Combined effects of nitrogen deposition and biochar application on emissions of N₂O, CO₂ and NH₃ from agricultural and forest soils

Abstract

Both nitrogen (N) deposition and biochar can affect the emissions of nitrous oxide (N₂O), carbon dioxide (CO₂) and ammonia (NH₃) from different soils. Here, we have established a simulated wet N deposition experiment to investigate the effects of N deposition and biochar addition on N₂O and CO₂ emissions and NH₃ volatilization from agricultural and forest soils. Repacked soil columns were subjected to six N deposition events over a 1-year period. N was applied at rates of 0 (N0), 60 (N60), and 120 (N120) kg Nh a⁻¹ yr⁻¹ without or with biochar (0 and 30 t ha⁻¹ yr⁻¹). For agricultural soil, adding N increased cumulative N₂O emissions by 29.8% and 99.1% (p < 0.05) from the N60 and N120 treatments, respectively as compared to without N treatments, and N120 emitted 53.4% more (p < 0.05) N₂O than the N60 treatment; NH₃ volatilization increased by 33.6% and 91.9% (p < 0.05) from the N60 and N120 treatments, respectively, as compared to without N treatments, and N120 emitted 43.6% more (p < 0.05) NH₃ than N60; cumulative CO₂ emissions were not influenced by N addition. For forest soil, adding N significantly increased cumulative N₂O emissions by 141.2% (p < 0.05) and 323.0% (p < 0.05) from N60 and N120 treatments, respectively, as compared to without N treatments, and N120 emitted 75.4% more (p < 0.05) N₂O than N60; NH₃ volatilization increased by 39.0% (p < 0.05) and 56.1% (p < 0.05) from the N60 and N120 treatments, respectively, as compared to without N treatments, and there was no obvious difference between N120 and N60 treatments; cumulative CO₂ emissions were not influenced by N addition. Biochar amendment significantly (p < 0.05) decreased cumulative N₂O emissions by 20.2% and 25.5% from agricultural and forest soils, respectively, and increased CO₂ emissions slightly by 7.2% and NH₃ volatilization obviously by 21.0% in the agricultural soil, while significantly decreasing CO₂ emissions by 31.5% and NH₃ volatilization by 22.5% in the forest soil. These results suggest that N deposition would strengthen N₂O and NH₃ emissions and have no effect on CO₂ emissions in both soils, and treatments receiving the higher N rate at N120 emitted obviously more N₂O and NH₃ than the lower rate at N60. Under the simulated N deposition circumstances, biochar incorporation suppressed N₂O emissions in both soils, and produced contrasting effects on CO₂ and NH₃ emissions, being enhanced in the agricultural soil while suppressed in the forest soil.

Key words:

• Nitrogen deposition
• biochar
• nitrous oxide
• carbon dioxide
• NH₃ volatilization

INTRODUCTION

Anthropogenic  activities, including fossil fuel combustion and intensive agricultural activities, have more than doubled reactive nitrogen (N) emission levels compared to pre-industrial times, resulting in a high N deposition rate to ecosystems worldwide (Galloway et al. 2004). In China, the production of net reactive N increased from 9.2 to 56 Tg from 1910 to 2010; the change has been more dramatic here than anywhere (Cui et al. 2013). If current trends proceed, anthropogenic reactive N creation will increase to 63 Tg by 2050 in China (Cui et al. 2013), with obvious consequences for stimulated atmospheric N deposition in and near China (Liu et al. 2013). Excessive N has led to a series of environmental problems and destroyed the balance and health of ecosystems (Van der Heijden et al. 2000; Matson et al. 2002; Mo et al. 2008). For example, N additions can interact with elevated carbon dioxide (CO₂) (Schimel, 1995; Bragazza et al. 2006), increase nitrous oxide (N₂O) emissions (Jia et al. 2012b), lead to ammonia (NH₃) volatilization (Battye et al. 2003) and losses of nitrate (NO₃^–) (Fang et al. 2009), change the carbon (C) and N cycles of forest ecosystems (Fang et al. 2008), affect microbial activities (Kolb et al. 2009) and provide N to crops (Liu et al. 2013). Many studies have currently demonstrated the effects of N addition, including atmospheric N deposition and the application of N fertilizers, on N₂O, CO₂ and NH₃ emissions from cropland, forest and grassland ecosystem soils (Dormann and Woodin 2002; Xiong et al. 2002; Galloway et al. 2004; Xiong et al. 2007). N₂O is an important greenhouse gas that has contributed to a 298-fold higher global warming potential than CO₂ for a 100-year duration (IPCC, 2007). Although NH₃ is not considered a greenhouse gas, deposited NH₃ can play a secondary role as a source of N₂O, which contributes to global warming (Mosier et al. 1998). N deposition has reached 70 Tg N yr⁻¹ globally (Galloway et al. 2004). Therefore, changes in atmospheric N deposition may increase uncertainties in the greenhouse gas exchanges of ecosystems.

Biochar has attracted significant attention in recent years due to its potential positive effects on soil physical-chemical properties. Biochar is a stable aromatic substance produced by pyrolyzing biomass residues at 350–600°C in an environment that is completely or partially devoid of oxygen (Sohi et al. 2010). Biochar is typically alkaline, highly porous and has a high cation exchange capacity (CEC) and large specific surface area (Downie et al. 2009). Adding biochar to soils has been demonstrated to depress N₂O emissions and to sometimes affect CO₂ from soils by affecting soil pH, aeration (Yanai et al. 2007; Van Zwieten et al. 2010) and available NH₄⁺ and NO₃^– (Spokas et al. 2010). Furthermore, biochar can capture the NH₃ produced in anthropogenic emissions and make it bioavailable in soils (Taghizadeh-Toosi et al. 2012). The response of N₂O fluxes, CO₂ emissions and NH₃ volatilization from soils due to biochar addition in increasing atmospheric N deposition is not well understood.

Currently, studies on atmospheric N deposition predominantly focus on natural ecosystems, such as forest, grassland and water ecosystems, and are limited on agricultural ecosystems due to the chronic and substantial use of N fertilizers. The amount of N deposition was small and usually neglected in agricultural ecosystems as compared to the applied N fertilizers. However, atmospheric N deposition can also provide N for crops and induce a nutrient burden for agricultural ecosystems (He et al. 2010). And wet N deposition has even reached 94.1 kg ha⁻¹°yr⁻¹ in agricultural regions in China (Lv et al. 2007; Cui et al. 2010), so we should not ignore N deposition in agricultural ecosystems as well as natural ecosystems. Biochar addition can affect N transformation processes and greenhouse gas emissions in soils (Singh et al. 2010; Van Zwieten et al. 2010). Zhang et al. (2010) reported that biochar application reduced N₂O emissions in a paddy field. Clough et al. (2013) reported that biochar addition initially stimulated N₂O emissions in pasture soil in the presence of bovine urine. A range of different responses for CO₂ emissions from soil with biochar amendment has been reported in different soil (Kolb et al. 2009; Spokas and Reicosky 2009). The emissions of N₂O and CO₂ were affected by soil type, N and biochar. Thus, in the circumstances of increasing N deposition and biochar application, comparison studies on two representative soils, such as forest soil from natural ecosystems and agricultural soil from agro-ecosystems receiving a large amount of N fertilizer, are essential to investigate the combined effects of N deposition and biochar on N₂O fluxes, CO₂ emission and NH₃ volatilization from different ecosystems.

Therefore, we simulated six atmospheric N deposition events at three N rates over a 1-year period on two contrasting soils treated with and without biochar. The aims of the present study are: (1) to assess the effects of N deposition on N₂O, CO₂ and NH₃ fluxes from two contrasting soil types under biochar amendment circumstances, and (2) to examine the effects of biochar on N₂O, CO₂ and NH₃ emissions under simulated N deposition circumstances.

MATERIAL AND METHODS

Soil and biochar properties

Agricultural and forest soils classified as Anthrosol and Lixisol (FAO Soil Classification), respectively, were used in this study. The agricultural soil was collected before sowing at a depth of 0–20 cm from a farmer’s arable field (Gaoqiaomen town, Nanjing city, Jiangsu Province, China; 31°59'N, 118°51'E) that has been intensively cultivated with vegetables for more than 30 years. The forest soil was gathered from the top layer (0–20 cm) from Purple Mountain in Nanjing (32°04'N, 118°50'E) at an altitude of 215 m, which has the widest distribution of Pinus massoniana Lamb. trees. Both soils were passed through a 2-mm sieve after air-drying. The repacked soil column height was 34 cm and 38 cm for agricultural and forest soil, respectively to maintain the original soil bulk density (Table 1). Major soil properties, measured using standard methods (Lu 2000), are listed in Table 1; i.e., pH (H₂O) was determined by a PHS-3 C mv/pH detector (Shanghai Kangyi Inc. China); soil texture was determined by the pipette method; soil bulk density was measured by the cutting ring method; total carbon was measured by wet digestion with Sulphuric acid - Potassium (H₂SO₄-K₂Cr₂O₇); total N was analyzed by semi-micro Kjeldahl digestion; cation exchange capacity (CEC) was measured by the Dichromate Ammonium acetate (CH₃COONH₄) method. Soil moisture contents were determined by the oven drying method at 105°C, and then converted to water-filled pore space (WFPS) via the following equation:

(1)

where total soil porosity = [1–(soil bulk density (g cm⁻³)/2.65)], with 2.65 (g cm⁻³) being the assumed particle density of the soil.

Table 1 Characteristics of the two soils and biochar used in the experiments

Item	Agricultural soil	Forest soil	Biochar
Texture	Silt clay	Loam
sand, %	3.5	50.7
silt, %	67.9	36.3
clay, %	28.6	13.0
Total C, g kg^–1	11.6	40.1	467.0
Total N, g kg^–1	1.4	1.2	5.6
pH (1:2.5 H2O)	5.1	5.2	9.4
CEC, cmol kg^–1	31.2	27.1	24.1
Bulk density, g cm^–3	1.0	0.9
Surface area, m^2 g^–1			8.9
Ash content, %			20.8

C: carbon; N: nitrogen; CEC: cation exchange capacity.

Biochar supplied by the Sanli New Energy Company (Henan Province, China) was used in this experiment. The biochar was produced from wheat straw through pyrolyzation at 450°C lasting for 4.5 h in a vertical kiln made of refractory. The important characteristics of this biochar are provided in Table 1 and were determined as described by Zhang et al. (2010).

Experimental columns

The simulated N deposition experiment was performed in specially designed polyvinylchloride (PVC) columns (40 cm height, 15 cm inner diameter) with a leachate collection outlet. The top edge of each column had a groove (4 cm width) filled with water, which allowed a chamber to be attached to the column during gas sampling. Sampling chambers were cylindrical (height 20 cm, inside diameter 18 cm) and made of PVC. This set of devices was used to collect N₂O and CO₂. Another column was designed to collect NH₃. It was a PVC cylinder with the inside diameter and height of 10 and 20 cm, respectively. Two pieces of sponge (2 cm thickness) immersed with 15 mL glycerol phosphoric acid were installed in a separate PVC cylinder. The bottom of the cylinder was inserted 2 cm into the soil. The lower sponge was 5 cm away from the soil surface and used to absorb NH₃ volatilized from the soil in the cylinder. The upper sponge was used to absorb NH₃ from the atmosphere, thus preventing contamination of the lower sponge. The NH₃ absorbed by the lower sponge was used to determine the NH₃ volatilization rate.

Experimental design and treatments

The simulated wet N deposition experiment was performed in a greenhouse in Nanjing (Jiangsu Province, China). The local bulk N deposition rate was recorded at 60 kg N ha⁻¹ yr⁻¹ in 2010 (data unpublished). Since the trend of bulk N deposition is increasing in China (Liu et al. 2013), we choose two levels of N deposition 60 and 120 kg N ha⁻¹ yr⁻¹, in this study. The N deposition in the present study denotes wet deposition. For sequestrating CO₂, 30 t ha⁻¹ yr⁻¹ of biochar was adopted, which is similar to the rate by Jia et al. (2012a). Agricultural and forest soils were subjected to six treatments with three replications: N0 (zero N, zero biochar), N60 (60 kg N ha⁻¹ yr⁻¹, zero biochar), N120 (120 kg N ha⁻¹ yr⁻¹, zero biochar), BCN0 (zero N, 30 t ha⁻¹ yr⁻¹ biochar), BCN60 (60 kg N ha⁻¹ yr⁻¹, 30 t ha⁻¹ yr⁻¹ biochar), BCN120 (120 kg N ha⁻¹ yr⁻¹, 30 t ha⁻¹ yr⁻¹ biochar). N was added into the columns in the form of NH₄NO₃. To simulate wet N deposition, water was sprayed onto the floor of the columns by hand every 2 months from December 2011 to December 2012. The Ammonium nitrate (NH₄NO₃) was mixed with the water to make sure that the WFPS was 0.9 after the commencement of every N deposition event. The N0 treatment only received an equal amount of water along with the other treatments. All treatments were subjected to six N deposition events over a 1-year period and the designed N was equally distributed among the six deposition events. For biochar amended treatments, biochar was thoroughly mixed with 6 kg of soil at the rate of 30 t ha⁻¹ when we repacked the soil columns on December 11, 2011. Then, the N deposition event was simulated on the next day right after the column repacking.

N₂O and CO₂ flux measurements

The fluxes of N₂O and CO₂ were measured every 10 d and were collected more frequently at an interval of 2 d after the commencement of every N deposition event throughout the entire experimental period using a static opaque chamber. Gas samples were collected from 8:00 am to 12:00 pm using a syringe (25 mL volume) at 10, 20 and 30 min after the chambers were placed on the groove filled with water. The concentrations of N₂O and CO₂ were analyzed within 48 h using a gas chromatograph (Agilent 7890A; Agilent, China) equipped with an electron capture detector (ECD) for N₂O analysis and a hydrogen flame ionization detector (FID) for CH₄ analysis, after CO₂ was reduced by hydrogen to methane (CH₄) in a nickel catalytic converter at 375°C. Fluxes were determined using the slope of the N₂O and CO₂ concentrations vs time regression at 10, 20 and 30 min after chamber closure. Sample sets were accepted for flux calculation if their linear correlation coefficient was significant at the α = 0.05 level. If a linear correlation coefficient was not significant at the α = 0.05 level, the average of the two adjacent measurements was used instead. The overall disposal rate is lower than 5% in this study. The cumulative N₂O and CO₂ emissions were the product of the mean flux and the duration of the measurements.

Ammonia sampling and chemical analysis

Ammonia loss from the soils was determined using the ventilation method (Zhao et al. 2010). The phosphoglycerol-soaked sponge was replaced every 2 d after every N deposition event until NH₃ became undetectable. The phosphoglycerol-soaked sponges used to collect the NH₃ samples were immediately extracted with 300 mL potassium chloride (KCl) solution (1 mol L⁻¹) for 1 h. The concentration of ammonium nitrogen (NH₄⁺-N) was measured using the indophenol blue method at 625 nm (Solorzano, 1969) by ultraviolet spectrophotometry (HITACHI, UV-2900, Japan, with 0.005 absorbance of photometric accuracy).

Statistical analysis

All statistical analyses were performed using JMP v. 7.0 (SAS Institute Inc, USA, 2007). A two-way analysis of variance (ANOVA) was applied to analyze the effects of N, biochar and their interactions on cumulative N₂O, CO₂ and NH₃ emissions for each soil type (Table 2). Tukey’s multiple range tests were applied since the differences between the treatment means and the subject means were statistically significant at the α = 0.05 level as shown in Table 3.

Table 2 Two-way analysis of variance (ANOVA) for the effects of nitrogen (N) and biochar (BC) on nitrous oxide (N₂O) and carbon dioxide (CO₂) emissions and ammonia (NH₃) volatilization during the entire sampling period

			N2O-N (kg ha^−1 yr^−1)			CO2-C (kg ha^−1 yr^−1)			NH3-N (kg ha^−1  yr^−1)
	Factors	df	SS	F value	P value	SS	F value	p value	SS	F value	p value
Agricultural soil	N	2	149.91	41.79	< 0.0001	16737	1.40	0.2850	10.76	31.89	< 0.0001
	BC	1	19.84	11.06	0.0060	60743	10.13	0.0079	0.89	5.29	0.0402
	N × BC	2	4.32	1.21	0.3333	122694	10.23	0.0025	0.14	0.41	0.674
	Model	5	174.64	19.48	< 0.0001	193595	6.46	0.0039	12.7	15.05	< 0.0001
	Error	12
Forest soil	N	2	420.08	200.63	< 0.0001	177746	3.25	0.0745	6.11	38.756	< 0.0001
	BC	1	30.74	29.36	0.0002	1671284	61.10	< 0.0001	2.02	25.65	0.0003
	N × BC	2	8.95	4.28	0.0396	88740	1.62	0.2379	0.87	5.49	0.0202
	Model	5	477.73	91.27	< 0.0001	1894164	13.85	< 0.0001	9.55	1.91	< 0.0001
	Error	12

C, carbon; SS: the sum of squares; F value: the ratio of mean squares of two independent samples; p value: the index of differences between the control group and the experimental group. If p < 0.05 and p < 0.001, significant differences exists between the control group and the experimental group.

Table 3 The cumulative nitrous oxide (N₂O), carbon dioxide (CO₂) and ammonia (NH₃) emissions from the agricultural and forest soils during the entire sampling period

Treatment^†		Agricultural soil (kg ha^−1 yr^−1)			Forest soil (kg ha^−^1 yr^−1)
N	BC	N2O-N	CO2-C	NH3-N	N2O-N	CO2-C	NH3-N
Treatment effect (n = 3)
N0		7.54 ± 1.101^cd‡	1060.73 ± 36.25^b	1.83 ± 0.08^d	4.15 ± 0.09^e	1971.72 ± 133.41^a	2.64 ± 0.29^d
N60		9.73 ± 2.622^bc	1204.96 ± 110.63^ab	2.55 ± 0.53^cd	11.13 ± 0.55^c	2297.42 ± 186.96^a	3.89 ± 0.26^ab
N120		15.40 ± 0.632^a	1287.78 ± 42.04^ab	3.75 ± 0.24^ab	17.49 ± 1.66^a	2219.82 ± 123.91^a	4.42 ± 0.26^a
N0	BC	6.15 ± 1.113^d	1349.58 ± 170.19^a	2.38 ± 0.26^cd	3.33 ± 0.70^e	1453.38 ± 152.42^b	2.29 ± 0.16^d
N60	BC	8.05 ± 0.651^cd	1204.88 ± 46.65^ab	3.10 ± 0.44^bc	6.91 ± 0.44^d	1471.06 ± 99.08^b	2.95 ± 0.52^cd
N120	BC	11.87 ± 0.291^b	1254.11 ± 113.37^ab	4.36 ± 0.44^a	14.15 ± 0.25^b	1519.23 ± 10.00^b	3.25 ± 0.49^bc
N effect (n = 6)
N0 mean		6.85 ± 1.11^b	1205.16 ± 103.22^a	2.11 ± 0.17^b	3.74 ± 0.40^c	1712.55 ± 142.92^a	2.46 ± 0.23^b
N60 mean		8.89 ± 1.64^b	1204.92 ± 78.64^a	2.82 ± 0.49^b	9.02 ± 0.50^b	1884.24 ± 143.02^a	3.42 ± 0.39^a
N120 mean		13.64 ± 0.46^a	1270.95 ± 77.71^a	4.05 ± 0.34^a	15.82 ± 0.96^a	1869.53 ± 66.96^a	3.84 ± 0.37^a
BC effect (n = 9)
without biochar		10.89 ± 1.45^a	1184.49 ± 62.97^a	2.71 ± 0.28^b	10.92 ± 0.77^a	2162.99 ± 148.09^a	3.65 ± 0.27^a
with biochar		8.69 ± 0.69^b	1269.52 ± 110.07^a	3.28 ± 0.38^a	8.13 ± 0.46^b	1481.22 ± 87.17^b	2.83 ± 0.39^b

N: ammonium nitrate (NH₄NO₃) solution; BC: biochar; ^†N0, N60 and N120 represent the nitrogen (N) amendment at 0, 60, 120 kg N ha⁻¹ yr⁻¹; ^‡mean ± standard deviations; superscript lowercase letters within the same column between treatments indicate differences according to Tukey’s multiple range test (p < 0.05).

RESULTS

Dynamics of N₂O and CO₂ emissions over the six wet N deposition cycles

Obvious temporal variations were observed for N₂O emissions from all of the treatments (Fig. 1a, b). N₂O fluxes ranging from –5.6 to 1435.6 μg N m⁻² h⁻¹ and from –14.8 to 2244.2 μg N m⁻² h⁻¹ were observed for the agricultural soil and forest soil, respectively, across all treatments. Each N₂O emission peak was detected and then decreased along the drying continuum after all six N deposition events in all treatments (Fig. 1a, b, e). For all treatments, the largest soil N₂O emissions were observed during the fifth N deposition cycle from both the agricultural and forest soils. The lowest N₂O peaks appeared after the second and third N deposition event from agricultural and forest soils, respectively (Fig. 1a, b).

Figure 1 Dynamics of nitrous oxide (N₂O) emission rates from the (a) agricultural and (b) forest soils, carbon dioxide (CO₂) emission rates from the (c) agricultural and (d) forest soils, and (e) water-filled pore space (WFPS; the mean of all the soil samplings) and air temperature during samplings throughout the six nitrogen (N) deposition events (cycles) A: agricultural soil, F: forest soil, N: ammonium nitrate (NH₄NO₃) solution, BC: biochar, BCN: biochar and NH₄NO₃ solution, the bars indicate the standard error of the mean [± standard error (SE)] for the three replicates of each treatment.

During the six N deposition cycles, CO₂ fluxes varied over ranges of 3.7–52.0 mg C m⁻² h⁻¹ and 4.5–66.9 mg C m⁻² h⁻¹ from the agricultural and forest soils, respectively. The largest soil CO₂ emissions were observed during the second N deposition cycle from both the agricultural and forest soils (Fig. 1c–e). Then soil CO₂ emissions become lower towards the end of the deposition cycle. Compared to those from agricultural soil, CO₂ emissions were greater from the forest soil throughout all six cycles (Fig. 1c–e).

The effects of nitrogen deposition on N₂O, CO₂ and NH₃ emissions

In terms of cumulative N₂O emissions over the six deposition cycles of 1 year, N addition significantly increased N₂O emissions for both agricultural and forest soils (p < 0.001, Table 2). The cumulative N₂O emissions ranged from 6.2 to 15.4 kg N₂O-N ha⁻¹ yr⁻¹ for the agricultural soil, and from 3.3 to 17.5 kg N₂O-N ha⁻¹ yr⁻¹ for the forest soil (Table 3).

Adding N increased cumulative N₂O emissions from agricultural soil by 29.8% and 99.1% (p < 0.05) from the N60 and N120 treatments, respectively, as compared to without N treatments, and N120 emitted 53.4% more (p < 0.05) N₂O than the N60 treatment (Table 3). For forest soil, adding N significantly increased cumulative N₂O emissions by 141.2% (p < 0.05) and 323.0% (p < 0.05) from the N60 and N120 treatments, respectively, as compared to without N treatments, and N120 emitted 75.4% more (p < 0.05) N₂O than N60 (Table 3).

In the case of cumulative CO₂ emissions, the effect of N deposition was insignificant for both agricultural and forest soils (p > 0.05, Table 2). Over the six N deposition cycles, the cumulative CO₂ emissions ranged from 1060.7 to 1349.6 kg CO₂-C ha⁻¹ yr⁻¹ for the agricultural soil, and from 1453.4 to 2297.4 kg CO₂-C ha⁻¹ yr⁻¹ for the forest soil; cumulative CO₂ emissions were not influenced by N addition (Table 3).

Over the six N deposition cycles, the amounts of volatilized NH₃ ranged from 1.8 to 4.4 kg N ha⁻¹ yr⁻¹ and from 2.3 to 4.4 kg N ha⁻¹ yr⁻¹ for the agricultural and forest soils, respectively (Table 3). The effect of N deposition was significant for both agricultural and forest soils (p < 0.001, Table 2). As compared to without N treatments, N addition increased NH₃ losses from the agricultural soil in N60 and N120 treatments by 33.6% and 91.9% (p < 0.05), respectively; N120 emitted 43.6% more (p < 0.05) NH₃ than N60. For the forest soil, NH₃ volatilization increased by 39.0% (p < 0.05) and 56.1% (p < 0.05) from the N60 and N120 treatments, respectively, as compared to without N treatments, and there was no obvious difference between N120 and N60 treatments, with a slight increase of N120 by 12.3% (Table 3). Since NH₃ emissions became negligible after several days of N addition, NH₃ fluxes were generally collected within 10 d after each N deposition event and recorded every day after N addition. This is a different time step with those of N₂O and CO₂ collection, thus the temporal variability of NH₃ was not shown and the main focus was the quantity of NH₃ volatilization (Fig. 2, Table 3).

Figure 2 The cumulative ammonia (NH₃) emissions from the (a) agricultural and (b) forest soils with and without biochar addition during the six nitrogen (N) deposition events (cycles) A: agricultural soil, F: forest soil, N: ammonium nitrate (NH₄NO₃) solution, BC: biochar, BCN: biochar and NH₄NO₃ solution, C: every N deposition event, the bars indicate the standard error of the mean [± standard error (SE)] for the three replicates of each treatment for the sum of the six N deposition events.

The effect of biochar addition on N₂O, CO₂ and NH₃ emissions

In terms of cumulative N₂O emissions over the six deposition cycles of 1 year, the effect of biochar addition was significant for both agricultural (p < 0.01) and forest soils (p < 0.001, Table 2). As compared to no biochar treatments, biochar amendment decreased cumulative N₂O emissions from the agricultural soil by 18.4%, 17.3% and 22.9% (p < 0.05) under N amendment at 0, 60 and 120 kg N ha⁻¹ yr⁻¹, respectively, and by 19.8%, 37.9% (p < 0.05) and 19.1% (p < 0.05) from forest soil, respectively (Table 3). Biochar addition significantly decreased N₂O emission under N120 treatment in agricultural soil and under both N60 and N120 treatments in forest soil. As compared to no biochar addition, biochar amendment decreased cumulative N₂O emissions by 20.2% (p < 0.05) for the agricultural soil, and by 25.5% (p < 0.05) from forest soil (Table 3). Significant interaction between N and biochar addition was observed in forest soil (p < 0.05) but not in agricultural soil (Table 2).

Cumulative CO₂ emissions were significantly affected by biochar amendment to both agricultural soil (p < 0.01, Tables 2 and 3) and forest soil (p < 0.001, Tables 2 and 3). Biochar amendment significantly reduced the cumulative CO₂ emissions by 31.5% from the forest soil while slightly increasing them by 7.2% from the agricultural soil (Table 3).

Across the entire observed period, adding biochar to the agricultural soil significantly increased NH₃ volatilization by 21.0% relative to no biochar treatments (Tables 2 and 3). This stimulation due to biochar addition was consistently present for each N deposition event among all treatments for the agricultural soil (Fig. 2a). In contrast, adding biochar significantly decreased NH₃ losses by 22.5% (p < 0.05) relative to no biochar treatments from the forest soil (Tables 2 and 3). However, this inhibition effect was only observed after the second N deposition event among all treatments for the forest soil (Fig. 2b). Significant interactions between N and biochar were observed on cumulative NH₃ emissions in forest soil (p < 0.05, Table 2).

DISCUSSION

N₂O, CO₂ and NH₃ emissions affected by N deposition

The cumulative emissions of N₂O were significantly stimulated after the commencement of every wet N deposition event in our experiment (Fig. 1a, b). N deposition significantly increased N₂O emissions from both soils in all treatments (Table 3). The N deposition event was in favor of N₂O production, because it provided not only anaerobic conditions but also a substrate for denitrification (Dalal et al. 2003; Bolan et al. 2004; Ruser et al. 2006; Chen et al. 2013). The no-N treatments of the two soils also exhibited similar dynamics, with N₂O emission peaks after each wet deposition (only water, no N fertilizer) event (Fig. 1). The fluctuation from wet to dry soil conditions due to precipitation can generally increase C and N mineralization rates for a few days (Franzluebbers et al. 2000) by increasing the mineralization of soil organic matter (SOM) (Denef et al. 2001a, 2001b) and affecting the soil’s physical properties, such as aggregation (Mikha et al. 2005). When dry soil is wetted, the soil matric potential increases. Increasing soil matric potential caused the release of microbial C and N from intracellular solutes and cell lysis (Halverson et al. 2000; Mikha et al. 2005). This labile C and N could be utilized by soil microorganisms and induce the pulse of soil respiration (Bottner, 1985), which supports our findings of CO₂ emission peaks after each deposition event (Fig. 1c, d). Greater CO₂ emission peaks occurred to both soils right after the commencement of the second N deposition event, which may be related to the rapidly increasing temperature with suitable soil moisture conditions (Fig. 1c–e). Rapid increase in temperature can increase the temperature sensitivity of soil respiration due to the change of soil microbial community composition (Eliasson et al. 2005). CO₂ emissions were relatively low when the temperature was even high with low WFPS (Fig. 1). The low soil moisture can cause the reduction of the temperature sensitivity of soil microbes since the diffusion of extracellular enzyme and substrate, and thus contact between microbe and respiration substrates would decrease (Grogan and Jonasson 2005; Jassal et al. 2008).

In terms of cumulative CO₂ emissions, N produced an insignificant influence in both agricultural soil and forest soil (Tables 2 and 3). For the agricultural and forest soils, adding N significantly increased NH₃ volatilization (Tables 2 and 3), which is consistent with many studies (Hayashi et al. 2006; Pacholski et al. 2008). This result was predominantly due to the enhanced NH₄⁺ concentration after the application of N, irrespective of biochar addition.

N₂O, CO₂ and NH₃ emissions affected by biochar addition

Biochar addition significantly decreased N₂O emission in agricultural soil under the N120 treatment and in forest soil under both N60 and N120 treatments (Table 3), which is in accordance with the results of Augustenborg et al. (2011), Singh et al. (2010) and Wang et al. (2011). Three possible explanations may apply. First, biochar contains plentiful redox reactive organic and inorganic compounds and these redox systems can change from one steady state to another form of organic matter in soil, and are known as “electron shuttles” (Joseph et al. 2010). Biochar as an “electron shuttle” facilitates the transfer of electrons to soil-denitrifying microorganisms, which could enhance the conversion of N₂O to N₂ (Cayuela et al. 2013). Second, biochar acid buffer capacity and liming effect are also considered important factors to mitigate N₂O emissions (Cayuela et al. 2013). Third, free NH₄⁺ can be adsorbed by biochar particles due to an enhanced physical retention (Lehmann et al. 2007; Liang et al. 2006). Elevated adsorption is mainly caused by increase in charge density (CEC per unit surface area), or by increase in surface area (Atkinson et al. 2010).

However, we observed a stimulating effect by biochar incorporation on N₂O emissions for all treatments after the commencement of the first N deposition event (Fig. 1a, b). Concurrently, the CO₂ emissions from agricultural soil also increased with biochar addition during the first N deposition event (Fig. 1c), which is consistent with the results of Yoo and Kang (2012). These findings can be partly explained by the following facts: right after the incorporation of biochar into soils, volatile biochar compounds (aliphatic compounds) may act as decomposable organic C sources in soil and thus provide a readily available substrate for denitrifying microorganisms (Blagodatskaya and Kuzyakov 2008; Ameloot et al. 2013); due to the priming effect (soil microbial community has been stimulated due to the addition of various organic amendments), biochar incorporation with a high C/N ratio may stimulate soil microorganisms to decompose SOM (Kuzyakov et al. 2000; Bünemann et al. 2006; Smith et al. 2010; Nelissen et al. 2012).

The forest soil emitted more CO₂ than the agricultural soil (Table 3), possibly due to the soil’s intrinsic SOM and dissolved organic C content, which has been demonstrated by previous studies where cumulative CO₂ fluxes, soil SOM and dissolved organic C were significantly positively correlated (Augustenborg et al. 2011; Wang et al. 2011). In contrast to agricultural soil, adding biochar significantly decreased the average CO₂ emissions over all treatments from the forest soil (Table 2), which may be explained by the relatively higher SOM in forest soil (Table 1). Many studies have also demonstrated that biochar addition can reduce CO₂ emissions in high-SOM soil (Steiner et al. 2007; Kolb et al. 2009) due to the shifts in the soil microbial community toward fungi (Steiner et al. 2007; Glaser and Birk 2012). Compared to bacteria, fungi generally have lower temperature sensitivity and reduced soil respiration (Biasi et al. 2005). Moreover, biochar amendment may reduce CO₂ emissions due to the fact that biochar may increase microbial biomass in soil by the complexation of SOM with biochar particles, and yet simultaneously induce negative priming of native soil C mineralization (Liang et al. 2010). As for the agricultural soil, biochar incorporation increased CO₂ emissions at the rate of 0 kg N ha⁻¹ yr⁻¹ (Table 3), which was in agreement with Augustenborg et al. (2011) that CO₂ emissions increased in the low-SOM soil. Generally due to the priming effect, biochar incorporation may stimulate soil microorganisms to decompose SOM (Bünemann et al. 2006; Smith et al. 2010; Nelissen et al. 2012) and provide readily available substrate for denitrifying microorganisms (Blagodatskaya and Kuzyakov 2008; Ameloot et al. 2013) and mineralize the labile biochar C fraction by biotic or abiotic ways (Kolb et al. 2009; Zimmerman et al. 2011) in agricultural soil thus enhance CO₂ emissions. Therefore, the relative responses of CO₂ emissions to biochar amendment were affected by the soil type with different levels of SOM.

Similarly, the forest soil volatilized more NH₃ than the agricultural soil without biochar additions (Table 3). This outcome may be explained by the lower soil bulk density and reduced negative charge or CEC, which was induced by the fact that the clay content in the forest soil was much lower than in the agricultural soil (Huang and Zhang 1990; Tables 1 and 3). When biochar was applied, a significant reduction in average NH₃ emissions from forest soil was observed at the same N application rate except under no-N treatments for the entire sampling period (Fig. 2, Table 3), which may be explained by the fact that biochar can increase NH₃ and ammonium (NH₄⁺) retention (Liang et al. 2006). It is worth noting that biochar addition stimulated the average NH₃ volatilization over all treatments for the first N deposition event from the forest soil (Fig. 2). This outcome may be partially explained by the alkalinity of the biochar, which increased the soil pH and thus favored NH₃ emissions during the first N deposition event; however, NH₃ sorption on biochar surfaces can react with surface oxygen groups to form amines and amides (Seredych and Bandosz 2007) and potential chemical ring structure (Jansen and van Bekkum 1994) over time after the biochar was incorporated into the soil. Thus, a relatively long-term experiment is required to better understand the effects of biochar on N cycling.

In contrast to the forest soil, NH₃ emissions from the agricultural soil were obviously stimulated by biochar incorporation (Table 3). The contrasting effect of biochar may be related to the different soil conditions (Table 1). The agricultural soil consisted of 67.9% silt, 28.6% clay and 3.5% sand, which was limited in large soil pores (Table 1); thus, porous biochar addition can promote soil aeration, which may be conducive to NH₃ volatilization in the agricultural soil. However, this did not work for the forest soil with enough large soil pores (50.7% sand). The influence of biochar adsorption or aeration on NH₃ depended on the soil pore condition, NH₃ being adsorbed and reduced with large soil pores as seen in the forest soil, and aerated and promoted with less soil pores as seen in the agricultural soil in this study (Tables 1 and 3). More measurements for various soil types are required to elucidate the underlying mechanisms of biochar on NH₃ emissions.

Since an obvious interaction between N and biochar was observed on cumulative CO₂ emissions for agricultural soil and on cumulative N₂O emissions and NH₃ volatilization for forest soil (Table 2), more specific experiments are needed to understand the underlying mechanisms for the combined effects of N deposition and biochar amendment.

CONCLUSIONS

This simulated year-round wet N deposition experiment measured N₂O, CO₂ and NH₃ emissions from agricultural and forest soils with and without biochar amendment. Though no effect was found on CO₂ fluxes, adding N significantly increased cumulative N₂O and NH₃ emissions from agricultural and forest soils, and treatments receiving the higher N rate at N120 emitted obviously more N₂O and NH₃ than at the lower rate of N60, suggesting that the sources of soil N₂O and NH₃-N emissions would be further strengthened with atmospheric N deposition irrespective of biochar amendment. As compared to without biochar, biochar amendment to the soils effectively depressed N₂O emissions from both agricultural and forest soils, whereas it had different effects on CO₂ and NH₃ emissions for the agricultural and forest soils, both CO₂ and NH₃ emissions being enhanced in the agricultural soil while being suppressed in the forest soil. This study demonstrates that for the forest soil, biochar addition to soils may provide a potential tool for climate change mitigation to respond to atmospheric N deposition. Further studies on the C and N cycles of different land use types, including measurements of dissolved organic C contents and microbial activities, will be required to understand the responses of soils amended with biochar to atmospheric N deposition.

ACKNOWLEDGMENTS

We would like to thank Haiyan Zhao for her help with the chemical analysis. We also benefited greatly from the critical comments by editors and reviewers. This work was jointly supported by the National Science Foundation of China (41171238), the Ministry of Science and Technology (2013BAD11B01), the Doctoral Program of Higher Education of China (20110097110001), the Fundamental Research Funds for the Central Universities (KYZ201110) and the Priority Academic Program Development of Jiangsu Higher Education Institutions (PAPD).