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Article

Effects of Nitrogen Input and Aeration on Greenhouse Gas Emissions and Pollutants in Agricultural Drainage Ditches

by
Qisen Zhang
1,†,
Jingwei Wu
1,†,
Chenyao Guo
1,*,
Jing Wang
1,
Yanchao Zhao
1,
Qiangkun Li
2 and
Yawei Hu
2
1
State Key Laboratory of Water Resources and Hydropower Engineering Science, Wuhan University, Wuhan 430072, China
2
Key Laboratory of Ecological Environment Protection and Restoration of Yellow River Basin of Henan Province, Institute of Yellow River Hydraulic Research, Zhengzhou 450003, China
*
Author to whom correspondence should be addressed.
†
These authors contributed equally to this work.
Agronomy 2024, 14(2), 235; https://doi.org/10.3390/agronomy14020235
Submission received: 11 December 2023 / Revised: 5 January 2024 / Accepted: 12 January 2024 / Published: 23 January 2024
(This article belongs to the Special Issue Nitrogen Cycle in Farming Systems—2nd Edition)
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Abstract

:
Understanding the patterns of greenhouse gas emissions and the changes in pollution load in terrestrial freshwater systems is crucial for accurately assessing the global carbon cycle and overall greenhouse gas emissions. However, current research often focuses on wetlands and rivers, with few studies on agricultural drainage ditches, which are an important part of the agricultural ecosystem. Investigating the greenhouse gas emission patterns and pollution load changes in agricultural drainage ditches can help accurately assess the greenhouse effect of agricultural systems and improve fertilization measures in farmlands. This study explored the effects of nitrogen input and aeration on the pollution load and greenhouse gas emission processes in paddy field drainage ditches. The results showed that aeration significantly reduced the concentration of ammonium nitrogen (NH4+) in the water, decreased the emissions of nitrous oxide (N2O) and methane (CH4), and slightly increased the emission of carbon dioxide (CO2), resulting in an overall reduction of the global warming potential (GWP) by 34.02%. Nitrogen input significantly increased the concentration of ammonium nitrogen in the water, slightly reduced the emissions of N2O and CH4, and increased the CO2 emissions by 46.60%, thereby increasing the GWP by 15.24%. The drainage ditches reduced the pollution load in both the water and sediment, with the overall GWP downstream being 9.34% lower than upstream.

1. Introduction

The excessive use of chemical fertilizers and pesticides has increasingly drawn attention to the problem of agricultural non-point source pollution [1,2,3]. Agricultural drainage ditches are an important part of terrestrial freshwater systems and agricultural ecosystems, serving as the main conduit for non-point source pollution discharge, and their greenhouse gas emissions cannot be ignored [4,5]. Agricultural drainage ditches are akin to linear wetlands, and research into their pollution load and greenhouse gas emission patterns is still insufficient [6,7,8,9]. Current research often focuses on the processes of greenhouse gas emissions and the physicochemical properties of water bodies in wetlands and rivers [10,11]. Due to their dynamic aquatic environment, particularly considering the influence of water flow dynamics and alternating wet-dry conditions, the greenhouse gas emissions and pollutant migration processes in drainage ditches differ from those in wetlands and rivers [12,13,14,15,16,17,18]. Factors such as gas exchange processes caused by flowing water, the impact of vegetation, and the carbon and nitrogen content in sediments all affect the greenhouse gas emissions from drainage ditches [19,20]. Therefore, it is necessary to investigate the processes of pollution load and greenhouse gas emissions in agricultural drainage ditches.
Moreover, studies have shown that a large amount of greenhouse gases are emitted during the pollutant purification process [8,21], and these emissions are closely related to the nitrogen content in the soil [22,23]. Water quality and dissolved oxygen content are also important factors affecting greenhouse gas emissions [24,25,26]. Thus, research on the migration and transformation of nitrogen in drainage ditches and its impact on greenhouse gas emissions is urgently needed.
Aeration is a common water treatment technology that improves water quality by increasing dissolved oxygen content, promoting the decomposition of harmful substances, reducing pollution, and altering the redox conditions of the water, thereby lowering greenhouse gas emissions and enhancing the biological activity of the water body, improving its self-purification capacity [27]. The effectiveness of aeration has been demonstrated in the fields of wastewater treatment, artificial wetlands, and rivers [28]. Although existing research results show that aeration can reduce both greenhouse gas emissions and pollutant loads in drainage ditches [29], there is still a lack of systematic research on the regularities of aeration in reducing greenhouse gas emissions and pollutant loads in drainage ditches [30]. There is also relatively little research on the interaction between greenhouse gas emissions and the nitrogen migration and transformation processes in drainage ditches.
To sum up, the process of greenhouse gas emissions and pollutant migration and transformation in agricultural drainage ditches requires more in-depth research. Aeration technology, as an effective means of water quality purification, still needs further exploration regarding its application effects and impact mechanisms in drainage ditch systems. This project takes paddy field drainage ditches as the research subject and conducts field experiments on nitrogen input and aeration in the drainage ditches to investigate the impact mechanisms on greenhouse gas emissions and pollutant loads in the drainage ditches. By exploring the correlations among these factors, the aim is to provide a theoretical basis and technical support for ecological design and nitrogen fertilizer application management in agricultural drainage ditches.

2. Materials and Methods

2.1. Experimental Materials and Equipment

Cylindrical floating static dark boxes (20 cm in diameter and 20 cm in height) were used to collect gases emitted from the drainage ditches. A 2010plus gas chromatograph (Shimadzu, Japan) was employed to analyze the concentrations of relevant gases in the collected gas samples. Water-related parameters were measured using a HYDROLAB HL7 multiparameter water quality analyzer (Hach, Ames, IA, USA). The nitrogen content in water and sediment was determined using a UV2600 spectrophotometer (Shimadzu, Kyoto, Japan). Other equipment included a BT100-2J peristaltic pump (Longer, Nanjing, China) and a solar-powered aeration pump (Zhishang, Nanshan, China). The nitrogen source used was ammonium chloride (with a nitrogen content of 24%). The maximum aeration capacity of the solar-powered aeration pump is 9 L/min.

2.2. Experimental Methods and Analytical Techniques

The experiment was conducted at the Irrigation Experiment Station of Wuhan University located in the Qujialing Management Area of Jingshan County, Jingmen City, China. The experiments were carried out in the station’s drainage ditch on 16 September 2023, and 18 September 2023. The experimental setup is shown in Figure 1, with the specific experimental arrangement and treatments as follows.
A drainage ditch with a stable flow was selected. The cross-section of the drainage ditch was trapezoidal. During the experiment, the water depth was controlled at 25 ± 5 cm, and the flow rate of the drainage ditch was maintained at about 208 L/min. Two static boxes were placed in the middle section of the ditch. During the aeration experiment, the upstream static box was designated as the AU (Aeration Upstream) treatment, and the downstream static box as the AD (Aeration Downstream) treatment. The static boxes were spaced 10.0 m apart, with the aeration device positioned between them, ensuring that aeration would not affect the upstream static box and that bubbles generated by aeration would not enter the downstream static box. The aeration device consisted of a pump, air stones, perforated PVC pipes, and a switch; the aeration method involved continuous aeration from 22:00 on 15 September 2023 to 03:00 on 17 September 2023; the aeration rate was 9 L/min, which was approximately 1:23 the flow rate of the drainage ditch. For the nitrogen source input experiment, the upstream static box was designated as the NU (Nitrogen source input Upstream) treatment, and the downstream static box as the ND (Nitrogen source input Downstream) treatment. A peristaltic pump was placed approximately 5 m upstream of the static box, continuously pumping NH4Cl solution into the ditch. The peristaltic pump was operated from 15:00 on 17 September 2023 to 03:00 on 19 September 2023, during which a total of 6 kg of NH4Cl was injected.
The enclosed chamber-gas chromatography method was used to measure the emissions of CO2, CH4, and N2O from the drainage ditch. Gas samples were collected from each static box at 6:00, 8:00, 10:00, 12:00, 14:00, 16:00, 18:00, 21:00, and then at 0:00 and 3:00 the following day. For each treatment, gas samples were collected 10 times, with three replicate samples taken at each collection time. Gas was sampled at 0, 15, 30, and 45 min after sealing the static box, with one bag of gas collected at each time point. The formula for calculating the emission fluxes of CO2, CH4, and N2O is as follows:
F = ρ V A P P 0 T 0 T d C t dt
In the formula, F is the measured gas emission flux (mg/m2·h); V is the volume of air in the static box (m3); A is the box coverage area (m2); dCt/dt is the slope of the trend line of the gas concentration in the box with time during the observation time; ρ is the density of the measured gas in the standard state (kg/m3); T0 is the absolute air temperature under standard conditions (K); P0 is the air pressure under standard conditions (kPa); P is the air pressure at the sampling site (kPa); and T is the absolute temperature (K) at the time of sampling.
Additionally, the global warming potential (GWP) is commonly used to assess the impact of various greenhouse gases on a 100-year timescale, in comparison to an equivalent effect of CO2. The GWPs of methane (CH4) and nitrous oxide (N2O) are 25 and 298 times that of CO2, respectively [31]. The GWPs for each treatment were calculated using the following formula:
GWP(CO2/mg·m−2) = F(CO2) + 25 × F(CH4) + 298 × F(N2O)
When collecting gases, water samples directly beneath the static chamber were also collected, and at 8:00, 16:00, and 24:00 on the same day, sediment samples around the static chamber were collected using a soil auger. The total organic carbon content in the sediment was determined by the potassium dichromate oxidation method (HJ615-2011); the microbial biomass carbon content in the sediment was measured using the chloroform fumigation-extraction method [32]. Deionized water was added to the sediment for leaching and filtration, and the filtrate was analyzed for soluble organic carbon content using a carbon and nitrogen analyzer (Jena, Multi N/C 3100, Beijing, China). The COD content in the water was determined using the potassium dichromate method [33], and the concentrations of nitrate nitrogen and ammonium nitrogen in the water and sediment were measured using a spectrophotometer. The temperature, pH, ORP, and DO of the water were directly measured using a HYDROLAB HL7 multiparameter (Hach, Ny, USA) water quality analyzer. The following formula was used to calculate the variance of the data:
S2 = (1/n)[(x1 − m)2 + (x2 − m)2 + ... + (xn − m)2]
where n is the number of samples, xn is the value of each sample, and m is the average of n sample values.
All data were collated using Microsoft Excel 2021 and the average value and standard deviation were calculated. SPSS v.27.0 data analysis software was used for data correlation analysis, drawn with Origin v.2018.

3. Results

3.1. Characteristics of Nitrate Nitrogen/Ammonium Nitrogen Variation in Agricultural Drainage Ditches and Related Indicators

According to Figure 2, the concentration of ammonium nitrogen in the water bodies treated with AU and AD peaked between 12:00 and 14:00. Additionally, the concentration of ammonium nitrogen in the AU treatment was significantly higher than that of the AD treatment (p < 0.05). The trend of nitrate nitrogen levels in both AU and AD treatments first decreased and then increased, reaching the lowest values between 16:00 and 18:00. However, the difference in nitrate nitrogen concentrations between the AU and AD treatments was not significant (p > 0.05).
For the NU and ND treatments, the trend of ammonium nitrogen concentration in the water bodies was essentially the same, continuously rising during the day and stabilizing around 14:00, fluctuating around 12 mg/L. Additionally, between 0:00 and 3:00, there was a sharp decrease in the ammonium nitrogen concentration in the water bodies. The difference in ammonium nitrogen concentrations between the NU and ND treatments was not significant (p > 0.05). The nitrate nitrogen concentration in the water bodies for NU and ND was 0 mg/L from 6:00 to 3:00 the following day.
From Figure 3, the average daily temperature difference for all treatments does not exceed 1 °C. The pH value range for the AU and AD treatments was 7.28–7.77, with an average water pH value of 7.41 for AU and 7.47 for AD, indicating that aeration slightly increased the water pH value. The pH value range for the NU and ND treatments was 7.00–7.34, showing that the input of NH4Cl solution caused a slight decrease in water pH value.
There was a significant difference (p < 0.05) between the conductivity of water bodies treated with AU, AD, and those treated with NU, ND, with the conductivity of the NU and ND treatments being noticeably higher than that of the AU and AD treatments, indicating that nitrogen input can cause a significant increase in water conductivity. The average ORP value for the AU treatment was significantly higher (p < 0.05) than that of the AD treatment, with both the average and peak ORP values of the aeration treatment being markedly higher than the other treatments.
The DO peaks for both the AU and AD treatments occurred between 12:00 and 14:00, but the DO content for the AD treatment increased by 20.13% compared to the AU treatment. The DO trends and levels for NU and ND were essentially the same. The average DO content for the AU and AD treatments was significantly higher (p < 0.05) than that for the NU and ND treatments.

3.2. Diurnal Variation of Physicochemical Properties of Sediments in Agricultural Drainage Ditches

The sediment organic carbon and MBC (microbial carbon content) of the AU treatment was higher than that of AD. The sediment DOC (dissolved organic carbon) content of the AU treatment was consistently higher than that of AD. The trend in organic carbon content for NU differed from that of ND; the organic carbon content in NU continuously increased, while in ND, there was a slight initial decrease followed by an increase, with the sediment organic carbon content of ND being consistently lower than that of NU. The sediment MBC of ND was consistently lower than that of NU. The DOC content for both the NU and ND treatments decreased, with the average sediment DOC content of the NU treatment being significantly higher (p < 0.05) than that of the ND treatment (Figure 4).
The average soil ammonium nitrogen content in the AU treatment was significantly higher (p < 0.05) than that in the AD treatment. Similarly, the soil ammonium nitrogen content in the NU treatment was significantly higher than that in the ND treatment. Compared to the AU treatment, the average sediment ammonium nitrogen content in the NU treatment increased by 29.55%, with the difference being statistically significant (p < 0.05). The soil nitrate nitrogen content in the AU treatment was significantly higher than that in the AD treatment. The soil nitrate nitrogen content in the NU treatment was significantly higher (p < 0.05) than that in the ND treatment. Compared to the AU treatment, the average sediment nitrate nitrogen content in the NU treatment increased by 30.79%, with the difference being statistically significant (p < 0.05) (Figure 5).

3.3. Diurnal Variation of Greenhouse Gas Emission Fluxes from Farmland Drains

According to Figure 6, the N2O emission fluxes for the AU, AD, NU, and ND treatments were all at relatively low levels. The peak values for all treatments occurred between 18:00 and 24:00. The N2O emission flux for the AU treatment was generally greater than that of the AD treatment, with the average difference in N2O emission flux between the two reaching statistical significance (p < 0.05). The N2O emission flux for the NU treatment was slightly higher than that of the ND treatment. The N2O emission flux for the AU treatment was significantly higher than that for the NU treatment, and the average difference in N2O emission flux between these two treatments also reached statistical significance (p < 0.05).
According to Figure 7, the difference in CO2 emission flux between the AU and AD treatments was not significant (p > 0.05). The difference in CO2 emission flux between the NU and ND treatments was not significant (p > 0.05). However, there was a significant difference in the average CO2 emission flux between the AU and NU treatments (p < 0.05), with nitrogen input significantly increasing the CO2 emission flux. The maximum CO2 emission flux for all treatments occurred between 21:00 and 6:00, while the minimum values were recorded during the day between 12:00 and 16:00.
As shown in Figure 8, the difference in CH4 emission flux between the AU and AD treatments reached a highly significant level (p < 0.01). The difference in CH4 emission flux between the NU and ND treatments was not significant (p > 0.05). Additionally, although the average CH4 emission flux for AU and NU was similar, there was a significant difference in the trend of CH4 emission flux between the AU and NU treatments (p < 0.05). The AU and AD treatments showed the same trend with CH4 flux, continuously increasing during the day, peaking between 14:00 and 18:00, and then starting to decline, but a second emission peak occurred at night around 0:00. The NU and ND treatments had a broadly similar trend, peaking between 8:00 and 10:00, then continuously declining, and reaching the lowest values between 14:00 and 16:00, followed by a rise and a second peak at 18:00.
From Table 1, within a single day, comparing the AU and AD treatments, aeration reduced N2O emissions by 15.73%, reduced CH4 emissions by 71.92%, and increased CO2 emissions by 10.42%, resulting in an overall GWP reduction of 34.02%. Comparing the AU and NU treatments, the addition of NH4Cl reduced N2O emissions by 75.75%, reduced CH4 emissions by 8.86%, and increased CO2 emissions by 46.60%, leading to an overall GWP increase of 15.24%. Comparing the NU and ND treatments, the self-purification function of the drainage ditch reduced N2O emissions by 7.94%, reduced CH4 emissions by 16.54%, and reduced CO2 emissions by 4.03%, resulting in an overall GWP reduction of 9.34%. Overall, it can be seen that aeration mainly reduces greenhouse gas emission intensity by reducing CH4 emissions, while the addition of NH4Cl mainly increases greenhouse gas emission intensity by increasing CO2 emissions. The drainage ditch can reduce the emissions of all three greenhouse gases, but it has a stronger effect on reducing CH4 emissions.
It is evident that the changes in the three greenhouse gases are complex and are influenced by a variety of factors. To analyze the main factors causing the changes in the three greenhouse gases, a significance analysis of the impact of soil and water physicochemical properties on the emissions of the three greenhouse gases was conducted. Detailed results are shown in the table. According to Table 2, N2O emissions are highly correlated with water electrical conductivity, dissolved oxygen content in water, and readily available nitrogen content in water. N2O emissions are significantly negatively correlated with water electrical conductivity and ammonium nitrogen content (p < 0.05), and significantly positively correlated with nitrate nitrogen content (p < 0.05). CO2 emissions are highly correlated with water electrical conductivity, dissolved oxygen content in water, and readily available nitrogen content in water, showing a significant positive correlation with water electrical conductivity and ammonium nitrogen content (p < 0.05), and a significant negative correlation with dissolved oxygen content and nitrate nitrogen content (p < 0.05). It is evident that N2O and CO2 emissions are influenced by the same main factors, while CH4 emissions are highly correlated with soil organic carbon content, soil ammonium nitrogen content, and water oxidation-reduction potential (ORP), and are significantly negatively correlated with water ORP (p < 0.05).

4. Discussion

4.1. Effects of Aeration and Nitrogen Source Input on Physical and Chemical Properties of Drainage Water and Sediment

4.1.1. Effects of Aeration and Nitrogen Source Input on the Physical and Chemical Properties of Drainage Ditches

Comparing the ammonium nitrogen (NH4+) and nitrate nitrogen (NO3−) content in the water between the AU and AD treatments, it is observed that aeration reduced the content of NH4+ in the water by 22.64%, while the content of NO3− increased by 10.73%. This is because aeration can increase the dissolved oxygen (DO) content in the water, which is beneficial for the growth and activity of nitrifying bacteria [34,35]. These bacteria can oxidize NH4+ to NO3−, and thus aeration can promote the removal of NH4+. At the same time, aeration promotes the nitrification process, converting NH4+ to NO3− [36]; on the other hand, aeration inhibits the activity of denitrifying bacteria that operate under anaerobic or low-oxygen conditions [37], thereby slowing down the removal rate of NO3−. Therefore, aeration can increase the content of NO3− in the water. When considering the content of NH4+ and NO3− in the water across all four treatments, it is evident that compared to the NU and ND treatments, the NH4+ content is at the same level, and compared to the AU and AD treatments, its content has increased by 21.35–27.81 times. Studies have shown that an increase in NH4+ can cause an increase in NO3− in the water [38,39]; however, in this study, after the addition of NH4Cl, the nitrate nitrogen content was almost zero when monitored. This is because when NH4Cl was initially added to the water in the drainage ditch, the NH4+ content rapidly increased. As seen in the AU treatment, before the addition of the nitrogen source, the DO content in the ditch was 2–4 mg/L, under which conditions the nitrification process can proceed [40], allowing NH4+ to be converted to NO3−, causing an increase in NO3− content. However, this process consumes a large amount of dissolved oxygen in the water, reducing the DO content to 0–2 mg/L, transitioning from an aerobic to an anaerobic environment. At this level of dissolved oxygen, the water no longer supports the conversion of NH4+ to NO3−, and the denitrification process begins to occur, converting all NO3− to NO2− or N2, resulting in the NO3− content in the water being almost zero. It is clear that an excess input of NH4+ will not only fail to increase the NO3− concentration in the water but will also lead to the complete consumption of NO3−.
Aeration slightly increased the pH value of the water, which is due to the increased dissolved oxygen content in the water during the aeration process. This oxygenation fully oxidizes acidic reducing substances, such as hydrogen sulfide. Additionally, aeration promotes the transfer of dissolved CO2 in the water into the air, which can reduce the concentration of carbonates in the water, thereby raising the pH value. The nitrogen source used in the experiment was NH4Cl, which is an acidic salt. When it dissolves in water, it produces NH4+ and Cl−. NH4+ can further hydrolyze to form NH3 and H+. It is evident that the input of NH4Cl increases the concentration of H+ in the water, thus lowering the pH value [41]. Therefore, the addition of NH4Cl solution will cause the pH value of the water to decrease. There was no significant change in the electrical conductivity of the water between the AU and AD treatments, because electrical conductivity is mainly influenced by the ionic concentration in the water, and aeration mainly increases the dissolved oxygen content in the water without directly changing the ionic concentration. Compared to the AU and AD treatments, the electrical conductivity of the water in the NU and ND treatments increased significantly. This is because when NH4Cl is added to the water, it ionizes into NH4+ and Cl−, thereby increasing the ionic concentration in the water and the electrical conductivity [42].
Except for the AD treatment, which showed a significantly higher oxidation-reduction potential (ORP), the ORP levels in the other treatments were similar, suggesting that aeration can increase the water’s ORP by elevating dissolved oxygen content, thus promoting the breakdown of organic matter, sulfides, and other reductive substances, and enhancing the water’s oxidizing capacity [43]. Adding NH4Cl to the water did not significantly affect its ORP. The AD treatment had a notably higher DO than the AU treatment. However, NH4Cl addition significantly reduced DO levels, as NH4+ is oxidized to NO2− and NO3−, consuming dissolved oxygen in the process [44]. It is evident that although the addition of NH4Cl caused a reduction in the DO content of the water, it did not have a significant impact on its ORP. This is because ORP is a complex parameter that is affected not only by DO but also by other chemical substances in the water, such as various ions, organic matter, and microorganisms [45]. Therefore, even if the dissolved oxygen content decreases, the ORP may not show a significant change. Additionally, considering the changes in the concentrations of ammonium nitrogen and nitrate nitrogen in the water, the conversion of NH4+ to NO3− and then to N2 involves two steps: nitrification and denitrification. Nitrification is the oxidation of NH4+ to NO3−, an oxidizing process that consumes dissolved oxygen in the water and may decrease the ORP value. However, at the same time, the produced NO3− is a strong oxidizing agent, which could increase the ORP value [46]. Subsequently, denitrification occurs in the water, reducing the NO3− produced from the oxidation of NH4+ and the original NO3− in the water to N2, a reduction process that consumes the oxidants (NO3−) in the water, so it could decrease the ORP value. Therefore, ORP may rise and fall during this process, but by the time of monitoring, the ORP in the water had stabilized and did not show significant changes.

4.1.2. Effects of Aeration and Nitrogen Source Input on the Physical and Chemical Properties of Drainage Ditch Sediment

Aeration has been observed to reduce the organic carbon content in the sediment, as the organic carbon in the drainage ditch sediment mainly exists in a form that is difficult to decompose, due to the typically anaerobic or low-oxygen environment, which results in lower microbial activity. However, aeration increases the concentration of oxygen in the sediment, creating an environment suitable for the growth of aerobic microorganisms, thus promoting the decomposition of organic carbon [47,48].
Aeration can increase the oxygen content in soil, which is beneficial for the growth and reproduction of aerobic microorganisms, and may increase the content of MBC in the soil. However, the subject of this study is the sediment in drainage ditches, which has been covered by water for extended periods and is typically an anaerobic environment, predominantly inhabited by anaerobic microorganisms. Aeration is not conducive to their growth and reproduction, which could lead to a decrease in sediment MBC content. Compared to the AU treatment, the MBC content in the sediment increased by 34.43% under the NU treatment due to the input of NH4Cl, because nitrogen is an important nutrient for microbial growth. An appropriate amount of nitrogen fertilizer can provide the nitrogen required by microorganisms, promoting their growth and reproduction [49], thereby increasing the content of soil MBC.
In this study, aeration led to a decrease in the DOC content in the sediment, which may be due to the fact that before aeration, the sediment was an anaerobic environment where the organic carbon existed in forms that were difficult to decompose. Aeration created an environment suitable for aerobic microorganisms, and in a short period, both organic carbon and DOC were rapidly decomposed, resulting in a decrease in DOC content [50]. Compared to the AU treatment, the DOC content in the NU treatment initially increased and then decreased, which could be because ammonium chloride is a source of nitrogen, and its addition can stimulate the activity of soil microorganisms. The enhanced microbial activity might accelerate the decomposition of organic matter, releasing more DOC, thereby causing a short-term increase in DOC content. Over time, soil microorganisms may begin to use these increased levels of DOC as an energy source, leading to a decrease in DOC content.
Aeration promotes nitrification. It facilitates the oxidation of ammonium nitrogen by nitrifying bacteria into nitrate, with aeration providing the necessary oxygen for this transformation, thus accelerating the conversion of ammonium nitrogen to nitrate nitrogen [51]. At the same time, aeration increases the oxygen content in the surface layer of the sediment, while the deeper layers are less affected and remain anaerobic, creating an environmental gradient conducive to both nitrification and denitrification processes. This promotes the conversion of nitrate nitrogen into N2, indicating that aeration can simultaneously promote the conversion of ammonium nitrogen and the loss of nitrate nitrogen. The addition of NH4Cl also led to an increase in both ammonium and nitrate nitrogen in the sediment. Compared to the AU treatment, the average content of ammonium and nitrate nitrogen in the NU treatment increased by 14.98 mg/kg and 26.65 mg/kg, respectively. It is evident that although the nitrogen source added was in the form of ammonium, the increase in nitrate nitrogen was higher than that of ammonium. Combined with the changes in nitrogen in the water body observed in this experiment, it can be inferred that after the addition of the nitrogen source, a large amount of ammonium nitrogen was first converted into nitrate nitrogen and then adsorbed by the sediment.

4.2. Effects of Aeration and Nitrogen Source Input on Greenhouse Gas Emissions in Drainage Ditches

4.2.1. Effects of Aeration and Nitrogen Source Input on N2O Emission in Drainage Ditches

The N2O emissions from all four treatments were at low levels, which is because the continuous flooding environment is conducive to the formation of NO3−, the final product of denitrification, and not conducive to the formation of the intermediate product, N2O [52,53].
Compared to the AU treatment, the AD treatment reduced the flux of N2O emissions. This is because aeration weakened the reductive environment in the soil, leading to a decrease in denitrification activity and thus a reduction in N2O emissions. Additionally, aeration caused an increase in the pH of the water body. Simek pointed out that the optimal pH range for natural denitrification is between 6.6 and 8.3, and within this range, an increase in pH is favorable for increasing the proportion of N2 in the denitrification products [54]. Studies have shown that a lower pH can inhibit the activity of the N2O reductase enzyme, thereby increasing the proportion of N2O in the denitrification products [55]. Therefore, a weakly alkaline environment is conducive to the complete reduction of N2O to N2, which can also explain the observed decrease in N2O emissions following aeration.
Compared to the AU treatment, the NU treatment significantly reduced the N2O emission flux. The input of nitrogen would increase the nutrients in the water body, thereby promoting nitrification and denitrification processes in the drainage ditch sediment, and increasing N2O emissions [56]. However, the results of this experiment showed the opposite conclusion: the input of NH4Cl actually led to a decrease in N2O emissions from the drainage ditch. The analysis of the DO content in the water body showed that at the initial stage of NH4Cl input, the nitrification process consumed a large amount of dissolved oxygen in the water body. During monitoring, the DO content in both the water and sediment further decreased. Although the input of NH4Cl increased the content of ammonium nitrogen and nitrate nitrogen in the sediment, providing ample substrate for nitrification and denitrification processes, N2O is an intermediate product of both processes. At this time, the reductive environment in the sediment made the denitrification process more complete, further reducing N2O to N2. Therefore, the observed N2O emissions were significantly reduced, showing a significant negative correlation with water conductivity (p < 0.05) and a strong positive correlation with dissolved oxygen in the water body. Moreover, as previously mentioned, an excessive input of NH4+ not only does not increase the concentration of NO3− in the water body but can lead to the complete consumption of NO3−. The input of NH4+ will inevitably cause a significant increase in NH4+ in the water body, hence N2O emissions showed a significant negative correlation with the concentration of ammonium nitrogen in the water body (p < 0.05) and a significant positive correlation with the concentration of nitrate nitrogen (p < 0.05).

4.2.2. Effects of Aeration and Nitrogen Source Input on CO2 Emissions from Drainage Ditches

The emission of CO2 in drainage ditches primarily originates from soil respiration, plant respiration, the oxidation of CH4, and the microbial decomposition of organic matter [57]. A high dissolved oxygen content in the water usually indicates that the water body is undergoing sufficient photosynthesis; therefore, the dissolved oxygen content in the water body is highly negatively correlated with CO2 emissions. Some studies have shown that artificial aeration increases the flux of CO2 emissions [58], because enhanced aerobic metabolism leads to more organic carbon being converted into CO2. Additionally, when aeration is introduced to the water body, the CO2 dissolved in the water may escape into the atmosphere [59], thereby increasing the collected concentration of CO2. Overall, the increase in CO2 concentration during aeration is the combined effect of these factors.
Compared to the AU treatment, the NU treatment significantly increased the CO2 emission flux. The input of NH4CL into water causes acidification, and the solubility of CO2 in water is reduced, leading to an increase in CO2 emissions. At the same time, the process of denitrification leads to the microbial decomposition of organic matter, thereby producing CO2. It is evident that while denitrification consumes nitrate nitrogen in the water body, it also produces CO2 [60], hence the significant negative correlation between the two (p < 0.05). Furthermore, NH4+ can be oxidized by nitrifying bacteria into NO2− and NO3−, a process that consumes dissolved oxygen in the water and produces CO2. Therefore, the oxidation of NH4+ increases the production of CO2. Simultaneously, the input of NH4+ leads to a significant rise in water conductivity, increasing the nutrients in the water body and providing additional nutrition for microbial growth and reproduction, which could stimulate the decomposition of organic matter and microbial respiration in the water body, producing more CO2. Consequently, CO2 emissions showed a significant positive correlation with both the concentration of ammonium nitrogen in the water body and water conductivity (p < 0.05).

4.2.3. Effects of Aeration and Nitrogen Source Input on CH4 Emissions in Drainage Ditches

CH4 is produced by methanogenic archaea in anaerobic environments as they decompose organic matter, and soil organic carbon is an important substrate for the production of CH4 [61,62]; therefore, the emission of CH4 is highly positively correlated with the content of soil organic carbon, and soil ammonium nitrogen can promote the growth and reproduction of methanogenic archaea [63], indicating a high positive correlation between CH4 emissions and soil ammonium nitrogen content. After its production, more than 80% of CH4 is oxidized by aerobic methanotrophic bacteria before it is emitted into the atmosphere during its transport from the soil to the atmosphere [64]. It is evident that the higher the ORP of the water body, the more CH4 is oxidized, hence there is a significant negative correlation between CH4 emissions and the water body’s ORP (p < 0.05).
Comparing the AU and AD treatments, it is evident that aeration significantly reduces CH4 emissions. The reduction in CH4 emission flux due to aeration occurs through various pathways. Firstly, aeration weakens the reductive environment in the soil, and methane is mainly produced under strongly reductive conditions; thus, aeration reduces the production of methane. Secondly, CH4 emissions mainly occur by forming bubbles that rise to the surface and into the atmosphere, and the dissolved oxygen (DO) content in the water increases due to aeration, leading to a large amount of CH4 being oxidized during the emission process. Therefore, aeration results in a significant reduction in the amount of CH4 emitted into the atmosphere. Compared with the AU treatment, the sediment in the NU treatment has a higher content of nitrate nitrogen due to the input of NH4Cl. The increase in NO3− promotes denitrification, which in turn promotes the growth and reproduction of denitrifying bacteria. As denitrifying bacteria are also anaerobic, their growth and reproduction compete with methanogenic archaea for organic matter in the sediment, inhibiting the activity of methanogenic archaea and thereby suppressing the production of CH4 [65].

4.3. Migration and Transformation of Carbon and Nitrogen in Drainage Channels and Its Impact on Greenhouse Gas Emissions

Comparing the NU and ND treatments, it is evident that there are no significant differences in the trends or magnitudes of changes in pH, ORP, temperature, electrical conductivity, and DO between the upstream and downstream water bodies. Additionally, the content of ammonium nitrogen in the water only slightly decreases. However, the content of nitrate nitrogen and ammonium nitrogen in the upstream sediment is significantly higher than downstream. This is because, despite the sediment’s continuous adsorption of NH4+ [66], the consistent input of NH4Cl solution upstream means that the concentration of NH4+ ions in the water does not change much. The upstream sediment can preferentially adsorb NH4+, which is the substrate for nitrification, and its increased concentration can promote the nitrification process, thereby also increasing the concentration of NO3− in the sediment.
The content of organic carbon, microbial biomass carbon, and soluble organic carbon in the upstream sediment is higher than downstream. This is because, compared to downstream, the upstream part of the drainage ditch receives more sewage and sediment, hence more organic matter. Additionally, the water flow upstream is more turbulent, causing organic carbon to settle before it can be decomposed, whereas the slower flow downstream allows more time for microbial decomposition. Studies have shown that slower water flow downstream is conducive to microbial growth. The faster upstream flow is not favorable for the accumulation of microbial numbers. However, in this experiment, the microbial biomass carbon upstream is higher than downstream. This is due to the input of NH4Cl, which significantly increases the content of ammonium nitrogen and nitrate nitrogen in the sediment. Considering the nitrogen content in the upstream and downstream sediment, the increase is greater upstream, indicating that upstream microbes can preferentially utilize these nutrients, leading to a rapid increase in upstream microbial biomass. Conversely, downstream microbes receive relatively less NH4Cl, resulting in less growth. Similarly, the input of NH4Cl promotes an increase in the number of upstream microbes, and microbial metabolism and death lysis produce more dissolved organic carbon (DOC), hence the DOC content is also higher upstream compared to downstream.
Analyzing the greenhouse gas emission data from the NU and ND treatments, it can be seen that under the input of NH4Cl, the emission trends of N2O, CO2, and CH4 in the upstream and downstream sections of the drainage ditch are basically consistent. However, compared to upstream, the emissions of these three greenhouse gases downstream have decreased, with CH4 emissions showing a higher reduction. Considering the carbon and nitrogen content in the water and sediment, the high nitrogen content upstream provides an abundant nitrogen source, promoting denitrification and nitrogen removal processes, and generating more N2O. Downstream, with less nitrogen available, less N2O is produced compared to upstream. Additionally, the high organic matter content upstream provides ample substrates for methanogenic bacteria, resulting in more CH4 production. Downstream, with less organic matter, there is less CH4 production. Moreover, upstream microbial activity is higher, leading to greater CO2 production. Downstream, with fewer microbes and reduced activity, CO2 production is lower [67]. The mineralization of organic matter in upstream sediments is strong, producing more CO2. Downstream, with less organic matter, mineralization is weaker [68].
In summary, the drainage ditch has a certain self-purification capability. The growth and metabolism of microbes, algae, and other organisms in the ditch can degrade and absorb organic matter and nutrients brought in from upstream, reducing the material basis for the generation of greenhouse gases. Additionally, the sediment and deposits in the ditch have a certain adsorption effect on pollutants. Finally, as suspended substances move through the water, they continuously settle, reducing the amount of pollutants [69]. Therefore, the drainage ditch can reduce the pollutant content in the water and sediment. In this process, some pollutants are transformed into greenhouse gases and emitted into the atmosphere. At the same time, the concentration of pollutants downstream decreases, and the intensity of greenhouse gas emissions also declines. In this experiment, a 10 m length of drainage ditch can reduce the combined effect of greenhouse gas emissions by 9.34%.

5. Conclusions

This study focused on paddy field drainage ditches to explore the effects of nitrogen input and aeration on the emissions of greenhouse gases from agricultural drainage ditches and the concentrations of pollutants in the water. Through experimentation and theoretical analysis, the conclusions are as follows:
(1)
Aeration significantly reduced the concentration of ammoniacal nitrogen in the water and decreased the carbon and nitrogen content in the drainage ditch sediments to varying degrees. At the same time, it reduced the emissions of N2O and CH4, with a 72.39% reduction in CH4 emissions. Although there was a slight increase in CO2 emissions, the overall global warming potential was reduced by 34.02%.
(2)
Nitrogen input significantly raised the concentration of ammoniacal nitrogen in the water and the content of readily available nitrogen in the sediments, which had a certain suppressive effect on the emissions of N2O and CH4. However, it led to a 46.60% increase in CO2 emissions, resulting in an overall 15.24% increase in the GWP.
(3)
Drainage ditches play a role in reducing pollution loads, with the concentration of pollutants and the intensity of greenhouse gas emissions downstream being lower than upstream.
(4)
The emissions of N2O and CO2 in the paddy field drainage ditches were significantly related to the electrical conductivity of the water and the concentration of readily available nitrogen in the water, while CH4 had a significant relationship with the ORP of the water.
This experiment explored the diurnal variation of greenhouse gas emissions and water pollutant concentrations in drainage ditches. Future research should investigate the effects of different aeration intensities (amount of aeration, frequency), and various types and concentrations of pollutant load inputs on agricultural drainage ditches over longer time scales, such as one to two crop growing seasons.

Author Contributions

Q.Z.: data collection, data analysis and interpretation, experimentation, and writing—original draft. J.W. (Jingwei Wu): resources and funding acquisition. C.G.: conceptualization, writing—review and editing. Y.H.: funding acquisition. Q.L.: funding acquisition. J.W. (Jing Wang) and Y.Z.: assistance in experimentation. All authors have read and agreed to the published version of the manuscript.

Funding

This work was jointly supported by the National Natural Science Foundation of China (Grant Nos. 52379047 and 52209067), the National Key Research and Development Program of China (Grant No. 2021YFD1900804), a project funded by China Postdoctoral Science Foundation (2022M712467), and a project titled ‘Eco-hydrological model and flood and drought disaster risk assessment in the middle reaches of the Yellow River under changing environment’ (U2243226).

Data Availability Statement

All the data and codes used in this study can be requested by email to the corresponding author Chenyao Guo at [email protected].

Conflicts of Interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

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Figure 1. Layout of the experiment (the red dots represent the sampling sites). (a) Top view of static box arrangement; (b) Aeration test layout; (c) Nitrogen source input layout.
Figure 1. Layout of the experiment (the red dots represent the sampling sites). (a) Top view of static box arrangement; (b) Aeration test layout; (c) Nitrogen source input layout.
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Figure 2. Available nitrogen content in water. (a) AU- and AD-treated water ammonium nitrogen content; (b) AU- and AD-treated water nitrate nitrogen content; (c) NU- and ND-treated water ammonium nitrogen content.
Figure 2. Available nitrogen content in water. (a) AU- and AD-treated water ammonium nitrogen content; (b) AU- and AD-treated water nitrate nitrogen content; (c) NU- and ND-treated water ammonium nitrogen content.
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Figure 3. Physical and chemical properties of water body. (a) Water pH; (b) conductivity of water; (c) dissolved oxygen content; (d) water temperature; and (e) water REDOX potential.
Figure 3. Physical and chemical properties of water body. (a) Water pH; (b) conductivity of water; (c) dissolved oxygen content; (d) water temperature; and (e) water REDOX potential.
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Figure 4. Carbon content of sediment. (a) Organic carbon content in sediment; (b) microbial carbon content in sediment; and (c) sediment soluble carbon content.
Figure 4. Carbon content of sediment. (a) Organic carbon content in sediment; (b) microbial carbon content in sediment; and (c) sediment soluble carbon content.
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Figure 5. Available nitrogen content in sediment. (a) Ammonium nitrogen content in sediment; and (b) nitrate nitrogen content in sediment.
Figure 5. Available nitrogen content in sediment. (a) Ammonium nitrogen content in sediment; and (b) nitrate nitrogen content in sediment.
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Figure 6. N2O emission flux(mg·m−2·h−1). (a) N2O emission fluxes under the AU and AD treatments; and (b) N2O emission fluxes under the NU and ND treatments.
Figure 6. N2O emission flux(mg·m−2·h−1). (a) N2O emission fluxes under the AU and AD treatments; and (b) N2O emission fluxes under the NU and ND treatments.
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Figure 7. CO2 emission flux (mg·m−2·h−1). (a) CO2 emission fluxes under the AU and AD treatments; and (b) CO2 emission fluxes under the NU and ND treatments.
Figure 7. CO2 emission flux (mg·m−2·h−1). (a) CO2 emission fluxes under the AU and AD treatments; and (b) CO2 emission fluxes under the NU and ND treatments.
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Figure 8. CH4 emission flux (mg·m−2·h−1). (a) CH4 emission fluxes under the AU and AD treatments; and (b) CH4 emission fluxes under the NU and ND treatments.
Figure 8. CH4 emission flux (mg·m−2·h−1). (a) CH4 emission fluxes under the AU and AD treatments; and (b) CH4 emission fluxes under the NU and ND treatments.
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Table 1. Emissions of three greenhouse gases and GWP.
Table 1. Emissions of three greenhouse gases and GWP.
Type of GasEmission Flux (mg/m2·d)GWP (mg/m2·d)
AUADNUNDAUADNUND
N2O1.531.290.370.34456.54384.72110.71101.92
CO215,142.4516,720.5422,198.9621,304.7515,142.4516,720.5422,198.9621,304.75
CH4719.14201.95655.44547.0317,978.505048.6316,385.8813,675.75
The total33,577.4922,153.8838,695.5435,082.42
Table 2. The emission of three kinds of greenhouse gases and the significance analysis of soil and water physicochemical properties.
Table 2. The emission of three kinds of greenhouse gases and the significance analysis of soil and water physicochemical properties.
Physicochemical Properties of Soil and WaterIndexN2O EmissionCO2 EmissionCH4 Emission
Soil organic carbon content (g/kg)Pearson correlation 0.088−0.0650.940 ###
Significance 0.9120.9350.060
Soil ammonium nitrogen content (mg/kg)Pearson correlation −0.371 #0.417 #0.893 ###
Significance 0.6290.5830.107
Soil nitrate nitrogen content (mg/kg)Pearson correlation −0.640 ##0.727 ##0.377 #
Significance 0.3600.2730.623
Soil MCB content (mg/kg)Pearson correlation −0.354 #0.433#0.726 ##
Significance 0.6460.5670.274
Soil DOC content (mg/kg)Pearson correlation −0.1970.2870.676 ##
Significance 0.8030.7130.324
Conductivity of water (μs/cm)Pearson correlation −0.986 *###0.980 *###0.355#
Significance 0.0140.0200.645
Dissolved oxygen in water (mg/L)Pearson correlation 0.933 ###−0.919 ###−0.535 ##
Significance 0.0670.0810.465
ORP (mv)Pearson correlation 0.316 #−0.283−0.969 *###
Significance 0.6840.717·0.031
Nitrate nitrogen content in water (mg/L)Pearson correlation 0.973 *###−0.961 *###−0.417 #
Significance 0.0270.0390.583
Ammonium nitrogen content in water (mg/L)Pearson correlation −0.986 *###0.976 *###0.361 #
Significance 0.0140.0240.639
* At 0.05 level, the correlation was significant. # Low correlation (0.3 < r < 0.5); ## Moderate correlation (0.5 < r < 0.8); ### Highly correlated (r < 0.8).
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Zhang, Q.; Wu, J.; Guo, C.; Wang, J.; Zhao, Y.; Li, Q.; Hu, Y. Effects of Nitrogen Input and Aeration on Greenhouse Gas Emissions and Pollutants in Agricultural Drainage Ditches. Agronomy 2024, 14, 235. https://doi.org/10.3390/agronomy14020235

AMA Style

Zhang Q, Wu J, Guo C, Wang J, Zhao Y, Li Q, Hu Y. Effects of Nitrogen Input and Aeration on Greenhouse Gas Emissions and Pollutants in Agricultural Drainage Ditches. Agronomy. 2024; 14(2):235. https://doi.org/10.3390/agronomy14020235

Chicago/Turabian Style

Zhang, Qisen, Jingwei Wu, Chenyao Guo, Jing Wang, Yanchao Zhao, Qiangkun Li, and Yawei Hu. 2024. "Effects of Nitrogen Input and Aeration on Greenhouse Gas Emissions and Pollutants in Agricultural Drainage Ditches" Agronomy 14, no. 2: 235. https://doi.org/10.3390/agronomy14020235

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