1 Introduction

Biomass burning emitted the second largest amount of CO2 (following fossil fuel consumption), accounting for approximately 84% of global carbon emissions in some regions (Su et al. 2023). Those carbon emissions could be reabsorbed during vegetation restoration, and are considered as a net zero carbon emission (Jiang et al. 2024; Santin et al. 2016). However, the incomplete burning of biomass may produce considerable amounts of carbonaceous solid residues, which is known as pyrogenic carbonaceous materials (PCMs) and a heterogeneous continuum ranging from slightly carbonized fractions to highly condensed aromatic components. It is reported that approximately 114–383 Tg of PCMs are generated annually from incomplete biomass burning (Chen et al. 2022b). Because of their recalcitrant nature, PCMs are a potentially stable carbon source for centuries to millennia (Gao et al. 2023).

Studies have shown that PCMs generated intentionally or unintentionally are widely distributed in environmental matrices, especially in soils. The input of PCMs into soils promoted soil carbon sequestration and mitigated climate change (Bolan et al. 2022). However, PCMs are generally overlooked in previous studies on carbon budget and carbon flux assessment, resulting in an inaccurate understanding of PCMs in the global carbon cycle (Coppola et al. 2022; Santin et al. 2016). The possible reason are changes in the properties of PCMs after they enter the environment, or the lack of accurate methods to quantify PCMs. Besides carbon storage, PCMs play various roles in soil ecosystems, such as (1) improving soil structure and enhancing soil function; (2) adsorbing or degrading pollutants, which is beneficial for soil productivity and ecosystem health. Specifically, the released carbon-, nitrogen-, or phosphorus-containing nutrients from PCMs can provide nutrients for plants and animals; the well-developed pore structure of PCMs can act as microbial shelter; the rich functional groups and redox properties of PCMs can immobilize or degrade pollutants. However, some toxic and harmful substances are also present in PCMs. Some (such as heavy metals) are enriched, while some, such as environmentally persistent free radicals (EPFRs) and polycyclic aromatic hydrocarbons (PAHs) are newly formed during biomass combustion. Although some reviews explored the application of PCMs in carbon storage and environmental remediation, they mainly focus on the preparation, characterization and modification of PCMs emphasizing their functional performance (Bolan et al. 2022; He et al. 2021; Hu et al. 2022). The measuring methods, the risk of endogenous pollutants, and the role of PCMs in soil ecosystems are always neglected.

In this review, we first compared different qualitative and quantitative analysis methods of PCMs, then analyzed their roles in carbon storage and ecological systems. The generation mechanism and environmental risks of contaminants (focusing on EPFRs and PAHs) associated with PCMs were also discussed. Finally, some research defects and perspectives were discussed.

2 Residual pyrogenic carbonaceous materials from incomplete biomass burning

Incomplete biomass burning produces considerable amounts of PCMs, which are the products from natural fires, unintentional human activities, and purposefully engineering activities (Pignatello et al. 2017). In the study of PCMs, terms such as biochar, charcoal, carbon black, brown carbon or black carbon, are often used in different research areas. Although they are semantically overlapped, these terms are differentiated from each other based on production conditions and main application purpose. For example, carbon black is a black powder synthesized artificially by partial combustion or thermal decomposition of gaseous or liquid hydrocarbons under controlled conditions. Its composition is nearly pure elemental carbon, widely used in tires, rubber products, inks, coating production (Kim et al. 2018). Brown carbon is defined as light-absorbing organic matter because it is light brown in color and strongly absorbs ultraviolet light. It contains various origins, including pyrolysis products and tarry materials from combustion, humus and humic-like substances from soil or biogenic emissions (Feng et al. 2013; Li et al. 2023b). According to the carbonization degree, PCMs can be divided into slightly charred biomass, char, charcoal and highly condensed soot (Fig. 1a) (Masiello 2004). Char and charcoal are solid residues formed during the combustion of biomass. More specifically, charcoal is usually produced by the pyrolysis of wood, while char is often referred to a substance with a lesser degree of pyrolysis than charcoal, typically as a product of vegetation fire (Han et al. 2020). Soot is produced from gas phase condensation. During the combustion, the carbonized residues (charred materials) and condensates (soot) are formed simultaneously and coexist in environmental matrices. All the components of continuum have the characteristics of high carbon content, high aromatization components and structural heterogeneity. Their particle size, biochemical reactivity and migration capacity vary with the degree of carbonization (Masiello 2004).

Fig. 1
figure 1

Properties (a) and quantification methods (b) of the pyrogenic carbonaceous materials continuum. Modified based on Masiello (2004), Poot et al. (2009), and Wagner et al. (2021). The gray color in Panel (b) indicates that PCMs can only be partially detected, while black color represents that PCMs can be completely detected by the methods. BPCAs stands for benzene polycarboxylic acids

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2.1 Isolation and identification of PCMs

The accurate evaluation of environmental effects for PCMs, such as carbon sequestration and pollutant control, requires precise determination of PCMs properties and content. However, after entering the environmental matrices, PCMs would mix with soil components. The complexity of environmental matrices poses difficulties to measure PCMs mostly because it is difficult to separate PCMs from the environmental matrices. Nevertheless, researchers have developed several PCMs determination methods mainly including thermal oxidation (Gelinas et al. 2001), chemical oxidation (Song et al. 2002), molecular marker (Glaser et al. 1998) and optical methods (Kralovec et al. 2002) (Fig. 1b, Table 1). The thermal and chemical methods focus on the thermal and chemical stability of PCMs. Specific aromatic cluster components could be identified by molecular method, and transmittance/reflectance/the light absorption of a specific component of PCMs is evaluated by optical methods. Obviously, each method focuses on the determination of one or a few major fractions of PCMs.

Table 1 Main analytical methods for PCMs in diverse environmental matrices
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The chemo-thermal oxidation method (CTO), thermos-gravimetric analysis (TGA) and hydrogen pyrolysis (HYPY) are the representative thermal oxidation methods in the determination of PCMs (Table 1). In CTO method, samples are pretreated with HCl and HF to remove inorganic carbon components, and then the non-PCMs organic fractions are oxidized by controlling pyrolysis temperature and heating time in an oxygen-rich atmosphere. The remaining carbon contents are determined as PCMs by elemental analysis. TGA method can distinguish the specific fraction of PCMs in environmental matrices based on the mass loss during pyrolysis at different temperatures. For example, Cuypers et al. (2002) suggested that mass loss at 680–730 °C may be due to coal and soot in soil and sediment. HYPY method is a new technique developed in recent year for the PCMs determination (Geng et al. 2023), which mainly involves rapid catalytic decomposition of samples under high hydrogen pressure (> 15 bar). The carbon of undecomposed fraction (usually from a highly aromatic structure > 7 benzene ring) is defined as PCMs (Cotrufo et al. 2016). Among them, CTO method is the most widely used due to its economy and simplicity of operation. Adequate removal of inorganic carbon and non-PCMs fractions is essential for accurate quantification of PCMs, otherwise, the PCMs content can be overestimated. Gelinas et al. (2001) demineralized the samples with a mixture of HCl and HF and removed hydrolyzable organic matter with trifluoroacetic acid (TFA) and HCl. They found that PCMs content of the original samples was three orders of magnitude higher than that of demineralized samples, which suggest a possible vast variation of this method. In addition, PCMs are continuum containing different compositions, and their pyrolysis potentials vary significantly with heating temperatures and durations. Elmquist et al. (2006) measured the PCMs content of soot, wood biochar and grass biochar at pyrolysis temperatures of 60–600 °C for 18 h. Some non-PCM fractions may be carbonized instead of being removed at pyrolysis temperatures lower than 375 °C, resulting in overestimation of PCMs. They suggested that CTO method is more accurate in measuring PCMs generated at high temperatures. To date, this method has been widely used in the determination of PCMs in soils and sediments (Dan et al. 2022; Eckdahl et al. 2022). Compared with other thermal oxidation technologies, CTO method will still be used as the main technology in PCMs determination for a long time due to its simple operation and low cost (Table 1). It should be noted that, since the thermal stability of each component in PCMs varies, select the appropriate pyrolysis temperature and pyrolysis time is particularly important for the accurate quantification of PCMs.

Chemical reagents with different oxidation strengths are used to remove non-PCMs organic fractions, which is known as chemical oxidation methods. Based on the types of oxidizing reagents, the chemical oxidation methods include hydrogen peroxide/potassium hydroxide (H2O2/KOH) mixture oxidation, HNO3 oxidation, hypochlorite (NaClO) oxidation and potassium dichromate/sulfuric acid (K2Cr2O7/H2SO4) mixture oxidation. Similar to CTO method, before the determination of PCMs, the inorganic mineral components in samples need to be removed by HCl and HF. Compared with other chemical methods, H2O2/KOH and NaClO oxidation methods are rarely used because the highly condensed non-PCMs organic fractions such as kerogen are difficult to be oxidized by H2O2/KOH. HNO3 and K2Cr2O7/H2SO4 oxidation are more widely used, but for highly stable non-PCMs organic fractions, the removal efficiency of HNO3 oxidation is not sufficient, which often leads to an overestimation of PCMs (Middelburg et al. 1999). The oxidation rule of K2Cr2O7/H2SO4 mixture solves this problem well. Wolbach and Anders (1989) found that, depending on the maturation degree, the half-life of kerogen at 50 °C in K2Cr2O7/H2SO4 mixture was 6–180 h, and the half-life of PCMs (soot, charcoal) was generally longer than 600 h. So as long as the reaction time is controlled, the kerogen material in the sample can be fully removed. For different types of PCMs, the half-life varies significantly under a certain heating temperature: the half-life of wood biochar is less than 340 h, and that of straw biochar is less than 1040 h (Bird and Grocke 1997), while the commercial charcoal and soot have half-lives of 547–637 h and 4000–5000 h, respectively (Masiello et al. 2002). Therefore, the reliability of the quantification of PCMs by this method depends mainly on the carbonization degree and material sources.

Optical methods mainly include microscopy, thermal optical reflection/transmission (TOR/TOT) and mid infrared spectroscopy-partial least-squares regression (MIR-PLSR). The elemental composition and its distribution on PCMs surface can be characterized by scanning electron microscope coupled with energy dispersive spectrometer. However, it is difficult to identify very small PCMs particles such as soot (Table 1). TOR/T determines the decomposition of organic carbon or PCMs by monitoring the reflection or transmittance of samples on the filter in the pyrolysis process (Chow et al. 2006; Phairuang et al. 2020). The pyrolysis is divided into two main stages, where organic carbon is decomposed onto four carbon fractions defined as OC1, OC2, OC3 and OC4 in pure He atmosphere at 120, 250, 450 and 550 °C, then stable PCMs are oxidized to BC1, BC2 and BC3 in a 2% O2/98% He atmosphere at three different temperatures (typically was 550, 700 and 800 °C). All carbon components were eventually reduced to methane, and the total amount of PCMs was determined by flame ion detector as the sum of BC1, BC2 and BC3 minus anthropogenic PCMs formed during the conversion between He atmosphere and O2/He atmosphere. BC1 was identified as char, BC2 and BC3 as soot fraction (Fang et al. 2019; Han et al. 2007a, b). This method was only used for the determination PCMs in aerosol in the early stage, and then was extended to the determination of soil and sediment by Han et al. (2007b). The MIR-PLSR is a new technology applied to the PCMs determination in recent years (Table 1). Combining mid infrared spectroscopy and multivariate data analysis, rapid detection of large-scale samples can be achieved without pretreatment such as demineralization. The PCMs content determined by other methods is used to verify and calibrate the accuracy of the model developed by MIR-PLSR method (Bornemann et al. 2008; Cotrufo et al. 2016). Cotrufo et al. (2016) pointed out that this method has a high accuracy in predicting environmental samples with low PCMs content. Except for the widespread application in characterizing PCMs morphological properties, the quantitative determination of PCMs by optical methods is limited due to its complicated analysis process (TOR/T method).

Under high temperature and high pressure, the condensed aromatic structures of PCMs could be oxidized by HNO3 into benzene polycarboxylic acids (BPCAs) with different number of carboxyl group substitutions, such as benzene tricarboxylic acids (B3CAs), benzene tetracarboxylic acids (B4CAs), benzene pentacarboxylic acid (B5CA) and mellitic acid (B6CA). The total amount of BPCAs can reflect the content of PCMs to a certain extent, and the proportions of the monomers reflect the condensation degree of PCMs (Chang et al. 2022). This method has been improved continuously since it was established and widely used in the determination of PCMs in soil, sediment, aerosol and water (Table 1). Briefly, inorganic carbon and polyvalent cations such as Fe3+ and Al3+ were removed by trifluoroacetic acid pretreatment. The treated samples were oxidized into BPCAs by concentrated HNO3 at high temperature and high pressure (170 ºC, 8 h). After that, internal standard (citric acid) was added into samples and further purified by cation exchange resin. Qualitative and quantitative analyses were performed using gas chromatography-mass spectrometry (GC–MS) (Glaser et al. 1998). In later studies, researchers stated that for environmental samples with high organic matter content, a sample mass of less than 5 mg of organic carbon is prerequisite to ensure the complete oxidation of PCMs into BPCAs (Brodowski et al. 2005b). In addition, Schneider et al. (2010) pointed out that phthalic acid is more stable than citric acid as an internal standard for BPCAs analysis, and the derivatization time of 24 h ensures the complete derivatization of BPCAs. How to simplify the complicated experimental steps is the main technical challenge when applying the BPCAs method in the determination of PCMs. In subsequent studies, high performance liquid chromatographic (HPLC) was gradually adopted because it eliminated the steps of sample purification and derivatization, and reduced the loss of PCMs due to complicated processing steps (Sun et al. 2021; Wiedemeier et al. 2013). In addition to quantitative determination of PCMs, the individual BPCA distribution characteristics are often used to describe the properties of PCMs, especially the aromatic condensation degree. B3CAs, B4CAs and B5CA are easily generated by the oxidation of benzene ring on the edges of aromatic cluster, while B6CA is only related to the benzene ring in the center of aromatic cluster. Therefore, the proportion of B6CA reflected the aromatic condensation degree of PCMs. The higher B6CA/BPCA values generally correlated with higher aromaticity or condensation. Some researchers inferred the burning temperature of forming PCMs and traced its sources based on the ratio of B5CA/B6CA and B6CA/B4CAs: the ratio of B5CA/B6CA < 0.8 indicates PCMs originated from domestic fire, 0.8–1.4 referred to grass burning and 1.3–1.9 for forest ground fires (Boot et al. 2015; Wolf et al. 2013). The ratios of B6CA/B4CA < 2, > 2, and > 7 were related to PCMs from grass, urban soils, and fossil fuels, respectively (Chang et al. 2018a; Lehndorff et al. 2014). In order to more accurately trace the source of PCMs, stable isotope carbon (13C) and radioisotope carbon (14C) techniques are also increasingly applied combining the BPCAs method (Wagner et al. 2021). It should be noted that the influence of interfering materials such as kerogen, humic acid, coal and other non-pyrogenic condensed organic matter should be fully excluded, and it remains to be answered what is the individual BPCA composition generated by those interfering materials (containing aromatic structures) and how to distinguish them from BPCAs formed from PCMs.

Based on the above discussions, each detection method has its own limitations. Thermal oxidation method is suitable for determination of PCMs with high condensation degree (Hammes et al. 2007). For environmental samples in complex matrices such as soils and sediments, the removal of inorganic carbon and non-PCMs organic matter is required, and it is necessary to ensure sufficient contact between the sample and oxygen in the pyrolysis process. Otherwise, the organic matter encapsulated by inorganic minerals and the carbonization of organic matter will lead to the overestimation of PCMs. The chemical oxidation has a wider determination range for PCM continuum than the thermal oxidation method, but the operation process is more complicated, involving multi-step extraction and separation process. In addition, it is likely to cause the artificial loss of PCMs during the operation process. The whole PCM continuum can be characterized by optical methods, among which microscopy and MIR-PLSR need to be combined with other techniques to quantify PCMs. The TOR/T method is interfered by dark-colored organic matter such as melanoid and coal (petrogenic carbon), which may result in overestimation of PCMs. BPCAs method cannot only measure the relative content of PCMs, but also determine its condensation degree based on individual BPCAs distribution characteristics. However, the non-PCMs, e.g., kerogen, coal and humic acids can also be oxidized by HNO3 to form BPCAs. Since a part of PCMs structures are transformed into BPCAs, a conversion factor is required to estimate PCMs content according to BPCAs concentration, but the conversion factors vary greatly depending on PCMs properties, from 1.5 to 4.5 (Brodowski et al. 2005b; Glaser et al. 1998; Schneider et al. 2011). Chang et al. (2018b) plotted the conversion factor over the B5CA/BPCAs for biochars and observed a significantly positive correlation. Obviously, conversion factor is not a constant but related to the properties and sources of PCMs. It is not practical to select a proper conversion factor without knowing the sources of the PCMs.

2.2 PCMs sources and the carbon pool

PCMs distribute widely in soils, water, sediments, as well as air and other matrices, and are considered as an important component of the global carbon cycle (Chen et al. 2022a; Fu et al. 2023).There are diverse sources of PCMs, including natural and inadvertent anthropogenic emissions. The natural wildfire (e.g., forest and grassland fire) produces most of the PCMs and more than 400 million hectares of land worldwide experienced fire every year (Abney et al. 2017). Kuhlbusch and Crutzen (1995) estimated that the global PCMs production from vegetation fires was in the range of 50–270 Tg yr−1 and 80% of PCMs were solid residues. To ensure a reliable estimation of PCMs production, we should always bear in mind that the role of grasslands and forests is different. Bowring et al. (2022) reported that ~ 80% of burning occurs in grassland-savannah-dominated regions. Seasonally dry savannas and grasslands accounted for 90% and 82% of the burned area in the Northern and Southern Hemispheres, respectively (Giglio et al. 2013). However, the wet rainforests and peat forests are less susceptible to natural wildfires in the absence of human interference. Thus, PCMs production should be low in this case.

Compared with the components released into atmosphere, PCMs produced from vegetation fires are mainly solid residues, while the PCMs discharged from inadvertent anthropogenic sources (fossil fuel and residential fuel) are mostly emitted into the atmosphere (Bird et al. 2015; Ronkko et al. 2023). Nevertheless, Han et al. (2009) found that biomass burning contributed 42% of atmospheric PCMs, which is higher than those contributed by fossil fuels (38%) and bio-fuel (20%). However, with the development of human society, agriculture management (slash-and-burn, prescribed forestry burning), cooking and heating (cookstoves, woodstoves, furnaces, fireplaces), and industrial emissions such as metallurgy (industrial/commercial boilers), and power generation (mobile and stationary diesel engines, fuel-based electric utility generating units) have unconsciously increased PCMs from bio-fuel and fossil fuel sources.

When PCMs enter into environmental matrices, the resistant structures prevents them from biological and abiotic degradation. The residence time of PCMs ranges from years to millennia corresponding to their form of continuum. The pyrogenic products are more stable than their precursors; therefore, PCMs is regarded as a typical stable carbon sink (Meng et al. 2022; Qi et al. 2017). Bird et al. (2015) calculated PCMs content in 0–100 cm soil, ranged from 54–109 Pg accounted for 3.8–7.7% of global SOC, which was lower than that of 5–15% estimated by Hockaday et al. (2007). The estimated PCMs amount in terrestrial sediments is 29–87 Tg yr−1 from erosion and redeposition under the assumption that sediment contained 1% SOC. In addition, after removing lithogenic graphite, the stock of PCMs decreased from 2000–5000 Pg to 400-1200 Pg for coastal ocean and 400–1000 Pg to 80–240 Pg for deep ocean (Bird et al. 2015). The accuracy of the above estimates of PCMs pools may need to be re-assessed due to methodological differences, natural variability and limited spatial coverage. In addition, the contribution of different types of PCMs to the carbon cycle needs to be evaluated due to their different stability.

3 Effects of PCMs on soil ecological functions

3.1 Soil carbon sequestration and soil fertility

When wildfire occurred in forest ecosystem, roughly 1% of the above-ground biomass was converted to PCMs (116–385 Tg C) each year (Ohlson et al. 2009; Santin et al. 2015). The sequestration of those PCMs in soil not only mitigated climate change but also increased soil organic carbon storage. Researchers found that dissolved organic carbon in soil can be adsorbed by PCMs, which decreases the available carbon substrate for CH4 production, resulting in the abatement of soil-borne CH4 emission (Cai et al. 2018). A meta-analysis of 24 studies using isotopic approaches (Wang et al. 2016) reported the mean residence time (MRT) of 97% PCMs content was 556 years, while 3% PCMs were labile or semi-labile with an MRT of 108 days. The recalcitrance and rapid formation of PCMs allow SOM formation at a much faster rate than through humification (Alexis et al. 2007). Therefore, PCMs produced from biomass burning may be an effective approach to increase carbon sequestration in a relatively short time. In addition, PCMs could reduce the mineralization rate of native SOC (Fig. 2), a process known as negative priming (Tian et al. 2022a). However, negative and positive priming may coexist in PCMs-containing soils. Generally, positive priming occurred in the first several years after PCMs amendment, and negative priming effect followed after this initial period (Ding et al. 2018). Several pathways need to be considered to understand the negative priming effect of PCMs on native SOC (Fig. 2). The toxic components in PCMs would inhibit the growth and activity of microorganisms (Lyu et al. 2016), and the labile carbon components of PCMs also could be utilized as carbon sources instead of native SOC (Chen et al. 2021). More importantly, PCMs can shelter microbial residuals or even unstable carbon components, thus hindering the decomposition of those SOC by microorganisms or microbial extracellular enzymes. PCMs could combine with organic matter and active mineral phases in soils, such as silicoaluminates and iron oxides, to form PCMs-mineral-SOC complex (Zhang et al. 2023), further mediating the formation and stability of soil aggregates, which is also essential in protecting native SOC (Sun et al. 2020). It should be noted that the quantitative identification of PCMs and other organic matter in soils, such as humus, is crucial for exploring the priming effect of PCMs on native SOC. However, due to the lack of effective methods to distinguish different components of organic matter, previous researchers mostly evaluated the changes in SOC before and after the input of PCMs based on the total carbon. Weng et al. (2017) attempted to apply 13CO2 labeled ryegrass-PCMs to investigate the priming effect of PCMs on native SOC. Compared with unplanted PCMs-amended soil, they found that total SOC content increased by 10% with decreased mineralization between 8.2 and 9.5 years in planted PCMs-amended soil. The contribution of direct PCMs input and negative priming effect to native SOC caused by PCMs cannot be distinguished. From the perspective of soil carbon sequestration, on the basis of distinguishing between the input of PCMs and the changes of native SOC, adopting regulatory strategies to enhance the negative priming effect of PCMs on native SOC requires in-depth exploration.

Fig. 2
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The role and mechanisms of PCMs in facilitating soil carbon sequestration, inhibiting soil nitrogen loss and promoting soil ecological functions

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In addition to carbon cycling, PCMs are also associated with the behavior of other elements or nutrients. The mineral elements such as N, P, K, Mg, Ca, Mn, and Zn in biomass would be enriched in PCMs during biomass pyrolysis. These elements will be introduced in soil as PCMs enter soil. The endogenous nutrients of soil would also be adsorbed on PCMs and slowly released, which is beneficial for plants growth. However, the understanding of the geochemical processes of these elements as affected by PCMs should consider various aspects. For example, N cycling is controlled by PCMs in different processes. On one hand, the specific surface area, pore structure and surface functional groups of PCMs are beneficial to the adsorption of NH4 +, NH3 and NO3 − on PCMs, decreasing the volatilization and leaching of N (Cao et al. 2023). At the same time, PCMs may significantly increase the abundance and activity of N-fixing microorganisms (symbiotic azotobacter and autogenous azotobacter) or nitrogenase activity, thereby promoting soil microbial N fixation. More importantly, the changes in soil physicochemical properties after PCMs addition could alter the abundance of nitrifying bacteria and denitrifying bacteria depending on PCMs properties and soil conditions. For example, PCMs adsorbed chemical compounds, such as tannins, which inhibit microbial activity and promote nitrification (Uusitalo et al. 2008). The ammonium in PCMs could reduce the competition between nitrifying bacteria and vegetation, thus increasing nitrification (Knicker 2007). The increased pH after PCMs addition may also benefit the digestion (Yu et al. 2021). On the contrary, PCMs may increase the abundance and activity of heterotrophic bacteria by supplying carbon resources (Ahirwar et al. 2018), and these heterotrophic bacteria may compete with nitrifying bacteria for NH4 + and inhibit the nitrification process (Hanan et al. 2016). PCMs could also release nitrification inhibitors or toxic substances, e.g., α-pinene and PAHs, which may inhibit nitrification. Simultaneously, the improved soil aeration after PCMs input could inhibit denitrification. Although the increased soil pH promoted denitrification, it mainly enhanced the nitrous oxide reductase activity of denitrifying bacteria, promoting the reduction of N2O to N2 (Weldon et al. 2019), thereby reducing greenhouse gas emissions. The adsorption of NH4 + and NO3 − on PCMs limited the substrate for denitrification. Obviously, the enhanced nitrification and inhibited denitrification accelerated soil N flow rate and decreased N loss, respectively. The above proposed processes suggest that PCMs are involved in both nitrification and denitrification, and the apparent effect is the balance between these two processes depending on PCM properties, soil texture, and the interaction duration. For instance, Nessa et al. (2021) reported that the pyrolysis temperature of PCMs significantly increased cumulative nitrification, peaked between 600 and 700 °C, which was related to the adsorption of NH4 +-N and NO3 −N, and the 65% soil water holding capacity (WHC) showed a higher cumulative nitrification compared with those in 50% WHC. When the soil moisture content reached 83%, the PCMs cannot not reduce the denitrification rate by improving the soil aeration (Yanai et al. 2007).

3.2 Soil structure and soil ecological functions

The input of PCMs into soils can enhance soil pH, reduce soil density, increase soil porosity, water retention and cation exchange capacity (CEC) (Fig. 2) (Santos et al. 2022; Singh et al. 2022). Besides the pristine alkalic pH, the negatively charged surface oxygen-containing functional groups of PCMs, such as phenols, carboxyl, and hydroxyl groups, could bind H+ from soil solution, and thus further increase soil pH (Gul et al. 2015). In addition, the rich oxygen-containing functional groups and alkaline elements, e.g., K, Ca, and Mg, will enhance soil CEC (Chen et al. 2019c), which will even increase with PCMs oxidation due to the increased negative charges (Mao et al. 2012). Studies also suggested that the water holding capacity of coarse soils could be enhanced after PCMs input, which may be attributed to the high porosity, large specific surface area, and hydrophilic domains of PCMs (Razzaghi et al. 2020). However, this effect is not significant for clay and organic-rich soils because these two types of soils have relatively high specific surface areas and microporosity, as well as strong water retention ability. More importantly, the addition of PCMs into soils would promote the formation of soil aggregates and enhance their stability (Lu et al. 2014), and the formation mechanisms were described in other reviews (Li et al. 2018).

The improvement of soil structure and physicochemical properties after PCMs input would inevitably lead to the enhanced soil ecological functions. Researches have suggested that PCMs promoted the restoration of surface soil ecosystems, especially those with low organic matter content, through fixing harmful substances and improving the microorganism living environment (Pingree and DeLuca 2017). For instance, the reduced soil bulk density after PCMs input increased soil porosity, promoting the mobility of water and air in soil (Saito 1990; Usowicz et al. 2020), thereby enhancing the colonization of arbuscular mycorrhiza and the growth of their host plants (Ezawa et al. 2002). Currently, more research is focused on exploring the effects of PCMs on the growth and development of plants and microorganisms by adding PCMs into farmland soil. They found that PCMs could promote plant growth by increasing the potential difference between root membrane and soil, improving the efficiency of nutrient transport (Zhou et al. 2016). In return, the growing root system increased uptake of nutrients, especially nitrogen and phosphorus (Schmidt et al. 2021). Other researchers reported that PCMs provide a shelter and nutrient sources for microorganisms (Zheng et al. 2022), which is conducive to enhancing microbial activity and abundance. Obviously, most of the current studies on the impacts of PCMs on ecological functions are based on laboratory simulation and focus on the agricultural field, lacking an assessment of the overall ecosystem function after actual wildfires occur in the field.

4 Impacts of PCMs on pollutant fates

4.1 The removal of organic pollutants mediated by PCMs

PCMs may strongly interact with various pollutants, which will play a significant role in pollutant behavior or their control. The apparent interactions include sorption and redox reactions. The adsorption mechanisms (Fig. 3a) of organic pollutants on PCMs include electrostatic interaction, hydrogen bonding, π-π interaction, pore filling, and hydrophobic interaction (Liao et al. 2023; Qiu et al. 2022). Partitioning is also involved in the adsorption process on non-carbonized phase of PCMs. Higher specific surface areas, microporous structures, and aromatic domains of PCMs may facilitate their adsorption to non-polar compounds. Carboxylic acid, nitro, and ketone groups on PCMs surface can act as electron acceptors to form π-π electron-donor–acceptor interactions with aromatic molecules, thereby facilitating their interactions. Studies have reported that π-aromatic PCMs systems acted as electron acceptors when prepared below 500 °C, but could act as a donor when prepared above 500 °C (Zheng et al. 2013). On the contrary, electrostatic interaction and hydrogen bonding may contribute to the adsorption of polar compounds, e.g., antibiotics and pesticides. Luo et al. (2021) studied the mechanism of antibiotics adsorption on PCMs as affected by pH value in system. These polar compounds exhibit different forms including cation, anion or neutral because of their various pKa (Chang et al. 2016). When PCMs surface is negatively charged at pH higher than their point zero charge (pHpzc), the sorption of cationic compounds was greatly promoted through electrostatic attraction (Yang et al. 2023). Obviously, the adsorption mechanisms vary significantly with the change of pollutants properties. Unfortunately, the relationships between removal mechanisms and properties of pollutants and PCMs have not been established, which cannot meet the needs of predicting pollutant behavior and designing PCMs.

Fig. 3
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Adsorption and degradation (a) of organic pollutants by PCMs, and the adsorption and redox reactions of heavy metals on PCMs (b)

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It was observed that not only the adsorption, but also the degradation plays a very important role in removing organic pollutants. PCMs can transfer electrons to oxygen molecules and induce the generation of ROS such as superoxide anion free radical (•O2−) and hydroxyl radical (•OH) in liquid phase, which are highly effective in reacting with organic pollutants (Fang et al. 2014). Meanwhile, organic pollutants adsorbed on the PCMs can be degraded directly (Yang et al. 2016). In addition, PCMs could be used as a catalyst in photocatalytic and advanced oxidation (AOPs) reactions. In photocatalytic degradation, the electrons in the PCMs carbon defects (valence band) can be excited by visible light, and then move to the oxygen-containing functional groups (conduction band), radical •O2− and h+ are simultaneously formed in the valence band and efficient degradation occurs (Xiao et al. 2020). Photo-generated electrons can also react with the surface-adsorbed oxygen to generate •O2− and •OH to degrade organic pollutants (Li et al. 2019; Zhang et al. 2019). During the AOPs catalytic reactions, both free radical and non-radical pathway were involved. PCMs could act as a catalyst to activate hydrogen peroxide/persulfate to generate •OH, sulfate radicals (SO4•−) and O2•−(Yu et al. 2022a). Oxygen-containing functional groups, defect sites, EPFRs and transition metal elements on PCMs are all possible effective compositions (Pan et al. 2018). Different from free radical pathway, non-radical pathway is a surface reaction that occurs on PCMs surface. The non-radical pathway includes surface electron-transfer, surface activated complex, and generation of singlet oxygen (Huang et al. 2021b). Thus, in the presence of PCMs, EPFRs, ROS and electron-transfer mediated redox reactions may be involved in organic pollutant degradation (Fig. 3a). However, some key issues need to be sorted out: (1) It is still unconcluded what are the contributions of different degradation pathways to the organic pollutant degradation; (2) It is essential to distinguish the degradation of organic pollutants through ROS generated by PCMs and through electron transfer with PCMs structures, which is fundamental for a targeted PCMs property manipulation.

4.2 The immobilization of heavy metals mediated by PCMs

PCMs can also effectively immobilize heavy metals. The physical sorption, chemical sorption (ion exchange, complexation), and other interaction mechanisms (redox reactions and precipitation) are involved in immobilization process (Fig. 3b). PCMs surface is negatively charged because of the abundant oxygen-containing functional groups, such as carboxyl, carbonyl and hydroxyl, which can facilitate the electrostatic attractions with positively charged metals (Mukherjee et al. 2011). In addition, surface oxygen-containing functional groups contribute to CEC, and thus, the high CEC is favorable for selective exchange between the target cations and the ionizable protons/cations on the surface of PCMs. The negatively charged sites occupied by Na+, K+, Ca2+, Mg2+ are prone to be the main sites for cation exchange on PCMs (Lian and Xing 2017). Furthermore, the oxygen-containing functional groups were demonstrated to bind with heavy metals through metal–ligand interactions and complexes formed are stable and unlikely to exchange with other cations in solution (Mahesh et al. 2022). For heavy metals with variable valence, such as Cr and As, the redox reaction usually was involved in immobilization process. Cr (III) and As (V) generally have a lower toxicity and mobility than Cr (VI) and As (III), respectively (Tian et al. 2022b; Zheng et al. 2021). Therefore, the reduction of Cr (VI) to Cr (III) and oxidation of As (III) to As (V) are highlighted in their immobilization. Electron-donating groups on PCMs, such as carboxyl groups, phenolic groups and amino groups are the dominant moieties in Cr (VI) reduction (Zhang et al. 2018, 2020), while electron-accepting groups (e.g., quinone groups, phenolic—OH) play a dominant role in oxidation of As (III) (Tian et al. 2022b; Zhong et al. 2019). Considering that the redox function of PCMs depends on the environmental condition and the surface functionality of the PCMs, it is of great importance to selectively apply or specifically design biochars to detoxicate these heavy metals. For example, the transition of oxygen-containing functional groups at low temperatures to conjugated aromatic structure at high temperatures increased the reduction of Cr (VI) (Xu et al. 2020).

For PCMs prepared at high temperatures, the alkaline character can lead to the precipitation of metal ions (Cao and Harris 2010). Some mineral components, e.g., calcite, calcium anhydrite and hydroxyapatite released from PCMs can combine with heavy metal ions to form precipitates. A high removal of heavy metals was found on PCMs with high phosphate and carbonate contents, which was dominated by precipitation (Cao et al. 2009). Functional groups facilitate dominantly immobilization of heavy metals on PCMs. Introducing tailormade functional groups on PCMs for specific heavy metals is a potential strategy in heavy metals control.

5 PCMs-associated contaminants

5.1 Formation mechanisms and emission of persistent free radicals in PCMs

Environmentally persistent free radicals, as emerging pollutants, may cause DNA damage or oxidative stress through direct contact or entering the organism (Mahne et al. 2012; Squadrito et al. 2001). The reactivity of EPFRs is affected by its composition and the location of the sites. For example, active EPFR species are mainly on the surface of particles (Gehling et al. 2014). Unlike small molecular free radicals such as hydroxyl radicals, which have lifetimes of only a few picoseconds, EPFRs can stably exist on/in combustion-generated PMs for several days or even months (Chen et al. 2019b; Vejerano et al. 2018). Such a long half-life of EPFRs enabled their potential risks associated with PMs emission, transportation, and photo/oxidation aging, especially for fine and ultrafine PMs that can be suspended in the atmosphere much longer than coarse particles. In the real world, the persistency of EPFRs in PMs is also determined by environmental factors. The surface-bound EPFRs decay rapidly as they are easily oxidized in air, while EPFRs embedded inside the particles remain intact because oxygen is difficult to diffuse into the particles and react with them (Vejerano et al. 2011). Zhao et al. (2021) reported that the half-life of EPFRs in biomass-burning PM2.5 was about 100 days under UV light. Chen et al. (2019b) measured a half-life of longer than 1,000 days under dark conditions. The half-life of metal-mediated EPFRs was longer on metals oxides with higher oxidation potentials (Vejerano et al. 2011). The half-life of EPFRs is generally calculated by pseudo first-order kinetics model (Kiruri et al. 2014; Yang et al. 2017b), but there is no model that can directly predict the PM-EPFRs decay by comprehensively taking the above-mentioned factors into account.

EPFRs in PM from biomass combustion were primarily carbon-centered adjacent to oxygen atoms (Tian et al. 2009; Zhao et al. 2022), such as phenoxyl-type radicals which have longer lifetimes than semiquinone-type radicals (Kiruri et al. 2014). To date, the formation mechanism of EPFRs in combustion-generated PM is mainly based on the analysis of by-products in different combustion zones (Fig. 4) (Cormier et al. 2006; Vejerano et al. 2018). In the fuel (pre-flame) zone, EPFRs detected in original biomass mainly depend on the contents of lignin (He et al. 2014; Tao et al. 2020a), but most of them would be destroyed during combustion. In the flame zone, EPFRs are formed by the low-barrier hydrogen-abstraction/ejection reactions with hydrocarbon species and subsequently involved in the nucleation and growth of soot (Commodo et al. 2021; Johansson et al. 2018); and EPFRs generated by chemical bond breaking may be stabilized inside unburned-char (Liao et al. 2014). In the post-flame and cooling zone, the organic species in the gas phase, such as phenols and PAHs could be physically or chemically adsorbed to metal nuclei, and EPFRs (metal-mediated) are readily generated during the electron transfer between organic precursors and metal oxides (Dellinger et al. 2007; Pan et al. 2019). At the same time, EPFRs are formed and embedded in the condensable volatiles generated during the pyrolysis of biomass (Tao et al. 2020b), and these condensable components may be coated on the surface of existing particles/fly ash. It is noted that this conceptual process is mainly based on knowledge from fine particle studies, while its formation in coarse particles (eg. PM2.5-PM10) with different particle size, surface properties and chemical compositions is still not well explored. It is proposed that the EPFRs in unburned-char are mainly distributed in the matrix of coarse particles, while the metal-mediated EPFRs would be mainly formed on the surface of fine particles or be embedded in the interior of the aggregated fine particles. Some previous studies have found that the metal-mediated EPFR generation could be facilitated in the presence of substituted benzenes with strong chemisorption and be affected by the type, crystal form or size of metal oxides (Vejerano et al. 2011; Yang et al. 2017a), but these studies were carried out in single-chemical systems, which could not be extended to explain the generation of EPFRs in biomass combustion systems with various by-products.

Fig. 4
figure 4

Conceptual diagram of EPFRs formation mechanism in combustion-generated fine PM depending on combustion zones. EPFRs are readily generated in the post-flame and cooling zone. Modified from Cormier et al. (2006)

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Some recent studies quantified emissions of EPFRs from incomplete biomass burning (Zhao et al. 2022, 2021), both in laboratories simulated and field-measured combustions. Zhao et al. (2022) found that the EFs of EPFR from crop residues burning were significantly higher than those from fuelwood based on field measurement, but the reason has not been well explained given the multiple influence of stove types, fire management measures, and modified combustion efficiency (MCE). It is inferred that the combustion rate of crop residues is faster than that of fuelwood, which results in faster oxygen-deficiency in stove and thus lower MCE, leading to the increase of EPFRs emission. Further laboratory simulated combustion need to be set up under targeted experimental conditions to clarify the variation of EPFRs emissions. Moreover, available studies, though limited, on the EFs of EPFRs from incomplete biomass burning are mainly on residential sector. Few studies were carried out on biomass emissions in open burning sources such as in-field crop residue burning and forest fires. Open fire burns on a large scale, and its energy release rate and turbulence increase compared to indoor combustion, resulting in the increased emission of coarse particles containing ash and soil material (McMeeking et al. 2009), and the obvious difference in EPFRs formation.

5.2 Formation of polycyclic aromatic hydrocarbons in PCMs

Polycyclic aromatic hydrocarbons are inevitably produced during biomass pyrolysis and thus also accompanied with PCMs formation. The decomposition, unimolecular cyclization, aromatization of biomass and subsequent recombination, re-condensation reactions of pyrolysis vapors are the key processes for forming PAHs (Buss and Masek 2014; Godlewska et al. 2021). However, the specific components vary significantly depending on the pyrolysis temperature. Generally, low pyrolysis temperatures (< 500 ºC) produced low molecular weight (2–3 rings) PAHs via carbonization and aromatization, while the high molecular weight PAHs appeared through recombination reactions by free radical pathway under high pyrolysis temperatures > 500 ºC (Wang et al. 2017). Due to their carcinogenicity, teratogenicity and mutagenicity, PAHs are classified as organic pollutants, with potential harm to the ecological safety and human health.

PAHs formed during biomass pyrolysis are generally adsorbed on PCMs and move with the migration of PCMs (Fig. 5). The content of PAHs in PCMs reached up to 172 mg/kg (Godlewska et al. 2021). The wildfire-derived PAHs sometimes contributed most of PAHs in a specific environment (Berthiaume et al. 2021). Compared with unburned forest soils, PAHs contents increased by 205% in soils associated with wildfire (Yang et al. 2022), which originated from both biomass and SOM pyrolysis (Fig. 5).

Fig. 5
figure 5

The generation and transport of PAHs after wildfire. PAHs are generated in both biomass and SOM pyrolysis. Gas-phase and particle-phase PAHs will deposit into soils through dry and wet depositions. The surface adsorbed and sequestered PAHs may be released into the environment, posing a threat to ecological systems

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Due to their strong hydrophobicity, PAHs are mostly PCMs-associated, and thus their toxicity to organisms is not significant (Yang et al. 2022). It should also be noted that PAHs are generated during PCMs formation, and may be adsorbed on the surface or embedded/sequestered in PCMs matrix. The surface adsorbed PAHs are involved in the physiochemical equilibrium with the aqueous phase, but the embedded PAHs may not. It is therefore essential to quantify PAHs in different forms of interactions with PCMs, especially the dissociable PAHs. Investigators have established various methods to quantitatively separate PAHs particles with different interaction strengths, such as sequential extraction (Li et al. 2023a), thermal desorption analysis (Jia et al. 2023), and biological mimic extraction (Maletic et al. 2019). These methods are also applicable in assessing PAHs mobility or risks in PCMs. The presence of dissolved organic matter (especially those released from PCMs) in the aqueous phase may greatly enhance the solubility of PAHs. But how the distribution of these different forms of PAHs affected by DOM is not investigated.

6 Research perspectives

6.1 Development of standard analysis system for PCMs identification and quantification

Although various PCMs determination methods have been established, each method focused on one or a few PCMs fractions (Kerre et al. 2016). It is necessary to fully investigate the properties and sources of PCMs for selecting a suitable PCMs determination method. However, the heterogeneity and diverse sources of PCMs bring a big challenge for their determination. How to distinguish non-PCMs organic matter from PCMs in complex environmental matrices is another urgent problem to be solved. Until now, most researchers have focused on the precise quantification of PCMs (Barton and Wagner 2022; Hardy et al. 2022), and some key points are overlooked. Firstly, the analysis of PCMs sources is confusing, and there is an urgent need to investigate and organize the source data of PCMs worldwide so as to provide a basis for speculating the formation conditions and judging the properties of PCMs. Secondly, the environmental impacts from PCMs are generally affected by their physicochemical properties rather than contents. For example, PCMs with a high condensation degree and aromaticity have a greater adsorption ability for hydrophobic organic contaminants (Chang et al. 2022). The development and refinement of methods to characterize PCMs properties is particularly important for predicting their environmental implications. And, though the structure of PCMs is stable, their physicochemical properties will be changed after long-term environmental exposure (Hyvaluoma et al. 2023). Clarifying how to dynamically monitor the properties of PCMs during their turnover at the molecular level is helpful to understand their interaction with other substances and the role in the global carbon cycle. In addition, various determination methods and operation steps resulted in different measurement results (Murano et al. 2021). Carrying out comparative research on a variety of methods in multiple laboratories based on standard reference materials and unifying their PCMs determination values, is conductive to providing a valuable reference for assessment of PCMs in the global scale.

6.2 The application of PCMs in pollutant controls in field

Although PCMs have shown promising potential in pollutant removal, the reported performance was obtained based on simulation experiments in laboratory under controlled conditions. The effectiveness of PCMs in the real environment need to be evaluated systematically. To make the experiment results more suitable for applying in real conditions, some researchers evaluated the impact of environmental factors such as pH, temperature, coexisted chemicals (anion, cation, and NOM) on the removal of pollutants by PCMs (Huang et al. 2021a; Yu et al. 2022a, b). However, it seems that these factors may not be mathematically overlaid to estimate PCMs–pollutant interactions in the real environment. For example, Hernandez-Abreu et al. (2020) explored bisphenol A adsorption on PCMs in real hospital wastewater. Compared with the simulated wastewater, a sharp reduction of 35–48% was observed for the adsorption because of the influence of pH, salt content and NOM. Similarly, when PCMs were applied in soils, the availability of PCMs catalytic potential for PAHs degradation was only 7.6–13.4% (Mazarji et al. 2022). Obviously, the performance of PCMs in removing pollutants from the real environment is not as effective as that in the simulated studies. Besides, the investigators have noticed that some coupling processes, such as microorganism and photochemistry, may enhance the removal of pollutants by PCMs. In PCMs–microorganism system, PCMs can provide nutrients and colonization sites for microorganisms and promote their growth; on the other hand, PCMs as electron shuttles, mediate the electron transfer between microorganisms and pollutants, and promote the biotransformation of pollutants adsorbed on PCMs (Mukherjee et al. 2022; Xiang et al. 2022). In addition, PCMs–pollutant interactions may be coupled with other chemical processes, such as photocatalytic reactions and redox reactions (Fu et al. 2016; Xiao et al. 2020). Considering the long-term biotic and abiotic aging of PCMs, the physicochemical properties as well as the chemical reactivity of PCMs may change substantially in the environment. It is may not be possible to include all the possible parameters in a model to describe PCMs–pollutant interactions. A practical method should be developed to specifically compare and extract the key parameters related to PCMs-pollutant interactions in a specific environmental system, which will be different in different systems.

6.3 Persistent free radicals, as well as other emerging pollutants, from incomplete biomass burning

EPFRs are more like a structural character than a chemical composition of PCMs. Although no consensual conclusion has been reached, some investigators suggested that EPFRs may partly reflect the reactivity of PCMs. Thus, estimation of EPFRs generation will be helpful to evaluate the environmental implications of EPCMs. EFs dataset of EPFR in biomass burning supports macroscopic estimates of EPFRs in aerosol. For better controlling FPFRs emissions, more systematic work, combining with in situ on-line monitoring technologies, is needed to explore the EPFRs generation mechanisms in the biomass burning, as well as other combustion sources. In particular, identifying the competitive or coupling role of various by-products in the formation of EPFRs in the combustion systems is a challenging, but crucial task. Some studies used the equivalent number of cigarettes to assess the risk of EPFRs in PM2.5 due to the similar adverse health effects of EPFRs (Chen et al. 2019a; Gehling and Dellinger 2013); however, their exposure pathways, the co-existing chemicals, as well as the species of EPFRs are different between cigarette smoke and ambient PM2.5. Additionally, it is still unclear how the measured EPR signals correlate to EPFRs toxicity. The major question is how to quantify the activity of EPFRs. Some investigators proposed a solvent extraction to describe the activity of EPFRs (Guo and Richmond-Bryant 2021). However, this method is experiential and the samples should be measured individually, which would not help the estimation. Thus, it remains unclear whether there are direct approaches to accurately identifying and evaluating the real risk of EPFRs in PM. Biological toxicity exposure test combined with epidemiological studies may provide a perspective for PM-EPFRs risk assessment.