Sustainable scale-up of negative emissions technologies and practices: where to focus

The large-scale implementation of afforestation and reforestation practices—currently spurred by global initiatives like the trillion tree campaign (Trillion tree campaign 2021)—could reverse the global trend of forest loss, with average deforestation rates of 10 Mha a⁻¹ between 2015 and 2020 (FAO and UNEP 2020). Afforestation is the conversion of land that was previously not covered by forest to forestland, whereas reforestation takes place on land that has been recently deforested; i.e. afforestation implies a LUC, whereas reforestation occurs in an area classified as forest (FAO 2018).

Forests can maintain CDR for decades before it declines as the trees mature (Houghton et al 2015). Recent analyses cap the global CDR potential of forestation at 2.6 Gt a⁻¹ over 2025–2055 (Austin et al 2020) and 5.6 Gt a⁻¹ (average over 21st century, Favero et al 2020), with the largest potential being realised in the tropics (Doelman et al 2020). On the other hand, improved management practices in existing forests (e.g. reduced logging or extended rotations) could lead to the additional removal of up to 2 Gt a⁻¹ CO₂-eq (Smith et al 2020).

The advantage of forestation over other NETPs is that it is easy to deploy and cost-competitive (The Royal society 2009), with costs ranging between 5 and 53 US$₂₀₂₀ t⁻¹ CO₂ removed (Fuss et al 2018). Nevertheless, the carbon sequestered in forests could return to the atmosphere due to human-induced LUC or unexpected perturbations such as pests, droughts or fires (Fuss et al 2018).

Agricultural residues and natural fibres like cork or hemp can be used as insulation materials, thereby sequestering the carbon contained in them (Kriegh et al 2021, Shen et al 2022a). Furthermore, the mechanical properties of certain engineered wood products—such as glued laminated timber (glulam) and cross-laminated timber—allow them to replace steel and concrete as structural materials (D'Amico et al 2021). Bamboo—a fast growing grass with properties comparable to those of steel and concrete—is also an appealing material because of its ability to sequester carbon rapidly and thrive in degraded lands (Project Drawdown 2021a). The main climate benefits of these products stem from the substitution of construction materials (National Academies of Sciences Engineering and Medicine 2019, Shen et al 2022a); the GHGs emitted throughout the life cycle of wood and bamboo products are lower than those associated with the functionally equivalent amounts of steel and concrete (Upton et al 2008, Sathre and Gustavsson 2009, Sathre and O'Connor 2010, Zea Escamilla et al 2016). Nevertheless, the duration of the carbon sequestration in these materials is usually temporal, dependent on their lifetime and end-of-life treatment.

It has been estimated that deploying timber as a construction material could sequester up to 0.04–2.5 Gt a⁻¹ CO₂ by 2050 (Churkina et al 2020) at a low cost (McLaren 2012). Updating building codes to guarantee higher energy efficiencies in buildings—which would favour wood over other materials with worse insulating properties—could help to reach the full potential of wood as a construction material (Wimmers 2017).

Wood can be harvested and buried under anaerobic conditions or stored above-ground (Zeng 2008, Zeng et al 2013). It has been estimated that this CDR strategy could sequester 3.7–11 Gt a⁻¹ CO₂ (Zeng et al 2013). The results of the field tests conducted to date (Carbon lockdown 2022) show that this NETP can significantly slow biomass decomposition, although it does not completely halt it (Adair et al 2010). Despite the low cost of this NETP (9–33 $ t⁻¹ CO₂) and its easy implementation (Zeng 2008), some of its potential adverse side-effects—including nutrient lockup in the stored biomass, and disturbance of the local biodiversity—could hinder its large-scale deployment (Zeng 2008).

Roots are usually the main contributor to plant SCS; they are five times more likely than above-ground litter to be stabilised as soil organic matter (Jackson et al 2017). Plants with deeper and larger roots—e.g. perennial vegetation such as trees and grasses—could remove on the order of 1 Gt a⁻¹ CO₂-eq (Paustian et al 2016). However, these crops should be planted on low-carbon soils, otherwise the carbon loss due to the cultivation and land conversion processes could offset the carbon gains associated with the root system and the above-ground plant residues (Whitaker et al 2018).

While the roots and above-ground plant residues are left in the field to increase the SOC pool, biomass can be harvested and used for bioenergy generation or as a construction material (Shen et al 2022b), which provides the cultivation of these crops with an economic edge over other NETPs. Another advantage of extensive root systems is that they improve the plants' ability to retain nutrients and water, providing resistance to droughts and fertiliser runoff (Kell 2011).

Current research efforts focus on developing crops with larger and deeper roots through selective breeding or genetic engineering techniques (Kell 2011, Paustian et al 2019). Most notably, the Land Institute is working on the perennialisation of cereal grains and other annual crops, some of which are already commercialised (The Land Institute 2021).

The carbon present in soil amendments (such as compost produced from municipal organic waste) contributes to increasing the SOC levels. Furthermore, organic amendments can improve the soil properties and nutrient availability, which stimulates plant productivity and additional carbon uptake (Paustian et al 2019). The CDR capacity of this NETP depends on the original fate of the materials—e.g. manure application does not sequester additional carbon compared to a baseline scenario where it is left in the field (Powlson et al 2011, Leifeld et al 2013)—, and on the decomposition rate of the added carbon. The latter is typically slower than that of fresh plant residues and dependent on the soil properties and amendment characteristics (Diacono and Montemurro 2011, Paustian et al 2016).

Spreading light-coloured residues such as cereal straw on the soil could provide cooling benefits by increasing the surface albedo (Smith 2016). By contrast, dark soil amendments like compost could have the opposite effect (Meyer et al 2012). Furthermore, the application of organic matter to the soil may increase nitrous oxide emissions (Smith et al 2001), which could offset the climate benefits of carbon sequestration. Moreover, the leaching of the nutrients contained in the organic amendments could pollute water bodies, whereas the heavy metals and organic pollutants in the composted organic wastes could pose a risk for human health (Cobo et al 2018).

The practices under this category seek to maximise the SOC inputs from plant roots and residues in grazing lands by maximising forage production or controlling grazing intensity (Paustian et al 2019).

Introducing more productive species with deeper roots, such as legumes, and adjusting the water and fertiliser inputs to the plants' demand can increase forage production (Smith et al 2008). On the other hand, techniques to manipulate the grazing intensity include decreasing the number of animals in a given area (to prevent overgrazing), and implementing rotational or multi-paddock grazing systems, which allow the land that has already been grazed to recover (Project Drawdown 2021b).

The main co-benefits of these practices are the improvements in biological diversity and soil properties (Bossio et al 2020). However, the carbon sequestration rates are highly location-dependent, i.e. subject to the climate, soil and vegetation characteristics (Conant et al 2017). It has been estimated that the global sequestration potential of grazing lands ranges between 0.3 and 1.4 Gt a⁻¹ CO₂-eq (Henderson et al 2015).

Maximising the time during which soil is covered by vegetation can increase the SOC stocks and simultaneously enhance soil quality and fertility. Fallow frequency can be reduced by planting seasonal cover crops—whose global CDR potential is estimated to be 0.4 Gt a⁻¹ CO₂-eq (Bossio et al 2020)—and diversifying crop rotations, specifically by introducing perennials, legumes and species that produce large amounts of residues (National Academies of Sciences Engineering and Medicine 2019, Paustian et al 2019). The costs of these practices vary greatly across locations and methods; the available estimates in the literature range from 23 to 147 $ t⁻¹ CO₂-eq (Tang et al 2016).

Zhou and Flynn (2005) assessed the implications of enhancing the solubility pump by cooling the ocean surface water, which would increase downwelling currents. They concluded that it is unlikely that this technology will ever become a competitive carbon sequestration method; cooling 1 Mm³ s⁻¹ of seawater from 6 °C to 0 °C would require a heat flux of 25 TW and capture only 35 Mt a⁻¹ CO₂, while the estimated costs could range between 258 and 5826 $ t⁻¹ CO₂. Moreover, they emphasised the need for modelling research, suggesting that upwelling currents releasing more CO₂ than the amount captured could offset downwelling currents. To the best of the authors' knowledge, experimental downwelling has never been conducted before.

Lovelock and Rapley (2007) proposed to use vertical pipes to pump up nutrient-rich deep waters to fertilise algae in the ocean surface and increase the transfer rate of organic carbon to the deep ocean via the biological pump. A 300 m long pipe was deployed by Maruyama et al (2011) to investigate the effects of artificial upwelling driven by the difference in salinity and temperature at both ends of the pipe. They found that the chlorophyll concentration at the pipe outlet was much higher than in the surrounding surface water, which suggests an increase in the CO₂ absorption rate. Nonetheless, a recent analysis shows that approximately 70% of the carbon exported to the deep ocean after increasing ecosystem productivity is transported back to the surface ocean within 50 years (Siegel et al 2021).

Earth system simulations reveal that this NETP could contribute to cooling the Earth's surface—due to the lower temperature of the ocean's bottom waters—and enhancing terrestrial carbon storage. However, discontinuing the upwelling could lead to rapid warming (Oschlies et al 2010, Keller et al 2014). Furthermore, the decrease in evaporation and evapotranspiration rates would lead to a decline in precipitations. Other side-effects include increased acidification, and a reduction in sea-ice loss (Keller et al 2014).

The global CDR potential of artificial upwelling is likely limited to 50–100 Mt a⁻¹ CO₂ (Koweek 2022); it has been estimated that upwelling 1 Mm³ s⁻¹ of seawater could only sequester 59 Mt a⁻¹ CO₂ (Lenton and Vaughan 2009). However, the results of the model developed by Yool et al (2009) showcased that pumping up water in some regions could lead to net CO₂ emissions to the atmosphere.

This marine NETP relies on the reaction of alkaline substances with the CO₂ dissolved in the seawater. The enhanced weathering (EW) reactions that occur as a result of adding synthetic chemicals or minerals to the ocean are shown below (Harvey 2008, Renforth 2019). Here, Me represents a divalent cation, typically calcium or magnesium.

$$\begin{align}\text{Me}{\left( \text{OH} \right)_2} + 2\text{C}{\text{O}_2} \to \text{M}{\text{e}^{2 + }} + 2\text{HCO}_3^ - \end{align} \tag{ R1 }$$

$$\begin{align}\text{MeSi}{\text{O}_3} + 2\text{C}{\text{O}_2} + {\text{H}_2}\text{O} \to \text{M}{\text{e}^{2 + }} + 2\text{HCO}_3^ - + \text{Si}{\text{O}_2}\end{align} \tag{ R2 }$$

$$\begin{align} & \text{M}{\text{e}_2}\text{Si}{\text{O}_4} + 4\text{C}{\text{O}_2} + 4{\text{H}_2}\text{O} \to 2\text{M}{\text{e}^{2 + }} + 4\text{HCO}_3^ - \nonumber\\ & \quad + {\text{H}_4}\text{Si}{\text{O}_4}\end{align} \tag{ R3 }$$

$$\begin{align}\text{MeC}{\text{O}_3} + \text{C}{\text{O}_2} + {\text{H}_2}\text{O} \to \text{M}{\text{e}^{2 + }} + 2\text{HCO}_3^ - \end{align} \tag{ R4 }$$

While the alkaline materials react with the dissolved CO₂ quite rapidly, the subsequent transfer of atmospheric CO₂ to the surface ocean is slower (National Academies of Sciences Engineering and Medicine 2021). Moreover, although the residence time of bicarbonate ions in the ocean is virtually permanent, an increase in the alkalinity levels would lead to mineral carbonation reactions that produce solid carbonate minerals and release part of the previously captured CO₂ (Renforth 2019):

$$\begin{align}\text{Me}^{2 + } + 2\text{HCO}_3^ - \to \text{MeC}{\text{O}_3} + \text{C}{\text{O}_2} + {\text{H}_2}\text{O}\end{align} \tag{ R5 }$$

The reaction of acidic species (other than carbonic acid) with the produced bicarbonate ions could also lead to the emission of the captured CO₂, in accordance with reaction (R6) (Zhang et al 2022):

$$\begin{align}{\text{H}^ + } + \text{HCO}_3^ - \to \text{C}{\text{O}_2} + {\text{H}_2}\text{O}\end{align} \tag{ R6 }$$

On the other hand, the nutrients released during the dissolution of the minerals could stimulate biological productivity, leading to additional carbon sequestration (Hartmann et al 2013, Hauck et al 2016). Nevertheless, the side-effects related to the dissolution of the trace metals present in the minerals on the marine biota are still largely unknown (Meysman and Montserrat 2017, Bach et al 2019, Gore et al 2019).

3.2.2.1.1. Direct alkalinisation

3.2.2.1.2. Coastal enhanced weathering

3.2.2.1.3. Electrochemical enhanced weathering

Eisaman et al (2012) described an experimental set up to extract up to 60% of the carbon dissolved in seawater with bipolar electrodialysis membranes. This process is based on the premise that returning the CO₂-depleted water to the ocean would draw the absorption of more atmospheric CO₂ into the seawater. The dilute acid and base products derived from the electrodialysis process are used to alter the pH equilibrium in the processed seawater and extract the dissolved carbon. Two process configurations were proposed: in the acid process, the produced acid is used to lower the pH of the treated water and convert the dissolved inorganic carbon to CO₂ gas, which should be sequestered or mineralised to attain carbon negative emissions. In the base configuration, the pH of the processed water is increased with the produced base, which causes the precipitation of the dissolved inorganic carbon as calcium carbonate. Before returning the water to the ocean, the pH is restored with the produced chemicals (Eisaman et al 2018). The associated energy consumption of the acid and base configurations is 11.3 and 15.8 GJ t⁻¹ CO₂ extracted, respectively (Eisaman et al 2018). Capture costs between 393 and 637 $ t⁻¹ of extracted CO₂ were estimated (Eisaman et al 2018).

This is an incipient research area with potential for novel technological developments; an experimental prototype of an electrolytic cation exchange process capable of extracting 92% of the CO₂ present in the seawater and simultaneously produce hydrogen has also been presented (Willauer et al 2014).

Iron availability limits the photosynthesis rate in around a third of the open ocean (Emerson 2019), whereas nitrogen and phosphorus are the limiting nutrients in the rest of the ocean (Williamson et al 2012). Hence, fertilising the ocean with site-specific deficient nutrients could lift the constraint on the biological pump.

Several small-scale iron fertilisation experiments have been carried out, confirming that iron promotes biomass productivity and CO₂ drawdown from the atmosphere in many regions (Williamson et al 2012, Emerson 2019). Nonetheless, field experiments on phosphorus fertilisation registered a slight decrease in phytoplankton and biomass chlorophyll, suggesting additional limitations to nutrient availability (Williamson et al 2012).

The costs of ocean fertilisation are dependent on the regional conditions; it has been estimated that the CDR costs via macronutrient (nitrogen and phosphorus) and iron fertilisation are 24 and 519 $ t⁻¹ CO₂, respectively (Harrison 2013, Jones 2014). The cost difference between the two strategies can be mainly attributed to the low sequestration efficiency of the latter.

Zahariev et al (2008) calculated that the maximal CDR rate that could be achieved by means of iron fertilisation would be below 3.6 Gt a⁻¹. In contrast, the maximum CO₂ sequestration potential of combined nitrogen and phosphorus fertilisation has been estimated to be 5.5 Gt a⁻¹ (Harrison 2017). However, the permanence of the carbon sequestered in the deep ocean by enhancing the upper ocean biological productivity is very low: on average, only 32% of the carbon would remain sequestered after 50 years, and it would drop to 25% after 100 years (Siegel et al 2021).

According to Williamson et al (2012), the potential unintended impacts of large-scale ocean fertilisation may include:

• The production of gases that may affect the climate, such as nitrous oxide, methane and dimethyl sulphide (DMS). The latter can increase cloud condensation and reflectivity (albedo), leading to a cooling effect (Wingenter et al 2007).
• Far-field effects on primary productivity due to the depletion of non-limiting nutrients.
• Decrease of oxygen levels.
• Biodiversity impacts. Large-scale ocean fertilisation will likely change the relative abundance of different species.
• Increased acidification of deep waters.

Blue carbon refers to the carbon sequestered in coastal and marine ecosystems.

3.2.3.2.1. Coastal blue carbon

3.2.3.2.2. Macroalgae cultivation

Metzger and Benford (2001) proposed collecting crop residues and depositing them (ballasted with stone) on the ocean floor as a carbon sequestration method. In a later work, the carbon sequestration efficiency of this practice was estimated to be 92.5% (Strand and Benford 2009). Thus, assuming a global generation of crop residues of 4.98 Gt a⁻¹ with a 40% carbon content, 6.75 Gt CO₂ could be sequestered annually, at the cost of 117 $ t⁻¹ CO₂ (Strand and Benford 2009). The authors suggest that lignocellulosic materials are highly stable in the marine environment, yet it is hard to foresee the associated environmental impacts due to the lack of experimental studies. Biochar may be more suitable for this application than raw biomass, given that it is more resistant to microbial degradation and it would not require ballasting because of its higher density (Miller and Orton 2021).

Fermentation is the process whereby microorganisms transform sugars into ethanol and CO₂. As reaction (R7) shows, one third of the carbon in the glucose molecules subjected to fermentation is transformed into CO₂ that can be subsequently sequestered; the remaining carbon composes the ethanol and will be released as CO₂ when the ethanol is used as fuel. Certain substances with a high sugar content, such as the juice extracted from raw sugarcane or sugar beets, can be directly fermented (Laude et al 2011, Milão et al 2019). By contrast, lignocellulosic materials must previously undergo an enzymatic hydrolysis process, where polysaccharides are broken down to glucose and other fermentable sugars (Xiros et al 2013).

$$\begin{align}{\text{C}_6}{\text{H}_{12}}{\text{O}_6} \to 2{\text{C}_2}{\text{H}_5}\text{OH} + 2\text{C}{\text{O}_2}\end{align} \tag{ R7 }$$

A purification unit is required to meet the commercial specifications of ethanol. Although several technologies such as membranes or molecular sieves could help reduce the energy consumption of the ethanol separation process, this step is typically carried out in a distillation unit, which accounts for 50–80% of the total energy input (Logsdon 2004). Therefore, the main CO₂ source in a conventional ethanol plant is the cogeneration unit, designed to supply heat and power to the system (Carminati et al 2019).

The CO₂ produced in the fermentation unit accounts for 11%–13% of the carbon emissions of an ethanol plant (Koornneef et al 2012); hence, sequestering this CO₂ stream alone does not make the overall process carbon negative (Laude et al 2011). However, most fermentation-BECCS plants—including the largest one to date, with a capacity of 1 Mt a⁻¹ CO₂—only capture the CO₂ released in the fermentation process (Global CCS Institute 2020). This CO₂ stream just needs to be dehydrated and compressed (Moreira et al 2016), whereas capturing the CO₂ produced in the cogeneration unit—which is diluted in the flue gas stream—is more energy-intensive and costly (Laude et al 2011, Bello et al 2020). The overall costs are likely to range between 21 and 185 $ t⁻¹ CO₂ captured, with the upper bound corresponding to the capture of CO₂ from both the fermentation and cogeneration units (Fuss et al 2018). The CDR potential of fermentation-BECCS could reach 1 Gt a⁻¹ by 2050 (Koornneef et al 2012).

Under anaerobic conditions, bacteria can degrade certain types of biomass to biogas, which is mainly composed of methane (50%–70%) and CO₂ (Thrän et al 2014). The latter must be captured—via CO₂ absorption in a liquid phase, pressure swing adsorption or membranes—prior to injecting the methane into the natural gas grid or using it as fuel in vehicles (Thrän et al 2014). Anaerobic digestion and biogas upgrading are well-established processes; in 2012 there were 277 biogas upgrading plants linked to anaerobic digesters, but none of them sequestered the separated CO₂ (Thrän et al 2014).

By 2050, 2.7 Gt a⁻¹ CO₂-eq could be removed from the atmosphere through the anaerobic digestion of energy crops, agricultural residues, municipal solid waste, sewage and manure (Koornneef et al 2013), although sequestering the CO₂ produced by burning the methane could provide additional negative emissions. A fraction of the biomass carbon content—contingent on the feedstock—is lost in the digestate, i.e. the remaining material that cannot be further degraded. This material is rich in nutrients and can be used for soil amendment, but it may require further treatment because of its high water content (WRAP 2012).

The CDR potential of this NETP is modest relative to other BECCS processes because feedstocks with high lignin contents cannot be subjected to anaerobic digestion; this technology is particularly well-suited for biogenic wastes such as the organic fraction of municipal solid waste, sewage sludge or manure (Koornneef et al 2013). Since these biomass resources are decentralised, the size of anaerobic digesters is typically small, which can increase the costs of connecting these facilities to the natural gas and CO₂ pipelines (Koornneef et al 2013). The total costs of anaerobic digestion BECCS, which are highly dependent on the biomass source, could range between 235 and 557 $ t⁻¹ CO₂ (IEA GHG 2013).

Theoretically, a global CDR potential of 10.4 Gt a⁻¹ could be sustainably achieved by 2050 with combustion-BECCS (Koornneef et al 2012). As biomass is burnt to generate heat and power, all the carbon in the feedstock is transformed into CO₂, whereas the non-combustible components of the biomass are transferred to the ashes. Several post-combustion capture processes (membrane separation, adsorption, etc) can separate the diluted CO₂ from the flue gas generated in the boiler. The chemical absorption of CO₂ into a monoethanolamine (MEA) solution followed by a desorption step is the most widespread of these processes (Bui et al 2018), and it can achieve CO₂ capture rates above 90% (Brandl et al 2021). Nevertheless, the associated energy penalty is substantial; e.g. the net energy efficiency of a 75 MW biomass power plant could drop from 36% to 23% when a CO₂ capture unit is added (IEA GHG 2009). The use of advanced commercial solvents based on amines blends could help reduce the energy penalty (Bui et al 2018).

Post-combustion capture BECCS is still in the pilot and demonstration phase, although several projects are expected to start sequestering on the order of Mt a⁻¹ CO₂ from biomass combustion plants within this decade (Global CCS Institute 2019, 2020). Capture costs between 224 and 266 $ t⁻¹ CO₂ have been estimated for BECCS based on post-combustion capture (IEA GHG 2009). However, retrofitting existing combustion and waste-to-energy plants with CCS could lower the CDR costs; sequestering the biogenic CO₂ generated in the energy recovery processes of kraft pulp mills could lead to 135 Mt a⁻¹ of negative CO₂ emissions (Kuparinen et al 2019) at a cost of 62–79 $ t⁻¹ (Onarheim et al 2017). Similarly, the incineration of the organic fraction of municipal solid waste produced worldwide coupled with CCS could produce negative emissions at the Gt scale (Pour et al 2018).

Oxy-combustion is an alternative technology to post-combustion capture where the feedstock comes into contact with pure oxygen instead of air, thereby releasing a stream of highly concentrated CO₂, which only has to be cooled in order to remove the water vapour prior to compression and storage (Cabral et al 2019). Although the electricity consumed in the air separation unit is considerable, this technology can improve the combustion efficiency (Cabral et al 2019), generating more energy per unit of biomass than BECCS based on post-combustion capture (Mac Dowell and Fajardy 2016) at a lower CDR cost (Bhave et al 2017), estimated to be 155 $ t⁻¹ CO₂ (Cabral et al 2019). Notwithstanding this, the feasibility of oxy-combustion BECCS has not been demonstrated at the pilot scale yet (Cabral et al 2019).

In BECCS processes based on chemical looping combustion, the biomass feedstock reacts with the oxygen in a metal oxide, generating a gas stream consisting mostly of CO₂ and water with a low energy penalty. The reduced metal oxide is then transported to another reactor where it is regenerated with air (Adánez et al 2018). This technology has been tested at semi-industrial scale, but the complete conversion of the fuel is hard to achieve due to the formation and loss of char (Adánez et al 2018). Furthermore, a third reactor is needed to fully oxidise the gas, increasing the CDR costs, which could range between 171 and 200 $ t⁻¹ CO₂ (Keller et al 2019). A novel variant of this technology, chemical looping with oxygen uncoupling—in which the fuel reacts with the oxygen released by the metal oxide—, could achieve carbon capture rates over 99%, but it is still in early development stages (Cormos 2017).

Biomass can be partially oxidised with a gasifying agent (air, oxygen or steam) to generate high yields of synthesis gas (syngas), mainly composed of carbon monoxide and hydrogen. Part of the biomass feedstock is transformed into biochar, which is usually burnt to provide the energy required to sustain the endothermic gasification reactions. The gasification process also produces a liquid fraction of heavy hydrocarbons (tars), which must be removed to prevent equipment fouling and plugging (McKendry 2002).

Steam is used to increase the CO₂ and hydrogen concentration in the gas phase through the water-gas-shift reaction (WGSR). The produced CO₂ is subsequently captured via pressure swing adsorption, absorption or membrane separation processes. Given the high concentration of CO₂ in the gas, this pre-combustion capture method consumes less energy than the alternative post-combustion capture (Shahbaz et al 2021).

After the CO₂ separation, hydrogen can be compressed and used as fuel off-site, or directly oxidised to generate heat and electricity. BECCS based on integrated gasification combined cycles (IGCCs)—which deploy both gas and steam turbines—can achieve higher energy efficiencies than combustion-BECCS at the expense of higher capture costs, between 175 and 364 $ t⁻¹ CO₂ (Bhave et al 2017). Approximately half of the biomass carbon content can be recovered from the syngas, (Rhodes and Keith 2005) but higher CDR efficiencies could be achieved by capturing the CO₂ generated in the biochar combustion and gas turbine. Hence, the technical CDR potential of IGCC-BECCS could amount to 10.4 Gt a⁻¹ by 2050 (Koornneef et al 2012). On the other hand, coupling biomass gasification with fuel cells could potentially double the power produced by IGCCs (Sadhukhan et al 2010), but this technology is still in its infancy (Bhave et al 2017).

A range of biofuels—including methane, methanol and dimethyl ether—can be synthesised after adjusting the carbon monoxide to hydrogen ratio in the syngas via the WGSR and removing the CO₂ (Gassner and Maréchal 2009, Sikarwar et al 2017). Notably, syngas can be converted into syncrude via the Fischer–Tropsch (FT) process, and further refined to produce synthetic gasoline, diesel or jet fuel, which would allow sequestering 57–74% of the biogenic carbon in the feedstock (Kreutz et al 2020), with costs ranging between 368 and 524 $ t⁻¹ CO₂ captured (Larson et al 2020). The CDR potential of BECCS based on gasification and subsequent methanation could amount to 3.5 Gt a⁻¹ CO₂-eq (Koornneef et al 2013), whereas the FT-BECCS pathway could sequester up to 5.8 Gt a⁻¹ CO₂-eq by 2050 (Koornneef et al 2012). However, these figures may substantially drop (below 0.1 Gt a⁻¹ CO₂-eq) if LUC emissions are factored in (Hanssen et al 2020).

Although the feasibility of biomass gasification has already been tested at large scale (e.g. GoBiGas biomass-to-methane demonstration plant, Larsson et al 2018) and even reached commercialisation (Enerkem waste-to-methanol plant, Enerkem 2022) these facilities do not include CCS.

Heating biomass in oxygen-free or oxygen-restricted environments triggers the pyrolytic reactions that yield bio-oil, biochar and pyrogas. Fast pyrolysis systems, which require high temperatures (around 500 °C) and short residence times (below 2 s), can maximise the production of bio-oil (Perkins et al 2018). Although bio-oil is composed of highly oxygenated hydrocarbons, it can be directly combusted to generate heat and power for stationary applications (Czernik and Bridgwater 2004). However, using it as a drop-in transportation fuel requires upgrading via hydrotreatment and hydrocracking processes, which can be partially carried out with the hydrogen derived from the pyrogas. These processes produce a mixture of lighter hydrocarbons with a low oxygen content (Peters et al 2015b).

Under fast pyrolysis conditions, approximately 48% of the biomass carbon content is distributed between the biochar and pyrogas (Schmidt et al 2019). These products can be burnt to provide the heat required by the system, releasing the carbon as CO₂ that can be subsequently captured and stored (Meerman and Larson 2017). Hence, we estimate the maximum technical CDR potential of fast pyrolysis as 48% of the CDR potential of combustion-BECCS, i.e. 5 Gt a⁻¹ (Koornneef et al 2012). It has been estimated that the costs of this BECCS configuration (including bio-oil upgrading) equipped with MEA post-combustion capture could amount to 404 $ t⁻¹ CO₂ (Meerman and Larson 2017).

Several companies (ENSYN, btg bioliquids) have successfully commercialised the fast pyrolysis technology (Perkins et al 2018), but we are not aware of any bio-oil projects integrating the combustion of biochar and pyrogas with CCS. The company Charm Industrial has developed an alternative CDR pathway whereby bio-oil is not used to deliver energy, but injected into geological reservoirs. They offer this CDR service at a price of 600 $ t⁻¹ CO₂-eq (Charm Industrial 2021).

Processing biomass in an aqueous medium at high temperatures and pressures (250 °C–374 °C and 40–220 bar) leads to the disintegration of the polymers in the biomass and the production of bio-crude (Elliott et al 2015). This bio-crude is more viscous, less dense and less oxygenated than pyrolysis bio-oil. It can directly substitute heavy fuel oil, but using the bio-crude as a transportation fuel requires a previous hydrotreatment process (Elliott et al 2015).

The hydrothermal liquefaction of biomass also generates biochar and gas—mainly composed of CO₂ and hydrogen (Lozano et al 2020). Part of the hydrogen required for the hydrotreatment can be sourced from this gas stream, whereas the remaining combustible gases can be burnt along with the biochar to provide heat to the system. Although the process is almost self-sufficient, an external supply of hydrogen and energy is needed (Lozano et al 2020).

Altogether, the gas and solid products contain 26–32% of the carbon in the biomass (Lozano et al 2020, SundarRajan et al 2020). Thus, capturing and sequestering this carbon could contribute to the removal of up to 3.3 Gt a⁻¹ CO₂-eq, estimated as a fraction of the CDR potential of combustion-BECCS in 2050 (Koornneef et al 2012). This is a conservative assumption because hydrothermal liquefaction does not require a drying pretreatment and therefore can process a wide range of biomass feedstocks with high moisture contents (Tekin et al 2014).

Currently, hydrothermal liquefaction is at pilot and demonstration scale, with companies such as Licella and Genifuel at the forefront (Lozano et al 2020). However, none of these facilities incorporate carbon capture units. The estimated cost of this technology ranges between 583 and 1365 $ t⁻¹ CO₂ captured (Lozano et al 2020).

The atmospheric concentration of methane is low compared to CO₂ (1.9 versus 410 ppm in 2019, IPCC 2021), but its global warming potential is 34 and 86 times that of CO₂ for 100- and 20 year time horizons, respectively (Myhre et al 2013). Oxidising 1 Gt methane would prevent an increase in the global surface temperature of 0.21 ± 0.04 °C (Abernethy et al 2021). Furthermore, the averted formation of tropospheric ozone would entail benefits for human health and ecosystems.

The minimum thermodynamic energy required to separate methane from ambient air is almost five times that of CO₂ per unit mass (Boucher and Folberth 2010). However, expressed per t CO₂ equivalent, the minimum energy needed to separate atmospheric methane is 7 and 19 times lower than that of CO₂, considering 100- and 20 year time horizons (Jackson et al 2021). Moreover, the combustion of the recovered methane could provide 28% of the energy required for the separation (Boucher and Folberth 2010). Some zeolites and porous polymeric networks have been preliminarily identified as promising materials for methane capture because of their selectivity and sorption capacities (Kim et al 2013, Jackson et al 2019). Nonetheless, the feasibility of deploying copper-doped zeolites to oxidise methane at atmospheric concentrations has recently been demonstrated (Brenneis et al 2021), which could render the need to develop an intermediate separation step unnecessary.

Bioreactors exploiting the natural capability of methanotrophs to metabolise methane at atmospheric concentrations could be an alternative pathway to oxidise atmospheric methane. However, this strategy has only been proven for methane concentrations of about 300 ppm in the animal husbandry sector (Stolaroff et al 2012); it has been estimated that the average atmospheric methane concentration is too low to enable the survival of the methanotrophs within the bioreactor (Yoon et al 2009).

The application of photocatalytic coats on the walls of buildings with enclosed animals has been demonstrated to be an effective measure to oxidise methane (de Richter et al 2017), but relatively modest reduction rates of nitrous oxide have been achieved under similar conditions (Ming et al 2016). The costs of applying photocatalysis on surfaces could be as low as 166 $ t⁻¹ CO₂-eq by 2030 (Ming et al 2022). On the other hand, chlorofluorocarbons, and hydrochlorofluorocarbons have been successfully degraded to halogens, acid halides, and CO₂ in experimental photocatalytic reactors (de Richter et al 2016).

Another method proposed to photocatalytically degrade non-CO₂ GHGs is the injection of iron salt aerosols with a chloride content in the troposphere. Solar radiation would promote the formation of chlorine radicals, which could subsequently oxidise methane, VOCs, and carbon black (Dietrich Oeste et al 2017). Furthermore, the deposition of soluble iron over soils and oceans could lead to increased terrestrial and marine primary productivity (Ming et al 2021). Costs between 2 and 54 $ t⁻¹ CO₂-eq have been estimated for this NETP (Ming et al 2022).

Relative to a baseline without GGR, this NETP can avert damage to human health and ecosystems through two pathways: the decrease in tropospheric ozone formation (due to the oxidation of methane and VOCs), and the prevention of stratospheric ozone depletion, linked to chlorofluorocarbons and hydrochlorofluorocarbons (Huijbregts et al 2017).