The effect of the nitrification inhibitor dicyandiamide (DCD) on nitrous oxide and methane emissions after cattle slurry application to Irish grassland

Agriculture, Ecosystems and Environment 199 (2014) 339–349

Contents lists available at ScienceDirect

Agriculture, Ecosystems and Environment journal homepage: www.elsevier.com/locate/agee

The effect of the nitrification inhibitor dicyandiamide (DCD) on nitrous oxide and methane emissions after cattle slurry application to Irish grassland E. Cahalan a,b , M. Ernfors a,c , C. Müller b,d , D. Devaney e, R.J. Laughlin f , C.J. Watson f , D. Hennessy g , J. Grant h , M.I. Khalil b , K.L. McGeough f , K.G. Richards a, * a

Teagasc, Johnstown Castle Environmental Research Centre, Co., Wexford, Ireland School of Biology and Environmental Science, University College Dublin, Belfield Dublin 4, Ireland c Department of Biosystems and Technology, Swedish University of Agricultural Sciences, P.O. Box 104, SE 23053 Alnarp, Sweden d Department of Plant Ecology, Justus-Liebig University Giessen, Heinrich-Buff-Ring 26, 35392 Giessen, Germany e Department for Environment, Food and Rural Affairs, 17 Smith Square, London SW1P 3JR, United Kingdom f Agri-Food and Biosciences Institute (AFBI), Newforge Lane, Belfast BT9 5PX, Northern Ireland g Teagasc, Animal and Grassland Research and Innovation Centre, Co., Cork, Moorepark, Fermoy, Ireland h Statistics and Applied Physics, Research Support Team, Teagasc, Ashtown Dublin 15, Ireland b

A R T I C L E I N F O

A B S T R A C T

Article history: Received 10 August 2013 Received in revised form 12 September 2014 Accepted 21 September 2014 Available online xxx

Application of cattle slurry to grassland can lead to gaseous losses of nitrogen (N). The dynamics of ammonia (NH3) emissions are well documented, but emissions of nitrous oxide (N2O) and other trace gases from land application of cattle slurry, are not as well understood. Nitrification inhibitors such as dicyandiamide (DCD) help to retain soil N in the ammonium (NH4+) form, which is expected to result in reduced N loss and increased N use efficiency. Emissions of methane (CH4) could potentially be affected by DCD, since the enzyme ammonia monooxygenase (AMO), which DCD is thought to inhibit, can oxidise CH4 as well as NH3. The objective of this study was to investigate the impacts of DCD on N2O and CH4 emissions from grassland soils after slurry application. At one experimental site, Johnstown Castle (JC), County Wexford, Ireland, slurry was applied in March, June and October for two years, at a rate of 33 m3 ha
1, by either bandspread (BS) or splashplate (SP) application methods, with and without DCD. A second site at Hillsborough (HB), County Down, Ireland, was treated in the same way for one year only, using the SP application method. Emissions of N2O and CH4 were measured using static chambers. Over the entire experiment DCD significantly reduced cumulative N2O emissions by 47 and 70% at both sites. Slurry spreading method had no significant effect on direct N2O emissions and there was no effect of DCD on CH4 emissions throughout the experiment. Using an emission factor for indirect N2O emissions of 1%, modelled losses of NH3 through volatilisation, resulted in an estimated 13 times greater indirect than direct N2O emissions. The results suggest that, under typical Irish field conditions, DCD can be used to decrease direct N2O emissions, without increasing CH4 emissions, and bandspreading can be used as a method of decreasing NH3 volatilisation, without increasing direct N2O or CH4 emissions. Since direct N2O emissions were relatively small, targeting indirect N2O emissions through mitigation of NH3 volatilisation and nitrate (NO3
) leaching could be an effective way of decreasing total N2O emissions from slurry application at these sites. ã 2014 Elsevier B.V. All rights reserved.

Keywords: Dicyandiamide (DCD) Nitrous oxide Cattle slurry Spreading method Trace gases

1. Introduction Manure applications are sources of the trace gases nitrous oxide (N2O) and methane (CH4) to the atmosphere. These two gases

* Corresponding author. E-mail address: [email protected] (K.G. Richards). http://dx.doi.org/10.1016/j.agee.2014.09.008 0167-8809/ ã 2014 Elsevier B.V. All rights reserved.

contribute to climate change (IPCC, 2006) and N2O is also involved in stratospheric ozone depletion (IPCC, 2013), which highlights the need for mitigation. It is estimated that agricultural sources account for 84 and 47% of the global anthropogenic N2O and CH4 emissions, respectively (Scheehle and Kruger, 2006), and regionally the contributions can be even larger. In Ireland, where this study was conducted, it is estimated that agriculture accounts for 29.1% of national greenhouse gas (GHG) emissions, of which 35.6%

340

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349

are derived from fertiliser application and animal deposition (McGettigan et al., 2010). The nitrification inhibitor dicyandiamide (DCD) has proved to be efficient in decreasing N2O emissions when applied in the field with mineral fertiliser or urine (Zaman et al., 2009; Smith et al., 2008; Di et al., 2007; Di and Cameron, 2002; Weiske et al., 2001). Laboratory studies have also shown that DCD can substantially reduce N2O emissions after slurry application (Hatch et al., 2005; Tao et al., 2008; Merino et al., 2002), although in some studies no effect has been observed (Pereira et al., 2010; Mkhabela 2006a,b). There are a limited number of field studies on the effects of DCD on N2O emissions after application of animal slurry. Merino et al. (2002) found a reduction in N2O emissions of 60% over 20 days after cattle slurry application to grassland on a gleysol in northern Spain. In a study by Vallejo et al. (2006) on a clay loam in central Spain, N2O emissions were 23% lower when pig slurry was applied with DCD, compared to pig slurry alone, over the 30 days after application to a potato crop. To our knowledge, no one has studied the effects of DCD on N2O emissions over several seasons, after cattle slurry application to grassland. Spreading techniques aimed at reducing NH3 emissions such as bandspreading, can potentially reduce N2O emissions, due to decreases in indirect emissions (Brink et al., 2001). Indirect emissions of N2O, as a result of deposited NH3 from volatilisation, can come from ground or surface water and soils (IPCC, 2006). At the same time, there is a risk that these techniques could enhance direct emissions of N2O due to increased anaerobic bacterial activity (Rodhe et al., 2006; Perälä et al., 2006). Emissions of methane (CH4) from Irish agriculture mostly originate from livestock, but can also be emitted from or taken up by soils (Chan and Parkin, 2001). The use of DCD could potentially increase CH4 emissions, since the enzyme ammonia monooxygenase, which DCD is thought to inhibit, can oxidise CH4 as well as NH3 (Bedard and Knowles, 1989). Decreases in CH4 emissions due to DCD addition have been reported from flooded soils (Lindau et al., 1993; Ghosh et al., 2003), but there do not appear to be any reports on the effect on DCD amended slurry on CH4 emissions from grassland. We hypothesised that the efficacy of DCD in affecting greenhouse gas emissions after slurry spreading is influenced by season and spreading method. The objectives of the current study were: (1) to quantify and compare the effects of the nitrification inhibitor DCD on N2O and CH4 emissions after cattle slurry application to Irish grassland soils, using bandspread or splashplate spreading techniques, in three different seasons and; (2) to

determine if N2O and CH4 emissions were affected by the slurry spreading technique. 2. Materials and methods 2.1. Experimental setup There were two experimental sites. The first site was a grassland site at Johnstown Castle (JC) Research Centre, County Wexford, Ireland (5218 0 N; 6 30 0 W), where experiments were conducted from March 2009 to October 2010. The long-term (1990–2010) average rainfall and air temperature for the site were 1036 mm and 10.4  C, respectively (Met Eireann, 2012). The soil was a loam with imperfect drainage. The sward at the site was predominantly perennial ryegrass (Lolium perenne). White clover (Trifolium repens) and broad-leaved docks (Rumex obtusifolius) were removed from the sward using MCPA AMINE 500 herbicide (Nufarm, UK Limited) in October 2008, before the experiment started. There were six treatments: (1) cattle slurry (CS), bandspread (BS) application method, with (+) DCD; (2) CS, BS, without (
) DCD; (3) CS, splashplate (SP) application method, +DCD; (4) CS, SP,
DCD; (5) H2O, + DCD and (6) H2O,
DCD. The treatments were applied to plots 3m  2 m in area and each treatment was replicated six times in a randomised block design. Application was carried out in March, June and October of 2009 and 2010, which corresponded to the times of slurry application commonly used on Irish grassland farms. The second site was also grassland and located at Hillsborough (HB), County Down, Northern Ireland (54 46 0 N; 6 08 0 W), where experiments were carried out from March 2010 to October 2010. The long-term (1995–2010) average rainfall and daily air temperature for the site were 916 mm and 9.3  C, respectively (AFBI, Climatological Database, 2012). The soil was a sandy clay loam with moderate drainage. The sward composition was similar to that at JC and was managed in the same way. Only treatments 3 and 4 were used. At both sites, cattle slurry was locally sourced and applied at a rate of 33 m3 ha
1, which is the usual rate for silage production in Ireland. DCD was mixed into the slurry at a rate of 15% of the slurry NH4+-N content, just prior to slurry application. The DCD application rate was based on the NH4+-N content of the slurry, as slurry NH4+ contents varied between sites and application timings. This allowed for an equal ratio of DCD and NH4+ to be applied at all application timings. The rate of DCD application ranged between 3.8 and 10.3 kg ha
1. The control plots (H2O)

Table 1 Application date, dry matter %, pH, Total N % and NH4+-N content of slurries used at the Johnstown Castle and Hillsborough sites. Application date

Dry matter (%)

pH

Total N (%)

NH4+-N (mg kg
1)

Mar-09 Jun-09 Oct-09 Mar-10 Jun-10 Oct-10

6.04 7.07 5.70 7.56 7.12 6.48

7.13 7.17 7.36 – 7.97 7.48

0.31 0.37 0.31 0.25 0.25 0.29

1537 2081 1711 1554 775 1100

Mean SD

6.66 0.71

7.42 0.34

0.30 0.05

1460 461

Hillsborough

Mar-10 Jun-10 Oct-10 Mean SD

5.62 5.85 3.16 4.75 2.34

8.27 7.47 7.52 6.42 3.00

0.37 0.33 0.18 0.25 0.12

1960 1671 1068 1287 528

Both Sites

Mean SD

6.07 1.29

7.55 0.39

0.30 0.06

1495 433

Johnstown Castle

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349

received water in an equal volume to that in the slurry and DCD was applied at the same rate as for the slurry treatments. Cattle slurry or water (for control treatments) was applied manually

341

using watering cans. For the bandspread treatment, slurry was applied in parallel lines, 20 cm apart. For the splashplate treatments an attachment was fitted to the nozzle of the watering can,

Fig. 1. (a) Daily rainfall and temperature at grass level, at the Johnstown Castle site. (b) Daily rainfall and temperature at grass level, at the Hillsborough site.

342

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349

 16 mm). Each syringe was flushed three times with headspace air before sampling. A 10 ml sample was withdrawn from the chamber and injected to over-pressurise a 7 ml pre-evacuated glass vial. At JC, headspace samples were taken at 0 and 25 min in 2009. In 2010, headspace samples were taken at 25 min and an average of 5 samples of ambient air was used as a t0 value for all chambers. Headspace samples at HB were taken at 40 min. Sample vials were stored for c.7–14 days before analysis using a Varian 3800 gas chromatograph (GC) (Varian, Inc., Palo Alto, CA, USA) equipped with an electron capture detector (ECD) for N2O and a flameionisation detector (FID) for CH4. The GC limits of detection for N2O and CH4 were 0.019 and 0.735 ppm, respectively. Emissions were calculated using the difference in gas concentrations at t0 and t25 at JC and between ambient and t40 at HB. Nitrous oxide sampling occurred every day (between 11.00 and 12.00pm) for the first 10 days after slurry application and then every 2–4 days until emissions from treated plots returned to ambient concentrations. At both sites the increases in N2O and CH4 concentrations in the static chambers were checked for linearity before the experiment

to simulate the coverage typical of a conventional splashplate slurry tanker. At HB, treatments were weighed out separately for application within chambers to help reduce spatial variability. 2.2. Trace gas measurements Emissions of both N2O (JC and HB) and CH4 (JC) were measured using the static chamber method. At JC, the stainless steel chambers used were 40 cm  40 cm wide and 10 cm high. Collars were permanently inserted into the centre of each plot to a depth of 10 cm. Each collar had a water-filled trough that the chamber lid was placed in during sampling to prevent gas exchange with the external atmosphere. At HB, round PVC chambers (40 cm diameter) with stainless steel lids were deployed. The PVC base was permanently inserted into the ground to a depth of 5 cm. A gastight seal was made using a rubber gasket and the lids were clipped into position. Emission samples were collected from the chamber headspaces through butyl rubber septa, using a 20 ml polypropylene syringe equipped with a 25 gauge Luer lock needle (0.5 mm

Soil NH4+-N concentration (kg ha-1)

120

A

D

BS BS DCD SP SP DCD H2O H2O DCD

100 80 60 40 20 0

Soil NH4+-N concentration (kg ha-1)

120 B

E

C

F

100 80 60 40 20 0

Soil NH4+-N concentration (kg ha-1)

120 100 80 60 40 20 0

5

10

15

20

Days after application

25

30

5

10

15

20

25

30

Days after application

Fig. 2a. Soil NH4+-N concentrations (0–10 cm) at Johnstown Castle. (A) March 2009; (B) June 2009; (C) October 2009; (D) March 2010; (E) June 2010 and (F) October 2010. BS = bandspread applied slurry, BS DCD = bandspread slurry with DCD, SP = splashplate applied slurry, SP DCD = splashplate slurry with DCD, H2O = control and H2O and DCD = control with DCD. BS SP H2O BS DCD SP DCD H2O+DCD

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349

Soil NO3--N concentration (kg ha-1)

10

343

D

A

BS BS DCD SP SP DCD H2O H2O DCD

8

6

4

2

0

Soil NO3--N concentration (kg ha-1)

10 E

B 8 6 4 2 0

Soil NO3--N concentration (kg ha-1)

10 C

F

8 6 4 2 0

5

10

15

20

25

Days after application

30

5

10

15

20

25

30

Days after application

Fig. 2b. Soil NO3– N concentrations (0–10 cm) at Johnstown Castle. (A) March 2009; (B) June 2009; (C) October 2009; (D) March 2010; (E) June 2010 and (F) October 2010. BS = bandspread applied slurry, BS DCD = bandspread slurry with DCD, SP = splashplate applied slurry, SP DCD = splashplate slurry with DCD, H2O = control and H2O DCD = control with DCD.

began. This was achieved by taking 15 emission samples over a 60 min period, from four representative chambers placed over plots treated with similar manures used in this experiment. Whilst fitting a non-linear function to the data, R2 values were not increased compared to using a linear function. Cumulative gas fluxes were obtained for each chamber and application date, at JC and HB, by integrating individual observations over time, using a trapezoidal approach. Cumulative fluxes were calculated over 24 days after treatment application for N2O and over 10 days after treatment application for CH4, covering the periods from treatment application until a few days after background emissions were reached, in all treatments. Measurements continued up to 80 days post treatment application, to ensure background levels were truly reached. 2.3. Slurry analysis Total N in the slurry samples at each application were determined by Kjeldahl digestion with concentrated sulphuric acid, using copper as a catalyst (MAFF, 1981). Slurry pH was

determined by an Orion pH meter Model 420A+. Slurry dry matter (DM) content was determined by drying a representative slurry sample at 105  C until a constant weight was obtained (after 24–48 h). The NH4+-N and total N contents of the slurries used in this study varied considerably over the six application timings (Table 1), with the rate of NH4+-N applied varying between 25.6 and 68.7 kg ha
1. The slurries used had a mean DM content of 6.1% and a mean pH of 7.5 (Table 1). 2.4. Soil mineral nitrogen Soil mineral N (NH4+-N and NO3
N) contents were determined once per week concurrently with gas sampling, at the JC site only. For each plot, approximately 6–8 soil cores (60 mm diameter) were collected using an auger, to a depth of 10 cm. The soil was sieved through a 3.5 mm screen. A 20 g subsample was shaken for 1 hour with 50 ml, 2 M KCl and the supernatant was filtered through a Whatman no. 2 filter paper, before being stored at 4  C. Samples were analysed for mineral N concentrations using an Aquakem 600 discrete analyser (Thermo Electron OY, Vantaa,

344

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349

Finland). Each soil sample was extracted less than 24 h after being removed from the field. 2.5. Meteorological data Daily rainfall (mm) and mean daily minimum grass level temperature ( C) (Fig. 1a,b) along with mean daily wind speed (m
1 s
1) and mean daily solar radiation (J cm
2) were obtained from weather stations less than 1 km from each experimental site.

across the application timing factor. Initial soil mineral N content and meteorological data were tested as covariates in the analysis. A significance level of 0.05 was used, unless otherwise stated and residual checks were made to ensure that the assumptions of the analysis were met. Where appropriate log transformation was used and results were back-transformed to the measurement scale. 3. Results 3.1. Meteorological data

2.6. Ammonia emissions Losses of NH3 were modelled after each slurry application using the ammonia loss from field-applied animal manure (ALFAM) model (Søgaard et al., 2002). Ammonia volatilisation rates from slurry amended with DCD were assumed to be the same as without DCD, as previous workers have reported that DCD had no significant effect on NH3 volatilisation rates (Mkhabela et al., 2006a,b; Tao et al., 2008). Indirect N2O emissions, associated with N volatilised from the experimental plots being redeposited in another place, were estimated as being 1% of the total NH3 emissions (Bourdin et al., 2014; IPCC, 2006). Meteorological data for the first 24 h following slurry application were used in the model along with slurry total ammoniacal nitrogen (TAN) and slurry DM content. Non-volatilised N was defined as the slurry NH4+-N application rate minus modelled NH3-N losses. 2.7. Statistical methods The data were analysed using the MIXED procedure in SAS v. 9.3 (SAS, 2011). Fixed factors in the model were block, application timing, application method and DCD. Repeated measurements were taken into account using an unstructured correlation model

At JC, total annual rainfall was 41% (389 mm) above and 16% (152 mm) below the 30 year average in 2009 and 2010, respectively. Mean annual grass level temperatures were 0.8  C and 2.1  C below the 30 year average in 2009 and 2010, respectively. In 2009, grass level temperatures were on average 1.7  C higher than in 2010 and rainfall was 37% higher in 2009 than in 2010 (Fig. 1a). At HB, annual total rainfall and grass level temperatures in 2010 were similar to the 30 year average. In 2010, annual grass level temperatures were on average 4.9  C lower at JC than at HB and annual rainfall was 18% (144 mm) higher at JC than at HB (Fig. 1a,b). Daily average wind speed was up to 5 times greater at JC than at HB in 2010. 3.2. Soil mineral N Mineral N concentrations from each treatment, over the six sampling dates can be seen in Figs. 2a and 2b. Concentrations of NH4+-N ranged from to 4.3 kg NH4+-N ha
1 within the H2O treatment in March 2009 to 84 kg NH4+-N ha
1 within BS DCD in June 2009. Concentrations of NO3– N were negligible following the June and October 2009 applications but they peaked at 27 kg

Table 2 Type 3 fixed effects for N2O cumulative fluxes (g N ha
1) over 24 days post application for each site. Mean direct N2O fluxes and confidence intervals (g N ha
1) for the main factors slurry application method (slurry), with (+DCD) and without DCD (Control) and slurry application timing (Month Year) at JC and HB. Site: JC

Site: HB

Effect

Num DF

Den DF

F value

Pr > F

Num DF

Den DF

F value

Pr > F

Block Month Year Slurry Slurry*Month Year DCD Month Year*DCD Slurry*DCD Slurry*Month Year*DCD

5 5 1 5 1 5 1 5

15 15 15 15 15 15 15 15

1.35 32.25 0.07 0.31 49.7 2.57 1.13 0.19

0.3 <.0001 0.79 0.9 <.0001 0.071 0.3 0.96

3 2

3 3

2.39 1.13

0.2466 0.4309

1 2

3 3

27.62 1.85

0.0134 0.2992

Mean direct N2O fluxes (g N ha
1) and 95% confidence interval (C.I.) Site: JC

Site: HB

Slurry SP BS

Flux 50.3a 51.6a

DCD

Flux

C.I.

DCD

Flux

C.I.

Control

37.1d

32.4–42.5

+DCD

73.0d

43.4–122.67

C.I. 44.0–57.6 45.0–59.0

Slurry SP

Month Year

Flux

C.I.

Month Year

Mar-09 Jun-09 Oct-09 Mar-10 Jun-10 Oct-10

79.4ef 103.5e 40.7fg 34.7g 39.8g 37.8g

63.7–99.0 82.5–129.9 28.7–57.8 27.2–44.2 30.4–52.2 32.0–44.6

Mar-10 Jul-10 Oct-10

126.6e 118.0e 160.4e

90.5–177.2 47.9–290.6 113.6–226.6

Each site (JC, HB) was analysed individually due to variation in the experimental treatments at each site. Mean N2O fluxes followed by the same letter are not significantly different.

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349

NO3– N in the BS treatment following the June 2010 application (Figs. 2a and 2b).

345

respectively, at both sites. At JC, daily N2O emissions measured for single chambers ranged from
5 to 36 g N2O-N ha
1,
5 to 35 g N2O-N ha
1 and
4 to 9 g N2O-N ha
1, for the SP, BS and H2O treatments respectively. At HB, daily N2O emissions ranged from 0.1 to 63 g N2O-N ha
1 within the SP treatment. Mean cumulative N2O fluxes at both sites were low ranging from 34.7 to 103.5 g N ha
1 at JC to 73.0 to 245.5 g N ha
1 at HB. At JC there were significant effects of application timing (P < 0.0001) and DCD

3.3. Gaseous emissions Emissions of N2O and CH4 from the slurry treated plots peaked within the first 120 and 24 h respectively, and fell to ambient levels within 14–21 days and 24–48 h after slurry application

N2O-N Emissions (g N ha-1 d-1)

40 A

D

B

E

C

F

BS BS DCD SP SP DCD H2O H2O DCD

30

20

10

0

N2O-N Emissions (g N ha-1 d-1)

40

30

20

10

0

N2O-N Emissions (g N ha-1 d-1)

40

30

20

10

0

0

5

10

15

20

Days after Application

25

30

0

5

10

15

20

25

30

Days after Application

Fig. 3a. Temporal nitrous oxide emissions at Johnstown Castle. (A) March 2009; (B) June 2009; (C) October 2009; (D) March 2010; (E) June 2010 and (F) October 2010. BS = bandspread applied slurry, BS DCD = bandspread slurry with DCD, SP = splashplate applied slurry, SP DCD = splashplate slurry with DCD, H2O = control and H2O DCD = control with DCD. BS SP H2O BS DCD SP DCD H2O+DCD

346

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349

N ha
1 at the March and June 2009 application timings and these were significantly higher than the other four applications which ranged from 34.7 to 40.7 g N ha
1.There was no significant effect of slurry application method, at JC, with mean cumulative emissions of 50.3 and 51.6 g N ha
1 for SP and BS, respectively. At HB there was a significant effect of DCD on cumulative N2O fluxes and there were no other significant effects (Table 2). Over the three application timings at HB, DCD significantly reduced cumulative emissions by 70% from 245.5 to 73 g N ha
1 for the control and +DCD treatments, respectively (Figs. 3a and 3b). Average daily methane fluxes from the slurry treatments ranged from
30.8 g CH4-C ha
1 to 202 g CH4-C ha
1 d
1at JC and 0 to 1283 g CH4-C ha
1 d
1 at HB for the six application dates (Figs. 4a and 4b). There were no significant effects of DCD addition, spreading method or season of application on cumulative CH4 emissions.

40

N2O-N Emissions (g N ha-1 d-1)

A 30

20

10

0

3.4. Ammonia losses

40

N2O-N Emissions (g N ha-1 d-1)

B

The output from the ALFAM model (Table 3) showed that the highest relative NH3 losses occurred during the June slurry applications in both years and at both sites, with between 51 and 100% of the TAN being lost 24 h after application. The lowest total-N relative losses through NH3 volatilisation occurred in October, with the BS treatment, at JC (30% of applied N) and in March at HB (36% of applied N). Ammonia relative losses at the HB site, as estimated by the ALFAM model, were slightly lower than at JC (Table 3).

30

20

10

4. Discussion 4.1. N2O fluxes

0

N2O-N Emissions (g N ha-1 d-1)

C

SP SP DCD

60

20

0 0

5

10

15

20

25

30

Days after Application Fig. 3b. Temporal nitrous oxide emissions at Hillsborough. (A) March 2010; (B) June 2010 and (C) October 2010. SP = splashplate applied slurry, SP DCD = splashplate slurry with DCD. SP SP DCD

(P < 0.0001) on cumulative N2O fluxes, there were no other significant interactions observed (Table 2). Over the six application timings at JC, DCD significantly reduced N2O fluxes by 47% from 69.9 to 37.1 g N ha
1 for the control and +DCD treatments, respectively. At JC cumulative N2O fluxes were 79.4 and 103.5 g

The cumulative N2O emissions over 24 days were relatively small (<130 g N2O-N ha
1 at JC and <380 g N2O-N ha
1 at HB) compared to previous studies on slurry amended grassland (Chadwick et al., 2000; Dittert et al., 2001; Merino et al., 2002), even though similar slurry application rates were used. Some negative N2O emissions were observed, especially in March and October when soil moisture contents were highest. The largest negative N2O fluxes were observed to have occurred after the October 2009 application during a very wet period when 174 mm rainfall was recorded in the 24 days following slurry application. These negative fluxes are most likely to be due to N2O reduction to N2 under the extreme wet period coupled with low NO3
availability (Fig. 2b) which was reported by Chapuis-Lardy et al. (2007) to favour N2O consumption. Direct N2O emissions as a percentage of N applied (average 0.22% in JC, 0.67% in HB) were both well below the annual 1% default direct emission factor used by the IPCC (IPCC, 2006). The significantly higher fluxes of N2O in March and June 2009, at JC were most likely due to a larger volume of rainfall combined with higher soil temperatures, as these are two of the main factors driving N2O emissions (Hénault et al., 1998; Skiba and Smith, 2000). In June and October, the NH4+-N content of the slurry was higher in 2009 than in 2010, contributing to higher N2O emissions. However, in March, the N2O emissions were higher in 2009 despite similar additions of NH4+-N in the two years, which suggests that climatic factors were important in determining the emissions. There were generally higher N2O emissions at HB (relative to the NH4+ application rate) when compared to JC (Table 2). The larger estimated values of volatilisation at JC were mainly due to higher wind speeds at JC compared to HB, as the HB site was more sheltered than JC. Higher temperatures at HB compared to JC in 2010, along with a larger volume of rainfall in March and higher available N due to lower volatilisation would have contributed to the larger N2O losses seen in HB.

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349

The estimated indirect emissions of N2O from the two sites were often larger than the direct emissions, especially in the SP treatment (Table 3). These results suggest that, for typical Irish field conditions, targeting indirect N2O emissions could be at least as important as decreasing direct emissions from slurry application. 4.1.1. Effects of DCD The results from this experiment clearly shows that DCD can reduce N2O emissions from slurry application, which is in

347

agreement with Vallejo et al. (2006), Pain et al. (1994) and Merino et al. (2002). The largest percentage reduction occurred on the poorly drained HB site which is likely to be a combination of higher N2O emissions and slower degradation of DCD at lower soil temperatures (Kelliher et al., 2008; Rajbanshi et al., 1992). The direct N2O emission reductions of 47 and 70% at the two sites are within the range of 33 to 82% reported by Di and Cameron (2002), Smith et al. (2008) and Singh et al. (2009). It has been shown in previous studies that DCD can be effective in decreasing nitrate (NO3
) leaching (Di and Cameron, 2002; Singh

1400

CH4-C Emissions (g C ha-1 d-1)

1200

A

D

B

E

C

F

BS BS DCD SP SP DCD H2O H2O DCD

1000 800 600 400

200 150 100 50 0 -50 1400

CH4-C Emissions (g C ha-1 d-1)

1200 1000 800 600 400

200 150 100 50 0 -50 1400

CH4-C Emissions (g C ha-1 d-1)

1200 1000 800 600 400

200 150 100 50 0 -50 0

2

4

6

Days after Application

8

10

0

2

4

6

8

10

Days after Application

Fig. 4a. Temporal methane emissions at Johnstown Castle. (A) March 2009; (B) June 2009; (C) October 2009; (D) March 2010; (E) June 2010 and (F) October 2010. BS = bandspread applied slurry, BS DCD = bandspread slurry with DCD, SP = splashplate applied slurry, SP DCD = splashplate slurry with DCD, H2O = control and H2O DCD = control with DCD. BS SP H2O BS DCD SP DCD H2O+DCD

348

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349 Table 3 Total ammoniacal nitrogen (TAN) application rate, NH3 volatilisation lossa and indirect N2O emissions for splashplate and bandspread slurry, at each application date and site.

1400 A

CH4-C Emissions (g C ha-1 d-1)

1200 1000

Splashplate

800 600 400

200 150 100 50 0

Bandspread

Johnstown Castle

NH3 lossa Indirect TAN NH3 TAN (kg ha
1) lossa (kg N ha
1) (kg N ha
1) N2Ob (kg N2O(kg ha
1) N ha
1)

Indirect N2Ob (kg N2ON ha
1)

Mar-09 Jun-09 Oct-09 Mar-10 Jun-10 Oct-10 Mean SD

50.7 68.7 56.5 51.3 25.6 50.7 50.6 12.8

34.5 49.7 31.4 31.8 26.8 32.2 34.4 7.2

0.35 0.5 0.31 0.32 0.27 0.32 0.35 0.07

0.2 0.29 0.18 0.18 0.16 0.19 0.2 0.04

Hillsborough Mar-10 Jun-10 Oct-10 SD

64.7 55.1 38.4 10.9

23.2 28.2 14.5 5.7

0.23 0.28 0.15 0.05

50.7 68.7 56.5 51.3 25.6 50.7 50.6 12.8

19.9 28.7 18.1 18.3 15.5 18.6 19.9 4.2

-50 1400 B

CH4-C Emissions (g C ha-1 d-1)

1200 1000 800

a NH3 volatilisation losses were obtained using the ammonia loss from Fieldapplied Animal Manure (ALFAM) model (Søgaard et al., 2002). b Indirect N2O emissions were estimated from the modelled NH3 volatilisation loss values, using an emission factor of 0.01 (IPCC, 2006).

600 400

200 150

effect of DCD on N2O emissions after slurry spreading to these types of grassland sites.

100 50 0

4.1.2. Influence of spreading method Contrary to the results of Rodhe et al. (2006) and Perälä et al. (2006), the use of bandspreading compared to the splashplate application method did not significantly affect direct N2O emissions at any application time. Spreading method did reduce the modelled indirect N2O emissions from NH3 volatilisation by 43% from 350 to 200 g N ha
1. This is almost double the reductions observed by Bourdin et al. (2014) who observed a 22% decrease in indirect N2O emissions from BS compared to SP applied slurry. A review by Webb et al. (2010) concluded that, while some studies have seen slight increases in direct N2O emissions as a result of NH3 controlling application methods, the increased benefit from reducing NH3 emissions outweighs any small increase in direct emissions.

-50 1400 C

CH4-C Emissions (g C ha-1 d-1)

1200

SP SP DCD

1000 800 600 400

200 150 100 50

4.2. CH4 fluxes

0 -50 0

2

4

6

8

10

Days after Application Fig. 4b. Temporal methane emissions at Hillsborough. (A) March 2010; (B) June 2010 and (C) October 2010. SP = splashplate applied slurry, SP DCD = splashplate slurry with DCD. SP SP DCD

et al., 2009). Leaching of NO3
can contribute to indirect N2O emissions, but these emissions could not be estimated in this study, since leaching of NO3
was not measured. The soils were not free-draining, meaning leaching would have been limited (van Es et al., 2006). However, considering the generally low direct N2O emissions, it is possible that the effect of DCD on indirect emissions through NO3
leaching makes up a significant part of the total

The finding that between 75 and 90% of the CH4 emissions occurred in the first 24 h after slurry application was in line with the results of Rodhe et al. (2006). Wulf et al. (2002) and Chadwick and Pain (1997) found that CH4 emissions after slurry application are short-lived and are derived from the slurry itself. Despite the peak in emissions in the first 24 h, cumulative CH4 fluxes were small and the soil acted as either a small sink or source of CH4 thereafter. Under aerobic conditions generally a net CH4 uptake is observed, associated with the methanotrophic activity in soils, which can be an important sink for CH4 in the atmosphere or generated in-situ (Amaral et al., 1998). Cumulative CH4 emissions were in a similar range as those reported by Perälä et al. (2006) (
290 to 50 g CH4-C ha
1). If the CH4 fluxes were mainly derived from the slurry, this explains why DCD did not affect CH4 emissions following slurry application to grassland in this study. The effect on CH4 emissions of DCD inclusion in animal diets (Welten et al., 2014) or addition during slurry storage warrants further investigation.

E. Cahalan et al. / Agriculture, Ecosystems and Environment 199 (2014) 339–349

5. Conclusions Dicyandiamide mixed into slurry prior to application to grassland was effective at lowering direct N2O emissions, by between 47 and 70% over 9 slurry application timings at two contrasting sites. Emissions of CH4 were not affected by DCD and were generally low and short-lived. Although direct N2O emissions were low (<0.3 kg N ha
1 over the first 24 days), they are still of environmental significance and can contribute to anthropogenic GHG emissions. Spreading method had a major impact on reducing indirect N2O emissions, resulting from NH3 being re-deposited to land, which are of concern as they were greater than direct N2O emissions. The combination of DCD and bandspreading are effective mitigation technologies for reducing total N2O emissions from the land application of cattle slurry. Acknowledgements The authors would like to thank the laboratory and field staff at both Teagasc Johnstown Castle and AFBI Hillsborough with their assistance on this project. This research was financially supported under the National Development Plan, through the Research Stimulus Fund (Grant RSF07519), administered by the Department of Agriculture, Food and the Marine. References Amaral, J., Ren, T., Knowles, R., 1998. Atmospheric methane consumption by forest soils and extracted bacteria at different pH values. Appl. Environ. Microbiol. 64, 2397–2402. Bedard, C., Knowles, R., 1989. Physiology, biochemistry and specific inhibitors of CH4, NH4+ and CO oxidation by methanotrophs and nitrifiers. Microbiol. Rev. 53, 68–84. Brink, C., Kroeze, C., Klimont, Z., 2001. Ammonia abatement and its impact on emissions of nitrous oxide and methane in Europe -Part 1: method. Atmos. Environ. 35 (36), 6299–6312. Bourdin, F., Sakrabani, R., Kibblewhite, M.G., et al., 2014. Effect of slurry dry matter content, application technique and timing on emissions of ammonia and greenhouse gas from cattle slurry applied to grassland soils in Ireland. Agric. Ecosyst. Environ. 188, 122–133. Chadwick, D.R., Pain, B.F., 1997. Methane fluxes following slurry applications to grassland soils: laboratory experiments. Agric. Ecosyst. Environ. 63, 51–60. Chadwick, D.R., Pain, B.F., Brookman, S.K.E., 2000. Nitrous oxide and methane emissions following application of animal manures to grassland. J. Environ. Qual. 29, 277–287. Chapuis-Lardy, L., Wrage, N., Metayz, A., et al., 2007. Soils, a sink for N2O? A review. Global Change Biol. 13, 1–17. Chan, A.S.K., Parkin, T.B., 2001. Effect of land use on methane flux from soil. J. Environ. Qual. 30, 786–797. Di, H.J., Cameron, K.C., 2002. The use of a nitrification inhibitor, dicyandiamide (DCD), to decrease nitrate leaching and nitrous oxide emissions in a simulated grazed and irrigated grassland. Soil Use Manage. 18, 395–403. Di, H.J., Cameron, K.C., Sherlock, R.R., 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. Dittert, K., Bol, R., King, R., et al., 2001. Use of a novel nitrification inhibitor to reduce nitrous oxide emissions from 15N labelled dairy slurry injected into the soil. Rapid Commun. Mass Spectrom. Rev. 15, 1291–1296. Ghosh, S., Majumdar, D., Jain, M.C., 2003. Methane and nitrous oxide emissions from an irrigated rice of North India. Chemosphere 51, 181–195. Hatch, D., Trindade, H., Cardenas, L., et al., 2005. Laboratory study of the effects of two nitrification inhibitors on greenhouse gas emissions from a slurry-treated arable soil: impact of diurnal temperature cycle. Biol. Fertil. Soils 41 (4), 225–232. Hénault, C., Devis, X., Page, S., et al., 1998. Nitrous oxide emissions under different soil and land management conditions. Biol. Fertil. Soils 26, 199–207. IPCC (Intergovernmental Panel on Climate Change), 2006. Guidelines for National Greenhouse Gas Inventories, prepared by the National Greenhouse Gas Inventories Programme. In: Eggleston, H.S, Buendia, L., Miwa, K. et al. (Eds.) Vol. 4, Chap. 11, N2O emissions from managed soils and CO2 emissions from lime and urea application, IGES, Hayama, Japan, 2006.

349

IPCC: Climate Change, 2013. The Physical Science Basis, Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change, edited by: Stocker et al. Cambridge University Press, Cambridge, UK and New York, NY, USA, 2013,1535 pp Kelliher, F.M., Clough, T.J., Clark, H., et al., 2008. The temperature dependence of dicyandiamide (DCD) degradation in soils: a data synthesis. Soil Biol. Biochem. 40, 1878–1882. Lindau, C.W., Bollich, P.K., Delaune, R.D., et al., 1993. Methane mitigation in flooded Louisiana rice fields. Biol. Fertil. Soils 15, 174–178. McGettigan M., Duffy P., Hyde B. et al., 2010. Irelands Greenhouse Gas Emissions in 2009. Environmental Protection Agency (EPA), Wexford. http://www.epa.ie/ downloads/pubs/air/airemissions/GHG_19902009_Provisional_2011.pdf. Merino, P., Estavillo, J.M., Graciolli, L.A., et al., 2002. Mitigation of N2O emissions from grassland by nitrification inhibitor and Actilith F2 applied with fertilizer and cattle slurry. Soil Use Manage. 18, 135–141. Ministry of Agriculture, Fisheries and Food, 1981. The Analysis of Agricultural Materials. Her Majesty’s Stationary Office, London, England. Mkhabela, M.S., Gordon, R., Burton, D., et al., 2006. Ammonia and nitrous oxide emissions from two acidic soils of Nova Scotia fertilised with liquid hog manure mixed with or without dicyandiamide. Chemosphere 65, 1381–1387. Mkhabela, M., Gordon, R., Burton, D., et al., 2006. Effect of lime, dicyandiamide and soil water content on ammonia and nitrous oxide emissions following application of liquid hog manure to a marshland soil. Plant Soil 284, 351–361. Pain, B.F., Misselbrook, T.H., Rees, Y.J., 1994. Effects of nitrification inhibitor and acid addition to cattle slurry on nitrogen losses and herbage yields. Grass Forage Sci. 49, 209–215. Perälä, P., Kapuinen, P., Esala, M., et al., 2006. Influence of slurry and mineral fertiliser application techniques on N2O and CH4 fluxes from a barley field in southern Finland. Agric. Ecosyst. Environ. 117 (1), 71–78. Pereira, J., Fangueiro, D., Chadwick, D.R., et al., 2010. Effect of cattle slurry pre-treatment by separation and addition of nitrification inhibitors on gaseous emissions and N dynamics: a laboratory study. Chemosphere 79, 620–627. Rajbanshi, S.S., Benckiser, G., Ottow, J.C.G., 1992. Effects of concentration, incubation temperature, and repeated applications on degradation kinetics of dicyandiamide (DCD) in model experiments with a silt loam soil. Biol. Fertil. Soils 13, 61–64. Rodhe, L., Pell, M., Yamulki, S., 2006. Nitrous oxide, methane and ammonia emissions following slurry spreading on grassland. Soil Use Manage. 22 (3), 229–237. SAS Institute Inc., 2011. SAS/STAT1 9.3 User’s Guide. Cary, NC: SAS Institute Inc. Scheehle E.A., Kruger D., 2006. Global anthropogenic methane and nitrous oxide emissions, Energy J., Multi-Greenhouse Gas Mitigation and Climate Policy, 2006 (special issue 3). Singh, J., Saggar, S., Bolan, N.S., 2009. Influence of dicyandiamide on nitrogen transformation and losses in cow urine amended soil cores from grazed pasture. Anim. Prod. Sci. 49, 253–261. Skiba, U., Smith, K.A., 2000. The control of nitrous oxide emissions from agricultural and natural soils. Chemosphere – Global Change Sci. 2, 379–386. Smith, L.C., deKlein, C.A.M., Catto, W.D., 2008. Effect of dicyandiamide applied in a granular form on nitrous oxide emissions from a grazed dairy pasture in Southland, New Zealand. New Zeal. J. Agric. Res. 51, 387–396. Søgaard, H.T., Sommer, S.G., Hutchings, N.J., et al., 2002. Ammonia volatilization from field-applied animal slurry – the ALFAM model. Atmos. Environ. 36, 3309–3319. Tao, X., Matsunaka, T., Sawamoto, T., 2008. Dicyandiamide application plus incorporation into soil reduces N2O and NH3 emissions from anaerobically digested cattle slurry. Aust. J. Exp. Agric. 48 (2), 169–174. Vallejo, A., Skiba, U., García-Torres, L., et al., 2006. Nitrogen oxides emission from soils bearing a potato crop as influenced by fertilization with treated pig slurries and composts. Soil Bio. Biochem. 38, 2782–2793. van Es, H.M., Sogbedji, J.M., Schindelbeck, R.R., 2006. Effect of manure application timing, crop, and soil type on nitrate leaching. J. Environ. Qual. 35, 670–679. Webb, J., Pain, B., Bittman, S., et al., 2010. The impacts of manure application methods on emissions of ammonia, nitrous oxide and on crop response – a review. Agric. Ecosyst. Environ. 137, 39–46. Weiske, A., Benckiser, G., Ottow, J.C.G., 2001. Effect of the new nitrification inhibitor DMPP in comparison to DCD on nitrous oxide (N2O) emissions and methane (CH4) oxidation during 3 years of repeated applications in field experiments. Nutr. Cycl. Agroecosyst. 60, 57–64. Welten, B.G., Ledgard, S.F., Luo, J., 2014. Administration of dicyandiamide to dairy cows via drinking water reduces nitrogen losses from grazed pastures. J. Agri. Sci doi:http://dx.doi.org/10.1017/S0021859613000634. Wulf, S., Maeting, M., Clemens, J., 2002. Application technique and slurry co-fermentation effects on ammonia, nitrous oxide and methane emissions after spreading: II. Greenhouse gas emissions. J. Environ. Qual. 31, 1795–1801. Zaman, M., Saggar, S., Blennerhassett, J.D., et al., 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.