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Published online 24 October 2007
Published in J Environ Qual 36:1883-1894 (2007)
DOI: 10.2134/jeq2007.0175
© 2007 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
677 S. Segoe Rd., Madison, WI 53711 USA
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TECHNICAL REPORTS

Waste Management

Soil Water Nitrate and Ammonium Dynamics under a Sewage Effluent–Irrigated Eucalypt Plantation

S. J. Livesleya,c,*, M. A. Adamsb and P. F. Griersonc

a School of Forest and Ecosystem Science, The Univ. of Melbourne, 500 Yarra Blvd., Melbourne, VIC 3122, Australia
b School of Biological, Earth and Environmental Sciences, The Univ. of New South Wales, Biological Sciences Bdg., Randwick, NSW 2031, Australia
c Ecosystems Research Group, School of Plant Biology, The Univ. of Western Australia, 35 Stirling Hwy., Crawley WA 6009, Australia

* Corresponding author (sjlive{at}unimelb.edu.au).

Received for publication April 10, 2007.

   ABSTRACT
TOP
 NOTES
 ABSTRACT
INTRODUCTION
 Materials and Methods
 Results
 Discussion
 Conclusions
 REFERENCES
 
Managed forests and plantations are appropriate ecosystems for land-based treatment of effluent, but concerns remain regarding nutrient contamination of ground- and surface waters. Monthly NO3–N and NH4–N concentrations in soil water, accumulated soil N, and gross ammonification and nitrification rates were measured in the second year of a second rotation of an effluent irrigated Eucalyptus globulus plantation in southern Western Australia to investigate the separate and interactive effects of drip and sprinkler irrigation, effluent and water irrigation, irrigation rate, and harvest residues retention. Nitrate concentrations of soil water were greater under effluent irrigation than water irrigation but remained <15 mg L–1 when irrigated at the normal rate (1.5–2.0 mm d–1), and there was little evidence of downward movement. In contrast, NH4–N concentrations of soil water at 30 and 100 cm were generally greater under effluent irrigation than water irrigation when irrigated at the normal rate because of direct effluent NH4–N input and indirect ammonification of soil organic N. Drip irrigation of effluent approximately doubled peak NO3–N and NH4–N concentrations in soil water. Harvest residue retention reduced concentrations of soil water NO3–N at 30 cm during active sprinkler irrigation, but after 1 yr of irrigation there was no significant difference in the amount of N stored in the soil system, although harvest residue retention did enhance the "nitrate flush" in the following spring. Gross mineralization rates without irrigation increased with harvest residue retention and further increased with water irrigation. Irrigation with effluent further increased gross nitrification to 3.1 mg N kg–1 d–1 when harvest residues were retained but had no effect on gross ammonification, which suggested the importance of heterotrophic nitrification. The downward movement of N under effluent irrigation was dominated by NH4–N rather than NO3–N. Improving the capacity of forest soils to store and transform N inputs through organic matter management must consider the dynamic equilibrium between N input, uptake, and immobilization according to soil C status, and the effect changing microbial processes and environmental conditions can have on this equilibrium.

Abbreviations: SOM, soil organic matter


   INTRODUCTION
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
Materials and Methods
 Results
 Discussion
 Conclusions
 REFERENCES
 
THE application of municipal effluent from sewage treatment to tree plantations has increased in many countries as an alternative to disposal in rivers, estuaries, or the ocean and as a means of increasing tree productivity. Managed natural forests (Meding et al., 2001) and plantations (Polglase et al., 1995) provide an effective land-based system for effluent treatment because of their capacity to transpire large volumes of waste water while retaining or transforming the nutrients contained within, thereby preventing pollution of surface and ground water (Myers et al., 1995; Speir et al., 1999; Barton et al., 2005; Sharma and Ashwath, 2006).

A major concern in the land-based treatment of effluent has been the contamination of groundwater with nitrate nitrogen (NO3–N) (Sopper and Kerr, 1979; Monnet et al., 1995; Roygard et al., 2001; Singleton et al., 2001). Nitrate is more susceptible to leaching and therefore more likely to contaminate groundwater than ammonium nitrogen (NH4–N) and many other nutrients (Barber, 1984). Understanding the causal factors of N leaching and extrapolating between studies of N leaching is problematic because of the many confounding factors that influence N dynamics, including soil type, climate, effluent chemistry, tree species, growth stage, irrigation rate, and regime (Polglase et al., 1995; Bond, 1998).

Under land-based treatment of effluent, NO3–N and NH4–N concentrations in soil water may increase as a direct result of inorganic N in the effluent or indirectly through improved soil water status and increased soil organic matter (SOM) mineralization (Stanford and Epstein, 1974; Myers et al., 1982). Separating the importance of these contributory factors is critical (Polglase et al., 1995). Regardless of the N source (effluent or SOM), available inorganic N is often rapidly immobilized through microbial uptake if a suitable C substrate is available to the population (Vance and Chapin, 2001). In the first rotation of a plantation, SOM conditions and microbial activity may improve slightly as litter accumulates (Mendham et al., 2002). In the second rotation, SOM conditions, microbial activity, and C substrate supply may change dramatically depending on whether the harvest residues are retained or removed (Pilar et al., 2001). Increased SOM and C substrate supply can in turn greatly increase the capacity of a forest soil to transform and retain added inorganic N in labile or recalcitrant organic forms (Tietema et al., 1998; Vance and Chapin, 2001), which could influence the long-term management of harvest residues (removal, retention, burning) in effluent irrigated forest systems.

The mechanism by which effluent is distributed can have a profound effect on the efficiency of water and nutrient uptake, nutrient loss (gaseous and leaching), and the potential for nutrient pollution. Under drip irrigation, effluent N is applied to a localized area beneath the dripper, whereas sprinkler irrigation distributes effluent N more uniformly (Haynes and Swift, 1987), better exploiting the conversion and retention capacity of the soil and microbial biomass. Preferential (macropore) flow increases with surface ponding under flood irrigation or high rates of sprinkler irrigation and is greatly reduced under drip or low rates of sprinkler irrigation (Jaynes and Rice, 1993). Regardless of the distribution mechanism, preferential flow may occur during, or after, large rainfall events (Seyfried and Rao, 1987; Magesan et al., 1999). The spatial distribution and concentration of nutrient solutes applied is therefore important, but the interaction between the distribution mechanism of the effluent and the spatial variability in saturated hydraulic conductivity has received little research attention (Wang et al., 1997).

This study investigated the comparative effect of (i) effluent or potable water irrigation, (ii) irrigation distribution mechanism (drip or sprinkler), (iii) rate of application (normal or double), and (iv) harvest residue retention on soil water NO3–N and NH4–N concentrations in the topsoil (30 cm) and subsoil (100 cm). The relative importance of effluent N (direct) and enhanced N mineralization (indirect) to soil N dynamics was investigated by separately irrigating areas with effluent waste water and potable drinking water. It was hypothesized that NO3–N and NH4–N concentrations of soil water would be greater under effluent irrigation than water irrigation or rain fed because of direct effluent N input and enhanced N mineralization. Similarly, drip irrigation would lead to greater concentrations of NO3–N and NH4–N in soil water than sprinkler irrigation because of the point application and limits on soil buffer capacity. Concurrently, harvest residue retention would decrease NO3–N and NH4–N concentrations under sprinkler irrigation but not under drip irrigation.

Soil water samples were collected at 30 and 100 cm depths on a monthly basis throughout 2001 from Eucalyptus globulus plantations that were drip or sprinkler irrigated with effluent or potable water and with and without harvest residues. Rain-fed areas served as a control. Gross rates of nitrification and ammonification were measured in six key treatments using 15N isotope pool dilution, and the accumulation of NO3–N and NH4–N in the soil system (soil and harvest residue) was measured after the first year of irrigation treatment.


   Materials and Methods
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Materials and Methods
Results
 Discussion
 Conclusions
 REFERENCES
 
In 1993–94, an effluent irrigated plantation was established 10 km north of Albany, a rural coastal town in southwestern Western Australia. The plantation provides full disposal of secondary treated sewage effluent for Albany (~50,000 people) in a two-stage process (overland flow through a pasture to a holding dam followed by drip irrigation throughout a 400-ha plantation). Albany has a Mediterranean climate, receiving 800 mm yr–1 (Fig. 1 ). The surface soil type is Yellow Duplex (Northcote, 1979), otherwise classed as a Petroferric Chromosol (Isbell, 1996) or a Typic Palexuralf (USDA, 1999), consisting of a fine sandy loam overlying a fragmented ferruginous duricrust at about 15 to 30 cm depth above a massive, mottled light clay. Clay content in the surface soil was 15% (Table 1 ), increasing to ~50% in the B horizon beyond a depth of 30 cm. The surface pH was slightly acidic (5.3), and where harvest residues were retained there was an increase in the C/N ratio of the mineral soil and a reduction in bulk density (Table 1). In total, 400 ha of previous pasture land was planted with Eucalyptus globulus (subsp. globulus, Maiden) in double tree rows with an approximately 5 x 2 m spacing to a density of 2000 stems ha–1. A double tree row consisted of two ripped and mounded lines, 1 m apart, into which E. globulus seedlings were planted in a staggered pattern. Irrigation with semi-treated sewage waste water commenced in 1995 through drip lines placed in furrows either side of a double tree row (~4.5 m apart) at a rate of 1.5 to 2.0 mm d–1, except during winter months when greater rainfall resulted in soil water content approaching field capacity. Total N concentration of the tree irrigated effluent typically ranged between 10 and 40 mg L–1 (75% NH4–N, 15% NO3–N, and 10% organic N (Adams et al., 2001). In June 1999, a 4-ha area was harvested and regrown from coppice. In the first year of regrowth, this area received no irrigation. The harvest residue (foliage and branch wood) was mulched and redistributed in July 1999 to produce nine groups of three harvest residue treatments, each group consisting of three 20 x 8 m areas of "double harvest residue," "normal harvest residue," and "harvest residue removed." Each residue treatment area spanned four double tree rows and contained ~32 coppiced root stumps. Superimposed on these harvest residue treatments were drip and sprinkler irrigation systems, such that four groups of harvest residue treatments were irrigated through drippers and another four through sprinklers. The remaining group of three harvest residue treatments was rain fed. Sprinklers (capacity 2.1 mm h–1) with a 3.5-m radius throw were installed in a 7 x 7 m grid pattern for normal irrigation rates in two groups and at 7 x 3.5 m in another two groups to sprinkler irrigate at twice the normal rate. Similarly, two of the drip irrigated groups were drip irrigated at twice the normal rate (x2) by laying two drip lines within a furrow. Finally, each pair of like-irrigated groups was further divided so that one group was irrigated with effluent and the other with potable water. When adjacent areas were irrigated at different rates, trenches were dug to 0.8 m, lined with double plastic, and back-filled to prevent lateral flow. The irrigation treatments began operation in November 2000 and continued until June 2001, effluent or potable water being applied in one irrigation event each day to achieve approximate daily application rates (Fig. 1). The total N concentration of the effluent irrigated over the plantation was generally about 20 mg L–1, but in the autumn months of April, May, and June, this increased to up to 40 mg L–1 (Fig. 1). Proportionally, the effluent N is dominated by NH4–N (75%), with NO3–N representing ~15% and 10% being organic N.


Figure 1
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  		Fig. 1. Monthly rainfall, mean drip- and sprinkler-irrigated effluent rates, and effluent total N concentration during 2001 at the Albany Effluent Irrigation Treatment Farm, Albany, Western Australia.

 

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  		Table 1. Basic surface soil properties at the Albany Effluent Irrigation Treatment Farm, Albany, Western Australia.

 
Lysimeter Installation
Suction cup lysimeters were installed in March 2000 to depths of 30 and 100 cm. Lysimeters were constructed from porous ceramic cups (50 mm wide) attached to appropriate lengths of 50 mm outer diameter PVC pipes, capped and sealed in a method similar to that described by Talsma et al. (1979). Lysimeters were washed before use with 1 M HCl, followed by repeat rinsing with deionized water. Holes slightly wider than the lysimeter were augered using a hydraulic drill rig and finished to the correct depth using a hand auger. The lysimeters were carefully installed, and the hole was repacked with soil and two bentonite plugs, one just above the ceramic cup and one just below the soil surface to prevent preferential vertical flow. Lysimeters were installed in all the sprinkler-irrigated treatments, in drip-irrigated effluent treatments with single and no harvest residues, and in drip-irrigated water treatments with single and no harvest residues irrigated at the normal rate.

Lysimeters were strategically installed to provide a spatially representative sample of free draining soil water at 30 and 100 cm depths. In sprinkler irrigated treatments, three lysimeters were installed to a depth of 30 cm in furrow, mid-row (between furrows), and intermediate locations. Two lysimeters were installed to a depth of 100 cm in furrow and mid-row locations. In drip-irrigated treatments, four lysimeters were installed to a depth of 30 cm, two in furrow and two in mid-row locations. Two lysimeters were installed to a depth of 100 cm in furrow and mid-row locations.

Soil Water and Soil Sampling
Monthly samples of soil water were collected by evacuating each lysimeter to –50 kPa so that water held at a soil matric potential less than –50 kPa drained into the ceramic cup. This soil water sample was pumped out of the lysimeter and refrigerated at 4°C before analysis. Soil water solutions were analyzed colorimetrically to determine NO3–N and NH4–N (µg mL–1) concentrations using a continuous flow, Technicon auto analyzer II (Saskatoon, Canada) (McLeod, 1992).

In November 2001, 5 mo after irrigation had ceased, harvest residue and mineral soil samples were collected in the sprinkler and drip-irrigated effluent treatments with and without harvest residue and from the sprinkler-irrigated water treatment without harvest residues. In drip-irrigated treatments, soil cores (0–20 cm) were collected within a furrow at 0, 25, and 50 cm from a dripper point and perpendicular to the inter-row at 50, 100, and 200 cm from a dripper point. In sprinkler-irrigated treatments, soil cores (0–20 cm) were collected along the inter-row at 50, 100, 150, and 200 cm from a sprinkler and perpendicular to the inter-row, from the sprinkler to the furrow at 50, 100, 150, and 200 cm from a sprinkler. All soil samples were a composite of three replicate cores, and all cores were separated into 0- to 5-, 5- to 10-, and 10- to 20-cm depths. Soil samples were refrigerated at 4°C and were processed within 7 d. Each composite soil sample was homogenized; then, 10 g of unsieved soil was extracted with 50 mL of 1 M KCl for 1 h on an end-to-end shaker. For treatments with harvest residues, similar composite harvest residues samples were collected and 10-g subsamples extracted with KCl. The KCl extracts were filtered (Whatman 40) and analyzed colorimetrically to determine NO3–N and NH4–N as before.

15N Isotope Pool Dilution
One-meter-square quadrats were selected in areas with and without harvest residues in rain-fed and sprinkler irrigated water and effluent treatments (normal irrigation rates). In each quadrat, 16 PVC rings were placed on the surface, with eight in each half. Each PVC ring indicated the position of subsequent soil cores (50 mm diameter and 40 mm depth). Eight soil cores on one half of the quadrat were injected with 6 mL of 99 atom % 15(NH4)2SO4 (28 mg N L–1), and the other eight received 6 mL of 99 atom % Na15NO3 (28 mg N L–1). Multiple injections were made with a bidirectional needle to evenly distribute the 15N throughout the soil volume. Cores were not hammered into the ground until the time of sampling to maintain root integrity and minimize disturbance. Each 6-mL injection equated to an addition of 168 µg N.

After 24 h, four cores from each quadrat that received 15NH4–N injections and four cores that received 15NO3–N cores were sampled. This process was repeated 120 h after injection. All soil samples were stored on ice in the field and refrigerated at 4°C before processing within 7 d. Each soil sample was thoroughly homogenized, and 40 g of unsieved soil were extracted with 200 mL of 1 mol L–1 KCl for 1 h on an end-to-end shaker and filtered (Whatman 40) directly into 500-mL glass jars. A 5-mL subsample was removed to colorimetrically determine NH4–N and NO3–N concentrations. Subsamples of soil were taken to determine gravimetric water content.

The KCl extracts were immediately prepared for 15N isotopic ratio analysis through sequential microdiffusion of NH4–N then NO3–N onto acidified GF/C glass fiber filter discs (Brooks et al., 1989; Murphy et al., 2003). The diffusion discs were dried over anhydrous sulfate and analyzed on a Europa (Cheshire, UK) Tracer Mass Isotope Analyser linked to a Europa Roboprep combustion analyzer. Rates of gross ammonification and gross nitrification were calculated according to a zero order 15N pool dilution equation (Barraclough, 1991),

Formula
where A is the labeled pool (in this case NH4–N, mg kg–1), t is incubation time (days), and m is the rate of ammonification. The asterisk indicates atom percentage excess 15N, the subscript 1 indicates the first time of sampling, the subscript 2 indicates the second time of sampling, and {theta} is the rate at which the labeled (NH4–N) pool changed size.

Data Analysis and Presentation
In sprinkler treatments, soil water NO3–N and NH4–N concentrations at 30 cm were the mean of three sample positions (furrow, mid-row, and an intermediate location), whereas at 100 cm they were the mean of two sample positions (furrow and mid-row). In the drip line treatments, soil water NO3–N and NH4–N concentrations at 30 cm were the mean of four samples (two furrow and two mid-row), whereas at 100 cm they were the mean of two samples (furrow and mid-row).

The two-dimensional sampling of mineral soil (and residue) NO3–N and NH4–N in November 2001 enabled the amount of N stored in each of the five treatments measured to be calculated and compared. Weighted averages are presented according to the surface area and depth each sample represented within the furrow and mid-row sections. The effect of (i) harvest residue retention and (ii) irrigation type on estimated rates of gross nitrification and ammonification were tested within irrigation treatments using nonpaired, heteroscedastic, two-tailed Student's t tests (n = 16). Significant differences in the spatial distribution of stored NO3–N and NH4–N (mg kg–1) in the topsoil (0–5 cm) were investigated using a homoscedastic, two-tailed t test.

The two-dimensional spatial distribution of soil inorganic N (and harvest residue) was presented using contour fill figures (Deltagraph; Red Rock Software, Salt Lake City, UT). Treatments with harvest residue have contours above the soil surface (y axis = 0) to represent inorganic N in the harvest residue.


   Results
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results
Discussion
 Conclusions
 REFERENCES
 
There was insufficient soil water during January and February to provide a sample via the lysimeters. In rain-fed treatments, lysimeters did not provide a soil water sample until May, with the break of the dry season (Fig. 1).

Impact of Irrigation Rate on Soil Water N
Sprinkler irrigating effluent at twice the normal rate increased soil water NO3–N concentrations at 30 cm during or immediately after the period of irrigation regardless of the harvest residue treatment (Fig. 2 ). However, at 100 cm, this increase in soil water NO3–N concentration was only obvious in soils without harvest residues (Fig. 3 ). Doubling the rate of sprinkler irrigated effluent decreased the measured concentrations of NH4–N in soil water below those under normal irrigation (Fig. 4 and 5 ). Similarly, doubling the rate of sprinkler irrigated water did not increase soil water NO3–N relative to normal irrigation rates. However, for soil without harvest residues, doubling the rate of sprinkler irrigated water greatly increased the concentration of NH4–N in soil water at 30 (up to 12.4 mg L–1) and 100 cm (up to 9.5 mg L–1). Drip irrigating effluent at twice the normal rate, more than doubled the measured N concentrations in soil water at 30 cm, with NO3–N increasing from 8.6 to 22.7 mg NO3–N L–1 and NH4–N increasing from 12.5 to 22.9 mg NH4–N L–1. At 100 cm, drip irrigating effluent at twice the normal rate only increased soil water NO3– above that when irrigated at the normal rate in soils without harvest residues, from <2.0 to >8.0 mg NO3–N L–1 (Fig. 3).


Figure 2
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  		Fig. 2. Monthly soil water NO3–N concentrations at 30 cm under two rates of effluent and water irrigation and for two modes of application with and without harvest residues at the Albany Effluent Irrigation Treatment Farm, Albany, Western Australia, during 2001.

 

Figure 3
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  		Fig. 3. Monthly soil water NO3–N concentrations at 100 cm under two rates of effluent and water irrigation and for two modes of application with and without harvest residues at the Albany Effluent Irrigation Treatment Farm, Albany, Western Australia, during 2001.

 

Figure 4
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  		Fig. 4. Monthly soil water NH4–N concentrations at 30 cm under two rates of effluent and water irrigation and for two modes of application with and without harvest residues at the Albany effluent irrigation treatment farm, Albany, Western Australia, during 2001.

 

Figure 5
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  		Fig. 5. Monthly soil water NH4–N concentrations at 100 cm under two rates of effluent and water irrigation and for two modes of application with and without harvest residues at the Albany Effluent Irrigation Treatment Farm, Albany, Western Australia, during 2001.

 
Impact of Irrigation Method on Soil Water N
There was no significant difference in soil water NO3–N or NH4–N concentrations between drip or sprinkler irrigated effluent treatment when irrigated at the normal rate. However, when effluent was drip irrigated at twice the normal rate, the measured peak concentrations of soil water NO3–N and NH4–N at 30 cm were markedly greater (22.7 and 22.9 mg L–1, respectively) than those under sprinkler irrigation (13.0 and 8.2 mg L–1, respectively) (Fig. 2 and 4), but there was no difference in NO3–N and NH4–N concentrations between irrigation methods at 100 cm (Fig. 3 and 5).

Impact of Effluent Irrigation on Soil Water N
When the irrigation systems were active, water-irrigated treatments (drip or sprinkler) had similar soil water NO3–N concentrations to those measured in the rain-fed treatments (≤1.0 mg L–1) at 30 and 100 cm (Fig. 2 and 3). Therefore, the large concentrations of NO3–N measured in soil water under effluent irrigation (drip or sprinkler) were a direct result of effluent N input, not a result of enhanced nitrification of existing organic/inorganic N. In contrast, the large NH4–N concentrations measured in soil water under effluent irrigation were partly created by enhanced ammonification of existing organic N because NH4–N concentrations under sprinkler irrigation increased with water irrigation, particularly when applied at double the normal rate (Fig. 4).

Impact of Harvest Residue Retention on Soil Water N
At a depth of 30 cm, there was a positive effect (reduced soil water NO3–N concentration) from the retention of harvest residues when effluent was sprinkler irrigated at the normal rate. When effluent was sprinkler irrigated at double the normal rate, large peak soil water NO3–N concentrations (4.8–13.0 mg L–1) were measured in August, and the beneficial effect (i.e., reduced soil water NO3–N) of harvest residue retention was still apparent, although NO3–N concentrations had increased (Fig. 2). When effluent was drip irrigated, the retention of harvest residues had no obvious mitigating effect on measured soil water NO3–N concentrations (Fig. 2). At a depth of 100 cm, soil water NO3–N concentrations differed little according to the retention or removal of harvest residues when effluent was irrigated at the normal rate (Fig. 3). When irrigation rates were doubled, the retention of harvest residues seemed to reduce the downward movement of soil NO3–N. However, the increase in subsoil water NO3–N in treatments without harvest residues was small, remaining less than 0.5 under sprinkler and 1.5 mg L–1 under drip irrigation (Fig. 3). For soil water NH4–N concentrations, there was no clear positive or negative effect from the retention of harvest residues during or after active irrigation (Fig. 4 and 5).

At the onset of spring (October through November), several months after irrigation had ceased, soil water NO3–N concentrations increased slightly at 30 cm and increased severalfold at 100 cm, from <0.5 mg L–1 at the end of winter to 2.0 to 4.8 mg L–1 in November (Fig. 2 and 3). This production and downward movement of NO3–N in the spring after irrigation had ceased several months earlier was most noticeable in treatments that had retained harvest residues and especially so in treatments that had been previously sprinkler irrigated with water or effluent. There was no comparable increase in soil water NH4–N concentration in the spring; in fact, NH4–N concentrations continued to decrease toward the end of the 2001 (Fig. 4 and 5).

Spatial Distribution of Stored Soil NO3–N and NH4–N
Mineral soil NO3–N concentrations in furrows of drip-irrigated effluent treatments were greater beneath a dripper than mid-way between drippers (Fig. 6 ). In contrast, mineral soil NH4–N concentration was noticeably less beneath a dripper than mid-way between dripper (Fig. 7 ). The retention of harvest residues significantly increased (p ≤ 0.05) soil NO3–N concentration in the upper 5 cm but had no significant effect on stored soil NH4–N in this mineral soil layer.


Figure 6
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  		Fig. 6. The spatial distribution of soil NO3–N (mg kg–1) with depth and distance along a drip line (A and B) and from a dripper to the mid-row (A–C) in drip-irrigated effluent treatments with and without harvest residues at the Albany Effluent Irrigation Treatment Farm, Albany, Western Australia. Treatments with harvest residue have a layer above the soil surface to represent NO3–N in the harvest residue.

 

Figure 7
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  		Fig. 7. The spatial distribution of soil NH4–N (mg kg–1) with depth and distance along a drip line (A and B) and from a dripper to the mid-row (A–C) in drip-irrigated effluent treatments with and without harvest residues at the Albany Effluent Irrigation Treatment Farm, Albany, Western Australia. Treatments with harvest residue have a layer above the soil surface to represent NH4–N in the harvest residue.

 
In sprinkler-irrigated effluent treatments, mineral soil (0–5 cm) NO3–N concentration was generally greater in the furrow (0.38 mg kg–1) than in the mid-row area (0.19 mg kg–1), but there was no significant effect from harvest residue retention. Mineral soil (0–5 cm) NH4–N concentration was generally greater outside of the furrow, but there was no significant effect from harvest residue retention.

In the sprinkler-irrigated water treatment without harvest residues, mineral soil NO3–N concentrations in the furrow were between 0.13 and 0.21 mg kg–1 throughout the 20-cm profile, considerably greater than the NO3–N concentrations in the mid-row area, which decreased from <0.14 mg kg–1 in the upper 5 cm to <0.05 mg kg–1 at 10 to 20 cm. Mineral soil NH4–N concentrations in the same treatment ranged between 0.4 and 0.9 mg kg–1 in the upper 5 cm but were generally greater in the furrow.

Stored NO3–N and NH4–N
The retention of harvest residues did not increase the total amount of NO3–N or NH4–N stored in treatments that had been sprinkler or drip irrigated with effluent. Inorganic N in the harvest residues represented <10% of the soil system (0–20 cm) NO3–N but >25% of NH4–N. Sprinkler irrigation of effluent, as compared with drip irrigation and regardless of harvest residue retention, led to a greater amount of soil stored NO3–N, but the increase was not significant (Table 2 ). Sprinkler irrigation of effluent, as compared with sprinkler irrigation of potable water, led to greater NO3–N and NH4–N in the upper 20 cm of mineral soil, but the difference was not significant (Table 2).


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  		Table 2. Stored NO3–N and NH4–N (g ha–1) in the upper 20 cm of soil (and harvest residue) in sprinkler- and drip-irrigated treatments of effluent and water with or without harvest residues at the Albany Effluent Irrigation Treatment Farm, Albany, Western Australia, in November 2001.

 
Gross Rates of Nitrification
Estimated rates of nitrification in rain-fed treatments were <0.5 mg N kg–1 d–1 but were significantly two-fold greater (p ≤ 0.01) in soil with harvest residues than in soil without harvest residue (Table 3 ). In water-irrigated treatments, estimated rates of nitrification were <1.0 mg N kg–1 d–1, and there was no significant effect from the retention of harvest residues. Estimated rate of nitrification in effluent irrigated soil with harvest residues was 3.1 mg N kg–1 d–1 and almost fivefold greater (p ≤ 0.001) than soil without harvest residues (0.51 mg N kg–1 d–1).


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  		Table 3. Gross nitrification, gross ammonification, and soil water content in the upper 4 cm of soil in rain-fed and sprinkler-irrigated effluent and water treatments, with or without harvest residues, at the Effluent Irrigation Treatment Farm, Albany, Western Australia.

 
Gross nitrification in soils irrigated with water was between two- and fourfold greater (p ≤ 0.001) than soils in rain-fed treatments, depending on harvest residue retention (Table 3). When harvest residues were retained, irrigation with effluent significantly increased gross nitrification by threefold, as compared with sites irrigated with water (p ≤ 0.01).

Gross Rates of Ammonification
Estimated rates of gross ammonification in rain-fed treatments with harvest residues (1.6 mg N kg–1 d–1) were significantly greater (p ≤ 0.01) than those for rain-fed soil without harvest residues (0.5 mg N kg–1 d–1). In water-irrigated treatments, the estimated rate of gross ammonification was greater without harvest residues (2.1 mg N kg–1 d–1) than with harvest residues (1.3 mg N kg–1 d–1), but not significantly so. Similarly, in the effluent-irrigated treatments, the estimated rate of gross ammonification was greater (p ≤ 0.01) without harvest residues (2.3 mg N kg–1 d–1) than with harvest residues (1.0 mg N kg–1 d–1) (Table 3).

Irrigation with water significantly (p ≤ 0.01) increased rates of ammonification in soil without harvest residues, as compared with like rain-fed treatments, but there was no significant increase in gross rates of ammonification when soil with harvest residues were water irrigated (Table 3). Irrigation of effluent did not increase rates of gross ammonification beyond those for comparable water-irrigated treatments.


   Discussion
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results
 Discussion
Conclusions
 REFERENCES
 
Little soil water was recovered from the lysimeters during the dry summer months of January and February, which indicated that little free-draining soil water was available; therefore, the opportunity for downward leaching of NO3–N or NH4–N would have been minimal.

Soil Water NO3–N Dynamics
Irrigating effluent (drip or sprinkler) at the normal rate produced mean concentrations of NO3–N in soil water that were generally <10 mg L–1 at 30 cm and <5.0 mg L–1 at 100 cm (Fig. 2 and 3). The guideline limit for NO3–N in drinking water is 50 mg L–1, to prevent methemoglobinemia in infants, although the generally accepted limit in surface drinking water is 10 mg L–1, with groundwater or well water frequently reaching 50 mg L–1 (WHO, 1998). In a study comparable to ours, Polglase et al. (1995) reported similar peak concentrations in soil water NO3–N at a depth of 50 cm in the second (6.0–9.0 mg L–1) and third (<6.0 mg L–1) years of an effluent-irrigated pine plantation at Wagga Wagga in eastern Australia. In our study, doubling the rate of effluent irrigation had a large and rapid effect on peak concentrations of soil water NO3–N at 30 cm; concentrations exceeded 10 mg L–1 under double-sprinkler irrigation and exceeded 20 mg L–1 under double-drip irrigation. These results contrast with Polglase et al. (1995), where there was no apparent increase in soil water NO3–N concentrations from doubling the rate of effluent irrigation until the third year of irrigation.

Effluent irrigation greatly increased the concentrations of soil water NO3–N measured at 30 cm, in comparison to those measured under water irrigation, but did not seem to stimulate the downward movement of NO3–N (Fig. 2 and 3). Although there were large differences between soil water NO3–N concentrations measured at 30 and 100 cm, soil water concentrations at 100 cm remained <2.0 mg L–1, whereas irrigation was active, which suggests that a large amount of the NO3–N in soil water was transformed (denitrified, immobilized) or taken up by E. globulus or groundcover plants, thereby reducing its downward movement. Polglase et al. (1995) reported little difference between soil water NO3–N concentrations at 50 and 100 cm; however, they concluded that effluent irrigation of the pine plantation did not promote leaching of NO3–N beyond that observed under water irrigation. In a similar study, effluent irrigation of an E. grandis plantation in New South Wales, Australia, produced peak soil water NO3–N concentrations of up to 150 mg L–1 at 50 cm and up to 60 mg L–1 at 100 cm in the fourth year of irrigation, and NO3–N leaching was clearly stimulated by effluent application (Snow et al., 1999). Other studies have shown that irrigating tree-based systems with effluent can increase NO3–N leaching but that the concentration of NO3–N in the leachate can be acceptably small (<10 mg L–1) when applied at suitable rates (Speir et al., 1999; Paavolainen et al., 2000) and increases considerably (i.e., >10 mg L–1) when over-irrigated (Sopper and Kerr, 1979; Kim and Burger, 1997). Soil water NO3–N was expected to be more susceptible to leaching and downward movement than NH4–N (Ramos and Kücke, 2001), but this was not the case. The monthly sampling frequency was insufficient to detect short-lived preferential flow events, but if NO3–N had moved down the soil profile to a depth of 100 cm, it would have to be rapidly reduced to NH4–N or taken up by the tree roots to avoid detection in any of the monthly samplings. The small concentrations of NO3–N measured in soil water at 100 cm at our study site suggest that the downward movement of NO3–N was small and that daily effluent irrigation rates (~2.0 mm d–1) seemed to approximately equal the uptake requirements of the plantation and buffer capacity of the soil (Fig. 3).

Soil Water NH4–N Dynamics
The concentrations of NH4–N measured in soil water while the irrigation systems were active were larger than expected and at 100 cm were several times greater than comparable concentrations of NO3–N (Fig. 4 and 5). During 2000, the total N concentration of the effluent ranged between 13 and 40 mg L–1, with ~75% being NH4–N, 15% NO3–N, and 10% organic N (Adams et al., 2001). The application of effluent (drip or sprinkler) clearly increased the measured concentrations of NH4–N in soil water at 30 and 100 cm beyond that measured under water irrigation. Therefore, we concluded that application of effluent seemed to promote the downward movement of NH4–N. Most studies investigating effluent irrigation and leaching losses have focused on NO3–N (Singleton et al., 2001), but in certain situations NH4–N or organic N may be the dominant form of N to be leached. For example, irrigation of dairy shed effluent to large (75 cm deep), undisturbed lysimeters with growing grass promoted leaching losses of NH4–N beyond those of NO3–N in high (25 cm) water table situations (Singleton et al., 2001), and it was suggested that preferential macropore flow was responsible. Similarly, Silva et al. (2000) investigated leaching losses after surface applications of cow urine to soils in New Zealand and observed that leaching losses greatly increased when soil water conditions reached field capacity (0 kPa) as macropore flow was initiated and the main form of N to be leached was NH4–N. It is generally accepted that in structured soils, preferential flow of solutes depends on the location of the solute within (or on) aggregates and on the timing and continuity of rainfall or irrigation (Kluitenberg and Horton, 1990; McLay et al., 1991). However, if the soil residence time of applied NH4–N is long enough, transformation and immobilization can occur, and losses can be greatly reduced. In another lysimeter study, Magesan et al. (1999) investigated solute leaching losses in two New Zealand soils irrigated with NH4–N rich effluent (40–55% N was NH4–N) over 3 yr. The measured mean concentration of NO3–N in the recovered leachate was 7.5 to 33.9 mg L–1, whereas NH4–N remained <1.0 mg L–1. These results suggest that preferential flow had not occurred before the NH4–N had been transformed to NO3–N (i.e., residence times in the profile were considerable). Simulated heavy rainfall and bromide tracer studies by Magesan et al. (1999) indicated that preferential flow did not occur in one of the soil types and only occurred under field capacity conditions in the other. At our study site in Albany, it is possible that when effluent irrigation coincided with a period of high rainfall, preferential flow may have occurred and NH4–N moved down through the upper 1-m profile before nitrification to NO3–N, uptake, or immobilization could occur. The potential for preferential flow to occur would have increased as the soil profile wetted up through the winter and into spring. The importance of preferential flow with regard to N movement (NO3–N, NH4–N, and organic N) requires further research analysis and management consideration (Kohne and Gerke, 2005).

There was no clear effect of doubling effluent irrigation rates on soil water NH4–N concentrations or differences between sprinkler or drip irrigation, probably because of the spatially variable nature of preferential flow. The blocky nature of the duricrust layer, the ripping of this duricrust layer and soil profile to 0.8 m, and the development and decay of tree root channels after the first rotation could promote preferential flow along vertical and tortuous routes.

The Proportional Importance of N and Bulk Water in Effluent
Irrigating with water increased soil water NH4–N concentrations at 30 and 100 cm beyond those measured under rain-fed treatments (Fig. 4 and 5). This suggests that a large amount of NH4–N was produced through the mineralization of soil or surface organic matter and that mineralization rates increased in response to increased soil water conditions. Therefore, it follows that the large concentrations of NH4–N measured in soil water under effluent irrigation may have resulted from a combination of (i) indigenous soil N mineralization and (ii) direct NH4–N addition from the effluent. Other irrigation studies found no difference in net mineralization rates between soil irrigated with effluent or water (Polglase et al., 1995). Similarly, in our study, the gross rate of ammonification measured in effluent-irrigated soil was not significantly greater than that under water irrigation (Table 3). However, when effluent was irrigated (drip or sprinkler) at the normal rate, thereby minimizing the priming effect of water, the greatest soil water NH4–N concentrations at 30 and 100 cm were seen under harvest residues (i.e., with increased organic inputs).

Soil Water NO3–N and NH4–N Dynamics and the Effect of Harvest Residue Retention
In effluent treatments that were sprinkler irrigated, the retention of harvest residues clearly reduced concentrations of NO3–N in soil water in the topsoil (Fig. 2) and thereby prevented the downward movement of NO3–N during active irrigation. In effluent treatments that were drip irrigated, harvest residue retention did not reduce concentrations of NO3–N at 30 cm but did seem to prevent downward movement of NO3–N to 100 cm (Fig. 3). The beneficial effects (immobilization and/or denitrification) from the retention of harvest residues seemed to be greater under sprinkler irrigation, probably as effluent was applied over a wider area. However, the beneficial effects of harvest residue retention during active irrigation were offset by an increase in NO3–N concentration and downward movement at the onset of spring (August through September), especially in sprinkler-irrigated treatments (Fig. 2 and 3). Polglase et al. (1995) also observed a small increase in soil water NO3–N at the onset of spring, before effluent irrigation recommenced in the New South Wales pine plantation. At our study site, the increase in soil water NO3–N at 30 cm was matched by large increases at 100 cm, especially in sprinkler-irrigated treatments (with effluent or water) (Fig. 3). This suggests that mode of irrigation and the addition of N in effluent promoted the spring "nitrate flush" and the subsequent downward movement of NO3–N when the soil profile was at its wettest (Gonçalves and Carlyle, 1994). Sprinkler irrigation distributes water or effluent over a much larger area than drip irrigation, thereby increasing water availability and promoting microbial activity over a greater area. Combined with the retention of harvest residues, sprinkler irrigation would promote conditions that would have supported a greater microbial population, a wider C/N ratio of substrates, and therefore increased immobilization of effluent N. However, as environmental and management conditions changed, the potential for microbial re-mineralization and N loss increased. The retention and incorporation of harvest residues in pine plantations has previously been shown to increase N immobilization, microbial biomass, and microbial activity (Pilar et al., 2001).

Soil Water NO3–N and NH4–N and the Method of Irrigation (Drip or Sprinkler)
Drip irrigation of effluent produced greater peak concentrations of NO3–N in soil water than sprinkler irrigation (Fig. 2 and 3). It is difficult to compare this observation with the results of others studies investigating leaching losses of fertilizer N under drip and sprinkler irrigation systems because of differences in experimental design, scale, soil type, species, and climate and their frequently contradictory outcomes (Hicklenton and Cairns, 1995; Colangelo and Brand, 2001).

In our study, concentrations of NO3–N and NH4–N in soil water under sprinkler irrigation were similar regardless of spatial location, whereas large spatial heterogeneity was observed in NO3–N and NH4–N concentrations under drip irrigation (data not presented). There is limited literature on the effects of irrigation method on solute transport and soil heterogeneity (Wang et al., 1997). Sprinkler systems can produce a regular and unimodal wetting front in simulated homogenous soils (Hamdi et al., 1994), but in drip-irrigated systems the wetting front is focused beneath the point of application (Strabbioli and Turci, 1995; Riga and Carpentier, 1999). Wang et al. (1997) used a two-dimensional solute transport model to investigate the comparative effects of sprinkler, drip, and furrow irrigation and concluded that sprinkler irrigation was the least likely to cause ground water contamination. In our study, the greatest concentrations of NH4–N in soil water under drip irrigation (30 cm) were in the furrow, whereas at 100 cm the greatest concentrations were in the mid-row (data not presented). The reason for this spatial dichotomy with depth was not clear but may be due to the redirection of percolating water by tortuous macropore pathways or root channels. Studies of preferential flow have generally compared ponding and nonponding irrigation systems (e.g., flood irrigation and sprinkler irrigation) and have reported that ponding promotes preferential flow (Jaynes and Rice, 1993; Chen et al., 2002). To fully explain our observations, further field research is required to investigate the interaction between distribution mechanisms (uniform and localized) and spatial heterogeneity in soil hydraulic conductivity (preferential flow) and solute leaching (Wang et al., 1997).

Accumulation of NO3–N and NH4–N in the Soil Profile
Effluent irrigation can significantly increase total N stored within a soil profile, according to effluent composition, loading rates, and solution chemistry (Schipper et al., 1996; Cameron et al., 1997; Degens et al., 2000), However, some studies have reported that stored soil N decreased after effluent irrigation (Harris and Urie, 1983; Falkiner and Smith, 1997). In our study, effluent irrigation of soil without harvest residues did not significantly increase accumulated inorganic N above that measured under water irrigation (Table 2). Similarly, the retention of harvest residues did not significantly increase the accumulation of inorganic N in effluent-irrigated (drip and sprinkler) soil. Microbially immobilized N was not measured in the soil or harvest residues, but it was likely that harvest residue retention enhanced microbial biomass immobilization (Pilar et al., 2001) as the C and N status of the soil system was improved (Vance and Chapin, 2001).

In drip-irrigated treatments, 5 mo after irrigation had ceased, the largest accumulations of soil NH4–N were in the furrows as expected but were not directly beneath drippers; rather, they were midway between drippers. In contrast, the largest accumulations of soil NO3–N in the furrow were directly beneath dripper points. During drip irrigation of effluent, there is often an accumulation of NH4–N in the saturated zone beneath a dripper point, which subsequently retards nitrification (Haynes and Swift, 1987; Parchomchuk et al., 1993). In the unsaturated soil surrounding the dripper, nitrification activity and nitrifying bacteria numbers often increase (Haynes and Swift, 1987; Laher and Avnimelech, 1980). This situation may be reversed once irrigation ceases and environmental conditions promote net mineralization, as was observed in the soil sampled in November 2001 (Table 2).

Estimated Rates of Gross Nitrification and Ammonification
Rates of gross nitrification and ammonification estimated in this study (0.1–3.1 mg NO3–N kg–1 d–1 and 0.5–2.3 mg NH4–N kg–1 d–1) are comparable to rates obtained using 15N isotope pool dilution in other natural forest or managed plantation soils (Verchot et al., 2001; Garcia-Montiel and Binkley, 1998; Pedersen et al., 1999; Breuer et al., 2002, Ste-Marie and Houle, 2006; Burton et al., 2007). Our study is the first attempt to measure gross mineralization rates using 15N isotope pool dilution in an effluent-irrigated plantation. Other studies investigating the effects of continued effluent irrigation on N cycling processes in forest stands have only been able to report an increase in "net" mineralization or "potential" mineralization (Kim and Burger, 1997; Speir et al., 1999).

The retention of harvest residue significantly increased gross nitrification and ammonification rates under rain-fed conditions, probably by providing a carbon substrate for microbial community growth or by improving soil moisture conditions. The balance between N mineralization and immobilization is closely linked to C substrate availability and the C/N ratio of the available organic matter (Recous et al., 1999; Cookson et al., 2006). Soil water status is also a major constraint to gross mineralization (Murphy et al., 1998a,b; Jamieson et al., 1999), and separating the importance of a soil moisture effect from a C-substrate effect is important to understand N mineralization and cycling in ecosystems with litter accumulation or harvest residue retention. At our study site, irrigation with water significantly increased gross nitrification above that in rain-fed treatments, with or without harvest residues, whereas ammonification only significantly increased in soil without harvest residues (Table 3). Similarly, concentrations of NH4–N in soil water also did not increase under normal water irrigation (Fig. 4) but did when the rate of water irrigation was doubled, which might suggest that ammonification was still water limited when irrigated at the normal rate (1.5–2.0 mm d–1).

Irrigation with effluent significantly increased gross nitrification rates above those measured under water irrigation but only in soil with harvest residues. This suggests a C substrate effect because the soil moisture effect had been eliminated. Nitrification is most commonly dominated by autotrophic nitrifiers that use CO2 as a C substrate and gain energy from the oxidation of inorganic NH4–N to NO3–N (Schimel and Bennett, 2004). However, when C is not limiting, heterotrophic nitrifiers, which use organic C as an energy source, can outcompete autotrophic nitrifiers for available NH4–N while using organic N compounds (Tietema and Wessel, 1992). The significant increase in the gross rate of nitrification when an excess supply of N was accompanied by an abundant supply of C suggests that heterotrophic nitrification has increased its proportional contribution to NO3–N production and may even be dominant (Muller et al., 2003). Furthermore, estimated rates of gross nitrification only exceeded gross ammonification in effluent-irrigated soil with harvest residues, which would be possible if the importance of heterotrophic nitrification had increased because NH4–N is not produced as an intermediate step toward NO3–N (Burton et al., 2007).


   Conclusions
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results
 Discussion
 Conclusions
REFERENCES
 
Effluent irrigation of forest systems requires careful consideration of the changing environmental conditions, the soil water and nutrient holding capacity, and preferential flow of NO3–N and NH4–N. Contamination of groundwater under effluent irrigation may not necessarily occur through the downward movement of nitrate but may occur through other N forms, such as NH4–N or organic N. Opportunities to improve the capacity of a forest soil system to store and transform N inputs through organic matter addition or retention must take into consideration the delicate balance and dynamic equilibrium between N input, uptake, and immobilization and the effect changing microbial processes and environmental conditions can have on this equilibrium.


   ACKNOWLEDGMENTS
 
This study was funded by an ARC Linkage grant with the cooperation of the Water Corporation of Western Australia. This study built on the research efforts of Michael Coote and was greatly assisted by Kate Bowler, Kent Heard, Matt Todd, Andrew Bussau, Abe Hitofumi, Peter Landman, Kelly Whyte, Mike Kemp, Doug Ford, and Alistar Grigg. Assistance with Mass Spectrometry was provided by Lidia Bednarek of the West Australian Biogeochemistry Centre at UWA. We would like to recognize the consultation and assistance provided by Gunnar Horn and Ken Eade of the Water Corporation and the operations assistance of the Forest Products Commission, Albany, WA.


   NOTES
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 NOTES
ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results
 Discussion
 Conclusions
 REFERENCES
 
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   REFERENCES
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 NOTES
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results
 Discussion
 Conclusions
 REFERENCES
 





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