Editing IPCC:AR6/WGII/Chapter-3 (section)

=== 3.5.5 Supporting and Regulating Services ===

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Ocean and coastal regulating services are detailed in Table 3.25. The economic value of all regulating ecosystem services in 2015 was estimated at 29.1 trillion USD, with water- and climate-regulating services contributing the most ( [[#Balasubramanian--2019|Balasubramanian, 2019]] ).

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==== 3.5.5.1 Habitat Creation and Maintenance, and Larval Dispersal ====

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Climate impacts have already altered ocean and coastal habitats ( [[#3.4.2|Section 3.4.2]] ; Table 3.26; [[#Gissi--2021|Gissi et al., 2021]] ) in ways that have led to species range shifts, biodiversity changes, phenology changes and regime shifts ( [[#3.4.3|Section 3.4.3]] ) from the surface ocean to the seafloor ( ''very high confidence'' ) (see Box 3.3; Figure 3.22). Continued ocean and coastal habitat impacts are projected, and their severities will depend on emissions scenario and co-occurring drivers ( [[#3.4.3|Section 3.4.3]] ; [[#Qiu--2019|Qiu et al., 2019]] ) or extremes (e.g., [[#Babcock--2019|Babcock et al., 2019]] ). Warming and physical circulation are projected to change larval dispersal, a habitat-related service ( [[#Bashevkin--2020|Bashevkin et al., 2020]] ), but identifying probable outcomes remains challenging owing to the high variability among species, locations and recruitment ( [[#Schilling--2020|Schilling et al., 2020]] ; [[#King--2021|King et al., 2021]] ; [[#Le%20Corre--2021|Le Corre et al., 2021]] ; [[#Raventos--2021|Raventos et al., 2021]] ). Climate risks to habitat can be decreased by reducing non-climate drivers, preserving ecosystems or restoring habitat (Sections 3.6.2, 3.6.3.2). Risk to larval dispersal cannot be meaningfully addressed at scale by human-implemented adaptations; instead, declines in this service will pressure natural systems to adapt via physiological plasticity or evolution ( [[#3.3|Section 3.3]] ; [[#Bashevkin--2020|Bashevkin et al., 2020]] ).

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[[File:4b47d3faf512f2e6f6e91c7717101878 IPCC_AR6_WGII_Figure_3_022.png]]

'''Figure 3.22 |''' '''Observed global influence of climate-induced drivers on ecosystem services.''' Symbols show whether the observed impact of the climate-induced drivers on a group of ecosystem services is positive (beneficial), negative (detrimental) or mixed (usually resulting from location, the presence of interacting drivers or changing effects over time). The ‘observed impact’ indicates the total effect of all climate-induced drivers on a specific ecosystem service, using expert judgement based on summary statements throughout [[#3.5|Section 3.5]] . Tick marks represent the presence of co-occurring drivers non-climate drivers that affect the service. No assessment indicates that not enough evidence is available to assess the direction of impact.

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==== 3.5.5.2 Climate Regulation and Air Quality ====

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Climate regulation by the ocean depends on physical and biogeochemical processes (Sections 3.2–3.4) that create, move, and store heat, water vapour and other climate-active compounds including CO 2 , methane and dimethyl sulphide (WGI AR6 Chapter 6; [[#Szopa--2021|Szopa et al., 2021]] ). Over the 21st century, ocean heat and CO 2 uptake will continue (WGI AR6 SPMB4.1, B5.1; [[#IPCC--2021b|IPCC, 2021b]] ) and sea ice loss from warming will allow some additional CO 2 uptake ( [[#Armstrong--2019|Armstrong et al., 2019]] ), but the ocean will take up a smaller fraction of CO 2 emissions as atmospheric CO 2 concentrations rise ( ''high confidence'' ) (Table 3.26; WGI AR6 SPM B4.1; [[#IPCC--2021b|IPCC, 2021b]] ).

There is ''very limited evidence'' on climate-driven air-quality changes in the coastal zone. Increased humidity decreases the lifetime of ozone and increases particulate matter and indoor mould levels (USGCRP, 2016), potentially affecting near-shore air quality. However, coastal-zone air pollution can enhance coastal-climate impacts by increasing the risk of acid rain, which worsens ocean acidification (nitrogen oxides, sulphur oxides and mercury; [[#Doney--2010|Doney, 2010]] ; [[#Northcott--2019|Northcott et al., 2019]] ).

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==== 3.5.5.3 Provision of Freshwater, Maintenance of Water Quality and Regulation of Pathogens ====

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The salinities of many estuaries, deltas, coastal freshwater aquifers and soils around the world are increasing, and this decrease in water quality is endangering human health and agricultural yields ( ''very high confidence'' ) ( [[#3.4.2.4|Section 3.4.2.4]] ; Table 3.26; [[#Bindoff--2019a|Bindoff et al., 2019a]] ; [[#Bouderbala--2019|Bouderbala, 2019]] ; [[#Rahman--2019|Rahman et al., 2019]] ; [[#Naser--2020|Naser et al., 2020]] ; [[#Rakib--2020|Rakib et al., 2020]] ; [[#Mastrocicco--2021|Mastrocicco and Colombani, 2021]] ). Coastal salinisation is attributed to regionally varying combinations of climate-induced drivers, like SLR and storm-related flooding by seawater, and non-climate drivers, like water withdrawal and land-use changes ( ''very high confidence'' ) ( [[#Islam--2019|Islam et al., 2019]] ; [[#Rahman--2019|Rahman et al., 2019]] ; [[#Paldor--2021|Paldor and Michael, 2021]] ). Monitoring-related adaptations ( [[#3.6.2.2|Section 3.6.2.2.2]] ), including advances in modelling and monitoring, are providing decision-relevant, regional-scale information ( [[#Colombani--2016|Colombani et al., 2016]] ; [[#Mukhopadhyay--2019|Mukhopadhyay et al., 2019]] ; [[#Slama--2020|Slama et al., 2020]] ; [[#Corwin--2021|Corwin, 2021]] ). For example, new projections indicate which drinking-water intake stations on China’s Pearl River Estuary will be unable to meet demands by 2100 due to SLR and drought ( [[#Wang--2021|Wang and Hong, 2021]] ), while others show that SLR effects on seawater intrusion into the coastal aquifer in Kerala, India, under both RCP4.5 and RCP8.5 scenarios are negligible ( [[#Sithara--2020|Sithara et al., 2020]] ). Salinisation-associated changes may disproportionately burden women responsible for securing drinking water and fuel, such as in the Indian Sundarbans ( [[#Mukhopadhyay--2019|Mukhopadhyay et al., 2019]] ). Salinisation will continue to endanger coastal water and soil quality in the future ( ''high confidence'' ) ( [[#Islam--2019|Islam et al., 2019]] ; [[#Paldor--2021|Paldor and Michael, 2021]] ), but the evidence assessed above shows that subsequent impacts to human health and agriculture will depend heavily on regional variations in environment and human behaviour ( ''medium confidence'' ).

Together, climate-induced and non-climate drivers can mobilise toxins and contaminants in ways that harm human and marine species health ( ''very high confidence'' ) (see Box 3.2), and climate change is altering these relationships ( ''high confidence'' ) (Table 3.26; [[#Bindoff--2019a|Bindoff et al., 2019a]] ). Under warming or ocean acidification, marine molluscs exposed to pharmaceuticals via wastewater experience more detrimental biological consequences or greater bioaccumulation ( ''limited evidence, high agreement'' ) ( [[#Costa--2020a|Costa et al., 2020a]] ; [[#Costa--2020b|Costa et al., 2020b]] ; [[#Dionísio--2020|Dionísio et al., 2020]] ; [[#Freitas--2020|Freitas et al., 2020]] ; [[#Kibria--2021|Kibria et al., 2021]] ). Physical circulation, temperature and biogeochemical characteristics ( [[#Bowman--2020|Bowman et al., 2020]] ; [[#Liu--2020a|Liu et al., 2020a]] ; [[#Liu--2020b|Liu et al., 2020b]] ; [[#Sun--2020|Sun et al., 2020]] ; [[#Zhang--2020b|Zhang et al., 2020b]] ) control the ubiquitous oceanic distribution of methylmercury, and ocean acidification- and warming-driven changes in planktonic speciation and interactions can promote additional food-web bioaccumulation of methylmercury ( [[#Tada--2020|Tada and Marumoto, 2020]] ; [[#Wu--2020b|Wu et al., 2020b]] ; [[#Zhang--2020b|Zhang et al., 2020b]] ; [[#Zhang--2021a|Zhang et al., 2021a]] ). Interactions among drivers also matter: temperature plus overfishing increased tissue methylmercury concentrations in Atlantic bluefin tuna from the 1970s to the 2000s more than the decreases in the late 1990s and 2000s from lower environmental mercury levels ( [[#Schartup--2019|Schartup et al., 2019]] ). This appears true for persistent organic pollutants as well, but their bioaccumulation is related more to temperature effects on animal behaviour than on pollutant dynamics ( [[#Houde--2019|Houde et al., 2019]] ; [[#Wagner--2019|Wagner et al., 2019]] ; [[#Kalia--2021|Kalia et al., 2021]] ). By 2100 under RCP8.5, productivity changes and community structure shifts are expected to increase methylmercury concentrations in polar oceans and high-latitude phytoplankton and decrease it in low latitudes ( [[#Zhang--2021a|Zhang et al., 2021a]] ). The estimated average global cost of mercury-related health effects by 2050, mainly from seafood consumption during 2010–2050, will be 19 trillion USD (2020), assuming a 3% discount rate, if methylmercury emissions are not reduced ( [[#Zhang--2021b|Zhang et al., 2021b]] ).

Since previous assessments, evidence has increased that climate impacts, such as warming, extreme weather and SLR, are increasing the geographic spread and risk of marine-borne human pathogen outbreaks, including ''Vibrio'' spp. ( ''very high confidence'' ) (Table 3.26; [[#Bindoff--2019a|Bindoff et al., 2019a]] ; [[#Logar-Henderson--2019|Logar-Henderson et al., 2019]] ; [[#Froelich--2020|Froelich and Daines, 2020]] ; [[#Montánchez--2020|Montánchez and Kaberdin, 2020]] ; [[#Semenza--2020|Semenza, 2020]] ; [[#Ferchichi--2021|Ferchichi et al., 2021]] ). Climate change affects at least 30 human pathogens with aquatic-system infection routes (e.g., ingestion of contaminated water or seafood, or contact with wounds; Table 3.SM.2; Cross-Chapter Box ILLNESS in Chapter 2; [[#Nichols--2018|Nichols et al., 2018]] ). Conditions favourable for ''Vibrio cholerae'' are increasing globally, which raises the risk to humans (Cross-Chapter Box ILLNESS in Chapter 2). Increased storm-related flooding and SLR further increase human encounters with ''Vibrio'' spp. ( [[#Froelich--2020|Froelich and Daines, 2020]] ). Aquatic diseases, particularly ''Vibrio'' spp., have caused large economic losses in aquaculture by decreasing the quality or survival of cultured species ( [[#Lafferty--2015|Lafferty et al., 2015]] ; [[#Novriadi--2016|Novriadi, 2016]] ). Temperature-based model projections show that all Canadian shellfish beds will experience conditions that promote high risk of ''Vibrio'' spp. growth by 2100 for both RCP4.5 and RCP8.5 scenarios ( [[#Ferchichi--2021|Ferchichi et al., 2021]] ). Climate-induced drivers may increase ''Vibrio'' spp. loads in seafood species: laboratory-simulated heatwaves increase ''Vibrio'' spp. abundance in Pacific oyster ( ''Crassostrea gigas'' ) ( [[#Green--2019|Green et al., 2019]] ) and simulated ocean acidification increases hard clam ( ''Mercenaria mercenaria'' ) susceptibility to ''Vibrio'' spp. infection ( [[#Schwaner--2020|Schwaner et al., 2020]] ). Projected increases in temperature, extreme and variable rainfall conditions, coastal flooding and SLR ( [[#3.2|Section 3.2]] ; Cross-Chapter Box SLR in Chapter 3) strongly increase the risk of frequent and severe aquatic human pathogen outbreaks in ocean and coastal areas that will continue to harm human health and cause economic losses ( ''high confidence'' ) (Cross-Chapter Box ILLNESS in Chapter 2; [[#Froelich--2020|Froelich and Daines, 2020]] ; [[#Semenza--2020|Semenza, 2020]] ; [[#Ferchichi--2021|Ferchichi et al., 2021]] ). [[#3.6.3.1.5|Section 3.6.3.1.5]] assesses human adaptations to increasing risk of marine-borne pathogens.

Climate-driven changes in temperature, salinity (from ice melt and precipitation changes), deoxygenation and ocean acidification can alter dynamics of infectious diseases that target ocean and coastal species by increasing hosts’ susceptibility or pathogens’ abundance or virulence ( ''high confidence'' ) ( [[#Burge--2020|Burge and Hershberger, 2020]] ; [[#Byers--2021|Byers, 2021]] ). Coral and urchin diseases have increased over time driven by warming-related declines in organism recovery and survival or immunity ( ''medium confidence'' ) ( [[#Cohen--2018|Cohen et al., 2018]] ; [[#Tracy--2019|Tracy et al., 2019]] ). Seagrass and sea star wasting disease outbreaks have occurred under combinations of ocean warming or MHWs and non-climate drivers (e.g., eutrophication, bottom trawling), but attribution of these outbreaks to specific drivers is still not resolved ( [[#Harvell--2019|Harvell et al., 2019]] ; [[#Jakobsson-Thor--2020|Jakobsson-Thor et al., 2020]] ; [[#Krause-Jensen--2021|Krause-Jensen et al., 2021]] ). Disease outbreaks threaten marine biodiversity, species that create habitat or dampen wave action, and keystone species ( [[#Harvell--2020|Harvell and Lamb, 2020]] ). Attributing observed changes in marine disease patterns to climate remains extremely difficult owing to interacting climate and non-climate drivers ( [[#Burge--2020|Burge and Hershberger, 2020]] ) and lack of baseline data ( [[#Tracy--2019|Tracy et al., 2019]] ). Projected increases in the frequency, duration and intensity of warming events would reduce survival and recovery of some species from hot events, reduce immunity of other species to pathogens, extend poleward ranges of some pathogens and increase infection risk when host species congregate in scarce habitat ( [[#Cohen--2018|Cohen et al., 2018]] ). Pathogens that target ocean and coastal organisms may themselves be sensitive to future climate conditions or subsequent ecosystem changes, which challenges development of projections ( [[#Cohen--2018|Cohen et al., 2018]] ; [[#Burge--2020|Burge and Hershberger, 2020]] ).

New examples have illustrated how toxic HABs interfere with regulating, provisioning ( [[#3.5.3|Section 3.5.3]] ) and cultural ecosystem services ( [[#3.5.6|Section 3.5.6]] ) in interconnected ways ( ''limited evidence, high agreement'' ). A massive toxic ''Pseudo-nitzschia'' spp. bloom in 2013–2016 along the USA West Coast triggered Dungeness crab, rock crab and razor clam fishery closures to protect human consumers (Sections 3.6.2, 3.6.3.1.5; [[#McCabe--2016|McCabe et al., 2016]] ), and this disproportionately harmed fishers, especially small-vessel owners, and fishing-support service industries, primarily through lost revenue ( [[#Ritzman--2018|Ritzman et al., 2018]] ; [[#Moore--2019|Moore et al., 2019]] ; [[#Trainer--2019|Trainer et al., 2019]] ; [[#Jardine--2020|Jardine et al., 2020]] ; [[#Moore--2020a|Moore et al., 2020a]] ). Toxic ''Alexandrium'' spp. blooms promoted by climate-driven coastal extremes (e.g., MHWs, stratification, runoff) in Tasmania, Australia, in 2012 and Chile in 2016 caused fish kills, shellfish product recalls, substantial economic losses, and human sickness and death ( [[#Trainer--2019|Trainer et al., 2019]] ). The Chile event caused an estimated loss of 800 million USD in the farmed salmon industry ( [[#Díaz--2019|Díaz et al., 2019]] ) and resulted in a series of large, long-lasting regional protests calling for national aid ( [[#Delgado--2019|Delgado et al., 2019]] ). New evidence, however, suggests that the perceived global increase in harmful algal blooms results from better monitoring and more detrimental bloom impacts, rather than a climate-linked mechanism ( [[#Hallegraeff--2021|Hallegraeff et al., 2021]] ).

Natural and engineered systems have long been used effectively to manage precipitation and wastewater safely (see Box 4.5), and maintaining and enhancing them is a key nature-based adaptation strategy for coastal communities ( [[#3.6.2.3|Section 3.6.2.3]] ; Cross-Chapter Paper 2). Estimated values of water purification and stormwater management provided by coastal ecosystems are in the hundreds to thousands of USD per hectare [e.g., 272 Euro per 0.01 km 2 yr –1 from the Mediterranean’s sandy coastline ( [[#Hérivaux--2018|Hérivaux et al., 2018]] ); 1100–2800 USD per 0.01 km 2 yr –1 from the state of Maryland, USA ( [[#Campbell--2020b|Campbell et al., 2020b]] ); 600 USD per 0.01 km 2 yr –1 in Zhuzhou City, China ( [[#Zhan--2020|Zhan et al., 2020]] )]. Both wild and cultured organisms also provide filtration services. Seagrasses’ ability to purify water is well recognised by coastal residents and ocean resource users in tropical and temperate locations ( [[#Ambo-Rappe--2019|Ambo-Rappe et al., 2019]] ; [[#Quevedo--2020|Quevedo et al., 2020]] ; [[#Heckwolf--2021|Heckwolf et al., 2021]] ; [[#McKenzie--2021a|McKenzie et al., 2021a]] ). Globally, aquacultured shellfish remove an estimated 49,000 tonnes of nitrogen and 6000 tonnes of phosphorus from coastal waters, worth a potential 1.20 billion USD, and they may help improve existing engineered wastewater treatment systems ( [[#van%20der%20Schatte%20Olivier--2020|van der Schatte Olivier et al., 2020]] ). Climate change, especially episodic extreme rains and RSLR ( [[#Romero-Lankao--2014|Romero-Lankao et al., 2014]] ), is challenging management and design of wastewater and stormwater systems ( ''high confidence'' ) ( [[#Flood--2011|Flood and Cahoon, 2011]] ; [[#Trtanj--2016|Trtanj et al., 2016]] ; [[#Hummel--2018|Hummel et al., 2018]] ; [[#Kirshen--2018|Kirshen et al., 2018]] ; [[#Nazarnia--2020|Nazarnia et al., 2020]] ; [[#Reznik--2020|Reznik et al., 2020]] ; [[#McKenzie--2021b|McKenzie et al., 2021b]] ) and the integrity of coastal landfills ( [[#Beaven--2020|Beaven et al., 2020]] ). Without substantial adaptation that addresses projected wastewater management challenges and community needs ( [[IPCC:Wg2:Chapter:Chapter-4#4.2.6|Section 4.2.6.1]] ; [[#Kirshen--2018|Kirshen et al., 2018]] ; [[#Kirchhoff--2019|Kirchhoff and Watson, 2019]] ; [[#Kool--2020|Kool et al., 2020]] ; [[#Nazarnia--2020|Nazarnia et al., 2020]] ; [[#Hughes--2021|Hughes et al., 2021]] ), coastal water quality in many areas will decrease because of more frequent or severe releases of untreated wastes ( ''high confidence'' ) ( [[#Flood--2011|Flood and Cahoon, 2011]] ; [[#Hummel--2018|Hummel et al., 2018]] ; [[#Hughes--2021|Hughes et al., 2021]] ; [[#McKenzie--2021b|McKenzie et al., 2021b]] ), and this will have harmful consequences for human and coastal ecosystem health ( ''high confidence'' ) ( [[IPCC:Wg2:Chapter:Chapter-4#4.2.6|Section 4.2.6.1]] ; Cross-Chapter Box ILLNESS in Chapter 2; [[#Bindoff--2019a|Bindoff et al., 2019a]] ).

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==== 3.5.5.4 Regulation of Physical Hazards ====

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Coastal ecosystems physically protect people and property from storms and flooding, and climate change threatens this protection function (Figure 3.22; Table 3.26). Increasingly detailed models show how warm-water coral reefs ( [[#Reguero--2019|Reguero et al., 2019]] ; [[#Storlazzi--2019|Storlazzi et al., 2019]] ; [[#Reguero--2021|Reguero et al., 2021]] ) mangroves ( [[#Blankespoor--2017|Blankespoor et al., 2017]] ; [[#Menéndez--2020|Menéndez et al., 2020]] ; [[#Trégarot--2021|Trégarot et al., 2021]] ) and wetlands ( [[#Sun--2020|Sun and Carson, 2020]] ) prevent billions of US dollars of direct and indirect damage to private and public property and shield millions of people from flooding each year. Protection by mangroves provides more economic benefits in higher-income nations and shields more people in lower-income nations ( [[#Menéndez--2020|Menéndez et al., 2020]] ). Seagrasses ( [[#James--2020|James et al., 2020]] ; [[#James--2021|James et al., 2021]] ), kelp ( [[#Morris--2020b|Morris et al., 2020b]] ; [[#Zhu--2020|Zhu, 2020]] ), suspended shellfish aquaculture ( [[#Gentry--2020|Gentry et al., 2020]] ; [[#Zhu--2020a|Zhu et al., 2020a]] ), oyster reefs ( [[#Chowdhury--2019|Chowdhury et al., 2019]] ), coastal wetlands ( [[#Möller--2019|Möller, 2019]] ; [[#Keimer--2021|Keimer et al., 2021]] ) and sandy coastlines ( [[#3.4.2.6|Section 3.4.2.6]] ) [[#Hérivaux--2018|Hérivaux et al., 2018]] ) also measurably decrease wave energy. Non-climate drivers [e.g., invasive species ( [[#James--2020|James et al., 2020]] ), sediment-supply changes ( [[#Ganju--2019|Ganju, 2019]] ; [[#Ladd--2019|Ladd et al., 2019]] ; [[#Ilia--2020|Ilia, 2020]] ), erosion and storm damage ( [[#Mehvar--2019|Mehvar et al., 2019]] ; [[#Bacopoulos--2021|Bacopoulos and Clark, 2021]] )], acting together with climate-induced drivers and associated impacts [e.g., SLR (Cross-Chapter Box SLR in Chapter 3), changes in plant biodiversity ( [[#3.5.2|Section 3.5.2]] ; [[#Lee%20Smee--2019|Lee Smee, 2019]] ; [[#Silliman--2019|Silliman et al., 2019]] ; [[#Schoutens--2020|Schoutens et al., 2020]] ), MHWs ( [[#3.4.3|Section 3.4.3.7]] ) and acidification ( [[#3.4.2.1|Section 3.4.2.1]] )], compromise physical protection by coastal ecosystems ( ''very high confidence'' ). (See Cross-Chapter Box SLR in [https://www.ipcc.ch/report/ar6/wg2/chapter/chapter-3 Chapter 3] and Sections 3.6.3.1 and 3.6.3.2.2 for assessment of adaptations that address this ecosystem service.)

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==== 3.5.5.5 Regulation of Carbon Cycling in Ocean and Coastal Ecosystems ====

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Current and future total carbon storage and cycling in the ocean are governed by past and future CO 2 emissions trajectories (Table 3.26), but regional ocean and coastal carbon stocks and cycling vary over time and space due to processes being altered by climate, including ocean circulation, sea ice cover, coastal upwelling and thermal stratification ( [[#3.2.2.3|Section 3.2.2.3]] ); ocean primary production and export (Sections 3.2.3, 3.4.4); and marine ecosystem biodiversity ( ''high confidence'' ) ( [[#3.5.2|Section 3.5.2]] ; Figure 3.22). Quantifying regional carbon fluxes and stocks is still challenging and relies on indirect measures (e.g., [[#Fennel--2019|Fennel et al., 2019]] ; [[#Clay--2020|Clay et al., 2020]] ), especially in coastal ecosystems where drivers interact. Carbon cycling and storage co-occurs with other regulating services such as habitat provision, water-quality maintenance and coastal protection ( [[#Ouyang--2018|Ouyang et al., 2018]] ), particularly in vegetated coastal ecosystems (see Box 3.4). Adaptations to support regional carbon cycling and storage generally focus on area-based management and conservation ( [[#3.6.3.2|Section 3.6.3.2]] ), but interventions to enhance ocean carbon storage are being explored for mitigation (WGIII AR6 Chapter 7).

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'''Box 3.4 | Blue Carbon Ecosystems'''
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Climate change and other anthropogenic drivers, including eutrophication, land-use changes and overexploitation, directly and indirectly threaten blue carbon ecosystems (Annex II: Glossary). Commonly considered blue carbon ecosystems include vegetated coastal ecosystems (Sections 3.4.2.3–3.4.2.5), whose mangroves, salt marshes and seagrass beds host rooted, vascular plants known to store large amounts of carbon for long periods and to be amenable to management ( [[#Lovelock--2019|Lovelock and Duarte, 2019]] ). Other ocean and coastal taxa, including rooted or floating macroalgae (e.g., non-vascular multicellular kelp or seaweed genera such as ''Macrocystis'' spp., ''Sargassum'' spp. or ''Laminaria'' spp. ( [[#Filbee-Dexter--2020|Filbee-Dexter and Wernberg, 2020]] ), phytoplankton and even pelagic fauna (e.g., finfish or whales; [[#Chami--2019|Chami et al., 2019]] ), have also been proposed as blue carbon ecosystems. Terrestrial vascular-plant-derived material can also carry and store significant amounts of carbon in marine environments ( [[#Cragg--2020|Cragg et al., 2020]] ). There is increasing evidence about the coverage and carbon content of macroalgal, planktonic and faunal taxa, but ''low agreement'' about their long-term carbon-storage potential and manageability ( [[#Alongi--2018b|Alongi, 2018b]] ; [[#Wernberg--2018|Wernberg and Filbee-Dexter, 2018]] ; [[#Lovelock--2019|Lovelock and Duarte, 2019]] ; [[#Ortega--2019|Ortega et al., 2019]] ; [[#Pfister--2019|Pfister et al., 2019]] ; [[#Queirós--2019|Queirós et al., 2019]] ; [[#Filbee-Dexter--2020a|Filbee-Dexter et al., 2020a]] ; [[#Gallagher--2020|Gallagher, 2020]] ; [[#Mariani--2020|Mariani et al., 2020]] ; [[#Thorhaug--2020|Thorhaug et al., 2020]] ; [[#van%20Son--2020|van Son et al., 2020]] ; [[#Bach--2021|Bach et al., 2021]] ; [[#Bayley--2021|Bayley et al., 2021]] ; [[#Cavanagh--2021|Cavanagh et al., 2021]] ; [[#Frontier--2021|Frontier et al., 2021]] ; [[#Martin--2021|Martin et al., 2021]] ; [[#Pedersen--2021|Pedersen et al., 2021]] ; [[#Weigel--2021|Weigel and Pfister, 2021]] ). This section focuses on the array of ecosystem services and adaptation opportunities provided by vegetated coastal blue carbon ecosystems, where consensus and evidence are most abundant. Mitigation potential of blue carbon ecosystems is assessed with land-based mitigation options in WGIII AR6 [[IPCC:Wg2:Chapter:Chapter-7#7.4|Section 7.4]] .

Carbon storage and burial in mangroves, salt marshes and seagrass meadows (see Table Box 3.4.1) help regulate ocean and coastal carbon cycling and may contribute to nature-based mitigation, although regional estimates vary widely based on climatic and edaphic conditions (WGIII AR6 [[IPCC:Wg2:Chapter:Chapter-7#7.4|Section 7.4]] ). In addition, coastal vegetated ecosystems provide substantial and interdependent regulating, provisioning and cultural ecosystem services. These services include: (a) disproportionately high biodiversity per unit area ( [[#Pörtner--2021a|Pörtner et al., 2021a]] ); (b) abundant habitat ( [[#3.5.5.1|Section 3.5.5.1]] ) and nurseries for aquatic, terrestrial, aerial and microbial species; (c) natural filtration of waste and stormwater runoff into the coastal ocean (Sections 3.5.5.3, 4.2.7; Cross-Chapter Box ILLNESS in Chapter 2); (d) coastal protection ( [[#3.5.5.4|Section 3.5.5.4]] ; [[#Ouyang--2018|Ouyang et al., 2018]] ; [[#Quevedo--2020|Quevedo et al., 2020]] ); (e) food and natural materials (Sections 3.5.3, 3.5.4); and (f) support for tourism, livelihoods and cultural activities ( [[#3.5.6|Section 3.5.6]] ). Global estimates of services provided by coastal blue carbon ecosystems depend on the quality of available mapping, which is currently best developed for mangroves ( [[#Macreadie--2019|Macreadie et al., 2019]] ), and improving for salt marshes and seagrasses ( [[#McOwen--2017|McOwen et al., 2017]] ; [[#McKenzie--2020|McKenzie et al., 2020]] ; [[#Young--2021|Young et al., 2021]] ).

'''Table Box 3.4.1 |''' Estimates of organic carbon storage and burial rates in mangroves, salt marshes and seagrass meadows a

{| class="wikitable"
|-
!
! Mangroves

! Salt marshes

! Seagrass meadows

|-
| Carbon stocks (MgC ha –1 )

| 856 ± 64.2 [79–2208] ( [[#Kauffman--2020|Kauffman et al., 2020]] )

| 317.2 ± 38.2 [27–1900] ( [[#Alongi--2018c|Alongi, 2018c]] )

| 139.7 [9.1–628] ( [[#Fourqurean--2012|Fourqurean et al., 2012]] ; [[#Alongi--2018d|Alongi, 2018d]] )

|-
| Carbon burial rate (g C m –2 yr –1 )

| 194 ± 30 [6.2–1722] ( [[#Wang--2020|Wang et al., 2020]] )

| 168 ± 14 [1.2–1167.5] ( [[#Wang--2020|Wang et al., 2020]] )

| 220.7 ± 40.2 [–2094 to 2124] ( [[#Alongi--2018d|Alongi, 2018d]] )

|-
| Global carbon burial rate (TgC yr –1 )

| 41 ( [[#Wang--2020|Wang et al., 2020]] )

| 12.63 ( [[#Wang--2020|Wang et al., 2020]] )

| 35.31 ( [[#Alongi--2020|Alongi, 2020]] )

|-
| Global areal coverage (Mha)

| 13.7 ( [[#Richards--2020|Richards et al., 2020]] )

| 5.5 ( [[#McOwen--2017|McOwen et al., 2017]] )

| 16 ( [[#McKenzie--2020|McKenzie et al., 2020]] )

|}

(a) Estimates are the mean ± 95% confidence interval, where available (indicating the ''extremely likely'' range) and range. Carbon stocks for mangroves include above- and below-ground storage up to 3 m depth (sampling period 2007–2017). The estimates for salt-marsh and seagrass stocks are soil stocks up to 1 m depth (observations spanning 1983–2016 for salt marshes and until 2016 for seagrass meadows). Date ranges for the burial rates are: 1989–2020, 1975–2020 and 1956–2016 for mangroves, salt marshes and seagrass meadows, respectively.

Coastal vegetated ecosystems are vulnerable to harm from multiple climate-induced and non-climate drivers, and together these have reduced wetland area globally ( ''high confidence'' ) ( [[#3.4.2.5|Section 3.4.2.5]] ) and endangered the services provided by these ecosystems ( ''high confidence'' ). Loss of coastal vegetated ecosystems changes biodiversity (Sections 3.5.2, 3.4.2.3–3.4.2.5; [[#Numbere--2019|Numbere, 2019]] ; [[#Parreira--2021|Parreira et al., 2021]] ), increases risk of damage and erosion from SLR and storms (Sections 3.4.2.3–3.4.2.5; Cross-Chapter Box SLR in Chapter 3; [[#Galeano--2017|Galeano et al., 2017]] ) and impacts provisioning (Sections 3.5.3–3.5.4; [[#Li--2018b|Li et al., 2018b]] ; [[#Maina--2021|Maina et al., 2021]] ). These changes also strongly determine the quantity and longevity of blue carbon storage ( ''high confidence'' ) ( [[#Macreadie--2019|Macreadie et al., 2019]] ; [[#Lovelock--2020|Lovelock and Reef, 2020]] ). Specific site characteristics and ecosystem responses to climate change will determine future local blue carbon storage or loss ( ''high confidence'' ) (see Table Box 3.4.2). For instance, poleward migration of mangroves to areas dominated by salt marshes is expected to increase carbon storage ( [[#Kelleway--2016|Kelleway et al., 2016]] ); however, this change in the dominant vegetation and associated faunal changes can modify carbon stocks and sequestration, as well as other ecosystem services ( [[#Martinetto--2016|Martinetto et al., 2016]] ; [[#Kelleway--2017|Kelleway et al., 2017]] ; [[#Smee--2017|Smee et al., 2017]] ; [[#Macreadie--2019|Macreadie et al., 2019]] ; [[#Macy--2019|Macy et al., 2019]] ). Landward range expansion of mangroves, marshes and seagrass in response to gradual RSLR can enhance carbon sequestration ( [[#3.4.2.5|Section 3.4.2.5]] ; Cross-Chapter Box SLR in Chapter 3; [[#Macreadie--2019|Macreadie et al., 2019]] ), but coastal squeeze can limit this ( [[#Phan--2015|Phan et al., 2015]] ; [[#Schuerch--2018|Schuerch et al., 2018]] ) and RSLR can either submerge and bury or erode and release stored blue carbon ( [[#3.4.2.5|Section 3.4.2.5]] ; [[#Macreadie--2019|Macreadie et al., 2019]] ; [[#Lovelock--2020|Lovelock and Reef, 2020]] ). Gains and losses of mangrove habitat area (and therefore carbon storage) projected for nations under RCP4.5 and RCP8.5 depend primarily on the combination of SLR rate, adaptation scenario (including coastal development) and island or continental status ( [[#Lovelock--2020|Lovelock and Reef, 2020]] ). The influence of warming, MHWs and acidification on seagrass meadows ( [[#Kendrick--2019|Kendrick et al., 2019]] ; [[#Strydom--2020|Strydom et al., 2020]] ), and associated coralligenous reefs ( [[#Zunino--2019|Zunino et al., 2019]] ), suggests that future warming and especially MHWs will cause more widespread loss of services from these ecosystems ( [[#3.4.2.5|Section 3.4.2.5]] ). Loss of blue carbon ecosystems will not only halt carbon storage but also release stored carbon: emissions after 2000 due to global mangrove deforestation have been estimated at 23.5–38.7 Tg Cyr –1 ( [[#Ouyang--2020|Ouyang and Lee, 2020]] ). Mitigation estimates for avoided conversion and restoration of coastal wetlands and the implications of the impacts of climate change are assessed in WGIII AR6 [[IPCC:Wg2:Chapter:Chapter-7#7.4|Section 7.4]] .

<div id="_idContainer093" class="Box_Header-continued"></div>

Box 3.4

To date, initiatives aiming to restore coastal wetland ecosystems primarily address ecosystem characteristics other than carbon storage ( [[#Herr--2017|Herr et al., 2017]] ; [[#de%20los%20Santos--2019|de los Santos et al., 2019]] ; [[#Lovelock--2019|Lovelock and Duarte, 2019]] ; [[#Friess--2020a|Friess et al., 2020a]] ). But recovery of coastal vegetated ecosystems is expected to bring back the full suite of ecosystem services they provide, not just carbon storage ( ''medium confidence'' ) ( [[#Marbà--2015a|Marbà et al., 2015a]] ; [[#Burden--2019|Burden et al., 2019]] ; [[#Friess--2020a|Friess et al., 2020a]] ), making coastal restoration a low-risk action that offers both adaptation and mitigation benefits ( [[#Steven--2020|Steven et al., 2020]] ; [[#Gattuso--2021|Gattuso et al., 2021]] ). Successful restoration requires using appropriate plant species in suitable environmental settings ( [[#Wodehouse--2019|Wodehouse and Rayment, 2019]] ; [[#Friess--2020a|Friess et al., 2020a]] ) with favourable geomorphology and biophysical conditions ( [[#Cameron--2019|Cameron et al., 2019]] ; [[#Ochoa-Gómez--2019|Ochoa-Gómez et al., 2019]] ) and considering social, economic, policy and operational constraints ( [[#3.6.3.2.2|Section 3.6.3.2.2]] ; Cross-Chapter Box NATURAL in Chapter 2), now and in the future ( ''high confidence'' ) ( [[#Duarte--2020|Duarte et al., 2020]] ; [[#Lovelock--2020|Lovelock and Reef, 2020]] ). Nevertheless, restored spaces may not store carbon at rates equal to those of undisturbed spaces ( [[#Yang--2020|Yang et al., 2020]] ), and it may take decades to determine or achieve carbon-storage outcomes of restoration ( [[#Sasmito--2019|Sasmito et al., 2019]] ; [[#Duarte--2020|Duarte et al., 2020]] ; [[#Oreska--2020|Oreska et al., 2020]] ). Integration improves efforts to restore or conserve coastal wetland ecosystems to accomplish both adaptation and mitigation outcomes ( [[#Steven--2020|Steven et al., 2020]] ). Government-led conservation of blue carbon ecosystems as part of national and subnational climate strategies (e.g., [[#Friess--2020a|Friess et al., 2020a]] ; [[#Kelleway--2020|Kelleway et al., 2020]] ; [[#Wedding--2021|Wedding et al., 2021]] ) benefits from coordination with private activities, such as incentivising conservation with payments for ecosystem services ( [[#Muenzel--2018|Muenzel and Martino, 2018]] ; [[#Friess--2020a|Friess et al., 2020a]] ). Moreover, successful area-based protection measures consider both environmental and social issues ( [[#3.6.3.2|Section 3.6.3.2]] ). Continued integration and alignment of policies at international to local levels ( [[#3.6|Section 3.6.5]] ) will also support achieving the adaptation and mitigation benefit of blue carbon spaces ( [[#Friess--2020a|Friess et al., 2020a]] ; [[#Steven--2020|Steven et al., 2020]] ; [[#Wu--2020a|Wu et al., 2020a]] ).

'''Table Box 3.4.2 |''' Examples of vegetated blue carbon ecosystem carbon-storage gains and losses in response to climate-induced drivers, and key actions contributing to maintained and/or increased carbon storage a

{| class="wikitable"
|-
!
! Mangroves

! Salt marshes

! Seagrasses

|-
| ''Sea level rise''

|
|-
| Landward expansion by vegetation

| +C

| +C

| +C

|-
| Coastal squeeze

| −C

| −C

| −C

|-
| Loss of low-lying or submerged land or vegetation

| −C

| −C

| −C

|-
| Human adaptation to increase accommodation space

| +C

| +C

|
|-
| ''Extreme storms''

|
|-
| Erosion, loss of area, subsidence

| −C

| −C

| 0 to −C

|-
| Enhanced sedimentation

| +C

| +C

| +C

|-
| Vegetation damage and mortality

| −C to +C

|
| −C

|-
| ''Warming''

|
|-
| Increased productivity

| +C

|
| +C

|-
| Vegetation mortality

|
| −C

|-
| Increased decomposition of soil

| −C

| −C to +C

|
|-
| Poleward expansion of mangroves

| +C

| −C

|
|-
| Poleward expansion of seagrasses

|
| +C

|-
| Poleward expansion of bioturbators

| ∆C

|
|-
| Change in dominant species

| ∆C

|
|-
| ''Rising concentrations of atmospheric CO'' 2

|
|-
| Increased productivity of some species

| ∆C

| ∆C

| +C

|-
| Biodiversity loss

|
| −C

|-
| ''Altered precipitation''

|
|-
| Vegetation mortality

| −C

|
|-
| Reduced productivity

| −C

| −C

|
|-
| Increased productivity

| +C

|
| +C

|-
| Increased remineralisation

| −C

| −C

|
|-
| Low-salinity events

|
| 0 to −C

|-
| ''Key actions to sustain blue carbon storage''

|
|-
| Protect ecosystems

| X

| X

| X

|-
| Develop alternative livelihoods

| X

|
|-
| Provide space for landward migration

| X

| X

|
|-
| Restore hydrological connections

| X

| X

|
|-
| Maintain or restore sediment supply

| X

| X

|
|-
| Restore ecosystems

| X

|
| X

|-
| Plant indigenous species

|
| X

|
|-
| Reduce nutrient inputs

|
| X

|}

(a) ‘+C’ indicates potential positive effects on blue carbon stocks, ‘−C’ indicates potential negative effects, ‘0’ indicates no effects and ‘∆C’ indicates positive potential or negative effects. Effects on carbon stocks are from [[#3.4.2.5|Section 3.4.2.5]] , [[#Macreadie--2019|Macreadie et al. (2019)]] , [[#Lovelock--2020|Lovelock and Reef (2020)]] and [[#Wang--2020|Wang et al. (2020)]] . Key actions to sustain blue carbon storage are from [[#Duarte--2020|Duarte et al. (2020)]] and [[#Wedding--2021|Wedding et al. (2021)]] .

<div id="_idContainer094" class="Box_Header-continued"></div>

Box 3.4

<div id="3.5.6" class="h2-container"></div>

<span id="cultural-services"></span>