Two materials were prepared for examination. The first was a mortar made with CEM I and <2 mm sand, which acted as our reference material. The second was another mortar containing CEM I, sand and three biomass ashes, where ashes partially replaced CEM I at 7 and 15% w/w (total weight).

Three types of industrial wood-derived biomass residues sourced in the UK were investigated: English oak (Quercus rober), beech (Fagus sylvatica) and English lime (Tilia × europaea) (oak and lime hereafter, respectively). These woody residues were combusted in a Muffle furnace at 800 ± 25°C with a residence time of 4 h to obtain carbon free ashes. These resulting ashes were a concentrated reactive mineral suite particular to each biomass.

The biomass ashes were characterised for selected physical properties (e.g., surface area and ash content) and their chemical (total carbon, elemental oxides and mineral phase) properties. The biomass ashes and their mortar mixtures were examined for their simple physico-chemical properties, elemental and phase-chemistry, compressive strength development and density. The surface area of ashes was determined (Micromeritics Gemini V2.00), and total carbon was analysed by CHN analysis (FLASH EA 1112 Series). The bulk elemental composition was determined by X-ray fluorescence spectrometry (Bruker S2 Puma). The crystalline and amorphous phases in both CEM I and biomass ashes were determined by using X-ray diffraction (Bruker D8 Advance using Cu Kα radiation generated at 40 kV voltage and 40 mA current, between 2° and 80° 2θ). The interpretation of diffractograms was aided by DIFFRACplus EVA software (Bruker AXS), an internal corundum standard and Rietveld refinement using TOPAS V4.1 (Bruker AXS).

Both CEM I and biomass ashes were tested for their reactivity to pure CO₂ at 15% moisture (w/w) and a pressure of ~2 bar. The ashes were exposed to CO₂ for four-separate cycles in a closed pressurised carbonation chamber, with the first three cycles extending to one hour each, and the fourth cycle being 24 h. The uptake of CO₂ in ashes was determined on weight gain (% w/w) basis, following Hills et al. (2020). The theoretical CO₂ uptake was also calculated using stoichiometry following Steinour equation (Huntzinger et al., 2009; Nam et al., 2012). The theoretical uptake was compared with the measured CO₂ uptake in this study.

The mix design consisted of CEM I (CEMEX, United Kingdom), 2 mm sand (S Walsh and Sons) and biomass ash (as a cement replacement at 7 and 15% w/w), and a control with CEM I and 2 mm sand only (Table 2). The cylindrical monoliths (20 mm in height and diameter) were fabricated by hand tamping the mixtures in 3 layers. Monolithic specimens of this size are not covered by civil engineering standards but are appropriate for research purposes. Hand tamping produces relatively low-density products that were pervious to CO₂ gas. The monoliths were divided into two halves, with one being allowed to mature for 28 days in sealed plastic bags, whereas the other was carbonated in 100% CO₂ at 2 bar pressure for 24 h prior to curing. The wood ash containing monolithic samples are given the acronym O-7, B-7, and L-7 for the 7% and O-15, B-15, and L-15 for the 15% CEM I substitutions.

The monoliths were tested in triplicate for their unconfined compressive strengths after 1, 7, and 28 days after being formed (ELE International, 65KN compression testing apparatus). The monoliths were tested for their mechanical properties (including density and unconfined compressive strength) and CO₂-reactivity. Dry density of oven dried monolith was determined using mass/volume relationship, with the volume calculated for each individual monolith using the diameter and length relationship measured using digital callipers. For each ash, 36 monoliths were broken plus 18 for the control (includes both carbonated and non-carbonated monoliths), making 126 monoliths in total.

The indicative CO₂ uptake in monoliths was calculated from the calcite content in carbonated monoliths at 28 days using an atomic mass balance approach. It should be recognised that some of the carbonate detected in “uncarbonated” monoliths might have arisen from atmospheric carbonation during materials handling and preparation. The CO₂ uptake in carbonated monoliths includes both atmospheric and added CO₂.

The biomass ashes and both uncarbonated and carbonated biomass ash-containing monolithic products were investigated by XRD and electron microscopy (Zeiss, Sigma 300VP, with Oxford Instruments X-max^N Energy Dispersive System, with an accelerating voltage of 10 kV and working distance between 8 to 10 mm). The capture of back-scattered electron micrographs and EDAX analysis was performed on polished resin-impregnated blocks.

The physical and chemical characteristics of the biomass ashes are given in Table 3. The ash content was highest in lime ash, being 3.80%, followed by beech (1.26%) and oak (0.83%) (w/w total weight) ashes. The beech ash presented the highest surface area at 4.53 m²/g followed by the lime and oak ashes at 3.28 and 3.8 m²/g, respectively. The total carbon content of oak ash was highest at 8.21 g/kg, followed by lime and beech ashes, at 5.47 and 2.80 g/kg, respectively. The composition of ashes as oxide equivalents is given in Table 4 and shows the calcium oxide (CaO) content was highest in oak (77% w/w), followed by lime (58%) and beech (45%). Calcium oxide is the key mineral indicator of the CO₂ reactivity in ashes and its presence is critical for carbonation to proceed. Lime ash contained the highest silica content at 37% w/w followed by beech (8%) and oak ash (3%).

After each successive cycle of exposure to carbon dioxide gas, a gradual individual decrease in the amount of CO₂ uptake was observed, as mineral phases progressively carbonated. For oak ash, the first hour CO₂ uptake was 25.63 g/kg, which declined to 0.61 g/kg after the fourth cycle of carbonation. In contrast, beech as reacted with approximately the same amount of CO₂ over all four consecutive cycles, ranging between 16.75 and 17.62 g/kg. The cumulative CO₂ uptake in ashes after 24 h is given in Table 5. The key major mineral phases of uncarbonated and carbonated biomass ashes are given in Tables 6, 7.

The mechanical properties of monoliths, including density and compressive strength, are shown in Table 8. Although our primary interest is in comparing carbonated monoliths to elucidate the influence of ashes on compressive strength development, the data for the uncarbonated monoliths is available in the Appendix 1. The results show an increase in compressive strength development over time, both during and after carbonation, with the latter indicating that some residual hydraulic activity in CEM I may have been present. The control sand-cement mix achieved a compressive strength of 740 KPa at 28 days. Compared to control, the compressive strengths of all ash-containing composites were slightly lower at 1, 7, and 28 days of age, except for the L-15 composite which was approximately 9% higher than the control.

The key major mineral phases of uncarbonated and carbonated biomass ash containing monoliths matured after 7 days of curing are given in Table 9. Calcite was the only crystalline carbonate-containing phase detected in carbonated monoliths. The calcite content of O-7, B-7, and L-7, and O-15, B-15, and L-15 was 16.76 and 18.83, 11.28, and 14.12, and 14.61 and 16.26% w/w, respectively. The results for 28 days old samples are not presented. The effects of atmospheric carbonation upon these porous materials after 7 days of age was significant despite being sealed cured in snap-shut plastic bags, with the effect of reducing the discrepancies observed earlier between carbonated and uncarbonated specimens.

The backscattered composite electron-micrograph of carbonated oak ash is given in Figure 1. Figure 1A shows angular polymineralic sand grains (grey tones) and porosity (black). The composite is very porous and the matrix containing carbonate-able CEM I and oak ash is found primarily coating the individual sand-sized aggregate grains. Figure 1B is an element map of the same sample showing the distribution of Ca (green). The image clearly shows where the carbonate-able cement is located as it forms a lace-like distribution pattern around the surfaces of the sand-sized aggregate particles.

The potential of ashes to mineralise CO₂ was calculated theoretically, using the Steinour equation (Huntzinger et al., 2009; Nam et al., 2012) which uses the stoichiometry of a mineral “system” to predict the maximum possible carbon (% w/w) “uptake”. The approach assumes that all the mineral oxides in our ashes are available for carbonation (Kashef and Ghoshal, 2013). However, it should be noted that this equation and modifications thereof are not appropriate to all potentially carbonate-able wastes and the predictions are normally greater than can be achieved under laboratory conditions.

The Steinour equation considers that the (hydr-)oxides of Ca, Mg, Na, and K undergo carbonation, while the corresponding carbonates as well as sulphur and chlorine compounds (expressed as CaCO₃, SO₃, and KCl) lower the CO₂ uptake (Hutzinger et al., 2009; Schnabel et al., 2021). Vassilev and Vassileva (2020) studied the theoretical CO₂ uptake of 141 biomass ashes and reported the mean value to be 33.99% (340 kg CO₂/t of biomass ash). However, our calculations for given biomass ashes indicate a much higher theoretical uptake, following Nam et al. (2012). Our calculated theoretical CO₂-uptakes were 69.31, 57.35, and 51.41% (w/w) for the oak, beech and lime ashes, respectively. However, the experimentally derived CO₂ uptake upon 24 h of CO₂ exposure was 32, 52, and 43% (w/w), respectively (Table 4). Schnabel et al. (2021) also reported a lower CO₂ uptake in biomass bottom and fly ashes than predicted and referred to the theoretical values as overestimations.

The beech ash contains less calcium oxide but higher K₂O, MgO, P₂O₅, and Fe₂O₃. One of the consequences of this is that the diffractogram for beech ash was very noisy showing “peaks” that were difficult to identify. Further, the presence of CaO appears lower than expected, and may be due to the presence of both amorphous and crystalline forms. Indeed, the amorphous phase content of beech ash is 70.93% w/w, approximately 31–35% higher than the other two ashes. Interestingly, despite the complex mineralogy of this ash, upon carbonation the diffractogram is more resolved and comparable to the other two ashes, supporting our view that amorphous CO₂-reactive phases were indeed present. The presence of butschliite (K₂Ca(CO₃)₂) indicates the presence of other carbonate phases and portlandite indicates that the hydroxides were formed, inhibiting carbonation.

The carbonation of mineral systems can be regarded essentially completed in three distinct phases involving 8 steps (Maries, 1985; Nam et al., 2012). In our dry carbonation “system” CO₂ gas diffuses and dissolves into a thin film of water (pore fluid) and ionises to H₂CO₃, leading to a drop in pH. As pH drops due to the neutralisation of pore fluid containing dissolved CaO, CaCO₃ is precipitated in pore space present in ash (Tripathi et al., 2020).

According to Vassilev et al. (2021), the formation of secondary carbonates in biomass ashes is due to solid–gas reactions between the CO₂ and Ca, Mg, K, and Na oxyhydroxides generated by thermal degradation. However, CO₂ uptake potential of biomass ashes depends on the decomposition/combustion temperature of biomass varying between 600–900°C (Vassilev and Vassileva, 2020). During the thermal degradation of biomass between 500 to 900°C, transformation of minerals and their phases present in biomass takes place. This includes combustion of residual char, dehydration and dehydroxylation of biomass minerals, decomposition of carbonates, and initial and partial decomposition of sulphates followed by formation of new active or partially active secondary Ca-, Mg-, K-, and Na-bearing minerals, and phases with potential to react with CO₂ gas (Vassilev et al., 2014).

The carbonation of calcia-containing phases can result in more complex carbonate mineral products other than calcium carbonate (CaCO₃). By way of example, ankerite (Ca(Mg,Fe)(CO₃)₂) and fairchildite (K₂Ca(CO₃)₂) may form upon carbonation. In the present study, the CO₂ mineralised was confirmed by X-ray diffractometry by the presence of calcite and other carbonate containing phases, such as fairchildite (oak ash) and butschiite (beech ash) (Tables 6, 7). The observed maximum CO₂ uptake recorded after exposure of biomass ashes to CO₂ was in the order: beech, lime, and oak (Table 5), which was not in accordance with the calcium oxide content, present in the following decreasing order: oak, lime, and beech (Table 4). It is possible that the unintended formation of portlandite in oak and beech ashes (shown in the XRD in Table 7), limited its carbonation on exposure to CO₂ gas. It is also possible to achieve higher rates of carbonation during prolonged exposure of portlandite to CO₂ gas (Tripathi et al., 2020), but in the present work the carbonation step was limited to 24 h. It is also possible that in these ashes, high amorphous phases consist of carbonates. Vassilev et al. (2021) also reported formation of different carbonate containing phases in biomass ashes.

The mineralisation of CO₂ in the biomass ashes was aided by their finely divided nature, their relatively high reaction surface available (Filho et al., 2009; Castel et al., 2016) and their calcium oxide, and other CO₂-reactive phases present. However, the reaction could be very slow, if CaO is converted into portlandite (Ca(OH)₂) and not “quickly” then to CaCO₃, as might happen if too much water in the sample limited the diffusion of CO₂. The nature of individual biomass may also have influenced carbonate formation (Tripathi et al., 2020).

The backscattered electron-micrographs taken from polished sections of carbonated monolithic samples given in Figure 1 further confirm the carbonation potential of these mixtures through their uniformly distributed Ca content. The grey-scale backscattered electron micrograph (a) showed the fine aggregate grains coated with the carbonate-able matrix as identified by the distribution of Ca in the element map (b). As carbonation proceeds cementation of the fine aggregate particles within the newly carbonated matrix occurs, augmented by the enhanced (CO₂ gas permeable) porosity of the matrix thereby, following the same lace-like distribution pattern.

The compressive strength of all ash-containing carbonated monoliths was higher for all additions at 1, 7, and 28 days of age when compared to their non-carbonated analogues. By way of an example, the 7% ash w/w substitutions gave compressive strength increases in the range 82–672% (Appendix 1). At 28 days of age carbonated monoliths containing O-7, B-7, and L-7 displayed compressive strength increases of 109, 28, and 138%, respectively. For 15% w/w ash, the increase was 210, 97, and 119% at 28 days of age, respectively.

Compressive strength development in ash containing monoliths at early age is ascribed to carbonate cementation (mainly calcite) as described above (Usta et al., 2023). However, in CEM I-only monoliths, strength is not mainly imparted by carbonate formation upon carbonation (calcite formation was lower than for biomass ash containing monoliths), but also by hydration. This indicates that the CEM I was not fully carbonated under the conditions prevailing, and that compressive strength gain was augmented by enhanced hydraulic activity (on exposure to CO₂) from the Le Chatelier principle, where the development of CSH is “forced” by a change in pH (Ali et al., 2015).

The results clearly indicate the conversion of calcium oxide and portlandite to calcite after exposure to CO₂. Calcite contributes to an increase in early strength development, which in this case increased over 28 days for both 7 and 15% ash containing mixes. We observed a lower compressive strength in the O-15 and B-15 monoliths containing, relative to those containing 7% ash (a 3–7% reduction) over 28 days. However, the compressive strength of the L-15 monoliths was 14.5% higher that the L-7 counterpart. The higher compressive strength of lime ash monoliths could be due to the presence of the higher silica content of this ash (Table 4), which may be reactive/pozzolanic in nature.

The European standard for cement (EN 197-1, 2011) reports the use of blended or Portland composite cements in construction. These cements include coal fly ash, natural volcanic materials, burnt oil shale, blast furnace steel slag, zeolites, ceramic wastes, etc., and are used as substitutes for “common” cement clinker (Canpolat et al., 2004; Taylor et al., 2006; Maschowski et al., 2020; Zeng et al., 2022). Based on their physico-chemical properties, these blended materials impart properties to the cement in different ways (Schneider et al., 2011; Sigvardsen and Ottosen, 2019). However, under right conditions and with proper product design, blended cements can be formulated for different applications (Bentz, 2010).

In the present study, it is observed that both 7 and 15% replacements of CEM I by biomass ashes in monoliths have attained/exceeded the selective mechanical properties compared to the CEM I only control (Appendix 1). Thus, the results obtained indicate that biomass ashes have potential to directly replace CEM I up to 15% w/w. Whilst recognising this is a limited study, further investigation of a broader range of key properties of mortar/concrete such as workability, water absorption and other transport properties is required.

In our earlier work (Tripathi et al., 2019, 2020) we showed that a 10% substitution of biomass ash to CEM I might be beneficial in terms of strength development and the level of substitution on CO₂e emissions. In the present work, we show that a 7–15% w/w substitution of the 3 ashes of interest is also beneficial. However, the strengths of oak and beech ash containing monoliths are slightly higher that the CEM I-only controls with a lower replacement of 7% w/w, indicating the optimal dosage rate is yet to be identified.

From the calcite content of monoliths, the CO₂ uptake achieved during carbonation was calculated (Table 9). The CO₂ uptake was derived using the stoichiometry of this phase, i.e., via the relative atomic weights of CaO and CO₂. The CO₂ uptake for O, B, and L at 7 and 15% substitution of CEM I was 7.37 and 8.29, 4.96 and 6.22, and 6.43 and 7.15% w/w, respectively. The corresponding unconfined compressive strengths recorded at 28 days were all within the range of ~80–94% of the control, except for L-15, which was 107% of the control.

One tonne of quick lime has the theoretical capacity to react with 799 kg of CO₂ to form calcite (comprising 56% CaO and 44% CO₂). With other secondary minerals including Ca-, Mg-, K-, and Na-bearing oxides, hydroxides, silicates, phosphates, and inorganic amorphous material, biomass ash can react with additional CO₂ gas (Vassilev et al., 2021). Due to carbonate cementation of biomass-contained minerals, replacing reactive calcium oxide (normally present in cement phases) as the fundamental raw material in cement (in a carbonation application) is potentially attractive. Low carbon materials utilising waste are employed elsewhere as geopolymer cements, blended cements, and magnesium cements (Lippiatt et al., 2020).

The conceptual diagram shown in Figure 2 summarises our observations on the potential of biomass ashes to act as a cement replacement in a carbonated system. This figure considers the currently available data for biomass ashes from biomass energy plants, following Zhai et al. (2021) and assumptions are made to quantify the CO₂ emissions from the burning of biomass. The figure also incorporates CO₂ saving potential (assuming 35–40% CO₂ uptake in biomass ashes), from our data given in Table 5, following Tripathi et al. (2020) and Hills et al. (2020). This also includes the CO₂ avoidance potential based on assumptions made for CO₂ emissions associated with transportation and landfilling, follows Zheng et al. (2022). This high-level calculation indicates that 150 Mt of CO₂-reactive mineral containing biomass ashes (derived from the 3 Gt of biomass used for energy production, from Zhai et al. (2021)) have potential to:

Although the amount of CO₂ emitted during incineration cannot be compensated by the mineralisation of CO₂ in ashes or that saved by their replacement for cement, the CO₂ offsets that can be achieved by low-carbon products are significant and much higher than the calculations presented in this paper. This study does not include the total CO₂ emissions and environmental implications of biomass burning for energy plants and has only considered landfilling of biomass ashes. Similarly, the implications of developing low-carbon materials to partially replace cement and natural aggregates are significant but have not been taken into account in this paper (see Figure 2).

Ash transport has environmental and economic implications, and we have undertaken the example from Zheng et al. (2022), where an emission of 0.1618 kg of CO₂ equivalent (CO₂e) per tonne-mile was calculated for freight transport of a (full) 15 tonne load and a 0% loading on the return trip. Bastian (2020) reported that the cost of transporting ash for the bioenergy industry in the UK is around £4 million, depending on the ash properties, with cost of fly ash being significantly higher than bottom ash. Bastian (2020) also reported that in the UK, the cost of ash disposal to a distance of 140–170 miles (including transport, landfilling and landfill tax) in two biomass plants ranged from £140–210 per tonne.

With the significant quantity of ash residue generated from biomass energy production, their ultimate disposal is energy intensive, involving transport to a landfill site, landfill engineering costs, site closure/maintenance and further potential emissions. Landfilling requirements can be minimised if ash can be used for its self-cementing properties, or in a blended system (Tripathi et al., 2020). In the UK and elsewhere, landfill capacity is challenged, costs are rising, and alternative management scenarios are needed for the 21st century (Eyre-Walker, 2018). The valorisation of biomass ash residues contributes to sustainable environmental and economic goals. Hope et al. (2017) performed a cost analysis for biomass to energy ash disposal in Canada and reported the landfill cost to be $77 per tonne. However, this was lower than the cost of ash application to forest sites ($92 per tonne) (Hope et al., 2017).