Introduction

Climate change and food demand have become urgent concerns in recent years. Hence, how to mitigate greenhouse gas (GHG) emissions and increase agricultural productivity simultaneously by maintaining soil organic matter (SOM) and improving soil quality have now become the key considerations for agricultural scientists (Agegnehu et al. 2016; Mandal et al. 2016; Yoo et al. 2016; Zhang and Ok 2014; Zhang et al. 2016). Recently, the impacts of biochar application on the emission of GHGs have been addressed in an increasing number of studies (Awad et al. 2016a; Bolan et al. 2012; Chen et al. 2014; Zhang and Ok 2014). In addition, Mekuria et al. (2014) insisted that compost can be effective in enhancing SOM and agricultural productivity. Both biochar and compost have the positive effects on soil fertility and agricultural productivity as well C sequestration (Hussain et al. 2015).

Biochar is a solid product derived from biomass pyrolysis such as agriculture and forestry residues (Jien et al. 2015; Lu et al. 2014). Furthermore, biochar is a porous material containing high levels of C and various functional groups (Ahmad et al. 2016). Aged biochar is known to have higher cation exchange capacity (CEC) than fresh one as the result of weathering (Cheng et al. 2006; Cheng et al. 2008). On the basis of these specific properties, the addition of biochar to agricultural soils has emerged as a feasible strategy for enhancing soil water-holding capacity (Basso et al. 2013; Busscher et al. 2010; Kammann et al. 2011), effectively improving soil quality (Jien and Wang 2013; Vaccari et al. 2011; Yuan et al. 2011; Zhao et al. 2014), stabilize SOM and nutrient retention (Clough et al. 2013; Ventura et al. 2013; Awad et al. 2012, 2013), sequester organic carbon (OC) (Laird 2008), and reduce greenhouse gases emissions (Stewart et al. 2013; Zheng et al. 2012; Woolf et al. 2010; Gaunt and Lehmann 2008; Kang et al. 2016a, b).

In the past 5 years, the co-application of biochar and organic amendments has been recognized to be an effective management method for improving the effects of the separate, individual application of each amendment (Abujabhah et al. 2016). Liu et al. (2012) reported that co-application of biochar with compost had the synergistic positive effect on SOM content, nutrients levels, and water-storage capacity. In addition, the co-application of biochar with compost has been shown to improve plant growth, nutrient uptake, plant health, increasing the yield of the maize, peanut, and grape (Agegnehu et al. 2015; Schmidt et al. 2014; Hussain et al. 2015). According to the studies, the major role of compost is to increase SOM and the major role of biochar is to stimulate microbial activities, facilitating the decomposition of SOM and stabilizing soil properties. However, several studies with inconsistent findings including facilitation and inhibition of SOM decomposition in soils amended with biochar and organic amendments have been published (Fernández et al. 2014; Qayyum et al. 2014). Furthermore, most of these previous studies have been conducted in arid, Mediterranean, and temperate regions, but not in tropical or subtropical regions. In tropical or subtropical regions, high annual rainfall (> 2000 mm) and temperature (> 20 °C) usually result in acidic soils and the soils with less fertility, and these degraded soils might be amended by biochars.

Except for stabilization of SOM, biochar incorporation may induce N immobilization in soils. Few studies have reported that N immobilization occurred in biochar-amended soils (Lehmann et al. 2003, 2015; Rondon et al. 2007). The sorption of SOM might produce a protective mechanism and decrease N availability. Chen et al. (2014) indicated that short-term N immobilization occurred throughout adsorption of N on the surface of the biochar by a laboratory incubation experiment. The N immobilization after biochar application seems reduce plant N take in the experiment conducted by Chen et al. (2014). On the contrary, Mandal et al. (2016) displayed that co-application of biochar with fertilizer has been shown to increase plant N uptake due to the enhanced retention of NH4 by biochar. Also, Van Zwieten et al. (2010) reported that co-application of biochar and fertilizer increased wheat yield up to 250% with increased N uptake. However, the mechanisms of biochar on N immobilization were not fully understood yet.

From practical viewpoint, compost as a common organic fertilizer increases the contents of OC and nutrients, contributing to improved soil physicochemical and biological characteristics (Rehman et al. 2016; Teutscherova et al. 2017). In contrast, there is few addition of nutrients (N, P, and K) by biochar application to soil compared with the compost application; however, it can effectively change nutrient transformation and microbial composition (Zhang and Ok 2014; Teutscherova et al. 2017). Therefore, we hypothesized that co-application of biochar and compost can have the synergistic positive effect on physiochemical properties of soil, plant nutrient uptake, greenhouse gas emission, and soil microbial activity (Awad et al. 2012; Rehman et al. 2016). To clarify the interaction between biochar and compost, the major objective of this study is to determine the effects of rice husk biochar (RHB) and bagasse compost application on CO2 emission and nitrogen immobilization in a tropical agricultural soil by performing a short-term (56 days) incubation experiment. The results of this study should provide some useful information regarding the management of croplands by using biochar and compost application.

Materials and methods

Soil collection and biochar production

Surface soil samples (0–15-cm depth) were collected from a cultivated land (22° 31′ 44.9″, 120° 29′ 35.9″) in Pingtung County, Taiwan. Choutseulun (Ct) soil is the slate alluvial sediment along streams in southern Taiwan belonging to a major soil type which occupied approximately 50% of the area of rural soils in Pingtung County, southern Taiwan. The soil samples were air-dried, sieved through a 2-mm screen, and stored in covered plastic containers at 25 °C. The soil could be classified as Typic Eutrudepts by the Soil Taxonomy (Soil Survey Staff 2014).

The biochar used in this study was produced from rice husks, which is a major agricultural waste produced 0.3 million tons per year in Taiwan. The biochar in this study was supplied by the Industrial Technology Research Institute (ITRI) of Taiwan. Before charring, the rice husk was dried at 60 °C for 24 h to <10% moisture and cut to a particle size of 2 cm. For pyrolysis, the samples were placed in a tubular furnace (ITRI, Tainan, Taiwan) equipped with a corundum tube (32 mm, diameter, 700 mm, length) with a N2 purge (1 L/min flow rate) to ensure an oxygen-free atmosphere. Heat treatments were performed in the temperature of 700 °C. The heating rate was 5 °C min−1. The temperature was maintained for 2 h before cooling to an ambient temperature under N2 flow. The produced rice hull biochar was denoted as RHB-700 later in this study. After pyrolysis, the biochar was ground to enable it to pass through a 2-mm sieve, ensuring that all the biochar used in the experiments exhibited similar particle sizes.

Analytical methods

The pH of the soil samples and the biochar were determined in a mixture with deionized water (1:1 w/v for soil; 1:10 w/v for biochar), using a glass electrode (Sparks et al. 1996). Electrical conductivity (EC) was measured on the saturation paste extracts of soils, using a conductivity meter (Rhoades 1982). Soil particle-size distribution was determined with the pipette method (Gee and Bauder 1986). Cation exchange capacity (CEC) was determined using the ammonium acetate method (pH 7.0) (Sumner and Miller 1996). Organic carbon content was determined by the wet oxidation method (Nelson and Sommers 1996). Total N was measured using the Kjeldahl procedure (Bremner and Mulvaney 1982). Inorganic N was extracted with 2 M KCl (1:10 w/v); the concentrations of NH4 +–N and NO3 −–N were estimated by steam distillation (Mulvaney 1996).

Readily oxidizable carbon (ROC) was determined using a method proposed by Blair et al. (1995). Air-dried soil samples containing 15 mg of C were weighed into centrifuge tubes and reacted with 333 mM KMnO4 for 1 h at 25 °C. After centrifugation, the supernatants were diluted at a ratio of 1:250 with deionized water. The absorbance of the diluted samples and standards was recorded using a split-beam spectrophotometer at 565 nm. The change in the KMnO4 concentration was used to estimate the amount of C oxidized assuming that 1 mM KMnO4 was consumed in the oxidation of 0.75 mM or 9 mg of C. The KMnO4-C fraction, suggested by Blair et al. (1995), encompasses all the organic components that can be readily oxidized by KMnO4, including labile humic material and polysaccharides (Conteh et al. 1997), and accounts for 5–30% of total organic carbon. All chemical analyses were conducted in triplicate. The selected properties of the soil and biochar are summarized in Table 1.

Table 1 Characteristics of the soil, cow manure compost, and the rice hull biochars (RHBs) used in the experiment
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Modified fast wetting in water was used to measure the aggregate stability of 2 mm air-dried aggregates (35 g) (Kemper and Rosenau 1986). A 4-cm amplitude was applied to a nest of sieves (>2000, 1000–2000, 500–1000, 250–500, 250–106, and <106 mm) immersed in a container of tap water (101 mS/cm) during the 5 min of vertical movement. Soil that remained in each sieve after wet shaking was carefully removed, and the mean weight diameter (MWD) of the aggregate size was calculated as follows:

$$ \mathrm{MWD}=\sum_{\mathrm{i}=1}^{\mathrm{n}}{x}_i{w}_i $$
(1)

where n is the number of sieves, x i is the diameter, and w i is the weight. In this study, different sizes, <0.25, 0.25–2, and >2 mm, were used to denote micro-aggregates, meso-aggregates, and macro-aggregates, respectively.

Incubation experiment

Twenty-five grams of each air-dried soil sample was placed in small plastic cups. Commercial cow manure compost (MC) was added as a substrate to the soil for each treatment at a rate of 20 t ha−1. The biochar was then thoroughly mixed with the soils at 0, 2, and 4% (w/w) (0, 40, and 80 t ha−1, respectively). The experimental design consisted of six treatments for each soil with triplicate: (1) control, (2) manure compost (MC), (3) RHB-2%, (4) RHB-4%, (5) RHB-2% + MC, (6) RHB-4% + MC. Deionized water (DI water) was added to the soils to achieve 60% water-holding capacity. Each cup of the treated soil was placed in a wide-mouth plastic jar with a plastic vessel containing 10 mL of 1 N NaOH solution. The jars were then sealed. The jars without treated soils were used as blanks. After 3, 7, 14, 21, 28, 42, and 56 days, the emitted CO2 was measured in a nondestructive set of triplicate by titrating the NaOH solution with 0.5 N HCl before adding an excessive amount of BaCl2. The jars were then sealed again for incubation until the next measurement. The incubation experiment was performed in the dark at 25 ± 2 °C.

Mineralization of carbon and N was studied in separate incubation experiments. The treatments of N incubation experiment were the same with C incubation one. The 200 g soil was filled into each beaker and added DI water to the soils to achieve 60% water-holding capacity. Eighteen beakers in total were put in the dark room with 25 °C, and 5 g soil was sampled at 0, 3, 7, 14, 21, 28, 42, and 56 days, respectively, for inorganic N analysis. Mineralization of N was measured on the basis of changes in inorganic N in the soil during the incubation period. The inorganic N was extracted using 2 mol L−1 KCl (1:10 w/v) under identical incubation conditions and on the same days as the destructive set. The concentrations of NH4 +–N and NO3 − –N were estimated using steam distillation using MgO and the Devarda’s alloy (Keeney and Nelson 1982).

Soil micromorphology

The third batch incubation set was conducted for micromorphological observations. Kubiena boxes (8 cm × 8 cm × 3 cm) were used to collect undisturbed blocks from the tested soils. The same solid mixture of each treatment was placed in a pot with a size larger than the Kubiena box; the incubation process described was then applied. After the 56-day incubation, the soil blocks were collected using Kubiena boxes. Thin sections of 30-μm thickness were then prepared following air drying by using a microtome by Spectrum Petrographics Inc. (Washington, USA). The thin sections were then used for observing the structure and distribution of organic matters among the soil particles under a polarized microscope (Leica DM EP, TX, USA).

Calculations and statistical analysis

A single first-order equation was used to describe C mineralization kinetics. Single first-order kinetic equation is one of the most commonly used equations for organic C mineralization in soils (De Neve et al. 1996, 2003). In this equation, it is assumed that C mineralization is proportional to the amount of mineralizable C at any time (t), it can be written as:

$$ {\mathrm{C}}_t={\mathrm{C}}_0\left(1-{\mathrm{e}}^{\hbox{--} kt}\right) $$
(2)

where C t is the amount of C at time t, C0 is the potentially mineralizable C at time 0, and k is the rate constant.

Statistical analysis

All treatments were conducted in triplicate during incubation. The triplicate data were subjected to mean separation analysis by using a one-way ANOVA test with a significance level of 푃 = 0.05. Significant differences between mean values were identified using Duncan’s test. The statistical analyses were performed using IBM SPSS Statistics, version 22.

Results

Basic properties of soil, biochar, and compost

The studied soil was sandy loam and characterized by slightly acidic and low amounts (<1%) of soil organic carbon (SOC) and CEC (Table 1). Table 1 also displays low total N content (<0.20%), low available P, and exchange K contents in this soil. The pH of the RHB-400 was approximately 8.0. The EC was ≦0.6 dS m−1. The CEC is slightly higher in RHB-400 compared with the soil. The contents of total C and total N in RHB-400 were 67.0 and 0.35%, respectively (Table 1). Table 1 indicates that the exchangeable K in RHB-400 was 7.01 g kg−1, which is higher than the soil. Besides, less ROC was also found in the RHB (3.10 g kg−1). As compared with the RHB, the compost was a fertile amendment, which total C, total N, available P, and exchange K are higher than those in the biochar. The ROC content in compost (22.9 g kg−1) is higher than that in RHB.

Changes of soil properties after biochar incorporation

After applying RHB-400 and compost to the soil and incubating for 56 days, the amended soils had a significantly higher soil pH (at least 0.5 units) than the control samples (Fig. 1). Initially, the highest pH, approximately 7.4, was found in the treatment of RHB-4% + MC; the pH of RHB-2% + MC, RHB-4%, and RHB-2% was approximately 7.1, the pH of MC alone was approximately 6.8, and the control exhibited the lowest pH. With incubation time, the pH gradually deceased for all treatments. All treatments exhibited a pH of approximately 6.4 at the end of the incubation (Fig. 1).

Fig. 1
figure 1

Dynamic changes of pH for all treatments during incubation in this study

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Regarding soil physical properties, the formations of macro-aggregate and meso-aggregate were obviously found as incorporation of biochar and compost after 28 days, and macro-aggregate was grew continually until the end of the incubation (56 days) (Fig. 2). After 56 days, the proportion of macro-aggregate increased by 34.3, 28.6, and 44.7% in the treatments of biochar only, compost only, and co-application (RHB + MC), respectively, which were significantly (p < 0.05) increased compared with the controls.

Fig. 2
figure 2

Variances of proportion of weight in soil aggregate with different sizes at 1, 4, and 8 weeks in this study: a the soil without any amendments (control soil), b the soil amended with only manure compost (1%, w/w), and c the soil amended with manure compost (1%, w/w) and 2% (w/w) RHB

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Carbon dioxide releases in the amended soils

The CO2 evolution rates and cumulative quantity of CO2 released in this study are presented in Fig. 3. For all treatments, the CO2 emission increased to maximum on the third day of incubation and then stabilized at a low level to the end of incubation (Fig. 3a, b). Although the pattern of the CO2 evolution rate was similar among the treatments, the maximum CO2 evolution rates (during the first 3 days of incubation) were markedly higher (p < 0.05) in compost treatments with or without RHB. A low level of CO2 evolution rate was maintained in control, and cumulative CO2 emission also exhibited the lowest amount (225 mg CO2-C kg soil−1) throughout the incubation period compared with other treatments (Fig. 3c). Compared with the control, RHB treatments slightly increased the cumulative CO2 emission amount during the incubation time, and the differences between the control and RHB treatments were statistically significant (p < 0.05) (Fig. 3). The cumulative CO2 emission amounts increased with an increased application rate of biochar (Fig. 3). For compost-added treatments (Fig. 3d), the maximum amount of cumulative CO2 emission was observed in the compost treatment (MC). Meanwhile, the co-application of RHB + MC treatments led to a significant reduction in cumulative CO2 emission over the 56 days of incubation, compared with the MC treatment. The quantity of cumulative CO2 emission with RHB-2% and RHB-4% treatments in the compost-treated soil did not differ significantly (p < 0.05).

Fig. 3
figure 3

CO2 evolution rate (a, b) and cumulative amount of CO2-C evolution (c, d) for the Sp soil amended with compost and biochar. The vertical error bars indicate the standard deviation. O (control): without compost and biochar; O + C: only compost (application rate of 20 t ha−1)

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For further comparison of the degree of CO2 emission among all treatments, the amounts of cumulative CO2 emission are expressed as a percentage of the initial C evolving to CO2 during the incubation (Table 2). The application of MC exhibited the highest proportion of total C mineralized (14%) in the soil (Table 2). The application of biochar produced proportions of only 0.74 and 0.62% of the total C mineralized in soil treated with RHB-2% and RHB4% without the addition of compost, respectively. In comparison, RHB + MC treatments exhibited on an average 3.2% (RHB-2% + MC) and 2.3% (RHB-4% + MC) of total mineralized C in the soil. When compost was applied, the RHB-4% treatment led to a significantly lower proportion of total C mineralized in soil than that of the RHB-4%. Regarding the treatment of co-application (O + RHB + MC), the net evolved CO2 was significantly lower in RHB + MC treatment than the net CO2 evolved in soil treated with RHB and MC.

Table 2 The percent of total C mineralized as CO2 for each treatment at the end of the incubation
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Effect of biochar treatment on soil inorganic nitrogen

The variations in the concentration of inorganic N (NH4 +-N + NO3 −-N) during incubation are illustrated in Fig. 4. The maximum amount of NH4 +-N was observed at 3 days of incubation; thereafter, the NH4 +-N concentrations in the soils declined over time for all treatments (Fig. 4a, b). The application of RHB-2% resulted in slightly higher NH4 +-N concentrations in the soil at 3 days of incubation than did the application of RHB-4%; however, the two treatments of biochar application did not differ significantly. The effects of the biochar application rate on the NH4 +-N concentrations in soil treated with biochar and compost (RHB + MC) were similar to those of biochar application rate on the soil treated solely with biochar.

Fig. 4
figure 4

Concentration of NH4 +-N (a, b) and NO3 −-N (c, d) in the Sp soil amended with compost and biochar over 56 days of incubation. The vertical error bars indicate the standard deviation

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For all treatments, the concentration of NO3 −-N gradually increased during the first 14 days of incubation; thereafter, it gradually decreased as the incubation period progressed (Fig. 4c, d). The NO3 −-N in soil amended with RHB alone did not differ significantly (p < 0.05) from that in the control soil at 56 days of incubation. However, the concentration of NO3 −-N in the soil treated with RHB + MC was significantly (p < 0.05) lower than that of the control treatment at 56 days of incubation.

Discussion

Carbon dioxide emissions from soils amended with the compost and biochar

To mitigate of global warming and land sustainability, the Paris Agreement (2015) proposed agricultural management practices aimed at increasing the amount of C in soils by 0.4% per year and reducing CO2 concentration to 350 ppm. According to our study, the co-application of compost and biochar (RHB + MC) could be a suitable agricultural practice for effectively sequestering SOC and mitigating CO2 emission from farming. In our study, co-applying RHB (2 or 4%) and manure compost (1%) (MC) to the soil significantly reduced CO2 emission by 14.7–19.6% compared with that in the control soil with MC. In contrast to our previous study (Jien et al. 2015), the co-application of RHB and bagasse compost into three soils with different textures notably increased cumulative CO2 emissions. The difference of ROC content in the added fresh organic matters between this study and our previous study (Jien et al. 2015) might be responsible for the different levels of CO2 emission. Culman et al. (2012) mentioned that ROC is a major labile OC in the soils, and which is could be a new and quick method to quantify soil labile OC. In addition, Hassen et al. (2016) further indicated that ROC might be a major part of SOC to dominate CO2 emission in soils. They also provided that the overall effect of labile C fractions on the CO2 emission was in the order light fraction of organic carbon (LFOC) > particulate organic carbon (POC) > readily oxidizable carbon (ROC) > dissolved organic carbon (DOC) > microbial biomass carbon (MBC) > reducing sugar carbon (RSC) > readily mineralicable carbon (RMC) > total organic carbon (TOC). In this study, the ROC content was approximately 36 g kg−1 in the bagasse compost mentioned in Jien et al. (2015), and the ROC was approximately 22 g kg−1 for the cow manure compost in this study.

Regarding mineralized C proportion, the soil with only compost exhibited the highest proportion of total mineralized C, whereas the soil with only biochar exhibited a very low mineralization rate (Table 2). Low mineralization rates of biochar have been previously reported. Hamer et al. (2004) also reported only 0.3–0.8% of C mineralization for oak-wood and crop residue biochars in 60 days of incubation. The microbial degradation of fresh biochar when applied to soil may occur and led to the release CO2 evolution as reported in a study by Teutscherova et al. (2017). Table 2 presents that the “net” CO2 evolved after the co-application of biochar and compost (“RHB + MC”-“control (O)”) was lower than the sum of the net CO2 evolved after biochar only (RHB-O) and compost only (MC-O) treatments, suggesting the stabilizing effect of biochar on fresh organic compost. Awad et al. (2016b) revealed that oak wood biochar exerted a minor effect on SOM mineralization while decreased the decomposition of 14C-labeled alfalfa residues in the soil.

To elucidate the interaction between compost and biochar, the expected values and measured values of cumulative CO2 emission in the treatments involving co-application (RHB + MC) were compared. The expected values were calculated with the values from the applications of only compost and only biochar as follows: RHB + MC-control (O). Therefore, the differences between the expected and measured values could be attributed to the effect of co-application. The single first-order equation was successfully used to predict C and N mineralization (Smith et al. 1980). We also fitted the results into the first-order equation and estimated the maximum mineralizable C pool as illustrated in Fig. 4. The cumulative CO2 emission curves of the measured values are obviously lower than those of the expected values, indicating that the mineralizable C pool (labile C) might have decreased because of co-application of RHB + MC in the soil. An interactive mechanism between biochar and compost were presented in this study, and this deduction is also supported by our micromorphological observations (Fig. 5), in which the added compost was clearly adsorbed on the biochar. Our results for the stabilization of C mineralization of added compost in the presence of biochar are consistent with the findings of Keith et al. (2011), who also observed the stabilization of labile organic matter decomposition through its rapid incorporation into organomineral fractions in biochar-treated soil. We supposed that the co-application of biochar with composts may be a more favorable means to sequester C in soils than the application of only biochar.

Fig. 5
figure 5

The expected and measured cumulative CO2 emission in the Sp soil amended with the compost at the rate of 20 t ha−1 and biochar at the rates of 2 and 4%. Expected values were calculated from the treatments of control (O), compost only (O + C), and biochar only (O + RHB). Fitted max: the maximum value estimated via first-order kinetic model fitting

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Inorganic N release from soils amended with compost and biochar

In addition to mitigating CO2 production, biochar had a notable effect on the rate of N turnover in the soil. For all treatments, the NH4 +-N concentration became negligible within the first 7 days of incubation (Fig. 4a, b). The volatilization of NH3 in agricultural soils is common, particularly at alkaline pH and high concentrations of NH4 +-N (Robertson and Groffman 2007). The soil pH significantly increased with biochar addition because of the alkaline minerals in the biochar but not to a level (pH > 8) that would result in considerable NH3 volatilization from the soils. Thus, ammonia volatilization in our incubation experiment should be negligible. This study consistently indicated that the NH4 +-N concentration rapidly decreased after 7 days of incubation; the increased NO3 −-N concentration at the beginning of the incubation suggested that the NH4 +-N was rapidly transformed to NO3 −-N. The macro-aggregate formation after the RHB application may cause higher N nitrification than control soil because of the improvement of soil physical properties and therefore enhanced oxidation of NH4 +-N to NO3 −-N (DeLuca et al. 2006). This finding also corresponded with the declining pH, indicating nitrification during incubation. However, the increased NO3 −-N concentrations from biochar treatments with and without compost were temporary; the reduction of NO3 −-N occurred after 21 days and persisted throughout the 56-day incubation period (Fig. 4c, d). Our result was also agreed with Nelson et al. (2011) who showed that biochar application decreased NO3 −-N recovery by 5–10 mg kg−1 after 56-day incubation, and they also suggested that additional N supply is necessary in biochar-amended soils.

In this study, the decrease in the amounts of NO3 −-N with an increase in biochar quantity could be attributed to N immobilization, which might induce from microbes and physical protection of SOM by the biochar in this study. However, Chen et al. (2014) suggested that microbial immobilization might not occur in biochar-amended soil, because biochar itself contains minimal labile C. In our study, the ROC content of biochar was relatively low (3.1 g kg−1) (Table 1). Chen et al. (2014) indicated that microbial N immobilization may not have occurred in biochar treatment because biochar contains little available C. They further suggested that the N immobilization in the biochar-added soil was likely caused by the biochar sorption over native SOM (Cross and Sohi, 2011; Zimmerman et al. 2011). Chen et al. (2014) and Jien et al. (2015) indicated that the sorption over native SOM could produce a protective mechanism for reducing the mineralization of native SOM and could cause the negative N release. Biochars pyrolyzed from higher charring temperatures are characterized by high adsorption over SOM because of their highly porous structure and positive charge (Cheng and Lehmann 2009). Our observations, made using a polarized microscope (Fig. 5), support this inference. The added compost was visibly adsorbed around the biochar or protected by the aggregates formed by soil–biochar–compost, suggesting that it is unsusceptible to soil microbes. Chen et al. (2014) indicated another possible factor responsible for the induction of N immobilization, namely a toxic effect inhibiting the degradation of SOM. However, this toxic effect appears more frequently in low-temperature (<300 °C) biochars than in high-temperature biochars (Spokas et al. 2012), and is, therefore, less supportive of the results of this study.

Even so, we could not still exclude the induced N immobilization by microbes, which Fig. 4 shows that mineralized NO3 −-N increased in the co-applied treatments compared with treatments of the control and MC-amended only before 21 days. Obvious decline of NO3 −-N contents occurred only on co-applied treatments after 28 days indicating possible microbial N immobilization. However, this study concluded that inorganic N contents reduced as co-application of the biochar and the cow manure compost after a short-term incubation, which might be resulted from adsorption of the biochar and physical protection of the compost by soil aggregates or biochar rather than microbial N immobilization (Figs. 2, 4d, and 6). Many studies revealed that applying fresh compost with high C/N ratio to the soil can lead to immobilization of N (Benito et al. 2005: Teutscherova et al. 2017).

Fig. 6
figure 6

Microstructural observations in the biochar- and compost-amended soil by using a polarized microscope (plane polarized light (PPL))

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Conclusions

CO2 emission amounts and mineralized N after the application of RHB + MC decreased in tropical cultivated soils compared to soil treated with only compost. Biochar application can benefit to C sequestration in the rural soils in Taiwan through mutual interaction of biochar and organic matter including natural SOM and added compost. The further in situ experiments should be conducted in other subtropical or tropical regions to confirm the mechanism. Beside, this study further indicated obvious N immobilization in the soil amended with the biochar and manure compost simultaneously, which might prevent rapid decomposition of the compost from microbes, and then elongate N supplies for crops in subtropical and tropical regions. Our research provided an insight into the interactions between biochar and added manure compost, and proved an evidence for C sequestration in the biochar-amended soil. However, further research is needed to determine the priming effects and identify specific microbial groups and their contributions to biochar-amended soil, particular in subtropical and tropical regions.