Cation incorporation into biochar affects alkalinity-driven carbon dioxide removal accounting

Biochar production from biomass may be a scalable and cost-effective means of atmospheric carbon dioxide removal. Current accounting frameworks, however, focus primarily on persistent organic carbon and overlook how pyrolysis alters the return of biomass-derived ions to soils and waters. Here, we quantify a first-order alkalinity-related adjustment to biochar CDR by comparing the inorganic carbon cycling implications of biochar production with a biomass-decay counterfactual. During plant growth, net cation uptake generates soil acidity, which can consume bicarbonate alkalinity and shift the carbonate system toward dissolved CO 2, causing CO 2 emissions upon re-equilibration with the atmosphere. Biomass decay can eventually reverse this acid-base perturbation by returning cations to the biogeochemical cycle. Biochar formation delays or prevents part of this return flux by retaining cations and anions within the biochar matrix, making permanent or prolonging what would otherwise have been a temporary CO 2 source. We estimate the resulting change in inorganic carbon cycling from the cation–anion charge equivalent imbalance associated with the retained elemental inventory relative to recalcitrant organic-carbon storage. Across observations reporting complete Ca, Mg, K, Na, N, and S data (n = 85), the mean alkalinity-related reduction in effective CDR is 3.2% when the retained elemental inventory is assumed to be released in proportion to biochar degradation. If cations and anions are instead retained preferentially as organic carbon is lost, the mean reduction increases to 3.9 and 5.1% when 82 and 63% of organic carbon persists, respectively. The fraction of observations showing a reduction in effective CDR of ≥5% similarly increases from 18.8% under proportional release to 21.2 and 22.4% under the two complete element retention scenarios. Individual biochars span a wide range of compositions and durabilities through time, including cases in which the inorganic carbon effect instead increases estimated net CDR.

Carbon dioxide removal (CDR) is viewed as an essential part of meeting international climate goals (IPCC, 2023; Lehmann et al., 2021). Most pathways that keep peak climate warming below 2°C will require a dramatic and rapid increase in CDR deployment in the coming decades as a supplement to aggressive decarbonization (IPCC, 2023; Lehmann et al., 2021). Although there is an increasing focus on government procurement programs, to date most CDR projects have been developed and funded through voluntary carbon markets (VCMs). However, there have recently been a series of articles strongly critiquing the efficacy of VCMs (Gill-Wiehl et al., 2024; West et al., 2023). Projects aiming to scale organic carbon storage (e.g., forest and soil management) have proven particularly contentious (Wu et al., 2023). In step with a reevaluation of the efficacy of VCMs and a crash in prices for carbon credits, there have been more vocal calls for investment in “durable” CDR-pathways for which carbon is removed for timescales of at least hundreds of years. There has been a particular focus on the production of biochar (Lefebvre et al., 2023; Lehmann et al., 2021), the persistent form of solid, carbonaceous residue of anoxically pyrolyzed organic matter.

Biochar projects can be implemented in most working lands and potentially deployed at scale. For instance, it has been proposed that biochar has the potential to account for three gigatons (Gt = 10 9 tons) of CDR annually (Lefebvre et al., 2023), out of roughly 5–10 Gt/year of CDR necessary by 2050 to meet climate goals (IPCC, 2023). Assuming a nominal price of $100 per ton of biochar derived CDR, this process could rapidly become a >$100 billion dollar/year industry. As CDR marketplaces grow, there has been a proliferation of companies aiming to remove carbon via biochar application to soil. These companies strive to commercially produce biochar as an affordable and effective soil amendment in a wide range of agricultural settings. Currently, there are more than 250 biochar suppliers listed on the largest CDR supplier database, and as of May 2026, these companies have pledged to remove ~16 million tons of carbon dioxide via biochar with more than 1.4 million tons of CDR having already been delivered (CDR, 2026). Furthermore, biochar CDR sales increased nearly three orders of magnitude between 2020 through 2025 (CDR, 2026).

Existing CDR crediting protocols frame biochar as a CDR strategy based upon the comparison of two scenarios. In the first scenario, biomass waste is pyrolyzed to produce biochar and amended to soil as an alternative to disposal, meaning organic carbon is transformed into stable, recalcitrant structures and stored over the long term in a soil environment. In the second (counterfactual) scenario, biomass waste is left to decompose back into the soil as a method of disposal or otherwise lost (e.g., via burning) and the organic carbon is released back into the environment. With respect to biogeochemical storage in soil, this comparison is therefore centered primarily on the fate and persistence of organic carbon, alongside the life-cycle emissions associated with biochar production and deployment (Etter et al., 2025; Schimmelpfennig and Glaser, 2012).

Biomass growth and decay also involve the cycling of inorganic nutrients that affect soil acid–base balance (Marschner, 2011; Tang and Rengel, 2003). In particular, net cation uptake generates soil acidity, which can be reversed when biomass decomposes and associated net cation charge equivalents are returned to the soil. By stabilizing biomass and retaining part of its elemental inventory, biochar can delay or prevent this reversal and thereby preserve an alkalinity deficit that would otherwise be transient. Because acidity release can promote CO 2 release upon carbonate-system re-equilibration, biochar production may impose an inorganic carbon-removal penalty alongside its primary organic-carbon storage benefit. Here, we quantify this first-order alkalinity-related adjustment of carbon storage and compare its magnitude with the durable organic-carbon storage conventionally credited to biochar.

Plant cation cycling and soil acidification

Plants acquire mineral nutrients from the soil solution, many of which are taken up as positively or negatively charged ions. Major nutrient cations include K +, Ca 2+, and Mg 2+, while important nutrient anions include NO 3 −, H 2 PO 4 − / HPO 4 2 −, and SO 4 2 − (Marschner, 2011). Depending on metabolic and structural needs, living plants, including crops, maintain charge balance within their tissue by absorbing and releasing cations and anions (Cunningham, 1964; De Wit et al., 1963). These species become constituent parts of the biomass.

Net cation uptake—the uptake of more positive than negative charge—is compensated by the release of charge-equivalent protons from roots, whereas net anion uptake is accompanied by proton uptake and/or release of base equivalents (Hinsinger et al., 2003; Tang and Rengel, 2003). Plants commonly take up more cation than anion charge during growth, making net cation uptake a well-established source of rhizosphere and soil acidification (De Wit et al., 1963; Hinsinger et al., 2003; Tang and Rengel, 2003). In agricultural systems, this process contributes to long-term soil acidification, particularly where biomass containing excess cation charge is removed through harvest rather than returned to the soil (Tang and Rengel, 2003).

However, in the biomass-decay counterfactual considered here this growth-associated acidification is not necessarily permanent. Decomposition returns the elemental and acid-base inventory stored in plant material to the soil system, and incorporation of plant residues with high excess-cation or ash alkalinity can increase soil pH as these base equivalents are released (Sparling et al., 1999; Tang and Rengel, 2003). Thus, for the first-order growth–decay cycle considered here, decomposition of biomass tends to reverse the acid-base perturbation associated with plant growth.

Effect of biochar pyrolysis on carbon and elemental storage

Biochar is produced by pyrolyzing biomass, an anoxic thermochemical process that converts part of biomass carbon into persistent structures resistant to oxidation and degradation (Joseph et al., 2021). This stabilization of biomass-derived carbon is biochar's primary CDR function, extending carbon residence times in soil far beyond those of the original biomass (Lefebvre et al., 2023; Lehmann et al., 2021). Biochar can also affect soil fertility, pH, water retention, greenhouse-gas emissions, and soil organic-matter dynamics, although these broader effects are not the focus here (Huang et al., 2023; Joseph et al., 2021; Lehmann et al., 2021; Van Zwieten et al., 2015; Wang et al., 2016).

Pyrolysis also transforms and redistributes the biomass elemental inventory. Cations (e.g., Na, K, Ca, Mg) and anions or anion-forming elements (e.g., N, S, P, Cl) can remain embedded in biochar in organometallic complexes, oxides, carbonates, carbonaceous structures, or micropores, depending on feedstock and pyrolysis conditions (Buss et al., 2022; Joseph et al., 2021; Nan et al., 2021; Xu et al., 2017). These mineral- and cation-associated phases contribute to carbon retention and biochar stability (Buss et al., 2022; Nan et al., 2021). In contrast, exchangeable ions are transiently sorbed to charged surfaces and remain responsive to surrounding porewater chemistry (Joseph et al., 2021). Analytical protocols may remove this exchangeable pool by leaching or extraction before quantifying the more strongly retained elemental inventory considered here. The non-extractable and non-leachable ions are the subject of this study.

Relative to biomass decomposition, pyrolysis therefore alters two coupled return fluxes: organic carbon and the biomass elemental inventory. Some biomass constituents are volatilized or partitioned into ash, whereas others remain in the biochar. Nitrogen, for example, can be lost as N 2 O or NO x during pyrolysis or quenching (Li, 2024), while plant-derived cations can be preferentially retained. Consequently, the retained cation-to-anion charge ratio can vary among biochars as a function of feedstock composition and pyrolysis conditions (Buss et al., 2022; Joseph et al., 2021; Nan et al., 2021).

Inorganic CDR implications of biochar generation

The effects of biomass growth and degradation on soil acid–base balance also have consequences for inorganic carbon cycling. In bicarbonate-bearing soils and waters, acidity consumes carbonate alkalinity and shifts carbonate-system speciation toward dissolved CO 2:

Addition of acidity therefore shifts this equilibrium toward the left, titrating bicarbonate to CO 2:

and can ultimately lead to CO 2 evasion as the aqueous system re-equilibrates with the atmosphere. This response does not need to occur where protons are released: acidity can be buffered or transported through soils and hydrological networks before encountering bicarbonate-rich waters, so carbonate-system re-equilibration may occur locally or downstream (DeVries, 2022; Drever, 1982; Zeebe and Wolf-Gladrow, 2001). Over a given carbon accounting horizon, biochar production can therefore delay or prevent reversal of an acidity effect that would otherwise tend to be transient during a complete biomass growth–decay cycle. Although biochar cation and anion storage and charge balance in soil waters are well-established concepts (Buss et al., 2022; Drever, 1982; Nan et al., 2021; Xu et al., 2017; Zeebe and Wolf-Gladrow, 2001), to our knowledge their implications for alkalinity-related biochar CDR accounting have not been quantified. An analogous charge-equivalent framework is widely used for enhanced weathering, where alkalinity generated by mineral dissolution contributes to CDR while acidity inputs consume alkalinity and reduce creditable removal (Beerling et al., 2020; Clarkson et al., 2024; Suhrhoff et al., 2026a).

Here, we apply a first-order inorganic carbon-accounting framework to biochar by comparing biochar production with a biomass-decay counterfactual while holding other background acid–base inputs constant. Our quantitative estimates therefore apply specifically to biomass decay as the counterfactual considered here; actual biochar projects may involve alternative biomass fates, for which the magnitude and timing of the adjustment would require project-specific evaluation. We term the resulting difference the “alkalinity-driven CDR adjustment” and compare it with durable organic-carbon storage. Using a novel database pairing biochar carbon content with incorporated cation and anion concentrations, we quantify the magnitude and variability of this adjustment arising from the net elemental charge retained in biochar over the accounting horizon.

Biochar database and elemental data

Based on a literature review of diverse types of biochar (see data SI), we collected information about elemental composition of base cations (i.e., Ca 2+, Mg 2+, K +, Na +), anions (i.e., N and S when provided), and the biochar's percent organic carbon. Manure- and animal-litter-derived biochars were retained in the dataset. For these feedstocks, we assume that the measured cation–anion imbalance pre-dominantly reflects elements originally incorporated into biomass and subsequently transferred through animal feed, such that the same biomass-growth–decay acid–base framework can be applied. Elemental concentrations were converted to molar concentrations and then to charge equivalents according to the valence assigned to each element. Biochar's cation concentrations in the database are from studies that first leached out the exchangeable fraction of cations and then carried out elemental analysis of the biochar. Elemental concentrations were derived from either X-ray fluorescence (XRF) analysis or elemental analysis following acid digestion (Abdel-Fattah et al., 2015; Alburquerque et al., 2016; Al-Rabaiai et al., 2024; Altikat et al., 2024; Al-Wabel et al., 2013; Andelini et al., 2023; Anjum et al., 2014; Boraah et al., 2023; Brantley et al., 2016; Butnan et al., 2015; Cantrell et al., 2012; Drané et al., 2024; Enders et al., 2012; Fernández et al., 2014; Jia et al., 2012; Katuwal et al., 2022; Kwoczynski et al., 2024; Lataf et al., 2022; Li and Shangguan, 2018; Macdonald et al., 2014; Mohanty et al., 2013; Naeem et al., 2014; Novak et al., 2013; Noyce et al., 2017; Nwajiaku et al., 2018; Obey et al., 2022; Omotade et al., 2020; Singh et al., 2010; Subedi et al., 2016; Tepecik et al., 2024; Usman et al., 2015; Wang K. et al., 2020; Wang et al., 2013; Williams and Thomas, 2023; Wrobel-Tobiszewska et al., 2015; Xiao et al., 2018; Yang et al., 2023; Zhang et al., 2020; Zhao et al., 2018). Because the charge-equivalent calculation is based on total elemental inventories rather than the molecular species present after pyrolysis, post-pyrolysis molecular speciation was not used in the calculation. Acid digestion provides total elemental concentrations irrespective of molecular speciation, while XRF similarly quantifies elemental composition of the analyzed solid. We restricted the literature compilation to studies published from 2005 onward; this criterion had little effect on database size because most biochar characterization studies were published after this date. Following Woolf et al. (2021), pyrolysis products containing less than 10 wt.% organic carbon were excluded from the database.

Net cation uptake impact on CDR

Excess cation charge of biochar was calculated by subtracting anionic charge (nitrogen, sulfur) from cationic charge (calcium, magnesium, potassium, sodium):

where n x is the molar concentration of element x in the biochar (mol kg −1 biochar), and Q net is expressed as mol charge equivalents kg −1 biochar. Importantly, Q net does not represent a net electrical charge of the biomass or biochar. Rather, it represents the net ionic charge imbalance associated with uptake of the quantified cations and anions from soil porewaters during plant growth. Electroneutrality is maintained during uptake through compensating processes, including proton release or uptake by roots (De Wit et al., 1963; Hinsinger et al., 2003; Tang and Rengel, 2003); thus, Q net is an acid–base accounting balance calculated from the measured elemental inventory under the ionic-equivalent assumptions specified below and is used as a first-order proxy for the net ionic imbalance associated with biomass nutrient uptake and return. Positive values indicate an excess of retained cation charge, whereas negative values indicate an excess of retained anion-equivalent charge.

For this first-order accounting, Ca, Mg, K, and Na are treated as Ca 2+, Mg 2+, K +, and Na +, respectively. Total reported N and S are assigned nitrate- and sulfate-equivalent charge in the biomass-decay counterfactual, corresponding to one negative charge equivalent per mole of N and two negative charge equivalents per mole of S. This represents an end-member acid–base accounting assumption regarding the forms in which biomass-derived N and S ultimately enter the aqueous biogeochemical cycle, rather than an assumption about their molecular forms within biochar which is not relevant to this analysis. Subsequent processes—including nitrogen transformations, nutrient re-uptake and leaching, cation exchange processes, and mineral precipitation and dissolution (Calabrese et al., 2022; Jahanzad et al., 2016; Redin et al., 2014; Suhrhoff et al., 2026a)—can generate or consume acidity and therefore modify the ultimate inorganic carbon response but are not assessed here.

Chlorine, a plant micronutrient (Broadley et al., 2012), and phosphorus were excluded from the database because both are inconsistently reported in the biochar elemental-composition literature. Because Cl and P are not represented, calculated net cation charge may be biased positive where these elements contribute materially to the retained anion-equivalent inventory; the available dataset does not permit the magnitude of this potential bias to be quantified systematically, but since these are minor components of biomass (Marschner, 2011) we expect the impact of this accounting choice to be low.

The measured Q net describes the initial cation-anion equivalent imbalance associated with the reported elemental inventory of the biochar. To account explicitly for release of this inventory as biochar mineralizes, we define r ion as the fraction of the initial equivalent imbalance released back to the biogeochemical cycle over the 100-year accounting horizon, and f ion as the fraction that remains withheld in the biochar-associated pool:

The fraction of the initial equivalent charge imbalance that is withheld was converted to an equivalent alkalinity-related CO 2 adjustment. For this first-order calculation, consistent with standard accounting in enhanced weathering CDR quantification (Clarkson et al., 2024; Suhrhoff et al., 2026b) and the carbonate-system assumption described in Section 1.3, one mole of retained excess cation equivalents was treated as one equivalent of foregone alkalinity export due to plant-based soil acidification and approximated as one mole of foregone CO 2 removal. The alkalinity-related CO 2 adjustment remaining at the accounting horizon was therefore calculated as:

where M C O 2 = 44.009 g mol −1. Positive values represent foregone CO 2 removal relative to the biomass-decay counterfactual, whereas negative values represent an increase in calculated CO 2 removal. This formulation separates the initial elemental equivalent imbalance from the fraction of that imbalance that remains withheld over the accounting horizon.

Organic-carbon persistence and overall adjustment of biochar CDR

The overall effect of the alkalinity-driven adjustment on net biochar CDR depends on the persistence of both components of the accounting: the net elemental charge retained over the accounting horizon and the fraction of organic carbon that remains unmineralized. The relative adjustment is therefore controlled by the amount of persistent net elemental charge per unit of persistent organic carbon. To quantify the latter, we calculated durable organic-CDR from the measured organic-carbon content and the fraction assumed to remain unmineralized over 100 years. Following Woolf et al. (2021), we evaluated two end-member persistence assumptions: 63% (low recalcitrance) and 82% (high recalcitrance). We denote this organic-carbon persistence fraction as f rec. For each biochar, we estimate CO 2, org as:

where C org is the measured organic-carbon content of the biochar (wt.%), f rec is either 0.63 or 0.82 (Woolf et al., 2021), and M C = 12.011 g mol −1. The resulting CO 2 org is expressed as grams of CO 2 -equivalent durably stored per kilogram of biochar. The alkalinity-related CDR adjustment (A CDR) was then expressed relative to durable organic-carbon storage. Substituting the expressions above makes explicit that the relative adjustment depends on the net elemental charge-to-organic-carbon ratio, scaled by their respective persistence fractions:

Accordingly, a positive percentage represents a reduction in net CDR relative to organic-carbon storage alone, whereas a negative value represents an increase in calculated CDR under the first-order accounting assumptions, representing an excess of anion charge compared to cation content.

Biochar degradation scenarios

Long-term constraints on the relative release rates of biochar organic carbon and matrix-associated elemental inventories are limited. We therefore evaluate three bracketing persistence scenarios. In the first, proportional-release scenario, release of the cation-anion equivalent imbalance scales proportionally with organic carbon loss from biochar degradation, such that r ion = 1– f rec and therefore f ion = f rec. Under this assumption, the persistence fractions cancel from the relative CDR adjustment:

Consequently, the proportional-release scenario gives the same relative alkalinity-related CDR adjustment for f rec = 0.63 and f rec = 0.82, although the absolute amounts of durable organic-carbon storage and the remaining alkalinity-related CO 2 adjustment both scale with f rec.

There is a mechanistic basis to consider proportional release a conservative lower-bound assumption for the magnitude of the relative adjustment. Mineral cations and mineral-associated phases can participate in organometallic associations, protective phases, and other structures that enhance carbon retention and biochar stability (Buss et al., 2022; Joseph et al., 2021; Nan et al., 2021; Xu et al., 2017). Because cations may contribute to biochar stability through cation bridging, carbon that is less strongly stabilized by these interactions may be preferentially degraded, leading to preferential retention of cations in the residual biochar. Hence, we evaluate two additional scenarios that together with the proportional release scenario bracket the plausible range of biochar degradation processes in which f ion = 1 over 100 years while organic carbon persists at either f rec = 0.63 or f rec = 0.82 (see Table 1).

Primary summary statistics were calculated for observations reporting Ca, Mg, K, Na, N, and S together (n = 85). This complete-element subset minimizes directional bias from omission of individual represented ion pools and is therefore used for the main estimates under each persistence scenario. To characterize the sensitivity of inferred adjustments to the incomplete elemental reporting common in the source literature, we additionally analyzed three broader datasets: (i) all observations using whatever subset of Ca, Mg, K, Na, N, and S was reported for each biochar (n = 215); (ii) observations reporting both N and S, with the cation-equivalent balance calculated from whichever of Ca, Mg, K, and Na were reported (n = 121); and (iii) observations reporting N, with the cation-equivalent balance calculated from whichever of Ca, Mg, K, and Na were reported (n = 194). Summary values were reported as means ± sample standard deviations and medians. To evaluate sensitivity to unequal representation among source studies, observations in the complete-element primary subset were additionally averaged within study before summarizing across the represented studies. Analytical measurement uncertainty was not propagated because measurement uncertainties are not reported consistently across the source literature.

Our analysis suggests that net cation incorporation during biochar production can have a non-negligible effect on the overall CDR associated with biochar amendment to soil. Our database comprises 215 biochar observations drawn from 39 source publications, some of which are review papers compiling data from additional studies. Our primary analysis was restricted to the 85 observations reporting Ca, Mg, K, Na, N, and S together, representing eight source studies. Under the proportional-release scenario (f ion = f rec), the mean A CDR was 3.2 ± 7.4%, with a median of 0.74% and a range from −7.8 to 39.1% (Figure 1). Assuming complete retention of the measured cation–anion imbalance over the accounting horizon (f ion = 1) increased the mean A CDR to 3.9 ± 9.0% at f rec = 0.82 and 5.1 ± 11.7% at f rec = 0.63, with corresponding medians of 0.91 and 1.18% and ranges of −9.5 to 47.7% and −12.3 to 62.1%, respectively (Figure 1). The proportion of observations with A CDR ≥5% was 18.8, 21.2, and 22.4% under proportional release, complete elemental retention with f rec = 0.82, and complete elemental retention with f rec = 0.63, respectively (Figure 2a).

For a specific example, a grapevine-residue biochar containing 77.2% organic carbon contained 17.38 mol of base-cation charge and 1.67 mol of anion-equivalent charge per kilogram of biochar, giving Q net = 15.71 mol charge equivalents kg −1. This corresponds to an initial alkalinity-related CO 2 adjustment of 691 gCO 2 eq kg −1 biochar. At f rec = 0.82 and 0.63, durable organic-carbon storage is 2,319 and 1,782 gCO 2 eq kg −1 biochar, respectively. Under proportional release, the remaining alkalinity-related CO 2 adjustment scales to 567 and 436 gCO 2 eq kg −1, yielding the same relative A CDR of 24.4% at both persistence assumptions. If the measured elemental imbalance is instead completely retained (f ion = 1), the full 691 gCO 2 eq kg −1 remains withheld, increasing A CDR to 29.8% at f rec = 0.82 and 38.8% at f rec = 0.63. In contrast, a eucalyptus-wood biochar containing 69.7% organic carbon contained 0.67 mol of base-cation charge and 0.17 mol of anion-equivalent charge per kilogram, giving Q net = 0.50 mol charge equivalents kg −1 and an initial alkalinity-related adjustment of only 22.1 gCO 2 eq kg −1. Its A CDR is consequently 0.87% under proportional release and 1.06 and 1.37% under complete elemental retention at f rec = 0.82 and 0.63, respectively.

A biochar with a negative A CDR contains more assigned anion-equivalent charge than base-cation charge and, under the first-order accounting assumptions used here, would therefore increase calculated CDR relative to accounting based on recalcitrant organic carbon alone. Within the complete six-element subset, 25 of 85 observations (29.4%) had negative A CDR. Because the three persistence scenarios scale the same underlying cation–anion balance without changing its sign, this fraction is identical among scenarios. Only two observations (2.4%) had A CDR < -5% in each of the three scenarios, whereas the positive tail extended to 39.1, 47.7, and 62.1% under proportional release and the two complete-element-retention scenarios, respectively (Figures 1, 2).

The estimated magnitude and distribution of A CDR were sensitive to elemental-data completeness (Figures 2, 3). When all 215 observations were analyzed using whatever subset of Ca, Mg, K, Na, N, and S was reported for each biochar, mean A CDR values were 3.8 ± 7.2%, 4.6 ± 8.8%, and 6.0 ± 11.5% under proportional release, complete elemental retention with f rec = 0.82, and complete elemental retention with f rec = 0.63, respectively; corresponding medians were 1.18, 1.44, 1.88%. In this available-element analysis, 26.0, 30.2, 34.0% of observations had A CDR ≥5%, compared with 18.8, 21.2, and 22.4% in the complete six-element subset (Figure 2). The N+S subset, in which Q net was calculated from whichever cations were reported (n = 121), yielded lower mean A CDR values of 2.52, 3.07, 4.00%, whereas observations reporting N together with any reported cations (n = 194) yielded means of 3.69, 4.51, 5.86% for the same three persistence scenarios (Figure 3).

Our primary analysis is conducted at the individual-biochar level because multiple observations reported within a single study generally represent distinct biochars, feedstocks, pyrolysis conditions, or experimental treatments and therefore contain genuinely different compositional information. To nevertheless evaluate sensitivity to unequal representation among source studies, we additionally averaged the complete six-element observations within study before summarizing across the eight represented studies. Study-level mean A CDR values were 3.5 ± 8.5% under proportional release, 4.3 ± 10.4% under complete elemental retention with f rec = 0.82, and 5.6 ± 13.5% under complete elemental retention with f rec = 0.63, with corresponding medians of 1.98, 2.41, and 3.14%. Study-level means ranged from −7.6 to 21.7%, −9.3 to 26.4%, and −12.0 to 34.4%, respectively. In each scenario, two of eight studies (25%) had a mean A CDR ≥5%. The study-level analysis therefore retains the broad variability and positive mean adjustment observed at the row level, indicating that the result is not solely driven by studies contributing large numbers of observations.

Magnitude, persistence, and reversibility of the alkalinity-related CDR adjustment

We present a data compilation and a simple framework to demonstrate that consideration of cation and anion incorporation into biochar can affect biochar CDR estimates when inorganic carbon cycling is considered alongside persistent organic-carbon storage. The compilation allows us to provide an initial, first order, assessment of the impact of net cation incorporation in biochar on CDR accounting over the 100-year persistence horizon considered here. Our results indicate that excluding this inorganic contribution can lead to either over- or under-estimation of CDR relative to accounting based on persistent organic carbon alone, depending on the net retained charge. However, the consistently positive mean adjustment across persistence scenarios suggests that current biochar CDR accounting approaches that omit this effect are, on average, likely to slightly overestimate net CDR.

The consistency between row- and study-level estimates indicates that the central result is not driven solely by studies contributing large numbers of observations. The differences among our three scenarios highlight that the alkalinity-related CDR adjustment depends on the relative persistence of organic carbon and the relative rate at which the associated elemental inventory is released. In this context, the relative extent to which cations are released upon biochar degradation is an important area of future research. While analyses based on more limited elemental reporting allow us to utilize a larger fraction of the existing data with findings generally showing similar tendencies between subsets, the differences among these estimates also demonstrate that more complete reporting of biochar cation and anion inventories would be valuable in future studies.

A process not explored here is long-term (>100 years) aging of the recalcitrant biochar matrix, which may eventually return part of the stored carbon and elemental inventory to the biogeochemical cycle. The alkalinity-related adjustment may therefore partly or fully reverse if retained cations and anions are released beyond the accounting horizon, representing a delay rather than a permanent reduction in CDR over the full lifetime of the biochar. However, this timing remains relevant (Cabiyo et al., 2026). In the biomass-decay counterfactual, decomposition returns net cation charge and associated base equivalents, allowing recovery of the growth-associated alkalinity deficit within the shorter timescale of plant growth and decomposition. Delaying this return of net cation charge and associated base equivalents through biochar formation correspondingly delays that alkalinity recovery and associated CO 2 removal to a future, higher-CO 2 atmosphere, lowering the overall climate benefit of biochar at the time of biochar production. Thus, even where the adjustment is ultimately reversible, its timing affects the climate benefit and time-bounded accounting of biochar CDR.

An additional process that may influence inorganic carbon cycling is the evolution of biochar surface charge during aging. As biochar surfaces oxidize, new cation-exchange sites can develop, altering cation retention and associated acid–base dynamics (Li et al., 2019; Wang L. et al., 2020). Increased adsorption of base cations at newly developed exchange sites (Qadir et al., 2024) could further modify the alkalinity-related CDR adjustment because, where these sites are initially protonated, their deprotonation and subsequent occupation by base cations can release H+ to soil solution. This represents an additional source of acidity distinct from the delayed return of biomass-derived net cation charge quantified here and could therefore increase the magnitude of the alkalinity-related adjustment. Conversely, cation desorption from exchange sites would also affect base-cation mobility and associated alkalinity dynamics. The magnitude of these effects will depend on the development and protonation state of exchange sites, soil pH, cation availability, and the degree of cation saturation during biochar aging, and is not quantified in our present framework.

Additional processes modifying the first-order alkalinity adjustment

Beyond the relative persistence of organic carbon, the retained elemental inventory, and aging-dependent changes to biochar surface chemistry considered above, several additional biogeochemical processes could modify the magnitude or even the direction of the first-order alkalinity-related CDR adjustment. These processes act either on the subsequent fate of the released elemental inventory or through broader changes to nutrient cycling and requirements, soil buffering, and mineral reactions following biochar deployment.

While cations can form secondary phases upon release or be lost to cation exchange (Calabrese et al., 2022; Kanzaki et al., 2025) and N can undergo subsequent transformations such as denitrification (Chiaravalloti et al., 2023), timescales of elemental release upon biochar degradation and post-release processes modulating these remain poorly constrained. Nevertheless, these processes can substantially alter the fate, timing, and ultimate impacts of the associated alkalinity perturbation during transport through soils and waters. Additionally, total elemental N also does not constrain either the ionic form through which N was acquired during plant growth or the form in which it is returned during decomposition. Because nitrate carries negative charge, treating all measured N as nitrate-equivalent represents a deliberately conservative choice with respect to the net-cation signal quantified here. Uptake as ammonium or neutral organic N would increase the inferred net cation-equivalent imbalance and, all else equal, strengthen the alkalinity related downward adjustment of net CDR. We do not attempt to account for subsequent N cycling following biomass decomposition, including ammonification, nitrification, denitrification, or biological re-uptake, and their associated effects on acidity and alkalinity (Bolan et al., 1991). Rather, our framework isolates the direct first-order consequence of stabilizing biomass and thereby altering the return of its elemental acid–base inventory relative to biomass decomposition.

A further consequence of diverting biomass from decomposition to biochar production is delayed or reduced nutrient recycling. If biomass that would otherwise decompose is converted to persistent biochar, part of its N and P inventory may be withheld from near-term return to soil, potentially increasing external nutrient requirements where these inputs are needed to maintain fertility and productivity. Any resulting increase in mineral-fertilizer demand would constitute an additional life-cycle carbon cost relative to decomposition (Menegat et al., 2022; Walling and Vaneeckhaute, 2020). Biochar also affects N and P cycling through numerous other processes whose magnitude and direction depend on feedstock, pyrolysis conditions, soil properties, and time since application (Gao et al., 2019; Gul and Whalen, 2016). We do not quantify these effects here, but they represent an additional life-cycle consequence of the biochar vs. biomass decay counterfactual that would further reduce net project level CDR.

As biochar CDR projects scale up and expand to a wider range of settings, such as highly acidic (Waheed et al., 2026), calcareous, arid, or submerged soils, biochar's net cation incorporation will interact with soil-specific buffering and transport processes, depending not only on the biomass feedstock and pyrolysis conditions as explained earlier, but also on soil pH, mineralogy, exchange capacity, hydrology, and background alkalinity. The impact of soil variability on the alkalinity-related CDR adjustment and its fate considering downstream processes is an important area for future soil science and aqueous geochemistry research. Dissolution of biochar-associated ash and mineral phases can also return retained base cations to the biogeochemical cycle. To the extent that this represents release of the biomass-derived net cation inventory quantified here, it constitutes a reversal of the alkalinity-related adjustment and is represented at first order by our elemental-release scenarios, including the proportional-release case. However, ash dissolution may also generate alkalinity through mineral phases or processes not captured by the retained cation–anion balance, which could further modify net CDR. Stacking biochar and enhanced weathering as a means of CDR (Buss et al., 2021; Sohng et al., 2025; Meyer zu Drewer et al., 2026) has also been proposed and tested experimentally, further emphasizing the need to resolve interacting organic and inorganic carbon fluxes. We present a data compilation and a simple framework to demonstrate that consideration of cation and nitrogen uptake by biochar can affect the overall CDR efficiency estimates of biochar deployments. This indicates that estimates of long term carbon removal in existing frameworks can be incorrect. However, it is already common to overestimate long-term biochar CDR rates—in cases where biochar preferentially binds Ca and Mg ions. Our data compilation allows us to provide an initial assessment of the impact of base cation trapping in biochar on carbon removal effectiveness at the moment of biochar deployment and payment transfer between biochar credit buyers and suppliers.

A process not explored in this study is the potential impact of biochar's cation and anion exchange capacity (CEC) on base cation cycling and bicarbonate production. Depending on how biochar is aged prior to deployment and ages after deployment, additional cation adsorption at biochar exchange sites and thus a decreased cation flux could further reduce overall carbon removal efficiency (Li et al., 2019; Wang K. et al., 2020). Additionally, the long-term (>100 years) aging of the recalcitrant biochar matrix will eventually release the stored carbon, cations, and nitrogen back into the environment, a consideration also not included in this analysis. While we know that cations can form secondary phases in porewater upon release and that nitrate will enter the aqueous system in the form of nitrate and eventually undergo denitrification, little to no data describing the mechanisms and timescales of the release of these species from biochar exist in the literature. Constraining changes in cation sorption as biochar ages within compost or soil also represents an important area of future work and should be further investigated in projects in which biochar is being used to credit carbon removal.

Carbon crediting and the verification provided by carbon removal crediting platforms (e.g., Verra, Isometric, and Puro) represent the currency of VCMs, allowing purchasers to pay carbon removal suppliers at set prices determined by the life cycle analysis of biochar projects. Although current verification methodologies used to issue biochar carbon credits consider greenhouse gases emitted during feedstock transport, biochar production, transport, and application, and changes to carbon removal permanence as biochar ages (Etter et al., 2025; Schimmelpfennig and Glaser, 2012), no current verification methodologies consider the CDR inefficiency introduced by cationic charge trapping after biochar has been added to soil and the fate of the cations is unaccounted for. The life cycle assessment of these projects may be robust, but this represents a flaw in the carbon accounting of this system (Nordahl et al., 2024). Ultimately, assuming the goal of CDR is carbon removal from the atmosphere, the costs of CDR from biochar have been underestimated and CDR potentials correspondingly overestimated. However, the magnitude of this effect will likely vary considerably between projects, with limited effect on many projects. Therefore, we suggest that biochar CDR quantification frameworks should be updated to consider base cation trapping. Additional research on the long-term effect of this process will be critical for accurate and responsible scaling of carbon removal through biochar amendment to soil.

Biochar CDR is typically quantified primarily from persistent organic-carbon storage, together with project-level life-cycle emissions and permanence considerations. Our results show that changes to inorganic carbon cycling can introduce an additional, and in some cases material, adjustment to net CDR that mostly reflects a net reduction in CDR. The magnitude of this effect depends on biochar composition, the relative persistence of organic carbon and retained elemental inventories, and the biogeochemical context in which biochar is deployed. Where this adjustment is material, it therefore has direct implications for how much CDR can defensibly be attributed to a biochar project and, consequently, how that removal is quantified and credited.

This is particularly relevant because biochar projects are increasingly evaluated through carbon-crediting methodologies on the VCM that translate project-level carbon accounting into issued CDR credits. Current verification frameworks used by platforms such as Verra, Isometric, and Puro account for greenhouse-gas emissions associated with feedstock transport, biochar production, transport, and application, as well as changes in CDR permanence as biochar ages (Etter et al., 2025; Schimmelpfennig and Glaser, 2012). However, to our knowledge current biochar verification methodologies do not explicitly account for the fate of incorporated net cations and the associated alkalinity-related CDR adjustment. The life cycle assessment of these projects may be robust, but this potentially represents an omitted geochemical term in carbon accounting (Nordahl et al., 2024). If the alkalinity-driven adjustment identified here is material under a project's specific feedstock, soil conditions, and counterfactual, omitting it could bias estimates of net CDR and associated crediting. However, the magnitude of this effect will likely vary considerably between projects.

Therefore, we suggest that biochar CDR quantification frameworks should consider the geochemical impacts of net cation incorporation into the biochar matrix and associated inorganic carbon cycling. More broadly, robust carbon accounting requires consideration of biogeochemical effects beyond the process that defines the primary CDR pathway: just as enhanced weathering assessments must account for potential effects on organic carbon cycling (Suhrhoff et al., 2026b; Vienne et al., 2026), biochar assessments should likewise account for consequences for inorganic carbon cycling. Durable biochar CDR is therefore not defined by carbon persistence alone, but by the full biogeochemical consequences of stabilizing that carbon including impacts on inorganic carbon cycling.