature (SST) data for
the Maldives (01/09/15 to 31/09/16). Daily data were extracted and replotted from the NOAA Coral Reef Watch
site (http://coralreefwatch.noaa.gov/vs/gauges/maldives.php), accessed 03/10/16; (C,D) Maps showing satellite
derived SST anomaly data for the Indian Ocean; (C) 1st March 2016, (D) 1st May 2016. Plots derived from
NOAA Coral Reef Watch. 2016, updated daily. NOAA Coral Reef Watch Daily Global 5-km Satellite Virtual
Station Time Series Data for Indian Ocean, Mar. 1, 2016 and May 1, 2016. College Park, Maryland, USA: NOAA
Coral Reef Watch. Data set accessed 2016-10-03 at http://coralreefwatch.noaa.gov/vs/index.php.
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the decline exceeded 70% (mean 75.6 ± 3.2%; SupplementaryTable1). Much of this decline was driven by wide-
spread mortality of branching and tabular Acropora spp. (SupplementaryFigure1), the relative abundance of
which has declined signicantly at all sites (p < 0.01; SupplementaryTable2), and with average % abundance
of Acropora spp. declining by almost an order of magnitude across sites (Jan 2016: 14.72 ± 5.04%, Sept 2016:
1.19 ± 0.46). Observational data indicates that extensive Acropora spp. mortality occurred to depths of ~5–6 m
at all sites. Whilst many massive and sub-massive morphology taxa (including Porites spp.) survived the bleach-
ing (albeit with some partial mortality evident), fundamental changes to the ecology of these shallow fore-reef
habitats have occurred. As of early September 2016 most dead Acropora colonies remain in living position
(SupplementaryFigure1), but declines in substrate rugosity (mean decline 10.1 ± 2.0%) are already becoming evi-
dent (Jan 2016 mean 2.6 ± 0.3, range: 2.5 to 2.8; Sept 2016 mean 2.3 ± 0.1%, range: 2.2 to 2.3%; Fig.2B), although
signicant declines have only occurred at Kadumaigala and Kagahlaa (p < 0.05, SupplementaryTable2).
Whilst the structural complexity of these shallow fore-reef habitat sites is thus far relatively unchanged, wide-
spread coral mortality has resulted in a very signicant decline (mean decline 157.5 ± 30.9%) in the net carbonate
budgets at all ve sites (p < 0.05; SupplementaryTable2), shiing from strongly net positive (Jan 2016, mean
5.92 ± 2.2 G, range: 3.6 to 8.6G; Fig.2C) to strongly net negative (Sept 2016 mean 2.96 ± 1.06 G, range: 1.9
to 4.7G; Fig.2C). ese substantial carbonate budget declines reect two interacting variables. Firstly, meas-
ured rates of coral carbonate production have declined signicantly at all sites (p < 0.05, SupplementaryTable2)
(Jan 2016 mean 8.43 ± 2.08 G, range: 6.4 to 11.3 G; Sept 2016 mean 1.83 ± 0.47 G, range: 1.5 to 2.6 G), an average
decline of 78.0 ± 3.1% (Fig.2D). At the same time, measured rates of parrotsh bioerosion have increased sub-
stantially (+ 139.5 ± 59.7%; Fig.2E). ese increases are signicant at all sites (p < 0.05, SupplementaryTable2;
Jan 2016 mean 1.76 ± 0.24 G, range: 1.4 to 2.1 G; Sept 2016 mean 4.15 ± 0.99 G, range: 3.4 to 5.8 G).
Our data also indicate marked post-warming shis in the proportional contributions made to coral carbonate
production rates by dierent groups of coral taxa. Pre-coral bleaching (Jan 2016) coral carbonate production
was dominated by branching, corymbose and tabular species of Acropora (mainly A. cytherea, A. digitifera, A.
muricata, A. lamarcki), which collectively accounted for between 52 ± 15.3% and 65.2 ± 13.7% of coral carbonate
production (Fig.3A). Post-bleaching, the contribution of Acropora spp. declined on average by 70.9 ± 5.6%,
and was signicant at all sites (p < 0.01, SupplementaryTable2) (Jan 2016 mean 57.2 ± 13.7%, range: 30.0 to
Figure 2. Comparisons between key ecological, structural and carbonate budget metrics across reef sites
in Gaafu Dhaalu atoll, Maldives between January 2016 and September 2016. Box (median and 50% quartile)
and whisker (95% quantile) plots showing dierences in; (A) coral cover, (B) rugosity, (C) net budget, (D) coral
production, and (E) parrotsh erosion. MAH – Mahutigala, KAN – Kandahalagala, KOD – Kodehutigalaa,
KAD – Kadumaigala, KAF – Kagahlaa.
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81.0%; Sept 2016 mean 16.6 ± 12.2%, range: 0–32.2%). In contrast, proportional contributions by other more
resilient massive and sub-massive taxa (Porites spp., Favia spp.) more than doubled (Jan 2016 mean 19.7 ± 16.8%,
range: 1.0 to 54.0%; Sept 2016 mean 47.6 ± 27.8%, range: 0–95.0%), and was signicant at all sites (p < 0.05,
SupplementaryTable1), except Mahutigala (p = 0.42) and Kodehutigalaa (p = 0.13) (Fig.3B). Proportional con-
tributions to coral G by other non-Acropora branching (mainly Pocillopora spp.) taxa also increased (Jan 2016,
mean 4.5 ± 0.6%, range: 0 to 28.0%; Sept 2016, mean 12.2 ± 14.1%, range: 0–48.0%) (Fig.3C), as did contribu-
tions from encrusting taxa (Jan 2016, mean 5.6 ± 4.2%, range: 1.0 to 16.0%; Sept 2016, mean 10.0 ± 12.9%, range:
1.0–61.0%) (Fig.3D), but overall contributions to coral G by these taxa remain relatively small, and measured
increases are non-signicant (SupplementaryTable2).
Based on the carbonate budgets measured around the ve study sites in January 2016, calculated rates of ver-
tical reef growth within the shallow fore-reef habitats averaged 4.2 mm yr1 (range: 1.4 to 9.1 mm yr1; Fig.4A).
Our data show, however, that following the warming event of mid-2016, very signicant reductions in reef growth
Figure 3. Proportional contributions to coral carbonate production rates (kg CaCO3 m2 yr1) across
reef sites in Gaafu Dhaalu atoll, Maldives between January 2016 and September 2016. Box (median and
50% quantile) and whisker (95% quartile) plots showing dierences in; (A) Acropora (all branched and tabular
species), (B) Massive and sub-massive taxa (including Porites), (C) Non-Acropora branched taxa, including
Pocillopora spp., and (D) Encrusting taxa. MAH – Mahutigala, KAN – Kandahalagala, KOD – Kodehutigalaa,
KAD – Kadumaigala, KAF – Kagahlaa.
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potential occurred across all sites (p < 0.001; SupplementaryTable2), with mean rates at all sites now net nega-
tive i.e., net erosional (average 0.43 mm yr1, range: 0.15 to 1.88 mm yr1) (Fig.4A). is points to not only
a rapid loss of growth potential, but also the likelihood that the structural complexity of these reefs, which in the
immediate aermath of the event has only marginally declined (see Fig.2B), will progressively be diminished.
Discussion
e major and prolonged period of elevated SST that caused coral bleaching in the Maldives in 2016 resulted from
the recent strong El Niño that originally begun forming in the central Pacic region in June 201419. is mid-
2014 warming event ultimately is thought to have initiated the much larger scale warming that spread across the
world’s oceans, and caused widespread coral bleaching, through late 2015 and mid-2016. Indeed, such has been
the scale of resultant bleaching that the event has been formally designated as the “ird global coral bleaching
event”20. Peak temperatures in the Maldives occurred between March and May 2016 (Fig.1), and our data from
the southern Maldives suggests that this resulted in very signicant ecological changes. Coral cover declined by
an average 75% and the overall impact of this on the carbonate budgets of the shallow fore-reef habitats on these
reefs has been profound. Carbonate budgets have reduced by an average of 157%, and all sites now have net neg-
ative budgets. Critically, in terms of the wider geographic relevance of these ndings, we note that our coral cover
decline data are consistent with ndings from rapid assessments undertaken during mid-2016 across the wider
Maldives, which also indicated between 60 and 90% coral mortality on shallow reefs21. is suggests that geo-
graphically widespread reductions in reef carbonate budgets and thus in reef growth potential, of the magnitude
we report here, are highly likely to have occurred.
A key driver of these declines has been the mass mortality of branched and tabular Acropora spp., the cover
of which has declined by an average 91%. is not only has major budget implications, but also major ecolog-
ical implications because of the high habitat complexity and diversity such corals provide within these shal-
low fore-reef habitats22. Interestingly, we note that mortality of another important shallow water fore-reef taxa
(Pocillopora), which is oen badly impacted by bleaching6,22, suered variable levels of mortality. Some colonies
clearly survived this event, whilst others show evidence of only partial mortality, but the proportional contribu-
tion of this taxa to coral carbonate production increased slightly at most sites (Fig.3). However, despite these
taxa specic responses, rates of overall coral carbonate production declined markedly (by more than 300%),
from a pre-event average of 8.4 G to a post-bleaching rate of 1.8 G. We also note that rates of bioerosion by par-
rotsh increased signicantly across sites (+ 139%) in the pre- and post-bleaching interval. is, we hypothesise,
must reect an increase in the exploitation of newly available, lamentous algal covered dead coral substrate by
Figure 4. Coral reef accretion rates. (A) Changes in rates of reef accretion (mean mm yr1 + / 1 standard
deviation) between atoll interior reef sites in Gaafu Dhaalu atoll, Maldives between January 2016 and September
2016. All sites are at 2 m depth on the fore-reef slopes. Comparative data from Acropora-dominated sites around
the Chagos Archipelago, central Indian Ocean (all 9 m depth) (in 23), and from the Caribbean (in 12), and
recalculated based on a conservative assumption of 20% of framework being removed by physical processes.
(B–D) e linear regression and 95% condence intervals for the relationships between: (B) Coral cover and net
carbonate production (G); (C) Coral cover and reef accretion potential (mm yr1); (D) Acropora spp. cover and
reef accretion potential (mm yr1); and (E) Acropora spp. cover pre-bleaching (Jan 2016) and the magnitude of
the subsequent reduction in the net carbonate budget (G).
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parrotsh at these sites. us the carbonate budgets of these reefs have been strongly impacted by both a rapid
decline in coral carbonate production and by an increase in substrate bioerosion rates.
e net eect of these budget changes has been to drive very substantial declines in shallow fore-reef growth
potential, which have declined from an average 4.2 mm yr1 to 0.4 mm yr 1. We note that pre-bleaching rates
are similar to those measured at Acropora-dominated sites around the remote Chagos Archipelago (central Indian
Ocean) in mid-2015 (mean: 2.9 mm yr1, range: 0.9 to 6.3 mm yr1; Fig.4A)23, albeit with the Chagos data being
derived from slightly deeper water (9 m depth) fore-reef habitats, and with communities their dominated by tab-
ular, rather than branched Acropora species. Notwithstanding these inter-site dierences, the similarities in site
average rates are clear, and suggest that potential growth rates in the order of ~3 to 4 mm yr1 might be considered
realistic for many “healthy”, shallow-water (<10 m depth) Acropora–dominated reefs in the central Indian Ocean,
a rate supported by metadata analyses of shallow water Indian Ocean accretion rates24.
A key implication of the major post-bleaching changes we report is that, at least at present, the shallow
fore-reef habitats on these Maldivian reefs have shied from states dened by strong growth potential to a situ-
ation dominated by net framework erosion and breakdown. is has potentially very signicant implications in
terms of the capacity of these reefs to continue acting as breakwater structures for the reef islands that they sup-
port. In the short-term (the next few years) the breakdown and loss of shallow fore-reef surface structural com-
plexity will likely reduce the eectiveness of these reefs to reduce wave energy propagation across the reef ats25,26.
is is perhaps an especially important issue in these settings because of the role that the surrounding reefs play
in modulating wave energy regimes around islands27 and thus the short-term morphodynamic responses of the
islands they underpin28,29. However, in the medium-term (decades), any continuation of the current states of low
(negative) budget states will also progressively impinge upon the capacity of the reefs to match any increases in
sea-level. Given the rapid increases in wave energy propagation that follow the increased submergence of reefs3,30
the impacts for island stability in the region may thus be signicant.
As at other sites where relationships between coral cover and carbonate budgets have been examined14,23, we
note a strong positive correlation between both net carbonate production rates and reef accretion (Fig.4B,C). In
our datasets the average coral cover threshold for the maintenance of positive carbonate budgets is around 12%
(Fig.4B), and for net positive accretion potential around 8% (Fig.4C). ese thresholds are broadly comparable
to those measured from multiple sites around Chagos further south in the Indian Ocean23, and thus appear to be
consistent not only across the Indian Ocean region, but are also comparable to those derived for the Caribbean
region12. It is important to emphasise, however, that these % cover data reect that measured as a function of the
true 3 dimensional cover of the reefs, and thus equate to higher % cover thresholds where data is collected using
linear point or photo/video quadrat type methods.
In the present study we also note that the abundance of fast growing Acropora species is an especially impor-
tant control on budgets and accretion potential. Acropora cover and accretion rate are strongly correlated
(Fig.4D), and there is a strong positive relationship between the % cover of Acropora on these Maldivian reefs,
and the magnitude of decline in their carbonate budgets post-bleaching (Fig.4E). In other words, reefs with the
highest initial cover of Acropora experienced the largest magnitudes of decline in carbonate budgets and thus in
accretion potential. is has important implications that relate to on-going discussions around ecological “win-
ners” and “losers” under severe climate stress22,31–33. As we move into a period where major SST anomaly events
are predicted to increase in frequency and intensity34, the extent to which reefs are colonised by either “winner”
or “loser” taxa will have direct relevance for predicting the long-term capacity of reefs to sustain periods of strong
positive budgets and growth potential. For example, communities dominated by branched, corymbose and/or
tabular Acropora species (depending on setting and energy exposure) typically dominate in shallow-water Indian
Ocean reef habitats. ese same species also drive high rates of coral framework production and result in high
potential accretion rates (even factoring for substantial amounts of annual framework export due to physical
breakage and export; ref. 23). However, it also well established that these same species are amongst the most
vulnerable to SST-driven stress32, and this has two key implications. Firstly, it is likely that it is the most “pristine
reefs (i.e., those with high branched/tabular Acropora cover) which will be the most susceptible to future bleach-
ing events, and thus the most susceptible to future rapid declines in carbonate production and reef accretion.
Secondly, it must also be assumed that if bleaching events become more severe and frequent then the capacity
(and timescales necessary for) the recovery of Acropora communities will progressively diminish, potentially to
the point that even the most remote and protected reefs will struggle to recover from disturbances in terms of
their budgets and growth potential.
An important issue now will therefore be whether and when these reefs may recover, both ecologically and
in terms of their carbonate budgets. Evidence from other sites in the region which were strongly impacted by the
1997/98 bleaching event suggest some grounds for optimism, especially where the reefs are relatively remote or
isolated from high levels of direct human disturbance. For example, the reefs around the remote Chagos archi-
pelago (to the south of the Maldives) showed an impressive capacity for recovery from the 1997/98 mass coral
bleaching event. Coral cover on most reefs around Chagos had recovered to pre-bleaching levels by around
201035 i.e., within about 12 years, and those reefs (as of May 2015) were characterised by strongly positive car-
bonate budgets and high reef accretion potential23. Many reefs in the Maldives were also severely impacted by
the 1997/98 event22,36, but subsequent monitoring across a number of atolls showed that whilst encrusting and
massive taxa dominated the early post-bleaching communities, it is now the case, as at our study sites, that these
had been replaced by species of Acropora and Pocillopora. e timescales over which this recovery occurred var-
ied between sites, but recovery of coral cover and communities to pre-1998 levels/states was generally evident by
around 2009 to 2012, depending on location22,36,37. Given these past recovery trajectories and the fact that local
populations of key reef building corals (mainly below 5–6 m water depth) have survived this event the potential
for the reefs to recovery must, in theory at least, be good. Indeed, based on past recovery trajectories a best esti-
mate might be that aer a period of reef framework breakdown, that recovery might be feasible with ~10 years.
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However, there will be three key factors that inuence future recovery. e rst, as outlined above, will be
the frequency and magnitude of future SST warming events, and which are predicted to increase in the near
future15,16,38. e second will be the occurrence of crown-of-thorns starsh (COTS) outbreaks, which impacted
some reefs in the central region of the Maldives in 201522, although very few COTS were observed at our study
sites in either sampling period in 2016. ird, will be the impacts of increasing direct anthropogenic stressors
across the region. ese include the impacts associated with further rapid expansion of resorts and any associated
increases in sedimentation or eutrophication stress, and shing pressures, and which may impede natural trajec-
tories of coral community recover from, or resilience to, climate-driven stressors. For the Maldives in particular,
any factors that result in long-term or more frequent suppression of shallow fore-reef carbonate budgets and reef
growth potential may have profound implications because of the close spatial proximity between the reefs and the
adjacent low-lying reef islands that dene the archipelago. In particular, any reduction in reef structure and loss of
growth potential will have likely implications for wave energy propagation across the lagoons, and for the capacity
of these reefs to track IPCC projections of future sea level rise, both factors that will threaten shoreline stability.
Methods
Surveys were conducted during January and September 2016 on ve atoll interior reefs in the southern Maldivian
atoll of Gaafu Dhaalu; Mahutigala, Kandahalagala, Kagahlaa, Kodehutigalaa and Kadumaigala. All surveys were
conducted along the fore-reef slope 2 m depth contour on the south-western margins of each reef. At each site
5 replicate survey lines (10 m long) were established running parallel to the reef crest, with a spacing of 5 m
between transects. To quantify substrate composition, reef rugosity, and gross carbonate production and erosion,
and thus to determine net carbonate budgets (G, where G = kg CaCO3 m2 yr1) we used a previously modied
version of the ReefBudget methodology13, as described in Perry et al.23. ese metrics are calculated as a function
of the 3-dimensional surface of the reefs. For the benthic assessments we measured the distance within each
linear 1 m covered by each category of benthic cover beneath a 10 m guide line using a separate exible tape.
All overhangs, vertical surfaces and horizontal surfaces below the line were surveyed, with the following groups
recorded: scleractinian corals to the genera and morphological level e.g., Acropora branching, Porites massive etc.
(see SupplementaryTable3 for a summary of coral cover data); crustose coralline algae (CCA) including CCA
below macroalgal or so coral cover; turf algae; eshy macroalgae; non encrusting coralline algae (e.g., Halimeda
spp., articulated coralline algae); sediment; bare substrate (e.g., limestone pavement); sediment; rubble; and other
benthic organisms. Substrate rugosity was calculated as total reef surface divided by linear distance (a completely
at surface would therefore have a rugosity of ref. 1).
Following the modied version of the ReefBudget approach described by Perry et al.23, we then used the mor-
phology and size of individual coral colonies in combination with genera specic skeletal density (g cm3) and linear
growth rates (cm year1) across each transect to estimates carbonate production rates in kg CaCO3 m2 year1.
In most cases we used mean regional growth rates and densities for each coral genera (see SupplementaryTable4
for a summary of rates used). ese data were then combined with geometric transformations based on colony
morphology to give a growth rate for each colony for the area under the transect line (taking a transect line width
of 1 cm): massive colonies were assumed to be hemispherical in cross-section; encrusting, foliose and plating
colonies, as well as colonies of crustose coralline algae (CCA) grow primarily from their margins such that the
rate of radial expansion decreases rapidly with increasing colony size39. To accommodate for this we thus make an
assumption that the growth rate across the internal (older) portion of such colonies is much lower – factored here
at a rate of 10% the rate of marginal expansion to account for secondary skeletal inlling; for branching colonies,
the proportion of the colony area of growing branch tips was assumed to be growing at published rates, and the
remainder of the colony at 10% of these rates (see ref. 23).
To calculate the production for a single transect over a year, the following equation was then used:
=++…+
=
CP CP CP CP
(1)
j
i
n
n
1
12
where CPj is the total carbonate production of both corals and crustose coralline algae for transect j in kg CaCO3
yr1. To estimate the production rate of the reef, we then used the following equation:
=GprodCP
(2)
jj
l
10000
where Gprodj is the carbonate production rate of both corals and crustose coralline algae for transect j in kg
CaCO3 m2 year1, and l is the transect length in centimetres.
To quantify rates of biological substrate erosion we used three approaches for dierent elements of the bioe-
roding communities. Rates of macro- and microendolithic bioerosion were based on published rates per unit area
derived from experimental studies in the Indo-Pacic region, and these were applied to all available dead car-
bonate substrate available to bioeroding sponges, including that covered by macroalgae or algal turf and live coral
cover and so corals23. To quantify echinoid bioerosion rates, a census of urchin abundance, size and species com-
position within 10 × 2 m belt transects was conducted along each benthic transect line. Typically these data would
then be applied to published relationships between urchin species, test size and erosion rate, but no urchins were
encountered during our surveys and they are thus assumed to exert little or no inuence on carbonate erosion
rates. Finally, to calculate bioerosion by parrotsh the species-size-life phase abundances of bioeroding parrotsh
at each site were calculated at each site based on eight 30 m × 4 m belt transects, with all surveys completed by the
same experienced observer (K.M.M.). Biomass of individual sh was then calculated using estimated length data
and length-weight relationships and multiplied by abundance of the species or family of the sh (see ref. 23). To
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Scientific RepoRts | 7:40581 | DOI: 10.1038/srep40581
calculate parrotsh bioerosion rates by each individual sh we then used a model based on total length and life
phase to predict the bite rates (bites hr1) based on published data for that species or for similar sized species with
the same feeding functional group. We then used the following equation to calculate species specic erosion rates
for the median value within each size class:
.=...∗
−− vs br dBioerosion rate(kg indyr)365 (3)
prop
11
where v is bite volume (cm3), sprop is the proportion of bites leaving scars, br is bite rate (bites day1) and d is sub-
stratum density (kg cm3), here taken to be 1.49 aer Morgan & Kench (2012).
To assess changes in the accretion potential of reefs (mm yr1) pre- and post-bleaching we converted our net
production rate estimates to potential accretion rates (mm yr1), using an approach previously applied to other
Indian Ocean and Caribbean reefs12,23. Specically, we estimated the maximum accretion potential of each reef
as a function of the net carbonate production rate of the site (calculated as gross production less gross erosion
rate) and assumed that a proportion of the bioeroded framework (that is converted to sediment) is also rein-
corporated back into the accumulating reef structure. is proportion is calculated as the sum of 50% of the
parrotsh-derived sediment (as a highly mobile bioeroder which defecates randomly over the reef), as well as any
sediment produced by urchins and by macroborer erosion. To keep our estimates conservative we worked on the
assumption that only 50% of this bioerosional sediment yield is actually incorporated back into the reef (based on
data in ref. 40), and excluded any sediment generation by other benthic sediment producers. Finally, we made an
allowance for variations in the porosity of the accumulating reef framework as follows: 30% for head and massive
coral dominated assemblages, 70% for branched and tabular dominated assemblages, and 50% for mixed coral
assemblages (based on data in ref. 41). A loss factor to account for natural framework removal through physical
processes was also included based on framework production and removal rates calculated from atoll interior reefs
in the Maldives42,43, such that we assumed that 20% of the annual framework produced was removed from these
relatively sheltered reef settings. T-tests were then used to test for signicance of dierence between net and gross
production, erosion and accretion rates pre- and post-bleaching at each site.
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Acknowledgements
Research was supported through a Leverhulme Research Fellowship (RF-2015-152) to CTP. We thank the LaMer
Small Island Research Centre, Faaresmathooda for assistance with eld work and logistics. e reviewers and
Editor of Scientic Reports are thanked for their useful and supportive comments on an earlier dra of the text.
Author Contributions
C.T.P. was responsible for the project design, data collection, interpretation and writing. K.M.M. was responsible
for data collection, interpretation and writing.
Additional Information
Supplementary information accompanies this paper at http://www.nature.com/srep
Competing nancial interests: e authors declare no competing nancial interests.
How to cite this article: Perry, C. T. and Morgan, K. M. Bleaching drives collapse in reef carbonate budgets and
reef growth potential on southern Maldives reefs. Sci. Rep. 7, 40581; doi: 10.1038/srep40581 (2017).
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Supplementary resource (1)

... Bleached corals are physiologically damaged and nutritionally comprised, and prolonged bleaching often leads to extensive coral mortality and reductions in live coral cover or abundance across large regions (e.g., Baird and Marshall, 2002;Hughes et al., 2018b). Impacts from coral bleaching events can be long lasting (e.g., Burt et al., 2011) and may alter species composition (Pratchett et al., 2011;Hughes et al., 2018b), recruitment (Burt and Bauman, 2020) and calcification rates (Perry and Morgan, 2017) resulting in the homogenization of coral assemblages (e.g., Bento et al., 2016) and loss of key ecological functions (Alvarez-Filip et al., 2013). Notably, climate induced bleaching events are already altering the structure and function of coral reefs (Hughes et al., 2018b) thereby jeopardizing critical ecosystem goods and services that reefs provide (Pratchett et al., 2014;Harris et al., 2018). ...
... Maintaining structural complexity is also a key factoring determining reef recovery following a disturbance (Graham et al., 2015). However, as corals die and their skeletons erode, the structural complexity of reef habitats decline (Alvarez-Filip et al., 2009), which can have important consequences for the biodiversity (Graham and Nash, 2013) productivity (Darling et al., 2017) and associated ecosystem services of coral reefs (Alvarez-Filip et al., 2013;Perry and Morgan, 2017). Moreover, shifts in the composition and dominance of coral assemblages can result in rapid losses in coral community calcification rates and reef rugosity (Alvarez-Filip et al., 2013). ...
... Reduction in and/or the loss of reef structural complexity following severe bleaching was initially reported to occur over timescales of 3-5 years (e.g., Graham et al., 2007;Alvarez-Filip et al., 2011b), but is now occurring more rapidly (Eakin et al., 2019). Recent bleaching studies report that the structure of both individual coral skeletons (Leggat et al., 2019) and entire reefs (Couch et al., 2017;Magel et al., 2019) are rapidly eroding over short timeframes (i.e., months to <1 year) causing abrupt shifts on reefs toward net erosion (Perry and Morgan, 2017). For example, Couch et al. (2017) reported a 99% reduction in structural complexity on bleached reefs in the Northwestern Hawaiian Islands <1 year after bleaching. ...
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Increasing incidence of severe coral bleaching events caused by climate change is contributing to extensive coral losses, shifts in species composition and widespread declines in the physical structure of coral reef ecosystems. With these ongoing changes to coral communities and the concomitant flattening of reef structural complexity, understanding the links between coral composition and structural complexity in maintaining ecosystem functions and processes is of critical importance. Here, we document the impacts of the 2016 global-scale coral bleaching event on seven coral reefs in Singapore; a heavily degraded, turbid reef system. Using a combination of field-based surveys, we examined changes in coral cover, composition and structural complexity before, during and after the 2016 bleaching event. We also quantified differential bleachingresponses and mortality among coral taxa and growth forms using a bleaching response index. Elevated SSTs induced moderate to severe coral bleaching across reefs in Singapore in July 2016, but low overall coral mortality (~12% of colonies). However, we observed high bleaching prevalence and post-bleaching mortality of the three most abundant coral genera (Merulina, Pachyseris and Pectinia), all generalists, declined significantly across reefs between March and November 2016. Four months post-bleaching (November 2016), small-scale structural complexity declined across all Singaporean reefs and no moderately complex reefs remained. Importantly, reductions in structural complexity occurred across reefs with a large range of live coral cover (19–62%) and was linked to the loss of dominant coral genera with low-profile foliose-laminar growth forms which resulted in flatter, less structurally complex reefs. And while generalist coral taxa remain highly competitive within Singapore’s reef environment, they may not have the capacity to maintain structural complexity or ensure the persistence of other reef functions, even within communities with high coral cover. The widespread loss of structurally complexity on Singapore’s degraded coral reefs may further impair ecosystem functioning, potentially compromising the long-term stability of reef biodiversity and productivity.
... These biological processes, which largely reflect the abundance of carbonate producing and eroding taxa, are inherently susceptible to ecological disturbances. Indeed, both gradual and steep declines in net carbonate budgets and resultant reef accretion potential have been reported globally due to chronic stress (Perry et al. 2018;Molina-Hern andez et al. 2020) and following coral bleaching events (Perry and Morgan 2017;Manzello et al. 2018;Lange and Perry 2019). ...
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Benthos is an encompassing term used to classify organisms found on, in, or in close contact to the bottom region of bodies of water. Benthic ecology, on the other hand, is the study of the relationships of benthos and their unique and diverse environment. This ecosystem is an abundant and valuable source of ecosystem goods and services that help sustain the ecosystem's healthy balance. The benthic infauna is the assemblage of organisms living within the seafloor, while the benthic epifauna is those living on or attached to the seafloor. These organisms are continuously threatened by various anthropogenic activities like bottom trawl fishing, mine tailing pollution, land reclamation and conversion, destruction of coralline and macroalgal communities, overexploitation, climate change, and many more. Due to their susceptibility and sensitivity, these organisms are often regarded as excellent bioindicators and biomonitors of environmental changes. This chapter attempts to introduce the complex world of the benthos and their vital role in the complex marine ecosystem. This chapter also presents current studies to understand further the ecology of benthos and the recent advances in research and technology.
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Coral cover on Caribbean reefs has declined rapidly since the early 1980's. Diseases have been a major driver, decimating communities of framework building Acropora and Orbicella coral species, and reportedly leading to the emergence of novel coral assemblages often dominated by domed and plating species of the genera Agaricia, Porites and Siderastrea. These corals were not historically important Caribbean framework builders, and typically have much smaller stature and lower calcification rates, fuelling concerns over reef carbonate production and growth potential. Using data from 75 reefs from across the Caribbean we quantify: (i) the magnitude of non-framework building coral dominance throughout the region and (ii) the contribution of these corals to contemporary carbonate production. Our data show that live coral cover averages 18.2% across our sites and coral carbonate production 4.1 kg CaCO3 m−2 yr−1. However, non-framework building coral species dominate and are major carbonate producers at a high proportion of sites; they are more abundant than Acropora and Orbicella at 73% of sites; contribute an average 68% of the carbonate produced; and produce more than half the carbonate at 79% of sites. Coral cover and carbonate production rate are strongly correlated but, as relative abundance of non-framework building corals increases, average carbonate production rates decline. Consequently, the use of coral cover as a predictor of carbonate budget status, without species level production rate data, needs to be treated with caution. Our findings provide compelling evidence for the Caribbean-wide dominance of non-framework building coral taxa, and that these species are now major regional carbonate producers. However, because these species typically have lower calcification rates, continued transitions to states dominated by non-framework building coral species will further reduce carbonate production rates below ‘predecline’ levels, resulting in shifts towards negative carbonate budget states and reducing reef growth potential.