Published June 24, 2014

Journal of Environmental Quality

TECHNICAL REPORTS Atmospheric Pollutants and Trace Gases

Soil Nitrous Oxide Emissions after Deposition of Dairy Cow Excreta in Eastern Canada Philippe Rochette,* Martin H. Chantigny, Noura Ziadi, Denis A. Angers, Gilles Bélanger, Édith Charbonneau, Doris Pellerin, Chang Liang, and Normand Bertrand

E

Urine and dung deposited by grazing dairy cows are a major source of nitrous oxide (N2O), a potent greenhouse gas that contributes to stratospheric ozone depletion. In this study, we quantified the emissions of N2O after deposition of dairy cow excreta onto two grassland sites with contrasting soil types in eastern Canada. Our objectives were to determine the impact of excreta type, urine-N rate, time of the year, and soil type on annual N2O emissions. Emissions were monitored on sandy loam and clay soils after spring, summer, and fall urine (5 and 10 g N patch-1) and dung (1.75 kg fresh weight dung-1) applications to perennial grasses in two successive years. The mean N2O emission factor (EF) for urine was 1.09% of applied N in the clay soil and 0.31% in the sandy loam soil, estimates much smaller than the default Intergovernmental Panel on Climate Change (IPCC) default value for total excreta N (2%). Despite variations in urine composition and in climatic conditions, these soil-specific EFs were similar for the two urine-N application rates. The time of the year when urine was applied had no impact on emissions from the sandy loam soil, but greater EFs were observed after summer (1.59%) than spring (1.14%) and fall (0.55%) applications in the clay soil. Dung deposition impact on N2O emission was smaller than that of urine, with a mean EF of 0.15% in the sandy loam soil and 0.08% in the clay soil. Our results suggest (i) that the IPCC default EF overestimates N2O emissions from grazing cattle excreta in eastern Canada by a factor of 4.3 and (ii) that a regionspecific inventory methodology should account for soil type and should use specific EFs for urine and dung.

xcreta deposited on pastures by grazing animals

are a source of nitrous oxide (N2O), a potent greenhouse gas that is involved in the destruction of stratospheric ozone (Ravishankara et al., 2009). In Canada, annual N2O emissions resulting from the deposition of urine and dung by farm animals on paddocks, ranges, and pastures are estimated at 8 Gg N2O–N or 11.5% of national agricultural N2O emissions (Rochette et al., 2008). Production of N2O in soils occurs during the nitrification and denitrification processes (Barnard et al., 2005). The N2O yield of denitrification is often higher than that of nitrification (Beauchamp, 1998), and denitrification has been identified as the main source of N2O in urine-treated sandy (de Klein and van Logtestijn, 1994) and clayey (Monaghan and Barraclough, 1993) soils. Amounts of N2O emitted from soils are generally proportional to N inputs but also depend on soil redox potential and organic C availability (Beauchamp, 1998). Accordingly, several factors affecting the soil redox potential, such as gaseous diffusion, water content, and compaction, have been shown to affect N2O emissions after urine application to grassland soils. For example, higher emissions have been reported from poorly than from well-drained soils (de Klein et al., 2003; Li and Kelliher, 2005; Di et al., 2007; Dixon et al., 2010), in compacted than in intact soils (van Groenigen et al., 2005a,b; Bhandral et al., 2007; Uchida et al., 2008), in finethan in coarse-textured soils (Di and Cameron, 2006; Clough et al., 1998), and in soils with higher water contents (Clough et al., 2004; Thomas et al., 2008). Climatic conditions at the time of application are major drivers of soil water content and aeration, and higher N2O emissions have been reported when urine was applied during wet than during dry periods (Allen et al., 1996; Yamulki et al., 1998; Anger et al., 2003; van Groenigen et al., 2005a; Luo et al., 2008; Zaman et al., 2009; Zaman and Nguyen, 2012) and in winter than in summer (Luo et al., 2008). However, the proportion of denitrified urine N that is emitted as N2O decreases at low redox potentials,

Copyright © American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America. 5585 Guilford Rd., Madison, WI 53711 USA. All rights reserved. No part of this periodical may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher.

P. Rochette, M.H. Chantigny, N. Ziadi, D.A. Angers, G. Bélanger, and N. Bertrand, Agriculture and Agri-Food Canada, Soils and Crops Research and Development Centre, 2560 Hochelaga Blvd, Québec, G1V 2J3; É. Charbonneau and D. Pellerin, Dep. of Animal Sciences, 2425 rue de l’Agriculture, Laval Univ., Québec, G1V 0A6; C. Liang, Environment Canada, 9th floor, 200 Sacré-Coeur, Gatineau, K1A 0H3. Assigned to Associated Editor Søren Petersen.

J. Environ. Qual. 43:829–841 (2014) doi:10.2134/jeq2013.11.0474 Received 27 Nov. 2013. *Corresponding author ([email protected]).

Abbreviations: EF, emission factor; IEM, ion exchange membrane; IPCC, Intergovernmental Panel on Climate Change; WFPS, water-filled pore space.

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and smaller N2O emissions have been reported in heavily compacted (Šimek et al., 2006) or water-saturated soils (Allen et al., 1996). Accordingly, the proportion of urine N lost as N2O was not always clearly related to soil types (Di et al., 2007; Singh et al., 2008; van der Weerden et al., 2011) or time of the year (Šimek et al., 2006; Kelley et al., 2008). Effects of time of the year and soil type may therefore vary depending on local climatic conditions, management practices, and soil properties. The impact of environmental conditions on N2O emissions after deposition of excreta by grazing animals is reflected in the large variability in N2O emission factor (EF) estimates that vary between 0.1 and 4% for urine N and between 0 and 0.7% for dung N (Oenema et al., 1997; de Klein et al., 2001; de Klein et al., 2003; van Groenigen et al., 2005b; van der Weerden et al., 2011). The default EF proposed for national inventories of N2O emission for total excreta N (urine + dung) by grazing dairy cattle is 2% (EF3) (IPCC, 2006), based on a summary of literature data (de Klein et al., 2010). However, nearly all studies reporting these soil N2O flux measurements were performed in Western Europe, the United Kingdom, and New Zealand, and it is unclear how their results apply to other regions. Therefore, the use of the Intergovernmental Panel on Climate Change (IPCC) default EF is a source of uncertainty for the Canadian national inventory of greenhouse gas emissions, and in-country measurements are needed to help determine how this EF applies to Canadian conditions. The objectives of this study were (i) to quantify emissions of N2O after deposition of dairy cow urine and dung onto two grassland sites with contrasting soil types in eastern Canada and (ii) to determine the impact of excreta type, urine-N rate, time of the year, and soil type on annual emissions.

Materials and Methods The study was conducted from September 2009 to October 2011 on two sites located at Agriculture and Agri-Food Canada experimental farms near Québec City, Canada (46°47¢ N, 71°08¢ W). The soil at the Harlaka site was a Kamouraska clay (fine, mixed, frigid, Typic Humaquept; texture: 0.58 g clay g-1, 0.08 g sand g-1; bulk density: 1.05 g cm-3; pH: 5.7; cation exchange capacity: 46 meq 100 g-1), and the soil at the Chapais site was a Saint-Pacôme sandy loam (loamy, mixed, frigid Umbric Dystrochrept; texture: 0.19 g clay g-1, 0.70 g sand g-1; bulk density: 1.27 g cm-3; pH: 5.2; cation exchange capacity: 20 meq 100 g-1). The experimental areas on the two soil types were established 3 yr earlier with timothy (Phleum pratense L.) but

also had some nonseeded species. Grasses were cut to a 0.05-m height in late May, late June, late August, and late October to simulate periodic grazing. The same experiment was conducted twice on the two soil types, each time with application of excreta in spring, summer, and fall. The four excreta treatments were unamended control plots, circular (0.1 m2) dairy cow urine patches at 5 g N patch-1 (U5) and 10 g N patch-1 (U10), and circular (0.1 m2) dairy cow dungs at 1.75 kg fresh weight (FW) dung-1 (except in fall 2009, when 1.50 kg FW dung-1 was applied) (Table 1). For each application, urine and dung were deposited on recently cut grasses (0.05 m). The impact of excreta on soil N dynamics often exceeds the area covered by the patch (Saarijärvi and Virkajärvi, 2009). In this study, the size of the patches was selected to ensure that all patch-induced impact on soil N2O emissions would be accounted for within the chamber area (0.3025 m2). The U10 treatment was achieved by applying the required volume of fresh urine (0.9–1.4 L patch-1), and the U5 treatment consisted of the same volume of diluted urine (1:1, urine:water ratio). A split-plot experimental design was used with time of application as main plots and excreta treatments as subplots with four replicates. Times of urine and dung applications for the two runs of the experiment were spring (31 May 2010 and 6 June 2011), summer (5 July 2010 and 4 July 2011), and fall (28 Sept. 2009 and 27 Sept. 2010). Subplots were separated by 1-m-wide buffer strips, and new subplots were used for each application to avoid cumulative effects of applications. Each subplot was divided into three subsubplots (1 m × 1 m). One sub-subplot was used for gas flux measurements, and soil temperature, mineral N content, and water content were monitored on the other two. One patch of either urine or dung was deposited in the center of the chamber base for gas flux measurements (see details of chamber below), and two patches were deposited in each of the other two subsubplots for monitoring of soil mineral N.

Collection of Urine and Dung For each time of excreta application, six to eight Holstein dairy cows (milk yield: 22.3 ± 6.7 kg d−1; body weight: 678 ± 95.5 kg; days in milk: 261 ± 54.2 d [mean ± SD]) kept in a tiestall barn were fed ad libitum fresh forages with three meals of concentrate feeds. Diets were formulated to be similar to a typical pasture diet (net energy of lactation: 1.57 ± 0.08 Mcal kg−1 DM; crude protein: 18.9 ± 0.97% DM; % rumen degradable protein requirement met: 134 ± 7.7% [mean ± SD]) and were fed for at

Table 1. Nitrogen content and application rate of the dairy cattle urine and dung for the spring, summer, and fall applications. Urine† Time of application Fall 2009 Spring 2010 Summer 2010 Fall 2010 Spring 2011 Summer 2011

U5

Dung‡ U10

pH

g N L-1

g N patch-1

pH

g N L-1

g N patch-1

8.2 8.2 8.2 8.5 8.4 8.0

3.62 3.91 5.25 3.09 4.15 5.59

5.07 4.89 4.73 4.33 5.19 4.92

8.2 8.4 8.3 8.1 9.0 8.0

6.86 7.75 10.4 7.56 7.45 9.48

9.61 9.69 9.36 10.58 9.31 8.34

pH g N kg FW-1 7.1 6.0 6.8 7.3 7.1 7.2

4.42 3.76 3.55 4.05 3.55 3.93

mg NH4–N kg FW-1

mg NO3–N kg FW-1

% DM

g N dung-1

183 204 66 313 201 233

0.69 2.33 2.36 3.31 1.96 1.60

14.5 14.5 17.0 14.1 10.5 11.8

6.63 6.58 6.21 7.09 6.21 6.88

† U5, urine at a target rate of 5 g N patch-1; U10, urine at a target rate of 10 g N patch-1. The area of urine and dung patches was 0.1 m2. ‡ Dung: feces at 1.75 kg fresh weight (FW) dung-1 (except in fall 2009 when 1.50 kg FW dung-1 was applied; target N rate of 6.5 g N dung-1). 830

Journal of Environmental Quality

least 14 d. Fresh forage was cut from a field containing pasture forage species (Trifolium repens L., Bromus riparius Rehmanns, and Phleum pratense L.) once a day at 0830 h and fed twice a day (0900 and 1400 h) to the cows. Concentrate meals were fed to meet cow nutritional requirements according to NRC (2001). Catheters (24 mm, 75-mL balloon; Bardex) were installed into the bladder of each cow on the Day 12 after initiation of the diet. At 0900 h on Days 13 and 14, catheters were connected to a container using polyvinyl chloride tubing. Urine was collected from the containers at 1200, 1500, and 1800 h on Days 13 and 14. Light mineral oil (90 mL) was added in containers after each collection to prevent contact with air, and containers were stored on ice to limit urine deterioration. Urine from all cows was mixed, stored by day in 50-L containers, and refrigerated as soon as possible after each collection time. On Day 14, urine was transferred to one container without dilution (50:50, Day 13 urine/Day d14 urine) and to one container with dilution (50:25:25, distilled water/Day 13 urine/Day 14 urine). Fecal excretion was collected using the same schedule as for urine. Feces of all cows were pooled by collection time in a plastic container, mixed, and weighed before being refrigerated. Feces in the plastic containers were covered with a polyvinyl plastic to prevent contact with air during refrigeration. On Day 14, feces were pooled proportionally to the fecal excretion at each collection time to assure uniformity between dungs applied on each subplot. When excretion of feces exceeded the needs, the feces from the earlier collection times were discarded. Sample bags of dung were refrigerated as soon as the pools were made.

Soil Mineral N Monitoring Soil mineral N (nitrate [NO3-] and ammonium [NH4+]) fluxes were measured using ion exchange membranes (IEMs) of 6.0 × 2.5 cm in size (Ziadi et al., 1999). The membranes were washed with distilled water to remove impurities, shaken 30 min with 0.5 mol L‒1 HCl, and rinsed three or four times with distilled water. The IEMs were then agitated in 1 mol L‒1 NaCl solution for 1 h on a lateral shaker (120 oscillations min-1); membranes with anion exchange sites to trap NO3- (Type AR 204, Ionics, Durpro) were saturated with Cl-, whereas membranes with cation exchange sites to trap NH4+ (Type CR 67, Ionics, Durpro) were saturated with Na+. Two IEMs (one of each type) were inserted (6–8 cm depth) underneath each excreta patch (except in sub-subplots where soil N2O emissions were measured) using a small trowel to open a narrow slot that minimized soil disturbance. The slot was closed by hand to ensure good contact between the soil and the IEM surface. The IEMs were deployed and replaced with new ones every 3 to 7 d for the first 3 wk after excreta deposition and at 7- to 21-d intervals thereafter. The monitoring periods started on the day excreta were deposited and ended when soil froze in the fall. Therefore, monitoring periods varied from 43 to 54 d after fall applications from 148 to 177 d after spring applications. Collected IEMs were washed with distilled water in the field to eliminate attached soil particles and placed in individual tubes containing 25 mL of 1 mol L‒1 NaCl to extract adsorbed NH4+ or NO3-. The concentrations of NO3- and NH4+ in the extracts were determined using an automated colorimeter and Lachat

methods 12-107-04-1B and 12-107-06-2A, respectively (Model QuickChem 8000 FIA+, Lachat Instruments). Cumulative soil NH4+ and NO3- fluxes measured by IEMs were reported in mg N cm-2 by dividing the individual NH4+ and NO3- concentrations by the IEM surface area (15 cm2) and summing the values over the monitoring period.

Gas Flux Measurements Soil-surface N2O fluxes were measured from September 2009 to November 2011 during snow-free periods using non–flow-through, non–steady-state chambers (Rochette and Bertrand, 2008). One acrylic frame (0.55 m × 0.55 m × 0.14 m height) was inserted to a depth of 0.1 m in every experimental plot and left in place for the next 12 mo. After application of excreta (one circular [0.1 m2] patch in the center of area covered by the frame), measurements were made twice a week during the first 2 wk, once a week during the following month, and every second week until the onset of snow cover. Measurements usually covered for a 12-mo period after applications. At each sampling date, a vented and insulated acrylic chamber (0.2 m in height) was tightly fitted to the frame. Air samples (20 mL) were collected from the chamber headspace at 0, 8, 16, and 24 min after chamber deployment through a rubber septum, using a syringe, and transferred immediately to 12-mL preevacuated glass vials (Exetainers, Labco). Time of sampling was between 0900 and 1200 h. Within 10 d of collection, the gas samples were analyzed for N2O concentration using a gas chromatograph (Model 3800 or 3600, Varian Inc.) equipped with an electron capture detector. Soil N2O fluxes were calculated using equations proposed by Rochette and Bertrand (2008) with an estimated minimum measureable flux of 0.025 mg N2O  m-2  h-1. Cumulative N2O–N losses were obtained by linearly interpolating emission rates between sampling dates during the approximately 12-mo periods after excreta application. Nitrous oxide EFs were calculated as the cumulative N2O–N emitted in treated plots corrected for emissions from the control plots and expressed per unit of urine or dung N deposited. Rainfall amounts and air and soil (5 cm) temperatures were recorded using automated weather stations at the experimental sites (Table 2). Soil water-filled pore space (WFPS) (0–10 cm) was calculated using soil bulk density, and volumetric water content (TDR probes) was measured on the same sampling days as the gas fluxes.

Data Analyses The effects of excreta treatments and time of application on cumulative N2O emissions and EFs were tested using the MIXED procedure of SAS with the restricted maximum likelihood method (SAS Institute, 1999). The model included time of application and excreta treatments as fixed effects, and runs of the experiment were considered as random effects. Mean values were compared using LSMEANS with a Tukey correction for multiple comparisons at p < 0.05. In an analysis that included site as a fixed factor, large differences in treatment effects between sites were noted, with a highly significant (p < 0.001) interaction between site and treatment. Consequently, the statistical analysis was performed for each site separately.

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Table 2. Monthly air and soil temperatures, rainfall, and soil water content during the experiment conducted at two sites with contrasting soil types. Year/month

Tair†

Sandy loam soil Tsoil Hsoil

———— °C ———— 2009  Oct.  Nov. 2010  May  June  July  Aug.  Sept.  Oct. 2011  May  June  July  Aug.  Sept.  Oct.

Clay soil Rain

Tair

Tsoil

%

mm

———— °C ————

Hsoil

Rain

%

mm

4.6 2.5

–‡ –

34.7 28.0

112 71

4.3 1.7

6.0 2.6

45.2 36.1

100 78

13.5 17.4 21.9 19.4 13.9 7.0

12.8 18.1 21.9 19.2 15.1 7.9

14.6 24.9 16.5 11.6 25.7 30.9

66 81 47 89 203 86

12.7 16.9 21.1 18.4 13.3 6.2

13.6 17.1 21.4 18.9 15.0 7.6

22.2 36.6 32 25.6 41.5 43.9

65 122 39 129 195 93

11.7 16.8 20.8 18.7 15.4 9.7

11.6 17.9 22.1 20.0 16.9 11.1

23.0 19.6 20.9 24.6 25.7 25.9

127 82 134 151 131 101

10.4 16.3 20.2 18.4 15 7.4

11.3 17.5 21.3 19.2 17.2 12.2

30.4 31.7 33.9 34.1 35.4 40.0

108 124 175 125 131 101

† Hsoil, volumetric soil water content (0–10 cm); Tair, air temperature; Tsoil, soil temperature (5 cm). ‡ Missing data.

Results and Discussion Temporal Pattern of N2O Emissions

Emissions were near zero from unamended control plots on both soil types at all excreta application times (Fig. 1 and 2). On treated plots, the temporal pattern of emissions varied among application times, with maximum fluxes observed between 1 and 22 d after urine application. The duration of the period during which emissions were increased compared with the control plots (estimated as the time to cumulate 90% of total emissions) was longer in the sandy loam (27–103 d) than in the clay soil (10–61 d). Our values compare well with durations of the urine-induced emissions from 30 d (Flessa et al., 1996; Bhandral et al., 2007) to 60 d (Clough et al., 2004), but values as high as 120 d (Clough et al., 1998), 180 d (Di et al., 2010), and 365 d (de Klein et al., 2003) have been reported. In this study, emissions that occurred more than 30 d after urine and dung applications were not significantly greater than zero, and their omission would have resulted in a mean underestimation of 12% in EF estimates (Tables 3 and 4). When averaged over all application times, treatmentinduced emission rates were twice greater during the first 10 d than between 11 and 30 d after application, thereby resulting in similar overall means of cumulative emissions in both intervals for all treatments (Tables 3 and 4). Greater emission rates shortly after excreta application are in line with the rapidly decreasing soil N concentration with time in response to plant uptake, microbial transformations, adsorption on soil particles, and volatilization (Deene and Middlekoop, 1992; Wachendorf et al., 2008). However, the relative contribution of the first two intervals (0–10 d and 11–30 d) varied greatly among application times in response to variations in climatic conditions. For example, emissions after 10 d on the clay soil were very low from urine-treated plots during a dry spell 832

in summer 2010 and were highest in summer 2011 after a 35-mm rainfall (Fig. 1; Table 2). The temporal distribution of treatment-induced emissions was similar in the two soils, indicating that, although soil type influenced the magnitude of emissions, it had little impact on their temporal patterns, which were governed by climatic conditions.

Cumulative N2O Emissions

Mean cumulative emissions from the control plots during the whole measurement periods were 0.29 and 0.30 kg N ha-1 at the clay and sandy loam sites, respectively, with values ranging from 0.12 to 0.46 kg N ha-1 among application times (data not shown). These values are lower than the mean background values observed on agricultural soils (1.0 kg N ha-1) (Bouwman, 1996; Rochette et al., 2008) but are in agreement with previous reports of similar low emissions from unfertilized perennial grasses in eastern Canada (Rochette et al., 2004). The tight N cycle of perennial crops systems (Randall et al., 1997) probably limited the availability of soil mineral N for nitrification and denitrification, thereby limiting the potential for N2O production and emission.

Urine At the clay site, treatment-induced cumulative emissions during the whole measurement periods were greater (p < 0.05) for the urine treatment than for the dung treatment at all application times (Table 4), in agreement with previous reports of the impact of excreta deposition on N2O emissions from pastures (de Klein et al., 2003; van der Weerden et al., 2011). At the sandy loam site, annual treatment-induced cumulative emissions from the urine-treated plots were on average 3.3 and 4.0 times smaller than at the clay site for the U10 and U5 treatments, respectively (Table 3). In contrast with the clay soil, emissions from the U5 plots were similar to the dung plots at the spring and summer application times. Journal of Environmental Quality

Fig. 1. Rainfall and soil N2O fluxes after the deposition of dairy cattle urine and dung on a clay soil cropped to perennial grasses in the fall of 2009 and 2010 and in the spring and summer of 2010 and 2011 at different target N rates (Control, unamended; Dung, 6.5 g N dung−1; U5, 5 g urine N patch−1; U10, 10 g urine N patch−1). Error bars are standard deviations (n = 4). Urine and dung patches covered one third of chamber area.

This was attributed to larger emissions from dung plots and lower emissions from U5 plots in the sandy loam soil as compared with the clay soil.

Despite differences in magnitude between soil types and time of year, cumulative emissions during the whole measurement periods increased linearly with the urine-N rate on both soils

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Fig. 2. Rainfall and soil N2O fluxes after the deposition of dairy cattle urine and dung on a sandy loam soil cropped to perennial grasses in the fall of 2009 and 2010 and in the spring and summer of 2010 and 2011 at different target N rates (Control, unamended; Dung, 6.5 g N dung−1; U5, 5 g urine N patch−1; U10, 10 g urine N patch−1). Error bars are standard deviations (n = 4). Urine and dung patches covered one third of chamber area.

(Tables 3 and 4; Fig. 3). There are few reports of the response of N2O emissions to urine-N rate. Linear relationships have been previously reported for rates up to 74.6 g N m-2 patch in the Netherlands (van Groenigen et al., 2005b) and up to 60 g 834

N m-2 patch in New Zealand (Dai et al., 2013), but similar emissions were also reported at 10, 25, and 100 g N m-2 patch under laboratory conditions (Clough et al., 2003). The increase in N2O emissions with increasing urine-N rate indicates that Journal of Environmental Quality

soil mineral N availability was limiting N2O production during the experiment, and the linear response suggests that there was no disproportionate increase in soil N excess when urine N was increased from 50 to 100 g N m-2 patch. This is in agreement with the high capacity of perennial grasses to value N inputs (Randall et al., 1997).

Dung Compared with the high urine application rate (U10), the mean increases in N2O emissions resulting from dung application were 3 times smaller in the sandy loam soil and 21 times smaller in the clay soil (Tables 3 and 4). This result is in agreement with most previous reports (Yamulki et al., 1998; van Groenigen et al., 2005b) and can be explained by the lower amounts of N in dung and by the slower rate at which mineral N is made available during the mineralization of dung organic-N compounds (e.g., proteins) as compared with the rapid release of NH4–N after hydrolysis of urea-N in urine.

The influence of soil type on N2O emissions after dung application was opposite to urine treatments, with values twice greater on the sandy loam than on the clay soil (Tables 3 and 4). Allen et al. (1996) also observed greater emissions after deposition of cattle dung on a well-drained than on a moderately drained soil in the United Kingdom. In the present study, it is unclear why dung would induce greater N2O emissions when deposited on a better-aerated soil. Explaining this observation would probably require a detailed study of the C and N transformations not only within the dung itself but also in the underlying soil because substrates may leach out of the dung and influence soil N2O production at greater depths. For example, we observed that the dung disappeared from the soil surface approximately 30% faster when deposited on the sandy loam than on the clay soil, suggesting a more rapid mineralization of organic N with associated greater potential for N2O production. The impact of dung on soil N2O emissions has received less attention than urine because of its lower

Table 3. Treatment-induced cumulative N2O emissions on the sandy loam soil during intervals after application of dairy cow urine and dung to grassland in spring, summer, and fall at different target N rates. Time of application

Spring 2010

Spring 2011

Spring mean

Summer 2010

Summer 2011

Summer mean

Fall 2009

Fall 2010

Fall mean

Overall mean

Treatment†

U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung

0–10 d

N2O emissions 11–30 d >30 d

Total

———————————— mg N patch−1‡ ———————————— 18.6 8.4 7.8 35.3 10.5 2.5 9.3 23.0 2.8 3.7 5.0 11.7 16.1 8.1 23.6 -1.2 9.3 4.7 8.2 -5.9 6.2 10.2 13.9 -2.5 17.4a§ 8.2ab 3.3*¶ 29.5a 9.9b 3.6b 1.7* 15.6b 4.5c* 7.0a 1.2* 12.8b 11.5 11.2 3.1 26.1 3.1 1.6 0.3 5.0 2.5 1.9 2.8 -1.2 21.1 22.6 43.5 -0.9 4.3 2.2 0.3 6.9 7.4 3.1 0.6 11.3 16.3a 16.9a 1.1* 34.8a 3.7b 1.9b 0.3* 6.0b 5.0b 2.5b 7.1b -0.3* 13.3 14.9 15.2 43.8 7.1 7.1 8.7 23.3 1.6 5.9 10.5 18.3 5.6 8.1 0.9 14.8 3.4 4.0 4.7 12.0 0.9 2.2 2.5 -0.6 9.5a 11.5a 8.1* 29.3a 5.3b 5.6b 6.7* 17.6b 1.2 b* 4.0b 5.0* 10.4b 14.4a 12.2a 4.1* 31.2a 6.3b 3.7b 2.9* 13.1b 3.6c 4.5b 2.0* 10.1b

Emission factor % applied N 0.36 0.47 0.18 0.25 0.16 0.22 0.31a 0.31a 0.20a 0.28 0.11 0.05 0.52 0.14 0.16 0.40a 0.12b 0.11b 0.46 0.46 0.28 0.14 0.28 0.04 0.30ab 0.37a 0.16b 0.34a 0.27a 0.15b

* Significant at the 0.05 probability level. † Dung, 6.5 g N dung-1; U5, 5 g urine N patch-1; U10, 10 g urine N patch-1. ‡ Cumulate emissions were corrected for emissions on unamended control plots. § Within a column for a given time of application, means with different letters differ at p £ 0.05. ¶ Values are not significantly different from zero (p > 0.05). www.agronomy.org • www.crops.org • www.soils.org 835

mineral N content. However, the reports of greater emissions for dung than for urine (Wachendorf et al., 2008; Virkajärvi et al., 2010) and reports of interactions between emissions and soil type (Allen et al., 1996; this study) may justify further investigation. Our results also agree with previous reports of a greater stimulation of N2O emission by solid organic amendments in coarse- than in fine-textured soils (Chantigny et al., 2010; Pelster et al., 2012).

Excreta Application Time Cumulative emissions from dung-treated plots were similar at all application times. In urine-treated plots, the application time had little impact on the magnitude of the emissions from the sandy loam (Table 3). In contrast, treatment-induced emissions on the clay soil were 3 times greater in summer than in fall, with intermediate values in the spring for both urine application rates (Table 4; Fig. 3). In regions with mild winters, most studies reported increases in emissions in cooler/wetter compared with

warmer/drier periods of the year. In New Zealand, Luo et al. (2008) reported 0.72 and 1.59% of applied N lost in winter but nearly none in summer. Similarly, fall emissions were 3.8 (Zaman and Nguyen, 2012), 1.4 (Zaman et al., 2013), and 1.5 times (Di and Cameron, 2003) greater than in spring and twice greater than in summer (Zaman et al., 2009). Similar observations were reported in western Europe, where higher emissions were reported in fall (Yamulki et al., 1998; van Groenigen et al., 2005b) and winter (Allen et al., 1996) than in summer. Most of these differences between seasons were attributed to higher soil water contents in cooler periods of the year, which resulted in greater N2O production by denitrification (Zaman et al., 2009; Luo et al., 2008; Zaman and Nguyen, 2012; Luo et al., 2000). This is supported by small differences in emissions between spring and summer applications when grasses were irrigated (Kelly et al., 2008). Similarly, Virkajärvi et al. (2010) reported small differences between June and August applications of urine

Table 4. Treatment-induced cumulative N2O emissions on the clay soil during intervals after application of dairy cow urine and dung to grassland in spring, summer, and fall at different target N rates. Time of application

Treatment†

0–10 d

N2O emissions 11–30 d >30 d

Spring 2010

Spring 2011

Spring mean

Summer 2010

Summer 2011

Summer mean

Fall 2009

Fall 2010

Fall mean

Overall mean

U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung U10 U5 dung

Total

———————————— mg N patch ‡ ———————————— 42.5 22.9 66.2 -0.3 14.9 6.2 17.0 -4.7 1.2 0.6 -3.7 -1.9 63.9 51.8 19.5 137.3 77.8 14.6 91.4 -2.5 5.3 6.8 5.4 -6.8 53.2a§ 37.4a 9.6*¶ 101.7a 46.3a 10.4a 54.2b -3.6* 3.3 b 3.7b* 1.7c -5.3* 73.2 9.9 15.5 100.2 25.7 0.6 23.9 -2.5 5.9 1.6 1.2 8.5 42.5 146.3 189.9 -1.6 35.0 91.8 0.0 128.8 6.8 1.2 7.2 -0.9 57.8a 78.1a 7.0* 145.1a 30.4b 46.2a 76.4b -1.2* 6.4c 1.4b* 0.2* 7.9c 10.9 18.6 41.5 72.5 9.9 7.8 5.9 23.9 0.9 2.8 3.5 -0.3 10.9 47.4 54.5 -5.0 10.2 22.9 26.1 -7.4 0.9 2.2 3.7 6.9 10.9a 33.0a 18.3* 63.5a 10.1a 15.3a 25.0b -0.8* 0.3 b* 1.6b* 3.3* 5.2c 40.6a 49.5a 11.6* 103.4a 28.9a 24.0b 51.9b -1.9* 3.3b 2.2b* 4.9c -0.6* −1

Emission factor % applied N 0.73 0.44 0.04 1.53 1.86 0.17 1.13a 1.15a 0.11c 1.02 0.41 0.06 2.28 2.63 0.11 1.65a 1.52a 0.09c 0.75 0.46 0.04 0.48 0.51 0.04 0.61a 0.48a 0.04c 1.13a 1.05a 0.08c

* Significant at the 0.05 probability level. † Dung, 6.5 g N dung-1; U5, 5 g urine N patch-1; U10, 10 g urine N patch-1. ‡ Cumulate emissions were corrected for emissions on unamended control plots. § Within a column for a given time of application, means with different letters differ at p £ 0.05. ¶ Values are not significantly different from zero (p > 0.05). 836

Journal of Environmental Quality

and dung to a medium-textured soil under the cool Northern Europe conditions. Our results suggest that N2O production in the sandy loam was limited by relatively good soil aeration conditions that reduced denitrification activity. In this context, the impact of the time of application was small because seasonal climatic variations only occasionally increased soil WFPS above levels at which denitrification is favored (Fig. 4). In the clay soil, where soil water content was greater, it appears that the higher summer soil temperature was the main driving factor (Table 2).

Soil Type Greater emissions from the clay than from the sandy loam soil are in agreement with previous reports of greater emissions from fine-textured than coarse-textured, urine-treated soils (Clough et al., 1998; Di and Cameron, 2006). The mineral N transformations that produce N2O in agricultural soils are largely controlled by the availability of N substrates and by the aeration status of the soil. Accordingly, we explored how soil type influenced these two factors in the present study. Cumulative N2O emissions from all experimental treatments increased linearly with cumulative mineral N (NH4–N + NO3–N) extracted from the IEMs (Fig. 5). This indicates that as more soil mineral N was available, the fraction emitted as N2O remained

Fig. 3. Relationship between cumulative soil N2O emissions and urine-N rate after the deposition of dairy cattle urine on sandy loam (open symbols) and clay (solid symbols) soils in the fall of 2009 and 2010 and in the spring and summer of 2010 and 2011. Urine patches covered one third of chamber area. Each value is the average of four replicates.

constant, in agreement with the linear relationship between N2O emissions and urine-N rates discussed previously. Those relationships were mostly influenced by values from the urine-

Fig. 4. Soil water-filled pore space (0–10 cm) and soil temperature (5 cm) at the (a) clay and (b) sandy loam sites from September 2009 to November 2011 (Control, unamended; Dung, 6.5 g N dung−1; U5, 5 g urine N patch−1; U10, 10 g urine N patch−1). www.agronomy.org • www.crops.org • www.soils.org 837

Emission Factors

Fig. 5. Relationship between cumulative soil N2O emissions and cumulative mineral N (NH4+ + NO3−) flux on soil ion exchange membranes after the deposition of dairy cattle urine and dung on sandy loam and clay soils. Values are presented for the four replicates of both runs of the experiment at the three times of excreta application. Urine and dung patches covered one third of chamber area.

treated plots as soil mineral N and N2O flux values for plots receiving dungs were generally small. Extending conclusions to situations where dung was deposited must therefore be made with caution. In fact, it is uncertain whether the N2O produced from dung N originated from N transformations that occurred in the soil under the dung or from the dung itself. Ion exchange membranes are increasingly used as an estimator of the availability of soil nutrients for plant roots (Ziadi et al., 1999). Our results suggest that they could also provide useful information regarding N availability for N2O production processes. However, the relationships were specific to each soil type as greater emissions were measured in the clay than in the sandy loam soil despite lower N availability (Fig. 5). This indicates that other factors, such as soil aeration status, influenced the response of N2O production and emission to soil mineral N availability. Denitrification has been identified as the main source of N2O in urine-treated soils (Monaghan and Barraclough, 1993; Orwin et al., 2010; Baily et al., 2012). Accordingly, increased emissions have been reported when WFPS of an excreta-treated silt loam soil was above 60% (Clough et al., 2004; Bhandral et al., 2007), a level above which denitrification is favored (Linn and Doran, 1984). In this study, the soil aeration status clearly favored denitrification in the clay soil where WFPS values during the 30 d after excreta applications exceeded 60% at four out of six application times (Fig. 4). In contrast, WFPS on the sandy loam exceeded the 60% threshold only in fall 2010, suggesting that the differences in soil N2O emissions between soil types were largely the result of differences in their water holding capacity, which directly influenced their aeration status. This is in agreement with reports of emissions 7 times higher in a silt loam above than below field capacity (Thomas et al., 2008) and up to 100 times greater in a silt loam incubated at 70 than at 30% WFPS (Orwin et al., 2010).

838

The impact of the excreta treatments on N2O EFs reflected differences between cumulative emissions discussed in the previous section. Mean estimates for dairy cattle urine were three times greater in the wetter clay (1.09%; from 0.41 to 2.63%) than in the drier sandy loam (0.31%; from 0.11 to 0.52%) soil (Tables 3 and 4). This is in agreement with previous reports of increasing EFs in urine-treated soils with decreasing soil aeration in response to differences in texture (Clough et al., 1998; Di and Cameron, 2006), compaction (Uchida et al., 2008; Thomas et al., 2008), drainage (de Klein et al., 2003; Li and Kelliher, 2005; Luo et al., 2008), or soil water content (Thomas et al., 2008). The time of the year for excreta deposition had no impact on the EF on the sandy loam soil, but mean EF values were greater in summer (1.59%) than in spring (1.14%) and fall (0.55%) in the clay soil (Tables 3 and 4; Fig. 3). There are several reports of seasonal variations in EF estimates, and most were related to differences in soil water content, with greater EFs occurring during the wetter periods of the year (Allen et al., 1996; Yamulki et al., 1998; Anger et al., 2003; van Groenigen et al., 2005b; Luo et al., 2008; Zaman et al., 2009; Zaman and Nguyen, 2012). In the present study, the greater EF estimates in summer were not related to higher WFPS values, and it is hypothesized that temperature and organic C availability were the main controlling factors under the generally humid conditions observed in this clay soil (Table 2). Accordingly, a greater impact of urine application on soil N2O emissions at higher temperatures was attributed to enhanced denitrification in response to increased amounts of plant-derived C (Uchida et al., 2011). More research is needed to elucidate the relative contributions of climatic conditions and plant factors to the seasonal variations in urine-induced N2O emissions in pastures grown on fine-textured soils. Nearly constant mean EFs were observed for both urine rates on the sandy loam (0.30%) and clay soil (1.10%) (Tables 3 and 4) as a result of the linear increase in urine-induced N2O emissions with increasing urine-N rate (Fig. 3). This has practical implications for EF-based national soil N2O emission inventories for two reasons. First, because the U10 treatment is representative of the high range of urine-N loading on pasture soils, our results suggest that one EF per soil type could be used for a wide range of urine deposition rates in eastern Canada. Second, the overlap of urine patches that increases urine-N load may not result in disproportionate increases in N2O emissions and associated biases when estimates are obtained using the same EF. The mean EF estimates for dung were one order of magnitude smaller than for urine (Tables 3 and 4). These EF values for dung are similar to those reported by Yamulki et al. (1998) (0.19%) in the United Kingdom, Maljanen et al. (2007) (0.28%) in Finland, and van der Weerden et al. (2011) (0–0.17%) in New Zealand but are smaller than those observed by Flessa et al. (1996) (0.5%) in Germany and Virkajärvi et al. (2010) (1.4%) in Finland. The application time did not affect the EF for dung. However, the impact of soil type on N2O emission after dung application was opposite to that of urine, with losses accounting for 0.08% of applied N in the clay soil and 0.15% in the sandy loam soil (Tables 3 and 4). Journal of Environmental Quality

For its national inventory of greenhouse gas emissions, Canada is currently using the IPCC default Tier I (IPCC, 2006) methodology for estimating N2O emission from the N excreted by grazing dairy cattle. The IPCC approach proposes that 2% of N in urine and feces is lost as N2O, an EF value that is greater than that observed in 22 of the 24 situations (time of application × soil × rate) that we studied (Tables 3 and 4). We used EFs derived in this study to calculate total emissions for eastern Canada based on the amounts of deposited N by grazing dairy cows (Environment Canada, 2013) and the fractions of total land under perennial forage crops classified as having coarse-, medium-, and finetextural soils (Agriculture and Agri-Food Canada, 2008) (Table 5). Assuming that the excreted N is partitioned 75% in urine and 25% in dung (Lantinga et al., 1987) and that EF for the medium soil textural class is intermediate between coarse and fine classes (urine: 0.70%; dung: 0.12%), total annual emissions in eastern Canada were estimated at 65.1 Mg N2O–N yr-1, a value 4.3 times smaller than values obtained using the default IPCC methodology (277.4 Mg N2O–N yr-1) (Table 5). The combination of the higher amounts of deposited N and of greater EFs resulted in emissions much higher for urine (93% of total) than for dung. Emission data from two sites during 2 yr may not justify a revision of the IPCC default EF. However, our results emphasize the need for additional field research for two reasons. First, the eastern Canada humid climatic conditions are conducive to relatively high emissions, especially from clay soils (Gregorich et al., 2005). That the EFs at the fine clay soil site were on average half the IPCC value (even at high urine-N rates) suggests that mean emissions in eastern Canada are probably not greater than those reported in the present study. Second, considering that the soil N2O emissions are on average one order of magnitude lower in the drier central Prairie region than in eastern Canada (Rochette et al., 2008), the overestimation of N2O emissions after deposition of dairy

cattle excreta by the IPCC default methodology is probably even greater for that region.

Summary and Conclusions Excreta deposited by grazing dairy cattle are a significant source of C and N substrates, which increase soil N2O production and emission. The excreta-induced emissions have been shown to be influenced by environmental conditions at the time of deposition. Urine application increased emissions for up to 103 d in the sandy loam and 61 d in the clay soil. The mean N2O EF for urine was greater in the clay soil (1.09% of applied N) than in the sandy loam soil (0.31%), most likely in response to a lower aeration status in the clay soil that favored denitrification. Despite variations in urine composition and in climatic conditions, these soil-specific EFs were similar for the two urine-N application rates. The time of the year when urine was applied had no impact on emissions from the sandy loam soil, but greater losses were observed after summer than spring and fall in the clay soil. The N2O emissions with the dung were smaller than those of urine, with an EF of 0.15% in the sandy loam and 0.08% in the clay soil. Inventory estimates of annual N2O emissions from excreta N deposited by grazing dairy cows in eastern Canada based on EFs derived in this study were 4.3 times smaller than those obtained using the IPCC default methodology. Our results therefore suggest that the contribution of grazing dairy cows to the Canadian inventory of soil N2O emissions is currently overestimated. They also indicate that the current methodology for the inventory of soil N2O emissions after deposition of excreta by grazing dairy cows in eastern Canada should be modified to account for soil type and should use distinct EFs for urine and dung.

Acknowledgments This study was funded by the Pollutant Inventories and Reporting Division of Environment Canada and by the Organic Cluster Program of Agriculture and Agri-Food Canada. The amounts of nitrogen deposited

Table 5. Inventory of N2O emissions from grazing dairy cow urine† and dung in provinces of eastern Canada estimated using the emission factor proposed by the default Intergovernmental Panel on Climate Change methodology (2%) and those derived in this study.‡ Province§

Deposited N

Mg N NL 116.0 PEI 508.9 NS 212.6 NB 371.9 QC 6,883.0 ON 5,777.5 Total per soil type Total per excreta type Grand total

Soil textural classes¶ Coarse

Medium

Fine

———— ha ha-1 ———— 0.34 0.66 0 0.93 0.07 0 0.35 0.65 0 0.42 0.56 0.02 0.42 0.37 0.22 0.35 0.63 0.02

This study Urine/soil textural class Dung/soil textural class Coarse Medium Fine Coarse Medium Fine ————————————— kg N2O–N ————————————— 90 399 3 15 21 0 1,083 186 0 177 10 0 171 721 0 28 38 0 360 1,081 64 59 57 1 6,594 13,153 12,147 1,081 691 260 4,582 19,040 1,102 751 1,001 24 12,880 34,581 13,317 2,111 1,818 285 60,778 4,338 65,115

IPCC, all N, all soils# kg N2O–N 2,320 10,178 4,252 7,438 137,661 115,551

277,399

† Total excreted N was distributed among urine (75%) and dung (25%) (Lantinga et al., 1987). ‡ Emission factors on coarse-, medium-, and fine-textured soils were 0.31, 0.70, and 1.09% for urine N, and 0.15, 0.12, and 0.08%, respectively, for dung N. § NB, New Brunswick; NL, Newfoundland-Labrador; NS, Nova Scotia; ON, Ontario; PEI, Prince Edward Island; QC, Quebec. ¶ Proportion of land in eastern Canada under perennial crops that is classified as having a coarse, medium, or fine soil textural class. # IPCC, Intergovernmental Panel on Climate Change. www.agronomy.org • www.crops.org • www.soils.org 839

by grazing dairy cows and the proportion of agricultural soil in different textural classes in Eastern Canada were provided by Devon Worth. The authors thank Johanne Tremblay, Nicole Bissonnette, Gabriel Lévesque, Alain Larouche, Marie-Ève Tremblay, Sylvie Michaud, and Danielle Mongrain for assistance in statistical analysis and field and laboratory work.

References Agriculture and Agri-Food Canada. 2008. National soil database (NSDB). http://sis.agr.gc.ca/cansis/nsdb/ecostrat/1999report/texture.html (accessed 21 Feb. 2014). Allen, A.G., S.C. Jarvis, and D.M. Headon. 1996. Nitrous oxide emissions from soils due to inputs of nitrogen from excreta return by livestock on grazed grassland in the U.K. Soil Biol. Biochem. 28:597–607. doi:10.1016/0038-0717(95)00186-7 Anger, M., C. Hoffmann, and W. Kühbauch. 2003. Nitrous oxide emissions from artificial urine patches applied to different N-fertilized swards and estimated annual N2O emissions for differently fertilized pastures in an upland location in Germany. Soil Use Manage. 19:104–111. doi:10.1079/ SUM2003175 Baily, A., C.J. Watson, R. Laughlin, D. Matthews, K. McGeough, and P. Jordan. 2012. Use of the 15 N gas flux method to measure the source and level of N2O and N2 emissions from grazed grassland. Nutr. Cycling Agroecosyst. 94:287–298. doi:10.1007/s10705-012-9541-x Barnard, R., P.W. Leadley, and B.A. Hungate. 2005. Global change, nitrification and denitrification: A review. Global Biogeochem. Cycles 19. file:///C:/ Users/Robert/Downloads/Barnard%20et%20al%202005%20GBC.pdf (accessed 4 Apr. 2014). Beauchamp, E.G. 1998. Nitrous oxide emission from agricultural soils. Can. J. Soil Sci. 77:113–123. Bhandral, R., S. Saggar, N.S. Bolan, and M.J. Hedley. 2007. Transformation of nitrogen and nitrous oxide emission from grassland soils as affected by compaction. Soil Tillage Res. 94:482–492. doi:10.1016/j.still.2006.10.006 Bouwman, A.F. 1996. Direct emission of nitrous oxide from agricultural soils. Nutr. Cycling Agroecosyst. 46:53–70. doi:10.1007/BF00210224 Chantigny, M.H., P. Rochette, D.A. Angers, S. Bittman, K. Buckley, D. Massé, G. Bélanger, N. Eriksen-Hamel, and M.-O. Gasser. 2010. Soil nitrous oxide emissions following band-incorporation of fertilizer nitrogen and swine manure. J. Environ. Qual. 39:1545–1553. doi:10.2134/jeq2009.0482 Clough, T.J., F.M. Kelliher, R.R. Sherlock, and C.D. Ford. 2004. Lime and soil moisture effects on nitrous oxide emissions from a urine patch. Soil Sci. Soc. Am. J. 68:1600–1609. doi:10.2136/sssaj2004.1600 Clough, T.J., S.F. Ledgard, M.S. Sprosen, and M.J. Kear. 1998. Fate of 15 N labelled urine on four soil types. Plant Soil 199:195–203. doi:10.1023/A:1004361009708 Clough, T.J., R.R. Sherlock, M.N. Mautner, D.B. Milligan, P.F. Wilson, C.G. Freeman, and M.J. McEwan. 2003. Emission of nitrogen oxides and ammonia from varying rates of applied synthetic urine and correlations with soil chemistry. Aust. J. Soil Res. 41:421–438. doi:10.1071/SR02105 Dai, Y., J.D. Hong, K.C. Cameron, and J.-Z. He. 2013. Effects of nitrogen application rate and a nitrification inhibitor dicyandiamide on ammonia oxidizers and N2O emissions in a grazed pasture soil. Sci. Total Environ. 465:125–135. doi:10.1016/j.scitotenv.2012.08.091 Deene, P.J.A.G., and N. Middlekoop. 1992. Effects of cattle dung and urine on nitrogen uptake and yield of perennial ryegrass. Neth. J. Agric. Sci. 40:469–482. de Klein, C.A.M., L. Barton, R.R. Sherlock, K.C. Cameron, and T.J. van der Weerden. 2001. Nitrous oxide emissions from agricultural soils in New Zealand: A review of current knowledge and directions for future research. J. R. Soc. N. Z. 31:543–574. doi:10.1080/03014223.2001.9517667 de Klein, C.A.M., R.J. Eckard, and T.J. van der Weerden. 2010. Nitrous oxide emissions from the N cycle in livestock agriculture: Estimation and mitigation. In: K.A. Smith, editor, Nitrous oxide and climate change. Earthscan Publications, London, UK. p. 107–142. de Klein, C.A.M., R.R. Sherlock, Z. Li, and R.P. Littlejohn. 2003. Estimating a nitrous oxide emission factor for animal urine from some New Zealand pastoral soils. Aust. J. Soil Res. 41:381–399. doi:10.1071/SR02128 de Klein, C.A.M., and R.S.P. van Logtestijn. 1994. Denitrification and N2O emission from urine-affected grassland soil. Plant Soil 163:235–241. doi:10.1007/BF00007973 Di, H.J., and K.C. Cameron. 2003. Mitigation of nitrous oxide emissions in spray-irrigated grazed grassland by treating the soil with dicyandiamide, a nitrification inhibitor. Soil Use Manage. 19:284–290. doi:10.1079/ SUM2003207 840

Di, H.J., and K.C. Cameron. 2006. Nitrous oxide emissions from two dairy pasture soils as affected by different rates of a fine particle suspension nitrification inhibitor, dicyandiamide. Biol. Fertil. Soils 42:472–480. doi:10.1007/s00374-005-0038-5 Di, H.J., K.C. Cameron, and R.R. Sherlock. 2007. Comparison of the effectiveness of a nitrification inhibitor, dicyandiamide, in reducing nitrous oxide emissions in four different soils under different climatic and management conditions. Soil Use Manage. 23:1–9. doi:10.1111/j.1475-2743.2006.00057.x Di, H.J., K.C. Cameron, R.R. Sherlock, J.-P. Shen, J.-Z. He, and C.S. Winefield. 2010. Nitrous oxide emissions from grazed grassland as affected by a nitrification inhibitor, dicyandiamide, and relationships with ammoniaoxidizing bacteria and archaea. J. Soils Sediments 10:943–954. doi:10.1007/s11368-009-0174-x Dixon, E.R., R.J. Laughlin, C.J. Watson, and D.J. Hatch. 2010. Evidence for the production of NO and N2O in two contrasting subsoils following the addition of synthetic cattle urine. Rapid Commun. Mass Spectrom. 24:519–528. doi:10.1002/rcm.4348 Environment Canada. 2013. National Inventory Report. 2011. Greenhouse gas sources and sinks in Canada. 1990-2011. Greenhouse Gas Division, Environment, Canada. Submission to the United Nations Framework Convention on Climate Change. http://ec.gc.ca/Publications/default. asp?lang=En&xml=A07ADAA2-E349-481A-860F-9E2064F34822 (accessed 21 Feb. 2014). Flessa, H., P. Dörsch, F. Beese, H. König, and A.F. Bouwman. 1996. Influence of cattle wastes on nitrous oxide and methane fluxes in pasture land. J. Environ. Qual. 25:1366–1370. doi:10.2134/jeq1996.00472425002500060028x Gregorich, E.G., P. Rochette, A.J. VandenBygaart, and D.A. Angers. 2005. Greenhouses gas contributions of agricultural soils and potential mitigation practices in Eastern Canada. Soil Tillage Res. 83:53–72. doi:10.1016/j.still.2005.02.009 IPCC. 2006. N2O emissions from managed soils and CO2 emissions from lime and urea application: Guidelines for national greenhouse gas inventories. In: 2006 IPCC guidelines for national greenhouse gas inventories. Volume 4: Agriculture, forestry and other land use. IPCC, Geneva, Switzerland. p 11.1–11.54. Kelly, K.B., F.A. Phillips, and R. Baigent. 2008. Impact of dicyandiamide application on nitrous oxide emissions from urine patches in northern Victoria, Australia. Aust. J. Exp. Agric. 48:156–159. doi:10.1071/ EA07251 Lantinga, E.A., J.A. Keuning, J. Greenwold, and P.J.A.G. Deenen. 1987. Distribution of excreted nitrogen by grazing cattle and its effect on sward quality, herbage production and utilization: Animal manure on grassland and fodder crops. Fertilizer or waste? Dev. Plant Soil Sci. 30:103–117. Li, Z., and F.M. Kelliher. 2005. Determining nitrous oxide emissions from subsurface measurements in grazed pasture: A field trial of alternative technology. Aust. J. Soil Res. 43:677–687. doi:10.1071/SR04106 Linn, D.M., and J.W. Doran. 1984. Aerobic and anaerobic microbial populations in no-till and plowed soils. Soil Sci. Soc. Am. J. 48:794–799. doi:10.2136/ sssaj1984.03615995004800040019x Luo, J., S.B. Lindsey, and S.F. Ledgard. 2008. Nitrous oxide emissions from animal urine application on a New Zealand pasture. Biol. Fertil. Soils 44:463–470. doi:10.1007/s00374-007-0228-4 Luo, J., R.W. Tillman, and P.R. Ball. 2000. Nitrogen loss through denitrification in a soil under pasture in New Zealand. Soil Biol. Biochem. 32:497–509. doi:10.1016/S0038-0717(99)00179-0 Maljanen, M., M. Martikkala, H.T. Koponen, P. Virkajärvi, and P.J. Martikainen. 2007. Fluxes of nitrous oxide and nitric oxide from experimental excreta patches in boreal agricultural soil. Soil Biol. Biochem. 39:914–920. doi:10.1016/j.soilbio.2006.11.001 Monaghan, R.M., and D. Barraclough. 1993. Nitrous oxide and dinitrogen emissions from urine-affected soil under controlled conditions. Plant Soil 151:127–138. doi:10.1007/BF00010793 National Research Council. 2001. Nutrient requirements of dairy cattle. 7th ed. Natl. Acad. Press, Washington, DC. Oenema, O., G.L. Velthof, S. Yamulki, and S.C. Jarvis. 1997. Nitrous oxide emissions from grazed grasslands. Soil Use Manage. 13:288–295. doi:10.1111/j.1475-2743.1997.tb00600.x Orwin, K.H., J.E. Bertram, T.J. Clough, L.M. Condron, R.R. Sherlock, M. O’Callaghan, J. Ray, and D.B. Baird. 2010. Impact of bovine urine deposition on soil microbial activity, biomass, and community structure. Appl. Soil Ecol. 44:89–100. doi:10.1016/j.apsoil.2009.10.004 Pelster, D.E., M.H. Chantigny, P. Rochette, D.A. Angers, C. Rieux, and A. Vanasse. 2012. Nitrous oxide emissions respond differently to mineral and organic nitrogen sources in contrasting soil types. J. Environ. Qual. 41:427–435. doi:10.2134/jeq2011.0261 Journal of Environmental Quality

Randall, G.W., D.R. Huggins, M.P. Russelle, D.J. Fuchs, W.W. Nelson, and J.L. Anderson. 1997. Nitrate losses through subsurface tile drainage in conservation reserve program, alfalfa, and row crop systems. J. Environ. Qual. 26:1240–1247. doi:10.2134/jeq1997.00472425002600050007x Ravishankara, A.R., J.S. Daniel, and R.W. Portmann. 2009. Nitrous oxide (N2O): The dominant ozone-depleting substance emitted in the 21st century. Science 326:123–125. doi:10.1126/science.1176985 Rochette, P., D.A. Angers, G. Bélanger, M.H. Chantigny, D. Prévost, and G. Lévesque. 2004. Emissions of N2O from alfalfa and soybean crops in Eastern Canada. Soil Sci. Soc. Am. J. 68:493–506. doi:10.2136/ sssaj2004.4930 Rochette, P., and N. Bertrand. 2008. Soil-surface gas emissions. In: M. Carter and E.G. Gregorich, editors, Soil sampling and methods of analysis. CRC Press, Boca Raton, FL. p. 851–861. Rochette, P., D. Worth, E. Huffman, A. Brierley, B.G. McConkey, J. Yang, J. Hutchinson, R.L. Desjardins, R. Lemke, and S. Gameda. 2008. Estimation of N2O emissions from agricultural in Canada. II: 1990-2005 inventory. Can. J. Soil Sci. 88:655–669. doi:10.4141/CJSS07026 Saarijärvi, K., and P. Virkajärvi. 2009. Nitrogen dynamics of cattle dung and urine patches on intensively managed boreal pasture. J. Agric. Sci. 147:479–491. doi:10.1017/S0021859609008727 SAS Institute. 1999. SAS/STAT user’s guide. Version 8. SAS Inst., Cary, NC. Šimek, M., P. Brůček, J. Hynšt, E. Uhlĩrová, and S.O. Petersen. 2006. Effects of excretal returns and soil compaction on nitrous oxide emissions from a cattle overwintering area. Agric. Ecosyst. Environ. 112:186–191. doi:10.1016/j.agee.2005.08.018 Singh, J., S. Saggar, D.L. Giltrap, and N.S. Bolan. 2008. Decomposition of dicyandiamide (DCD) in three contrasting soils and its effect on nitrous oxide emission, soil respiratory activity, and microbial biomass: An incubation study. Aust. J. Soil Res. 46:517–525. doi:10.1071/SR07204 Thomas, S.M., M.H. Beare, G.S. Francis, H.E. Barlow, and D.I. Hedderley. 2008. Effects of tillage, simulated cattle grazing and soil moisture on N2O emissions from a winter forage crop. Plant Soil 309:131–145. doi:10.1007/ s11104-008-9586-4 Uchida, Y., T.J. Clough, F.M. Kelliher, J.E. Hunt, and R.R. Sherlock. 2011. Effects of bovine urine, plants and temperature on N2O and CO2 emissions from a sub-tropical soil. Plant Soil 345:171–186. doi:10.1007/ s11104-011-0769-z Uchida, Y., T.J. Clough, F.M. Kelliher, and R.R. Sherlock. 2008. Effects of aggregate size, soil compaction, and bovine urine on N2O emissions from a pasture soil. Soil Biol. Biochem. 40:924–931. doi:10.1016/j. soilbio.2007.11.007

van der Weerden, T.J., J. Luo, C.A.M. de Klein, C.J. Hoogendoorn, R.P. Littlejohn, and G.J. Rys. 2011. Disaggregating nitrous oxide emission factors for ruminant urine and dung deposited onto pastoral soils. Agric. Ecosyst. Environ. 141:426–436. doi:10.1016/j.agee.2011.04.007 van Groenigen, J.W., P.J. Kuikman, W.J.M. de Groot, and G.L. Velthof. 2005a. Nitrous oxide emissions from urine-treated soil as influenced by urine composition and soil physical conditions. Soil Biol. Biochem. 37:463–473. doi:10.1016/j.soilbio.2004.08.009 van Groenigen, J.W., G.L. Velthof, F.J.E. Van Der Bolt, A. Vos, and P.J. Kuikman. 2005b. Seasonal variation in N2O emissions from urine patches: Effects of urine concentration, soil compaction and dung. Plant Soil 273:15–27. doi:10.1007/s11104-004-6261-2 Virkajärvi, P., M. Maljanen, K. Saarijärvi, J. Haapala, and P.J. Martikainen. 2010. N2O emissions from boreal grass and grass-clover pasture soils. Agric. Ecosyst. Environ. 137:59–67. doi:10.1016/j.agee.2009.12.015 Wachendorf, C., C. Lampe, F. Taube, and K. Dittert. 2008. Nitrous oxide emissions and dynamics of soil nitrogen under 15N-labeled cow urine and dung patches on a sandy grassland soil. J. Plant Nutr. Soil Sci. 171:171– 180. doi:10.1002/jpln.200625217 Yamulki, S., S.C. Jarvis, and P. Owen. 1998. Nitrous oxide emissions from excreta applied in a simulated grazing pattern. Soil Biol. Biochem. 30:491–500. doi:10.1016/S0038-0717(97)00145-4 Zaman, M., and M.L. Nguyen. 2012. How application timings of urease and nitrification inhibitors affect N losses from urine patches in pastoral system. Agric. Ecosyst. Environ. 156:37–48. doi:10.1016/j.agee.2012.04.025 Zaman, M., S. Saggar, J.D. Blennerhassett, and J. Singh. 2009. Effect of urease and nitrification inhibitors on N transformation, gaseous emissions of ammonia and nitrous oxide, pasture yield and N uptake in grazed pasture system. Soil Biol. Biochem. 41:1270–1280. doi:10.1016/j.soilbio.2009.03.011 Zaman, M., S. Zaman, M.L. Nguyen, T.J. Smith, and S. Nawaz. 2013. The effect of urease and nitrification inhibitors on ammonia and nitrous oxide emissions from simulated urine patches in pastoral system: A two-year study. Sci. Total Environ. 465:97–106. doi:10.1016/j.scitotenv.2013.01.014 Ziadi, N., R.R. Simard, G. Allard, and J. Lafond. 1999. Field evaluation of anion exchange membranes as a N soil testing method for grasslands. Can. J. Soil Sci. 79:281–294. doi:10.4141/S98-062

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