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Article

Mitigation of GHG Emissions from Soils Fertilized with Livestock Chain Residues

by
Alessandra Lagomarsino
1,*,
Massimo Valagussa
2,
Carla Scotti
3,
Lamberto Borrelli
3,
Claudia Becagli
1 and
Alberto Tosca
4
1
Centro di Ricerca Agricoltura e Ambiente, Consiglio per la Ricerca in Agricoltura e l’Analisi dell’Economia Agraria, via di Lanciola 12/A, 50125 Firenze, Italy
2
MAC Minoprio Analisi e Certificazioni SRL, 22070 Vertemate, Italy
3
Centro di Ricerca Zootecnia e Acquacoltura, Consiglio per la Ricerca in Agricoltura e l’Analisi dell’Economia Agraria, 20075 Lodi, Italy
4
Fondazione Minoprio, Viale Raimondi 54, 22070 Vertemate, Italy
*
Author to whom correspondence should be addressed.
Agronomy 2022, 12(7), 1593; https://doi.org/10.3390/agronomy12071593
Submission received: 31 May 2022 / Revised: 27 June 2022 / Accepted: 27 June 2022 / Published: 30 June 2022
(This article belongs to the Special Issue Organic Waste Recycling as Fertilising Practice in Agronomic Ecosystems)
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Versions Notes

Abstract

:
Using livestock residues as alternative fertilizers is a sustainable practice which recycles nutrients that would otherwise be lost. However, organic fertilizers may have a large impact on N2O emissions, offsetting the beneficial effects of C sequestration. After four years from biochar application, greenhouse gas fluxes were monitored for two years from a Maize field fertilized with digestate, slurry, or urea, with and without biochar. The objectives of the present study were to assess (i) the climate feedback of using residues from the livestock chain as alternative fertilizers and (ii) the contribution of biochar in mitigating GHGs emissions, while increasing the organic C in soil. Digestate was shown to have the highest impact on CO2 and N2O emissions from soil, with respect to mineral fertilization (+29 and +142%), more than slurry (+21 and −5%), whereas both residues positively affected CH4 uptake (+5 and +14%, respectively). The maximum N2O peaks occurred between 7–20 days after fertilization, accounting for 61% of total emissions, on average. Biochar was effective in reducing N2O emissions derived from mineral fertilization and digestate (−54% and −17%, respectively). An excess of labile organic matter and N induced the highest CO2 emissions and N2O peaks, independent of—or even triggered by—biochar. Mitigation of GHG emissions, from soils fertilized with livestock chain residue, can be obtained using biochar, but with limitations dependent on (i) the quantity of organic matter added, (ii) its quality, and (iii) the time from application: those aspects that deserve further investigations.

1. Introduction

Currently, the agriculture sector accounts for 30–35% of global Greenhouse Gas (GHG) emissions, and it is the largest contributor to anthropogenic non-CO2 GHGs [1,2], largely due to methane (CH4) emissions from livestock and rice cultivation, and nitrous oxide (N2O) emissions from fertilized soils. It is estimated that approximately 60% of the global anthropogenic N2O emissions is from cultivated soil [3,4], either as a direct consequence of synthetic N fertilizers, crop residues and manure, or indirectly from leached N compounds, such as nitrates [5,6,7,8]. Given its high global warming potential, N2O emissions represent the highest risk of negating the beneficial effects of increasing organic matter (OM) inputs to enhance C sequestration [9].
The use of N-fertilizers directly influences the amount of NH4+ or NO3− available in soil, which, in turn, affects nitrification [10,11,12] and denitrification [13,14,15] activities. It has also been shown that the intensity of N2O emissions is related to N fertilizer rate [16]; soil N excess, such as N amounts not available to plants, lead to increasing N2O emissions [17].
In the vision of circular economy, using crop or livestock residues is a sustainable strategy, improving soil organic matter and recycling nutrients that otherwise would be lost from the soil system. Organic fertilization includes different types of organic matter, such as animal manures, slurry, crop residues, green manures, and compost [18]. Depending on the type of residue, organic fertilizers generally have a large impact on N2O emissions, even when compared to mineral fertilizers [19]. To predict N2O emission, the quality and quantity of crop residues being applied to agricultural fields should be linked to soil properties and pedo-climatic conditions [20,21,22]. The application method is significant as well, with the highest emissions occurring with surface application, in comparison to techniques that rely upon incorporation further into the soil [19,23].
Anaerobic digestion of agricultural materials and sludges is a promising strategy for generating products that can be applied to enhance the productivity of agricultural soils, being rich in plant available nutrients [24], but only few relevant studies have been conducted [25,26,27]. Overall, a short-term increase in N2O emissions is reported, depending on soil texture, moisture, and temperature [28], as well as digestate composition, soil type, and NH4+ availability [27]. The increase can be significantly reduced by adjusting the timing of application, controlling pH, and combining digestate with other amendments, such as biochar [29].
Biochar addition to soil has been proposed as a method to increase soil C storage and reduce N2O emissions from soil on a global scale [30,31,32]. Biochar has been reported to decrease the denitrification process [33,34], while positively affecting nitrification [35]. Several possible mechanisms driving reduction in N2O emissions from soil after biochar application have been proposed, including: (i) change of soil structure, improving soil aeration, and/or reducing soil moisture, which inhibit denitrification through enhancing O2 concentration [33,35,36]; (ii) sorption of inorganic-N substrates no longer available for nitrification and denitrification processes [37,38]; (iii) increase in soil pH that may stimulate the N2O reductase enzyme activity, increasing N2 production and N2/N2O ratio [36,39,40]; (iv) the capacity of biochar to act as electron acceptor through Mn(IV) and Fe(III) surface sites, therefore promoting the last step of denitrification pathway [33,41]. However, large gaps of knowledge exist on the real mitigation capacity, depending on the type of applied fertilizers and the processes involved.
Actually, there is a good understanding regarding the effect of specific soil management strategies on either C sequestration or non-CO2 GHG fluxes [42]. Nevertheless, integrated and comprehensive investigations, considering the trade-off between C sequestration and the outputs, including the three GHGs, are still scarce [43]. The objectives of the present study were to assess (i) the climate feedback of using residues from livestock chain (slurry and digestate) as alternative fertilizers, (ii) the contribution of biochar in mitigating GHGs emissions, while increasing the organic C in the soil. With our approach, we intend to cover knowledge gaps on the mitigation potential of the interactions between livestock chain residues and biochar, assessing the dynamics of mineral N lost into the groundwater or in the atmosphere.

2. Materials and Methods

2.1. Experimental Site

The experimental site was established in 2018 at Cascina Baroncina (Lodi, Po Valley, Italy; 45°17′25′′ N−9°29′43′′ E−81.5 m asl) on a sandy loam soil with a sub-acid reaction. Main soil characteristics are reported in Table S1. The Lodi area lies in a “nitrate vulnerable zone” (according to the 91/676/EEC Directive against pollution caused by nitrate from agricultural sources) in which a maximum of 170 kg ha−1 of N, available for plants from livestock effluents, is allowed. The daily air temperature (mean, max, and min values), rainfall, and irrigation interventions during the 2020 and 2021 growing seasons are reported in Figure S1. The crop rotation system is represented by silage maize–Italian ryegrass characterizing the intensive dairy cattle farming regions of the Po Valley (Figure S2). The Pioneer hybrid 1547 class 600 was sown in June (Table 1) and harvested after 90 and 99 days in 2020 and 2021, respectively. The Italian ryegrass cv. Star was sown at the beginning of October and harvested in May.
The treatments investigated in the trial involved the application of biochar from pyrolysis processes in 2018, at a dose of 20 t ha−1 of dry matter (DM), and incorporated it into the soil by ploughing 30 cm depth. Biochar main characteristics are reported in Table S2.
Starting from 2018, and for the following 4 years, digestate and slurry were spread in springtime each year and incorporated into the soil by ploughing within 24 h. Main characteristics are reported in Table 2. Urea distribution was split, with one third at sowing and the remaining part at one in-season application (Table 1). The amount of fertilizer supplied was calculated to provide 170 kg ha−1 N available for a plant, considering 100% of urea and of the N-NH4 fraction of the organic fertilizers, as well as 50% of the organic N fraction (Table 2).
In the case of mineral fertilization, 80 kg ha−1 P2O5 and 180 kg ha−1 K2O were also distributed at sowing. The Italian ryegrass was grown without fertilization to highlight possible residual nutrients present into the soil after the maize cropping.
Within the field, 2 blocks were arranged in strips, having received the same agronomical treatment for 20 years. In each block, six plots were established, considering mineral fertilization, digestate, and slurry with and without biochar. In each plot, 2 chambers were installed as sub-replicates, for a total of 24 chambers (2 blocks × 2 sub−replicates × 3 fertilization treatments × with and without biochar).

2.2. Crop and Soil Analyses

Crop yield has been assessed at harvest, manually cutting a subsample of 5 m2 (about 30 plants) for each plot, which was chopped and dried at 60 °C to measure the dry matter percentage of silage.
Post-harvest soil samples were taken from each individual parcel (0–30 cm deep), dried at 40 °C in a ventilated oven, and sieved at 2 mm. Total organic C was determined on a homogenized fine fraction (500 μm), with an elemental analyzer (Flash Smart NC Soil, Thermo Fisher, Waltham, MA, USA), after subtracting inorganic C from total C.
Extractable mineral N was determined after collecting the samples at a 0–30 cm depth and transporting them to the laboratory in a refrigerated container. Samples have been collected in the first six GHGs monitoring events of 2020 and in the first 6 weeks in 2021. The NH4 and NO3 fractions of mineral N were determined by extraction with 2M KCl and subsequent distillation (UDK 132, Velp Scientific, Usmate Velate, Italy), with the addition of Devarda for the determination of nitric N. Samples moisture, which was also determined to express the data on dry matter.
Cumulated NH4 and NO3 values were assessed by means of ion-exchange resin lysimeters and were assembled using pure resin PMB 101-3 (with an ion exchange capacity of 1.3 eq/L min). Additionally, 15 g of resin were enclosed in a nylon bag with sufficient fine mesh to prevent material from escaping. The nylon bag was placed between two layers of glass beads (2–3 mm diameter) and enclosed in a polyvinyl chloride pipe section (5.3 cm diameter and 3 cm height), all enclosed in a mesh net. In 2020, at sowing, in each plot, a lysimeter was placed at 40 cm deep and subsequently recovered at harvest. In 2021, two lysimeters for plot (excluding those with slurry) were placed at sowing and recovered at harvest. The collected lysimeters were opened in the laboratory, and N was extracted by washing the resin with 2 M KCl extraction solution (ratio 1:10), in 500-mL Erlenmeyer flasks on an orbital shaker, at 100 rpm for 1 h. Extracts were filtered (Whatman no. 42 filters) and analyzed for NO3 and NH4 concentration through distillation (UDK 132, Velp Scientific), with the addition of Devarda for the determination of nitric N.

2.3. GHG Fluxes

During each maize growing season, 17 gas samplings were performed (see Table 1 for starting dates) by means of the closed chamber method described in [44], for measuring soil CO2, CH4, and N2O fluxes. Gas sampling was performed between 9.30 and 13.30 to minimize changes in soil CO2 effluxes associated with diurnal cycles [45]. During each gas sampling event, chambers were closed for 30 min with four gas samplings (at 0, 10, 20, and 30 min). Additionally, 25 mL of headspace gas were collected with air-tight 30 mL propylene syringes and were immediately pressurized into pre-evacuated 12 mL glass Exetainer® vials (Labco Ltd., Buckinghamshire, UK). Soil temperature and moisture were also measured at each time. Soil temperature was measured with a probe (HI 9043 Hand-Held Thermometer, Hanna Instruments, Woonsocket, RI, USA) and volumetric soil moisture was measured with a ML3 ThetaProbe Soil Moisture Sensor (Delta-T Device, Cambridge, UK) at a depth of 5 cm.
Gas samples were analyzed within 4 weeks of collection using a GC-2014 gas chromatograph (Shimadzu Scientific, Kyoto, Japan) with a thermal conductivity detector for CO2, 63Ni electron capture detector for N2O and flame ionization detector for CH4. Chamber gas concentrations were converted to mass per volume unit using the Ideal Gas Law and measured chamber air temperatures and volumes. Fluxes of CO2, N2O and CH4 were calculated using the slope of linear regression of gas concentration vs. chamber closure time and the enclosed soil surface area. Fluxes were set to zero if the change in gas concentration during chamber enclosure fell below the minimum detection limit of GC, and flux values were rejected (i.e., treated as missing data) if they passed the detection test but had a coefficient of determination (R2) < 0.90. Estimates of cumulative CO2, N2O and CH4 fluxes for each field replicate were based on linear interpolation between sampling dates during each growing season [46].

2.4. Statistical Analysis

Mixed model analysis of variance (ANOVA) module of Statistica package (StatSoft Inc., Tulsa, OK, USA) and Fisher’s LSD Post-Hoc test were performed to evaluate the effects of fertilizers and biochar on CO2, N2O and CH4 fluxes, SOIL TOC, crop yield, NH4 and NO3 availability in soil in 2020 and 2021. Univariate Tests of Significance for CO2, N2O and CH4, with Sigma-restricted parameterization for each year was performed for cumulative values.

3. Results

3.1. Crop Yield and Soil Organic C

Overall, 11% increase in maize yield was observed in 2021 with respect to the previous year, although it was not significant (Table 3). Among treatments, the mineral fertilization showed the lowest values with respect to all other treatments in 2020, and the highest with respect to mineral fertilization plus biochar in 2021.
Soil TOC (Table 3) remained stable in the 2 years of measurement, with a slight not significant increase in 2021. Among treatments, there was a clear effect of biochar addition in 2020, with all fertilizers, and in 2021, with slurry and mineral fertilization. No significant differences among fertilizers were observed.

3.2. NH4 and NO3 Dynamics and Accumulation in Soil

Inorganic N showed slight oscillations in the first 2 weeks of 2020 and in the first month of 2021 growing seasons, with NO3 values about two times higher than those of NH4 (Figure 1). Higher NH4 values were observed with digestate than mineral, 2–3 weeks after fertilization in 2020. Mineral fertilization + biochar showed the highest NH4 content after the top dress fertilization in 2021 growing season. Digestate + biochar showed higher NO3 values immediately and 1 month after initial fertilization in 2021. Mineral fertilization + biochar showed lower values of NO3 in one event of each growing season.
Cumulative NH4 reached the maximum values with mineral fertilization + biochar, which is significantly different from those without biochar. NO3 values did not show significant differences, although a peak was observed with digestate + biochar (Table 3).

3.3. GHGs Fluxes

CO2 emissions (Figure 2) showed a decreasing trend in the two growing seasons, with the maximum emissions observed in the first 10 days after digestate and slurry application, followed by a slow decrease. A significant correlation with temperature (p < 0.05) was observed. The increase in CO2 emissions with digestate was significant in four events in the second half of the 2020 growing season and throughout 2021 (15 out of 17 events).
Slurry induced an increase in CO2 emissions in seven events during the 2020 growing season and in eight events at the beginning and at the end of the 2021 growing season (Table 4). Biochar significantly reduced CO2 emissions from soils fertilized with digestate in most of the 2021 growing season. CO2 emissions from soils fertilized with slurry increased with biochar at the beginning of the 2021 growing season and decreased at the end.
Overall, cumulative CO2 emissions (Figure 3 and Table 5) were higher from soil with digestate and slurry than from soil with mineral fertilization in both years, with the largest increase in 2020 with slurry and in 2021 with digestate. The effect of biochar was significant in 2021, with a reduction in CO2 emissions from soil with digestate and an increase with slurry.
CH4 uptake was prevalent throughout the 2020 growing season and for most of the monitoring events in 2021. CH4 emissions were observed in the first two events, after fertilization in digestate and slurry + biochar plots, and in two events in the second half of the growing season with digestate (Figure 4, Table 4). CH4 uptake increased significantly with digestate, with respect to mineral, in five events in 2020, and with slurry in two events in 2020 and three events in 2021. Biochar increased CH4 uptake with slurry in four events in 2020, as well as with mineral fertilization in two events in 2020 and one event in 2021. CH4 emissions decreased with biochar, after fertilization with digestate, and increased after fertilization with slurry in 2021.
Overall, cumulative CH4 uptake was higher with digestate in 2020 and with slurry in 2021, with respect to mineral fertilization. Conversely, biochar increased CH4 uptake with slurry in 2021 and with digestate in 2021 (Figure 3).
N2O emissions showed a decreasing trend in both seasons, with peaks occurring 5–10 days after organic and mineral fertilization events (Figure 5).
In 2020, the highest peaks were observed with digestate, with and without biochar, 6–8 times higher than in 2021. The high heterogeneity in the field plots reduced the significance, and N2O emissions with digestate were higher than those with mineral fertilization in the first event of 2020 and in the first four events of 2021 (Table 4). After the top dress fertilization in 2021, an inverse trend was observed, with higher emissions from mineral fertilization than digestate or slurry in three and four events, respectively. Biochar + digestate induced an increase in N2O emissions in the first half of 2020 (two significant events), and a decrease in the second half of 2020 (two significant events) and in eight events of 2021. Biochar + slurry induced a significant increase in three events at the beginning of 2021. Biochar with mineral fertilization significantly reduced N2O emissions after the top dress fertilization in both years.
Cumulative N2O emissions were generally lower in 2021 than 2020, with the exception of those from slurry fertilized soil (Figure 3). Digestate induced a higher increase, with respect to mineral fertilization, in both growing seasons. The reduction in N2O emissions with biochar was significant with digestate in 2021 and with mineral fertilization in both growing seasons.
N2O emissions showed a significant linear relation with the effective N (N-NH4 + ½ N org), added with digestate and slurry (Figure 6 top), and a significant exponential regression with NO3 availability in soil (Figure 6, bottom).

4. Discussion

4.1. Fertilizer Types

Using livestock residues is a sustainable practice, which improves soil organic matter and recycles nutrients that, otherwise, would be lost from the soil system. However, depending on the type of residue, negative feedback to climate may occur, mainly mediated by increased N2O emissions [19]. The residues used in our study were different in composition between them, being digestate N richer than slurry, and between years. This also influenced the application doses, which varied between years depending on the effective N availability, which, in turn, affected the input of organic matter to soil. The higher NH4 content of digestate than slurry is commonly reported [47], and it is considered positive in a fertilizer, being immediately available to plants. However, evidence on negative environmental feedbacks through NH3 [48] and N2O [49] emissions, as well as N leaching [50], are rising. In both years of the present study, the increase in CO2 and N2O emissions from soils fertilized with digestate with respect to mineral fertilization, was higher than that from soils fertilized with slurry. Ambiguous results on the impact of digestate and slurry on N2O emissions are also reported in literature because of the large variability of residues composition and amounts. Higher [51] and lower [52,53] N2O emissions from digestate than slurry were found, as well as no effects [54,55]. N2O emissions negatively correlate with the C/N ratio of residues [56], which also explain the higher rates of N2O emissions from digestate (N richer) than slurry found in our study. Baral et al. [57] suggested that nitrification in manure hotspots was one of the factors controlling N2O emissions, either directly or indirectly, supplying substrate for denitrification. The higher initial NH4 content of digestate can have therefore triggered N2O emissions, as suggested by [49] and confirmed by the positive correlation with the added effective N (N-NH4 + ½ N org) we found. Moreover, an increase in NO3 content in soil due to nitrification might have occurred, which is reported as inhibitory on N2O reduction during coupled denitrification by several authors, especially when combined with a simultaneous depletion of readily available C [53,58]. This effect was confirmed by the exponential increase in N2O emissions with cumulated NO3 found in our study.
The N2O peaks and the highest distance between organic and mineral fertilizers occurred at the beginning of the growing seasons, within 7 and 24 days from the application, and fluxes gradually decreased afterwards. This pattern was consistent with several other studies, as reported by [53]. After this transient increase, a reduction in N2O from digestate and slurry with respect to mineral fertilization occurred, although not large enough to compensate the initial increase. This temporal variability might also explain the mixed effects found in literature.
In the first few days after application to soil, digestate, and slurry caused CH4 emissions too, which were visible in 2021 (4 days after fertilizers application) but not in 2020 (8 days after fertilizers application). As already found in other studies, maximum values of CH4 emissions were observed immediately after digestate [59] and slurry [54,60] application to the soil, followed by a sharp decrease and then uptake. The rapid and transient increase might be related to (i) fresh and easily decomposable organic substrate with (ii) high water content, which provided optimal conditions for CH4 emissions. After the initial flush, CH4 uptake was prevalent and further enhanced by digestate in 2020 and slurry in 2021. In the context of mineral upland environments, agricultural soils play an important role as CH4 sink [61,62,63]. Interactive effects of fertilizers on CH4 emission are complex and sometimes contradictory, depending on the nature of the fertilizer, the quantity applied, and the method of application [64]. Mineral N fertilizer application can also have a negative effect on the CH4 oxidation capacity of arable soils [65,66]. In addition, the impact of organic fertilizers on C cycling was significant in the short-term. In fact, soil organic C was not affected by organic fertilizer’s addition, implying that the external input of organic matter was rapidly decomposed and lost, as CO2 in the atmosphere or leached dissolved organic matter, as evident from the 17 to 42% increase in CO2 emissions, with respect to mineral fertilization. Indeed, other studies did not observe effects on TOC after digestate or slurry application to soil [67,68], and a priming effect was hypothesized, after organic residue application [69,70], that explains higher CO2 losses.

4.2. Biochar Contribution

The positive effect of biochar on C sequestration in soil is widely demonstrated through (i) increased stabilization of soil organic C, (ii) deepening of soil organic C distribution, and (iii) higher crop yield and above/belowground productivity, as reviewed by [71]. After 3 and 4 years from biochar application, a significant increase in soil organic C was found, independent of fertilizer type, even without significant effects on maize yield. The increase seemed, therefore, to be the result of a direct effect of C added with biochar, which was then rapidly stabilized into the soil.
The impact of biochar on GHG fluxes is more uncertain, varying among soil and crop types, biochar feedstock source, and pyrolysis temperature, as synthesized by the meta-analysis of [72]. Additionally, differences among the three gases were found, with several studies reporting an initial increase in CO2 emissions [73], a decrease in N2O fluxes [32], and no effects on CH4 fluxes [72]. These general trends are only partly confirmed by results found in the present study, which were dependent on the type of fertilizer, its quantity, and the time from application. A clear decrease in CO2 emissions was observed only in the fourth year from biochar application, thus confirming the suppression of mineralization, once the labile C was consumed [73]. In addition, the decrease was observed with digestate, whereas even an increase was observed with slurry. This difference suggests that the larger amount of fresh labile C, supplied with slurry rather than with digestate, might have favored microbial activity and organic C mineralization.
An overall decrease in N2O emissions with biochar is reported for several climates and soil types [32] and a suppressive effect of biochar on N2O emissions, induced by N fertilization, was typically reported [74]. However, this effect was variable, depending on the application period and the land-use [32]. A biochar-mediated reduction in N2O emissions, from soil amended with digestate, was observed by [29], and our study confirmed this decrease, although only after four years from the beginning of application. This temporal variability, coupled with the different applied doses of effective N, being added in 2020, is higher by 20% than that added in 2021. Therefore, biochar seemed more effective if below a certain threshold of N availability. After that, the excess of N not fitting the plant demand triggered N2O emissions [19,75,76]. The N2O emissions from soil fertilized with slurry seemed to confirm this pattern. In fact, in 2021, an almost two times higher amount of slurry was added to the soil, with respect to 2020, inducing N2O peaks, which are even higher with biochar. The effect of biochar was found to be negligible on N2O emissions in organic C rich soils [32], and excessive N fertilization was found to increase NO3 content in soil, since biochar has a limited capacity to entrap it after a certain threshold [77]. Therefore, we can hypothesize that, when labile organic matter and N were in excess, as in the case of digestate in 2020 and slurry in 2021, the impact of biochar on N2O emissions was null or even negative.
Similarly, CH4 fluxes responded differently to biochar depending on the quantity and quality of livestock residues. Decreased CH4 emissions from livestock residues, during composting [78,79] and after application to soil [80], were observed by several authors. However, other studies observed no effects of biochar and manure on CH4 fluxes from aerobic soils [81,82,83]. In our study, the only event of CH4 emissions was within the first week from residue application, in 2021, when an increase was observed with slurry + biochar. In that case, the slurry amount added to soil was almost twice that from the previous year, adding a large quantity of fresh and labile organic matter, which was promptly degraded in soils treated with biochar that are supposed to host a more active microbial community [84]. After that unique event, CH4 uptake increased significantly in soils treated with biochar, suggesting a major role of CH4 oxidation that is possibly driven by a better soil structure and aeration [35].

5. Conclusions

The hypothesis of higher N2O emissions from livestock residues compared to mineral fertilization was confirmed by our study after four annual applications. Digestate induced higher N2O emissions than slurry because of a larger N content, while it showed a higher CH4 uptake. The highest N2O peaks occurred between 1 and 3 weeks after fertilization, in concomitance with not limiting moisture and optimal NO3 availability.
We provided evidence that biochar was effective in reducing the environmental impact of organic fertilizers from the livestock chain, which even resulted in a synergy with the positive effect on C sequestration. However, the effect on CO2 and N2O emissions were dependent on (i) the quantity of organic matter added, (ii) its quality, and (iii) the time from application. An excess of labile organic matter and N induced higher CO2 emissions and peaks of N2O, independent of—or even triggered by—biochar. Indeed, the mitigation effect of biochar was evident below a certain threshold of organic fertilizer. Above that, no effect (digestate), or even an initial priming effect (slurry), was observed.
The results of the present study may contribute to policy recommendations at the EU level, which are at the base of sustainable agriculture improvement in the near-future, including climate change mitigation and water pollution reduction. We also highlighted the needs of identifying specific threshold for the quantity of livestock residues applied to soil to limit N losses.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/agronomy12071593/s1, Figure S1: Mean and maximum temperature (left axis), precipitation and irrigation (right axis) during 2020 and 2021 maize growing seasons; Table S1: Main soil characteristics of the experimental field at the beginning of the trial in 2018. Figure S2. Aerial image of the experimental trial at Cascina Baroncina (45°17’25’’ N–9°29’43’’ E). Source: Google Earth. Table S2. Biochar characterization according to Italian (ICH-AR, Italian Biochar Association), EU (EBC, European Biochar Certificate) and international (IBI, International Biochar Initiative) prescriptions.

Author Contributions

Conceptualization, A.L. and M.V.; methodology, M.V. and C.S.; field management, L.B.; data collection, M.V., C.S., L.B., A.L., C.B. and A.T.; resources, A.T.; writing—original draft preparation, A.L.; writing—review and editing, A.L., C.S., M.V. L.B., A.T. and C.B.; funding acquisition, A.T. All authors have read and agreed to the published version of the manuscript.

Funding

This research activity has been funded by Regione Lombardia with PSR 2014–2020, Operazione 1.2.01 of FEASR (Fondo europeo per l’agricoltura e lo sviluppo rurale), Projects INFOCHAR (2018–2019) and N-CONTROL (2020–2022, CUP E65F20001200007), and with POR FESR 2014–2020, Azione I.1.b.1.3, Project AGRI HUB (2020–2022).

Conflicts of Interest

The authors declare no conflict of interest.

References

  1. Syakila, A.; Kroeze, C. The global nitrous oxide budget revisited. Greenh. Gas Meas. Manag. 2011, 1, 17–26. [Google Scholar] [CrossRef]
  2. Saunois, M.; Stavert, A.R.; Poulter, B.; Bousquet, P.; Canadell, J.G.; Jackson, R.B.; Raymond, P.A.; Dlugokencky, E.J.; Houweling, S.; Patra, P.K.; et al. The global methane budget 2000–2017. Earth Syst. Sci. Data 2020, 12, 1561–1623. [Google Scholar] [CrossRef]
  3. Stehfest, E.; Bouwman, L. N2O and NO emission from agricultural fields and soils under natural vegetation: Summarizing available measurement data and modeling of global annual emissions. Nutr. Cycl. Agroecosystems 2006, 74, 207–228. [Google Scholar] [CrossRef]
  4. Reay, D.S.; Davidson, E.A.; Smith, K.A.; Smith, P.; Melillo, J.M.; Dentener, F.; Crutzen, P.J. Global agriculture and nitrous oxide emissions. Nat. Clim. Change 2012, 2, 410–416. [Google Scholar] [CrossRef]
  5. Davidson, E.A. The contribution of manure and fertilizer nitrogen to atmospheric nitrous oxide since 1860. Nat. Geosci. 2009, 2, 659–662. [Google Scholar] [CrossRef]
  6. IPCC. Climate Change 2013: The Physical Science Basis. In Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change; Stocker, T.F., Qin, D., Plattner, G.-K., Midgley, P.M., Eds.; Cambridge University Press: Cambridge, UK, 2013. [Google Scholar]
  7. Kammann, C.; Ratering, S.; Eckhard, C.; Müller, C. Biochar and hydrochar effects on greenhouse gas (carbon dioxide, nitrous oxide, and methane) fluxes from soils. J. Environ. Qual. 2012, 41, 1052–1066. [Google Scholar] [CrossRef]
  8. De Vries, F.T.; Van Groenigen, J.W.; Hoffland, E.; Bloem, J. Nitrogen losses from two grassland soils with different fungal biomass. Soil Biol. Biochem. 2011, 43, 997–1005. [Google Scholar] [CrossRef]
  9. Lugato, E.; Leip, A.; Jones, A. Mitigation potential of soil carbon management overestimated by neglecting N2O emissions. Nat. Clim. Change 2018, 8, 219–223. [Google Scholar] [CrossRef]
  10. Mosier, A.R. Exchange of gaseous nitrogen compounds between agricultural systems and the atmosphere. Plant Soil 2001, 228, 17–27. [Google Scholar] [CrossRef]
  11. Khalil, K.; Mary, B.; Renault, P. Nitrous oxide production by nitrification and denitrification in soil aggregates as affected by O2 concentration. Soil Biol. Biochem. 2004, 36, 687–699. [Google Scholar] [CrossRef]
  12. Liu, X.J.; Mosier, A.R.; Halvorson, A.D.; Zhang, F.S. Tillage and nitrogen application effects on nitrous and nitric oxide emissions from irrigated corn fields. Plant Soil 2005, 276, 235–249. [Google Scholar] [CrossRef]
  13. Carmo, J.B.; Filoso, S.; Zotelli, L.C.; De Sousa, E.R.; Pitombo, L.M.; Duarte-Neto, P.J.; Vargas, V.P.; Andrade, C.A.; Gava, G.J.; Rossetto, R.; et al. Infield greenhouse gas emissions from sugarcane soils in Brazil: Effects from synthetic and organic fertilizer application and crop trash accumulation. GCB Bioenergy 2012, 5, 267–280. [Google Scholar] [CrossRef]
  14. Ruser, R.; Flessa, H.; Russow, R.; Schmidt, G.; Buegger, F.; Munch, J.C. Emission of N2O, N2 and CO2 from soil fertilized with nitrate: Effect of compaction, soil moisture and rewetting. Soil Biol. Biochem. 2006, 38, 263–274. [Google Scholar] [CrossRef]
  15. Hellebrand, H.J.; Scholz, V.; Kern, J. Fertiliser induced nitrous oxide emissions during energy crop cultivation on loamy sand soils. Atmos. Environ. 2008, 42, 8403–8411. [Google Scholar] [CrossRef]
  16. Zou, J.; Huang, Y.; Zheng, X.; Wang, Y. Quantifying direct N2O emissions in paddy fields during rice growing season in mainland China: Dependence on water regime. Atmos. Environ. 2007, 41, 8030–8042. [Google Scholar] [CrossRef]
  17. McSwiney, C.P.; Robertson, G.P. Nonlinear response of N2O flux to incremental fertilizer addition in a continuous maize (Zea mays L.) cropping system. Glob. Change Biol. 2005, 11, 1712–1719. [Google Scholar] [CrossRef]
  18. Aguilera, E.; Lassaletta, L.; Sanz-Cobena, A.; Garnier, J.; Vallejo, A. The potential of organic fertilizers and water management to reduce N2O emissions in Mediterranean climate cropping systems. A review. Agric. Ecosyst. Environ. 2013, 164, 32–52. [Google Scholar] [CrossRef] [Green Version]
  19. Zhou, M.; Zhu, B.; Wang, S.; Zhu, X.; Vereecken, H.; Brüggemann, N. Stimulation of N2O emission by manure application to agricultural soils may largely offset carbon benefits: A global meta-analysis. Glob. Change Biol. 2017, 23, 4068–4083. [Google Scholar] [CrossRef] [PubMed]
  20. Chen, H.; Li, X.; Hu, F.; Shi, W. Soil nitrous oxide emissions following crop residue addition: A meta-analysis. Glob. Change Biol. 2013, 19, 2956–2964. [Google Scholar] [CrossRef]
  21. Wang, X.; Zou, C.; Gao, X.; Guan, X.; Zhang, W.; Zhang, Y.; Shi, X.; Chen, X. Nitrous oxide emissions in Chinese vegetable systems: A meta-analysis. Environ. Pollut. 2018, 239, 375–383. [Google Scholar] [CrossRef]
  22. Xia, L.; Lam, S.K.; Wolf, B.; Kiese, R.; Chen, D.; Butterbach-Bahl, K. Trade-offs between soil carbon sequestration and reactive nitrogen losses under straw return in global agroecosystems. Glob. Change Biol. 2018, 24, 5919–5932. [Google Scholar] [CrossRef] [PubMed]
  23. Velthof, G.L.; Kuikman, P.J.; Oenema, O. Nitrous oxide emission from animal manures applied to soil under controlled conditions. Biol. Fertil. Soils 2003, 37, 221–230. [Google Scholar] [CrossRef]
  24. Alburquerque, J.A.; de la Fuente, C.; Campoy, M.; Carrasco, L.; Nájera, I.; Baixauli, C.; Caravaca, F.; Roldán, A.; Cegarra, J.; Bernal, M.P. Agricultural use of digestate for horticultural crop production and improvement of soil properties. Eur. J. Agron. 2012, 43, 119–128. [Google Scholar] [CrossRef]
  25. Möller, K. Effects of anaerobic digestion on soil carbon and nitrogen turnover, N emissions, and soil biological activity. A review. Agron. Sustain. Dev. 2015, 35, 1021–1041. [Google Scholar] [CrossRef]
  26. Johansen, A.; Carter, M.S.; Jensen, E.S.; Hauggard-Nielsen, H.; Ambus, P. Effects of digestate from anaerobically digested cattle slurry and plant materials on soil microbial community and emission of CO2 and N2O. Appl. Soil Ecol. 2013, 63, 36–44. [Google Scholar] [CrossRef]
  27. Dietrich, M.; Fongen, M.; Foereid, B. Greenhouse gas emissions from digestate in soil. Int. J. Recycl. Org. Waste Agric. 2020, 9, 1–19. [Google Scholar]
  28. Fiedler, S.R.; Augustin, J.; Wrage-Mönnig, N.; Jurasinski, G.; Gusovius, B.; Glatzel, S. Potential short-term losses of N2O and N2 from high concentrations of biogas digestate in arable soils. Soil 2017, 3, 161–176. [Google Scholar] [CrossRef] [Green Version]
  29. Martin, S.L.; Clarke, M.L.; Othman, M.; Ramsden, S.J.; West, H.M. Biochar-mediated reductions in greenhouse gas emissions from soil amended with anaerobic digestates. Biomass Bioenergy 2015, 79, 39–49. [Google Scholar] [CrossRef]
  30. Woolf, D.; Amonette, J.E.; Street-Perrott, F.A.; Lehmann, J.; Joseph, S. Sustainable biochar to mitigate global climate change. Nat. Commun. 2010, 1, 56. [Google Scholar] [CrossRef] [Green Version]
  31. Spokas, K.A.; Cantrell, K.B.; Novak, J.M.; David, W.; Archer, D.W.; Ippolito, J.A.; Collins, H.P.; Boateng, A.A.; Lima, I.M.; Lamb, M.C.; et al. Biochar: A synthesis of its agronomic impact beyond carbon sequestration. J. Environ. Qual. 2012, 41, 973–989. [Google Scholar] [CrossRef] [PubMed]
  32. Borchard, N.; Schirrmann, M.; Cayuela, M.L.V.; Kammann, C.; Wrage-Mönnig, N.; Estavillo, J.M.; Fuertes-Mendizábal, T.; Siguah, G.; Spokas, K.; Ippolito, J.A.; et al. Biochar, soil and land-use interactions that reduce nitrate leaching and N2O emissions: A meta-analysis. Sci. Total Environ. 2019, 651, 2354–2364. [Google Scholar] [CrossRef] [PubMed]
  33. Cayuela, M.L.; Sánchez-Monedero, M.A.; Roig, A.; Hanley, K.; Enders, A.; Lehmann, J. Biochar and denitrification in soils: When, how much and why does biochar reduce N2O emissions? Sci. Rep. 2013, 3, 1732. [Google Scholar] [CrossRef] [Green Version]
  34. Case, S.D.; McNamara, N.P.; Reay, D.S.; Stott, A.W.; Grant, H.K.; Whitaker, J. Biochar suppresses N2O emissions while maintaining N availability in a sandy loam soil. Soil Biol. Biochem. 2015, 81, 178–185. [Google Scholar] [CrossRef] [Green Version]
  35. Case, S.D.; McNamara, N.P.; Reay, D.S.; Whitaker, J. The effect of biochar addition on N2O and CO2 emissions from a sandy loam soil–the role of soil aeration. Soil Biol. Biochem. 2012, 51, 125–134. [Google Scholar] [CrossRef]
  36. Van Zwieten, L.; Singh, B.; Joseph, S.; Kimber, S.; Cowie, A.; Chan, K.Y. Biochar and emissions of non-CO2 greenhouse gases from soil. In Biochar for Environmental Management: Science and Technology; Lehmann, J., Joseph, S., Eds.; Routledge: London, UK, 2009; pp. 259–282. [Google Scholar]
  37. Clough, T.J.; Condron, L.M.; Kammann, C.; Müller, C. A review of biochar and soil nitrogen dynamics. Agronomy 2013, 3, 275–293. [Google Scholar] [CrossRef] [Green Version]
  38. Taghizadeh-Toosi, A.; Clough, T.J.; Sherlock, R.R.; Condron, L.M. Biochar adsorbed ammonia is bioavailable. Plant Soil 2012, 350, 57–69. [Google Scholar] [CrossRef]
  39. Singh, B.P.; Hatton, B.J.; Singh, B.; Cowie, A.L.; Kathuria, A. Influence of biochars on nitrous oxide emission and nitrogen leaching from two contrasting soils. J. Environ. Qual. 2010, 39, 1224–1235. [Google Scholar] [CrossRef]
  40. Šimek, M.; Jíšová, L.; Hopkins, D.W. What is the so-called optimum pH for denitrification in soil? Soil Biol. Biochem. 2002, 34, 1227–1234. [Google Scholar] [CrossRef]
  41. Kappler, A.; Wuestner, M.L.; Ruecker, A.; Harter, J.; Halama, M.; Behrens, S. Biochar as an electron shuttle between bacteria and Fe(III) minerals. Environ. Sci. Technol. Lett. 2014, 1, 339–344. [Google Scholar] [CrossRef]
  42. Brevik, E.C. Soils and climate change: Gas fluxes and soil processes. Soil Horiz. 2012, 53, 12–23. [Google Scholar] [CrossRef] [Green Version]
  43. Guenet, B.; Gabrielle, B.; Chenu, C.; Arrouays, D.; Balesdent, J.; Bernoux, M.; Bruni, E.; Caliman, J.P.; Cardinael, R.; Chen, S.; et al. Can N2O emissions offset the benefits from soil organic carbon storage? Glob. Change Biol. 2021, 27, 237–256. [Google Scholar] [CrossRef]
  44. Adviento-Borbe, M.A.; Pittelkow, C.M.; Anders, M.; van Kessel, C.; Hill, J.E.; McClung, A.M.; Linquist, B.A. Optimal fertilizer nitrogen rates and yield-scaled global warming potential in drill seeded rice. J. Environ. Qual. 2013, 42, 1623–1634. [Google Scholar] [CrossRef] [Green Version]
  45. Davidson, E.A.; Belk, E.; Boone, R.D. Soil water content and temperature as independent or confounded factors controlling soil respiration in temperate mixed hardwood forest. Glob. Change Biol. 1998, 4, 217–227. [Google Scholar] [CrossRef] [Green Version]
  46. Adviento-Borbe, M.A.; Kaye, J.P.; Bruns, M.A.; McDaniel, M.D.; McCoy, M.; Harkcom, S. Soil greenhouse gas and ammonia emissions in long-term maize-based cropping systems. Soil Sci. Soc. Am. J. 2010, 74, 1623–1634. [Google Scholar] [CrossRef]
  47. Arthurson, V. Closing the global energy and nutrient cycles through application of biogas residues to agricultural land-potential benefits and drawbacks. Energies 2009, 2, 226–242. [Google Scholar] [CrossRef] [Green Version]
  48. Riva, C.; Orzi, V.; Carozzi, M.; Acutis, M.; Boccasile, G.; Lonati, S.; Tambone, F.; D’Imporzano, G.; Adani, F. Short-term experiments in using digestate products as substitutes for mineral (N) fertilizer: Agronomic performance, odours, and ammonia emission impacts. Sci. Total Environ. 2016, 547, 206–214. [Google Scholar] [CrossRef] [PubMed]
  49. Petrova, I.P.; Pekrun, C.; Möller, K. Organic matter composition of digestates has a stronger influence on N2O emissions than the supply of ammoniacal Nitrogen. Agronomy 2021, 11, 2215. [Google Scholar] [CrossRef]
  50. Chu, H.; Fujii, T.; Morimoto, S.; Lin, X.; Yagi, K.; Hu, J.; Zhang, J. Community structure of ammonia-oxidizing bacteria under long-term application of mineral fertilizer and organic manure in a sandy loam. Appl. Environ. Microbiol. 2007, 73, 458–491. [Google Scholar] [CrossRef] [PubMed] [Green Version]
  51. Yoshida, H.; Nielsen, M.P.; Scheutz, C.; Jensen, L.S.; Christensen, T.H.; Nielsen, S.; Bruun, S. Effects of sewage sludge stabilization on fertilizer value and greenhouse gas emissions after soil application. Acta Agric. Scand. Sect. B Soil Plant Sci. 2015, 65, 506–516. [Google Scholar]
  52. Rodhe, L.K.K.; Ascue, J.; Willén, A.; Persson, B.V.; Nordberg, Å. Greenhouse gas emissions from storage and field application of anaerobically digested and non-digested cattle slurry. Agric. Ecosyst. Environ. 2015, 199, 358–368. [Google Scholar] [CrossRef]
  53. Köster, J.R.; Cardenas, L.M.; Bol, R.; Lewicka-Szczebak, D.; Senbayram, M.; Well, R.; Giesemann, A.; Dittert, K. Anaerobic digestates lower N2O emissions compared to cattle slurry by affecting rate and product stoichiometry of denitrification—An N2O isotopomer case study. Soil Biol. Biochem. 2015, 84, 65–74. [Google Scholar] [CrossRef]
  54. Amon, B.; Kryvoruchko, V.; Amon, T.; Zechmeister-Boltenstern, S. Methane, nitrous oxide and ammonia emissions during storage and after application of dairy cattle slurry and influence of slurry treatment. Agric. Ecosyst. Environ. 2006, 112, 153–162. [Google Scholar] [CrossRef]
  55. Clemens, J.; Trimborn, M.; Weiland, P.; Amon, B. Mitigation of greenhouse gas emissions by anaerobic digestion of cattle slurry. Agric. Ecosyst. Environ. 2006, 112, 171–177. [Google Scholar] [CrossRef]
  56. Pilegaard, K.; Skiba, U.; Ambus, P.; Beier, C.; Brüggemann, N.; Butterbach-Bahl, K.; Dick, J.; Dorsey, J.; Duyzer, J.; Gallagher, M.; et al. Factors controlling regional differences in forest soil emission of nitrogen oxides (NO and N2O). Biogeosciences 2006, 3, 651–661. [Google Scholar] [CrossRef] [Green Version]
  57. Baral, K.R.; Labouriau, R.; Olesen, J.E.; Petersen, S.O. Nitrous oxide emissions and nitrogen use efficiency of manure and digestates applied to spring barley. Agric. Ecosyst. Environ. 2017, 239, 188–198. [Google Scholar] [CrossRef]
  58. Senbayram, M.; Chen, R.; Budai, A.; Bakken, L.; Dittert, K. N2O emission and the N2O/(N2O + N2) product ratio of denitrification as controlled by available carbon substrates and nitrate concentrations. Agric. Ecosyst. Environ. 2012, 147, 4–12. [Google Scholar] [CrossRef]
  59. Czubaszek, R.; Wysocka-Czubaszek, A. Emissions of carbon dioxide and methane from fields fertilized with digestate from an agricultural biogas plant. Int. Agrophysics 2018, 32, 29. [Google Scholar] [CrossRef] [Green Version]
  60. Eickenscheidt, T.; Freibauer, A.; Heinichen, J.; Augustin, J.; Drösler, M. Short-term effects of biogas digestate and cattle slurry application on greenhouse gas emissions affected by N availability from grasslands on drained fen peatlands and associated organic soils. Biogeosciences 2014, 11, 6187–6207. [Google Scholar] [CrossRef] [Green Version]
  61. Dutaur, L.; Verchot, L.V. A global inventory of the CH4 sink. Global Biogeochem. Cycles 2007, 21, GB4013. [Google Scholar] [CrossRef]
  62. Kirschke, S.; Bousquet, P.; Ciais, P.; Saunois, M.; Canadell, J.G.; Dlugokencky, E.J.; Bergamaschi, P.; Bergmann, D.; Blake, D.R.; Bruhwiler, L.; et al. Three decades of global methane sources and sinks. Nat. Geosci. 2013, 6, 813. [Google Scholar] [CrossRef]
  63. Biernat, L.; Taube, F.; Loges, R.; Kluss, C.; Reinsch, T. Nitrous oxide emissions and methane uptake from organic and conventionally managed arable crop rotations on farms in Northwest Germany. Sustainability 2020, 12, 3240. [Google Scholar] [CrossRef] [Green Version]
  64. Dalal, R.C.; Allen, D.E.; Livesley, S.J.; Richards, G. Magnitude and biophysical regulators of methane emission and consumption in the Australian agricultural, forest, and submerged landscapes: A review. Plant Soil 2008, 309, 43–76. [Google Scholar] [CrossRef]
  65. Hütsch, B.W.; Webster, C.P.; Powlson, D.S. Long-term effects of nitrogen fertilization on methane oxidation in soil of the broadbalk wheat experiment. Soil Biol. Biochem. 1993, 25, 1307–1315. [Google Scholar] [CrossRef]
  66. Ullah, S.; Frasier, R.; King, L.; Picotte-Anderson, N.; Moore, T.R. Potential fluxes of N2O and CH4 from soils of three forest types in Eastern Canada. Soil Biol. Biochem. 2008, 40, 986–994. [Google Scholar] [CrossRef]
  67. Möller, K. Effects of biogas digestion on soil organic matter and nitrogen inputs, flows and budgets in organic cropping systems. Nutr. Cycl. Agroecosyst. 2009, 84, 179–202. [Google Scholar] [CrossRef]
  68. Barłóg, P.; Hlisnikovský, L.; Kunzová, E. Effect of digestate on soil organic carbon and plant-available nutrient content compared to cattle slurry and mineral fertilization. Agronomy 2020, 10, 379. [Google Scholar] [CrossRef] [Green Version]
  69. Fontaine, S.; Mariotti, A.; Abbadie, L. The priming effect of organic matter: A question of microbial competition? Soil Biol. Biochem. 2003, 35, 837–843. [Google Scholar] [CrossRef]
  70. Abubaker, J.; Risberg, K.; Pell, M. Biogas residues as fertilisers—Effects on wheat growth and soil microbial activities. Appl. Energy 2012, 99, 126–134. [Google Scholar] [CrossRef]
  71. Lorenz, K.; Lal, R. Biochar application to soil for climate change mitigation by soil organic carbon sequestration. J. Plant Nutr. Soil Sci. 2014, 177, 651–670. [Google Scholar] [CrossRef] [Green Version]
  72. He, Y.; Zhou, X.; Jiang, L.; Li, M.; Du, Z.; Zhou, G.; Shao, J.; Wang, X.; Xu, Z.; Bai, S.H.; et al. Effects of biochar application on soil greenhouse gas fluxes: A meta-analysis. GCB Bioenergy 2017, 9, 743–755. [Google Scholar] [CrossRef]
  73. Mukherjee, A.; Lal, R. Biochar impacts on soil physical properties and greenhouse gas emissions. Agronomy 2013, 3, 313–339. [Google Scholar] [CrossRef] [Green Version]
  74. Barnard, R.; Leadley, P.W.; Hungate, B.A. Global change, nitrification, and denitrification: A review. Glob. Biogeochem. Cycles 2005, 19, GB1007. [Google Scholar] [CrossRef]
  75. Cooper, R.J.; Wexler, S.K.; Adams, C.A.; Hiscock, K.M. Hydrogeological controls on regional-scale indirect nitrous oxide emission factors for rivers. Environ. Sci. Technol. 2017, 51, 10440–10448. [Google Scholar] [CrossRef]
  76. Tian, L.; Zhu, B.; Akiyama, H. Seasonal variations in indirect N2O emissions from an agricultural headwater ditch. Biol. Fertil. Soils 2017, 53, 651–662. [Google Scholar] [CrossRef]
  77. Hagemann, N.; Joseph, S.; Schmidt, H.-P.; Kammann, C.I.; Harter, J.; Borch, T.; Young, R.B.; Varga, K.; Taherymoosavi, S.; Elliott, K.W.; et al. Methane emissions and associated microbial activities from paddy salt-affected soil as influenced by biochar and cow manure addition. Appl. Soil Ecol. 2020, 152, 103531. [Google Scholar]
  78. Chowdhury, M.A.; de Neergaard, A.; Jensen, L.S. Composting of solids separated from anaerobically digested animal manure: Effect of different bulking agents and mixing ratios on emissions of greenhouse gases and ammonia. Biosyst. Eng. 2014, 124, 63–77. [Google Scholar] [CrossRef]
  79. Vu, Q.D.; de Neergaard, A.; Tran, T.D.; Hoang, H.T.T.; Vu, V.T.K.; Jensen, L.S. Greenhouse gas emissions from passive composting of manure and digestate with crop residues and biochar on small-scale livestock farms in Vietnam. Environ. Technol. 2015, 36, 2924–2935. [Google Scholar] [CrossRef]
  80. Nguyen, S.H.; Nguyen, H.D.T.; Hegarty, R.S. Defaunation and its impacts on ruminal fermentation, enteric methane production and animal productivity. Livest. Res. Rural. Dev. 2020, 32, 4. [Google Scholar]
  81. Kammann, C.; Ippolito, J.; Hagemann, N.; Borchard, N.; Cayuela, M.L.; Estavillo, J.M.; Fuertes-Mendizabal, T.; Jeffery, S.; Kern, J.; Novak, J.; et al. Biochar as a tool to reduce the agricultural greenhouse-gas burden–knowns, unknowns and future research needs. J. Environ. Eng. Landsc. Manag. 2017, 25, 114–139. [Google Scholar] [CrossRef]
  82. Wang, J.; Pan, X.; Liu, Y.; Zhang, X.; Xiong, Z. Effects of biochar amendment in two soils on greenhouse gas emissions and crop production. Plant Soil 2012, 360, 287–298. [Google Scholar] [CrossRef]
  83. Abagandura, G.O.; Chintala, R.; Sandhu, S.S.; Kumar, S.; Schumacher, T.E. Effects of biochar and manure applications on soil carbon dioxide, methane, and nitrous oxide fluxes from two different soils. J. Environ. Qual. 2019, 6, 1664–1674. [Google Scholar] [CrossRef]
  84. Palansooriya, K.N.; Wong, J.T.F.; Hashimoto, Y.; Huang, L.; Rinklebe, J.; Chang, S.X.; Bolan, N.; Wang, H.; Ok, Y.S. Response of microbial communities to biochar-amended soils: A critical review. Biochar 2019, 1, 3–22. [Google Scholar] [CrossRef] [Green Version]
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Figure 1. NH4 (top) and NO3 (bottom) content of soil at 0–30 cm depth during the first six monitoring events of 2020 and during the first 6 weeks of 2021 maize growing seasons in the 6 treatments (digestate, slurry, and mineral fertilization with and without biochar). Different letters indicate significant differences between treatments for each sampling event.
Figure 1. NH4 (top) and NO3 (bottom) content of soil at 0–30 cm depth during the first six monitoring events of 2020 and during the first 6 weeks of 2021 maize growing seasons in the 6 treatments (digestate, slurry, and mineral fertilization with and without biochar). Different letters indicate significant differences between treatments for each sampling event.
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Figure 2. C-CO2 emissions from soil during 2020 and 2021 maize growing seasons for the 6 treatments (digestate, slurry, and mineral fertilization with and without biochar). Significative biochar effects on emissions are reported for each date from digestate (d), slurry (s), and mineral (m) fertilization.
Figure 2. C-CO2 emissions from soil during 2020 and 2021 maize growing seasons for the 6 treatments (digestate, slurry, and mineral fertilization with and without biochar). Significative biochar effects on emissions are reported for each date from digestate (d), slurry (s), and mineral (m) fertilization.
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Figure 3. Cumulative fluxes of CO2 (top), CH4 (middle), and N2O (bottom) from soil in 2020 and 2021 maize growing seasons for the six treatments (digestate, slurry and mineral fertilization with and without biochar). Error bars represent standard errors for each year.
Figure 3. Cumulative fluxes of CO2 (top), CH4 (middle), and N2O (bottom) from soil in 2020 and 2021 maize growing seasons for the six treatments (digestate, slurry and mineral fertilization with and without biochar). Error bars represent standard errors for each year.
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Figure 4. C-CH4 fluxes from soil during the 2020 and 2021 maize growing seasons for the six treatments (digestate, slurry, and mineral fertilization, with and without biochar). Significative biochar effects on emissions are reported for each date from digestate (d), slurry (s), and mineral (m) fertilization.
Figure 4. C-CH4 fluxes from soil during the 2020 and 2021 maize growing seasons for the six treatments (digestate, slurry, and mineral fertilization, with and without biochar). Significative biochar effects on emissions are reported for each date from digestate (d), slurry (s), and mineral (m) fertilization.
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Figure 5. N-N2O fluxes from soil during 2020 and 2021 maize growing seasons for the six treatments (digestate, slurry, and mineral fertilization, with and without biochar). Significative biochar effects on emissions are reported for each date from digestate (d), slurry (s), and mineral (m) fertilization.
Figure 5. N-N2O fluxes from soil during 2020 and 2021 maize growing seasons for the six treatments (digestate, slurry, and mineral fertilization, with and without biochar). Significative biochar effects on emissions are reported for each date from digestate (d), slurry (s), and mineral (m) fertilization.
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Figure 6. Linear regression between N2O cumulative emissions and effective N, added by digestate and slurry in the 2020 and 2021 maize growing seasons (top), and exponential regression between N2O cumulative emissions and NO3 cumulated in the soil in the 2020 maize growing season (bottom).
Figure 6. Linear regression between N2O cumulative emissions and effective N, added by digestate and slurry in the 2020 and 2021 maize growing seasons (top), and exponential regression between N2O cumulative emissions and NO3 cumulated in the soil in the 2020 maize growing season (bottom).
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Table
Table 1. Crop management, GHGs monitoring, and soil sampling events during the maize growing seasons 2020 and 2021.
Table 1. Crop management, GHGs monitoring, and soil sampling events during the maize growing seasons 2020 and 2021.
20202021
Fertilization (33% urea; 100% organic) and ploughingJune 23June 7
Harrowing and sowingJune 24June 8
Pre-emergency weedingJune 25June 9
Static chambers installationJune 29June 10
Gas monitoring17 events17 events
N mineral sampling6 events6 events
Irrigation3 events6 events
Post emergency weedingJuly 17July 6
In-season (top-dress) urea fertilization (67%)July 15July 7
End of gas samplingAugust 31August 5
Table
Table 2. Characteristics, added doses, and N inputs of digestate (solid + liquid fractions) and slurry for the 2020 and 2021 growing seasons. N added is the sum of N-NH4 + N org. Effective N is the sum of N-NH4 + 0.5 N org.
Table 2. Characteristics, added doses, and N inputs of digestate (solid + liquid fractions) and slurry for the 2020 and 2021 growing seasons. N added is the sum of N-NH4 + N org. Effective N is the sum of N-NH4 + 0.5 N org.
DigestateSlurry
2020202120202021
Moisture content (%)92.193.194.694.7
Norg content (g kg−1)2.431.614.321.64
N-NH4 content (g kg−1)2.571.480.980.46
Effective N (%)0.380.230.310.13
Added dose (Mg ha−1)547445133
N added (Kg ha−1)270229239279
Effective N added (Kg ha−1)205170140173
Table
Table 3. Crop yield, soil TOC content, and cumulated NH4 and NO3 at the end of 2020 and 2021 maize growing seasons. Standard deviation is reported in brackets. Different letters (a, b, c, etc.) indicate significant differences between treatments and years.
Table 3. Crop yield, soil TOC content, and cumulated NH4 and NO3 at the end of 2020 and 2021 maize growing seasons. Standard deviation is reported in brackets. Different letters (a, b, c, etc.) indicate significant differences between treatments and years.
YieldTOCN-NH4N-NO3
Gg ha−1Gg ha−1Kg ha−1Kg ha−1
2020Digestate16.4 (1) a,b,c44 (0.4) d,e,f6.7 (2.3) b5.7 (1.0) a
Digestate + biochar16.9 (1) a,b,c49 (2) c,g7.1 (0.1) b50.1 (47) a
Slurry14.9 (0.2) a,b,c39 (0.8) f7.3 (0.7) b10.2 (6.3) a
Slurry + biochar14.9 (1) a,b,c49 (0.6) c,d,g7.7 (1.1) b3.7 (2.2) a
Mineral13.6 (1) c39 (1) f6.2 (1.8) b11.6 (5.7) a
Mineral + biochar14.2 (0.8) b,c56 (3) a,b13.0 (2.0) a14.4 (4.2) a
2021Digestate16.6 (0.5) a,b,c46 (3) d,e,g8.6 (0.5) b2.9 (1.1) a
Digestate + biochar18.6 (3) a,b50 (2) c,g10.2 (0.2) a,b29.5 (3.6) a
Slurry16.9 (0.2) a,b,c46 (0.8) d,e,g
Slurry + biochar15.6 (3) a,b,c59 (3) a
Mineral19.2 (1) a42 (1) e,f10.3 (0.7) a,b27.6 (19.4) a
Mineral + biochar14.0 (1) c54 (1) b,c11.2 (1.2) a,b21.5 (12.6) a
Table
Table 4. Percentage effects of organic fertilizers (digestate or slurry vs. Mineral fertilization) and biochar (in combination with digestate, slurry, and mineral fertilization) on CO2, CH4, and N2O fluxes during 2020 and 2021 maize growing seasons. DAF: Days After Fertilization. Significances of the effects are reported at * p < 0.05 (Univariate ANOVA).
Table 4. Percentage effects of organic fertilizers (digestate or slurry vs. Mineral fertilization) and biochar (in combination with digestate, slurry, and mineral fertilization) on CO2, CH4, and N2O fluxes during 2020 and 2021 maize growing seasons. DAF: Days After Fertilization. Significances of the effects are reported at * p < 0.05 (Univariate ANOVA).
2020 2021
DAF8101315202224293437434549596265694791115212428313538424649525659
DAY1/73/76/78/713/715/717/722/727/730/75/87/811/821/824/827/831/811/614/616/618/622/628/61/75/78/712/715/719/723/726/729/72/85/8
CO2
Dig. vs Min.241011141823211215 *27 *92231 *22343126 *55 *70 *62 *71 *51 *34 *41 *48 *21 *34 *3435 *31 *32 *31 *2138 *
Slu. vs Min.423527 *43 *37 *2751 *419 *2224 *182611162839 *29 *30 *20 *23 *15569−61193130 *27 *42 *2029 *
Bioch-Dig1966115752−414−3−1−17−21−12−6−19 *−24 *−27 *−26 *−22 *−20 *−26 *−27 *−17 *−22 *−13−14−18 *−15 *−11−33 *−34 *
Bioch-Slu12032−126 *−4−8 *630−1−4−1−1−12−20 *17 *15 *15 *13 *15 *1017 *911033 *−10−11−13 *−3−22 *−23 *
Bioch-Min−14−13−7−11−7−50−7−13 *−10−3−10−6−13−21−1123 *−5−3−9−6−10−12−9−5−16−1121−14−6−6−4−28−10
CH4
Dig. vs Min.7503125634 *−15−659 *2167 *273428235 *115 *39−920−2029192329−12−1124406235−230−176231−1348
Slu. vs Min.35−1937−3−26−2−28 *−13310311268 *−2197 *3025−13020677 *−1230−6−1146 *572 *25146−10−331979
Bioch-Dig444171309434573 *468 *11815−90 *−84 *−22−13−36−1−135−66 *−70 *−75 *−75 *−71 *−66 *−21−52 *−52 *−65−74−57−39−58−16−29
Bioch-Slu5229887−2118−19−23−13−151−38−42−22−1931−12−4841164 *151 *177 *150−1169594211−59 *−35−1277−164−63
Bioch-Min−29−38−53−29−53−44−28−53−44−33−76−73−51−84 *−65 *21−51−20−7−14−1577−50−1610−12−22−41−91 *−64−70−53−63−39
N2O
Dig. vs Min.2701 *2004954322123442015−224511939912−68 *215−16546 *914 *796 *597 *259533−241037 *28−70 *−38−54 *−33−54 *−52
Slu. vs Min.3261481867164116−43−71−84−84−56−79 *−56 *−11−2727138039120870−36−12−39−53 *−40 *170 *−67 *−45−63 *−44−59 *−46
Bioch-Dig444171309434573 *468 *11815−90 *−84 *−22−13−36−1−135−66 *−70 *−75 *−75 *−71 *−66 *−21−52 *−52 *−65−74−57−39−58−16−29
Bioch-Slu5229887−2118−19−23−13−151−38−42−22−1931−12−4841164 *151 *177 *150−1169594211−59 *−35−1277−164−63
Bioch-Min−29−38−53−29−53−44−28−53−44−33−76−73−51−84 *−65 *21−51−20−7−14−1577−50−1610−12−22−41−91 *−64−70−53−63−39
Table
Table 5. Percentage effects of organic fertilizers (digestate or slurry vs. Mineral fertilization) and biochar (in combination with digestate, slurry, and mineral fertilization) on cumulative CO2, CH4, and N2O fluxes in 2020 and 2021 maize growing seasons. Significances of the effects are reported at p < 0.1 #, p < 0.05 *, p < 0.01 **, and p < 0.01 *** (Univariate Analysis of Variance).
Table 5. Percentage effects of organic fertilizers (digestate or slurry vs. Mineral fertilization) and biochar (in combination with digestate, slurry, and mineral fertilization) on cumulative CO2, CH4, and N2O fluxes in 2020 and 2021 maize growing seasons. Significances of the effects are reported at p < 0.1 #, p < 0.05 *, p < 0.01 **, and p < 0.01 *** (Univariate Analysis of Variance).
20202021
Organic vs. Mineral Fertilizer (%)Biochar Effect (%)Organic vs. Mineral Fertilizer (%)Biochar Effect (%)
CO2CH4N2OCO2CH4N2OCO2CH4N2OCO2CH4N2O
Digestate17 #+31 *+202 #0+8+24+42 ***−8+63 #−21 ***+58 *−59 *
Slurry25 *+11−180+24 **−6+18 *+16 #+11+6 **−15+49
Mineral−−−−8 *+7−57 **-−−−8 #+10−50 #
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Lagomarsino, A.; Valagussa, M.; Scotti, C.; Borrelli, L.; Becagli, C.; Tosca, A. Mitigation of GHG Emissions from Soils Fertilized with Livestock Chain Residues. Agronomy 2022, 12, 1593. https://doi.org/10.3390/agronomy12071593

AMA Style

Lagomarsino A, Valagussa M, Scotti C, Borrelli L, Becagli C, Tosca A. Mitigation of GHG Emissions from Soils Fertilized with Livestock Chain Residues. Agronomy. 2022; 12(7):1593. https://doi.org/10.3390/agronomy12071593

Chicago/Turabian Style

Lagomarsino, Alessandra, Massimo Valagussa, Carla Scotti, Lamberto Borrelli, Claudia Becagli, and Alberto Tosca. 2022. "Mitigation of GHG Emissions from Soils Fertilized with Livestock Chain Residues" Agronomy 12, no. 7: 1593. https://doi.org/10.3390/agronomy12071593

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