Seven desert dust samples were examined, comprising Saharan dust, Arizona Test Dust (ATD, nominal 0–3 µm fraction), Luochuan loess (LC loess), Taklimakan dust (TK dust), Gobi dust (GB dust), Qinghai dust (QH dust), and Tibetan Plateau dust (Tibet dust). Saharan dust was collected at the Cape Verde Islands (Chen et al., 2020), ATD was purchased from Powder Technology, Inc., USA, and the last five dust samples were collected from different regions in China. Desert dust samples used in this work were collected from topsoil in corresponding regions, and the volume-equivalent diameters were in the range of 1.1–87.2 µm (Tang et al., 2019; Chen et al., 2020).

Among the 28 power plant coal fly ash samples we studied (Table S3), 26 samples were provided by large coal-fired power plants from 25 provinces in China (Li et al., 2026), one sample was the certified reference material (GBW08401) provided by the Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences (Li et al., 2022), and one sample (CFA-FDU) was provided by Fudan University (Wu et al., 2023). These samples were obtained from baghouse rows or electrostatic precipitators in coal power plants (Li et al., 2026), and their volume equivalent diameters were in the range of 16.9–67.6 µm (Table S3).

The 21 steelwork fly ash samples (Table S4) were provided by large and medium-sized steelwork plants located in different regions in China (Li et al., 2026), and their volume-equivalent diameters were in the range of 4.6 to 176.4 µm (Table S4). These samples were emitted from several major steelwork processes, including coking (n=3), sintering (n=5), blast furnace (n=2), converter (n=2), post-steelmaking (n=1), and casting (n=1). Samples could be provided by the same plant or obtained from the same/similar steelwork processes, while there were no two samples obtained from the same process in the same plant. Among the 21 samples, 12 samples were collected from baghouses or electrostatic precipitators, 2 blast furnace fly ash samples were obtained from gravity settlers, and the other 7 samples were collected from chimney outlets. All the samples were dried in an oven (65 °C) for 24 h, and stored in sealed vessels for further analysis.

We employed the device depicted in Fig. 1 to generate and collect biofuel burning aerosols, and more details can be found elsewhere (Li et al., 2026). It should be noted that biofuel burning experiments this work conducted were designed to simulate domestic biofuel burning, and may not be representative of wildfires (Hamilton et al., 2022). As shown in Fig. 1, biofuel was burned in a commercial cook stove, and the smoke generated from biofuel burning passed through a horizontal metal chimney (200 cm in length, 30 cm in inner diameter) and then entered a chamber (50 cm in height, 45 cm in diameter). Air in the chamber was then drawn through a filter bag using a vacuum cleaner to collect aerosol particles. Being different from our previous work (Li et al., 2026) which used a medium volume aerosol sampler to collect PM_2.5 samples, the present study employed a vacuum cleaner to collect aerosol particles emitted from biofuel burning, and the purpose was to obtain a sufficient amount of Fe-containing particles for Fe isotopic analysis. Since no size selection was applied, the samples represent total suspended particles (TSP) from biofuel burning. Biofuel examined in this work contained no apparent soil particles and was burned in a stove, and the influence of Fe contained in soil particles was minimized.

Prior to each sample collection, a new filter bag was used for the vacuum cleaner and the sampling tubing was cleaned with ultrapure water, in order to avoid cross-contamination between samples. Each sample was collected over a period of 6–12 h, with final sample mass of 2–5 g. After sampling, filter bags loaded with particles were placed in a desiccator at room temperature for 48 h to remove moisture, and then stored in sealed vessels for further analysis.

This work examined 11 types of biofuel commonly found in China, consisting of 5 crop straws (wheat, maize, peanut, soybean, and rice) and 6 woods (Chinese fir, birch, pine, lychee, apple, and pear). One aerosol sample was collected for each biofuel, and thus in total 11 biofuel burning aerosol samples were collected.

Moreover, this study measured the Fe isotope compositions of another 4 anthropogenic samples. They were a municipal waste fly ash sample (MWFA-FDU) provided by Fudan University (Ding et al., 2019), a certified reference material for waste incineration fly ash (BCR-615) from the Institute for Reference Materials and Measurements (IRMM), a heavy oil bottom ash sample provided by Fudan University (Fu et al., 2012), and a certificated reference material for urban particulate matter (NIST 1648a) from the National Institute of Standards and Technology (NIST). Their volume-equivalent diameters were 115.9, 21.2, 15.4 and 5.9 µm, respectively.

Due to low Fe content and high organic matter content, biofuel burning aerosol samples were first heated in a muffle furnace at 900 °C for at least 90 min to remove organic matters they contained; after being cooled down to room temperature, for each sample, 0.5 g of the sample was weighed and then used for subsequent analysis. All the other samples were heated in a muffle furnace at 600 °C for at least 60 min; after being cooled down, for each sample, 10–35 mg was weighed and then used for further analysis.

As detailed in the Supplement (Sect. S1.2), desert dust, power plant coal fly ash, municipal waste fly ash, oil bottom ash, and urban particulate matter samples were digested by a sequential protocol at 120 °C, using (1) HNO3-HF, (2) H2O2-HNO3, (3) Aqua Regia, and (4) HCl, respectively. Steelwork fly ash and biofuel burning aerosol samples contained substantial amounts of carbonaceous or refractory materials, and thus required microwave-assisted digestion to achieve complete dissolution. As a result, these samples were digested using the following sequential digestion protocol, namely (1) pre-digestion in HNO3-H2O2 at room temperature, (2) microwave-assisted digestion in HNO3-HF, (3) heated dissolution in Aqua Regia, and (4) heated dissolution in HCl. After digestion, the digestates were diluted with 2.5 molL-1 HCl and then subjected to centrifugation, and the resulting solutions were used for the subsequent chemical purification.

Chemical purification, which minimizes matrix effects and enhances the accuracy of Fe isotope analysis, is an essential step for high-precision Fe isotope analysis (Conway et al., 2013; Zhang et al., 2025). In this study, chemical purification of Fe was achieved using a single column filled with AGMP-50 cation exchange resin (Bio-Rad, USA), as described in the Supplement (Sect. S1.3), and further details could be found elsewhere (Zhu et al., 2020). In brief, the resin column was sequentially cleaned with 6 molL-1 HNO3 (10 mL), 6 molL-1 HCl (15 mL), and ultrapure water (5 mL), respectively. It was then conditioned with 0.2 molL-1 HCl (1 mL), and after that 100 µL sample solution was loaded into the resin column. Subsequently, HCl-HF (both at 0.2 molL-1, 3 mL) and HCl-HF solution (which contained 0.2 and 0.5 molL-1 HCl and HF, 1 mL) were sequentially added into the resin column to elute matrix elements. Finally, 7 mL HCl-HF solution (which contained 0.2 and 0.5 molL-1 HCl and HF) was added into the resin column to elute Fe, and the Fe-containing fraction was collected into a clean PFA beaker.

The Fe-containing fraction was then evaporated to dryness, dissolved with 200 µL concentrated HNO3, and evaporated again to dryness at 120 °C. The obtained residue was then dissolved in 2 % HNO3 for Fe isotope measurement. For the purification procedure employed in this study, the Fe recovery was >98 % and the procedural blank was <6 ng (Zhu et al., 2020). This procedural blank was negligible when compared to the amount of Fe our samples contained (>300 ng).

For the 28 power plant coal fly ash samples we investigated, δ56Fe ranged from -0.20 ‰ to +0.57 ‰ (Table S3), showing considerable variation, and the median and average values were found to be +0.28 ‰ and +0.26 ‰ ± 0.18 ‰ (1σ) (Table 3). The three coal fly ash samples Mead et al. (2013) investigated had δ56Fe values of +0.22 ‰ ± 0.10 ‰, +0.22 ‰ ± 0.06 ‰, and +0.61 ‰ ± 0.08 ‰, with the average being +0.35 ‰ ± 0.23 ‰. Li et al. (2022) found δ56Fe to be +0.05 ‰ ± 0.08 ‰, +0.20 ‰ ± 0.07 ‰, and +0.75 ‰ ± 0.01 ‰ for one Chinese and two American coal fly ash samples, with an average value of +0.33 ‰ ± 0.37 ‰. Our measured δ56Fe values have large overlaps with these reported by the two previous studies (Mead et al., 2013; Li et al., 2022), although the average δ56Fe we reported is slightly lower.

As shown in Table 3, our work suggests that the δ56Fe values have substantial overlaps for power plant coal fly ash (-0.20 ‰ to +0.57 ‰) and desert dust (+0.01 ‰ to +0.30 ‰), while the average δ56Fe is higher for coal fly ash (+0.26 ‰ ± 0.18 ‰) than desert dust (+0.14 ‰ ± 0.10 ‰). This implies that coal fly ash Fe could be isotopically heavier (at least not lighter) than desert dust Fe.

The 21 steelwork fly ash samples we studied showed large variation in δ56Fe (-1.16 ‰ to +0.43 ‰) (Table S4), with average and median values being +0.00 ‰ ± 0.42 ‰ (1σ) and +0.09 ‰. Among the 21 samples, δ56Fe were measured to be +0.39 ‰, +0.42 ‰, and +0.43 ‰ for the three coking fly ash samples, with an average of +0.42 ‰ ± 0.04 ‰ (1σ), being much higher than the other 18 steelwork fly ash samples. This difference can be attributed to the fact that coking process (during which coal is pyrolyzed in an oxygen-free environment to produce coke) uses coal as the main raw material (Keboletse et al., 2021), and thus Fe composition of the generated fly ash is different from that produced in other steelwork processes. In fact, the average δ56Fe (+0.42 ‰ ± 0.04 ‰, 1σ) we determined for coking fly ash is rather close to that for power plant coal fly ash (+0.26 ‰ ± 0.18 ‰, 1σ), implying that they originated from the same raw material (i.e. coal).

For the other 18 steelwork fly ash samples we examined, δ56Fe ranged from -1.16 ‰ to +0.29 ‰, and the median and average values were determined to be +0.04 ‰ and -0.07 ‰ ± 0.41 ‰ (1σ) (Table 3). The average δ56Fe value was reported to be +0.08 ‰ ± 0.24 ‰ (n=10) (Flament et al., 2008) and -0.12 ‰ ± 0.08 ‰ (n=4) for steelwork fly ash (Maters et al., 2022), aligning fairly well with our result. Compared to desert dust, although a few samples were enriched in isotopically heavier Fe, most steelwork fly ash in general shows slightly lower or similar δ56Fe values. Therefore, aerosol Fe emitted by steelworks may be a potential source of isotopically light Fe observed in the tropospheric aerosols.

This study determined the Fe isotopic composition of 11 biofuel burning aerosol samples (Table S5). As mentioned in Sect. 2.1.2, biofuel burning experiments conducted in this work were designed to simulate domestic biofuel burning, and may not be representative of wildfires which can also entrain soil particles into the atmosphere (Hamilton et al., 2022; Bunnell et al., 2025). The lowest and highest δ56Fe values (-1.23 ‰ versus +0.21 ‰) were found for pear wood and soybean straw burning aerosol, respectively. As shown in Table 3, the median and average δ56Fe were determined to be -0.15 ‰ and -0.28 ‰ ± 0.39 ‰ (1σ), lower than that of desert dust (+0.14 ‰ ± 0.10 ‰). Most biofuel burning aerosol samples we examined exhibited δ56Fe below 0 ‰, showing significant enrichment of isotopically lighter Fe. In contrast, a previous study (Kurisu and Takahashi, 2019) found average δ56Fe to be +0.09 ‰ ± 0.03 ‰ for reed residual ash (and +0.08 ‰ ± 0.10 ‰ for reed), being very similar to the UCC and thus indicating no obvious enrichment of lighter Fe.

Plants use Strategy I or II to absorb Fe from soil (Kobayashi and Nishizawa, 2012). It was suggested that compared to the soil where plants are grown, Strategy I plants are usually enriched in isotopically lighter Fe while Strategy II plants are relatively enriched in heavier Fe (Guelke and Von Blanckenburg, 2007). Eight Strategy I plants, namely peanut, soybean, birch, lychee wood, apple wood, pear wood, fir, and pine, were examined in our work, and the average δ56Fe was determined to be -0.22 ‰ ± 0.47 ‰ (1σ, n=8) for the corresponding biofuel burning aerosols. We also investigated three Strategy II plants, namely wheat, corn, and rice straw, and the average δ56Fe was determined to be -0.33 ‰ ± 0.28 ‰ (1σ, n=3) for the corresponding biofuel burning aerosol. Based on the data our work obtained, our work suggests that δ56Fe values of biofuel burning aerosols generated from Strategy I plants are not significantly different from those for Strategy II plants (p>0.05). Our result seems to deviate from what was reported by Guelke and Von Blanckenburg (2007), probably because biofuel we studied were grown from different soils which may have different Fe isotopic compositions (Johnson et al., 2020). In addition, the number of biofuel burning samples analyzed in our study is relatively limited, and further measurements are required to confirm our finding.

For the two municipal incineration fly ash samples we examined, δ56Fe were measured to be +0.32 ‰ ± 0.06 ‰ and +0.08 ‰ ± 0.13 ‰ (Table 3), respectively, with an average value of +0.20 ‰ ± 0.12 ‰ (1σ). Kurisu et al. (2016b) found δ56Fe to be -0.08 ‰ ± 0.09 ‰ and -0.10 ‰ ± 0.03 ‰ for municipal incineration bottom and fly ash from Japan, slightly lower than our result. Li et al. (2022) found δ56Fe to be +0.10 ‰ ± 0.08 ‰ for a European municipal incineration fly ash sample (BCR-176R), close to that obtained in our current study. Overall, available studies suggest relatively large variations in δ56Fe for municipal incineration fly ash, ranging from -0.10 ‰ to +0.32 ‰.

The δ56Fe value was measured to be +0.38 ‰ ± 0.13 ‰ for the heavy oil bottom ash sample we examined (Table 3), similar to that for oil fly ash (+0.30 ‰ ± 0.17 ‰, n=4) reported by Mead et al. (2013). NIST 1648a and NIST 1649a are both certificated urban particulate matter provided by NIST. The δ56Fe was measured by our work to be +0.08 ‰ ± 0.13 ‰ for NIST 1648a (Table 3), in good agreement with those reported for NIST 1649a (+0.00 ‰ ± 0.03 ‰ and +0.01 ‰ ± 0.12 ‰, respectively) by two previous studies (Beard et al., 2003a; Mead et al., 2013).

Figure 3 summarizes the Fe isotopic composition of natural desert dust and several important types of anthropogenic particles our work investigated. Mead et al. (2013) proposed biomass burning as the most likely source of isotopically light Fe observed in the troposphere, since plant matter is the sole material known to contain light Fe. Nevertheless, at that time no measurement of δ56Fe endmember values was available for biomass burning aerosol. Here we found that δ56Fe values ranged from -1.23 ‰ to +0.21 ‰ (average: -0.28 ‰ ± 0.39 ‰, 1σ) for the 11 biofuel burning aerosol samples examined in this work, lower than natural desert dust. Thus, our work suggests that biofuel burning aerosol could be a potentially important source of light Fe in the troposphere, providing experimental results to support what Mead et al. (2013) proposed. The lowest δ56Fe we reported for domestic biofuel burning was -1.23 ‰, while δ56Fe could be lower than -3 ‰ for soluble Fe in tropospheric aerosols (Kurisu et al., 2016a; Hsieh and Ho, 2024). This indicates that biofuel burning alone cannot explain the very light Fe observed for tropospheric aerosols; in other words, additional sources/processes are needed.

In addition to domestic biofuel burning, there are other types of biomass burning, such as wildfire and open burning of agricultural straws. Our study represents domestic biofuel burning well. On the other hand, soil and biomass Fe both contribute substantially to aerosol Fe emitted by wildfires (Kurisu and Takahashi, 2019; Hamilton et al., 2022). Therefore, the δ56Fe signatures of wildfire aerosol Fe likely reflect a mixture of soil and biomass endmembers, distinct from those we reported for domestic biofuel burning.

The 28 power plant coal fly ash samples we examined had δ56Fe in the range of -0.20 ‰ to +0.57 ‰ (average: +0.26 ‰ ± 0.18 ‰, 1σ), indicating that power plant coal fly ash is usually enriched in heavier Fe. As a result, we suggest that not all the anthropogenic aerosol Fe is isotopically lighter than desert dust Fe, in contrast to what previous studies usually assumed. Tropospheric aerosol Fe typically exhibits similar or lower δ56Fe values relative to desert dust; nevertheless, higher values have been reported in a few studies (Labatut et al., 2014; Bunnell et al., 2025; Camin et al., 2025), the causes of which are still under debate. Based on δ56Fe endmember values our work reported, we suggest that coal-fired power plant emission can potentially (at least partly) explain the higher δ56Fe values observed for aerosol Fe in the troposphere.

The average δ56Fe was -0.07 ‰ ± 0.41 ‰ (1σ) for the 18 steelwork fly ash samples we examined (excluding the three samples collected from the coking process). Compared to desert dust, most steelwork fly ash samples had slightly lower or similar δ56Fe (ranging from -1.16 ‰ to +0.29 ‰) (Fig. 3), and thus may contribute to lighter Fe observed for tropospheric aerosols.

Compared to desert dust, municipal incineration fly ash, heavy oil combustion bottom ash, and urban particulate matter our work examined had similar or higher δ56Fe (Fig. 3), but the small sample sizes preclude us from drawing further conclusions.

Biofuel burning aerosol samples this work used were aerosol particles emitted by biofuel burning in a commercial stove, thereby being of direct atmospheric relevance. Instead of aerosol particles, fly ash samples were used to represent particulate matters emitted by coal-fired power plants and steelwork plants. Their volume mean diameters were mostly a few tens of µm (Li et al., 2026), considerably larger than these for tropospheric aerosols (typically a few µm or less). Kurisu et al. (2016b) found δ56Fe to be much lower for TSP than fly ash collected in an incinerator, possibly due to enrichment of lighter Fe in aerosol particles caused by the evaporation-condensation process in combustion. Lighter Fe preferentially evaporates at high temperatures and subsequently condenses to form Fe-containing nanoparticles, and thus particles with smaller size will be enriched in lighter Fe (Kurisu et al., 2016b; Kurisu et al., 2019). This hypothesis was further supported by a following study (Kurisu et al., 2019), which found that δ56Fe was much lower in the fine mode (<1.3 µm) than the coarse mode (>1.3 µm) for aerosol particles collected at a site which was ∼4 km from a steel plant. If the hypothesis proposed by Kurisu et al. (2016b) is correct, then the actual δ56Fe endmember values for coal-fired power plant and steelwork emission could be lower than these reported in our work.

It is noted that an earlier study (Flament et al., 2008) did not fully support what was proposed by Kurisu et al. (2016b). In fact, Flament et al. (2008) observed no significant Fe fractionation during steelwork processes, and found the average δ56Fe of total Fe to be +0.14 ‰ ± 0.11 ‰ (1σ, n=6) for aerosol particles collected at a site which was ∼5 km from a major steelwork plant, being similar to (or even slightly larger than) that for desert dust Fe. Ambient aerosol particles Flament et al. (2008) investigated may also contain dust particles, and thus the δ56Fe values they reported for ambient aerosols may not fully represent those for aerosol particles emitted by steelwork. On the other hand, if aerosol Fe emitted by steelwork was significantly lighter than dust Fe, one would expect ambient aerosol particles Flament et al. (2008) collected at a site close to a major steelwork plant to exhibit δ56Fe values lower than +0.1 ‰. Further measurements of δ56Fe values of aerosol particles emitted by anthropogenic sources, especially size-resolved aerosol particles, can be very valuable.

Among the 18 steelwork fly ash samples we examined (excluding the three coking fly ash samples), 11 samples were fly ash particles retained in dust collectors, with δ56Fe in the range of -1.11 ‰ to +0.20 ‰; the other 7 samples were aerosol particles emitted into the atmosphere (the last seven samples in Table S4), with δ56Fe in the range of -1.16 ‰ to +0.29 ‰. Fe in the 7 aerosol samples was not isotopically lighter than that in the 11 fly ash samples retained by dust collectors. Nevertheless, this is not necessarily incompatible with the Fe fractionation hypothesis, because these samples were not collected from a single plant and the variation of δ56Fe in raw materials may also play a role.