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Nitrogen transformation in a denitrification layer irrigatedwith dairy factory effluent

L.A. Schippera,�, A. McGillb

aDepartment of Earth and Ocean Sciences, University of Waikato, Private Bag 3105, Hamilton, New ZealandbLandcare Research NZ Ltd., Private Bag 3127, Hamilton, New Zealand

a r t i c l e i n f o

Article history:

Received 15 December 2007

Received in revised form

24 January 2008

Accepted 28 January 2008

Available online 14 February 2008

Keywords:

Denitrification

Land-based effluent treatment

Nitrogen

Wastewater

Effluent

nt matter & 2008 Elsevie.2008.01.033

thor. Tel.: +64 7 858 3700; fschipper@waikato.ac.nz (

a b s t r a c t

Adoption of land-based effluent treatment systems can be constrained by the costs and

availability of land. Sufficient land area is needed to ensure nitrate leaching from applied

effluent is minimised. One approach to decrease required land area is to enhance N

removal by denitrification. Layers of organic matter (100 mm thick) were installed below

topsoil of a site irrigated with dairy factory effluent. These ‘‘denitrification’’ layers were

tested to determine whether they could decrease nitrate leaching by increasing

denitrification. Four plots (10� 10 m2 each) were constructed with a denitrification layer

installed at 300 mm below the surface, and N losses were measured in leachate using

suction cups every 3 weeks for 19 months. N in leachate was compared with 4 control plots.

Denitrifying enzyme activity, nitrate concentrations, and carbon availability were

measured in samples collected from the denitrification layers. These measurements

demonstrated that denitrification occurred in the layer; however, denitrification rates were

not sufficiently high to significantly decrease nitrate leaching. Total N leaching was

296 kg N ha�1 from control plots and 238 kg N ha�1 from plots with denitrification layers; a

total of 798 kg N ha�1 was applied in effluent. More than 50% of the leached N to 40 cm was

as organic N, presumably due to bypass flow. Other studies have demonstrated that thicker

denitrification layers (more than 300 mm) can reduce nitrate leaching from small-scale

septic tank drainage fields but this study suggests that it is probably not practical to use

denitrification layers at larger scales.

& 2008 Elsevier Ltd. All rights reserved.

1. Introduction

Application of effluent to land can reduce direct discharge of

nutrients (particularly N and P) into water ways. However,

excess loadings of effluent can still result in high concentra-

tions of nitrate leaching to groundwater and then onto

surface waters (Cameron et al., 1997). Consequently, effluent

application rates are often constrained by N concentrations in

effluent. Where N concentrations are high, additional land

area is required and the cost of additional land is often the

greatest impediment for industries or municipalities choos-

r Ltd. All rights reserved.

ax: +64 7 858 4964.L.A. Schipper).

ing land application over mechanical treatment and dis-

charge to surface water. Therefore, there is a need to improve

N removal rates in the soil/plant systems if land-based

effluent treatment systems are to be encouraged.

Mechanisms for reducing nitrate leaching from land treat-

ment systems include plant harvest and removal (Cameron

et al., 1997), N immobilisation into the organic matter pool

(Degens et al., 2000), and denitrification (Barton et al., 1999a;

Lowrance et al., 1998; Singleton et al., 2001). Of these,

denitrification is the only mechanism whereby applied N is

converted to N gas completing the N cycle. Denitrification is

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WAT E R R E S E A R C H 4 2 ( 2 0 0 8 ) 2 4 5 7 – 2 4 6 42458

an anaerobic process so that in well-drained soils (i.e. those

often used in land treatment systems) it is generally

constrained by excess oxygen (Tiedje, 1988). In practice, this

means that denitrification rates in well-drained soils are

limited by the amount and distribution of particulate organic

matter (Parkin, 1987), which has the dual role of providing

both a carbon source for denitrifiers and an anaerobic

environment. Few studies have measured denitrification

rates in land-based effluent treatment systems but rates can

vary widely from less than 5 kg ha�1 yr�1 (Barton et al., 1999a)

to between 49 and 239 kg ha�1 yr�1 (Lowrance et al., 1998)

depending on N loading rate and soil type.

Little work has been undertaken on identifying approaches

to enhance denitrification at the site of application, although

denitrification walls have been shown successful at removing

nitrate from shallow groundwater leaving a land effluent

treatment system (Schipper et al., 2005). Singleton et al. (2001)

demonstrated that controlled drainage could increase deni-

trification in poorly drained soil irrigated with effluent.

Robertson et al. (2000) investigated the feasibility of adding

a layer of organic matter under the topsoil to increase N

removal as nitrate leached from septic tank drainage fields.

They reported reduction in nitrate leaching from around

30 mg N L�1 to less than 5 mg N L�1 at one site and from more

than 100 mg N L�1 to generally less than 10 mg N L�1 at a

second site during a 6-year study. The organic layer was

placed below the topsoil to allow mineralisation and nitrifica-

tion of applied organic N and ammonium to nitrate prior to

entering the organic layer (Fig. 1). Whether this approach

might also be useful in larger land treatment systems is

untested. Hydraulic and nutrient loads can be considerably

higher than in septic drainage fields. The economic feasibility

of adding large amounts of organic matter as layers to land

treatment systems would depend on the additional N removal

achieved and the subsequent financial saving accrued from

not needing to purchase further land.

Effluent applicationOrganic N, NH4

+, NO3-

Organic N, NH4+, NO3

-

NO3-

NO3-

Leaching

Mineralisationand nitrification

Denitrification

N2, N2O, NO

Topsoil

Denitrification layer

Subsoil

Fig. 1 – Conceptual diagram of a denitrification layer and

postulated N transformations. Not to scale.

The objective of this study was to determine whether a

denitrification layer could decrease N leaching from a land

treatment system. Denitrification layers were constructed

within a paddock irrigated with dairy factory effluent.

Leaching of N species was monitored for 19 months below

plots with a denitrification layer and below control plots to

determine whether denitrification layers were a possible

approach to improve the performance of this land treatment

system. To determine whether the organic layer provided a

suitable environment for denitrification, denitrifying enzyme

activity, carbon availability and nitrate concentrations were

also measured in samples taken from the layer.

2. Methods

2.1. Study site and construction

The study was conducted at the Bardowie dairy farm (371 570S,

1751 270E), Cambridge, North Island, New Zealand. The

pastures on the farm were spray irrigated with effluent from

the nearby Hautapu Dairy Factory. Several soils types are

represented on the farm and this experiment was conducted

in a paddock dominated by Horotiu sandy loam (New Zealand

classification: Typic Orthic Allophanic Soil; US Soil taxonomy

(Great group): Vitric hapludand). This was a freely draining

soil with 11% carbon and 1.3% N associated with its relatively

high content of allophanic clays, low bulk density of

0.49 g cm�3, high total porosity 78% and high unsaturated

infiltration rates (K�40) of 50 mm h�1 (Sparling et al., 2001).

Sand:silt:clay contents were: 34:49:17 (Degens et al., 2000).

Further biological, chemical and physical properties are given

in Sparling et al. (2001). The paddock was surveyed by auger to

ensure that the soil was uniform in the selected area for the

trial.

Paddocks were irrigated on a semi-regular basis depending

on effluent flows from the adjacent dairy factory. Effluent

consisted predominantly of cleaning fluids and by-products

of milk processing. Effluent composition is given in Degens

et al. (2000), but in summary, chemical oxygen demand

averaged 6350 g m�3, total P averaged 93 g m�3, and the pH

averaged 6.2. Further details on N content and composition

are reported in the current study below. Effluent was irrigated

using fixed rotating sprinklers on a diamond grid with

approximately 20 m spacing between rows. This design

ensured even coverage of the paddock with effluent. Typically,

50 mm of effluent was applied during a 4-day period every 18

days with 15–25 mm applied on days 1, 2 and 4; day 3 was not

irrigated to allow the previously applied effluent to drain. On

occasion, when effluent volumes from the factory were low,

effluent was only irrigated for the first 2 days. Staff from the

farm collected a sample of effluent for each irrigation event

and these samples were analysed for nitrate and nitrite using

flow injection analysis (cadmium reduction) and total kjel-

dahl N using phenol/hypochlorite colorimetry following

digestion with sulphuric acid and copper sulphate catalyst

(Blakemore et al., 1987). Paddocks were grazed by a herd of

cows (stocking rate of 3.5 cows ha�1) about every 2 weeks and

generally before irrigation.

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Before construction of the denitrification layers, two blocks

(each 55 m long�10 m wide) were established in the paddock

in the centre of three rows of irrigators. Each block was

subdivided into four plots (each 10�10 m2, with a 5 m buffer

between plots). Two plots in each block had denitrification

layers installed (described below) and two remained undis-

turbed (control). Treatments were randomly assigned within

blocks. In February 2004, the denitrification layers were

constructed by removing the top 30 cm of soil from a

10�10 m2 plot using a large bucket excavator. Then 10 m3 of

garden chip (generally sold as a garden mulch, approximate

dry weight of 3 T) was spread across each excavated plot,

about 10 cm thick. The garden chip was a mixture of freshly

chipped branches and stem material with some leaf material

generally less than 10 mm in length and 3 mm in thickness.

Topsoil was replaced over the top of the organic material and

reseeded with ryegrass (Lolium presenne L.) and white clover

(Trifolium repens L.).

The soil on the control plots was not disturbed because the

objective of the study was to compare the effectiveness of an

installed denitrification layer for reducing nitrate leaching

against current practice. This meant that the experimental

design would not separate effects of topsoil disturbance and

addition of organic matter on the amount of N leaching. If the

layer was successful at reducing N losses then further work

would be conducted to determine exact mechanisms for N

removal and also measure potential adverse effects, such as

nitrous oxide emissions.

2.2. Leachate collection

Suction cups for collecting leachate were constructed using a

porous ceramic cup (Soilmoisture Ceramics, Australia, 40 mm

diameter) glued onto a PVC pipe (40 cm in length). Two plastic

tubes passed through the cap, one tube ran to bottom of the

ceramic cup for sampling of liquid. The second tube extended

just inside the cap and was used to apply a vacuum.

Suction cups were installed in November 2004 into each

plot (4 cups per plot). First a pit (30�30 cm2) was dug to below

the organic matter layer with four short narrow side trenches

(40 cm long�10 cm wide and 20 cm deep) radiating away from

the pit in each corner. The trenches allowed a Dutch auger to

drill into the opposite side of the pit below the organic layer at

approximately 201. Auger holes were about 40 cm in length

and the same diameter as the suction cups. A small amount

of water was added to some subsurface soil, mixed to a slurry,

and used to coat the ceramic cup before insertion into the

hole. The cup was about 40 cm below the surface. One cup

was installed into each wall and suction tubes were run into a

central PVC container (approximately 25 cm diameter and

40 cm high) through holes in sidewalls. Tubes passing through

the holes of the central container were sealed using expand-

ing foam. The container was closed with a screw cap, which

was flush with the soil surface and protected the container

from treading damage.

For the first 5 months after installation, suction (�30 kPa)

was periodically applied to all cups using a syringe and water

samples were discarded to avoid disturbance effects. Sample

collection was timed to occur immediately after the effluent

irrigation. A suction (�30 kPa) was applied to each suction cup

after the first 2 days of irrigation and before the final day of

irrigation (day 4). The day after the last irrigation, leachate

was collected from the suction cups using a syringe and

stored frozen until analysis. Nitrate and ammonium were

analysed using standard auto-analyser techniques (Blakemore

et al., 1987) and organic N determined following persulfate

digestion (Sparling et al., 1996) by auto-analyser.

2.3. Measurement of spray cover

To test whether plots received the same amounts of effluent

during the course of the experiment, total hydraulic loads

were measured on each of the plots. On day 3 of each

irrigation event, 4 buckets (30 cm diameter) were placed in a

square (5 m apart) centred on the middle of each plot.

Following a single day of irrigation, volume of effluent

collected in each of the buckets was measured. Volumes of

effluent were summed for each plot during the trial and a

t-test used to determine whether there were significant

differences in hydraulic load between plots with and without

denitrification layers.

2.4. Denitrification layer analyses

On three occasions (January, July 2005 and June 2006) samples

of organic matter were taken from the denitrification layer

and measurements made of denitrifying enzyme activity,

carbon availability, moisture content, and nitrate and ammo-

nium concentrations. Samples were taken from the layer

using a spade, transferred into plastic bags, and transported

to the laboratory for analysis within an hour.

Nitrate and ammonium were immediately extracted with

100 mL of KCl (2 M). The KCl extract was filtered (Advantec 5C)

and the filtrate was analysed for nitrate and ammonium by

standard auto-analyser techniques (Blakemore et al., 1987).

Denitrifying enzyme activity (DEA) was measured as de-

scribed by Schipper et al. (2005). Briefly, soil (10 g fresh weight)

was incubated shaking in airtight bottles (100 mL) at 28 1C.

Soils were amended with a 20 mL solution containing glucose

(0.2 g L�1), potassium nitrate (0.1 g N L�1), and chlorampheni-

col (0.12 g L�1). The headspace was replaced with N gas, and

10 mL acetylene added to inhibit reduction of N2O to N2. After

15 and 75 min, headspace samples were removed and

analysed for N2O using a gas chromatograph (Philips

PU4410, Cambridge, UK) equipped with an electron capture

detector. Available C was determined from the accumulation

of CO2 in the headspace of bottles (610 mL) that were statically

incubated at 25 1C for 7 days (Schipper et al., 2005). Carbon

availability was measured under aerobic (air headspace) and

anaerobic (N gas headspace) conditions. CO2 was analysed

using an infra-red gas analyser (Series 225, Analytical

Development Ltd. Corporation, Hoddesdon, England). Bio-

chemical results are presented on an oven dry-weight basis.

2.5. Data analysis

A simple daily water balance was used to estimate the

volume of leaching from the field site. A leaching event was

considered to occur when the sum of daily rainfall and

irrigation minus losses through evapotranspiration was

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WAT E R R E S E A R C H 4 2 ( 2 0 0 8 ) 2 4 5 7 – 2 4 6 42460

greater than the capacity of the soil storage. Rainfall and

Class A pan-evaporation were measured daily on-site. Poten-

tial evapotranspiration was estimated to be 0.8 of Class A

pan-evaporation (Doorenbos and Pruitt, 1984). Available soil

water reservoir was determined from the depth to the suction

cup (0.40 m) and measured available soil water data by

horizon. The resulting reservoir of available soil water was

93 mm. Actual daily evapotranspiration was determined

using a crop factor of 1.0 for pasture and an empirical

relationship that estimates AET based on the percentage of

available water in the soil reservoir (Allen et al., 1998).

Mass of N species leaching for each plot was calculated as

the product of concentration of the N species in the leachate

collected from the suction cups and the calculated leaching

volume determined from the water balance. The leaching

volume used was the sum of the daily leaching volumes

spanned by the suction cup sampling dates; this was

approximately 9 days before and 9 days after the sampling

of suction cup. These leaching losses were summed for each

plot to determine cumulative N losses over the period of the

study. For one control plot, samples from suction cups were

only intermittently available and it was not possible to

calculate cumulative losses. Consequently, for data analysis

the number of control plots was 3 whereas there were four

plots with a denitrification layer. T-tests were used to

determine whether there were significant differences be-

tween plots with and without denitrification layers.

3. Results

Total rainfall at the site was 2710 mm during the 19-month

study (Fig. 2). Irrigation events were generally 18 days apart

except in May when there were larger gaps between irrigation

events as less milk was processed through the factory. As

expected, evaporation rates were highest in mid-summer

(December–February) and lowest in winter (July–September)

(Fig. 2). Consequently, calculated drainage was generally

lower in summer (November–March), particularly in 2005,

when rainfall was low in January through March (Fig. 2).

The N loading in effluent during the study period was

870 kg ha�1. There were no significant differences in effluent

hydraulic loads between control plots (average hydraulic load

of 750 mm) and plots with a denitrification layer (average

hydraulic load of 767 mm). Total N loads in effluent were

798 kg N ha�1 applied evenly throughout the 2-year study

period (Fig. 3), which is considerably higher than most dairy

farms would receive through fertiliser additions and N

fixation by clover. In comparison, in the Waikato region, dairy

farms applied about 113 kg N ha�1 y�1 and N fixation would

have added a further 120 kg ha�y�1 (Parfitt et al 2006). The

average N concentration in the effluent was 75 g N m�3,

predominantly as total kjeldahl N (average 94%) with the

remainder being nitrate and nitrite.

For both control plots and plots with a denitrification layer,

N leaching was greatest in winter months (May–August) for

both years (Fig. 4). In the first year, losses were predominantly

as organic N, whereas in the second year, nitrate accounted

for a greater proportion of N losses (Fig. 4a and b). Ammonium

losses were a small component of the total N leached (Fig. 4c).

The only significant differences in total N leaching between

control plots and denitrification layer plots occurred in the

first few months of monitoring. This difference was primary

due to less organic N leaching through the denitrification

layer plots (Fig. 4). Cumulative N leaching for the study period

was initially less in the denitrification layer plots than control

plots (Fig. 3). However, by the end of the study, cumulative

leaching in the controls was 296 kg ha�1 yr�1 (standard

error ¼ 28, n ¼ 3) and not significantly different from plots

with denitrification layers with a cumulative leaching of

238 kg ha�1 yr�1 (standard error ¼ 58, n ¼ 4) (Fig. 3). This

leaching loss was about 30–38% of total N applied in effluent.

Denitrifying enzyme activity and available carbon (aerobic

and anaerobic) in the denitrification layer declined with time

(Table 1). Nitrate and ammonium was always present in the

organic material.

4. Discussion

Biochemical measurements (DEA, available carbon and ni-

trate concentrations) suggested that conditions were con-

ducive for denitrification in the garden chip layer (Table 1).

Denitrification is generally limited by available carbon, nitrate

concentration and sufficient anaerobic sites. Nitrate was

present in extracted samples collected from the garden chip

layer and was generally greater than 5 mg kg�1, which is the

concentration below which denitrification can become nitrate

limited (Barton et al., 1999b). Carbon availability was similar

to that measured in denitrification walls where nitrate

removal was shown to be due to denitrification for more

than 7 years (Schipper and Vojvodic-Vukovic, 2001; Schipper

et al., 2005), suggesting carbon was not limiting. However,

unlike the denitrification wall, the denitrification layer was

not permanently water saturated and denitrification could be

constrained by excess oxygen in pores. In aerobic environ-

ments, such as well-drained soils, denitrification occurs in

anaerobic microsites, such as pieces of organic matter

(Parkin, 1987) or in aggregates (Sexstone et al., 1985). Garden

chip material was nearly saturated when sampled with a

moisture content of about 80% (w/w), which would reduce

penetration of oxygen and allow formation of anaerobic

microsites within chips where denitrification could be active.

Supporting evidence for this was the initially high DEA, which

is a surrogate measure of the size of the active microbial

population capable of denitrification reflecting antecedent

conditions. Like carbon availability, DEA values were similar

to those measured in the denitrification walls (Schipper et al.,

2005). DEA declined during the 1.5-year sampling period (from

over 1400 to less than 40 ng N g�1 h�1), probably due to

declines in available carbon as labile fractions of garden chip

degraded. Declines in DEA were also measured with time in

denitrification walls as carbon availability decreased (Schip-

per and Vojvodic-Vukovic 2001). DEA may also have been less

in the last two samplings due to seasonal effects. The last two

samplings were taken during winter when soil temperatures

at 30 cm were about 8 1C and much less than during summer

when temperatures are around 19 1C.

While there was strong evidence that denitrification

occurred in the garden chip matrix, the rate was not sufficient

ARTICLE IN PRESS

Dai

ly ra

infa

ll (m

m)

0

20

40

60

80

100

Irrig

atio

n (m

m)

05

1015202530

Dai

ly e

vapo

ratio

n (m

m)

02468

1012

Date (2005-2006)Ja

nFeb Mar Apr

May Ju

n JulAug Sep Oct

Nov Dec Jan

Feb Mar AprMay Ju

n JulAug Sep Oct

Nov

Cal

cula

ted

mon

thly

dra

inag

e (m

m)

0

50

100

150

200

250

300

Fig. 2 – Rainfall, irrigation, evaporation and hydraulic leaching losses from the studied land treatment site.

WA T E R R E S E A R C H 4 2 ( 2 0 0 8 ) 2 4 5 7 – 2 4 6 4 2461

to cause a significant decrease in nitrate leaching below the

denitrification layer when compared with the control plots

(Fig. 4). A possible reason for this lack of nitrate removal was

insufficient contact between the organic material and leach-

ing effluent for sufficient denitrification. The garden chip

layer was approximately 50 mm thick after soil had been

placed back on top, and assuming a porosity of 50% (v/v),

would give water storage in the chip layer of 25 mm. Effluent

loading rates and leaching events (Fig. 2) were often greater

than this, suggesting a retention time of less than 1 day in the

layer. In comparison, Robertson et al. (2000) conducted field

studies of two organic layers below septic tank discharges

measuring 70–80% reduction in N. Their organic layers were

350 and 500 mm thick, with lower loading rates ranging from

3 to 40 mm d�1 so that the retention time within the layer was

probably in the order of several days. The organic matter

layers tested by Robertson et al. (2000) were also mixed with

silty material to assist water storage and enhance anaerobic

conditions to support denitrification. Bedessem et al. (2005)

also demonstrated high rates of nitrate removal in an organic

layer below a septic tank discharge: 67% removal in an organic

matter layer compared with a removal of 31% in control

columns. As with Robertson et al. (2000), the layers tested

were significantly thicker, ranging from 300 to 900 mm.

Effluent application rates were about 40 mm d�1, so their

retention times would be several days, depending on porosity.

While the field study used denitrification layers with a higher

organic matter contents than these previous studies, the

layers were thinner and short contact time between organic

layer and leachate may have restricted nitrate removal.

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Date (2005-2006)

0.00

0.05

0.10

0.15

0.20

0.25

0.30

0.35

Jan0

10

20

30

40

50

60

70

80

90

100

May Sep Jan May Sep Jan

Jan May Sep Jan May Sep Jan

Tota

l NH

4 le

achi

ng (k

gN h

a-1)

Tota

l NO

3 le

achi

ng (k

gN h

a-1)

Fig. 4 – N species in leachate below plots with a denitrification la

(B) organic N, (C) ammonium (note different y-axis scale) and (D

Date (2005-2006)

Jan

Cum

ulat

ive

Tota

l N (

kg N

ha-1

)

0

200

400

600

800Leaching below control plotsLeaching below denitrification layerCumulative effluent load

May Sep Jan May Sep Jan

Fig. 3 – Cumulative N applied in effluent and N leaching

beneath control plots (three replicates) and plots with

denitrification layers (four replicates). Error bars are one

standard error.

WAT E R R E S E A R C H 4 2 ( 2 0 0 8 ) 2 4 5 7 – 2 4 6 42462

The high concentrations of organic N in the leachate

suggested there was also significant bypass flow of

leachate through the topsoil to the suction cups. These soils

are well-structured loams with surface unsaturated conduc-

tivities (K�40) of 75 mm h�1 considered highly suitable for

irrigation (Sparling et al., 2001). However, bypass flow can

occur in these soils even when instantaneous loading

rates are below this unsaturated conductivity (McLeod

et al., 1998). This occurs when the soil becomes pugged by

animal treading, which results in localised surface films of

water in hoof-prints allowing macropore flow to occur

(McLeod et al., 1998). For denitrification layers to remove N

from effluent, leachate–N needs to be converted to nitrate

before entering the organic matter layer. Consequently, the

effectiveness of the denitrification layers can be limited by

the rate of mineralisation and subsequent nitrification in soil

above the denitrification layer. Bedessem et al. (2005) also

concluded that N removal was limited by conversion of

organic N and ammonium to nitrate in soils above the organic

layer.

0

20

40

60

80

100

Date (2005-2006)

0

20

40

60

80

100

Control plots Denitrification layer plots

Sep

Sep

Jan May Jan May Sep Jan

Jan May Jan May Sep Jan

Tota

l N le

achi

ng (k

g ha

-1)

Org

anic

N le

achi

ng (k

gN h

a-1)

yer (four replicates) and control (three replicates). (A) nitrate,

) total N. Error bars are 1 standard error.

ARTICLE IN PRESS

Table 1 – Biochemical properties of the denitrification layer

Soil property Units January 2005 July 2005 June 2006

Average SDa Average SD Average SD

Moisture % (w/w) 80.5 27.5 86.9 22.2 78.2 8.0

Nitrate mg N g�1 0.7 1.3 17.0 10.9 19.5 15.9

Ammonium mg N g�1 40.3 31.6 4.7 2.0 11.3 9.1

C availability +O2 mg C g�1 h�1 13.0 3.9 11.0 8.3 4.9 1.6

C availability �O2 mg C g�1 h�1 4.1 1.3 4.2 0.6 1.9 0.6

DEAb ng N g�1 h�1 1410 1903 72 86 35 35

There were four replicates per sampling time.a Standard deviation.b Denitrifying enzyme activity.

WA T E R R E S E A R C H 4 2 ( 2 0 0 8 ) 2 4 5 7 – 2 4 6 4 2463

Over half the N leached was as organic N rather than

nitrate, which is generally considered the most mobile form

of N. Other studies of effluent irrigation onto land have also

noted significant organic N leaching below the topsoil

(Singleton et al., 2001; Sparling et al., 2006). In these studies,

losses of organic N were attributed to bypass flow through

topsoil. The fate of the organic N below the topsoils is not well

understood. In tile-drained soils with a near-surface water-

table, direct losses of organic N to surface water have been

reported (e.g., Monaghan and Smith, 2004). It is possible

that in soil with a deeper watertable, the organic N is

mineralised in the vadose zone before reaching ground-

water, so that only nitrate is detected in groundwater. In

monitoring the performance of land-based treatment sys-

tems, therefore, it is important to measure both organic N and

nitrate in leachate and possibly also in groundwater. As most

of the treatment of applied effluent occurs in topsoil,

irrigation practices that minimise bypass flow need to be

developed and applied (McLeod et al., 2001; Monaghan and

Smith, 2004).

Despite nitrate leaching being similar below denitrification

layers and control plots there was an apparent initial

difference in cumulative N losses (Fig. 4). This difference

was predominantly due to organic N leaching being greater in

the control plots. This suggested the garden chip layer

initially retained organic compounds in applied effluent but

either became saturated or released organic N during

decomposition. After September 2005, there were no differ-

ences between treatments, and cumulative losses between

treatments narrowed becoming non-significant by the end of

the trial (Fig. 3).

A simple N budget for the site can be constructed. Total

inputs of N were about 400 N kg ha�1 y�1 coming predomi-

nantly from applied effluent, with no N fertiliser being added.

N fixation was assumed to be 0 as clover was outcompeted by

grasses because of high N inputs. Atmospheric deposition in

New Zealand is low, generally less than 5 kg N ha�1 y�1 (Parfitt

et al., 2006). Outputs include leaching which averaged across

all plots was 134 kg N ha�1 y�1, and product off-takes, primar-

ily in milk, about 70 kg N ha�1 y�1. Net N immobilisation (long-

term storage in the organic matter pool) at the site was

measured between 1975 and 1997 at about 95 kg N ha�1 y�1

(Degens et al., 2000). But it was unclear if this rate of N

storage would continue because the organic matter at the

site may have become saturated with N (Schipper et al.,

2004). N unaccounted for then ranges between 106 and

201 kg N ha�1 y�1 as calculated by difference between inputs

and outputs depending on whether net N immobilisation is

assumed to continue or have stopped. The amounts of N lost

by ammonium volatilisation from effluent irrigated site sites

are difficult to estimate; however, much of the unaccounted N

could be attributed to denitrification. In a review of studies of

fertilised and irrigated pastures, denitrification rates averaged

113 kg N ha�1 y�1 (Barton et al., 1999a). Denitrification rates

could even be greater at the Bardowie farm because of the

high organic matter content of the Horotiu soil (10%C) and

loading of available carbon (particularly lactose) in the

effluent (Degens et al., 2000), which could stimulate deni-

trification.

5. Conclusions

Denitrification was supported in an organic layer receiving

effluent; however, rates were insufficient to decrease nitrate

leaching significantly from a large-scale land treatment

system. This lack of removal was likely due to the short

residence time of leachate in the layer. Other studies have

demonstrated that thicker layers (300–500 mm) can success-

fully remove nitrate below septic tank drainage fields with

areas in the order several m2 (e.g., Robertson et al., 2000,

Bedessem et al., 2005). However, the practicality of incorpor-

ating such thick organic layers below larger land treatment

systems, which can be many hectares in size, is probably

limited. Bypass flow occurred through the denitrification

layer, reducing the mineralisation and nitrification in topsoils

above the denitrification layer, which may have also con-

strained N removal. Bypass flow allowed considerable trans-

port of organic N below the topsoil, which is often not

measured in an assessment of environmental impacts of land

application of effluents but probably should be. There is a

continuing need to test and identify practical management

practices that enhance denitrification of excess nitrate at the

site of application before leaching.

ARTICLE IN PRESS

WAT E R R E S E A R C H 4 2 ( 2 0 0 8 ) 2 4 5 7 – 2 4 6 42464

Acknowledgements

Thanks to staff from Hautapu Dairy factory for access to the

site, assistance with irrigation and grazing scheduling, and

provision of weather and effluent data. Greg Barkle is thanked

for assistance with calculating water balances. Maja Vojvodic-

Vukovic is thanked for biochemical analysis. Malcolm McLeod

is thanked for help with site selection. John Russell, Greg

Barkle, Bryan Stevenson, Graham Sparling and Anne Austin

are thanked for helpful comments on the manuscript. Two

anonymous reviewers are thanked for their comments, in

particular the suggestion to incorporate a simple nitrogen

budget for the site. This work was funded by FRST Contract

C09X0304 and the University of Waikato.

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