Dynamics of CO2 Emission and Biochemical Properties of a Sandy Calcareous Soil Amended with Conocarpus Waste and Biochar

Pedosphere 25(1): 46–56, 2015 ISSN 1002-0160/CN 32-1315/P c 2015 Soil Science Society of China ° Published by Elsevier B.V. and Science Press

Dynamics of CO2 Emission and Biochemical Properties of a Sandy Calcareous Soil Amended with Conocarpus Waste and Biochar Mohamed EL-MAHROUKY1 , Ahmed Hamdy EL-NAGGAR1,2 , Adel Rabie USMAN1,3,∗ and Mohammad Al-WABEL1 1 Soil Science Department, College of Food & Agriculture Sciences, King Saud University, P. O. Box 2460, Riyadh 11451 (Saudi Arabia) 2 Department of Soil Science, Faculty of Agriculture, Ain Shams University, P.O. Box 68, Hadayek Shobra, Cairo 11241 (Egypt) 3 Department of Soils and Water, Faculty of Agriculture, Assiut University, Assiut 71526 (Egypt)

(Received April 25, 2014; revised September 8, 2014)

ABSTRACT Biochar is a carbon-rich product obtained by biomass pyrolysis and considered a mean of carbon sequestration. In this research, a sandy calcareous soil from the Farm of the College of Food & Agriculture Sciences, King Saud University, Saudi Arabia, was amended with either woody waste of Conocarpus erectus L. (CW) or the biochar (BC) produced from CW at rates of 0 (control), 10, 30 and 50 g kg−1 . The effects of the amendments on soil pH, dissolved organic carbon (DOC), microbial biomass carbon (MBC), CO2 emission and metabolic quotient (qCO2 ) of the sandy calcareous soil were studied in a 60-d incubation experiment. The results showed that the addition of CW led to a significant decrease in soil pH compared to the control and the addition of BC. The CO2 -C emission rate was higher in the first few days of incubation than when the incubation time progressed. The cumulative CO2 -C emission from the soil amended with CW, especially at higher rates, was higher (approximately 3- to 6-fold) than that from the control and the soil amended with BC. The BC-amended soil showed significant increases in CO2 -C emission rate during the first days of incubation as compared to the non-amended soil, but the increase in cumulative CO2 -C emission was not significant after 60 d of incubation. On the other hand, CW applications resulted in considerably higher cumulative CO2 -C emission, MBC and DOC than the control and BC applications. With the exception of 0 day (after 1 h of incubation), both CW and BC applications led to lower values of qCO2 as compared to the control. The power function kinetic model satisfactorily described the cumulative CO2 -C emission. Generally, the lowest values of CO2 emission were observed in the soil with BC, suggesting that the contribution of BC to CO2 emission was very small as compared to that of CW. Key Words:

dissolved organic carbon, metabolic quotient, microbial biomass carbon, power function kinetic model, soil pH

Citation: El-Mahrouky, M., El-Naggar, A. H., Usman, A. R. and Al-Wabel, M. 2015. Dynamics of CO2 emission and biochemical properties of a sandy calcareous soil amended with Conocarpus waste and biochar. Pedosphere. 25(1): 46–56.

Many agricultural areas in the arid or semi-arid regions have light-textured sandy soils, which are characterized by high CaCO3 and pH and low organic matter content. Soil additives such as organic amendments are needed to improve soil functions and thus increase soil fertility and productivity (Goyal et al., 1993; Pascual et al., 1997; Usman et al., 2004a). Organic amendments can enhance soil properties such as structure, water-holding capacity and nutrient status. However, incorporation of organic amendments leads to an increase in carbon dioxide (CO2 ) emission from soil as a result of rapid soil organic matter decomposition. Recently, it has been suggested that transforming organic materials to biochar is considered a mean for carbon sequestration and a successful strategy for mitigation of global warming and greenhouse gases (GHG) emissions (Lehmann et al., 2006, 2009). Biochar is pro∗ Corresponding

duced by pyrolysis of organic materials. Due to its aromatic structure, biochar is chemically and biologically more stable than the initial biomass (Lehmann et al., 2009). Recent studies showed that, in addition to reducing emission of CO2 , applying biochar to the soils increased their pH, cation exchange capacity (CEC), and level of highly stable organic carbon (OC) (Novak et al., 2009; Hossain et al., 2010). It has also been reported that biochar application to soil can improve soil fertility and quality as well as plant growth (Chan et al., 2008; Van Zwieten et al., 2010; Yuan et al., 2011). Ibrahim et al. (2013) showed that biochar application improved soil water-holding capacity and significantly increased water-stable aggregates. As reported in the literature, the variations in biochar impact on CO2 emission may be attributed to variations in biochar raw feedstocks, pyrolysis con-

author. E-mail: [email protected]; [email protected].

CO2 EMISSION AND PROPERTIES OF BIOCHAR-AMENDED SOIL

ditions, biochar composition, biochar application rate and soil type (Kolb et al., 2009; Spokas and Reicosky, 2009; Luo et al., 2011; Case et al., 2012). In Saudi Arabia, Conocarpus erectus L. (button mangrove) is an evergreen tree with hard wood and huge aboveground biomass. Transforming Conocarpus waste biomass into biochar may be considered as a potential alternative tool for waste management and recycling (Al-Wabel et al., 2013). On the other hand, there is currently no information available about the effect of biochar produced from Conocarpus waste on CO2 emission and biochemical properties of high-pH sandy calcareous soils of arid regions. The objective of this research was: i) to characterize biochar produced from Conocarpus waste and ii) to study the effects of increasing application rates of Conocarpus waste and its biochar on the dynamics of CO2 -C emission and changes in biochemical properties of a sandy calcareous soils. MATERIALS AND METHODS Conocarpus waste and biochar characteristics Conocarpus waste (CW) was collected from the campus of King Saud University, Saudi Arabia, dried and then chopped to small pieces. The wood pieces were pyrolyzed in a stainless steel cylinder container (30 cm radius, 60 cm length) in an outdoor pyrolysis reactor for 150 min at a temperature of 400 ± 10 ◦ C. The pH values of CW and the biochar (BC) produced were measured at a ratio of 1:10 (Al-Wabel et al., 2013). Total C, H and N were measured by a CHN analyzer (series II, Perkin Elmer, USA). Their surface structure was investigated using a scanning electron microscope (SEM) (Inspect S50, EFI, Netherlands). X-ray diffraction analysis (XRD-7000, Shimadzu, Japan) was also carried out to identify any crystallographic structure in the CW and BC samples. Moreover, the Fourier transformation infrared analysis was applied using a Nicolet 6700 Fourier transform infrared (FTIR) spectrometer (Thermo Electron Scientific Instruments Corporation, USA) in the wavenumber range of 500–4 000 cm−1 to characterize the surface functional groups for CW and BC. Incubation experiment The soil used for the incubation experiment was sampled at the Farm of the College of Food & Agriculture Sciences, King Saud University, Saudi Arabia. The soil samples were air dried and ground to pass through a 2-mm sieve. The physico-chemical properties of the soil were measured according to standard me-

47

thods (Sparks, 1996). Soil particle size distribution was determined by the pipette method (Gee and Bauder, 1994). Soil pH was measured using a glass electrode in a saturation paste. The electrical conductivity (EC) was measured in the extracts of soil saturation paste. Soil calcium carbonate content was determined using a calcimeter. The soil organic matter content was measured according to Nelson and Sommers (1996). The CW and BC samples were ground to fine powder and added to the soil at four rates of 0 (control), 10, 30 and 50 g kg−1 as treatments. The soil mixtures (100 g) were put in glass vessels (250 mL) and ultrapure water was added to adjust each soil mixture to 70% of field capacity before incubation. A soil without addition of CW and BC was also incubated as the control. For CO2 emission determination during the incubation, triplicate small vials with 10 mL of 1 mol L−1 NaOH solution were placed in vessels to trap CO2 (Black, 1965). After the addition of NaOH, the vessels were closed air-tight and incubated for 60 d at 30 ◦ C. Three vessels containing only small vials with 10 mL of NaOH solution without soil were used as blank. The vessels were opened after 0 (1 h), 3, 7, 15, 30, 45 and 60 d to replenish the NaOH solution for CO2 trapping so that the mineralization process was not inhibited by a lack of O2 . CO2 emission during the incubation was trapped in 1 mol L−1 NaOH and the excess NaOH was titrated with 0.1 mol L−1 HCl after addition of BaCl2 (Black, 1965). CO2 -C emission rate (µg C g−1 soil d−1 ) during incubation was calculated using the following equation: CO2 -C emission rate = (T − V )− (T − B)(NE /W )d

(1)

where T is the total volume (mL) of NaOH at the start of incubation; V is the volume (mL) of HCl to titrate NaOH in treatment; B is the volume (mL) of HCl to titrate NaOH in blank; N is the normality of the HCl to titrate NaOH; E (E = 6) is the equivalent weight of carbon; W is the soil weight and d is the days between every two sampling times. For the determination of soil microbial biomass C (MBC), pH and dissolved organic carbon (DOC), each treatment was replicated twenty one times. Three replicates of each treatment were sampled after 0 (1 h), 3, 7, 15, 30, 45 and 60 d of incubation to measure the soil MBC, pH and DOC. The bottles were also opened at each specific sampling time to avoid a lack of O2 . Soil MBC was measured by the fumigation-extraction technique (Vance et al., 1987). At each sampling time, the soil samples (5 g) of each treatment were

48

M. EL-MAHROUKY et al.

cations are reported. Statistical analysis was performed using the software Statsoft (StatSoft Inc., 1995). Differences of means between treatments were tested by analysis of variance (ANOVA) and subsequent posthoc comparisons of means using least significance difference (LSD) test at P = 0.05.

fumigated with ethanol-free chloroform. Then, the soil samples were extracted with 0.5 mol L−1 K2 SO4 for 30 min. The non-fumigated soil samples were extracted similarly. Microbial biomass C was calculated as follow: MBC = (OC extracted from fumigated soils − OC extracted from non-fumigated soils) ÷ kEC (2)

RESULTS AND DISCUSSION

where the kEC (kEC = 0.38) is a calibration factor. The metabolic quotient (qCO2 ) a specific respiration activity parameter, was also calculated and expressed as the CO2 -C produced per unit of MBC and per unit of time (mg CO2 -C h−1 g−1 MBC) (Anderson and Domsch, 1993). At each sampling time during incubation, the soil was destructively sampled and soil pH was measured with a glass electrode using a soil-to-water ratio of 1:1. The extractable forms of C (MBC and DOC) in the soil were measured by an oxidation with K2 Cr2 O7 and a subsequent back-titration of the unreduced dichromate (Vance et al., 1987).

Characteristics of soil, CW and BC used The soil used had a particle size distribution of 90.26% sand, 5.76% silt and 3.98% clay. The pH of the soil was 7.88 and the EC was 3.98 dS m−1 . The OC and CaCO3 contents were 1.4 and 101.7 g kg−1 , respectively. The raw feedstock CW had an acidic pH of 5.26, whereas the biochar produced had an alkaline pH of 9.61 (Table I). The C, O, H and N contents all increased in the BC samples (761.8, 186.7, 25.3 and 4.2 g kg−1 ), as compared to those in the raw feedstock samples (449.6, 458.2, 54.1 and 6.2 g kg−1 ). The atomic ratios of O/C, H/C and (O+N)/C were also calculated in this study (Table I). These ratios were lower in the BC samples than in the raw feedstock samples, which are in agreement with the results of Al-Wabel et al. (2013). The decline in the O/C ratio in the BC samples was attributed to dehydration reactions, which resulted in less hydrophilic biochar surface (Ahmad et al., 2012; Al-Wabel et al., 2013). In addition, the lower H/C ratio in BC suggested that the BC samples had a lower content of organic substances, which were highly carbonized and had a higher aromaticity. As it is expected, the higher (O+N)/C ratio of the CW samples indicated that CW contained higher polar functional groups than the biochar samples, suggesting an increase in aromaticity and a decrease in polarity of biochar. The FTIR analysis of the raw feedstock showed a high peak at about 3 400 cm−1 , suggesting the presence of O–H stretching and strong hydrogen bonding in the raw feedstock (Fig. 1). In addition, appearance of some peaks at 700–1 600 cm−1 suggested the presence of cellulosic and ligneous constituents in the raw feedstock. Vibrations lie in the range of 1 500–1 640 cm−1 , which were mainly responsible for lignin in the raw

Kinetic models Four different kinetic models, first order, second order, Elovich and power function models, were applied for describing the cumulative CO2 -C emission from the sandy calcareous soil with or without CW and BC (Sparks, 1989). Their linearised forms are given in Eqs. 3–6, respectively: lnCt = lnC0 − kt

(3)

1/Ct = 1/C0 + kt

(4)

Ct = (1/k)ln(Co k) + (1/k)lnt

(5)

lnCt = lnCo + klnt

(6)

where Ct is the amount of carbon mineralized or the remaining substrate carbon concentration after incubation time t (min); Co is the potentially mineralizable carbon or the initial substrate carbon concentration at t = 0 and k is the mineralization or rate constant. Statistical analysis The means and standard deviations of three repliTABLE I

Characteristics of biochar (BC) and the raw feedstock Conocarpus waste (CW) used as soil ammendments Amendment CW BC

pH 5.26 9.61

C 449.6 761.8

O g 458.2 186.7

H

N

O/C

H/C

(O+N)/C

54.1 25.3

6.2 4.2

1.02 0.24

0.12 0.03

1.03 0.25

kg−1

CO2 EMISSION AND PROPERTIES OF BIOCHAR-AMENDED SOIL

feedstock. The FTIR analysis of the BC samples showed peaks at 3 431, 1 440 and 1 608 cm−1 . The BC peak at 3 431 cm−1 also suggested the presence of O– H stretching and strong hydrogen bonding in BC. The peaks at 1 440 and 1 608 cm−1 were evidence for O– H or C–O stretching vibration of phenol and aromatic ring C=C in BC, respectively. The obtained XRD spectra also indicated differences between the raw feedstock CW and its BC (Fig. 1). It was found that the raw feedstock samples had the main peaks at 5.80 and 3.89 ˚ A, indicating the presence of organic compounds and representing crystal structure of cellulose in CW wood (Wang et al., 2009; Ertas and Alma, 2010; Kim et al., 2012). In the raw feedstock samples, there was evidence for the presence of whewellite (CaC2 O4 ·H2 O), as indicated by the appearance of a peak at 5.80 ˚ A. However, these peaks were diminished in the biochar samples, indicating the decomposition of cellulose and calcium oxalate. In an experiment to study the effect of pyrolysis temperature on changes in characteristics and chemical composition of biochar produced from CW, Al-Wabel et al. (2013) found that the peaks of cellulose and calcium oxalate in CW and the biochar produced at 200 ◦ C were diminished when increasing the pyrolysis temperature to 400–800 ◦ C. The results also indicates that whewellite structures of the raw feedstock samples disappeared in the biochar samples and were replaced by calcite, as indicated by the appearance of a peak at 3.02, 2.48, 2.09, 1.91 and 1.87 ˚ A. The SEM image revealed that the BC produced was subjected to clear structural alteration as compared with its raw feedstock CW (data not shown), indicating that cellulose loss and porosity increased in the biochar samples.

49

Changes in soil pH and DOC Fig. 2 shows the influence of applying CW and BC on soil pH and DOC during the incubation experiment. The soil pH was in alkaline range (7.20–8.32) for all treatments during incubation. The application of CW at 10, 30 and 50 g kg−1 caused significant decreases in soil pH. This decline in the values of soil pH can be explained by the acidic organic decomposable products present in the raw feedstock CW. Contrary to the effects induced by the raw feedstock, BC application raised the values of soil pH. The effect of BC in increasing the soil pH may have been due to the alkalinity of BC and the release of base cations into the soil (Nguyen and Lehmann, 2009). Numerous studies have found that BC is an efficient material for increasing soil pH and inducing liming effect (Nguyen and Lehmann, 2009; Ameloot et al., 2013). The soil treated with the raw feedstock CW showed the highest concentrations of DOC (Fig. 2). The concentrations of DOC were also affected by BC, especially at higher application rates. Applying BC at 50 g kg−1 increased the concentration of DOC compared to the control, but these increases were significant only during the first few days of incubation. Generally, the results showed that this readily available pool of OC in the soil treated with CW and BC decreased with increasing incubation time, mainly attributing to mineralization by soil microorganisms. It is very important to note that, though the significant increases in DOC in the soil treated with BC occurred during the first few days of incubation, the increases were not significant when the incubation time progressed, despite of the very high content of total carbon (C) (761.8 g kg−1 )

Fig. 1 Fourier transform infrared (FTIR) spectroscopy (a) and X-ray diffraction (XRD) (b) analyses for biochar (BC) and the raw feedstock Conocarpus waste (CW).

50

M. EL-MAHROUKY et al.

Fig. 2 Effects of biochar (BC) and the raw feedstock Conocarpus waste (CW) at different application rates on pH (a) and dissolved organic carbon (DOC) (b) of a sandy calcareous soil during incubation. Vertical bars represent standard deviation of the mean (n = 3). CK is the control without BC and CW. BC10, BC30 and BC50 are the rates of BC at 10, 30 and 50 g kg−1 , respectively. CW10, CW30 and CW50 are the rates of CW at 10, 30 and 50 g kg−1 , respectively.

respectively. However, the rate increased and ranged from 3.1 to 243.1 µg C g−1 soil d−1 , 4.3 to 371.0 µg C g−1 soil d−1 and 5.1 to 499.0 µg C g−1 soil d−1 in the soil treated with CW of 10, 30 and 50 g kg−1 , respectively. In the treated and non-treated soil, the CO2 C emission rates were observed to be highest at the beginning of incubation, but tended to decline when the incubation time progressed. These decreases were sharp during the first few days of incubation. The decreases of the CO2 -C emission rates with time could be attributed to a high portion of the readily labile carbon fraction being more available to soil microorganisms at the beginning. However, this bioavailable pool of carbon gradually decreased with time, resulting in lower microbial activity (Usman et al., 2004a, b, 2005). Applying CW resulted in higher significant increases in CO2 -C emission rate compared to the control

in the biochar samples. Therefore, our results suggested that carbon in the BC samples was fixed and recalcitrant more than in the raw feedstock samples. Overall, our findings are consistent with Jones et al. (2011), who found that biochar addition to soil resulted in the short-term significant increases in soil DOC. Changes in soil CO2 emission The results in Table II showed that the addition of BC and the raw feedstock CW led to significant increases in CO2 -C emission rate compared to the unamended control. The values of CO2 -C emission rate in the control soil ranged from 1.8 to 66.0 µg C g−1 soil d−1 during incubation. For the soil treated with BC of 10, 30 and 50 g kg−1 , the rate ranged from 1.8 to 85.7 µg C g−1 soil d−1 , from 1.8 to 145.7 µg C g−1 soil d−1 and from 1.9 to 184.1 µg C g−1 soil d−1 , TABLE II

Effects of biochar (BC) and the raw feedstock Conocarpus waste (CW) at different application rates on CO2 -C emission rate of a sandy calcareous soil during incubation Treatmenta)

LSD0.05 b)

CO2 -C emission rate 1h

Day 3

Day 7

Day 15 g−1

CK CW10 CW30 CW50 BC10 BC30 BC50 LSD0.05

66.0 243.1 371.0 499.0 85.7 145.7 184.1 43.7

13.0 55.1 97.8 142.0 14.6 17.3 19.1 2.8

6.3 23.5 27.0 33.5 6.0 6.8 6.6 3.9

µg C 1.8 6.2 8.1 12.1 1.9 2.0 2.0 1.4

soil

Day 30

Day 45

Day 60

1.9 4.7 5.8 10.0 2.0 1.8 1.9 1.0

1.8 3.6 5.2 7.6 1.8 1.8 2.0 0.9

1.8 3.1 4.3 5.1 1.9 1.9 2.0 0.8

d−1 27.7 27.7 13.9 2.6 1.9 2.5 13.9

a) CK is the control without BC and CW; BC10, BC30 and BC50 are the rates of BC at 10, 30 and 50 g kg−1 , respectively; CW10, CW30 and CW50 are the rates of CW at 10, 30 and 50 g kg−1 , respectively. b) Least significance difference at P = 0.05.

CO2 EMISSION AND PROPERTIES OF BIOCHAR-AMENDED SOIL

and BC-treated soil throughout incubation. Moreover, cumulative CO2 -C emission showed higher increases of 200%, 336% and 524% with CW added at 10, 30 and 50 g kg−1 , respectively, compared to non-treated soil (Fig. 3). The addition of CW determined higher soil CO2 -C emission rate and cumulative CO2 -C emission due to the fact that the added organic substances induced higher easily decomposable organic compounds, resulting in an important loss of OC possibly due to the intense microbial degradation of its labile C fractions. Woody materials such as CW have high contents of lignin, cellulose and pectin with high values of C:N ratio, which might provide high CO2 emission. The high concentration of DOC induced by the raw feedstock CW could also be responsible for a high microbial activity, resulting in a higher cumulative CO2 -C evolution. In the current study, a significant correlation between DOC and CO2 -C emission rate (R2 = 0.90– 0.98) indicated that this fraction of labile C was the most important source of energy for microorganisms. Several researchers attributed the high microbial activity in organic waste-amended soil to the high level of water-soluble C of the amended soil (Pascual et al., 1999; Usman et al., 2004a). Moreover, it is also possible that organic acids released during organic material decomposition may result in soil calcium carbonate dissolution, contributing to CO2 efflux from calcareous

51

soils (Ramnarine et al., 2012). In contrast to applying CW, applying BC resulted in significantly different CO2 -C emission rate from that of the control only during the first three days of incubation and with higher application rates of 30 and 50 g kg−1 . Except for these first few days of incubation, there were no significant differences in the CO2 -C emission rates between the control soil and the three rates of BC applied to the soil. Moreover, applying BC at 10, 30 and 50 g kg−1 showed higher increases of 6%, 12% and 20% in cumulative CO2 -emission, respectively, compared to the control (Fig. 3). These increases were significant only for the highest application rate of 50 g kg−1 . The small and non-significant increases in cumulative CO2 -emission induced by adding biochar indicated that the contribution of BC to CO2 emission was lower compared to that of the raw feedstock. Similarly, other studies (Kuzyakov et al., 2009; Smith et al., 2010; Wang et al., 2011; Case et al., 2012) have observed small or no significant impacts of biochar addition on CO2 emission. In an incubation experiment conducted by Spokas and Reicosky (2009) to investigate the impacts of 16 different biochars from different feedstock materials on net CO2 emission in agriculture soil, five of the biochars increased, three reduced and eight had no significant impact on the CO2 respiration. In the current study, the significant increase in CO2 -C

Fig. 3 Effects of biochar (BC) and the raw feedstock Conocarpus waste (CW) at different application rates on cumulative CO2 -C emission from a sandy calcareous soil during incubation. Vertical bars represent standard deviation of the mean (n = 3). Same letter(s) above the bars indicate no significant differences between treatments at P = 0.05. CK is the control without BC and CW. BC10, BC30 and BC50 are the rates of BC at 10, 30 and 50 g kg−1 , respectively. CW10, CW30 and CW50 are the rates of CW at 10, 30 and 50 g kg−1 , respectively.

52

M. EL-MAHROUKY et al.

emission from soil at the beginning of the incubation with biochar addition may be attributed to the presence of a labile C fraction in biochar. The addition of biochar to a soil can infer a significant amount of labile C, which can be easily utilized by soil microorganisms as an energy source, resulting in increased microorganism activity and subsequently in increased respiration of the soil (Smith et al., 2010; Jones et al., 2011; Bruun et al., 2012). Deenik et al. (2010) and Zimmerman (2010) found a positive relationship between the labile fraction of organic matter in biochar and the CO2 emission. The importance of this C fraction depends on biochar raw feedstock and pyrolysis conditions as well as on the soil properties (Calvelo Pereira et al., 2011; Bruun et al., 2012). A release of CO2 -C can thus be expected in the short period from the decomposition of this fraction (Smith et al., 2010). In this study, there was evidence for the presence of labile C as the CO2 -C emission rates from soil amended with biochar were significantly higher than those from the control soil during the first three days of incubation. In addition to C lability in biochar, there are several other reasons reported to be responsible for increasing CO2 -C emission from biochar-amended soil, including internal microporosity, the presence of carbonate in the ash, the presence of nutrients in the ash and the abiotic oxidation of biochar (Cheng et al., 2006; Jones et al., 2011).

Kinetics of cumulative CO2 -C emission A number of studies suggested that kinetic models can be used to describe cumulative CO2 -C emission from soils (Murwira et al., 1990; Riffaldi et al., 1996). In this study, linearised forms of five models were used for qualitative comparison of cumulative CO2 -C emission data from the soil treated and nontreated with CW and BC. The amount of cumulative CO2 -C emission from the sandy calcareous soil was poorly described by the first (R2 = 0.369–0.498) and second order (R2 = 0.221–0.249) models, but better by the power function model (Table III). The Elovich model best described the cumulative CO2 -C emission from the soil treated with CW. The kinetic study indicated that the best model and its parameters for soil cumulative C mineralization strongly depended on the type of amendments. Additionally, the quantities of soil amendments added affected significantly the obtained parameters from each model. The Co values in the power function and Elovich models were much higher in the presence of CW than in the presence of BC or in the control. This could be explained by higher cumulative CO2 -C emission and thus higher microbial activity in the raw feedstock CW than BC. Generally, increasing the application rate of CW or BC led to increases in the Co values in the po-

TABLE III Parametersa) of kinetic models for cumulative CO2 -C emission from a sandy calcareous soil as affected by biochar (BC) and the raw feedstock Conocarpus waste (CW) at different application rates Treatmentb)

First-order model

Second-order model

k

Co

R2

CK CW10 CW30 CW50 BC10 BC30 BC50

−0.042±0.0062c) −0.037±0.0014 −0.035±0.0004 −0.036±0.0003 −0.040±0.0013 −0.035±0.0014 −0.033±0.0018

20.8±4.9ad) 84.4±3.2d 135.0±2.2e 187.9±2.0f 24.1±1.8a 32.2±1.2b 36.8±2.2c

0.482 0.399 0.369 0.370 0.485 0.487 0.498

Treatment

Elovich model k

Co

R2

k

Co

R2

0.050±0.0011 0.016±0.0005 0.011±0.0004 0.008±0.0002 0.048±0.0009 0.045±0.0034 0.043±0.0036

207±27a 975±33c 1 704±29d 2 373±24e 234±30a 310±22b 353±30b

0.818 0.927 0.957 0.953 0.829 0.862 0.863

0.55±0.09 0.53±0.02 0.52±0.01 0.53±0.00 0.41±0.04 0.37±0.03 0.35±0.03

18.4±5.0a 69.8±3.5d 108.8±2.7e 151.4±1.4f 30.0±5.0b 38.8±2.8c 43.4±3.4c

0.989 0.968 0.953 0.953 0.989 0.986 0.984

CK CW10 CW30 CW50 BC10 BC30 BC50 a) k

k

Co

R2

−0.0028±0.0027 −0.0007±0.0001 −0.0005±0.0006 −0.0004±0.0000 −0.0021±0.0001 −0.0013±0.0002 −0.0010±0.0001

8.0±3.9a 30.3±3.8d 47.6±1.9e 62.5±0.1f 10.2±0.1a 16.4±0.1b 20.4±1.7c

0.224 0.216 0.211 0.211 0.229 0.242 0.249

Power function model

is the mineralization or rate constant; Co is the potentially mineralizable carbon or the initial substrate concentration. is the control without BC and CW; BC10, BC30 and BC50 are the rates of BC at 10, 30 and 50 g kg−1 , respectively; CW10, CW30 and CW50 are the rates of CW at 10, 30 and 50 g kg−1 , respectively. c) Mean±standard deviation. d) Means followed by the same letter in a column are not significantly different at P = 0.05. b) CK

CO2 EMISSION AND PROPERTIES OF BIOCHAR-AMENDED SOIL

wer function and Elovich models. It was observed that the lowest values of the rate constant (k) in the power function model were found for BC, especially at higher application rates. In the simple Elovich model, it was observed that the values of Co increased and those of k decreased with increasing application rates of CW and BC, suggesting an increase in C mineralization rate. It could generally be concluded that the parameters derived from the power function model in all cases and those from the Elovich model only in the case of CW may appear to be useful indicators of C mineralization. Changes in soil MBC The soil amended with BC and the raw feedstock CW showed a significant increase in MBC throughout incubation (Table IV). The values of soil MBC for the control, the treatments with BC and the treatments with CW were 4.4–18.3, 7.0–43.4 and 25.1–144.0 µg C g−1 soil, respectively. The values of MBC in the soil amended with CW were higher than those of the control and soil amended with BC. During the incubation experiment, the highest increases in MBC were found for the soil with the highest application rate of CW. Increased soil MBC with organic amendments concurs with the results of several other researchers (Goyal et al., 1993; Pascual et al., 1997; Usman et al., 2004a), who suggested that the increase in microbial biomass due to the application of organic additives is mainly

53

due to the microbial biomass present in these organic additives and the addition of substrate C, which can be responsible for the increased number of microorganisms in the organically-amended soil and the subsequently stimulated microbial activity. The results of the current study also showed that applying BC at high rates significantly increased MBC, but these increases were lower than those by applying CW. Generally, the values of MBC of the soil treated with CW showed the highest increase on day 7 for the application rate of 10 g kg−1 and on day 15 for the application rates of 30 and 50 g kg−1 , but they tended to decline with incubation time. However, in the soil treated with BC, the highest values of MBC were recorded on day 0 (after 1 h of incubation) and then these values tended to decline with incubation time. A possible explanation for decreasing MBC with incubation time may be that the availability of decomposable compounds was greater at the beginning and the organic matter tended to be stabilized with increasing incubation time. Several studies found that the soil MBC increased following organic additions and declined when the incubation time progressed (Chander et al., 1995; Usman et al., 2004a, 2013). The values of soil MBC were significantly higher throughout incubation in the CW treatments compared to the control. The values of MBC in the soil treated with CW at 10, 30 and 50 g kg−1 increased by 2.6- to 15.3-fold 2.2- to 17.3-fold and 3.6- to 21.8-fold, respectively, as

TABLE IV Effects of biochar (BC) and the raw feedstock Conocarpus waste (CW) at different application rates on microbial biomass C (MBC) and metabolic quotient (qCO2 ) of a sandy calcareous soil during incubation Item

Treatmenta)

MBC (µg C g−1 soil)

CK CW10 CW30 CW50 BC10 BC30 BC50 LSD0.05 CK CW10 CW30 CW50 BC10 BC30 BC50 LSD0.05

qCO2 (mg CO2 -C h−1 g−1 MBC)

a) CK

LSD0.05 b)

Incubation time 1h

Day 3

Day 7

Day 15

Day 30

Day 45

Day 60

18.3 48.5 40.7 66.6 25.4 43.4 40.6 12.7 143.2 209.6 383.7 312.5 142.3 146.7 188.9 69.0

5.6 57.1 72.9 127.2 11.5 39.7 34.3 18.2 99.5 40.4 57.4 46.9 53.4 18.2 23.4 21.3

4.4 67.2 75.8 107.8 7.8 21.6 29.9 19.1 65.2 14.7 15.0 13.3 32.8 13.4 9.5 24.5

6.6 65.3 94.3 144.0 12.8 19.8 30.5 20.0 13.7 4.0 3.6 3.5 4.1 4.4 3.0 7.1

8.4 47.8 79.6 137.0 7.0 16.9 29.5 18.3 10.4 4.1 3.4 3.0 11.8 4.9 2.8 5.1

8.4 35.1 74.7 107.5 13.0 13.7 23.6 13.7 10.7 4.4 3.0 3.1 56.0 5.6 3.7 5.8

6.9 25.1 45.0 82.3 10.7 15.2 22.1 14.5 11.5 5.2 4.2 2.6 8.2 5.4 3.8 4.0

5.9 9.4 24.5 20.0 5.0 9.1 9.8 44.0 29.5 36.5 10.8 17.7 36.5 6.0

is the control without BC and CW; BC10, BC30 and BC50 are the rates of BC at 10, 30 and 50 g kg−1 , respectively; CW10, CW30 and CW50 are the rates of CW at 10, 30 and 50 g kg−1 , respectively. b) Least significance difference at P = 0.05.

54

compared with those of the control soil. The soil MBC was significantly higher throughout incubation with biochar applied at 50 g kg−1 and only during the first few days of incubation with biochar at 30 kg−1 than that of the control soil, whereas the soil MBC in the case of BC at 10 g kg−1 did not significantly differ from that of the control soil. The values of MBC in the soil treated with biochar at 30 and 50 g kg−1 increased by 1.6- to 7.0-fold and 2.2to 6.8-fold, respectively, as compared with those in the control soil. Previously, it has been demonstrated that soil microbial biomass increases significantly with increasing application rates of charcoal (Steiner et al., 2008; Kolb et al., 2009). More recently, Zhang et al. (2014) also found that biochar application increased soil MBC significantly and the effect size increased with the biochar application rate. However, Dempster et al. (2012) reported contrasting results and suggested that the decreases in MBC led to less C mineralization and thus decreased CO2 -C emission. The variations in biochar impact on microbial activity and biomass may be attributed to variations in biochar raw feedstocks, pyrolysis conditions, biochar composition, biochar application rate and soil type (Kolb et al., 2009; Luo et al., 2011). It was observed that the significant differences in DOC in the treated soil were reflected in the microbial response. The addition of CW resulted in higher DOC, microbial activity and MBC than the addition of BC, which may have resulted from greater DOC in CW as a readily available source of C for microorganisms. The lower values of these parameters with addition of BC as compared to CW suggested that microbial biomass and activity may be limited by C availability in BC. We could attribute the short-term significant increases in both microbial biomass and activity in the soil during the first few days of incubation to the significant increases in substrate availability following BC addition, as indicated by significant increases in DOC. It has been reported that the increased microbial biomass due to biochar application may be explained by the availability of an easily decomposable fraction of biochar (Steiner et al., 2008; Kolb et al., 2009; Novak et al., 2009). The results of the current study showed that the highest application rate of biochar caused a significant increase in soil microbial biomass C throughout incubation, although the increases in DOC and CO2 -C emission of the biochar-amended soil were significant during the first few days and become not significant when the incubation time progressed compared to the control. The increased soil microbial biomass with addition of biochar may be not only due to the increased

M. EL-MAHROUKY et al.

amounts of labile organic matter, especially when the biochar applied is relatively recalcitrant to microbial decay, but also due to the biochar being a preferred habitat for microorganisms, which can encourage a larger population (Liang et al., 2008; Steiner et al., 2008). On the other hand, applying biochar to soils may also cause a beneficial alteration for properties and improve soil fertility of the soil; thus, a possible stimulation of soil microorganisms may lead to increases in the efficiency of microorganisms for utilizing nutrients in their biomass (Steinbeiss et al., 2009). Generally, it could be suggested that the responses of soil microbial biomass and activity to biochar addition may be different from those to the addition of its raw feedstock. Changes in qCO2 In this study, the responses of qCO2 seemed to be strongly affected by the organic treatments (Table IV). The values of qCO2 varied from 10.4 to 143.2 mg CO2 -C h−1 g−1 MBC for the control soil during incubation, but were in the ranges from 4.0 to 209.6 mg CO2 -C h−1 g−1 MBC, 3.0 to 383.7 mg CO2 -C h−1 g−1 MBC and 2.6 to 312.5 mg CO2 -C h−1 g−1 MBC for the soil with CW at 10, 30 and 50 g kg−1 , respectively, and from 4.1 to 142.3 mg CO2 -C h−1 g−1 MBC, 4.4 to 146.7 mg CO2 -C h−1 g−1 MBC, and 2.8 to 188.9 mg CO2 -C h−1 g−1 MBC for the soil with BC at 10, 30 and 50 g kg−1 , respectively. At the beginning of incubation (day 0), the control soil had an initial qCO2 value of 143.2 mg CO2 -C h−1 g−1 MBC. This value increased with the addition of BC and CW, but the increase was only significant for CW. The highest increase in qCO2 was found at day 0 in the CW-amended soil, which may be mainly due to the high proportion of easily biodegradable compounds in the raw feedstock CW. However, as incubation progressed, the values of qCO2 in the soil amended with CW and BC at higher application rates were significantly lower than those of the control soil. In all treatments, the values of qCO2 tended to decrease as incubation time increased, indicating the lower CO2 -C evolution per unit of MBC. Except for the first few days, there were no significant differences in qCO2 between CW and BC. It is very interesting to note that although the soil treated with BC showed a very lower soil CO2 -C production when compared to the soil treated with CW, the qCO2 values of the soil treated with BC did not differ significantly with those of the soil treated with CW. Therefore, our results suggested that applying BC to the soil can lead to a possible increase in microbial C use efficiency and a decrease in C turnover. Similarly, Jin (2010) reported that soils treated with high application rates of BC had

CO2 EMISSION AND PROPERTIES OF BIOCHAR-AMENDED SOIL

a lower soil basal respiration and thus lower values of qCO2 , suggesting a greater metabolic efficiency and decreases in C turnover after BC addition. Liang et al. (2010) also observed lower qCO2 values in a charcoal-rich soil and suggested that microorganisms may be present in a less active fraction within the pore space formed in BC compared to non-amended soil. However, Kolb et al. (2009) observed an increase in soil respiration and qCO2 following the addition of the manure-based BC. This can be explained by the high nutrient contents of the manure-based BC as well as the presence of a significant amount of labile C in BC. In this context, there are several factors that might affect soil respiration activity and C mineralization, including the nature and composition of BC, nutrient contents and availability of BC and the amount of BC added to the soil (Kolb et al., 2009; Steinbeiss et al., 2009; Luo et al., 2011). CONCLUSIONS The soil CO2 -C emission, DOC and MBC were higher in the treatments with CW compared to the control and the treatments with BC. However, the soil CO2 -C emission was slightly increased or unaffected by BC addition, indicating that the contribution of BC to CO2 -C emission was very small compared to that of the raw feedstock CW. Applying BC at higher rates significantly increased MBC and these increases were lower than those caused by applying CW. Further research is required to study long-term impacts of biochar produced from various raw feedstocks and under different pyrolysis conditions on CO2 emission, biochemical properties and fertility of high-pH soils in arid regions. ACKNOWLEDGEMENT This work was supported by the NSTIP Strategic Technologies Program (No. ENV1592-11) in the Kingdom of Saudi Arabia. REFERENCES Ahmad, M., Lee, S. S., Dou, X, Mohan, D., Sung, J. K., Yang, J. E. and Ok, Y. S. 2012. Effects of pyrolysis temperature on soybean stover- and peanut shell-derived biochar properties and TCE adsorption in water. Bioresource Technol. 118: 536–544. Al-Wabel, M. I., Al-Omran, A., El-Naggar, A. H., Nadeem, M. and Usman, A. R. A. 2013. Pyrolysis temperature induced changes in characteristics and chemical composition of biochar produced from conocarpus wastes. Bioresource Technol. 131: 374–379. Ameloot, N., De Neve, S., Jegajeevagan, K., Yildiz, G., Buchan, D., Funkuin, Y. N., Prins, W., Bouckaert, L. and Sleutel,

55

S. 2013. Short-term CO2 and N2 O emissions and microbial properties of biochar amended sandy loam soils. Soil Biol. Biochem. 57: 401–410. Anderson, T.-H. and Domsch, K. H. 1993. The metabolic quotient for CO2 (qCO2 ) as a specific activity parameter to assess the effects of environmental conditions, such as pH, on the microbial biomass of forest soil. Soil Biol. Biochem. 25: 393–395. Black, C. A. 1965. Methods of Soil Analysis. American Society of Agronomy, Inc., Madison. Bruun, E. W., Ambus, P., Egsgaard, H. and Hauggaard-Nielsen, H. 2012. Effects of slow and fast pyrolysis biochar on soil C and N turnover dynamics. Soil Biol. Biochem. 46: 73–79. Calvelo Pereira, R., Kaal, J., Camps Arbestain, M., Pardo Lorenzo, R., Aitkenhead, W., Hedley, M., Mac´ıas, F., Hindmarsh, J. and Mac´ıa-Agull´ o, J. A. 2011. Contribution to characterisation of biochar to estimate the labile fraction of carbon. Org. Geochem. 42: 1331–1342. Case, S. D. C., McNamara, N. P., Reay, D. S. and Whitaker, J. 2012. The effect of biochar addition on N2 O and CO2 emissions from a sandy loam soil—The role of soil aeration. Soil Biol. Biochem. 51: 125–134. Chan, K. Y., Van Zwieten, L., Meszaros, I., Downie, A. and Joseph, S. 2008. Using poultry litter biochars as soil amendments. Soil Res. 46: 437–444. Chander, K., Brookes, P. C. and Harding, S. A. 1995. Microbial biomass dynamics following addition of metal-enriched sewage sludges to a sandy loam. Soil Biol. Biochem. 27: 1409–1421. Cheng, C. H., Lehmann, J., Thies, J. E., Burton, S. D. and Engelhard, M. H. 2006. Oxidation of black carbon by biotic and abiotic processes. Org. Geochem. 37: 1477–1488. Deenik, J. L., McClellan, T., Uehara, G., Antal, M. J. and Campbell, S. 2010. Charcoal volatile matter content influences plant growth and soil nitrogen transformations. Soil Sci. Soc. Am. J. 74: 1259–1270. Dempster, D. N., Gleeson, D. B., Solaiman, Z. M., Jones, D. L. and Murphy, D. V. 2012. Decreased soil microbial biomass and nitrogen mineralisation with Eucalyptus biochar addition to a coarse textured soil. Plant Soil. 354: 311–324. Ertas, M. and Alma, M. H. 2010. Pyrolysis of laurel (Laurus nobilis L.) extraction residues in a fixed-bed reactor: Characterization of bio-oil and bio-char. J. Anal. Appl. Pyrol. 88: 22–29. Gee, G. W. and Bauder, J. W. 1994. Particle-size Analysis. In Klute, A. (ed.) Methods of Soil Analysis. Part 1. Physical and Mineralogical Methods. 3rd Ed. SSSA and ASA, Madison. pp. 377–382. Goyal, S., Mishra, M. M., Dhankar, S. S., Kapoor, K. K. and Batra, R. 1993. Microbial biomass turnover and enzyme activities following the application of farmyard manure to field soils with and without previous long-term applications. Biol. Fert. Soil. 15: 60–64. Hossain, M. K., Strezov, V., Chan, K. Y. and Nelson, P. F. 2010. Agronomic properties of wastewater sludge biochar and bioavailability of metals in production of cherry tomato (Lycopersicon esculentum). Chemosphere. 78: 1167–1171. Ibrahim, H. M., Al-Wabel, M. I., Usman, A. R. A. and AlOmran, A. 2013. Effect of Conocarpus biochar application on the hydraulic properties of a sandy loam soil. Soil Sci. 178: 165–173. Jin, H. 2010. Characterization of microbial life colonizing biochar and biochar-amended soils. Ph.D. Dissertation, Cornell University, Ithaca.

56

Jones, D. L. Murphy, D. V., Khalid, M., Ahmad, W., EdwardsJones, G. and DeLuca, T. H. 2011. Short-term biocharinduced increase in soil CO2 release is both biotically and abiotically mediated. Soil Biol. Biochem. 43: 1723–1731. Kim, K. H., Kim, J.-Y., Cho, T.-S. and Choi, J. W. 2012. Influence of pyrolysis temperature on physicochemical properties of biochar obtained from the fast pyrolysis of pitch pine (Pinus rigida). Bioresource Technol. 118: 158–162. Kolb, S. E., Fermanich, K. J. and Dornbush, M. E. 2009. Effect of charcoal quantity on microbial biomass and activity in temperate soils. Soil Sci. Soc. Am. J. 73: 1173–1181. Kuzyakov, Y., Subbotina, I., Chen, H. Q., Bogomolova, I. and Xu, X. L. 2009. Black carbon decomposition and incorporation into soil microbial biomass estimated by 14 C labeling. Soil Biol. Biochem. 41: 210–219. Lehmann, J., Czimczik, C., Laird, D. and Sohi, S. 2009. Stability of biochar in the soil. In Lehmann, J. and Joseph, S. (eds.) Biochar for Environmental Management. Science and Technology, Earthscan, Sterling. pp. 183–205. Lehmann, J., Gaunt, J. and Rondon, M. 2006. Bio-char sequestration in terrestrial ecosystems: a review. Mitig. Adapt. Strat. Glob. Chang. 11: 403–427. Liang, B. Q., Lehmann, J., Sohi, S. P., Thies, J. E., O’Neill, B., Trujillo, L., Gaunt, J., Solomon, D., Grossman, J., Neves, E. G. and Luiz˜ ao, F. J. 2010. Black carbon affects the cycling of non-black carbon in soil. Org. Geochem. 41: 206–213. Liang, B. Q., Lehmann, J., Solomon, D., Sohi, S., Thies, J. E., Skjemstad, J. O., Luiz˜ ao, F. J., Engelhard, M. H., Neves, E. G. and Wirick, S. 2008. Stability of biomass-derived black carbon in soils. Geochim. Cosmochim. Acta. 72: 6069–6078. Luo, Y., Durenkamp, M., De Nobili, M., Lin, Q. and Brookes, P. C. 2011. Short term soil priming effects and the mineralisation of biochar following its incorporation to soils of different pH. Soil Biol. Biochem. 43: 2304–2314. Murwira, H. K., Kirchmann, H. and Swift, M. J. 1990. The effect of moisture on the decomposition rate of cattle manure. Plant Soil. 122: 197–199. Nelson, D. W. and Sommers, L. E. 1996. Total carbon, organic carbon, and organic matter. In Sparks et al. (eds.) Methods of Soil Analysis. Part 3. Chemical Methods. SSSA and ASA, Madison. pp. 961–1010. Nguyen, B. T. and Lehmann, J. 2009. Black carbon decomposition under varying water regimes. Org. Geochem. 40: 846– 853. Novak, J. M., Lima, I., Xing, B., Gaskin, J. W., Steiner, C., Das, K. C., Ahmedna, M., Rehrah, D., Watts, D. W., Busscher, W. J. and Schomberg, H. 2009. Characterization of designer biochar produced at different temperatures and their effects on a loamy sand. Ann. Environ. Sci. 3: 195–206. Pascual, J. A., Garc´ıa, C. and Hernandez, T. 1999. Lasting microbiological and biochemical effects of the addition of municipal solid waste to an arid soil. Biol. Fert. Soil. 30: 1–6. Pascual, J. A., Garc´ıa, C., Hernandez, T. and Ayuso, M. 1997. Changes in the microbial activity of an arid soil amended with urban organic wastes. Biol. Fert. Soil. 24: 429–434. Ramnarine, R., Wagner-Riddle, C., Dunfield, K. E. and Voroney, R. P. 2012. Contributions of carbonates to soil CO2 emissions. Can. J. Soil Sci. 92: 599–607. Riffaldi, R., Saviozzi, A. and Levi-Minzi, R. 1996. Carbon mineralization kinetics as influenced by soil properties. Biol. Fert.

M. EL-MAHROUKY et al.

Soil. 22: 293–298. Smith, J. L., Collins, H. P. and Bailey, V. L. 2010. The effect of young biochar on soil respiration. Soil Biol. Biochem. 42: 2345–2347. Sparks, D. L. 1989. Kinetics of Soil Chemical Processes. Academic Press, San Diego. Sparks, D. L. 1996. Methods of Soil Analysis. American Society of Agronomy-Soil Science Society of America, Madison. Spokas, K. A. and Reicosky, D. C. 2009. Impacts of sixteen different biochars on soil greenhouse gas production. Ann. Environ. Sci. 3: 179–193. StatSoft Inc. 1995. Statistica for Windows (Computer Program Manual). StatSoft, Inc., Tulsa. Steinbeiss, S., Gleixner, G. and Antonietti, M. 2009. Effect of biochar amendment on soil carbon balance and soil microbial activity. Soil Biol. Biochem. 41: 1301–1310. Steiner, C., Das, K. C., Garcia, M., F¨ orster, B. and Zech, W. 2008. Charcoal and smoke extract stimulate the soil microbial community in a highly weathered xanthic Ferralsol. Pedobiologia. 51: 359–366. Usman, A. R. A., Almaroai, Y. A, Ahmad, A., Vithanage, M. and Ok, Y. S. 2013. Toxicity of synthetic chelators and metal availability in poultry manure amended Cd, Pb and As contaminated agricultural soil. J. Hazard. Mater. 262: 1022– 1030. Usman, A. R. A., Kuzyakov, Y. and Stahr, K. 2004a. Dynamics of organic C mineralization and the mobile fraction of heavy metals in calcareous soil incubated with organic wastes. Water Air Soil Poll. 158: 401–418. Usman, A. R. A., Kuzyakov, Y. and Stahr, K. 2004b. Effect of clay minerals on extractability of heavy metals and sewage sludge mineralization in soil. Chem. Ecol. 20: 123–135. Usman, A. R. A., Kuzyakov, Y. and Stahr, K. 2005. Effect of clay minerals on immobilization of heavy metals and microbial activity in a sewage sludge-contaminated soil. J. Soil Sediment. 5: 245–252. Van Zwieten, L., Kimber, S., Morris, S., Chan, K. Y., Downie, A., Rust, J., Joseph, S. and Cowie, A. 2010. Effects of biochar from slow pyrolysis of papermill waste on agronomic performance and soil fertility. Plant Soil. 327: 235–246. Vance, E. D., Brookes, P. C. and Jenkinson, D. S. 1987. An extraction method for measuring microbial biomass C. Soil Biol. Biochem. 19: 703–707. Wang, Z. Y., Cao, J. Q. and Wang, J. 2009. Pyrolytic characteristics of pine wood in a slowly heating and gas sweeping fixed-bed reactor. J. Anal. Appl. Pyrol. 84: 179–184. Wang, J. Y., Zhang, M., Xiong, Z. Q., Liu, P. L. and Pan, G. X. 2011. Effects of biochar addition on N2 O and CO2 emissions from two paddy soils. Biol. Fert. Soil. 47: 887–896. Yuan, J. H., Xu, R. K., Wang, N. and Li, J. Y. 2011. Amendment of acid soils with crop residues and biochars. Pedosphere. 21: 302–308. Zhang, Q. Z., Dijkstra, F. A., Liu, X. R., Wang, Y. D., Huang, J. and Lu, N. 2014. Effects of biochar on soil microbial biomass after four years of consecutive application in the north China plain. PLoS One. 9: e102062. Zimmerman, A. R. 2010. Abiotic and microbial oxidation of laboratory-produced black carbon (biochar). Environ. Sci. Technol. 44: 1295–1301.