Gaseous losses of nitrogen (N), nitrous oxide (N₂O) denitrification and ammonia (NH₃) volatilization, reduce fertilizer-N use efficiency and may cause environmental degradation [1]. Global estimates suggest approximate N losses of 0.5% - 2% and 10% - 18% of initial N content through denitrification and volatilization respectively [2] [3]. Lack of studies regarding the temperature sensitivity of gaseous losses of N makes it difficult to model how changing spatial variability of crop, soil, and water management practices will impact the environment [4].

Soil temperature has significant control over N mineralization [5], denitrification [6], and volatilization [7]. Temperature sensitivity of the biological processes is generally expressed as the function of the increase in metabolic rate with 10˚C rise in temperature or Q₁₀. For most modeling approaches, Q₁₀ value was assumed to be close to 2, irrespective of soil type, climate and management practices [6] [8]. However, researchers reported a wide range of Q₁₀ values ranging from 1 to 17.1 for denitrification [9], 1.4 to 5.0 for volatilization [10], and 1.67 to 2.43 for soil N mineralization [11].

A laboratory incubation study was conducted to determine the Q₁₀ value of N₂O and NH₃ flux, volatilization and N mineralization for eight soil samples collected across agricultural systems of the United States. If Q₁₀ value is not affected by climate, soil type, or cropping system, measurements of Q₁₀ will be equal to 2 regardless of soil evaluated. To test this hypothesis, we measured cumulative N₂O and NH₃ flux and net N mineralization with incubation temperature, and temperature sensitivity or Q₁₀ of N₂O and NH₃ flux and net N mineralization at 10˚C, 20˚C, and 30˚C. We then calculated the temperature sensitivity, or Q₁₀ of N₂O and NH₃ flux and net N mineralization for these agricultural soils.

Surface soil samples of 0 - 15 cm depth were collected from eight agricultural fields across the United States (Figure 1, Table 1). Soil samples were air-dried

*Inorganic nitrogen-ammonium (NH+ 4) and nitrate (NO− 3) concentrations; ^†Annual average.

and grounded to pass through 2 mm sieve. Soil pH and electrical conductivity were measured of 1:2.5 soil slurry with Oakton PC700 pH and EC meter. Soil organic carbon and total N were determined by automated dry combustion method [12]. Soil inorganic N concentration was measured by extracting soils with 2 M KCl and determining NH+ 4 and NO− 3 concentrations using Timberline Ammonia (TL-2800) analyzer (Boulder, CO). Field capacity (at 0.33 bar) was determined using the pressure plate apparatus as described by [13].

Soil samples were incubated at 10˚C, 20˚C and 30˚C using an incubation chamber. For incubation, 30 g soils moistened at field capacity level were placed in a 1-L clear jar (Table 1). One granule of urea (~40 mg) was weighed and added on the soil surface. Water loss was compensated by adding water based on the difference in jar weight. The cap of jar was fitted with the gas sampling port (butyl rubber septum) to sample headspace air and a metal wire attached to the cap to hold a 50 mL clear plastic beaker filled with 15 mL of 0.5 M phosphoric acid to trap NH₃ emission from soils. A total of 36 jars (8 sites × 4 replication + 4 blanks) were incubated for 91 days at each temperature. Headspace air was sampled approximately on days 1, 3, 6, 9, 12, 16, 19, 23, 26, 30, 34, 37, 40, 44, 47, 50, 56, 63 70, 77, 84, and 91 for N₂O flux and until day 47 for NH₃ flux. On each observation day, first headspace air sample was collected using a 10 mL syringe, followed by the removal of the acid trap, then the jar was aerated for half an hour, and soil moisture was readjusted to field capacity and then jars were capped and returned to the incubator.

The N₂O concentration of headspace air samples was determined using a Shimadzu GC-2014 (Shimadzu Scientific Instruments Inc., Houston, TX) fitted with 63Ni-electron capture detector. The GC oven was operated at 80˚C and ECD was operated at 325˚C, and N₂ carrier gas was supplied at 20 PSI. Instrument was calibrated using analytical N₂O standards of: 0, 1, 5, 50, 100, 500, and 1000 µmole∙ml⁻¹. Compound peak was recorded and analyzed with Lab Solutions software (LabSolutions, Atlanta, Georgia). The N₂O concentration was converted to mass unit using ideal gas equations and expresses as micrograms of N₂O produced between sampling days per kg of soil [14] [15]. Soil-emitted NH₃ was trapped and replaced with fresh phosphoric acid solution at the same intervals as N₂O flux measured. The collected acid solution was extracted with 25 mL of 2 M KCl with half an hour shaking the mixture in reciprocal shaker [14]. The extracts were then analyzed for NH+ 4 concentrations using an automated ammonia analyzer (TL 2800, Timberline Instruments, Boulder, CO). The amount of volatilization during each incubation interval was expressed in the form of microgram NH₃ per gram soil. Cumulative NH₃-N loss (mg NH₃-N kg soil) during the entire incubation was computed from the summation of NH₃ emission during all sampling periods.

After 91 days of incubation, soil samples from each jar were analyzed for inorganic N concentration (NH+ 4 and NO− 3) for each incubation temperature. Percent of net N mineralized during incubation was calculated using the following equation.

Net   N   mineralization % = ( Initialsoil-N + Urea-N ) − Final   soil-N Initialsoil-N + urea-N × 1 00

Urea-N was calculated by multiplying 0.46 with the weight of urea granule. The effect of temperature on of N₂O and NH₃ flux, and net N mineralization% was evaluated by determination of the parameter E_a in the logarithmic form of the Arrhenius equation:

ln ( k ) = ln ( A ) − E a / R T

where k is the rate of N₂O and NH₃ flux, A is the preexponential constant, E_a is the activation energy (kJ∙mol⁻¹), R is the gas constant (8.314 J∙mol⁻¹∙K⁻¹) and T is the absolute temperature in Kelvin (K). The activation energy was calculated from the slope (−E_a/R) of the linear regression in the plot of log of N₂O and NH₃ flux rate vs. the inverse incubation temperature, Q₁₀ value was calculated as

Q 10 = exp [ ( E a / R ) ( 1 / ( T 1 + 10 ) − 1 / T 1 ) ]

T₁ = 293˚C equivalent to 20˚C

Cumulative N₂O and NH₃ flux at each incubation temperature, net N mineralization percentage and Q₁₀ values were compared for different sites using the completely randomized design (CRD) with a mean separation at 95% significance level using SAS 9.4. For each site, incubation temperature effect on cumulative N₂O and NH₃ flux were also determined using CRD with a mean separation at 95% significance level. Correlation coefficient and regression analyses were conducted to determine the relationship between soil properties and Q₁₀ values using SAS 9.4.

Cumulative N₂O flux increased with temperature for most soils except those collected from Frenchville, Bismarck, and Pendleton (Table 2). At 10˚C, soils from Pendleton had the highest cumulative flux, but statistically similar to Frenchville and Bismarck; whereas, the lowest value was observed for soils from Blackville. At 20˚C, Pendleton soils had the highest cumulative N₂O flux, similar to Frenchville, and the lowest value was observed for soils from North Bend. At 30˚C, Frenchville had the highest cumulative flux, significantly higher than rest, and the lowest value was found for soils from Dickinson. Temperature sensitivity, Q₁₀ values of cumulative N₂O flux ranged between 0.23 at Bismarck, and 11.4 at Blackville. Soils from Jackson, TN had Q₁₀ value of 10.1, statistically similar to Blackville. The rest of the six sites had similar Q₁₀ values ranging between 0.23 - 2.14. Blackville and Jackson had lower soil organic C than other sites; low soil organic C or high recalcitrance of substrates should generally be more sensitive to temperature changes than that of more labile substrates, which could, in turn, increase the Q₁₀ value. Researchers have also found that additions of C and N substrates reduced Q₁₀ of N₂O due to increased soil microbial C and N use efficiency [15].

Increasing temperature reduced cumulative NH₃ flux except for Downer, and North Bend, sites (Table 2). Soils from Blackville had the highest and Frenchville, had the lowest cumulative NH₃ flux at all three temperatures. Soils from North Bend had the highest Q₁₀ value for NH₃, but similar to Downer, and Pendleton. For the rest of the sites, Q₁₀ value for NH₃ ranged from 0.63 to 0.70. Most researchers observed an increase in volatilization loss with temperature [7] [16]. Researchers [7] reported a two-fold increase when temperature increased from 5˚C to 25˚C but a threefold when temperature increased from 25˚C to 45˚C. They concluded that greatly enhanced NH₃ volatilization at 45˚C compared with 25˚C was related to the inhibition of nitrification at high temperature, which increased the supply of ammoniacal N for NH₃ volatilization for a prolonged time. Our maximum incubation temperature of (30˚C) was comparatively lower than the threshold for the inhibition of nitrification. Further, researcher [17] found that high temperatures (32˚C) increased the initial rates of NH₃-N loss and they were proportionally reduced at later stages; on the contrary, the lowest temperature (12˚C) resulted in the lowest initial NH₃-N loss rate but became highest for the last 76 hours.

*Different capital letters indicate significant differences among sites of the same incubation temperature and different small letters indicate significant differences among temperatures for the same site.

For all sites, net N mineralization was significantly lower at 20˚C than 10˚C, and 30˚C, this might be caused due to greater N immobilization at 20˚C. At all three temperatures, soils from Blackville had the highest, and Dickinson had the lowest N mineralization. Temperature sensitivity or Q₁₀ of net N mineralization varied from 0.96 to 1.00. Soils from Frenchville had the highest Q₁₀ and soils from Bismarck and Dickinson had the lowest Q₁₀. Other researchers found that Q₁₀ values of N mineralization varied from 1.03 to 11.89 with an average of 2.21 [11].

The Pearson relationship between soil organic C and total N showed a significant negative relationship with Q₁₀ value of N₂O (−0.82 and −0.72, respectively), but did not show any relationship with volatilization or N mineralization. Linear regression relationships showed that SOC and TN explained the 68 and 52 percent of the variation in Q₁₀ of N₂O. With the rise in each unit (g∙kg⁻¹) of SOC and total N, Q₁₀ value of N₂O declines by 0.67 and 6.0, respectively. Similarly, other researchers [15] also observed a significant inhibition of pulse N₂O emissions following C addition, they hypothesized that C addition facilitates the microbial growth and in turn accelerates N immobilization rate.

This study clearly indicates a wide variation in Q₁₀ for N₂O (0.23 to 11.4), and small variations in Q₁₀ for NH₃ (0.63 to 1.24) and for the net N mineralization (0.96 to 1.00). Distribution of soil organic C can explain the spatial variation of Q₁₀ for N₂O flux. Future research should explore the spatial variation in Q₁₀ for soils within sensitive regions.

The authors declare no conflicts of interest regarding the publication of this paper.

Chatterjee, A., De Jesus, A.F., Goyal, D., Sigdel, S., Cihacek, L.J., Farmaha, B.S., Jagadamma, S.., Sharma, L. and Long, D.S. (2020) Temperature Sensitivity of Nitrogen Dynamics of Agricultural Soils of the United States. Open Journal of Soil Science, 10, 298-305. https://doi.org/10.4236/ojss.2020.107016