Field slurry application in northeastern China for reducing greenhouse gases and ammonia emission

Abstract

Rapid development of the pig industry in China has led to numerous challenges in managing livestock manure and slurry. Field application of slurry has proven to be an economical, effective and environmentally beneficial approach to sustainable resource recycling globally. However, China remains largely dependent on fertilizer inputs from mineral sources, with limited adoption of slurry application practices. A common challenge with slurry field application is its typically higher emission of ammonia and greenhouse gases (GHGs) compared to mineral fertilizers. This work investigated selected treatments with specific ratios of pig slurry and mineral fertilizers aimed at reducing the use of mineral fertilizers and emission following basal fertilizer application and topdressing after maize planting specifically, the ratios of pig slurry included 30%, 50% and 100%. The methods of fertilizer application involved a comparison of acidified versus non-acidified pig slurry for field application, as well as a comparison between sprinkler and drip irrigation. The results showed that replacing 30% of mineral fertilizers with pig slurry (RC30) reduced total GHG emission by 62% and NH₃ emission by 60.4% compared to a full slurry substitution during field application. Meanwhile, the RC30 group recorded the lowest total NH₃ emission, totaling 5.08 kg·ha⁻¹, among all treatments using pig slurry. The acidification of pig slurry significantly reduced NH₃emission, decreasing them by 42.1% compared to the direct application of untreated pig slurry. Drip irrigation proved to be more effective in reducing total GHG emission compared to sprinkler irrigation. Drip irrigation reduced NH₃emission by 38.9%–42.6%, N₂O emission by 12.4%–18.6%, and GHG emission by 21.5%–34.7%. In summary, this study demonstrated that replacing 30% of mineral fertilizers with pre-acidified pig slurry, combined with drip irrigation, reduced GHG and NH₃ emission.

Graphical abstract

Keywords

Ammonia / emission reduction / greenhouse gas emission / slurry application

Highlight

	● Pre-acidified slurry (pH 6) reduced NH3 emission by 42%.
	● Partial slurry resulted in lower NH3 emission compared to full slurry application.
	● Drip irrigation reduced N2O and NH3 more effectively than sprinkler irrigation.
	● N2O emission increased after precipitation in slurry application.

1 Introduction

With the improvement of global living standards, livestock and poultry farming have increased significantly, especially in China. China’s livestock industry is gradually moving towards intensification and large scaling, currently accounting for over 40% of the global pork production^[1]. For example, multilayer farming is emerging in agriculture. The large-scale pig farming inevitably raises serious concerns about the management of manure slurry and wastewater^[2,^3]. Agriculture is characterized by low added value and high sensitivity to costs, requiring economically efficient treatment methods to promote recycling. After simple treatment, using the manure and slurry from pig farms for field application is usually a cyclic and cost-effective method, especially considering China’s economic situation as a developing country^[4]. This method relies on with the principles of the circular economy both in China and globally, contributing to the achievement of carbon neutrality goals. Currently, China’s land cultivation heavily relies on fertilizers to provide nutrients to crops, leading to soil compaction and negative impacts on the land’s original ecological functions. As a result, modern agriculture has moved to using a combination of organic and mineral fertilizers^[5]. Using animal manure as organic fertilizer for field application is, in fact, an economical, straightforward and resource-recycling method. This can enhance soil nutrients, improve soil structure and promote crop growth due to its high nitrogen and phosphorus content^[6-^8]. However, numerous studies indicate that although slurry has a higher water content and lower nutrient concentration compared to manure, its use can still offer benefits in field applications. The urinary nitrogen in slurry is notably higher than in manure^[9], which consequently leads to relatively higher NH₃emission. However, careful coordination with inorganic fertilizers is required to meet the yield requirements of various crops.

The typical process of using manure and slurry for field application inevitably leads to increased NH₃ emission, nitrogen loss and greenhouse gas (GHG) emission^[10–^12]. Therefore, it is crucial to strike a balance between meeting the nutrient requirements of crops and mitigating the environmental impact. Different fertilizer application ratios and methods are crucial for GHG and NH₃ emission during plant growth. Therefore, it is necessary to study how to meet the reduction of GHGs and NH₃ while ensuring the nutritional needs of crops are fulfilled. Due to the limited availability of large machinery, drip and sprinkler irrigation are the primary methods of slurry application in China, as they are relatively low-cost. Studies from abroad primarily focus on reducing pollution in livestock and poultry farms in terms of environmental impact^[13,^14], with some studies indicating that pre-acidified slurry for field application can reduce NH₃ emission^[15,^16]. In contrast, domestic studies have placed greater emphasis on crop fields^[17–^19]. However, there are still insufficient field experiments and on-site experimental data, particularly concerning practices in China.

Therefore, this study investigated the impact of different ratios of pig slurry to mineral fertilizers, slurry application methods (drip and sprinkler irrigation) and the pre-acidified slurry measure on GHG and NH₃ emission from soil during maize production in the field. Given the similarity of farmland conditions to those in some other countries and the availability of suitable land for cultivation, this study selected Heilongjiang Province in northeastern China as a typical representative area. The findings provided technical support for reducing emission following slurry application while promoting the efficient use of slurry resources.

2 Materials and methods

2.1 Experimental site

The field experiment was conducted from May to August 2023 in an experimental field of Sihetun (46°25′N, 126°29′E) in Pingshan Town, Lanxi County, Suihua City, Heilongjiang Province, which experiences a cold temperate monsoon climate. The region has hot and rainy summers, with autumn rainfall accounting for about 16% of the annual precipitation. The soil type is black calcareous soil. The basic physical and chemical properties of the soil (0–20 cm deep) are presented in the Supplementary Materials (Table 1).

2.2 Slurry preparation and application methods

The slurry used in the experiment was sourced from pig manure processed at the Muyuan Breeding Farm in Pingshan Town, Lanxi County, Suihua City, Heilongjiang Province. Different ratios of mineral fertilizers and slurry were applied to the field in various treatment groups. The crop planted in the experimental field was maize. The basic physical and chemical properties of the pig slurry at basal fertilizer application and topdressing are presented in Table 2. Seven treatments were applied in the experiment, details are shown in Table 3.

Pig slurry is initially produced by collecting liquid pig manure from a slated floor, which is then pumped into a storage tank for solid-liquid separation. The solid fraction is used to produce organic fertilizer and the liquid portion is transferred to a covered anaerobic lagoon for fermentation. In northeastern China, the fermentation period lasts about 45 days. After fermentation, the pig slurry is moved to a storage tank until it is ready for field application. In the acidification slurry (AS) group, the pig slurry is pre-acidified before fertilizer application. The slurry is removed from the storage tank and acidified to about pH 6 using concentrated sulfuric acid.

Based on common fertilizer application methods used by local farmers in northern China, two distinct groups were established: sprinkler irrigation group (RC50S; replacing 50% mineral fertilizer) and drip irrigation group (RC50D; replacing 50% mineral fertilizer). The sprinkler system had a strip with five holes installed every 1.5 m, with each hole having a diameter of 25.4 mm. The drip irrigation system used a 16 mm drip tape, with a water emitter placed every 0.2 m. The row spacing for maize was 0.6 m, and the drip irrigation belts were laid between every two rows of maize, with one belt in each alternate space. Both the RC50S and RC50D groups received the same amount of fertilizer, with pig slurry replacing 50% of the mineral fertilizer. This 50% refers to replacing an equivalent nitrogen content from the mineral fertilizer by mass with pig slurry.

The fertilizer application methods and ratios were determined based on previous research, which identified key factors influencing gas emission following pig slurry application. These factors include pH^[15, ²⁰–^22], the ratios of pig slurry to mineral fertilizer^[23–^25] and the fertilizer application methods^[26–^29]. Acidification is widely recognized as an effective method for reducing NH₃ emission. Additionally, different ratios of slurry to mineral fertilizer and application methods significantly impact both GHG and NH₃ emission.

2.3 Sampling procedures

In the experiment, the static chamber method was used to monitor GHG emission and the indophenol blue colorimetric method to measure NH₃ concentrations. The base was positioned at one third of the center line of the area. The base of the static dark box remained in the experimental field throughout the entire period. A rubber hammer was used to drive the base into the soil until it was fully sealed, preventing any air from entering the static chamber. NH₃ was absorbed using a sponge absorption through an NH₃ volatilization absorption device (thick white pipe in Fig. 1(c)). The equipment used for gas sampling includes a static dark chamber with a base (50 cm × 50 cm × 50 cm), an automatic gas extractor, a 100 mL sampling gas bag, a vacuum pump, an anemometer and a hygrometer.

The sampling time for GHGs was from 8 to 10 a.m. In the first week after fertilizer application, samples were collected daily, then every 7 days. The entire sampling period was from May to August. Before sampling, the air bag was evacuated, and the base of the static dark box was filled with water. The automatic gas extractor and air bag were connected in series. The air bag valve was opened, and the automatic gas extractor was switched on extract air. The automatic gas extractor was set to sample at 0, 10, 20 and 30 min. Each gas bag collected 100 mL of gas, with a total of 4 samples collected per group. After sampling, the air bag valve was tightened and gas detection was completed within 5 days.

The NH₃ volatilization collection device (Fig. 1(a,b)) is inserted into soil about 10 cm. A sponge infiltrated with glycerol phosphate solution (50 mL phosphoric acid and 40 mL glycerol in 1 L) was placed in a straight pipe where about 10 cm above the ground, while another sponge was placed in the bent pipe at the other end to isolate the air. Each treatment plot had three devices. The different fertilizer application methods had different fertilizer application times. Drip irrigation took longer than sprinkler irrigation. The only difference between the RC50S and RC50D groups was the fertilizer application method. Due to the different application time of the same amount of pig slurry to the field, the impact on NH₃ volatilization also differed. With a longer application time, the amount of NH₃ volatilization would be higher. NH₃ emission capture began on the day of fertilizer application. NH₃ collection devices were placed at different locations within each treatment plot, with samples taken the following morning at 7 a.m. During sampling, the sponge in the straight pipe was replaced and placed in numbered, sealed plastic bags. New sponges were used as replacements. The sponge at the bent end was replaced every 7 days. After replacing the sponge at the bent end, the next NH₃ capture cycle began, with devices repositioned accordingly. The sponges in the numbered, sealed plastic bags were transported back to the laboratory, immersed in 400 mL of 1 mol/L KCl solution, and shaken for 1 h. The ammonia nitrogen in the leachate was determined using the indophenol blue colorimetric method.

2.4 Calculation of GHG and NH₃ emission

2.4.1 GHG emission fluxes

The gas emission fluxes of GHGs are calculated as:

(1)F=\rho \times \frac{V}{A}\times \frac{\mathrm{d}c}{\mathrm{d}t}\times \frac{273}{273+T}

where, F is the emission flux (mg·m^–2·h^–1), ρ is GHG density under standard conditions, the density is 0.714 kg·m^–2 for CH₄ and 1.96 kg·m^–2 for N₂O, respectively, V/A is the volume-to-area and the actual height of the box, A is the bottom area of the box, dc/dt is the rate of change in gas concentration inside the box per unit time, which can be expressed by the slope K of the gas curve corresponding to the four sampling time points (R² > 0.9), and T is the temperature inside the box during sampling.

Due to the limitations of fertilizer application methods in the field under drip irrigation conditions, fertilizer application cannot cover the entire planting area as comprehensively as with sprinkler irrigation. Therefore, there are two approaches for calculating the emission flux in drip irrigation plots. One is the area method, because the drip irrigation belts are laid at intervals calculated as:

(2)F_A=\frac{F+F_{TF}}{2}

where, A is the area of maize farm in m², F is the emission flux (mg·m^–2·h^–1) and F_TF is the mineral fertilizer (TF) group emission flux (mg·m^–2·h^–1).

The other approach is the single-hole emission calculation method, where the flux is based on the number of drip irrigation holes per unit area that discharge:

(3)F_H=\frac{F\times 10000}{0.5\times 0.5\times 2\times 1.2\times 0.2}

where, F_H is the single-hole emission calculation (mg·m^–2·h^–1), F is the emission flux (mg·m^–2·h^–1), 0.5 is the side length of the static box (m), 10000 is for conversion (m² →ha), 1.2 is the spacing between drip irrigation tapes (m), 0.2 is the distance between the two holes on belt (m).

2.4.2 Cumulative GHG emission

The cumulative emission of GHGs are calculated as:

(4)Q={\mathit{\Sigma }}_i^n\left(\frac{F_{i-1}+F_i}{2}\times 24\times \frac{D_i}{1000}\right)

where, Q is the cumulative emission (kg·ha^–1); F_i is the emission flux of the i-th sampling, and D_i is the number of days between the (i–1)-th and i-th sampling.

2.4.3 The relevant calculations for NH₃ emission

The NH₄⁺-N content is calculated using the following formula:

(5){\text{NH}}_4^{+}-{\mathrm{N}}_{\text{content}}=\frac{m}{A}\times \frac{V_1}{V_2}\times \frac{ts}{D}\times {10}^{-5}

where, m is the content of NH₄⁺-N in the color-developing solution (μg), V₁ is the volume of the leachate (mL), ts is the sampling multiple; V₂ is the volume of the color-developing solution (mL); A is the cross-sectional area of the capture device (m²); and D is the duration of each consecutive capture (day).

2.5 Meteorological conditions

Meteorological conditions, including wind speed, humidity, temperature and precipitation, were monitored throughout the entire experimental period. Wind speed and humidity were measured at the time of air sampling. The anemometer used to measure wind speed was model YW-561M, and the hygrometer used for measuring humidity was model PM6252A. Precipitation was measured using a rain gauge.

2.6 Statistical analysis

A significant analysis of collected data was performed based on the student t-test by using Origin 2022 (OriginLab, Northampton, MA USA) and GraphPad Prism (v9.5.1, GraphPad Software, Boston, MA, USA). The threshold of statistical significance was set at P < 0.05.

3 Results and discussion

3.1 Daily NH₃ emission fluxes and cumulative NH₃ emission

NH₃ sampling was conducted in two phases, following basal fertilizer application and topdressing. After fertilizer application was completed, NH₃ was collected for about 20 days in each phase. NH₃ emission peak occurred in the first week following fertilizer application. As shown in Fig. 2(a), the raw slurry replacement (RS) group had the highest NH₃ emission rate of 4.93 kg·ha^–1·d^–1 immediately after fertilizer application. Following basal fertilizer application (Fig. 2(a)), no significant differences in NH₃ emission rates were observed between the treatment groups (P > 0.05). Following topdressing (Fig. 2(b)), the initial RS group again had the highest NH₃ emission rate, reaching 4.20 kg·ha^–1·d^–1. Despite this peak, no significant differences in NH₃ emission rates were observed between the treatment groups following topdressing (P > 0.05). These data indicate that NH₃ emission in soil were primarily concentrated in the few days immediately following fertilizer application.

As shown in Fig. 2(c), following basal fertilizer application, the RS group had the highest cumulative NH₃ emission, reaching 8.23 kg·ha^–1. The cumulative NH₃ emission following topdressing was added to the basal fertilizer emission, with the RS group again having the highest cumulative NH₃ emission flux, reaching 12.8 kg·ha^–1. The NH₃ volatilization from the RS group was significantly higher than that from the TF (traditional fertilization) group (P < 0.01). This was likely because organic nitrogen in the slurry underwent nitrification more readily than inorganic nitrogen from mineral fertilizers commonly used in traditional fertilization systems. Compared to the RS group, it was clear that acidification treatment significantly reduced NH₃ emission (P < 0.01). The total cumulative emission for the AS and RS groups were 7.4 and 12.8 kg·ha^–1, respectively. The AS group reduced NH₃ emission by 42.1% compared to the RS group. A previous study found that pre-acidification of pig slurry for field application reduced the emission of both GHGs and NH₃, which aligned with our experimental results^[30]. Also, a study acidified five types of slurry, achieving a reduction of over 60% in NH₃ emission for all types except pig slurry, which showed little to no reduction^[31]. However, in this experiment, pig slurry demonstrated a reduction in emission.

Replacing 30% of the mineral fertilizer with pig slurry for field application (RC30) resulted in the lowest cumulative NH₃ emission, at 5.08 kg·ha^–1, a 60.4% reduction compared to full replacement (RS), with no significant impact on maize yield (P > 0.05). When comparing the RC50S and RC50D groups, significant differences in NH₃ emission were observed for both sprinkler irrigation and drip irrigation (P < 0.05). The total cumulative NH₃ emission for the RC50D(H) group was 3.79 kg·ha^–1, while the RC50D(A) group had 3.56 kg·ha^–1, compared to 6.20 kg·ha^–1 for the RC50S group, respectively, where H and A denote the hole-based and area-based calculation methods. Drip irrigation reduced NH₃ emission by 38.9%–42.6% compared to sprinkler irrigation. Previous studies indicated that disc injection reduced average NH₃ emission by 38% compared to trailing shoe application ^[32].

3.2 GHG emission fluxes

The fluxes of methane emission differed from those of nitrous oxide, with CH₄ showing negative absorption (Fig. 3). This was likely due to factors such as soil moisture and pore structure, which caused CH₄ emission to exhibit a negative absorption state.

Following basal fertilizer application (Fig. 3(a)), only the RS and AS groups had slurry applied. Acidification during this period had no significant impact on N₂O emission (P > 0.05). However, when comparing the RS and TF groups following basal fertilizer application, a significant difference in N₂O emission was observed (P < 0.05). This demonstrates that replacing mineral fertilizer with slurry resulted in significantly higher N₂O emission. Following topdressing (Fig. 3(a)), no significant difference was found between the RS and AS groups (P > 0.05), indicating that acidification had no significant influence on N₂O emission. There was no significant difference between the RS, RC30 and RC50S groups (P > 0.05), indicating that there was no significant difference in N₂Oemission when only mineral fertilizers were applied. The RC50S and RC50D groups, however, demonstrated that the fertilizer application practice significantly impacted N₂O emission (P < 0.05). Drip irrigation was found to reduce nitrous oxide emission. Following topdressing, N₂O emission increased sharply. Following topdressing, slurry was used to replace mineral fertilizers in several groups. Since the slurry contains substantial amounts of organic nitrogen, this resulted in an increase in N₂O emission. Additionally, abundant precipitation and significant rainfall before and after topdressing led to a sudden variation in N₂O emission around the topdressing phase.

Following basal fertilizer application (Fig. 3(b)), no significant differences were observed between the RS and AS groups (P > 0.05), indicating that acidification had no significant influence on soil CH₄ emission. No significant variance was found between the RS and TF groups (P > 0.05), indicating that the application of slurry and mineral fertilizers had no notable impact on CH₄ emission. The RC30, RC50S and TF groups used varying amounts of mineral fertilizer at the basal fertilizer application stage (Fig. 3(b)), but no significant effect was detected between the groups (P > 0.05). This indicates that the amount of mineral fertilizer applied had no significant effect on soil CH₄ emission. Following topdressing, a significant difference was found between the RS and AS groups (P < 0.05), indicating that acidification significantly affected soil CH₄ emission at this stage. In contrast to the negligible difference in CH₄ emission following basal fertilizer application, it is likely that the growth stage of maize influenced gas emission. Both sprinkler irrigation and drip irrigation significantly impacted CH₄ emission (P < 0.05). The CH₄ emission measured with drip irrigation were higher than those with sprinkler irrigation, with the latter group having a negative absorption state following topdressing.

3.3 Cumulative GHG emission and GHG emission expressed in CO₂ equivalent

Following basal fertilizer application (Fig. 4(a)), no significant differences in cumulative soil N₂O emission were observed between the AS and RS groups (P > 0.05). However, a significant difference was found between the RS and TF groups (P < 0.05), indicating that slurry substitution for mineral fertilizers significantly increased N₂O emission. This increase may be attributed to the high organic nitrogen concentration in the slurry. Following topdressing (Fig. 4(a)), no significant difference in cumulative soil N₂O emission was found between the AS and RS groups (P > 0.05), indicating that acidification does not significantly affect N₂O emission. Additionally, no significant differences were observed between the RS, RC50S and RC30 groups (P > 0.05), indicating that the usage quantity of mineral fertilizers has no significant influence on the emission of N₂O.

From Fig. 4(a), it can be observed that the RC50S group had the highest cumulative N₂O emission, while the TF group had the lowest, at 5.12 and 0.90 kg·ha^–1 respectively. The pre-acidification treatment reduced total N₂O emission by 2.1%. Malique also studied the impact of acidified slurry on N₂O emission in a mountain grassland system in southern Germany^[33]. Overall, the differences between the acidified and non-acidified slurry treatments were not significant. The 30% and 50% substitution ratios increased total N₂O emission by 9.6% and 20.2%, respectively, compared to full slurry substitution. A study found that biogas slurry partially substituted for mineral fertilizer decreased the AOB-amoA, nirK and nirS copies by 28.4%–4.4%^[34].

After topdressing, N₂O emission increased significantly. Drip irrigation in our study leads to a reduction by 12.4%–18.6% compared to sprinkler irrigation. The amount of pig slurry used during topdressing was noticeably higher than the basal fertilizer. The increase of pig slurry was positively correlated with N₂O emission. There was frequent heavy rainfall before and after the topdressing. Therefore, the sudden increase in N₂O emission was likely related to the rise in soil moisture. Also, at the jointing stage, plant roots released organic compounds, such as root exudates, as they grow. These organic compounds served as a carbon source for soil microbes, affecting their activity and altering the soil microbial community composition. Some root exudates could stimulate specific microbes, promoting nitrification and denitrification processes to increase N₂O emission^[35–^37].

Drip irrigation delivers water and fertilizer directly to the soil around the roots, potentially leading to elevate nitrogen concentrations in localized areas. In nitrogen-rich environments, soil microbes accelerated nitrification and denitrification processes, potentially leading to increased N₂O emission^[38]. Drip irrigation moistens only specific areas, creating alternating wet and dry cycles in the soil around the roots. Under moist conditions, microbial activity increases, promoting nitrogen transformation and producing N₂O^[39]. Under dry conditions, nitrogen is less likely to be further denitrified into N₂O, reducing the amount of N₂O released into the atmosphere. Drip irrigation makes soil water distribution uneven and led to oxygen-deficient zones, promoting denitrification, which is a key pathway for N₂O production. In sprinkler irrigation, the water distribution is relatively even, ensuring sufficient soil oxygen, which helps to inhibit denitrification^[40]. Drip irrigation systems require precise control of water volume. If irrigation frequency and amounts are insufficient, it may lead to low soil moisture, which can increase microbial nitrogen transformation and consequently raise N₂O emission.

Following basal fertilizer application (Fig. 4(b)), no significant difference was observed in the cumulative CH₄ emission from soil between the AS and RS groups (P > 0.05). However, a highly significant difference was found between the RS and TF groups (P < 0.001), indicating that the substitution of mineral fertilizer with slurry leads to a substantial increase in CH₄ emission. This phenomenon is likely attributable to the high organic carbon concentration in slurry. Additionally, significant differences in both the amount of mineral fertilizer applied and soil CH₄ emission were observed between the RC30, RC50S and TF groups (P < 0.05), indicating that inorganic nitrogen from mineral fertilizers also plays a role in influencing CH₄ emission. Following topdressing (Fig. 4(b)), a significant difference was observed between the RS and AS groups (P < 0.001), indicating that acidification has a remarkable influence on soil CH₄ emission. Acidification can markedly reduce CH₄ emission. Compared to the RS group, pre-acidified pig slurry effectively reduced CH₄ emission by 104%. The cumulative emission of the experimental group subjected to pre-acidification treatment (AS group) presented a state of negative absorption. In Fig. 4(b), it is evident that the RS group had the highest cumulative CH₄ emission, while the AS group had the lowest, at 3.85 and 0.15 kg·ha^–1, respectively. The RS group had significant discrepancies from the RC30 and RC50S groups (P < 0.05), indicating that the ratio of slurry substituting for mineral fertilizers has a pronounced influence on soil CH₄ emission. The RC50S group reduced CH₄emission by 103% compared to the full substitution group (RS). When comparing the RC50S and RC50D groups, it was clear that CH₄ emission of sprinkler irrigation for slurry application were lower than drip irrigation (P < 0.05). Sprinkler irrigation reduced CH₄ emission by about 107% compared to drip irrigation.

The equivalent emission of each GHG and the total GHG CO₂ emission equivalent for the Harbin Lanxi experimental farm during maize production from May to August 2023 are presented in the figure below (Fig. 5). In Fig. 5, it is clear that the TF group had the lowest total N₂O emission in CO₂ equivalents, while the highest was the RC50S group, at 267 and 1530 kg·ha^–1 CO₂ eqv, respectively. The RC50D group had the lowest total CO₂ emission, while the RC50S group had the highest, at 4980 and 1650 kg·ha^–1 CO₂ eqv, respectively. The AS group had the lowest total CH₄ emission in CO₂ equivalent, while the RS group had the highest, at 3.82 and 96.4 kg·ha^–1 CO₂ eqv, respectively. The RC50D group had the lowest total GHG emission in CO₂ equivalent, while the RC50S group had the highest, at 3590 and 3170 kg·ha^–1 CO₂ eqv, respectively. The total GHG emission of pure mineral fertilizers group was significantly lower than pig slurry, whether partial or complete substitute of mineral fertilizers in field applications.

However, under two calculation approaches, the emission range of drip irrigation is approximately at the level of that when only mineral fertilizers are used. This demonstrates that drip irrigation can exert a remarkable effect in reducing greenhouse gas emission. Among the groups combining the pig slurry and mineral fertilizers for field application, the RC30 group, which replaced 30% of mineral fertilizers with pig slurry, had the lowest total GHG emission, reducing by 96.4% compared to 50% substitution with pig slurry and 62.0% compared to fully submission group. The total GHG emission for the RC50S group was 3170 kg·ha^–1 CO₂ eqv, whereas the RC50D(A) group was 713 kg·ha^–1 CO₂ eqv and RC50D(H) group was 3590 kg·ha^–1 CO₂ eqv. Drip irrigation leads to a reduction of total GHG emission by more than 9.1%–6.1% compared to sprinkler irrigation.

3.4 Influence of meteorological factors

Since the experiment was conducted during spring and summer, the average daily temperature gradually increased over time, exceeding 20 °C (Fig. 6). Rising temperatures directly influenced GHG emission, particularly NH₃ emission (P < 0.001), with higher temperatures being associated with an increase in NH₃ emission. Wind speed had a significant effect on NH₃ (P < 0.001), CH₄ (P < 0.001), and N₂O (P < 0.001) emission. Relative humidity also significantly influenced NH₃ emission (P < 0.001) and GHG emission (P < 0.001) (Fig. 6).

The impact of temperature and precipitation (Fig. 6) on N₂O emission was highly significant (P < 0.01). N₂O emission significantly increased under both pre- and post-rainfall conditions, as well as during high-temperature weather (P < 0.01). Using SPSS software, a significant correlation between temperature and N₂O emission was found for the CT group (r = −0.524, P < 0.05) and the SCT group (r = 0.535, P < 0.05). This indicates that temperature had a significant negative correlation with N₂O emission when slurry was used as a complete substitute for mineral fertilizer. Precipitation also significantly influenced N₂O emission (P < 0.01). Correlation analysis indicated that N₂O emission from agricultural soils were significantly correlated with climatic factors such as temperature and precipitation^[41,^42]. Petersen et al.^[43]used isotopes to investigate seasonally distinct sources of N₂O in acid organic soil and found that N₂O emission were strongly related to water table dynamics, individual rain events, slope position, and soil nitrogen availability. Their experiments, which included periods II and III characterized by repeated rainfall, showed increased N₂O emission after rainfall events, consistent with the results of our experiment. N₂O emissions were linked to soil water content, with precipitation altering soil oxygen levels and moisture content, while temperature fluctuations affected microbial activity and plant growth within the soil. Both factors significantly influenced N₂O emission during maize production. The temporal dynamics of N₂O direct emission rates were closely tied to increases in soil moisture^[44,^45].

During the snow-covered period, when the climate was dry and soil water content was low, N₂O production primarily resulted from nitrification. After rainfall, when soil water content was high, N₂O production was mainly attributed to denitrification. At medium soil water content, about half of the N₂O emission originated from nitrification, while the other half came from denitrification^[46–^48]. Some studies have shown that the interactions between core bacterial and archaeal members across different groups, as well as within-group interactions of core bacterial members, collectively contribute to N₂O emission and their temperature sensitivity^[49]. By absorbing water through their roots, plants reduce soil moisture, which allows more oxygen to enter and improves soil aeration. In well-aerated conditions, nitrification becomes more active, while denitrification is suppressed, leading to lower N₂O production since denitrification typically generates more N₂O in low-oxygen environments.

3.5 Changes in soil properties

The basic physical and chemical properties of the soil at depths of 0–20 and 20–40 cm, after fertilizer application and maize planting, are presented for different experimental groups in Table 4. For treatments labeled RC50D, the lowercase letters “a” and “b” denote different soil depth layers, where “a” represents 20–40 cm and “b” represents 0–20 cm.

After fertilizer application in the different experimental groups, no significant changes were observed in soil pH (P > 0.05), total nitrogen (P > 0.05) or available humus nitrogen (P > 0.05). Soil NO₃-N and the C/N ratio were identified as the most influential soil properties affecting N₂O emission, as determined through random forest analysis ^[34]. Under varying fertilizer application conditions, soil electrical conductivity increased, with the AS group showing the highest increase at 47.8%. The conductivity of the NF group decreased by 8.2%. The AS group also had the highest sulfate concentration at 219 mg·kg^–1, while the TF group had the lowest at 28.7 mg·kg^–1. Additionally, the TF group had the lowest concentration of available phosphorus and potassium at 6.5 and 2.8 mg·kg^–1, respectively. At soil depths of 0–20 cm and 20–40 cm, the available potassium concentration decreased, with the RC30 and TF groups showing the most significant reductions of 55.2% and 55.5%, respectively (P < 0.05). Sodium concentration also decreased notably, with the NF and TF groups showing reductions of 57.1% and 45.2%, respectively. The total salt concentration of the soil increased by 100%, with the RC30 group showing the highest increase. In contrast, the total salt concentration of the NF and TF groups decreased by 35.7% and 21.4%, respectively.

The AS group had the highest yield per unit area (Table 5) and the NF group the lowest. This indicates that different fertilizer application ratios and methods can reduce the loss of essential nutrients following fertilizer application, minimize post-fertilizer application GHG and NH₃ emission, and maintain crop yield.

3.6 Economic cost analysis of drip irrigation

3.6.1 Core irrigation system

Initial equipment investment accounted for 60%–80% of total cost. The unit price of drip irrigation pipe ranges from 0.3 to 2 yuan·m^–1, depending on the material (e.g., PE pipe and built-in drip tape) and specifications (e.g., flow rate and wall thickness). The application rate is 1200–2250 m·ha^–1, with a cost of 360–4500 yuan·ha^–1. Head hub includes water pumps, filters, fertilizing tanks and pressure gauges. A small system (covering up to 4 ha) costs 5000–10,000 yuan, while large systems (covering roughly 7–20 ha) can cost tens of thousands of yuan, distributed to about 1500–7500 yuan·ha^–1. Auxiliary pipe fittings include main pipes, branch pipes, valves, connectors, etc., cost about 750–3000 yuan·ha^–1.

3.6.2 Land and installation costs

If the land is not regular, additional costs of 750–2250 yuan·ha^–1 will be required for modifications (such as ridge formation and channel repair). From 750 to 2250 yuan·ha^–1 will be needed for labor, with costs potentially higher for complex terrains (such as mountainous areas or greenhouses).

3.6.3 Operational and maintenance costs

Energy consumption includes water pump electricity or fuel costs, depending on irrigation frequency and land size, 750–3000 yuan·ha^–1·yr^–1. Drip irrigation tape (with a lifespan of 1–3 years) needs to be replaced regularly, with an average annual cost of 450–1500 yuan·ha^–1; filter cartridges and other accessories have an annual cost of 300–750 yuan·ha^–1. Daily inspection, adjustments, and cleaning of blockages, with an average annual cost of 450–1500 yuan·ha^–1 for labor costs.

4 Conclusions

In summary, using 30% slurry as a replacement for mineral fertilizers and acidification resulted in the lowest cumulative NH₃ emission, reducing emission by 60.4% compared to full slurry application. Replacing 30% of mineral fertilizers with pig slurry for field application led to the lowest total GHG emission. Climate conditions, particularly precipitation, were crucial factors influencing GHG and NH₃ emission, with precipitation significantly affecting N₂O emission. Wind speed and humidity also played significant roles in influencing GHG and NH₃ emission. The field application of acidified slurry increased soil total nitrogen content, which benefited crop yield and substantially reduced NH₃emission. The optimal fertilizer application strategy was to replace 30% of mineral fertilizers with pre-acidified slurry via drip irrigation. This approach effectively reduced both GHG and NH₃ emissions. Further targeted experiments are needed to compare and refine the choice between drip and sprinkler irrigation for field applications. It is also essential to select the most accessible and economical fertilizer application tools for local farmers.

References

[1] Zhong B, Zhao L G. The growth effect of industrial agglomeration of hog industry in China. Research of Agricultural Modernization, 2020, 41(2): 341−350 (in Chinese)

[2] She L, Jiang S, Jiang C H, Li X, Liu W R, Su W X, Wu G Y. Progress and prospect for environmental management of livestock and poultry breeding in China. Journal of Agro-Environment Science, 2021, 40(11): 2277−2282 (in Chinese)

[3] Yang B Q, Cui Y Estimation of total pollution of livestock and poultry breeding in Beijing. Chinese Journal of Agricultural Resources and Regional Planning, 2021, 42(5): 77−81 (in Chinese)

[4] Wang S G, Tian X. Can the integrated crop-livestock production improve the environmental efficiency of small-scale hog farming? Analysis of hog farmers in Jiangsu Province. Journal of China Agricultural University, 2021, 26(2): 199−210 (in Chinese)

[5] Du W Y, Tang S, Wang H. The status of organic fertilizer industry and organic fertilizer resources in China. Soil and Fertilizer Sciences in China, 2020, (3): 210−219 (in Chinese)

[6] Liu L Y, Xu Y, Zhu S H, Gao Y, Zheng X Q. Meta-analysis on the responses of soil bulk density to supplementation of organic fertilizers in croplands in China. Journal of Agricultural Resources and Environment, 2021, 38(5): 867−873 (in Chinese)

[7] Wen C Y, Xiong Y H, Wang P, Zhang Z Y, Yao X Y, Peng Z Q, Chen J, Wu Y S. Effects of reduced chemical fertilizer with organic fertilizer application on yield and rice quality in high-quality indica rice. Soils, 2023, 55(2): 280−287 (in Chinese)

[8] Xu L L, Wang X J. Effects of organic manure replacing chemical fertilizer with equal nitrogen on soil nutrients and enzyme activities. Soil and Fertilizer Sciences in China, 2023, (1): 23−29 (in Chinese)

[9] Fu M, Jiang X S, He Z P, Liu Y, Chen T B. N excretion coefficient based on the whole process of pig production and manure N excretion in China. Chinese Journal of Eco-Agriculture, 2020, 28(11): 1799−1810 (in Chinese)

[10] Doyeni M O, Stulpinaite U, Baksinskaite A, Suproniene S, Tilvikiene V. . Greenhouse gas emissions in agricultural cultivated soils using animal waste-based digestates for crop fertilization. The Journal of Agricultural Science, 2021, 159(1−2): 23–30

[11] Schils R L M, Postma R, van Rotterdam D, Zwart K B. . Agronomic and environmental consequences of using liquid mineral concentrates on arable farms. Journal of the Science of Food and Agriculture, 2015, 95(15): 3015–3024

[12] Zhang L, Pan S F, Ouyang Z Y, Canadell J G, Chang J F, Conchedda G, Davidson E A, Lu F, Pan N Q, Qin X Y, Shi H, Tubiello F N, Wang X K, Zhang Y Z, Tian H Q. . Global nitrous oxide emissions from livestock manure during 1890–2020: an IPCC tier 2 inventory. Global Change Biology, 2024, 30(5): e17303

[13] Moreno-García B, Guillén M, Quílez D. . Greenhouse gas emissions as affected by fertilization type (pig slurry vs. mineral) and soil management in Mediterranean rice systems. Agronomy, 2020, 10(4): 493

[14] Swoboda P, Hamer M, Stotter M, Döring T F, Trimborn M. . Effects of rock powder additions to cattle slurry on ammonia and greenhouse gas emissions. Atmosphere, 2021, 12(12): 1652

[15] Finzi A, Vazifehkhoran A H, Dinuccio E, Ambrosini R, Provolo G. . Acidification of livestock slurry and digestate to reduce NH₃ emissions: predicting needed H₂SO₄ dosage and pH trends over time based on their chemical-physical composition. Biosystems Engineering, 2024, 240: 1–13

[16] Pedersen J, Feilberg A, Nyord T. . Effect of storage and field acidification on emissions of NH_3, NMVOC, and odour from field applied slurry in winter conditions. Journal of Environmental Management, 2022, 310: 114756

[17] Wang G Y. Effects of Application of Cattle Farm Slurry on Alfalfa Growth and Nitrogen and Phosphorus Balance. Beijing: Chinese Academy of Agricultural Sciences, 2021 (in Chinese)

[18] Li L H Characteristics of Rhizosphere Process and Farmland Phosphorus Balance in Response to Dairy Cattle Slurry Returning of Spring Maize in Northeast of China. Dissertation for the Master’s Degree. Harbin: Northeast Agricultural University, 2023 (in Chinese)

[19] Holatko J, Hammerschmiedt T, Kucerik J, Kintl A, Baltazar T, Malicek O, Latal O, Brtnicky M. . Fertilisation of permanent grasslands with digestate and its effect on soil properties and sustainable biomass production. European Journal of Agronomy, 2023, 149: 126914

[20] Ellersiek N C T, Olfs H W. . An incubation system for the simulation of ammonia emissions from soil surface-applied slurry: effect of pH and acid type. Agronomy, 2024, 14(5): 1078

[21] Langley-Randall J, Jones D L, Cotton J, Williams J R, Chadwick D R. . Slurry acidification is as effective as slurry injection at reducing ammonia emissions without increasing N₂O emissions: a short-term mesocosm study. Geoderma Regional, 2024, 37: e00791

[22] Stevens R J, Laughlin R J, Frost J P. . Effect of acidification with sulphuric acid on the volatilization of ammonia from cow and pig slurries. The Journal of Agricultural Science, 1989, 113(3): 389–395

[23] Du C Y, Lu H Y, Xiong Y Z, Sun X, Sun X M, Pu J X, Zhang N M Effects of combined application of biogas slurry and chemical fertilizer on peach growth and soil physical and chemical properties for two consecutive years. Journal of Agricultural Science and Technology, 2023, 25(8): 165−175 (in Chinese)

[24] He Z J, Ding B X, Pei S Y, Cao H X, Liang J P, Li Z J. . The impact of organic fertilizer replacement on greenhouse gas emissions and its influencing factors. Science of the Total Environment, 2023, 905: 166917

[25] Kong F, Li Q, Yang Z, Chen Y. . Does the application of biogas slurry reduce soil N₂O emissions and increase crop yield?—A systematic review. Journal of Environmental Management, 2023, 342: 118339

[26] Hamad A A A, Wei Q, Xu J Z, Hamoud Y A, He M, Shaghaleh H, Wei’ Q, Li X T, Qi Z M. . Managing fertigation frequency and level to mitigate N₂O and CO₂ emissions and NH₃ volatilization from subsurface drip-fertigated field in a greenhouse. Agronomy, 2022, 12(6): 1414

[27] Han H H, Gao R, Cui Y L, Gu S X. . A semi-empirical semi-process model of ammonia volatilization from paddy fields under different irrigation modes and urea application regimes. Agricultural Water Management, 2022, 272: 107841

[28] Silva A A, Carvalho M, Coutinho J, Vasconcelos E, Fangueiro D. . Dairy slurry application to stubble-covered soil: a study on sustainable alternatives to minimize gaseous emissions. Agriculture, 2022, 12(7): 1021

[29] Zheng J, Qi X Y, Shi C, Yang S H, Wu Y. . Tomato comprehensive quality evaluation and irrigation mode optimization with biogas slurry based on the combined evaluation model. Agronomy, 2022, 12(6): 1391

[30] Emmerling C, Krein A, Junk J. . Meta-analysis of strategies to reduce NH₃ emissions from slurries in European agriculture and consequences for greenhouse gas emissions. Agronomy, 2020, 10(11): 1633

[31] Wagner C, Nyord T, Vestergaard A V, Hafner S D, Pacholski A S. . Acidification effects on in situ ammonia emissions and cereal yields depending on slurry type and application method. Agriculture, 2021, 11(11): 1053

[32] Pedersen J, Nyord T. . Effect of low-dose acidification of slurry digestate on ammonia emissions after field application. Atmospheric Environment: X, 2023, 17: 100205

[33] Malique F, Wangari E, Andrade-Linares D R, Schloter M, Wolf B, Dannenmann M, Schulz S, Butterbach-Bahl K. . Effects of slurry acidification on soil N₂O fluxes and denitrification. Journal of Plant Nutrition and Soil Science, 2021, 184(6): 696–708

[34] Shi Y L, Rahaman M A, Zhang Q W, Zhan X Y, Zheng L. . Effects of partial substitution of chemical fertilizer with biogas slurry on nitrous oxide emissions and the related nitrifier and denitrifier in a saline-alkali soil. Environmental Technology & Innovation, 2022, 28: 102900

[35] Ai C, Liang G Q, Wang X B, Sun J W, He P, Zhou W. . A distinctive root-inhabiting denitrifying community with high N₂O/(N₂O+N₂) product ratio. Soil Biology and Biochemistry, 2017, 109: 118–123

[36] Rummel P S, Well R, Pfeiffer B, Dittert K, Floßmann S, Pausch J. . Nitrate uptake and carbon exudation—do plant roots stimulate or inhibit denitrification. Plant and Soil, 2021, 459(1−2): 217–233

[37] Timilsina A, Oenema O, Luo J F, Wang Y Y, Dong W X, Pandey B, Bizimana F, Zhang Q, Zhang C, Yadav R K P, Li X X, Liu X P, Liu B B, Hu C S. . Plants are a natural source of nitrous oxide even in field conditions as explained by ¹⁵N site preference. Science of the Total Environment, 2022, 805: 150262

[38] Liu J Y, Qiu H S, Zhang W Z, Hu Z Z, Lü M C, Wang Y. Effects of irrigation types and biochar amendment on soil nitrifying and denitrifying functional microorganisms. Journal of Agro-Environment Science, 2023, 42(10): 2254−2263 (in Chinese)

[39] Kuang W N, Gao X P, Tenuta M, Zeng F J. . A global meta-analysis of nitrous oxide emission from drip-irrigated cropping system. Global Change Biology, 2021, 27(14): 3244–3256

[40] Fentabil M M, Nichol C F, Neilsen G H, Hannam K D, Neilsen D, Forge T A, Jones M D. . Effect of micro-irrigation type, N-source and mulching on nitrous oxide emissions in a semi-arid climate: an assessment across two years in a Merlot grape vineyard. Agricultural Water Management, 2016, 171: 49–62

[41] Li J Q, Meng B, Yang X C, Cui N, Zhao T H, Chai H, Zhang T, Sun W. . Suppression of AMF accelerates N₂O emission by altering soil bacterial community and genes abundance under varied precipitation conditions in a semiarid grassland. Frontiers in Microbiology, 2022, 13: 961969

[42] Lu Y Y, Huang Y, Zheng X H. N₂O emission factor for agricultural soils. Chinese Journal of Applied Ecology, 2005, 16(7): 1299−1302 (in Chinese)

[43] Petersen S O, Well R, Taghizadeh-Toosi A, Clough T J. . Seasonally distinct sources of N₂O in acid organic soil drained for agriculture as revealed by N₂O isotopomer analysis. Biogeochemistry, 2020, 147(1): 15–33

[44] Ferrara R M, Carozzi M, Decuq C, Loubet B, Finco A, Marzuoli R, Gerosa G, Di Tommasi P, Magliulo V, Rana G. . Ammonia, nitrous oxide, carbon dioxide, and water vapor fluxes after green manuring of faba bean under Mediterranean climate. Agriculture, Ecosystems & Environment, 2021, 315: 107439

[45] Portela S I, Araujo P I, Restovich S B, Della Chiesa T, Ponsa J M, Peñas Ballesteros A. . Reuse of dairy cattle slurry as an opportunity to recycle nitrogen and reduce its environmental impact. Ecología Austral, 2023, 33(2): 411–426

[46] Bouteldja M, Malek I, Posta K, Kampfl G, Fóti S, Pintér K, Nagy Z, Balogh J. . Temporal variability and drivers of nitrous oxide emissions from central Hungarian croplands: field and lab experiments. Eurasian Soil Science, 2021, 54(8): 1183–1195

[47] Huang G H, Chen G X, Han B, van Cleemput O. Relationships between soil water content and N₂O production. Chinese Journal of Applied Ecology, 1999, 10(1): 53−56 (in Chinese)

[48] Pärn J, Verhoeven J T A, Butterbach-Bahl K, Dise N B, Ullah S, Aasa A, Egorov S, Espenberg M, Järveoja J, Jauhiainen J, Kasak K, Klemedtsson L, Kull A, Laggoun-Défarge F, Lapshina E D, Lohila A, Lõhmus K, Maddison M, Mitsch W J, Müller C, Niinemets Ü, Osborne B, Pae T, Salm J O, Sgouridis F, Sohar K, Soosaar K, Storey K, Teemusk A, Tenywa M M, Tournebize J, Truu J, Veber G, Villa J A, Zaw S S, Mander Ü . Nitrogen-rich organic soils under warm well-drained conditions are global nitrous oxide emission hotspots. Nature Communications, 2018, 9(1): 1135

[49] Xiao X, Delgado-Baquerizo M, Shen H Y, Ma Z Y, Zhou J Z, Sun B, Liang Y T. . Microbial interactions related to N₂O emissions and temperature sensitivity from rice paddy fields. mBio, 2023, 14(1): e03262–22

RIGHTS & PERMISSIONS

The Author(s) 2026. Published by Higher Education Press. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0)