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

== 5.5 Risk-reduction Responses and their Governance ==

<span id="ocean-based-mitigation"></span>
=== 5.5.1 Ocean-based Mitigation ===

<div id="section-5-5-1-1context-for-blue-carbon-and-overview-assessment"></div>

<span id="context-for-blue-carbon-and-overview-assessment"></span>
==== 5.5.1.1 Context for Blue Carbon and Overview Assessment ====

<div id="section-5-5-1-1context-for-blue-carbon-and-overview-assessment-block-1"></div>

There is political and scientific agreement on the need for a wide range of mitigation actions to avoid dangerous climate change (UNEP, 2017 <sup>[[#fn:r1616|1616]]</sup> ; IPCC, 2018 <sup>[[#fn:r1617|1617]]</sup> ). Opportunities to reduce emissions by the greater use of ocean renewable energy are identified in Section 5.4.2.3.2. Here, in accordance with the approved scoping of this report, the assessment of mitigation options is limited to the management of natural ocean processes, that is, requiring policy intervention, with a focus on ‘blue carbon’. Natural processes ''per se'' , although important to the climate system and the global carbon cycle, are not a mitigation response. Two management approaches are possible: first, actions to maintain the integrity of natural carbon stores, thereby decreasing their potential release of greenhouse gases, whether caused by human or climate-drivers; and second, through actions that enhance the longterm (century-scale) removal of greenhouse gases from the atmosphere by marine systems, primarily by biological means.

These mitigation approaches match those proposed using terrestrial natural processes (Griscom et al., 2017 <sup>[[#fn:r1618|1618]]</sup> ), with extensive afforestation and reforestation included in all climate models that limit future warming to 1.5⁰C (de Coninck et al., 2018). As on land, reliable carbon accounting is a critical consideration (Grassi et al., 2017 <sup>[[#fn:r1619|1619]]</sup> ), together with confidence in the longterm security of carbon storage. The feasibility of climatically-significant (and societally acceptable) mitigation using marine natural processes therefore depends on a robust quantitative understanding of how human actions can affect the uptake and release of greenhouse gases from different marine environments, interacting with natural biological, physical and chemical processes. Whilst CO 2 is the most important greenhouse gas, marine fluxes of methane and nitrous oxide can also be important, for both coastal regions and the open ocean (Arévalo-Martínez et al., 2015 <sup>[[#fn:r1620|1620]]</sup> ; Borges et al., 2016 <sup>[[#fn:r1621|1621]]</sup> ; Hamdan and Wickland, 2016 <sup>[[#fn:r1622|1622]]</sup> ).

The term ‘blue carbon’ was originally used to cover biological carbon in all marine ecosystems (Nellemann et al., 2009 <sup>[[#fn:r1623|1623]]</sup> ). Subsequent use of the term has focused on carbon-accumulating coastal habitats structured by rooted plants, such as mangroves, tidal salt marshes and seagrass meadows, that are relatively amenable to management (McLeod et al., 2011 <sup>[[#fn:r1624|1624]]</sup> ; Pendleton et al., 2012 <sup>[[#fn:r1625|1625]]</sup> ; Thomas, 2014 <sup>[[#fn:r1626|1626]]</sup> ; Macreadie et al., 2017a <sup>[[#fn:r1627|1627]]</sup> ; Alongi, 2018 <sup>[[#fn:r1628|1628]]</sup> ; Windham-Myers et al., 2019 <sup>[[#fn:r1629|1629]]</sup> ; Lovelock and Duarte, 2019 <sup>[[#fn:r1630|1630]]</sup> ). Comparisons across the full range of freshwater and saline wetland types are assisted by standardised approaches (Nahlik and Fennessy, 2016 <sup>[[#fn:r1631|1631]]</sup> ; Vázquez-González et al., 2017 <sup>[[#fn:r1632|1632]]</sup> ). Seaweeds (macroalgae) can also be considered as coastal blue carbon (Krause-Jensen and Duarte, 2016 <sup>[[#fn:r1633|1633]]</sup> ; Krause-Jensen et al., 2018 <sup>[[#fn:r1634|1634]]</sup> ; Raven, 2018 <sup>[[#fn:r1635|1635]]</sup> ), however, because of differences in their carbon processing, their climate mitigation potential is assessed separately within Section 5.5.1.2 below.

In the open ocean, the biological carbon pump is driven by the combination of photosynthesis by phytoplankton and downward transfer of particulate carbon by a variety of processes (Henson et al., 2010 <sup>[[#fn:r1636|1636]]</sup> ; DeVries et al., 2017 <sup>[[#fn:r1637|1637]]</sup> ); it results in large-scale transfer of around 10 GtC yr -1 carbon from near-surface waters to the ocean interior (Boyd et al., 2019 <sup>[[#fn:r1638|1638]]</sup> ). Most of this carbon is respired in the mesopelagic and contributes to the 37,000 GtC inventory of DIC, with around ~0.1 GtC yr -1 eventually being permanently removed in deep sea sediments (Cartapanis et al., 2018 <sup>[[#fn:r1639|1639]]</sup> ). In addition, the microbial carbon pump (Jiao et al., 2010 <sup>[[#fn:r1640|1640]]</sup> ) produces refractory dissolved organic molecules throughout the water column at a rate of around 0.4 GtC yr -1 (Jiao et al., 2014b <sup>[[#fn:r1641|1641]]</sup> ), which due to their residence time of hundreds to thousands of years maintain the 700 GtC inventory of dissolved organic carbon in the ocean (Jiao et al., 2010 <sup>[[#fn:r1642|1642]]</sup> ; Jiao et al., 2014a <sup>[[#fn:r1643|1643]]</sup> ; Legendre et al., 2015 <sup>[[#fn:r1644|1644]]</sup> ; Jiao et al., 2018a <sup>[[#fn:r1645|1645]]</sup> ). The natural removal of carbon by the various carbon pumps is closely balanced by upwelling and outgassing, with the ocean a moderate source of CO 2 under pre-industrial conditions (Ciais et al., 2013 <sup>[[#fn:r1646|1646]]</sup> ). The mitigation potential of managing natural processes in the open ocean is only briefly assessed here (Section 5.5.1.3).

Gattuso et al. (2018) <sup>[[#fn:r1647|1647]]</sup> provide an overview assessment of the environmental, technical and societal feasibilities of using a range of ocean management actions to reduce climate change and its impacts. Their results for nine actions based on natural processes are summarised in Figure 5.23, also including marine renewable energy (wind, wave and tidal) for comparison. Eight semi-quantitative criteria were used to assess each action: maximum potential effectiveness by 2100 in reducing climatic drivers (ocean warming, ocean acidification and SLR), assuming full theoretical implementation; technological readiness and lead time to full potential effectiveness (subsequently combined as technical feasibility); duration of benefits; co-benefits; trade-offs (originally described as dis-benefits); cost-effectiveness; and governability (capability of implementation, and management of any associated conflicts). Here, governability is considered as a constraint (governability challenges) reversing the scoring scale used by Gattuso et al. (2018) <sup>[[#fn:r1647|1647]]</sup>

Global measures (circles in Figure 5.23) can be regarded as mitigation, reducing drivers; local measures (rectangles), are primarily EbA, reducing impacts (Section 5.5.2), although they may also contribute to mitigation; two actions were considered at both scales. Gattuso et al. (2018) did not consider the effects of actions on ocean oxygenation, notwithstanding the importance of deoxygenation as a component of climate change. Additional detail is given in SM5.4.

<div id="section-5-5-1-1context-for-blue-carbon-and-overview-assessment-block-2"></div>

<span id="figure-5.23"></span>
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'''Figure 5.23'''

<span id="summary-of-potential-benefits-and-constraints-of-ocean-based-risk-reduction-options-using-natural-processes-from-literature-based-expert-assessments-by-gattuso-et-al.-2018.-mitigation-effectiveness-was-quantified-relative-to-representative-concentration-pathway-rcp8.5-assuming-maximum-theoretical-implementation-with-reduction-of-climate-related-drivers-considered-at-either-global-or-local-100-km2-scale-shown-as-circles-or-rectangles"></span>
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'''Summary of potential benefits and constraints of ocean-based risk-reduction options using natural processes, from literature-based expert assessments by Gattuso et al. (2018). Mitigation effectiveness was quantified relative to Representative Concentration Pathway (RCP)8.5, assuming maximum theoretical implementation, with reduction of climate-related drivers considered at either global or local (&lt;100 km2) scale, shown as circles or rectangles […]'''
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[[File:a072007cf5429aa42751d06d8f804f17 SROCC_Ch05_Figure5.23-scaled.jpg]]

Summary of potential benefits and constraints of ocean-based risk-reduction options using natural processes, from literature-based expert assessments by Gattuso et al. (2018). Mitigation effectiveness was quantified relative to Representative Concentration Pathway (RCP)8.5, assuming maximum theoretical implementation, with reduction of climate-related drivers considered at either global or local (&lt;100 km2) scale, shown as circles or rectangles respectively. Impact reduction, co-benefits and trade-offs are in the context of eight sensitive marine ecosystems and ecosystem services. ‘Technical issues to overcome’ is based on scores for technological readiness, lead time for full implementation and duration of effects. Cost is based on USD per tonne of CO2 either not released or removed from the atmosphere (for global measures) or per hectare of coastal area with action implemented (for local measures). ‘Governance challenges’ shows the potential difficulty of implementation by the international community. NA, not assessed. Additional information on scoring methods is given in SM5.4, Tables SM5.9a and SM5.9b.

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<div id="section-5-5-1-2climate-mitigation-in-the-coastal-ocean"></div>

<span id="climate-mitigation-in-the-coastal-ocean"></span>
==== 5.5.1.2 Climate Mitigation in the Coastal Ocean ====

<div id="section-5-5-1-2climate-mitigation-in-the-coastal-ocean-block-1"></div>

<span id="opportunities-and-challenges-relating-to-coastal-carbon"></span>
===== 5.5.1.2.1 Opportunities and challenges relating to coastal carbon =====

Estuaries, shelf seas and a wide range of other intertidal and shallow-water habitats (Section 5.3) play an important role in the global carbon cycle through their primary production by rooted plants, seaweeds (macroalgae) and phytoplankton, and also by processing riverine organic carbon. However, the natural carbon dynamics of these systems have been greatly changed by human activities (Regnier et al., 2013 <sup>[[#fn:r1648|1648]]</sup> ; Cloern et al., 2016 <sup>[[#fn:r16|16]]</sup> 49; Day and Rybczyk, 2019 <sup>[[#fn:r1650|1650]]</sup> ) ( ''high confidence'' ). Direct anthropogenic impacts include coastal land-use change (Ramesh et al., 2015 <sup>[[#fn:r1651|1651]]</sup> ; Li et al., 2018a <sup>[[#fn:r1652|1652]]</sup> ); indirect effects include increased nutrient delivery and other changes in river catchments (Jiao et al., 2011 <sup>[[#fn:r1653|1653]]</sup> ; Regnier et al., 2013 <sup>[[#fn:r1654|1654]]</sup> ), and marine resource exploitation in shelf seas (Bauer et al., 2013 <sup>[[#fn:r1655|1655]]</sup> ). There is ''high confidence'' that these human-driven changes will continue, reflecting coastal settlement trends and global population growth (Barragán and de Andrés, 2015 <sup>[[#fn:r1656|1656]]</sup> ) .

Policy recognition of the mitigation benefits of coastal ecosystems requires quantitative information on their actual and potential carbon uptake and storage at the local and national scale, within an international framework for carbon accounting (Crooks et al., 2011 <sup>[[#fn:r1657|1657]]</sup> ; Hejnowicz et al., 2015 <sup>[[#fn:r1658|1658]]</sup> ). Such methods are being developed for coastal habitats structured by rooted plants (Needelman et al., 2018 <sup>[[#fn:r1659|1659]]</sup> ; Troxler et al., 2018 <sup>[[#fn:r1660|1660]]</sup> ; Needelman et al., 2019 <sup>[[#fn:r1661|1661]]</sup> ), considered here as ‘coastal vegetation’, linked to protocols for verification of longterm carbon removal and financial incentives (Crooks et al., 2011 <sup>[[#fn:r1662|1662]]</sup> ; Hejnowicz et al., 2015 <sup>[[#fn:r1663|1663]]</sup> ) and building on techniques used for managing terrestrial carbon sinks (Ahmed and Glaser, 2016b <sup>[[#fn:r1664|1664]]</sup> ; Aziz et al., 2016 <sup>[[#fn:r1665|1665]]</sup> ). Proposals to apply carbon accounting to seaweeds, the water column and shelf sea sediments (Krause-Jensen and Duarte, 2016 <sup>[[#fn:r1666|1666]]</sup> ; Zhang et al., 2017 <sup>[[#fn:r1667|1667]]</sup> ) are less well-developed.

<div id="section-5-5-1-2climate-mitigation-in-the-coastal-ocean-block-2"></div>

<span id="coastal-vegetation-mangrove-salt-marsh-and-seagrass-ecosystems"></span>
===== 5.5.1.2.2 Coastal vegetation: mangrove, salt marsh and seagrass ecosystems =====

Mangrove, salt marsh and seagrass habitats are widely recognised as blue carbon ecosystems with mitigation potential (Chmura et al., 2003 <sup>[[#fn:r1668|1668]]</sup> ; Duarte et al., 2005 <sup>[[#fn:r1669|1669]]</sup> ; Kennedy et al., 2010 <sup>[[#fn:r1670|1670]]</sup> ; McLeod et al., 2011 <sup>[[#fn:r1671|1671]]</sup> ). Although covering only ~0.1% of the Earth’s surface, these three ecosystems together have been estimated to support 1–10% of global marine primary production (Duarte et al., 2017 <sup>[[#fn:r1672|1672]]</sup> ). More than 150 countries contain at least one of these ecosystems; 71 countries contain all three (Herr and Landis, 2016 <sup>[[#fn:r1673|1673]]</sup> ), and 74 countries mention such coastal wetlands (five specifically as blue carbon) in their Nationally Determined Contributions (NDCs) to the Paris Agreement (Martin et al., 2016a <sup>[[#fn:r1674|1674]]</sup> ; Gallo et al., 2017 <sup>[[#fn:r1675|1675]]</sup> ).

These three vegetated coastal habitats are characterised by high, yet variable, organic carbon storage in their soils and sediments on a per unit area basis ( ''high confidence'' ). In the humid tropics, mangrove below-ground organic carbon is typically 500–1000 tC ha –1 (Donato et al., 2011 <sup>[[#fn:r1676|1676]]</sup> ; Alongi and Mukhopadhyay, 2015 <sup>[[#fn:r1677|1677]]</sup> ; Howard et al., 2017 <sup>[[#fn:r1678|1678]]</sup> )), although only ~50 tC ha –1 in arid regions (Almahasheer et al., 2017 <sup>[[#fn:r1679|1679]]</sup> ). Australian salt marshes show particularly wide variation in organic carbon storage, ranging from 15–1000 tC ha –1 (top 1 m) with mean of 165 tC ha –1 (Kelleway et al., 2016 <sup>[[#fn:r1680|1680]]</sup> ; Macreadie et al., 2017b <sup>[[#fn:r1681|1681]]</sup> ). For seagrass meadows, storage values are typically 400–1600 tC ha –1 but can exceed 2000 tC ha –1 (Serrano et al., 2014 <sup>[[#fn:r1682|1682]]</sup> ). These accumulations have occurred over decadal to millennial time scales (McKee et al., 2007 <sup>[[#fn:r1683|1683]]</sup> ; Lo Iacono et al., 2008 <sup>[[#fn:r1684|1684]]</sup> ). Such blue carbon stock values are similar to freshwater wetlands and peat, but higher than for most forest soils (Laffoley and Grimsditch, 2009 <sup>[[#fn:r1685|1685]]</sup> ; Pan et al., 2011 <sup>[[#fn:r1686|1686]]</sup> ) ( ''high confidence'' ).

When vegetated coastal ecosystems are disturbed, a proportion of their stored carbon is released back to the atmosphere, along with other greenhouse gases (Marba and Duarte, 2009 <sup>[[#fn:r1686|1686]]</sup> ; Duarte et al., 2010 <sup>[[#fn:r1686|1686]]</sup> ; Pendleton et al., 2012 <sup>[[#fn:r1688|1688]]</sup> ; Lovelock et al., 2017 <sup>[[#fn:r1689|1689]]</sup> ). Globally, around 25–50% of vegetated coastal habitats have already been lost or degraded due to coastal agricultural developments, urbanisation and other human disturbance during the past 100 years (McLeod et al., 2011 <sup>[[#fn:r1690|1690]]</sup> ). The highest historical losses (60–90%) have occurred in Europe and China (Jickells et al., 2015 <sup>[[#fn:r1691|1691]]</sup> ; Gu et al., 2018 <sup>[[#fn:r1692|1692]]</sup> ; Li et al., 2018a <sup>[[#fn:r1693|1693]]</sup> ). Current losses are estimated at 0.2–3.0% yr -1 , depending on vegetation type and location (FAO et al., 2014; Alongi and Mukhopadhyay, 2015 <sup>[[#fn:r1694|1694]]</sup> ; Atwood et al., 2017 <sup>[[#fn:r1695|1695]]</sup> ) ( ''medium confidence'' ). Associated global carbon emissions are estimated at 0.04–0.28 GtC yr –1  (Pendleton et al., 2012 <sup>[[#fn:r1696|1696]]</sup> ); 0.06–0.61 GtC yr –1 (Howard et al., 2017 <sup>[[#fn:r1697|1697]]</sup> ); 0.10–1.46 GtC yr –1 (Lovelock et al., 2017 <sup>[[#fn:r1698|1698]]</sup> ); and 0.007 GtC yr –1 (mangroves only) (Taillardat et al., 2018 <sup>[[#fn:r1699|1699]]</sup> ). This range of values reflects uncertainties regarding the global rate of habitat loss, and the proportion of carbon remineralised to CO 2 .

Mitigation through emission reduction can therefore be achieved by habitat protection, to greatly reduce or end the human-driven loss of mangrove, salt marsh and seagrass ecosystems. Such action could potentially produce nationally-significant mitigation (&gt;1% of fossil fuel emissions) for several countries (Taillardat et al., 2018 <sup>[[#fn:r1700|1700]]</sup> ). However, there are still many uncertainties in quantifying carbon release due to habitat degradation and loss (Lovelock et al., 2017 <sup>[[#fn:r1701|1701]]</sup> ), and hence in determining emission reductions. Furthermore, this mitigation option is not available to those countries where habitat loss is not currently occurring, for example, in Bangladesh (Taillardat et al., 2018 <sup>[[#fn:r1702|1702]]</sup> ). Since legal structures already exist in many countries to protect coastal wetlands, the main policy need may be the enforcement of national regulation and site-specific MPAs (Miteva et al., 2015 <sup>[[#fn:r1703|1703]]</sup> ; Herr et al., 2017 <sup>[[#fn:r1704|1704]]</sup> ; Howard et al., 2017 <sup>[[#fn:r1705|1705]]</sup> ).

The alternative mitigation approach using coastal blue carbon ecosystems is to enhance the natural carbon uptake of such habitats, not only by increasing their spatial coverage through habitat restoration and new habitat creation, but also by taking management measures to maximise the carbon uptake and storage for existing coastal ecosystems. Such measures include reducing anthropogenic nutrient inputs and other pollutants; restoring hydrology, by removing barriers to tidal flow and sediment delivery; and reinstating predators (to reduce carbon loss caused by some bioturbators) (Macreadie et al., 2017a <sup>[[#fn:r1706|1706]]</sup> ). Per unit area of habitat created, restored or rehabilitated, such actions may offer high rates of carbon removal: widely-quoted values are 226 ± 39 gC m -1 yr -1 for mangroves, 218 ± 24 gC m -1 yr -1 for salt marsh and 138 ± 38 gC m -1 yr -1 for seagrass ecosystems (McLeod et al., 2011 <sup>[[#fn:r1707|1707]]</sup> ; Isensee et al., 2019 <sup>[[#fn:r1708|1708]]</sup> ).

Around 90 restoration and rehabilitation projects for mangroves have been documented (López-Portillo et al., 2017 <sup>[[#fn:r1709|1709]]</sup> ), with associated development of a range of restoration evaluation methods (Zhao et al., 2016a <sup>[[#fn:r1710|1710]]</sup> ). Salt marsh restoration is reviewed by Adam (2019) <sup>[[#fn:r1711|1711]]</sup> and seagrass restoration by van Katwijk et al. (2016). Consistent conclusions, supported by other studies (Bayraktarov et al., 2016 <sup>[[#fn:r1712|1712]]</sup> ; Wylie et al., 2016 <sup>[[#fn:r1713|1713]]</sup> ) are that: natural regeneration increases the likelihood of longterm survival; higher success rates are achieved with strong stakeholder engagement; and it is critical that the (human) factors causing original loss and degradation have been properly addressed ( ''high confidence'' ).

Quantification of the climatic benefits of such actions is, however, not straightforward. Measurements of carbon burial rates show high site-specific variability, being strongly affected by a wide range of environmental factors for mangroves (Adame et al., 2017 <sup>[[#fn:r1714|1714]]</sup> ; Schile et al., 2017 <sup>[[#fn:r1715|1715]]</sup> ), seagrasses (Lavery et al., 2013 <sup>[[#fn:r1716|1716]]</sup> ) and salt marshes (Kelleway et al., 2017b <sup>[[#fn:r1717|1717]]</sup> ). The reliable determination of sediment accumulation rates is a key consideration, with associated uncertainties not fully reflected in the McLeod et al. (2011) <sup>[[#fn:r1718|1718]]</sup> estimates given above. In particular, geochemical-based studies have indicated that seagrass carbon burial may have been greatly overestimated (Johannessen and Macdonald, 2016 <sup>[[#fn:r1719|1719]]</sup> ). These issues are contentious (Johannessen and Macdonald, 2018a <sup>[[#fn:r1720|1720]]</sup> ; Johannessen and Macdonald, 2018b <sup>[[#fn:r1721|1721]]</sup> ; Macreadie et al., 2018 <sup>[[#fn:r1722|1722]]</sup> ; Oreska et al., 2018 <sup>[[#fn:r1723|1723]]</sup> ); their scientific resolution is highly desirable. Additional complexities relating to the mitigation role of coastal blue carbon ecosystems include the following:

* Emissions of other greenhouse gases also need to be taken into account (Keller, 2019b <sup>[[#fn:r1724|1724]]</sup> ). Methane release from mangrove habitats can reduce the scale of their climatic benefits by 18–22% (Adams et al., 2012 <sup>[[#fn:r1725|1725]]</sup> ; Chen and Ganapin, 2016 <sup>[[#fn:r1726|1726]]</sup> ; Chmura et al., 2016 <sup>[[#fn:r1727|1727]]</sup> ; Rosentreter et al., 2018 <sup>[[#fn:r1728|1728]]</sup> ; Cameron et al., 2019 <sup>[[#fn:r1729|1729]]</sup> ) and nitrous oxide and methane together may offset salt marsh CO 2 uptake by 24–31% (Adams et al., 2012 <sup>[[#fn:r1730|1730]]</sup> ). Nitrous oxide emissions are strongly affected by nutrient loading (Chmura et al., 2016 <sup>[[#fn:r1731|1731]]</sup> ); under pristine conditions, mangroves can provide a sink rather than a source (Maher et al., 2016 <sup>[[#fn:r1732|1732]]</sup> ). Note that values of the ‘offset’ depend on the metrics used for determining CO 2 equivalents.
* Carbonate formation, releasing CO 2 , may also reduce the benefits of carbon storage by similar proportions (Howard et al., 2017 <sup>[[#fn:r1733|1733]]</sup> ; Macreadie et al., 2017a <sup>[[#fn:r1734|1734]]</sup> ; Kennedy et al., 2018 <sup>[[#fn:r1735|1735]]</sup> ; Saderne et al., 2019 <sup>[[#fn:r1736|1736]]</sup> ).
* Lateral transfers are not well-quantified. Whilst some of the carbon stored in coastal marine sediments may be recalcitrant carbon from terrestrial or atmospheric sources (and should therefore be excluded) (Chew and Gallagher, 2018 <sup>[[#fn:r1737|1737]]</sup> ), export of dissolved organic carbon, inorganic carbon and alkalinity may be considered as additional sequestration (Maher et al., 2018 <sup>[[#fn:r1738|1738]]</sup> ; Santos et al., 2019 <sup>[[#fn:r1739|1739]]</sup> ).
* The permanence of vegetated coastal systems, even if well-protected, cannot be assumed under future temperature regimes (Ward et al., 2016 <sup>[[#fn:r1740|1740]]</sup> ; Duke et al., 2017 <sup>[[#fn:r1741|1741]]</sup> ; Jennerjahn et al., 2017 <sup>[[#fn:r1742|1742]]</sup> ; Nowicki et al., 2017 <sup>[[#fn:r1743|1743]]</sup> )
* Responses to future SLR are also uncertain and complex (Kirwan and Megonigal, 2013 <sup>[[#fn:r1744|1744]]</sup> ; Spencer et al., 2016 <sup>[[#fn:r1745|1745]]</sup> ) . However, impacts are not necessarily negative: carbon sequestration capacity may increase where totally new habitats are created (Barnes, 2017 <sup>[[#fn:r1746|1746]]</sup> ), or if mangroves replace salt marshes (Kelleway et al., 2016 <sup>[[#fn:r1747|1747]]</sup> ).

In summary, a combination of both conservation and restoration of mangrove, salt marsh and seagrass habitats can contribute to national mitigation effort for those countries with relatively large coastlines where such ecosystems naturally occur (Murdiyarso et al., 2015 <sup>[[#fn:r1748|1748]]</sup> ; Atwood et al., 2017 <sup>[[#fn:r1749|1749]]</sup> ). However, the associated current uncertainties in quantifying relevant carbon storage and flows are expected to be problematic for reliable measurement, reporting and verification ( ''high confidence'' ).

At the global scale, synthesis studies have estimated the potential additional sequestration achieved by cost effective coastal blue carbon restoration as ~0.05 GtC yr –1 (Griscom et al., 2017 <sup>[[#fn:r1750|1750]]</sup> ) and 0.04 GtC yr –1 (National Academies of Sciences, Engineering, and Medicine, 2019), assuming that a relatively high proportion of vegetated ecosystems can be re-instated to their 1980–1990 extents. These values compare to current net anthropogenic emissions from all sources of 10.0 GtC yr –1 (Le Quéré et al., 2018), and are consistent with the ‘very low’ scores by (Gattuso et al., 2018) for the climate mitigation benefits of conserving and restoring coastal vegetation (Figure 5.23). Coastal ecosystem restoration could theoretically achieve higher sequestration, around ~0.2 GtC yr-1 (Griscom et al., 2017 <sup>[[#fn:r1752|1752]]</sup> ), but would be challenging, because of the semi-permanent and on-going nature of most coastal land-use change, such as human settlement, conversion to agriculture and aquaculture, shoreline hardening and port development (Gittman et al., 2015 <sup>[[#fn:r1753|1753]]</sup> ; Li et al., 2018a <sup>[[#fn:r1754|1754]]</sup> ).

Restoration costs could also be an important constraint for large-scale application. Based on published data from 246 observations, Bayraktarov et al. (2016) <sup>[[#fn:r1755|1755]]</sup> estimated median total costs for restoration of one hectare of mangrove, salt marsh and seagrass habitat to be ~2,508, 151,129 and 383,672 respectively, in 2010 USD. For each ecosystem, there was high variability in costs according to the economy of the country where the restoration projects were carried out, and the restoration technique applied. Assessment of coastal conservation and restoration costs is also given in Section 4.4.2.3, in Box 5.5 (in the context of coral reef restoration costs) and Section 5.5.2.5.

Measures to protect and restore coastal blue carbon habitats provide many other societal benefits in addition to climate regulation (Section 5.4.1). In particular, there is ''high confidence'' that coastal wetlands benefit local fisheries, enhance biodiversity, give storm protection, reduce coastal erosion, improve water quality and support local livelihoods (Costanza et al., 2008 <sup>[[#fn:r1756|1756]]</sup> ; Spalding et al., 2014 <sup>[[#fn:r1757|1757]]</sup> ). Coastal ecosystems may keep pace with sufficiently gradual SLR, and may be more cost-effective in flood protection than hard infrastructure like seawalls (Temmerman et al., 2013 <sup>[[#fn:r1758|1758]]</sup> ; Möller, 2019 <sup>[[#fn:r1759|1759]]</sup> ). Coastal blue carbon can therefore be considered as a ‘no regrets’ mitigation option at the national level in many countries, in addition to (not a replacement for) more effective mitigation measures. Additional research is needed over the full range of environmental conditions to improve knowledge and understanding of the complex carbon dynamics of coastal vegetation and associated systems, to enable well-quantified and cost-effective carbon sequestration enhancement (Vázquez-González et al., 2017 <sup>[[#fn:r1760|1760]]</sup> ; Windham-Myers et al., 2019 <sup>[[#fn:r1761|1761]]</sup> ).

<div id="section-5-5-1-2climate-mitigation-in-the-coastal-ocean-block-3"></div>

<span id="seaweeds-macroalgae"></span>
===== 5.5.1.2.3 Seaweeds (macroalgae) =====

Seaweeds do not directly transfer carbon to marine sediments, unlike the rooted coastal vegetation considered above (Howard et al., 2017 <sup>[[#fn:r1762|1762]]</sup> ). Nevertheless, seaweed detritus can deliver carbon to sedimentary sites (Hill et al., 2015 <sup>[[#fn:r1763|1763]]</sup> ) and may provide a source of refractory dissolved organic (Krause-Jensen and Duarte, 2016 <sup>[[#fn:r1764|1764]]</sup> ). Recent studies indicate that globally important amounts of carbon may be involved in these processes (Krause-Jensen and Duarte, 2016 <sup>[[#fn:r1765|1765]]</sup> ; Krause-Jensen et al., 2018 <sup>[[#fn:r1766|1766]]</sup> ; Smale et al., 2018 <sup>[[#fn:r1767|1767]]</sup> ). There is, however, currently ''low confidence'' that enhancement of natural seaweed production can provide a significant mitigation response, due to large uncertainties relating to sequestration duration and effectiveness. Such considerations relate to transport pathways, the fate of material transported to deeper water, and the timescales of its subsequent return to the atmosphere over decadal to century timescales.

Seaweed aquaculture is inherently more manageable as a mitigation response (N‘Yeurt et al., 2012 <sup>[[#fn:r1768|1768]]</sup> ; Chung et al., 2013 <sup>[[#fn:r1769|1769]]</sup> ; Chung et al., 2017 <sup>[[#fn:r1770|1770]]</sup> ; Duarte et al., 2017 <sup>[[#fn:r1771|1771]]</sup> ). If linked to biofuel or biogas production (N‘Yeurt and Iese, 2014 <sup>[[#fn:r1772|1772]]</sup> ; Moreira and Pires, 2016 <sup>[[#fn:r1773|1773]]</sup> ; Sondak et al., 2017 <sup>[[#fn:r1774|1774]]</sup> ), there would be potential to reduce emissions (as an alternative to fossil fuels); if also linked to carbon capture and storage (Hughes et al., 2012 <sup>[[#fn:r1775|1775]]</sup> ), it may be possible to achieve negative emissions (net CO 2 removal from the atmosphere). Full life cycle analyses are needed to assess the energy efficiency of such approaches, and the viability of scaling them up to climatically-important levels, taking account of associated environmental and socioeconomic implications.

A different mitigation option using seaweeds relates to their use as a dietary supplement for ruminants to suppress methane production. In vitro studies have given promising results (Dubois et al., 2013 <sup>[[#fn:r1776|1776]]</sup> ; Machado et al., 2016 <sup>[[#fn:r1777|1777]]</sup> ; Machado et al., 2018 <sup>[[#fn:r1778|1778]]</sup> ). However, because the potential scale of real-world benefits have yet to be quantified, there is ''low confidence'' in this approach as a mitigation option.

<div id="section-5-5-1-2climate-mitigation-in-the-coastal-ocean-block-4"></div>

<span id="land-sea-integrated-eco-engineering"></span>
===== 5.5.1.2.4 Land-sea integrated eco-engineering =====

Land-based nutrient management could, in theory, be used to enhance carbon storage in coastal seas and deeper waters, by increasing the amount of refractory dissolved organic carbon (Jiao et al., 2011 <sup>[[#fn:r1779|1779]]</sup> ; Jiao et al., 2014b <sup>[[#fn:r1780|1780]]</sup> ; Jiao et al., 2018b <sup>[[#fn:r1781|1781]]</sup> ). This idea is supported by a statistical analysis of the relationship between organic carbon and nitrate in various natural environments (Taylor and Townsend, 2010 <sup>[[#fn:r1782|1782]]</sup> ) as well as by experimental results in estuarine and offshore waters (Yuan et al., 2010 <sup>[[#fn:r1783|1783]]</sup> ; Jiao et al., 2011 <sup>[[#fn:r1784|1784]]</sup> ; Jiao et al., 2014b <sup>[[#fn:r1785|1785]]</sup> ). Delivery of nutrients from agricultural fertilisers and sewage discharge to coastal waters may currently promote the microbial breakdown of river-derived terrestrial dissolved organic carbon, reducing carbon storage (Liu et al., 2014 <sup>[[#fn:r1786|1786]]</sup> ). Thus reducing nutrient inputs in the future may expand carbon storage by favouring the microbial carbon pump, in addition to the multiple co-benefits of reduced nutrient loads related to HABs, oxygenation and ocean acidification (Miranda et al., 2013 <sup>[[#fn:r1787|1787]]</sup> ; Jiao et al., 2018a <sup>[[#fn:r1788|1788]]</sup> ; Zhang et al., 2018 <sup>[[#fn:r1789|1789]]</sup> ). Although there is some evidence for the impact of dissolved organic carbon variations on global scale climate (Rothman et al., 2003 <sup>[[#fn:r1790|1790]]</sup> ) the benefits of this approach have yet to be determined quantitatively and uncertainties remain regarding the longevity of removal and associated carbon accounting (measurement, reporting and verification). Until such issues are better resolved, there is ''low confidence'' that stimulation of refractory dissolved organic carbon production could provide an operational long-term mitigation measure.

<div id="section-5-5-1-2climate-mitigation-in-the-coastal-ocean-block-5"></div>

<span id="control-of-sediment-disturbance-enhanced-weathering-and-other-geochemical-approaches"></span>
===== 5.5.1.2.5 Control of sediment disturbance, enhanced weathering and other geochemical approaches =====

Anthropogenic sediment disturbance, through fishing, dredging and the installation of offshore structures, affects the security of carbon storage in shelf sea sediments (Hale et al., 2017 <sup>[[#fn:r1791|1791]]</sup> ). Management of such activities might therefore increase carbon retention, over relatively large areas of shelf seas (Avelar et al., 2017 <sup>[[#fn:r1792|1792]]</sup> ; Luisetti et al., 2019 <sup>[[#fn:r1793|1793]]</sup> ). However, there is a lack of data and understanding of the complex processes that affect carbon storage in the potentially mobile fraction of marine sediments (van de Velde et al., 2018 <sup>[[#fn:r1794|1794]]</sup> ); exceptions are provided by Hu et al. (2016) <sup>[[#fn:r1795|1795]]</sup> and Diesing et al. (2017) <sup>[[#fn:r1796|1796]]</sup> . Due to these uncertainties, there is currently ''low confidence'' that control of sediment disturbance can be used for climate mitigation.

There is theoretically greater potential for carbon removal by ‘enhanced weathering’ using mineral additions to coastal waters (and the open ocean) (Rau, 2011; Renforth and Henderson, 2017 <sup>[[#fn:r1797|1797]]</sup> ). These approaches are based on increasing the naturally-occurring uptake of CO 2 by carbonates (e.g., calcite and dolomite) or silicate minerals (such as olivine). Such rock-weathering currently sequesters ~0.25 GtC yr –1 , on land and at sea (Taylor et al., 2015) and provides the longterm control of atmospheric CO 2 concentrations. It could be enhanced by adding ground minerals to beaches (Montserrat et al., 2017 <sup>[[#fn:r1801|1801]]</sup> ) or the sea surface. Other geochemical approaches for adding alkalinity that are less directly based on natural processes (Rau et al., 2012 <sup>[[#fn:r1802|1802]]</sup> ; GESAMP, 2019 <sup>[[#fn:r1803|1803]]</sup> ) are not considered here.

Enhanced weathering methods might be used to reduce local impacts, for example, for coral reefs (Albright et al., 2016b <sup>[[#fn:r1798|1798]]</sup> ; Feng et al., 2016 <sup>[[#fn:r1799|1799]]</sup> ), as well as contributing to wider mitigation of climate change. However, their climatic benefits would be difficult to quantify, with other constraints on their development and deployment relating to the governance, cost and uncertain environmental impacts of large-scale application (Gattuso et al., 2018 <sup>[[#fn:r1800|1800]]</sup> ). The combination of these factors results in ''low confidence'' that enhanced weathering can provide a viable and acceptable climate mitigation approach.

<div id="section-5-5-1-3climate-mitigation-in-the-open-ocean"></div>

<span id="climate-mitigation-in-the-open-ocean"></span>
==== 5.5.1.3 Climate Mitigation in the Open Ocean ====

<div id="section-5-5-1-3climate-mitigation-in-the-open-ocean-block-1"></div>

Recent reviews of the scope for using natural processes in the open ocean for climate mitigation are provided by Keller (2019a) and GESAMP (2019). The summary assessment given here is limited to direct and indirect biologically-based approaches, consistent with the scoping of this report and the major governance constraints on the large-scale application of open ocean interventions.

Current NPP by marine phytoplankton is estimated to be 58 ± 7 GtC yr –1 (Legendre et al., 2015), similar to terrestrial primary production and around 6 times greater than anthropogenic emissions (Le Quere et al. (2016). However, over 99% of the biologically-fixed carbon returns to the atmosphere over a range of timescales (Cartapanis et al., 2018).

The direct method of increasing marine productivity involves adding land-derived nutrients that may currently limit primary production, particularly iron. This approach has been investigated experimentally, by modelling and by observations of natural system behaviour (Keller et al., 2014a; Bowie et al., 2015; Tagliabue et al., 2017). The 13 experimental studies to date (seven in the Southern Ocean, five in the Pacific, and one in the sub-tropical Atlantic) have shown that primary production can be, but is not always, enhanced by the addition of iron (Boyd et al., 2007; Yoon et al., 2016; GESAMP, 2019).

The difficulties arise in demonstrating the time-scale of additional carbon removal, and in obtaining information on the consequences of the fertilisation for other marine ecosystem components, including ocean acidification and other potential side-effects (Williamson and Turley, 2012). Modelling studies (Aumont and Bopp, 2006) indicate that the climatic benefits could be relatively short-lived. Furthermore, public and political acceptability for ocean fertilisation is low (Williamson et al., 2012; Boyd and Bressac, 2016; Williamson and Bodle, 2016; Fuentes-George, 2017; McGee et al., 2018). Ocean iron fertilisation is regulated by the London Protocol, with amendments prohibiting such action unless constituting legitimate scientific research authorised under permit (see Section 5.5.4.1). There are additional governance constraints for the Southern Ocean where ocean iron fertilisation is theoretically considered to be most effective (Robinson et al., 2014).

Open ocean fertilisation by macro-nutrients (e.g., nitrate) has also been proposed, with modelled potential for gigaton-scale carbon removal (Harrison, 2017). Similar technical and governance considerations apply with regard to the quantification of mitigation benefits, the monitoring of potential adverse impacts, and the political acceptability of large-scale deployment. This approach would also involve higher costs, because of the much greater quantities of nutrients required (Williamson and Turley, 2012).

The indirect method of enhancing marine productivity uses physical devices to increase upwelling, thereby increasing the supply of a wide range of naturally-occurring nutrients from deeper water. This technique risks releasing additional CO 2 to the atmosphere, reducing its potential for climate mitigation (Bauman et al., 2014). There may also be other undesirable climatic consequences, including disruption of regional weather patterns and long-term warming rather than cooling, if enhanced upwelling is deployed at large scale (Kwiatkowski et al., 2015).

Because of the many technical, environmental and governance issues relating to marine productivity enhancement, by either direct fertilisation or upwelling, there is ''low confidence'' that such open ocean manipulations provide a viable mitigation measure.

<span id="ocean-based-adaptation"></span>
=== 5.5.2 Ocean-based Adaptation ===

<div id="section-5-5-2ocean-based-adaptation-block-1"></div>

The AR5 concluded, with ''high agreement'' but ''limited evidence'' , that climate change impacts on coastal human settlements and communities could be reduced through coastal adaptation activities (Wong et al., 2014a). The limited evidence of the context-specific application of adaptation principles to support the assessment was highlighted as a knowledge gap for future research. This assessment reports progress made with developing such evidence and assesses human adaptation response to climate change in ecosystems, coastal communities and marine environments.

Components of human adaptation responses include risk assessment, risk reduction, and pathways towards resilience (Cross-Chapter Box 2; Chapter 1.6). Residual risk remains where hazard, vulnerability and exposure intersect, subsequent to an adaptation pathway response. Here we focus on adaptation responses within ecosystems and in human systems, as framed in Chapter 1, and defined by:

* ''Nature-based'' or ''ecosystem-based adaptation'' (5.5.2.1). The use of biodiversity and ecosystem services as part of an overall adaptation strategy to help people to adapt to the adverse effects of climate change. EbA uses the range of opportunities for the sustainable management, conservation, and restoration of ecosystems to provide services that enable people to adapt to the impacts of climate change (Narayan et al., 2016; Moosavi, 2017).
* ''Human systems – Built environment adaptation'' (5.5.2.3.1) Adaptation solutions pertaining to coastal built infrastructure and the systems that support such infrastructure (Mutombo and Ölçer, 2016; Forzieri et al., 2018).
* ''Human systems – Socioinstitutional adaptation'' (5.5.2.3) Adaptation responses within human social, governance and economic systems and sectors (Oswald Beiler et al., 2016; Thorne et al., 2017). This includes, but is not limited to ''community-based adaptation'' by coastal communities (5.5.2.3.2) based on empowering and promoting the adaptive capacity of communities, through appropriate use of context, culture, knowledge, agency, and community preferences (Archer et al., 2014; Shaffiril et al., 2017)

To avoid duplication, detailed consideration of adaptation responses to SLR and extreme events (including heat waves, and compound and cascading events) are avoided here, as they are covered by Chapter 4 and Chapter 6, respectively. Tables 5.7 and 5.8 provide a summary assessment of climate change impacts, human adaptation response and benefits in ecosystems and human systems respectively. Details of the assessed literature are in SM Table 5.7. Climate drivers and impacts reported in the adaptation literature are consistent with those reported in Sections 5.2 and 5.3. Physical impacts include the disruption of physical coastal processes, like sediment dynamics, leading to, for example, erosion, flooding and coastal infrastructure damage (see Tables 5.7 and 5.8). Ecological impacts include the loss of ecosystems and biodiversity (Sections 5.2.3, 5.2.4, 5.3), which affected provision of ecosystem services, like coastal protection or food provision. The most commonly reported non-climate human drivers are growing human coastal populations (Elliff and Silva, 2017; van Oppen et al., 2017a; Gattuso et al., 2018) with poorly planned or managed urban development (Barbier, 2015; Wigand et al., 2017), land use change (Robins et al., 2016), loss of ecosystems (Runting et al., 2017), socioeconomic vulnerability (Broto et al., 2015; Bennett et al., 2016) of many coastal communities, ineffective governance and knowledge gaps for implementation.

<div id="section-5-5-2-1ecosystem-based-adaptation"></div>

<span id="ecosystem-based-adaptation"></span>
==== 5.5.2.1 Ecosystem-based Adaptation ====

<div id="section-5-5-2-1ecosystem-based-adaptation-block-1"></div>

This section assesses adaptation response in coastal ecosystems, beginning with biological adaptation in species, and followed by a summary assessment of EbA as a response to climate change.

<div id="section-5-5-2-1ecosystem-based-adaptation-block-2"></div>

<span id="biological-adaptation"></span>
===== 5.5.2.1.1 Biological adaptation =====

There are many studies on biological climate change adaptation responses (Crozier and Hutchings, 2014 <sup>[[#fn:r1844|1844]]</sup> ; Miller et al., 2017 <sup>[[#fn:r1845|1845]]</sup> ; Diamond, 2018 <sup>[[#fn:r1846|1846]]</sup> ). Sections 5.2.3 and 5.3.3 discuss three main types of biological adaptation, broadly defined: evolutionary (genetic) adaptation through natural selection; phenotypic plasticity (acclimatisation), within an organism’s lifetime; and individual or population mobility towards more favourable conditions. There are, however, expected to be limits to such natural adaptation, and large variations between species and populations (Gienapp and Merilä, 2018 <sup>[[#fn:r1847|1847]]</sup> ).

An accurate understanding of climate change impacts upon species, their sensitivity and adaptive capacity and consequent ecological effects (considering both indirect as well as direct impacts) is used to estimate extinction risk, so that an appropriate management response can be developed (Butt et al., 2016 <sup>[[#fn:r1848|1848]]</sup> ). EbA takes these complex interactions into account (Hobday et al., 2015 <sup>[[#fn:r1849|1849]]</sup> ), including the disruptive impacts of alien invasive species (Ondiviela et al., 2014 <sup>[[#fn:r1850|1850]]</sup> ; Wigand et al., 2017 <sup>[[#fn:r1851|1851]]</sup> ) . Effective adaptation action, therefore, contains a broader consideration than historical conservation practices ( ''medium evidence, high agreement'' ), including the development of international collaborations and databases to improve ocean-scale understanding of climate change impacts (Okey et al., 2014 <sup>[[#fn:r1852|1852]]</sup> ; Young et al., 2015 <sup>[[#fn:r1853|1853]]</sup> ). A key knowledge gap relates to the critical thresholds for irreversible change for species (Powell et al., 2017 <sup>[[#fn:r1854|1854]]</sup> ).

<span id="table-5.7"></span>
<!-- START IMG -->
<!-- TABLE IMG -->
<!-- IMG TITLE -->
'''Table 5.7'''
<!-- IMG CAPTION -->
Summary of reported Adaptation responses (A), the Impacts (I) they aimed to address, and the expected Benefits (B) in coastal ecosystems within Physical, Ecological, Social, Governance, Economic and Knowledge categories. For further details of impacts on ecosystems see Section 5.3. Legend: a + sign indicates ''robust evidence'' , a triangle indicates ''medium evidence'' and an underline indicates ''limited evidence'' . Dark blue cells indicate ''high agreement'' , blue indicates ''medium agreement'' and light blue indicates either ''low agreement'' (denoted by presence of a sign) if sufficient papers were reviewed for an assessment or no assessment (if less than three papers were assessed per cell). The papers used for this assessment can be found in SM5.5.

[[File:7d39cdbf3aab3eb42035a9db2c3f1367 table5.7-a.png]]<br />
[[File:995d701e2b1eb3c4eb9b8539bb2d5ab5 table5.7-b.png]]<br />
[[File:2b4df234366282f1ef81dd3b8f263a19 table5.7-c.png]]<br />
[[File:2493b73e56fe1a2359445f0abd7692fc table5.7-d.png]]

<!-- END IMG -->
<div id="section-5-5-2-1ecosystem-based-adaptation-block-3"></div>

<span id="adaptation-in-coral-reefs"></span>
===== 5.5.2.1.2 Adaptation in coral reefs =====

Coral reefs are currently threatened by the continuous global degradation of warm water coral reef ecosystems and the failure of traditional conservation actions to revive most of the degrading reefs (Rinkevich, 2008 <sup>[[#fn:r1855|1855]]</sup> ; Miller and Russ, 2014 <sup>[[#fn:r1856|1856]]</sup> ). Interventions to rehabilitate degraded coral reef ecosystems can be categorised as preventive (‘passive’ restoration) or adaptive (‘active’ restoration) (Miller and Russ, 2014 <sup>[[#fn:r1857|1857]]</sup> ; Linden and Rinkevich, 2017 <sup>[[#fn:r1858|1858]]</sup> ) (see Box 5.5).

Inspired by silviculture (forestation) approaches to terrestrial ecosystem restoration, studies (Rinkevich, 1995 <sup>[[#fn:r1859|1859]]</sup> ; Rinkevich, 2005 <sup>[[#fn:r1860|1860]]</sup> ; Rinkevich, 2006 <sup>[[#fn:r1861|1861]]</sup> ; Rinkevich, 2008 <sup>[[#fn:r1862|1862]]</sup> ; Bongiorni et al., 2011 <sup>[[#fn:r1863|1863]]</sup> ) have proposed a two step restoration strategy for warm water coral reefs termed gardening of denuded coral reefs. In the first step, a large pool of coral colonies (derived from coral nubbins and fragments, and from sexually derived spat) are farmed in underwater nurseries, preferably on mid-water floating devices installed in sheltered zones, in which coral material can be cultured for up to several years. In the second step, nursery-grown coral colonies, together with recruited associated biota, are transplanted to degraded reef sites (Shafir and Rinkevich, 2008 <sup>[[#fn:r1864|1864]]</sup> ; Mbije et al., 2010 <sup>[[#fn:r1865|1865]]</sup> ; Shaish et al., 2010b <sup>[[#fn:r1866|1866]]</sup> ; Shaish et al., 2010a <sup>[[#fn:r1867|1867]]</sup> ; Bongiorni et al., 2011 <sup>[[#fn:r1868|1868]]</sup> ; Horoszowski-Fridman et al., 2011 <sup>[[#fn:r1869|1869]]</sup> ; Linden and Rinkevich, 2011 <sup>[[#fn:r1870|1870]]</sup> ; Mbije et al., 2013 <sup>[[#fn:r1871|1871]]</sup> ; Cruz et al., 2014 <sup>[[#fn:r1872|1872]]</sup> ; Chavanich et al., 2015 <sup>[[#fn:r1873|1873]]</sup> ; Horoszowski-Fridman et al., 2015 <sup>[[#fn:r1874|1874]]</sup> ; Lirman and Schopmeyer, 2016 <sup>[[#fn:r1875|1875]]</sup> ; Montoya Maya et al., 2016 <sup>[[#fn:r1876|1876]]</sup> ; Ng et al., 2016; Lohr and Patterson, 2017 <sup>[[#fn:r|]]</sup> ; Rachmilovitz and Rinkevich, 2017 <sup>[[#fn:r1878|1878]]</sup> ). Active restoration of coral reefs, while still in its infancy and facing a variety of challenges (Rinkevich, 2015b <sup>[[#fn:r1879|1879]]</sup> ; Hein et al., 2017 <sup>[[#fn:r1880|1880]]</sup> ), has been suggested to potentially improve the ecological status of degraded coral reefs and the socioeconomic benefits that the reefs provide (Rinkevich, 2014 <sup>[[#fn:r1881|1881]]</sup> ; Rinkevich, 2015b <sup>[[#fn:r1882|1882]]</sup> ; Linden and Rinkevich, 2017 <sup>[[#fn:r1883|1883]]</sup> ).

Ecological engineering approaches may promote coral reef adaptation (Rinkevich, 2014 <sup>[[#fn:r1884|1884]]</sup> ; Forsman et al., 2015 <sup>[[#fn:r1885|1885]]</sup> ; Coelho et al., 2017 <sup>[[#fn:r1886|1886]]</sup> ; Horoszowski-Fridman and Rinkevich, 2017 <sup>[[#fn:r1887|1887]]</sup> ; Linden and Rinkevich, 2017 <sup>[[#fn:r1888|1888]]</sup> ; Rachmilovitz and Rinkevich, 2017 <sup>[[#fn:r1889|1889]]</sup> ). They also include: augmenting functional diversity, including that of the microbiome (Casey et al., 2015 <sup>[[#fn:r1890|1890]]</sup> ; Horoszowski-Fridman and Rinkevich, 2017 <sup>[[#fn:r1891|1891]]</sup> ; Shaver and Silliman, 2017 <sup>[[#fn:r1892|1892]]</sup> ); transplantating whole habitats (Shaish et al., 2010b <sup>[[#fn:r1893|1893]]</sup> ; Gómez et al., 2014 <sup>[[#fn:r1894|1894]]</sup> ); and enhancing genetic diversity (Iwao et al., 2014 <sup>[[#fn:r1895|1895]]</sup> ; Drury et al., 2016 <sup>[[#fn:r1896|1896]]</sup> ; Horoszowski-Fridman and Rinkevich, 2017 <sup>[[#fn:r1897|1897]]</sup> ). Active restoration can contribute to reef rehabilitation in all major reef regions (Rinkevich, 2014 <sup>[[#fn:r1898|1898]]</sup> ; Rinkevich, 2015b <sup>[[#fn:r1899|1899]]</sup> ). However, there is ''limited evidence'' on how resistant these manipulated corals are to global change drivers (Shaish et al., 2010b <sup>[[#fn:r1900|1900]]</sup> ; Shaish et al., 2010a <sup>[[#fn:r1901|1901]]</sup> ) or how the nursery time affects biological traits like reproduction in coral transplants (Horoszowski-Fridman et al., 2011 <sup>[[#fn:r1902|1902]]</sup> ). Coral epigenetics may also be used as an adaptive management tool for reef rehabilitation ( ''low confidence'' ), as suggested by studies on coral adaptation (Brown et al., 2002 <sup>[[#fn:r1903|1903]]</sup> ; Horoszowski-Fridman et al., 2011 <sup>[[#fn:r1904|1904]]</sup> ; Palumbi et al., 2014 <sup>[[#fn:r1905|1905]]</sup> ; Putnam and Gates, 2015 <sup>[[#fn:r1906|1906]]</sup> ; Putnam et al., 2016 <sup>[[#fn:r1907|1907]]</sup> ).

Research on active coral reef restoration (Box 5.5) suggests the potential to help rehabilitate degraded coral reefs, provided that the underlying drivers of the impacts are mitigated ( ''high confidence'' ). Ongoing and new research in active coral reef restoration may further improve active reef restoration outcomes (Box 5.5) ( ''low confidence'' ). However, these coral reef restoration options may be ineffectual if global warming exceeds 1.5 o C relative to pre-industrial levels (Hoegh-Guldberg et al., 2018 <sup>[[#fn:r1908|1908]]</sup> ; IPCC, 2018 <sup>[[#fn:r1909|1909]]</sup> ).

<div id="section-5-5-2-1ecosystem-based-adaptation-block-4" class="box"></div>

<span id="box-5.5-coral-reef-restoration-as-ocean-based-adaptation"></span>