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

== 12.3 Carbon Dioxide Removal ==

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Carbon dioxide removal (CDR) refers to a cluster of technologies, practices, and approaches that remove and sequester carbon dioxide from the atmosphere and durably store the carbon in geological, terrestrial, or ocean reservoirs, or in products. Despite the common feature of removing carbon dioxide, CDR methods can be very different ( [[#Smith--2017|Smith et al. 2017]] ). There are proposed methods for removal of non-CO 2 greenhouse gases such as methane ( [[#Jackson--2019|Jackson et al. 2019]] ; [[#Jackson--2021|Jackson et al. 2021]] ) but scarcity of literature on these methods prevents assessment here.

A number of CDR methods (e.g., afforestation/reforestation (A/R), bioenergy with carbon capture and storage (BECCS), soil carbon sequestration (SCS), biochar, wetland/peatland restoration and coastal restoration) are dealt with elsewhere in this report (Chapters 6 and 7). These methods are synthesised in [[#12.3.2|Section 12.3.2]] . Others, not dealt with elsewhere, – direct air carbon capture and storage (DACCS), enhanced weathering (EW) of minerals and ocean-based approaches including ocean fertilisation (OF) and ocean alkalinity enhancement (OAE) – are discussed in Sections 12.3.1.1 to 12.3.1.3 below (see also [[#IPCC--2019b|IPCC 2019b]] and AR6 WGI, [[IPCC:Wg3:Chapter:Chapter-5#5.6|Section 5.6]] ). Some methods, such as BECCS and DACCS, involve carbon storage in geological formations, which is discussed in Chapter 6. The climate system and the carbon cycle responses to CDR deployment and each method’s physical and biogeochemical characteristics such as storage form and duration are assessed in Chapters 4 and 5 of the AR6 WGI report.

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=== Cross-Chapter Box 8 | Carbon Dioxide Removal: Key Characteristics and Multiple Roles in Mitigation Strategies ===

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'''Authors:''' Oliver Geden (Germany), Alaa Al Khourdajie (United Kingdom/Syria), Christopher Bataille (Canada), Göran Berndes (Sweden), Holly Jean Buck (the United States of America), Katherine Calvin (the United States of America), Annette Cowie (Australia), Kiane de Kleijne (the Netherlands), Jan Christoph Minx (Germany), Gert-Jan Nabuurs (the Netherlands), Glen P. Peters (Norway/Australia), Andy Reisinger (New Zealand), Pete Smith (United Kingdom), Masahiro Sugiyama (Japan)

Carbon dioxide removal (CDR) is a necessary element of mitigation portfolios to achieve net zero CO 2 and GHG emissions both globally and nationally, counterbalancing residual emissions from hard-to-transition sectors such as industry, transport and agriculture. CDR is a key element in scenarios that limit warming to 2°C (&gt;67%) or lower, regardless of whether global emissions reach near-zero, net zero or net-negative levels (Sections 3.3, 3.4, 3.5 and 12.3). While national mitigation portfolios aiming at net zero or net-negative emissions will need to include some level of CDR, the choice of methods and the scale and timing of their deployment will depend on the ambition for gross emissions reductions, how sustainability and feasibility constraints are managed, and how political preferences and social acceptability evolve ( [[#12.3.3|Section 12.3.3]] ). This box gives an overview of CDR methods, presents a categorisation based on the key characteristics of removal processes and storage timescales, and clarifies the multiple roles of CDR in mitigation strategies. The term ‘negative emissions’ is used in this report only when referring to the net emissions outcome at a systems level (e.g., ‘net negative emissions’ at global, national, sectoral or supply chain levels).

'''Categorisation of the m''' '''ain CDR methods'''

CDR refers to anthropogenic activities that remove CO 2 from the atmosphere and store it durably in geological, terrestrial, or ocean reservoirs, or in products. It includes anthropogenic enhancement of biological, geochemical or chemical CO 2 sinks, but excludes natural CO 2 uptake not directly caused by human activities. Increases in land carbon sink strength due to CO 2 fertilisation or other indirect effects of human activities are not considered CDR (see Glossary). Carbon capture and storage (CCS) and carbon capture and utilisation (CCU) applied to CO 2 from fossil fuel use are not CDR methods as they do not remove CO 2 from the atmosphere. CCS and CCU can, however, be part of CDR methods if the CO 2 has been captured from the atmosphere, either indirectly in the form of biomass or directly from ambient air, and stored durably in geological reservoirs or products (Sections 11.3.6 and 12.3).

There are many different CDR methods and associated implementation options (Cross-Chapter Box 8, Figure 1). Some of these methods (including afforestation and improved forest management, wetland restoration and soil carbon sequestration (SCS)) have been practised for decades to millennia, although not necessarily with the intention of removing carbon from the atmosphere. Conversely, methods such as direct air carbon capture and storage (DACCS), bioenergy with carbon capture and storage (BECCS) and enhanced weathering are novel, and while experience is growing, their demonstration and deployment are limited in scale. CDR methods have been categorised in different ways in the literature, highlighting different characteristics. In this report, as in AR6 WGI, the categorisation is based on the role of CDR methods in the carbon cycle, that is, on the removal process ( ''land-based biological'' ; ''ocean-based biological'' ; ''geochemical'' ; ''chemical'' ) and on the timescale of storage ( ''decades to centuries'' ; ''centuries to millennia'' ; ''ten thousand years or longer'' ). The time scale of storage is closely linked to the storage medium: carbon stored in ocean reservoirs (through enhanced weathering, ocean alkalinity enhancement or ocean fertilisation) and in geological formations (through BECCS or DACCS) generally has longer storage times and is less vulnerable to reversal through human actions or disturbances such as drought and wildfire than carbon stored in terrestrial reservoirs (vegetation, soil). Furthermore, carbon stored in vegetation or through SCS has shorter storage times and is more vulnerable than carbon stored in buildings as wood products; as biochar in soils, cement and other materials; or in chemical products made from biomass or potentially through direct air ( [[#Fuss--2018|Fuss et al. 2018]] ; [[#Minx--2018|Minx et al. 2018]] ; [[#NASEM--2019|NASEM 2019]] ) capture ( [[IPCC:Wg3:Chapter:Chapter-11#11.3.6|Section 11.3.6]] ; AR6 WGI, Figure 5.36). Within the same category (e.g., land-based biological CDR) options often differ with respect to other dynamic or context-specific dimensions, such as mitigation potential, cost, potential for co-benefits and adverse side effects, and technology readiness level (Table 12.6).

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'''Cross-Chapter Box 8, Figure 1 | Carbon dioxide removal taxonomy.''' '''Methods are categorised based on removal process (grey shades) and storage medium (for which timescales of storage are given, yellow/brown shades).''' Main implementation options are included for each CDR method. Note that specific land-based implementation options can be associated with several CDR methods, for example, agroforestry can support soil carbon sequestration and provide biomass for biochar or BECCS. Source: adapted from [[#Minx--2018|Minx et al. (2018)]] .

'''Roles of CDR in mitiga''' '''tion strategies'''

Within ambitious mitigation strategies at global or national levels, CDR cannot serve as a substitute for deep emissions reductions but can fulfil multiple complementary roles: it can (i) further reduce net CO 2 or GHG emission levels in the near-term; (ii) counterbalance residual emissions from hard-to-transition sectors, such as CO 2 from industrial activities and long-distance transport (e.g., aviation, shipping), or methane and nitrous oxide from agriculture, in order to help reach net zero CO 2 or GHG emissions in the mid-term; (iii) achieve and sustain net-negative CO 2 or GHG emissions in the long-term, by deploying CDR at levels exceeding annual residual gross CO 2 or GHG emissions (Sections 2.7.3 and 3.5).

In general, these roles of CDR are not mutually exclusive and can exist in parallel. For example, achieving net zero CO 2 or GHG emissions globally might involve some countries already reaching net-negative levels at the time of global net zero, allowing other countries more time to achieve this. Equally, achieving net-negative CO 2 emissions globally, which could address a potential temperature overshoot by lowering atmospheric CO 2 concentrations, does not necessarily involve all countries reaching net-negative levels ( [[#Rajamani--2021|Rajamani et al. 2021]] ; [[#Rogelj--2021|Rogelj et al. 2021]] ) (Cross-Chapter Box 3 in Chapter 3).

Cross-Chapter Box 8, Figure 2 shows these multiple roles of CDR in a stylised ambitious mitigation pathway that can be applied to global and national levels. While such mitigation pathways will differ in their shape and exact composition, they include the same basic components: CO 2 emissions from fossil sources, CO 2 emissions from managed land, non-CO 2 emissions, and various forms of CDR. Cross-Chapter Box 8, Figure 2 also illustrates the importance of distinguishing between gross CO 2 removals from the atmosphere through deployment of CDR methods and the net emissions outcome (i.e., gross emissions minus gross removals).

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'''Cross-Chapter Box 8, Figure 2 | Roles of CDR in global or national mitigation strategies.''' Stylised pathway showing multiple functions of CDR in different phases of ambitious mitigation: (1) further reducing net CO 2 or GHG emissions levels in near-term; (2) counterbalancing residual emissions to help reach net zero CO 2 or GHG emissions in the mid-term; (3) achieving and sustaining net-negative CO 2 or GHG emissions in the long-term.

CDR methods currently deployed on managed land, such as afforestation or reforestation and improved forest management, lead to CO 2 removals already today, even when net emissions from land use are still positive, for example, when gross emissions from deforestation and draining peatlands exceed gross removals from afforestation or reforestation and ecosystem conservation (Sections 2.2 and 7.2; Cross-Chapter Box 6 in Chapter 7). As there are currently no removal methods for non-CO 2 gases that have progressed beyond conceptual discussions ( [[#Jackson--2021|Jackson et al. 2021]] ), achieving net zero GHG implies gross CO 2 removals to counterbalance residual emissions of both CO 2 and non-CO 2 gases, applying 100-year global warming potential (GWP100) as the metric for reporting CO 2 -equivalent emissions, as required for emissions reporting under the Rulebook of the Paris Agreement (Cross-Chapter Box 2 in Chapter 2).

Net zero CO 2 emissions will be achieved earlier than net zero GHG emissions. As volumes of residual non-CO 2 emissions are expected to be significant, this time-lag could reach one to several decades, depending on the respective size and composition of residual GHG emissions at the time of net zero CO 2 emissions. Furthermore, counterbalancing residual non-CO 2 emissions by CO 2 removals will lead to net-negative CO 2 emissions at the time of net zero GHG emissions (Cross-Chapter Box 3 in Chapter 3).

While many governments have included A/R and other forestry measures in their NDCs under the Paris Agreement ( [[#Moe--2018|Moe and Røttereng 2018]] ; [[#Fyson--2019|Fyson and Jeffery 2019]] ; [[#Mace--2021|Mace et al. 2021]] ), and a few countries also mention BECCS, DACCS and enhanced weathering in their mid-century low emission development strategies ( [[#Buylova--2021|Buylova et al. 2021]] ), very few are pursuing the integration of a broad range of CDR methods into national mitigation portfolios so far ( [[#Schenuit--2021|Schenuit et al. 2021]] ) (Box 12.1). There are concerns that the prospect of large-scale CDR could, depending on the design of mitigation strategies, obstruct near-term emissions reduction efforts ( [[#Lenzi--2018|Lenzi et al. 2018]] ; [[#Markusson--2018|Markusson et al. 2018]] ), mask insufficient policy interventions ( [[#Geden--2016|Geden 2016]] ; [[#Carton--2019|Carton 2019]] ), might lead to an overreliance on technologies that are still in their infancy ( [[#Anderson--2016|Anderson and Peters 2016]] ; [[#Larkin--2018|Larkin et al. 2018]] ; [[#Grant--2021|Grant et al. 2021]] ), could overburden future generations ( [[#Lenzi--2018|Lenzi 2018]] ; [[#Shue--2018|Shue 2018]] ; [[#Bednar--2019|Bednar et al. 2019]] ) might evoke new conflicts over equitable burden-sharing ( [[#Pozo--2020|Pozo et al. 2020]] ; [[#Lee--2021|Lee et al. 2021]] ; [[#Mohan--2021|Mohan et al. 2021]] ), could impact food security, biodiversity or land rights ( [[#Buck--2016|Buck 2016]] ; [[#Boysen--2017|Boysen et al. 2017]] ; [[#Dooley--2018|Dooley and Kartha 2018]] ; [[#Hurlbert--2019|Hurlbert et al. 2019]] ; [[#Dooley--2021|Dooley et al. 2021]] ), or might be perceived negatively by stakeholders and broader public audiences (Royal Society and Royal Academy of Engineering 2018; [[#Colvin--2020|Colvin et al. 2020]] ). Conversely, without considering different timescales of carbon storage ( [[#Fuss--2018|Fuss et al. 2018]] ; [[#Hepburn--2019|Hepburn et al. 2019]] ) and implementation of reliable measurement, reporting and verification of carbon flows ( [[#Mace--2021|Mace et al. 2021]] ), CDR deployment might not deliver the intended benefit of removing CO 2 durably from the atmosphere. Furthermore, without appropriate incentive schemes and market designs ( [[#Honegger--2021b|Honegger et al. 2021b]] ), CDR implementation options could see under-investment. The many challenges in research, development and demonstration of novel approaches, to advance innovation according to broader societal objectives and to bring down costs, could delay their scaling up and deployment ( [[#Nemet--2018|Nemet et al. 2018]] ). Depending on the scale and deployment scenario, CDR methods could bring about various co-benefits and adverse side effects (see below). All this highlights the need for appropriate CDR governance and policies ( [[#12.3.3|Section 12.3.3]] ).

The volumes of future global CDR deployment assumed in IAM-based mitigation scenarios are large compared to current volumes of deployment, which presents a challenge since rapid and sustained upscaling from a small base is particularly difficult ( [[#de%20Coninck--2018|de Coninck et al. 2018]] ; [[#Nemet--2018|Nemet et al. 2018]] ; [[#Hanna--2021|Hanna et al. 2021]] ). All Illustrative Mitigation Pathways (IMPs) that limit warming to 2°C (&gt;67%) or lower use some form of CDR. Across the full range of similarly ambitious IAM scenarios (scenario categories C1 to C3; see [[IPCC:Wg3:Chapter:Chapter-3#3.3|Section 3.3]] ), the reported annual CO 2 removal from AFOLU (mainly A/R) reaches 0.86 [0.01–4.11] GtCO 2 yr –1 by 2030, 2.98 [0.23–6.38] GtCO 2 yr –1 by 2050, and 4.19 [0.1–6.91] GtCO 2 yr –1 by 2100 (values are the medians and bracketed values denote the 5–95th percentile range [[#footnote-003|1]] ). The annual BECCS deployment is 0.08 [0–1.09] GtCO 2 yr –1 , 2.75 [0.52–9.45] GtCO 2 yr –1 , and 8.96 [2.63–16.15] GtCO 2 yr –1 for these years, respectively. The annual DACCS deployment eaches 0 [0–0.02] GtCO 2 yr –1 by 2030, 0.02 [0–1.74] GtCO 2 yr –1 by 2050, and 1.02 [0–12.6] GtCO 2 yr –1 by 2100 (Figure 12.3). [[#footnote-002|2]] Reported cumulative volumes of BECCS, CO 2 removal from AFOLU, and DACCS reach 328 [168–763] GtCO 2 , 252 [20–418] GtCO 2 , and 29 [0–339] GtCO 2 for the 2020–2100 period, respectively. Reaching the higher end of CDR volumes is subject to issues regarding their feasibility (see below), especially if achieved with only a limited number of CDR methods. Recent studies have identified some drivers for large-scale CDR deployment in IAM scenarios, including insufficient representation of variable renewables, a high discount rate that tends to increase initial carbon budget overshoot and therefore inflates usage of CDR to achieve net-negative emissions at later times, omission of CDR methods aside from BECCS and A/R ( [[#Emmerling--2019|Emmerling et al. 2019]] ; [[#Hilaire--2019|Hilaire et al. 2019]] ; [[#Köberle--2019|Köberle 2019]] ), and limited deployment of demand-side options ( [[#Grubler--2018|Grubler et al. 2018]] ; [[#van%20Vuuren--2018|van Vuuren et al. 2018]] ; [[#Daioglou--2019|Daioglou et al. 2019]] ). The levels of CDR in IAMs in modelled pathways would change depending on the allowable overshoot of policy targets such as temperature or radiative forcing and the costs of non-CDR mitigation options ( [[#Johansson--2020|Johansson et al. 2020]] ; [[#van%20der%20Wijst--2021|van der Wijst et al. 2021]] ) ( [[IPCC:Wg3:Chapter:Chapter-3#3.2.2|Section 3.2.2]] ).

While many CDR methods are gradually being explored, IAM scenarios have focused mostly on BECCS and A/R ( [[#Tavoni--2013|Tavoni and Socolow 2013]] ; [[#Fuhrman--2019|Fuhrman et al. 2019]] ; [[#Rickels--2019|Rickels et al. 2019]] ; [[#Calvin--2021|Calvin et al. 2021]] ; [[#Diniz%20Oliveira--2021|Diniz Oliveira et al. 2021]] ). Although some IAM studies have also included other methods such as DACCS ( [[#Chen--2013|Chen and Tavoni 2013]] ; [[#Marcucci--2017|Marcucci et al. 2017]] ; [[#Realmonte--2019|Realmonte et al. 2019]] ; [[#Fuhrman--2020|Fuhrman et al. 2020]] ; [[#Akimoto--2021|Akimoto et al. 2021]] ; [[#Fuhrman--2021a|Fuhrman et al. 2021a]] ), enhanced weathering ( [[#Strefler--2021|Strefler et al. 2021]] ), SCS and biochar ( [[#Holz--2018|Holz et al. 2018]] ) there is much less literature compared to studies on BECCS ( [[#Hilaire--2019|Hilaire et al. 2019]] ). A large-scale coordinated IAM study on BECCS (‘EMF-33’) has been conducted ( [[#Muratori--2020|Muratori et al. 2020]] ; [[#Rose--2020|Rose et al. 2020]] ) but none exists for other CDR methods. A recent review proposes a combination of various CDR methods ( [[#Fuss--2018|Fuss et al. 2018]] ) but more in-depth literature on such a portfolio approach is limited ( [[#Strefler--2021|Strefler et al. 2021]] ). A multi-criteria analysis has identified pathways with CDR portfolios different from least-cost pathways often dominated by BECCS and A/R ( [[#Rueda--2021|Rueda et al. 2021]] ).

At the national and regional levels, the role of land-based biological CDR methods has long been analysed, but there is little detailed techno-economic assessment of the role of other CDR methods. There is a small but emerging literature providing such assessments for developed countries ( [[#Kraxner--2014|Kraxner et al. 2014]] ; [[#Baik--2018|Baik et al. 2018]] ; [[#Daggash--2018|Daggash et al. 2018]] ; [[#Patrizio--2018|Patrizio et al. 2018]] ; [[#Sanchez--2018|Sanchez et al. 2018]] ; [[#Breyer--2019|Breyer et al. 2019]] ; [[#Kato--2019|Kato and Kurosawa 2019]] ; [[#Larsen--2019|Larsen et al. 2019]] ; [[#McQueen--2020|McQueen et al. 2020]] ; [[#Bistline--2021|Bistline and Blanford 2021]] ; [[#García-Freites--2021|García-Freites et al. 2021]] ; [[#Jackson--2021|Jackson et al. 2021]] ; [[#Kato--2021|Kato and Kurosawa 2021]] ; [[#Negri--2021|Negri et al. 2021]] ) while the literature outside developed countries is limited ( [[#Alatiq--2021|Alatiq et al. 2021]] ; [[#Fuhrman--2021b|Fuhrman et al. 2021b]] ; [[#Weng--2021|Weng et al. 2021]] ).

In IAMs, CDR is contributed mainly by the energy sector (through BECCS) and AFOLU (through A/R) (Figure 12.3). IAMs are starting to include other CDR methods, such as DACCS and enhanced weathering ( [[#12.3.1|Section 12.3.1]] ), which are yet to be attributed to specific sectors in IAMs. Following IPCC guidance for UNFCCC inventories, A/R and SCS are reported in land use, land-use change and forestry (LULUCF), while BECCS would be reported in the sector where the carbon capture occurs, that is, the energy sector in the case of electricity and heat production, and the industry sector for BECCS linked to manufacturing (e.g., steel or hydrogen) ( [[#Tanzer--2020|Tanzer et al. 2020]] ; Bui et al. 2021; [[#Tanzer--2021|Tanzer et al. 2021]] ).

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'''Figure 12.3 | Sequestration through three predominant CDR methods: BECCS, CO''' 2 '''removal from AFOLU (mainly A/R), and DACCS (upper panels) annual sequestration and (lower panels) cumulative sequestration.''' The IAM scenarios described in the figure correspond to those that limit warming to 2°C (&gt;67%) or lower. The black line in each of the upper panels indicates the median of all the scenarios in categories C1 to C3. Hinges in the lower panels represent the interquartile ranges while whiskers extend to 5th and 95th percentiles. The IMPs are highlighted with colours, as shown in the key. The number of scenarios is indicated in the header of each panel. The number of scenarios with a non-zero DACCS value is 146.

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=== 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

|}

<div id="12.3.1.2" class="h3-container"></div>

<span id="enhanced-weathering"></span>
==== 12.3.1.2 Enhanced Weathering ====

<div id="h3-2-siblings" class="h3-siblings"></div>

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 ====

<div id="h3-3-siblings" class="h3-siblings"></div>

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.

<div id="12.3.1.4" class="h3-container"></div>

<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.

<div id="_idContainer009ee" class="_idGenObjectStyleOverride-1"></div>

[[File:7546fafe5ddda8b2adef134392bc12dd IPCC_AR6_WGIII_Figure_12_4.png]]

'''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.

<div id="12.3.2" class="h2-container"></div>

<span id="consideration-of-methods-assessed-in-sectoral-chapters-ar-biochar-beccs-soil-carbon-sequestration"></span>
=== 12.3.2 Consideration of Methods Assessed in Sectoral Chapters: A/R, Biochar, BECCS, Soil Carbon Sequestration ===

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'''Status:''' BECCS, afforestation/reforestation (A/R), soil carbon sequestration (SCS) and biochar are land-based biological CDR methods ( [[#Smith--2016|Smith et al. 2016]] ). BECCS combines biomass use for energy with CCS to capture and store the biogenic carbon geologically ( [[IPCC:Wg3:Chapter:Chapter-6#6.4.2.6|Section 6.4.2.6]] ); A/R and SCS involve fixing atmospheric carbon in biomass and soils, and biochar involves converting biomass to biochar and using it as a soil amendment. These CDR methods can be associated with both co-benefits and adverse side effects ( [[#Smith--2016|Smith et al. 2016]] ; [[#Hurlbert--2019|Hurlbert et al. 2019]] ; [[#Mbow--2019|Mbow et al. 2019]] ; [[#Olsson--2019|Olsson et al. 2019]] ; [[#Schleicher--2019|Schleicher et al. 2019]] ; [[#Smith--2019b|Smith et al. 2019b]] ; [[#Babin--2021|Babin et al. 2021]] ; [[#Dooley--2021|Dooley et al. 2021]] ) (Sections 7.4 and 12.5).

Among CDR methods, BECCS and A/R are most commonly selected by IAMs to meet the requirements of scenarios that limit warming to 2°C (&gt;67%) or lower. This is partially because of the long lead time required to refine IAMs to include additional methods and update techno-economic parameters. Currently, few IAMs represent SCS or biochar ( [[#Frank--2017|Frank et al. 2017]] ). Given the removal potential of SCS and biochar and some potential co-benefits, more efforts should be made to include these methods within IAMs, so that their mitigation potential can be compared to other CDR methods, along with possible co-benefits and adverse side effects ( [[#Smith--2016|Smith et al. 2016]] ; [[#Rogelj--2018|Rogelj et al. 2018]] ) ( [[#12.5|Section 12.5]] ).

'''Potential:''' The technical potential for BECCS by 2050 is estimated at 0.5–11.3 GtCO 2 -eq yr –1 (Table 7.3). These potentials do not include avoided emissions resulting from the use of heat, electricity and/or fuels provided by the BECCS system, which depend on substitution patterns, conversion efficiencies, and supply chain emissions for the BECCS and substituted energy systems (Box 7.7). The mitigation effect of BECCS also depends on how deployment affects land carbon stocks and sink strength ( [[IPCC:Wg3:Chapter:Chapter-7#7.4.4|Section 7.4.4]] ).

As detailed in Chapter 7, the technical potential for gross removals realised through A/R in 2050 is 0.5–10.1 GtCO 2 -eq yr –1 , and for improved forest management the potential is 1–2.1 GtCO 2 -eq yr –1 (including both CDR and emissions reduction). Technical potential for SCS in 2050 is estimated to be 0.6–9.4 GtCO 2 -eq yr –1 , for agroforestry it is 0.3–9.4 GtCO 2 -eq yr –1 , and for biochar it is 0.2–6.6 GtCO 2 -eq yr –1 . Peatland and coastal wetland restoration have a technical potential of 0.5–2.1 GtCO 2 -eq yr –1 in 2050, with an estimated 80% of the potential being CDR. Note that these potentials reflect only biophysical and technological conditions and become reduced when factoring in economic, environmental, socio-cultural and institutional constraints (Table 12.6).

'''Costs:''' Costs across technologies vary substantially ( [[#Smith--2016|Smith et al. 2016]] ) and were estimated to be USD15–400 tCO 2 –1 for BECSS, USD0–240 tCO 2 –1 for A/R, –USD45 to +USD100 tCO 2 –1 for SCS and USD10–345 tCO 2 –1 for biochar. [[#Fuss--2018|Fuss et al. (2018)]] estimated abatement cost ranges for BECCS, A/R, SCS and biochar to be 100–200, 5–50, 0–100, and 30–120 tCO 2 -eq −1 respectively, corresponding to 2100 potentials. Ranges for economic potential (&lt;USD100 tCO 2 –1 ) reported in [[IPCC:Wg3:Chapter:Chapter-7|Chapter 7]] are 0.5–3.0 GtCO 2 yr –1 (A/R); 0.6–1.9 GtCO 2 yr –1 (improved forest management); 0.7–2.5 GtCO 2 yr –1 (SCS); 0.4–1.1 GtCO 2 yr –1 (agroforestry); 0.3–1.8 GtCO 2 yr –1 (biochar); and 0.2–0.8 GtCO 2 yr –1 (peatland and coastal wetland restoration).

'''Risks, impacts, and co-benefits:''' a brief summary of risks, impacts and co-benefits is provided here and more detail is provided in [[IPCC:Wg3:Chapter:Chapter-7|Chapter 7]] and [[#12.5|Section 12.5]] . A/R and biomass production for BECCS and biochar potentially compete for land, water and other resources, implying possible adverse outcomes for ecosystem health, biodiversity, livelihoods and food security ( ''medium evidence'' , ''high agreement'' ) ( [[#Smith--2016|Smith et al. 2016]] ; [[#Heck--2018|Heck et al. 2018]] ; [[#Hurlbert--2019|Hurlbert et al. 2019]] ; [[#Mbow--2019|Mbow et al. 2019]] ) (Chapter 7). SCS requires the addition of nitrogen and phosphorus to maintain stoichiometry of soil organic matter, leading to a potential risk of eutrophication ( [[#Fuss--2018|Fuss et al. 2018]] ). Apart from possible negative effects associated with biomass supply, adverse side effects from biochar are relatively low if the biomass is uncontaminated ( [[#Tisserant--2019|Tisserant and Cherubini 2019]] ).

Possible climate risks relate to direct and/or indirect land carbon losses (A/R, BECCS, biochar), increased N 2 O emissions (BECCS, SCS), saturation and non-permanence of carbon storage (A/R, SCS) ( [[#Jia--2019|Jia et al. 2019]] ; [[#Smith--2019b|Smith et al. 2019b]] ) (Chapter 7), and potential CO 2 leakage from deep geological reservoirs (BECCS) (Chapter 6). Land cover change associated with A/R and biomass supply for BECCS and biochar may cause albedo changes that reduce mitigation effectiveness ( [[#Fuss--2018|Fuss et al. 2018]] ; [[#Jia--2019|Jia et al. 2019]] ). Potentially unfavourable albedo change resulting from biochar use can be minimised by incorporating biochar into the soil ( [[#Fuss--2018|Fuss et al. 2018]] ) (Chapter 7).

Concerning co-benefits, A/R and biomass production for BECCS or biochar could improve soil carbon, nutrient and water cycling ( ''robust evidence'' , ''high agreement'' ), and contribute to market opportunities, employment and local livelihoods, economic diversification, energy security, and technology development and transfer ( ''medium evidence'' , ''high agreement'' ) ( [[#Fuss--2018|Fuss et al. 2018]] ) (Chapter 7). It may contribute to reduction of other air pollutants, health benefits, and reduced dependency on imported fossil fuels. A/R can improve biodiversity if native and diverse species are used ( [[#Fuss--2018|Fuss et al. 2018]] ). For biochar, additional co-benefits include increased crop yields, reduced drought impacts, and reduced CH 4 and N 2 O emissions from soils ( [[#Joseph--2021|Joseph et al., 2021]] ) ( [[IPCC:Wg3:Chapter:Chapter-7#7.4.5.2|Section 7.4.5.2]] ). SCS can improve soil quality and resilience and improve agricultural productivity and food security ( [[#Frank--2017|Frank et al. 2017]] ; [[#Smith--2019b|Smith et al. 2019b]] ).

'''Role in mitigation pathways:''' Biomass use for BECCS in 2050 is 61 EJ yr –1 (13–208 EJ yr –1 , 5–95th percentile range) in scenarios limiting warming to 1.5°C (&gt;50%) with no or limited overshoot (C1, excluding traditional energy). This corresponds to 5.3 GtCO 2 yr –1 (1.1–18 GtCO 2 yr –1 ) CDR, if assuming 28 kg C GJ –1 biomass carbon content and 85% capture rate in BECCS systems. In scenarios that limit warming to 2°C (&gt;67%) (C3), biomass use for BECCS in 2050 is 28 EJ yr –1 (0–96 EJ yr –1 , 5–95th percentile range), corresponding to 2.4 GtCO 2 yr –1 (0–8.3 GtCO 2 yr –1 ) CDR. Cumulative CO 2 removal from AFOLU (mainly through A/R), as reported from models, in the period 2020 to 2100 is 262 GtCO 2 (17–397 GtCO 2 ) and 209 GtCO 2 (20–415 GtCO 2 ) in C1 and C3 scenarios, respectively (5–95th percentile range).

Uncertainties remain in two main areas: the availability of land and biomass, which is affected by many factors ( [[#Anandarajah--2018|Anandarajah et al. 2018]] ) (Chapter 7), and the role of other mitigation measures including CDR methods other than A/R and BECCS. Strong near-term climate change mitigation to limit overshoot, and deployment of CDR methods other than A/R and BECCS, may significantly reduce the contribution of these CDR methods in scenarios limiting warming to 1.5°C or 2°C ( [[#Köberle--2019|Köberle 2019]] ; [[#Hasegawa--2021|Hasegawa et al. 2021]] ).

'''Trade-offs and spillovers:''' Some land-based biological CDR methods, such as BECCS and A/R, demand land. Combining mitigation strategies has the potential to increase overall carbon sequestration rates ( [[#Humpenöder--2014|Humpenöder et al. 2014]] ). However, these CDR methods may also compete for resources ( [[#Frank--2017|Frank et al. 2017]] ). Land-based mitigation approaches currently propose the use of forests (i) as a source of woody biomass for bioenergy and various biomaterials and (ii) for carbon sequestration in vegetation, soils, and forest products. Forests are therefore required to provide both provisioning (biomass feedstock) and regulating (carbon sequestration) ecosystem services. This multifaceted strategy has the potential to result in trade-offs ( [[#Makkonen--2015|Makkonen et al. 2015]] ). Some land-based mitigation options could conflict with biodiversity goals, e.g., A/R using monoculture plantations can reduce species richness when introduced into (semi-)natural grasslands ( [[#Smith--2019a|Smith et al. 2019a]] ; [[#Dooley--2021|Dooley et al. 2021]] ). When trade-offs exist between biodiversity protection and mitigation objectives, biodiversity is typically given a lower priority, especially if the mitigation option is considered risk-free and economically feasible ( [[#Pörtner--2021|Pörtner et al. 2021]] ). Approaches that promote synergies, such as sustainable forest management, reducing deforestation rates, cultivation of perennial crops for bioenergy in sustainable farming practices, and mixed-species forests in A/R, can mitigate biodiversity impacts and even improve ecosystem capacity to support biodiversity while mitigating climate change ( [[#Pörtner--2021|Pörtner et al. 2021]] ) ( [[#12.5|Section 12.5]] ). Systematic land-use planning could help to deliver land-based mitigation options that also limit trade-offs with biodiversity ( [[#Longva--2017|Longva et al. 2017]] ) (Cross-Working Group Box 3: Mitigation and Adaptation via the Bioeconomy, in this chapter).

Status, costs, potentials, risk and impacts, co-benefits, trade-offs and spillover effects and the role in mitigation pathways of A/R, biochar, SCS, peatland and coastal wetland restoration, agroforestry and forest management are summarised in Table 12.6. See also [[#12.5|Section 12.5]] .

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=== 12.3.3 CDR Governance and Policies ===

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As shown in Cross-Chapter Box 8 in this chapter, CDR fulfils multiple functions in different phases of ambitious mitigation: (i) further reducing net CO 2 or GHG emission levels in the near term; (ii) counterbalancing residual emissions (from hard-to-transition sectors like transport, industry, or agriculture) to help reach net zero CO 2 or GHG emissions in the mid term; (iii) achieving and sustaining net-negative CO 2 or GHG emissions in the long term. While inclusion of emissions and removals on managed land (LULUCF) is mandatory for developed countries under UNFCCC inventory rules ( [[#Grassi--2021|Grassi et al. 2021]] ), not all Annex I countries have included land-based biological removals when setting domestic mitigation targets in the past, but updated NDCs for 2030 indicate a shift, most notably in the European Union ( [[#Gheuens--2021|Gheuens and Oberthür 2021]] ; [[#Schenuit--2021|Schenuit et al. 2021]] ). The early literature on CDR governance and policy has been mainly conceptual rather than empirical, focusing on high-level principles (see the concerns listed in the introduction to [[#12.3|Section 12.3]] ) and the representation of CDR in global mitigation scenarios ( [[IPCC:Wg3:Chapter:Chapter-3#3.2.2|Section 3.2.2]] ). However, with the widespread adoption of net zero targets and the recognition that CDR is a necessary element of mitigation portfolios to achieve net zero CO 2 or GHG emissions, countries with national net-zero emissions targets have begun to integrate CDR into modelled national mitigation pathways, increase research, development and demonstration (RD&amp;D) efforts on CDR methods, and consider CDR-specific incentives and policies ( [[#Honegger--2021b|Honegger et al. 2021b]] ; [[#Schenuit--2021|Schenuit et al. 2021]] ) (Box 12.1). Nevertheless, this increasing consideration of CDR has not yet extended to net-negative targets and policies to achieve these. While the use of CDR at levels that would lead to net negative CO 2 or GHG emissions in the long term has been assumed in most global mitigation scenarios that limit warming to 1.5°C, net-negative emissions trajectories and BECCS as the main CDR method modelled to achieve these have not been mirrored by corresponding UNFCCC decisions so far ( [[#Fridahl--2017|Fridahl 2017]] ; [[#Mohan--2021|Mohan et al. 2021]] ). Likewise, only a few national long-term mitigation plans or legal acts entail a vision for net-negative GHG emissions ( [[#Buylova--2021|Buylova et al. 2021]] ), for example Finland, Sweden, Germany and Fiji.

For countries with emissions targets aiming for net zero or lower, the core governance question is not whether CDR should be mobilised or not, but which CDR methods governments want to see deployed by whom, by when, at which volumes and in which ways ( [[#Minx--2018|Minx et al. 2018]] ; [[#Bellamy--2019|Bellamy and Geden 2019]] ). The choice of CDR methods and the scale and timing of their deployment will depend on the respective ambitions for gross emissions reductions, how sustainability and feasibility constraints are managed, and how political preferences and social acceptability evolve ( [[#Bellamy--2018|Bellamy 2018]] ; [[#Forster--2020|Forster et al. 2020]] ; [[#Fuss--2020|Fuss et al. 2020]] ; [[#Waller--2020|Waller et al. 2020]] ; [[#Clery--2021|Clery et al. 2021]] ; [[#Iyer--2021|Iyer et al. 2021]] ; [[#Rogelj--2021|Rogelj et al. 2021]] ). As examples of emerging CDR policymaking at (sub-)national levels show, policymakers are beginning to incorporate CDR methods beyond those currently dominating global mitigation scenarios, that is, BECCS and afforestation/reforestation ( [[#Bellamy--2019|Bellamy and Geden 2019]] ; [[#Buylova--2021|Buylova et al. 2021]] ; [[#Schenuit--2021|Schenuit et al. 2021]] ; [[#Uden--2021|Uden et al. 2021]] ) (Box 12.1). CDR policymaking is faced with the need to consider method-specific timescales of CO 2 storage, as well as challenges in MRV and accounting, potential co-benefits, adverse side effects, interactions with adaptation and trade-offs with SDGs ( [[#Dooley--2018|Dooley and Kartha 2018]] ; [[#McLaren--2019|McLaren et al. 2019]] ; [[#Buck--2020|Buck et al. 2020]] ; [[#Honegger--2020|Honegger et al. 2020]] ; [[#Brander--2021|Brander et al. 2021]] ; [[#Dooley--2021|Dooley et al. 2021]] ; [[#Mace--2021|Mace et al. 2021]] ) (Table 12.6). Therefore, CDR governance and policymaking are expected to focus on responsibly incentivising RD&amp;D and targeted deployment, building on both technical and governance experience with already widely practised CDR methods like afforestation/reforestation ( [[#Lomax--2015|Lomax et al. 2015]] ; [[#Field--2017|Field and Mach 2017]] ; [[#Bellamy--2018|Bellamy 2018]] ; [[#Carton--2020|Carton et al. 2020]] ; [[#VonHedemann--2020|VonHedemann et al. 2020]] ), as well as learning from two decades of slow-moving CCS deployment ( [[#Buck--2021|Buck 2021]] ; [[#Martin-Roberts--2021|Martin-Roberts et al. 2021]] ; [[#Wang--2021|Wang et al. 2021]] ). For some less well-understood methods and implementation options, such as ocean alkalinisation or enhanced weathering, investment in RD&amp;D can help in understanding the risks, rewards, and uncertainties of deployment ( [[#Nemet--2018|Nemet et al. 2018]] ; [[#Fajardy--2019|Fajardy et al. 2019]] ; [[#Burns--2020|Burns and Corbett 2020]] ; [[#Goll--2021|Goll et al. 2021]] ).

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<span id="box-12.1-case-study-emerging-cdr-policy-research-and-development-in-the-united-kingdom"></span>
=== Box 12.1 | Case Study: Emerging CDR Policy, Research and Development in the United Kingdom ===

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Climate change mitigation policies in the UK have been motivated since 2008 by a domestic, legally-binding framework. This framework includes a 2050 target for net zero greenhouse gas emissions, interim targets and an independent advisory body called the Climate Change Committee ( [[#Muinzer--2019|Muinzer 2019]] ). It has led successive UK governments to publish mitigation plans to 2050, causing policy to be more forward looking ( [[#Averchenkova--2021|Averchenkova et al. 2021]] ).

The UK’s targets include emissions and removals from LULUCF. In 2008 the target for 2050 was an economy-wide net emissions reduction of at least 80% below 1990 levels. Even the first government plans to achieve this target proposed deployment of removal methods, specifically afforestation and wood in construction, increased soil carbon and BECCS ( [[#HM%20Government--2011|HM Government 2011]] ).

Box 12.1

Adoption of the Paris Agreement in 2015 caused the government to change the legislated 2050 target to a reduction of at least 100% (i.e., net zero). Since then, removal of CO 2 and other greenhouse gases has received greater prominence as a distinct topic. The most recent national plan (published October 2021) proposes deployment not only of the methods mentioned above, but also DACCS, biochar and enhanced weathering. The government has committed to amend accounting of UK targets to include a wider range of removal methods beyond LULUCF, and set a target of 5 MtCO 2 yr –1 from methods such as BECCS, DACCS and enhanced weathering by 2030. It is consulting on markets and incentives for deployment, and exploring new requirements for MRV ( [[#HM%20Government--2021|HM Government 2021]] ).

In parallel to these policy developments, the UK funds research into technical, environmental and social aspects of removal ( [[#Lezaun--2021|Lezaun et al. 2021]] ). Research on some elements (e.g., forestry, CCS, soils, bioenergy) have been funded for well over a decade, but the first programme dedicated to greenhouse gas removal ran during 2017–2021. This has been followed by two new programmes with greater focus on demonstration, totalling GBP100 million over four years ( [[#HM%20Government--2021|HM Government 2021]] ). A wide variety of methods is supported in these programmes, covering approaches such as CO 2 capture from seawater and capture of methane from cattle, in addition to those included already in national mitigation scenarios.

Deployment of removal methods has lagged behind expectations, as national targets for tree planting are not being met and infrastructure for CO 2 transport and storage is not yet in place ( [[#Climate%20Change%20Committee--2021|Climate Change Committee 2021]] ). While public awareness around carbon removal is low, studies indicate support in general, provided it is perceived as enhancing rather than impeding action to reduce emissions ( [[#Cox--2020a|Cox et al. 2020a]] ).

Since the enhancement of carbon sinks is a form of climate change mitigation ( [[#Honegger--2021a|Honegger et al. 2021a]] ), CDR governance challenges will in many respects be similar to those around emissions reduction measures, as will policy instruments like RD&amp;D funding, carbon pricing, tax or investment credits, certification schemes, and public procurement (Sections 13.4, 13.6, 14.4 and 14.5). Effectively integrating CDR into mitigation portfolios can build on already existing rules, procedures and instruments for emissions abatement ( [[#Torvanger--2019|Torvanger 2019]] ; [[#Fridahl--2020|Fridahl et al. 2020]] ; [[#Zakkour--2020|Zakkour et al. 2020]] ; [[#Honegger--2021b|Honegger et al. 2021b]] ; [[#Mace--2021|Mace et al. 2021]] ; [[#Rickels--2021|Rickels et al. 2021]] ). Additionally, to accelerate RD&amp;D and to incentivise CDR deployment, a political commitment to formal integration into existing climate policy frameworks is required ( ''robust evidence'' , ''high agreement'' ) ( [[#Lomax--2015|Lomax et al. 2015]] ; [[#Geden--2018|Geden et al. 2018]] ; [[#Honegger--2018|Honegger and Reiner 2018]] ; [[#VonHedemann--2020|VonHedemann et al. 2020]] ; [[#Schenuit--2021|Schenuit et al. 2021]] ). To avoid CDR being misperceived as a substitute for deep emissions reductions, the prioritisation of emissions cuts can be signalled and achieved with differentiated target setting for reductions and removals ( [[#Geden--2019|Geden et al. 2019]] ; [[#McLaren--2019|McLaren et al. 2019]] ). Similarly, sub-targets are conceivable for different types of CDR, to prioritise preferred methods according to characteristics such as removal processes or timescales of storage ( [[#Smith--2021|Smith 2021]] ).

IPCC guidance on quantifying removals is available for land-based biological CDR methods ( [[#IPCC--2006|IPCC 2006]] , 2019), but has yet to be developed for other CDR methods (Royal Society and Royal Academy of Engineering 2018). Challenges with development of estimation algorithms, data collection, and attribution between sectors and countries will need to be overcome ( [[#Luisetti--2020|Luisetti et al. 2020]] ; [[#Wedding--2021|Wedding et al. 2021]] ). Trusted methodologies for MRV, required to enable private sector participation. will need to address the permanence, leakage, and saturation challenges with land- and ocean-based biological methods ( [[#Mace--2021|Mace et al. 2021]] ). Protocols that also capture social and ecological co-benefits could encourage the adoption of biological CDR methods such as SCS, biochar, A/R and blue carbon management ( ''robust evidence'' , ''high agreement'' ) ( [[#VonHedemann--2020|VonHedemann et al. 2020]] ; [[#Macreadie--2021|Macreadie et al. 2021]] ).

Private capital and companies, impact investors, and philanthropy will play a role in technical demonstrations and bringing down costs, as well as creating demand for carbon removal products on voluntary markets, which companies may purchase to fulfil corporate social responsibility-driven targets ( [[#Friedmann--2019|Friedmann 2019]] ; [[#Fuss--2020|Fuss et al. 2020]] ; [[#Joppa--2021|Joppa et al. 2021]] ). Niche markets can provide entry points for limited deployment of novel CDR methods ( [[#Cox--2019|Cox and Edwards 2019]] ), but targeting currently existing revenue streams by using CO 2 captured from the atmosphere in Enhanced Oil Recovery and other utilisation routes ( [[#Mackler--2021|Mackler et al. 2021]] ; [[#Meckling--2021|Meckling and Biber 2021]] ) is contested, and highlights the importance of choosing appropriate system boundaries when assessing supply chains ( [[#Tanzer--2019|Tanzer and Ramírez 2019]] ; [[#Brander--2021|Brander et al. 2021]] ). While the private sector will play a distinct role in scaling CDR, governments will need to commit to developing infrastructure for the transport and storage of CO 2 , including financing, permitting, and regulating liabilities ( [[#Sanchez--2018|Sanchez et al. 2018]] ; [[#Mace--2021|Mace et al. 2021]] ; [[#Mackler--2021|Mackler et al. 2021]] ).

International governance considerations include global technology transfer around CDR implementation options ( [[#Batres--2021|Batres et al. 2021]] ); land use change that could affect food production and land condition and cause conflict around land tenure and access ( [[#Dooley--2018|Dooley and Kartha 2018]] ; [[#Hurlbert--2019|Hurlbert et al. 2019]] ; [[#Milne--2019|Milne et al. 2019]] ); and efforts to create sustainable and just supply chains for CDR ( [[#Fajardy--2020|Fajardy and Mac Dowell 2020]] ; [[#Tan--2021|Tan et al. 2021]] ), such as resources used for BECCS, enhanced weathering, or ocean alkalinisation. International governance would be particularly important for methods posing transboundary risks, especially for ocean-based methods. Specific regulations have so far only been developed in the context of the London Protocol, an international treaty that explicitly regulates ocean fertilisation and allows parties to govern other marine CDR methods like ocean alkalinity enhancement ( [[#GESAMP--2019|GESAMP 2019]] ; [[#Burns--2020|Burns and Corbett 2020]] ; [[#Boettcher--2021|Boettcher et al. 2021]] ) ( [[IPCC:Wg3:Chapter:Chapter-14#14.4.5|Section 14.4.5]] ).

Engagement of civil society organisations and publics will be important for shaping CDR policy and deployment ( ''medium evidence'' , ''high agreement'' ). Public awareness of CDR and its role in national net zero emissions strategies is generally very low ( [[#Cox--2020a|Cox et al. 2020a]] ), and perceptions differ across countries and between methods ( [[#Bertram--2020|Bertram and Merk 2020]] ; [[#Spence--2021|Spence et al. 2021]] ; [[#Sweet--2021|Sweet et al. 2021]] ; [[#Wenger--2021|Wenger et al. 2021]] ). When awareness increases, social processes will shape political attitudes on CDR ( [[#Shrum--2020|Shrum et al. 2020]] ), as will efforts to frame particular CDR methods as ‘natural’ or ‘technological’ ( [[#Osaka--2021|Osaka et al. 2021]] ), and the policy instruments chosen to support CDR ( [[#Bellamy--2019|Bellamy et al. 2019]] ). Lack of confidence in CDR implementation options from both publics and investors, and lack of trust in project developers ( [[#Cox--2020b|Cox et al. 2020b]] ) have hampered support for CCS ( [[#Thomas--2018|Thomas et al. 2018]] ) and are expected to affect deployment of CDR methods with geological storage ( [[#Gough--2019|Gough and Mander 2019]] ). On local and regional scales, CDR projects will need to consider air and water quality, impacts to human health, energy needs, land use and ecological integrity, and local community engagement and procedural justice. Bottom-up and community-driven strategies are important for deploying equitable carbon removal projects ( [[#Batres--2021|Batres et al. 2021]] ; [[#Hansson--2021|Hansson et al. 2021]] ).

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