Editing IPCC:AR6/WGII/Chapter-14 (section)

==== 14.5.2.2 Adaptation: Current State, Barriers and Opportunities ====

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Emerging technologies and cooperative marine management are approaches to facilitate adaptation but require coordination and investment for implementation ( ''high confidence'' ) ( [[#Gattuso--2018|Gattuso et al., 2018]] ; [[#Miller--2018|Miller et al., 2018]] ; [[#Holsman--2019|Holsman et al., 2019]] ; [[#Karp--2019|Karp et al., 2019]] ). Advancements in oceanographic and ecological nowcasting and forecasting tools (i.e., O 2 , pH, temperature, aragonite saturation state, sea ice conditions) can reduce climate impacts by supporting fisheries and aquaculture adaptation along US coasts ( [[#14.5.4|Section 14.5.4]] ; [[#Cooley--2015|Cooley et al., 2015]] ; [[#Irby--2015|Irby et al., 2015]] ; [[#Siedlecki--2015|Siedlecki et al., 2015]] ; [[#Siedlecki--2016|Siedlecki et al., 2016]] ; [[#Siddon--2017|Siddon and Zador, 2017]] ). Forecasts and warnings reduce human exposure to HAB toxins in the Great Lakes, the west coast of Florida, east coast of Texas and the Gulf of Maine ( [[#Anderson--2019|Anderson et al., 2019]] ).

Ocean management that utilises a portfolio of nested, multi-scale, climate-informed and ecosystem-based management approaches in North American waters can increase the resilience of marine ecosystems by addressing multiple stressors simultaneously ( ''high confidence'' ) ( [[#Marshall--2018|Marshall et al., 2018]] ; [[#Holsman--2019|Holsman et al., 2019]] ; [[#Smale--2019|Smale et al., 2019]] ; [[#Holsman--2020|Holsman et al., 2020]] ). Integrated ecosystem assessments ( [[#Foley--2013|Foley et al., 2013]] ; [[#Levin--2014|Levin et al., 2014]] ) are increasingly used to provide strategic advice and context for harvest allocations and bycatch avoidance ( [[#Zador--2017|Zador et al., 2017]] ) as well as early warnings of ecosystem-wide change (e.g., sentinel species, ecological indicators) ( [[#Cavole--2016|Cavole et al., 2016]] ; [[#Hazen--2019|Hazen et al., 2019]] ; [[#Moore--2019|Moore and Kuletz, 2019]] ). Dynamic ocean management policies may improve resilience of marine species and ecosystems to climate ( ''medium confidence'' ) ( [[#Hyrenbach--2000|Hyrenbach et al., 2000]] ; [[#Maxwell--2015|Maxwell et al., 2015]] ; [[#Dunn--2016|Dunn et al., 2016]] ; [[#Tommasi--2017a|Tommasi et al., 2017a]] ; [[#Tommasi--2017b|Tommasi et al., 2017b]] ; [[#Hazen--2018|Hazen et al., 2018]] ; [[#Wilson--2018|Wilson et al., 2018]] ; [[#Holsman--2019|Holsman et al., 2019]] ; [[#Karp--2019|Karp et al., 2019]] ). New proactive and rapid management approaches have been developed to minimise impacts of increasingly frequent entanglements of protected species, caused by climate-driven changes in prey and fishery activities ( [[#Corkeron--2018|Corkeron et al., 2018]] ; [[#Meyer-Gutbrod--2018|Meyer-Gutbrod et al., 2018]] ). Dynamic closure areas are being used to address these issues and reduce loggerhead turtle bycatch in Hawaiian shallow-set longline fisheries ( [[#Howell--2015|Howell et al., 2015]] ; [[#Lewison--2015|Lewison et al., 2015]] ), blue whale ship-strike risk in near-real time ( [[#Hazen--2017|Hazen et al., 2017]] ; [[#Abrahms--2019b|Abrahms et al., 2019b]] ) and bycatch of multiple top predator species in a west coast drift gillnet fishery ( [[#Hazen--2018|Hazen et al., 2018]] ).

Improved coordination and planning at multiple scales will be important for marine species conservation and recovery as species redistribute across fishery areas, marine protected zones, and international and jurisdictional boundaries ( [[#14.5.4|Section 14.5.4]] ; Cross-Chapter Box MOVING PLATE in Chapter 5; [[#Pinsky--2018|Pinsky et al., 2018]] ; [[#Karp--2019|Karp et al., 2019]] ). Indigenous Peoples’ co-management with federal and state partners of marine resources and protected species is an important approach ( [[#14.5.4|Section 14.5.4]] ; Chapters 5 and 6; CCP6; [[#Galappaththi--2019|Galappaththi et al., 2019]] ).

Securing broodstocks for rebuilding and supplementation can be challenging for marine populations already in decline (e.g., blue king crab in Alaska, steelhead salmon in Puget Sound, white abalone in California, groundfish in the northeast USA and Canada) ( [[#14.5.4|Section 14.5.4]] ; Table SM14.8). Marine protected areas can attenuate climate impacts through trophic redundancy, preserving ecological processes, biodiversity and climate refugia ( [[#Roberts--2017a|Roberts et al., 2017a]] ; [[#Schoen--2017|Schoen et al., 2017]] ), although benefits decrease after mid-century (or sooner for high-latitude marine protected areas) as species reach their thermal limit, unless coupled with GHG mitigation ( [[#Bruno--2018|Bruno et al., 2018]] ). Transport, relocation and cultivation of resistant breeds of salmon, oysters, corals, marine mammals and other keystone species, as well as hatchery supplementation of impaired populations of fish and shellfish, are species conservation and recovery methods that will be in greater demand under climate change, although unintended environmental impacts must be considered. Options for protecting and restoring coral reefs to prevent loss of ecosystem function are under development with Florida reef species ( [[#Gattuso--2018|Gattuso et al., 2018]] ; [[#National%20Academies%20of%20Sciences--2019|National Academies of Sciences, 2019]] ). An emerging approach for financing the protection of reefs involves re-categorising reefs as ‘natural infrastructure’ which has allowed for use of insurance to rebuild lost reefs ( [[#Storlazzi--2019|Storlazzi et al., 2019]] ).

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'''Box 14.2 | Wildfire in North America'''
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'''Recent Observations, Attribution to Climate Change and Projections'''
Anthropogenic climate change has led to warmer and drier conditions (i.e., fire weather) that favour wildland fires in North America ( ''high confidence'' ) (see AR6, WGI, Chapter 12, [[#Ranasinghe--2021|Ranasinghe et al., 2021]] ). In response, increased burned area in recent decades in western North America has been facilitated by anthropogenic climate change ( ''medium confidence'' ). Annual numbers of large wildland fires and area burned have risen in the past several decades in the western USA ( [[#USGCRP--2017|USGCRP, 2017]] ; [[#USGCRP--2018|USGCRP, 2018]] ), and area burned has increased in Canada (although the number of large fires has declined slightly recently) ( [[#Gauthier--2014|Gauthier et al., 2014]] ; [[#Natural%20Resources%20Canada--2018|Natural Resources Canada, 2018]] ; [[#Hanes--2019|Hanes et al., 2019]] ). Attribution studies have reported that climate change increased burned area in Canada (1959–1999) ( [[#Gillett--2004|Gillett et al., 2004]] ) as well as the western USA (1984–2015) ( [[#Abatzoglou--2016|Abatzoglou and Williams, 2016]] ) and California (1972–2018) ( [[#Williams--2019a|Williams et al., 2019a]] ). Decreased precipitation was the primary climate-change cause of increased burned area in the western USA, with warming a secondary influence (Holden et al. 2018), whereas warming (through aridity) was most important in a California study ( [[#Williams--2019a|Williams et al., 2019a]] ). A drier atmosphere (including reduced precipitation) has been linked to climate change through altered large-scale atmospheric circulation, which then facilitated greater burned area in the western USA ( [[#Zhang--2019c|Zhang et al., 2019c]] ). Through anomalous warm and dry conditions, anthropogenic climate change contributed to the extreme fires of 2016 ( [[#Kirchmeier-Young--2019|Kirchmeier-]] [[#Young--2019|Young et al., 2019]] ; [[#Tan--2019|Tan et al., 2019]] ) in western Canada and the extreme fire season in 2015 in Alaska ( [[#Partain--2017|Partain et al., 2017]] ). These studies did not include human activities that influence fire–climate relationships ( [[#Syphard--2017|Syphard et al., 2017]] ).

Warming has led to longer fire seasons ( [[#Westerling--2016|Westerling, 2016]] ) and drier fuels ( [[#Williams--2019a|Williams et al., 2019a]] ). Warmer and drier fire seasons in the western USA during 1985–2017 have contributed to greater burned area of severe fires ( [[#Parks--2020|Parks and Abatzoglou, 2020]] ). Simultaneity in fires increased during 1984–2015 ( [[#Podschwit--2020|Podschwit and Cullen, 2020]] ), challenging firefighting effectiveness and resource sharing. In Mexico, fires have been correlated with dry conditions ( [[#Kent--2017|Kent et al., 2017]] ; [[#Marin--2018|Marin et al., 2018]] ; [[#Zuniga-Vasquez--2019|Zuniga-Vasquez et al., 2019]] ). Wildland fire activity in the grasslands of the US Great Plains has increased during the past several decades ( [[#Donovan--2017|Donovan et al., 2017]] ) related to antecedent precipitation or aridity that affected fuel quantity ( [[#Littell--2009|Littell et al., 2009]] ).

Climate change is projected to increase fire activity in many places in North America during the coming decades (see also AR6, WGI, Chapter 12, [[#Ranasinghe--2021|Ranasinghe et al., 2021]] ) ( [[#Boulanger--2014|Boulanger et al., 2014]] ; [[#Williams--2016|Williams et al., 2016]] ; [[#Halofsky--2020|Halofsky et al., 2020]] ), via longer fire seasons ( [[#Wotton--1993|Wotton and Flannigan, 1993]] ; [[#USGCRP--2017|USGCRP, 2017]] ), long-term warming ( [[#Villarreal--2019|Villarreal et al., 2019]] ; [[#Wahl--2019|Wahl et al., 2019]] ) and increased lightning frequency in some areas of the USA and Canada ( ''medium confidence'' ) ( [[#Romps--2014|Romps et al., 2014]] ; [[#Finney--2018|Finney et al., 2018]] ; [[#Chen--2021|Chen et al., 2021]] ). Unusually extensive and severe fires have occurred in the Arctic tundra during recent extremely warm and dry years, suggesting that continued warming may increase the probability of such fires in the future ( [[#Hu--2015|Hu et al., 2015]] ). In drier non-forest ecosystems in the western USA, fires are limited by fuel availability and vegetation productivity; warming will decrease productivity, leading to lower burned area ( [[#Littell--2018|Littell et al., 2018]] ).

'''Impacts on Natural Systems'''
Although fire is a natural process in many North American ecosystems, increases in burned area and severity of wildland fires have had significant impacts on natural ecosystems ( ''medium confidence'' ). The length of streams and rivers impacted by fire has increased in the USA along with burned area (Ball et al. 2021). Mega-fires can cause major changes in the structure and composition of ecosystems, particularly where human alterations are significant ( [[#Stephens--2014|Stephens et al., 2014]] ; [[#Loehman--2020|Loehman et al., 2020]] ). Unusually severe fires may have led to the conversion of forest to grassland in the southwest USA ( [[#Haffey--2018|Haffey et al., 2018]] ). Recent warming and drying have limited post-fire tree seedling and shrub establishment, limiting ecosystem recovery ( [[#Davis--2019|Davis et al., 2019]] ; [[#O’Connor--2020|O’Connor et al., 2020]] ; [[#Rodman--2020|Rodman et al., 2020]] ). In boreal forests, soil carbon is being lost through increasingly severe or frequent fires ( [[#Walker--2019|Walker et al., 2019]] ).

Projected future fire activity will continue to affect ecosystems and alter their structure and function ( ''medium confidence'' ) ( [[#Coop--2020|Coop et al., 2020]] ; [[#Loehman--2020|Loehman et al., 2020]] ). Increased fire activity ( [[#Stevens-Rumann--2018|Stevens-Rumann et al., 2018]] ; [[#Stevens-Rumann--2019|Stevens-Rumann and Morgan, 2019]] ; [[#Turner--2019a|Turner et al., 2019a]] ; [[#Cadieux--2020|Cadieux et al., 2020]] ), further warming and drying that stresses tree seedlings, and model projections of stand-replacing fires at the forest–non-forest boundary in the western USA ( [[#Parks--2019|Parks et al., 2019]] ) have raised the possibility of shifts in species composition or vegetation type ( [[#Halofsky--2020|Halofsky et al., 2020]] ). These projections suggest high variability in ecosystem responses depending on interactions between vegetation type, moisture stress, disturbances regimes and human alterations ( [[#Hurteau--2008|Hurteau et al., 2008]] ; [[#Kitzberger--2017|Kitzberger et al., 2017]] ; [[#Littell--2018|Littell et al., 2018]] ; [[#Hurteau--2019|Hurteau et al., 2019]] ; [[#Loehman--2020|Loehman et al., 2020]] ; [[#O’Connor--2020|O’Connor et al., 2020]] ).

'''Impacts on Human Systems'''
Increased fire activity, partly attributable to anthropogenic climate change, has had direct and indirect effects on mortality and morbidity, economic losses and costs, key infrastructure, cultural resources and water resources ( ''medium confidence'' ), although other factors, such as increasing populations in the wildland–urban interface, have also contributed. During 2000–2018, significant fire events claimed 315 lives in the USA ( [[#NOAA--2019|NOAA, 2019]] ); the economic impacts (e.g., capital, health, indirect losses from economic disruption) from the 2018 California fires were 149 billion USD ( [[#Wang--2021|Wang et al., 2021]] ). Poor air quality from fires caused increased respiratory distress ( ''very high confidence'' ); exposure extends long distances from the fire source ( [[#14.5.6.3|Section 14.5.6.3]] ). In addition to public and private property damage and loss, fires have caused irretrievable losses from archaeological and historical sites ( [[#Ryan--2012|Ryan et al., 2012]] ). Post-fire conditions have created unanticipated challenges for communities’ water supply operations ( [[#Bladon--2014|Bladon et al., 2014]] ; [[#Návar--2015|Návar, 2015]] ; [[#Martin--2016|Martin, 2016]] ) by altering water quality and availability ( [[#Smith--2011|Smith et al., 2011]] ; [[#Bladon--2014|Bladon et al., 2014]] ; [[#Robinne--2020|Robinne et al., 2020]] ) or public safety by increasing exposure to mass wasting events after extreme rainfall events ( [[#Cui--2019|Cui et al., 2019]] ; [[#Kean--2019|Kean et al., 2019]] ). California utilities have proactively shut down parts of their electricity grid to reduce risk of fire during extreme weather, and substantial numbers of people will be increasingly vulnerable to this action in the coming decades ( [[#Abatzoglou--2020|Abatzoglou et al., 2020]] ).

In the USA, annual costs of federal wildland fire suppression have increased by a factor of 4 since 1985 ( [[#USGCRP--2018|USGCRP, 2018]] ) and were 1.5–3 billion USD during 2016–2020 ( [[#NIFC--2021|NIFC, 2021]] ). Annual costs of fire protection in Canada have risen two- to threefold from 1970 to 2017, to $1.0–1.4 billion CAD during 2015–2017 (considering the 2017 CAD value) ( [[#Natural%20Resources%20Canada--2021|Natural Resources Canada, 2021]] ). In one of its worst fire seasons, British Columbia expended over 500 million CAD in 2017 for fire suppression ( [[#Natural%20Resources%20Canada--2018|Natural Resources Canada, 2018]] ). The number of days of synchronous fire danger is expected to double in the western USA by 2051–2080, thereby increasing demands on fire suppression resources ( [[#Abatzoglou--2021|Abatzoglou et al., 2021]] ).

The 2016 Fort McMurray fire ranks as the costliest natural disaster in Canada to date (3 billion CAD in insured damages) ( [[#Mamuji--2018|Mamuji and Rozdilsky, 2018]] ; [[#IBC--2020|IBC, 2020]] ). More than 88,000 people were evacuated; many were not aware of the high pre-existing fire risk and had limited warning to prepare and leave ( [[#McGee--2019|McGee, 2019]] ). The community subsequently required extensive social support and experienced mental health challenges ( [[#Government%20of%20Alberta--2016|Government of Alberta, 2016]] ; [[#Cherry--2017|Cherry and Haynes, 2017]] ; [[#Mamuji--2018|Mamuji and Rozdilsky, 2018]] ; [[#Brown--2019a|Brown et al., 2019a]] ; [[#McGee--2019|McGee, 2019]] ). Although a broad recovery plan was developed ( [[#Regional%20Municipality%20of%20Wood%20Buffalo--2016|Regional Municipality of Wood Buffalo, 2016]] ), reconstruction and economic recovery has been slow ( [[#Mamuji--2018|Mamuji and Rozdilsky, 2018]] ).

Wildland fire was identified as a top climate-change risk facing Canada ( [[#Council%20of%20Canadian%20Academies--2019|Council of Canadian Academies, 2019]] ) and poses a challenge to communities and fire management ( [[#Coogan--2019|Coogan et al., 2019]] ). Projected area burned in Canada using RCP2.6 will increase annual fire suppression costs to 1 billion CAD the by end of century (60% increase relative to 1980–2009) and to 1.4 billion CAD using RCP8.5 (119% increase) ( [[#Hope--2016|Hope et al., 2016]] ). In the USA, cumulative costs of fire response through 2100 are projected to be 23 billion USD (considering the 2015 USD value) yr −1 under RCP8.5 ( [[#EPA--2017|EPA, 2017]] ). Lower-emissions scenarios reduce these future cumulative costs by 55 million USD ( [[#EPA--2017|EPA, 2017]] ) to 7–9 billion USD (considering the 2005 USD value) ( [[#Mills--2015a|Mills et al., 2015a]] ). Fire increases from future warming will reduce timber supply in eastern Canada ( [[#Gauthier--2015|Gauthier et al., 2015]] ; [[#Chaste--2019|Chaste et al., 2019]] ) and increase post-fire sedimentation in watersheds of the western USA ( [[#Sankey--2017|Sankey et al., 2017]] ).

'''Adaptation'''
Wildland fire risks are not equitably distributed as they intersect with exposure and socioeconomic attributes (e.g., age, income, ethnicity) to influence vulnerability and adaptive capacity ( ''medium confidence'' ) ( [[#Wigtil--2016|Wigtil et al., 2016]] ; [[#Davies--2018|Davies et al., 2018]] ; [[#Palaiologou--2019|Palaiologou et al., 2019]] ). Individuals in rural areas, low-income neighbourhoods and immigrant communities, as well as renters in California, had less capacity to prepare for and recover from fire ( [[#Davies--2018|Davies et al., 2018]] ). In the USA, 29 million people live in areas with significant potential for wildfires and 12 million are socially vulnerable ( [[#Davies--2018|Davies et al., 2018]] ). In Canada, there are 117 million ha (14% of total land area) of wildland–human interface, and 96% of populated places have some wildland–urban interface within 5 km ( [[#Johnston--2018|Johnston and Flannigan, 2018]] ).

There is growing recognition of the need to shift fire management and suppression activities to co-exist with more fire on the landscape. This includes widespread use of prescribed fire across landscapes to increase ecological and community-based resilience ( ''high agreement, medium evidence'' ) ( [[#Schoennagel--2017|Schoennagel et al., 2017]] ; [[#McWethy--2019|McWethy et al., 2019]] ; [[#Tymstra--2020|Tymstra et al., 2020]] ). Otherwise, the unprecedented combination of increased human exposure and size of recent mega-fires creates community risks that may exceed conventional operational and forest management response capacity and budgets ( [[#Podur--2010|Podur and Wotton, 2010]] ; [[#Wotton--2017|Wotton et al., 2017]] ; [[#Loehman--2020|Loehman et al., 2020]] ; [[#Moreira--2020|Moreira et al., 2020]] ; [[#Parisien--2020|Parisien et al., 2020]] ) particularly with ongoing population and infrastructure expansion into the wildland–urban interface ( [[#Canadian%20Council%20of%20Forest%20Ministers--2016|Canadian Council of Forest Ministers, 2016]] ; [[#Coogan--2019|Coogan et al., 2019]] ).

Climate-informed post-fire ecosystem recovery measures (e.g., strategic seeding, planting, natural regeneration), restoration of habitat connectivity and managing for carbon sequestration (e.g., soil conservation through erosion control, preservation of old growth forests, sustainable agroforestry) are critical to maximise long-term adaptation potential and reduces future risk through co-benefits with carbon mitigation ( [[#Davis--2019|Davis et al., 2019]] ; [[#Hurteau--2019|Hurteau et al., 2019]] ; [[#Coop--2020|Coop et al., 2020]] ; [[#Stewart--2021|Stewart et al., 2021]] ). Innovation in and scaling up the use of prescribed fire and thinning approaches are contributing to pre- and post-fire resilience goals, including use of Indigenous Peoples burning practices that are receiving a new level of awareness (see Box 14.1; [[#Kolden--2019|Kolden, 2019]] ; [[#Marks-Block--2019|Marks-Block et al., 2019]] ; [[#Long--2020b|Long et al., 2020b]] ).

The tools FireSmart Canada [[#footnote-002|1]] , Firewise USA [[#footnote-001|2]] and Think-Hazard Mexico [[#footnote-000|3]] were devised to reduce fire risks and create fire-resilient communities. They provide design guidance at building, lot, subdivision and community scales, and instruct citizens on creating defensible space ( [[#National%20Fire%20Protection%20Association--2013|National Fire Protection Association, 2013]] ; [[#Firesmart%20Canada--2018|Firesmart Canada, 2018]] ). Implementation has been fragmented and variable as it depends on voluntary uptake by individuals, businesses and communities across a range of adaptive capacities and fire-exposed landscapes ( [[#Smith--2016a|Smith et al., 2016a]] ). Many vulnerable groups do not have access to financial or physical resources to reduce fire risk ( [[#Collins--2009|Collins and Bolin, 2009]] ; [[#Palaiologou--2019|Palaiologou et al., 2019]] ).

Although innovative, holistic approaches to wildland fire management are becoming more common across North America, broader application is necessary to address the growing risks ( ''medium confidence'' ). A social–ecological perspective blends ecosystem complexity, scale and processes into land-use planning along with community values, perception and capacities as well as institutional arrangements ( [[#Smith--2016a|Smith et al., 2016a]] ; [[#Spies--2018|Spies et al., 2018]] ). A risk assessment perspective expands from short-term, reactive fire response to landscape-scale, long-term prevention, mitigation, and preparedness with community and practitioner engagement ( [[#Coogan--2019|Coogan et al., 2019]] ; [[#Sherry--2019|Sherry et al., 2019]] ; [[#Johnston--2020|Johnston et al., 2020]] ; [[#Tymstra--2020|Tymstra et al., 2020]] ).

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[[#footnote-002-backlink|1]] 4 See [http://www.firesmartcanada.ca www.firesmartcanada.ca]

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[[#footnote-001-backlink|2]] 5 See [http://www.nfpa.org www.nfpa.org]

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[[#footnote-000-backlink|3]] 6 See https://thinkhazard.org

Box 14.2

Box 14.2

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'''Box 14.3 | Marine Heatwaves'''
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Marine heatwaves are periods of discrete anomalously high (compared with a 30-year history) sea surface temperatures that persist for a minimum 5 d but up to several months ( [[#Hobday--2016|Hobday et al., 2016]] ; [[#Frölicher--2018|Frölicher et al., 2018]] ; [[#Holbrook--2019|Holbrook et al., 2019]] ; [[#Laufkötter--2020|Laufkötter et al., 2020]] ). There have been MHWs attributed to climate change in every marine system of North America including large areas of the Northwest Atlantic (2012), Caribbean Sea (2015), Bering Sea (2016–2018) and central through Northeast Pacific (2013–2016) (NOAA, 2018; [[#Holbrook--2019|Holbrook et al., 2019]] ; [[#Smale--2019|Smale et al., 2019]] ). Such MHW events have affected kelp forests ( [[#Arafeh-Dalmau--2019|Arafeh-Dalmau et al., 2019]] ), corals ( [[#Eakin--2018|Eakin et al., 2018]] ), seagrasses, bottom-dwelling organisms, marine birds ( [[#Loredo--2019|Loredo et al., 2019]] ; [[#Smale--2019|Smale et al., 2019]] ), mammals ( [[#Suryan--2021|Suryan et al., 2021]] ), fish and shellfish, and marine-dependent human communities ( [[#Huntington--2020|Huntington et al., 2020]] ; [[#Fisher--2021|Fisher et al., 2021]] ; [[#Suryan--2021|Suryan et al., 2021]] ). Increased sea temperatures directly increase metabolic demand and change productivity and behaviour of fish species ( [[#Stock--2017|Stock et al., 2017]] ; [[#Free--2019|Free et al., 2019]] ) as well as induce rapid redistribution of species poleward and to deeper, colder waters ( [[#Pecl--2017|Pecl et al., 2017]] ; [[#Rheuban--2017|Rheuban et al., 2017]] ; [[#Crozier--2019|Crozier et al., 2019]] ; [[#Stevenson--2019|Stevenson and Lauth, 2019]] ; [[#Yang--2019|Yang et al., 2019]] ; [[#Barbeaux--2020|Barbeaux et al., 2020]] ; [[#Cheung--2020|Cheung and Frölicher, 2020]] ). In the Pacific, from the Baja Peninsula to the Bering Sea, there is evidence of widespread shifts in coastal biota and multi-trophic-level starvation of seabirds and whales from combined metabolic demand and reduced prey quality associated with protracted MHWs across multiple regions ((CCP6); [[#Sydeman--2015|Sydeman et al., 2015]] ; Duffy- [[#Anderson--2019|Anderson et al., 2019]] ; [[#Sanford--2019|Sanford et al., 2019]] ; [[#Smale--2019|Smale et al., 2019]] ; Suryan et al. 2021). The distribution of two economically important North American species, Bering Sea Pacific cod ( [[#Pinsky--2013b|Pinsky et al., 2013b]] ; [[#Stevenson--2019|Stevenson and Lauth, 2019]] ; [[#Barbeaux--2020|Barbeaux et al., 2020]] ; [[#Spies--2020|Spies et al., 2020]] ) and American lobster ( [[#Rheuban--2017|Rheuban et al., 2017]] ), have shifted north. The MHW-induced loss of coral reefs across tropical North American waters has varied in severity regionally. For instance, in 2015 and 2016, extensive, severe bleaching affected more than 30% of corals off the southeast USA and a large proportion of US Hawaiian Islands, but had moderate to no impact off the Mexican Yucatan Peninsula ( [[#Frieler--2013|Frieler et al., 2013]] ; [[#Weijerman--2015a|Weijerman et al., 2015a]] ; [[#Weijerman--2015b|Weijerman et al., 2015b]] ; [[#Cinner--2016|Cinner et al., 2016]] ; [[#van%20Hooidonk--2016|van Hooidonk et al., 2016]] ; [[#Hughes--2018|Hughes et al., 2018]] ; [[#Sully--2019|Sully et al., 2019]] ; [[#Williams--2019b|Williams et al., 2019b]] ). Some reefs are exhibiting recovery following efforts focused at reducing non-climate stressors (e.g., overfishing, nutrient pollution and tourism use). Such MHWs are increasing in intensity and frequency ( [[#Hobday--2016|Hobday et al., 2016]] ; [[#Smale--2019|Smale et al., 2019]] ) with the largest increases in frequency and spatial coverage projected for the Gulf of Mexico, US southern east coast and US Pacific Northwest ( [[#Ranasinghe--2021|Ranasinghe et al., 2021]] ) and pose a key risk to marine systems in North America ( [[#14.5.2|Section 14.5.2]] ; Chapters 3, 16).

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