Open Access
How to translate text using browser tools
15 October 2021 Nitrous oxide production and nitrogen transformations in a soil amended with biosolids
Carmen C. Roman-Perez, Guillermo Hernandez-Ramirez
Author Affiliations +

The application of organic amendments to agricultural soils enables the recycling of nutrients, further reducing the inputs of synthetic fertilizers for crop production. However, the production of N2O emissions is a concern that arises from such a practice. A 35 d incubation experiment was conducted with soils receiving three contrasting types of biosolids — mesophilic anaerobic digested (BM), composted (BC), and alkaline-stabilized (BA) — at four water-filled pore spaces (WFPS): 28%, 40%, 52%, and 64%. A zero-N-addition control was also evaluated. Across all the three types of biosolids, N2O production increased with soil moisture content, with BM and BC producing the overall highest N2O fluxes. The most intense pulses of N2O production were exhibited by BC at the beginning of the incubation. The highest cumulative N2O production was found with 64% WFPS and from BC- (409 μg N2O–N·kg−1 soil) or BM-amended soils (390 μgN2O–N·kg−1 soil), which produced more than four and two times the emissions from the control and BA-amended soils at 64% WFPS, respectively. We also found the highest nitrification rates in the BM- and BC-amended soils. The total N2O production was exponentially associated with the NO3–N concentration present at the end of the experiment (R2 = 0.83). Changes in the concentration of the soil available N indicated the occurrence of mineralization, nitrification, and denitrification over the incubation. These results provided insight into the interacting responses of N2O production to soil moisture contents, biosolids treatment stabilization and properties, and soil N availability.


Agricultural soils are considered to be one of the major contributors to anthropogenic N2O emissions, as they are the source of about 60% of the global N2O emissions (De Rosa et al. 2016; Chai et al. 2020; Roman-Perez and Hernandez-Ramirez 2021). This figure is mainly caused by the increased application of synthetic N fertilizers to meet the increasing demand for food and biofuels (Kim et al. 2013; Smith 2017; Thilakarathna et al. 2020). Concerns around N2O emissions arise from its high global warming potential, which is 298 times that of CO2 (in a 100 yr timeframe), and its role as a stratospheric-ozone depleting substance (Wang et al. 2013; Charles et al. 2017).

Common pathways of N2O production in soils are during the process of nitrification (oxidation of NH4+ → NO2 → NO3) via hydroxylamine (NH2OH) oxidation (NH3 → NH2OH → NO → N2O) and denitrification (reduction of NO3 → N2O → N2) under limiting conditions of oxygen (Hernandez-Ramirez et al. 2009; T. Zhu et al. 2013; Shcherbak et al. 2014). Soil water content is considered one of the main controllers of soil N2O production (Jäger et al. 2011; T. Zhu et al. 2013), as it regulates oxygen availability for microbes and affects gas diffusivity (Schaufler et al. 2010; Butterbach-Bahl et al. 2013; Cardoso et al. 2017).

Other factors influencing N2O production are nitrogen input, organic carbon availability, pH, and temperature (Snyder et al. 2009; Oertel et al. 2016). Previous studies have shown an exponential increase of N2O emissions with higher N application rates (Kim et al. 2013; Shcherbak et al. 2014). Additionally, availability of organic C increases microbial respiration, which reduces the oxygen concentration within the soil air, promoting the creation of anaerobic microsites, which further lead to N2O production (Velthof et al. 2003; Jäger et al. 2011). Available organic C also acts as electron donor during denitrification, favoring the release of N2O emissions (Guenet et al. 2021). Moreover, the C:N ratio of organic amendments is an indicator of the mineralization or immobilization rates of N in the soil. Organic amendments with narrower C:N ratios will mineralize the organic N faster, leading to a faster release of NH4+ and NO3, which further serve as substrates for N2O emissions (Jones et al. 2007; T. Zhu et al. 2013; De Rosa et al. 2016). Denitrification is also affected by the soil pH; values between 6 and 8 are considered optimum for the occurrence of denitrification (Coyne 2008). However, N2O fluxes are higher at lower pH because the nitrous oxide reductase is inhibited while the reduction of NO3, NO2, and NO is favored at pH below 7 (Hu et al. 2015). The optimum temperature range for denitrification is from 28 to 37 °C (Coyne 2008), with increasing N2O emissions with temperature up to ∼37 °C (Oertel et al. 2016).

Biosolids, or treated sewage sludge, are solids, semisolids, or liquid residues that are by-products of municipal wastewater treatment plants (Wijesekara et al. 2016; Torri et al. 2017). Globally, a large amount of biosolids (1 × 108 Mg) is generated every year (Thangarajan et al. 2013). With an increasing global population, by 2050, this amount is expected to increase by 75% (17.5 × 107 Mg·yr−1) for an estimated population of 9.6 billion at a production rate of 50 g·person−1·day−1 on a dry basis (Wijesekara et al. 2016). Sustainable biosolids management is a major challenge because of the large quantities generated and their subsequent disposal (Haynes et al. 2009; Braguglia et al. 2015).

Land application of biosolids is considered to be the most economical and advantageous management method because the nutrients can be recycled and utilized by crops; concurrently, the dependence on synthetic fertilizers can be reduced (Christie et al. 2001; Rigby et al. 2009; Brown et al. 2011; Lu et al. 2012). However, there is also risk for increased field N2O emissions when biosolids are land-applied relative to applying synthetic N fertilizers (Pu et al. 2010; Wijesekara et al. 2016; Roman-Perez et al. 2021). Although, when accounting for the elevated energy costs of industrial production of synthetic fertilizers, the lower N2O emissions from fields receiving synthetic N fertilizers are offset (Roman-Perez et al. 2021), leading to a reduced agricultural carbon footprint when replacing synthetic fertilizers with organic amendments (Sharma et al. 2017). Thus, understanding the factors that govern N2O emissions from the different available organic nitrogen sources is important to design management strategies to abate N2O emissions (Pu et al. 2010; X. Zhu et al. 2013).

Most existing studies have focused on soil N2O emissions from synthetic fertilizer use (Linzmeier et al. 2001; Barrena et al. 2017; Liu et al. 2017; Guardia et al. 2018), whereas studies assessing organic amendments have focused mostly on applications of manure (Velthof et al. 2003; Yang et al. 2003; Dalal et al. 2009) or compost (X. Zhu et al. 2013; Zhu-Barker et al. 2015), and only a few studies have assessed N2O from biosolids applications (Inubushi et al. 2000; Pu et al. 2010; Yoshida et al. 2015, Roman-Perez et al. 2021). Therefore, there is still a knowledge gap regarding the effect of biosolids on N2O fluxes, particularly those comparing divergent types of biosolids and stabilization methods (e.g., anaerobic digestion, alkaline stabilization, or composting) result in products with contrasting properties. These biosolid properties, as well as soil characteristics, need to be taken into account to minimize agricultural N2O emissions while maintaining or improving crop yields. To our knowledge, controlled studies comparing N2O emissions from a wide range of contrasting biosolids and under multiple soil water contents are not yet available in the literature. The objective of the present study was to quantify the effects of biosolids’ properties, multiple soil moisture contents, and their potential interactive effect on soil N dynamics and N2O production. We hypothesized that biosolids with an increased availability of organic C and N, in combination with increasing soil moisture, would lead to faster N dynamics (mineralization and nitrification) and higher N2O emissions.


Soil collection

Soil samples were taken from the 0–15-cm topsoil layer at the Ellerslie Research Station (53°25'13"N, 113°33'03"W), in Edmonton, Alberta, Canada. The soil is an Orthic Black Chernozem according to the Canadian soil classification system (the U.S. soil taxonomy equivalent is Typic Cryoboroll). The typical crop rotation in the sampled field included annual crops of wheat (Triticum aestivum L.), canola (Brassica napus L.), and barley (Hordeum vulgare L.) (Roman-Perez and Hernandez-Ramirez 2021). The soil texture is silty clay loam with 327 g·kg−1 clay, 511 g·kg−1 silt, and 162 g·kg−1 sand. The soil properties are as follows: pH 6.5, 54.8 g·kg−1 organic carbon, and 4.6 g·kg−1 total N. Soil samples were collected prior to the beginning of the growing season, in mid-May 2018. The samples were stored for 4 wks at field moisture conditions at 4 °C until the beginning of the experiment.

Biosolids collection, experimental design, and treatment preparation

Three types of biosolids were evaluated — mesophilic anaerobic digested biosolid (BM), composted biosolid (BC), and alkaline-stabilized biosolid (BA) — as they are representative of common biosolids treatments and their characteristics contrast widely. The BM and BC were provided by the Edmonton Waste Management Centre; BA was sourced from the Banff Wastewater Treatment Facility, both located within Alberta, Canada. The biosolids were obtained at each corresponding facility and transported to the Edmonton Research Station of University of Alberta (Roman-Perez et al. 2021). Composite subsamples of each type of fresh biosolids were taken from the field piles for conducting the incubation experiment. As preliminary steps, we determined the water content (as gravimetric weight loss), total carbon (TC), and total nitrogen (TN) concentrations (dry combustion with a ThermoScientific, Flash 2000 Organic Elemental Analyzer) for each of the three biosolids. The TN concentration and gravimetric water content were used to calculate the N application rates of fresh biosolids.

Prior to treatment preparation, the soil was mixed and sieved to 8 mm. The soil was then air-dried to a water content of ∼28% water-filled pore space (WFPS). The experiment was a two-factor factorial design, with biosolid type (control 0 N, BM, BA, and BC) and soil moisture content (28%, 40%, 52%, and 64% WFPS) as the factors. The selected WFPS levels are common soil moisture contents over the growing season in Central Alberta. Five replicates of the 16 treatment combinations were prepared to measure N2O production (three replicates) and soil NH4+–N and NO3–N concentrations (two replicates) for destructive sampling on day 7 of the incubation.

The soil microcosms were prepared by placing 0.83 kg of soil (dry mass basis) into plastic containers 10 cm in height and with an 11.5 cm inner diameter. The soil microcosms were preincubated for 3 d at room temperature of 20–22 °C to create favorable conditions for microbial activity (Lin and Hernandez-Ramirez 2020, Thilakarathna and Hernandez-Ramirez 2021). After the preincubation period, the corresponding soils were mixed with each type of biosolid at an N rate of 122 mg TN·kg−1 soil (equivalent to a field application rate of 192 kg N·ha−1), whereas the controls did not receive any biosolids. Each soil microcosm was packed to a bulk density of 1.03 g·cm−3 and watered to the treatment WFPS. All microcosms were checked daily and maintained at their respective WFPS by adjusting the weight loss to the total weight with Milli-Q ultrapure water. Over the 35 d incubation period, the microcosms remained at room temperature (22 °C on average). To allow gas exchange, the lids of the soil microcosms were perforated (six small holes per lid).

Nitrous oxide emissions

To measure N2O fluxes, we used a non-steady-state chamber system (12 autochambers Eosense eosAC) connected to a thermoelectrically cooled, mid-infrared quantum cascade laser absorption spectroscope system (QCLAS, Aerodyne Research Inc., Billerica, USA). The coupled systems allowed the recirculation of the gas samples between the chamber headspace (2.4 L) at ambient pressure and the QCLAS analytical cell at a pressure of 30 Torr at a flow rate of 1.61 standard L·min−1. Chambers were closed for 5 min, during which the N2O concentrations were obtained at 1 Hz resolution and recorded with TDLWintel software. The TDLWintel software also controlled the QCLAS system. After the 5 min enclosure time, the chamber system was allowed to return to ambient concentrations prior to the next measurement. The QCLAS system was calibrated daily with reference gas standards as well as working standards of synthetic N2O at 0.5 and 1.9 μmol·mol−1, and ultrahigh purity dinitrogen (N2) for background absorption spectra subtraction. Ambient temperature and pressure were recorded with a temperature data logger (HOBO UX100, Onset, Bourne, USA) and a barometric pressure meter (Testo 511, West Chester, USA). The first N2O measurement was conducted 3 h after the treatments had been applied, then on incubation days 1, 2, 3, 4, 5, 6, 7, 9, 11, 13, 15, 18, 21, 24, 27, 31, and 35.

Flux calculation

Nitrous oxide production rates (μg N2O–N·kg−1 soil·day−1) were calculated by applying a modified ideal gas law as follows:


where slope is the linear regression coefficient during the 5 min when the chamber was enclosed (nL·L−1·s−1), 2M is the mass of two atoms of N in a mole of N2O (28.01 g N·mol−1), V is the volume of the chamber headspace (L), P is the pressure of the chamber headspace (atm), R is the universal gas constant (atm·nL·K−1·nmol−1), T is the temperature in the chamber headspace (K), and soil mass is on a dry weight basis (kg).

Cumulative fluxes of N2O were calculated via linear interpolation of consecutive flux measurements over the entire incubation period (Thilakarathna and Hernandez-Ramirez 2021).

Inorganic nitrogen concentration

The patterns of NH4+–N and NO3–N concentration in the soil throughout the incubation period were assessed by analyzing soil samples from prior to the treatment application, through destructive sampling on day 7 (two replicates); and on day 35, in all the incubated soils (three replicates). Soil available N (NO3–N and NH4+–N) was extracted from 5 g soil samples with 50 mL of a 2M KCl solution, shaken in a reciprocal shaker for 30 min, and filtered with Whatman 42 filter paper (Fisher Scientific, Pittsburg, USA) (Chai et al. 2020; Roman-Perez and Hernandez-Ramirez 2021). The colorimetric method was applied to the filtrates with a Thermo Gallery Plus Beermaster Autoanalyzer. The colorimetric method used vanadium chloride to reduce the NO3–N to nitrite (NO2–N); thus, NO3–N concentrations include both NO3 and the preexisting NO2–N in the samples. The NH4+–N was measured using the salicylate-hypochlorite method.

Nitrification rates were calculated as the linear regression coefficients of changes in nitrate concentrations over the incubation period (Lin and Hernandez-Ramirez 2021). Net mineralization was inferred from the net changes in available N concentration (i.e., combined NO3–N and NH4+–N) from the beginning to the end of the incubation period (i.e., net mineralization = final available N − initial soil available N).

We determine an exponential relationship between cumulative N2O emissions over the entire incubation versus NO3–N on day 35 across all the treatment combinations using the function N2O = aeb(NO3–N) with the Euler’s number as a base while a and b were fitting parameters.

Statistical analyses

Statistical analyses were performed with R Studio software (R Core Team 2019). Regression analyses were performed to test the relationships between nitrification rates and WFPS. Biosolid type, WFPS, and their interaction (biosolid type × WFPS) were included in an ANOVA model for the N2O fluxes and available N concentrations. Following significant ANOVAs (alpha critical value of 0.05), Tukey’s honest significant difference test was performed for pairwise comparisons of the treatments. Data were Box–Cox transformed when needed to meet the assumptions of normality and homoscedasticity. The standard errors of the means are presented as error terms.


Biosolids’ properties

The properties of the three biosolids strongly differed between stabilization methods. The total N content of BM was five and two times the N contents of BA and BC, respectively (Table 1). The amount of TN in the form of NH4+–N was 19.4%, 16.1%, and 6.2% for BM, BA, and BC, respectively (Table 1). Across the biosolid types, BA had the highest pH (>12) caused by the addition of alkaline materials during its production. Similar to the results of N contents, BM also showed the highest TC concentration (283.3 g C·kg−1), which was slightly higher than that of BC (278.4 g C·kg−1) and more than twice the amount in BA (122.5 g C·kg−1) (Table 1).

Table 1.

Characterization of the three assessed biosolids ± standard error of the mean.


Nitrous oxide fluxes

Daily N2O production rates increased with soil moisture and biosolids additions (Fig. 1a1d). Overall, peak N2O fluxes occurred shortly after the addition of biosolids (ranging from the first 3 h to day 3 of the incubation), with BC as the biosolid that presented the highest fluxes at each WFPS (Fig. 1a1d). BC-amended soils peaked rapidly 3 h after the beginning of the incubation, but quickly dropping back to the basal fluxes by day 4. For BM- and BA-amended soils, relatively high N2O production took place mainly within 9 d from the beginning of the incubation, particularly at 52% and 64% WFPS (Fig. 1c, d). After this initial high activity, daily N2O fluxes remained relatively low and constant until the end of the incubation.

Fig. 1.

N2O–N fluxes (μg N·kg−1 soil·day−1) at (a) 28%, (b) 40%, (c) 52%, and (d) 64% water-filled pore space (WFPS) for all biosolid-amended soils and the untreated control over a 35 d incubation period. BM, mesophilic anaerobic-digested biosolid; BA, alkaline-stabilized biosolid; BC, composted biosolid. Note the different y-scales across some of the panels. [Colour online.]


Cumulative N2O emissions ranged from 14.19 to 409.01 μg N2O–N·kg−1 soil for the control at 28% WFPS and the BC-amended soil at 64% WFPS, respectively. Relative to the controls, biosolids additions increased N2O production by 4.7, 3.8, and 2.1 times on average across all moisture contents for BM-, BC-, and BA-amended soils, respectively. Statistical analysis of the cumulative N2O showed a significant interaction between biosolid addition and WFPS (P < 0.05) (Fig. 2). As part of this WFPS × biosolid addition interaction, significant effects of biosolid additions were found at 64% WFPS, where BM- and BC-amended soils were fourfold higher than the control (P < 0.05) (Fig. 2). Moreover, in the BM- and BC-amended soils, N2O emissions at 64% WFPS were significantly higher than the emissions at 28% WFPS by 12 and 22 times, respectively (P < 0.05) (Fig. 2). At 28%, 40%, and 52% WFPS, BM-amended soils had the highest cumulative N2O fluxes; conversely, at 64% WFPS, the BC-amended soil’s emissions were numerically higher than those from BM by 19 μg N2O–N·kg−1 soil, but these two treatment combinations were not significantly different from each other (P > 0.05) (Fig. 2).

Fig. 2.

Cumulative N2O–N fluxes (μg N·kg−1 soil) at 28%, 40%, 52%, and 64% water-filled pore space (WFPS) for all biosolid-amended soils and the untreated control over a 35 d incubation period. BM, mesophilic anaerobic-digested biosolid; BA, alkaline-stabilized biosolid; BC, composted biosolid. [Colour online.]


Ammonium and nitrate concentrations

The initial NH4+–N and NO3–N concentrations in the soil were 4.39 and 13.61 mg N·kg−1, respectively (Figs. 3 and 4). Over the incubation period, NO3–N increased (Fig. 4), whereas NH4+–N exhibited increasing and decreasing fluctuations (Fig. 3). In general, NO3–N concentrations increased with increasing soil moisture content (Fig. 4); the opposite trend was observed for NH4+–N concentrations (Fig. 3).

Fig. 3.

Changes in NH4+–N concentration (mg N·kg−1) during the incubation period at (a) 28%, (b) 40%, (c) 52%, and (d) 64% water-filled pore space (WFPS) for all biosolid-amended soils and the untreated control. Vertical scales are different. BM, mesophilic anaerobic digested biosolid; BA, alkaline-stabilized biosolid; BC, composted biosolid. [Colour online.]


Fig. 4.

Changes in NO3–N concentration (mg N·kg−1) during the incubation period at (a) 28%, (b) 40%, (c) 52%, and (d) 64% water-filled pore space (WFPS) for all biosolid-amended soils and the untreated control. BM, mesophilic anaerobic-digested biosolid; BA, alkaline-stabilized biosolid; BC, composted biosolid. [Colour online.]


Both biosolid addition and moisture levels had separate significant effects on NH4+–N as well as on NO3–N (P < 0.05) at days 7 and 35. The interaction between soil water and biosolids addition was significant only for NO3–N on day 35 of the incubation (P < 0.05). When we compared the available N among soils receiving biosolids, BM-amended soils showed the highest NH4+–N and NO3–N accumulation on both days 7 and 35 of the incubation (Figs. 3 and 4). In the case of NH4+–N, on day 7, the control soils produced significantly less than the BM-amended soils at all WFPS (P < 0.05), whereas on day 35, the control soils produced significantly less than the BM-amended soils at 40%, 52%, and 64% WFPS (P < 0.05). By the end of the incubation, NO3–N in the control soils was significantly lower than from the biosolids-amended soils (P < 0.05), except for BC at 28% WFPS. In addition, at 40%, 52%, and 64% WFPS, BM-amended soils had significantly higher NO3–N concentrations than BA- and BC-amended soils (P < 0.05).

Nitrification rates

Nitrification rates varied from 0.11 to 1.21 mg NO3–N·kg−1 soil·day−1 for the control soil at 28% WFPS and the BM-amended soil at 52% WFPS, respectively. We observed faster nitrification rates with increasing soil moisture up to 52% WFPS in the BM- and BC-amended soils (nonlinear relationships, R2-BM = 0.996, R2-BC = 0.999; Fig. 5), up to 55% WFPS in the BA-amended soils (non-linear relationship, R2-BA = 0.990; Fig. 5), and 64% in the control soils (linear relationship, R2 = 0.998; Fig. 5).

Fig. 5.

Nitrification rates for all biosolid additions and the untreated control as a function of moisture content expressed as water-filled pore space (WFPS). Exponential fitting and equations are provided for each biosolid treatment combination. Linear fitting and the equation are presented for the control treatment combinations. BM, mesophilic anaerobic-digested biosolid; BA, alkaline-stabilized biosolid; BC, composted biosolid. [Colour online.]


Additionally, net mineralization over the incubation period increased with soil moisture for all biosolid types and the control, reaching more than double the initial available N concentration (i.e., as indicated by combined NO3–N and NH4+–N): an increase from 18 (4 mg N·kg−1 from NH4+–N plus 14 mg N·kg−1 from NO3–N, day 0) to 37.6 mg N·kg−1 (1 plus 36.6, day 35) at 64% WFPS. Of all the biosolids-amended soils, those receiving BM showed the highest mineralization rates at each WFPS.

The concentration of available N in the BM-amended soils increased by threefold at 28% WFPS and fourfold at 40%, 52%, and 64% WFPS by the end of the incubation (on day 35) (Figs. 3 and 4). Moreover, we observed an increase in cumulative N2O emissions with NO3–N on day 35 (Fig. 6) across all the treatment combinations (exponential fit, R2 = 0.825; Spearman’s correlation ρ = 0.938, P < 0.001).

Fig. 6.

Cumulative N2O–N production (μg N·kg−1 soil) as a function of NO3–N concentration (mg N·kg−1) at the end of the incubation period across all treatment combinations (nitrogen sources × water-filled pore space). Exponential fitting and the equation are presented. BM, mesophilic anaerobic-digested biosolid; BA, alkaline-stabilized biosolid; BC, composted biosolid. [Colour online.]



The interacting effect between the biosolids type and moisture content on N2O emissions shows that N2O production depends on both the presence of available N as a substrate and favorable soil moisture conditions to promote the availability of the essential soluble C and N nutrients for microbial activity (Banerjee et al. 2016). In our study, a wide range of conditions for N2O production and emission were covered by using three contrasting biosolids and four WFPS levels. In addition to contributing directly with some inorganic N, biosolids are rich sources of organic N that undergoes through mineralization to further release mainly NH4+, which is available for plant uptake (Rigby et al. 2016; Sharma et al. 2017). Therefore, N substrate was not likely to be a limiting factor in the biosolids-amended soils in our experiment, while the increasing soil water contents represented a range of moist, aerobic conditions (i.e., 28–64% WFPS) (Roman-Perez and Hernandez-Ramirez 2021). This range of soil moisture might have favored an increased mineralization (Curtin et al. 2012).

As expected, most of the results showed higher daily and cumulative N2O emissions with increasing soil moisture content from 28% to 64% WFPS (Figs. 1 and 2). Similar results have been found in several studies applying either synthetic fertilizer (Bateman and Baggs 2005; Mathieu et al. 2006) or organic amendments (T. Zhu et al. 2013; X. Zhu et al. 2013). The response of N2O emissions to increasing soil moisture can be explained by increases in the accessibility and mobility of substrates, which can be associated with solute diffusion across the soil pore network (Curtin et al. 2012; Lin and Hernandez-Ramirez 2020; Roman-Perez and Hernandez-Ramirez 2021). The range of soil moisture contents in our experiment was favorable for nitrification, as has been reported previously by Bateman and Baggs (2005) and Linn and Doran (1984). In their studies, they reported <60% WFPS as optimum soil moisture for nitrification, since the diffusion of both nutrients and O2 is not limited at this WFPS. The occurrence of fast nitrification in our study could have increased the availability of NH2OH and NO2–N substrates for the production of N2O emissions (Roman-Perez and Hernandez-Ramirez 2021). Mineralization of the organic N added in the biosolids might have further promoted nitrification in the aerobic domains of the soil while denitrification might have occurred in the anaerobic microsites located within the aggregates in the soil microcosms (García-Ruiz et al. 2012; T. Zhu et al. 2013; Guardia et al. 2018).

The differences in N2O emissions across the three biosolids-amended soils could be related to the amount of nitrogen and carbon in the different biosolids (Table 1) and their availabilities. BM had a higher total N concentration than either BC and BA (Table 1); and the organic N in BM can be mineralized during the digestion process. Then BM has more available N in comparison with BA and BC, (higher proportion of NH4+–N concentration in BM, Table 1) (Rigby et al., 2009, 2016; Yoshida et al. 2015). In the case of BA, the addition of alkaline materials and the resulting N losses (NH3 volatilization) during the lime stabilization treatment led to lower N contents, whereas the addition of woody bulking agents during composting led to more recalcitrant N in BC (Yoshida et al. 2015; Rigby et al. 2016; Roman-Perez et al. 2021). Addition of biosolids also provides available organic C, which increases microbial respiration and depletes the O2; likely favoring N2O production in the anaerobic microsites (Gentile et al. 2008; Zhu-Barker et al. 2015; Guenet et al. 2021). In our study, the C contents in BM (283.3 g C·kg−1) and BC (278.4 g C·kg−1) were similar, and more than double the C content of BA (122.5 g C·kg−1) (Table 1). Therefore, even though the soil moisture contents represented aerobic conditions, some anaerobic hotspots might have been created within the soil microcosms shortly after the biosolids addition. These relationships of N and C supplies across biosolids also explain the higher N2O emissions activity, in particular, at the beginning of the incubation. In a field study, Roman-Perez et al. (2021) found how the addition of biosolids produced under mesophilic anaerobic digestion (BM) led to higher N2O emissions than BC and BA since BM had greater N content and moisture and narrower C:N ratio. These results are in agreement with our present incubation results; moreover, across biosolid-amended soils, Roman-Perez et al. (2021) typically found lower N2O emissions from fields receiving BC, and these BC effects was congruently mirrored by their higher crop biomass and N use efficiency in comparison to BM and BA additions.

Additionally, we observed an exponential increase in cumulative N2O emissions in association with the NO3–N concentrations present on day 35 (R2 = 0.825, Fig. 6). This nonlinear linkage reveals that soils with increased capacity for N2O production also accumulate the most NO3, in particular under overall aerobic moisture conditions, which probably did not favor further reduction of the produced NO3 to N2O. Most of the total N2O emissions at 52% and 64% WFPS were produced within the first 4, 9, or 13 d of the incubation for the BC-, BA-, and BM-amended soils, respectively, whereas NO3–N accumulated over the entire incubation period (Fig. 4). Moreover, as the NH4+–N concentrations showed an overall decrease over time, this could indicate a reduction in the mineralization rate, an increase in the nitrification rate, or both effects occurring simultaneously toward the end of the incubation period. Our results for daily N2O emissions also help to explain the dynamics of the available N. Since most of the cumulative N2O emissions were produced within the first 4 to 13 d of the incubation as noted above, with a great increase in NO3–N by day 7, it could be assumed that the NH4+–N produced was rapidly nitrified to NO3–N, as reported by Inubushi et al. (2000) and He et al. (2017) after applying organic amendments. In the study by Inubushi et al. (2000), soils receiving sewage sludge compost (at rates of 10 and 20 Mg DM·ha−1·yr−1) showed a N2O peak at the same time as the NH4+–N concentrations (day 30 of a 90-d incubation experiment). Moreover, He et al. (2017) reported a decrease in NH4+–N concentrations concurrently with an increase in NO3–N in soils receiving grass clover biomass or cattle manure at a rate of 10 g DM·kg−1 over a 20-d incubation experiment. Therefore, in our experiment, most N2O was probably emitted during the nitrification process because the aerobic conditions might have resulted in more N2O coming from nitrification than from denitrification (Inubushi et al. 2000; Bateman and Baggs 2005; Liu et al. 2017; Mekala and Nambi 2017). In addition, part of the newly produced NO3–N might have been reduced to N2O in the anaerobic microsites via denitrification. This is because NO3–N could have been required as an alternative electron acceptor once decomposition of the organic C provided by the biosolids caused O2 depletion (Wrage et al. 2001; Jäger et al. 2011). After this period of high microbial activity, we observed a decline in N2O emissions (Fig. 1) as well as in NH4+–N concentrations (Fig. 3), although NO3–N concentrations remained high (Fig. 4). This could be explained by the fact that in the beginning of the incubation, the rapidly mineralizable C was probably depleted (Gentile et al. 2008), leading to a subsequent decrease in the mineralization rate, thus lowering the amount of NH4+–N produced, whereas most of the NO3–N already produced remained in the soil. The reduced NH4+–N supply from the soil and biosolids might have caused lower nitrification rates, which explains the lower N2O emissions after the organic matter from the biosolids became depleted, as the moisture conditions were not favorable for NO3–N reduction to N2O (Bateman and Baggs 2005; Mathieu et al. 2006; Zhu-Barker et al. 2015). The changes in microbial processes once C substrate is depleted might have induced N immobilization in the soil, as indicated by an overall decrease in NH4+–N concentrations (Fig. 3). Nevertheless, N2O can also be produced fron other processes such as chemodenitrification and nitrifier denitrification (NH3 → NO2 followed by NO → N2O → N2); however, the contribution of those processes to N2O emissions during our experiment might have been minimal since favorable conditions were not provided. For example, acidic pH (<5) and high NO2 concentrations are needed for chemodenitrification, which is most common in forest soils; whereas nitrifier denitrification occurs under concurrent conditions of O2 limitations, low organic C, and high N availability (Butterbach-Bahl et al. 2013; Hu et al. 2015).

Nitrification reached optimal rates at a soil moisture content of 52% WFPS in the BM- and BC-amended soils and at 55% WFPS in the BA-amended soils (Fig. 5). Nitrification rates declined when the soil moisture was beyond these WFPS levels. As noted above, previous studies (Linn and Doran 1984; Bateman and Baggs 2005) have reported 60% WFPS to be the optimum conditions for nitrification; however, those previous studies used synthetic N fertilizers. Our results suggested a slightly different pattern for optimum conditions, with a shift toward a lower WFPS with organic amendments. The declining nitrification rates in the biosolids-amended soils while the WFPS was still lower than 60% might be related to the fact that biosolids provided readily degradable organic C, which was likely used by microbes, leading to faster consumption and depletion of soil O2 at a lower WFPS than previously thought. This increased use of O2 by microbial respiration can preclude or reduce the aerobic conditions that would have favored the nitrification process (Gentile et al. 2008; Zhu-Barker et al. 2015; Guenet et al. 2021). The feedback effect of adding biosolids with narrow C:N ratios would have presumably shifted the optima of nitrification rates toward less moist soil conditions. It is noted that BA-amended soils showed less variation in their nitrification rates across the soil moisture levels, probably because of the wider C:N ratio and the more recalcitrant C than in BM and BC (Table 1, Figs. 3, 4, and 5). These postulations are in line with the results of Rigby et al. (2009) from a 90-d field study in which lime-treated and mesophilic anaerobic-digested biosolids were applied. Their results showed that 20% of the TN applied with the mesophilic anaerobic-digested biosolid was recovered as NO3– N + NO2–N, whereas only 10% of the TN was recovered in the lime-treated biosolids plots after 20 d of their experiment.

An additional explanation for the high N2O fluxes observed could be the occurrence of a positive N2O priming effect, which refers to the stimulation of soil organic matter (SOM)-derived N2O production as triggered by the addition of labile N (Roman-Perez and Hernandez-Ramirez 2021). The underlying notion is that the labile N addition alters the decomposition and mineralization of preexisting SOM by microbes, commonly increasing (positive priming) the soil-available N derived specifically from preexisting SOM (Fiorentino et al. 2019). This positive priming effect of preexisting SOM can be also stimulated by new additions of labile organic C (Fiorentino et al. 2019). Since the addition of biosolids in our experiment provided both C and N, it is probable that positive priming of SOM occurred, which further led to a positive N2O priming effect. The occurrence of a positive N2O priming effect on a Black Chernozem soil was recently reported by Roman-Perez and Hernandez-Ramirez (2021). In their study, they found a positive N2O priming effect with increasing soil moisture after adding labile N (urea), which resulted in a 19% primed N2O flux, with even more N2O coming from SOM than from the added urea-N (59% vs. 41%, respectively). Additional research is needed to evaluate the effects of biosolids amendments on the decomposition and mineralization of preexisting SOM and the further release of primed N2O. A study of available N dynamics (mineralization–immobilization) in soils receiving biosolids is also needed to estimate the potential value of biosolids as an N source for crop production. Furthermore, analysis of microbial biomass N could also be meaningful, as the microbial population and dynamics play a role in biosolids N turnover (Rigby et al. 2009).

Certain study limitations can be noted. This incubation was specifically done under controlled laboratory conditions, including only one soil, packed to a single bulk density, without plants growing, and under general aerobic conditions. Furthermore, the effect of biosolids’ physical properties on the soil physical properties such as bulk density, porosity, and water holding capacity was not considered. Several studies have shown that the addition of biosolids decreases the bulk density while increasing the porosity and water-holding capacity (Gardner et al. 2010; Sharma et al. 2017). However, these studies were conducted at field conditions and changes were measured after 1 yr of the application of biosolids. Therefore, we believe no major changes in the referred soil properties occurred due to the addition of biosolids in the microcosms. Nevertheless, we acknowledge that the pore size distribution of the soil microcosms might have been affected after biosolids additions. Therefore, there is a need to further evaluate the effect of applying biosolids on N2O emissions under a wider range of soil moisture or wetting–drying cycles while taking into consideration the differences on the pore size distribution, as this controls the gas exchange between the soil and atmosphere (Balaine et al. 2013) as well as the pore continuity (van der Weerden et al. 2012). The study of van der Weerden et al. (2012) found that N2O emissions are reduced with increased pore continuity when macropores (>60 μm diameter) become drained. Thus, soils with a greater proportion of macropores would release less N2O. In our study, the higher moisture of added BM might have reduced the volume of aerated macropores to a greater proportion than both BA and BC, which in turn might have favored the occurrence of N2O emissions over a longer period in the BM-amended soils when compared with BC at each water-filled pore space (Fig. 1). This hypothesis points out to the important influence of soil physical properties on N2O emissions, which should be considered for future studies.


This study shows how the addition of contrasting biosolid types, a range of soil moisture contents, and their interaction significantly impacted N2O production. As expected, N2O emissions were enhanced by increasing soil moisture and N additions from different types of biosolids, with BM and BC amendments producing the highest N2O emissions, pointing to the influence of the forms and availability of N and C as well as the C:N ratio on the microbial processes that generate N2O emissions (i.e., nitrification and denitrification). As expected, soil N dynamics were enhanced by increasing soil moisture and organic-N additions, with BM and BC amendments producing the highest cumulative N2O emissions, pointing to the influence of the forms and availability of N and C as well as the C:N ratio on nitrification and denitrification. Thus, these biosolid properties should be taken into account when designing practices for land application of biosolids in agricultural systems. For example, the addition of a biosolid with higher N, C, and NH4+ contents and a lower C:N ratio has the potential to increase N2O emissions; therefore, caution should be taken when applying biosolids with such characteristics. Nevertheless, our results substantiate the need for additional research to assess the temporal variation of available N as well as the net N mineralization induced by biosolid amendments to identify the biosolid rates to be applied in fields. Land application of biosolids should aim to prevent N losses to the environment and optimize nutrient recovery by plants.

Competing Interests

The authors declare there are no competing interests.


The present study was funded by the Agriculture and Agrifood Canada Program — Agricultural Greenhouse Gases Program (AGGP2 - 033), the Natural Sciences and Engineering Research Council of Canada (NSERC Discovery, 2018-05717), and the Canadian Foundation for Innovation (CFI, John Evans Leadership Fund 32860). We also acknowledge the Edmonton Waste Management Centre and Walker Industries for providing the biosolids. We further acknowledge the support from Sisi Lin, Jichen Li, Emanuel Geier, Sanat Kanekar, and Melissa Wheatley.



Balaine, N., Clough, T.J., Beare, M.H., Thomas, S.M., Meenken, E.D., and Ross, J.G. 2013. Changes in relative gas diffusivity explain soil nitrous oxide flux dynamics. Soil Sci. Soc. Am. J. 77(5): 1496–1505. Scholar


Banerjee, S., Helgason, B., Wang, L., Winsley, T., Ferrari, B.C., and Siciliano, S.D. 2016. Legacy effects of soil moisture on microbial community structure and N2O emissions. Soil Biol. Biochem. 95: 40–50. Scholar


Barrena, I., Menéndez, S., Correa-Galeote, D., Vega-Mas, I., Bedmar, E.J., González-Murua, C., and Estavillo, J.M. 2017. Soil water content modulates the effect of the nitrification inhibitor 3,4-dimethylpyrazole phosphate (DMPP) on nitrifying and denitrifying bacteria. Geoderma. 303(May): 1–8. Scholar


Bateman, E.J., and Baggs, E.M. 2005. Contributions of nitrification and denitrification to N2O emissions from soils at different water-filled pore space. Biol. Fertil. Soils 41(6): 379–388. Scholar


Braguglia, C.M., Coors, A., Gallipoli, A., Gianico, A., Guillon, E., Kunkel, U., et al. 2015. Quality assessment of digested sludges produced by advanced stabilization processes. Environ. Sci. Pollut. Res. 22(10): 7216–7235. Scholar


Brown, S., Kurtz, K., Bary, A., and Cogger, C. 2011. Quantifying benefits associated with land application of organic residuals in Washington State. Environ. Sci. Technol. 45(17): 7451–7458. Scholar


Butterbach-Bahl, K., Baggs, E.M., Dannenmann, M., Kiese, R., and Zechmeister-Boltenstern, S. 2013. Nitrous oxide emissions from soils: How well do we understand the processes and their controls? Philos Trans. R Soc. Lond. B Biol. Sci. 368(1621). Scholar


Cardoso, A.d.S., Quintana, B.G., Janusckiewicz, E.R., Brito, F., Morgado, E.da S., et al. Reis, al. 2017. N2O emissions from urine-treated tropical soil: Effects of soil moisture and compaction, urine composition, and dung addition. Catena. 157(May): 325–332. Scholar


Chai, L.L., Hernandez-Ramirez, G., Dyck, M., Pauly, D., Kryzanowski, L., Middleton, A., et al. 2020. Can fertigation reduce nitrous oxide emissions from wheat and canola fields? Sci. Total Environ. 745: 141014. Scholar


Charles, A., Rochette, P., Whalen, J.K., Angers, D.A., Chantigny, M.H., and Bertrand, N. 2017. Global nitrous oxide emission factors from agricultural soils after addition of organic amendments: A meta-analysis. Agric. Ecosyst. Environ. 236(3): 88–98. Scholar


Christie, P., Easson, D.L., Picton, J.R., and Love, S.C.P. 2001. Agronomic value of alkaline-stabilized sewage biosolids for spring barley. Agron. J. 93(1): 144–151. Scholar


Coyne, M.S. 2008. Biological denitrification. Nitrog. Agric. Syst. (02): 201–253. Scholar


Curtin, D., Beare, M.H., and Hernandez-Ramirez, G. 2012. Temperature and moisture effects on microbial biomass and soil organic matter mineralization. Soil Sci. Soc. Am. J. 76(6): 2055–2067. Scholar


Dalal, R.C., Gibson, I.R., and Menzies, N.W. 2009. Nitrous oxide emission from feedlot manure and green waste compost applied to vertisols. Biol. Fertil. Soils 45(8): 809–819. Scholar


De Rosa, D., Rowlings, D.W., Biala, J., Scheer, C., Basso, B., McGree, J., and Grace, P.R. 2016. Effect of organic and mineral N fertilizers on N2O emissions from an intensive vegetable rotation. Biol. Fertil. Soils 52(6): 895–908. Scholar


Fiorentino, N., Sánchez-Monedero, M.A., Lehmann, J., Enders, A., Fagnano, M., and Cayuela, M.L. 2019. Interactive priming of soil N transformations from combining biochar and urea inputs: A 15N isotope tracer study. Soil Biology and Biochemistry 131(June 2018): 166–175. Scholar


García-Ruiz, R., Gómez-Muñoz, B., Hatch, D.J., Bol, R., and Baggs, E.M. 2012. Soil mineral N retention and N2O emissions following combined application of 15N-labelled fertiliser and weed residues. Rapid Commun. Mass Spectrom. 26(20): 2379–2385. Scholar


Gardner, W.C., Broersma, K., Naeth, A., Chanasyk, D., and Jobson, A. 2010. Influence of biosolids and fertilizer amendments on physical, chemical and microbiological properties of copper mine tailings. Can. J. Soil Sci. 90(4): 571–583. Google Scholar


Gentile, R., Vanlauwe, B., Chivenge, P., and Sixa, J. 2008. Interactive effects from combining fertilizer and organic residue inputs on nitrogen transformations. Soil Biol. Biochem. 40(9): 2375–2384. Scholar


Guardia, G., Vallejoa, A., Cardenas, L.M., Dixon, E.R., and García-Marcoa, S. 2018. Fate of 15N-labelled ammonium nitrate with or without the new nitrification inhibitor DMPSA in an irrigated maize crop. Soil Biol. Biochem. 116(May 2017): 193–202. Scholar


Guenet, B., Chenu, C., Arrouays, D., Balesdent, J., Bernoux, M., Bruni, E., et al. 2021. Can N2O emissions offset the benefits from soil organic carbon storage? Global Change Biol. 27(2): 237–256. Scholar


Haynes, R.J., Murtaza, G., and Naidu, R. 2009. Chapter 4 inorganic and organic constituents and contaminants of biosolids. Implications for land application. 1st edn, Advances in Agronomy. Elsevier Inc. Scholar


He, M., Ma, W., Zelenev, V.V., Khodzaeva, A.K., Kuznetsov, A.M., Semenov, A.M., et al. 2017. Short-term dynamics of greenhouse gas emissions and cultivable bacterial populations in response to induced and natural disturbances in organically and conventionally managed soils. Appl. Soil Ecol. 119(August): 294–306. Scholar


Hernandez-Ramirez, G., Brouder, S.M., Smith, D.R., and Van Scoyoc, G.E. 2009. Greenhouse gas fluxes in an eastern corn belt soil: weather, nitrogen source, and rotation. J. Environ. Qual. 38(3): 841–854. Scholar


Hu, H.W., Chen, D., and He, J.Z. 2015. Microbial regulation of terrestrial nitrous oxide formation: Understanding the biological pathways for prediction of emission rates. FEMS Microbiol. Rev. 39(5): 729–749. PMID:25934121. Google Scholar


Inubushi, K., Goyal, S., Sakamoto, K., Wada, Y., Yamakawa, K., and Arai, T. 2000. Influences of application of sewage sludge compost on N2O production in soils. Chemosp. Glob. Chan. Sci. 2(3–4): 329–334. Scholar


Jäger, N., Stange, C.F., Ludwig, B., and Flessa, H. 2011. Emission rates of N2O and CO2 from soils with different organic matter content from three long-term fertilization experiments-a laboratory study. Biol. Fertil. Soils 47(5): 483–494. Scholar


Jones, S.K., Reesa, R.M., Skibab, U.M., and Ball, B.C. 2007. Influence of organic and mineral N fertiliser on N2O fluxes from a temperate grassland. Agric. Ecosyst. Environ. 121(1–2): 74–83. Scholar


Kim, D.G., Hernandez-Ramirez, G., and Giltrap, D. 2013. Linear and nonlinear dependency of direct nitrous oxide emissions on fertilizer nitrogen input: A meta-analysis. Agric. Ecosyst. Environ. 168: 53–65. Scholar


Lin, S., and Hernandez-Ramirez, G. 2020. Nitrous oxide emissions from manured soils as a function of various nitrification inhibitor rates and soil moisture contents. Science of the Total Environment. 738(3): 139669. Scholar


Lin, S., and Hernandez-Ramirez, G. 2021. Nitrogen turnover and N2O production in incubated soils after receiving field applications of liquid manure and nitrification inhibitors. Can. J. Soil Sci. 101(2). Scholar


Linn, D.M., and Doran, J.W. 1984. Effect of water-filled pore space on carbon dioxide and nitrous oxide production in tilled and nontilled soils. Soil Sci. Soc. Am. J. 48(6): 1267–1272. Scholar


Linzmeier, W., Gutser, R., and Schmidhalter, U. 2001. Nitrous oxide emission from soil and from a nitrogen-15-labelled fertilizer with the new nitrification inhibitor 3,4-dimethylpyrazole phosphate (DMPP). Biol. Fertil. Soils 34(2): 103–108. Scholar


Liu, R., Hayden, H.L., Suter, H., Hu, H., Lam, S.K., He, J., et al. 2017. The effect of temperature and moisture on the source of N2O and contributions from ammonia oxidizers in an agricultural soil. Biol. Fertil. Soils 53(1): 141–152. Scholar


Lu, Q., He, Z.L., and Stoffella, P.J. 2012. Land application of biosolids in the USA: A review. Appl. Environ. Soil Sci. 2012. Scholar


Mathieu, O., Hénault, C., Lévêque, J., Baujard, E., Milloux, M.-J., and Andreux, F. 2006. Quantifying the contribution of nitrification and denitrification to the nitrous oxide flux using 15N tracers. Environ. Pollut. 144(3): 933–940. Scholar


Mekala, C., and Nambi, I.M. 2017. Understanding the hydrologic control of N cycle: Effect of water filled pore space on heterotrophic nitrification, denitrification and dissimilatory nitrate reduction to ammonium mechanisms in unsaturated soils. J. Contam. Hydrol. 202(February): 11–22. Scholar


Oertel, C., Matschullat, J., Zurba, K., Zimmermann, F., and Erasmi, S. 2016. Greenhouse gas emissions from soils—A review. Chemie der Erde 76(3): 327–352. Scholar


Pu, G., Barry, G., and Bell, M. 2010. Gaseous nitrogen losses following soil amendment with Biosolids under controlled conditions. J. Residuals Sci. Technol. 7(4): 209–217. Google Scholar


R Core Team. 2019. R: A Language and Environment for Statistical Computing. R Foundation for Statistical Computing, Vienna, Austria. Google Scholar


Rigby, H., Perez-Viana, F., Cass, J., Rogers, M., and Smith, S.R. 2009. The influence of soil and biosolids type, and microbial immobilisation on nitrogen availability in biosolids-amended agricultural soils - Implications for fertiliser recommendations. Soil Use Manage. 25(4): 395–408. Scholar


Rigby, H., Clarke, B.O., Pritchard, D.L., Meehan, B., Beshah, F., Smith, S.R., and Porter, N.A. 2016. A critical review of nitrogen mineralization in biosolids-amended soil, the associated fertilizer value for crop production and potential for emissions to the environment. Sci. Total Environ. 541: 1310–1338. Scholar


Roman-Perez, C.C., and Hernandez-Ramirez, G. 2021. Sources and priming of nitrous oxide production across a range of moisture contents in a soil with high organic matter. J. Environ. Qual. 2006(July 2020): 94–109. Scholar


Roman-Perez, C.C., Hernandez-Ramirez, G., Kryzanowski, L., Puurveen, D., and Lohstraeter, G. 2021. Greenhouse gas emissions, nitrogen dynamics and barley productivity as impacted by biosolids applications. Agric. Ecosyst. Environ. 320(February): 107577. Scholar


Schaufler, G., Kitzler, B., Schindlbacher, A., Skiba, U., Sutton, M.A., and Zechmeister-Boltenstern, S. 2010. Greenhouse gas emissions from European soils under different land use: Effects of soil moisture and temperature. Eur. J. Soil Sci. 61(5): 683–696. Scholar


Sharma, B., Sarkar, A., Singh, P., and Singh, R.P. 2017. Agricultural utilization of biosolids: A review on potential effects on soil and plant grown. Waste Manag. 64: 117–132. PMID:28336334. Google Scholar


Shcherbak, I., Millar, N., and Robertson, G.P. 2014. Global meta-analysis of the nonlinear response of soil nitrous oxide (N2O) emissions to fertilizer nitrogen. Proc. Nat. Acad. Sci. USA 111(25): 9199–9204. Scholar


Smith, K.A. 2017. Changing views of nitrous oxide emissions from agricultural soil: key controlling processes and assessment at different spatial scales. Eur. J. Soil Sci. 68(2): 137–155. Scholar


Snyder, C.S., Bruulsema, T.W., Jensen, T.L., and Fixen, P.E. 2009. Review of greenhouse gas emissions from crop production systems and fertilizer management effects. Agric. Ecosyst. Environ. 133(3–4): 247–266. Scholar


Thangarajan, R, Bolan, N.S., Tian, G., Naidu, R., and Kunhikrishnan, A. 2013. Role of organic amendment application on greenhouse gas emission from soil. Science of the Total Environment. 465: 72–96. Scholar


Thilakarathna, S.K., Hernandez-Ramirez, G., Puurveen, D., Kryzanowski, L., Lohstraeter, G., Powers, L.-A., et al. 2020. Nitrous oxide emissions and nitrogen use efficiency in wheat: Nitrogen fertilization timing and formulation, soil nitrogen, and weather effects. Soil Sci. Soc. Am. J. (May 2020): 1–18. Scholar


Thilakarathna, S.K., and Hernandez-Ramirez, G. 2021. Primings of soil organic matter and denitrifcation mediate the effects of moisture on nitrous oxide production, soil nitrogen, and weather effects. Soil Biol. Biochem. (April 2021): 108166. Google Scholar


Torri, S.I., Corrêa, R.S., and Renella, G. 2017. Biosolid application to agricultural land—A contribution to global phosphorus recycle: A review. Pedosphere 27(1): 1–16. Scholar


Van Der Weerden, T.J., Kelliher, F.M., and De Klein, C.A.M. 2012. Influence of pore size distribution and soil water content on nitrous oxide emissions. Soil Res. 50(2): 125–135. Scholar


Velthof, G.L., Kuikman, P.J., and Oenema, O. 2003. Nitrous oxide emission from animal manures applied to soil under controlled conditions. Biol. Fertil. Soils 37(4): 221–230. Scholar


Wang, L., Du, H., Han, Z., and Zhang, X. 2013. Nitrous oxide emissions from black soils with different pH. J. Environ. Sci. (China) 25(6): 1071–1076. Scholar


Wijesekara, H., Bolan, N.S., Kumarathilaka, P., Geekiyanage, N., Kunhikrishnan, A., Seshadri, B. et al. 2016. Biosolids enhance mine site rehabilitation and revegetation, environmental materials and waste: Resource recovery and pollution prevention. Elsevier Inc. Scholar


Wrage, N., Velthof, G.L., Beusichem, M.L., and Oenema, O. 2001. Role of nitrifier denitrification in the production of nitrous oxide. Soil Biol. Biochem. 33(12–13): 1723–1732. Scholar


Yang, X.M., Drury, C.F., Reynolds, W.D., Tan, C.S., and McKenney, D.J. 2003. Interactive effects of composts and liquid pig manure with added nitrate on soil carbon dioxide and nitrous oxide emissions from soil under aerobic and anaerobic conditions. Can. J. Soil Sci. 83(4): 343–352. Scholar


Yoshida, H., Nielsen, M.P., Scheutz, C., Jensen, L.S., Christensen, T.H., Nielsen, S., et al. 2015. Effects of sewage sludge stabilization on fertilizer value and greenhouse gas emissions after soil application. Acta Agric. Scand. Sect. B Soil Plant Sci. 65(6): 506–516. Scholar


Zhu-Barker, X., Doane, T.A., and Horwath, W.R. 2015. Role of green waste compost in the production of N2O from agricultural soils. Soil Biol. Biochem. 83: 57–65. Scholar


Zhu, T., Zhang, J., Yang, W., and Cai, Z. 2013. Effects of organic material amendment and water content on NO, N2O, and N2 emissions in a nitrate-rich vegetable soil. Biol. Fertil. Soils 49(2): 153–163. Scholar


Zhu, X., Silva, L.C.R., Doane, T.A., Wu, N., and Horwath, W.R. 2013. Quantifying the effects of green waste compost application, water content and nitrogen fertilization on nitrous oxide emissions in 10 agricultural soils. J. Environ. Qual. 42(3): 912–918. Scholar
© 2021 The Author(s).
Carmen C. Roman-Perez and Guillermo Hernandez-Ramirez "Nitrous oxide production and nitrogen transformations in a soil amended with biosolids," Canadian Journal of Soil Science 102(2), 505-518, (15 October 2021).
Received: 28 May 2021; Accepted: 12 October 2021; Published: 15 October 2021
émissions d’oxyde nitreux
espace interstitiel empli d’eau
nitrification rate
nitrous oxide emissions
taux de nitrification
Back to Top