1 Introduction

Fossil fuel combustion, forest fires, and biochar application have led to large amounts of black carbon (BC) entering the natural environment (Coppola et al., 2019). BC is heterogeneous in nature and ranges from slightly charred fractions to highly condensed aromatic components formed through incomplete combustion at high temperatures. It is estimated that approximately 60 Pg of BC has been produced by landscape fires since 1750 (Jones et al., 2019). On a global scale, biomass burning and fossil fuel combustion are expected to produce 114–383 Tg yr−1 and 2–29 Tg yr−1 of BC (Coppola et al., 2019), respectively. Owing to the large quantities and recalcitrant nature of this carbon-rich material (Masiello and Druffel, 1998; Preston and Schmidt, 2006; Reisser et al., 2016), BC represents a significant but previously ignored carbon pool in the global carbon budget (Jones et al., 2019; Woolf et al., 2010).

In the current review, dissolved black carbon (DBC) is defined as the water-soluble fraction (generally refers to fraction of < 0.45 μm) extracted from BC. DBC is one of the most labile and mobile fractions in the BC continuum (Wang et al., 2013), and it can easily be released via soil infiltration and surface runoff. Globally, DBC represents approximately 10% of the global fluxes of dissolved organic matter (DOM) from rivers to oceans (Jaffe et al., 2013), which is a significant pathway for BC emission to the sea. Moreover, the fact that more than 2% of oceanic DOM exerts a heat-induced molecular signature also suggests a considerable DBC fluxes from land to oceans (Dittmar and Paeng, 2009; Jaffe et al., 2013; Mannino and Rodger Harvey, 2004). As such, the biogeochemical behaviors of DBC may be particularly significant in regulating greenhouse gas emissions and stabilizing organic matter.

The molecular structures of DBC are the basis for understanding their biogeochemical behaviors. However, due to the varied parent solids (e.g., biomass types) and formation conditions (e.g., production temperature) and the heterogeneous nature of BC (Alan Roebuck et al., 2017), the molecular structures of different DBC vary greatly (Bostick et al., 2020). The diversity, complexity and heterogeneity of DBC make the prediction of their environmental behaviors and effects complicated. Once DBC enters the environment, it will inevitably undergo various environmental processes. Photodegradation is one of the most important pathways for DBC transformation and is also a relatively well-studied process (Fu et al., 2016). DBC contains many aromatic structures, and it may be one of the more photoactive fractions in the DOM pool (Fu et al., 2016). Biodegradation is another important pathway for DBC transformation. Although recalcitrant in nature, DBC can be utilized by microbes via co-metabolism of other labile carbon source (Qi et al., 2020). In contrast, DBC, as a refractory carbon source, may also affect the constitution and succession of microbial communities (Hockaday et al., 2006; Zhou et al., 2021). Similar to well-studied DOM, DBC can affect the fate and transport of pollutants via sorption processes, redox reactions or radical-driven processes (Fu et al., 2016; Smith et al., 2013). However, the understanding of the aforementioned environmental behaviors of DBC is still far from robust, which is partly attributed to the lack of a well-organized analytical system.

In this work, qualitative and quantitative analytical methods for DBC were discussed in detail. In addition, the research on the molecular structures and environmental behaviors (e.g., photodegradation, interaction with microorganisms and pollutants) of different DBC was critically reviewed. The current knowledge gaps and future perspectives were put forward. This work can be useful for future studies on the environmental behaviors and effects of DBC.

1.1 Analytical methods for DBC

Until now, information regarding the environmental behaviors of DBC in aquatic environments has been far from robust, which is in part due to the lack of a well-recognized and consensus method. Recent studies have gradually uncovered the photodegradation of DBC under specified laboratory conditions. However, DBC in real aquatic environments is far more complicated as it is well-dispersed in the natural DOM (NDOM) pool. The lack of consensus regarding standard separation and quantification methods results in considerable uncertainties on the sources and fates of DBC in natural aquatic systems. The combined utilization of advanced technologies can help improve our understanding of the environmental transformations of DBC.

1.1.1 Chemothermal oxidation (CTO) method

At present, CTO is popular for isolating BC from environmental samples (Gustafsson et al., 2001; Zencak et al., 2007), and it can similarly determine the abundance of DBC in the DOM pool (Qi et al., 2020). The quantitative principle of the CTO method is based on the determination of the highly condensed aromatic components in DBC. First, the environmental DOM sample is extracted using ultrafiltration/lyophilization (Mannino and Rodger Harvey, 2004) and solid-phase extraction (Wang et al., 2016; Xu et al., 2016). Carbonate is removed by chemical pretreatment. The recovered DOM sample is thermally oxidized at 375 °C for 2 h to remove the non-DBC DOM. Then, the residue is oxidized again at 850 °C for 2 h, and the DBC abundance is determined by the resultant CO2 generated during this process. Several reports have shown that DBC concentrations in river and oceanic waters determined by the CTO method resemble those determined by the benzene polycarboxylic acid method (Coppola et al., 2018; Dittmar and Koch, 2006; Dittmar et al., 2012b; Wagner et al., 2019; Wang et al., 2016), indicating that CTO is a potentially efficient quantitative method. In addition, Qi et al. (2020) excluded the possibility of additional DBC generation from DOM during thermal oxidation at 375 °C, which further confirmed the accuracy of measured CTO results. However, questions remain for DBC of relatively low production temperature (e.g., 450 °C), which contained abundant non-condensed components but very little residues for CTO determination (Hammes et al., 2007; Nguyen et al., 2004). Another deficiency concerns the positive bias due to the samples charring during the analysis (Hammes et al., 2007). Thus, influencing factors and the effects of environmental matrix should be considered when applying this method (Wagner et al., 2018).

1.1.2 Benzene polycarboxylic acid (BPCA) method

The BPCA method is another quantification analytical method that was originally applied to determine the abundance of BC in soil environments. The quantitative principle of the BPCA method is based on the detection of polysubstituted benzene carboxylic acids containing 3‒6 carboxylic acid groups (B3CA, B4CA, B5CA, and B6CA, respectively), which are converted by thermo-chemically oxidizing condensed aromatic compounds (Dittmar, 2008; Schneider et al., 2010; Ziolkowski et al., 2011). The distribution patterns of polysubstituted benzene carboxylic acids appropriately indicate the condensed aromaticity of BC and DBC. The BPCA aromatic condensation index, which represents the average number of carboxyl groups in DOM, is usually used as an indicator of condensed aromatic cluster sizes (Bostick et al., 2018; Ziolkowski and Druffel, 2010).

Glaser et al. (1998) first developed the BPCA method and employed it to measure artificially produced BC. Hockaday et al. (2007) used the BPCA method to track the export of BC in DOM from soils. Dittmar (2008) then optimized the oxidation temperature and duration of the BPCA method and successfully applied it to determine both the concentration and the structural properties of DBC. The BPCA method gradually became the primary DBC quantification technology, and was widely applied to track DBC in river (Stubbins et al., 2015), stream (Ding et al., 2013; Jaffé et al., 2012), and marine waters (Coppola and Druffel, 2016), as well as to investigate the photochemistry (Stubbins et al., 2012b) and biochemistry (Bostick et al., 2021) of DBC. For example, Stubbins et al. (2012b) quantified oceanic DBC using benzene polycarboxylic acid markers and found that DBC was more photolabile than colored DOM. Bostick et al. (2021) further found that photodegradation and biodegradation altered the BPCA composition of BC leachates. Specifically, photodegradation process profoundly removed high-substituted B5CA and B6CA, whereas biodegradation process promoted the loss of low-substituted B3CA and B4CA. Liu et al. (2022) compared the photodegradation patterns of DBC and NDOM released from composted Rice Straw. They observed a much higher yields of B3CA, B4CA, B5CA, and B6CA from DBC (2.81 mg C/L) than NDOM (0.43 mg C/L). The photodegradation process decreased the abundance of all polysubstituted benzene carboxylic acids, with larger losses of high-substituted B5CA (50.28%) and B6CA (51.51%) relative to the low-substituted B3CA (34.04%) and B4CA (42.45%). This indicated a more significant removal of high-condensed aromatic carbon in DBC, which was consistent with FT-ICR-MS results. Such results demonstrated that biogeochemical processing may play a significant role in reworking the BPCA composition of DBC. Moreover, distinct differences were noted for the BPCA compositions of BC solid and corresponding DBC. Therefore, although the BPCA composition can be utilized to distinguish different DOM pools, the mismatch of BPCA composition between DBC and parent solid, as well as the dynamically changing BPCA distribution of DBC after irradiation or microbial reworking, suggest the uncertainty of BPCA method in assessing the source of DBC.

1.1.3 Compound-specific radiocarbon analysis (CSRA)

Compared to BPCA method, CSRA is an efficient technique for dating and tracing sources of DBC. This analytical method mainly includes two steps: 1) oxidizing DBC into BPCA, as mentioned above, and 2) determining the 14C abundance of BPCA using accelerator mass spectrometry (Wei et al., 2017). Based on the distinctive 14C ages of the DBC produced from biomass burning (modern) and fossil fuel combustion (ancient) (Coppola and Druffel, 2016; Coppola et al., 2018; Masiello and Druffel, 1998; Ziolkowski and Druffel, 2010), the CSRA method can be applied to verify the relative contribution of these two sources. In addition, comparing the distinctive 14C ages between riverine DBC (modern) (Coppola et al., 2018; Wang et al., 2016) and oceanic DBC (ancient, > 20,000 years) (Coppola and Druffel, 2016; Ziolkowski and Druffel, 2010) helps to better elucidate the transport and export of DBC from land to sea, as changes in 14C ages suggest significant contributions from microbial and photochemical degradation processes (Qi et al., 2020). However, limitations remain in BPCA-specific CSRA method, as it is hard to distinguish whether the 14C-depleted DBC result from the contemporary input of pre-aged DBC sources (e.g., fossil fuel combustion products) or from the in-situ aging of DBC (Coppola and Druffel, 2016).

1.1.4 Fourier transform ion cyclotron resonance mass spectrometry (FT-ICR-MS)

FT-ICR-MS has unsurpassed mass-resolution abilities and can provide molecular-level identification of DBC. The primary competitive advantage of FT-ICR-MS is that it does not require the prior separation of individual molecules before identification (Sleighter and Hatcher, 2007). With ultrahigh mass-resolution spectrometry data, FT-ICR-MS techniques can identify more than 1000 molecules at a time by assigning CaHbNcOd molecular formulas to each spectral peak. In addition, it can also distinguish between molecules of nearly identical nominal mass (Koch et al., 2007; Kujawinski and Behn, 2006).

To visualize the molecular compositions of DBC, van Krevelen diagrams for CHO and CHON molecules are often applied to describe the types of DBC compounds. The H/C atomic ratio is related to the aromaticity of DBC molecules. To evaluate the bulk composition of DBC, different characteristic parameters have been calculated. The modified aromaticity index (AImod) can be utilized to determine the bulk aromaticity of DBC, as well as the relative abundance of aliphatic (AImod < 0.5), aromatic (0.5 < AImod < 0.67), and condensed aromatic (AImod > 0.67) molecules (Hao et al., 2018; Osterholz et al., 2016a). In natural environment, molecular formulas of AImod > 0.67 are generally defined as part of the DBC pool (Koch and Dittmar, 2006). The DBE index is applied to analyze the number of double bonds and rings in a molecule, which manifests the degree of unsaturation and aromaticity of DBC molecules (Stenson et al., 2003). The nominal oxygen state of carbon (NOSC) is used to evaluate the degree of oxidation and molecular polarity or hydrophobicity of different components (Riedel et al., 2012).

FT-ICR-MS techniques have also been applied to investigate the shifts in DBC molecular structural compositions under different biogeochemical processes like photodegradation and biodegradation. Stubbins et al. (2010) used FT-ICR-MS to identify DBC in Congo River water and found that the DBC undergoes faster photodegradation than the other structural components of DOM. Ward et al. (2014) found that the condensed aromatics in DBC were preferentially degraded via photo-oxidation. Latch et al. (2003) determined the conversion of DBC molecules into smaller molecular weights components by the photo-oxidation of the higher molecular weight fractions. As such, FT-ICR-MS showed great advantages in uncovering the molecular change mechanism of DBC during environmental processes. Compared with photodegradation, few studies have focused on the biogeochemical processes of DBC. Zhou et al. (2021) showed that the successions of microbial communities were closely associated with the aromaticity of DOM on the basis of FT-ICR-MS analysis. Similar to NDOM, the involvement of DBC in the DOM pool can potentially alter the community compositions and functions of microorganisms (Hao et al., 2018). In return, microbial activities can change the molecular structure of DBC (Qi et al., 2020), a phenomenon that lacks supporting evidence from FT-ICR-MS analysis (Chen et al., 2022; Goranov et al., 2022).

As different molecules vary in their ionization efficiencies, the FT-ICR-MS technique is only qualitative or semiquantitative (Kujawinski et al., 2004). To avoid the substantial suppression of analyte signals, solid-phase extraction (SPE) is generally applied to concentrate and desalt DBC samples (Schmidt et al., 2003). Although SPE can guarantee a relatively high recovery rate for DBC (Li et al., 2017), the selective loss of certain compounds, such as biopolymers, CHNO, and CHOP molecules, is inevitable (Chen et al., 2016; Zhang et al., 2021b). Besides, FT-ICR-MS shows poor reproducibility and the commonly used electrospray ionization (ESI) mode is susceptible to interference of solvents and impurities (Enke, 1997). Several studies suggested that ESI mode may not be able to measure large aromatics and condensed aromatics in DBC of high pyrolysis temperature (Wozniak et al., 2020). Also, the ions transportation efficiency depends on the optimization of ion optics, the vacuum state of ion cyclotron resonance cell, and related parameters (Marshall et al., 1998). Additionally, considerable uncertainties exist in selecting one proper formula from the multiple formula assignments. To alleviate this problem, the formula screening can be conducted according to following rules (Liu et al., 2015; Shi et al., 2021; Vetere and Schrader, 2017). In most cases, CHO components were more reliable than CHON components (Herzsprung et al., 2014, 2015). In addition, ESI– mode favors detection of acidic functional groups, while ESI + mode facilitates the measurement of CHON, CHONS, amines and amino sugars (Lin et al., 2012; Ohno et al., 2016; Sleighter and Hatcher, 2007). Moreover, FT-ICR-MS technology is one of the most expensive and time-consuming methods, which partly limits its widespread use (Dittmar and Koch, 2006). It should be noted that the combined application of multiple techniques can help to alleviate the deficiency of a single method. BPCA and FT-ICR-MS can reflect similar reactivity patterns of DBC (e.g., the preferential photodegradation of B5CA and B6CA by BPCA (Stubbins et al., 2012b) corresponding with the condensed aromatic structures by FT-ICR-MS (Stubbins et al., 2010)), thus complementing each other (Liu et al., 2022). Also, solid-state 13C nuclear magnetic resonance shows good potential in quantitative investigations of DBC structures, and the combination of 13C NMR and FT-ICR-MS makes it possible to determine the condensed aromatic structures containing nitrogen and sulfur (Abdul Jameel et al., 2019; Helms et al., 2015; Hertkorn et al., 2016; Li et al., 2015; Porto et al., 2019). Several studies have used 13C NMR to quantify the structural changes in DBC derived from different parent solids and to compare their functional constituents with NDOM (Huang et al., 2021; Tian et al., 2019). Fu et al. (2016) first applied 13C NMR to distinguish the structural changes in DBC under photoirradiation and observed major decreases in aromatic and methyl peaks after irradiation. Other techniques, such as GC–MS, can be applied for qualitative assessments of less aromatic DBC and used as supplementary tools for the accurate determination of possible molecules. For example, Hao et al. (2018) used GC–MS to verify the potential molecules identified via FT-ICR-MS and found the concentrations of specific molecules as a function of their hydrothermal carbonization temperatures, suggesting the potential of GC–MS to uncover structural changes in DBC. However, GC–MS are restricted to volatile organic compounds with low molecular weights, and they cannot identify polar organic compounds with high molecular weights. As a result, the information for more than 90% of DBC molecules is lost.

1.2 Chemical and molecular structures of DBC

Owing to the varied organic matter sources and formation processes, the chemical and structural properties of DBC significantly differ from those of well-studied NDOM. Herein, we summarized the structural characteristics of DBC, which is a crucial basis for understanding its environmental behavior.

1.2.1 Optical spectral characteristics of DBC

DBC derived from different parent solids exhibits various UV–Vis spectral features. Several studies have suggested that DBC released from bamboo biochar exhibits broad, featureless, and declining spectral characteristics, which are similar to those of humic substances (Chen et al., 2002; Qu et al., 2016; Wei et al., 2017). Fu et al. (2016) and Bostick et al. (2020) also showed featureless UV–vis spectra for DBC released from bamboo and oak biochars generated at 650 °C. However, DBC released from oak biochars generated at 400 °C (Bostick et al., 2020) and bamboo hydrochars (Hao et al., 2018) displayed characteristic absorption peaks. Some spectral parameters were applied to indirectly infer the structural characteristics of the DBC. It is generally acknowledged that DBC is more aromatic than NDOM, as indicated by SUVA254 values. Fu et al. (2016) concluded that DBC may have smaller molecular weights than DOM, as the E2/E3 parameters for DBC (5.47) were higher than those of commonly studied DOM (3.06 ~ 4.77) (Fu et al., 2016). However, some studies have shown that the E2/E3 values of NDOM may exceed 8.6 (Helms et al., 2008), while those of DBC may be as low as 1.7 ~ 2.9 (Bostick et al., 2020). Such results may drive a widely divergent conclusion. Thus, the information provided by UV–vis is still limited, and comparisons between different studies are constrained. In addition, owing to the disparities in DBC sources, no unanimous conclusions about DBC spectral characteristics can be reached, at least for the time being.

Three-dimensional fluorescence (EEM) spectra can be used as an indicator of DBC component characteristics. Table 1 summarizes the diverse fluorescence peaks and corresponding components of DBC. In general, DBC produced at higher temperatures has simpler EEM spectra. In addition, the EEM spectra of DBC are simpler than those of NDOM, as the latter often contain several peaks in different EEM regions (Chen et al., 2003). To obtain more compositional information about DBC, EEM-PARAFAC has become increasingly popular (Chen et al., 2020; Han et al., 2021; Yamashita et al., 2021). However, this method requires large sample quantities, and the effectiveness of component separation is hard to guarantee. Moreover, the comparison of PARAFAC model between different studies is hard. One useful solution is to upload the PARAFAC model to the OpenFluor online software (Murphy et al., 2014), which facilitates the searching of the similar components reported in previous records.

Table 1 Fluorescence peaks and corresponding components of DBC derived from BC of different biomass types and generated at different production temperatures
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1.2.2 Chemical structures of DBC

NDOM are mainly composed of aliphatic structures (> 50%), as indicated by 13C NMR analysis (Fig. 1). Owing to the intensive dehydration and condensation reactions during thermal treatment, both biochar- and hydrochar-derived DBC generally possess more abundant aromatic structures but fewer aliphatic structures than NDOM. Typically, hydrochar-derived DBC contains more aliphatic structures than biochar-derived DBC does. In addition, biochar-derived DBC is also more abundant in carboxy or aldehyde functional groups than NDOM, while hydrochar-derived DBC barely contains these structures. As the parent material’s thermal maturity increases, the proportions of aromatic and alkyl components in biochar-derived DBC increase, while those of oxygenated functional groups decrease (Bostick et al., 2018). With respect to hydrochar-derived DBC, the proportion of O-alkyl structures, aromatic C-O and aldehyde decrease, while those of carboxyl and aromatic C increase (Hao et al., 2018). This summary confirmed the higher aromaticity of DBC than NDOM and further suggested that pyrolysis technique can produce more aromatic structure than hydrothermal carbonization treatment.

Fig. 1
figure 1

Relative abundance of different functional groups in NDOM, biochar-derived DBC (B-DBC) and hydrochar-derived DBC (H-DBC) by 13 C NMR (Cao et al., 2016; Fu et al., 2016; Hao et al., 2018; Huang et al., 2021; Qu et al., 2016; Tian et al., 2019). The spectra was divided into the following regions: alkyl C (0–45 ppm), methoxyl C (45–63 ppm), carbohydrate C (63–108 ppm), aryl C (108–148 ppm), O-aryl C (148–165 ppm), COO/NC = O C (165–190 ppm) and C = O C (190–220 ppm) (Chen et al., 2020)

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1.2.3 Molecular structural characteristics of DBC

With respect to molecular structures, DBC is generally composed of condensed aromatic compounds, aliphatic hydrocarbons, thermally treated biopolymer oligomers, and low molecular weight components (Bostick et al., 2018; Fu et al., 2016). As indicated by Fig. 2a, the average molecular weights of different DOM follow a decreasing order of biochar-derived DBC, riverine NDOM and hydrochar-derived DBC. In addition, biochar-derived DBC exhibits higher double bond equivalents (DBEs), O/C ratios and AImod but lower H/C ratios than riverine NDOM (Fig. 2b–e). As the production temperature increases, the molecular structures of DBC shift to lower molecular weights (Hao et al., 2018). The molecular diversity and H/C ratios of biochar-derived DBC decrease, while the O/C ratios and DBE increase (Huang et al., 2021). For hydrochar-derived DBC, however, higher production temperatures often lead to higher molecular diversity (Hao et al., 2018). When the temperature exceeds a certain limit, however, the molecular diversity may exhibit a slight decrease, which can be potentially attributed to the repolymerization of small organic compounds (Hao et al., 2018).

Fig. 2
figure 2

Differences in the molecular structures, including a) average molecular weight, b) DBE, c) H/C atomic ratio, d) O/C atomic ratio and e) AImod, between riverine NDOM, biochar-derived DBC (B-DBC) and hydrochar-derived DBC (H-DBC) by FT-ICR-MS (Chen et al., 2022; Hao et al., 2018; Huang et al., 2020, 2021; Seidel et al., 2014; Stubbins et al., 2010; Wagner et al., 2015; Ward et al., 2014; Yan et al., 2022)

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Although the chemical properties of DBC vary with the solid structures of the parent material (e.g., biomass type and production parameters), the common aromatic characteristics of DBC can be used as a marker for their content in environmental samples. Kim et al. (2004) first applied FT-ICR-MS to identify the structural characteristics of DBC in aqueous environments. The differentiation of DBC in environmental DOM samples is based on the consistency of their molecular elemental compositions (H/C < 1) with solid BC (0.27 < H/C < 1.3) and their differences from NDOM (H/C > 1). In addition, DBE characteristics are also applied to distinguish DBC (DBE/C > 1) from NDOM. Some studies also used the BPCA method to differentiate DBC in environmental DOM samples (Mori et al., 2021). The distinction was based on the fact that DBC molecules contain a higher average quantity of aromatic rings (> 7) (Dittmar and Koch, 2006) than NDOM (< 5) (Hockaday et al., 2007).

1.3 Environmental behaviors of DBC

1.3.1 Photochemical degradation of DBC

DBC is acknowledged as one of the more photoactive components in the DOM pool (Fu et al., 2016; Stubbins et al., 2012b), and its photodegradation rate may far exceed that of colored DOM (Stubbins et al., 2012b). It has been estimated that the photochemical half-life of oceanic DBC is less than 800 years, far below its previously assumed apparent age of 1,500 years (Stubbins et al., 2012b). As such, photochemical degradation may be one of the significant pathways for DBC mineralization and alteration in surface waters (Fu et al., 2016; Stubbins et al., 2012b, 2010; Wagner et al., 2018; Ward et al., 2014).

There are two different stages of DBC photodegradation (Bostick et al., 2020). The most photoactive fractions are lost in the initial 2 days and roughly correspond to the aryl-C by 1H NMR and condensed aromatic carbons (ConACs) by the BPCA and FT-ICR-MS method (Bostick et al., 2020; Yan et al., 2022). The less photolabile fraction of non-condensed components (small aromatic compounds, aliphatic compounds including low-molecular-weight acids, and alcohols), however, is likely to be photo-mineralized by radical-driven propagation reactions generated by the mineralization of aromatic components (Fig. 3). These results indicated that photo-mineralization may be chemically selective and act on different photoactive components with the extension of irradiation time. Moreover, the carbon loss sequence of DBC produced at different temperatures varied with photoirradiation time. Yan et al. (2022) reported that the mineralization ratio of 300 °C DBC was higher than that of 450 °C DBC during 24 h of photoirradiation. Chen et al. (2022) also showed that 300 and 450 °C DBC degraded faster than 650 °C DBC after 24 h of irradiation. Bostick et al. (2020) showed that the photodegradation ratios of 250 and 525 °C DBC were higher than those of 400 and 650 °C during the initial 24 h, respectively, while those of 525 and 650 °C DBC were still higher than those of 250 and 400 °C DBC. Such results indicated that the photodegradation pattern of DBC was quite complex during the first stage, potentially attributed to the heterogeneous reactivity of different subcomponents in DBC. After a long period of photoirradiation, however, a similar pattern that the mineralization ratio of DBC increased with production temperature was observed in different studies (Bostick et al., 2020; Yan et al., 2022), suggesting the overall higher photolability of DBC produced at higher temperatures. However, due to the use of different feedstock for DBC production and the application of varied illumination intensity, the photodegradation pattern of DBC was still largely unknown.

Fig. 3
figure 3

Pathways for reactive species generation from DBC under photoirradiation and potential photo-transformation mechanisms of pollutants mediated by DBC. P represents pollutants

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Photodegradation processes have a preference for specific components, such as the chromophoric fractions and aromatic substances. Many studies have reported a rapid decrease in the fluorescence peak intensity of EEM spectra (Bostick et al., 2020; Fu et al., 2016), indicating that fluorescent DBC fractions are highly susceptible to light irradiation. Bostick et al. (2020) further pointed out that some fluorescent components are especially photolabile. For example, humic acid-like fluorescent compounds were particularly susceptible to loss, and their degradation ratio was even higher than that of less condensed fulvic acid-like fluorescent compounds. Aromatic components have also been acknowledged as photoactive fractions in DBC, but their half-lives are longer than those of the fluorescent fractions (Fu et al., 2016). Stubbins et al. (2010) first reported direct evidence for the photodegradation of DBC like molecules in Congo River and showed that more than 90% of aromatic compounds potentially associated with BC were lost based on FT-ICR-MS. Ward et al. (2014) also revealed a nearly complete loss of aromatic compounds in DBC after 17 h of sunlight irradiation. Several studies also observed major losses in benzene polycarboxylic acids after irradiation and suggested that ConACs have especially poor photo-resistance (Liu et al., 2022; Spencer et al., 2009; Stubbins et al., 2012a, 2012b). Given that the SUVA254 value is linearly correlated with the aromatic content by 13C NMR (Weishaar et al., 2003), the decrease in SUVA254 during irradiation observed in previous studies also suggests a major loss of the aromatic components of DBC (Bostick et al., 2020). In addition to aromatic components, photo-transformation also preferentially targets aliphatic hydrocarbons containing methyl functional groups (Fu et al., 2016), as well as polysubstituted olefinic moieties (Goranov et al., 2020). However, Yan et al. (2022) suggested the removal of condensed aromatics and carbohydrate in DBC during photoirradiation, while Chen et al. (2022) reported the photo-mineralization of lignin-like components. The mismatch may be due to the use of different feedstock types for DBC production and the varied intensities of photobleaching treatments.

During irradiation, the molecular diversity and average molecular weights of DBC show a decreasing trend (Chen et al., 2022; Stubbins et al., 2010; Wagner and Jaffé, 2015; Yan et al., 2022). The apparent photo-induced transformation in DBC relative abundance from higher molecular weight fractions to lower molecular fractions can be attributed to the combined processes of complete mineralization to CO2 and structural photo-alteration (Wagner and Jaffé, 2015). The photo-transformation of DBC was dominated by photo-oxidation (> 68%), with photo-mineralization contributing only approximately 8 ~ 13% as the subsidiary reaction (Ward et al., 2014). On the one hand, the selective removal of DBC with higher molecular weights and the preservation and generation of DBC with lower molecular weights by irradiation may generate a secondary DBC pool with potentially enhanced bioavailability (Bruun et al., 2008). Also, the lower molecular weight fractions may exhibit priming effects on DBC biodegradation (Woods et al., 2010). On the other hand, some of the biolabile subcomponents in DBC can also be degraded under photoirradiation (Bittar et al., 2015). Thus, the mixing effects by photodegradation and biodegradation processes may significantly alter the turnover rate of DBC and merit further investigation.

DBC structure may be one of the most important factors influencing its photolability. Although several studies have revealed the varied ConAC content in DBC from different sources (Bostick et al., 2018; Fu et al., 2016; Norwood et al., 2013; Ward et al., 2014), few studies have investigated their corresponding photolability. Bostick et al. (2020) compared the photolability of DBC derived from oak biochars at different charring temperatures and found that DBC produced at higher temperatures was more photoactive than that at lower temperatures. Our recent work also supported this result (Yan et al., 2022). Beyond charring temperature, the effects of other factors, such as biomass type (e.g., livestock manure with high protein and traditional lignocellulosic biomass), deserve further investigation.

Exposure to sunlight can also enhance the release of DBC from its parent solid. Previous studies have shown that the photo-dissolution of particulate organic matter is a significant process for generating DOM (Mayer et al., 2006, 2012; Pisani et al., 2011; Shank et al., 2011). Similarly, photo-dissolution may contribute to DBC generation from solid BC in aquatic environments (Alan Roebuck et al., 2017). In a natural aqueous environment, photo-dissolution from BC and photodegradation of DBC can take place simultaneously, making the photochemical process more complicated. By comparing the amount of DBC generated from BC with and without irradiation, Alan Roebuck et al. (2017) showed that photo-dissolution is a potential source of DBC in aquatic systems. However, the simultaneous or further photodegradation of newly generated DBC by photo-dissolution is poorly understood.

Overall, the potential mechanism for DBC photodegradation can be summarized as two pathways: (1) direct photodegradation and (2) radical-driven propagation reactions (Fig. 3). During the direct photodegradation process, the most photolabile fractions undergo primary photoreactions under sunlight irradiation of a prescribed wavelength and often generate the radicalized excited triplet states of DOM (3DOM*) (McNeill and Canonica, 2016), thus inducing the generation of singlet oxygen and superoxide (Fu et al., 2016; Zhou et al., 2018). These generated substances will further oxidize DOM through radical-driven reaction propagation. During this process, quenched radicals will be constantly regenerated and drive the transformation of DBC into lower molecular fractions, similar to the transformation of large fulvic and humic acid chromophores into smaller organic acids (Goldstone et al., 2002; Pullin et al., 2004).

1.3.2 Interaction between DBC and microorganisms

To date, few studies have investigated the interaction between DBC and microbes, which is in fact one of the most significant processes for DBC transformation and microbial community succession in aquatic environments. On the one hand, microbes can alter the molecular composition of DBC via catabolism and anabolism (Bostick et al., 2021; Chen et al., 2022; Goranov et al., 2022). On the other hand, DBC provides a unique habitat as a refractory carbon source and may thus affect the constitution and succession of the microbial communities (Chen et al., 2022; Zhou et al., 2021).

BC has long been considered as a refractory fraction of the organic carbon pool and exhibits chemical stability and resistance to biodegradation. Its soluble fraction (DBC), however, is unlikely to be recalcitrant to microbial degradation. Qi et al. (2020) revealed the biodegradability of DBC via laboratory charcoal leaching experiments and a field study on DBC 14C ages in rivers and oceans. They found that the DBC in rivers is predominantly young via 14C ages, and no DBC accumulated in the microbe-active water during long-term leaching experiments. Norwood et al. (2013) subjected fresh DBC leached from charcoal to microbial incubation and found the removal of more than 50% of the total DBC within one month. Similarly, high biomineralization ratio (30%–70%) of biochar-derived DBC in 2‒3 months were also reported in several studies (Bostick et al., 2021; Chen et al., 2022; Goranov et al., 2022). Such results provide direct evidence for the instability of DBC. However, the potential mechanisms driving the biomineralization of DBC are still unclear. One explanation is that DBC shows heterogeneous reactivity and the highly biolabile subcomponents in DBC may be easily degraded and even exhibit priming effects on bulk DBC biodegradation (Bianchi Thomas, 2011). Further study should underline the effects of biolabile fraction on DBC biodegradation possibly through subjecting DBC of varied aromaticity to microbial incubation.

Several studies have investigated the impact of DOM on the composition of microbial communities. Zhou et al. (2021) demonstrated that the source aromaticity of DOM influenced microbial community succession in aquatic environments. DOM with higher aromaticity exhibited much stronger environmental filtering of the microbial community. DBC is more aromatic than most DOM in nature and will inevitably change the microbial community structures in aquatic environments. Osterholz et al. (2016b) observed a strong correlation between DOM and total bacterial community composition in a pelagic marine system. In the DOM pool, some condensed aromatic components (representing DBC) partially correlated with DNA-based bacterial community composition, supporting the significant role of DBC in microbial activities. Some recent studies further provided direct evidence for microbial community variations related to DBC molecular structures (i.e., aliphatic compounds and condensed polycyclic aromatics) and incubation time (Chen et al., 2022; Zhang et al., 2020a). Moreover, Proteobacteria, which comprised a huge variety of nitrogen-fixing bacteria, proved to be the predominant phylum in DBC (Chen et al., 2022). Research hitherto has only concerned about the community succession in DBC. Thus, further investigations regarding more aspects of DBC mineralization, such as the measurement of functional genes, should be conducted to better uncover the biogeochemical process of DBC.

In return, microbes also play a crucial role in regulating DBC transformation. Goranov et al. (2022) reported the microbial diversification of DBC molecules after 10 days of incubation, while Chen et al. (2022) demonstrated a decreased molecular diversity after 56 days of incubation. One possibility for this mismatch is the use of different feedstock types for DBC production and the choice of different photoirradiation intensity. Also, the different incubation time may also serve as an explanation. It is reasonable to speculate that the molecular composition of DBC may first become diversified and then less diverse. Further work should focus on the molecular shifts of DBC at different incubation time scales.

DBC is both photodegradable and biodegradable in nature. After photoirradiation, the aromaticity and bioavailability of DBC were constantly altered. The combined effects of photoirradiation and microbial utilization are of great significance as they are closely related to the cycling and turnover of DBC in an aquatic environment, where DBC is exposed to continuous photoirradiation and microbial utilization. Until recently, a few studies started to explore the effects of photodegradation on the biomineralization of DBC. An increased biomineralization ratio of DBC after photoirradiation was commonly observed (Bostick et al., 2021; Chen et al., 2022; Goranov et al., 2022). One probability is that the highly aromatic and recalcitrant subcomponents in DBC are decomposed into smaller sized aromatic clusters with higher bioavailability. Another explanation is that the photogenerated small molecules have priming effects on the further biodegradation of recalcitrant structures.

Hydrochar-derived DOM is also a type of DBC in a broad sense. Hao et al. (2018) compared the growth of cyanobacteria in DBC solutions extracted from a thermal series of bamboo hydrochar and determined that DBC samples with > 210 °C production temperatures exhibited a certain toxicity to cyanobacteria, with DBC produced at 300 °C causing the highest inhibition. However, 180 °C hydrochar-derived DBC and 500 °C biochar-derived DBC promoted the growth of cyanobacteria. Similarly, Zhang et al. (2021a) also revealed the general phytotoxicity of a thermal series of biochar-derived DBC on the germination of cabbage seeds. These results suggest that the production parameters of biochars and hydrochars should be optimized to guarantee safety toward aquatic systems by considering their toxicity toward aquatic organisms.

1.3.3 Interaction between DBC and pollutants

In recent decades, DOM has been proven to play a key role in the fate and transport of pollutants (Leal et al., 2013; Wenk et al., 2011; Zhou et al., 2017). Similarly, DBC can affect the transport and photodegradation of contaminants via sorption and photoproduction of reactive intermediates. Table 2 summarizes the available records for interactions between DBC and contaminants.

Table 2 Effects of DBC on the fate and transport of pollutants
Full size table

First, DBC exhibits a high adsorption capacity for hydrophobic pollutants. Their condensed aromatic structure has high hydrophobicity and can adsorb hydrophobic organic pollutants, especially hydrophobic organic contaminants, via hydrophobic interactions (Chin et al., 1997; Gauthier et al., 1987; Kopinke et al., 2001; Perminova et al., 1999). Besides, the pseudomicellar conformation of DBC could promote the binding interaction between DBC and hydrophobic organic contaminants (Fu et al., 2018a). Also, DBC contains abundant polar functional groups (e.g., carboxyl and hydroxyl groups) and can adsorb polar organic contaminants through hydrogen bonding and electrostatic interaction (Klaus et al., 1998; Senesi, 1992; Yamamoto et al., 2003). Furthermore, DBC can coordinate with heavy metals via complexation (Stubbins et al., 2010; Wu et al., 2011; Xia et al., 1997). Moreover, as DBC is defined as the water-soluble fraction of < 0.45 μm, it may exert some overlap with the definition of nano BC (< 1 μm) (Liu et al., 2018; Song et al., 2019) and holds pollutants well (Ramezanzadeh et al., 2021; Zhu et al., 2021). Tang et al. (2016) demonstrated that hydrophobic interactions were the main interaction mechanisms between DBC and polycyclic aromatic hydrocarbons (PAHs). Xu et al. (2020) revealed that coexisting Pb2+, Cu2+, Cd2+ and Zn2+ ions enhanced the aggregation and deposition of DBC via complexation and cation–π interactions, which helped to understand the fate of DBC and coexisting pollutants in environmental and engineering systems. Ramezanzadeh et al. (2021) found that nano BC was more effective at cadmium sorption in contaminated sandy soils than macro BC. Mahmoud et al. (2020) showed that nano BC was also a promising sorbent for phosphorus mainly through surface precipitation, potentially via complexation reactions between phosphorus and the mineral component in DBC.

Second, DBC can participate in the redox conversion processes of pollutants in the environment as a reactant or an electron transport carrier. Similar to NDOM, DBC contains a variety of redox-active organic molecules (e.g., those bearing quinoid, aromatic and thiol moieties). These species can act as reductants, reactive intermediaries or electron shuttles and affect the chemical and biological redox-mediated reactions of pollutants. Graber et al. (2014) demonstrated that DBC possessed significant redox activity and can promote the reduction and solubilization of Mn and Fe. In addition, DBC generated at lower production temperatures had a stronger redox ability. Dong et al. (2014) revealed that DBC could either act as an electron donor or an acceptor and efficiently reduce Cr(VI) or oxidize As(III), respectively.

Third, DBC can generate a large variety of reactive intermediates (e.g., 3DBC*, ·OH, 1O2, and O2–) under photoirradiation, which can promote not only its own photodegradation but also the photo-transformation of important pollutants (Fig. 3). It has been reported that the photo-reactivity of a DOM is closely associated with its aromaticity and molecular weight. DOM with higher aromaticity can absorb photons more efficiently (Boyle et al., 2009). Compared with NDOM, DBC has lower molecular weights and higher aromaticity (Qu et al., 2016; Zhou et al., 2018) and has proven to be one of the more photoactive components in the DOM pool. Thus, DBC may act as an efficient driver for the photo-transformation of contaminants. Fang et al. (2017) suggested that DBC contributed to the photodegradation of diethyl phthalate via the generation of •OH and 1O2. Serelis et al. (2021) compared the effects of biochar- and hydrochar-derived DBC on metribuzin photodegradation and observed that both char-derived DBC enhanced metribuzin photolysis, with hydrochar-derived DBC presenting a stronger optical filter effect and biochar-derived DBC exerting higher ability to generate •OH and 1O2. Zhou et al. (2018) found that DBC could significantly enhance the photo-transformation of 17β-estradiol by generating 3DBC* species. Tian et al. (2019) also proposed the same mechanism for chlortetracycline photo-transformation via 3DBC* attack. With respect to the photo-transformation of Ag+ to nano Ag particles, the photo-reduction efficiency of DBC is higher than (Liu et al., 2021) or in opposition to those of NDOM (Wan et al., 2021). Furthermore, Wang et al. (2020) investigated the effects of DBC on the photodegradation of 24 pharmaceutically active compounds and demonstrated that DBC could accelerate the photodegradation of such contaminants that were susceptible to one-electron oxidation by 3DBC*. Moreover, both the laboratory prepared and naturally occurring DBC are expected to contain mineral compositions, which could serve as a natural photocatalyst mediating the photo-redox reactions of pollutants (Fu et al., 2018b).

In addition, DBC can also affect the environmental behaviors of some emerging contaminants. For example, DBC can be strongly adsorbed on the surfaces of nano-Ag, facilitating DBC transport in porous media (Wang et al., 2021). In addition, DBC can enhance the destabilization and deposition of polystyrene by reducing electrostatic repulsion, which is in contrast to the effect of humic acid (Xu et al., 2021).

As one of the most labile and mobile fractions in the BC continuum (Wang et al., 2013), DBC can easily be released via soil infiltration and surface runoff (Dittmar et al., 2012a; Jaffe et al., 2013). The potential role of mobile DBC as a carrier of contaminants may query the role of BC as long-term remediation materials. Thus, more efforts should be made to uncover the interaction mechanisms between DBC and pollutants to better evaluate the applicability of BC in immobilizing pollutants.

2 Future perspectives

To better understand the environmental behavior of DBC, further efforts should be made to develop new technologies of higher precision and resolution, as well as jointly apply existing advanced techniques, such as combined utilization of FT-ICR-MS and BPCA, GC–MS and 13C-NMR. Available studies have mainly focused on the photodegradation processes of DBC. Information regarding the microbial transformations of DBC molecules, the impact of DBC on microflora and expression of C metabolism- and N-related functional genes, and the transformation and degradation of coexisting pollutants regulated by DBC, however, is far from robust. In addition, the lack of direct comparisons between the environmental transformation processes for DBC and NDOM, such as the different degradability, the preferential degradation or utilization of different molecular during photodegradation and biodegradation, and the different role of ROS plays in regulating their photo-transformation, increase the complexity of distinguishing the environmental behaviors of DBC. Photodegradation and biodegradation processes occur simultaneously in real environments. As such, both processes should be considered in evaluating the stability of DBC. After photoirradiation, the aromaticity, molecular weights and methyl functional groups of DBC decrease. Therefore, it is plausible to hypothesize that the residual DBC may be more bioavailable. Only a few studies started to focus on the biodegradation of photobleached DBC; however, the simultaneous photodegradation and biodegradation of DBC still merits further investigation. In recent years, the selective protection of different components in DOM by minerals has been widely reported. Similarly, after selectively adsorbed to minerals, the subsequent biodegradation of the presumably protected DBC fractions is largely unknown and may contribute to the understanding of the fate of DBC in real environment.