Temperature response of ammonia and greenhouse gas emission from manure amended silty clay soil

ABSTRACT

Soil temperature plays an important role in organic matter decomposition, thus likely to affect ammonia and gaseous emission from land application of manure. An incubation experiment was conducted to quantify ammonia and greenhouse gas (GHG) (N₂O, CO₂ and CH₄) emissions from manure and urea applied at 215 kg N ha⁻¹ to Fargo-Ryan silty clay soil. Soil (250 g) amended with solid beef manure (SM), straw-bedded solid beef manure (BM), urea only (UO), and control (CT) were incubated at 5, 10, 15, and 25 °C for 31 days at constant 60% water holding capacity (WHC). The cumulative GHGs and NH₃ emission generally increased with temperature and highest emission observed at 25 °C. Across temperature levels, 0.11–1.3% and 0.1–0.7% of the total N was lost as N₂O and NH₃, respectively. Cumulative CO₂ emission from manure was higher than UO and CT at all temperatures (P < 0.05). Methane accounted for <0.1% of the total C (CO₂ + CH₄) emission across temperatures. The Q₁₀ values (temperature sensitivity coefficient) derived from Arrhenius and exponential models ranged 1.5–3.7 for N₂O, 1.4–6.4 for CO₂, 1.6–5.8 for CH₄, and 1.4–5.0 for NH_3. Our results demonstrated that temperature significantly influences NH₃ and GHG emissions irrespective of soil amendment but the magnitude of emission varied with soil nutrient availability and substrate quality. Overall, the highest temperature resulted in the highest emission of NH₃ and GHGs.

KEYWORDS:

• Activation energy
• beef manure
• greenhouse gas
• soil temperature
• temperature sensitivity
• urea

Introduction

Soil temperature and its interaction with agricultural management practices such as addition of commercial (e.g. urea) and organic (e.g. livestock manure) N fertilisers, moisture, and soil characteristics [texture, cation exchange capacity (CEC), organic matter (OM), C:N etc.] Have been reported to induce gaseous production from soils (Chadwick et al. 2011; Ren et al. 2017).

Organic matter decomposition has often been overlooked and considered as quantitatively negligible during winter period when soil temperature is close to or below 0°C (Chantigny et al. 2002). Nevertheless, significant levels of microbial activity have been reported in soils in a colder environment (Dorland and Beauchamp 1991; Singurindy et al. 2009). Dorland and Beauchamp (1991) further showed that denitrifiers could function at -2 °C in an unfrozen soil. Carbon and N dynamics are closely linked to decomposition of animal manure and that the significance and extent of decomposition across variable temperature ranges are still poorly understood (Petersen et al. 2013). Kirschbaum (1995) suggested that a 1°C increase in temperature could ultimately lead to a loss of over 10% of soil organic C in regions of the world with an annual mean temperature of 5°C whereas the same temperature increase would lead to a loss of only 3% of soil organic C for a soil at 30°C. These findings further highlight the temperature effects on the soil respiration, N decomposition, nitrification and denitrification, and the subsequent production of CO₂ and N₂O.

The temperature sensitivity (Q₁₀) is a factor by which the rate of decomposition increases with a 10 °C rise in temperature and it has been widely used to depict the responses of soil gaseous emission rate to temperature changes (Davidson et al. 2006). Since Q₁₀ values of various agrosystems are not yet known, a default value of 2.0 has been used in exponential functions to simulate the effect of changing the temperature on gaseous emission rate (Li et al. 2015). A large variation in Q₁₀ values of N₂O (2.01–3.48) (Li et al. 2013), CO₂ (1.9–18.2) (Chen et al. 2010), and CH₄ (1.3–28) (Van Hulzen et al. 1999) are reported in literature, suggesting apparent uncertainty of Q₁₀ estimation. These variations indicate that Q₁₀ relationship cannot be explained on the simple basis of increased molecular agitation at higher temperatures. Attention has shifted to exponential and Arrhenius models to determine Q₁₀ (Lloyd and Taylor 1994; Davidson et al. 2006), that are more precise on reaction rates especially with laboratory incubation data. Furthermore, with Arrhenius equation, the decomposition rate and the temperature response of reactions are determined as activation energy (E_a), the minimum amount of energy required for the reaction to occur (Lloyd and Taylor 1994). The soil respiration rate increases with increase in temperature. However, the kinetic theory says that Q₁₀ will be higher in soil organic matter that are resistant to decomposition (e.g. cellulose), than those decompose easily (e.g. glucose). In general, organic substrate have complex molecular structures and have higher E_a and hence higher Q_10. However, environmental constraints (water availability and anaerobic vs. anaerobic sites) on organic matter decomposition can obscure the Q₁₀ by reducing substrate availability, often causing the Q₁₀ to be less than expected. Moreover, temperature affects the chemical processes of soil organic matter adsorption and desorption onto mineral surfaces, but little is known about the activation energies of these processes (Davidson and Janssens 2006).

Fargo Clay soils of North Dakota are productive soils with high organic matter content and undergoes cycles of cold winters to warm summers annually. In terms of livestock, recently, North Dakota beef cattle number increased to 6% from 2016 to 2017 (USDA-NASS 2017) and with this increase in beef cattle, the need for better manure management is essential. There is also limited information on the application of bedded manure as the use of manure bedding can influence the nutrient availability given the range of bedding types, application and incorporation timing, and climatic factors (Petersen et al. 2007). In addition, the influence of temperature on gaseous emission from soils amended with N-fertilisers may vary with variations in climate and management practices (Bowden et al. 1998; Jenkins and Adams 2010). The objectives of this study were to (1) assess the effects of temperature variations (5, 10, 15, and 25°C) on N₂O, CO₂, CH₄, and NH₃ emissions from urea and manure amended soils and (2) investigate temperature sensitivity (Q₁₀) models of decomposition on organic substrate addition using exponential and Arrhenius models.

Materials and methods

Soil and manure preparation

Five composite samples of surface soil (0–15 cm) from different areas within the filed were collected from the North Dakota State University (NDSU) research farm (46° 55’ 15” N, 96° 51’ 31” W) located in Fargo, North Dakota, USA. The soil is classified as Fargo-Ryan silty clay (fine, smectitic, frigid Typic Epiaquerts) with 0–1 percent slope (Soil Survey Staff 2016). Soils were then air-dried, ground down to pass a 2-mm sieve, and analyzed for their physical and chemical properties (Table 1). Soil pH and electrical conductivity (EC) were determined electrometrically in 1:2.5 soil/water extract (Thomas 1996) using Accumet AB pH metre (Fisher Scientific, Hampton, NH); soil texture (3.7% sand, 44.3% silt, and 52.1% clay) by hydrometer method (Elliott et al. 1999); soil organic matter (OM) by loss on ignition method (Combs and Nathan 1988); CEC by sodium acetate method (Chapman 1965); soil -N and -N concentration by KCL extract method (Maynard et al. 2008); soil water holding capacity (WHC) was determined following the procedure by Bowden et al. (1998); and total carbon and nitrogen results were obtained from the NDSU Soil Testing Laboratory, Fargo, North Dakota.

Table 1. Physical and chemical properties of soil (0–15 cm) and manure used in the study.

Properties	Soil	Solid manure	Bedded manure
pH	8.1	8.3	8.7
Electrical conductivity (EC) (dSm^−1)	0.82	5.9	4.8
NO3-N (mg kg^−1)	11.2
NH4-N (mg kg^−1)	2.2	873.3	703.3
Sand (%)	3.6
Silt (%)	43.8
Clay (%)	52.6
Cation exchange capacity (CEC) (cmolc kg^−1)	41.1
Organic matter (OM) (%)	6.3	25.8	23.2
Total C (%)	3.3	18.7	28.2
Total N (%)	0.28	1.1	0.72
C:N	11.8	17	39.2
Dry matter (%)		54	38
Bulk density (g cm^−3)	1.1

Manure used in this study consisted of (i) solid beef manure (SM) and (ii) solid beef manure with wheat straw bedding or bedded manure (BM). Representative samples of SM and BM were collected from the manure stockpile at the NDSU Beef Cattle Research Complex in Fargo, ND. Manures were then air-dried, ground down to pass a 2-mm sieve, and analyzed for their physical and chemical properties (Table 1). Manure pH, EC, NH₄-N, total C, and total N were determined using the same procedure used to determine soil properties. Additionally, the NDSU Soil Testing Laboratory, Fargo, North Dakota, determined manure dry matter content (% DM) for this study.

Experimental set-up

Emissions of N₂O, CO₂, CH₄, and NH₃ from manure N-sources as well as untreated urea (460 g N kg⁻¹) applied to Fargo silty clay soils were studied under controlled laboratory conditions. A set of 64 1-L clear glass canning jars (area = 46.6 cm²), each filled with 250 g of sieved soil were used to monitor gaseous emissions following the procedure described by Mukome et al. (2013) and Awale and Chatterjee (2017). Each of the canning jar lids were fitted with gas sampling port (butyl rubber septum) on the top surface. The bottom surface of each jar lids were attached to a metal wire and modified as a 50-mL cup-holder to capture NH₃ emission (discussed later).

The canning jars were placed in polystyrene boxes (to reduce temperature fluctuation) and 16 of the total 64 jars were then placed in each of four incubators set at 5, 10, 15, and 25 °C (±0.2 °C) and allowed to equilibrate and reactivate soil microorganisms for one week (Harrison-Kirk et al. 2013). Soil moisture inside each jar was maintained at 40% WHC during pre-incubation period. After one week of pre-incubation, 16 jars placed inside each incubator were assigned to four treatments and four replications. The four treatments used in this study included: (1) solid beef manure (SM), (2) solid beef manure with wheat straw bedding (BM), (3) urea only (UO), and (4) control (CT) with only the soil. The treatment assigned jars, except CT, were then amended by uniformly broadcasting 23 g of SM, 26 g of BM, and 218 mg of UO (equivalent to 215 kg N ha⁻¹) and incorporated immediately using a glass rod stirrer. The N application rates were determined based on the typical N application rate for a corn (Zea mays L.) as recommended for the Eastern North Dakota. The soil inside the jars was compacted to the height of 4.9 cm to achieve the field bulk density (determined by gravimetric method) of 1.1 g cm⁻³. The WHC of the soil was maintained at 60% by uniformly adding additional water to achieve 117 mL of deionised water over the soil surface inside each jar. A set of 16 jars were then placed inside the respective incubators set at 5, 10, 15, and 25 °C and were removed for gas sample collection on days 1, 2, 3, 4, 7, 10, 14, 17, 21, 24, 28, and 31 after treatment additions. For rest of the 31-day incubation period, the soil moisture inside each jar was maintained to 60% WHC by weighing each jar on sampling days and adding deionised water as a fine spray to replace the lost weight, if needed.

Gas sampling and analysis

On each sampling day, the headspace air inside each jar was mixed by pumping a polypropylene syringe three times, and finally, a 30-mL headspace air sample was collected for N₂O, CO₂, and CH₄ analysis. Headspace air samples were collected by inserting gas tight Luer lock syringe into the jar through lid septum, and immediately transferring it into a pre-evacuated 20-mL gas serum vials. Prior to sampling, gas serum vials were over-pressurized to eliminate air diffusion into them and to facilitate the subsequent removal of gas samples for analysis.

The gas samples were analyzed for N₂O, CO₂, and CH₄ within 24 h of their collection using gas chromatograph (GC) (Model No. 8610C, SRI Instruments, Torrance, CA) paired with a 10-vial Autosampler (SRI part# 8690-0047, SRI Instruments, Torrance, CA). The GC oven was operated at 60 °C and was fitted with electron capture detector (ECD) for N₂O and flame ionisation detector (FID) for CO₂ and CH₄ detection. The ECD and the FID were operated at 350 and 300 °C, respectively, and N₂ carrier gas was supplied at 20 PSI for the ECD whereas air and H₂ were supplied to FID/methanizer at 20 PSI. Compound peaks were recorded and analyzed with PeakSimple Chromatography Data System Software (Ver. 3.72; SRI Instruments, Torrance, CA, USA). A detailed description of the GHG measurement using SRI Instruments GC is described by Borhan et al. (2011). In addition, before each measurement, analytical gas standards (0, 1, 5, 50, 100, 500, and 1000 ppm for N₂O; 0, 500, 1000, 3000, 5000 ppm for CO₂; and 0, 4, 10, 30, 100 ppm for CH₄; Scotty Specialty Gases) were included to construct standard calibration curve. The gas concentrations were then analyzed following the procedure outlined by Mukome et al. (2013) and Awale and Chatterjee (2017). Briefly, the concentrations of N₂O, CO₂, and CH₄ were converted into mass units assuming ideal gas relations and expressed as micrograms N₂O-N, milligrams CO₂-C, and micrograms CH₄-C produced between sampling dates per kilogram of soil, respectively. Cumulative μg N₂O-N, mg CO₂-C, or μg CH₄-C emission per kg⁻¹ soil was calculated by adding the gas emissions during each sampling period. Daily and cumulative emissions were calculated as N₂O-N, CO₂-C, and CH₄-C, however for simplicity flux is herein referred to as N₂O, CO₂, and CH₄, respectively.

Ammonia (NH₃) volatilisation losses were measured immediately following the gas sampling using phosphoric acid (H₃PO₄) traps that were placed inside the headspace of the screw-top canning jars (Khan et al. 2001). Briefly, after treatment application, 50 mL clear plastic cup containing 15 mL of 0.5 M H₃PO₄ were placed in the cup-holder attached to the lid and the lid was then placed in the jar. After removal of NH₃ traps or after gas sampling the jars were kept open for ∼1 h for aeration. The collected traps were extracted with 50 mL of 2 M KCL solution and jars were replaced with freshly prepared 15 mL H₃PO₄ solution to facilitate NH₃ trapping until the next sampling day. The extracts were then analyzed for concentrations using an Automated Timberline TL2800 Ammonia Analyzer (Timberline Instruments, Boulder, CO). Daily NH₃ volatilisation loss (μg NH₃–N kg^–1 moist soil d^–1) were calculated by dividing the NH₃-N emitted between the sampling dates by the elapsed time. In addition, NH₃ volatilisation losses during each sampling periods were summed to obtain the cumulative NH₃-N losses (μg NH₃–N kg^–1) during entire 31-day incubation.

The emission factor (%EF), percent of N emitted as N₂O-N or NH₃-N from the applied N treatments was calculated using the following equation:where C_fert and C_control are the cumulative N₂O or NH₃ emissions from fertilised N treatments and control, respectively (Gagnon et al. 2011).

Statistical analysis

A completely randomised design (CRD) in a split-plot arrangement with four replications was used in this study. The four temperature regime (5, 10, 15, and 25 °C) were the whole-plot treatments and the four N fertilisers (SM, BM, UO, and CT) were subplot treatments. The effects of soil temperature and N-fertilisers on daily and cumulative N₂O, CO₂, CH₄, and NH₃ emission were analyzed using the PROC MIXED procedure of SAS (SAS Institute 2014). Soil temperature and N-fertiliser were considered fixed effects, and replication and interaction with replication were considered random effects. Residuals were evaluated for homogeneity of variance and normality using the UNIVARIATE procedure of SAS. Analysis of variance with a test of Tukey–Kramer Method at p-value ≤ 0.05 was used to determine the significant differences of means within and across temperature levels.

Regression analysis was conducted by fitting both, an exponential (ln soil flux vs. T) and Arrhenius functions (ln soil flux v. 1/T) (Jenkins and Adams 2010). Equations are shown below for soil flux measures over four temperature levels.(1) Where F is the soil flux rate (µg N₂O-N kg⁻¹ soil, mg CO₂-C kg^-1 soil, µg CH₄-C kg⁻¹ soil, and µg NH₃-N kg⁻¹ soil), a and b are the rate constants, and T is the temperature in Kelvin (K).(2) Where A is a constant, E_a is the activation energy of the process (kJ mol⁻¹), R is the universal gas constant (8.314 J mol⁻¹ K⁻¹), and T is the temperature in Kelvin (K). To determine E_a, natural logarithm (ln) of soil flux was plotted against 1/T to generate a slope (−E_a/R). The parameter E_a in the Arrhenius equation is directly related to the temperature sensitivity (Q₁₀) variation.

The Q₁₀ was then calculated from the derived parameter for both exponential and modified Arrhenius function accordingly:(3) (4)

Results and discussion

All fitted regressions were highly significant (P < 0.05) and the r² values indicate the strong fit of both Arrhenius (r² =0.89–0.95) or exponential functions (r² =0.90–0.96).

Temperature effect on N₂O emissions

This study was focused on direct effects of variable temperature on soil N₂O emission. In presence of either organic or inorganic N fertiliser. The daily mean N₂O emissions at 5, 10, 15, and 25 °C ranged from 1 to 76, 1 to 146, 1 to 217, and 1 to 704 μg N₂O-N kg⁻¹ d^-1, respectively, with the highest daily emission from urea only (UO) amended soil at all temperature regimes. The analysis of treatment × day showed significant differences in N-treatments at all temperature levels for the first 6 days of incubation (P < 0.05). Fluxes of N₂O during 1st day after soil application were up to 6-folds higher for the manured soil than for urea applied soil, across temperature gradient (Figure 1). Thereafter, fluxes from manured soils gradually decreased for rest of the incubation period. The sharp peak on 1st day is likely due to enhanced denitrification of soil by the addition of easily degradable organic substrates. Velthof et al. (2003) also reported a peak in N₂O emission on the first days after manure application. Moreover, volatile fatty acids from the manure are metabolised within a few days by soil bacteria, increasing denitrification and/or immobilisation of N (Kirchmann and Lundvall 1993). In addition, the amount of manure applied to soil was over 100 times the amount of urea that was applied to the soil that likely created anaerobic microsites. It would also explain the relatively slow start of N₂O emission from urea.

Figure 1. Daily soil N₂O fluxes after N fertilizers [SM (solid beef manure), straw-bedded solid beef manure (BM), urea only (UO), and control (CT)] application on silty clay soils at 5, 10, 15, and 25°C incubation temperatures. Vertical bars are standard errors (n = 4). *Indicates any significant (P ≤ 0.05) differences between treatments at the day. Please note the large differences in y-axis scaling.

Contrary to N₂O emission trend of manured soil, urea applied soils gradually increased to its peak level on day-7 for soils at 5 and 10 °C and on day-4 for soils at 15 and 25 °C (Figure 1). Of the total incubation time, the daily emission from urea-applied soils were higher for 58–83% of the time compared to manured soils. A large portion of manure N in our study was in the organic form (Table 1) and organic-N requires mineralisation followed by nitrification to form manure derived NO₃^- pool for denitrification (Chadwick et al. 2011). Previous studies (Macleant and McRae 1987; Halvin et al. 2013) predicted the hydrolysis of urea in the range of 83-90% whereas organic substance mineralisation rates are reported in the range of 16-53% (Hernandez-Ramirez et al. 2009). In this study, rapid hydrolysis of urea to NH₄⁺ most likely influenced the higher N₂O emission across all temperature regimes with emission in the decreasing order of 25 > 15 > 10 > 5°C. In addition, manure induced NH₃ volatilisation loss following the incorporation most likely reduced the pool of N available for N₂O emission.

The cumulative soil N₂O emission and the N₂O emission factors within and across all four temperature regimes and treatments are summarised in Tables 2 and 3, respectively. Compared to CT, cumulative N₂O emissions from the SM, BM, and UO applied soils during the entire incubation period were significantly higher (all P < 0.05; Table 2) by up to 17 folds. Low availability of N substrate in control soil compared to fertilised soils most likely influenced N₂O emissions differently.

Table 2. Cumulative GHGs (N₂O, CO₂ and CH₄) emission and NH₃ volatilization loss from N fertilizers over 31 days of incubation at 5, 10, 15, and 25°C.

N fertilizer source	5 °C	10 °C	15°C	25°C
	Cumulative N2O-N (μg N2O-N kg^−1 soil)
Solid beef manure	351 ± 31 aA	495 ± 23 abAB	725 ± 36 abB	1569 ± 81 aC
Bedded manure	289 ± 49 aA	409 ± 59 bAB	452 ± 20 bB	1199 ± 77 aC
Urea	447 ± 57 aA	673 ± 22 aAB	848 ± 38 aB	3083 ± 97 bC
Control	43 ± 6 bA^†	47 ± 5 cA	62 ± 4 cA	174 ± 9 cB
	Cumulative CO2-C (mg CO2-C kg^−1 soil)
Solid beef manure	278 ± 27 aA	450 ± 30 aB	517 ± 29 aB	3260 ± 118 aC
Bedded manure	238 ± 11 aA	343 ± 9 aAB	462 ± 30 aB	2407 ± 60 aC
Urea	150 ± 15 bA	227 ± 20 bB	347 ± 14 bC	1377 ± 101 bD
Control	141 ± 7 bA	206 ± 25 bA	316 ± 19 bB	1189 ± 95 bC
	Cumulative CH4-C (μg CH4-C kg^−1 soil)
Solid beef manure	4.3 ± 2 aA	10.2 ± 1 aA	22.1 ± 2 aB	52.8 ± 3 aC
Bedded manure	6.5 ± 1 aA	10.7 ± 1 aA	27.2 ± 2 aB	61.5 ± 6 aC
Urea	2.2 ± 0 aA	3.2 ± 1aA	6.8 ± 2 bAC	11.1 ± 4 bC
Control	1.4 ± 1 aA	2.7 ± 3 aA	5.0 ± 4 bAC	7.8 ± 11 bC
	Cumulative NH3-N (μg NH3-N kg^−1 soil)
Solid beef manure	272 ± 5 aB	319 ± 1 aAB	461 ± 16 aA	1039 ± 108 aC
Bedded manure	254 ± 5 aB	340 ± 20 aAB	442 ± 6 aA	977 ± 61 aC
Urea	397 ± 31 bA	494 ± 36 bA	1309 ± 123 bB	2193 ± 122 bC
Control	92 ± 18 cA	100 ± 2 cA	224 ± 1 cB	498 ± 6 cC

^†Values are means ± standard errors (n = 4). Different lowercase letters within a column and different uppercase letters across a row indicate significant differences at 0.05 significance level.

Table 3. Emission factors (%) of N₂O and NH₃ from N fertilizers estimated from the cumulative emission of 31 days of incubation at 5, 10, 15, and 25 °C.

N fertilizer source	5°C	10°C	15°C	25°C
	N2O Emission factor (EF, %)
Solid beef manure	0.14	0.2	0.29	0.62
Bedded manure	0.11	0.16	0.17	0.46
Urea	0.18	0.28	0.35	1.29
	NH3 Emission factor (EF, %)
Solid beef manure	0.09	0.09	0.10	0.23
Bedded manure	0.07	0.10	0.09	0.20
Urea	0.13	0.17	0.46	0.69

No significant difference in the total N₂O emission was observed between the three N-treated soils at 5 °C. A significant difference in cumulative N₂O emission from all soil was observed only at 25 °C where the highest emissions were observed for all treatments under study (Table 2).High temperature in moist soil can cause an increase in microbial activity increasing the activity of denitrifiers in soil. This was most likely the reason for the higher concentration of N₂O at 25 °C. Chantigny et al. (2002) reported similar result where maximum N₂O accumulations occurred when O₂ content fell below 6%, whereas the time of complete N₂O disappearance in the jars corresponded fairly well with the time of depletion.

When comparing manure-types, no significant differences were found on cumulative N₂O emission from SM and BM treated soils at all temperature regime. The presence of high and readily available NH₄-N in UO compared to manure may have favoured higher nitrification in UO amended soil thus producing higher , a source for denitrification (). Moreover, soils were at 60% WHC and both nitrification and denitrification can occur simultaneously at ∼60% WHC (Linn and Doran 1984; Uchida et al. 2013). Urea treated soil had significantly higher N₂O emission than BM at T ≥ 10 °C by a factor of 1.5–2.6 across temperature regime while it was significantly higher than SM only at 25 °C. Because urease is an extracellular enzyme produced by microorganisms, high temperatures can increase microbial growth and urease production. Moyo et al. (1989) found that increasing temperature from 5 to 45 °C greatly increased urease activity. Furthermore, lower cumulative N₂O emission from manure amended soil compared to UO can be attributed to the higher C:N ratio in SM and BM, which may have promoted N immobilisation and/or slow mineralisation. The cumulative N₂O emission were negatively correlated with the C:N ratio in organic amendments in other studies (Baggs et al. 2000; Huang et al. 2004) as well. Thus, it can be concluded that at ∼ 60% WHC levels, soil temperature was the main factor determining the N₂O emission from the N-treated soils and its effects are manifested with the increasing temperature regimes.

The values of EFs of N₂O were in the range of 0.11–1.29% (Table 3). Our results indicated that the EF values are proportional to the increase in temperature, with the highest value of 1.29% from UO applied soil at 25 °C. The low loss of fertiliser N (EF = 0.11–0.16) were expectedly observed at low temperatures (5–10 °C). Across all temperature regime, EF followed the decreasing trend of UO > SM > BM. The obtained values from our study agree with the annual EF values of the agricultural soils (0.1–7.3%) (Bouwman 1996; Dobbie and Smith 2003).

The relationship between soil N₂O emission and temperature were approximated by both Arrhenius and exponential functions (Table 4). Temperature response of N₂O production fitted to an Arrhenius function in the range 5–25 °C generated apparent activation energies (E_a) between 49.5 and 66.7 kJ mol⁻¹ (Table 4). This value corresponds to a Q₁₀ of ∼2, which indicates that the reaction rates doubled for every 10 °C rise in temperature. Our temperature response for the 5–25 °C range are similar to those observed by others; the activation energy of disappearance by denitrification has been estimated to be 28–76 kJ mol⁻¹ in a temperate soil (Holtan-Hartwig et al. 2002) to 47–89 kJ mol⁻¹ for riparian soils (Maag et al. 1997). The reaction rate of the higher activation energy in treatments increased faster while undergoing the same extent of temperature rise. The Q₁₀ derived from both Arrhenius and exponential equations were in the range of 1.94 to 2.74 (Table 5). Arrhenius model showed a slight decrease in Q₁₀ with increasing temperature [calculated using equation (4) at 5, 10, 15, and 25 °C] (Table 5).

Table 4. Fitted parameters describing Arrhenius and two-parameter exponential function for soil GHG emission (N₂O-N, CO₂-C, and CH₄-C) and NH₃ volatilization loss measured over four temperature regime.

N fertilizer source	Arrhenius Equation			Two-parameter Exponential
		N2O-N
	A	Ea (kJ mol^−1)	r^2	a	b	r^2
Solid beef manure	e^28.4	52.2	0.98	1.7 × 10^−6	0.08	0.98
Bedded manure	e^27	49.5	0.91	7.8 × 10^−5	0.07	0.91
Urea	e^34.8	66.7	0.95	1.0 × 10^−8	0.1	0.96
Control	e^25.4	50.5	0.89	4.6 × 10^−7	0.07	0.9
		CO2-C
Solid beef manure	e^41.8	84.2	0.91	1.4 × 10^−12	0.12	0.92
Bedded manure	e^39.9	80.2	0.93	2.3 × 10^−12	0.12	0.94
Urea	e^38.3	77.4	0.96	2.3 × 10^−11	0.11	0.96
Control	e^36.9	74.2	0.96	1.4 × 10^−10	0.11	0.97
		CH4-C
Solid beef manure	e^40.2	90.2	0.84	2.9 × 10^−13	0.13	0.83
Bedded manure	e^37.7	82.9	0.97	6.5 × 10^−12	0.12	0.96
Urea	e^23.6	53.1	0.62	4.2 × 10^−3	0.08	0.61
Control	e^23.2	53.0	0.51	1.7 × 10^−2	0.08	0.51
		NH3-N
Solid beef manure	e^25.8	47.1	0.96	2.1 × 10^−5	0.07	0.97
Bedded manure	e^25.5	46.3	0.98	5.7 × 10^−6	0.07	0.98
Urea	e^32.9	62.3	0.92	3.9 × 10^−8	0.09	0.92
Control	e^32.1	64.2	0.92	2.2 × 10^−8	0.09	0.93

Note: Regression analysis was conducted using linearized data for both Arrhenius (soil NH₃ vs. 1/T) and exponential functions (ln soil NH₃ vs. T) for all reported data. All regressions were significant at P < 0.05.

Table 5. Temperature dependence of soil GHG emission (N₂O-N, CO₂-C, and CH₄-C) and NH₃ volatilization loss across N fertilizer sources.

N fertilizer source	Arrhenius Equation (Q10)^‡				Exponential Q10
	5°C	10°C	15°C	25°C
	N2O-N
Solid beef manure	2.20 ± 0.1 abA	2.14 ± 0.1 abAB	2.09 ± 0.1 aB	1.99 ± 0.1 aC	2.14 ± 0.1 a
Bedded manure	2.15 ± 0.3 aA	2.09 ± 0.2 aAB	2.04 ± 0.2 aB	1.94 ± 0.2 aC	2.09 ± 0.2 a
Urea	2.74 ± 0.2 bA	2.65 ± 0.2 bB	2.56 ± 0.2 aC	2.4 ± 0.1 aD	2.65 ± 0.2 b
Control	2.15 ± 0.2 aA^†	2.09 ± 0.1aA	2.04 ± 0.1 aAB	1.95 ± 0.1 aB	2.1 ± 0.1 a
	CO2-C
Solid beef manure	3.55 ± 0.1 aA	3.40 ± 0.1 aB	3.26 ± 0.1 aC	3.02 ± 0.1 aB	3.41 ± 0.1 a
Bedded Manure	3.33 ± 0.1 aA	3.20 ± 0.1 aB	3.07 ± 0.1 aC	2.86 ± 0.0 aD	3.21 ± 0.1 a
Urea	3.21 ± 0.2 aA	3.09 ± 0.2 aB	2.97 ± 0.2 aC	2.77 ± 0.2 aD	3.09 ± 0.2 a
Control	3.1 ± 0.2 aA	2.94 ± 0.2 aB	2.84 ± 0.1 aC	2.65 ± 0.1 aD	2.95 ± 0.1 a
	CH4-C
Solid beef manure	3.97 ± 0.5 aA	3.78 ± 0.5 aAB	3.61 ± 0.4 aB	3.32 ± 0.4 aC	3.77 ± 0.5 a
Bedded manure	3.52 ± 0.4 aA	3.37 ± 0.3 aA	3.23 ± 0.3 aAB	2.99 ± 0.3 aC	3.35 ± 0.3 a
Urea	2.56 ± 0.8 aA	2.46 ± 0.7 aB	2.38 ± 0.7 aC	2.23 ± 0.6 aD	2.44 ± 0.7 a
Control	2.63 ± 0.8 aA	2.53 ± 0.8 aAB	2.44 ± 0.7 aAB	2.29 ± 0.7 aB	2.52 ± 0.8 a
	NH3-N
Solid beef manure	2.03 ± 0.1 aA	1.99 ± 0.1 acAB	1.94 ± 0.1 bB	1.86 ± 0.1 aC	1.99 ± 0.1 a
Bedded manure	2.01 ± 0.1 aA	1.96 ± 0.1 aAB	1.92 ± 0.1 bB	1.84 ± 0.1 aC	1.96 ± 0.1 a
Urea	2.56 ± 0.2 bA	2.48 ± 0.2 bcB	2.41 ± 0.2 abC	2.27 ± 0.2 aD	2.47 ± 0.2 b
Control	2.65 ± 0.2 abA	2.56 ± 0.2 bB	2.48 ± 0.2 aC	2.33 ± 0.2 aD	2.56 ± 0.2 b

^†Values are means ± standard errors (n = 4). Different lowercase letters within a column and different uppercase letters across a row indicate significant differences at 0.05 significance level.

^‡Q₁₀ calculated according to equation (3) for two-parameter exponential and equation (4) for Arrhenius model.

The Q₁₀ for N₂O production not only expressed the temperature effect but rather a combined effect of temperature and the development of anaerobic microsites in the soils under incubation. Our findings further illustrate that temperature regulates the soil denitrification both directly and indirectly, the latter by influencing the availability of O₂, , N₂O and C substrates. In summary, both organic amendments (SM and BM) and UO appeared to stimulate microbial activity in warm as well as cold soils, indicating that C and N transformations can occur throughout the winter and summer period in soils. Our results suggest that N₂O production can occur even at a temperature close to 5 °C in Fargo-clay soils.

Temperature effects on CO₂ emission

In general, rates of respiration increased across temperature gradient from 5 °C to 25 °C throughout the incubation period (Figure 2). In agreement with previous studies (Brooks et al. 1997; Chantigny et al. 2002), soil respiration occurred at significant rates in our study even at low temperature (5 °C). The daily average CO₂ emission varied from 1.8–79.4, 2.8–104.6, 2.7–133.2, and 11.7–641.8 mg CO₂-C kg⁻¹ soil d⁻¹ at 5, 10, 15, and 25 °C, respectively. The CO₂ flux rate was large right after manure application, which was most likely due to rapid decomposition of labile and refractory organic carbon in manure. At low temperatures, the CO₂ fluxes have been linked to labile organic carbon because at higher temperature refractory organic carbon are likely to increases (Bais et al. 2006). However, we did not investigate the changes in soil organic matter over time in this study. In our study, 25 °C temperature significantly increased daily average CO₂ flux compared with treatments at temperatures <25 °C (Figure 2). A marked decrease in C mineralisation rates was previously reported when amended soils change from aerobic to anaerobic condition (Chantigny et al. 2002). Our results suggest that soils slowly progressed towards anaerobic environment after the initial peak of CO₂ emission. The progressive decrease after the peak CO₂ emission from all soils was observed at all temperature regimes during the incubation, which was likely caused by a gradual shortage of easily available C from the amendments.

Figure 2. Daily soil CO₂ fluxes after N fertilizers [SM (solid beef manure), straw-bedded solid beef manure (BM), urea only (UO), and control (CT)] application on silty clay soils at 5, 10, 15, and 25°C incubation temperatures. Vertical bars are standard errors (n = 4). *Indicates any significant (P ≤ 0.05) differences between treatments at the day. Please note the large differences in y-axis scaling.

The cumulative CO₂ emission followed the trend of 25 > 15 > 10 > 5 °C, with the highest emission from SM followed by the BM treated soil at all temperature levels. The fertiliser application rate in this study was based on N based and the total C differ between SM and BM. Amendments at 25 °C resulted in significantly greater emissions than amendments at 5, 10, and 15 °C (Table 2). Moreover, cumulative CO₂ emission from SM and BM were significantly higher than that in UO and CT at all temperature regime, whereas there was no difference in the emission comparison between UO and CT only (Table 2). When comparing the manure-type, CO₂ emission from SM treated soil was 16.8, 31.2, 11.9, and 35.4% higher than BM treated soil at 5, 10, 15, and 25 °C, respectively, however no significant differences in cumulative CO₂ emission was observed.

The differences in cumulative CO₂ emission from UO and manure-N sources can be attributed to lower microbial activity in UO treated soil at low C:N ratio. Several studies showed a return of straw or addition of organic manure to soil can result in substantial increases in the soil respiration rate (Ding et al. 2007; Mapanda et al. 2011). Therefore, manure with and without straw bedding may have greatly increased the number of substrates for soil microorganisms, subsequently increasing SOM decomposition, further aided by the increasing temperature levels. The low CO₂ emission in UO applied soil may also be due to the suppression of the respiration of native soil organic C. Several other possible mechanisms were suggested in the literature. First, a decrease in pH caused by nitrification may inhibit microbial activity (Kuzyakov et al. 2000). Second, soil microbial population may have been adversely affected by the increase in solute concentration (DeForest et al. 2004). Third, high N levels in N-abundant soils repressed the synthesis and activity of certain enzymes (Carreiro et al. 2000). Our study showed the response of CO₂ emission to UO is most likely temporary; however, long-term study is suggested to investigate whether there exist differences in soil respiration between UO and CT soil on a long-term basis.

There are still uncertainties associated with modelling the strong temperature dependence of soil respiration. Some researchers (Lloyd and Taylor 1994; Thierron and Laudelout 1996) strongly recommended Arrhenius equation because it gives evenly distributed residual variances across the temperature range. Buchmann (2000), however, found that an exponential equation could more accurately explain the observed relationships. In our study, Q₁₀ values calculated based on the Arrhenius model and exponential function ranged between 2.65–3.55 and 2.95–3.41, respectively (Table 5), and within the range of 2.0 to 3.9 generally given for bulk soil respiration (Raich and Schlesinger 1992; Ding et al. 2007). Moreover, our r² values from Arrhenius and exponential equations indicate that 92–97% of the variation in soil CO₂ emission under the current incubation conditions can be explained by the change in temperature (Table 4). The Arrhenius equation further showed that Q₁₀ of CO₂ emission was highest under cold temperatures suggesting temperature increase to have a larger impact on CO₂ emissions in cold areas (e.g. northern latitudes or mountain areas). Overall, this study showed that at 60% WHC level in Fargo-clay soil of the RRV flux-temperature relationships alone can be used to reasonably estimate CO₂ emission. However, as indicated by the variable rate constants CO₂ from manure can also be attributed to differences in C input as indicated by total C and C:N ratio of manure treatments used. Our data support the kinetic theory of activation energy and previous findings of low respiration and carbon concentration with increasing Q₁₀ (Jenkins and Adams 2010).

The apparent activation energy for respiration (E_a) was at 84, 80, 77, and 74 kJ mol⁻¹ in SM, BM, UO, and CT, respectively (Table 4). This suggests that the energy required for SOM decomposition is related to substrate quality and that higher E_a associated with the breakdown of recalcitrant substrates result in a greater Q₁₀ at low temperatures. To a certain extent, E_a reflected soil substrate and soil microbial activity, and that higher energy is required to initiate a reaction in complex organic compounds. We found that the difference in the Q₁₀ values as result of the added substrate are not coincidental. Additionally, the significantly higher Q₁₀ values for CO₂ than N₂O (P < 0.05) might indicate that the impacts of global warming on the emission of CO₂ from soils were higher than on the emissions of N₂O in the silty clay soils of the RRV.

Temperature effects on CH₄ emissions

Daily variations of soil CH₄ emission and uptake for soils incubated at 5, 10, 15, and 25 °C are presented in Figure 3. During the 31-day soil incubation study, the average CH₄ emission ranged from −0.45 to 1.33, −0.21 to 2.20, −3.39 to 8.16, and −1.85 to 14.83 µg kg⁻¹ soil d⁻¹ at 5, 10, 15, and 25 °C, respectively, where negative values indicate CH₄ consumption and positive values indicate the net CH₄ emission from the soil under the current incubation condition. Previous studies (Hutsch 2001; Le Mer and Roger 2001) show CH₄ production in the anaerobic zones of submerged soils by methanogens and the oxidation into CO₂ by methanotrophs in the aerobic zones of wetland and upland soils. The significant differences in daily average CH₄ flux were most likely due to higher emission from manure treated soil, most noticeably observed with increasing temperature regime (Figure 3). Organic matter addition is one of the factors that favour CH₄ emission from soils (Le Mer and Roger 2001).

Figure 3. Daily soil CH₄ fluxes after N fertilizers [SM (solid beef manure), straw-bedded solid beef manure (BM), urea only (UO), and control (CT)] application on silty clay soils at 5, 10, 15, and 25°C incubation temperatures. Vertical bars are standard errors (n = 4). *Indicates any significant (P ≤ 0.05) differences between treatments at the day. Please note the large differences in y-axis scaling.

The cumulative CH₄ emission across temperature regime was in the range of 1.4–61.5 μg CH₄-C kg⁻¹, with the highest emission from BM treated soil at 25 °C (Table 2). Moller et al. (2004) also showed an increase in livestock-based CH₄ production from the use of straw bedded manure due to higher volatile solids (VS) content of straw-bedded manure. Combined across all temperature regime, manure treated soil had significantly higher emissions (P < 0.05) compared to UO and CT soils. Further analysis of temperature × treatment showed no significant differences in CH₄ emission for treatments at 5 and 10 °C. Similarly, at 15°C, only BM treated soil showed significantly higher emission than CT soils while there were no significant differences between other treatments. The lack of significant differences at T < 15°C even when the cumulative emission from UO and CT were on average 4-folds lower than manure treated soil was probably due to high variability in the CH₄ emission data. The literature on CH₄ emissions from soils as the consequence of CH₄ production, consumption and transport show large spatial variability (Bartlett and Harriss 1993; Van Den Pol-Van Dasselaar et al. 1999). However, at 25 °C, both SM and BM treated soils had significant CH₄ emission compared to UO treated and CT soils (Table 2).

In this study, CH₄ emission was weakly influenced by temperature with CH₄ accounting for less than 0.1% of the total C (CO₂ + CH₄) emission at all temperature regime and for all treatments. The major portion (83%) of the total CH₄ emission came from manure sources suggesting that CH₄ originated primarily from the decomposition of straw and manure rather than SOM sources.

The Q₁₀ from Arrhenius equation ranged from 2.2 to 4.0, with highest Q₁₀ from SM amended soil at 5 °C (Table 5). The decrease in Q₁₀ values (higher Q₁₀ at low temperature and vice-versa) from Arrhenius equation have been observed in many other studies (Lloyd and Taylor 1994; Schipper et al. 2014). Kirschbaum (1995) reported the Q₁₀ of organic matter decomposition decrease with the increase in temperature (4.5 at 10°C and 2.5 at 20°C). The Q₁₀ from two-parameter exponential equation were at 3.8, 3.4, 2.4, and 2.5 for SM, BM, UO, and CT respectively, and within the range of abovementioned studies.

The two parameter exponential functions and E_a calculated are summarised in Table 4. The values of E_a were at 90, 83, 53, and 53 kJ mol⁻¹ in SM, BM, UO, and CT, respectively. Apparent activation energy in this study comprised temperature response of both, CH₄ emitting and CH₄ oxidising microbial communities present in the surface layers of the soil. Moreover, E_a also comprised the temperature response of the physical transport of CH₄ into the atmosphere. Thus, E_a of CH₄ emission measured the temperature response of the overall reaction suggesting that recalcitrant organic matter might have higher E_a for decomposition. Our initial prediction from this study is that soils of RRV can be both the source or sink to CH₄ emission across all temperature regime; however, the interaction of organic matter and factors such as moisture content needs further evaluation.

Temperature effects on NH₃ volatilisation loss

Daily NH₃ volatilisation loss of N-fertiliser treatments from soils incubated at four different temperature levels are presented in Figure 4. The 31-day average NH₃ emission at 5, 10, 15, and 25 °C ranged from 7.7 to 33.1, 8.4 to 41.1, 18.7 to 109.1, and 41.5 to 178.25 µg NH₃ kg⁻¹ soil d⁻¹, respectively, indicating an increase in NH₃ volatilisation loss with the increase in temperature levels across all treatments. Maximum ammonia loss rates did not develop until 10 days after urea application at 5 and 10°C in all soils, presumably due to delayed urea hydrolysis and lower temperature. A lower urease activity would be expected in soil incubated at 5 °C compared to 25 °C because of the decrease in general biological activity at low temperatures. Temperature effect on NH₃ emission reported by Clay et al. (Clay et al. 1990) showed a linear increase in volatilisation of NH₃ in soil with an increase in temperature. Soil warming at higher temperature possibly increased the turbulence in the soil surface inside the jar increasing the transport of NH₃ away from the surface and into the headspace inside the incubation jars. Soils in this study were maintained at 60% WHC by adding additional water to balance the lost water through evaporation; however, the drying (evaporation) rate was not measured. Temperature increases the rate of diffusion in the soil water and air, increasing the volatilisation from the soil surface as shown in a number of studies (Freney et al. 1983; Harrison and Webb 2001; Halvin et al. 2013).

Figure 4. Daily soil NH₃ fluxes after N fertilizers [SM (solid beef manure), straw-bedded solid beef manure (BM), urea only (UO), and control (CT)] application on silty clay soils at 5, 10, 15, and 25°C incubation temperatures. Vertical bars are standard errors (n = 4). *Indicates any significant (P ≤ 0.05) differences between treatments at the day. Please note the large differences in y-axis scaling.

The cumulative NH₃ volatilisation loss ranged from 92 µg NH₃ kg⁻¹ in CT soils at 5 °C to 2193 µg NH₃ kg⁻¹ in UO amended soil at 25 °C (Table 2). Ammonia volatilisation losses were influenced significantly by temperature × N-fertiliser with the highest cumulative NH₃ emission from UO treated soil at all temperature levels. The UO amended soil alone accounted for 39, 39, 54, and 46% of the total NH₃ volatilisation loss at 5, 10, 15, and 25 °C, respectively. The increase in urea hydrolysis rate with increasing temperature most likely led to the more rapid formation of , pushing the equilibrium between and NH₃ towards NH₃. In contrast, Sommer et al. (2004) suggested that volatilisation continues for a longer period at low temperature compared to high temperature, consequently, the total loss is most related to soil properties and other climatic variables.

When comparing manure-type differences (SM vs BM), NH₃ volatilisation loss in SM was 7.1, 6.6, 4.3, and 6.7% higher than BM treated soil within 5, 10, 15, and 25 °C, respectively, however, the difference was not significant (within the column, Table 2). Dry matter (DM) content in solid manure vary between 31 and 67% and higher dry matter content in manure are linked to increases the NH₃ volatilisation potential of manure (Sommer and Hutchings 2001; Huijsmans et al. 2003). In our study, SM had 41% higher DM compared to BM (Table 1). The equilibrium between and NH₃ in manure are also regulated by the pH levels, with higher pH (7–10) associated with the increase in the rate of reaction and subsequent formation of NH₃ (Rhoades et al. 2008). Rhoades et al. (2008) also reported an increase in NH₃ volatilisation loss from the feedlot surface with increased ambient temperature from spring to summer. In the current study, SM and BM pH levels were at 8.3 and 8.7, respectively, which probably influenced similar emissions.

When comparing individual treatment across temperature levels (e.g. SM at 5, 10, 15, and 25°C), effects of temperature were more evident with an increase in temperature (Table 2). For soils with manure amendments, significantly higher NH₃ volatilisation loss was observed only when the temperature difference was ≥ 10 °C (5 vs.15°C, 5 vs. 25°C, and 10 vs. 25°C). However, for CT and UO amended soils, only 5 vs 10 °C showed no significant difference in NH₃. These results suggest temperature as a major factor that can increase the rate of N loss through volatilisation. High temperature not only increases urease activity and formation of and OH^- in the soil solution (Lai and Tabatabai 1992) but also increases the conversion of to NH₃ and the diffusion of NH₃ from the aqueous phase to the air phase volatilisation (Sander 1999). The influence on NH₃ volatilisation loss from soils under variable soil moisture contents have been discussed in previous studies (Liu et al. 2007; Yan et al. 2016), however, it is to be noted that our study shows the effect of temperature at 60% WHC only.

Across N treatments and temperature levels, NH₃ volatilisation loss was in the range of 0.1–0.7% of the total applied N with the most loss from UO treated soils at 25 °C (Table 3). Others (Jarecki et al. 2008; Cameron et al. 2013) have reported much higher NH₃ volatilisation losses than what we observed. The combined effects of manure and soil properties (silty clay with fine texture, high organic matter and high CEC) most likely reduced the overall cumulative NH₃ volatilisation loss from the current study. In addition, N-treatments used in this study were immediately incorporated into soil that likely increased the volume of soil to retain and reduced the total NH₃ emission. Immediate incorporation of surface-applied N fertilisers have potential to reduce the total NH₃ volatilisation loss by 60–90% (Huijsmans et al. 2003; Agnew et al. 2010).

Both, Arrhenius and exponential functions were used to approximate the relationship between soil NH₃ loss and temperature (Table 5). All fitted regressions were highly significant (P < 0.05) and the r² values indicate the strong fit of both, Arrhenius or exponential functions (r² =0.92–0.98). Temperature response of NH₃ emission fitted to an Arrhenius function in the range 5–25 °C generated apparent activation energies (E_a) between 47.1 and 64.2 kJ mol⁻¹, suggesting variation in the minimum amount of energy required to ensure that the reaction will occur varies with temperature and N-source (Table 4). The variation in E_a values between different treatments indicates differences in desorption potential of the N-amended soils and increased urease activity. In our study, the Q₁₀ derived from Arrhenius and exponential function was in the range of 1.8 to 3.0 and 1.96 to 2.56, respectively. These results indicate that Q₁₀ and E_a could be used as indices for N management for crop production to minimise NH₃ emissions.

This study examined the greenhouse gas and ammonia emissions response to variable temperature (5, 10, 15, and 25 °C) from soil amended with manure and urea under incubation. Our results demonstrated that temperature significantly influences ammonia and GHG emissions irrespective of soil amendment but the magnitude of emission varied with soil nutrient availability and substrate quality. Overall, the highest temperature resulted in the highest emission of ammonia and GHGs. Arrhenius and exponential functions were used to approximate the relationship between soil gaseous emissions (N₂O, CO₂, CH₄, and NH₃) and temperature. Soil temperature explained the seasonal variation in soil fluxes by 51–97% using Arrhenius and exponential models. The Q₁₀-values for the soils incubated at higher temperatures tended to be lower than those at lower temperatures, suggesting that the response of microbial respiration, nitrification and denitrification may differ due to the difference in the choice of N-fertiliser and climate. Thus, temperature can influence soil and manure properties such as OM, moisture, and C:N ratio. We recommend in the future to conduct a long term lab studies to address these changes with destructive sampling as well as long-term field experiment to document the rate and extent of emissions. The findings from this study could be useful in developing alternative N management strategies to conventional urea application for mitigating N losses in arable systems within the RRV.

Acknowledgement

The authors thank their colleague Dr. MS Borhan, Research Specialist at Agricultural and Biosystems Engineering, North Dakota State University, Fargo, for proofreading, and providing valuable suggestions on the manuscript. We also thank the reviewers for providing constructive feedback on this manuscript. This research did not receive any specific grant from funding agencies in the public, commercial, or not-for-profit sectors.

Disclosure statement

No potential conflict of interest was reported by the authors.

Notes on contributors

Suresh Niraula is a doctoral candidate and pursuing his degree in Environmental and Conservation Science at the North Dakota State University. His research focus is on the manure management, especially gaseous emissions from the land application of solid beef manure and manure with bedding. Suresh joined NDSU in Fall of 2015 and plans to graduate in the Spring of 2018.

Dr Shafiqur Rahman, Associate Professor, Agricultural and Biosystems Engineering, North Dakota State University. He received his PhD from University of Manitoba, Winnipeg, Canada, in 2004. Dr. Rahman's research focuses on natural resource management including animal nutrient management (best management practices, composting, and anaerobic digestion), monitoring runoff water quality from livestock production facilities, measurement & mitigation of odor, pollutant gases & GHG from livestock production facilities.

Dr Amitava Chatterjee is Assistant Professor of Soil Science at North Dakota State University. He earned his doctoral degree from University of Wyoming in 2007. His research interest is focused on crop production and nutrient dynamics. He teaches undergraduate and graduate level courses in soil fertility.