Next Article in Journal
Comparative Evaluation of Technologies at a Heavy Metal Contaminated Site: The Role of Feasibility Studies
Next Article in Special Issue
Slow Pyrolysis of Quercus cerris Cork: Characterization of Biochars and Pyrolysis Volatiles
Previous Article in Journal
Socio-Economic Impact of the Imposed Lockdowns in Food Chains: A Case Study in Cyprus
Previous Article in Special Issue
Torrefaction Upgrading of Heterogenous Wastes Containing Cork and Chlorinated Polymers
Font Type:
Arial Georgia Verdana
Font Size:
Aa Aa Aa
Line Spacing:
Column Width:

Effects of Biochar on the C Use Efficiency of Soil Microbial Communities: Components and Mechanisms

Department of Civil Engineering, Architecture, Environmental and Mathematics (DICATAM), University of Brescia, Via Branze 43, 25123 Brescia, Italy
Department of Agronomy, Food, Natural Resources, Animals and Environment (DAFNAE), University of Padua, Viale dell’Università 16, 35020 Legnaro, Italy
Author to whom correspondence should be addressed.
Environments 2022, 9(11), 138;
Original submission received: 3 August 2022 / Revised: 12 October 2022 / Accepted: 15 October 2022 / Published: 4 November 2022
(This article belongs to the Special Issue Preparation and Application of Biochar)


Biochar production and incorporation into soil is gaining momentum as a sustainable strategy for climate change mitigation, supported by ever increasing reports of significant carbon (C) sequestration in soil and reduction in greenhouse gas (GHG) emissions from the amended soils. With the progression in biochar testing and use, there is also emerging evidence that biochar induces C sequestration in soil, and that it may not be solely caused by its inherent chemical stability, but also by the complex microbially driven processes and an increase in C use efficiency (CUE) through soil microbial metabolism. This evidence contradicts the current paradigm that sees the microbial CUE decrease during the degradation of recalcitrant material due to thermodynamic constraints, as observed only in several short-term and pilot-scale trials. As the CUE in soil results from interactions between several abiotic and biotic factors, in this paper we examine the link between the biochar properties, soil physico-chemical properties and microbial physiology to explain the CUE increase reported for biochar-amended soils. Based on the large body of physico-chemical literature, and on the high functional diversity and metabolic flexibility of soil microbial communities, we hypothesize that the long-term stabilization of biochar-borne C in the soil systems is not only controlled by its inherent recalcitrance, but also by the cooperative actions of improved soil status and increased microbial CUE. Given that the current knowledge on this specific aspect is still poor, in this feature paper we summarize the state of knowledge and examine the potential impact of biochar on some factors contributing to the whole-soil CUE. We conclude that, beside its inherent recalcitrance, biochar weathering and oxidation in soil create physical and chemical conditions that can potentially increase the microbial CUE. While these processes stabilize the microbial processed C in soil and increase soil fertility, more data from long-term field trials are needed to model the relationship between the CUE and the MRT of biochar-borne C. Based on our hypotheses and relying upon analysis of the available literature, we also suggest possible research approaches that may contribute to filling the gaps in the current knowledge on the topic.

1. Biochar: Chemical Stability Influences the Carbon Use Efficiency (CUE) in Soil

Biochar is the product of thermochemical decomposition of biomass at temperatures ranging from 300 to 1000 °C in the absence of oxygen, either by pyrolysis or gasification, and can be used for extracting energy from biomass [1]. These processes decompose the organic biopolymers into small gaseous molecules such as methane (CH4), hydrogen (H2), carbon monoxide (CO), and carbon dioxide (CO2), condensable vapours (tars and oils), and a solid phase as a by-product termed biochar. Biochar production is maximized by pyrolysis at low temperatures (e.g., 300–450 °C) and slow heating rates and residence times (hours), whereas gasification is maximized at high temperatures (≥800 °C) and rapid heating rates (seconds), with the biochar production rates of slow and fast pyrolysis being 20–30% and 4–8%, respectively. Moreover, the biochar qualities resulting from pyrolysis and gasification are also different; for the aims of this paper, it is worthwhile to underline that slow pyrolysis results in less recalcitrant biochar with high porosity as determined by the Brunauer-Emmett-Teller (BET) method and coarse texture. Conversely, gasification results in highly recalcitrant biochar with low BET values and finer texture. In fact, gasification pyrolysis requires finely ground feed of specific biomasses with a narrow range of initial moisture content, a precisely controlled temperature and effective cooling of the vapour phase in order to achieve efficient recovery of the oily product. Therefore, it is a less flexible process mainly aimed at energy production, producing highly inert biochar with a sandy-dusty texture, whereas slow pyrolysis produces manipulable biochar which can be incorporated into soil. For this reason, biochar from fast pyrolysis has been used less than slow pyrolysis for the purposes of achieving agronomic and environmental goals such as climate change mitigation [2].
Based on the potential long mean residence time (MRT) of biochar in soil, the so-called charcoal vision [3] is nowadays considered as a strategy to offset a significant share of anthropogenic C emissions through C sequestration in soil. In addition to its climate mitigation potential, incorporation of biochar into soil also brings significant improvement in soil fertility as it enhances the cation exchange capacity (CEC), neutralizes acidic pH values, changes the soil color and increases the soil thermal capacity, increases water retention, and immobilizes inorganic and organic pollutants [4]. Concurrence of these factors explains the frequently reported increases in crop yields [5,6,7], especially in highly degraded soils. Owing to its soil-improving and beneficial effects, biochar has been admitted into the new EU Regulation on Fertilizers (EU 1009/2019) under the product of pyrolysis component material category (CMC). The ever-increasing mass of information about the positive effects of biochar parallels to the historical evidence of the transformation of Amazonian dark earth soils (Terra Preta dos Indios), which maintain significantly higher pH values and fertility compared with the surrounding soils even after millennia or centuries from the charcoal burial [8].
Biochar-induced positive effects on soils and related ecosystem services are related to its stability. Initially, potential biochar stability in the environment was estimated through the process parameters such as pyrolysis temperature and feedstock types, but such parameters are currently no longer considered valid. Conversely, chemical properties such as O/C molar ratio [9], the H/Corg ratio value [10], and both the H/Corg and O/Corg ratio values are nowadays considered better descriptors of biochar stability [11,12]. In fact, these parameters reflect the labile C/recalcitrant C ratio value, and are well correlated to the results of thermal/chemical oxidation resistance tests. The molecular structure of biochar is predominantly aromatic, and its inherent stability depends upon this level molecular arrangement. Aromatic substances can form either amorphous phases, in which the aromatic substances are randomly organized, or crystalline phases, in which the aromatic structures form ordered condensed sheets, as observed in the pyrolysis and pyrogasification processes, respectively. Because more aromatic and more condensed molecular structures are supposed to be more resistant to chemical and biological degradation [2], the evaluation of both aromaticity and degree of aromatic condensation, for example by Nuclear Magnetic Resonance (NMR) spectroscopy, is increasingly used as a chemical indicator of biochar stability [13]. Based on the microbial degradative mechanisms we describe in the following paragraphs, in our opinion the quantification of benzene polycarboxylic acids (BPCA) and molar H/C ratio, indicating the degree of molecular unsaturation and H deficiency [14], can be reliable indicators of potential microbial attack on biochar C. The analysis of hydroxylic and carbonylic functional groups, nonaromatic C branches, or N-containing functional groups using Fourier transform infrared spectroscopy (FTIR) [15], or wet chemistry oxidation assays [15,16], provides additional and complementary indications of biochar stability towards potential microbial attack.
Biochar persistence in soil can be also evaluated by batch incubation and C mineralization experiments, in which both biotic and abiotic degradations occur. As compared to the chemical tests, that provide indirect stability indications, the biochar stability, assessed by soil incubation tests, is increasingly adopted to obtain direct evidence of persistence of different biochar in soil. Significant correlations between biochar C stability, C mineralization rate, the degree of aromaticity and molecular condensation of biochar have been reported [10]. A meta-analysis of biochar incubation experiment reported that the MRT of biochar, added at a rate of 3% and having an initial recalcitrant C proportion of 97%, has an MRT in soil spanning from 100–1000 years [17].
Despite the large number of reports on the topic, in our opinion, the following question is still open: is biochar chemical stability sufficient to forecast its persistence in microbiologically soils? This question led us to a twofold reflection on the microbial C use efficiency (CUE) in biochar-amended soils: a chemical and a microbiological one. From the chemical point of view, microorganisms have a nominal cellular C oxidation state close to the average sugar’s C oxidation state. For this reason, they mainly derive energy and produce cell components from anabolic pathways fueled by more reduced C substrates. Classic microbiological experiments have shown lower CUE of single isolates grown on substrates with higher C oxidization degrees (e.g., mono- or bicarboxylic acids), and peak CUE values for microorganisms grown on C substrates with an oxidation state (e.g., sugars) close to that of microbial cells [18]. However, for the highly diverse and metabolically flexible soil microbial communities, the relationship between the oxidation state of C substrates and CUE may depend on more factors that the relative C oxidation states of substrates [19]. In terms of microbial physiology, soil microbial communities have been hypothesized to be characterized by the ‘metabolic infallibility’, i.e., soil microorganisms have the potential capability to use any known substance as an energy source, even one of high molecular complexity and recalcitrance, especially if limiting environmental conditions are alleviated. Though this concept has been debated for decades [20], the microbial metabolic infallibility principle has received confirmation from an increasing body of evidence [21]. This concept is highly relevant for biochar-amended soils, as it implies that under suitable conditions microorganisms can oxidize any substance which is theoretically capable of being oxidized [22]. Since its earlier formulation of the metabolic infallibility concept, environmental microbiologists have demonstrated that microbial degradation processes have two main components, the genetic-based metabolic potential of microorganisms and the chemical recalcitrance of the substrate molecules [20]. The aim of this review is to answer to this question by critically binding the knowledge on the biochar properties to the main soil properties, the fundamental microbial physiology and microbial ecology concepts.
Thermodynamically, soils are open systems that exchange energy and matter with other natural systems, constituted by reactive solid phases and host to large and highly diverse microbial communities characterized by the highest known diversity and metabolic flexibility among the terrestrial ecosystems [23]. The result of soil microbial oxidative activity is that biochar aging in soil turns its properties from inert hydrophobic and neutrally charged matter to increasingly hydrophilic negatively charged solid phases [24], thus increasing its potential interactions with soil microorganisms. We hypothesize that the reported capacity of biochar in stimulating soil microbial activity could be ascribed to the fact that microorganisms possess metabolic degradative potential because black C is ubiquitous to terrestrial and aquatic environments due to natural or human-induced fire and anthropogenic environmental enrichment with organic xenobiotics [25]. Nevertheless, better understanding of the chemical and biochemical mechanisms requires the evaluation of the changes in main soil properties induced by the biochar, treated in Section 3, the alteration of the C:N:P stoichiometry, treated in details in Section 4, and the composition of the soil microbial community, treated in Section 5.

2. Microbial CUE: Definition for Soil Systems and Changes in Biochar-Amended Soils

The simplest definition of carbon use efficiency (CUE) of microorganisms is the ratio between microbial biomass C production and C taken up by microorganisms (1)
microbial CUE = µ/Cuptake
where µ is the microbial biomass growth and C uptake is the sum of C immobilized by microbial growth, plus respired C. The CUE of soil microorganisms represents the efficiency of microbial biomass production resulting from the mineralization of the soil organic matter (SOM). Determination of the microbial CUE can be used to predict the energy flow among trophic levels in different ecosystems, across different soil management and different soil types [26] and represents an integrated parameter that can predict the microbial C turnover in the biogeochemical cycle of the SOM [27]. This highly meaningful soil ecological parameter can be determined with different methods, based on the measurement of soil respiration and variations in the soil microbial biomass C pool. Early CUE estimations were based on in vitro analysis of microbial biomass yields upon exposure to selected C sources. This ‘black box’ approach mainly takes into account the specific C respiration (qCO2) during the microbial growth (µ) phase (i.e., qCO2/µ), which is an expression of the maintenance energy of microbial cells exposed to different C sources [28], not an expression of the microbial CUE.
The CUE values, calculated with by the mass-based approach as in Equation (1), only results in CUE values > 0, with values typically in the order of 0.60 [29,30]. Conversely, theoretically the CUE can also have negative values if C assimilation is lower than the respired C, or when microbial biomass declines due to mortality [31].
As mentioned above, incorporation of biochar also leads to changes in soil properties that affect the CUE. However, from a literature search using SCOPUS as a database in spite of the ca. 11,500 papers retrieved using ‘soil’ and ‘biochar’ as search criteria and ca. 1300 paper matching ‘soil’ and ‘CUE’ as search criteria, only 12 papers matched the keywords ‘soil’ and ‘biochar’ and ‘CUE’ as search criteria. This literature search highlights the current poor knowledge on the CUE as a key process that controls the biochar C stability in soil.
In structured, open, and highly biodiverse systems like soils, the CUE is not only a function of C availability, but also of the microbial community composition [32]. Moreover, CUE measurements integrate several biotic and abiotic factors such as SOM quality and solubility in the soil solution, availability of other macronutrients than C, soil temperature and moisture levels [33,34]. This only to make mention the major factors, the effects of which will be discussed in detail below. For soil microorganisms, this means that the SOM quality is not the only factor that determines the maximal CUE. Rather, other environmental properties altered by biochar amendment also influence microbial activity and proliferation. In a greenhouse trial, the soil organic C (SOC) mineralization and the microbial community of a soil amended with 20 and 40 t ha−1 of biochar Zhang et al. [35] showed the increase in SOC and recalcitrant C, the reduction in SOC mineralization and the temperature sensitivity (Q10) value [35], confirming previous reports on this topic [36]. Concerning the importance of organic substrates quality, microbial CUE is higher when the C availability is limited but, thermodynamically, decomposition of recalcitrant C should reduce the microbial CUE due to the high metabolic costs of the oxidative catabolic steps [37].
Changes in quality and rates of organic substrates upon undergoing biochar aging could induce changes in soil microbial communities, either as individuals or microbial physiological groups, potentially altering the kinetics of C assimilation pathways [38,39]. In enzyme-mediated uptake pathways, C uptake rates are expressed as a saturating function of substrate concentration, well represented by a Michaelis–Menten model [40]. Theoretically, C uptake for soil microorganisms energetically covers the metabolic costs of the substrate uptake mechanisms [41] and the optimal substrate saturation should not vary. Under these conditions, low C availability should decouple catabolic and anabolic pathways and globally lead to low CUE of soil microbial communities [26]. At a microbial community level, the CUE varies depending on major ecological constraints, namely C sources type and bioavailability, relative abundance of different microbial physiological groups, soil moisture level and temperature, and generally decreases with soil depth [42,43]. At an individual level, the CUE values vary in different physiological states of microorganisms, being lower during the C assimilation stage and higher during the exponential growth phase [44]. This is because, a short time after exposure to energy substrates, microorganisms respire C and thus synthesize extracellular enzymes for SOM decomposition and cell membrane transport proteins. Whereas, in the logarithmic growth phase, the assimilated C is maximally allocated into a new biomass [40,45].
Moreover, in terms of metabolic flexibility, under nutrient-limiting conditions and/or other environmental constraints, soil microorganisms may activate degradation but not growth metabolic pathways. These cometabolic pathways involved in biochar modification due to the release of nonspecific enzymes are capable of chemically modify recalcitrant organics that, though not directly capable of supporting microbial proliferation, can be oxidized or reduced in the presence of SOM-derived substrates providing metabolic energy [46]. For example, cometabolism has been demonstrated as a mechanism of transformation of polycyclic aromatic hydrocarbons (PAHs) and other organic xenobiotics [47]. Based on the above considerations, we hypothesize that in the presence of biochar the microbial physiological groups capable of utilizing biochar-borne recalcitrant substances are positively selected, and their capability to use the biochar-derived substances and native similar organic substances may explain the higher microbial CUE reported for biochar-amended than nonamended soils.

3. Biochar-Induced Temperature and Moisture Effects on Soil CUE

Temperature and moisture are environmental factors that influence microbial CUE [48]. Incorporation of biochar into soil significantly changes soil color and water retention properties. Reduced reflectance of biochar-amended soils increases soil temperature due to changes of soil albedo [49]. By definition, the albedo values range from 0 to 1, and the range is between 0.1–0.2 for dark soils and between 0.4–0.5 for light-colored soils [50], with either geographical, daily and seasonal variations. Beside the incident radiation, soil thermal capacity is also increased by soil moisture content, SOM content, and particle size distribution [50]. The biochar-related increase in water retention has a cooperative effect with albedo, especially in sandy soils that drain and dry out faster than clayey soils, as the specific heat of water in moist soil is ca. 5 times higher than in dry soil [51]. With few exceptions, long-term field trials show that biochar increases the water retention, and higher water retention in dry periods may reduce the accumulation of osmolytes [52] that generally increase the C:N ratio values of the microbial biomass and the apparent CUE values [53,54]. Microbial CUE is generally reduced upon an increase in soil temperature [26], mainly due to the faster acceleration of microbial respiration processes than microbial growth responses [48]. Although at the community level, microorganisms adapt to increased temperatures in terms of species composition, the link between biochar-induced changes in microbial community composition and thermal adaptation of microbial communities still needs to be assessed, and reliable information can be obtained only from the analysis of soils from long-term field trials.
Higher soil temperature can reduce the activation energy of SOM decomposition [34,55], though SOM activation energy depends on its molecular complexity [56] and increases upon the number of enzymatic steps required for substrate modification and decomposition [57]. Under different temperatures, changes of CUE values in the presence of molecular complex substrates are generally less pronounced than those recorded during the decomposition of low molecular weight organic compounds (LMWOCs) [58], and biochar generally has lower temperature sensitivity than native SOM [57]. In this regard, higher mean temperature and more constant moisture levels may facilitate microbial oxidative enzymes synthesis and release, and in cooperation with the nonspecific enzymatic mechanisms, they may reduce the activation energy for microbial respiration [59]. Overall, these mechanisms can make microbial oxidation of biochar in soil less dependent on its inherent thermal stability and more dependent on microbial enzymatic activity. However, to our knowledge, no experiments aiming at determining the changes in the activation energy of SOM of biochar-amended soils have been conducted.
Biochar stability also depends on soil texture and its eventual association with soil minerals, or minerals deliberately associated with biomass feedstocks for producing different biochar types [60]. Because microbial activity occurs in hot spots mainly present in soil aggregates, the diffusion of biochar-borne substrates into the aggregates can possibly determine an ‘abiotic gate’ limiting their decomposition, at least shortly after soil amendment [61]. Limitations in the use of insoluble pools of biochar-borne C by soil microorganisms, due to physical and chemical protection in soil aggregates, can be alleviated when the decomposition process initiated by the synthesis of enzymes and sustained by the subsequent formation of more hydrophilic and soluble C pools, increase the microbial accessibility to organic substrates. Physico-chemical mechanisms occurring in soil such as sorption, diffusion, and occlusion into aggregates exert additional control on biochar stability in soil, because they increase the microbial energy investment in enzyme synthesis for C acquisition. These mechanisms, that depend on the properties of the soil solid phases and the soil structure complexity, along with surface hydrophobicity and molecular recalcitrance of biochar, control the biochar C transfer from stable to more labile pools. In our opinion, the biochar decomposition rate in soil is co-controlled by the diffusion of LMWOCs from the biochar particles surface towards soil aggregates driven by moisture (Figure 1), and from their sorption onto organic and inorganic soil solid phases [62].
These considerations let us also to propose that physical aspects are important for future formulation of biochar-based fertilizers, which are supposed to be more efficient for crop nutrition [63] but may not contribute to the maintenance of a porous soil structure as compared to organic amendments such as compost [64].
Overall, we support the idea that, by increasing soil thermal capacity and water retention, biochar confers resilience to soils allowing more constant microbial activity, attenuating the seasonal variations or eventual environmental drought stressful conditions, and enhancing the C stabilization through the ‘Microbial Carbon Pump’ mechanism [65,66]. Such fluctuations are particularly broad in agricultural soils, where biochar can be incorporated, because microbial CUE also decreases with soil depth due to energetic limitations [43], for example due to unfavourable C:N and C:P ratios. Biochar incorporation in the deeper horizons for an effective C storage may also increase the CUE in the long term owing to the release of LMWOCs (Figure 1) and let microorganisms living deeper soil layers reach CUE values similar to those of microorganisms of the surface horizons. To our knowledge, there is no information on the effects of the biochar on the CUE in deeper soil horizons, and if proven, such a change may become an additional factor stabilizing organic C in the subsoil.

4. Biochar Effects on C Availability, Nutrient Stoichiometry and CUE in Soil

Several studies have demonstrated that biochar amendment improves the soil habitat for microorganisms, and that the positive effects on microbial activity are mainly related to biochar and properties of soil [6,67,68]. These results provide a ground base for a critical evaluation of the reported increase in microbial CUE in biochar-amended soils. In a laboratory incubation experiment, the higher CUE of biochar-amended soils was attributed to the biochar protection of hydrophilic SOM towards microbial mineralization, whereas no such effect was observed for hydrophobic SOM [69]. Soil amendment with aged biochar increased microbial CUE and significantly decreased the biomass turnover time compared to the amendment with fresh biochar. However, such effects were observed in a sandy, as opposed to in a clayey, soil [70,71]. This result could be explained by the fact that clay soils better stabilize SOM than sandy soils [72]. A support to this hypothesis is a study of the short-term effects of biochar by pyrolyzed maize straw pyrolysis at different temperatures showed that biochar changed the microbial CUE, inducing the release of more biochar-borne LMWOCs [73]. These results indicated that, while biochar stimulates the mineralization of LMWOCs, it may reduce the decomposition of native SOM and thus increase the potential C storage in soil (Figure 1). The two above-mentioned studies confirmed that changes in pH value, availability of primary nutrients and SOM quality, induced by fresh or aged biochar, were all covariates that influenced microbial CUE in biochar-amended soils [74].
However, though higher CUE is increasingly reported for biochar-amended than unamended soils, considering only the C availability, higher microbial biomass or activity in biochar-amended soils may not necessarily result in higher CUE.

4.1. Biochar Changes the Soil N and P Contents and the C:N:P Stoichiometry

A large number of studies have demonstrated that, though microbial metabolism in soil is limited by C availability, microbial CUE can be also limited by the availability of other nutrients, primarily N and P [75,76]. The soil C:N:P stoichiometric ratio only significantly varies in different soil types or after major changes in soil use and management. Additionally, the physiological effects of nutrient stoichiometry, as the main forces that drive the biochar transformation and regulate SOM stability in soil, have received experimental confirmation. In highly degraded soils, P availability is considered the primary limitation on microbial growth [77]. In severely degraded soils, colimitation of microbial activity by N and P have been also reported [78], and this may explain why in degraded soils the potential of biochar to increase soil microbial biomass and stimulate microbial activity is stronger than in more healthy soils.
Soil C:N:P stoichiometry affects microbial CUE in soil because, while C substrates are oxidized for energy production in the catabolic pathways, N is needed for protein synthesis and P is required for nucleotide synthesis and ribosome activity [77,79]. For these reasons, nutrient stoichiometry is also considered in SOM turnover models e.g., [26,80]. Due to the high condensation degree of C structures, biochar incorporation into soil increases the C:N:P ratio of amended soils. However, the literature data show that the biochar amendment of soils increases total concentration and availability of N, P and K in the long-term [5,81]. Biochar enriches in phosphate P during carbonization [82] or retains P on its surface due to the presence of sorption sites [83,84]. Higher P availability in biochar-amended soils may be a major factor in improving microbial CUE because the microbial biomass C:P ratio varies more than the biomass C:N ratio. The C:N:P stoichiometry in soils also influences the nutrient availability in soil as soil microorganisms tend to regulate the SOM stoichiometry by releasing nutrients absorbed in excess [85]. Therefore, C:N:P stoichiometry is a main parameter conditioning the activity of soil microorganisms during the SOM decomposition and nutrient mineralization, and significant variations of soil stoichiometry can induce changes in soil microbial diversity and soil microbial activity, because it can make one of the nutrients limiting [86,87]. For these reasons, P availability influences microbial CUE, and the use of decomposition models indicates that the CUE should decrease upon an increase in the C:P ratio of the SOM [88]. However, a decrease in microbial CUE concerns forest litter, which typically has C:P ratios in the order of 1000 on a mass basis [89], whereas in biochar-amended soils the C:P ratio values are in the order of 100, i.e., in the typical stoichiometric range of the microbial biomass C:P ratio [88]. Moreover, soils amended with biochar increase the abundance of P-solubilizing bacteria in a forest soil [90], and only in soils where biochar reduces P availability may it impair microbial activity of amended soils [91]. Accordingly, CUE in biochar-amended soils may have an opposite trend as compared to that of native SOM, as for the latter the C:N ratio moves below the N immobilization–mineralization threshold as C becomes increasingly recalcitrant upon decomposition, thus limiting microbial activity. Differently, biochar aging in soil increases C lability and the C:N ratio may not decrease, leading to a faster microbial growth and increase in the CUE to a plateau level, with timing and increments depending on the biochar quality, soil type and soil use and management (Figure 2).
Overall, we hypothesize that the few reports on high CUE in biochar-amended soils could be due to the reaching of a C:N:P ratio that favors metabolically efficient microbial communities.

4.2. Determination of Soil Microbial Biomass Homeostatic C:N:P Ratios in Biochar-Amended Soils

To date, studies of biochar-amended soils have seldom focused on the release of nutrient limitations to microorganisms, which is crucial to understand the adaptation of microbial metabolism in soil. Relying on the concept that microorganisms have an homeostatic C:N:P composition, altered availabilities of N and P, induced by biochar amendment in soil, have the potential to favour microbial communities that can initially cope with the stoichiometrically altered substrates [80]. Such changes could be estimated by the measurement of nutrient assimilation efficiency by soil microbial communities [92] or maximum microbial growth efficiency [93]. Such estimations can be complicated in soil because it is a structured environment and biochar movement can cause the creation of hot spots of nutrients, or transfer nutrients along the soil profile and also between distinct microbial populations. Decomposition studies showed that SOM residues with high C:N ratio (e.g., conifers wood) indicate that microbial CUE decreases upon an increasing in the C:N ratio, something which normally leads to N immobilization [94]. Limited use of insoluble pools of biochar-borne C by soil microorganisms due to physical and chemical protection into soil aggregates, as discussed in Section 3, can be overcome when the decomposition process is initiated by the synthesis of enzymes and the subsequent formation of more hydrophilic and soluble C pools that can be used by microorganisms.

5. Microbial Community Composition and CUE in Biochar-Amended Soils

While at a single-species level CUE depends on cellular metabolism and on the production of cell structures and storage compounds, at the community level variation in CUE is related to the soil microbial communities composition, changes of dominant microbial species and relative abundance of microbial groups, exemplified by the fungi:bacteria biomass ratio. Differently from natural organic polymers that are mineralized after depolymerization and solubilization processes [95], biochar-C can be used by soil microorganisms after oxidative and hydrolytic reactions that should be theoretically restricted to a subset microbial physiological group capable of performing the highly demanding reactions (Figure 1). Such decomposer microorganisms, characterized by low CUE values, should be adapted to low substrate concentrations and be slow growing [95]. Based on these assumptions, initial steps of biochar use should be characterized by a drop of CUE related to biochar recalcitrance and to the relative abundance of degradative keystone microbial species (Figure 2).
Biochar incorporation into soil significantly increased both Gram-negative and Gram-positive bacteria and decreased the fungi:bacteria ratio [70]. Fungi, which have higher C molar ratio than bacteria, could be mainly involved in the ‘char dominated phase’ because theoretically they have a higher C-demand and slower growth rate than bacteria [96]. Moreover, an increase in fungal activity during early biochar decomposition may justify a high CUE due to their large C:N and C:P ratios, which imply larger C demands [97]. However, different fungal phylogenetic groups have different C:N:P ratio values [98]. In particular, the C:N:P ratio for free living fungi varies broadly with lower C:P and N:P ratios found in Ascomycota as compared to Basidiomycota fungi, whereas lower C:N ratios and higher N:P ratios were found in mycorrhizae than in saprotrophic fungi [98]. Interestingly, fungal C:N:P: ratios were correlated with environmental parameters such as moisture and temperature, and all parameters varied by biochar amendment of soil (cfr. Section 2). A field experiment showed that after 3 years biochar amendment in an alkaline soil cultivated with soybean, Ascomycota significantly decreased and Mortierellomycota significantly increased, whereas no significant changes in the bacterial community was observed [99]. Overall, changes in bacterial and fungal communities were associated with soil properties such as SOC and TN [99].
Bacteria also exhibit highly flexible metabolism under various environmental conditions e.g., [100], and may contribute to biochar oxidation.
Later, when a significant share of biochar has been weathered, the CUE values of the soil microbial community should approach a plateau level (Figure 2). While the length of the char-dominated phase mainly depends on the biochar type, soil properties and environmental conditions, the new CUE level should mainly depend on the microbial community composition. Changes in soil microbial community structure have been reported [101,102], but the effect of biochar on total microbial diversity and functional gene abundance and expression are still controversial. Changes in the soil microbial community after biochar amendment were observed, and an increase in the abundance of soil microbial communities in a dose-related relationship has been reported [103]. An increase in fungal abundance after biochar application in an alkaline soil was observed, highlighting that the effect of biochar is related to different types of biochar and soil properties [99]. The addition of biochar (10–15% w:w) induced modification of the microbial community structure, and a significant increase in the richness and diversity index of total microbes [64]. Differently, no significant effect on microbial community structures and extracellular enzyme activities was observed in a short-term experiment, in which the biochar application (22 t ha−1) was compared with manure amendment [104]. This result could be attributed to the relatively low application rate of biochar or to site-specific environmental conditions.
Current information on microbial groups potentially active in biochar-amended soils is inconclusive. Future studies based on stable isotope probes may clarify the main biosynthetic pathways and improve the models applicable to microbial communities of biochar-amended soils, especially considering long-term trials. For example, laboratory incubation studies, based on pulsed additions of large amounts of LMWOCs, may allow us to quantify the maintenance energy and growth respiration of specific substrates released during biochar aging. These substrates could be representative of in situ microbial metabolism activated by LMWOCs release, and also analyze the influence of environmental factors such as soil temperature and moisture level.

5.1. Polycyclic Aromatic Hydrocarbon Degraders: The Chemical Gate Operators

Specialized polycyclic aromatic hydrocarbon (PAH) degraders may be among the first microbial groups attacking the biochar C. In a pot experiment, the biochar amendment increased the PAH-ring hydroxylating dioxygenase (PAH-RHD) genes coding for the enzyme involved in the initial step of the microbial degradation of PAHs [105]. In the same soils, the increase in K, P and N availability in biochar-amended soil was also highlighted [105]. These results from short-term laboratory or greenhouse scale experiments are in agreement with evidence from the Amazonian dark earth (ADE), anthropogenic dark soil horizons formed by the deposit of organic materials that date back from ca 8000 years before the Columbus age [106]. Such deposits turned infertile Oxisols into highly fertile hotspots, mainly due to the stabilization of organic matter as a result of incompletely combusted biomass, similar to the biochar [107]. An ADE soil had similar bacterial community compositions, as well as a significantly greater species richness than a pristine forest soil [108]. The relative abundance and diversity of the biphenyl dioxygenase (bph) gene involved in aromatic hydrocarbon degradation in an ADE and a nonanthropogenic-adjacent soil, and of the bacterial genera harbouring the bph genessuch as Streptomyces, Sphingomonas, Rhodococcus, Mycobacterium, Conexibacter and Burkholderia with known aromatic hydrocarbon degradation capacity, were more abundant and diverse in the ADE than non-ADE soil [109]. The bph and thp genes-coding dioxygenase and aromatic ring-hydroxylating dioxygenases have been identified in genomic and plasmidic DNA of aromatic hydrocarbon-degrading bacteria also detected ADE soils [110]. Overall, evidence from short- and long-term studies demonstrates that biochar oxidation is energetically permissible, but that it likely occurs in microbial communities hosting physiological groups of microorganisms that use catabolic pathways to obtain energy. The PAH-degrading microorganisms could constitute one of the operational groups responsible for the metabolic infallibility of soil microbial communities.

5.2. Enzyme Activity: The Toolbox

Changes in enzyme activity profile from biopolymer hydrolase to condensed C oxidase allow us to hypothesize that, in biochar-amended soils, the respiration rate should be lower than thermal oxidation. Activation energy for the mineralization of recalcitrant biochar matter by microbial communities, which is greater than the electromotive force of the cell oxidative metabolism, is lowered due to the multiple enzymatic steps needed to transform the organic C into CO2 [111]. In fact, the relation between temperature dependency for chemical and microbial oxidation of SOM was studied, and the release of specific oxidative enzymes was considered as a factor that makes the microbial decomposition capacity relatively independent of the SOM recalcitrance [112]. This hypothesis should be tested in different soils, especially from long-term trials, because under natural conditions the temperature sensitivity of microbial respiration is far higher than chemical oxidization, and other main soil factors (e.g., C:N ratio, pH value, salinity) influence microbial activity, not only SOM chemical oxidability. If proven operating, this mechanism could explain the higher CO2 release generally reported for biochar-amended soils compared to nonamended soils.
An alternative hypothesis that may explain higher abundance of dioxygenase, phenol oxidase enzyme activities and relevant encoding genes could be the need to mitigate the impact of toxic compounds. In fact, increased expression of genes and enzyme activities involved in the degradation of PAH has been reported for contaminated soils [113]. In this light, other well-characterized genes involved in the oxidation and detoxification of recalcitrant C compounds, such as the naphthalene dioxygenase genes (nahA), may be involved in the biochar modification, as such genes are ubiquitous in soil. In particular, the larger abundance of nahA genes in the rhizosphere, i.e., in the soil portion modified by the plant roots, may explain the faster biochar C turnover in soils under permanent grass, which can be considered as an entirely rhizospheric soil. A rapid increase in the relative abundance of dioxygenase genes in PAH polluted soils amended with biochar has been frequently reported e.g., [101,114,115], although no effects have been also reported e.g., [116,117]. While biochar used in agriculture may contain low levels PAHs, amendment generally does not induce soil toxicity to microorganisms [118]. In addition to oxidase, hydrolytic enzymes also play fundamental role in CUE because catalyze reactions necessary for soil organic C decomposition and mineralization of N, P and S [119]. Recent research has shown that also the profile of the hydrolytic enzyme activity changes in long-term biochar-amended soils [81]. Soil enzymatic activity is an integral component of the whole CUE as it regulates the nutrient availability to soil microorganisms [120], and such changes in substrate utilization can potentially affect the enzymatic stoichiometry, and in turn the biogeochemical cycle of nutrients in soil [80,121]. According to the economic theory of soil enzymatic activity [40,122], microorganisms synthesize and release extracellular enzymes to decompose SOM into LMWOCs [123]. The release of LMWOCs in biochar-amended soils can also be enhanced by the increase in arylesterase enzyme activity that cleaves carboxylic branches from aromatic compounds. This change is important because it was estimated that LMWOCs microbial uptake and mineralization accounts for a significant share (up to 30%) of soil CO2 emission from soil [124]. Thus, the LMWOCs pool should be included in models of SOC dynamics. Shifts in soil enzyme activities have been reported from both short-term laboratory/greenhouse experiments [125] and long-term field trials [126]. Because it can be considered as the last enzymic step prior to utilization of LMWOCs as energy sources by microorganisms actively degrading biochar. The ratios of oxidative and carboxyl esterase to hydrolytic enzyme activity as proxies for SOM were proposed [127]. Here, we propose the following enzyme activity ratio as a proxy for describing the potential biochar degradability in soil, as a potential C source to soil microorganisms (Equation (2)), with the highest value indicating faster potential microbial C turnover and microbial CUE in biochar-amended soils.
Biochar C use = f [(oxidase + carboxyl-esterase)/hydrolase]
As mentioned above, microbial CUE is assumed to be lower during the use of recalcitrant substrates due to the high metabolic energy investment in enzymes and cell uptake systems, but in the soil environment the CUE may non increase owing to the accumulation of extracellular oxidase and hydrolase enzymes. This hypothesis converges with those deriving from the degradation of fresh biomass polymers, as the available C molecular can be used to offset a significant part of the energy dissipation due to the initial substrate recalcitrance (Figure 2). This change in CUE can be realized by adaptation of microbial groups capable of using biochar-derived C substrates after the first oxidative reactions that enrich the condensed C in carbonyl, carboxyl- and phenolic functional groups, open the aromatic rings and cleave the lateral branches, releasing LMWOCs (Figure 2). Release of oxidase and hydrolase enzymes coupled with higher mean temperature, moisture, and N and P availability, can improve C acquisition and maintain the C:N:P stoichiometric balance of adapted microbial consortia. In these conditions, microbial taxa with high CUE reduce energy investments for enzyme synthesis, especially for catalysing the first steps of biochar degradation. Possibly, the greater CUE in biochar-amended soils is reached after the biochar-borne C structure modification releases sufficient LMWOCs, and less degradative enzymes need to be produced per unit of biomass formed [128]. This CUE adaptation stage at microbial community level may be responsible for the C storage in biochar-amended soils due to sorption of SOM of microbial origin [129], and microbial colonization of biochar habitats.
In amended soils the chemical modification of biochar induces modification of the community metabolism and the stimulation of a potential metabolic mechanisms controlling microbial nutrition and microbial energy (Figure 3).
The above hypothesized CUE adaptation based on the utilization of low molecular weight carbonyl- and acyl- compounds requires the oxidation and modification of the poly- and branched aromatic molecules, formed during the biomass pyrolysis. This preliminary step can be conducted by various bacterial and fungal groups peroxidases releasing extracellular peroxidases, phenol oxidases and laccases enzymes [130,131]. It is important to underline that, independently on the involved microbial species, the mentioned oxidative enzymes are those involved in the degradation of lignin and other biopolymers. The convergence of molecular structure during the biochar oxidation and SOM decomposition will result in an optimization of the enzymatic efforts of soil microorganisms which will benefit from additional C sources from the same enzyme activities (Figure 2).

5.3. P and N Mineralizing Microorganisms: The Helpers

The above discussed importance of nutrient stoichiometry should be reflected in changes of microbial physiological groups involved in N and P turnover. In particular, the release of abundant oxidized LMWOCs from biochar decomposition should influence the diversity and relative abundance of functional genes coding for the enzymes involved in the N and P cycles, but evidence on the relevant microbial groups is recent and still conflicting. In a short-term field trial an increased total soil organic C, the decrease in extractable nitrate pool and the reduction in gross rates of organic N transformation were reported, whereas the increase in ammonia-oxidizer bacterial and archaeal populations that paralleled a two-fold increase in the gross nitrification rate were observed [132]. Soil urease activity paralleled to the increased abundance of the PAH-degrading bacteria in biochar-amended soils was reported [101]. No or variable effects of biochar addition on ammonia-oxidizing Archaea and ammonia oxidizer and nitrite-reducing bacteria were reported [133,134]. The bacterial community involved in the Dissimilatory Nitrate Reduction to Ammonium (DNRA) was found not significantly affected by amendment with biochar of neutral and alkaline soils [135], whereas biochar amendment was reported that it could increase both potential nitrification and denitrification activity in acidic soils as the biochar alkalinity could create a favorable environment for the activity of nitrifying and denitrifying communities [136]. In a pot experiment biochar amendment of an acidic soil stimulated both nitrification and denitrification processes, reduced N2O emissions which paralleled the increased nosZ gene expression, indicating an increased tendency of reduction in N2O to N2 [121]. Differently, biochar addition reduced the relative abundance of ammonia oxidizers and the nitrification potential of soil [137]. The addition of leaf and woodchip biochar showed the increase in the abundance of P-solubilizing bacteria and the diversity of soil bacterial community in the forest soil [90]. These inconsistent results could be due to differences in plant species, biochar characteristics, soil types and biochar application periods [138,139]. Biochar plus chemical fertilizers increased the abundance of bacteria in the rhizosphere [140]. It is important to underline that most of the studies on changes in diversity and abundance N cycling microbial populations in biochar-amended soils have been conducted in short-term microcosm experiments e.g., [141,142,143,144,145,146].

5.4. Sulfur Reducing Bacteria: The Stone Guest?

When compared to N, P and K, little information on total S and sulphate in biochar-amended soils is currently available. Sulfur is essential for plant growth in lower amounts as compared to N, P and K, and an increase due to biochar dose may release large amounts of sulfates, triggering sulfate-reducing bacteria. Such bacteria are known to degrade aromatic hydrocarbons with various degree of condensation and molecular complexity, and under optimal conditions they can mineralize completely aromatic C up to CO2. Sulfate-reducing bacteria have been detected in hydrocarbon-contaminated environments, and their degradative activities result in release LMWOCs (e.g., acetate, propionate) which are mineralized by the same bacteria or other bacterial groups [147]. Microbial acetogenic activity of sulfate-reducing bacteria was enhanced by biochar added to a wastewater treatment sludge [148]. Higher emissions of acetylene, propene, and C4 aldehydes from long-term biochar-amended soils as compared to nonamended soils were observed, but no information on the source of such LMWOCs was investigated [126]. Therefore, the potential role of sulfate reducing bacteria in biochar-amended soils remains obscure. Sulfate-reducing microorganisms can occur in association with fermentative, syntrophic microorganisms, such as methanogenic bacteria with other metabolic pathways involving Fe (III), Mn (IV), NO3-N reduction [149]. Archaea can also degrade aromatic hydrocarbons into acetate followed by acetate syntrophic conversion into CH4 and CO2 [150]. Soil amended with biochar could increase the number of sulfur-mobilizing bacteria, which may explain the enhancement of sulfur cycle functions [151]. The combined use of ferrous sulfate and biochar significantly reduced emissions of NH3, N2O, CO2, and CH4 and increased of the richness of bacterial communities during composting of pig manure and straw mixture more than single ferrous sulfate and biochar [152]. Biochar acting as an electron shuttle in the microbial reduction in Fe (III) in ferrihydrite has been reported [153]. In our opinion, the biochar-mediated S and Fe transformations and the involved microbial groups, deserve higher attention in future studies to better understand the biochar influence on nutrients biogeochemistry with the complex mechanisms illustrated in Figure 4, not only on microbial CUE.
Summarizing, in biochar-amended soils after the char-dominated phase initiated by specialized microbial physiological groups, two major and related features can be anticipated: (i) an abundance of microbial species can process the biochar-borne LMWOCs and (ii) microbial CUE of the soil tends to a new stable value, typical for each specific soil (Figure 2). Duration of the different processes, depicted in Figure 2, will depend on the biochar type, pedoclimatic conditions, and soil management, but in our opinion reaching new equilibria may require decades. The increased CUE values reported shortly after fresh additions of biochar to soil may result from the use of labile C associated with biochar, formed onto the biochar surface during the condensation of syngas during the cooling phase of the biochar preparation.

6. Conclusions and Research Needs

We support the idea that biochar amendment can improve the fertility of degraded and intensively cultivated soils and can lead to significant C sequestration. Nevertheless, we hypothesize that, in the long term, stabilization of biochar C in soil is not solely due to biochar’s inherent chemical stability, but also to the stabilization of microbial processed C released by the oxidation and cleavage of biochar C. This process should be initiated by soil microbial physiological groups possessing the suite of functional genes for biochar oxidation. We hypothesize that microbial activation is facilitated by higher thermal capacity, greater water retention, and larger C N and P availability induced by biochar addition. Biochar porosity also increase the biological space and create favorable conditions for microbial activity, stabilization of extracellular oxidase and hydrolase enzymes and adaption of microbial groups with relatively high CUE values. To test these hypotheses, we suggest the following research objectives:
Determine the theoretical CUE of biochar-amended soils and under different environmental conditions, soil types and management. Determination of the thermodynamic maxima of different biochar types can also allow the improved determination the limiting or unlocking effects of increased nutrient availability;
Models operating on finer time scales (days to seasons) should consider the effects of changing SOM molecular composition, multi-element stoichiometric constraints, and microbial community physiology. In addition, environmental drivers should be implemented on data from long-term field trials to predict the CUE of biochar-amended soils under different management. The effects of larger N and P availability as driving forces, and changes in enzymatic activity, should be tested, particularly in arable soils amended with biochar and under chemical fertilization, in order to determine the stability and MRT of the biochar-borne C. An interesting technique that could be tested to assess the extent of biochar stability in soil could be the reverse stable isotope labelling [154];
Perform hypothesis-driven metagenomic research to reveal presence, activity, and evolution of microbial metabolic pathways in long-term biochar-amended soils. Complementary proteomic and metabolomic studies may help to elucidate the hypothesized pathway of release of LMWOCs from biochar, and better estimate the C partition in CO2, microbial and SOM pools, also analyzing the changes of the 13C signature of biochar and SOM;
Adopt imaginative direct observation approaches to observe the surface of weathered biochar, also extracted from amended soils, to describe the formation of the ‘charrosphere’ and its relations with biochar biodegradation and the release of LMWOCs into the soil solution.

Author Contributions

Conceptualization, L.G. and G.R.; writing—original draft preparation, L.G. and G.R.; writing—review and editing, L.G. and G.R.; visualization, L.G. and G.R.; All authors have read and agreed to the published version of the manuscript.


This research received no external funding.


The authors are grateful to three anonymous Reviewers and the Journal’s Editor for their highly constructive comments.

Conflicts of Interest

The authors declare no conflict of interest.


  1. Anca-Couce, A. Reaction mechanisms and multi-scale modelling of lignocellulosic biomass pyrolysis. Prog. Energy Combust. Sci. 2016, 53, 41–79. [Google Scholar] [CrossRef]
  2. Lehmann, J.; Joseph, S. Biochar for Environmental Management: Science and Technology, 2nd ed.; Earthscan: London, UK, 2015. [Google Scholar]
  3. Laird, A.D. The charcoal vision: A win-win-win scenario for simultaneously producing bioenergy, permanently sequestering carbon, while improving soil and water quality. Agron. J. 2008, 100, 178–181. [Google Scholar] [CrossRef]
  4. Lehmann, J.; da Silva, J.P.; Steiner, C.; Nehls, T.; Zech, W.; Glaser, B. Nutrient availability and leaching in an archaeological anthrosol and a ferralsol of the central Amazon Basin: Fertilizer, manure and charcoal amendments. Plant Soil 2003, 249, 343–357. [Google Scholar] [CrossRef]
  5. Jeffery, S.; Verheijen, F.G.A.; van der Velde, M.; Bastos, A.C. A quantitative review of the effects of biochar application to soils on crop productivity using meta-analysis. Agric. Ecosyst. Environ. 2011, 144, 175–187. [Google Scholar] [CrossRef]
  6. Wong, J.; Hashimoto, Y.; Huang, L.; Rinklebe, J.; Chang, S.; Bolan, N.; Wang, H.; Ok, Y. Response of microbial communities to biochar-amended soils: A critical review. Biochar 2019, 1, 3–22. [Google Scholar]
  7. Jaiswal, A.; Alkan, N.; Elad, Y.; Sela, N.; Philosoph, A.M.; Graber, E.R.; Frenkel, O. Molecular insights into biochar-mediated plant growth promotion and systemic resistance in tomato against Fusarium crown and root rot disease. Sci. Rep. 2020, 10, 13934. [Google Scholar] [CrossRef]
  8. Sambroek, W.G. Amazonian soils: A reconnaissance of the soils of the Brazilian Amazon Valley. Ph.D. Thesis, Wageningen University, Wageningen, The Netherlands, February 1966. [Google Scholar]
  9. Spokas, K.A. Review of the stability of biochar in soils: Predictability of O: C molar ratios. Carbon Manag. 2010, 1, 289–303. [Google Scholar] [CrossRef][Green Version]
  10. Budai, A.; Zimmerman, A.R.; Cowie, A.L.; Webber, J.B.W.; Singh, B.P.; Glaser, B.; Masiello, C.A.; Andersson, D.; Shields, F.; Lehmann, J.; et al. Biochar Carbon Stability Test Method: An Assessment of Methods to Determine Biochar Carbon Stability; Technical Report for International Biochar Initiative: Canandaigua, NY, USA, September 2013. [Google Scholar]
  11. ASTM D1762-84; Standard Test Method for Chemical Analysis of Wood Charcoal. ASTM International: West Conshohocken, PA, USA, 2021.
  12. EBC. European Biochar Foundation—European Biochar Certificate—Guidelines for a Sustainable Production of Biochar; European Biochar Foundation (EBC): Arbaz, Switzerland, 2012. [Google Scholar]
  13. Wiedemeier, D.B.; Brodowski, S.; Wiesenberg, G.L.B. Pyrogenic molecular markers: Linking PAH with BPCA analysis. Chemosphere 2015, 119, 432–437. [Google Scholar] [CrossRef][Green Version]
  14. McBeath, A.V.; Wurster, C.M.; Bird, M.I. Influence of feedstock properties and pyrolysis conditions on biochar carbon stability as determined by hydrogen pyrolysis. Biomass Bioenergy 2015, 73, 155–173. [Google Scholar] [CrossRef]
  15. Cross, A.; Sohi, S.P. A method for screening the relative long-term stability of biochar. GCB Bioenergy 2013, 5, 215–220. [Google Scholar] [CrossRef]
  16. Keiluweit, M.; Nico, P.S.; Johnson, M.G.; Kleber, M. Dynamic molecular structure of plant biomass-derived black carbon (biochar). Environ. Sci. Technol. 2010, 44, 1247–1253. [Google Scholar] [CrossRef]
  17. Wang, J.; Xiong, Z.; Kuzyakov, Y. Biochar stability in soil: Meta-analysis of decomposition and priming effects. GCB Bioenergy 2016, 8, 512–523. [Google Scholar] [CrossRef][Green Version]
  18. Gommers, P.J.F.; van Schie, B.J.; van Dijken, J.P.; Kuenen, J.G. Biochemical limits to microbial growth yields: An analysis of mixed substrate utilization. Biotechnol. Bioeng. 1998, 82, 86–94. [Google Scholar]
  19. Gunina, A.; Smith, A.R.; Kuzyakov, Y.; Jones, D.L. Microbial uptake and utilization of low molecular weight organic substrates in soil depend on carbon oxidation state. Biogeochemistry 2017, 133, 89–100. [Google Scholar] [CrossRef][Green Version]
  20. Alexander, M. Biochemical ecology of soil microorganisms. Annu. Rev. Microbiol. 1964, 18, 217–250. [Google Scholar] [CrossRef]
  21. O’Malley, M.A.; Walsh, D.A. Rethinking microbial infallibility in the metagenomics era. FEMS Microbiol. Ecol. 2021, 97, fiab092. [Google Scholar] [CrossRef]
  22. Gale, E.F. The Chemical Activities of Bacteria, 3rd ed.; Academic Press Inc.: Cambridge, MA, USA, 1951. [Google Scholar]
  23. Tiedje, J.M.; Cho, J.C.; Murray, A.; Treves, D.; Xia, B.; Zhou, J. Soil teeming with life: New frontiers for soil science. In Sustainable management of Soil Organic Matter; Rees, R.M., Ball, B.C., Campbell, C.D., Watson, C.A., Eds.; CABI: Wallingford, UK, 2001. [Google Scholar]
  24. Tomczyk, A.; Sokołowska, Z.; Boguta, P. Biochar physicochemical properties: Pyrolysis temperature and feedstock kind effects. Rev. Environ. Sci. Bio/Technol. 2020, 19, 191–215. [Google Scholar] [CrossRef][Green Version]
  25. Schmidt, M.W.I.; Noack, A.G. Black carbon in soils and sediments: Analysis, distribution, implications, and current challenges. Glob. Biogeochem. Cycles 2000, 14, 777–793. [Google Scholar] [CrossRef]
  26. Manzoni, S.; Taylor, P.; Richter, A.; Porporato, A.; Ågren, G.I. Soil carbon and nitrogen mineralization: Theory and models across scales. New Phytol. 2012, 196, 79–91. [Google Scholar] [CrossRef]
  27. Six, J.; Frey, S.D.; Thiet, R.K.; Batten, K.M. Bacterial and fungal contribution to carbon sequestration in agroecosystems. Soil Sci. Soc. Am. J. 2006, 70, 555–569. [Google Scholar] [CrossRef]
  28. Pirt, S.J. Principles of Microbe and Cell Cultivation; Wiley: Hoboken, NJ, USA, 1985. [Google Scholar]
  29. Payne, W.J.; Wiebe, W.J. Growth yield and efficiency in chemosynthetic microorganisms. Ann. Rev. Microbiol. 1978, 32, 155–183. [Google Scholar] [CrossRef]
  30. Narang, A. The steady states of microbial growth on mixtures of substitutable substrates in a chemostat. J. Theor. 1998, 190, 241–261. [Google Scholar] [CrossRef]
  31. Wang, G.S.; Post, W.M. A theoretical reassessment of microbial maintenance and implications for microbial ecology modeling. FEMS Microbiol. Ecol. 2012, 81, 610–617. [Google Scholar] [CrossRef][Green Version]
  32. Bölscher, T.; Wadsö, L.; Börjesson, G.; Herrmann, A.M. Differences in substrate use efficiency: Impacts of microbial community composition, land use management, and substrate complexity. Biol. Fertil. Soils 2016, 52, 547–559. [Google Scholar] [CrossRef]
  33. Tiemann, L.K.; Billings, S.A. Changes in variability of soil moisture alter microbial community C and N resource use. Soil Biol. Biochem. 2011, 43, 1837–1847. [Google Scholar] [CrossRef]
  34. Wagai, R.; Kishimoto-Mo, A.W.; Yonemura, S.; Shirato, Y.; Hiradate, S.; Yagasaki, Y. Linking temperature sensitivity of soil organic matter decomposition to its molecular structure, accessibility, and microbial physiology. Glob. Change Biol. 2013, 19, 1114–1125. [Google Scholar] [CrossRef]
  35. Zhang, X.; Zhang, Q.; Zhan, L.; Xu, X.; Bi, R.; Xiong, Z. Biochar addition stabilized soil carbon sequestration by reducing temperature sensitivity of mineralization and altering the microbial community in a greenhouse vegetable field. J. Environ. Manag. 2022, 313, 114972. [Google Scholar] [CrossRef]
  36. Zhou, H.; Zhang, D.; Wang, P.; Liu, X.; Cheng, K.; Li, L.; Zheng, J.; Zhang, X.; Zheng, J.; Crowley, D.; et al. Changes in microbial biomass and the metabolic quotient with biochar addition to agricultural soils: A Meta-analysis. Agric. Ecosyst. Environ. 2017, 239, 80–89. [Google Scholar] [CrossRef][Green Version]
  37. Blagodatskaya, E.; Kuzyakov, Y. Mechanisms of real and apparent priming effects and their dependence on soil microbial biomass and community structure: Critical review. Biol. Fertil. Soils 2008, 45, 115–131. [Google Scholar] [CrossRef]
  38. Button, D.K. Nutrient-limited microbial-growth kinetics—Overview and recent advances. Antonie Van Leeuwenhoek 1993, 63, 225–235. [Google Scholar] [CrossRef]
  39. Hobbie, J.E.; Hobbie, E.A. Amino acid cycling in planktonic and soil microbes studied with radioisotopes: Measured amino acids in soil do not reflect bioavailability. Biogeochemistry 2012, 107, 339–360. [Google Scholar] [CrossRef]
  40. Schimel, J.P.; Weintraub, M.N. The implications of exoenzyme activity on microbial carbon and nitrogen limitation in soil: A theoretical model. Soil Biol. Biochem. 2003, 35, 549–563. [Google Scholar] [CrossRef]
  41. Sinsabaugh, R.L.; Follstad Shah, J.J. Integrating resource utilization and temperature in metabolic scaling of riverine bacterial production. Ecology 2010, 91, 1455–1465. [Google Scholar] [CrossRef]
  42. Dijkstra, P.; Thomas, S.C.; Heinrich, P.L.; Koch, G.W.; Schwartz, E.; Hungate, B.A. Effect of temperature on metabolic activity of intact microbial communities: Evidence for altered metabolic pathway activity but not for increased maintenance respiration and reduced carbon use efficiency. Soil Biol. Biochem. 2011, 43, 2023–2031. [Google Scholar] [CrossRef]
  43. Spohn, M.; Klaus, K.; Wanek, W.; Richter, A. Microbial carbon use efficiency and biomass turnover times depending on soil depth—implications for carbon cycling. Soil Biol. Biochem. 2016, 96, 74–81. [Google Scholar] [CrossRef]
  44. Robinson, C. Heterotrophic bacterial respiration. In Microbial Ecology of the Oceans; Kirchman, D.L., Ed.; Wiley: Hoboken, NJ, USA, 2008; pp. 299–334. [Google Scholar]
  45. Vetter, Y.A.; Deming, J.W.; Jumars, P.A.; Krieger-Grockett, B.B. A predictive model of bacterial foraging by means of freely released extracellular enzymes. Microb. Ecol. 1998, 36, 75–92. [Google Scholar] [CrossRef]
  46. Hazen, T.C. Cometabolic Bioremediation. In Handbook of Hydrocarbon and Lipid Microbiology; Timmis, K.N., Ed.; Springer: Berlin/Heidelberg, Germany, 2010; pp. 2505–2514. [Google Scholar]
  47. Mahendra, S.; Alvarez-Cohen, L. Kinetics of 1,4-dioxane biodegradation by monooxygenase-expressing bacteria. Environ. Sci. Technol. 2006, 40, 5435–5442. [Google Scholar] [CrossRef]
  48. Frey, S.D.; Lee, J.; Melillo, J.M.; Six, J. The temperature response of soil microbial efficiency and its feedback to climate. Nat. Clim. Change 2013, 3, 395–398. [Google Scholar]
  49. Genesio, L.; Miglietta, F.; Lugato, E.; Baronti, S.; Pieri, M.; Vaccari, F.P. Surface albedo following biochar application in durum wheat. Environ. Res. Lett. 2012, 70, 14025. [Google Scholar] [CrossRef]
  50. Baumgardner, M.F.; Sylva, L.F.; Biehl, L.L.; Stoner, E.R. Reflectance Properties of Soils. Adv. Agron. 1985, 38, 1–44. [Google Scholar]
  51. Brutsaert, W. Evaporation into the Atmosphere; D. Reidel Publishing Company: Dordrecht, Holland, 1982. [Google Scholar]
  52. Boot, C.M.; Schaeffer, S.M.; Schimel, J.P. Static osmolyte concentrations in microbial biomass during seasonal drought in a California grassland. Soil Biol. Biochem. 2013, 57, 356–361. [Google Scholar] [CrossRef]
  53. Uhlırova, E.; Elhottova, D.; Trıska, J.; Santruckova, H. Physiology and microbial community structure in soil at extreme water content. Folia Microbiol. 2005, 50, 161–166. [Google Scholar] [CrossRef]
  54. Herron, P.M.; Stark, J.M.; Holt, C.; Hooker, T.; Cardon, Z.G. Microbial growth efficiencies across a soil moisture gradient assessed using 13C-acetic acid vapor and 15N-ammonia gas. Soil Biol. Biochem. 2009, 41, 1262–1269. [Google Scholar] [CrossRef]
  55. von Stockar, U.; Marison, I.W. The definition of energetic growth efficiencies for aerobic and anerobic microbial growth and their determination by calorimetry and other means. Thermochim. Acta 1993, 229, 157–172. [Google Scholar] [CrossRef]
  56. Davidson, E.A.; Janssens, I.A. Temperature sensitivity of soil carbon decomposition and feedbacks to climate change. Nature 2006, 440, 165–173. [Google Scholar] [CrossRef][Green Version]
  57. Bosatta, E.; Ågren, G.I. Soil organic matter quality interpreted thermodynamically. Soil Biol. Biochem. 1999, 31, 1889–1891. [Google Scholar] [CrossRef]
  58. Öquist, M.G.; Erhagen, B.; Haei, M.; Sparrman, T.; Ilstedt, U.; Schleuche, J.; Nilsson, M.B. The effect of temperature and substrate quality on the carbon use efficiency of saprotrophic decomposition. Plant Soil 2017, 414, 113–125. [Google Scholar] [CrossRef][Green Version]
  59. Ågren, G.I.; Wetterstedt, J.A.M. What determines the temperature response of soil organic matter decomposition? Soil Biol. Biochem. 2007, 39, 1794–1798. [Google Scholar] [CrossRef][Green Version]
  60. Woolf, D.; Amonette, J.E.; Street-Perrott, F.A.; Lehmann, J.; Joseph, S. Sustainable biochar to mitigate global climate change. Nat. Commun. 2010, 1, 1–9. [Google Scholar] [CrossRef][Green Version]
  61. Resat, H.; Bailey, V.; McCue, L.A.; Konopka, A. Modeling microbial dynamics in environments: Growth on soil carbon sources. Microb 2012, 63, 883–897. [Google Scholar] [CrossRef]
  62. Müller, T.; Höper, H. Soil organic matter turnover as a function of the soil clay content: Consequences for model applications. Soil Biol. Biochem. 2004, 36, 877–888. [Google Scholar] [CrossRef]
  63. Zheng, J.; Han, J.; Liu, Z.; Xia, W.; Zhang, X.; Li, L.; Liu, X.; Bian, R.; Cheng, K.; Zheng, J.; et al. Biochar compound fertilizer increases nitrogen productivity and economic benefits but decreases carbon emission of maize production. Agric. Ecosyst. Environ. 2017, 241, 70–78. [Google Scholar] [CrossRef]
  64. Chen, K.; Peng, J.; Li, J.; Yang, Q.; Zhan, X.; Liu, N.; Han, X. Stabilization of soil aggregate and organic matter under the application of three organic resources and biochar-based compound fertilizer. J. Soils Sediments 2020, 20, 3633–3643. [Google Scholar] [CrossRef]
  65. Miltner, A.; Bombach, P.; Schmidt-Brücken, B.; Kästner, M. SOM genesis: Microbial biomass as a significant source. Biogeochemistry 2012, 111, 45–55. [Google Scholar] [CrossRef]
  66. Liang, C.; Schimel, J.P.; Jastrow, J.D. The importance of anabolism in microbial control over soil carbon storage. Nat. Microbiol. 2017, 2, 17105. [Google Scholar] [CrossRef]
  67. Mia, S.; Dijkstra, F.A.; Singh, B. Long-term aging of biochar: A molecular understanding with agricultural and environmental implications. Adv. Agron. 2017, 141, 1–51. [Google Scholar]
  68. Yang, K.; Jiang, Y.; Wang, J.; Cai, X.; Wen, Z.; Qiu, Z.; Qiao, G. Tobacco straw biochar improved the growth of Chinese cherry (Prunus pseudocerasus) via altering plant physiology and shifting the rhizosphere bacterial community. Sci. Hortic. 2022, 303, 111244. [Google Scholar] [CrossRef]
  69. Zheng, X.; Li, X.; Singh, B.P.; Wei, L.; Huang, L.; Huang, Y.; Huang, Q.; Chen, X.; Su, Y.; Liu, Z.; et al. Biochar protects hydrophilic dissolved organic matter against mineralization and enhances its microbial carbon use efficiency. Sci. Total Environ. 2021, 795, 148793. [Google Scholar] [CrossRef]
  70. Liu, Z.; Zhu, M.; Wang, J.; Liu, X.; Guo, W.; Zheng, J.; Bian, R.; Wang, G.; Zhang, X.; Cheng, K.; et al. The responses of soil organic carbon mineralization and microbial communities to fresh and aged biochar soil amendments. GCB Bioenergy 2019, 11, 1408–1420. [Google Scholar] [CrossRef][Green Version]
  71. Pei, J.; Li, J.; Mia, S.; Singh, B.; Wu, J.; Dijkstra, F.A. Biochar aging increased microbial carbon use efficiency but decreased biomass turnover time. Geoderma 2021, 382, 114710. [Google Scholar] [CrossRef]
  72. Singh, B.; Sarkar, S.; Churchman, J.; Bolan, N.; Mandal, S.; Menon, M.; Purakayastha, T.J.; Beerling, D.J. Stabilization of soil organic carbon as influenced by clay mineralogy. Adv. Agron. 2018, 148, 33–84. [Google Scholar]
  73. Cheng, H.; Hill, P.W.; Bastami, M.S.; Jones, D.L. Biochar stimulates the decomposition of simple organic matter and suppresses the decomposition of complex organic matter in a sandy loam soil. GCB Bioenergy 2017, 9, 1110–1121. [Google Scholar] [CrossRef]
  74. Silva-Sánchez, A.; Soares, M.; Rousk, J. Testing the dependence of microbial growth and carbon use efficiency on nitrogen availability, pH, and organic matter quality. Soil Biol. Biochem. 2019, 134, 25–35. [Google Scholar] [CrossRef]
  75. Cleveland, C.C.; Liptzin, D. C:N:P stoichiometry in soil: Is there a “Redfield ratio” for the microbial biomass? Biogeochemistry 2007, 85, 235–252. [Google Scholar] [CrossRef]
  76. Sinsabaugh, R.L.; Manzoni, S.; Moorhead, D.L.; Richter, A. Carbon use efficiency of microbial communities: Stoichiometry, methodology and modelling. Ecol. Lett. 2013, 16, 930–939. [Google Scholar] [CrossRef]
  77. Waring, B.G.; Weintraub, S.R.; Sinsabaugh, R.L. Ecoenzymatic stoichiometry of microbial nutrient acquisition in tropical soils. Biogeochemistry 2013, 117, 101–113. [Google Scholar] [CrossRef]
  78. Xiao, L.; Liu, G.; Li, P.; Li, Q.; Xue, S. Ecoenzymatic stoichiometry and microbial nutrient limitation during secondary succession of natural grassland on the Loess Plateau, China. Soil Tillage Res. 2020, 200, 104605. [Google Scholar] [CrossRef]
  79. Brown, J.H.; Gillooly, J.F.; Allen, A.P.; Savage, V.M.; West, G.B. Toward a metabolic theory of ecology. Ecology 2004, 85, 1771–1789. [Google Scholar] [CrossRef]
  80. Sinsabaugh, R.L.; Follstad Shah, J.J. Ecoenzymatic stoichiometry and ecological theory. Annu. Rev. Ecol. Evol. Syst. 2012, 43, 313–342. [Google Scholar] [CrossRef][Green Version]
  81. Lasota, J.; Babiak, T.; Błońsk, E. C:N:P stoichiometry associated with biochar in forest soils at historical charcoal production sites in Poland. Geoderma. Reg. 2022, 28, e00482. [Google Scholar] [CrossRef]
  82. de la Rosa, J.M.; Paneque, M.; Miller, A.Z.; Knicker, H. Relating physical and chemical properties of four different biochars and their application rate to biomass production of Lolium perenne on a Calcic Cambisol during a pot experiment of 79 days. Sci. Total Environ. 2014, 499, 175–184. [Google Scholar] [CrossRef][Green Version]
  83. Mia, S.; Dijkstra, F.A.; Singh, B. Aging induced changes in biochar’s functionality and adsorption behavior for phosphate and ammonium. Environ. Sci. Technol. 2017, 51, 8359–8367. [Google Scholar] [CrossRef]
  84. Hong, C.; Lu, S. Does biochar affect the availability and chemical fractionation of phosphate in soils? Environ. Sci. Pollut. Res. 2018, 25, 8725–8734. [Google Scholar] [CrossRef]
  85. Vitousek, P.M.; Farrington, H. Nutrient limitation and soil development: Experimental test of a biogeochemical theory. Biogeochemistry 1997, 37, 63–75. [Google Scholar] [CrossRef]
  86. Cherif, M.; Loreau, M. Stoichiometric constraints on resource use, competitive interactions, and elemental cycling in microbial decomposers. Am. Nat. 2007, 169, 709–724. [Google Scholar] [CrossRef]
  87. Ahmed, I.U.; Mengistie, H.K.; Godbold, D.L.; Sanden, H. Soil moisture integrates the influence of land-use and season on soil microbial community composition in the Ethiopian highlands. Appl. Soil Ecol. 2019, 135, 85–90. [Google Scholar] [CrossRef]
  88. Manzoni, S.; Trofymow, J.A.; Jackson, R.B.; Porporato, A. Stoichiometric controls on carbon, nitrogen, and phosphorus dynamics in decomposing litter. Ecol. Monogr. 2010, 80, 89–106. [Google Scholar] [CrossRef][Green Version]
  89. McGroddy, M.E.; Daufresne, T.; Hedin, L.O. Scaling of C:N:P stoichiometry in forests worldwide: Implications of terrestrial Redfield-type ratios. Ecology 2004, 85, 2390–2400. [Google Scholar] [CrossRef]
  90. Zhou, C.; Heal, K.; Tigabu, M.; Xia, L.; Hu, H.; Yin, D.; Ma, X. Biochar addition to forest plantation soil enhances phosphorus availability and soil bacterial community diversity. For. Ecol. Manag. 2020, 455, 117635. [Google Scholar] [CrossRef]
  91. Mahendra, S.; Alvarez-Hong, C.; Lu, S. Biochar, Ochre, and Manure Maturation in an Acidic Technosol Helps Stabilize As and Pb in Soil and Allows Its Vegetation by Salix triandra. Environment 2022, 9, 87. [Google Scholar]
  92. Frost, P.C.; Benstead, J.P.; Cross, W.F.; Hillebrand, H.; Larson, J.H.; Xenopoulos, M.A.; Yoshida, T. Threshold elemental ratios of carbon and phosphorus in aquatic consumers. Ecol. Lett. 2006, 9, 774–779. [Google Scholar] [CrossRef]
  93. Doi, H.; Cherif, M.; Iwabuchi, T.; Katano, I.; Stegen, J.C.; Striebel, M. Integrating elements and energy through the metabolic dependencies of gross growth efficiency and the threshold elemental ratio. Oikos 2010, 119, 752–765. [Google Scholar] [CrossRef]
  94. Allison, S.D.; Vitousek, P.M. Responses of extra cellular enzymes to simple and complex nutrient inputs. Soil Biol. Biochem. 2005, 37, 937–944. [Google Scholar] [CrossRef]
  95. Shen, J.; Bartha, R. Metabolic efficiency and turnover of soil microbial communities in biodegradation tests. Appl. Environ. Microbiol. 1996, 62, 2411–2415. [Google Scholar] [CrossRef][Green Version]
  96. Newell, S.Y.; Fallon, R.D. Toward a method for measuring instantaneous fungal growth rates in field samples. Ecology 1991, 72, 1547–1559. [Google Scholar] [CrossRef]
  97. Keiblinger, K.M.; Hall, E.K.; Wanek, W.; Szukics, U.; Hammerle, I.; Ellersdorfer, G.; Böck, S.; Strauss, J.; Sterflinger, K.; Richter, A.; et al. The effect of resource quantity and resource stoichiometry on microbial carbon-use-efficiency. FEMS Microbiol. Ecol. 2010, 73, 430–440. [Google Scholar] [CrossRef]
  98. Zhang, J.; Elser, J.J. Carbon:Nitrogen:Phosphorus stoichiometry in fungi: A meta-analysis. Front. Microbiol. 2017, 8, 1281. [Google Scholar] [CrossRef]
  99. Gao, W.; Gao, K.; Guo, Z.; Liu, Y.; Jiang, L.; Liu, C.; Liu, X.; Wang, G. Different responses of soil bacterial and fungal communities to 3 years of biochar amendment in an alkaline soybean soil. Front. Microbiol. 2021, 12, 630418. [Google Scholar] [CrossRef]
  100. Scott, J.T.; Cotner, J.B.; LaPara, T.M. Variable stoichiometry and homeostatic regulation of bacterial biomass elemental composition. Front. Microbiol. 2012, 3, 42. [Google Scholar] [CrossRef][Green Version]
  101. Zhao, X.; Miao, R.; Guo, M.; Shang, X.; Zhou, Y.; Zhu, J. Biochar enhanced polycyclic aromatic hydrocarbons degradation in soil planted with ryegrass: Bacterial community and degradation gene expression mechanisms. Sci. Total Environ. 2022, 838, 156076. [Google Scholar] [CrossRef]
  102. Hagemann, N.; Harter, J.; Behrens, S. Elucidating the impacts of biochar applications on nitrogen cycling microbial communities. In Biochar Application; Ralebitso-Senior, T.K., Orr, C.H., Eds.; Elsevier: Amsterdam, The Netherlands, 2016; pp. 163–198. [Google Scholar]
  103. He, X.; Xie, H.; Gao, D.; Khashi, M.; Rahman, U.; Zhou, X.; Wu, F. Biochar and intercropping with potato–onion enhanced the growth and yield advantages of tomato by regulating the soil properties, nutrient uptake, and soil microbial community. Front. Microbiol. 2021, 12, 695447. [Google Scholar] [CrossRef] [PubMed]
  104. Elzobair, K.A.; Stromberger, M.E.; Ippolito, J.A.; Lentz, R.D. Contrasting efects of biochar versus manure on soil microbial communities and enzyme activities in an Aridisol. Chemosphere 2016, 142, 145–152. [Google Scholar] [CrossRef] [PubMed]
  105. Song, L.; Niu, X.; Zhang, N.; Li, T. Effect of biochar-immobilized Sphingomonas sp. PJ2 on bioremediation of PAHs and bacterial community composition in saline soil. Chemosphere 2021, 279, 130427. [Google Scholar] [CrossRef]
  106. Tsai, S.; O’neill, B.; Cannavan, F.S.; Saito, D.; Falcao, N.P.S.; Kern, D.C.; Grossman, J.; Thies, J. The Microbial World of Terra Preta. In Amazonian Dark Earths: Wim Sombroek’s Vision; Woods, W.I., Teixeira, W.G., Lehmann, J., Steiner, C., Winkler Prins, A., Rebellato, L., Eds.; Springer: Berlin/Heidelberg, Germany, 2009. [Google Scholar]
  107. Glaser, B. Prehistorically modified soils of central Amazonia: A model for sustainable agriculture in the twenty-first century. Philos. Trans. R. Soc. B 2007, 362, 187–196. [Google Scholar] [CrossRef] [PubMed][Green Version]
  108. Kim, J.-S.; Sparovek, G.; Longo, R.M.; De Melo, W.J.; Crowley, D. Bacterial diversity of terra preta and pristine forest soil from the Western Amazon. Soil Biol. Biochem. 2007, 39, 684–690. [Google Scholar] [CrossRef]
  109. de Lima Brossi, M.J.; Mendes, L.W.; Gomes Germano, M.; Barbosa Lima, A.; Tsai, S.M. Assessment of Bacterial bph Gene in Amazonian Dark Earth and Their Adjacent Soils. PLoS ONE 2014, 9, e99597. [Google Scholar]
  110. Germano, M.G.; Cannavan, F.S.; Mendes, L.W.; Lima, A.B.; Teixeira, W.G.; Pellizzari, V.H.; Tsai, S.M. Functional diversity of bacterial genes associated with aromatic hydrocarbon degradation in anthropogenic dark earth of Amazonia. Pesqui. Agropecu. Bras. 2012, 47, 654–664. [Google Scholar] [CrossRef]
  111. Sierra, C.A. Temperature sensitivity of organic matter decomposition in the Arrhenius equation: Some theoretical considerations. Biogeochemistry 2012, 108, 1–15. [Google Scholar] [CrossRef]
  112. Leifeld, J.; von Lützow, M. Chemical and microbial activation energies of soil organic matter decomposition. Biol. Fertil. Soils 2014, 50, 147–153. [Google Scholar] [CrossRef][Green Version]
  113. Ding, G.C.; Heuer, H.; Zuhlke, S.; Spiteller, M.; Pronk, J.G.; Heister, K.; Kögel-Knabner, I.; Smalla, K. Soil type-dependent responses to phenanthrene as revealed by determining the diversity and abundance of polycyclic aromatic hydrocarbon ring-hydroxylating dioxygenase genes by using a novel PCR detection system. Appl. Environ. Microbiol. 2010, 76, 4765–4771. [Google Scholar] [CrossRef][Green Version]
  114. Storey, S.; Ashaari, M.M.; McCabe, G.; Harty, M.; Dempsey, R.; Doyle, O.; Clipson, N.; Doyle, E.M. Microbial community structure during fluoranthene degradation in the presence of plants. J. Appl. Microbiol. 2014, 117, 74–84. [Google Scholar] [CrossRef] [PubMed]
  115. Zhu, F.; Storey, S.; Ashaari, M.M.; Clipson, N.; Doyle, E. Benzo(a)pyrene degradation and microbial community responses in composted soil. Environ. Sci. Pollut. Res. 2016, 24, 5404–5414. [Google Scholar] [CrossRef] [PubMed]
  116. Louvel, B.; Cébron, A.; Leyval, C. Root exudates affect phenanthrene biodegradation bacterial community and functional gene expression in sand microcosms. Int. Biodeterior. 2011, 65, 947–953. [Google Scholar] [CrossRef]
  117. Ahmad, M.; Yang, Q.S.; Zhang, Y.Y.; Ling, J.; Sajjad, W.; Qi, S.H.; Zhou, W.G.; Zhang, Y.; Lin, X.C.; Zhang, Y.H.; et al. The distinct response of phenanthrene enriched bacterial consortia to different PAHs and their degradation potential: A mangrove sediment microcosm study. J. Hazard. Mater. 2019, 380, 120863. [Google Scholar] [CrossRef] [PubMed]
  118. Maienza, A.; Baronti, S.; Cincinelli, A.; Martellini, T.; Grisolia, A.; Miglietta, F.; Renella, G.; Stazi, S.R.; Vaccari, F.P.; Genesio, L. Biochar improves the fertility of a Mediterranean vineyard without toxic impact on the microbial community. Agron. Sustain. Dev. 2017, 37, 47. [Google Scholar] [CrossRef][Green Version]
  119. Adamczyk, B.; Kilpeläinen, P.; Kitunen, V.; Smolander, A. Potential activities of enzymes involved in N, C, P and S cycling in boreal forest soil under different tree species. Pedobiologia 2014, 57, 97–102. [Google Scholar] [CrossRef]
  120. Trasar-Cepeda, C.; Leiros, M.C.; Gil-Sotres, F. Hydrolytic enzyme activities in agricultural and forest soils. Some implications for their use as indicators of soil quality. Soil Biol. Biochem. 2008, 40, 2146–2155. [Google Scholar] [CrossRef][Green Version]
  121. Xu, H.J.; Wang, X.H.; Li, H.; Yao, H.Y.; Su, J.Q.; Zhu, Y.G. Biochar impacts soil microbial community composition and nitrogen cycling in an acidic soil planted with rape. Environ. Sci. Technol. 2014, 48, 9391–9399. [Google Scholar] [CrossRef]
  122. Allison, S.D.; Wallenstein, M.D.; Bradford, M.A. Soil-carbon response to warming dependent on microbial physiology. Nat. Geosci. 2010, 3, 336–340. [Google Scholar] [CrossRef]
  123. Sinsabaugh, R.L.; Hill, B.H.; Follstad Shah, J.J. Ecoenzymatic stoichiometry of microbial organic nutrient acquisition in soil and sediment. Nature 2009, 462, 795–798. [Google Scholar] [CrossRef]
  124. van Hees, P.A.W.; Jones, D.L.; Finlay, R.; Godbold, D.L.; Lundström, U.S. The carbon we do not see—The impact of low molecular weight compounds on carbon dynamics and respiration in forest soils: A review. Soil Biol. Biochem. 2005, 37, 1–13. [Google Scholar] [CrossRef]
  125. Song, X.; Razavi, B.S.; Ludwig, B.; Zamanian, K.; Gunina, A. Combined biochar and nitrogen application stimulates enzyme activity and root plasticity. Sci. Total Environ. 2020, 735, 139393. [Google Scholar] [CrossRef]
  126. Giagnoni, L.; Maienza, A.; Baronti, S.; Vaccari, F.P.; Genesio, L.; Taiti, C.; Martellini, T.; Scodellini, R.; Cincinelli, A.; Costa, C.; et al. Long-term soil biological fertility, volatile organic compounds and chemical properties in a vineyard soil after biochar amendment. Geoderma 2019, 344, 127–136. [Google Scholar] [CrossRef]
  127. Takriti, M.; Wild, B.; Schnecker, J.; Mooshammer, M.; Knoltsch, A.; Lashchinskiy, N.; Eloy Alves, R.J.; Gentsch, N.; Gittel, A.; Mikutta, R.; et al. Soil organic matter quality exerts a stronger control than stoichiometry on microbial substrate use efficiency along a latitudinal transect. Soil Biol. Biochem. 2018, 21, 212–220. [Google Scholar] [CrossRef][Green Version]
  128. Williams, E.K.; Jones, D.L.; Sanders, H.R.; Benitez, G.V.; Plante, A.F. Effects of 7 years of field weathering on biochar recalcitrance and solubility. Biochar 2019, 1, 237–248. [Google Scholar] [CrossRef][Green Version]
  129. Mukherjee, A.; Zimmerman, A.R.; Hamdan, R.; Cooper, W.T. Physicochemical changes in pyrogenic organic matter (biochar) after 15 months of field aging. Solid Earth 2014, 5, 693–704. [Google Scholar] [CrossRef][Green Version]
  130. Kuwahara, M.; Glenn, J.K.; Morgan, M.A.; Gold, M.H. Separation and characterization of two extracelluar H2O2-dependent oxidases from ligninolytic cultures of Phanerochaete chrysosporium. FEBS Lett. 1984, 169, 247–250. [Google Scholar] [CrossRef][Green Version]
  131. Sanchez, C. Lignocellulosic residues: Biodegradation and bioconversion by fungi. Biotechnol. Adv. 2008, 27, 185–194. [Google Scholar] [CrossRef]
  132. Prommer, J.; Wanek, W.; Hofhansl, F.; Trojan, D.; Offre, P.; Urich, T.; Schleper, C. Biochar decelerates soil organic nitrogen cycling but stimulates soil nitrification in a temperate arable field trial. PLoS ONE 2014, 9, e86388. [Google Scholar]
  133. Hou, L.; Zhang, L.; Chen, X.; Li, X.; Zhang, Z.; Lin, Y.B. The ben efits of biochar: Enhanced cadmium remediation, inhibited precursor production of nitrous oxide and a short-term disturbance on rhizosphere microbial community. Environ. Pollut. 2021, 272, 116040. [Google Scholar] [CrossRef]
  134. Fuertes-Mendizábal, T.; Huérfano, X.; Vega-Mas, I.; Torralbo, F.; Menéndez, S.; Ippolito, J.A.; Kammann, C.; Wrage-Mönnig, N.; Cayuela, M.L.; Borchard, N.; et al. Biochar reduces the efficiency of nitrification inhibitor 3,4-dimethylpyrazole phosphate (DMPP) mitigating N2O emissions. Sci. Rep. 2019, 9, 2346. [Google Scholar] [CrossRef] [PubMed][Green Version]
  135. Wang, A.; Li, X.; Hao, X.; Luo, X.; Chen, W.; Huang, Q. Ammonia level influences the assembly of dissimilatory nitrate reduction to ammonia bacterial community in soils under different heavy metal remediation treatments. Sci. Total Environ. 2022, 838, 156393. [Google Scholar] [CrossRef] [PubMed]
  136. Ji, C.; Li, S.; Geng, Y.; Miao, Y.; Zou, J. Differential responses of soil N2O to biochar depend on the predominant microbial pathway. Appl. Soil Ecol. 2019, 145, 103348. [Google Scholar] [CrossRef]
  137. Song, Y.; Zhang, X.; Ma, B.; Chang, S.X.; Gong, J. Biochar addition affected the dynamics of ammonia oxidizers and nitrification in microcosms of a coastal alkaline soil. Biol. Fertil. Soils 2014, 50, 321–332. [Google Scholar] [CrossRef]
  138. Sun, J.; He, F.; Zhang, Z.; Shao, H.; Xu, G. Temperature and moisture responses to carbon mineralization in the biochar-amended saline soil. Sci. Total Environ. 2016, 569–570, 390–394. [Google Scholar] [CrossRef]
  139. Yao, Q.; Liu, J.; Yu, Z.; Li, J.; Jin, J.; Liu, X.; Wang, G. Three years of biochar amendment alters soil physiochemical properties and fungal community composition in a black soil of northeast China. Soil Biol. Biochem. 2017, 110, 56–67. [Google Scholar] [CrossRef]
  140. Ibrahim, M.M.; Zhang, H.; Guo, L.; Chen, Y.; Heiling, M.; Zhou, B.; Mao, Y. Biochar interaction with chemical fertilizer regulates soil organic carbon mineralization and the abundance of key C-cycling-related bacteria in rhizosphere soil. Eur. J. Soil Biol. 2021, 106, 103350. [Google Scholar] [CrossRef]
  141. Bai, S.H.; Reverchon, F.; Xu, C.Y.; Xu, Z.; Blumfield, T.J.; Zhao, H.; Van Zwieten, L.; Wallace, H.M. Wood biochar increases nitrogen retention in field settings mainly through abiotic processes. Soil. Biol. Biochem. 2015, 90, 232–240. [Google Scholar] [CrossRef]
  142. Mierzwa-Hersztek, M.; Klimkowicz-Pawlas, A.; Gondek, K. Influence of poultry litter and poultry litter biochar on soil microbial respiration and nitrifying bacteria activity. Waste Biomass Valorization 2018, 9, 379–389. [Google Scholar] [CrossRef]
  143. Ahmad, Z.; Mosa, A.; Zhan, L.; Gao, B. Biochar modulates mineral nitrogen dynamics in soil and terrestrial ecosystems: A critical review. Chemosphere 2021, 278, 130378. [Google Scholar] [CrossRef]
  144. Feng, Y.; Du, H.; Wulandari, T.; Poinern, G.E.J.; Jiang, Z.-T.; Fawcett, D.; Hassan, N.; Xue, L.; Yang, L. Hydrochar amendments stimulate soil nitrous oxide emission by increasing production of hydroxyl radicals and shifting nitrogen functional genes in the short term: A culture experiment. Chemosphere 2022, 302, 134771. [Google Scholar] [CrossRef] [PubMed]
  145. Zhang, X.; Zhang, J.; Song, M.; Dong, M.; Xiong, Z. N2O and NO production and functional microbes responding to biochar aging process in an intensified vegetable soil. Environ. Pollut. 2022, 307, 119491. [Google Scholar] [CrossRef] [PubMed]
  146. Zhang, Y.; Zhao, H.; Hu, W.; Wang, Y.; Zhang, H.; Zhou, X.; Fei, J.; Luo, G. Understanding how reed-biochar application mitigates nitrogen losses in paddy soil: Insight into microbially-driven nitrogen dynamics. Chemosphere 2022, 295, 133904. [Google Scholar] [CrossRef] [PubMed]
  147. Kuever, J.; Konneke, M.; Galushko, A.; Drzyzga, O. Reclassification of Desulfobacterium phenolicum as Desulfobacula phenolica comb. nov and description of strain Sax (T) as Desulfotignum balticum gen. nov., sp nov. Int. J. Syst. Evol. Microbiol. 2001, 51, 171–177. [Google Scholar] [CrossRef] [PubMed][Green Version]
  148. Wu, X.; Zhou, Y.; Liang, M.; Lu, X.; Chen, G.; Zan, F. Insights into the role of biochar on the acidogenic process and microbial pathways in a granular sulfate-reducing up-flow sludge bed reactor. Bioresour. Technol. 2022, 355, 127254. [Google Scholar] [CrossRef]
  149. Pannekens, M.; Kroll, L.; Müller, H.; Mbow, F.T.; Meckenstock, R.U. Oil reservoirs, an exceptional habitat for microorganisms. New Biotechnol. 2019, 49, 1–9. [Google Scholar] [CrossRef]
  150. Gieg, L.M.; Fowler, S.J.; Berdugo-Clavijo, C. Syntrophic biodegradation of hydrocarbon contaminants. Curr. Opin. Biotechnol. 2014, 27, 21–29. [Google Scholar] [CrossRef]
  151. Fox, A.; Kwapinski, W.; Griffiths, B.S.; Schmalenberger, A. The role of sulfur- and phosphorus-mobilizing bacteria in biochar-induced growth promotion of Lolium perenne. FEMS Microbiol. Ecol. 2014, 90, 78–91. [Google Scholar] [CrossRef][Green Version]
  152. Chen, P.; Zheng, X.; Cheng, W. Biochar combined with ferrous sulfate reduces nitrogen and carbon losses during agricultural waste composting and enhances microbial diversity. Process Saf. Environ. 2021, 162, 531–542. [Google Scholar] [CrossRef]
  153. Kappler, A.; Wuestner, M.L.; Ruecker, A.; Harter, J.; Halama, M.; Behrens, S. Biochar as an Electron Shuttle between Bacteria and Fe (III) Minerals. Environ. Sci. Technol. Lett. 2014, 1, 339–344. [Google Scholar] [CrossRef]
  154. Schulte, S.; Köster, D.; Jochmann, M.; Meckenstock, R. Applying reverse stable isotope labeling analysis by mid-infrared laser spectroscopy to monitor BDOC in recycled wastewater. Sci. Total Environ. 2019, 665, 1064–1072. [Google Scholar] [CrossRef] [PubMed]
Figure 1. Organic matter decomposition in soil (a) and biochar decomposition in soil (b) processes mainly controlled by hydrolase and oxo-reductase enzymatic activities. The bold lines indicate the decomposition processes stimulated by the biochar amendment in soil. LMWCOs are the low molecular weight organic compounds.
Figure 1. Organic matter decomposition in soil (a) and biochar decomposition in soil (b) processes mainly controlled by hydrolase and oxo-reductase enzymatic activities. The bold lines indicate the decomposition processes stimulated by the biochar amendment in soil. LMWCOs are the low molecular weight organic compounds.
Environments 09 00138 g001
Figure 2. CUE in unamended soil, in soil amended with biochar produced by pyrolysis (pyrolysis biochar) and in soil amended with biochar produced by gasification (gasification biochar).
Figure 2. CUE in unamended soil, in soil amended with biochar produced by pyrolysis (pyrolysis biochar) and in soil amended with biochar produced by gasification (gasification biochar).
Environments 09 00138 g002
Figure 3. The chemical modification of biochar and organic substances in biochar amended soils and the potential metabolic mechanisms.
Figure 3. The chemical modification of biochar and organic substances in biochar amended soils and the potential metabolic mechanisms.
Environments 09 00138 g003
Figure 4. Hypotetical pathway of S and Fe reductions by electron shuttling via biochar.
Figure 4. Hypotetical pathway of S and Fe reductions by electron shuttling via biochar.
Environments 09 00138 g004
Publisher’s Note: MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

Share and Cite

MDPI and ACS Style

Giagnoni, L.; Renella, G. Effects of Biochar on the C Use Efficiency of Soil Microbial Communities: Components and Mechanisms. Environments 2022, 9, 138.

AMA Style

Giagnoni L, Renella G. Effects of Biochar on the C Use Efficiency of Soil Microbial Communities: Components and Mechanisms. Environments. 2022; 9(11):138.

Chicago/Turabian Style

Giagnoni, Laura, and Giancarlo Renella. 2022. "Effects of Biochar on the C Use Efficiency of Soil Microbial Communities: Components and Mechanisms" Environments 9, no. 11: 138.

Note that from the first issue of 2016, this journal uses article numbers instead of page numbers. See further details here.

Article Metrics

Back to TopTop