PACIFIC MARINE CLIMATE CHANGE REPORT CARD
Science Review 2018: 132-158
Effects of Climate Change on Corals Relevant to the Pacific
Leo X.C. Dutra, CSIRO, Australia, and The University of the South Pacific (USP), Fiji; Michael D.E. Haywood,
CSIRO, Australia; Shubha Shalini Singh, USP, Fiji; Susanna Piovano, USP, Fiji; Marta Ferreira, USP, Fiji,
University of Porto, Portugal; Johanna E. Johnson, C2O Pacific, Cairns and Vanuatu, James Cook University,
Australia; Joeli Veitayaki, USP, Fiji; Stuart Kininmonth, USP, Fiji, University of Oslo, Norway; Cherrie W. Morris,
In the Pacific island region, anthropogenic-induced ocean warming is impacting coral reefs through
thermal coral bleaching (Adjeroud et al., 2009; Cumming et al., 2000; Davies et al., 1997; Kleypas et al.,
2015; Lovell et al., 2004; Mangubhai, 2016; Obura and Mangubhai, 2011; Rotmann, 2001) and by
reducing coral calcification rates (high confidence) (De'ath et al., 2009; Nurse et al., 2014). Ocean
acidification is also affecting calcification rates (low confidence) (Barros and Field, 2014; IPCC, 2014;
Johnson et al., 2016b), while tropical cyclones (TC) are becoming more intense (low confidence) (Elsner
et al., 2008; Nott and Walsh, 2015), causing widespread coral damage (Johnson et al., 2016a;
Mangubhai, 2016). Accelerating climate change is currently the strongest driver affecting coral dynamics
(Aronson and Precht, 2016) exacerbating other non-climate pressures, such as pollution from land-
based activities and coastal development, eutrophication, overfishing, crown-of-thorns starfish
outbreaks, direct physical impacts (e.g. tourism-related impacts, trampling and anchor damage), and
coral diseases, which together are causing changes in coral communities not previously recorded (high
confidence) (Brown et al., 2017a; Hoffmann, 2002; Johnson, 2017; Maata and Singh, 2008; Maynard et
al., 2015b; Mayor, 1924; Morris et al., 2008; Morrison, 1999; Movono et al., 2018; Singh et al., 2009;
Sykes and Morris, 2009; Zann, 1994). Predicted increase in atmospheric carbon dioxide (CO2)
concentrations will continue to increase ocean temperatures and acidity (high confidence) (Gattuso et
al., 2015; Newman et al., 2018) as well as increase the intensity of TC (low confidence) (Elsner et al.,
Climate pressures will exacerbate non-climate pressures (high confidence) (Aronson and Precht, 2016;
Sheridan et al., 2014; Wild et al., 2011), thus increasing the frequency and severity of coral bleaching,
disease incidence, and mortality (high confidence) (IPCC, 2014; Langlais et al., 2017; Reisinger et al.,
2014; Ruiz-Moreno et al., 2012; Sheridan et al., 2014). Declines in coral reef habitat will cause negative
social-economic and ecological effects (high confidence) (IPCC, 2014; Morrison, 1999). Traditional
Coral reefs are the dominant coastal habitat in the
tropical Pacific, representing more than 25% of reefs
globally – nearly 66,000 km2 (Wilkinson, 2004;
Wilkinson, 2008). Many Pacific Islands and Territories
(PICTs), including French Polynesia, Kiribati and
Palau have at least double the reef area than land area
(SPC data). Corals are the fundamental reef
ecosystem engineers because they construct the
framework that supports over 600 species of calcifying
corals, 4,000 species of fish, as well as a high diversity
of invertebrates, macroalgae and marine megafauna
totalling approximately 830,000 species of multi-
cellular animals and plants worldwide, or 32 percent of
all named marine species (Allen, 2008; Fisher et al.,
2015; Wilkinson, 2008). Coral reefs form a mosaic of
habitats along with mangroves and seagrasses, which
sustains a diversity of organisms, greater fish
productivity, a variety of industries (e.g. fisheries and
tourism), also more effectively protecting the coastline
against erosion (Veitayaki et al., 2017) when these
habitats are adequately inter-connected (Guannel et
al., 2016; Moberg and Folke, 1999; Zann, 1994).
The distribution of corals is limited by physical and
biological factors. Physical factors include water
temperature, pH, light, turbidity/sedimentation, salinity,
and water depth (Aronson and Precht, 2016), while
biological factors include predation (e.g. the Crown-of-
Thorns-Starfish (COTS) and Drupella spp.; both well-
known coralivores (Berthe et al., 2016; Bruckner et al.,
2017)), intra and inter-specific competition (e.g.
competition between corals and algae, and between
different coral species for space), reproductive and
regenerative capacity, and the ability to cope with
pollutants, nutrients and sediments (Guinotte et al.,
2003; Kleypas et al., 1999; Rogers, 1990). This means
that alterations to any of these factors can seriously
threaten the existence of corals.
Growing human populations in most PICTs, along with
changes in lifestyle put enormous pressure on coral
reefs and connected habitats. In addition to an
increasing demand for food, people also want
increased standards of living and, to do so, they build
infrastructure such as piped water and sewerage
facilities, but also better housing, roads and a range of
foodstuffs and consumer goods. Despite the beneficial
effects on corals of sewerage systems and better
housing, increased land clearing and alteration of
riverine and littoral vegetation for construction, waste
disposal, and mining, negatively affect water quality,
coral reefs and connected habitats, leading to the
decline of corals and associated organisms (Morrison,
1999; Remling and Veitayaki, 2016; Veitayaki and
Holland, in press; Zann, 1994).
Since the 1970s there has been a growing body of
evidence showing the deterioration of coral
populations around the world and in the Pacific Ocean
due to both climatic and non-climatic drivers.
Widespread temperature-related bleaching events
have been recorded world-wide and in the Pacific
Ocean (Hughes et al., 2017b; Lough, 2012;
Mangubhai, 2016; Obura and Mangubhai, 2011).
Cyclones are becoming stronger (low confidence
(Elsner et al., 2008)), oceans are becoming more
acidified (medium confidence) (Barros and Field, 2014;
Bates et al., 2014; Dore et al., 2009; IPCC, 2014;
Johnson et al., 2016b), and sea level is gradually rising
(Dean and Houston, 2013; Jevrejeva et al., 2009).
These events cause negative effects on coral reefs
and adjacent ecosystems upon which corals depend
(e.g. seagrasses and mangroves) (Albert et al., 2017;
Guannel et al., 2016; Hassenruck et al., 2015), and are
expected to increase in the future due to increased
carbon dioxide (CO2) emissions and consequent
global warming (high confidence) (Gattuso et al.,
systems of resource governance in the Pacific are often associated with healthy reefs (Hoffmann, 2002)
and coastal communities depend heavily on coral reefs for food, income and livelihoods. Therefore,
management and adaptation options must consider and build on the regional diversity of governance
systems to enable community-based initiatives and cross-sectoral cooperation, taking into account
traditional knowledge that can inform sustainable solutions to problems (Aswani et al., 2017; Morrison,
1999; Remling and Veitayaki, 2016; Veitayaki, 2014) and allow the involvement of a broader section of
the community. Such an inclusive approach will offer enhanced opportunities to develop and implement
measures to reduce non-climate pressures and develop early warning systems, to identify potential
refuges for coral reef communities (Karnauskas and Cohen, 2012) and also to test coral farming,
assisted colonisation and shading or other climate mitigation techniques (Barros and Field, 2014; Coelho
et al., 2017).
The most important non-climatic drivers affecting
Pacific corals include overfishing (Hughes et al., 2007;
Jackson et al., 2001; Rasher et al., 2012), human-
induced pollution, shore modification (Bell et al., 2011),
nutrification (Wiedenmann et al., 2013), increase in
sedimentation and turbidity (Dadhich and Nadaoka,
2012; De'ath and Fabricius, 2010; Haywood et al.,
2016; Sheridan et al., 2014), and increase in the
incidence of COTS outbreaks (Berthe et al., 2016;
Hughes et al., 2017a; Uthicke et al., 2015). Other
impacts include recreational and tourism-related
impacts, such as the construction of resorts, water
catchment dams, roads, and even man-made islands,
along with trampling, tourism handling, and anchor
damage, which have brought substantial ecological
shifts in coral communities over time, but also
unprecedented social, economic and cultural changes
within Pacific communities (Ayala, 2005; Engelhardt,
2005:176; Kininmonth et al., 2014; Morris et al., 2008;
Morris, 2009; Movono et al., 2018; Sykes and Morris,
2009). Climatic drivers can work synergistically with
other non-climate pressures such as nutrient
enrichment to further impact corals (Chazottes et al.,
2017; Wiedenmann et al., 2013; Wooldridge et al.,
2017). Coral growth depends on the intrinsic immune
response and repair mechanisms in reef-building
corals (D'Angelo et al., 2012). Increase in temperature
and nutrient enrichment, along with ocean acidification
can affect corals’ immune response, thus increasing
the susceptibility of corals to diseases and affecting
coral survivability and growth (Aronson and Precht,
2016; Hassenruck et al., 2015; Maynard et al., 2015a).
Moreover, synergistic effects between ocean
acidification (low aragonite saturation) and nutrient
loading are far more effective at driving coral
macrobioerosion than ocean acidification alone
(DeCarlo et al., 2015). This has important
consequences for corals and reefs because it
diminishes coral growth rates, thus compromising reef
Coral responses to climatic and non-climatic pressures
are similar and include bleaching (expulsion of
zooxanthellae that live in their tissue (Aronson and
Precht, 2016; Chazottes et al., 2017)), reproductive
and growth impairments (Albright and Mason, 2013;
Fabricius et al., 2017; Sheridan et al., 2014), and coral-
algal phase shifts (from coral-dominated to algae-
dominated reefs) (Done, 1992; Hughes et al., 2007).
The end result of sustained stresses on corals is a
simplification of coral community structure and
reductions in live coral cover (Bruno and Selig, 2007)
and species coral trait diversity (Darling et al., 2013),
which have resulted in habitat losses (Aronson and
Precht, 2016; Brown et al., 2017a). The rate and extent
of coral loss in the Pacific is greater than expected
(Bruno and Selig, 2007) and have caused negative
social-ecological consequences, such as decreased
fish catches and coastal protection, and biodiversity
loss (Adam et al., 2014). These consequences are
very relevant to PICT due to the reliance of the majority
of their population on coastal and marine resources for
cultural reasons, income and food (Morrison, 1999).
It is clear from the literature that the widespread and
deleterious effects of non-climate drivers on corals,
climate change is currently the strongest driver
affecting coral dynamics (Aronson and Precht, 2016)
exacerbating non-climate pressures, which together
are causing unprecedented changes in coral
communities and connected habitats (high confidence)
(Brown et al., 2017a; Hoffmann, 2002; Johnson, 2017;
Maata and Singh, 2008; Maynard et al., 2015b; Mayor,
1924; Morrison, 1999; Singh et al., 2009; Zann, 1994).
Scope and aims of the review
This review will discuss impacts of climate change on
corals according to standardised metrics. We use the
Representative Concentration Pathways (RCP)
scenarios, which provide socio-economic,
technological, land-use, emissions of greenhouse
gases, and environmental descriptions of possible
futures. RCP are used as input for climate model runs
and as a basis for assessment of possible climate
impacts and mitigation options (van Vuuren et al.,
This chapter is limited in scope to a desktop literature
review on the impacts of climate change on corals
relevant to the Pacific Island Countries and Territories
(PICTs). The review also deals with non-climate
drivers because of the synergistic effects they have
with climate drivers affecting Pacific corals. The
geographical focus of the review centres on Fiji,
Kiribati, Nauru, Papua New Guinea, Samoa, Solomon
Islands, Tonga, Tuvalu, and Vanuatu. The review is
structured as follows: firstly, we provide a general
overview of observed changes on corals associated
with non-climate drivers; secondly, we present
observed changes on coral populations related to
climate drivers, followed by the observed synergistic
effects of non-climate and climate drivers on corals;
thirdly, we offer the qualitative level of confidence in
the science for current and future changes on corals;
and, lastly we offer pressing knowledge gaps, risks
and socio-economic impacts and an overview on
adaptation options relevant to the Pacific islands.
What is Already Happening?
1. Observed changes: non-climate pressures on
Pollution, overfishing and direct physical impacts have
been degrading coastal coral reefs in the Pacific for
centuries (Hughes et al., 2017a). Globally, over eighty-
percent of marine pollution originates from land-based
wastewater, sediment and nutrients delivered via
waterways (UNEP 2008). In PICTs, most of the
pollutants relevant to corals are associated with
sewage, domestic and industrial waste, marine debris,
ship-based pollution (e.g. oil, antifouling paints,
dumped or abandoned waste from fishing vessels)
(Richardson et al., 2017), agricultural fertilizers,
sediment, pesticides and organochlorine compounds,
trace metals, and a range of emerging pollutants such
as pharmaceuticals, personal care products, and
microplastics (Lovell et al., 2004; Morrison, 1999;
Nuttall and Veitayaki, 2015; South et al., 2004). These
pollutants are known to have adverse effects on corals,
even when released at low levels (Haynes and
Johnson, 2000; Pinto et al., 2010; United Nations
Environment Programme, 2017). Below we present a
summary of the effects of each of these on corals.
Physical impacts: trampling, anchor damage,
aquarium trade, and dynamite impacts
Tourism-related and fishing impacts, such as trampling
and coral breakage due to reef gleaning and coral
manipulation by SCUBA and skin divers and fishers
are known to have decreased live coral cover and
growth, which causes also a reduction in fish
abundance (Rodgers and Cox, 2003; Sulu, 2004;
Tilmant, 1987). The problem is aggravated when large
numbers of tourists are taken to the reefs during
snorkelling activities, such as the case in Fiji (e.g.
Yasawa and Mamanuca Islands) (Lovell and Sykes,
2004:34). The construction of tourism infrastructure
(e.g. resorts, roads, dams, artificial islands) also
influence catchment dynamics and may increase
sediment and nutrient runoff into coral reefs (refer to
land-based runoff section).
Anchor damage impacts are associated with the
deployment and retrieval of anchors, as well as the
movement of the anchor rode, which can cause
damage to corals, seagrasses and other benthic
ecosystems (Kininmonth et al., 2014). Overall, data
suggests that anchor damage in the Southwest Pacific
is relatively low (Sulu, 2004), but it is recognised that
in reefs heavily used for subsistence fishing, anchors
can cause some damage to small sections of the reef
due to the use of ‘reef hook’ grapple anchors and
concrete blocks (Lovell and Sykes, 2004:34).
The aquarium trade is an important economic activity
in the Pacific (Morris, 2009; Sulu, 2004). It involves the
collection of fish and corals and is largely carried out
using best practices on a small scale and does not
pose a major threat to corals. However, the collection
of ‘live rock’ involves the breakage of the reef
framework in order to collect coralline algae, which are
also exported. This is a more damaging activity that
needs to be properly managed (Morris et al., 2008;
Morris, 2009; Sykes and Morris, 2009).
Dynamite fishing exists but is not frequently used in the
Pacific. Its use has been diminishing in PICTs due to
its relative high costs, legislation in place to ban it, and
growing awareness about the damage it causes on
marine life (Morris, 2009; Samuelu and Sapatu, 2009;
Sykes and Morris, 2009).
Both tourism-related and anchor damage impacts
have caused ecological shifts in coral communities,
also affecting social, economic and cultural issues in
the Pacific (Ayala, 2005; Engelhardt, 2005:176;
Kininmonth et al., 2014; Morris et al., 2008; Morris,
2009; Movono et al., 2018; Sykes and Morris, 2009). If
properly managed such impacts can be mitigated or
contained (Hall, 2001).
Domestic and Industrial waste
Sewage disposal leads to increase in nutrients and
organic matter (Dubinsky and Stambler, 1996).
Nutrients act as fertilisers for macroalgal growth, which
can outcompete corals for space, while organic matter
stimulates proliferation of oxygen-consuming
microbes. These may negatively impact corals and
other reef organisms, either directly by anoxia, or by
related hydrogen sulphide production (Dubinsky and
Stambler, 1996). Sewage contamination was
considered a serious problem in Pohnpei (Federated
States of Micronesia) (Morrison, 1999), Fanga’uta
Lagoon (Tonga) (Morrison, 1999; Zann, 1994), and
Suva Lagoon (Fiji) (Morrison et al., 2013; Naidu et al.,
1991; Veitayaki, 2010; Zann, 1994), leading to
increased macro and turf algae biomass, high
abundances of herbivorous sea urchins (Echinometra
matheae and Tripneustes gratillus) and detrital-feeding
holothurians, and decline of corals (Zann, 1994). Other
areas in Fiji with known water quality problems are
Nadi Bay and Lautoka harbours (from shipping,
domestic and industrial wastes, organic wastes from a
sugar mill) and Levuka harbour (discharges from a
tuna cannery) (Zann, 1994). In the 1990s lead,
pesticides, hydrocarbons, chemical wastes from
industries, and PCBs from old transformers were
considered as potentially serious problems in Tonga
(Zann, 1994), while in Port Vila (capital of Vanuatu) low
organochlorine contamination in sediments was
reported (Morrison, 1999).
Marine debris and plastics
Large concentration of marine debris and plastics
(mainly sourced from land) occur in the centre of the
Pacific Ocean Gyres (Howell et al., 2012; Maes and
Blanke, 2015; Martinez et al., 2009). In the ocean,
plastic litter will fragment in smaller particles termed
microplastics (particles less than 5 mm, even though
no scientific definition has been established) (Hall et
al., 2015; Reichert et al., 2018). In the early 1990s,
marine debris, including plastic waste and litter, was
identified as a major and growing hazard to fish,
marine mammals, turtles, sea birds, coral reefs, and
even human activities in PICTs (Morrison, 1999; Pichel
et al., 2007; Veitayaki, 2010; Zann, 1994). However,
little action has been taken to deal with this threat.
Plastic marine debris threaten corals (Reichert et al.,
2018) because they ingest particles such as
microplastics leading to either
malnourishment/starvation or the plastic may be toxic
to the corals (Hall et al., 2015). This can cause
negative effects on coral health, through bleaching and
tissue necrosis (Reichert et al., 2018). It is known that
entanglement with plastic debris increases coral
susceptibility to diseases (Lamb et al., 2018).
However, coral responses to microplastic ingestion
vary due to different cleaning mechanisms and feeding
interaction (Lamb et al., 2018; Reichert et al., 2018).
The condition of microplastics (presence/absence of
microfilms) also influence coral ingestion rates and
thus, their response to microplastics (Allen et al.,
2017). Microplastic pollution occurs all over the Pacific
with records including New Caledonia and Chesterfield
Islands (Maes and Blanke, 2015), as well as Fiji’s main
Island Viti Levu, where all samples collected from the
water column, sediment traps, fish guts and bottom
sediments between 2015-17 contained microplastic
(Ferreira, M. unpublished data).
Ship-based pollution (e.g. oil and antifouling
Sea transport that relies on fossil fuels is crucial for
PICT as it moves people, goods and resources
(Holland et al., 2014; Newell et al., 2017; Nuttal et al.,
2014). Oil spills are therefore a constant risk in the
region. The effects of oil spills (and dispersants used
in clean-up operations) on corals include extensive
death of corals and prolonged cessation of growth and
reproduction, through the decrease in the number of
female gonads per polyp, impairment of reproduction
due to planulae abortion, change in tentacle motion,
and in mucus excretion and photosynthesis of
zooxanthellae, defective reproductive tissue,
degeneration and loss of zooxanthellae, and
degeneration of mucus excretion cells and muscle
bundles, inhibition of planulae settlement and
metamorphosis (Dubinsky and Stambler, 1996).
The exposure of corals to antifouling paints (Tributyl tin
(TBT) and copper-based) can cause decreased growth
rates and reduced abundance of branching species
(Dubinsky and Stambler, 1996), inhibit coral
fertilization and larval metamorphosis (Negri and
Heyward, 2001). Antifoulant contamination is thought
to cause major mortality of resident coral communities
and have a negative impact on the recovery of adult
populations to other stressors (Smith et al., 2003).
Suva (capital of Fiji) is the main commercial, shipping
and industrial centre of the South Pacific (Morrison et
al., 2013), where TBT from antifouling paints, and
other pollutants have contributed to the decline of coral
reefs in the region since the 1990s (Zann, 1994). In
fact, in the 1990s, Suva Harbour had the highest levels
of TBT in sediments among PICTs (Lovell et al., 2004;
In PICTs, the modernisation of traditional villages via
transition from semi-subsistence to commercial-
oriented activities is often associated with the
increased and unregulated use of pesticides and other
farming chemicals, such as fertilisers, along with land
clearing, changes in land use, introduced trees, and
infrastructure construction, which increase sediment
and nutrient runoff into coastal waters thereby causing
coral reef decline (McCauley et al., 2012; Nuttall and
Veitayaki, 2015; Remling and Veitayaki, 2016;
Veitayaki, 2018). Agricultural runoff threatens
approximately 25% of the total global reef area with
further increases projected for the coming decades
(Fowler et al., 2013). Rangeland grazing, mining and
logging (extensive clearing), along with land clearing
for urban development (housing, roads, etc.) are the
main contributors of sediment, while intensive cropping
(e.g. sugarcane) and horticulture are the main
contributors of nutrients, herbicides and pesticides
(e.g. carbaryl, naphthol, chloropyrifos and
organochlorine pesticides) (Dubinsky and Stambler,
1996; Erftemeijer et al., 2012). Most of these pollutants
are delivered through waterways during periods of high
flows, and flood plumes often carry elevated
concentrations of several pollutants, simultaneously
exposing near-shore reef systems to toxic
combinations of chemicals (van Dam et al., 2011).
Increased sedimentation can negatively affect corals
through direct burial, which depletes oxygen and
leaves corals in darkness, abrasion and by reducing
light penetration, which is essential for photosynthesis
by zooxanthellae (Dubinsky and Stambler, 1996;
Rogers, 1990). This results in reduced skeletal growth
rates, species diversity, and recruitment, and
reproductive impairment (Haywood et al., 2016;
Rogers, 1990; Wicks et al., 2012). Increased sediment
runoff may also expose corals to a variety of other
contaminants, such as nutrients and heavy metals,
which interfere with coral fertilisation (copper and zinc
sulphates), larvae development and cause coral
bleaching (Dubinsky and Stambler, 1996). Poor
agricultural land-use practices (e.g. land clearing and
slash-and-burn agriculture, logging, and slope
cultivation of sugar cane and ginger) have resulted in
serious erosion of topsoil in many agricultural areas,
leading to increased sedimentation in coral reef areas
(Zann, 1994). For example, coral reefs in inner Vava'u
(Tonga) have been affected by sediment run-off from
coastal roads and earth works and the construction of
causeways (Zann, 1994). The almost total clearing of
Upolu's coastal forests (Samoa) for copra (coconut)
plantations between 1870 and 1960 caused serious
soil erosion of steep cultivated slopes in catchments of
Western Samoa; the death of some of the leeward
reefs in Western Samoa may date from this period
(Lovell et al., 2004; Zann, 1994).
There has been some research to determine whether
or not Marine Protected Areas (MPAs) have the ability
to offer some protection to corals from indirect
stressors, such as increased sedimentation, through
enhanced resilience. However, a five-year study in the
Solomon Islands comparing the response of fish and
corals to increased sedimentation both inside and
outside MPAs showed management afforded little
benefit (Halpern et al., 2013).
Release of excess nutrients into coastal waters causes
eutrophication (Cloern, 2001), resulting in macroalgae
proliferation, algal blooms (Kuffner and Paul, 2004),
and the creation of hypoxic or anoxic ‘dead zones’
(Diaz and Rosenberg, 2008), which can kill large
numbers of organisms, such as fish (Lotze et al. 2006).
Nutrients, herbicides and pesticides can kill or impair
corals (through abnormalities, blemishes, necrotic
patches, and increased incidence of diseases) (Al-
Moghrabi, 2001; Bruno et al., 2003; Furby et al., 2014;
Haapkyla et al., 2011; Kuta and Richardson, 2002;
Redding et al., 2013; Sato et al., 2016; Voss and
Richardson, 2006), and have been linked to declines
in coral cover in reef ecosystems around the world
(Dubinsky and Stambler, 1996; Restrepo et al., 2016).
The direct effects of nutrients (nitrogen (N) and
phosphorous (P)) on corals are diverse. Nutrients can
increase a) the amount of chlorophyll a within
zooxanthellae cells, b) zooxanthellae density within
coral hosts, and c) chlorophyll a density per unit area
of coral (Shantz and Burkepile, 2014). N reduces
calcification and increases photosynthetic rates, while
P alone can increase calcification with little effect on
photosynthetic rates. The effects of nutrients on coral
communities depend on taxa, enrichment source and
nutrient identity (Shantz and Burkepile, 2014). Poor
water quality is also related to COTS outbreaks
(Berthe et al., 2016; De'ath et al., 2012; Vanhatalo et
al., 2017). In Fiji, increased erosion associated with
land clearing, and poor agriculture practice has led to
increased nutrient levels resulting in seagrass and
macroalgae expansion in less flushed, leeward reefs
(Zann, 1994). The coral decline in the Fanga’uta
Lagoon (Tonga) was associated with high levels of
DDT (Morrison, 1999; Zann, 1994) and eutrophication,
which also contributed to the advance of seagrasses
and mangroves (Zann, 1994).
Coral reef overfishing reduces the abundance and size
of fish with consequences for reproduction. Effects of
overfishing on coral reefs include reduction in: a) live
coral cover, b) calcareous and coralline algae cover, c)
diversity of substratum, and d) topographic complexity.
In countries that target herbivorous species (as is the
case in many PICTs), overfishing can also accelerate
bioerosion, facilitate the growth of macroalgae, and
reduce biodiversity and fish productivity (Doropoulos et
al., 2013; McManus and Polsenberg, 2004; Sandin et
al., 2008; Stuhldreier et al., 2015) (see Science Review
2018: Fish and Shellfish and Coastal Fisheries, and
Oceanic Fisheries papers).
Population growth, increased commercialisation,
promotion of fisheries development projects and the
live reef fish trade are currently increasing demand for
reef fish in PICTs and thus increasing reef fishing
pressure in the region (Cinner and McClanahan, 2006;
Nuttall and Veitayaki, 2015; Veitayaki, 1995, 2018;
Veitayaki and South, 1997). Available data indicate
that more than two thirds of the reef fishes and
invertebrates are taken for subsistence (Bell et al.,
2017; Zeller et al., 2015) and overfishing of reef
resources is common throughout the Pacific (Pinca et
al. 2010; see also Science Review 2018: Fish and
Shellfish and Coastal Fisheries, and Oceanic Fisheries
papers). The primary driver of the decline in reef
fisheries is small-scale local fisheries that target reef
fish and invertebrates for food, with the demand for fish
growing as populations increase in most PICTs (Bell et
al., 2017; Veitayaki and Ledua, 2016; Williams et al.,
Conclusion on effects of non-climate drivers on
It is widely acknowledged that non-climate drivers are
damaging corals all over the world (Halpern et al.,
2008; Rasher et al., 2012). Increased land-based
pollution (sediments, nutrients, chemicals) is a key
driver of coral degradation worldwide (Rogers, 1990)
and in the Pacific (Lal, 2016), along with overfishing
(Jackson et al., 2001) and increase in the incidence of
COTS outbreaks (Berthe et al., 2016). In addition to
causing the decline of coral communities, these non-
climate drivers also limit the recovery capacity of coral
assemblages (Aronson and Precht, 2016; Jackson et
al., 2001). The synergistic effects of overfishing,
excessive nutrients and sediments, and pollution on
corals are extremely negative (Dubinsky and Stambler,
1996; Haywood et al., 2016; Rasher et al., 2012;
Rogers, 1990; Wicks et al., 2012).
One of the major concerns in PICTs is the lack of
baseline data and long-term monitoring programmes
addressing water quality and (over)fishing effects on
corals for adequate system understanding on cause-
effect relationships and proper assessment of non-
climate impacts on corals and management actions
(Zann, 1994). It is important to note that by effectively
managing non-climate impacts on corals, communities
can boost the capacity of reef ecosystems to survive
climate change (Hughes et al., 2017a).
2. Observed changes: climate pressures on
Despite the detrimental effects of non-climate drivers
on coral communities, accelerating climate change is
currently the strongest driver affecting coral dynamics
(Aronson and Precht, 2016). Sea level rise, high ocean
temperatures, ocean acidification and the synergistic
effects that climate drivers have with one another and
with non-climate drivers are currently causing
widespread negative effects on PICT corals. For
example, calcification rates are affected by both ocean
pH and temperature. Current maximum calcification
rates are just 2-3 oC below the maximum temperature
corals can withstand before thermal bleaching
(Evenhuis et al., 2015). When corals bleach,
calcification is further supressed because
photosynthetic products from the zooxanthellae are
essential for the calcification process (Evenhuis et al.,
Sea level rise
Despite the evident negative effect of sea level rise
(SLR) on Pacific islands and atolls (Nurse et al., 2014),
it may also provide some opportunities to corals. SLR
effects on corals are complex and not straightforward
to understand because corals grow vertically and, in
principle, additional depth provides extra
‘accommodation space’, in which corals could expand
in intertidal areas and also colonise new inundated
areas – provided that suitable substrate is available
(van Woesik et al., 2015; Woodroffe and Webster,
2014) – thus increasing live coral cover (Albert et al.,
2017). Evidence from Tuvalu strongly suggest that
Pacific islands can persist on reefs under SLR rates
around 4mm.yr-1 (commensurate with RCP 2.6
scenario) (Kench et al., 2018). However, it is unclear
whether islands will continue to maintain their sizes
under rising seas of up to 0.82m by 2100 (RCP 8.5)
(Kench et al., 2018). Under such scenario it is more
likely than not that rising seas will inundate coastal
areas and promote coastal erosion (Barros and Field,
2014; Kench et al., 2018), thus increasing turbidity and
sedimentation in coastal waters negatively affecting
corals and other reef organisms (Brown et al., 2017a;
Brown et al., 2017b; De'ath and Fabricius, 2010). In
addition to this, a recent study comparing the accretion
rates of coral reefs in the Indian and Atlantic Oceans
showed that without sustained ecological recovery,
very few reefs in either ocean would be able to track
the projected sea level rise under RCP4.5 or RCP8,
potentially resulting in reef submergence. This will lead
to changes in wave energy regimes, increasing
sediment mobility, shoreline change and island
overtopping (Perry et al., in press). Human-induced
SLR is associated with the melting of polar ice sheets
and water expansion as a result of increased water
temperatures (IPCC, 2014), which itself cause
negative effects to corals (see below) and have
already caused the displacement of PICT people
Increase in Ocean Temperatures
Sea surface temperature (SST) plays an integral role
in the growth of corals by influencing their skeletal
density, extension and calcification rate (Carricart-
Ganivet, 2004). Scientists have high confidence that
anthropogenic-induced ocean warming is impacting
coral reefs through thermal coral bleaching (Adjeroud
et al., 2009; Cumming et al., 2000; Davies et al., 1997;
Kleypas et al., 2015; Lovell et al., 2004; Mangubhai,
2016; Obura and Mangubhai, 2011; Rotmann, 2001)
and by increasing bioerosion (Chaves-Fonnegra et al.,
2017), reducing coral calcification rates (De'ath et al.,
2009; Nurse et al., 2014) and influencing mass coral
spawning events over large geographical areas (Keith
et al., 2016).
Coral bleaching impacts coral species differently
depending on the species of the coral host affected
and its zooxanthellae symbiont (Hoadley et al., 2015).
This means that coral species and their symbionts
respond and adapt differently to rising SST (Barshis et
al., 2013; Berkelmans and van Oppen, 2006; Oliver
and Palumbi, 2011). Bleaching also affects colony size
(favours smaller size corals), the time of coral
spawning (Paxton et al., 2016), slows down swimming
of coral larvae and reduces the number of viable
recruits (Singh, 2018).
Coral reef bleaching events have increased
dramatically since the early 1980s. Most of these have
been caused by anomalous increases in SST, reduced
cloud cover, higher than average air temperatures and
higher than average atmospheric pressures (Glynn,
1983; Hoegh-Guldberg, 1999; Hoegh-Guldberg et al.,
2007; Hughes et al., 2003; McGowan and Theobald,
2017; Nunn, 1990), which are associated with El Niño
or La Niña events. Six distinct peaks in bleaching have
occurred worldwide between 1976 and 2016 (Le
Nohaic et al., 2017; Oliver et al., 2009). The 1997-98
El Niño and the 1998-99 La Niña were the most severe
and spatially extensive on record (Oliver et al., 2009).
The recent severe thermal coral bleaching events in
2015-16 have caused extensive habitat and
biodiversity losses around the world and in the Pacific
(Anthony et al., 2017; Gardner et al., 2017; Le Nohaic
et al., 2017). Interestingly, the south-western Pacific
islands were little affected by the mass coral bleaching
resulting from the 1998 El Niño – possibly due to
under-reporting (Oliver et al., 2009) – but were heavily
impacted by the subsequent strong La Niña event in
2000 (Cumming et al., 2000) and the 2015-16 El Niño
(Mangubhai, 2016; Wolanski et al., 2017).
Some examples of coral bleaching for locations
relevant to this review include the northern and
southern parts of Papua New Guinea (PNG) in early
1998 (Goreau et al., 2000), Lihir Island group in March
1996 (Done, 1996), February-March 2001 (Rotmann,
2001) and February 2006 (Flynn, 2006). Also in PNG,
coral bleaching was recorded in the Milne Bay
Province in February 1996 (Davies et al., 1997).
Thermal coral bleaching was registered throughout the
Fijian archipelago during the first half of 2000
(Cumming et al., 2000), and around Gau Island and
the Vatu-i-ra seascape in 2015-16 (Mangubhai, 2016).
In 2016, increased thermal stress reduced oxygen
concentrations in seawater promoting mass fish kills
on the Coral Coast of Fiji, where water temperatures
were as high as 35 °C (Singh, 2018). In Samoa,
bleaching and extensive coral mortality were recorded
at Ofu Island in 1994 and 2015 (Craig et al., 2001;
Thomas and Palumbi, 2017), Palolo Deep Marine Park
(Western Samoa) in 1987-1988, and a major episode
occurred around Upolu and Savai’i between April and
July 1991 (unpublished data) (Zann, 1994). In Kiribati,
the greatest thermal stress event reported in the coral
reef literature occurred at the Phoenix Islands in
January 2003, which persisted for 4 years causing
extensive coral bleaching and mortality (Obura and
Mangubhai, 2011). Still in Kiribati, extensive coral
bleaching occurred at the Gilbert Islands in 2004 and
2009 (Carilli et al., 2012). In Moorea (French
Polynesia), extensive coral bleaching and mass
mortality were registered in 1978, 1982, 1998, and
2003 (Riegl et al., 2015). Seawater temperatures as
high as 30 °C were reported in Vanuatu causing
widespread fish kills (Johnson et al., 2016a).
SST acts synergistically with nutrients and sediments
amplifying bleaching effects and also influencing the
recovery period from bleaching. As a result, corals
currently sheltered from effects of excessive nutrients
and elevated temperatures will be the most at risk from
ocean acidification (Riegl et al., 2015).
Ocean acidification is the increase of partial pressure
of CO2 and associated decline in seawater pH (Enochs
et al., 2016) (see Science Review 2018: Ocean
Acidification paper). For scleractinian corals, probably
the most significant consequence of ocean
acidification is the decrease in the concentration of
carbonate ions (CO32-). As coral’s skeletons are made
from the mineral phase of calcium carbonate, called
aragonite, the saturation state of aragonite (arg) is
often related to rates of calcification (Evenhuis et al.,
2015). Increases in atmospheric CO2 concentrations
are negatively correlated with arg, which decreases
the rates of coral calcification (Pandolfi et al., 2011).
Present-day aragonite saturation levels are postulated
to be close to the point where it can weaken coral
skeletons and slow coral growth. Weaker reef systems
will be far more susceptible to other pressures
including bioerosion, eutrophication, coral disease,
intense storms and bleaching because coral skeletons
become more fragile (Meissner et al., 2012; Nuttall and
Veitayaki, 2015; van Hooidonk et al., 2014).
Our review did not identify any field studies pointing to
conclusive cause-effect relationships attributable to
anthropogenic ocean acidification and weakening of
coral skeletons in the Pacific Ocean. However, there is
emerging evidence of the impacts of ocean
acidification on coral reefs (Barros and Field, 2014). It
is likely that, in many places, responses of corals to
more acidified waters are not yet outside natural
variability and may be influenced by local issues and
mechanisms (e.g. circulation patterns and pollution)
(Barros and Field, 2014).
Field observations of reefs growing in more acidified
waters suggest that reduced pH promotes changes in
the abundance and structure of coral communities
(Enochs et al., 2015; Enochs et al., 2016; Fabricius et
al., 2011), threatening the ecosystem functioning of
coral reefs (Done, 1992; Hughes et al., 2007; Mumby,
2009; Wild et al., 2011). However, the literature offers
mixed responses of corals to more acidified waters.
For instance, a reef in Palau (Nikko Bay) growing in
waters with pH of 7.8 has coral cover 2 to 6 times
higher than other coral reefs around the island growing
in waters with higher pH (Golbuu et al., 2016), whereas
in PNG a recent decrease of 0.1 units in pH may have
affected coral diversity and recruitment (Fabricius et
al., 2011), and the abundance of crustose coralline
algae (CCA) – an inducer of coral larvae settlement
(Fabricius et al., 2015). These negative effects on
corals most likely facilitate macroalgae growth,
causing a shift from a coral-dominated to algal-
dominated state (Enochs et al., 2015). In PNG massive
Porites dominate reefs with pH > 7.8 and no distinct
evidence of calcification exists in places with pH ≤ 7.7
(Fabricius et al., 2011), which is consistent with a
switch from net calcification to net dissolution of coral
reefs at a pH of roughly 7.8 (Enochs et al., 2016).
Similar effects of more acidified waters on corals were
also registered in waters of Japan and the Northern
Mariana Islands (Enochs et al., 2015; Enochs et al.,
2016; Fabricius et al., 2011).
Several laboratory studies suggest that more acidified
waters impair calcification and accelerate the
dissolution of coral skeletons thus weakening coral
skeletons, and triggering stress-response
mechanisms, which affect the rates of tissue repair,
feeding rate, reproduction, and early life-stage survival
(D'Angelo et al., 2012; Enochs et al., 2015; Fabry et
al., 2008; Kroeker et al., 2010). Possible responses of
reef building corals to reduced calcification include a)
decreased linear extension rate and skeletal density
(Cooper et al., 2008), b) the maintenance of physical
extension rate, but reduced skeletal density, leading to
greater erosion (Szmant and Gassman, 1990), and c)
maintenance of linear extension and density but
greater investment of energy diverting resources from
other processes such as reproduction (Albright and
Mason, 2013; Szmant and Gassman, 1990).
In PICTs, cyclones regularly contribute to changes in
island geomorphology (Duvat and Pillet, 2017),
causing extensive damage to corals by exacerbating
nearshore biogeochemical changes associated with
land-use, runoff, and human activities (Levin et al.,
2015). These include physical damage through wave
action which breaks corals (with branching and foliose
growth forms being more vulnerable to wave damage),
and associated flood events which cause substantial
soil erosion leading to increased sediment and nutrient
runoff into coral reefs (Veitayaki, in press), which
creates a non-optimal environment for coral recovery
(Gardner et al., 2005; Guillemot et al., 2010; Zann,
1994). Since most of the coral fragments are washed
into unfavourable environments, and these
scleractinian corals take years to reach spawning age,
breaking them off often results in the death of the coral
fragment (Cheal et al., 2002; Gardner et al., 2005).
Reefs stressed by pollution, overfishing and other
human influences may be slower to recover after
cyclones (Zann, 1994).
Despite the fact that cyclones are part of ocean-
atmosphere tropical dynamics, there is evidence of a
correlation between an increase in SST and
associated increase in the intensity of tropical
cyclones, and increased coral stress (Elsner et al.,
2008; Nott and Walsh, 2015; Terry, 2007). The
proportion of intense tropical cyclones (category 4-5)
versus weaker cyclones (category 1-2) has increased
substantially in the last 40 years, which coincides with
most of the current anthropogenic warming (Holland
and Bruyere, 2014).
The physical destruction associated with cyclones has
been identified as an important feature of coral reef
ecosystems, because it opens up space in the reef for
new settlement which increases biodiversity (Adjeroud
et al., 2009; Connell, 1978; Connell, 1997). Some of
the observed impacts of high-intensity cyclones in
PICTs include the loss of coral reef and mangrove
areas after TC Pam in Vanuatu (Johnson et al.,
2016a), and TC Winston in Fiji (Mangubhai, 2016). In
1991 cyclone Val caused widespread damage to the
coral reefs around American Samoa and a subsequent
increase in coral recruitment following the cyclone
(Mundy, 1996). Widespread coral destruction was
seen in New Caledonia after TC Erica (1993), where
the reefs took more than 3 years to recover (Guillemot
et al., 2010). Although the same trend of coral
destruction was expected on all reefs that lay in the
path of the category 5 TC Winston in Fiji, this was not
the case. Winston, caused widespread physical
damage to coral reefs around the Vatu-i-ra passage
(Mangubhai, 2016), and Navakavu reefs near Suva
(K.J. Brown and L.X.C Dutra, pers. obs.) resulting in a
severe decline in fish and coral assemblages, while in
other sites surveyed in the Coral Coast (Singh, 2018),
there was little or no damage to the reefs. The coral
reefs that were not directly on the path of the eye of the
cyclone but that have been affected indirectly, seemed
to have recovered better and more rapidly (Singh,
Conclusions on impacts of climate drivers on
Simultaneous climate change drivers, such as sea-
level rise, ocean warming, acidification and more
frequent high-intensity cyclones, act together with local
changes (e.g., pollution, eutrophication, overfishing)
(high confidence), leading to interactive, complex and
amplified impacts for species and ecosystems, which
complicates marine management regimes (Barros and
Field, 2014; Valmonte-Santos et al., 2016). The ability
of corals to adapt naturally to changes in community
composition and structure appears limited and
insufficient to offset the detrimental effects of climate
change (Barros and Field, 2014). Our review indicates
that the current understanding of climate change
effects on corals and adaptation mechanisms is very
limited and requires additional field and laboratory
studies to fully comprehend adaptation mechanisms,
tolerance and ecological consequences of climate
drivers, especially under low emission scenarios
What Could Happen?
Future climate change and impacts on corals
The extent of impacts of climate change on corals will
depend on how close reality will be to future RCP CO2
emission scenarios (Gattuso et al., 2015). However,
common across all RCP scenarios is that atmospheric
CO2 concentrations will continue to rise, at least in the
coming decades, thus increasing ocean temperatures
and acidity (high confidence) (see Science Review
2018: Fish and Shellfish and Coastal Fisheries, and
Oceanic Fisheries, Sea Level Rise, Sea Surface
Temperature, Ocean Acidification and Extreme Events
papers for predictions, Evenhuis et al., 2015; Gattuso
et al., 2015; Hoegh-Guldberg et al., 2011; IPCC,
2014). Increased temperatures are also expected to
increase the intensity of TC (low confidence) (Elsner et
al., 2008). By 2100, the greatest impacts from climate
change on corals will likely occur as the result of
increased temperature stress, ocean acidification,
increased sedimentation and turbidity from more
extreme events (e.g. rainfall and cyclones), rising sea
level, and physical damage from the combination of
rising sea levels and more severe cyclones (Johnson
et al., 2017; United Nations Environment Programme,
The future is uncertain but PICT coral reefs will most
likely substantially change – rather than completely
disappear – as a result of climate change and
associated changes in the sea- and landscape, which
will impact livelihoods (Hughes et al., 2003). As most
of the impacts of climate change on corals from the
literature are related to increased SST, ocean
acidification and cyclone damage, we detail these
three major threats below, followed by a review on
synergistic effects of climate and non-climate drivers
Sea surface temperature
Higher SST predicted in all RCP scenarios will
increase the extent, frequency and severity of coral
bleaching events (van Hooidonk et al., 2016), the
incidence of coral diseases, and the potential for mass
coral mortality (high confidence) (IPCC, 2014; Langlais
et al., 2017; Maynard et al., 2015a; Reisinger et al.,
2014; Ruiz-Moreno et al., 2012; Sheridan et al., 2014).
In addition, increased SST can alter growth patterns,
body size, immune defence, feeding behaviour, and
reproductive success of corals (Bonesso et al., 2017;
Riegl et al., 2015; Ruiz-Moreno et al., 2012), and may
even affect mesophotic, deep-sea and cold-water coral
distribution (although evidence of the direct effect of
climate change is less clear) (Hoegh-Guldberg et al.,
2017). Changes are expected to become substantial
from 2050 (or earlier) under a high emission scenario
(Figure 1). Hence, higher SST will potentially lead to
declines in coral populations, which will cause
unprecedented changes in coral reefs and associated
habitats (high confidence) (Figure 1, Hughes et al.,
2017a; IPCC, 2014; Morrison, 1999). A similar
phenomenon occurred in Espiritu Santo (Vanuatu)
during the early Holocene (Beck et al., 1997).
Figure 1. The vulnerability of marine communities to species loss
due to warming in the Pacific Ocean. Proportion of fish and
invertebrate species in present-day communities likely to exceed
their upper realised thermal limit by 2025 (A) and 2115 (B) based
on regional IPCC warming rates (RCP8.5 scenario). Sites with
confidence scores <2.5 were excluded from most ecoregion means
(details in Stuart-Smith et al., 2015).
The literature offers, however, some potential benefits
of warmer waters to corals. For example, although
highly speculative, in the short term, rising average
temperatures could promote coral growth and
compensate partially for long-term shifts in aragonite
concentration, as some corals (e.g. Porites) have
shown an increase in calcification rates associated
with an increase in ocean temperature (~0.10 °C per
decade) between 1990 and 2010 (Hughes et al.,
2017a). Warmer waters may also facilitate the
expansion of the geographical range of corals into
higher latitudes, and cause shifts in species
composition in response to differences in susceptibility
to climate and non-climate drivers (Hughes et al.,
2017a; Woodroffe et al., 2010).
Ocean acidification is expected to increase in the
future under all RCP scenarios (IPCC, 2014).
Variability in responses due to local environmental
conditions and species composition is likely to be
substantial (Chan and Connolly, 2013), which along
with lack of evidence of contracting geographical range
of calcifying species towards the Equator (Hughes et
al., 2017a) contribute to the low confidence on the
effects of ocean acidification on corals (Barros and
Field, 2014; IPCC, 2014; Johnson et al., 2016b).
Recent modelling suggests that under RCP2.6, CO2
concentrations are expected to be around 450 ppm by
2100, where pH will likely be between 7.9 and 8.1 (an
average decline of 0.07 pH units from Present day
mean ocean pH (Gattuso et al., 2015)) in most tropical
oceanic waters. For medium- to high-emission
scenarios (RCP4.5, 6.0, and 8.5), ocean acidification
poses far more risks to coral reefs. For example, under
RCP8.5, CO2 concentration in the sea is expected to
reach 450 ppm by 2030 – 2050 (Salvat and Allemand,
2009) and a decline of 0.33 units of pH is expected by
2100 (Gattuso et al., 2015), which would seriously
threaten corals and other calcifying organisms. Such
an increase in ocean acidification is expected to impact
on coral physiology (calcification rates, ability to repair
tissues and growth), behaviour (feeding rate),
reproduction (early life-stage survival, timing of
spawning) as well as weaken calcified structures, and
alter coral stress-response mechanisms (Fabricius et
al., 2015; Fabricius et al., 2011), with impacts on
population dynamics of individual species from
phytoplankton to animals (medium to high confidence).
The vulnerability of reef-building corals to ocean
acidification has potentially detrimental consequences
for fisheries and livelihoods (Barros and Field, 2014).
Sea level rise and cyclones
Rising sea level would allow minimum shore zone for
mangrove development or shore stability; with an early
period of increased coastal erosion and destabilizing
sedimentation as catchment streams and shorelines
adjusted to new and changing base levels, thus
negatively affecting corals (Dutra and Haworth, 2008).
Therefore, widespread degradation of corals is
expected due to coastal erosion associated with SLR.
The observed destruction that high intensity tropical
cyclones have caused to corals in PICT is well known
(Adjeroud et al., 2009; Guillemot et al., 2010;
Mangubhai, 2016). However, the observed increase in
Category 4–5 cyclones may not continue at the same
rate with future global warming. Following an initial
climate increase in intense cyclone proportions a
saturation level will be reached beyond which any
further global warming will have little effect (Holland
and Bruyere, 2014).
Synergistic climate and non-climate drivers
Climate change impacts, such as increased SST and
sea level, and changes in ocean chemistry, ocean
circulation, rainfall and storm patterns are expected to
compound existing anthropogenic pressures on coral
reefs from habitat degradation due to development,
resource extraction, overfishing and pollution.
Interactions between climate and localized stressors
such as pollution are expected to create particularly
damaging synergies, adding to concerns about coral
reefs globally (Bridge et al., 2013). For example, corals
exposed to nutrients, turbidity, sedimentation, or
pathogens have been shown to be more susceptible to
thermal bleaching, or less able to survive and recover
a bleaching episode, or an acute disturbance, such as
a cyclone (Hoegh-Guldberg et al., 2007). This is
because fertilization and larval recruitment in corals
are particularly sensitive to environmental conditions
(Lam et al., 2015), and because macroalgal growth
rates increase in nutrient-rich waters, thus
outcompeting corals (McCook, 1999).
Predicted changes in drought and flood regimes may
exacerbate nearshore biogeochemical changes
associated with land-use, runoff, and human activities
(Levin et al., 2015), increasing nutrient and sediment
loads to corals. Warmer waters will potentially increase
the incidence and extent of COTS outbreaks in the
Pacific (Zann, 1994) because it likely 'pushes' well-fed
larvae faster to settlement (Uthicke et al., 2015).
Ocean warming and acidification may also enhance
the success of COTS juveniles, because more
acidified waters facilitate growth and feeding rates on
coralline algae (Kamya et al., 2016; Kamya et al.,
2017; Uthicke et al., 2015) although the affects have
been shown to be variable depending upon the genetic
makeup of the COTS (Sparks et al., 2017). Warmer
and more acidified waters can also facilitate the
poleward migration of COTS (Kamya et al., 2017)
following the expected migration of corals, but in
contrast to their coral prey, COTS juveniles appear to
be highly resilient to future ocean change (Kamya et
Coral bleaching is exacerbated by ocean acidification,
which impairs coral calcification and the recovery
potential of coral species (Anthony et al., 2017). Coral
bleaching also facilitates bioerosion by sponges
(Carballo et al., 2013). Plastic pollution can act alone
or synergistically with increased SST and more
acidified waters to increase the incidence of coral
diseases (Lamb et al., 2018). Coastal reef fisheries are
generally considered overexploited in most PICT
(Matthews et al., 1998; Valmonte-Santos et al., 2016),
which negatively affect coral reef ecology due to
ongoing pressures as a result of human population
growth, which further increase fishing effort and catch,
thus facilitating macroalgal growth.
The literature offers very little information on climate
change impacts on corals associated with a low-
emission scenario (RCP2.6), due to technical and
other issues (Hughes et al., 2017a). Climate change
impacts on corals are often tested in the laboratory
using values associated with a high emission scenario
(RCP8.5) (Hughes et al., 2017a), which are
Higher SST will increase the incidence of thermal coral
bleaching with some Pacific nations experiencing
annual severe bleaching before 2040 (e.g. Nauru,
Guam, Northern Marians Islands) and all nations
experiencing severe annual bleaching by 2050 under
RCP8.5 (van Hooidonk et al., 2016). This trend is
already evident with the largest coral bleaching event
in the Pacific documented in many nations during the
2016 summer, and the Great Barrier Reef
experiencing two successive severe bleaching events
in 2016 and 2017.
While a lower emissions scenario (RCP4.5) adds 11
years to the global average year of severe annual
bleaching, 75% of reefs will still experience annual
bleaching before 2070. Spatial patterns of bleaching
are variable, with high latitude reefs faring better than
reefs near the equator (van Hooidonk et al., 2016).
Only significant mitigation of greenhouse gas
emissions (RCP2.6) will delay the onset of annual
bleaching until 2100.
The area of the Pacific Ocean with suitable chemistry
for coral growth will decrease by 15% by 2050, with
conditions declining to marginal and only some areas
in the Central Pacific being ‘adequate’ for coral growth
by 2100 (Johnson et al., 2017; Lenton et al., 2015). By
2050, Ω is expected to be ~3 (Johnson et al., 2016b),
which will result in poor conditions for coral growth
across much of the region (Chan and Connolly, 2013;
Hoegh-Guldberg et al., 2007).
The combined effects of increased coral bleaching and
ocean acidification are expected to reduce live coral
cover by 50–75% by 2050, and as much as 90% by
2100 using RCP8.5 (Freeman et al., 2013; Hoegh-
Guldberg et al., 2011). Global reef habitat loss was
predicted to be 60% by 2100 using RCP4.5(Freeman
et al., 2013). This will intensify the competition
between reef-building corals and non-calcareous
macroalgae as coral cover, growth and calcification
continue to decrease (Cooper et al., 2008; De'ath et
al., 2009). Macroalgae are likely to outcompete corals
for space and become a dominant feature of reefs by
2100 (De'ath et al., 2012).
Coral reefs and adjacent ecosystems such as
mangroves and seagrasses are intrinsically linked
through sediment capture, localised pH buffering and
linked life cycles of some fish species (Albert et al.,
2017; Atkinson et al., 2016). Together, the three
habitats substantially improve coastline protection than
any one habitat alone (Guannel et al., 2016).
Therefore, any declines in any of the three ecosystems
will most likely affect the other two. Mangrove areas in
some locations are projected to decline due to sea-
level rise (where there are barriers to landward
migration) (Veitayaki et al., 2017), more intense
cyclones, and changes in rainfall patterns (Waycott,
2011). Seagrasses have already shown sensitivity to
the effects of increased turbidity due to flooding
(McKenzie et al. 2012) and their area is expected to
decrease by 5–35% by 2050 due to increased runoff
from more extreme rainfall as well as increases in
cyclone intensity (Waycott, 2011).
Recent research has demonstrated that the
combination of ocean acidification (low aragonite
saturation) and nutrient loading is ten times more
effective at driving coral macrobioerosion than ocean
acidification alone (DeCarlo et al., 2015). This has
significant implications for corals and reefs, resulting in
slower growth rates and compromised reef structures.
The Pacific island region will experience similar
changes and, if emissions are not seriously curbed and
restricted to the levels of a ‘low-emission’ scenario
(RCP2.6) coral calcification is expected to decline
along with reef area and structural integrity.
Despite the heterogeneity in responses of coral to
ocean acidification (Comeau et al., 2014), even a slight
decrease in aragonite saturation level may weaken the
skeletons of calcareous organisms and shells. In this
state, reef systems will be far more susceptible to other
pressures including eutrophication, coral disease,
storms and bleaching, which are also projected to
increase in frequency due to climate change (Meissner
et al., 2012; van Hooidonk et al., 2014). The interaction
between ocean acidification and locally high nutrient
loading also accelerates coral reef loss (DeCarlo et al.,
The synergistic and cumulative impacts of extreme
events caused by climate change, such as thermal
coral bleaching, floods and tropical cyclones, and the
chronic impacts of poor water quality, will continue to
be major drivers of future reef degradation in the
Pacific Ocean (Johnson et al., 2013). Impacts of
climate change drivers in isolation, or in combination
with non-climate drivers, include reduced areas of
living coral, increased macroalgal cover, compromised
reef structure, higher incidence of COTS and disease
outbreaks, reduced species diversity, and lower fish
What is already happening
The scores are based on the significant body of
experimental and in situ evidence and scientific
consensus on how corals are: (i) sensitive to different
climate variables; (ii) responding to changing climate
variables, particularly to increased SST, intense
storms and reduced pH; and (iii) how they are likely to
respond to future climate change. The consensus is
very high with minimal disagreement that climate
change is happening and affecting corals in PICT, in
particular with regards to evidence of increased
thermal-induced coral bleaching and COTS outbreaks,
and increased incidence of high-intensity cyclones.
The data used in global projections is still strong
however, with limitations with regards to a) data
specific for PICTs to couple regional and global scale
models, b) lack of baseline data for climatic and non-
climatic drivers. Overall, this results in substantial
uncertainties around the level and extent of impacts of
ocean acidification. As instrumentation is deployed
and effects are enhanced this will rapidly change.
What could happen in the future
The future remains relatively uncertain because it is
still impossible to predict the exact amount of CO2
emissions in the next decades, and thus the exact
extent of impacts associated with higher SST, ocean
acidification, and cyclones. Most of the studies about
coral responses to climate change use values from
‘worse case scenarios’ (RCP8.5) and very few deal
Amount of evidence (theory /
observations / models) modelled)
Amount of evidence (theory /
observations / models) modelled)
with projections associated with a low-emission
scenario (RCP2.6). This leaves important knowledge
gaps on how corals can cope with or adapt to human-
induced climate change under a low emission
1. Socioecological linkages to complex changes
in environment. This is a recent trend but there
is relatively high consensus about socio-
economic and governance issues affecting
resource sustainability and adaptation.
2. Species tolerance within a multidimensional
non-stationary niche. This was a priority
identified by the authors of this chapter.
Models of the likely shifts in coral species distributions,
and the effects of climate-induced changes in species
composition on ecosystems and maintaining coral-
dominated reefs to identify refugia and potential areas
of habitat losses and to assist in regional and
community planning. There is high consensus on the
influence of climate change on coral reef ecosystems
but there is still high uncertainty on how the reef
ecosystem will cope with these changes, especially
when non-climate drivers are considered.
Our review identified that climate pressures will
exacerbate non-climate pressures (e.g. pollution and
overfishing) (high confidence) (Aronson and Precht,
2016; Sheridan et al., 2014; Wild et al., 2011) and
negatively affect corals. Predicted coral decline will
cause unprecedented socio-economic changes with
consequences to human populations from across the
Pacific Ocean in coral reef and associated habitats
(high confidence) (Hughes et al., 2017a; IPCC, 2014;
There are significant socio-economic implications in
PICTs from the combined effects of population growth
and reef degradation due to climate change (Johnson
et al., 2016b). Coral degradation directly affects the
livelihoods of Pacific Islanders through reduced local
income and inflow of foreign exchange due to the
decline of fisheries. Current projections anticipate
declines in reef fisheries productivity of as much as 10-
20% in the western Pacific under climate change due
primarily to habitat degradation (Pratchett et al. 2011,
Bell et al. 2016). Decline in fisheries will increase food
insecurity (given the high levels of dependency Pacific
Islanders have on reef fishes for subsistence) (Zeller
et al., 2015). Reef degradation will also affect reef-
based tourism (Sajjad et al., 2014), reduce coastal
protection, increase the risk of property damage
(Guannel et al., 2016), and promote negative effects
on aesthetics, cultural connections between traditional
communities and their marine environments, spiritual
values, and other ecosystem services that contribute
to human wellbeing (Hughes et al., 2017a; Johnson et
A regional assessment of all 22 PICTs (Johnson et al.,
2016b) found that nations with large reef areas and
predominately coastal communities had the highest
relative vulnerability to climate change impacts on
reefs because of high reef to land area, dependence
of household incomes on coastal fisheries (for food
and livelihoods), aquaculture (for jobs) and tourism (for
jobs and contribution to GDP), and low education
standards. Based on this assessment, the most
vulnerable PICTs to climate change are: Solomon
Islands, Kiribati, Papua New Guinea (PNG), Federated
States of Micronesia (FSM), Tonga and Tuvalu. The
PICTs that had the lowest relative vulnerability to
climate change impacts on reefs and the goods and
services they provide are: Niue, Commonwealth of the
Northern Mariana Islands (CNMI), Tokelau, New
Caledonia and Guam.
The history of coral reef evolution shows that they are
resilient and adaptable to environmental changes, for
example in sea level and sedimentation (Albert et al.,
2017; Camoin and Webster, 2015; Done et al., 2007).
Similarly, human communities in PICTs have been
adapting to changes and managing reef ecosystems
through traditional practices, such as temporal
closures and seasonal bans, for centuries (Reenberg
et al., 2008; Remling and Veitayaki, 2016; The Locally-
Managed Marine Area (LMMA) Network, 2014;
Veitayaki, 2014; Veitayaki and Holland, in press).
However, the capacity of the coupled human-reef
ecosystems to persist and adapt is under threat
because non-climate threats will be exacerbated with
changes in climate (Bellwood et al., 2004; Hoffmann,
Acclimatisation or adaptation of corals to climate
change depends on intrinsic characteristics of both
coral and zooxanthellae species (Dandan et al., 2015),
as well as on governance and other social-economic
characteristics, such as education and poverty. For the
sake of presentation and summarising information we
divide adaptation options in two main groups. These
are: a) direct interventions to reduce non-climate
stresses, and b) socio-economic, governance, and
technological interventions, which are discussed in
Direct interventions to reduce non-climate
Decisions about the most appropriate options for
PICTs to adapt to the effects of climate change must
take into account the many other drivers that affect the
ability of coral reefs to maintain their structure and
function, and support local communities (Gillett and
Cartwright, 2010). A range of practical adaptation
options have been identified by Bell et al (2013; 2011)
and Johnson (2013), unless otherwise specified.
These include managing land use and vegetation in
catchments to reduce the transfer of sediments and
nutrients into coral reefs, the prevention of marine
pollution, and improve waste management. Also, as
corals are impacted by climate change, their
functioning and recovery mechanisms will depend on
the abundance and diversity of fish that support coral
recruitment and settlement, such as herbivores. At the
same time, sustainable reef fisheries depend on
healthy coral reef habitats to support fish and
invertebrates. Actively managing reef fisheries via
reducing fishing effort and improving enforcement on
gear, catch restrictions and quantity, and enforcement
of non-take areas (Edgar et al., 2014) are therefore
essential to improve overall coral health and support
reef resilience (Clua and Legendre, 2008; Jackson et
al., 2001). This can be enhanced through the
combination of Natural Resource Management and
aquaculture techniques to reduce wild fish catches,
increase the supply of fish to the wild (seeding), and
support food production to PICT (Dey et al., 2016).
Direct manipulation of species, via the alteration of the
composition of ecosystems to increase the proportion
of species that are more tolerant to climate drivers,
such as heat-tolerant species that bleach and die at
higher temperatures or species that recover faster, can
be potentially beneficial to corals in a changing world
(Hughes et al., 2017a). Assisted colonisation and
shading have been proposed but remain untested at
scale (Barros and Field, 2014).
Mangrove and seagrass habitats re-vegetation
facilitate their migration landward as sea level rises,
while the removal of barriers to migration, such as
infrastructure is beneficial to corals due to their
connectedness (Albert et al., 2017; Guannel et al.,
2016). Assisted recovery of corals through measures
that facilitate genetic adaptations and/or physical
interventions, such as coral farming and
transplantation may also be necessary to improve
coral conditions across the Pacific (Hughes et al.,
2017a; van Oppen et al., 2015)}. Improving water
quality to minimise COTS outbreaks, and developing
alternative control measures, could prevent further
coral decline, but would probably only be successful in
localised areas (Nakamura et al., 2016). Such
strategies can, however, only be successful if climatic
conditions are stabilised, as losses due to bleaching
and cyclones will otherwise increase (De'ath et al.,
Direct socio-economic, governance, and
Future human and ecosystem adaptation can only be
achieved when population growth is addressed in
some PICTs (Butler et al., 2014). Reducing poverty or
encouraging shifts in social norms, improving
governance systems, and using modern approaches
to reef conservation that merge traditional knowledge
and practices with scientific and technological
advances can produce positive social and ecological
outcomes (Hughes et al., 2017a; Veitayaki, 2014;
Veitayaki and Holland, in press). Hughes et al. (2017a)
propose that options include the introduction of
technological innovations such as the introduction of
social norms to reduce pollution and harmful practices
on coral reefs, and encourage voluntary compliance
with formal and informal environmental rules. In social
systems, shifts in the composition of society — for
example, through enhanced education and a reduction
in poverty — can also increase resilience to strong
drivers such as climate change (Hughes et al.,
Climate change can be partially offset by collective
action and cooperative measures in local and regional
management interventions that help minimise pollution
and overfishing, control human population size,
consumption and access to markets (McCook et al.,
2010). Cooperation between governments,
organisations and communities will improve human-
reef resilience to climate change (Morrison, 1999).
At the community level, adaptation must build on local
knowledge and be integrated with science, which will
be of great value for environmental management,
particularly when scientific data are lacking (Aswani
and Ruddle, 2013; Leon et al., 2015). In fact, reefs with
traditional systems of resources management in PICTs
are healthier because the harvest of marine resources
is better controlled (Hoffmann, 2002). When
conservation measures are in line with traditional
governance structures and values, local coastal
communities are more likely to engage in and adopt
conservation programs (Walter and Hamilton, 2014).
The interdependences between corals, non-climate
and climate drivers point to the critical importance of
an ecosystem-based approach to fisheries
management throughout the Pacific (Welch 2016). The
use of climate-informed, community-based ecosystem
approaches to fisheries management (CEAFM) (SPC
2010, Heenan et al. 2015) is essential to maintain the
role of key functional groups on reefs. CEAFM
approaches should be based on primary fisheries
management (Cochrane et al. 2011) intended to
protect reef habitats and keep the harvest of reef fish
and invertebrates within sustainable bounds.
Any intervention in PICTs should therefore pay
attention to people's development aspirations,
potential social, economic and environmental benefits,
local dynamics of village governance, social rules and
protocols, and traditional forms of knowledge that can
inform sustainable solutions (Remling and Veitayaki,
2016). It is important for policy makers to understand
the cultural influences that have helped shape current
social norms and decision-making processes, and the
ways in which adaptations to climate change can be
developed and sustained (Nunn, 2009). It is extremely
important that in the future PICT governments should
take on ownership of the climate-change adaptation
process to a greater degree than they do at present
The Pacific offers several examples of local initiatives
that have been successful in improving coral
communities, but it is still unclear how to bring together
top-down global-scale climate change mitigation and
bottom-up local initiatives to maximise the perceived
benefits of improving the condition of coral ecosystems
in the Pacific region. A clear lesson from the literature
is that tenure systems and associated political systems
differ substantially across PICTs (Aswani et al., 2017).
As a result, one cannot simply extrapolate adaptation
measures and strategies similarly across the Pacific.
Local environmental, social and governance contexts
must be considered when rolling out adaptation
programmes in the Pacific.
It is also important that climate change must be
addressed in terms of current pressing challenges to
livelihoods, future risks and how to address these
(Remling and Veitayaki, 2016; Veitayaki and Holland,
in press). Building capacity is essential to improve local
understanding about the complexities involved in
climate change and adaptation, as well as to help
communities better prepare for the future. Such
capacity should be built around practical discussions
on receding shorelines and processes to rehabilitate
coastal habitats and protect local forests, water
catchment areas and food sources. Coastal
communities in PICTs understand well that only a
healthy environment can support their basic needs for
food and clean water in the long term, and have drawn
connections to broader environmental changes such
as climate change (Remling and Veitayaki, 2016;
Veitayaki, in press).
Please cite this document as:
Dutra, L.X.C., Haywood, M.D.E., Singh, S.S.,
Piovano, S., Ferreira, M., Johnson, J.E., Veitayaki, J.,
Kininmonth, S., Morris, C.W. (2018) Impacts of
Climate Change on Corals Relevant to the Pacific
Islands. Pacific Marine Climate Change Report Card:
Science Review 2018, pp 132-158.
The views expressed in this review paper do not
represent the Commonwealth Marine Economies
Programme, individual partner organisations or the
Foreign and Commonwealth Office.
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