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

==== 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>