1 Introduction

Biochar is the solid product of pyrolyzing biomass under the temperatures of 300–700 °C with limited oxygen (Lehmann and Joseph 2015). Biochar includes artificial biochar and natural biochar (usually formed by wildfire), and has a large range of physicochemical properties that are determined by pyrolysis condition and feedstock type. Due to its high alkalinity, porous structure, stable C content and surface area, biochar is widely applied for soil fertility improvement, soil carbon sequestration, crop promotion and contaminated soil remediation (Caporaso et al. 2012; Lehmann and Joseph 2012).

Over the last decade, the effects of biochar on soil microbial growth and communities, and the underlying mechanisms have been widely reported (Lehmann et al. 2011; Muhammad et al. 2014; Palansooriya et al. 2019; Yu et al. 2020) because soil microorganisms carry out a range of important ecosystem functions, such as soil nutrient cycling and plant growth (Fierer 2017). Soil microorganisms contribute greatly to decomposition of soil organic C, nitrification, denitrification and phosphorous solubilization, and these processes are associated with soil fertility maintenance and global climate change (Fierer 2017). Furthermore, microorganisms interact with plant roots to enhance their nutrient uptake capacity and increase plant systemic resistance to environmental perturbation (e.g., pathogen invasion and drought), and finally benefit plant health and growth (Philippot et al. 2013). Bacteria are most diverse in terrestrial ecosystems and are susceptible to the availability of C sources and pH changes (Dai et al. 2018b). Most soil nutrient cycles, such as carbohydrate metabolism, C fixation, nitrification, denitrification and P solubilization, are primarily mediated by bacterial communities. Soil fungi, heterotrophic microorganisms, are saprophytic, parasitic or pathogenic (Madigan et al. 2008). They can either decompose plant residues or influence plant health and production by forming arbuscular mycorrhiza or cause diseases (Elmer and Pignatello 2011; Liu et al. 2015). Soil fauna, a major component of soil food webs are the consumers of bacteria, fungi, algae and nematodes, and are as parasites of plants and animals (Geisen et al. 2018). The fauna can exert top-down control effects and are affected by both soil properties (e.g., moisture, clay content and nutrient status) and associated bacterial and fungal communities (Xiong et al. 2018; Oliverio et al. 2020). Overall, any changes in soil properties, such as nutrient status, pH and texture, would alter the microbial abundance, diversity and communities and related functions. The effects of biochar on soil microbial biomass and communities have been widely reported across different biochar pyrolysis temperatures, biochar feedstocks, soil type and experimental condition.

The number of literature involved in biochar effects on soil biota increased rapidly from year 2011 to 2020 (Fig. 1). Previous reviews have widely reported the responses of soil biota and mediated nutrient cycling to biochar addition, while the relationships between biochar physicochemical properties and microbial responses, i.e. underlying mechanisms, still need further summary. In addition, previous reviews/meta-analysis only synthesized the responses of microbial community to biochar mainly based on the traditional techniques (e.g., PLFA and DGGE). With the recently rapid development of analytical techniques (e.g., high-throughput sequencing), we can observe the diversity and abundances of microorganisms at a higher taxonomic resolution (e.g., bacteria and fungi) in biochar-amended soils. The roles of fungal and protist keystones in biochar-soil–plant continuum and specific biogeochemical processes such as P solubilization in biochar-amended soils should not be negligible, while most studies have focused on bacterial roles and the N cycling process. Therefore, this review attempts to resolve the above deficiencies. It aims to (1) summarize biochar unique physicochemical properties, such as porous structure, pH, labile C and electron shuttle, that play important roles in microbial growth, (2) identify the relationships between biochar properties and soil bacterial, fungal and fauna communities, and (3) synthesize the effects of biochars on microbial-mediated soil N and P transformation processes and underlying mechanisms. Finally, the review highlights the issues related to research methodology and the need of further research.

Fig. 1
figure 1

The number and proportion of literatures involved in biochar effects on soil biota from year 2011 to 2020. The histograms presenting the number of biochar publications were corresponding to the left axis and the lines presenting the proportion of biochar publications were corresponding to the right axis

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2 Relationship between biochar properties and microbial responses

Biochar can change soil physicochemical properties, which in turn alters soil microbial communities and related functions via following ways (Fig. 2). First, biochar can serve as a potential porous habitat for microbial growth and protection from predators. Second, it provides carbon source (in particular labile C) and mineral nutrients for microorganisms. Third, biochar changes soil basic properties such as increase in soil pH. Furthermore, it acts as an electron shuttle for microbial metabolisms (Dai et al. 2019; Saquing et al. 2016; Wu et al. 2018).

Fig. 2
figure 2

The special physicochemical properties (i.e., pH, labile C, porous structure, electron shuttle and ash) of biochars that play important roles in altering the activity, diversity and community structure of bacteria, fungi and fauna

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2.1 Porous structure

Biochar has porous aromatic structures that present high surface area and adsorption capacity. They can adsorb water, labile C and mineral nutrients from adjacent soils. The appropriate porous structure (i.e. macropores) can protect less competitive microorganisms from predator grazing (e.g., protists). Thus, microorganisms may use biochar particles as a habitat to colonize, grow and reproduce (Fig. 2). For instance, using a scanning electron microscope, Luo et al. (2013) observed the rich microbial colonization (i.e., diverse microbial shape and size) on the porous structure of lignocellulose-based biochar derived from Miscanthus giganteus. Dai et al. (2017a, b) showed that diverse microorganisms grew in manure-based biochar particles and that Actinobacteria with hyphae and spores were the dominant taxon in biochar particles, regardless of soil type. These studies also indicated that the presence of microorganisms is independent of biochar type and soil environments, as pyrolyzed biochar particles are always porous.

In general, the porosity and surface area of biochar increase with increasing pyrolysis temperature (Table 1) (Dai et al. 2013; Suliman et al. 2017). The microbial colonization would be more diverse in biochars pyrolyzed at high temperatures. This was consistent with more diverse microbial communities in manure biochar pyrolyzed at 700 °C compared to at 300 °C (Dai et al. 2017a, b). While the labile C reduces as pyrolysis temperature increases, biochar C availability also accounts greatly for microbial colonization (Table 1), which was supported by the higher number of microbes colonized in low-temperature biochar (Luo et al. 2013). Overall, the microbial colonization has resulted from the overall effects of porous structure and labile C content. Furthermore, the effects of porous structure were not apparent in aged biochar as soil succession, showed by sparse microbes being colonized in both external and internal surfaces of wood biochar aged for three years (Quilliam et al. 2013). This suggests that aging may reduce the microbial colonization in biochar particles, as the porous structure may collapse during aging and the adsorbed nutrients may be largely consumed by microorganisms.

Table 1 The effects of pyrolysis temperature and feedstock type on biochar physicochemical properties that influence microbial communities
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2.2 Labile C

The labile C, which can serve as C substrates for microbial growth and metabolism, increases with decreasing pyrolysis temperature and is higher in lignocellulose-based biochar than manure-based ash-rich biochar (Table 1) (Cross and Sohi 2011; Dai et al. 2016b). The labile C can directly increase the microbial population and alter microbial communities by supplying energy sources such as carbohydrates. Some other bio-oils and condensates, such as polycyclic aromatic hydrocarbons, xylenol in labile C, are toxic to most microorganisms (Freddo et al. 2012), while they may selectively stimulate the growth of specific microorganisms (Table 2).

Table 2 Current research status and future perspectives for biochar effects on soil microbial communities
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Dai et al. (2019) used acetone to extract the major fraction of labile C from manure biochar and found that acetone-extracted C largely affected microbial biomass, community structure and microbial-mediated N cycling processes in both farmland and forest soils (Dai et al. 2018a, 2019). These effects were significantly greater than the aromatic C and ashes of biochar, supporting the importance of labile C in determining microbial communities. Moreover, isotopic tracing methods also verified that the labile C fraction of biochar was highly associated with microbial respiration and community succession in different soils, regardless of biochar type and incubation condition (Luo et al. 2013, 2017; Watzinger et al. 2014). Although the aliphatic C (volatile matter) accounts for 13.2% to 70.0% of biochar (Enders et al. 2012), microorganisms only utilize a small amount of the volatile matter (less than 5%), and the mineralization rate of labile C dramatically decreased to a minimum level during one year of incubation (Nguyen et al. 2014; Watzinger et al. 2014). For instance, only 4.3% of the ryegrass biochar C was mineralized by microorganisms after 158 days of incubation and only 0.45% of the biochar C was incorporated into microbial biomass (Maestrini et al. 2014). This implies that the effects of labile C on microbial communities are usually short-lived and diminishing over time. We point out that ash can supply inorganic nutrients for microbial growth and form organo-mineral bonds as the major driver for the microbial-metabolite stabilization (Fig. 2). However, the maximum content of ash from biochar only caused a small change in microbial communities (Dai et al. 2018a), again emphasizing the importance of labile C in determining microbial communities.

2.3 Biochar pH

Biochar is usually alkaline (pH > 7.0), resulting in a large increase in soil pH, especially for acid soils. Its alkalinity increases with increasing pyrolysis temperature and in most cases the manure-based biochars have higher pH compared to lignocellulose-based biochars (Table 1) (Dai et al. 2013; Enders et al. 2012). In general, soil microbial growth, community changes and relevant functions were highly sensitive to pH changes. An increase in bacterial growth and a decrease in fungal growth were observed with increasing soil pH at the gradient pH range of 4.0 to 8.3 (Rousk et al. 2009). In addition, the diversity and community structure of bacteria and the relative abundance of dominant phyla revealed by 16S sequencing were positively associated with soil pH, while the fungi and protist were not responsive to soil pH (Oliverio et al. 2020; Rousk et al. 2010). The abundances of bacterial phyla are also sensitive to pH, and for example the abundances of Acidobacteria are negatively related to soil pH (Madigan et al. 2008). Furthermore, the abundances of microbial genes involved in soil nutrient cycling, such as nitrification (e.g., amoA genes) and P mineralization (e.g., phoD genes), were affected by soil pH (Dai et al. 2020; DeForest and Otuya 2020).

Therefore, the effects of biochar on soil microbial communities, in particular bacterial communities, via increasing soil pH have been widely emphasized. For instance, the biochars with higher pH contributed greatly to the increases in bacterial diversity (Zhou et al. 2020). Also, the variances in bacterial communities after biochar addition were dominantly determined by soil pH rather than other properties such as organic C, regardless of agricultural soils or forest soils (Dai et al. 2016a; Wang et al. 2020). Wildfire-produced biochar consistently showed the similar trends for microbial evolution/succession (Zackrisson et al. 1996), as microbiomes in biochar significantly interact with those in adjacent soils, and this interaction is ascribed to niche differentiation between habitat specialists in biochar and soil (Zhang et al. 2021).

2.4 Electrochemical properties

The electrochemical properties of biochar have received an increasing attention, due to its (1) redox-active moieties that enable biochar to donate and accept electrons and (2) graphite-like aromatic structures for electron transfer (Sun et al. 2017). Low pyrolysis temperatures (e.g., 400–500 °C) induce redox-active moieties (e.g., quinone and hydroquinone) dominated by electron donating, while high temperatures (e.g., 650–800 °C) induce electron-accepting moieties and graphite-like aromatic structure (Table 1) (Kluepfel et al. 2014; Yu et al. 2015; Sun et al. 2018). Thus, biochar can act as an electron shuttle for microorganisms, which participate in redox reactions, such as nitrification and denitrification. The examples include the stimulated microbial reduction of Fe(III) oxyhydroxide mineral ferrihydrite by biochar (Kappler et al. 2014) and the suppression of N2O emission in biochar-amended soils (Harter et al. 2014). The oxidized biochar by H2O2 increased soil N2O emission rates, again indicating the suppressive effect of biochar redox-active moieties on N2O emissions from soil denitrification (Yuan et al. 2019).

3 Biochar effects on soil microbial activity, biomass and enzyme activities

3.1 Biochar effects on soil CO2 evolution and microbial biomass

Biochar addition significantly affects soil CO2 evolution and microbial biomass by altering soil C substrate quality and quantity, while the effects are short-term and decrease over time as the C substrates are gradually consumed by microorganisms. For example, the application of corn and nutshell biochars significantly enhanced the cumulative microbial respiration and C-use efficiencies during short-term incubation (Khadem and Raiesi 2017; Xu et al. 2018b). The transitory and large increase in soil CO2 evolution after biochar addition is mainly attributed to: (1) increased microbial respiration using labile C or toxic organic compounds as C substrates; (2) increased decomposition priming effect of native organic C pools in soil; and/or (3) the metabolism of microbial necromass by other microorganisms (Lehmann and Joseph 2015).

By contrast, some studies have reported that biochar has inhibiting effects or no effects on soil CO2 evolution (Zhou et al. 2017; Li et al. 2018; Chen et al. 2019). For instance, biochar application to a subtropical bamboo plantation reduced soil microbial heterotrophic respiration (Li et al. 2018). This was attributed to increased aromatic C content and the decreased activities of β-glucosidase and cellobiohydrolase which are responsible for decomposition of carbohydrates and celluloses (Li et al. 2018). No changes in soil respiration in temperate forests and reduced microbial respiration in paddy soils with the application at 40 t ha−1 also supported the inhibiting effects of biochar on soil CO2 evolution (Zhou et al. 2017; Chen et al. 2019). The decrease in CO2 evolution over time in incubation and field experiments is mainly attributed to: (1) the enhancement of C recalcitrancy and the decrease of labile C pool (Chen et al. 2019), and (2) the presence of toxic compounds which are only utilized by a minority of microorganisms.

Our review showed that biochar addition significantly increased microbial biomass C and N (Fig. 3), which was consistent with the study of Pokharel et al. (2020). The microbial biomass C and N in biochar-amended soils were 455 and 24.0 mg kg−1, while the biomass C and N in non-amended soils were 412 and 20.2 mg kg−1, respectively (Fig. 3). In addition, we pointed out that biochar produced at low temperatures had larger increasing effects on microbial biomass than those at high temperatures due to its higher content of labile C (Table 1), since biochar provides C substrates for microbial growth. This was supported by the increased microbial biomass C and N increased with the application of biochar pyrolyzed at 350 °C but were not affected by the biochars pyrolyzed at high temperatures of 500 and 700 °C (Guo et al. 2020). In another study, increasing application rate of wheat-straw biochar (350–550 °C) increased the amount of labile C which in turn increased microbial biomass C (Chen et al. 2016). Although biochars produced at high pyrolysis temperatures provide more living space for microorganisms, the abundant labile C substrates in biochars with low pyrolysis temperatures are the basic biogenic resource for the majority of microbial growth. Thus, we suppose the more important role of C substrates than porous structure in increasing microbial biomass.

Fig. 3
figure 3

The effects of biochar on soil microbial biomass C and N, N mineralization rate, nitrification rate, microbial alpha diversity and the relative abundances of soil dominant bacterial and fungal taxa. “*”and “-”represent the significant and non-significant differences in microbial parameters between control and biochar-amended soils at p ≤ 0.05, respectively. The values above the black box are the average values of the microbial parameters, and the n represents the sample size. The references that provide the data are shown in the Supporting Information

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3.2 Biochar effects on soil microbial enzyme activities

Biochar affects the activities of soil extracellular enzymes which are responsible for organic C degradation and other important enzymes for N mineralization or P solubilization. Such effects vary with biochar property, soil type and enzyme type (Zhang et al. 2019; Pokharel et al. 2020). A meta-analysis investigated the biochar effects on the activities of soil C-degrading enzymes and showed that biochar reduced the activities of total enzymes involved in C cycling by 6.3% (Zhang et al. 2019). The deceases were larger when biochar addition rates and pyrolysis temperature increased (Zhang et al. 2019). These C-degrading enzymes include α-glucosidase, β-cellobiosidase and β-glucosidase (Chen et al. 2016). On the contrary, Infan et al. (2019) indicated that biochar at the application rate of 1% C (w/w) stimulated the activities of dehydrogenase, which was consistent with the increased dehydrogenase activities by 19.8% in a meta-analysis (Pokharel et al. 2020). Some studies revealed that the activities of organic C-degrading enzymes reduced while those of C:N-acquiring enzymes increased moderately with increasing pyrolysis temperature of biochar (Guo et al. 2020). We attributed the contrasting responses of C-cycling enzyme activities to the ratio of biochar labile C to recalcitrant C, which determines microbial C-use efficiency.

Contrasting effects of biochar on the activities of N mineralization or P solubilization enzymes were observed in our study (Fig. 3) and other studies (Chen et al. 2016, 2020; Zhang et al. 2017). Based on a meta-analysis, biochar significantly increased the activities of urease and alkaline phosphatase by 23.1% and 25.4% (Pokharel et al. 2020). Other individual observations also showed the significant increase in the activities of urease, nitrite reductase, N-acetylglucosaminidase and alkaline phosphatase in different soils following biochar additions (Chen et al. 2016, 2020; Zhang et al. 2017). However, the activities of N and P enzymes were related to the application rate and biochar type. For example, the addition of 10 mg kg−1 biochar stimulated the activities of alkaline phosphatase and urease, whereas 50 mg kg−1 inhibited alkaline phosphatase and invertase activities (Huang et al. 2017). Similarly, the activity of N acquisition (N-acq) enzyme increased with the application rate of 1% but was unaffected with the application rate of > 1%. In addition, biochars produced at pyrolysis temperature of 350–550 °C with a pH of > 10 and C/N ratio of < 50 increased the urease activity to a greater extent than those produced at other pyrolysis conditions (Pokharel et al. 2020).

4 Biochar effects on soil microbial diversity and community structure

4.1 Soil bacterial diversity and community structure

Overall effects of biochar on soil bacterial diversity and community structure depend on biochar type, soil type and agricultural managements, such as crop type and planting duration (Abujabhah et al. 2016; Dai et al. 2016a; Herrmann et al. 2019; Liu et al. 2018; Yu et al. 2018a). With biochar type, the bacterial diversity was higher in forest soils amended with leaf biochar than with woodchip biochar, probably attributed to the higher pH with the leaf biochar (Zhou et al. 2020). The Proteobacteria was the dominant taxon in biochar pyrolyzed at low temperature (300 °C), while the Chloroflexi tended to be more prevalent in biochar pyrolyzed at high temperature (700 °C) (Dai et al. 2017a, b). Soil type impacted the responses of bacterial diversity and community structure to biochar addition. For instance, biochars at addition rates of 20 and 40 t ha−1 increased the relative abundance of Betaproteobacteria and Deltaproteobacteria in the soil with pH 4.89 and soil organic C of 17.7 g kg−1, while decreased the abundance of Betaproteobacteria and increased the abundance of Chloroflexi in the soil with pH 5.99 and soil organic C of 20.1 g kg−1. No significant changes were found in the soils with pH 6.21 and soil organic C of 18.8 g kg−1 (Chen et al. 2015). A field experiment also showed that straw biochar increased the relative abundances of Proteobacteria and Chloroflexi in bulk soils while it enhanced the relative abundances of Gemmatimonadetes in rhizosphere soils (Cheng et al. 2019). Woolet and Whitman (2020) identified the positive responders, such as Nocardioides, Micromonospora, and Ramlibacter, to biochar at the genus level by re-analyzing publicly available raw data from 16S Illumina sequencing, while most bacterial taxa did not consistently respond to biochar addition across soil types.

Crop type and planting duration also impacted the responses of bacterial diversity and community structure to biochar addition. For example, the addition of biochar increased the abundances of bacterial (16S rRNA) gene, and Gram +, Gram − and Pseudomonas bacteria in the soils planted with mash bean but not with wheat (Azeem et al. 2020). In cotton-cropping soils, biochar amendment increased the abundances of Sphingomonas and Pseudomonas in biochar-amended cotton soils compared to the un-amended control, while the abundances of these bacteria decreased after continuous cropping for 11 years and 14 years (Han et al. 2017). Nguyen et al. (2018) compared the differences in bacterial diversity and communities between newly applied (1 year) and aged biochar (9 year), and observed that the abundances of Cyanobacteria, N2-fixation bacteria, nitrification and denitrification bacteria were lower after 9 years of biochar history although the similar bacterial diversity and community structure (e.g., Chao 1 and Shannon index) were found between 1 and 9 years of biochar amendment (Nguyen et al. 2018). Although the variable effects on bacterial diversity were reported, a new meta-analysis concluded that the increased diversity was generally observed in acidic and sandy soils with biochar pyrolyzed under low temperature and derived from nutrient-rich feedstocks, as well as with low application rates (Li et al. 2020). Furthermore, our statistical analyses collected the abundance of bacterial taxa from high-throughput sequencing from published literatures (references listed in Supporting Information) and showed that biochar significantly increased the bacterial alpha diversity and the relative abundances of Firmicutes and Proteobacteria and decreased the relative abundance of Acidobacteria by paired t test (Fig. 3). This synthesis gives a supplement to the understanding of bacterial communities based on the traditional techniques.

4.2 Soil fungal diversity and community structure

In general, the effects of biochar on fungal diversity and community structure have been widely assessed, while the soil dominant taxa of Basidiomycota and Ascomycota were not responsive to biochar addition (Fig. 3). At the functional perspective, fungi encompass three functional groups with saprotrophs, pathotrophs and symbiotrophs (Madigan et al. 2008). Among them, saprotrophs decompose soil organic matter and biochar particles, and the pathotrophs cause crop diseases. The symbiotrophs promote plant growth by forming the root-fungal mutualists in soil ecosystems.

Saprophytic fungi rely on C substrates as the energy source and hence biochar addition always enhances the growth of fungal saprotrophs. For instance, Dai et al. (2018a) found that the labile C of biochar as a microbial C source increased the abundance of saprotroph, enhancing their competitive capacity with pathotrophs and symbiotrophs, and finally decreased fungal diversity. The inorganic minerals and aromatic C had less effects compared to labile C. This indicates that C substance is likely more important to determine saprotroph growth rather than microbial habitats and mineral nutrients. Except the mineralization of biochar itself, biochar addition to soil may accelerate the decomposition of soil organic C (priming effects) mediated by fungal communities (Luo et al. 2011). For instance, the microbial succession altered from bacterial to fungal communities during soil incubation with biochar, and Sordariomycetes and Tremellomycetes contributed more to soil priming effects in a late phase (i.e., 40 day) compared to bacterial communities (Yu et al. 2018b). Some studies report that saprophytic fungi have the capacity to utilize the recalcitrant C fraction of biochar when the labile-C substances have been utilized (Dai et al. 2018a). Overall, saprophytic fungi play an important role in soil C biochemical cycling by decomposing both soil organic C and biochar C.

Biochar can reduce plant disease by suppressing soil-borne pathogens. Dai et al. (2018a) found that biochar favors the growth of saprotrophs over soil-borne fungal pathogens by providing labile C for saprotroph growth and enhance their competitive capacity to pathogens. The application of biochar for 3 years decreased the relative abundances of potential soil-borne plant pathogens (e.g. Fusarium) and suppressed crop diseases in a black soil (Yao et al. 2017). The infection of Fusarium on asparagus roots was suppressed by coconut biochar (Matsubara et al. 2002) and hardwood-dust biochar (Elmer and Pignatello 2011). Other studies involved in the soil-borne pathogen suppression caused by biochar and the pathogens varied from the Phytophthora spp. in tree seedlings (Zwart and Kim 2012) to Rhizoctonia solani in cucumber (Jaiswal et al. 2014). The possible mechanisms are described as follows. First, soil microorganisms promoted by biochar addition can compete the C resources with pathogens, produce toxic compounds to pathogens or parasitizing pathogens (Dai et al. 2018a; Graber et al. 2010). Second, biochar indirectly increases plant systemic resistance by supplying nutrients, improving root architecture and enhancing arbuscular mycorrhizal colonization (Lehmann and Joseph 2015).

Mycorrhizae are one of the ubiquitous symbiotrophs in soil–plant systems. They build a vast connection between plant roots and soils to increase plant nutrient uptake, such as N, P and water (Elmer and Pignatello 2011). Biochar has been considered as a micro-habitat for mycorrhizal growth and interaction with plant roots (Ameloot et al. 2015), while most studies have reported the negative effects or no effects of biochar on mycorrhizae communities. For instance, biochars significantly decreased the biomass of arbuscular mycorrhizal fungi (AMF) in a soil with low soil organic C (8.9 g C kg−1), while did not affect it in a soil with high soil organic C (16.1 g C kg−1). The poultry-litter biochar had larger decreasing effects on AMF biomass compared to pine-chip biochars (Ameloot et al. 2015). Similarly, mycorrhizal colonization in tallgrass roots was suppressed by royal-oak biochar addition (5.2%) in the first year of study and had no effects in the second or third year (Biederman et al. 2017). This trend was also reported by Elzobair et al. (2016), showing no effects of biochars on arbuscular mycorrhizal fungal colonization on corns in an Aridisol (Elzobair et al. 2016) and cowpea plants growing in Mollisol/Alfisol (Cobb et al. 2018). Furthermore, biochars derived from switchgrass, hardwood, or softwood feedstocks all reduced the colonization in roots of Allium porrum L. by the AMF (Han et al. 2016). By contrast, for potato plants, straw biochar significantly increased AMF root colonization, and plant N, P, and K uptake and plant biomass (Yang et al. 2020). Further work should be focused on understanding the underlying mechanisms and design of the targeted biochar for mycorrhizal colonization.

4.3 Soil fauna diversity and community structure

Soil fauna communities are the consumers of bacteria, fungi and other microorganisms. Biochar can change soil fauna communities and contribute greatly to soil nutrient cycling. For instance, biochar significantly increased the abundance of soil total nematodes and affected nematode community structure, with bacterivores dominant in the biochar treatments based on a 5-year field experiment (Liu et al. 2020b). The altered nematode community by biochar contributed to the improvement of soil quality and the enhancement of productivity in yellow cinnamon soil (Liu et al. 2020b). A farm case study indicated that straw biochar increased protist population, but had no effect on earthworms (Hansen et al. 2017). In a cultivated acidic soil, biochar only increased the flagellates abundance and had negative effects on the growth of nematodes, such as bacterivorous, fungivorous, and herbivorous and the amoebae (protists), indicating the detrimental effect on multitrophic levels of soil fauna (Liu et al. 2020a).

The mechanisms by which biochar affects soil fauna communities are related to (1) specific biochar properties, (2) soil physicochemical property changes and (3) changes in soil bacterial and fungal communities. Noyce et al. (2016) showed that the microorganisms living in or around the biochars had higher diversity than the adjacent soil, in particular, with the dominant taxa of Aveolata superphylum from protist, suggesting that biochar particles provide a unique habitat for soil protists. Biochar-induced changes in soil physicochemical properties, such as soil pH, C availability and moisture, may affect fauna (e.g., protist) growth and community succession (Geisen et al. 2018). This was supported by the Asiloglu et al. (2021a) revealing that the increase in the relative abundance of Stramenopiles in biochar-amended soils was highly correlated with the increases in soil total pore volume and C/N ratio. The increases in the relative abundances of protists, such as Alveolata, Amoebozoa and Excavata, were associated with increased soil pH and nutrient status (Asiloglu et al. 2021a). In addition, as the predators in food web, the shifts in bacterial and fungal communities after biochar addition would also affect the behaviors of fauna (Xiong et al. 2018). For instance, both poultry-litter and rice-husk biochars significantly altered soil bacterial communities and trophic interactions, resulting in the changes in protist communities and relevant N cycling processes (Asiloglu et al. 2021b).

5 Biochar effects on soil microbial-mediated nutrient transformation processes

5.1 Soil microbial-mediated N cycling

Biochar pH, labile C content, mineral N availability and electrochemical properties play an important role in mediating soil N cycling (Fig. 4). Biochars increase soil pH (Yuan et al. 2011) that can change the activity and community of nitrifiers and affect soil nitrification and N2O emission (Yuan et al. 2011; Liu et al. 2017; Yu et al. 2019). For example, Lin et al. (2017) found that straw biochar increased the abundance of the bacterial amoA genes (ammonia-oxidizing bacteria) in the nitrification process, due to the increased soil pH. The importance of biochar pH in determining N-cycling microorganisms is clearly supported by a char-sphere study. Yu et al. (2019) found that soil pH and bacterial amoA abundance increased in the proximity to the biochar surface, while pH decreased with increasing the distance from the biochar. However, some studies have reported that the archaeal amoA abundance was not pH-sensitive or negatively correlated with soil pH (Hu et al. 2014; Nicol et al. 2008; Ying et al. 2017; Yu et al. 2021). This indicates that pH changes caused by biochar effects are mainly attributed to changes in AOB abundance and community. Although biochar contains a small amount of NH4+, which is the substrate for nitrification, the concentration of NH4+ in biochar is approximately equivalent to soil NH4+ concentration (Enders et al. 2012), except for biochar derived from high-ammonium feedstocks. We assume that biochar addition does not change soil NH4+ availability and not strongly affect soil nitrification. Furthermore, due to the high adsorption capacity, biochar may adsorb NH4+ from soils when N fertilization is applied, and reduce soil NH4+ availability for nitrifiers compared to soils without biochar addition. This assumption has been verified by Yang et al. (2015) and Wang et al. (2015), reporting that biochar significantly decreased NH4+-N availability and thus decreased nitrification and the abundance of ammonia-oxidizing microorganisms, regardless of biochar and soil type (Wang et al. 2015; Yang et al. 2015).

Fig. 4
figure 4

The effects of biochar on soil nitrification, denitrification and P solubilization processes and the underlying mechanisms. “+”and “−”represent the stimulatory and inhibitory effects of biochar on nutrient transformation processes, respectively. “variable” and “/”represent the unknown and no effects of biochar on nutrient transformation processes, respectively

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Studies of the biochar effects on denitrification rates and denitrifier-mediated N2O emission are inconsistent, with the increased denitrification in some studies but decreased denitrification in others. Such discrepancies can be explained by the differences in labile C content and electrochemical properties of biochars used in different studies (Fig. 4). A meta-analysis showed that biochar overall increased the abundance of denitrification process and relevant genes of nirK, nirS and nosZ (Xiao et al. 2019), which was probably due to the amount of labile C in biochar. The labile C could be directly utilized by N-cycling heterotrophic denitrifiers. In addition, the labile C may stimulate the priming effects on decomposition of soil organic C, leading to a quick consumption of O2 and formation of anaerobic localized sites in soil (Harter et al. 2014). Both processes can increase soil denitrification and hence N2O emission. Dai et al. (2019) showed that biochars with high labile C contents stimulated soil denitrification and had greater stimulatory effects on the abundances for denitrification genes, such as nirK, nirS, and nosZ and also N2O emission, compared with biochar with low labile C. The close positive relationships between N2O emission and biochar-increased dissolved organic C (Feng et al. 2018) and biochar volatile matter (Subedi et al. 2016) also indicated the C–N coupled cycling in soil denitrification and N2O emission. As the labile C in biochar increased with decreasing pyrolysis temperature, we propose that the biochars produced at low pyrolysis temperatures (e.g., 300 °C) contribute to denitrification and N2O emission to a greater extent than those produced at high temperatures. By contrast, due to its electrochemical properties, biochars, especially produced at high pyrolysis temperature (e.g., 700 °C), were shown to reduce soil N2O emission (Kluepfel et al. 2014), as these biochars showed the greater electron-accepting capacities. Therefore, biochars at high pyrolysis temperatures can compete for electrons with NO3− during denitrification, whereas biochars produced at low pyrolysis temperatures act as electron donors and provide the electrons for NO3− reduction. This was also supported by the stimulated N2O emission under supply of low-temperature biochar (e.g., 300 °C) (Dai et al. 2019) and the inhibited N2O emission associated with the application of high-temperature biochar (e.g., 700 °C) (Harter et al. 2014, 2016). Although NO3− is the substrate for denitrifiers, due to the low content of NO3− in biochar and the negative charges in biochar surface, the NO3− in biochar does not significantly affect denitrification. Overall, the biochar with lower labile C content and higher electron-accepting capacity, usually produced at high pyrolysis temperatures, would induce low denitrification and N2O emission (Fig. 4). Here, we pointed out that the labile C of biochars mainly affected soil N2O production while the electron-accepting capacity stimulated the conversion of N2O to N2, in particular in anaerobic conditions. This indicated that biochar produced at high pyrolysis temperatures decreased the N2O/N2 emission ratio (Harter et al. 2014, 2016; Dai et al. 2019).

5.2 Soil microbial-mediated P cycling

Microorganisms participate in soil P cycling. They can solubilize inorganic P or release enzymes to mineralize organic P, and finally increase soil P availability (Khan et al. 2014). Biochar can affect soil P cycling through (1) directly providing labile P nutrient, especially for manure biochar (Novak et al. 2014); (2) improving soil P retention capacity to reduce P leaching (Liu et al. 2019; Yang et al. 2021) and (3) affecting the activities and communities of P-solubilizing microorganisms. In this section, we discuss the effects of biochar on soil microbial communities and in turn the feedbacks of altered microbial communities to (1) soil P availability and (2) soil P fractions (Fig. 2).

Biochar can increase soil P availability via altering microbial communities, as it can provide the suitable growth condition (i.e., porous habitat and C supply as discussed above) for microorganisms (Fig. 4). For instance, the addition of rice-husk biochar enhanced soil P availability (Olsen-P) and the activities of related enzymes by enhancing the growth of P-solubilizing bacteria, such as Thiobacillus, Pseudomonas, and Flavobacterium, in a forest soil (Zhou et al. 2020). Leaf biochar increased the abundances of P-solubilizing bacteria, such as Burkholderia-Paraburkholderia, Planctomyces, and Singulisphaera, thus boosting P availability in forest soils (Zhou et al. 2020). These changes may be attributed to increased soil pH and water-holding capacity (Liu et al. 2017). Furthermore, Gao and DeLuca (2020) indicated that biochar shifted microbial communities from bacterial to fungal communities while fungi played an important role in solubilizing soil-fixed P and interactions with plant roots to enhance P acquisition. The changes in fungal communities and soil P availability were highly associated with biochar-induced pH increase (Gao and DeLuca 2020). Regardless of P status in soil, biochar can increase soil total P concentrations and dithionite-citrate-bicarbonate (DCB)-extractable P, leading to the improved growth of crops (Xu et al. 2019). Further investigation should be focused on the contribution percentages of these properties to the increase of soil P availability via changing microbial activities and communities.

Biochar can change soil P fractions and increase the labile P fraction by altering microbial communities and relevant enzyme activities (Fig. 4). For instance, a microcosm experiment revealed that manure-based biochar changed soil P fractions by increasing soil orthophosphate and pyrophosphate and decreasing those of monoesters, in additional to increasing soil P availability (Jin et al. 2016). Biochar decreased the activity of acid phosphomonoesterase but increased the activities of alkaline phosphomonoesterase (responsible for mineralization of organic P), indicating that biochar may affect related microorganisms and their released enzymes by increasing soil pH (Jin et al. 2016). Xu et al. (2018a) showed that biochar application significantly increased fractions of H2O-soluble and NaOH-extractable inorganic P, and NaHCO3-extractable organic P, but decreased the NaHCO3‐extractable inorganic P fraction. This indicated that biochar addition assists microorganisms to solubilize the fixed P and increases microbial P immobilization. The P immobilization is transitory and can be released after microbial mortality.

6 Conclusion and perspectives

The application of biochar to soils has great potential to influence microbial communities and relevant functional processes and hence nutrient supply and plant growth enhancement. Biochar affects soil microbial growth, activity and communities by directly providing growth promoters (e.g. substrates and porous structures) for soil biota or indirectly changing soil basic properties (e.g. pH). The porous structure, labile C, high pH and electrochemical properties of biochar dominantly affect soil microbial abundance, communities and their mediated N and P cycling processes. Due to the high variances in biochar basic properties determined by pyrolysis temperature and feedstock type, the selection and application of biochar to alter microbial communities and mediate nutrient cycling should follow the specific agricultural demands. However, some key aspects still require further investigations (Table 2).

As the rapid development of high-throughput sequencing techniques, such as 16S and ITS sequencing, a large number of literatures report the detailed changes in microbial taxonomic and phylogenetic communities caused by biochar. However, the microbial abundances are relative and only reveal the potential or dormant microbial communities. We recommend the multi-omics (combing meta-transcriptomics, meta-proteomics and meta-metabolomics) to investigate the active microbial communities. Alternatively, the improved methods of 16S sequencing (by adding an internal standard of Aliivibrio fischeri) (Smets et al. 2016) or an internal standard strain (ISS) HAAQ-GFP (Yang et al. 2018) that can obtain the absolute abundance of microbial taxa are recommended.

The effects of biochar on soil virus and protist community are poorly understood, while virus and protists affect microbial bacterial and fungal communities by top-down control (Li et al. 2019). Soil viruses influence the ecology of soil biological communities by transferring genes, cause microbial mortality and consequently participate in soil biogeochemical cycles (Hurst et al. 1980). Soil protist communities are the predators of bacteria and fungi, and their communities are also affected by environmental perturbation. However, the interactions between communities of virus, protists, fungi and bacteria in microbial food-webs are still poorly understood. Recently, microbial network, especially the time-varying networks, has been reported to be a useful tool to capture microbial interactions during microbial community succession (Faust et al. 2015). Given the microbial key nodes in the networks, we can detect the responsive microbial keystones in biochar-amended soils and try to control the growth of these keystones. As the accurate predictions of keystones from microbial networks are still contradictory (Berry and Widder 2014; Weiss et al. 2016), combining the classic experimental validation of keystones with time-varying networks is recommended.

Most studies have focused on examining the effects of biochar on soil indigenous microbial communities, while the microbial communities, dominant taxa and their functions in biochar particles and char-sphere still require more investigations. The microorganisms that colonize on biochar particles participate in the biochar mineralization and affect biochar oxidation. They interact with soil indigenous microorganisms and consequently contribute to soil biogeochemical cycling (Dai et al. 2017a, b). The char-sphere is defined as the unique zone surrounding biochar particles with the properties, such as pH, porosity and nutrient status, differing from those of the bulk soil and the biochar particles (Pei et al. 2017; Quilliam et al. 2013). Investigation of the microbial communities and dominant taxa in biochar particles and along the char-sphere would provide a comprehensive understanding of microbial taxonomic and functional distributions from biochar particles to adjacent soils.

Previous studies have mainly focused on microbial-mediated individual nutrient cycling processes. In soil ecosystems, C, N and P cycling are usually integrated, as the C:N:P stoichiometry in microbial biomass are relatively constrained to a narrow range (Cleveland and Liptzin 2007). The environmental disturbances such as biochar addition may disrupt the coupling of microbial C, N and P cycling by changing soil nutrient statue. For example, the enrichments of C substrates may induce microbial demand for N, which leads to increased mineralization of organic N or acquirement of inorganic N from soils, or increased the mineralization of organic P (Mooshammer et al. 2014). Therefore, we point out the importance of the investigation of C, N and P coupled functional genes or processes in biochar-amended soils. This provides novel insights into soil nutrient management and plant growth after biochar amendments.

The alteration in soil microbial communities caused by biochar is linked tightly with agricultural food production. Plant rhizosphere is a special region where the complex interactions between plant root and microorganisms occur. This region provides a hotspot of microbial biomass and enzymatic activities (Mendes et al. 2013). The beneficial microorganisms including N2-fixing bacteria, growth-promoting bacteria (PGPR) and arbuscular mycorrhiza can enhance plant defense and nutrient uptake capacity, and finally increase soil productivity (Lugtenberg 2015). However, how to build a healthy microbiome in the rhizosphere, control the pathogen and regulate the PGPR promotion or arbuscular mycorrhizal colonization still needs further investigation.