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Impact of harvest residues on soil mineral nitrogen dynamics following clearfall harvesting of a hoop pine plantation in subtropical Australia
Forest Ecology and Management 179 (2003) 55–67
Impact of harvest residues on soil mineral nitrogen dynamics following clearfall harvesting of a hoop pine plantation in subtropical Australia T.J. Blumfielda,b,*, Z.H. Xua,b,c,1 a
Cooperative Research Centre for Sustainable Production Forestry, Griffith University, Nathan, Qld 4111, Australia b Faculty of Environmental Science, Griffith University, Nathan, Qld 4111, Australia c Queensland Forestry Research Institute, PO Box 631, Indooroopilly, Qld 4068, Australia Received 2 May 2002; accepted 4 September 2002
Abstract An alternative management strategy allowing post-harvest residues to remain as a blanket cover instead of being incorporated into windrows may prevent problems associated with lack of soil fertility in the inter-windrow space. Parallel, 2-year, in situ nitrogen (N) mineralisation studies were undertaken during the inter-rotation period following clearfall harvesting of a first-rotation hoop pine (Araucaria cunninghamii Aiton ex D. Don) plantation in subtropical Australia. We investigated the dynamics of ammonium N, nitrate N and nitrite N in the top 20 cm soil under a residue retention situation and under normal, operational conditions. Initially, ammonium N was the dominant form of soil mineral N but declined to <10 kg N ha1 at both sites after 12 sampling cycles of 28-day duration. Nitrate N levels remained at approximately 30 kg N ha1 despite seasonal fluctuations, throughout the 2-year sampling period. At the residue site there was no net N mineralisation in the soil until the 20th sampling cycle; approximately 110 kg N ha1 was mineralised at the end of the sampling period. This compared with approximately 300 kg N ha1 mineralised at the operational site for the same period. Approximately 100 and 220 kg N ha1 were lost through leaching at the residue and operational sites, respectively. Residue retention promoted N immobilisation, reducing the potential for N losses through leaching and denitrification in the critical inter-rotation and early establishment period. # 2002 Elsevier Science B.V. All rights reserved. Keywords: In situ mineralisation; Post-harvest residues; Soil mineral nitrogen dynamics; Windrows
1. Introduction *
Corresponding author. Present address: Cooperative Research Centre for Sustainable Production Forestry, Griffith University, Nathan, Qld 4111, Australia. Tel.: þ61-7-38756709; fax: þ61-7-38757459. E-mail addresses: [email protected] (T.J. Blumfield), [email protected] (Z.H. Xu). 1 Tel.: þ61-7-33629367; fax: þ61-7-38969628.
Nutrient cycling, in particular the dynamics of N within forested ecosystems, has been widely and intensively studied (Keeney, 1980; Carlyle, 1986; Attiwill and Adams, 1993; Nambiar, 1996; Ballard, 2000). However, it is becoming increasingly evident that some of the basic assumptions underlying such N
0378-1127/02/$ – see front matter # 2002 Elsevier Science B.V. All rights reserved. PII: S 0 3 7 8 - 1 1 2 7 ( 0 2 ) 0 0 4 8 5 - 1
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dynamics are based on research in northern hemisphere temperate and boreal forests, which was initiated after the onset of anthropogenic disturbance of the ‘natural’ system (Perakis and Hedlin, 2002). This may not, with certainty, be applied to southern hemisphere forests (van Breemen, 2002). Tropical and subtropical forests form 56% of the world’s total forest resources, while 45% of the world’s forests are to be found in South America, Africa and Oceania, within the southern hemisphere (FAO, 2001). Currently, there is a paucity of studies of N mineralisation in subtropical and tropical forests (Bubb et al., 1998). Research into the nutrient dynamics of these forest systems, particularly intensively managed plantation systems, therefore becomes essential if we are to formulate suitable strategies for sustainable and environmentally focused forest management. A major operational change within the hoop pine plantation estate in subtropical Australia, has been the decision to retain post-harvest residues on-site by forming the residues into windrows. This decision was a response to concerns regarding N losses following burning of post-harvest residues (Costantini et al., 1997), these N losses occurring both through volatilisation and through erosion (Ryan and Gilmour, 1985). However, operational decisions regarding the windrowing of retained residues were driven by the nature of the available machinery, costs and workplace health and safety. The effect that changes to the residue-management regime would have on N dynamics was not known. Soil mineral N fluxes in first-rotation (1R) hoop pine plantations of various ages (Bubb et al., 1998) indicated that the N mineralisation rate was the highest in the growing season and that ammonium was the dominant mineral N form. The net soil N mineralisation was 25.3, 30.9 and 45.8 kg N ha1 over a 7-month period for 3-, 20- and 62-year-old plantations, respectively. However, the dynamics of soil mineral N fluxes in the inter-rotation and early second rotation (2R) phase have not been studied. Some attention has been given to soil organic matter dynamics and N pools (Mao et al., 2002; Mathers et al., 2002) as well as nutrient loss mechanisms under a windrow residue-management regime (Pu et al., 2001), but the effect of these practices on soil mineral N fluxes in the tree planting area between the windrows has not been explored.
The purpose of this study was to gain an understanding of soil mineral N dynamics in the critical inter-rotation and early second rotation periods using the in situ core mineralisation technique; particularly in relation to residue management, potential N losses through denitrification and leaching (Pu et al., 2001; Johnston and Crossley, 2002), and N retention through immobilisation (Tietema et al., 1998; Puri and Ashman, 1999). Successful residue management can benefit the plantation through increased growth (Jones et al., 1999), reducing the need for fertiliser inputs, and benefiting the environment by reducing off-site effects.
2. Materials and methods 2.1. Site description Two sites were established immediately following clearfall harvesting of a 1R hoop pine plantation in the Imbil State Forest (268310 S, 1528380 E) in Queensland. Each had a slight to moderate slope (10–158) with an easterly aspect. The forest lies within the subtropical zone, with a predominating weather pattern of cool, dry winters and hot, wet summers. This pattern is subject to variation in response to an overlap between tropical and temperate weather systems (Costantini et al., 1997) and to the effects of weather patterns induced by the southern oscillation. There is a wide variation in annual rainfall (495–1964 mm, mean 1188 mm). The average daily temperature ranges from 19.4 to 30.5 8C (mean 25.0 8C) during October to March; from 6.9 to 21.5 8C with a mean of 14.2 8C during April to September (Bubb et al., 1998). Rainfall and temperature were measured daily at the Imbil Forestry Office, which is situated within 5 km of the sites and considered representative of the area. The data for these variables have been presented in 28day increments to correspond with the sampling periods. 2.2. Site preparation One site (designated ‘operational’) was in an area that had been established using normal operational practice; windrows of post-harvest residues were formed along the contours of the slope using a D6
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Table 1 Chemical properties of soil profiles for the two study areas Electrical conductivity (mS m1)
Total K (mg kg1)
Total N (%)
Organic C (%)
Available P (mg kg1)
Total P (mg kg1)
pH
Operational site 0–5 36.1 5–10 27.1 10–20 23.4 20–30 22.3 30–60 29.4 60–90 32.4
8.84 6.44 4.63 3.2 3.78 6.38
5234 5666 5596 6318 6492 6020
0.33 0.22 0.16 0.11 0.06 0.03
4.60 2.97 1.90 1.29 0.67 0.32
46.2 24.9 23.1 18.6 10.0 6.5
965 911 853 706 497 302
5.98 5.55 5.43 5.44 5.31 5.41
Residue site 0–5 5–10 10–20 20–30 30–60 60–90
6.05 4.62 4.62 4.20 3.25 4.68
7589 7601 7717 8858 9576 9047
0.33 0.19 0.17 0.12 0.04 0.03
4.84 2.56 2.12 1.31 0.46 0.24
47.1 29.9 28.4 23.3 13.2 4.8
948 881 868 678 514 362
5.73 5.31 5.05 5.02 5.28 5.53
Soil depth (cm)
CEC (cmol kg1)
32.5 27.5 24.6 22.8 24.4 28.7
bulldozer with shear blade. The windrows were approximately 15 m apart with most of the post-harvest residues and some mineral soil being incorporated. At the other site (designated ‘residue’), the postharvest residues were left to provide a reasonably uniform covering of residues formed mainly of branches, needles and bark, to a depth of approximately 5–10 cm. The larger pieces, mainly stem, were removed by hand to provide a safe working environment. Both sites were approximately 1 ha in size and separated into three blocks to allow for spatial variation across the site. The soils within the plantation area have been described by Webb and Tracey (1967) and are classed as Ferrosols (Isbell, 1996) or Oxisols (Soil Survey Staff, 1999); basic soil properties of samples taken from both sites at the onset of sampling, are presented in Table 1. 2.3. In situ N mineralisation In situ N mineralisation tubes were installed in the upper, middle and lower slope areas of each block, with ca. 4–5 m between each area, to ensure representative coverage of the block. In the operational area, the in situ N mineralisation tubes were installed in the inter-windrow areas with the upper and lower sets of tubes ca. 0.5 m from the windrows. The technique used was that described by Raison et al. (1987) for the sequential, in situ exposure of soils for the
purpose of studying fluxes of soil mineral N. The tubes used were constructed from PVC with a wall thickness of approximately 3 mm, a length of 25 cm and an internal diameter of 10 cm. The lower 20 cm were perforated with 10, 1 cm diameter holes to allow for a reasonable exchange of moisture and gases with the soil outside the tube (Bubb et al., 1998). At each sampling, the top of the soil was brushed free of any loose organic matter and three tubes were driven into the ground to a depth of 20 cm. One tube was immediately removed to provide baseline data for that sampling cycle. One tube had a PVC cap that covered and over-hung the open end of the tube to prevent N losses through leaching. Small wooden blocks raised the cap above the rim of the tube to allow the free passage of air. The remaining tube was not capped, allowing N leaching to occur. Both these tubes were left in situ for 28 days when they were removed and a further three tubes inserted to repeat the cycle. Sampling took place over 26 consecutive cycles, spanning 2 calendar years, commencing in April 1999. However, results from the first month are not included due to procedural errors. Weed growth was controlled by spraying with glyphosate at both sites, which is part of the normal operational maintenance for sites of this age. There was no growth of weeds inside the tubes during the sampling period. The residue site was not replanted during the course of the experiment. Whilst the operational site was replanted 8 months following
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commencement, the 5 m spacing and the age of the seedlings allowed a large enough space to be left between trees and tubes for zero N uptake by the seedlings to be a reasonable assumption. The tubes were stored in a cold room at 4 8C until analysis, usually within 2 or 3 days following sampling. The soil cores were extracted wholly from the tubes and divided into samples representing the 0–10 and 10–20 cm depths. Each sample was thoroughly mixed with stones being removed by hand, but the high clay content of the soil and number of samples to be prepared excluded the possibility of passing the soils through a sieve. Mineral N was extracted from the samples by shaking with 2 M KCl at a soil solution ratio of 1:10, followed by centrifuging at 2000 rpm for 20 min and finally filtering through Whatman 42 filter papers.
volatilisation had exceeded the rate of mineralisation for that period; for nitrate N and nitrite N, negative values were assumed to show losses through immobilisation or denitrification. Differences between the capped and uncapped cores were assumed to demonstrate N losses through leaching. Data for leaching were calculated using both soil depths to represent leaching events within the entire core. Data were analysed using the SPSS Base 10 System (SPSS, 1999). Only the data for soil moisture had a normal distribution, therefore non-parametric statistical analyses were performed on those parts of the data where normal distribution could not be assumed. Comparison of means used the Mann–Whitney U test and correlations were established using Spearman’s rho (rank correlation).
2.4. Chemical analysis
3. Results
The extractant was analysed colorimetrically for ammonium N, nitrate N and nitrite N using a Lachat flow injection analyser. Results were verified using external standards after every 11 samples. A minimum of 10% duplication of samples was employed to verify the accuracy and repeatability of the results. A subsample of each soil was dried at 105 8C for 24 h to determine soil moisture content. Soil properties were determined as reported by Xu et al. (1995).
3.1. Rainfall, temperature and soil moisture
2.5. Analysis of results Operational limitations necessitated the establishment of two residue treatment areas, each to be sampled using a stratified random sampling design (Dick et al., 1996). However, at each site the management history, slope, aspect and environmental conditions were the same. Comparisons are drawn between the effects of the different management techniques employed at the two sites using appropriate statistical analysis to obtain levels of significance (Oksanen, 2001). The difference in soil mineral N between the capped core and the baseline core was used to measure the N dynamics for the sampling period. The dynamics for a given N form are cumulative and, for ammonium N, where this value was negative, it was assumed that the rate of loss through immobilisation, nitrification or
The mean monthly rainfall and mean monthly maximum air temperature for the sampling period are given in Fig. 1a and b, respectively. Temperature is the mean for each day of the sampling period. The first 12 sampling cycles had an unusually wet winter (sampling cycles 3–9), whereas the second 12 sampling cycles were more typical. The air temperature shows the seasonal variation typical of the region. The residue site (Fig. 2a) had approximately 60% soil moisture at the 0–10 cm depth at the onset of the experiment. This level declined steadily to a little over 20% in the 21st sampling cycle when seasonal rains began. Compared with the 0–10 cm depth, the 10– 20 cm depth had significantly lower soil moisture content (P < 0:001) though the values converged toward the end of the sampling period. The operational site showed similar trends (Fig. 2b) though the starting soil moisture content of approximately 40% at the 0– 10 cm depth was lower than the residue site. In the first 12 sampling cycles, soil moisture content at the 0– 10 cm depth was significantly higher at the residue site than at the operational site (P < 0:05), but there was no significant difference between sites for the remainder of the sampling period (P > 0:05). There was no significant difference in soil moisture content between sites at the 10–20 cm depth.
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Fig. 1. Monthly rainfall (a) and mean monthly maximum air temperature (b) for the sampling period.
3.2. Ammonium N 3.2.1. Residue site Ammonium N was the principal form of mineral N immediately following harvest though levels at the 0– 10 cm depth declined from around 60 kg N ha1 at the onset to <10 kg N ha1 in the first 14 sampling cycles (Fig. 3a). The levels of ammonium N remained at <10 kg N ha1 for the balance of the sampling period. Ammonium N was significantly correlated with soil moisture at the 0–10 cm depth throughout the sampling period (P < 0:001, R2 ¼ 0:20, data not shown). Ammonium N was significantly less at the 10–20 cm than at the 0–10 cm depth during the first 12 sampling cycles (P < 0:001), but remained at similar levels for the remainder of the sampling period. At the 0–10 cm depth, Ammonium N dynamics (Fig. 3b) showed a strong negative response of approximately
100 kg N ha1 during the sampling period. At the 10–20 cm depth there was a significantly lesser negative response at 40 kg N ha1 (P < 0:001). 3.2.2. Operational site Ammonium N was the principal form of mineral N following harvest, but levels at the 0–10 cm depth declined from about 55 to <10 kg N ha1 in the 12 sampling cycles following harvest (Fig. 3c), and then remained at <10 kg N ha1 for the remainder of the period. Ammonium N was significantly less at the 10– 20 cm than at the 0–10 cm depth during the first 12 sampling cycles (P < 0:001), but then remained at similar levels. Ammonium N dynamics showed a negative response of around 30 kg N ha1 for the 0–10 cm depth and 10 to 20 kg N ha1 for the 10–20 cm depth (Fig. 3d). The difference in ammonium N dynamics between depths was not significant (P > 0:05).
Fig. 2. Moisture content at 0–10 and 10–20 cm soil depth for: (a) the residue site (mean standard error (M.S.E.) 0.79 and 0.44, respectively, n ¼ 447); (b) the operational site (M.S.E. 0.71 and 0.48, respectively, n ¼ 448).
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Fig. 3. Ammonium N fluxes with M.S.E. at both depths over the sampling period for: (1) the residue site showing: (a) ammonium N in the baseline incubation tubes with M.S.E. 1.18 (0–10 cm) and 0.66 (10–20 cm) (n ¼ 442); (b) cumulative ammonium N dynamics with M.S.E. 3.26 (0–10 cm) and 2.56 (10–20 cm) (n ¼ 444). (2) The operational site showing (c) ammonium N in the baseline incubation tubes with M.S.E. 1.72 (0–10 cm) and 0.88 (10–20 cm) (n ¼ 436); (d) cumulative ammonium N dynamics with M.S.E. 4.44 (0–10 cm) and 4.05 (10– 20 cm) (n ¼ 446).
3.2.3. Management Ammonium N dynamics were significantly lower at the residue site than at the operational site for both the 0–10 cm (P < 0:001) and 10–20 cm (P < 0:01) depth. 3.3. Nitrate N 3.3.1. Residue site Nitrate N at the 0–10 cm depth declined from around 45 kg N ha1 during the first three sampling cycles following harvest (Fig. 4a), to around
10 kg N ha1 until the 10th sampling cycle and then rose to approximately 20 kg N ha1 towards the end of the sampling period. Nitrate N was negatively correlated with rainfall at the 0–10 cm depth (P < 0:01, data not shown) and positively correlated with temperature at both the 0–10 cm (P < 0:01, data not shown) and 10–20 cm (P < 0:01, data not shown) depths. The difference in nitrate N between depths was significant (P < 0:001). Nitrate N dynamics at the 0–10 cm depth was positive (Fig. 4b), rising steadily to approximately 200 kg N ha1 at the end of the sampling period; at
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Fig. 4. Nitrate N fluxes with M.S.E. for both depths over the sampling period for: (1) the residue site showing (a) nitrate N in the baseline incubation tubes with M.S.E. 1.32 (0–10 cm) and 0.48 (10–20 cm) (n ¼ 436); (b) cumulative nitrate N dynamics with M.S.E. 8.27 (0– 10 cm) and 1.76 (10–20 cm) (n ¼ 446); (c) nitrate N losses through leaching with M.S.E. 9.03 (0-20 cm, n ¼ 225). (2) The operational site showing (d) nitrate N in the baseline incubation tubes with M.S.E. 1.05 (0–10 cm) and 0.60 (10–20 cm) (n ¼ 441); (e) cumulative nitrate N dynamics with M.S.E. 7.77 (0–10 cm) and 1.70 (10–20 cm) (n ¼ 444); (f) nitrate N losses through leaching with M:S:E: ¼ 7:38 (0–20 cm, n ¼ 225).
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the 10–20 cm depth there was no significant change in nitrate N dynamics. At the beginning of the sampling period about 50 kg nitrate N ha1 was lost through leaching, but there was no further significant leaching of nitrate N from the 0 to 20 cm depth until the onset of the wet season in the 21st sampling cycle when losses of nitrate N through leaching rose steadily (Fig. 4c).
but at the 10–20 cm depth the operational site had higher nitrate N overall (P < 0:05). Nitrate N dynamics at the 0–10 cm depth were significantly higher at the operational site (P < 0:001), but there was no significant difference between the sites at the 10–20 cm depth. Nitrate N losses through leaching were significantly higher at the operational site (P < 0:001). 3.4. Nitrite N
3.3.2. Operational site Nitrate N at the 0–10 cm depth declined from around 35 kg N ha1 during the first two sampling cycles following harvest (Fig. 4d) to <10 kg N ha1 at the 10th sampling cycle. Nitrate N then peaked at around 35 kg N ha1 in the wet warmer summer period (sampling cycles 10–15) and then remained at around 15– 20 kg N ha1. Nitrate N was not correlated with rainfall at either depth, but was significantly correlated with temperature at both depths (0–10 cm, P < 0:001; 10–20 cm, P < 0:001; data not shown). Nitrate dynamics were positive throughout the sampling period at the 0–10 cm depth with a cumulative value of 300 kg N ha1 at the end of the sampling period (Fig. 4e). Nitrate N losses through leaching were steady with a total of around 225 kg N ha1 lost by the end of the sampling period (Fig. 4f). 3.3.3. Management There was no significant difference in baseline nitrate N between the sites at the 0–10 cm depth (P > 0:05),
3.4.1. Residue site Nitrite N was detected during the warm, wet summer at the 0–10 cm depth and at the 10–20 cm depth between the 10th and 16th sampling cycles (Fig. 5a) though the levels, 2 and 4 kg N ha1, respectively, were very low. Nitrite N was significantly correlated with air temperature at both the 0–10 and 10–20 cm depths (P < 0:01, data not shown). Nitrite N dynamics and leaching were very low (data not shown). 3.4.2. Operational site Nitrite N was detected at low levels, 4 kg N ha1 at the 0–10 cm depth and 2 kg N ha1 at the 10–20 cm depth between the 10th and 16th sampling cycles (Fig. 5b). Nitrite N was significantly correlated with air temperature and soil moisture content at both the 0–10 and 10–20 cm depths (P < 0:01, data not shown). Nitrite N dynamics were at very low levels and the overall losses through leaching of nitrite N were very low (data not shown).
Fig. 5. Nitrite N at the 0–10 cm and the 10–20 cm depth for (a) residue site (M.S.E. 0.65 and 0.09, respectively, n ¼ 436); (b) operational site (M.S.E. 0.73 and 0.06, respectively, n ¼ 441).
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4. Discussion Friedel et al. (2000) questioned the validity of the in situ mineralisation technique, due to the abscission of fine roots by the core, suggesting that ionexchange resins mixed into the soil and then incubated in the laboratory would better simulate the N immobilisation by fine roots. However, such an approach denies the basic principles underlying in situ mineralisation: that the soils are undisturbed and that mineralisation occurs under actual field conditions. Knoepp and Swank (1995) compared the in situ N mineralisation technique with in situ buried bags, tension lysimeters and aerobic laboratory incubations as a means of estimating N mineralisation, highlighting that the in situ core method was superior as the procedure took into account the changes in sitespecific parameters such as soil moisture content and temperature. 4.1. Ammonium N Harvesting increases N mineralisation (Attiwill and Adams, 1993; Carlyle, 1994), with soil disturbance and increases in soil temperature following canopy removal stimulating microbially mediated N mineralisation. The resulting increase in plant available mineral N is therefore introduced into the system at a time when there are no trees to absorb the available nutrients (Johnston and Crossley, 2002). Our results show that, regardless of residue management, over 100 kg N ha1 were available in the first few sampling cycles following the harvest of a 1R hoop pine plantation; such levels of N loss could lead to a substantial decline in potential site productivity. Ammonium N is the initial product of N mineralisation and has the potential to be lost through volatilisation and leaching, to be immobilised by both biotic assimilation and abiotic fixation or transformed through nitrification. Ammonia volatilisation is normally associated with burning, a practice that has ceased in hoop pine plantations (Ryan and Gilmour, 1985); ammonium N leaching has been shown not to be a significant factor in soils where there is no N saturation (Harmsen and Kolenbrander, 1965). Comparison of baseline and capped core data in this study, suggests that either N immobilisation or nitrification was the principal process affecting ammonium N
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during the 12 sampling cycles when levels declined from around 60 to 10 kg N ha1. N fixation to the soil will account for some of the apparent immobilisation of ammonium N and can occur quite rapidly (Trehan, 1996; Berntson and Aber, 2000). However, it has been demonstrated that up to 60% of the ammonium N that is bound to the soil surface may also become available for biological assimilation (Schulten and Schnitzer, 1998). Microbial immobilisation of ammonium N is well documented (Tietema et al., 1998; Puri and Ashman, 1999; Mao et al., 2002), particularly when there is an adequate supply of available C. The residue site showed significantly higher soil moisture during the sampling period and it may be assumed that the conditions for microbial activity were enhanced through both physical protection from soil drying and a readily available C supply provided by the residue layer (Burket and Dick, 1997). Pu et al. (2001) reported that N immobilisation in the hoop pine plantations was significantly higher underneath residues than in areas where the residues had been removed. As both nitrification and N immobilisation are microbially dependent, both processes would be competing for the available ammonium N. 4.2. Nitrate N Unlike northern coniferous forest soils, which are regarded as having a low nitrification potential (Davidson et al., 1992), the onset of nitrification was immediate and durable in the hoop pine soils. Nitrate N formed approximately 40% of total mineral N at both sites in the first three sampling cycles following harvest, indicating that substantial nitrification was taking place. Nitrate N may be readily lost through leaching if it is not immobilised by the soil microbial biomass. It has been claimed that little immobilisation of nitrate N will occur in the presence of ammonium N (Attiwill and Adams, 1993) because, unlike nitrate N, ammonium N does not require reduction and is therefore more easily assimilated by the microbial biomass (Puri and Ashman, 1999). However, where there is a sufficient supply of available C, nitrate immobilisation may take place, even when rates of nitrification are high (Stark and Hart, 1997). Nitrate N dynamics showed a high level of nitrification (300 kg N ha1) at the 0–10 cm depth at the
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operational site. Ammonium N dynamics were close to zero, indicating that the rate of ammonium N production through mineralisation and the rate of transformation to nitrate N were equal throughout the period. With overall levels of ammonium N declining, this suggests that ammonium N was the factor limiting nitrate production. In a study of soils amended with straw, Recous et al. (1998) also concluded that the supply of ammonium ions was probably the factor limiting nitrification. The nitrate N dynamics at the residue site show a lower level of nitrification (200 kg N ha1) than the operational site, whilst the ammonium N dynamics were negative (100 kg N ha1) for the same period. The presence of organic matter can stimulate a rapid expansion in the soil microbial population (Dalias et al., 2002), and rates of both N mineralisation and immobilisation have been shown to be higher in soils amended with straw (Recous et al., 1998). The negative ammonium N dynamics and lower rates of nitrification at the residue site indicate that immobilisation was competing for ammonium N released through the mineralisation of organic matter, there may also have been some immobilisation of nitrate N which has been shown to take place when levels of soil C are high (Stark and Hart, 1997). Pu et al. (2001) also found immobilisation of N to be higher in areas under post-harvest residues compared to those areas where the residues had been removed. Whilst denitrification is another potential mechanism for loss of nitrate in forest soils, it mainly occurs in saturated soils under anaerobic conditions and the potential is increased by N saturation (Mohn et al., 2000). The soils in the study area are free draining and saturation is difficult to achieve (Pu et al., 2001). Though denitrification has been recorded in northern hemisphere soils at soil moisture contents <50% (Stevens et al., 1997), other studies have shown that even at 80% water-filled pore space, N2O emissions can be as low as 8 kg N ha1 over a 10-month period (Smith et al., 1998). Vor and Brumme (2002) found that in situ N mineralisation increases denitrification. However, the compaction and anaerobic conditions associated with their 40 cm long, 7.6 cm wide steel tubes would have been far less likely to occur with the 20 cm long, 10 cm wide, aerated tubes that were used in this study. While the potential for N losses through denitrification in these soils can be high (Pu
et al., 2001) and some minor denitrification may occur as soil aeration status changes with depth (Whalley et al., 1995), the conditions promoting significant denitrification occur only intermittently (Pu et al., 2001). Immobilisation of both ammonium N and nitrate N at the residue site would prevent some N loss through leaching. The N leaching losses were low at the residue site until the 21st sampling cycle at the onset of the warm, wet summer. It was observed that by this time, most of the foliage (which formed the finer residue) had decomposed, leaving the coarser and more recalcitrant fractions in this area. The exhaustion of the readily available C within the finer foliage would stimulate the release of N from within the microbial biomass with subsequent nitrification and leaching. Parfitt et al. (2001) found that N from decomposing litter was initially stored within the litter before subsequent release into the soil layer. Over 100 kg N ha1 were lost through leaching from the residue site in the final six sampling cycles. N losses through leaching at the operational site demonstrated a seasonal variation with rainfall, which suggests that little or no N immobilisation had occurred. 4.3. Nitrite N Nitrite N is not normally considered to be a significant part of the hoop pine forest mineral N pool (Fisher and Binkley, 2000). Data presented here demonstrate that under certain environmental conditions it can be present at levels up to 4 kg N ha1, which is below that considered toxic to plants and bacteria (Paul and Clark, 1996). There were no significant differences in baseline nitrite N between the sites. There was a weak, but significant correlation of nitrite N with air temperature and soil moisture between 13 and 26 sampling cycles. It has been previously reported that denitrification may occur in these soils when wetted (Pu et al., 2001) and as nitrite N is an intermediate stage in denitrification (Fisher and Binkley, 2000), it is reasonable to assume that some denitrification was occurring in the wet summer months. The intermittent nature and low levels of nitrite N agree with the previous conclusion that denitrification would occur only sporadically in these soils.
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4.4. Site management Residue retention favours microbial activity by acting as a physical protection, maintaining levels of soil moisture and providing a readily available supply of organic C. Barrett and Burke (2000) found higher rates of N immobilisation in areas under plants, which had greater levels of available C and higher rates of microbial activity, than in areas of open soil between the plants. Hart et al. (1994) also found that C availability exerted a strong influence on N immobilisation. Following their work on 1R hoop pine plantations, Bubb et al. (1998) suggested that there would be the potential for substantial losses of N due to leaching following clearfall. These studies have indicated that retaining residues as an even ground cover can prevent N leaching losses through N immobilisation in the first 18 months following harvesting. It is evident that hoop pine seedlings in 2R soils are N limited (Xu et al., 2002), but the extent to which young trees would be capable of gaining access to the available N within the soil would be governed by the development of the tree’s root system. It would therefore seem probable that, for the period covered by this study, the trees would not be disadvantaged by the immobilisation of available N. As suggested by Kushwaha et al. (2000), the presence of high C:N ratio residues may act as a slow release fertiliser through the immobilisation and eventual re-mineralisation of the mineral N. Similarly, Carlyle (1994) suggested that residue retention would have the capacity to allow for a steady increase in N availability without incurring the risk of leaching. Higher N immobilisation under residue retention may prevent N losses that would otherwise occur through volatilisation, leaching and denitrification.
5. Conclusion Soil mineral N flux is the result of a complicated and dynamic series of inter-related mechanisms that are highly responsive to environmental conditions, including human disturbance. At any moment in time, most of the processes described above may be operating; with changes in soil mineral N dictated by whichever process is dominant. The in situ core technique for studying soil N mineralisation was a
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useful tool for unravelling some of the complexities of mineral N dynamics following the severe disturbance of clearfall harvesting. Harvesting initiated a flush of N mineralisation at both sites that persisted beyond the 2-year sampling period. Whereas established 1R hoop pine plantations are known to have low rates of nitrification, harvesting disturbance increased the rate of nitrification and brought about a greater potential for N loss through both leaching and denitrification during the critical inter-rotation and early establishment phase of 2R hoop pine plantations. Retaining post-harvest residues on-site was initiated to try and mitigate some of the N losses that had been occurring under the previous residue-burning regime. Whilst windrowing of post-harvest residues retained the residues on-site, serious concerns regarding the potential for N losses through leaching and denitrification in the cleared inter-windrow spaces remained. Retaining the residues as mulch, whilst posing operational problems, has the potential for retaining mineral N on-site and available for the trees of the next rotation. Incidental effects of weed suppression and moisture retention would also benefit the young trees. Whilst the ultimate management decisions regarding residue retention would probably be based on economic factors, an operational trial coupled with a costbenefit analysis of alternative residue-management systems is recommended.
Acknowledgements T.J. Blumfield was supported in this work through a scholarship grant from the Cooperative Research Centre for Sustainable Production Forestry. We would like to acknowledge the invaluable help and support of Mr Paul Keay and thank Griffith University and the Queensland Forestry Research Institute for their technical help and support. We are grateful to Dr Chris Beadle for his constructive comments on earlier versions of the manuscript. References Attiwill, P.A., Adams, M.A., 1993. Tansley Review No. 50. Nutrient cycling in forests. New Phytol. 124, 561–582. Ballard, T.M., 2000. Impacts of forest management on northern forest soils. For. Ecol. Manage. 133, 37–42.
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Barrett, J.E., Burke, I.C., 2000. Potential nitrogen immobilization in grassland soils across a soil organic matter gradient. Soil Biol. Biochem. 32, 1707–1716. Berntson, G.M., Aber, J.D., 2000. Fast nitrate immobilization in N saturated temperate forest soils. Soil Biol. Biochem. 32, 151– 156. Bubb, K.A., Xu, Z.H., Simpson, J.A., Saffigna, P.G., 1998. In situ measurements of soil mineral–nitrogen fluxes in hoop pine plantations of subtropical Australia. NZ J. For. Sci. 28, 152– 164. Burket, J.Z., Dick, R.P., 1997. Long-term vegetation management in relation to accumulation and mineralisation of nitrogen in soils. In: Cadisch, G., Giller, K.E. (Eds.), Driven by Nature: Plant Litter Quality and Decomposition. CAB International, Wallingford. Carlyle, J.C., 1986. Nitrogen cycling in forested ecosystems. For. Abst. 47, 307–336. Carlyle, J.C., 1994. Opportunities for managing nitrogen uptake in established Pinus radiata plantations on sandy soils. NZ J. For. Sci. 24, 344–361. Costantini, A., Grimmet, J.L., Dunn, G.M., 1997. Towards sustainable management of forest plantations in south-east Queensland. 1. Logging and understorey residue management between rotations in steep country Araucaria cunninghamii plantations. Aust. For. 60, 213–225. Dalias, P., Anderson, J.M., Bottner, P., Couteaux, M.-M., 2002. Temperature responses of net nitrogen mineralization and nitrification in conifer forest soils incubated under standard laboratory conditions. Soil Biol. Biochem. 34, 691–701. Davidson, E.A., Hart, S.C., Firestone, M.K., 1992. Internal cycling of nitrate in soils of a mature coniferous forest. Ecology 73, 1148–1156. Dick, R.P., Thomas, D.R., Halvorson, J.J. (Eds.), 1996. Standardized Methods, Sampling and Sample Pretreatment. In: Methods for Assessing Soil Quality. Soil Science Society of America, Madison, pp. 107–121. FAO, 2001. State of the world’s forests 2001. Technical Report. Food and Agriculture Organisation of the United Nations, Rome. Fisher, R.F., Binkley, D., 2000. Ecology and Management of Forest Soils. Wiley, New York. Friedel, J.K., Herrmann, A., Kleber, M., 2000. Ion exchange resin– soil mixtures as a tool in net nitrogen mineralisation studies. Soil Biol. Biochem. 32, 1529–1536. Harmsen, G.W., Kolenbrander, G.J., 1965. Soil inorganic nitrogen. In: Bartholomew, W.V., Clark, F.E. (Eds.), Soil Nitrogen. American Society of Agronomy, Inc., Madison. Hart, S.C., Nason, G.E., Myrold, D.D., Perry, D.A., 1994. Dynamics of gross nitrogen transformations in an old-growth forest: the carbon connection. Ecology 75, 880–891. Isbell, R.F., 1996. The Australian Soil Classification. CSIRO Australia, Collingwood, Vic. Johnston, J.M., Crossley Jr., J.D.A., 2002. Forest ecosystem recovery in the southeast US: soil ecology as an essential component of ecosystem management. For. Ecol. Manage 155, 187–203. Jones, H.E., Madeira, M., Herraez, L., Dighton, J., Fabiao, A., Gonzalez-Rio, F., Fernandez Marcos, M., Gomez, C., Tome,
M., Feith, H., 1999. The effect of organic-matter management on the productivity of Eucalyptus globulus stands in Spain and Portugal: tree growth and harvest residue decomposition in relation to site and treatment. For. Ecol. Manage. 122, 73–86. Keeney, D.R., 1980. Prediction of soil nitrogen availability in forest ecosystems: a literature review. For. Sci. 26, 159–171. Knoepp, J.D., Swank, W.T., 1995. Comparison of available soil nitrogen assays in control and burned forested sites. Soil Sci. Soc. Am. J. 59, 1750–1754. Kushwaha, C.P., Tripathi, S.K., Singh, K.P., 2000. Variations in soil microbial biomass and N availability due to residue and tillage management in a dryland rice agroecosystem. Soil Till. Res. 56, 153–166. Mao, X.A., Xu, Z.H., Luo, R.S., Mathers, N.J., Zhang, Y.H., Saffigna, P.G., 2002. Nitrate in soil humic acids revealed by 14 N nuclear magnetic resonance spectroscopy. Aust. J. Soil Res. 40, 717–726. Mathers, N.J., Xu, Z.H., Berners-Price, S.J., Perera, M.C.S., Saffigna, P.G., 2002. Hydrofluoric acid pre-treatment for improving 13 C CPMAS NMR spectral quality of forest soils in south-east Queensland, Australia. Aust. J. Soil Res. 40, 655– 674. Mohn, J., Schurmann, A., Hagedorn, F., Schleppi, P., Bachofen, R., 2000. Increased rates of denitrification in nitrogen-treated forest soils. For. Ecol. Manage. 137, 113–119. Nambiar, E.K.S., 1996. Sustained productivity of forests is a continuing challenge to soil science. Soil Sci. Soc. Am. J. 60, 1629–1642. Oksanen, L., 2001. Logic in experiments in ecology: is pseudoreplication a pseudoissue. Oikos 94, 27–38. Parfitt, R.L., Salt, G.J., Saggar, S., 2001. Post-harvest residue decomposition and nitrogen dynamics in Pinus radiata plantations of different N status. For. Ecol. Manage. 154, 55– 67. Paul, E.A., Clark, F.E., 1996. Soil Microbiology and Biochemistry. Academic Press, San Diego. Perakis, S.S., Hedlin, L.O., 2002. Nitrogen loss from unpolluted South American forests mainly via dissolved organic compounds. Nature 415, 416–419. Pu, G.X., Saffigna, P.G., Xu, Z.H., 2001. Denitrification, leaching and immobilisation of 15 N-labelled nitrate in winter under windrowed harvesting residues in hoop pine plantations of 1–3 years old in subtropical Australia. For. Ecol. Manage. 152, 183– 194. Puri, G., Ashman, M.R., 1999. Microbial immobilization of 15 Nlabelled ammonium and nitrate in a temperate woodland soil. Soil Biol. Biochem. 31, 929–931. Raison, R.J., Connell, M.J., Khanna, P.K., 1987. Methodology for studying fluxes of soil mineral-N in situ. Soil Biol. Biochem. 19, 521–530. Recous, S., Aita, C., Mary, B., 1998. In situ changes in gross N transformations in bare soil after addition of straw. Soil Biol. Biochem. 31, 119–133. Ryan, P.A., Gilmour, D., 1985. Potential erosion and nutrient losses in the conversion to the second rotation hoop pine crop. Queensland Forestry Research Institute Technical Report. Queensland Forestry Research Institute, Gympie, Qld, 28 pp.
T.J. Blumfield, Z.H. Xu / Forest Ecology and Management 179 (2003) 55–67 Schulten, H.R., Schnitzer, M., 1998. The chemistry of soil organic nitrogen: a review. Biol. Fert. Soils. 26, 1–15. Smith, K.A., Thomson, P.E., Clayton, H., McTaggart, I.P., Conen, F., 1998. Effects of temperature, water content and nitrogen fertilisation on emissions of nitrous oxide by soils. Atmos. Environ. 32, 3301–3309. Soil Survey Staff, 1999. Soil Taxonomy: A Basic System of Soil Classification for Making and Interpreting Soil Surveys. US Department of Agriculture Soil Conservation Service, Washington. SPSS, 1999. SPSS Base 10 Application Guide. SPSS, Inc., Chicago. Stark, J.M., Hart, S.C., 1997. High rates of nitrification and nitrate turnover in undisturbed coniferous forests. Nature 385, 61–64. Stevens, R.J., Laughlin, R.J., Burns, L.C., Arah, J.R.M., Hood, R.C., 1997. Measuring the contributions of nitrification and denitrification to the flux of nitrous oxide from soil. Soil Biol. Biochem. 29, 139–151. Tietema, A., Emmett, B.A., Gundersen, P., Kjonaas, O.J., Koopmans, C.J., 1998. The fate of 15 N-labelled nitrogen deposition
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in coniferous forest ecosystems. For. Ecol. Manage. 101, 19– 27. Trehan, S.P., 1996. Immobilisation of 15 NH4 in three soils by chemical and biological processes. Soil Biol. Biochem. 28, 1021–1027. van Breemen, N., 2002. Natural organic tendency. Nature 415, 381–382. Vor, T., Brumme, R., 2002. N2O losses result in underestimation of in situ determinations of net N mineralization. Soil Biol. Biochem. 34, 541–544. Webb, L.J., Tracey, J.G., 1967. An ecological guide to new planting areas and site potential for hoop pine. Aust. For. 31, 224–239. Whalley, W.R., Dumitru, E., Dexter, A.R., 1995. Biological effects of soil compaction. Soil Till. Res. 35, 53–68. Xu, Z.H., Simpson, J.A., Osborne, D.O., 1995. Mineral nutrition of slash pine in subtropical Australia. I. Stand growth response to fertilization. Fert. Res. 41, 93–100. Xu, Z.H., Bubb, K.A., Simpson, J.A., 2002. Effects of nitrogen fertilisation and weed control on nutrition and growth of a 4year-old Araucaria cunninghamii plantation in subtropical Australia. J. Trop. For. Sci. 14, 213–222.
Impact of harvest residues on soil mineral nitrogen dynamics following clearfall harvesting of a hoop pine plantation in subtropical Australia T.J. Blumfielda,b,*, Z.H. Xua,b,c,1 a
Cooperative Research Centre for Sustainable Production Forestry, Griffith University, Nathan, Qld 4111, Australia b Faculty of Environmental Science, Griffith University, Nathan, Qld 4111, Australia c Queensland Forestry Research Institute, PO Box 631, Indooroopilly, Qld 4068, Australia Received 2 May 2002; accepted 4 September 2002
Abstract An alternative management strategy allowing post-harvest residues to remain as a blanket cover instead of being incorporated into windrows may prevent problems associated with lack of soil fertility in the inter-windrow space. Parallel, 2-year, in situ nitrogen (N) mineralisation studies were undertaken during the inter-rotation period following clearfall harvesting of a first-rotation hoop pine (Araucaria cunninghamii Aiton ex D. Don) plantation in subtropical Australia. We investigated the dynamics of ammonium N, nitrate N and nitrite N in the top 20 cm soil under a residue retention situation and under normal, operational conditions. Initially, ammonium N was the dominant form of soil mineral N but declined to <10 kg N ha1 at both sites after 12 sampling cycles of 28-day duration. Nitrate N levels remained at approximately 30 kg N ha1 despite seasonal fluctuations, throughout the 2-year sampling period. At the residue site there was no net N mineralisation in the soil until the 20th sampling cycle; approximately 110 kg N ha1 was mineralised at the end of the sampling period. This compared with approximately 300 kg N ha1 mineralised at the operational site for the same period. Approximately 100 and 220 kg N ha1 were lost through leaching at the residue and operational sites, respectively. Residue retention promoted N immobilisation, reducing the potential for N losses through leaching and denitrification in the critical inter-rotation and early establishment period. # 2002 Elsevier Science B.V. All rights reserved. Keywords: In situ mineralisation; Post-harvest residues; Soil mineral nitrogen dynamics; Windrows
1. Introduction *
Corresponding author. Present address: Cooperative Research Centre for Sustainable Production Forestry, Griffith University, Nathan, Qld 4111, Australia. Tel.: þ61-7-38756709; fax: þ61-7-38757459. E-mail addresses: [email protected] (T.J. Blumfield), [email protected] (Z.H. Xu). 1 Tel.: þ61-7-33629367; fax: þ61-7-38969628.
Nutrient cycling, in particular the dynamics of N within forested ecosystems, has been widely and intensively studied (Keeney, 1980; Carlyle, 1986; Attiwill and Adams, 1993; Nambiar, 1996; Ballard, 2000). However, it is becoming increasingly evident that some of the basic assumptions underlying such N
0378-1127/02/$ – see front matter # 2002 Elsevier Science B.V. All rights reserved. PII: S 0 3 7 8 - 1 1 2 7 ( 0 2 ) 0 0 4 8 5 - 1
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dynamics are based on research in northern hemisphere temperate and boreal forests, which was initiated after the onset of anthropogenic disturbance of the ‘natural’ system (Perakis and Hedlin, 2002). This may not, with certainty, be applied to southern hemisphere forests (van Breemen, 2002). Tropical and subtropical forests form 56% of the world’s total forest resources, while 45% of the world’s forests are to be found in South America, Africa and Oceania, within the southern hemisphere (FAO, 2001). Currently, there is a paucity of studies of N mineralisation in subtropical and tropical forests (Bubb et al., 1998). Research into the nutrient dynamics of these forest systems, particularly intensively managed plantation systems, therefore becomes essential if we are to formulate suitable strategies for sustainable and environmentally focused forest management. A major operational change within the hoop pine plantation estate in subtropical Australia, has been the decision to retain post-harvest residues on-site by forming the residues into windrows. This decision was a response to concerns regarding N losses following burning of post-harvest residues (Costantini et al., 1997), these N losses occurring both through volatilisation and through erosion (Ryan and Gilmour, 1985). However, operational decisions regarding the windrowing of retained residues were driven by the nature of the available machinery, costs and workplace health and safety. The effect that changes to the residue-management regime would have on N dynamics was not known. Soil mineral N fluxes in first-rotation (1R) hoop pine plantations of various ages (Bubb et al., 1998) indicated that the N mineralisation rate was the highest in the growing season and that ammonium was the dominant mineral N form. The net soil N mineralisation was 25.3, 30.9 and 45.8 kg N ha1 over a 7-month period for 3-, 20- and 62-year-old plantations, respectively. However, the dynamics of soil mineral N fluxes in the inter-rotation and early second rotation (2R) phase have not been studied. Some attention has been given to soil organic matter dynamics and N pools (Mao et al., 2002; Mathers et al., 2002) as well as nutrient loss mechanisms under a windrow residue-management regime (Pu et al., 2001), but the effect of these practices on soil mineral N fluxes in the tree planting area between the windrows has not been explored.
The purpose of this study was to gain an understanding of soil mineral N dynamics in the critical inter-rotation and early second rotation periods using the in situ core mineralisation technique; particularly in relation to residue management, potential N losses through denitrification and leaching (Pu et al., 2001; Johnston and Crossley, 2002), and N retention through immobilisation (Tietema et al., 1998; Puri and Ashman, 1999). Successful residue management can benefit the plantation through increased growth (Jones et al., 1999), reducing the need for fertiliser inputs, and benefiting the environment by reducing off-site effects.
2. Materials and methods 2.1. Site description Two sites were established immediately following clearfall harvesting of a 1R hoop pine plantation in the Imbil State Forest (268310 S, 1528380 E) in Queensland. Each had a slight to moderate slope (10–158) with an easterly aspect. The forest lies within the subtropical zone, with a predominating weather pattern of cool, dry winters and hot, wet summers. This pattern is subject to variation in response to an overlap between tropical and temperate weather systems (Costantini et al., 1997) and to the effects of weather patterns induced by the southern oscillation. There is a wide variation in annual rainfall (495–1964 mm, mean 1188 mm). The average daily temperature ranges from 19.4 to 30.5 8C (mean 25.0 8C) during October to March; from 6.9 to 21.5 8C with a mean of 14.2 8C during April to September (Bubb et al., 1998). Rainfall and temperature were measured daily at the Imbil Forestry Office, which is situated within 5 km of the sites and considered representative of the area. The data for these variables have been presented in 28day increments to correspond with the sampling periods. 2.2. Site preparation One site (designated ‘operational’) was in an area that had been established using normal operational practice; windrows of post-harvest residues were formed along the contours of the slope using a D6
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Table 1 Chemical properties of soil profiles for the two study areas Electrical conductivity (mS m1)
Total K (mg kg1)
Total N (%)
Organic C (%)
Available P (mg kg1)
Total P (mg kg1)
pH
Operational site 0–5 36.1 5–10 27.1 10–20 23.4 20–30 22.3 30–60 29.4 60–90 32.4
8.84 6.44 4.63 3.2 3.78 6.38
5234 5666 5596 6318 6492 6020
0.33 0.22 0.16 0.11 0.06 0.03
4.60 2.97 1.90 1.29 0.67 0.32
46.2 24.9 23.1 18.6 10.0 6.5
965 911 853 706 497 302
5.98 5.55 5.43 5.44 5.31 5.41
Residue site 0–5 5–10 10–20 20–30 30–60 60–90
6.05 4.62 4.62 4.20 3.25 4.68
7589 7601 7717 8858 9576 9047
0.33 0.19 0.17 0.12 0.04 0.03
4.84 2.56 2.12 1.31 0.46 0.24
47.1 29.9 28.4 23.3 13.2 4.8
948 881 868 678 514 362
5.73 5.31 5.05 5.02 5.28 5.53
Soil depth (cm)
CEC (cmol kg1)
32.5 27.5 24.6 22.8 24.4 28.7
bulldozer with shear blade. The windrows were approximately 15 m apart with most of the post-harvest residues and some mineral soil being incorporated. At the other site (designated ‘residue’), the postharvest residues were left to provide a reasonably uniform covering of residues formed mainly of branches, needles and bark, to a depth of approximately 5–10 cm. The larger pieces, mainly stem, were removed by hand to provide a safe working environment. Both sites were approximately 1 ha in size and separated into three blocks to allow for spatial variation across the site. The soils within the plantation area have been described by Webb and Tracey (1967) and are classed as Ferrosols (Isbell, 1996) or Oxisols (Soil Survey Staff, 1999); basic soil properties of samples taken from both sites at the onset of sampling, are presented in Table 1. 2.3. In situ N mineralisation In situ N mineralisation tubes were installed in the upper, middle and lower slope areas of each block, with ca. 4–5 m between each area, to ensure representative coverage of the block. In the operational area, the in situ N mineralisation tubes were installed in the inter-windrow areas with the upper and lower sets of tubes ca. 0.5 m from the windrows. The technique used was that described by Raison et al. (1987) for the sequential, in situ exposure of soils for the
purpose of studying fluxes of soil mineral N. The tubes used were constructed from PVC with a wall thickness of approximately 3 mm, a length of 25 cm and an internal diameter of 10 cm. The lower 20 cm were perforated with 10, 1 cm diameter holes to allow for a reasonable exchange of moisture and gases with the soil outside the tube (Bubb et al., 1998). At each sampling, the top of the soil was brushed free of any loose organic matter and three tubes were driven into the ground to a depth of 20 cm. One tube was immediately removed to provide baseline data for that sampling cycle. One tube had a PVC cap that covered and over-hung the open end of the tube to prevent N losses through leaching. Small wooden blocks raised the cap above the rim of the tube to allow the free passage of air. The remaining tube was not capped, allowing N leaching to occur. Both these tubes were left in situ for 28 days when they were removed and a further three tubes inserted to repeat the cycle. Sampling took place over 26 consecutive cycles, spanning 2 calendar years, commencing in April 1999. However, results from the first month are not included due to procedural errors. Weed growth was controlled by spraying with glyphosate at both sites, which is part of the normal operational maintenance for sites of this age. There was no growth of weeds inside the tubes during the sampling period. The residue site was not replanted during the course of the experiment. Whilst the operational site was replanted 8 months following
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commencement, the 5 m spacing and the age of the seedlings allowed a large enough space to be left between trees and tubes for zero N uptake by the seedlings to be a reasonable assumption. The tubes were stored in a cold room at 4 8C until analysis, usually within 2 or 3 days following sampling. The soil cores were extracted wholly from the tubes and divided into samples representing the 0–10 and 10–20 cm depths. Each sample was thoroughly mixed with stones being removed by hand, but the high clay content of the soil and number of samples to be prepared excluded the possibility of passing the soils through a sieve. Mineral N was extracted from the samples by shaking with 2 M KCl at a soil solution ratio of 1:10, followed by centrifuging at 2000 rpm for 20 min and finally filtering through Whatman 42 filter papers.
volatilisation had exceeded the rate of mineralisation for that period; for nitrate N and nitrite N, negative values were assumed to show losses through immobilisation or denitrification. Differences between the capped and uncapped cores were assumed to demonstrate N losses through leaching. Data for leaching were calculated using both soil depths to represent leaching events within the entire core. Data were analysed using the SPSS Base 10 System (SPSS, 1999). Only the data for soil moisture had a normal distribution, therefore non-parametric statistical analyses were performed on those parts of the data where normal distribution could not be assumed. Comparison of means used the Mann–Whitney U test and correlations were established using Spearman’s rho (rank correlation).
2.4. Chemical analysis
3. Results
The extractant was analysed colorimetrically for ammonium N, nitrate N and nitrite N using a Lachat flow injection analyser. Results were verified using external standards after every 11 samples. A minimum of 10% duplication of samples was employed to verify the accuracy and repeatability of the results. A subsample of each soil was dried at 105 8C for 24 h to determine soil moisture content. Soil properties were determined as reported by Xu et al. (1995).
3.1. Rainfall, temperature and soil moisture
2.5. Analysis of results Operational limitations necessitated the establishment of two residue treatment areas, each to be sampled using a stratified random sampling design (Dick et al., 1996). However, at each site the management history, slope, aspect and environmental conditions were the same. Comparisons are drawn between the effects of the different management techniques employed at the two sites using appropriate statistical analysis to obtain levels of significance (Oksanen, 2001). The difference in soil mineral N between the capped core and the baseline core was used to measure the N dynamics for the sampling period. The dynamics for a given N form are cumulative and, for ammonium N, where this value was negative, it was assumed that the rate of loss through immobilisation, nitrification or
The mean monthly rainfall and mean monthly maximum air temperature for the sampling period are given in Fig. 1a and b, respectively. Temperature is the mean for each day of the sampling period. The first 12 sampling cycles had an unusually wet winter (sampling cycles 3–9), whereas the second 12 sampling cycles were more typical. The air temperature shows the seasonal variation typical of the region. The residue site (Fig. 2a) had approximately 60% soil moisture at the 0–10 cm depth at the onset of the experiment. This level declined steadily to a little over 20% in the 21st sampling cycle when seasonal rains began. Compared with the 0–10 cm depth, the 10– 20 cm depth had significantly lower soil moisture content (P < 0:001) though the values converged toward the end of the sampling period. The operational site showed similar trends (Fig. 2b) though the starting soil moisture content of approximately 40% at the 0– 10 cm depth was lower than the residue site. In the first 12 sampling cycles, soil moisture content at the 0– 10 cm depth was significantly higher at the residue site than at the operational site (P < 0:05), but there was no significant difference between sites for the remainder of the sampling period (P > 0:05). There was no significant difference in soil moisture content between sites at the 10–20 cm depth.
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Fig. 1. Monthly rainfall (a) and mean monthly maximum air temperature (b) for the sampling period.
3.2. Ammonium N 3.2.1. Residue site Ammonium N was the principal form of mineral N immediately following harvest though levels at the 0– 10 cm depth declined from around 60 kg N ha1 at the onset to <10 kg N ha1 in the first 14 sampling cycles (Fig. 3a). The levels of ammonium N remained at <10 kg N ha1 for the balance of the sampling period. Ammonium N was significantly correlated with soil moisture at the 0–10 cm depth throughout the sampling period (P < 0:001, R2 ¼ 0:20, data not shown). Ammonium N was significantly less at the 10–20 cm than at the 0–10 cm depth during the first 12 sampling cycles (P < 0:001), but remained at similar levels for the remainder of the sampling period. At the 0–10 cm depth, Ammonium N dynamics (Fig. 3b) showed a strong negative response of approximately
100 kg N ha1 during the sampling period. At the 10–20 cm depth there was a significantly lesser negative response at 40 kg N ha1 (P < 0:001). 3.2.2. Operational site Ammonium N was the principal form of mineral N following harvest, but levels at the 0–10 cm depth declined from about 55 to <10 kg N ha1 in the 12 sampling cycles following harvest (Fig. 3c), and then remained at <10 kg N ha1 for the remainder of the period. Ammonium N was significantly less at the 10– 20 cm than at the 0–10 cm depth during the first 12 sampling cycles (P < 0:001), but then remained at similar levels. Ammonium N dynamics showed a negative response of around 30 kg N ha1 for the 0–10 cm depth and 10 to 20 kg N ha1 for the 10–20 cm depth (Fig. 3d). The difference in ammonium N dynamics between depths was not significant (P > 0:05).
Fig. 2. Moisture content at 0–10 and 10–20 cm soil depth for: (a) the residue site (mean standard error (M.S.E.) 0.79 and 0.44, respectively, n ¼ 447); (b) the operational site (M.S.E. 0.71 and 0.48, respectively, n ¼ 448).
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Fig. 3. Ammonium N fluxes with M.S.E. at both depths over the sampling period for: (1) the residue site showing: (a) ammonium N in the baseline incubation tubes with M.S.E. 1.18 (0–10 cm) and 0.66 (10–20 cm) (n ¼ 442); (b) cumulative ammonium N dynamics with M.S.E. 3.26 (0–10 cm) and 2.56 (10–20 cm) (n ¼ 444). (2) The operational site showing (c) ammonium N in the baseline incubation tubes with M.S.E. 1.72 (0–10 cm) and 0.88 (10–20 cm) (n ¼ 436); (d) cumulative ammonium N dynamics with M.S.E. 4.44 (0–10 cm) and 4.05 (10– 20 cm) (n ¼ 446).
3.2.3. Management Ammonium N dynamics were significantly lower at the residue site than at the operational site for both the 0–10 cm (P < 0:001) and 10–20 cm (P < 0:01) depth. 3.3. Nitrate N 3.3.1. Residue site Nitrate N at the 0–10 cm depth declined from around 45 kg N ha1 during the first three sampling cycles following harvest (Fig. 4a), to around
10 kg N ha1 until the 10th sampling cycle and then rose to approximately 20 kg N ha1 towards the end of the sampling period. Nitrate N was negatively correlated with rainfall at the 0–10 cm depth (P < 0:01, data not shown) and positively correlated with temperature at both the 0–10 cm (P < 0:01, data not shown) and 10–20 cm (P < 0:01, data not shown) depths. The difference in nitrate N between depths was significant (P < 0:001). Nitrate N dynamics at the 0–10 cm depth was positive (Fig. 4b), rising steadily to approximately 200 kg N ha1 at the end of the sampling period; at
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Fig. 4. Nitrate N fluxes with M.S.E. for both depths over the sampling period for: (1) the residue site showing (a) nitrate N in the baseline incubation tubes with M.S.E. 1.32 (0–10 cm) and 0.48 (10–20 cm) (n ¼ 436); (b) cumulative nitrate N dynamics with M.S.E. 8.27 (0– 10 cm) and 1.76 (10–20 cm) (n ¼ 446); (c) nitrate N losses through leaching with M.S.E. 9.03 (0-20 cm, n ¼ 225). (2) The operational site showing (d) nitrate N in the baseline incubation tubes with M.S.E. 1.05 (0–10 cm) and 0.60 (10–20 cm) (n ¼ 441); (e) cumulative nitrate N dynamics with M.S.E. 7.77 (0–10 cm) and 1.70 (10–20 cm) (n ¼ 444); (f) nitrate N losses through leaching with M:S:E: ¼ 7:38 (0–20 cm, n ¼ 225).
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the 10–20 cm depth there was no significant change in nitrate N dynamics. At the beginning of the sampling period about 50 kg nitrate N ha1 was lost through leaching, but there was no further significant leaching of nitrate N from the 0 to 20 cm depth until the onset of the wet season in the 21st sampling cycle when losses of nitrate N through leaching rose steadily (Fig. 4c).
but at the 10–20 cm depth the operational site had higher nitrate N overall (P < 0:05). Nitrate N dynamics at the 0–10 cm depth were significantly higher at the operational site (P < 0:001), but there was no significant difference between the sites at the 10–20 cm depth. Nitrate N losses through leaching were significantly higher at the operational site (P < 0:001). 3.4. Nitrite N
3.3.2. Operational site Nitrate N at the 0–10 cm depth declined from around 35 kg N ha1 during the first two sampling cycles following harvest (Fig. 4d) to <10 kg N ha1 at the 10th sampling cycle. Nitrate N then peaked at around 35 kg N ha1 in the wet warmer summer period (sampling cycles 10–15) and then remained at around 15– 20 kg N ha1. Nitrate N was not correlated with rainfall at either depth, but was significantly correlated with temperature at both depths (0–10 cm, P < 0:001; 10–20 cm, P < 0:001; data not shown). Nitrate dynamics were positive throughout the sampling period at the 0–10 cm depth with a cumulative value of 300 kg N ha1 at the end of the sampling period (Fig. 4e). Nitrate N losses through leaching were steady with a total of around 225 kg N ha1 lost by the end of the sampling period (Fig. 4f). 3.3.3. Management There was no significant difference in baseline nitrate N between the sites at the 0–10 cm depth (P > 0:05),
3.4.1. Residue site Nitrite N was detected during the warm, wet summer at the 0–10 cm depth and at the 10–20 cm depth between the 10th and 16th sampling cycles (Fig. 5a) though the levels, 2 and 4 kg N ha1, respectively, were very low. Nitrite N was significantly correlated with air temperature at both the 0–10 and 10–20 cm depths (P < 0:01, data not shown). Nitrite N dynamics and leaching were very low (data not shown). 3.4.2. Operational site Nitrite N was detected at low levels, 4 kg N ha1 at the 0–10 cm depth and 2 kg N ha1 at the 10–20 cm depth between the 10th and 16th sampling cycles (Fig. 5b). Nitrite N was significantly correlated with air temperature and soil moisture content at both the 0–10 and 10–20 cm depths (P < 0:01, data not shown). Nitrite N dynamics were at very low levels and the overall losses through leaching of nitrite N were very low (data not shown).
Fig. 5. Nitrite N at the 0–10 cm and the 10–20 cm depth for (a) residue site (M.S.E. 0.65 and 0.09, respectively, n ¼ 436); (b) operational site (M.S.E. 0.73 and 0.06, respectively, n ¼ 441).
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4. Discussion Friedel et al. (2000) questioned the validity of the in situ mineralisation technique, due to the abscission of fine roots by the core, suggesting that ionexchange resins mixed into the soil and then incubated in the laboratory would better simulate the N immobilisation by fine roots. However, such an approach denies the basic principles underlying in situ mineralisation: that the soils are undisturbed and that mineralisation occurs under actual field conditions. Knoepp and Swank (1995) compared the in situ N mineralisation technique with in situ buried bags, tension lysimeters and aerobic laboratory incubations as a means of estimating N mineralisation, highlighting that the in situ core method was superior as the procedure took into account the changes in sitespecific parameters such as soil moisture content and temperature. 4.1. Ammonium N Harvesting increases N mineralisation (Attiwill and Adams, 1993; Carlyle, 1994), with soil disturbance and increases in soil temperature following canopy removal stimulating microbially mediated N mineralisation. The resulting increase in plant available mineral N is therefore introduced into the system at a time when there are no trees to absorb the available nutrients (Johnston and Crossley, 2002). Our results show that, regardless of residue management, over 100 kg N ha1 were available in the first few sampling cycles following the harvest of a 1R hoop pine plantation; such levels of N loss could lead to a substantial decline in potential site productivity. Ammonium N is the initial product of N mineralisation and has the potential to be lost through volatilisation and leaching, to be immobilised by both biotic assimilation and abiotic fixation or transformed through nitrification. Ammonia volatilisation is normally associated with burning, a practice that has ceased in hoop pine plantations (Ryan and Gilmour, 1985); ammonium N leaching has been shown not to be a significant factor in soils where there is no N saturation (Harmsen and Kolenbrander, 1965). Comparison of baseline and capped core data in this study, suggests that either N immobilisation or nitrification was the principal process affecting ammonium N
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during the 12 sampling cycles when levels declined from around 60 to 10 kg N ha1. N fixation to the soil will account for some of the apparent immobilisation of ammonium N and can occur quite rapidly (Trehan, 1996; Berntson and Aber, 2000). However, it has been demonstrated that up to 60% of the ammonium N that is bound to the soil surface may also become available for biological assimilation (Schulten and Schnitzer, 1998). Microbial immobilisation of ammonium N is well documented (Tietema et al., 1998; Puri and Ashman, 1999; Mao et al., 2002), particularly when there is an adequate supply of available C. The residue site showed significantly higher soil moisture during the sampling period and it may be assumed that the conditions for microbial activity were enhanced through both physical protection from soil drying and a readily available C supply provided by the residue layer (Burket and Dick, 1997). Pu et al. (2001) reported that N immobilisation in the hoop pine plantations was significantly higher underneath residues than in areas where the residues had been removed. As both nitrification and N immobilisation are microbially dependent, both processes would be competing for the available ammonium N. 4.2. Nitrate N Unlike northern coniferous forest soils, which are regarded as having a low nitrification potential (Davidson et al., 1992), the onset of nitrification was immediate and durable in the hoop pine soils. Nitrate N formed approximately 40% of total mineral N at both sites in the first three sampling cycles following harvest, indicating that substantial nitrification was taking place. Nitrate N may be readily lost through leaching if it is not immobilised by the soil microbial biomass. It has been claimed that little immobilisation of nitrate N will occur in the presence of ammonium N (Attiwill and Adams, 1993) because, unlike nitrate N, ammonium N does not require reduction and is therefore more easily assimilated by the microbial biomass (Puri and Ashman, 1999). However, where there is a sufficient supply of available C, nitrate immobilisation may take place, even when rates of nitrification are high (Stark and Hart, 1997). Nitrate N dynamics showed a high level of nitrification (300 kg N ha1) at the 0–10 cm depth at the
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operational site. Ammonium N dynamics were close to zero, indicating that the rate of ammonium N production through mineralisation and the rate of transformation to nitrate N were equal throughout the period. With overall levels of ammonium N declining, this suggests that ammonium N was the factor limiting nitrate production. In a study of soils amended with straw, Recous et al. (1998) also concluded that the supply of ammonium ions was probably the factor limiting nitrification. The nitrate N dynamics at the residue site show a lower level of nitrification (200 kg N ha1) than the operational site, whilst the ammonium N dynamics were negative (100 kg N ha1) for the same period. The presence of organic matter can stimulate a rapid expansion in the soil microbial population (Dalias et al., 2002), and rates of both N mineralisation and immobilisation have been shown to be higher in soils amended with straw (Recous et al., 1998). The negative ammonium N dynamics and lower rates of nitrification at the residue site indicate that immobilisation was competing for ammonium N released through the mineralisation of organic matter, there may also have been some immobilisation of nitrate N which has been shown to take place when levels of soil C are high (Stark and Hart, 1997). Pu et al. (2001) also found immobilisation of N to be higher in areas under post-harvest residues compared to those areas where the residues had been removed. Whilst denitrification is another potential mechanism for loss of nitrate in forest soils, it mainly occurs in saturated soils under anaerobic conditions and the potential is increased by N saturation (Mohn et al., 2000). The soils in the study area are free draining and saturation is difficult to achieve (Pu et al., 2001). Though denitrification has been recorded in northern hemisphere soils at soil moisture contents <50% (Stevens et al., 1997), other studies have shown that even at 80% water-filled pore space, N2O emissions can be as low as 8 kg N ha1 over a 10-month period (Smith et al., 1998). Vor and Brumme (2002) found that in situ N mineralisation increases denitrification. However, the compaction and anaerobic conditions associated with their 40 cm long, 7.6 cm wide steel tubes would have been far less likely to occur with the 20 cm long, 10 cm wide, aerated tubes that were used in this study. While the potential for N losses through denitrification in these soils can be high (Pu
et al., 2001) and some minor denitrification may occur as soil aeration status changes with depth (Whalley et al., 1995), the conditions promoting significant denitrification occur only intermittently (Pu et al., 2001). Immobilisation of both ammonium N and nitrate N at the residue site would prevent some N loss through leaching. The N leaching losses were low at the residue site until the 21st sampling cycle at the onset of the warm, wet summer. It was observed that by this time, most of the foliage (which formed the finer residue) had decomposed, leaving the coarser and more recalcitrant fractions in this area. The exhaustion of the readily available C within the finer foliage would stimulate the release of N from within the microbial biomass with subsequent nitrification and leaching. Parfitt et al. (2001) found that N from decomposing litter was initially stored within the litter before subsequent release into the soil layer. Over 100 kg N ha1 were lost through leaching from the residue site in the final six sampling cycles. N losses through leaching at the operational site demonstrated a seasonal variation with rainfall, which suggests that little or no N immobilisation had occurred. 4.3. Nitrite N Nitrite N is not normally considered to be a significant part of the hoop pine forest mineral N pool (Fisher and Binkley, 2000). Data presented here demonstrate that under certain environmental conditions it can be present at levels up to 4 kg N ha1, which is below that considered toxic to plants and bacteria (Paul and Clark, 1996). There were no significant differences in baseline nitrite N between the sites. There was a weak, but significant correlation of nitrite N with air temperature and soil moisture between 13 and 26 sampling cycles. It has been previously reported that denitrification may occur in these soils when wetted (Pu et al., 2001) and as nitrite N is an intermediate stage in denitrification (Fisher and Binkley, 2000), it is reasonable to assume that some denitrification was occurring in the wet summer months. The intermittent nature and low levels of nitrite N agree with the previous conclusion that denitrification would occur only sporadically in these soils.
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4.4. Site management Residue retention favours microbial activity by acting as a physical protection, maintaining levels of soil moisture and providing a readily available supply of organic C. Barrett and Burke (2000) found higher rates of N immobilisation in areas under plants, which had greater levels of available C and higher rates of microbial activity, than in areas of open soil between the plants. Hart et al. (1994) also found that C availability exerted a strong influence on N immobilisation. Following their work on 1R hoop pine plantations, Bubb et al. (1998) suggested that there would be the potential for substantial losses of N due to leaching following clearfall. These studies have indicated that retaining residues as an even ground cover can prevent N leaching losses through N immobilisation in the first 18 months following harvesting. It is evident that hoop pine seedlings in 2R soils are N limited (Xu et al., 2002), but the extent to which young trees would be capable of gaining access to the available N within the soil would be governed by the development of the tree’s root system. It would therefore seem probable that, for the period covered by this study, the trees would not be disadvantaged by the immobilisation of available N. As suggested by Kushwaha et al. (2000), the presence of high C:N ratio residues may act as a slow release fertiliser through the immobilisation and eventual re-mineralisation of the mineral N. Similarly, Carlyle (1994) suggested that residue retention would have the capacity to allow for a steady increase in N availability without incurring the risk of leaching. Higher N immobilisation under residue retention may prevent N losses that would otherwise occur through volatilisation, leaching and denitrification.
5. Conclusion Soil mineral N flux is the result of a complicated and dynamic series of inter-related mechanisms that are highly responsive to environmental conditions, including human disturbance. At any moment in time, most of the processes described above may be operating; with changes in soil mineral N dictated by whichever process is dominant. The in situ core technique for studying soil N mineralisation was a
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useful tool for unravelling some of the complexities of mineral N dynamics following the severe disturbance of clearfall harvesting. Harvesting initiated a flush of N mineralisation at both sites that persisted beyond the 2-year sampling period. Whereas established 1R hoop pine plantations are known to have low rates of nitrification, harvesting disturbance increased the rate of nitrification and brought about a greater potential for N loss through both leaching and denitrification during the critical inter-rotation and early establishment phase of 2R hoop pine plantations. Retaining post-harvest residues on-site was initiated to try and mitigate some of the N losses that had been occurring under the previous residue-burning regime. Whilst windrowing of post-harvest residues retained the residues on-site, serious concerns regarding the potential for N losses through leaching and denitrification in the cleared inter-windrow spaces remained. Retaining the residues as mulch, whilst posing operational problems, has the potential for retaining mineral N on-site and available for the trees of the next rotation. Incidental effects of weed suppression and moisture retention would also benefit the young trees. Whilst the ultimate management decisions regarding residue retention would probably be based on economic factors, an operational trial coupled with a costbenefit analysis of alternative residue-management systems is recommended.
Acknowledgements T.J. Blumfield was supported in this work through a scholarship grant from the Cooperative Research Centre for Sustainable Production Forestry. We would like to acknowledge the invaluable help and support of Mr Paul Keay and thank Griffith University and the Queensland Forestry Research Institute for their technical help and support. We are grateful to Dr Chris Beadle for his constructive comments on earlier versions of the manuscript. References Attiwill, P.A., Adams, M.A., 1993. Tansley Review No. 50. Nutrient cycling in forests. New Phytol. 124, 561–582. Ballard, T.M., 2000. Impacts of forest management on northern forest soils. For. Ecol. Manage. 133, 37–42.
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Barrett, J.E., Burke, I.C., 2000. Potential nitrogen immobilization in grassland soils across a soil organic matter gradient. Soil Biol. Biochem. 32, 1707–1716. Berntson, G.M., Aber, J.D., 2000. Fast nitrate immobilization in N saturated temperate forest soils. Soil Biol. Biochem. 32, 151– 156. Bubb, K.A., Xu, Z.H., Simpson, J.A., Saffigna, P.G., 1998. In situ measurements of soil mineral–nitrogen fluxes in hoop pine plantations of subtropical Australia. NZ J. For. Sci. 28, 152– 164. Burket, J.Z., Dick, R.P., 1997. Long-term vegetation management in relation to accumulation and mineralisation of nitrogen in soils. In: Cadisch, G., Giller, K.E. (Eds.), Driven by Nature: Plant Litter Quality and Decomposition. CAB International, Wallingford. Carlyle, J.C., 1986. Nitrogen cycling in forested ecosystems. For. Abst. 47, 307–336. Carlyle, J.C., 1994. Opportunities for managing nitrogen uptake in established Pinus radiata plantations on sandy soils. NZ J. For. Sci. 24, 344–361. Costantini, A., Grimmet, J.L., Dunn, G.M., 1997. Towards sustainable management of forest plantations in south-east Queensland. 1. Logging and understorey residue management between rotations in steep country Araucaria cunninghamii plantations. Aust. For. 60, 213–225. Dalias, P., Anderson, J.M., Bottner, P., Couteaux, M.-M., 2002. Temperature responses of net nitrogen mineralization and nitrification in conifer forest soils incubated under standard laboratory conditions. Soil Biol. Biochem. 34, 691–701. Davidson, E.A., Hart, S.C., Firestone, M.K., 1992. Internal cycling of nitrate in soils of a mature coniferous forest. Ecology 73, 1148–1156. Dick, R.P., Thomas, D.R., Halvorson, J.J. (Eds.), 1996. Standardized Methods, Sampling and Sample Pretreatment. In: Methods for Assessing Soil Quality. Soil Science Society of America, Madison, pp. 107–121. FAO, 2001. State of the world’s forests 2001. Technical Report. Food and Agriculture Organisation of the United Nations, Rome. Fisher, R.F., Binkley, D., 2000. Ecology and Management of Forest Soils. Wiley, New York. Friedel, J.K., Herrmann, A., Kleber, M., 2000. Ion exchange resin– soil mixtures as a tool in net nitrogen mineralisation studies. Soil Biol. Biochem. 32, 1529–1536. Harmsen, G.W., Kolenbrander, G.J., 1965. Soil inorganic nitrogen. In: Bartholomew, W.V., Clark, F.E. (Eds.), Soil Nitrogen. American Society of Agronomy, Inc., Madison. Hart, S.C., Nason, G.E., Myrold, D.D., Perry, D.A., 1994. Dynamics of gross nitrogen transformations in an old-growth forest: the carbon connection. Ecology 75, 880–891. Isbell, R.F., 1996. The Australian Soil Classification. CSIRO Australia, Collingwood, Vic. Johnston, J.M., Crossley Jr., J.D.A., 2002. Forest ecosystem recovery in the southeast US: soil ecology as an essential component of ecosystem management. For. Ecol. Manage 155, 187–203. Jones, H.E., Madeira, M., Herraez, L., Dighton, J., Fabiao, A., Gonzalez-Rio, F., Fernandez Marcos, M., Gomez, C., Tome,
M., Feith, H., 1999. The effect of organic-matter management on the productivity of Eucalyptus globulus stands in Spain and Portugal: tree growth and harvest residue decomposition in relation to site and treatment. For. Ecol. Manage. 122, 73–86. Keeney, D.R., 1980. Prediction of soil nitrogen availability in forest ecosystems: a literature review. For. Sci. 26, 159–171. Knoepp, J.D., Swank, W.T., 1995. Comparison of available soil nitrogen assays in control and burned forested sites. Soil Sci. Soc. Am. J. 59, 1750–1754. Kushwaha, C.P., Tripathi, S.K., Singh, K.P., 2000. Variations in soil microbial biomass and N availability due to residue and tillage management in a dryland rice agroecosystem. Soil Till. Res. 56, 153–166. Mao, X.A., Xu, Z.H., Luo, R.S., Mathers, N.J., Zhang, Y.H., Saffigna, P.G., 2002. Nitrate in soil humic acids revealed by 14 N nuclear magnetic resonance spectroscopy. Aust. J. Soil Res. 40, 717–726. Mathers, N.J., Xu, Z.H., Berners-Price, S.J., Perera, M.C.S., Saffigna, P.G., 2002. Hydrofluoric acid pre-treatment for improving 13 C CPMAS NMR spectral quality of forest soils in south-east Queensland, Australia. Aust. J. Soil Res. 40, 655– 674. Mohn, J., Schurmann, A., Hagedorn, F., Schleppi, P., Bachofen, R., 2000. Increased rates of denitrification in nitrogen-treated forest soils. For. Ecol. Manage. 137, 113–119. Nambiar, E.K.S., 1996. Sustained productivity of forests is a continuing challenge to soil science. Soil Sci. Soc. Am. J. 60, 1629–1642. Oksanen, L., 2001. Logic in experiments in ecology: is pseudoreplication a pseudoissue. Oikos 94, 27–38. Parfitt, R.L., Salt, G.J., Saggar, S., 2001. Post-harvest residue decomposition and nitrogen dynamics in Pinus radiata plantations of different N status. For. Ecol. Manage. 154, 55– 67. Paul, E.A., Clark, F.E., 1996. Soil Microbiology and Biochemistry. Academic Press, San Diego. Perakis, S.S., Hedlin, L.O., 2002. Nitrogen loss from unpolluted South American forests mainly via dissolved organic compounds. Nature 415, 416–419. Pu, G.X., Saffigna, P.G., Xu, Z.H., 2001. Denitrification, leaching and immobilisation of 15 N-labelled nitrate in winter under windrowed harvesting residues in hoop pine plantations of 1–3 years old in subtropical Australia. For. Ecol. Manage. 152, 183– 194. Puri, G., Ashman, M.R., 1999. Microbial immobilization of 15 Nlabelled ammonium and nitrate in a temperate woodland soil. Soil Biol. Biochem. 31, 929–931. Raison, R.J., Connell, M.J., Khanna, P.K., 1987. Methodology for studying fluxes of soil mineral-N in situ. Soil Biol. Biochem. 19, 521–530. Recous, S., Aita, C., Mary, B., 1998. In situ changes in gross N transformations in bare soil after addition of straw. Soil Biol. Biochem. 31, 119–133. Ryan, P.A., Gilmour, D., 1985. Potential erosion and nutrient losses in the conversion to the second rotation hoop pine crop. Queensland Forestry Research Institute Technical Report. Queensland Forestry Research Institute, Gympie, Qld, 28 pp.
T.J. Blumfield, Z.H. Xu / Forest Ecology and Management 179 (2003) 55–67 Schulten, H.R., Schnitzer, M., 1998. The chemistry of soil organic nitrogen: a review. Biol. Fert. Soils. 26, 1–15. Smith, K.A., Thomson, P.E., Clayton, H., McTaggart, I.P., Conen, F., 1998. Effects of temperature, water content and nitrogen fertilisation on emissions of nitrous oxide by soils. Atmos. Environ. 32, 3301–3309. Soil Survey Staff, 1999. Soil Taxonomy: A Basic System of Soil Classification for Making and Interpreting Soil Surveys. US Department of Agriculture Soil Conservation Service, Washington. SPSS, 1999. SPSS Base 10 Application Guide. SPSS, Inc., Chicago. Stark, J.M., Hart, S.C., 1997. High rates of nitrification and nitrate turnover in undisturbed coniferous forests. Nature 385, 61–64. Stevens, R.J., Laughlin, R.J., Burns, L.C., Arah, J.R.M., Hood, R.C., 1997. Measuring the contributions of nitrification and denitrification to the flux of nitrous oxide from soil. Soil Biol. Biochem. 29, 139–151. Tietema, A., Emmett, B.A., Gundersen, P., Kjonaas, O.J., Koopmans, C.J., 1998. The fate of 15 N-labelled nitrogen deposition
67
in coniferous forest ecosystems. For. Ecol. Manage. 101, 19– 27. Trehan, S.P., 1996. Immobilisation of 15 NH4 in three soils by chemical and biological processes. Soil Biol. Biochem. 28, 1021–1027. van Breemen, N., 2002. Natural organic tendency. Nature 415, 381–382. Vor, T., Brumme, R., 2002. N2O losses result in underestimation of in situ determinations of net N mineralization. Soil Biol. Biochem. 34, 541–544. Webb, L.J., Tracey, J.G., 1967. An ecological guide to new planting areas and site potential for hoop pine. Aust. For. 31, 224–239. Whalley, W.R., Dumitru, E., Dexter, A.R., 1995. Biological effects of soil compaction. Soil Till. Res. 35, 53–68. Xu, Z.H., Simpson, J.A., Osborne, D.O., 1995. Mineral nutrition of slash pine in subtropical Australia. I. Stand growth response to fertilization. Fert. Res. 41, 93–100. Xu, Z.H., Bubb, K.A., Simpson, J.A., 2002. Effects of nitrogen fertilisation and weed control on nutrition and growth of a 4year-old Araucaria cunninghamii plantation in subtropical Australia. J. Trop. For. Sci. 14, 213–222.