Editing IPCC:AR6/WGIII/Chapter-12 (section)

==== 12.3.1.3 Ocean-based Methods ====

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The ocean, which covers over 70% of the Earth’s surface, contains about 38,000 gigatonnes of carbon, some 45 times more than the present atmosphere, and oceanic uptake has already consumed close to 30–40% of anthropogenic carbon emissions (Sabine et al. 2004; [[#Gruber--2019|Gruber et al. 2019]] ). The ocean is characterised by diverse biogeochemical cycles involving carbon, and ocean circulation has much longer timescales than the atmosphere, meaning that additional anthropogenic carbon could potentially be stored in the ocean for centuries to millennia for methods that increase deep ocean-dissolved carbon concentrations or temporarily bury the carbon; or essentially permanently (over ten thousand years) for methods that store the carbon in mineral forms or as ions by increasing alkalinity ( [[#Siegel--2021|Siegel et al., 2021]] ) (Cross-Chapter Box 8, Figure 1). A wide range of methods and implementation options for marine CDR have been proposed ( [[#Gattuso--2018|Gattuso et al. 2018]] ; [[#Hoegh-Guldberg--2018|Hoegh-Guldberg et al. 2018]] ; [[#GESAMP--2019|GESAMP 2019]] ). The most studied ocean-based CDR methods are ocean fertilisation, alkalinity enhancement (including electrochemical methods) and intensification of biologically-driven carbon fluxes and storage in marine ecosystems, referred to as ‘blue carbon’. The mitigation potentials, costs, co-benefits and trade-offs of these three options are discussed below. Less well studied are methods including artificial upwelling, terrestrial biomass dumping into oceans, direct CO 2 removal from seawater (with CCS), and sinking marine biomass into the deep ocean or harvesting it for bioenergy (with CCS) or biochar ( [[#GESAMP--2019|GESAMP 2019]] ). These methods are summarised briefly below. Potential climate response and influence on the carbon budget of ocean-based CDR methods are discussed in WGI AR6, Chapter 5.

One natural mechanism of carbon transfer from the atmosphere to the deep ocean is the ocean biological pump, which is driven by the sinking of organic particles from the upper ocean. These particles derive ultimately from primary production by phytoplankton and most of them are remineralised within the upper ocean with only a small fraction reaching the deep ocean where the carbon can be sequestered on centennial and longer timescales ''.'' Increasing nutrient availability would stimulate uptake of CO 2 through phytoplankton photosynthesis producing organic matter, some of which would be exported into the deep ocean, sequestering carbon. In areas of the ocean where macronutrients (nitrogen, phosphorus) are available in sufficient quantities (about 25% of the total area), the growth of phytoplankton is limited by the lack of trace elements such as iron. Thus, OF CDR can be based on two implementation options to increase the productivity of phytoplankton ( [[#Minx--2018|Minx et al. 2018]] ): macronutrient enrichment and micronutrient enrichment. A third option, highlighted in [[#GESAMP--2019|GESAMP (2019)]] , is based on fertilisation for fish stock enhancement, for instance, as naturally occurs in eastern boundary current systems. Iron fertilisation is the best-studied OF option to date, but knowledge so far is still inadequate to predict global ecological and biogeochemical consequences.

'''Status:''' OF has a natural analogue: periods of glaciation in the geological past are associated with changes in deposition of dust containing iron into the ocean. Increased formation of phytoplankton has also been observed during seasonal deposition of dust from the Arabian Peninsula and ash deposition on the ocean surface after volcanic eruptions ( [[#Achterberg--2013|Achterberg et al. 2013]] ; [[#Jaccard--2013|Jaccard et al., 2013]] ; [[#Olgun--2013|Olgun et al. 2013]] ; [[#Martínez-García--2014|Martínez-García et al. 2014]] ). OF options may appear technologically feasible, and enhancement of photosynthesis and CO 2 uptake from surface waters is confirmed by a number of field experiments conducted in different areas of the ocean, but there is scientific uncertainty about the proportion of newly-formed organic carbon that is transferred to deep ocean, and the longevity of storage ( [[#Blain--2008|Blain et al. 2008]] ; [[#Williamson--2012|Williamson et al. 2012]] ; [[#Trull--2015|Trull et al. 2015]] ). The efficiency of OF also depends on the region and experimental conditions, especially in relation to the availability of other nutrients, light and temperature ( [[#Aumont--2006|Aumont and Bopp 2006]] ). In the case of macronutrients, very large quantities are needed and the proposed scaling of this technique has been viewed as unrealistic ( [[#Williamson--2016|Williamson and Bodle 2016]] ).

'''Costs:''' Ocean fertilisation costs depend on nutrient production and its delivery to the application area ( [[#Jones--2014|Jones 2014]] ). The costs range from USD2 tCO 2 –1 for fertilisation with iron ( [[#Boyd--2008|Boyd 2008]] ) to USD457 tCO 2 –1 for nitrate ( [[#Harrison--2013|Harrison 2013]] ). Reported costs for macronutrient application at USD20 tCO 2 –1 ( [[#Jones--2014|Jones 2014]] ) contrast with higher estimates by ( [[#Harrison--2013|Harrison 2013]] ) reporting that low costs are due to overestimation of sequestration capacity and underestimation of logistical costs. The median of OF cost estimates, USD230 tCO 2 –1 ( [[#Gattuso--2021|Gattuso et al., 2021]] ) indicates low cost-effectiveness, albeit uncertainties are large.

'''Potentials:''' Theoretical calculations indicate that organic carbon export increases 2–20 kg per gram of iron added, but experiments indicate much lower efficiency: a significant part of the CO 2 can be emitted back the atmosphere because much of the organic carbon produced is remineralised in the upper ocean. Efficiency also varies with location ( [[#Bopp--2013|Bopp et al. 2013]] ). Between studies, there are substantial differences in the ratio of iron added to carbon fixed photosynthetically, and in the ratio of iron added to carbon eventually sequestered ( [[#Trull--2015|Trull et al. 2015]] ), which has implications both for the success of this strategy and its cost. Estimates indicate potentially achievable net sequestration rates of 1–3 GtCO 2 yr –1 for iron fertilisation, translating into cumulative CDR of 100–300 GtCO 2 by 2100 ( [[#Ryaboshapko--2015|Ryaboshapko and Revokatova 2015]] ; [[#Minx--2018|Minx et al. 2018]] ), whereas OF with macronutrients has a higher theoretical potential of 5.5 GtCO 2 yr –1 ( [[#Harrison--2017|Harrison 2017]] ; [[#Gattuso--2021|Gattuso et al. 2021]] ). Modelling studies show a maximum effect on atmospheric CO 2 of 15–45 parts per million volume in 2100 ( [[#Zeebe--2005|Zeebe and Archer 2005]] ; [[#Aumont--2006|Aumont and Bopp 2006]] ; [[#Keller--2014|Keller et al. 2014]] ; [[#Gattuso--2021|Gattuso et al. 2021]] ).

'''Risks and impacts:''' Several of the mesoscale iron enrichment experiments have seen the emergence of potentially toxic species of diatoms ( [[#Silver--2010|Silver et al. 2010]] ; [[#Trick--2010|Trick et al. 2010]] ). There is also (limited) evidence of increased concentrations of other GHGs such as methane and nitrous oxide during the subsurface decomposition of the sinking particles from iron-stimulated blooms ( [[#Law--2008|Law 2008]] ). Impacts on marine biology and food web structure are not well known, however OF at large scale could cause changes in nutrient distributions or anoxia in subsurface water ( [[#Fuhrman--1991|Fuhrman and Capone 1991]] ; [[#DFO--2010|DFO 2010]] ). Other potential risks are perturbation to marine ecosystems via reorganisation of community structure, enhanced deep ocean acidification ( [[#Oschlies--2010|]] [[#Oschlies--2010|Oschlies et al. 2010]] ) and effects on human food supply.

'''Co-benefits:''' Co-benefits of OF include a potential increase in fish biomass through enhanced biological production ( [[#Minx--2018|Minx et al. 2018]] ) and reduced ocean acidification in the short term in the upper ocean (by CO 2 removal), though it could be enhanced in the long term in the ocean interior (by CO 2 release) ( [[#Oschlies--2010|]] [[#Oschlies--2010|Oschlies et al., 2010]] ; [[#Gattuso--2018|Gattuso et al. 2018]] ).

'''Trade-offs and spillover effects:''' Potential drawbacks include subsurface ocean acidification and deoxygenation ( [[#Cao--2010|Cao and Caldeira 2010]] ; [[#Oschlies--2010|]] [[#Oschlies--2010|Oschlies et al., 2010]] ; [[#Williamson--2012|Williamson et al. 2012]] ); altered regional meridional nutrient supply and fundamental alteration of food webs ( [[#GESAMP--2019|GESAMP 2019]] ); and increased production of N 2 O and CH 4 ( [[#Jin--2003|Jin and Gruber 2003]] ; [[#Lampitt--2008|Lampitt et al. 2008]] ). Ocean fertilisation is considered to have negative consequences for eight SDGs, and a combination of both positive and negative consequences for seven SDGs ( [[#Honegger--2020|Honegger et al. 2020]] ).

CDR through ‘ocean alkalinity enhancement’ or ‘artificial ocean alkalinisation’ ( [[#Renforth--2017|Renforth and Henderson 2017]] ) can be based on: (i) the dissolution of natural alkaline minerals that are added directly to the ocean or coastal environments; (ii) the dissolution of such minerals upstream from the ocean (e.g., enhanced weathering, [[#12.3.1.2|Section 12.3.1.2]] ); (iii) the addition of synthetic alkaline materials directly to the ocean or upstream; and (iv) electrochemical processing of seawater. In the case of (ii), minerals are dissolved on land and the dissolution products are conveyed to the ocean through runoff and river flow. These processes result in chemical transformation of CO 2 and sequestration as bicarbonate and carbonate ions (HCO 3 – , CO 3 2– ) in the ocean. Imbalances between the input and removal fluxes of alkalinity can result in changes in global oceanic alkalinity and therefore the capacity of the ocean to store carbon. Such alkalinity-induced changes in partitioning of carbon between atmosphere and ocean are thought to play an important role in controlling climate change on timescales of 1000 years and longer (e.g., [[#Zeebe--2012|Zeebe 2012]] ). The residence time of dissolved inorganic carbon in the deep ocean is around 100,000 years. However, residence time may decrease if alkalinity is reduced by a net increase in carbonate minerals by either increased formation (precipitation) or reduced dissolution of carbonate ( [[#Renforth--2017|Renforth and Henderson 2017]] ) ''.'' The alkalinity of seawater could potentially also be increased by electrochemical methods, either directly by reactions at the cathode that increase the alkalinity of the surrounding solution that can be discharged into the ocean, or by forcing the precipitation of solid alkaline materials (e.g., hydroxide minerals) that can then be added to the ocean (e.g., [[#Rau--2013|Rau et al. 2013]] ; [[#La%20Plante--2021|La Plante et al. 2021]] ).

'''Status:''' OAE has been demonstrated by a small number of laboratory experiments (in addition to enhanced weathering, [[#12.3.1.2|Section 12.3.1.2]] ). The use of enhanced ocean alkalinity for carbon storage was first proposed by [[#Kheshgi--1995|Kheshgi (1995)]] who considered the creation of highly reactive lime that would readily dissolve in the surface ocean and sequester CO 2 . An alternative method proposed the dissolution of carbonate minerals (e.g., calcium carbonate) in the presence of waste flue gas CO 2 and seawater as a means capturing CO 2 and converting it to bicarbonate ions ( [[#Rau--1999|Rau and Caldeira 1999]] ; [[#Rau--2011|Rau 2011]] ). [[#House--2007|House et al. (2007)]] proposed the creation of alkalinity in the ocean through electrolysis. The fate of the stored carbon is the same for these proposals (i.e., HCO 3 – and CO 3 2– ions), but the reaction pathway is different. Enhanced weathering of silicate minerals such as olivine could add alkalinity to the ocean, for example, by placing olivine sand in coastal areas ( [[#Meysman--2017|Meysman and Montserrat 2017]] ; [[#Montserrat--2017|Montserrat et al. 2017]] ). Some authors suggest use of maritime transport to discharge calcium hydroxide (slaked lime) ( [[#Caserini--2021|Caserini et al. 2021]] ).

'''Costs:''' Techno-economic assessments of OAE largely focus on quantifying overall energy and carbon balances. Cost ranges are USD40–260 tCO 2 –1 ( [[#Fuss--2018|Fuss et al. 2018]] ). Considering life-cycle carbon and energy balances for various OAE options, adding lime (or other reactive calcium or magnesium oxide/hydroxides) to the ocean would cost USD64–260 tCO 2 –1 ( [[#Renforth--2013|Renforth et al. 2013]] ; Renforth &amp; Kruger 2013; [[#Caserini--2019|Caserini et al. 2019]] ). [[#Rau--2008|Rau (2008)]] and [[#Rau--2018|Rau et al. (2018)]] estimate that electrochemical processes for increasing ocean alkalinity may have a net cost of USD3–160 tCO 2 –1 , largely depending on energy cost and co-product (H 2 ) market value. In the case of direct addition of alkaline minerals to the ocean (i.e., without calcination), the cost is estimated to be USD20–50 tCO 2 –1 ( [[#Harvey--2008|Harvey 2008]] ; [[#Köhler--2013|Köhler et al. 2013]] ; [[#Renforth--2017|Renforth and Henderson 2017]] ).

'''Potentials:''' For OAE, the ocean theoretically has the capacity to store thousands of GtCO 2 (cumulatively) without exceeding pre-industrial levels of carbonate saturation ( [[#Renforth--2017|Renforth and Henderson 2017]] ) if the impacts were distributed evenly across the surface ocean. The potential of increasing ocean alkalinity may be constrained by the capability to extract, process, and react minerals ( [[#12.3.1.2|Section 12.3.1.2]] ); the demand for co-benefits (see below), or to minimise impacts around points of addition. Important challenges with respect to the detailed quantification of the CO 2 sequestration efficiency include nonstoichiometric dissolution, reversed weathering and potential pore water saturation in the case of adding minerals to shallow coastal environments ( [[#Meysman--2017|Meysman and Montserrat 2017]] ). [[#Fuss--2018|Fuss et al. (2018)]] suggest storage potentials of 1–100 GtCO 2 yr –1 . ( [[#González--2016|González and Ilyina 2016]] ) suggested that addition of 114 picomoles of alkalinity to the surface ocean could remove 3400 GtCO 2 from the atmosphere.

'''Risks and impacts:''' For OAE, the local impact of increasing alkalinity on ocean chemistry can depend on the speed at which the impacted seawater is diluted/circulated and the exchange of CO 2 from the atmosphere ( [[#Bach--2019|Bach et al. 2019]] ). Also, more extreme carbonate chemistry perturbations due to non-equilibrated alkalinity could affect local marine biota ( [[#Bach--2019|Bach et al. 2019]] ), although biological impacts are largely unknown. Air-equilibrated seawater has a much lower potential to perturb seawater carbonate chemistry. However, seawater with slow air-sea gas exchange, in which alkalinity increases, consumes CO 2 from the surrounding water without immediate replenishment from the atmosphere, which would increase seawater pH and saturation states and may impact marine biota ( [[#Meysman--2017|Meysman and Montserrat 2017]] ; [[#Montserrat--2017|Montserrat et al. 2017]] ). It may be possible to use this effect to ameliorate ocean acidification. Like enhanced weathering, some proposals may result in the dissolution products of silicate minerals (e.g., silicon, iron, potassium, nickel) being supplied to ocean ecosystems ( [[#Meysman--2017|Meysman and Montserrat 2017]] ; [[#Montserrat--2017|Montserrat et al. 2017]] ). Ecological and biogeochemical consequences of OAE largely depend on the minerals used. When natural minerals such as olivine are used, the release of additional Si and Fe could have fertilising effects ( [[#Bach--2019|Bach et al. 2019]] ). In addition to perturbations to marine ecosystems via reorganisation of community structure, potentially adverse effects of OAE that should be studied include the release of toxic trace metals from some deposited minerals ( [[#Hartmann--2013|Hartmann et al. 2013]] ).

'''Co-benefits:''' Intentional addition of alkalinity to the oceans through OAE would decrease the risk to ocean ecosystems caused by the CO 2 -induced impact of ocean acidification on marine biota and the global carbon cycle ( [[#Doney--2009|Doney et al. 2009]] ; [[#Köhler--2010|Köhler et al. 2010]] ; [[#Rau--2012|Rau et al. 2012]] ; [[#Williamson--2012|Williamson and Turley 2012]] ; [[#Albright--2016|Albright et al. 2016]] ; [[#Bach--2019|Bach et al. 2019]] ) ''.'' OAE could be jointly implemented with enhanced weathering ( [[#12.3.1.2|Section 12.3.1.2]] ), spreading the finely crushed rock in the ocean rather than on land. Regional alkalinisation could be effective in protecting coral reefs against acidification ( [[#Feng--2016|Feng et al. 2016]] ; [[#Mongin--2021|Mongin et al., 2021]] ) and coastal OAE could be part of a broader strategy for geochemical management of the coastal zone, safeguarding specific coastal ecosystems, such as important shellfisheries, from the adverse impact of ocean acidification ( [[#Meysman--2017|Meysman and Montserrat 2017]] ).

'''Trade-offs and spillover effects:''' There is a paucity of research on biological effects of alkalinity addition. The very few studies that have explored the impact of elevated alkalinity on ocean ecosystems have largely been limited to single species experiments ( [[#Cripps--2013|Cripps et al. 2013]] ; [[#Gore--2019|Gore et al. 2019]] ) and a constrained field study quantifying the net calcification response of a coral reef flat to alkalinity enhancement ( [[#Albright--2016|Albright et al. 2016]] ). The addition rate would have to be great enough to overcome mixing of the local seawater with the ambient environment, but not sufficient to detrimentally impact ecosystems. More research is required to assess locations in which this may be feasible, and how such a scheme may operate ( [[#Renforth--2017|Renforth and Henderson 2017]] ). The environmental impact of large-scale release of natural dissolution products into the coastal environment will strongly depend on the scale of olivine application, the characteristics of the coastal water body (e.g., residence time) and the particular biota present (e.g., coral reefs will react differently compared with seagrasses) ( [[#Meysman--2017|Meysman and Montserrat 2017]] ). Model simulations ( [[#González--2018|González et al. 2018]] ) suggest that termination of OAE implemented on a massive scale under a high CO 2 emission scenario (Representative Concentration Pathway 8.5) might pose high risks to biological systems sensitive to rapid environmental changes because it would cause a sharp increase in ocean acidification. For example, OAE termination would lead to a decrease in surface pH in warm shallow regions where vulnerable coral reefs are located, and a drop in the carbonate saturation state. However, other studies with lower levels of OAE have shown no termination effect ( [[#Keller--2014|Keller et al., 2014]] ).

The term ‘blue carbon’ was used originally to refer to biological carbon sequestration in all marine ecosystems, but it is increasingly applied to CDR associated with rooted vegetation in the coastal zone, such as tidal marshes, mangroves and seagrasses. Potential for carbon sequestration in other coastal and non-coastal ecosystems, such as macroalgae (e.g., kelp), is debated ( [[#Krause-Jensen--2016|Krause-Jensen and Duarte, 2016]] ; [[#Krause-Jensen--2018|Krause-Jensen et al., 2018]] ). In this report, blue carbon refers to CDR through coastal blue carbon management.

'''Status:''' In recent years, there has been increasing research on the potential, effectiveness, risks, and possibility of enhancing CO 2 sequestration in shallow coastal ecosystems (Duarte, 2017). About 20% of the countries that are signatories to the Paris Agreement refer to blue carbon approaches for climate change mitigation in their NDCs and are moving toward measuring blue carbon in inventories. About 40% of those same countries have pledged to manage shallow coastal ecosystems for climate change adaptation ( [[#Kuwae--2019|Kuwae and Hori 2019]] ).

'''Costs:''' There are large differences in the cost of CDR applying blue carbon management methods between different ecosystems (and at the local level). Median values are estimated as USD240, 30,000, and 7800 tCO 2 –1 , respectively for mangroves, salt marsh and seagrass habitats ( [[#Gattuso--2021|Gattuso et al. 2021]] ). Currently estimated cost effectiveness (for climate change mitigation) is very low ( [[#Siikamäki--2012|Siikamäki et al. 2012]] ; [[#Bayraktarov--2016|Bayraktarov et al. 2016]] ; [[#Narayan--2016|Narayan et al. 2016]] ).

'''Potentials:''' Globally, the total potential carbon sequestration rate through blue carbon CDR is estimated in the range 0.02–0.08 GtCO 2 yr –1 ( [[#Wilcox--2017|Wilcox et al. 2017]] ; National Academies of Sciences 2019). [[#Gattuso--2021|Gattuso et al. (2021)]] estimate the theoretical cumulative potential of coastal blue carbon management by 2100 to be 95 GtCO 2 , taking into account the maximum area that can be occupied by these habitats and historic losses of mangroves, seagrass and salt marsh ecosystems.

'''Risks and impacts:''' For blue carbon management, potential risks relate to the high sensitivity of coastal ecosystems to external impacts associated with both degradation and attempts to increase carbon sequestration. Under expected future warming, sea level rise and changes in coastal management, blue carbon ecosystems are at risk, and their stored carbon is at risk of being lost ( [[#Bindoff--2019|Bindoff et al. 2019]] ).

'''Co-benefits:''' Blue carbon management provides many non-climatic benefits and can contribute to ecosystem-based adaptation, also reducing emissions associated with habitat degradation and loss ( [[#Howard--2017|Howard et al. 2017]] ; [[#Hamilton--2018|Hamilton and Friess 2018]] ). Shallow coastal ecosystems have been severely affected by human activity; significant areas have already been deforested or degraded and continue to be denuded. These processes are accompanied by carbon emissions. The conservation and restoration of coastal ecosystems, which will lead to increased carbon sequestration, is also essential for the preservation of basic ecosystem services, and healthy ecosystems tend to be more resilient to the effects of climate change.

'''Trade-offs and spillover effects:''' Blue carbon management schemes should consist of a mix of restoration, conservation and areal increase, including complex engineering interventions that enhance natural capital, safeguard their resilience and the ecosystem services they provide, and decrease the sensitivity of such ecosystems to further disturbances.

'''Artificial upwelling:''' This concept uses pipes or other methods to pump nutrient-rich deep ocean water to the surface where it has a fertilising effect (see OF section). To achieve CO 2 removal at a Gt magnitude, modelling studies have shown that artificial upwelling would have to be implemented on a massive scale (over 50% of the ocean to deliver maximum rate of 10GtCO 2 yr –1 under RCP8.5) ( [[#Oschlies--2010|]] [[#Oschlies--2010|Oschlies et al., 2010]] , [[#Keller--2014|Keller et al. 2014]] ). Because the deep water is much colder than surface water, at massive scale this could cool the Earth’s surface by several degrees, but the cooling effect would cease as the deeper ocean warms, and would reverse, leading to rapid warming, if the pumping ceased ( [[#Oschlies--2010|]] [[#Oschlies--2010|Oschlies et al., 2010]] , [[#Keller--2014|Keller et al. 2014]] ).

Furthermore, the cooling would also severely alter atmospheric circulation and precipitation patterns ( [[#Kwiatkowski--2015|Kwiatkowski et al. 2015]] ). Several upwelling approaches have been developed and tested ( [[#Pan--2016|Pan et al., 2016]] ) and more R&amp;D is underway.

'''Terrestrial biomass dumping:''' There are proposals to sink terrestrial biomass (crop residues or logs) into the deep ocean as a means of sequestering carbon ( [[#Strand--2009|Strand and Benford 2009]] ). Sinking biochar has also been proposed ( [[#Miller--2021|Miller and Orton, 2021]] ). Decomposition would be inhibited by the cold and sometimes hypoxic/anoxic environment on the ocean floor, and absence of bacteria that decompose terrestrial lignocellulosic biomass, so storage timescale is estimated at hundreds to thousands of years ( [[#Strand--2009|Strand and Benford 2009]] ) ( [[#Burdige--2005|Burdige 2005]] ). Potential side effects on marine ecosystems, chemistry, or circulation have not been thoroughly assessed. Neither have these concepts been evaluated with respect to the impacts on land from enhanced transfer of nutrients and organic matter to the ocean, nor the relative merits of alternative applications of residues and biochar as an energy source or soil amendment (Chapter 7).

'''Marine biomass CDR options:''' Proposals have been made to grow macroalgae ( [[#Duarte--2017|Duarte et al., 2017]] ) for BECCS ( [[#N’Yeurt--2012|N’Yeurt et al. 2012]] ; [[#Duarte--2013|Duarte et al. 2013]] ; [[#Chen--2015|]] [[#Chen--2015|Chen et al., 2015]] ), to sink cultured macroalgae into the deep sea, or to use marine algae for biochar ( [[#Roberts--2015|Roberts et al., 2015]] ). Naturally-growing sargassum has also been considered for these purposes ( [[#Bach--2021|Bach et al., 2021]] ). [[#Froehlich--2019|Froehlich et al. (2019)]] found a substantial area of the ocean (about 48 million km 2 ) suitable for farming seaweed. [[#N’Yeurt--2012|N’Yeurt et al. (2012)]] suggested that converting 9% of the oceans to macroalgal aquaculture could take up 19 GtCO 2 in biomass, generate 12 Gt per annum of biogas, and the CO 2 produced by burning the biogas could be captured and sequestered. Productivity of farmed macroalgae in the open ocean could potentially be enhanced through fertilising via artificial upwelling ( [[#Fan--2020|Fan et al., 2020]] ) or through cultivation platforms that dive at night to access nutrient-rich waters below the, often nutrient-limited, surface ocean. If the biomass were sunk, it is unknown how long the carbon would remain in the deep ocean and what the additional impacts would be. Research and development on macroalgae cultivation and use is currently underway in multiple parts of the world, though not necessarily directly focused on CDR.

'''Extraction of CO''' 2 '''from seawater (with storage):''' CO 2 can be extracted by applying a vacuum, or by purging with a gas low in CO 2 ( [[#Koweek--2016|Koweek et al., 2016]] ). CO 2 stripping can also be accomplished by acidifying seawater with a mineral acid, or through electrodialysis and electrolysis, to convert bicarbonate ions (HCO 3 – ) to CO 2 ( [[#Willauer--2017|Willauer et al., 2017]] ; Eisaman et al., 2018; [[#Digdaya--2020|Digdaya et al., 2020]] ; [[#Eisaman--2020|Eisaman 2020]] ; Sharifian et al., 2021). The removal of CO 2 from the ocean surface leads to undersaturation in the water, thus forcing CO 2 to move from the atmosphere into the ocean to restore equilibrium. Electrochemical seawater CO 2 extraction has been modelled, prototyped, and analysed from a techno-economic perspective ( [[#Eisaman--2012|Eisaman et al., 2012]] ; [[#Willauer--2017|Willauer et al., 2017]] ; [[#de%20Lannoy--2018|de Lannoy et al., 2018]] ; Eisaman et al., 2018a; Eisaman et al., 2018b).

Status, costs, potentials, risk and impacts, co-benefits, trade-offs and spillover effects and the role in mitigation pathways of ocean-based approaches are summarised in Table 12.6.

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