Editing IPCC:AR6/WGII/Chapter-15 (section)

===== 15.3.3.1.3 Impacts on marine and coastal ecosystems =====

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Loss of marine and coastal biodiversity and ecosystem services is a key risk in small islands (see KR1 in Figure 15.5). Coral bleaching caused by elevated water temperatures is the most visible and widespread manifestation of a climate change impact on coastal ecosystems in most small islands but is far from being the only one (Sections 3.4.2.1 and [[IPCC:Wg2:Chapter:Chapter-5#5.3|Section 5.3.4]] ; [[#Spalding--2015|Spalding and Brown, 2015]] ; [[#Hoegh-Guldberg--2017|Hoegh-Guldberg et al., 2017]] ; [[#IPCC--2018|IPCC, 2018]] ; [[#Bindoff--2019|Bindoff et al., 2019]] ; [[#Sully--2019|Sully et al., 2019]] ). Severe coral bleaching, together with declines in coral abundance have been documented in many small islands, especially those in the Pacific Ocean and Indian Ocean (e.g., Guam, Fiji, Palau, Vanuatu, Chagos, Comoros, Mauritius, Seychelles, and the Maldives ( ''high confidence'' ) (Box 15.1; [[#Golbuu--2007|Golbuu et al., 2007]] ; [[#Woesik--2012|Woesik et al., 2012]] ; [[#Perry--2017|Perry and Morgan, 2017]] ; [[#Hughes--2018|Hughes et al., 2018]] ). During severe bleaching events, not only do reefs lose a significant amount of live coral cover, but they also experience a decrease in growth potential, and thus reef erosion surpasses reef accretion ( [[#Perry--2017|Perry and Morgan, 2017]] ). Median return time between two severe bleaching events has diminished steadily since 1980 and is now only 6 years (e.g., [[#Hughes--2017b|Hughes et al., 2017b]] ; [[#Hughes--2018|Hughes et al., 2018]] ) and is often associated with warm phase of ENSO events ( ''high confidence'' ) ( [[#Lix--2016|Lix et al., 2016]] ). Modelling of both bleaching and ocean acidification effects under future climate scenarios suggested that some Pacific small islands (e.g., Nauru, Guam, Northern Marianas Islands) will experience conditions that cause severe bleaching on an annual basis before 2040 and that 90% of the world reefs are projected to experience conditions that result in severe bleaching annually by 2055 ( ''medium confidence'' ) ( [[#van%20Hooidonk--2016|van Hooidonk et al., 2016]] ). Models are currently predicting the large-scale loss of coral reefs by mid-century under even low-emission scenarios. Even achieving emission reduction targets consistent with the ambitious goal of 1.5°C of global warming under the Paris Agreement will result in the further loss of 70–90% of reef-building corals compared to today, with 99% of corals being lost under warming of 2°C or more above the pre-industrial period ( ''high confidence'' ) ( [[#Hoegh-Guldberg--2018|Hoegh-Guldberg et al., 2018]] ).

Satellite data and local field studies at 3351 sites in 81 countries including small islands show that not all coral reefs are equally exposed to severe temperature stress events, and even similar coral reefs exposed to similar conditions show local and regional variation and species-specific responses ( [[#Sully--2019|Sully et al., 2019]] ). There is great variability in terms of sensitivity of corals to climate change, as also demonstrated in the Comoros Archipelago ( [[#Cowburn--2018|Cowburn et al., 2018]] ), in the Pacific ( [[#Fox--2019|Fox et al., 2019]] ; [[#Mollica--2019|Mollica et al., 2019]] ; [[#Romero-Torres--2020|Romero-Torres et al., 2020]] ) and globally ( [[#Sully--2019|Sully et al., 2019]] ; [[#McClanahan--2020|McClanahan et al., 2020]] ). It has been hypothesised that low-latitude tropical reefs bleached less than those in higher latitudes because: (a) of the geographical differences in species composition, (b) of the higher genotypic diversity at low latitudes, and (c) some corals were pre-adapted to thermal stress because of consistently warmer temperatures at low latitude prior to thermal stress events ( [[#Sully--2019|Sully et al., 2019]] ). However, latitudinal variation was not reported in other global surveys of coral bleaching occurrence ( [[#Donner--2017|Donner et al., 2017]] ; [[#Hughes--2017a|Hughes et al., 2017a]] ; [[#Hughes--2017b|Hughes et al., 2017b]] ; [[#McClanahan--2019|McClanahan et al., 2019]] ). [[#Ainsworth--2016|Ainsworth et al. (2016)]] and [[#Ateweberhan--2013|Ateweberhan et al. (2013)]] showed that coral bleaching can be mitigated by pre-exposure to elevated temperatures. Regionally, recovery is also highly variable. While some reefs in the Seychelles and Maldives were shown to recover to pre-disturbance levels of coral cover after previous bleaching events (Box 15.1; [[#Pisapia--2016|Pisapia et al., 2016]] ; [[#Koester--2020|Koester et al., 2020]] ), other reefs underwent seemingly permanent regime shifts toward domination by fleshy macro algae ( [[#Graham--2015|Graham et al., 2015]] ), or major declines in carbonate budgets, and thus the capacity of reefs to sustain vertical growth under rising sea levels ( [[#Perry--2017|Perry and Morgan, 2017]] ).

Despite their vital social and ecological value, substantial declines in seagrass communities have been documented in many small islands ( [[IPCC:Wg2:Chapter:Chapter-3#3.4.2.5|Section 3.4.2.5]] ; [[#Arias-Ortiz--2018|Arias-Ortiz et al., 2018]] ; [[#Kendrick--2019|Kendrick et al., 2019]] ; [[#Brodie--2020|Brodie et al., 2020]] ), including Fiji ( [[#Joseph--2019|Joseph et al., 2019]] ), Reunion Island ( [[#Cuvillier--2017|Cuvillier et al., 2017]] ), Bermuda, Cayman Islands, US Virgin Islands ( [[#Waycott--2009|Waycott et al., 2009]] ), Kiribati ( [[#Brodie--2020|Brodie et al., 2020]] ), Federated States of Micronesia, and Palau ( [[#Short--2016|Short et al., 2016]] ), but attribution of such declines to climatic influences remains weak ( ''low confidence'' ). The impact of climate change on seagrasses goes beyond the loss of seagrass but includes acceleration of seagrass decomposition ( [[#Kelaher--2018|Kelaher et al., 2018]] ), palatability ( [[#Jimenez-Ramos--2017|Jimenez-Ramos et al., 2017]] ) and the cumulative effect of warming and eutrophication ( [[#Ontoria--2019|Ontoria et al., 2019]] ). Seagrasses face a multitude of threats including physical disturbance and direct damage caused by rapidly growing human populations, declines in water quality, and coastal erosion ( [[#Short--2016|Short et al., 2016]] ). Experimental studies have shown increased mortality, leaf necrosis, and respiration when seagrasses are exposed to higher-than-normal temperatures ( [[#Hernan--2017|Hernan et al., 2017]] ). As such, seagrass meadows growing near the edge of their thermal tolerance are at risk from rising temperatures ( [[#Pedersen--2016|Pedersen et al., 2016]] ). In the Mediterranean, seagrass meadows are already showing signs of regression, which may have been aggravated by climate change ( ''high confidence'' ). Some studies suggest seagrasses have potential for acclimation and adaptation ( [[#Duarte--2018|Duarte et al., 2018]] ; [[#Ruiz--2018|Ruiz et al., 2018]] ; [[#Beca-Carretero--2020|Beca-Carretero et al., 2020]] ). Chefaoui et al. (2018) attempted to forecast the distribution of two seagrasses in the future, including around the islands of Cyprus, Malta, Sicily and the Balearic Islands. Under the worst-case scenario, ''Posidonia oceanica'' was projected to lose 75% of suitable habitat by 2050. Conversely, it has been suggested that seagrasses could actually benefit from an increase in anthropogenic CO 2 because of increased growth and photosynthesis ( [[#Hopley--2007|Hopley et al., 2007]] ; [[#Waycott--2011|Waycott et al., 2011]] ; [[#Sunday--2016|Sunday et al., 2016]] ; [[#Repolho--2017|Repolho et al., 2017]] ). However, [[#Collier--2017|Collier et al. (2017)]] argued that when faced with increased heat waves, thermal stress will rarely be offset by the benefit of elevated CO 2 and therefore that the widespread belief that seagrasses will be a ‘winner’ under future climate change conditions seems unlikely ''(low confidence'' ).

Since 2011, the Caribbean region has been experiencing unprecedented influxes of the pelagic seaweed ''Sargassum'' . These extraordinary sargassum ‘blooms’ have resulted in mass strandings of sargassum throughout the Lesser Antilles, with significant damage to coastal habitats, mortality of seagrass beds and associated corals ( [[#van%20Tussenbroek--2017|van Tussenbroek et al., 2017]] ), as well as consequences for fisheries and tourism. Whether or not such events are related to long-term climate change remains unclear; however, it has been suggested that the influx may be related to strong Amazon discharge, enhanced West African upwelling, together with rising seawater temperatures in the Atlantic ( ''low confidence'' ) ( [[#Oviatt--2019|Oviatt et al., 2019]] ; [[#Wang--2019|Wang et al., 2019]] ). Since 2011, the Pacific atoll nation of Tuvalu has also been affected by algal blooms, the most recent being a large growth of ''Sargassum'' on the main atoll of Funafuti, and this phenomenon has been related to anthropogenic eutrophication and high seawater temperatures ( [[#De%20Ramon%20N’Yeurt--2014|De Ramon N’Yeurt and Iese, 2014]] ).

Mangroves face serious risks from deforestation and unsustainable coastal development ( [[IPCC:Wg2:Chapter:Chapter-3#3.4.2.5|Section 3.4.2.5]] ; [[#Gattuso--2015|Gattuso et al., 2015]] ). Large-scale die-offs around many small islands suggest that mangroves face increased risks from climate change ( [[#Sippo--2018|Sippo et al., 2018]] ). Mangrove seaward edge retreat has been demonstrated in American Samoa and at Tikina Wai in Fiji, in Bermuda, West Papua, Grand Cayman and attributed to long-term SLR or tectonic subsidence ( [[#Ellison--1993|Ellison, 1993]] ; [[#Ellison--2005|Ellison, 2005]] ; [[#Gilman--2007|Gilman et al., 2007]] ; [[#Ellison--2015|Ellison and Strickland, 2015]] ). Inundation-related mortality of mangroves could, in theory, be mitigated if mangrove substrates can ‘keep up’ with rising sea level by accretion. Pacific Island studies using radionuclides (e.g., 210Pb, 137Cs) have suggested that most mangroves are keeping up with current rates of SLR ( [[#Alongi--2008|Alongi, 2008]] ; [[#MacKenzie--2016|MacKenzie et al., 2016]] ), while surface elevation tables (SETs) suggest otherwise. [[#Lovelock--2015|Lovelock et al. (2015)]] reported that nearly 70% of the mangroves monitored with SETs are not keeping up with current SLR rates. If SLR exceeds 6 mm yr –1 , mangroves may be unable to maintain their elevation relative to sea level, a threshold likely to be surpassed in the next 30 years under high emission scenarios ( [[#Ellison--1993|Ellison, 1993]] ; [[#Saintilan--2020|Saintilan et al., 2020]] ). In these worst-case scenarios, flooding would result in tree, root and rhizome death and an abrupt change in elevation through peat collapse ( [[#Krauss--2010|Krauss et al., 2010]] ; [[#Lang’at--2014|Lang’at et al., 2014]] ), creating a positive feedback loop between SLR and elevation loss. Geomorphology, hydrology, tidal range and suspended sediments are important factors that will determine if mangroves will survive increased rates of SLR ( [[#Lovelock--2015|Lovelock et al., 2015]] ; [[#Sasmito--2015|Sasmito et al., 2015]] ; [[#Rogers--2019|Rogers et al., 2019]] ). TCs can cause extensive damage to mangroves ( [[#Short--2016|Short et al., 2016]] ). While immediate physical damage is often considerable, trees can sometimes recover by re-foliating, re-sprouting or regenerating ( [[#Kauffman--2010|Kauffman and Cole, 2010]] ). Examples of substantive mangrove recovery include the regrowth of trees in the Bay Islands of Honduras following Hurricane Mitch (October 1998) ( [[#Fickert--2018|Fickert, 2018]] ) and in the Nicobar Islands, India, following the December 2004 Indian Ocean Tsunami ( [[#Nehru--2018|Nehru and Balasubramanian, 2018]] ).

Sandy beaches are an important ecosystem in small islands, with high socioeconomic as well as ecosystem services value ( [[#Ellison--2018|Ellison, 2018]] ). Turtles and many seabirds nest just above the high-water mark on sandy beaches or among sand dunes, but TCs, rising seas, storm surges and heavy rainfall as well as inappropriate coastal development can erode beaches ( [[#15.3.1|Section 15.3.1.2]] ) resulting in damage to nests and eggs ( [[#Fuentes--2011|Fuentes et al., 2011]] ). Beach-nesting turtle populations are projected to become threatened around many small islands as a result of future climate change (e.g., Bonaire – Netherlands Antilles ( [[#Fish--2005|Fish et al., 2005]] ), Bioko Island – Equatorial Guinea ( [[#Veelenturf--2020|Veelenturf et al., 2020]] ), Cyprus ( [[#Varela--2019|Varela et al., 2019]] ), Raine Island – Australia ( [[#Pike--2015|Pike et al., 2015]] )), although other populations such as those around the Cape Verde Islands are projected to remain relatively robust ( [[#Abella%20Perez--2016|Abella Perez et al., 2016]] ). Turtles are also threatened by temperature rise around some small islands as warmer temperatures on nesting beaches can lead to an unbalanced sex ratio in the population (e.g., St. Eustatius island, ( [[#Laloë--2016|Laloë et al., 2016]] )).

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