Nitrogen leaching after fertilising young Pinus radiata plantations in New Zealand

Forest Ecology and Management 280 (2012) 20–30

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Nitrogen leaching after fertilising young Pinus radiata plantations in New Zealand Murray Davis a,⇑, Graham Coker a, Michael Watt a, Doug Graham a, Stephen Pearce a, John Dando b a b

Scion, P.O. Box 29237, Christchurch, New Zealand Landcare Research, Private Bag 11052, Palmerston North, New Zealand

a r t i c l e

i n f o

Article history: Received 4 April 2012 Received in revised form 29 May 2012 Accepted 7 June 2012 Available online 7 July 2012 Keywords: Forest fertilisation Urea Nitrate-N Ammonium-N Organic-N Leaching

a b s t r a c t Although there is an opportunity to increase the productivity of Pinus radiata plantations in New Zealand through nitrogen (N) fertilisation, this treatment may reduce the quality of drainage water through leaching of nitrate-N. To improve our understanding of the effects of N fertilisation on potential N leaching, N concentrations in soil water below the root zone and potential losses were investigated after N fertiliser was applied as urea at 200 kg ha1 of N to 7–9 year-old operational plantations at 10 sites in New Zealand. A water balance model was used to estimate soil water drainage on a daily basis. Annual drainage ranged from 70 to 1,199 mm and was closely correlated with annual rainfall. Nitrate-N, ammonium-N and organic-N concentrations in unfertilised control plots ranged between 0.03–2.26 mg L1, 0.06– 0.49 mg L1 and 0.44–0.95 mg L1 respectively. Fertilisation significantly increased nitrate-N concentrations at four sites but did not affect ammonium-N or organic-N concentrations. In the 2 years after fertiliser application, fertiliser increased nitrate-N leaching at eight of the 10 sites by 0–15 kg ha1 (average 6.4 kg ha1 or 3.2% of the N applied). These values underestimate total nitrate-N loss as fertiliser enhanced leaching at two sites had not ceased after 2 years. Nitrogen fertilisation caused much greater increases in nitrate-N loss at a low productivity, high rainfall site (28 kg ha1of N), and a coastal sand site (90 kg ha1of N). To reduce losses at such sites, fertiliser should be applied at reduced rates in multiple applications. Factors that seem to have pre-disposed sites to nitrate-N leaching following N fertilisation include a pasture land-use history, the presence of a high component of the N-fixer Ulex europaeus in the understory, and soil C/N ratios of 15 or lower. Organic-N dominated leaching at 7 of the 10 sites, consistent with the pattern observed in unpolluted old-growth southern hemisphere forests, but mineral-N dominated at three sites, two of which had a recent land-use history of fertilised pasture. Ó 2012 Elsevier B.V. All rights reserved.

1. Introduction The forest industry in New Zealand is based on plantations of the exotic conifer Pinus radiata D. Don and currently contributes around 3% of the national gross domestic product. The forestry sector has a desire to improve profitability by doubling productivity on a per hectare basis (Anon, 2012). Forest productivity may be increased by use of high growth rate breeds, control of weed competition at forest establishment, control of insects and pathogens, and fertilisation. Many forest areas in New Zealand have nutrient levels that are below optimum for growth, and fertilisation is potentially an important tool for increasing plantation productivity (Will, 1985; Hunter et al., 1991). Nitrogen (N) is one of the key nutrients for which deficiencies have been recorded in New Zealand plantations, and studies have indicated the potential of N fertilisation to increase productivity of P. radiata plantations in a range of soils (Hunter, 1982; Hunter et al., 1985; Mead et al., 1984; Will, 1985; Woollons and Will, 1975). ⇑ Corresponding author. Tel.: +64 3 3642949; fax: +64 3 3642812. E-mail address: [email protected] (M. Davis). 0378-1127/$ - see front matter Ó 2012 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.foreco.2012.06.009

Although N concentrations in water draining from forest plantations are typically lower than that of other major land uses in New Zealand (Elliot et al., 2005; Hamilton, 2005; Larned et al., 2004) and elsewhere (e.g. Binkley et al., 1999), application of N fertiliser has the potential to reduce drainage water quality (Binkley et al., 1999). High nitrate-N concentrations are of concern because of possible risks to human health in drinking water and nutrient enrichment and increased productivity of aquatic ecosystems. High N loss is also of concern from the point of view of fertiliser use efficiency. In their review, Binkley et al. (1999) reported that N fertilisation generally increased both peak and average nitrateN concentrations in streamwater from forests as well as in soil-drainage water sampled beneath the majority of tree roots. Elevated nitrate-N concentrations may persist for at least a year after fertilisation in both stream water and root zone soil water (Binkley et al., 1999). Few studies of the effect of N fertilisation on the quality of streamwater or soil drainage water of P. radiata plantations have been undertaken in New Zealand. In two catchment-scale studies Neary and Leonard (1978), found aerially applied N fertiliser caused short-term increases in nitrate-N and total N concentrations in

M. Davis et al. / Forest Ecology and Management 280 (2012) 20–30

stream water, however some of the increase was due to fertiliser falling directly into stream beds. Contrasting effects of N fertiliser have been found on soil drainage water. On pumice soil, no N was found to have leached for a period of 3 years after application of 200 kg ha1 of N to 13-year-old forest (Worsnop and Will, 1980). In contrast, in a young coastal sand dune forest being re-established after harvest, Smith et al. (1994) found N application at 900 kg ha1 (applied quarterly at 50 kg ha1over 4.5 years) substantially increased soil water nitrate-N concentrations 6 months after application. Nitrate-N levels remained elevated throughout the monitoring period for more than 2 years after trial establishment. Although such high fertilisation rates would not be used operationally, the trial demonstrated the potential of sands to leach nitrate-N after N fertilisation. While pumice soils and coastal sands are important for forestry in New Zealand, further study of N fertiliser impacts on N leaching in a wider range of soil groups important for forestry is required. Moreover, plantations in New Zealand were historically established after clearing indigenous vegetation, but since the 1960s many plantations have been established in pastures during periods when economic returns from livestock farming have been poor. Such plantations would have higher labile soil N than soils of plantations established after clearing indigenous vegetation and may therefore be more prone to N-leaching after N-fertilisation. To better understand the effects of fertilisation on N concentrations in soil drainage water and potential leaching of N, concentrations and potential losses of N were investigated after N fertiliser was applied to 7–9 year-old stands in operational plantations varying in rainfall, prior land-use and soil characteristics at 10 sites across New Zealand. Nitrogen fertiliser is normally applied early rather than late in the rotation to New Zealand P. radiata plantations. Because of high demand for N, leaching following fertiliser application should be relatively limited in young stands. Further, compared to the northern hemisphere, atmospheric deposition of N is low in New Zealand (Parfitt et al., 2006). The hypothesis of this study is that as a result of high demand by the young crop and low atmospheric deposition rates, fertiliser application at standard rates should not greatly increase the risk of N leaching loss where the prior land-use was indigenous vegetation, but losses may be greater where plantations were established in pasture.

2. Materials and methods 2.1. Site, forest stand and soil description Ten trial sites spanning an 8° latitude range and 600 m elevation range were included in the study (Table 1, Fig. 1). The trial sites were selected from a group of 35 Long Term Site Productivity trial sites established to identify soil indicators influencing forest productivity (Watt et al., 2005, 2008). Six sites were located in the North Island while four were in the South Island (Fig. 1). At each site a single 40  40 m permanent sample plot was installed as part of the initial study. For the present N fertilisation experiment one additional plot receiving N fertiliser of the same size, and as similar as possible to the original plot in terms of soil, slope and aspect, was installed at each site. Topography at the selected sites ranged from relatively flat terrace, plain and plateau landscapes through to steep hill country with slopes up to 27°. The New Zealand climate is temperate and rainfall is relatively evenly distributed throughout the year. Precipitation varied threefold across the 10 sites, from 680 mm (site 7, central South Island) to 2,248 mm (site 3, central North Island) and mean annual temperature ranged from 8.9 °C (site 10, southeast coast South Island) to 13.3 °C (site 6, West Coast North Island) (Table 1). Soil pits were excavated as part of the original study and

21

a pedalogical description of the soils was made (Ross et al., 2009). Our sites encompassed six major soil orders and soil textures varied from well drained sands through to imperfectly drained silt loams and clay loams (Table 1). The forest stands were aged from 7 to 9 years at the time of trial establishment. Apart from site 6, which was first rotation, all sites were second or third (sites 5 and 8) rotation P. radiata plantation forests. Tree heights and diameters measured in winter 2009 varied threefold across sites and there was a nearly eightfold variation in stand basal area (Table 1). Stand basal areas were low at sites 2 and 8 because of low tree growth rates and at site 3 because the stand had been thinned at age 8, 1 year before the trial was established. Site 1 was previously an agro-forest, with low stand density, that had been planted in high productivity fertilised grazed pasture containing the N-fixer Trifolium repens as a dominant (Hawke and O’Connor, 1993; Yeates et al., 2000). The low tree density of the agro-forest (50 trees ha1) allowed continued growth of fertilised pasture and sheep grazing under the forest canopy. The plantation at site 6 was also planted into improved, fertilised pasture. Sites 6 and 10 had soil Ap horizons, indicating incorporation of organic matter as a result of cultivation or increased microbial activity associated with fertilising or manuring. The plantations at the remaining sites were planted directly after clearing indigenous forest or shrubland, or planted into indigenous grassland or into poor grassland that was reverting to shrubland. Estimates were made of the ground cover percentage of N-fixing species that were present at six sites. No N fixing species were observed at sites 1, 2, 4 and 10. Small amounts (5% or less of ground cover) of herbaceous legumes (Lotus and Lupinus spp.) were present at site 3 and site 5 in both control and fertilised plots. Cytisus scoparius (15%) was also present in the fertilised plot at site 3. At site 6 Lotus and C. scoparius were present (<5%) in the control plot, and a larger amount (20%) of Ulex europaeus was present in the fertilised plot. U. europaeus was also present at site 2 in control (5%) and fertilised (1%) plots, at site 7 (45% and 25% respectively) as well as site 8 (30% and 15% respectively). A small amount (<5%) of C. scoparius was present in both plots at site 9. Soil samples (0–10 cm) were randomly collected from the original (control) plots when the trees were 4 years old. Twenty samples were collected per plot and bulked, air dried and passed through a 2 mm sieve. The <2 mm fraction was analysed for pH, organic carbon (C), total N and total phosphorus (P) and exchangeable cations using the methods described in Blakemore et al. (1987). Soil total C and N were also determined in samples collected from individual horizons of the soil pits excavated for pedalogical description. Fine earth (<2 mm) bulk density was determined in the horizon samples as described by Ross et al. (2009) to allow calculation of soil C and N mass. Soil C and N mass was determined to a depth of 0.9 m except for sites 8–10 where impediments precluded determination of bulk density to that depth. Foliage was randomly collected from 10 trees from both control and fertilised plots in March 2010. Current season foliage was collected from second-order shoots in the upper third of the crown using a shotgun (Will, 1985). Samples were bulked by plot, placed in a paper bag, then sent to the laboratory and analysed for plant nutrients after oven drying (70 °C) by the methods described in Davis et al. (2007). 2.2. Fertiliser application and soil water sampling Four suction cup lysimeters were installed in each plot at each site in winter 2009. The lysimeters were placed below the majority of roots at a depth of 1.0 m at all sites except sites 4, 7 and 8 where stones, rocks or other impediment precluded deeper placement. Lysimeter depths at these sites averaged 0.93, 0.77 and 0.65 m,

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M. Davis et al. / Forest Ecology and Management 280 (2012) 20–30

Table 1 Site level variation in physiography, climate, stand structure, water balance input variables, soil variables and land use history. Climatic data was obtained from the Virtual Climate Station Network and averaged over the 2 years during which the water balance model was run. For leaf area index, Lt, and stand dimensions values shown are averaged across the two treatments. Site numbers shown in the table follow Fig. 1. Site 1

Site 2

Site 3

Site 4

Site 5

Site 6

Site 7

Site 8

Site 9

Site 10

Physiography Latitude (°) Longitude (°) Altitude (m) Landscape

38.06 176.34 345 Hill

38.08 176.02 570 Plateau

38.27 176.67 450 Hill

39.45 175.12 559 Hill

39.47 175.56 690 Plain

43.15 172.58 270 Hill

43.43 172.30 150 Terrace

46.06 170.10 213 Hill

46.42 169.46 256 Hill

Slope (°)

3

3

4

27

2

40.19 175.34 70 Sand dune 25

15

1

14

7

1,856

1,931

2,248

1,543

1,188

1,016

680

721

833

1206

13.2

11.8

12.7

12.5

10.1

13.3

11.7

11.8

9.75

8.9

7.36

7.16

7.15

6.94

7.33

7.38

6.44

6.25

6.21

6.45

15.8 298

5.7 116

12.7 188

11.9 192

9.5 180

12.0 165

11.5 132

5.4 83

9.4 111

8.4 147

25.5

5.1

7.8

30.9

19.7

21.5

26.5

3.9

24.8

15.0

Water balance and soil variables Average Lt (m2 m2) 3.87 Wmax  Wmin (mm) 63 Soil order Pumice Soil texture Sandy loam

4.48 174 Podzol Loamy sand

2.15 57 Recent Gravelly sand

2.71 41 Allophanic Loamy silt

3.83 124 Allophanic Sandy loam

3.58 50 Brown Sand

4.27 110 Pallic Silt loam

1.46 110 Brown Stony sandy loam

5.71 84 Brown Stony silt loam

3.79 132 Brown Clay loam

Land use prior to pine plantation

Indigenous forest

Indigenous shrubland

Reverting grassland

Indigenous grassland

Fertilised pasture

Reverting grassland

Indigenous shrubland

Reverting grassland

Reverting grassland

Climate Total annual rainfall (mm year1) Annual average air temperature (°C) Annual average PAR (MJ PAR m2 day1) Stand structure Tree height (m) Tree ground line diameter (mm) Stand basal area (m2 ha1)

Fertilised pasture

respectively. Lysimeters were installed using a soil auger to excavate a hole as described by Close et al. (2004), but with a smaller diameter auger (60 mm). Fine sand (<300 lm) was placed at the bottom of the hole and around the suction cup at the base of the lysimeter before backfilling with soil excavated from the hole. The top of the hole was sealed with a layer of bentonite. Nitrogen was applied as urea to the N treated plots in November 2009 at a rate of 200 kg ha1of N. Soil water was collected at approximately six weekly intervals beginning 3 months before fertiliser application to the +N plots. Sampling continued until December 2011 (2 years after fertilisation) at most sites, but ceased earlier at some sites where analysis indicated that nitrate-N concentrations in fertilised plots had returned to pre-fertiliser levels. Samples were collected by applying a 60 kPa vacuum to the lysimeter tubes and samples were collected the following day. Samples were returned to the laboratory, filtered (0.45 lm) and stored under refrigerated conditions prior to analysis. Nitrate-N was determined on all samples while ammonium-N and total Kjeldahl-N were determined on samples collected prior to fertiliser addition and on samples collected between 6 and 12 months after fertilisation (May–October 2010). Nitrate-N and ammonium-N were determined colorimetrically with an auto analyser using methods described in Eaton et al. (1999). Nitrate-N was determined after cadmium reduction to nitrite-N followed by sulphanilamide-NAD reaction. Ammonium-N was determined before and after Kjeldahl digestion by the nitroprusside catalysed indophenol reaction. Organic-N was determined as the difference between total Kjeldahl-N and ammonium-N. 2.3. Determination of water balance

Fig. 1. Map showing the location of the 10 study sites in the North and South Islands, New Zealand.

A daily water balance equation was used to calculate daily available root-zone water storage, Wi, in each of the 20 site/treatment combinations as

M. Davis et al. / Forest Ecology and Management 280 (2012) 20–30

W i ¼ W i1 þ Pi  Eti  Etwi  Eui  F i

ð1Þ

where Pi is rainfall, Eti is transpiration from the dry tree canopy, Etwi evaporation of intercepted rainfall from the tree canopy, Eui evaporation from the understory and soil, and Fi drainage from the root zone (Whitehead et al., 2001). Drainage from the root zone was assumed to be zero when Wi 6 Wmax, where Wmax is the maximum root-zone water storage, and equal to rainfall reaching the soil when Wi > Wmax. Transpiration from the tree canopy was calculated using the simple diffusion equation:

Et ¼ Dg st Lt

ð2Þ

where D is the air saturation deficit, gst is the average stomatal conductance for the tree canopy and Lt is the projected tree leaf area index. The inverse relationship between stomatal conductance and air saturation deficit was modelled using the following function (Lohammer et al., 1980):

Gst ¼ ug stmax =ð1 þ ðD  Dsmin Þ=D0t Þ

ð3Þ

where gstmax describes maximum tree stomatal conductance at low air saturation deficit, D0t is the sensitivity of gst to D, when D > Dsmin (value of D below which gst is constant, set at 0.5 kPa) and u will be defined later. Following the approach outlined in Landsberg and Waring (1997), Etw/P, was scaled linearly from 0 at Lt of 0, to a maximum of 0.30 at Lt of 3, using the parameter I (=0.10) in the following function:

Etw =P ¼ Lt I

ð4Þ

Evaporation from the understory and soil surface was calculated from the available energy beneath the tree canopy, Gg, using the expression:

Eu ¼ xuðssGg =½kðs þ cÞÞ

ð5Þ

where the coefficient s (=1.4) describes the degree of coupling of the soil surface with the air above the canopy (Kelliher et al., 1990), s is the slope of the relationship between saturated vapour pressure and temperature at a given air temperature, k is the latent heat of vaporisation, and c is the psychometric constant. Values for s, c, and k are temperature dependant and calculated from standard meteorological tables. Using Beers Law, Gg was calculated from (ekLtGa) where Ga is the available flux density above the tree canopy (assumed to be 70% of shortwave radiation) and k is the light extinction coefficient (assumed to be 0.5 for a spherical leaf angle distribution). The coefficient u was used to reduce Eu and gstmax as soil water storage declined. The value of the coefficient u was set to 1 at maximum values of fractional available root-zone water storage, Wf (=(Wi  Wmin)/(Wmax  Wmin)) of 1, and was not reduced until Wf declined to a threshold value (Wt) of 0.6 and 0.55, respectively, for Eu and gstmax. As Wf progressively declined below this threshold u was linearly reduced from 1 to 0 at minimum values of Wf (=0). Evaporation from the soil and transpiration were reduced to 50% of their potential rates on days when rain fell. Transpiration was also reduced by 50% on days when an air temperature <0 °C was recorded. Daily meteorological data required for the water balance model include total rainfall and solar radiation, mean air temperature, minimum air temperature, and average vapour pressure deficit. These data were obtained from the Virtual Climate Station Network (VCSN) developed by NIWA (National Institute of Water and Atmospheric Research Ltd.). Daily VCSN data were estimated for the whole of New Zealand on a 0.05 latitude/longitude grid (Tait et al., 2006; Tait, 2008; Tait and Liley, 2009). This spatial

23

interpolation used a thin-plate smoothing spline model (ANUSPLIN; Hutchinson, 2010), utilising two location variables (latitude and longitude) and, for some variables, a third ‘‘pattern’’ variable (usually elevation). Data were extracted from the VCSN grid point nearest to each site. The model also requires values for the parameters gstmax, D0t, available volumetric water content (Wmax  Wmin) and measurements of projected leaf area index, Lt. Values of gstmax and D0t were set to 0.1 mol m2 s1 and 0.5 kPa, respectively. Treatment level values of Lt at each site were measured in the field using a canopy analyser (LAI-2000, Li-Cor Inc., Lincoln, Nebraska, USA) over the first year of the study. At the North Island sites gravimetric soil samples were taken every 6 weeks between August 2009 and December 2010 and periodically thereafter. At the South Island sites, gravimetric soil samples were taken at six weekly intervals between August 2009 and October 2009 after which soil moisture was determined by time domain reflectometry (TDR). These gravimetrics and TDR measurements were used to define Wmax  Wmin. The data were rescaled to a fractional available root-zone water storage, Wf, and used to determine the accuracy of the water balance model at simulating seasonal changes in Wf. 2.4. Data analysis Mean concentrations of soil water nitrate-N, ammonium-N and organic-N in control and fertilised plots were determined for each sampling date. Site mean concentrations in control plots were determined over the whole sampling period, and in both control and fertilised plots over the period after fertiliser was applied, beginning January 2010. Analysis of variance was undertaken to determine the main effect of fertiliser on soil water nitrate-N, ammonium-N and organic-N concentrations after fertiliser application. For this analysis, individual sites formed the replicates. The T-test (assuming unequal variances) was used to determine, for individual sites, if soil water N concentrations in samples from fertilised plots differed significantly from samples in control plots. This comparison used the means of each sampling date within each plot after fertiliser application, beginning January 2010, as replicates. The mass of each N-form that was potentially leachable from below the root zone of control plots on an annual basis was calculated by multiplying the mean concentration of the particular Nform over the whole sampling period by the estimated annual drainage, determined from the water balance model. The mass of each N-form that was potentially leachable after fertiliser application was determined for both control and fertilised plots by multiplying the mean concentration of each N-form after fertiliser application by the amount of drainage that occurred after fertilisation, up until the time sampling ceased. As with the nutrient concentration data, the post-fertiliser application data from control and fertilised plots were subjected to analysis of variance, using sites as replicates, to determine the effect of fertiliser on potential leaching of nitrate-N, ammoniumN and organic-N. Simple linear correlation analysis was used to examine relationships between concentrations and flux of the different N forms and site variables in Tables 1 and 2. 3. Results 3.1. Soil analysis Analyses of the upper soil layer of control plots showed the soils to be moderately to strongly acid (Table 2). Soil C and N concentrations varied about threefold across sites and were closely correlated (r = 0.83, P < 0.01). Soil C/N ratios were low at sites 1, 4 and

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M. Davis et al. / Forest Ecology and Management 280 (2012) 20–30

Table 2 Soil pH and total C, N and P concentrations in the upper (0–10 cm) mineral soil of control plots and total C and N mass to 0.9 m depth, or less where shown, and N and P concentrations of P. radiata in control plots. Concentrations in 0–10 cm mineral soil samples and foliage samples are from 20 and 10 subsamples respectively, combined into one sample. C and N masses were determined from sampling horizons of a soil pit outside of the plots. Site

1

2

3

4

5

6

7

8

9

10

Soil pH C (%) N (%) C/N P (mg kg1) C (Mg ha1) N (Mg ha1) C/N C and N mass sample depth (m)

5.38 3.89 0.28 13.9 733 194 12.8 15.2 0.90

5.81 5.91 0.31 19.1 290 227 9.1 24.9 0.90

5.88 2.93 0.14 20.9 180 71 2.7 26.3 0.90

5.67 6.47 0.45 14.4 777 263 13.6 19.3 0.90

5.62 8.22 0.43 19.1 531 240 13.2 18.2 0.90

5.73 2.85 0.19 15.0 501 85 6.8 12.5 0.90

4.73 3.25 0.16 20.3 103 92 7.1 12.9 0.90

5.59 6.17 0.28 22.0 399 72 2.9 24.8 0.40

4.64 4.40 0.14 31.4 214 154 6.1 25.2 0.65

4.74 8.58 0.35 24.5 697 150 6.8 22.1 0.70

1.62 0.22

1.49 0.11

1.36 0.16

1.50 0.16

1.56 0.16

1.26 0.14

1.52 0.09

1.22 0.16

1.30 0.17

1.30 0.23

6; the low values at sites 1 and 6 reflect their recent fertilised pasture histories and pasture legume-N input. The high soil N concentration of site 4 suggests this site may also have received high Ninputs. Total P was correlated with N (r = 0.75, P < 0.05), but not C concentrations. Total P varied fourfold across sites and was particularly low at site 7, which was known to be P deficient. Sites 1, 2, 4 and 5 had high profile contents of C and N indicative of their Pumice, Allophanic and Podzol soil orders (Ross et al., 2009), while the Recent soil (site 3) and Brown shallow stony soil (site 8) had particularly low C and N contents and the Brown sand (site 6) had low C content. Profile C/N ratios were low at sites 1 and 6 which had a recent fertilised pasture history and at site 7 where there was a high (45%) cover of U. europaeus. 3.2. Water balance Modelled values of Wf closely corresponded to measured values. Good temporal correspondence between measured and modelled Wf was evident at all sites including two unfertilised treatments at sites representing the extremes in rainfall (Fig. 2). Mean annual site root zone soil water drainage estimated from the model ranged between 70 mm and 1199 mm (Table 3). Mean annual drainage across the 10 sites differed very little between unfertilised (468.5 mm) and fertilised plots (468.6 mm). There was a close positive linear relationship (P < 0.01) between drainage and rainfall (Fig. 3), with rainfall explaining almost all of the variance in drainage (R2 = 0.96). Extrapolation of the relationship to the x-axis showed drainage did not occur until annual rainfall exceeded 627 mm. 3.3. Soil water nitrate-N, ammonium-N and organic-N concentrations Mean soil water ammonium-N concentrations in unfertilised plots ranged between 0.06 and 0.15 mg L1, except for site 6 where the mean concentration was 0.49 mg L1 (Table 3). Mean soil water organic-N concentrations in unfertilised plots were generally higher than nitrate and ammonium-N concentrations, but varied over a smaller range (0.44–0.95 mg L1, Table 3). Analysis of variance showed that, across all sites, N application did not significantly affect soil water ammonium-N or organic-N concentrations (P = 0.24, and P = 0.77 respectively) and T-tests failed to reveal any significant differences at the individual site level. Mean soil water nitrate-N concentrations in control plots over the full sampling period ranged between 0.03 and 2.26 mg L1 (Table 3). Concentrations in control plots were generally stable, but changed over the sampling period at two sites (sites 1 and 8, Fig. 4). At site 1 concentrations were low initially (0.04 mg L1),

(a)

1.0

0.8

Normalised volumetric water content

Foliage N (%) P(%)

0.6

0.4

0.2

0.0 1.0

(b)

0.8

0.6

0.4

0.2

0.0 0

200

400

600

800

Days after 1-June-2009 Fig. 2. Relationship between modelled (dotted line) and measured (solid circles) normalised volumetric water content in unfertilised treatments at sites with the (a) highest and (b) lowest mean annual rainfall over the duration of the study.

but after June 2010 concentrations increased (average 3.67 mg L1). The opposite pattern occurred at site 8 where concentrations were initially high, (average 3.34 mg L1), but from July 2010 to October 2011, concentrations declined to an average of 0.11 mg L1. Analysis of variance showed that N fertilisation significantly increased average soil water nitrate-N concentrations from 0.64 to 5.64 mg L1 (P = 0.05). T-tests of individual sites showed N fertilisation did not significantly affect soil water nitrate-N concentrations at five sites (sites 1, 3, 5, 8, 9), significantly (P < 0.05) increased concentrations above values found in unfertilised plots at four sites (sites 2, 4, 7, 10), and apparently (but not significantly) increased concentrations at one further site (site 6) (Table 4). Nitrate-N concentrations increased rapidly (within 2 months) after fertilisation at two sites (sites 7 and 10, Fig. 4). Concentrations then

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M. Davis et al. / Forest Ecology and Management 280 (2012) 20–30

Table 3 Estimated annual drainage (mean of control and fertilised plots, mm) and mean (and standard error) soil water nitrate-N, ammonium-N and organic-N concentrations (mg L1) in control plots. 1

2

3

4

5

6

7

8

9

10

Control + Fert. Drainage

781

778

1,199

668

372

264

74

126

70

351

Control Nitrate-N Ammonium-N Organic-N

2.26 (0.791) 0.06 (0.010) 0.51 (0.275)

0.14 (0.021) 0.09 (0.024) 0.62 (0.308)

0.05 (0.011) 0.07 (0.008) 0.49 (0.307)

0.03 (0.004) 0.07 (0.008) 0.63 (0.343)

0.04 (0.009) 0.07 (0.014) 0.61 (0.282)

1.66 (0.463) 0.49 (0.192) 0.48 (0.279)

0.31 (0.093) 0.16 (0.113) 0.74 (0.242)

1.35 (0.573) 0.12 (0.041) 0.95 (0.231)

0.06 (0.015) 0.06 (0.020) 0.44 (0.288)

0.13 (0.040) 0.15 (0.060) 0.70 (0.231)

Annual drainage (mm)

Site

1500 y = 0.6657x - 418.38 R≤ = 0.96

1000 500 0 0

500

1000 1500 2000 Annual rainfall (mm)

2500

Fig. 3. Relationship between annual rainfall and annual soil water drainage.

declined at these sites to pre-fertiliser levels by 8 months after fertilisation at site 10, and 15 months after fertilisation at site 7. Site 6 appeared to follow a similar response pattern (Fig 4), however insufficient lysimeter samples were collected at this site to adequately characterise the response. Although the difference between unfertilised and fertilised plots at site 6 was not significant (P = 0.26), the highest concentration measured in fertilised plots (81 mg L1) was recorded there in a sample collected 7 months after fertilisation. Peak concentrations recorded at sites 7 and 10 were 39.3 and 8.6 mg L1 respectively, both samples being collected 2 months after fertiliser application. Nitrate-N concentrations at the other two sites to be significantly affected by fertiliser (sites 2 and 4, Fig. 4) showed a more delayed but ongoing response. The first increase in nitrate-N concentrations at these sites occurred 9 and 11 months after fertilisation respectively, and concentrations then remained elevated above pre-fertilisation concentrations for the remainder of the sampling period. Peak concentrations recorded for these sites (3.8 and 2.5 m mg L1 respectively) were lower than sites showing the more immediate response. 3.4. Leaching of nitrate-N, ammonium-N and organic-N Potential leaching of nitrate-N from the root zone of control plots (Table 5) was less than 1.7 kg ha1 year1 at eight of the 10 sites, but greatly exceeded this value at sites 1 (17.2 kg ha1 year1) and 6 (4.5 kg ha1 year1). At site 1, potential leaching increased over the sampling period from 0.2 kg ha1 year1 over the first 11 month period to June 2010, to 28.7 kg ha1 year1 over the final 18 months because of increasing nitrate-N concentrations after June 2010 (see above). Analysis of variance showed nitrogen application significantly increased potential leaching of nitrate-N on fertilised plots from 5.6 to 22.6 kg ha1 (P = 0.04). The increase was negligible (0–2 kg ha1) at four sites, ranged between 10–15 kg ha1 at four sites and was greater at two sites (sites 2 and 6, 28 and 91 kg ha1 respectively) (Table 6). The estimates for sites 6 and 9 are not robust because of the limited numbers of samples obtained at those sites. The increases at sites 1, 2 and 4 are underestimates as fertiliser enhanced leaching had not ceased 2 years after fertilisation at those sites. Average leaching due to fertiliser application for the first 2 years after fertiliser application amounted to 17.0 kg ha1.

Potential leaching of ammonium-N from the root zone of control plots was mostly less than 1 kg ha1 year1, with leaching at only one site being greater (Table 5). Potential leaching of organic-N ranged from 0.28 to 5.70 kg ha1 year1, and was generally greater than nitrate-N and ammonium-N combined. However at three sites (sites 1, 6 and 7) potential leaching of mineral-N exceeded that of organic-N. Nitrogen application did not significantly affect potential leaching of either ammonium-N or organic-N (P = 0.76 and 0.50 respectively). 3.5. Relationships with site variables A number of significant correlations between concentrations and potential leaching of nitrate, ammonium and organic-N and site variables were observed, however data plots showed these were invariably driven by a single data point. The only exception was for potential leaching of organic-N which was significantly correlated with both rainfall and drainage (both P < 0.01), which explained 83% and 89% of the variance, respectively (Fig. 5). 4. Discussion Average nitrate-N concentrations in soil drainage water collected from beneath the majority of roots in unfertilised managed forests is generally <1 mg L1, occasionally rising as high as 4– 5 mg L1 in areas of heavy atmospheric N deposition or where the forest ecosystem has been severely disturbed (Binkley et al., 1999). Nitrogen leaching from pine forest planted directly after clearing indigenous vegetation is normally low in New Zealand (Dyck et al., 1981; Knight and Will, 1977), however leaching may be greater where pine is planted into fertilised pasture (Dyck et al., 1987; Parfitt et al., 2002). Thus, land-use history has an important bearing on leaching rates from pine forest. At seven sites in this study, unfertilised control plots had average nitrate-N concentrations of less than 0.5 mg L1, consistent with unfertilised forests elsewhere. The marginally higher nitrate-N concentrations (1.3–2.3 mg L1) at the remaining three sites may be explained by recent pasture history of the sites (sites 1 and 6) and/or the presence of shrubby N-fixers in the forest understory (sites 6 and 8). Site 1 had a history of high quality fertilised pasture containing the N-fixer Trifolium repens prior to conversion to agroforestry with fertilised pasture continuing in the understory after conversion (Hawke and O’Connor, 1993; Yeates et al., 2000). This history would have ensured greater labile soil N and soil nitrate-N concentrations than forests established directly after removal of indigenous forest or shrubland, or into low quality grassland. Similarly, site 6 had a fertilised pasture history, although the control plot there had one rotation of pine where pasture legume N inputs would only have occurred in the early part of the plantation prior to canopy closure. The presence of the N-fixer U. europaeus may have contributed to the marginally elevated nitrate-N concentrations in the control plot at site 8. Enhanced nitrate-N concentrations in lysimeter soil water extracts have been observed under

26

M. Davis et al. / Forest Ecology and Management 280 (2012) 20–30

a

Fig. 4. Mean nitrate-N concentrations in control and fertilised plots from August 2009 to December 2011; (a) sites 1–5, (b) sites 6–10. Arrows show timing of fertiliser application. Note variation in scale of the y-axis at sites 6 and 7 from that of other sites.

U. europaeus shrubland in New Zealand (Dyck et al., 1983; Magesan et al., 2012). However, the presence of a relatively high ground

cover of U. europaeus did not result in enhanced nitrate-N concentrations in the control plot at site 7.

M. Davis et al. / Forest Ecology and Management 280 (2012) 20–30

27

Fig. 4 (continued)

Forests have a strong capacity to retain N after fertilisation (Aber et al., 1998; Neary et al., 2009). Estimated nitrate-N losses in the present study as a result of fertilisation in the 2 years following application ranged between 0 and 15 kg ha1 (0–7.5% of the N

applied) at eight of the 10 sites, however estimated leaching was substantially greater at two sites. Greatest losses (91 kg ha1, 45% of the N applied) were estimated for site 6 which is on a coastal sand soil in a moderate rainfall zone. Profile drainage at site 6

28

M. Davis et al. / Forest Ecology and Management 280 (2012) 20–30

Table 4 Mean nitrate-N concentrations (mg L1) in control and fertilised plots (and standard error), and T-test P values for the period after fertiliser application, beginning January 2010. Note that control plot values differ from those in Table 3 which are derived from the full sampling period. Site

1

2

3

4

5

6

7

8

9

10

Control Fertilised P

3.14 (1.026) 3.34 (0.596) 0.87

0.08 (0.009) 2.79 (0.146) <0.01

0.06 (0.031) 0.25 (0.103) 0.14

0.03 (0.005) 1.44 (0.363) 0.01

0.04 (0.011) 0.03 (0.005) 0.37

2.00 (0.699) 37.31 (23.243) 0.27

0.32 (0.117) 11.20 (4.382) 0.04

3.19 (1.416) 2.70 (0.639) 0.77

0.05 0.010) 0.14 (0.097) 0.40

0.18 (0.108) 3.73 (1.291) 0.04

Table 5 Potential annual leaching (kg ha1) of nitrate-N, ammonium-N, organic-N and total-N from the root zone of unfertilised control plots over the entire sampling period. Values were estimated by multiplying annual drainage by the mean annual concentrations of nitrate-N, ammonium-N, and organic-N (Table 3). Total-N is the sum of nitrate-N, ammonium-N and organic-N. Site

1

2

3

4

5

6

7

8

9

10

Nitrate-N Ammonium-N Organic-N

17.21 0.46 3.88

1.12 0.73 4.95

0.58 0.87 5.7

0.21 0.49 4.4

0.15 0.25 2.25

4.5 1.33 1.3

0.24 0.12 0.57

1.66 0.16 1.17

0.04 0.04 0.28

0.47 0.54 2.52

Total-N

21.55

6.8

7.15

5.1

2.65

7.12

0.92

2.99

0.35

3.53

Table 6 Potential nitrate-N leaching (kg ha1) from the root zone of control and N fertilised plots after fertiliser was applied and the period (months) over which potential leaching was measured. Values in the difference column indicate the total amount of nitrate-N potentially leached from a single application of urea (200 kg ha1of N) except for sites 1, 2 and 4 where N fertiliser enhanced leaching of nitrate-N had not ceased 2 years after the fertiliser was applied. 1

2

3

4

5

6

7

8

9

10

Control Fertilised Difference Period

43.2 54.7 11.5 24

1.5 29.9 28.4 24

1 2.7 1.7 20

0.4 12.5 12.1 24

0.3 0.2 -0.1 24

5.8 96.6 90.8 10

0.3 15.5 15.2 22

2.7 2.9 0.2 22

0.1 0.3 0.2 17

0.8 11.1 10.3 19

Organic-N leaching loss (kg ha-1year-1)

Site

10 8 6 4 2 0 0

500

1000 1500 2000 Annual rainfall (mm)

2500

Fig. 5. Relationship between rainfall and organic-N leaching loss.

was estimated to be the fourth lowest of the sites considered, but mean soil water nitrate concentrations following fertilisation were more than three times greater than recorded at any other site and greatly exceeded water quality standards for nitrate of (10 mg L1). Lysimeter samples were only able to be collected at site 6 on four occasions following fertilisation so the estimate of nitrate loss for the site is not robust. However Thomas and Mead (1992) reported that 30% of the 150 kg ha1 of N applied to 2-year-old seedlings on a coastal sand site in a low rainfall environment was lost via leaching. When split into three applications of 50 kg ha1, no N was detected below the root zone indicating that split application at low rates may be an important management measure to mitigate leaching on sites prone to high rates of leaching of nitrate-N. Leaching losses associated with fertiliser application were also high at site 2 where they amounted to 28 kg ha1 (14% of N applied). Mean nitrate-N concentrations were increased by fertiliser at this site to a much lesser extent than site 6, however rainfall was high at this site, and greater leaching can be ascribed to both high drainage and elevated nitrate concentrations. Further losses could be expected from site 2 as, although declining, loss of N had not ceased 2 years after fertiliser application. Land-use history plays an important role in determining the leaching response of northern hemisphere forest ecosystems

where N is added in the form of atmospheric deposition (Aber and Driscoll, 1997; Aber et al., 1998). The results of the present study support this hypothesis for added N in the form of fertiliser. Nitrogen fertilisation significantly increased root zone soil water nitrate-N concentrations at four sites and potential N leaching losses at these and two additional sites. Five of these six sites had a recent or earlier pasture history. To improve pasture productivity it has been standard practice in New Zealand for farmers to apply phosphate fertiliser (and other limiting nutrients) to stimulate clover (Trifolium spp.) growth and N-fixation. With the aid of aircraft this practice extended to hill country in the 1950s. Thus, pine plantations planted into pasture after this time may have had increased labile soil N that might pre-dispose forests to N leaching after fertiliser application. One site additionally had a high cover of the N-fixer U. europaeus in the understory, also likely to increase labile soil N (Dyck et al., 1983 Magesan et al., 2012). Four of these sites had low C/N ratios (12–15) either in the upper soil layer or through the whole profile, or both, indicative of a pasture history. One site (site 10) did not have low C/N ratios, though it did have an Ap horizon. The four sites where fertilisation did not increase leaching were second or third rotation pine sites that had originally been planted into indigenous shrubland (sites 3 and 8) or poor quality grassland (sites 5 and 9), all having C/N ratios in excess of 18. Of the latter grassland sites, one (site 5) was in the third rotation of pine and was originally planted in indigenous grassland (Roche, 1990) that would have had little or no legume component and not have been fertilised. The pasture history of site 9 is unknown, but the high soil C/N ratios of this site suggest it would similarly have had little or no legume-N input. The leaching response at site 2 did not follow the above pattern. Site 2 leached considerable nitrate-N following fertilisation, but did not have a pastoral history and had only a low amount of the N-fixer U. europaeus in the understory (although this may have been greater in the past) and additionally had high C/N ratios in both the topsoil and through the profile. Site 2 was a low productivity site, as

M. Davis et al. / Forest Ecology and Management 280 (2012) 20–30

evidenced by small tree dimensions, in a high rainfall environment. It is possible that the added fertiliser N exceeded the capacity of the plants and microflora, especially mycorrhizae, to absorb the additional N, resulting in increased nitrate-N leaching (Aber et al., 1998). Ammonium-N concentrations and losses were low and unaffected by fertiliser application, consistent with the results of other studies which indicate that fertiliser typically has very little impact on soil drainage water ammonium-N concentrations even after extreme rates of N addition (Binkley et al., 1999). Organic-N concentrations were similarly unaffected by fertiliser application. At all but three sites in our study organic forms dominated N concentrations and losses in control plots, the three sites (1, 6 and 8) with higher inorganic-N loss having a recent land-use history of pasture with high legume-N inputs, or the N-fixer U. europaeus present in the understory. The predominance of losses by organic-N is consistent with the pattern of N losses from temperate old-growth southern hemisphere forests (Hedin et al., 1995; McGroddy et al., 2008; Perakis and Hedin, 2001). In contrast, in old-growth northern hemisphere forests, which receive high inorganic-N inputs from atmospheric pollution, inorganic-N forms dominate losses. Both rainfall and drainage explained most of the variance in loss of organic-N in our study, confirming that hydrologic flux is the primary determinant of ecosystem organic-N loss (McGroddy et al., 2008; Worrall and Burt, 2007). Concentrations of all forms of N in control plots in our study were greater than those McGroddy et al. (2008) reported for old growth indigenous forests in New Zealand. The greater concentrations in the present study may be due to sample collection location (soil drainage water compared to first order streams in the study of McGroddy et al., 2008), or alternatively the higher concentrations of our samples may reflect that they are from young plantations and thus disturbed ecosystems, compared to the undisturbed old growth forests sampled by McGroddy et al. (2008). Studies of attenuation of N or other nutrients as they pass from the root zone to the point of emergence in catchments are rare (Binkley et al., 1999), however Parfitt et al. (2002) found no attenuation of nitrate-N between the root zone point of collection (0.6 m depth) and adjacent springs in a New Zealand P. radiata plantation. This finding suggests that the greater concentrations in our P. radiata plantations may be a product of ecosystem disturbance rather than sample point location. Further studies of concentration attenuation along the gradient from root zone to catchment streamwater emergence point are needed to confirm this. Nitrate-N concentrations in unfertilised control plots generally remained low throughout the sampling period but there were two exceptions. Concentrations at site 1 were low initially but began to climb towards the end of the first year. A similar pattern of increasing nitrate-N leaching with time has been observed elsewhere in New Zealand with P. radiata planted into pasture (Parfitt et al., 2002). This pattern may arise on high-N pasture sites because N demand by the developing canopy is likely to be high initially, leaving little N available for leaching. As canopy expansion is completed and the understory is shaded out, N demand will decline rapidly, allowing more N to become available for leaching. In contrast to site 1, concentrations were initially high at site 8 before declining and then remaining consistently low from the end of the first year. Site 8 is a low rainfall, low productivity site and had a substantial cover of U. europaeus in the understory which may have been fixing N in excess of initial crop demands (Augusto et al., 2005; Magesan et al., 2012; Watt et al., 2003), allowing N to become available for leaching. With crop growth and canopy expansion, demand for N would increase and suppression of U. europaeus would occur, both factors would contribute to a decrease in N availability for leaching (Augusto et al., 2005).

29

5. Conclusions Nitrogen fertiliser application to 7–9 year-old P. radiata stands at a standard application rate of 200 kg ha1 of N increased nitrate-N leaching at eight of the 10 study sites from between 0 and 15 kg ha1 (average 6.4 kg ha1 or 3.2% of the N applied) in the 2 years following fertilisation. These are slight underestimates because of ongoing leaching at two sites, but are ‘one-off’ loss rates that would occur once in a rotation, and are less than would occur annually from most agricultural land uses. Nitrogen fertilisation caused much greater increases in nitrate-N leaching on a low productivity, high rainfall site (28 kg ha1of N), and a coastal sand site (90 kg ha1of N). To reduce losses and increase efficiency of fertiliser use at such sites, fertiliser-N should be applied at reduced rates in multiple applications. Low loss rates at other sites suggest that the standard rate of 200 kg ha1 of N is not excessive. Land use history had an important bearing on the propensity of sites to leach nitrate-N after N fertilisation. Forests planted into pasture, even after one rotation of pine, were more likely to lose nitrate-N by leaching than forests planted into indigenous vegetation. This study confirmed that nitrate-N losses from unfertilised plantation P. radiata forests are low (<2 kg ha1) relative to other agricultural land uses, except where the forests have been planted into fertilised pasture. The results also indicate that, where forest has been planted into fertilised, productive pasture, more than one rotation of forest may be required to reduce nitrate-N loss to pre-pasture levels. Nitrogen fertiliser application did not affect soil water ammonium-N or organic-N concentrations or leaching of these N-forms. Soil water organic-N concentrations generally exceeded mineral-N concentrations, consistent with forests receiving low inputs of atmospheric N. This may change in future to mineral-N dominance as pastoral farming in New Zealand has become more reliant on fertilisation and less reliant on pasture legumes for provision of N (Parfitt et al., 2006), a change which is likely to cause increased inputs of atmospheric N to forests. This will in turn have implications for nitrate-N leaching from forests as well as forest productivity and fertilisation requirements. The results of the present study should be generally applicable to other southern-hemisphere countries, such as Australia, Chile and South Africa, with low inputs of atmospheric N. Acknowledgements This research was supported by the New Zealand Foundation for Research, Science and Technology through the Future Forest Research Programme ‘‘Protecting and Enhancing the Environment through Forestry’’ (Contract C04X0806). We thank Morkel Zaayman, Diana Unsworth and Kaye Eason (Scion) for carrying out the soil water nitrogen analyses. We also thank the following forest owners and managers for allowing the study to be conducted in their forests: Dennis Hocking; Ernslaw One; Hancock Natural Resource Group; Kaingaroa Timberlands Ltd.; OTPP New Zealand Forest Investment Ltd./P.F. Olsen Ltd.; Rayonier/Matariki Forests; Wenita Forest Products. References Aber, J.D., Driscoll, C.T., 1997. Effects of land use, climate variation and N deposition on N cycling and C storage in northern hardwood forests. Global Biogeochem. Cycles 11, 639–648. Aber, J., McDowell, W., Nadelhofer, K., Magill, A., Berntson, G., Kamakea, M., McNulty, S., Currie, W., Rustad, L., Fernandez, I., 1998. Nitrogen saturation in temperate forest ecosystems. Hypothesis revisited. BioScience 48, 921–934. Anon, 2012. New Zealand Forestry Science and Innovation Plan. N. Z. For. Owners Assoc., Wellington. Augusto, L., Crampon, N., Saur, E., Bakker, M., Pellerin, S., de Lavaissiere, C., Trichet, P., 2005. High rates of nitrogen fixation of Ulex species in the understory of

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