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Environmental behavior of engineered biochars and their aging processes in soil


In recent years, due to the broad application of biochars, the preparation, environmental behavior and aging processes of biochars have attracted wide attention globally, especially the modification of biochars. However, most of the studies only consider the improvement of biochar properties right after the modification, but neglect a complete evaluation of the long-term stability and eco-toxicity of these newly developed materials after entering the environment. With the development and utilization of biochars, engineered biochars (EngBCs) will soon enter the market, but its environmental risk still remains unclear. The literature does not provide adequate information on how aging of EngBCs will affect their properties, and indirectly impact the properties of soils (cycle of elements and organic matter). Therefore, this review paper summarizes the aging process and environmental risk of biochars, aiming at better understanding the interactions between EngBCs and soil components or pollutants. More importantly, this review is to point out the contradictory speculations of environmental behavior of EngBCs studied at the present stage. Due to the modification, the EngBCs stability may be significantly reduced. However, the formation of functional group on EngBCs will enhance their interaction with soil minerals to form biochars–mineral complex, and thus EngBCs could be protected. The impacts of EngBCs after entering the environment are also ambiguous. Therefore, understanding EngBCs environmental behavior is critical, which is helpful to reduce the potential risk and to produce EngBCs following the rule of sustainable development and safety to the environment.


In recent years, many research interests have been focused on the application and development of biochars to benefit agriculture production by providing nutrients and improving soil physical properties (Atkinson et al. 2010; El-Naggar et al. 2019; Schulz and Glaser 2012), soil/water remediation because of its large surface area and abundant surface functional groups (Ahmad et al. 2014; Huang et al. 2019; Tang et al. 2013), and carbon sequestration due to its high chemical and biochemical stability (Fang et al. 2014; Lehmann 2007). In most cases, pristine biochars have relatively low surface area and thus low adsorption capacity (Ahmad et al. 2014; Rajapaksha et al. 2016). Purposive “tailoring” of their properties enhanced their effectiveness in engineering applications (Rajapaksha et al. 2016). Biochar tailoring is also known as biochar modification. Massively produced biochars and their modified products are also referred to as engineered biochars (EngBCs). Available modification methods have been discussed in several review papers (Liu et al. 2012; Rajapaksha et al. 2016; Wang et al. 2018a; Xiong et al. 2017). The modification methods can be divided into four categories including chemical modifications (acid, alkali and oxidizer) (Boguta et al. 2019; Huff and Lee 2016), physical modifications (steam/gas, ball milling, and microwave) (Lyu et al. 2018; Morgan et al. 2017; Rajapaksha et al. 2015), impregnation with metal (magnetic amendment and nanoparticle modification) (Harikishore Kumar Reddy and Lee 2014; Tan et al. 2016) or organic sorbents (coating) (Zhou et al. 2013), and biological modifications (Frankel et al. 2016; Yao et al. 2011). The main purpose of these modification methods is to increase the adsorption selectivity or adsorption capacity by increasing the pore volume, surface area, and surface functional group. Modifications are performed either pre- and/or post-pyrolysis of the biochars (Sizmur et al. 2017). While pre-pyrolysis modifications involve the treatment of feedstock, post-pyrolysis modifications are more common and involve the treatment of biochars. After chemical modification, the functional groups, micropores as well as surface area and cation exchange capacity (CEC) of biochars can be improved, thereby leading to an enhanced sorption capacity to various pollutants (Goswami et al. 2016; Xue et al. 2012; Zhang et al. 2013). Physical modification alters their particular size and pore structure, and offers advantages over chemical modification since this modification is clean and easy to control with low cost (Shim et al. 2015). However, some research indicated that the physical modification (e.g., steam activation) could remove biochar surface functional group leading to an increase in aromaticity, and a neglect improvement in metal ion removal was identified (Shim et al. 2015). Metal oxide or organic sorbent impregnation modification has been used to facilitate chemical co-precipitation and electrostatic interactions, or enhance the CEC and surface area (Ahmad et al. 2014; Huang et al. 2019). It also can enable biochars to have some unique properties, such as magnetic driven ability and high selectivity in complex media. Biological modification of biochars usually is to produce biochars from the biomass pre-treated by bacterial conversion (Yao et al. 2015) or biochars post-treated by biofilm (Frankel et al. 2016). A few studies found that the biofilm covered on the carbon materials could significantly enhance the adsorption and degradation rate of organic chemicals (Scott et al. 1995). Biochars prepared from the anaerobic digestion residue exhibited promising adsorption removal potential for organic dye and heavy metals (Inyang et al. 2012; Sun et al. 2013).

Recently, the stability of pristine biochars (de la Rosa et al. 2018; Mia et al. 2017) with slow degradation rate (Kuzyakov et al. 2014; Leng et al. 2019) has attracted interests as a potential carbon sink capable of reducing greenhouse gas emissions. However, not enough attention has been paid toward the stability and property of EngBCs after modification. Many researchers only emphasized its capability and efficiency in contaminant removal, but neglected its stability changes (Bogusz et al. 2015; Tang et al. 2013). Due to the changed physicochemical properties of EngBCs, their stability definitely has been changed because most of the modification method involved harsh treatments (such as high temperature and low pH) (Iriarte-Velasco et al. 2016; Rajapaksha et al. 2016; Zhao et al. 2017). These property changes of biochars also can alter their interactions with contaminants (e.g., heavy metal and organic pollutants), as well as mineral and organic components of soil/sediments. These processes may have a serious impact on the cycle of carbon and nutrients in the environment, and on the mobility and bioavailability of various contaminants in soils and in biochars themselves. The eco-toxicological characteristics of these EngBCs may change significantly as a result of aging. For example, the aging of modified nano-ZnO or nano-Mg EngBCs might lead to the release of metal nanoparticles from EngBCs, interfering natural processes and affecting the biological diversity in soils or waters.

With the fast development of biochar technology, understanding EngBCs environmental behavior, which includes the long-term environmental stability, the interaction with contaminant and soil components and the toxicity during their lifetime, is important and urgent. Over the past decade, the application of biochars is sharply increased all over the world (Huang et al. 2019). In the near future, EngBCs will be readily available on the market and widely spread. Researchers not only need to identify efficient modification methods for specific engineering purposes, but also have to understand their structural transformation and alternation during aging in order to evaluate their eco-toxicity. However, current studies did not provide sufficient information related to the environmental risks of EngBCs. Researches of fundamental importance for safe use of EngBCs include: (1) determination of the toxicity of EngBCs themselves during or after the process of modification/activation; (2) exploration of their interaction mechanisms not only with the targeted contaminates, but also with other coexisted contaminants and natural soil/sediment components; (3) examination of their stability in soils as a result of various environmental processes during aging. As a long-term soil remediation agent, understanding the dynamics of EngBCs properties is critical for the real world application, which will optimize EngBCs production, reduce their potential risk, and ensure the safe and sustainable development of biochar technology.

EngBCs modification

In this section, the EngBCs modification methods using different reagents and processes will be briefly discussed.

Chemical modification

HNO3 is widely used as a chemical modification agent to increase the abundance of carboxylic groups on the biochars, which could significantly increase the removal efficiency of metal contaminants, such as Cu2+ (Hadjittofi et al. 2014). The adsorption capacity was pH dependent, confirming the active adsorption site of oxygen-containing functional groups on biochars surface. The H3PO4 treatment prior to pyrolysis of pine tree sawdust increased the total pore volume and the content of micropores (Chu et al. 2018; Zhao et al. 2017). The P–O bond formed in the C structure acted like a backbone to prevent micropores clap during the pyrolysis, which enhanced the adsorption capacity for organic pollutants (Chu et al. 2018). Oxidation with H2O2 is another common method used to modify biochars (Wang et al. 2015, 2016, 2018c). After H2O2 modification, the amount of carboxyl groups and cation exchanges sites on biochars can be significantly increased, which is beneficial to heavy metal removal. For example, the enhanced lead adsorption capacity can achieve 22.82 mg/g, which is 20 times higher than that of the unmodified biochars (Xue et al. 2012). H2O2 modification also reduced biochar ash content due to the increased water accessibility after carbon structure modification, which lead to the change of heavy metal sorption mechanism from carbonate/phosphate precipitation to complexation with carboxyl function group (Wang and Liu 2018). Alkali modification of biochars usually involves potassium hydroxide (KOH) and sodium hydroxide (NaOH) to increase oxygen-containing functional groups and the percentage of surface graphitic carbon, which can enhance the adsorption of organic pollutants by π–π interaction (Fan et al. 2010).

Physical modification

Biochar physical modification methods usually include steam/gas activation, microwave modification, and ball milling. Steam activation is a common physical modification method. It is often used to increase the porosity and remove the impurities to increase the specific surface area of EngBCs. Recent investigations showed that steam modified biochars enhanced the pharmaceutical (sulfamethazine) removal by 55% compared to the non-activated biochars (Rajapaksha et al. 2015). On the other hand, Shim et al. (2015) indicated that the steam activated EngBCs and pristine biochars had similar Cu2+ sorption capacity. They found that steam activation significantly increased the biochar’s surface area but decreased the amount of functional groups. Therefore, steam modified biochars seem to be more effective for removing non-polar organic contaminant than charged metal ions. Microwave modification is based on high-frequency electromagnetic waves at frequencies ranging from 300 MHz to 300 GHz (Huang et al. 2013; Wang et al. 2018a). This modification method can rapidly and efficiently distribute uniform internal energy to heat biomass without direct contact and create EngBCs with more function groups and high surface area than conventional pyrolysis process at low temperature (Mašek et al. 2013). Mohamed et al. (2016) indicated that microwave-induced catalytic pyrolysis can increase the water holding capacity (up to 18.5%), cation exchange capacity and fertility of sand soil in comparison with the biochar produced using the conventional pyrolysis in the same conditions. Compared with the above-mentioned physical treatment of biochars, ball milling is a relatively new non-equilibrium processing method to produce EngBCs, which mechanically reduces the grain size of biochars to ultrafine (nano-size) particles (Mahbub et al. 2014). Lyu et al. (2018) demonstrated that this technology could increase the amount of oxygen functional on the biochars surface, as well as the internal and external surface area. Their results showed an enhanced Ni2+ removal efficiency (230–650 mmol/kg) than untreated biochars (26–110 mmol/kg).

Surface modification by impregnating metal oxide or organic sorbent

Biochars coated by nanoparticles or organic sorbents recently have received a lot of attention. For example, Wang et al. (2017b) coated Al2O3 nanoparticles on biochar surfaces using thin film coating methods through atomic layer deposition. They found that the amount of methylene blue adsorbed onto the EngBCs was nearly three times higher than that of the uncoated biochars during the first hour of adsorption.

Due to the small size and low density of most pristine biochars, it is extremely difficult to remove and reuse them after their application in wastewater treatment or soil remediation. Therefore, engineers proposed an efficient strategy to magnetize the pristine biochars to reinforce the separation efficiency. For example, Michalekova-Richveisova et al. (2017) showed that iron-impregnated biochars can be easily recovered by an external magnetic field and showed high removal efficiency for anionic forms of phosphorus. Mohan et al. (2014) synthesized magnetic oak bark biochars and magnetic oak wood biochars, which were also recovered by a low strength external magnetic field after trapping lead and cadmium from aqueous solution through chemical co-precipitation and electrostatic interaction.

Biological modification

Biological pretreatment of biomass is not an intentional modification method of biochars production initially. Most researches proposed to use pyrolysis process to reduce the amount of biofuels residue, wastewater sludge and increase the energy efficiency (Gupta et al. 2009). Recently, several studies indicated that biochars converted from anaerobically digested biomass is a very effective sorbent of heavy metals (Inyang et al. 2011; Yao et al. 2011). The lead sorption capacity was close to 200 mmol/kg, which is comparable to commercial active carbon (Inyang et al. 2012). Biofilm-modified biochars is the post-biological modification method. After the biochar was covered by bacterial, the heavy metals and organic contaminants removal efficiency was significantly improved (Scott et al. 1995). A study using a biofilm-modified softwood bard biochars showed 87% of naphthenic acid removal and the Fe and Al sorption was four times higher than non-modified biochars (Frankel et al. 2016).

As we discussed above, most of the researches only focus on improving the properties of biochars, and using fast and controlled lab conditions to display their broad application potential in water and soil remediation. However, the long-term stability and eco-toxicity for EngBC environmental applications have not been well understood or explained. We speculated that with the increasing complexity of biochar modification, their long-term performance in the environment will become more and more unpredictable. Therefore, urgent research attention should be paid to identify how the EngBC modification methods would impact the long-term stability and eco-toxicity of EngBCs.

EngBCs aging

Although biochars are considered as a carbon sequester due to their relatively stable carbon structure, it actually undergoes long-term changes depending on the environmental conditions (Mia et al. 2017). Numerous studies demonstrated that biochars undergo physical and biochemical changes in the environment, commonly referred to as aging, through the biotic and abiotic processes occurring in soil (Rechberger et al. 2017, 2019; Trigo et al. 2014). For this reason, the physicochemical properties of biochars differ at the consecutive stages of its aging, and will be substantially different from those of the fresh ones (Hale et al. 2011). Recently, the structural changes of biochars in soil environment is explored, and several factors have been identified to control the aging process, including the interactions between biochars and soil constituents (such as dissolved organic and inorganic matter through physical embedding), microorganisms (biological process) and plant roots (chemical process) (Mia et al. 2017; Rechberger et al. 2017).

Biotic aging processes

Research showed that biotic and abiotic processes in the field caused primary changes in biochars structure including aromatic structure (stable C) and aliphatic C (reactive C) (Kaal et al. 2009). Biochar as soil amendment is considered as a great habitat for soil microorganisms (Atkinson et al. 2010). Therefore, biochars are closely related to biological processes. Some researches confirmed that the application of biochars to soil can change microorganisms’ carbon utilization profiles and population structure. Depending on soil type, the soil respiration could be increased or decreased accordingly (Spokas and Reicosky 2009). Some of the soil microorganisms may use biochars as their carbon source and thus the degradation or even mineralization of biochars is mediated by biological processes. Studies indicated that during the early aging process of biochars, aliphatic C was mineralized by soil microorganisms (Braadbaart et al. 2009; Kuzyakov et al. 2014). In biochar structures, aliphatic C is the connection of aromatic moieties or are present as side branches of the aromatic layers (de la Rosa et al. 2018). Therefore, during the aging process, the aromatic C could be released into the environment and the ratio of reactive C to stable C changes overtime. Hockaday et al. (2007) indicated that soil fungi were responsible for the chars degradation to form water soluble molecules, which resemble oxidized condensed aromatic ring structures. The benzene rings located at the structure edges become detached since they are attached to the larger aromatic moieties by a smaller number of bonds than the rings located inside the core. At the bond-breaking sites, additional functional groups appeared and thus surface-oxidized biochars are formed. The surface-oxidized biochars are becoming more and more resistant to further decomposition due to the loss of the most labile moieties (Bird et al. 1999). At the final aging stage, the degradation initiated on the surface gradually moves into the material internal structure, which leads to a decrease in the size of aromatic moieties even to only several (~ 6) conjugated benzene rings with numerous surface functional groups.

Besides the direct biological degradation/mineralization of biochars, the indirect biological processes also change the fate of biochars in soil, such as the excretion of extracellular enzyme. For example, saprophytes can potentially impact the persistence of biochars, because their invasive hyphal growth and extracellular enzymes enable them to create colonies in biochars pores, which might lead to the biochars fragmentation. Except extracellular enzyme, other secretions during microbial activities also will greatly alter the microenvironment of biochar and thus facilitate its aging. The active secretions include CO2, NOx, NH3, root-derived signal molecules and rhizospheric allelochemicals. CO2 and NH3 could alter the local pH environment near biochars’ surface. The application of biochars can alter the production and intensity of root-derived signal molecules and detoxify rhizopheric allelochemicals which hinder mycorrhizal establishment (Warnock et al. 2007), and therefore, the enhanced microorganism and plant root activity might accelerate the biochar aging process (Rechberger et al. 2017).

Some soil animals may intake biochars as their food source. It is quite unlikely that these animals will fully digest biochars. However, it is reasonable to speculate a completely different biochars particle in the feces of soil animals, because of the interactions between biochars and the digestion fluid or other chemicals in the secretions. These EngBCs properties changes have not been carefully investigated and thus their fate, transport and aging process in soil still remain unknown.

EngBCs can be considered as already aged biochars, especially modified through chemical oxidation. After chemical oxidation, the oxygen functional groups of biochars will be significantly increased (Hadjittofi et al. 2014; Wang and Liu 2018). High content of O- or N-containing functional groups can increase the reactivity of biochars, and reactive functional groups along with the non-aromatic C structures which constitutes the main part of the labile biochars fraction and is vulnerable to biotic degradations in soil. For these reasons, we can speculate that EngBC might be less stable than regular biochars at the early aging stage (Fig. 1).

Fig. 1

The possible aging process of Fe3O4-modified EngBCs. Contaminant removal performance of EngBC is significantly increased after Fe3O4 modification. However, along with the aging, the functions of biochars may be dramatically decreased

Abiotic aging and physical embedding processes

Abiotic aging of biochars usually involve heat and light, and these processes could lead to the decomposition of aliphatic C, which consequently may result in the release of volatile organic carbon (VOC) such as pyrazines, pyridines, pyrroles and furans (Mia et al. 2017). The chemical reactions among inorganic carbon, N and P inside the biochars may also accelerate the dissolution of the inorganic carbon in fresh biochars (Carreira et al. 2006). Biochar’s stability in soil may not attribute only to their chemical structure, but also to the reduced accessibility to soil microorganisms and/or their enzymes due to the physical embedding of biochars by other soil components (Fang et al. 2014), such as soil organic matter (SOM), mineral particles, microbe debris or excretions. For example, SOM can be sorbed to biochars surfaces and a significant proportion of biochars pore spaces may be blocked by SOM. The formation of these biochars–SOM complexes may increase with time, and therefore, this biochar’s aging process may reduce the sorption of organic contaminants. The oxygen-containing functional groups formed during biochars’ aging may also facilitate soil mineral adsorption and to form biochars–mineral complexes. The complex structure has been confirmed to be resistant for biodegradation.

The physical embedding also includes the particle–particle interactions. Some natural nano-sized particles may be mixed with biochar particles, wedging into their curvature surface, or even the porous structures. Because of their different expansion properties, biochars may be broken into smaller particles during the freeze–thawing cycles or dry–wet cycles (Liu et al. 2018).

In a word, the aging processes of biochars may be influenced by direct biological degradation (aliphatic C chain break up, aromatic carbon releasing and oxygen-containing functional groups formation), indirect biological degradation (microbial secretions enhanced root growth and soil animal digestion), and physical embedding with soil components. After their modification, the liable carbon structures, oxygen-containing functional groups and the metal additives of the EngBCs are changed significantly. Therefore, the stability of EngBCs is hard to predict based on our current knowledge. It depends on the modification method, the extent of the modification, regional climate character, biological activity, and soil components. The biotic and/or abiotic processes occurring in the soil primarily cause changes in the EngBCs structure. Consequently the physicochemical properties as well as the chemical reactivity of EngBCs will vary at different stages of aging and will markedly differ from the properties of freshly made biochars. The bioavailability and mobility of nutrients and contaminants in the environment are regulated by reactive surfaces in the soils. An increase in reactive surfaces on biochars in soil through the formation of functional groups with aging may facilitate interaction of aged biochars with soil minerals, nutrients and contaminants and might lead to a slower EngBCs degradation rate at the later aging process. Several studies focused on the aging of the pristine biochars, but no study was conducted on EngBCs aging. Therefore, future researches should examine EngBCs degradation and stability depending on a specific situation, which is crucial for the estimation of the residence time of EngBCs.

The interactions between EngBCs and contaminants or soil components as affected by their aging

Soil contamination by heavy metals and organic pollutants has become a serious environmental problem all over the world. The chemical or physical modifications lead to the increase of EngBCs surface functional group and pore volume, which can more efficiently and selectively remediate the polluted soils. The specifically enhanced adsorption ability increases the interactions between EngBCs and contaminants (Ahmad et al. 2014), soil organic matter, minerals, and nutrients in soil (Brodowski et al. 2006). Hadjittofi et al. (2014) modified biochars with HNO3 to increase the abundance of carboxylic groups on the surface to enhance Cu2+ adsorption. However, some strong interaction between EngBCs and soil constitutes may cause a disturbance of a number of natural processes in soils. We speculated that this strong adsorption force of EngBCs may reduce the availability of nutrients for plants and dissolved organic carbon for soil microorganisms.

When the biochars are aged, surface functional groups such as carboxylic and phenolic groups are formed (Fig. 2). Furthermore, the physical properties of biochars, particularly specific surface area (SSA) and pore volume, might firstly increase and then decrease as the extent of aging or chemical oxidation become more severe, consequently altering the apparent sorption. Researchers found that the formation of functional groups during biochar aging controls the heavy metal sorption. In addition, their results indicated that the accessibility of biochar pore network and reactive surfaces are not the key factors responsible for heavy metal sorption capacity (Rechberger et al. 2019). Some studies pointed out that extreme oxidation or aging can reduce SSA, thereby reducing the sorption of hydrophobic organic carbons (Villacanas et al. 2006; Zhu and Pignatello 2005). On the other hand, biochars surface oxidation may increase the sorption of other non-target organic carbons (such as hydrophilic and charged organic contaminants) through electrostatic interaction, π–π electron donor–acceptor (EDA), and H-bond mechanisms (Ahmad et al. 2014; Sun et al. 2012). The oxygen-containing functional groups on the biochars surface, would also increase the hydrophilicity of biochars. Hydrophilic interaction can increase water sorption by enhancing H-bonds between hydroxyl or nitro groups of organic contaminants and electron-rich functional groups of biochars, such as carbonyl group (Chen et al. 2015). Water molecules can also form clusters and attach onto the surface functional groups of biochars. These clusters often block organic carbons from entering the micropores and reduce contaminant removal efficiency. More importantly, it may lead to the decreased remediation effectiveness of hydrophobic organic pollutants due to the decreased surface hydrophobicity. Hence, the adsorption efficiency to the target contaminants could be significantly increased for EngBCs because of the specifically intensified properties, but this high adsorption also leads to their strong interactions with other untargeted compounds. In addition, during biochars aging, the interaction with contaminants could change overtime depending on the extent of biochars oxidation. For instance, the aged biochars would be less favorable for hydrophobic pollutants removal, but favor heavy metal immobilization.

Fig. 2

Schematic graph of possible sorption mechanisms between aged biochars and organic contaminants or soil components. The interactions between aged biochars and organic contaminants include electrostatic interaction, π–π interaction, H-bond mechanisms, hydrophobic effect, pore filling, partitioning and degradation. The interactions between aged biochars and soil components include electrostatic interaction, bridge and surface coverage

Some researches indicated that the interactions between biochars and organic pollutants, includes not only the adsorption process, but also degradation. Researchers demonstrated that the dissolved organic matter in biochars can induce the production of reactive oxygen components in liquid phase (superoxide anion free radical ·O2, hydroxyl radical ·OH) (Fang et al. 2017; Yang et al. 2016). However, after quenching these reactive free radicals in the liquid phase, the degradation of organic pollutants was not substantially reduced. They attributed the degradation to the direct reactions with environmental persistent free radicals (EPRRs) on the solid particles of biochars (Yang et al. 2016). Also, some researchers believe that the functional groups on carbon-based materials can catalyze the degradation of organic pollutants (Jiang et al. 2016). Thus, there is no consistent conclusion on the degradation mechanism of organic pollutants by biochars. After biochars modification, many researchers showed that the organic pollutant removal in water was significantly increased. But few studies confirmed whether the organic pollutants were degraded, but empirically attributed the removal efficiency to adsorption process.

Biochars characterization during their aging

Prior to using biochars for a particular application, it is essential to quantitatively describe biochar properties to ensure optimal outcomes for agricultural or environmental activities. Various methods are used for this purpose. Depending on different research purposes and resource availability, the characterization methods for biochars can be divided into four categories, including bulk elemental composition analysis, physicochemical analysis, surface analysis, and structure/molecular identification (Igalavithana et al. 2018). Bulk elemental composition analysis is usually conducted with an element analyzer to quantify O, C, N and S contents and inductively coupled plasma mass spectroscopy (ICP-MS) to measure metals. Physicochemical properties include pH, electrical conductivity, redox property, density and exchangeable cations. Surface properties are analyzed using microscopy technique, such as scanning electron microscopy and transmission electron microscopy, to identify surface morphology and BET analyzer to obtain the overall pore volume and specific surface area. The surface functional groups can be qualitatively identified using Fourier transform infrared spectroscopy (FT-IR) and X-ray photoelectron spectroscopy (XPS). Böehm titration is a quantitative method to identify the amount of acidic and basic functional groups on biochars’ surface. The structure and molecular analysis often used thermal gravimetric analysis (TGA) for thermal stability, Raman for aromaticity or nuclear magnetic resonance spectroscopy (NMR) for other specific structures.

Biochar’s characterization is generally motivated by the following two main purposes: (1) to understand the physical and chemical properties of biochars and evaluate if they fulfill the designed purpose; (2) to examine the eco-toxicity of biochars and estimate their environmental risk. The International Biochar Initiative (IBI) has clearly and strictly advised that biochars have to be characterized before utilizing as a soil amendment. However, many characterization methods only evaluated the pristine biochars before application in the field.

It is not practicable to directly measure biochar properties after their application in the environment, simply because it is extremely difficult to separate biochars from natural particles, especially carbonaceous particles. The previous mentioned characterization methods are usually qualitative and have large variation, which may not be sensitive enough to distinguish the biochar’s property change after aging. The promising techniques for a long-term observation of biochar aging usually involves isotope labeling and 13C NMR spectroscopy (de la Rosa et al. 2018; Dong et al. 2017). 14C-labeled biochars are used to trace its mineralization rate (decomposition to CO2) (Kuzyakov et al. 2014), but this method is relatively expensive and time consuming. 13C NMR spectroscopy is effective to quantify the biochars aromaticity and aromatic condensation degree (Leng et al. 2019), but large variation existed among different NMR instruments. The transformation of chemical compounds including lipids, polysaccharides and benzene poly-carboxylic acid (BPCA) can be determined by GC–MS. BPCAs are widely used as bio-markers to quantify black carbon, and recently extended to biochar and soil organic matter characterization (Chang et al. 2018a, b). For BPCA analysis, the complex polycyclic structures are chemically broken into small individual BPCA molecules, which can be quantified using gas chromatography–mass spectrometer (GC–MS). The ratio of individual BPCA monomers to the overall BPCA and the ratio of the overall BPCA to total organic carbon imply the degree of the aromaticity (Singh et al. 2017). However, this method cannot release all the stable C in the carbon matrix (maximum estimation of ~ 25% of the total stable carbon) and as the pyrolysis temperature increases, the percentage will be reduced (Singh et al. 2017). Therefore, in order to identify the biochar properties after their aging, different technologies need to combined, and a set of robust analysis on biochar stability will be needed in the future.

Eco-toxicological concerns of EngBCs

In spite of the positive effect of biochars on soil, some researchers started to pay attention to their eco-toxicology. The eco-toxicology of various types of EngBCs should be evaluated differently depending on the modification method, especially when considering their different aging processes. It is known that biochars often contain numerous contaminants (such as polycyclic aromatic hydrocarbons (PAHs) and heavy metals) (Oleszczuk et al. 2013; Wang et al. 2017a). These contaminants are strongly bound with biochars and may not pose great threat to living organisms initially. However, the modification or aging of biochars may release these strongly bound native contaminants, which subsequently may result in greater mobility and toxicity of the contaminants (Fig. 3).

Fig. 3

The eco-toxicity concerns of EngBCs. PAHs, heavy metals, and the imbedded nanoparticles in EngBCs may be released during biochar aging and pose an environmental risk. EngBCs modified by transition metals may make health risk even more complicated by inducing persistent free radical formation

Dissolvable contaminants

Many researches confirmed the presence of PAHs (Wang et al. 2017a) and trace heavy metal in EngBCs (Stefaniuk et al. 2016). The total concentration of PAHs is mainly dependent on the pyrolysis methods, temperature, and time as well as the biomass source (Hale et al. 2012; Oleszczuk et al. 2013; Stefaniuk et al. 2016; Wang et al. 2017a). The total PAHs concentrations in biochars were detected from below the detection limit to dozens of thousands μg/kg, some of which are bioavailable (Oleszczuk et al. 2013; Wang et al. 2017a). The dependence of PAH yield on the pyrolysis temperature still does not have unified conclusion yet. Some studies suggested that high pyrolysis temperature lead to high PAHs yield, but Devi and Saroha (2014) found that biochars produced at 400 °C and 500 °C have the highest PAHs yield compared with other temperatures. Feedstock has been found to be one of the primary factors that could influence the present of PAHs in biochars. Lignin-rich feedstock was approved to produce biochars with lower PAHs concentrations than pectin and cellulose dominated feedstock (Sharma and Hajaligol, 2003).

It is noteworthy that some researches already showed a clear evidence that biochar application led to an accumulation of PAHs in plants and pose a risk to human health (Wang et al. 2018b). Because of this health risk, many efforts were devoted to understand or reduce the total and bioavailable PAHs inside the biochars (Sharma and Hajaligol 2003; Zielinska and Oleszczuk, 2016). On the other hand, after biochars were applied in strongly PAHs contaminated soil, they often act as a sink to adsorb PAHs rather than as a PAHs releasing source (Mayer et al. 2016). The application of EngBCs can even further reduce the eco-toxicity of PAHs contaminated soils and the reduction efficiency depends on the method of modification. For example, Koltowski et al. (2017) showed that biochars activated by microwaves can reduce the eco-toxicity of the contaminated soil more effectively than the non-activated biochars, whereas biochars activated by CO2 often aggravated the situation of soil eco-toxicity. Steam activated biochars showed the ability of reducing freely dissolved and bioaccessible fraction of PAHs in soil (Koltowski et al. 2016). Unfortunately, most of the above-mentioned studies only focused on the short-term effects of newly prepared biochars, not on their toxicity changes after a long-term exposure in the environment. Only until recently, Oleszczuk and Koltowski (2018) published a paper to discuss the total and freely dissolved PAHs after various aging process to fill this research gap. Their research points out that in the case of three different aging processes (chemical, freeze–thaw cycles and biological aging), the contents of total and freely dissolved PAHs in all aged biochars were reduced. Biological aging led to the greatest reduction efficiency. Another study also confirmed that PAHs concentration in biochars decreased after 4 years field aging process (Sigmund et al. 2017). The authors referred to the sorption by the surrounding soil of low molecular weight PAHs. Both researches indicated that PAHs in biochars may be released during the biochar aging process. Therefore, we can speculate that the EngBCs, which is used as a PAHs sink, could become a PAHs source after being aged in the environment and the environmental risk during the whole biochars lifetime should be carefully evaluated. It should be noted that these studies are mostly focused on the pristine biochars. How the biochar modification alters PAHs speciation and releasing in EngBCs still remains to be systematically investigated.

Heavy metals in biochars mainly originate from two different sources. One is the high content of heavy metals in the feedstock, such as wastewater sludge. The other source is from the modification process of biochars. In order to improve the adsorption efficiency and catalysis activity of biochars, heavy metals, such as Fe, Zn, and Mg, were added purposely in the preparation process. It is generally believed that these EngBCs do not have the risk of releasing heavy metals, because biochars have high specific surface area and abundant surface functional groups. However, this assumption is based on freshly made EngBCs (Rechberger et al. 2019). After aged for a longtime, the surface functional groups of EngBCs would be eventually reduced in biotic and abiotic process, which might lead to the heavy metal release into the environment.

For some particle-embedded EngBCs, such as magnetic EngBCs with embedded Fe3O4 particles, different unique processes should be considered. These hybrid particles may be less strongly combined with biochar particles. The aging processes, including organic matter coating and thus dispersion, freeze–thaw cycles, as well as expansion and contraction because of the temperature change, will result in the detachment of these particles. These released nanoparticles may also have potential risks to the ecosystems, as summarized by several review papers (Devin et al. 2017; Mezni et al. 2018; Rana and Kalaichelvan, 2013).

Persistent free radicals

The persistent free radicals (PFRs) are the organic free radicals stabilized on or inside the biochars, which can induce the formation of reactive oxygen species (ROS) to effectively degrade organic containments (Fang et al. 2017; Pan et al. 2019; Ruan et al. 2019). This high reactive species may also pose a health risk due to its potential to cause biological damage, such as cardiopulmonary disease and cancer. Early researches believed that ROS generated by PFRs is the key reason responsible for their toxicity. If this is the case, it will be very easy to measure and evaluate the health risk in biochars driven system, and to eliminate the risk by adding ROS scavengers. However, recent studies showed that the activity of PFRs was not passivated in the presence of excessive ROS scavengers (Yang et al. 2016). This result suggested that not only the ROS generated by PFRs, but also the particles itself may have very high potential health risk (Pan et al. 2019).

EngBCs modified by transition metals may make this health risk even more complicated. Previous researches showed that transition metals are the key factor to enable the formation of PFRs, which were generated during the reactions between transition metals and organic chemical degradation byproducts (Lomnicki et al. 2008). When heavy metals were added to enhance biochars properties, PFRs are very likely to form during the pyrolysis process. However, no systematic description is reported on PFR formation in EngBCs, such as the intensity, type and activity of the PFRs generated inside the EngBCs. Therefore, future studies to evaluate the application of EngBCs should also consider the presence of the PFRs and properly address their environmental health risk.

Conclusions and perspectives

While the stability of the pristine biochars in soil environment has been investigated over the last decade, the stability of EngBCs after their application in the environment has not been paid enough of attention. Extrapolating the previous results on biochars aging, we can speculate that the EngBCs in soil will be slowly oxidized and the oxygen-containing functional groups will increase accordingly. The interaction between EngBCs and contaminants or soil components during EngBCs aging process also was not discussed efficiently, let alone their eco-toxicology. Therefore, to ensure the safe and sustainable application of EngBCs, the following questions have to be properly addressed:

  1. 1.

    Although EngBCs are considered as relatively stable carbonaceous materials, aging process will surely alter their properties after being applied in the environment. Extended study is needed to examine the aging processes and mechanisms as affected by the modification method, environmental chemical conditions as well as biological behaviors.

  2. 2.

    During the aging process of EngBCs, their interactions with contaminants or natural soil components will change continually. Thus the functions of EngBCs should be viewed dynamically. The effective time period of EngBCs should also be evaluated accordingly.

  3. 3.

    The aging process alters the physical, chemical and biological properties of EngBCs, which will alter the potential toxicity of biochars. The potential negative impacts of EngBCs should be evaluated carefully considering the aging processes, such as the release of toxic chemicals or particles.


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This research was supported by National Natural Science Foundation of China (41725016, U1602231, 41807377 and 41673098), a joint fund between NSFC-NCN (4181101459), and Yunnan Provincial Scientific Innovation Team of Soil Environment and Ecological Safety (2019HC008).

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Duan, W., Oleszczuk, P., Pan, B. et al. Environmental behavior of engineered biochars and their aging processes in soil. Biochar 1, 339–351 (2019). https://doi.org/10.1007/s42773-019-00030-5

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  • Biochar modification
  • Stability
  • Toxicity
  • Soil components
  • Organic contaminants