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

=== 12.3.1 CDR Methods Not Assessed Elsewhere in This Report: DACCS, Enhanced Weathering and Ocean-based Approaches ===

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This section assesses the CDR methods that are not carried out solely within conventional sectors and so are not covered in other parts of the report: direct air carbon capture and storage, enhanced weathering, and ocean-based approaches. It provides an overview of each CDR method: their costs, potentials, risks and impacts, co-benefits, and their role in mitigation pathways. Since these processes, approaches and technologies have medium to low technology readiness levels, they are subject to significant uncertainty.

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==== 12.3.1.1 Direct Air Carbon Capture and Storage (DACCS) ====

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Direct air capture (DAC) is a chemical process to capture ambient CO 2 from the atmosphere. Captured CO 2 can be stored underground (direct air carbon capture and storage, DACCS) or utilised in products (direct air carbon capture and utilisation, DACCU). DACCS shares with conventional CCS the transport and storage components but is distinct in its capture part. Because CO 2 is a well-mixed GHG, DACCS can be sited relatively flexibly, though its locational flexibility is constrained by the availability of low-carbon energy and storage sites. Capturing the CO 2 involves three basic steps: (i) contacting the air, (ii) capturing on a liquid or solid sorbent or a liquid solvent, and (iii) regeneration of the solvent or the sorbent (with heat, moisture and/or pressure). After capture, the CO 2 stream can be stored underground or utilised. The duration of storage is an important consideration; geological reservoirs or mineralisation result in removal for more than 1000 years. The duration of the removal through DACCU ( [[#Breyer--2019|Breyer et al. 2019]] ) varies with the lifetime of respective products ( [[#Wilcox--2017|Wilcox et al. 2017]] ; [[#Bui--2018|Bui et al. 2018]] ; [[#Fuss--2018|Fuss et al. 2018]] ; [[#Gunnarsson--2018|Gunnarsson et al. 2018]] ; Royal Society and Royal Academy of Engineering 2018; [[#Creutzig--2019|Creutzig et al. 2019]] ), ranging from weeks to months for synthetic fuels to centuries or more for building materials (e.g., concrete cured using mineral carbonation) ( [[#Hepburn--2019|Hepburn et al. 2019]] ). The efficiency and environmental impacts of DACCS and DACCU options depend on the carbon intensity of the energy input (electricity and heat) and other lifecycle assessment (LCA) considerations (Zimmerman 2018; [[#Jacobson--2019|Jacobson 2019]] ). See Chapters 6 and 11 for further details regarding carbon capture and utilisation. Another key consideration is the net carbon CO 2 removal of DACCS over its lifecycle ( [[#Madhu--2021|Madhu et al. 2021]] ). [[#Deutz--2021|Deutz and Bardow (2021)]] and [[#Terlouw--2021|Terlouw et al. (2021)]] demonstrated that the life-cycle net emissions of DACCS systems can be negative, even for existing supply chains and some current energy mixes. They found that the GHG intensity of energy sources is a key factor.

DAC options can be differentiated by the specific chemical processes used to capture ambient CO 2 from the air and recover it from the sorbent ( [[#Fasihi--2019|Fasihi et al. 2019]] ). The main categories are (i) liquid solvents with high-temperature regeneration, (ii) solid sorbents with low-temperature regeneration and (iii) regenerating by moisturising of solid sorbents. Other approaches such as electro-swing ( [[#Voskian--2019|Voskian and Hatton 2019]] ) have been proposed but are less developed. Compared to other CDR methods, the primary barrier to upscaling DAC is its high cost and large energy requirement ( ''high confidence'' ) ( [[#Nemet--2018|Nemet et al. 2018]] ), which can be reduced through innovation. It has therefore attracted entrepreneurs and private investments ( [[#IEA--2020b|IEA 2020b]] ).

'''Status:''' There are some demonstration projects by start-up companies and academic researchers, who are developing various types of DAC, including aqueous potassium solvent with calcium carbonation and solid sorbents with heat regeneration ( [[#NASEM--2019|NASEM 2019]] ). These projects are supported mostly by private investments and grants or sometimes serve utilisation niche markets (e.g., CO 2 for beverages, greenhouses, enhanced oil recovery). As of 2021, there are more than ten plants worldwide, with a scale of ktCO 2 yr –1 or smaller ( [[#Larsen--2019|Larsen et al. 2019]] ; [[#NASEM--2019|NASEM 2019]] ; [[#IEA--2020b|IEA 2020b]] ). Because of the fundamental difference in the CO 2 concentration at the capture stage, DACCS does not benefit directly from research, development and demonstration (RD&amp;D) of conventional CCS. Public RD&amp;D programmes dedicated to DAC have therefore been proposed ( [[#Larsen--2019|Larsen et al. 2019]] ; [[#NASEM--2019|NASEM 2019]] ). Possible research topics include development of new liquid solvents, novel solid sorbents, and novel equipment or system designs, and the need for third-party evaluation of techno-economic aspects has also been emphasised ( [[#NASEM--2019|NASEM 2019]] ). However, since basic research does not appear to be a primary barrier, both [[#NASEM--2019|NASEM (2019)]] and [[#Larsen--2019|Larsen et al. (2019)]] argue for a stronger focus on demonstration in the US context. Though the US and UK governments have begun funding DACCS research ( [[#IEA--2020b|IEA 2020b]] ), the scale of R&amp;D activities is limited.

'''Costs:''' As the process captures dilute CO 2 (~0.04%) from the ambient air, it is less efficient and more costly than conventional carbon capture applied to power plants and industrial installations (with a CO 2 concentration of ~10%) ( ''high confidence'' ). The cost of a liquid solvent system is dominated by the energy cost (because of the much higher energy demand for CO 2 regeneration, which reduces the efficiency) while capital costs account for a significant share of the cost of solid sorbent systems ( [[#Fasihi--2019|Fasihi et al. 2019]] ). The range of the DAC cost estimates found in the literature is wide (USD60–1000 tCO 2 –1 ) ( [[#Fuss--2018|Fuss et al. 2018]] ) partly because different studies assume different use cases, differing phases (first plant vs ''n'' th plant) ( [[#Lackner--2012|Lackner et al. 2012]] ), different configurations, and disparate system boundaries. Estimates of industrial origin are often on the lower side ( [[#Ishimoto--2017|Ishimoto et al. 2017]] ). [[#Fuss--2018|Fuss et al. (2018)]] suggest a cost range of USD600–1000 tCO 2 –1 for first-of-a-kind plants, and USD100–300 tCO 2 –1 as experience accumulates. An expert elicitation study found a similar cost level for 2050 with a median of around USD200 tCO 2 –1 ( [[#Shayegh--2021|Shayegh et al. 2021]] ) ( ''medium evidence'' , ''medium agreement'' ). [[#NASEM--2019|NASEM (2019)]] systematically evaluated the costs of different designs and found a range of 84–386 USD2015 tCO 2 –1 for the designs currently considered by active technology developers. This cost range excludes the site-specific costs of transportation or storage.

'''Potentials:''' There is no specific study on the potential of DACCS but the literature has assumed that the technical potential is virtually unlimited provided that high energy requirements could be met ( ''medium evidence'' , ''high agreement'' ) ( [[#Marcucci--2017|Marcucci et al. 2017]] ; [[#Fuss--2018|Fuss et al. 2018]] ; [[#Lawrence--2018|Lawrence et al. 2018]] ) since DACCS encounters fewer non-cost constraints than any other CDR method. Focusing only on the Maghreb region, [[#Breyer--2020|Breyer et al. (2020)]] reported an optimistic potential 150 GtCO 2 at less than USD61 tCO 2 –1 for 2050. [[#Fuss--2018|Fuss et al. (2018)]] suggest a potential of 0.5–5 GtCO 2 yr –1 by 2050 because of environmental side effects and limits to underground storage. In addition to the ultimate potentials, [[#Realmonte--2019|Realmonte et al. (2019)]] noted the rate of scale-up as a strong constraint on deployment. [[#Meckling--2021|Meckling and Biber (2021)]] discuss a policy roadmap to address the political economy for upscaling. More systematic analysis on potentials is necessary; first and foremost on national and regional levels, including the requirements for low-carbon heat and power, water and material demand, availability of geological storage and the need for land in case of low-density energy sources such as solar or wind power.

'''Risks and impacts:''' DACCS requires a considerable amount of energy ( ''high confidence'' ), depending on the type of technology, water, and make-up sorbents, while its land footprint is small compared to other CDR methods ( [[#Smith--2016|Smith et al. 2016]] ). Yet, depending on the source of energy for DACCS (e.g., renewables vs nuclear), DACCS could require a significant land footprint ( [[#NASEM--2019|NASEM 2019]] ; [[#Sekera--2020|Sekera and Lichtenberger 2020]] ). The theoretical minimum energy requirement for separating CO 2 gas from the air is about 0.5 GJ tCO 2 –1 ( [[#Socolow--2011|Socolow et al. 2011]] ). [[#Fasihi--2019|Fasihi et al. (2019)]] reviewed the published estimates of energy requirements and found that for the current technologies, the total energy requirement is about 4–10 GJ tCO 2 –1 , with heat accounting for about 80% and electricity about 20% ( [[#McQueen--2021|McQueen et al. 2021]] ). At a 10 GtCO 2 yr –1 sequestration scale, this would translate into 40–100 exajoules (EJ) yr –1 of energy consumption (32–80 EJ yr –1 for heat and 8–20 EJ yr –1 electricity), which can be contrasted with the current primary energy supply of about 600 EJ yr –1 and electricity generation of about 100 EJ yr –1 . For the solid sorbent technology, low-temperature heat could be sourced from heat pumps powered by low-carbon sources such as renewables ( [[#Breyer--2020|Breyer et al. 2020]] ), waste heat ( [[#Beuttler--2019|Beuttler et al. 2019]] ), and nuclear energy ( [[#Sandalow--2018|Sandalow et al. 2018]] ). Unless sourced from a clean source, this amount of energy could cause environmental damage ( [[#Jacobson--2019|Jacobson 2019]] ). Because DACCS is an open system, water lost from evaporation must be replenished. Water loss varies, depending on technology (including adjustable factors such as the concentration of the liquid solvent) as well as environmental conditions (e.g., temperate vs tropical climates). For a liquid solvent system, it can be 0–50 tH 2 O tCO 2 –1 ( [[#Fasihi--2019|Fasihi et al. 2019]] ). A water loss rate of about 1–10 tH 2 O tCO 2 –1 ( [[#Socolow--2011|Socolow et al. 2011]] ) would translate into about 10–100 GtH 2 O (10–100 km 3 ) to capture 10 GtCO 2 from the atmosphere. Some solid sorbent technologies actually produce water as a by-product, for example 0.8–2 tH 2 O tCO 2 –1 for a solid-sorbent technology with heat regeneration ( [[#Beuttler--2019|Beuttler et al. 2019]] ; [[#Fasihi--2019|Fasihi et al. 2019]] ). Large-scale deployment of DACCS would also require a significant quantity of materials, and energy to produce them ( [[#Chatterjee--2020|Chatterjee and Huang 2020]] ). Hydroxide solutions are currently being produced as a by-product of chlorine but replacement (make-up) requirement of such materials at scale exceeds the current market supply ( [[#Realmonte--2019|Realmonte et al. 2019]] ). The land requirements for DAC units are not large enough to be of concern ( [[#Madhu--2021|Madhu et al. 2021]] ). Furthermore, these can be placed on unproductive lands, in contrast to biological CDR. Nevertheless, to ensure that CO 2 -depleted air does not enter the air contactor of an adjacent DAC system, there must be enough space between DAC units, similar to wind power turbines. Considering this, [[#Socolow--2011|Socolow et al. (2011)]] estimated a land footprint of 1.5 km 2 MtCO 2 –1 . In contrast, large energy requirements can lead to significant footprints if low-density energy sources (e.g., solar PV) are used ( [[#Smith--2016|Smith et al. 2016]] ). For the issues associated with CO 2 utilisation and storage, see Chapter 6.

'''Co-benefits:''' While [[#Wohland--2018|Wohland et al. (2018)]] proposed solid sorbent-based DAC plants as a Power-to-X technology that could use excess renewable power (at times of low or even negative prices), such operation would add additional costs. Installations would need to be designed for intermittent operations (i.e., at low load factors) which would negatively affect capital and operation costs ( [[#Daggash--2018|Daggash et al. 2018]] ; [[#Sandalow--2018|Sandalow et al. 2018]] ) as a high time-resolution model suggests a high utilisation rate ( [[#Breyer--2020|Breyer et al. 2020]] ). Solid sorbent DAC designs can potentially remove more water from the ambient air than needed for regeneration, thereby delivering surplus water that would contribute to SDG 6 (clean water and sanitation) in arid regions ( [[#Sandalow--2018|Sandalow et al. 2018]] ; [[#Fasihi--2019|Fasihi et al. 2019]] ).

'''Trade-offs and spillover effects:''' Liquid solvent DACCS systems need substantial amounts of water ( [[#Fasihi--2019|Fasihi et al. 2019]] ), although much less than BECCS systems ( [[#Smith--2016|Smith et al. 2016]] ), which could negatively affect SDG 6 (clean water and sanitation). Although the high energy demand of DACCS could affect SDG 7 (affordable and clean energy) negatively through potential competition or positively through learning effects ( [[#Beuttler--2019|Beuttler et al. 2019]] ), its impact has not been thoroughly assessed yet.

'''Role in mitigation pathways:''' There are a few IAM studies that have explicitly incorporated DACCS. Stringent emissions constraints in these studies lead to high carbon prices, allowing DACCS to play an important role in mitigation. [[#Chen--2013|Chen and Tavoni (2013)]] examined the role of DACCS in an IAM (WITCH) and found that incorporating DACCS reduces the overall cost of mitigation and tends to postpone the timing of mitigation. The scale of capture goes up to 37 GtCO 2 yr –1 in 2100. [[#Akimoto--2021|Akimoto et al. (2021)]] introduced DACCS in the IAM DNE21+, and also found the long-term marginal cost of abatement is significantly reduced by DACCS. [[#Marcucci--2017|Marcucci et al. (2017)]] ran MERGE-ETL, an integrated model with endogenous learning, and showed that DACCS allows for a model solution for the 1.5°C target, and that DACCS substitutes for BECCS under stringent targets. In their analysis, DACCS captures up to 38.3 GtCO 2 yr –1 in 2100. [[#Realmonte--2019|Realmonte et al. (2019)]] modelled two types of DACCS (based on liquid and solid sorbents) with two IAMs (TIAM-Grantham and WITCH), and showed that in deep mitigation scenarios, DACCS complements, rather than substitutes, other CDR methods such as BECCS, and that DACCS is effective at containing mitigation costs. At the national scale, [[#Larsen--2019|Larsen et al. (2019)]] utilised the Regional Investment and Operations (RIO) Platform coupled with the Energy PATHWAYS model, and explicitly represented DAC in US energy systems scenarios. They found that in a scenario that reaches net zero emissions by 2045, about 0.6 GtCO 2 or 1.8 GtCO 2 of DACCS would be deployed, depending on the availability of biological carbon sinks and bioenergy. The modelling supporting the European Commission’s initial proposal for net zero GHG emissions by 2050 incorporated DAC, with the captured CO 2 used for both synthetic fuel production (DACCU) and storage (DACCS) ( [[#Capros--2019|Capros et al. 2019]] ). [[#Fuhrman--2021a|Fuhrman et al. (2021a)]] evaluated the role of DACCS across five shared socio-economic pathways with the GCAM modelling framework and identified a substantial role for DACCS in mitigation and a decreased pressure on land and water resources from BECCS, even under the assumption of limited energy efficiency improvement and conservative cost declines of DACCS technologies. The newest iteration of the World Economic Outlook by [[#IEA--2021b|IEA (2021b)]] deploys CDR on a limited scale, and DACCS removes 0.6 GtCO 2 in 2050 for its Net Zero CO 2 Emissions scenario.

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

'''Table 12.6 | Summary of status, costs, potentials, risk and impacts, co-benefits, trade-offs and spillover effects and the role in mitigation pathways for CDR methods.''' Technology readiness level (TRL) is a measure of maturity of the CDR method. Scores range from 1 (basic principles defined) to 9 (proven in operational environment). Author judgement ranges (assessed by authors in the literature) are shown, with full literature ranges shown in brackets.

{| class="wikitable"
|-
! '''CDR method'''

! '''Status (TRL)'''

! '''Cost (USD tCO''' 2 –1 ''')'''

! '''Mitigation Potential (GtCO''' 2 '''y''' '''r''' –1 ''')'''

! '''Risk and impacts'''

! '''Co-benefits'''

! '''Trade-offs and spillover effects'''

! '''Role in modelled mitigation pathways'''

! '''Section'''

|-
| DACCS

| 6

| 100–300 (84–386)

| 5–40

| Increased energy and water use

| Water produced (solid sorbent DAC designs only)

| Potentially increased emissions from water supply and energy generation

| In a few IAMs; DACCS complements other CDR methods

| 12.3.1.1

|-
| Enhanced weathering

| 3–4

| 50–200 (24–578)

| 2–4 (&lt;1–95)

| Mining impacts; air quality impacts of rock dust when spreading on soil

| Enhanced plant growth, reduced erosion, enhanced soil carbon, reduced soil acidity, enhanced soil water retention

| Potentially increased emissions from water supply and energy generation

| In a few IAMs; EW complements other CDR methods

| 12.3.1.2

|-
| Ocean alkalinity enhancement

| 1–2

| 40–260

| 1–100

| Increased seawater pH and saturation states may impact marine biota. Possible release of nutritive or toxic elements and compounds. Mining impacts

| Limiting ocean acidification

| Potentially increased emissions of CO 2 and dust from mining, transport and deployment operations

| No data

| 12.3.1.3

|-
| Ocean fertilisation

| 1–2

| 50–500

| 1–3

| Nutrient redistribution, restructuring of the ecosystem, enhanced oxygen consumption and acidification in deeper waters, potential for decadal-to-millennial-scale return to the atmosphere of nearly all the extra carbon removed, risks of unintended side effects

| Increased productivity and fisheries, reduced upper ocean acidification

| Subsurface ocean acidification, deoxygenation; altered meridional supply of macro-nutrients as they are utilised in the iron-fertilised region and become unavailable for transport to, and utilisation in, other regions, fundamental alteration of food webs, biodiversity

| No data

| 12.3.1.3

|-
| Blue carbon management in coastal ecosystems

| 2–3

| Insufficient data, estimates range from ~100 to ~10,000

| &lt;1

| If degraded or lost, coastal blue carbon ecosystems are likely to release most of their carbon back to the atmosphere; potential for sediment contaminants, toxicity, bioaccumulation and biomagnification in organisms; issues related to altering degradability of coastal plants; use of subtidal areas for tidal wetland carbon removal; effect of shoreline modifications on sediment redeposition and natural marsh accretion; abusive use of coastal blue carbon as means to reclaim land for purposes that degrade capacity for carbon removal

| Potential for many non-climatic benefits and can contribute to ecosystem-based adaptation, coastal protection, increased biodiversity, reduced upper ocean acidification; could potentially benefit human nutrition or produce fertiliser for terrestrial agriculture, anti-methanogenic feed additive, or as an industrial or materials feedstock

| If degraded or lost, coastal blue carbon ecosystems are likely to release most of their carbon back to the atmosphere. The full delivery of the benefits at their maximum global capacity will require years to decades to be achieved

| Not incorporated in IAMs, but in some bottom-up studies: small contribution

| 12.3.1.3, 7.4

|-
| BECCS

| 5–6

| 15–400

| 0.5–11

| Competition for land and water resources, to grow biomass feedstock. Biodiversity and carbon stock loss if from unsustainable biomass harvest

| Reduction of air pollutants; fuel security, optimal use of residues, additional income, health benefits and if implemented well can enhance biodiversity, soil health and land carbon

| Competition for land with biodiversity conservation and food production

| Substantial contribution in IAMs and bottom-up sectoral studies

| 7.4

|-
| Afforestation/reforestation

| 8–9

| 0–240

| 0.5–10

| Reversal of carbon removal through wildfire, disease, pests may occur.

Reduced catchment water yield and lower groundwater level if species and biome are inappropriate

| Enhanced employment and local livelihoods, improved biodiversity, improved renewable wood products provision, soil carbon and nutrient cycling. Possibly less pressure on primary forest

| Inappropriate deployment at large scale can lead to competition for land with biodiversity conservation and food production

| Substantial contribution in IAMs and also in bottom-up sectoral studies

| 7.4

|-
| Biochar

| 6–7

| 10–345

| 0.3–6.6

| Particulate and GHG emissions from production; biodiversity and carbon stock loss from unsustainable biomass harvest

| Increased crop yields and reduced non-CO 2 emissions from soil; resilience to drought

| Environmental impacts associated with particulate matter; competition for biomass resource

| In development – not yet in global mitigation pathways simulated by IAMs

| 7.4

|-
| Soil carbon sequestration in croplands and grasslands

| 8–9

| -45–100

| 0.6–9.3

| Risk of increased nitrous oxide emissions due to higher levels of organic nitrogen in the soil; risk of reversal of carbon sequestration

| Improved soil quality, resilience and agricultural productivity

| Attempts to increase carbon sequestration potential at the expense of production. Net addition per hectare is very small; hard to monitor

| In development – not yet in global mitigation pathways simulated by IAMs; in bottom-up studies: with medium contribution

| 7.4

|-
| Peatland and coastal wetland restoration

| 8–9

| Insufficient data

| 0.5–2.1

| Reversal of carbon removal in drought or future disturbance. Risk of increased methane emissions

| Enhanced employment and local livelihoods, increased productivity of fisheries, improved biodiversity, soil carbon and nutrient cycling

| Competition for land for food production on some peatlands used for food production

| Not in IAMs but some bottom-up studies with medium contribution

| 7.4

|-
| Agroforestry

| 8–9

| Insufficient data

| 0.3–9.4

| Risk that some land area lost from food production; requires high skills

| Enhanced employment and local livelihoods, variety of products, improved soil quality, more resilient systems

| Some trade-off with agricultural crop production, but enhanced biodiversity, and resilience of system

| No data from IAMs, but in bottom-up sectoral studies. with medium contribution

| 7.4

|-
| Improved forest management

| 8–9

| Insufficient data

| 0.1–2.1

| If improved management is understood as merely intensification involving increased fertiliser use and introduced species, then it could reduce biodiversity and increase eutrophication

| In case of sustainable forest management, it leads to enhanced employment and local livelihoods, enhanced biodiversity, improved productivity

| If it involves increased fertiliser use and introduced species, it could reduce biodiversity and increase eutrophication and upstream GHG emissions

| No data from IAMs, but in bottom-up sectoral studies with medium contribution

| 7.4

|}

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==== 12.3.1.2 Enhanced Weathering ====

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Enhanced weathering involves (i) the mining of rocks containing minerals that naturally absorb CO 2 from the atmosphere over geological timescales (as they become exposed to the atmosphere through geological weathering), (ii) the comminution of these rocks to increase the surface area, and (iii) the spreading of these crushed rocks on soils (or in the ocean/coastal environments; [[#12.3.1.3|Section 12.3.1.3]] ) so that they react with atmospheric CO 2 ( [[#Schuiling--2006|Schuiling and Krijgsman 2006]] ; [[#Hartmann--2013|Hartmann et al. 2013]] ; [[#Beerling--2018|Beerling et al. 2018]] ; [[#Goll--2021|Goll et al. 2021]] ). Construction waste and waste materials from mining can also be used as a source material for enhanced weathering. Silicate rocks such as basalt, containing minerals rich in calcium and magnesium and lacking metal ions such as nickel and chromium, are most suitable for enhanced weathering ( [[#Beerling--2018|Beerling et al. 2018]] ); they reduce soil solution acidity during dissolution, and promote the chemical transformation of CO 2 to bicarbonate ions. The bicarbonate ions can precipitate in soils and drainage waters as a solid carbonate mineral ( [[#Manning--2008|Manning 2008]] ), or remain dissolved and increase alkalinity levels in the ocean when the water reaches the sea ( [[#Renforth--2017|Renforth and Henderson 2017]] ). The modelling study by [[#Cipolla--2021|Cipolla et al. (2021)]] found that rate of weathering is greater in high rainfall environments, and was increased by organic matter amendment.

'''Status:''' Enhanced weathering has been demonstrated in the laboratory and in small-scale field trials (TRL 3–4) but has yet to be demonstrated at scale ( [[#Beerling--2018|Beerling et al. 2018]] ; [[#Amann--2020|Amann et al. 2020]] ). The chemical reactions are well understood ( [[#Manning--2008|Manning 2008]] ; [[#Gillman--1980|Gillman 1980]] ; [[#Gillman--2001|Gillman et al. 2001]] ), but the behaviour of the crushed rocks in the field and potential co-benefits and adverse side effects of enhanced weathering require further research ( [[#Beerling--2018|Beerling et al. 2018]] ). Small-scale laboratory experiments have calculated weathering rates that are orders of magnitude slower than the theoretical limit for mass transfer-controlled forsterite ( [[#Renforth--2015|Renforth et al. 2015]] ; [[#Amann--2020|Amann et al. 2020]] ) and basalt dissolution ( [[#Kelland--2020|Kelland et al. 2020]] ). Uncertainty surrounding silicate mineral dissolution rates in soils, the fate of the released products, the extent of legacy reserves of mining by-products that might be exploited, location and availability of rock extraction sites, and the impact on ecosystems remain poorly quantified and require further research to better understand feasibility ( [[#Renforth--2012|Renforth 2012]] ; [[#Moosdorf--2014|Moosdorf et al. 2014]] ; [[#Beerling--2018|Beerling et al. 2018]] ). Closely monitored, large-scale demonstration projects would allow these aspects to be studied ( [[#Smith--2019a|Smith et al. 2019a]] ; [[#Beerling--2020|Beerling et al. 2020]] ).

'''Costs:''' [[#Fuss--2018|Fuss et al. (2018)]] , in a systematic review of the costs and potentials of CDR methods including enhanced weathering, note that costs are closely related to the source of the rock and the technology used for rock grinding and material transport ( [[#Renforth--2012|Renforth 2012]] ; [[#Hartmann--2013|Hartmann et al. 2013]] ; [[#Strefler--2018|Strefler et al. 2018]] ). Due to differences in the methods and assumptions between studies, literature ranges are highly uncertain and range from USD15–40 tCO 2 –1 to USD3460 tCO 2 –1 ( [[#Köhler--2010|Köhler et al. 2010]] ; [[#Taylor--2016|Taylor et al. 2016]] ). [[#Renforth--2012|Renforth (2012)]] reported operational costs in the UK of applying mafic rocks (rocks with high magnesium and iron silicate mineral concentrations) of USD70–578 tCO 2 –1 , and for ultramafic rocks (rocks rich in magnesium and iron silicate minerals but with very low silica content – the low silica content enhances weathering rates) of USD24–123 tCO 2 –1 . [[#Beerling--2020|Beerling et al. (2020)]] combined a spatially resolved weathering model with a techno-economic assessment to suggest costs of between USD54–220 tCO 2 –1 (with a weighted mean of USD118–128 tCO 2 –1 ). [[#Fuss--2018|Fuss et al. (2018)]] suggested an author judgement cost range of USD50–200 tCO 2 –1 for a potential of 2–4 GtCO 2 yr −1 from 2050, excluding biological storage.

'''Potentials:''' In a systematic review of the costs and potentials of enhanced weathering, [[#Fuss--2018|Fuss et al. (2018)]] report a wide range of potentials ( ''limited evidence'' , ''low agreement'' ). The highest reported regional sequestration potential, 88.1 GtCO 2 yr −1 , is reported for the spreading of pulverised rock over a very large land area in the tropics, a region considered promising given the higher temperatures and greater rainfall ( [[#Taylor--2016|Taylor et al. 2016]] ). Considering cropland areas only, the potential carbon removal was estimated by [[#Strefler--2018|Strefler et al. (2018)]] to be 95 GtCO 2 yr −1 for dunite and 4.9 GtCO 2 yr −1 for basalt. Slightly lower potentials were estimated by [[#Lenton--2014|Lenton (2014)]] where the potential of carbon removal by enhanced weathering (including adding carbonate and olivine to both oceans and soils) was estimated to be 3.7 GtCO 2 yr –1 by 2100, but with mean annual removal an order of magnitude less at 0.2 GtC-eq yr –1 ( [[#Lenton--2014|Lenton 2014]] ). The estimates reported in [[#Smith--2016|Smith et al. (2016)]] are based on the potential estimates of [[#Lenton--2014|Lenton (2014)]] . [[#Beerling--2020|Beerling et al. (2020)]] estimate that up to 2 GtCO 2 yr –1 could be removed by 2050 by spreading basalt onto 35–59% (weighted mean 53%) of agricultural land of 12 countries. [[#Fuss--2018|Fuss et al. (2018)]] provide an author judgement range for potential of 2–4 GtCO 2 yr −1 for 2050.

'''Risks and impacts:''' Mining of rocks for enhanced weathering will have local impacts and carries risks similar to those associated with the mining of mineral construction aggregates, with the possible additional risk of greater dust generation from fine comminution and land application. In addition to direct habitat destruction and increased traffic to access mining sites, there could be adverse impacts on local water quality ( [[#Younger--2004|Younger and Wolkersdorfer 2004]] ).

'''Co-benefits:''' Enhanced weathering can improve plant growth by pH modification and increased mineral supply ( [[#Kantola--2017|Kantola et al. 2017]] ; [[#Beerling--2018|Beerling et al. 2018]] ), can enhance SCS in some soils ( [[#Beerling--2018|Beerling et al. 2018]] ) thereby protecting against soil erosion ( [[#Wright--1998|Wright and Upadhyaya 1998]] ), and increasing the cation exchange capacity, resulting in increased nutrient retention and availability ( [[#Gillman--1980|Gillman 1980]] ; [[#Baldock--2000|Baldock and Skjemstad 2000]] ; [[#Gillman--2001|Gillman et al. 2001]] ; [[#Manning--2010|Manning 2010]] ; [[#Guntzer--2012|Guntzer et al. 2012]] ; [[#Tubana--2016|Tubana et al. 2016]] ; [[#Yu--2017|Yu et al. 2017]] ; [[#Haque--2019|Haque et al. 2019]] ; [[#Smith--2019a|Smith et al. 2019a]] ). Through these actions, it can contribute to SDG 2 (zero hunger), SDG 15 (life on land) (by reducing land demand for croplands), SDG 13 (climate action) (through CDR), SDG 14 (life below water) (by ameliorating ocean acidification) and SDG 6 (clean water and sanitation) ( [[#Smith--2019a|Smith et al. 2019a]] ). To more directly ameliorate ocean acidification while increasing CDR and reducing impacts on land ecosystems, alkaline minerals could instead be directly added to the ocean ( [[#12.3.1.3|Section 12.3.1.3]] ). There are potential benefits in poverty reduction through employment of local workers in mining ( [[#Pegg--2006|Pegg 2006]] ).

'''Trade-offs and spillover effects:''' Air quality could be adversely affected by the spreading of rock dust ( [[#Edwards--2017|Edwards et al. 2017]] ), though this can partly be ameliorated by water-spraying ( [[#Grundnig--2006|Grundnig et al. 2006]] ). As noted above, any significant expansion of the mining industry would require careful assessment to avoid possible detrimental effects on biodiversity ( [[#Amundson--2015|Amundson et al. 2015]] ). The processing of an additional 10 billion tonnes of rock would require up to 3000 Terawatt-hours of energy, which could represent approximately 0.1–6 % of global electricity use in 2100. The emissions associated with this additional energy generation may reduce the net carbon dioxide removal by up to 30% with present-day grid average emissions, but this efficiency loss would decrease with low-carbon power ( [[#Beerling--2020|Beerling et al. 2020]] ).

'''Role in mitigation pathways:''' Only one study to date has included enhanced weathering in an integrated assessment model to explore mitigation pathways ( [[#Strefler--2021|Strefler et al. 2021]] ).

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

<div id="12.3.1.3" class="h3-container"></div>

<span id="ocean-based-methods"></span>
==== 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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<span id="feasibility-assessment"></span>
==== 12.3.1.4 Feasibility Assessment ====

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Following the framework presented in [[IPCC:Wg3:Chapter:Chapter-6#6.4|Section 6.4]] and Annex II, Part IV, Section 11, a multi-dimensional feasibility assessment of the CDR methods covered here is provided in Figure 12.4, taking into account the assessment presented in this section. Both DACCS and EW perform positively on the geophysical and technological dimensions while for ocean-based approaches performance is mixed. There is limited evidence to assess social-cultural, environmental/ecological, and institutional dimensions as the literature is still nascent for DACCS and EW, while these aspects are positive for blue carbon and mixed or negative for ocean fertilisation. On the economic dimension, the cost is assessed negatively for all CDR methods.

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'''Figure 12.4 | Summary of the extent to which different factors would enable or inhibit the deployment of the carbon dioxide removal methods DACCS, EW, ocean fertilisation and blue carbon management.''' Blue bars indicate the extent to which the indicator enables the implementation of the CDR method (E) and orange bars indicate the extent to which an indicator is a barrier (B) to the deployment of the method, relative to the maximum possible barriers and enablers assessed. An ‘X’ signifies the indicator is not applicable or does not affect the feasibility of the method, while a forward slash indicates that there is no or limited evidence whether the indicator affects the feasibility of the method. The shading indicates the level of confidence, with darker shading signifying higher levels of confidence. Supplementary Material 12.SM.B provides an overview of the factors affecting the feasibility of CDR methods and how they differ across contexts (e.g., region), time (e.g., 2030 versus 2050), and scale (e.g., small versus large), and includes a line of sight on which the assessment is based. The assessment methodology is explained in Annex II, Part IV, Section 11.

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