Editing IPCC:AR6/SROCC/Chapter-5 (section)

=== 5.3.2 Coastal Wetlands (Salt Marshes, Seagrass Meadows and Mangrove Forests) ===

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Coastal vegetated wetlands include salt marshes, mangrove forests and subtidal seagrass meadows ecosystems, considered to be the main ‘blue carbon’ habitats (Sections 5.4.1 and 5.5.1.1) (McLeod et al., 2011 <sup>[[#fn:r921|921]]</sup> ). AR5 WGII and SR15 concluded that wetland salinisation is occurring at a large geographic scale ( ''high confidence'' ); that rising water temperatures has led to shifts in plant species distribution ( ''medium confidence'' )(Wong et al., 2014b <sup>[[#fn:r923|923]]</sup> ); and that SLR and storms are causing wetland erosion and habitat loss, enhanced by human disturbances ( ''high confidence'' ) (Section 4.3.3.5.1) (Wong et al., 2014b <sup>[[#fn:r922|922]]</sup> ). This section assesses new evidence since AR5 and SR15 of observed climate impacts and future risks of these vegetated wetlands in terms of their role in supporting biodiversity and key ecosystem functions. The recent literature confirms and strengthens the SR15 conclusions (Section 5.3.7 and Figure 5.16).

Nearly 50% of the pre-industrial, natural extent of global coastal wetlands have been lost since the 19th century (Li et al., 2018a <sup>[[#fn:r924|924]]</sup> ). Such a reduction in wetlands is primarily caused by non-climatic drivers such as alteration of drainage, agriculture development, coastal settlement, hydrological alterations and reductions in sediment supply (Adam, 2002 <sup>[[#fn:r925|925]]</sup> ; Wang et al., 2014 <sup>[[#fn:r926|926]]</sup> ; Kroeger et al., 2017 <sup>[[#fn:r927|927]]</sup> ; Thomas et al., 2017 <sup>[[#fn:r928|928]]</sup> ; Li et al., 2018a <sup>[[#fn:r929|929]]</sup> ). However, large-scale mortality events of mangroves from ‘natural causes’ has also occurred globally since the 1960s; ~70% of this loss has resulted from low frequency, high intensity weather events, such as tropical cyclones (45%) and climatic extremes such as droughts, SLR variations and heat waves (Sippo et al., 2018 <sup>[[#fn:r930|930]]</sup> ) ( ''high confidence'' ). In Australia, the mangrove loss due to heat waves accounted for 22% of global mangrove forests (Sippo et al., 2018 <sup>[[#fn:r931|931]]</sup> ), with negative impacts on ecosystem biodiversity and the provisioning of services (Carugati et al., 2018 <sup>[[#fn:r932|932]]</sup> ; Saintilan et al., 2018 <sup>[[#fn:r933|933]]</sup> ) (Section 5.4). In coastal areas with sufficient sediment supply across the Indo-Pacific region, inland expansion of mangroves is occurring as a result of vertical accretion and root growth, allowing them to keep pace with current SLR (Lovelock et al., 2015 <sup>[[#fn:r934|934]]</sup> ). In seagrass meadows, temperature is the main limiting range factor, and over the past decades there have been several global die-off events (Hoegh-Guldberg et al., 2018 <sup>[[#fn:r935|935]]</sup> ). The vulnerability of seagrasses to warming varies locally depending on soil accretion and herbivory (El-Hacen et al., 2018 <sup>[[#fn:r936|936]]</sup> ; Marbà et al., 2018 <sup>[[#fn:r937|937]]</sup> ; Vergés et al., 2018 <sup>[[#fn:r938|938]]</sup> ) and on the population assemblages (e.g., expansion at high latitudes) (Beca-Carretero et al., 2018 <sup>[[#fn:r939|939]]</sup> ; Duarte et al., 2018 <sup>[[#fn:r940|940]]</sup> ). The compounding effects of heat waves, hypersaline conditions and increased turbidity and nutrient levels associated with floods have been shown to cause negative changes in the composition and biomass of co-occurring seagrass species (Nowicki et al., 2017 <sup>[[#fn:r941|941]]</sup> ; Arias-Ortiz et al., 2018 <sup>[[#fn:r942|942]]</sup> ; Lin et al., 2018 <sup>[[#fn:r943|943]]</sup> ) ( ''high confidence'' ). For example, in Shark Bay, Western Australia, a marine heat wave in austral summer 2010/2011 caused widespread losses (36% of area) of seagrass meadows, with negative implications for carbon storage (Arias-Ortiz et al., 2018 <sup>[[#fn:r944|944]]</sup> ). The poleward expansion of tropical mangroves into subtropical salt marshes as a result of increase in temperature has been also observed over the past half century on five continents (Saintilan et al., 2014 <sup>[[#fn:r945|945]]</sup> ; Saintilan et al., 2018 <sup>[[#fn:r946|946]]</sup> ) ( ''high confidence'' ); for example, in the Texas Gulf Coast (Armitage et al., 2015 <sup>[[#fn:r947|947]]</sup> ). The loss of open areas with herbaceous plants (salt marshes) reduces food and habitat availability for resident and migratory animals (Kelleway et al., 2017a <sup>[[#fn:r948|948]]</sup> ; Lin et al., 2018 <sup>[[#fn:r949|949]]</sup> ) (Section 5.4.1.2).

The ability of salt marshes to increase their elevation and withstand erosion under SLR depends on the development of new soil by the external supply of mineral sediments and organic accretion by local biota (Section 5.4.1, Figure 5.19) (Bouma et al., 2016 <sup>[[#fn:r950|950]]</sup> ). In some places, critical organic accretion rates are declining due to reduced plant productivity from stress by more frequent inundation, and increased plant and microbial respiration rates as a result of warming; consequently, the elevation of marshes from soil accretion is slower than the rate of rising sea level, resulting in reduction of salt marsh area (Carey et al., 2017 <sup>[[#fn:r951|951]]</sup> ; Watson et al., 2017b <sup>[[#fn:r952|952]]</sup> ). Vegetation loss rates were significantly negatively correlated with marsh elevation, suggesting inundation due to SLR since 1970 as the main driver, enhanced by storms and increased tidal range in back barrier marshes (Watson et al., 2017b <sup>[[#fn:r953|953]]</sup> ). Plant species that are more sensitive to higher temperatures and increases in saltwater intrusion were found to be less abundant and in some cases replaced by salinity-tolerant species (Janousek et al., 2017 <sup>[[#fn:r954|954]]</sup> ; Piovan et al., 2019 <sup>[[#fn:r955|955]]</sup> ). Plant community restructuring has resulted in biodiversity loss (Pratolongo et al., 2013 <sup>[[#fn:r956|956]]</sup> ; Raposa et al., 2017 <sup>[[#fn:r957|957]]</sup> ) and reduced above- and below-ground productivity (McLeod et al., 2011 <sup>[[#fn:r958|958]]</sup> ; Watson et al., 2017b <sup>[[#fn:r959|959]]</sup> ). As a result of tidal flooding, salt marsh soils do not dry out and high levels of carbon can accumulate under anaerobic conditions. This is coupled with generally low rates of methane emission which is strongly limited in saline marshes (Poffenbarger et al., 2011 <sup>[[#fn:r960|960]]</sup> ; Martin and Moseman-Valtierra, 2015 <sup>[[#fn:r961|961]]</sup> ; Kroeger et al., 2017 <sup>[[#fn:r962|962]]</sup> ; Tong et al., 2018 <sup>[[#fn:r963|963]]</sup> ) ( ''high confidence'' ).

Non-climatic human pressures on wetland ecosystems, including overfishing (Crotty et al., 2017 <sup>[[#fn:r964|964]]</sup> ), eutrophication (Legault II et al., 2018 <sup>[[#fn:r965|965]]</sup> ), and invasive species (Zhang et al., 2016 <sup>[[#fn:r966|966]]</sup> ), interact with climate change drivers and affect wetlands composition and structure, with the impacts varying between regions and species (Tomas et al., 2015 <sup>[[#fn:r967|967]]</sup> ; O’Brien et al., 2017; Pagès et al., 2017 <sup>[[#fn:r968|968]]</sup> ; York et al., 2017 <sup>[[#fn:r969|969]]</sup> ). The intensity of herbivory on seagrasses is expected to increase with global warming, particularly in temperate areas, because of the migration of tropical herbivores into temperate seagrass meadows (Hyndes et al., 2016 <sup>[[#fn:r970|970]]</sup> ; Vergés et al., 2018 <sup>[[#fn:r971|971]]</sup> ) ( ''medium confidence'' , Section 5.2.3.1.1). Warming also reduces the fitness of seedlings by increasing necrosis and susceptibility to consumers and pathogenic pressure while reducing establishment potential and nutritional (Olsen et al., 2016b <sup>[[#fn:r972|972]]</sup> ; Hernán et al., 2017 <sup>[[#fn:r973|973]]</sup> ). Because herbivores play a key role in modulating the biomass of plant communities, their more intense activity affects the provision of services in these ecosystems (Scott et al., 2018 <sup>[[#fn:r974|974]]</sup> ) (Section 5.4).

Globally, between 20−90% of existing coastal wetland area is projected to be lost by 2100 (Blankespoor et al., 2014 <sup>[[#fn:r975|975]]</sup> ; Crosby et al., 2016 <sup>[[#fn:r976|976]]</sup> ; Spencer et al., 2016 <sup>[[#fn:r977|977]]</sup> ), depending on different SLR projections under future emission scenarios. These projected changes vary regionally and between different types of wetlands. Gaining area may be possible, at least locally, if vertical sediment accretion occurs together with lateral re-accommodation (Brown et al., 2018b <sup>[[#fn:r978|978]]</sup> ; Schuerch et al., 2018 <sup>[[#fn:r979|979]]</sup> ) (Section 4.3.3.5.1). Local losses may also be higher; for example, in New England, where regional rates of SLR have been as much as 50% greater than the global average (from 1−5.83 mm yr -1 ; 1979-2015) (Watson et al., 2017a <sup>[[#fn:r980|980]]</sup> ) and where projections suggest that 40–95% of salt marshes will be submerged by the end of this century (Valiela et al., 2018 <sup>[[#fn:r981|981]]</sup> ). In some species of seagrasses, enhanced temperature-driven flowering (Ruiz-Frau et al., 2017 <sup>[[#fn:r982|982]]</sup> ) and greater biomass production in response to elevated CO 2 (Campbell and Fourqurean, 2018 <sup>[[#fn:r983|983]]</sup> ) may increase resilience to warming. Nevertheless, severe habitat loss (70%) of endemic species such as ''Posidonia oceanica'' is projected by 2050 with the potential for functional extinction by 2100 under RCP8.5 climate scenario. For ''Cymodosea nodosa'' , the species with the highest thermal optima (Savva et al., 2018 <sup>[[#fn:r984|984]]</sup> ), warming is expected to lead to significant reduction of meadows (46% under RCP8.5) in the Mediterranean, although potentially compensated in part by future expansion into the Atlantic (Chefaoui et al., 2018 <sup>[[#fn:r985|985]]</sup> ).

The mangrove habitats of small islands, with lack of rivers, steep topography, sediment-starved areas, groundwater extraction and coastal development, are particularly vulnerable to SLR. Although mangrove ecosystems may survive the increased storm intensity and sea levels projected until 2100 under RCP2.6 (Ward et al., 2016 <sup>[[#fn:r986|986]]</sup> ), for RCP8.5 they are only resilient up to 2050 conditions (Sasmito et al., 2016 <sup>[[#fn:r987|987]]</sup> ). Negative climate impacts will be exacerbated in cases where anthropogenic barriers cause further ‘coastal squeeze’ that prevents inland movement of plants and limits relocation of sediment ( ''medium confidence'' ) (Enwright et al., 2016 <sup>[[#fn:r988|988]]</sup> ; Borchert et al., 2018 <sup>[[#fn:r989|989]]</sup> ).

In conclusion, substantial evidence supports with ''high confidence'' that warming and salinisation of wetlands caused by SLR are causing shifts in the distribution of plant species inland and poleward, such as mangrove encroachment into subtropical salt marshes ( ''high confidence'' ) or seagrass meadows contraction at low latitudes ( ''high confidence'' ). Plants with low tolerance to flooding and extreme temperatures are particularly vulnerable and may be locally extirpated ( ''medium confidence'' ). The flooded area of salt marshes can become a mudflat or be colonised by more tolerant, invasive species, whose expansion is favoured by combined effects of warming, rising CO 2 and nutrient enrichment ( ''medium confidence'' ). The loss of vegetated coastal ecosystems causes a reduction in carbon storage with positive feedbacks to the climate system ( ''high confidence'' ) (Section 5.4.1.2). SLR and warming are expected to continue to reduce the area of coastal wetlands, with a projected global loss of 20–90% by the end of the century depending on emission scenarios. High risk of total local loss is projected under the RCP8.5 emission scenario by 2100 ( ''medium confidence'' ), especially if landward migration and sediment supply is constrained by human modification of shorelines and river flows ( ''medium confidence'' ).

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