Published as: O’ Flynn, C.J., Healy, M.G., Lanigan, G.J., Troy,... O. 2013. Impact of chemically amended pig slurry on greenhouse...

Published as: O’ Flynn, C.J., Healy, M.G., Lanigan, G.J., Troy,... O. 2013. Impact of chemically amended pig slurry on greenhouse...

Published as: O’ Flynn, C.J., Healy, M.G., Lanigan, G.J., Troy, S.M., Somers, C., Fenton,

O. 2013. Impact of chemically amended pig slurry on greenhouse gas emissions, soil properties and leachate. Journal of Environmental Management 128: 690-698.

http://dx.doi.org/10.1016/j.jenvman.2013.06.020

Impact of chemically amended pig slurry on greenhouse gas emissions, soil properties and leachate

Cornelius J. O’ Flynn a

, Mark G. Healy

a*

, Gary J. Lanigan b

, Shane M. Troy c

, Cathal Somers b

,

Owen Fenton b a

Civil Engineering, National University of Ireland, Galway, Co. Galway, Ireland. b

Teagasc, Environmental Research Centre, Johnstown Castle, Co. Wexford, Ireland. c

Scottish Rural College, Roslin Institute Building, Edinburgh, UK.

*Corresponding author. Tel: +353 91 495364; fax: +353 91 494507. E-mail address: [email protected]

Abstract

The effectiveness of chemical amendment of pig slurry to ameliorate phosphorus (P) losses in runoff is well studied, but research mainly has concentrated only on the runoff pathway

. The aims of this study were to investigate changes to leachate nutrient losses, soil properties and greenhouse gas (GHG) emissions due to the chemical amendment of pig slurry spread at 19 kg total phosphorus (TP), 90 kg total nitrogen (TN), and 180 kg total carbon (TC) ha

-1

. The amendments examined were: (1) commercial grade liquid alum (8% Al

2

O

3

) applied at a rate of 0.88:1 [Al:TP] (2) commercial-grade liquid ferric chloride (38% FeCl

3

) applied at a rate of

0.89:1 [Fe:TP] and (3) commercial-grade liquid poly-aluminium chloride (PAC) (10% Al

2

O

3

) applied at a rate of 0.72:1 [Al:TP].

Columns filled with sieved soil were incubated for 8 mo at

10 o

C and were leached with 160 ml (19 mm) distilled water wk

-1

.

All amendments reduced the

Morgan’s phosphorus and water extractable P content of the soil to that of the soil-only treatment, indicating that they have the ability to reduce P loss in leachate following slurry application. There were no significant differences between treatments for nitrogen (N) or carbon (C) in leachate or soil, indicating no deleterious impact on reactive N emissions or soil

C cycling.

Chemical amendment posed no significant change to GHG emissions from pig slurry, and in the cases of alum and PAC, reduced cumulative N

2

O and CO

2

losses. Chemical amendment of land applied pig slurry can reduce P in runoff without any negative impact on nutrient leaching and GHG emissions. Future work must be conducted to ascertain if more significant reductions in GHG emissions are possible with chemical amendments.

Keywords: pig slurry; P sorbing amendments; Water Framework Directive; nitrate

1. Introduction

The European Union Water Framework Directive (EU WFD) (European Commission (EC),

2000) aims to achieve ‘at least’ good ecological status for all water bodies, including rivers, lakes, groundwater, estuaries and coastal waters, in all member states by 2015. To meet this objective, Programmes of Measures (POM) must be implemented in all EU member states. In

Ireland, POM are enacted by the Nitrates Directive (European Economic Community, 1991), which, amongst other measures, limits the magnitude, timing and placement of inorganic fertilizer and organic manure applications to land.

In Ireland, as part of the National Action Programme (NAP) to address the requirements of the EU WFD, the maximum amount of livestock manure that may be spread on land, together with manure deposited by the livestock, cannot exceed 170 kg nitrogen (N) ha

-1

yr

-1

and 49 kg phosphorus (P) ha

-1

yr

-1

. This limit is dependent on grassland stocking rate and soil test phosphorus (STP; based on plant available Morgan’s P (Pm)). Soil P Index categories of 1

(deficient) to 4 (excessive) are used to classify STP concentrations in Ireland (Schulte et al.,

2010). Phosphorus losses from P Index 4 soils have the potential to become exported along the transfer continuum within a catchment, and may adversely affect surface and groundwater quality (Wall et al., 2011). The amount by which these limits can be exceeded will be reduced gradually to zero by January 1, 2017. These new regulations will have an impact on the pig industry in particular, as it is focused in relatively small areas of Ireland, and will, in effect, reduce the amount of land available for the application of pig slurry. This may lead to the need for pig slurry export, which is energetically questionable at distances over 50 km

(Fealy and Schroder, 2008).

Landspreading is currently the most cost effective treatment option for pig slurry in Ireland

(Nolan et al., 2012). Due to the high concentrations of pig farming in certain areas, in the midlands and south of the country especially, the constant application of pig slurry results in certain fields (those nearest the farm or the most suitable areas for spreading (Wall et al.,

2011)) becoming high in STP, which may take years-to-decades to be reduced to agronomically optimum levels (Schulte et al., 2010).

When applications of pig slurry are followed by rainfall events, incidental (short-term), diffuse transfers of P and N may occur in runoff. Losses of both P and N may also occur through leaching, which ultimately could have adverse consequences for water bodies

(McDowell and Sharpley, 2001; Fenton et al., 2011; Sophocleous, 2011). Karstified aquifers, which are overlain by free-draining soils, are particularly susceptible to groundwater pollution, as they have less attenuation potential than surface runoff pathways and there is a high potential for macropore flow of dissolved and particulate forms of P (Kramers et al.,

2012). In Ireland, karstified limestone covers approximately 20% of the area of the country

(Daly, 2005), and much pig farming is conducted in karst-covered areas.

Chemical amendment of pig slurry has been shown to be an effective means of reducing surface runoff of P and suspended sediment (SS) by numerous researchers (Smith et al.,

2001, 2004; Dou et al, 2003), but as yet, the role pig slurry amendments have to play in controlling leached losses has not been investigated. O’ Flynn et al. (2012a, 2012b) examined the effectiveness and feasibility of different chemical amendments, added to pig slurry, in reducing P, SS and metal concentrations in a series of laboratory studies, conducted first at bench scale (O’ Flynn et al., 2012a) and then using a laboratory rainfall simulator (O’ Flynn et al., 2012b). In the latter study, O’ Flynn et al. (2012b), found additions of alum, ferric chloride (FeCl

3

) and poly-aluminium chloride (PAC) reduced total phosphorus (TP) and SS losses in surface runoff, without posing a significant risk of metal losses.

Although there has been much work done on the chemical amendment of surface applied pig slurry, there is an absence of work investigating any potential negative impact that this may have on N and carbon (C) losses and on greenhouse gas (GHG) emissions. Brennan et al.

(2012) found in a plot study that chemical amendment of dairy cattle slurry with PAC reduced ammonium-N (NH

4

+

-N) runoff losses, but alum and lime led to increased NH

4

+

-N losses. All amendments reduced P losses in runoff, but had no effect on nitrate (NO

3

-

-N) runoff losses. The Intergovernmental Panel on Climate Change (IPCC) (2007) estimates that

agricultural activities, including land application of animal manures, account for about 20% of the anthropogenic global warming budget, with emissions principally comprised of methane (CH

4

) from enteric fermentation and manure management and nitrous oxide (N

2

O) from N application to soils. The EU 2020 Climate and Energy Package and its associated

Effort-Sharing Decision (Decision No 406/2009/EC) envisages reducing GHG emissions by

20% by 2020 across the whole of the EU. Whilst previous work has investigated the impact of chemical amendments to pig slurry to reduce P in runoff (O’ Flynn et al., 2012a, 2012b), no study has investigated the impact of chemical amendment of pig slurry on GHG emissions.

Therefore, the aims of this laboratory study were to investigate if chemical amendment of pig slurry: (1) reduced leached losses of N, P and C from a low P index soil (2) resulted in changes to soil properties at different time intervals during the study period and (3) led to a reduction in GHG emissions over 28 d from the time of application.

2. Materials and Methods

2.1 Slurry collection and characterisation

Pig slurry was taken from an integrated pig unit in Teagasc Research Centre, Moorepark,

Fermoy, Co. Cork, Rep. of Ireland in September 2011. The sampling point was a valve on an outflow pipe between two holding tanks, which were sequentially placed after a holding tank under the slats on which no bedding materials were used. To ensure a representative sample, this valve was turned on and left to run for a few minutes before taking a sample. The slurry was stored in a 25-L drum inside a cold-room fridge at 10 o

C prior to testing. The TP and total

nitrogen (TN) were determined using persulfate digestion. Ammonium-N was determined by adding 50 ml of slurry to 1L of 0.1M HCl, shaking for 30 min at 200 rpm, filtering through

No. 2 Whatman filter paper, and analysing using a nutrient analyser (Konelab 20, Thermo

Clinical Labsystems, Finland). Total carbon was measured using a nutrient analyser

(Biotector, BioTector Analytical Systems Ltd, Ireland). Slurry pH was determined using a pH probe (WTW, Germany). Dry matter (DM) content was determined by drying at 105 o

C for 24 h. The physical and chemical characteristics of the pig slurry used in this experiment and characteristic values of pig slurry from other farms in Ireland are presented in Table 1.

2.2 Pig slurry amendment

Amendments for the present study were chosen based on effectiveness of P sequestration and feasibility criterion (cost and potential environmental impediments) determined by O’ Flynn et al. (2012a, 2012b). The amendment rates, which were applied on a stoichiometric basis, were: (1) commercial grade liquid alum (8% Al

2

O

3

) applied at a rate of 0.88:1 [Al:TP] (2) commercial-grade liquid ferric chloride (38% FeCl

3

) applied at a rate of 0.89:1 [Fe:TP], and

(3) commercial-grade liquid PAC (10 % Al

2

O

3

) applied at a rate of 0.72:1 [Al:TP].

Amendments were added to slurry in a 100-ml plastic cup and mixed for 10 s. The compositions of the amendments used are shown in Table 2.

2.3 Soil collection and analysis

A sample of the plough layer (top 0.2 m) of an acid brown earth soil was collected from a tillage farm in Fermoy, Co. Cork, Republic of Ireland. The site is typical of a free draining soil, underlain by a karstified limestone aquifer. Tillage soil was chosen, as this type of soil is

often of a lower P index and is more suitable for the landspreading of pig manure. The soil was air-dried, sieved (<2 mm) and thoroughly mixed. Soil samples (n=3) were oven dried at

40 °C for 72 h, crushed to pass a 2 mm sieve and analysed for Morgan’s P (Pm, the national test used for the determination of plant available P in Ireland) using Morgan’s extracting solution (Morgan, 1941). Soil total carbon (TC) and TN were determined by high temperature combustion using a LECO Truspec CN analyser (LECO Corporation, St. Joseph,

MI, USA). Soil pH (n=3) was determined using a pH probe (WTW, Germany) and a 2:1 ratio of deionised water-to-soil. The STP of the sample used in the column and batch experiments was 3.21±0.29 mg L

-1

(making it a P index 2 soil according to S.I. No. 610 of 2010), total potassium (TK) of 41.8±3.00 mg L

-1

, TC of 1.84±0.05 %, TN of 0.19±0.00 %, C:N ratio of

9.87±0.22, a pH of 6.26±0.13, an organic matter (OM) content of 4.68±0.14%. A low range

STP tillage soil was chosen for this experiment to avoid the risk of background P from a high range STP soil ‘masking’ the effect of each treatment. A low range STP tillage soil was also chosen as present and future regulations will have the effect of making this type of land more preferable for pig slurry spreading in the future.

The particle size distribution was determined using a sieving and pipette method (B.S.1377-2;

British Standards Institution (BSI), 1990a) and the organic content of the soil was determined using the loss on ignition (LOI) test (B.S.1377-3; BSI, 1990b). The unstructured soil in the column and batch experiments consisted of 57% sand, 29% silt and 14% clay, giving it a sandy loam texture.

During any interaction with chemically amended slurry, the background soil P adsorption rate must also be considered and can be assessed in a batch experiment following the procedure outlined by Fenton et al. (2009). Ortho-phosphorus (PO

4

3-

-P) solutions (90 ml), synthesised

using dissolved potassium phosphate (KH

2

PO

4

) in distilled water, ranging in concentration from 4.1 to 28.9 mg P L

-1

, were added to 5 g samples of soil and shaken for 24 h using an end-over-end shaker. Samples were passed through 0.45-µm syringe filters prior to being analysed colorimetrically for dissolved reactive phosphorus (DRP) using a nutrient analyser

(Konelab 20, Thermo Clinical Labsystems, Finland). A Langmuir isotherm was used to estimate the mass of P adsorbed per mass of the soil (McBride, 2000):

C e x m

1

ab

C e b

[1] where C e

is the concentration of P in solution at equilibrium (mg L

-1

), x/m is the mass of P adsorbed per unit dry weight of soil (g kg

-1

), a is a constant related to the binding strength of molecules onto the soil, and b is the maximum adsorption capacity of the soil (g kg

-1

). In conjunction with the P adsorption capacity of the soil, the equilibrium P concentration of the soil (EPC

0

) (i.e. the point where no net desorption or sorption occurs) was derived using

(Olsen and Watanabe, 1957):

S′= k d

C−S

0

[2] where S′ is the mass of P adsorbed from slurry (mg kg

-1

), C is the final P concentration of the solution, k d

is the slope of the relationship between S′ and C, and S

0

is the amount of P originally sorbed to the soil (mg L

-1

). The mass of P adsorbed per unit dry weight of soil was

0.224 g P kg

-1

and the soil’s EPC

0

was 0.513 mg L

-1

.

Soil water holding capacity (WHC) was determined according to Cassel and Nielsen (1986).

Soil was placed on a funnel whose sides were covered with Whatman no. 2 filter paper, and distilled water was added to the soil until it became completely saturated. Saturated soil was weighed, oven-dried overnight at 105 o

C, and weighed again.

Water-filled pore space, which can impact on rates of denitrification in soil, was estimated in accordance with Haney and Haney (2010):

WFPS

WC n

*

b

[3] where ρ b

is bulk density and n is total porosity (mineral density was taken as 2.65 g cm

-3

).

Mineral N in soil (NH

4

+

-N, NO

3

-

-N and nitrite-N (NO

2

-

-N)) was determined at 0, 7 and 28 d after land application of pig slurry by adding 20 g of soil to 2 M KCl, shaking for 1 h, filtering through No. 2 Whatman filter paper, and testing using a nutrient analyser (Konelab

20, Thermo Clinical Labsystems, Finland). Extra soil columns (n=3 for each treatment) were set up to allow sampling after 7 d for soil mineral N.

2.4 Experimental columns

The experiment was conducted in 0.3-m-deep and 0.104-m-internal diameter columns with a perforated stop-end inserted at the base to ensure that the soil remained free draining. A 0.05m layer of gravel, with a grain size of 5 – 10 mm, was placed at the base of each column.

Sieved soil (< 2 mm), previously mixed with distilled water to achieve a water content (WC) of 26% (to replicate the average in situ field condition of the soil), was placed in 0.05 m-deep increments in each column, so as the average dry bulk density was approximately 1.1 g cm

-3

(equivalent to field conditions) and the total depth of soil was 0.2 m. At each depth increment, soil was pressed along the wall of the column to avoid preferential flow

(Bhupinder Singh, pers. comm.).

The following treatments were examined: (1) soil only with no slurry applied (2) soil with unamended slurry applied (the study control) and (3) soil receiving amended slurry. Slurry was spread at 19 kg TP, 90 kg TN, and 180 kg TC ha

-1

. Columns were stored in a controlled environment for 8 mo at 10 o

C at 75% humidity, based on typical climatic conditions in

Ireland (Walsh, 2012). All columns received 160 ml of distilled water wk

-1

, applied twice weekly in two 80 ml increments over 2 h. This is equivalent to 980 mm of rainfall yr

-1

, or 19 mm wk

-1

, which would be in the mid-range of average annual rainfall amounts in Ireland

(Walsh, 2012). This application rate remained constant for the duration of the study; however, actual rainfall rates will vary considerably over the course of a year. Drainage water leachate was collected in plastic containers via funnels positioned under the perforated stopend of each column.

2.5 Leachate collection and analysis

The leachate from each column was composited and sampled weekly. Upon collection, samples were weighed and a subsample was passed through a 0.45-µm filter and analysed colorimetrically for DRP, NO

2

-

, NH

4

+ and total oxidized nitrogen (TON) using a nutrient analyser (Konelab 20, Thermo Clinical Labsystems, Finland). Nitrate was calculated by subtracting NO

2

-

from TON. Filtered and unfiltered subsamples were tested for total dissolved phosphorus (TDP) and TP using acid persulfate digestion. Particulate phosphorus

(PP) was calculated by subtracting TDP from TP. Dissolved un-reactive phosphorus (DUP)

10 

was calculated by subtracting DRP from TDP. Total nitrogen, total organic carbon (TOC) and total inorganic carbon (TIC) were measured using a nutrient analyser (Biotector,

BioTector Analytical Systems Ltd, Ireland). Total carbon was calculated by adding TIC and

TOC. Leachate pH was determined using a pH probe (WTW, Germany). This addressed the first aim of the study.

2.6 Destructive soil sampling

Soil columns were destructed after 1, 2, 3, 6 and 8 mo (n=3 for each treatment, at each time period) and tested for WC, OM, pH, water extractable P (WEP), Pm, TN and TC. Before analyses, each column was divided into 3 layers (0 to 0.05 m, 0.05 to 0.1 m, and 0.1 to 0.2 m from the surface). Organic matter content of the soil was determined using the LOI test

(B.S.1377-3; BSI, 1990b). Soil pH was determined using a pH probe (WTW, Germany) and a

2:1 ratio of deionised water-to-soil. Water extractable P was measured by shaking 5 g of soil in 25 ml of distilled water for 30 min, filtering through a 0.45-μm syringe filter, prior to being analyzed colorimetrically for DRP (McDowell and Sharpley, 2001) using a nutrient analyser

(Konelab 20, Thermo Clinical Labsystems, Finland). Morgan’s P was determined using

Morgan’s extracting solution (Morgan, 1941). Soil TC and TN were determined for the middle layer only in each column (0.05 to 0.1-m-depth) by high temperature combustion using a LECO Truspec CN analyser (LECO Corporation, St. Joseph, MI, USA). This addressed the second aim of the study.

2.7 Greenhouse gas emissions

11 

Direct GHG emissions (N

2

O, carbon dioxide (CO

2

) and CH

4

) were analyzed over a 28-d period in accordance with Troy et al. (2013). Samples were taken on the day of slurry application (day 1) and subsequently on days 2, 3, 4, 5, 6, 7, 9, 11, 13, 15, 19, 23 and 28. The top of the PVC columns were sealed using a rubber stopper. A sample of the air in the headspace above the columns was taken through a rubber septum using a polypropylene syringe with a hypodermic needle. The sample was immediately transferred into a preevacuated 7-ml screw cap septum vial. Samples were taken at 0, 5, 10 and 20 min after the sealing of columns with a rubber stopper. After this period, the rubber stopper was removed.

Nitrous oxide, CO

2

and CH

4

concentrations were analysed using a gas chromatograph

(Varian CP 3800 GC, Varian, USA) fitted with a 63Ni electron capture detector (ECD) for

N

2

O analysis, a thermal conductivity detector (TCD) for CO

2

analysis and a flame ionization detector (FID) for CH

4

analysis. During the analysis, 0.7 ml of a sub-sample from each vial was drawn and injected first into a magnesium perchlorite (14-22 mesh) packed pre-column to remove any moisture, followed by a 3-m-long, 3-mm-outside diameter stainless steel column packed with Poropak Q (80/100 mesh). The column oven and injector temperature were both 60°C and the detector temperature was 350°C. Argon (BOC Gases, Ireland), flowing at 35 ml min

-1

, was used as a carrier gas. Samples were fed into the system by a

Combi-Pal automatic sampler (CTC Analysis, Switzerland) controlled by computer software.

Two-thirds of the injected sample was split to the ECD detector and one-third to the TCD and

FID in series. This allowed the simultaneous measurement of all three gases from the one sample. Areas under the peaks were integrated using Star Chromatography Workstation

(Varian, USA). Fluxes were calculated from the change in headspace concentration over measured period using:

dGas

10

x dt

V chamber

p

R

100

T

MW

10

y

1

A

[4]

12 

where: dGas is measured in ppm or ppb to get concentration at a certain point in time or ppm h

-1

or ppb h

-1

to get the change in concentration over time; 10 x

is a recalculation (10

-6

if starting from ppm or 10

-9

if starting from ppb); V

chamber

is the volume of the chamber used; p is atmospheric pressure; MW is the molecular weight either of N or N

2

O, depending of which compound in which the emissions are expressed; R is a gas constant, 8314 J mol

-1

K

-1

; T is temperature in Kelvin; 10 y

is a recalculation (10

3

if the results are expressed in mg or 10

6

if in

µg); and A is the area of the chamber. The fluxes were then converted into mg m

-2 d

-1

. Mean daily emissions rates were calculated for each replicate by interpolation of values in between the measurement days using arithmetic means (Velthof and Oenema, 1995; Flechard et al.,

2007). This addressed the third aim of the study.

2.8 Statistical analysis

The data was analysed in SPSS 20 (IBM, 2011) using a general linear model. Mean values of:

WC; OM; soil P, N and C species; soil pH; leachate P, N and C species; leachate pH; and

GHGs were analysed in a multivariate Tukey analysis when soil-only, slurry-only (the study control), and slurry treated with alum, PAC and FeCl

3

were applied. Data met the normal distributional assumptions required. Probability values of p>0.05 were deemed not to be significant.

3. Results

3.1. Water content, organic matter and soil pH

13 

The WHC of the soil was found to equate to a WC of 53%. In general, there were no significant differences observed in WC between treatments, apart from at 1 mo in the top soil layer, where the soil-only treatment had a WC of 30.33±0.24% (data not shown).

Comparatively, at the same time, slurry-only, alum, FeCl

3

and PAC treatments had WCs of

31.76±0.44%, 32.45±0.35%, 31.89±0.78%, and 32.13±0.39%. Water contents increased with depth: WCs in the top soil layer were generally between 30 and 33%, between 31 and 34% in the middle layer, and between 35 and 38% in the bottom layer. These equated to water-filled pore space (WFPS) values of between 56 and 62% in the top layer, between 58 and 64% in the middle layer, and between 65 and 72% in the bottom layer. Organic matter (generally between 4.3 and 4.7%) and soil pH (between 6 and 6.5) were not significantly affected by treatment, depth or time.

3.2. Nitrogen leachate and soil properties

There were no statistically significant differences between treatments for TN in soil (Table

3). No significant differences between treatments were observed for the N in leachate water, which mainly comprised NO

3

-

. The amount of NO

3

-

leached increased rapidly until wk 2, before it reduced gradually thereafter (Fig. 1 c). Approximately 95% of TN leached from the columns over the duration of the studies was in the form of NO

3

-

, with roughly 0.2% in the form of NO

2

-

and 0.3% in the form of NH

4

+

. The C:N ratio for all treatments at all destructive periods was between 9 and 10 (Table 3). Nitrite loads peaked between wks 10 and 26 (Fig. 1 b).

At all times, mineral N in soil comprised less than 2% of soil TN. Seven days after application, soil NH

4

+

was observed to be highest for the alum and FeCl

3

treatments (83.7

14 

and 79.3 g NH

4

+

-N kg

-1

soil, respectively). This compared with values of 44.0 and 48.9 g

NH

4

+

-N kg

-1

soil for soil-only and slurry-only, respectively.

3.3. Nitrous oxide emissions

Nitrous oxide emissions from the soil-only treatment remained fairly constant throughout the

28-d study (Fig. 2 a), with cumulative emissions of 22±8 mg N

2

O-N m

-2

. Application of pig slurry led to an increased cumulative release of N

2

O. Cumulative emissions across all Napplied treatments were high, ranging approximately from 60 to 200 mg N

2

O-N m

-2

. The highest cumulative losses of 188±86 mg N

2

O-N m

-2 was observed for FeCl

3

-amended slurry and this was the only treatment statistically significantly different (p=0.008) to soil-only, but was not statistically significantly different to any other treatment. Cumulative emissions from all treatments remained relatively constant between 4 and 7 d after application of slurry, at which point they increased more rapidly, although not significantly, and continued to rise until the end of the study. However, N

2

O losses from FeCl

3

–amended slurry were at all times greater than all other treatments. Alum and PAC-amended slurries both had less, but not statistically significantly different, N

2

O losses than unamended slurry, but more than soilonly.

3.4. Phosphorus leachate and soil properties

There were no significant differences in the quantity of P leached between treatments, with the majority of TP made up of TDP for all treatments. Particulate phosphorus comprised approximately 30% of the TP load in all cases.

15 

In general, there were no significant differences in levels of Pm and WEP between treatments in the bottom two soil layers (Table 3). However, in the top soil layer, application of unamended slurry resulted in increased Pm and WEP, which were significantly different

(p<0.05) to the soil-only columns at all destructive periods (Table 3). Levels of Pm and WEP in the top soil layer were both reduced by the application of amended slurry to levels not significantly different to soil-only columns (Table 3).

3.5. Carbon leachate

The average cumulative amount of TOC and TIC leached is shown in Fig. 3 a. The average

TC leached from the soil-only columns was 217.3 mg. This increased to 253 mg from columns with unamended slurry, with reduced amounts of TC leached from columns treated with amended slurry. However, there were no statistically significant differences for TC loads between treatments. There was an increase in loads of TC leached from wk 1 to wk 2 (Fig. 3 b); however, this was due to lower leachate volumes during wk 1 than wk 2, rather than any changes in concentration. The loads of TC leached then decreased after wk 2 until the end of the study, during which time there was no significant change in flows.

3.6. Carbon emissions

Emissions of CO

2

followed a similar trend to N

2

O emissions (Fig. 2 b). The soil-only treatment had the lowest emissions, with cumulative losses of 36±4 g CO

2

-C m

-2

. Losses increased upon application of slurry, but were only statistically significantly different

(p=0.008) in the case of FeCl

3

-amended slurry, which had cumulative losses of 106±23 g

CO

2

-C m

-2

. However, this was not statistically significantly different to any other unamended

16 

or amended slurry treatment. Alum and PAC-amended slurries had less, but not statistically significant different, losses than unamended slurry. Methane losses were highly variable (Fig.

2 c), but no treatment had significantly higher losses than the soil-only treatment. After 5 days, all treatments either gained or lost CH

4, with FeCl

3

–amended slurry acting overall as a net sink with cumulative losses of -13±7 mg CH

4

-C m

-2

, whilst PAC-amended slurry had cumulative losses of 13±6 mg CH

4

-C m

-2

.

4. Discussion

4.1 Nitrogen leachate and soil properties

Denitrification is the mainly microbial reduction of NO

3

-

-N to the gaseous products nitric oxide (NO), N

2

O, or inert di-nitrogen (N

2

). Some studies have shown that the highest rates of denitrification occur in the upper soil horizon (Kustermann et al., 2010; Jahangir et al., 2012), the extent of which depends on WC and WFPS. Soil WC can impact on many different soil processes such as mineralization, leaching, plant uptake and denitrification (Porporato et al.,

2003).

The early peak in NO

3

-

loss may be due to the drying and re-wetting during column construction, which could have caused a surge in microbial activity and C and N mineralisation (Van Gestel et al., 1991; Bengtsson et al., 2003). This may also have led to an early peak in leachate NH

4

+

(Fig. 1 a). Once rewetting was complete, WFPS levels were between 65 and 72% in the bottom layer. At WFPS levels of over 60%, denitrification may take place, releasing nitrogen gas (N

2

) and N

2

O into the atmosphere (Porporato et al., 2003).

Aerobic microbial activity and nitrification is also reduced in these anaerobic conditions

17 

where denitrification is facilitated (Poporato et al., 2003; Rivett et al., 2008). The fractions of

NO

2

-

, NO

3

-

and NH

4

+

in the leachate would seem to indicate that almost complete nitrification occurred, and also led to the drop in NO

3

-

levels after wk 2. This hypothesis was also supported by the C:N ratios present (Table 3). Soil with C:N ratios below 20 can be characterised as having a surplus of available NH

4

+

for nitrification (Bengtsson et al., 2003).

The peak in NO

2

-

between wks 10 and 26 may have been due to a delay in reduction of NO

2

during denitrification down to the preference of denitrifiers for NO

3

-

, even when both are present (Rivett et al., 2008).

High NH

3

+

volatilization may occur after land application of pig slurry, with over 60% of total losses occurring in the first 10 h after application (Gordon et al., 2001; Rochette et al.,

2001). It would appear in the current study that a large amount of volatilization occurred from both amended and unamended slurry treatments with little unvolatilised inorganic N remaining, which is in agreement with previous studies (Morvan et al., 1997; Hoekstra et al.,

2010; Hoekstra et al., 2011). Indeed, these rates of volatilization may represent a loss of 50-

80% of total ammoniacal nitrogen from landspread slurry over a 10-d period (Misslebrook et al., 2005a, 2005b; Meade et al., 2011). The slurry organic fraction was undetectable in leachate or soil (Table 3) due to the large background amounts of soil inorganic N, which was a result of the occurrence of mineralization. Unlike the present study, which found no significant difference between NO

3

-

losses from columns with and without slurry spread on them, Daudén et al. (2004) found that drainage NO

3

-

concentrations and loads consistently increased with increasing amount of N applied when landspreading pig slurry and mineral fertiliser between 275 and 1487.5 kg N ha

-1

. However, the spreading rate used by Daudén et al. (2004) was much higher than in the present study (90 kg N ha

-1

), and in that study, pig slurry was incorporated into soil to minimise volatilization losses.

18 

4.2 Nitrous oxide emissions

The increased cumulative release of N

2

O after slurry application was as expected (Velthof et al., 2003). The cumulative N

2

O emissions across all N-applied treatments represented a loss of between 1% and 3% of applied total ammoniacal N for a 28-d period. This was a higher emission factor than the IPCC default emission factor of 1% (IPCC, 2006). Generally, higher emission factors would not be associated with free-draining soil such as the one used in this study (Abdalla et al., 2009; Rafique et al., 2011). However, emission factors associated with slurry application have previously been observed to be higher than the default values and this may be related to the simultaneous application of a labile C source, which increases microbial activity (Dendooven et al., 1998; Sherlock et al., 2002). Nitrous oxide is produced by both nitrification and denitrification (Chadwick et al., 2011), and can be influenced by oxygen availability, soil WC, soil temperature, soil NO

3

-

and organic carbon content (Section 4.4)

(Velthof et al., 2003). The drying and rewetting of the soil during construction provided conditions which facilitated C and N mineralisation and denitrification, would also have facilitated N

2

O release to the atmosphere (Porporato et al., 2003).

The increase in N

2

O emissions associated with FeCl

3

addition may be explained as a result of ammonia volatilisation abatement. The difference in soil NH

4

+

levels between treatments 7 d after application was due to a reduction in volatilisation, possibly resulting from a reduction in slurry pH upon amendment addition. Previous work has observed that volatilisation may be reduced upon FeCl

3

addition, principally due to a reduction in slurry pH (Molloy and

Tunney, 1983).

19 

4.3 Phosphorus leachate and soil properties

Unlike previous runoff studies (O’ Flynn et al., 2012b), in which spreading of pig slurry led to a large increase in all types of P in runoff compared to runoff from soil-only, there were no significant differences in the quantity of P leached between treatments. The fraction of TP load made up of TDP was less when compared to O’ Flynn et al. (2012b), who found PP in runoff comprised, on average, 45% of TP. This is in agreement with McDowell et al. (2004), who found that more TP was lost as PP in overland than subsurface flow due to the higher kinetic energy and erosive power of high-frequency storms. Loss of P in subsurface flow is generally less than that in runoff, and will decrease as the degree of soil–water contact increases, due to sorption by P-deficient subsoils (Haygarth et al., 1998; McDowell et al.,

2004). Although a soil with a low Pm (3.21±0.29 mg L

-1

) was used in this experiment, its high adsorption capacity for P (0.224 g P kg

-1

) and low EPC

0

(0.513 mg L

-1

) facilitated adsorption of P during leaching.

The same amendments and application rates as used in the present study were also used by O’

Flynn et al. (2012a), who achieved reductions of between 95 and 99% in the WEP of slurry.

Dao (1999) amended stockpiled cattle manure with caliche, alum and flyash in an incubation experiment, and reported WEP reductions in amended manure, compared to the study control, of 21, 60 and 85%, respectively. Similarly, in a study that examined the effect of soil

P level in a silt loam soil which was incubated at 25°C, Kalbasi and Karthikeyan (2004) reported that applications of alum and FeCl

3

-amended slurry to soil decreased soil WEP. In the present study, due to the regular application of 160 ml water wk

-1

, which led to the downward leaching of P from the slurry, both Pm and WEP levels in the columns spread with unamended slurry reduced to levels closer, but still significantly different (p<0.05), to soil-

20 

only and amended slurry columns. This P was adsorbed by the soil’s high adsorption capacity for P, but was not detected by WEP or Pm analysis.

4.4 Carbon leachate and emissions

The decrease in loads of TC leached after wk 2 may have been due to the increased mineralization of C and N, which may have been the cause of increased losses of CO

2

to the atmosphere. This loss of CO

2

to the atmosphere may also be the reason that there were statistically no significant differences between treatments for TC in soil (Table 3). In addition, organic carbon can act as an electron donor to facilitate the occurrence of denitrification when anaerobic conditions are present (Rivett et al., 2008).

The addition of manure slurries to soil has been shown to cause an increase in microbial activity and CO

2

emissions (Bol et al., 2004; Dumale et al., 2009; Cayuela et al., 2010). The increased CO

2

losses from unamended or amended slurry treatments were in agreement with the hypothesis that these losses were the cause for no statistically significant differences between slurry treatments for TC in soil (Table 3).

After land application, CH

4

emissions are generally of minor importance compared to N

2

O emissions (Wulf et al., 2002a, 2002b), as CH

4

emissions from enteric fermentation and during slurry storage are much more important (Chadwick et al., 2000). This is due to CH

4 being produced by decomposition of OM in faecal matter under anaerobic conditions. After landspreading, OM is oxidised to CO

2

and H

2

O in the aerobic conditions present. Mineral grassland soils are known to generally be a CH

4

sink, due to either oxidation of CH

4

to CO

2 in soils or incorporation into microbial biomass, with uptake rates ranging from 0.5 – 3.3 mg

21 

CH

4

m

-2

d

-1

(Mosier et al., 1991; Dobbie et al., 1996; Saggar et al., 2008). The results from the present study show that no additional risk to CH

4

emissions is posed by the chemical amendment of pig slurry.

4.5 Outlook for use of chemical amendment as a mitigation measure

Increased intensification of pig farming activities, along with legislation reducing the amount of land onto which pig farmers may apply slurry, has meant that the pig industry is under increasing pressure to reconcile production and water quality objectives. Land application of pig slurry is currently the most cost-efficient method for its disposal. In Ireland, the pig industry is concentrated in a small number of areas, with typically high stocking rates.

Therefore, the disposal of slurry in a cost-effective and environmentally responsible way is a serious issue for farmers.

This study demonstrates that amendments previously selected on the basis of ability to reduce runoff P (O’ Flynn et al., 2012a,b), may be used without posing a negative impact on leachate, soil properties, and GHG emissions.

Based on the results of the current study and also previous work by the authors comparing cost (O’ Flynn et al., 2012a) and surface runoff losses (O’ Flynn et al., 2012b), PAC appears to be the most suitable amendment with which to chemically amend pig slurry. Ferric chloride resulted in increased N

2

O and CO

2

losses, whereas alum and PAC resulted in reduced, but not significantly different, losses to slurry-only. Poly-aluminium chloride performed best in overall removal of runoff P and SS (O’ Flynn et al., 2012b). There was little difference between leachate losses and soil effects from alum and PAC-amended slurry,

22 

although this study only included one soil type. The current study used a low STP soil so as to avoid the risk of background P from a high range STP soil ‘masking’ the effect of each treatment. However, future work must examine a wide variety of soil types, including high

STP soils. These amendments must also be examined at field-scale, and include repeated application and incorporation. Costs were comparable (O’ Flynn et al., 2012a), with estimated costs of amending and spreading amended slurry of €3.33 and €3.69 m

-3

for alum and PAC, respectively, in comparison to €1.56 m

-3

to spread unamended slurry.

In the current study, reductions were not adequate to satisfy the EU 2020 Climate and Energy

Package of reducing GHG emissions by 20% across the whole of the EU by 2020. It has however, been shown that some reductions are possible, and future work must be carried out to identify if more significant reductions in GHG emissions is possible at different application rates.

At present, there is no provision in legislation for chemical amendments to be used as a mitigation measure in the land application of pig slurry, but if they are to be utilised, a regulatory framework will need to be introduced by the relevant bodies.

5. Conclusions

Chemical amendment of land applied pig slurry can reduce P in runoff without any negative impact on nutrient leaching. Furthermore, there were no significant differences between treatments for N and C in leachate or soil, indicating no deleterious impact on reactive N emissions or soil C cycling. Chemical amendment posed no significant change to GHG emissions from pig slurry, and in the cases of alum and PAC, reduced cumulative N

2

O and

23 

CO

2

losses. Moreover, increased N

2

O emissions associated with FeCl

3

addition were likely to be due to a reduction in ammonia volatilisation, a theory supported by an increase in soil

NH

4

+

concentrations.

Acknowledgements

The first author gratefully acknowledges the award of the EMBARK scholarship from the

Irish Research Council to support this study. The authors would like to thank Dr. Raymond

Brennan, Dr. John Regan and Swann Lamarche.

24 

References

Abdalla, M., Jones, M., Smith, P., Williams, M., 2009. Nitrous oxide fluxes and denitrification sensitivity to temperature in Irish pasture soils. Soil Use Manage. 24, 376–388.

Bengtsson, G., Bengtson, P., Mansson, K.F., 2003. Gross nitrogen mineralization, immobilization, and nitrification rates as a function of soil C/N ratio and microbial activity. Soil Biol. Biochem. 35, 143-154.

Bol, R., Amelung, W., Friedrich, C., 2004. Role of aggregate surface and core fraction in the sequestration of carbon from dung in a temperate grassland soil. Eur. J. Soil Sci. 55, 71-77.

Brennan, R.B., Healy, M.G., Grant, J., Ibrahim, T.G., Fenton, O., 2012. Incidental phosphorus and nitrogen loss from grassland plots receiving chemically amended dairy cattle slurry. Sci. Tot. Environ. 441, 132-140.

British Standards Institution, 1990a. British Standard Methods of Test for Soils for

Civil Engineering Purposes. Determination of Particle Size Distribution. BS 1377.

BSI, London.

British Standards Institution, 1990b. Determination by Mass-loss on Ignition. British

Standard Methods of Test for Soils for Civil Engineering Purposes. Chemical and

Electrochemical Tests. BS 1377. BSI, London.

25 

Cassel, D.K., Nielsen, D.R., 1986. Field capacity and available water capacity, in: Klute, A.

(Ed.), Methods of soil analysis. Part 1. Second ed. Agron. Monogr. 9. ASA and SSSA,

Wisconsin. pp. 901–915.

Cayuela, M. L., Oenema, O., Kuikman, P.J., Bakker, R.R., van Groenigen, J.W., 2010.

Bioenergy by-products as soil amendments? Implications for carbon sequestration and greenhouse gas emissions. GCB Bioenergy. 2, 201–213.

Chadwick, D.R., Pain, B.F., Brookman, S.K.E., 2000. Nitrous oxideand methane emissions following application of animal manures to grassland. J. Environ. Qual. 29, 277–287.

Chadwick, D.R., Sommer, S., Thorman, R., Fangueiro, D., Cardenas, L., Amon, B.,

Misselbrook, T., 2011. Manure management: Implications for greenhouse gas emissions. J.

Anim. Feed Sci. 166-167, 514–531.

Daly, D. 2005. The characterisation and analysis of Ireland’s river basin districts: groundwater aspects, in: Proceedings of International Association of Hydrologists (Irish

Group) Seminar, Tullamore, Co. Offaly, Ireland. pp. 141-150.

Dao, T.H., 1999. Co-amendments to modify phosphorus extractability and nitrogen/phosphorus ration in feedlot manure and composted manure. J. Environ. Qual. 28,

1114–1121.

Daudén, A., Quílez, D., Vera, M.V., 2004. Pig slurry application and irrigation effects on nitrate leaching in Mediterranean soil lysimeters. J. Environ. Qual. 33, 2290–2295.

26 

Dendooven, L., Bonhomme, E., Merckx, R., Vlassak, K., 1998. N dynamics and sources of

N

2

O production following pig slurry application to a loamy soil. Biol. Fertil. Soils. 26, 224–

228.

Dobbie, K.E., Smith, K.A., Priemé, A., Christensen, S., Degorska, A., Orlanski, P., 1996.

Effect of land use on the rate of methane uptake by surface soils in northern Europe. Atmos.

Environ. 30, 1005-1011.

Dou, Z., Zhang, G.Y., Stout, W.L., Toth, J.D., Ferguson J.D., 2003. Efficacy of alum and coal combustion by-products in stabilizing manure phosphorus. J. Environ. Qual. 32, 1490–

1497.

Dumale, W. A., Miyazaki, T., Nishimura, T., Seki, K., 2009. CO2 evolution and short-term carbon turnover in stable soil organic carbon from soils applied with fresh organic matter.

Geophys. Res. Lett. 36, L01301.

European Economic Community, 1991. Council Directive of 12 December 1991 concerning the protection of waters against pollution by nitrates from agricultural sources (91/676/EEC).

Available at www.environ.ie/en/Environment/Water/WaterQuality/NitratesDirective/

(accessed 27.03.2013). Dep. of the Environ., Dublin, Ireland.

European Commission, 2000. Council Directive of 22 December 2000 establishing a framework for the Community action in the field of water policy (2000/60/EC). Available at www.wfdireland.ie (accessed 27.03.2013). Dep. of the Environ., Dublin, Ireland.

27 

European Commission, 2009. Decision No. 406/2009/EC of the European Parliament and of the Council of 23 April 2009 on the Effort of Member States to Reduce their Greenhouse Gas

Emissions to Meet the Community's Greenhouse Gas Emission Reduction Commitments up to 2020. Available at eurlex.europa.eu/LexUriServ/LexUriServ.do?uri=OJ:L:2009:140:0136:0148:EN:PDF (accessed

27.03.2013)

Fealy, R., Schroder, J., 2008. Assessment of manure transport distances and their impact on economic and energy costs. International Fertiliser Society Conference, Cambridge, 12

December, 2008.

Fenton, O., Healy, M.G., Rodgers, M., O’ Huallachain, D., 2009. Site-specific P absorbency of ochre from acid mine-drainage near an abandoned Cu-S mine in the Avoca-Avonmore catchment, Ireland. Clay Miner. 44, 113-123.

Fenton, O., Healy, M.G., Henry, T., Khalil, M.I., Grant, J., Baily, A., Richards, K.G., 2011.

Exploring the relationship between groundwater geochemical factors and denitrification potentials on a dairy farm in southeast Ireland. J. Ecol. Eng. 37, 1304– 1313.

Flechard, C., Ambus, P., Skiba, U., Rees, R.M., Hensen, A., van den Pol, A., Soussana, J.F.,

Jones, M., Clifton-Brwon, J., Raschi, A., Horvath, L., van Amstel, A., Neftel, A., Jocher, M.,

Ammann, C., Fuhrer, J., Calanca, P., Thalman, E., Pilegaard, K., Di Marco, C., Campbell, C.,

Nemitz, E., Hargreaves, K.J., Levy, P., Ball, B., Jones, S., van de Bulk, W.C.M., Groot, T.,

Blom, M., Gunnink, H., Kasper, G., Allard, V., Cellier, P., Laville, P., Henault, C., Bizouard,

28 

F., Jolivot, D., Abdalla, M., Williams, M., Baronti, S., Berretti, F., Grosz, B., Dominques, R.,

2007. Effects of climate and management intensity on nitrous oxide emissions in grassland systems across Europe. Agric. Eco. Environ. 121, 135–152.

Gordon, R., Jamieson, R., Rodd, V., Patterson, G., Harz, T., 2001. Effects of surface manure application timing on ammonia volatilization. Can. J. Soil Sci. 81, 525-533.

Haney, R.L., Haney, E.B., 2010. Simple and rapid laboratory method for rewetting dry soil for incubations. Comms. Soil Sci. Plant Anal. 41, 1493–1501.

Haygarth, P.M., Hepworth, L., Jarvis, C., 1998. Forms of phosphorus transfer in hydrological pathways from soil under grazed grasslands. Eur. J. Soil Sci. 49, 65-72.

Hoekstra, N. J., Lalor, S.T.J., Richards, K.G., O’Hea, N., Lanigan, G.J., Dyckmans, J.,

Schulte, R.P.O., Schmidt, O., 2010. Slurry

15

NH

4

-N recovery in herbage and soil: effects of application method and timing. Plant Soil. 330, 357–368.

Hoekstra, N. J., Lalor, S.T.J., Richards, K.G., O’Hea, N., Dungait, J., Schulte, R.P.O,

Schmidt, O., 2011. The fate of slurry N fractions in herbage and soil during two growing seasons following application, Plant Soil. 342, 83-96.

IPCC. 2006. IPCC Guidelines for National Greenhouse Gas Inventories. Institute for Global

Environmental Strategies for the IPCC, Kanagawa, Japan.

IPCC/WMO/UNEP, 2007. Climate Change 2007: Impacts, adaptation, and mitigation

29 

of climate change: scientific-technical analyses. Cambridge University Press,

Prepared by IPCC Working Group III, Cambridge, UK.

Jahangir, M.M.R., Khalil, M.I., Johnston, P., Cardenas, L.M., Hatch, D.J., Butler, M., Barrett,

M., O’ Flaherty, V., Richards, K.G., 2012. Denitrification potential in subsoils: A mechanism to reduce nitrate leaching to groundwater. Agric. Ecosyst. Environ. 147, 13-23.

Kalbasi, M., Karthikeyan, K.G., 2004. Phosphorus dynamics in soils receiving chemically treated dairy manure. J. Environ. Qual. 33, 2296-2305.

Kramers, G., Holden, N.M., Brennan, F., Green, S., Richards, K.G., 2012. Water content and soil type effects on accelerated leaching after slurry application. Vad. Zon. J. 11, 244-257.

Kustermann, B., Christen, O., Hulsgergen, K., 2010. Modelling nitrogen cycles of farming systems as basis of site- and farm-specific nitrogen management. Agric. Ecosyst. Environ.

135, 70–80.

McBride, M.B., 2000. Chemisorption and precipitation reactions, in: Sumner, M.E. (Ed.),

Handbook of Soil Science. CRC Press, Florida. pp. B265-B302.

McCutcheon, G.A., 1997. MSc Thesis, National University of Ireland, Dublin.

McDowell, R.W., Sharpley, A.N., 2001. Soil phosphorus fractions in solution: influence of fertiliser and manure, filtration and method of determination. Chemosphere. 45, 737-

748.

30 

McDowell, R.W. Biggs, B.J.F., Sharpley, A.N., Nguyen, L, 2004. Connecting phosphorus loss from agricultural landscapes to surface water quality. Chem. Ecol. 20, 1–40.

Meade, G., Pierce, K., O'Doherty, J.V., Mueller, C., Lanigan, G., Mc Cabe, T., 2011.

Ammonia and nitrous oxide emissions following land application of high and low nitrogen pig manures to winter wheat at three growth stages. Agric. Ecosyst. Environ. 140, 208-217.

Misselbrook, T. H., Nicholson, F.A., Chambers, B.J., 2005a. Predicting ammonia losses following the application of livestock manure to land. Bioresour. Technol. 96, 159-168.

Misselbrook, T. H., Nicholson, F.A., Chambers, B.J., Johnson, R.A., 2005b. Measuring ammonia emissions from land applied manure: an intercomparison of commonly used samplers and techniques. Environ. Pollut. 135, 389-397.

Molloy S.P., Tunney, H., 1983. A laboratory study of ammonia volatilization from cattle and pig slurry. Irish J. Agric. 22, 37-45.

Morgan, M.F., 1941. Chemical Soil Diagnosis by the Universal Soil Testing System.

Connecticut Agricultural Experimental Station Bulletin 450, New Haven, Connecticut.

Morvan, T., Leterme, P., Arsene, G.G., Mary, B., 1997. Nitrogen transformations after the spreading of pig slurry on bare soil and ryegrass using N-labelled ammonium. Eur. J. Agron.

7, 181–188.

31 

Mosier, A. R., Schimel, D., Valentine, D., Bronson, K., Parton, W., 1991. Methane and nitrous oxide fluxes in native fertilized and cultivated grasslands. Nature 350, 330-332.

Nolan, T., Troy, S.M., Gilkinson, S., Frost, P., Xie, S., Zhan, X., Harrington, C., Healy,

M.G., Lawlor, P.G., 2012. Economic analyses of pig manure treatment options in Ireland.

Bioresour. Technol. 105, 15-23.

O’Bric, C., 1992. MSc Thesis, National University of Ireland, Dublin 1992.

O’ Flynn, C.J., Fenton, O., Healy, M.G., 2012a. Evaluation of amendments to control phosphorus losses in runoff from pig slurry applications to land. Clean – Soil, Air, Wat. 40,

164–170.

O’ Flynn, C.J., Fenton, O., Wilson, P., Healy, M.G., 2012b. Impact of pig slurry amendments on phosphorus, suspended sediment and metal losses in laboratory runoff boxes under simulated rainfall. J. Environ. Man. 113, 78-84.

Olsen, S.R., Watanabe, F.S., 1957. A method to determine a phosphorus absorption maximum of soils as measured by the Langmuir isotherm. Soil Sci. Soc. Proc. 31, 144–149.

Porporato, A., Odorico, P.D., Laio, F., Rodriguez-Iturbe, I., 2003. Hydrologic controls on soil carbon and nitrogen cycles. I. Modelling scheme. Water Resour. 26, 45-58.

Rafique, R., Hennessy, D., Kiely, G., 2011. Nitrous oxide emission from grassland under different management systems. Ecosystems. 14, 563–582.

32 

Rivett, M.O., Buss, S.R., Morgan, P., Smith, J.W.N., Bemment, C.D., 2008. Nitrate attenuation in groundwater: A review of biogeochemical controlling processes. Water

Resour. 42, 4215-4232.

Rochette, P., Chantigny, M.H., Angers, D.A., Bertrand, N., Côté, D., 2001. Ammonia volatilization and soil nitrogen dynamics following fall application of pig slurry on canola crop residues. Can. J. Soil. Sci. 81, 515-523.

Saggar, S., Tate, K.R., Giltrap, D.L., Singh, J., 2008. Soil-atmosphere exchange of nitrous oxide and methane in New Zealand terrestrial ecosystems and their mitigation options: a review. Plant Soil. 309, 25-42.

Schulte, R.P.O., Melland, A.R., Fenton, O., Herlihy, M., Richards, K.G., Jordan, P., 2010.

Modelling soil phosphorus decline: Expectations of Water Frame Work Directive policies.

Environ. Sci. Pol. 13, 472-484.

Sherlock, R.R., Sommer, S.G., Khan, R.Z., Wood, C.W., Guertal, E.A., Freney, J.R., Dawson,

C.O., Cameron, K.C., 2002. Ammonia, Methane, and Nitrous Oxide Emission from Pig

Slurry Applied to a Pasture in New Zealand. J. Environ. Qual. 31, 1491-1501.

S.I. No. 610 of 2010. European Communities (good agricultural practice for protection of waters) regulations 2010. www.environ.ie/en/Legislation/Environment/Water/FileDownLoad,25133,en.pdf (accessed

27.03.2013).

33 

Smith, D.R., Moore Jr., P.A., Griffis, C.L., Daniel, T.C., Edwards, D.R., Boothe. D.L., 2001.

Effects of alum and aluminium chloride on phosphorus runoff from swine manure. J. Environ.

Qual. 30, 992-998.

Smith, D.R., Moore Jr., P.A., Maxwell, C.V., Haggard, B.E., Daniel, T.C., 2004. Reducing phosphorus runoff from swine manure with dietary phytase and aluminum chloride. J.

Environ. Qual. 33, 1048-1054.

Sophocleous, M., 2011. On understanding and predicting groundwater response time. Ground

Water. 50, 528-540.

Troy, S.M., Lawlor, P.G., O’ Flynn, C.J., Healy, M.G., 2013. Impact of biochar addition to soil on greenhouse gas emissions following pig manure application. Soil Biol. Biochem. 60,

173-181.

Tunney, H., 2000. Phosphorus needs of grassland soils and loss to water, in: Steenvoorden, J.,

Claessen, F., Willems, J. (Eds.), Agricultural effects on ground and surface waters: Research at the edge of science and society. IAHS, England, 273, pp. 63–69.

Van Gestel, M., Ladd, J.N., Amato, M., 1991. Carbon and nitrogen mineralization from two soils of contrasting texture and micro-aggregate stability: influence of sequential fumigation, drying and storage. Soil Biol. Biochem. 23, 313–322.

34 

Velthof, G.L., Oenema, O., 1995. Nitrous oxide fluxes from grassland in the Netherlands: I.

Statistical analysis of flux chamber measurements. Eur. J. Soil Sci. 46, 533–540.

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

Wall, D., Jordan, P., Melland, A.R., Mellander, P.E., Buckley, C., Reaney, S.M., Shortle, G.,

2011. Using the nutrient transfer continuum concept to evaluate the European Union Nitrates

Directive National Action Programme. Environ. Sci. Pol. 14, 664-674.

Walsh, S. 2012. A summary of climate averages for Ireland 1981-2010. Met Eireann,

Dublin, Ireland.

Wulf, S., Maeting, M., Clemens, J., 2002a. Application technique and slurry co-fermentation effects on ammonia, nitrous oxide, and methane emissions after spreading: I. Ammonia volatilization. J. Environ. Qual. 31, 1789–1794.

Wulf, S., Maeting, M., Clemens, J., 2002b. 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.

35 

Figure Captions

Figure 1. Average weekly loads of ammonium a) nitrite b) and nitrate c) leached column

-1

(± standard deviation).

Figure 2. Cumulative gaseous emissions of N

2

O-N (a) CO

2

-C (b) and CH

4

-C (c) from columns at each sampling period (± standard deviation).

Figure 3. Cumulative loads of total organic carbon (TOC) and total inorganic carbon (TIC) leached over the duration of the experiment a) and weekly loads of total carbon leached from columns b) (± standard deviation).

36 

Table 1. Physical and chemical characteristics of the pig slurry used in this experiment and characteristic values of pig slurry from other farms in Ireland.

TP TN TC

NH

4

+

-N pH DM Reference

800

(mg L

-1

4200

) (%)

620±32

2940

±156

5860

±80

1739 ±8 7.51 ± 0.08

3.02± 0.24

The present study

S.I. No. 610 of 2010

1630 6621 5.77

McCutcheon, 1997 a

900±7 4600±21 3.2±2.3

O’ Bric, 1991 a

TP, total P; TN, total N; TK, total K; DM, dry matter. a

Values changed to mg L

-1

assuming densities of 1 kg

L

-1 .

37 

Table 2. Characterisation of amendments used in this study (O’ Flynn et al., 2012a, 2012b)

Amendment Alum

Ferric

Chloride

PAC

FeCl

3

% O

3

1.25

1.0 – 3.0

Al %

Fe %

As mg <2.8 <1.0

Cd mg <3.4 <0.2

Cr mg <48 <2.0

Cu mg <65

Mn mg

Ni mg <48 <1.0

Pb mg <14 <2.0

Sb mg <2.8 <1.0

Se mg <2.8 <1.0

Hg mg <0.7 <0.2

38 

Table 3. Average soil phosphorus, nitrogen and carbon contents by sampling time and depth.

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Morgan's P

(mg L

-1

WEP

(mg kg

-1

 

 

 

 

TC

(%) 

 

TN

(%) 

 

)

 

C:N Ratio

 

 

 

 

Treatment

 

 

abc

Means in a row, which do not share a superscript, were significantly different (p< 0.05)

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Month Depth (mm) Soil Only Slurry

1 0-50

50-100

3.53

3.69

a

  

  

100-200

2 0-50

50-100

100-200

3 0-50

50-100

100-200

6 0-50

50-100

100-200

8 0-50

50-100

100-200

1 0-50

50-100

100-200

2 0-50

50-100

100-200

3 0-50

50-100

100-200

6 0-50

50-100

100-200

8 0-50

50-100

100-200

1 50-100

2 50-100

3 50-100

6 50-100

8 50-100

1 50-100

2 50-100

3 50-100

6 50-100

8 50-100

1 50-100

2 50-100

3 50-100

6 50-100

8 50-100

3.53

3.84

4.02

4.14

3.19

3.14

3.35

2.69

3.22

3.51

2.17

2.44

2.66

0.54

0.56

0.64

0.51

0.49

0.50

0.62

0.65

0.64

0.54

0.54

0.49

0.58

0.58

0.55

1.70

1.78

1.72

1.81

1.75

0.18

0.18

0.18

0.19

0.19

9.53

9.73

9.54

9.38

9.31

a a

7.79

3.80

3.99

a

6.12

a

4.03

a

4.31

a

6.28

a

3.17

a

3.55

a

4.60

a

3.41

a

3.67

a

3.42

a a

2.39

a

3.14

a

1.13

0.58

a

0.56

a

0.99

a

0.47

a

0.46

a

1.06

a

0.66

a

0.70

a

0.87

a

0.55

Alum c a

4.19

3.75

a

3.79

b

4.41

a

3.85

a

3.88

c

4.22

a

3.50

a

3.71

c

3.44

a

3.21

a

3.65

c

2.63

ab

2.67

a

0.51

a

0.79

a a

0.62

a

0.61

1.73

a

1.73

a

1.73

a

1.78

b

0.57

a

0.49

ab

0.39

b

0.69

a

0.61

a

0.65

b

0.60

a

0.50

a a

1.73

0.18

a

0.18

a

0.18

a

0.19

a a

0.19

a

9.39

9.89

a

9.61

a

9.43

a

9.35

a

0.47

b

0.55

a a

0.55

a

0.58

1.86

a

1.77

a

1.74

a

1.74

a a

1.73

0.20

a

0.18

a

0.18

a

0.18

a a

0.18

a

9.32

9.69

a

9.51

a

9.78

a

9.79

ab

FeCl

3

4.64

a

3.69

a

3.95

a

4.61

a

3.80

a

3.86

b

4.55

a

3.60

a

3.78

ab

4.18

PAC

b

a

b

4.40

ab

3.68

a

3.84

a

4.52

a

3.99

a

4.08

a

3.39

a

3.67

a a a a a a a a a a

0.57

0.45

b

0.43

0.71

0.67

0.63

0.63

0.52

0.47

0.56

0.53

ab

0.50

a

0.62

a

0.52

a

a

0.49

0.44

a

a

a

a

a

a

a

0.62

a

0.57

ab

a

a

a

a

0.45

ab

a

0.73

a

0.62

a a

0.57

a

1.76

a

0.56

1.69

a

a

a

1.74

a

a

1.68

a

1.84

a

1.79

a a

1.79

a

1.68

a

1.66

a

1.75

a

a

a a a

0.19

a

a

0.19

a

0.19

0.19

a

a

0.18

0.18

a a

0.18

a

0.18

a a a

0.18

a

0.18

a

9.07

a

a

9.30

a a

9.41

a

9.30

a a

9.80

a

9.78

10.04

a

a

9.32

a

9.39

9.76

a

a

a

a

c

a

a

a

a

a

a

a

4.28

b a

3.62

a

3.61

a

3.10

a

3.28

ab

3.00

bc

3.38

b

a

a

a

a

bc

3.52

ab a

3.01

b a

0.49

0.54

a

0.57

ab

2.95

ab

3.16

a a

3.38

0.57

a a

0.58

a

0.60

a

3.66

a

a

0.59

0.57

a

a

0.54

0.55

39 

c)

25

20

15

10

5

0

Fig 1.

a)

0.10

0.08

0.06

0.04

0.02

0.00

0 5 10 15 20 25 30 35

Week b)

0.05

0.04

0.03

0.02

0.01

0.00

0 5 10 15 20 25 30 35

Week

0 5 10 15 20 25 30 35

Week

40 

Fig 2.

a) b)

1 2 3 4 5 6 7 9 11 13 15 19 23 28

Day

c)

1 2 3 4 5 6 7 9 11 13 15 19 23 28

Day

1 2 3 4 5 6 7 9 11 13 15 19 23 28

Day

41 

Fig 3.

a)

b)

12

10

8

4

6

2

0

0 5 10 15 20 25 30 35

Week

42 

Was this manual useful for you? yes no
Thank you for your participation!

* Your assessment is very important for improving the work of artificial intelligence, which forms the content of this project

Download PDF

advertisement