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Dynamics of Nitrogen Speciation in Horticultural Soils in Suburbs of Shanghai, China
Pedosphere 20(2): 261–272, 2010 ISSN 1002-0160/CN 32-1315/P c 2010 Soil Science Society of China Published by Elsevier Limited and Science Press
Dynamics of Nitrogen Speciation in Horticultural Soils in Suburbs of Shanghai, China∗1 GE Ti-Da1,2,3 , HUANG Dan-Feng1,∗2 , P. ROBERTS2 , D. L. JONES2 and SONG Shi-Wei1 1 School
of Agriculture and Biology, Shanghai Jiao Tong University, Shanghai 201101 (China) of the Environment and Natural Resources, University of Wales, Bangor, Gwynedd LL57 2UW (UK) 3 Key Laboratory of Agro-ecological Processes in Subtropical Region, Institute of Subtropical Agriculture, Chinese Academy of Sciences, Hunan 410125 (China) 2 School
(Received September 11, 2009; revised January 5, 2010)
ABSTRACT Dissolved organic nitrogen (DON) represents a significant pool of soluble nitrogen (N) in soil ecosystems. Soil samples under three different horticultural management practices were collected from the Xiaxiyang Organic Vegetable and Fruit Farm, Shanghai, China, to investigate the dynamics of N speciation during 2 months of aerobic incubation, to compare the effects of different soils on the mineralization of 14 C-labeled amino acids and peptides, and to determine which of the pathways in the decomposition and subsequent ammonification and nitrification of organic N represented a significant blockage in soil N supply. The dynamics of N speciation was found to be significantly affected by mineralization and immobilization. DON, total free amino acids, and NH+ 4 -N were maintained at very low levels and did not accumulate, whereas NO− 3 -N gradually accumulated in these soils. The conversion of insoluble organic N to low-molecular-weight + (LMW) DON represented a main constraint to N supply, while conversions of LMW DON to NH+ 4 -N and NH4 -N to -N did not. Free amino acids and peptides were rapidly mineralized in the soils by the microbial community and NO− 3 consequently did not accumulate in soil. Turnover rates of the additional amino acids and peptides were soil-dependent and generally followed the order of organic soil > transitional soil > conventional soil. The turnover of high-molecular-weight DON was very slow and represented the major DON loss. Further studies are needed to investigate the pathways and bottlenecks of organic N degradation. Key Words:
amino acids, dissolved organic N, mineralization, N transformation, peptides
Citation: Ge, T. D., Huang, D. F., Roberts, P., Jones, D. L. and Song, S. W. 2010. Dynamics of nitrogen speciation in horticultural soils in suburbs of Shanghai, China. Pedosphere. 20(2): 261–272.
INTRODUCTION Dissolved organic nitrogen (DON) has been recognized as a potential source of nitrogen (N) for use by microorganisms and plants, and it plays an important role in N cycling in most ecosystems (Neff et al., 2003; Jones et al., 2004a; Chen et al., 2005a; Christou et al., 2006). However, plant and soil responses to DON have received much less attention compared with those to inorganic N (NO− 3 -N and + NH4 -N). This is primarily due to technical difficulties, as it is difficult to experimentally show direct plant uptake of DON or its ecological significance (Bhogal et al., 2000; Zhu and Carreiro, 2004; Jones et al., 2005; Kuzyakov and Jones, 2006). Recently, some clear experimental evidence to support the direct uptake of low-molecular-weight (LMW) DON, specifically amino acids, has been reported for plant roots and associated mycorrhizas (Streeter et al., 2000; Xu et al., 2004). Uptake of organic N species may reduce the reliance of plants on soil microorganisms for conversion of soil organic matter to inorganic N. The DON pool is generally mobile within the soil solution and can easily be transported out of ∗1 Project
supported by the National High Technology Research and Development Program (863 program) of China (No. 2006AA10A311), the National Natural Science Foundation of China (No. 40901124) and the Shanghai Leading Academic Discipline Program, China (No. B209). ∗2 Corresponding author. E-mail: [email protected].
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the soil and into groundwater by hydrological events and farming management (e.g., precipitation and + fertilization). However, compared with NO− 3 -N and NH4 -N, DON is less available and less likely to be lost by leaching or denitrification (Northup et al., 1995). DON consists of many compounds, including LMW compounds, such as amino acids, peptides, amino sugars, and urea, and high-molecular-weight (HMW) compounds, such as proteins, chlorophyll, DNA, and RNA (Antia et al., 1991). All DON forms enter the soil from a range of sources, including dry and wet deposition, throughfall, litterfall, root and microbial exudation, turnover of roots and organisms, urine and feces, and organic fertilizer additions (Kalbitz et al., 2000; Christou et al., 2006). The classical mineralization-immobilization turnover (MIT) theory suggests that the dominant pathway of soil-derived N in agricultural and many natural systems is the decomposition of insoluble organic N to NH+ 4 -N prior to microbial assimilation (Hadas et al., 1992). DON compounds that can enter into soil are free and polymeric amino acids (proteins and peptides) (Jones et al., 2005). Microbial enzymes involved in this process include oxidases, hydrolases, lyases, and deaminases (Jones et al., 2004a). These enzymes function endocellularly, in dead, autolysing cells, either free in solution or while absorbed to soil colloids. Barraclough (1997) has suggested an alternative route for assimilation of LMW DON. He suggests that the LMW DON can be used directly by the microbial cell (i.e., direct assimilation), where endogenous enzymes decompose these compounds. In this case, only the N in excess of microbial demand is released. Few studies have yet focused on the source of NH+ 4 -N from DON in agricultural soils (Murphy et al., 2000). Although soil organic matter can produce NH+ 4 -N by direct deamination, the main route for + NH4 -N production in soil appears to operate through extracellular enzymes that first convert insoluble organic N into DON (Zaman et al., 1999). If this DON pool has a LMW fraction, the transformation of proteins to amino acids, rather than amino acids to NH+ 4 -N, may be the major factor limiting N availability in low N soils (Jones and Kielland, 2002a). Subsequent turnover of this microbial population will lead to the production of NH+ 4 -N (Griffiths, 1994; Bonkowski et al., 2000). Previous studies carried out in grassland soils have indicated that the uptake of LMW DON by soil microorganisms in NO− 3 -N rich soils may primarily provide them with C to fuel respiration, rather than with N to satisfy their internal N demand (Jones et al., 2004a). However, to our knowledge, the role of DON in N cycling has not been intensively studied in Chinese horticultural soils. We hypothesize that there is a significant blockage of N decomposition pathways and subsequent ammonification and nitrification of organic N in soil N supply in different horticultural soils. There is also little information available on the dynamics of protein and microbial biomass yields as a function of time-dependent removal of amino acids and peptides. Therefore, the main aims of this study were: 1) to investigate the dynamics of N speciation and LMW DON mineralization in three soils subjected to common horticultural and agricultural practices in China; 2) to compare the effects of different soils on the mineralization of 14 C-labeled amino acids and peptides; and 3) to determine which of three possible degradation pathways, N to DON, DON to + − NH+ 4 -N, and NH4 -N to NO3 -N, may be the most significant bottleneck in N supply in horticultural soils. MATERIALS AND METHODS Soils Soil samples were collected from three sites in the Xiaxiyang Organic Vegetable and Fruit Farm, Jiading District, Shanghai, China (31◦ 4 N, 121◦ 24 E, 4 m above sea level). The climate is humid subtropical. Approximately 70% of the annual precipitation of 1 255 mm falls from May to September, the mean annual temperature is 17.5 ◦ C, and the total sunlight is 1 778 h. The sampling sites selected represented three typical horticultural production systems: The organic site (approximate 20 ha) was certified (Certification No. OF-3105/3106-931-243-2005) by the Organic Farming Development Center
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of China (OFDC) in June 2004 after a 2 year conversion period; the transitional site (about 27 ha) was certified in June 2003 (Certification No. V-3106-931-264-2003); and the conventional site was situated in a greenhouse close to the organic and transitional plots. The land is used to grow vegetable crops, typically in plastic tunnels. At each site, leaf and vegetable litters were removed from the surface, soil samples of 0–20 and 20–60 cm depths were randomly taken from an area of 6 m × 60 m in early winter (October 2006) using a 5 cm diameter stainless steel corer. The soil samples were stored field-moist at 4 ◦ C with dry ice (solid CO2 ) in CO2 permeable polypropylene bags until analysis. Selected properties of the soils from the organic, transitional, and conventional systems, i.e., organic soil (OS), transitional soil (TS), and conventional soil (CS), are shown in Table I. TABLE I Selected properties of the three horticultural soils Soil
Dominant vegetation
Organic Transitional
Tomato and strawberry Cauliflower
Conventional
Cantaloupe
a) Mean±standard
Depth
Moisture content
Ash content
cm 0–20 20–60 0–20 20–60 0–20 20–60
g kg−1 198.8±9.5a) 51.3±1.1 213.2±5.5 30.9±0.8 171.9±7.5 42.5±0.9 224.9±3.3 37.2±1.9 222.8±2.2 31.8±0.2 252.3±15.1 26.6±1.7
pHH2 O
7.2±0.1 7.9±0.1 7.4±0.1 8.0±0.1 7.7±0.1 7.7±0.1
Electrical conductivity
Respiration
dS m−1 0.26±0.01 0.13±0.00 0.14±0.01 0.08±0.01 0.17±0.01 0.14±0.00
μmol CO2 kg−1 h−1 269.4±12.5 225.9±10.9 248.4±15.9 207.9±20.7 209.4±19.5 195.3±16.7
error (n = 3).
Aerobic incubation and preparation of soil extracts Aerobic incubation of soils was as described in the study by Zhong and Makeschin (2003). In brief, the field-moist soil samples were broken by hand into aggregates 0.5–5.0 cm in diameter, and large stones and roots (> 1 cm in diameter) were removed. The samples 250 g each in 4 replicates were then aerobically incubated in 500-mL plastic cups covered with Parafilm punched with 6 holes for aeration and placed in a dark, climate-controlled chamber maintained at 20 ± 1 ◦ C. Soil moisture content was maintained at field capacity throughout the experiment by randomly selecting and weighing five plastic cups each day. If the water loss from these cups exceeded 5 g, all cups were weighed and brought back to the original moisture content. This was necessary two or three times a week. Water additions were made directly to the soil, using a 5-mL pipette to ensure that a uniform distribution of water was obtained. After 0, 7, 14, 21, 28, 42, and 56 d of incubation, the soil samples were destructively harvested and extracted. Soil extracts were prepared using the procedure described by Jones and Willett (2006). Briefly, 5.0 g incubated soil was shaken with 25 mL of either 1 mol L−1 KCl or 1/15 mol L−1 phosphate buffer (pH 7.0) consisting of 14.6 g L−1 Na2 HPO4 ·12H2 O and 3.5 g L−1 KH2 PO4 in 50 cm3 polypropylene tubes on a reciprocating shaker (Unimax 2010, Heidolph Elekrto Gmbh, Kelheim, Germany) at a speed of 250 r min−1 for 30 min. After shaking, the soil extracts were centrifuged at 8 000 g for 15 min, and the supernatant was recovered and stored in polypropylene bottles at −20 ◦ C until analysis. The dynamics of N speciation in 1 mol L−1 KCl extracts were very similar to those in 1/15 mol L−1 phosphate buffer extracts; therefore, only the results from 1 mol L−1 KCl extracts are presented for this study. Chemical analysis Soil pH and electrical conductivity were determined in 1:1 (v/v) soil:H2 O extracts using standard electrodes. Soil moisture and ash contents were determined gravimetrically by drying at 105 ◦ C and combustion at 450 ◦ C for 24 h, respectively. Total C and total N were determined with a CHN-2000
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analyzer (Leco Corporation, St Joseph, USA). Soil respiration was measured at 20 ◦ C using a CIRASIRGA soil respirometer (PP Systems, Hitchin, UK). Total dissolved N (TDN) in soil extracts was measured with a Shimadzu TOC-TN analyzer (Shimadzu Corp, Kyoto, Japan) as described by Jones et al. (2004a). The 1/15 mol L−1 phosphate buffer extracts were analyzed without dilution, but the 1 mol L−1 KCl extracts were diluted 10-fold with de-ionized water before measurement, to prevent a large amount of salts from precipitating onto the surface of the Pt/Al2 O3 catalysts and reducing catalyst effi+ ciency (Chen et al., 2005b). NO− 3 and NH4 were determined using the methods described in Mulvaney (1996) with a Versamax microplate reader (Molecular Devices Corp, Sunnyvale, USA). DON was calcu+ lated as the difference between TDN and the sum of NO− 3 -N and NH4 -N. Total free amino acid (TFAA) concentrations were determined on a Varian Cary Eclipse 96-well plate fluorescence spectrophotometer (Varian Inc., Palo Alto, USA) using the o-phthaldialdehyde/β-mercaptoethanol procedure of Jones et al. (2002b). Proteins in the soil extracts were determined by the Bradford method (Bradford, 1976) using bovine serum albumin as a standard. Mineralization of amino acids and peptides To assess the amino acid and peptide mineralization rates in the soils, mixtures of two uniformly labeled amino acids, 14 C-labeled valine (Val) and glutamate (Glu), and two labeled peptides, 14 Clabeled glutamate-phenylalanine (Glu-Phe) and valine-proline-proline (Val-Pro-Pro), from the St Louis Arc, USA, were added to the soil samples and their subsequent evolution to 14 CO2 was measured over 24 h, as described in Jones (1999). Soil (5 g) was weighed into 50 mL polypropylene tubes, 0.5 mL 9 990 MBq mmol−1 14 C-labeled Val, Glu, Glu-Phe, or Val-Pro-Pro solution was added dropwise to a final concentration of 0.5 mmol L−1 (2.7 kBq kg−1 ). To trap any evolved 14 CO2 , a polypropylene vial containing 1 mL of 1 mol L−1 NaOH was placed above the soil in the tubes, and then the tubes were hermetically sealed and maintained at 15 ◦ C. Any 14 C-labeled amino acids or peptides remaining in the soil at 0.5, 1, 3, 6, 12, and 24 h after substrate addition were removed by shaking in 10 mL of 0.5 mol L−1 ice-cold K2 SO4 for 10 min at 250 r min−1 , followed by recovery of the supernatant solution by centrifugation (18 000 g, 5 min) for scintillation counting. All radioactivity was determined by liquid scintillation counting using a Wallac 1414 counter and the aqueous compatible scintillation fluid Wallac Optiphase Hisafe 3 (EG & G Ltd., Milton Keynes, UK). The amount of 14 C immobilized in the microbial biomass (microbial biomass carbon) was calculated as the difference: 100 − (14 CO2 in NaOH trap + 14 C remaining in 0.5 mol L−1 K2 SO4 extracts). The free amino acid and peptide concentrations added to all soil samples were 8.4 mg N L−1 (0.5 mmol L−1 ). We chose these compounds as they were commercially available and they represented common peptides likely to occur in the soil environment, based on known protein structures and sequence data. The concentration of 0.5 mmol L−1 was used in accordance with previous studies which have indicated that this is the likely concentration of an individual amino acid in a typical plant cell and the likely concentration that might appear in the soil if a root cell was to burst (Jones et al., 2005). The concentration of 14 C-labeled amino acids remaining in the soil extracts was calculated by the formula: C = 14 C-AAsoil × V × Cs × 14 × 1.2/14 C-AAtotal where C is the concentration of 14 C-labeled amino acids remaining in the extracts (mg N L−1 ), 14 CAAsoil is the 14 C decayed per minute at a known time, V is the K2 SO4 volume added to the soil (mL), Cs is the substrate initial concentration (mmol L−1 ), 14 C-AAtotal is the 14 C decayed per minute in the substrate in the beginning, 14 is the atomic weight of N, and 1.2 is the transfer coefficient. Statistical and data analysis All data were expressed as the mean of four replicates ± standard error. One-way analyses of variance (ANOVA) with Tukey test were used to identify treatment differences. Analyses were run using SPSS
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14.0 for Windows (SPSS Inc., Chicago, Illinois, USA). To estimate the amino acid and peptide half-lives in soil, a first-order exponential decay equation was fitted to the amino acid and peptide mineralization data, using Sigmaplot 8.01, by a least-square optimization routine (Jones et al., 2004a). RESULTS Soil characteristics In all three soils, the levels of ash, electrical conductivity, and respiration at the depth of 0–20 cm were higher than those at the depth of 20–60 cm, whereas the moisture content at 0–20 cm was lower than that at 20–60 cm (Table I). Soil ash and respiration were the greatest in the OS and least in the CS. The EC value was the highest in the OS and least in the TS. No significant difference in pH was observed among the three horticultural soils (P > 0.05). Dynamics of inorganic N and DON mineralization during aerobic incubation The concentration of NO− 3 -N gradually increased in all three soils during the 56-d incubation. The increase in NO− 3 -N concentration was almost linear with incubation time in all soils (Fig. 1). The rates −1 −1 d ) > TS (0.92 mg N kg−1 d−1 ) > CS of NO− 3 -N increase followed the order of OS (0.96 mg N kg −1 −1 (1.07 mg N kg d ) although the amounts were in the contrary order (Fig. 1). The amount of NO− 3 -N in the soil extracts increased by 7.5 folds in the OS, 4.8 folds in the TS, and 3.5 folds in the CS by the end of the incubation.
+ −1 KCl extracts of three horticultural soils incubated Fig. 1 Dissolved organic N (DON), NO− 3 -N, and NH4 -N in 1 mol L aerobically without leaching for 56 days. All values represent means with standard errors shown by vertical bars (n = 4). In some cases, the error bars are not shown where the symbols are larger than the error bars.
−1 The amounts of NH+ in the OS, 1.9 and 4 -N in the extracts varied between 1.7 and 3.8 mg N kg −1 −1 3.7 mg N kg in the TS, and 2.8 and 5.0 mg N kg in the CS, but no overall change was observed in either soil throughout the incubation (Fig. 1). At the final sampling (56 d), the amount of inorganic N + −1 in the OS, 69.4 mg N kg−1 in the TS, and 88.0 mg N kg−1 (NO− 3 -N and NH4 -N) was 66.1 mg N kg in the CS. In the TS and CS, the concentrations of DON in the 1 mol L−1 KCl extracts did not increase significantly during the 2-month incubation; however, in the OS, the concentration of DON gradually increased during the incubation period, ending up being 1.6 folds greater at 56 d than that at 7 d (Fig. 1). Between 28 and 56 d in this soil, the rate of DON increase was 0.47 times that of the rate of NO− 3 -N accumulation (Fig. 1).
Dynamics of TFAA and protein concentrations during aerobic incubation The TFAA and protein concentrations during the 56-d aerobic incubation of the three soils are shown in Fig. 2. The TFAA concentrations in the CS and TS increased significantly (P ≤ 0.05) in 28 d,
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and then gradually decreased in the following 4 weeks. TFAA changes in the OS were much less with no significant differences observed between the beginning and the end of the incubation period. The TFAA concentrations in the CS and TS were significantly higher than that in the OS by the end of the incubation period (P ≤ 0.01). The concentrations of proteins followed the order of OS > TS > CS (Fig. 2). In the OS and TS, the protein concentrations increased during the initial 21 d, and then decreased in the following 5 weeks; however, there were no significant differences between the initial and the final concentrations of proteins in either soil (P ≤ 0.05). In contrast, in the CS, the concentration of proteins gradually decreased during incubation, with the concentration at 56 d being significantly (P ≤ 0.05) decreased to 17.6% that at 0 d (Fig. 2).
Fig. 2 Concentrations of total free amino acids (a) and proteins (b) in KCl extracts of three horticultural soils incubated aerobically without leaching for 56 days. All values represent means with standard errors shown by vertical bars (n = 4). In some cases, the error bars are not shown where the symbols are larger than the error bars.
Amino acid and peptide mineralization The depletion of 14 C-labeled free amino acids and peptides from the soils is shown in Fig. 3. Amino acid and peptide half-lives are shown in Table II. The evolution of 14 CO2 from the 14 C-labeled amino acids and peptides added to the soil was rapid (Fig. 3). After 24 h, less than 1% of the amino acids and peptides applied were recovered in 0.5 mol L−1 K2 SO4 extracts (Fig. 4). The patterns of amino acid and peptide removal and mineralization were identical in all three soils (Fig. 3). The microbial biomass yields as a function of time-dependent removal of amino acids and peptides from three soils are shown Fig. 4. Microbial yields appeared to be largely independent of amino acid and peptide types in all soil samples. Partition of peptide C into biomass was higher in the CS than TS and OS, whereas no distinct pattern resulted from amino acid addition (Fig. 4). In all three soils tested, the amino acid and peptide C was predominantly used in the production of new cell microbial biomass (76% ± 3%) rather than for respiration (24% ± 2%) as CO2 after 24 h (Figs. 3 and 4). DISCUSSION DON pools Results of this study showed clearly that DON constituted an important reservoir of soluble N in the horticultural soils, which was in agreement with the studies of Murphy et al. (2000), Jones and Kielland (2002a), and Jones et al. (2004a). Although the constituents of the DON pool (Antia et al., 1991) were not fully characterized in our study due to the current experimental techniques (Jones et al., 2005), our results indicated that the LMW free amino acids represented only a minor part of the total DON recovered in KCl extracts (< 3.2%, < 17.7%, and < 15.3% of the total DON in the OS, TS, and CS, respectively). In the TFAA pool comprised of > 15 compounds, the mean concentration of individual amino acids was on average 80.9 μg N L−1 in OS, 283.7 μg N L−1 in TS, and 308.9 μg N L−1 in CS. The
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Fig. 3 Time-dependent removal of two 14 C-labeled amino acids, valine (Val) and glutamate (Glu), and two 14 C-labeled peptides, glutamate-phenylalanine (Glu-Phe) and valine-proline-proline (Val-Pro-Pro), from the solutions of three horticultural soils and their subsequent utilization by the soil microbial biomass in the production of 14 CO2 . All values represent means with standard errors shown by vertical bars (n = 3). In some cases, the error bars are not shown where the symbols are larger than the error bars. TABLE II Half-lives of removal of two 14 C-labeled amino acids, valine (Val) and glutamate (Glu), and two 14 C-labeled peptides, glutamate-phenylalanine (Glu-Phe) and valine-proline-proline (Val-Pro-Pro), from three horticultural soils Soil
Glu
Glu-Phe
Val
Val-Pro-Pro
8.2±1.5 19.7±3.5 27.3±4.7
2.3±0.2 2.8±0.1 2.5±0.3
h Organic Transitional Conventional
3.4±0.3 5.5±0.7 5.2±0.2
5.6±0.7 11.5±1.7 9.8±1.2
Fig. 4 Microbial biomass yields as a function of time-dependent removal of two 14 C-labeled amino acids, valine (Val) and glutamate (Glu), and two 14 C-labeled peptides, glutamate-phenylalanine (Glu-Phe) and valine-proline-proline (ValPro-Pro), from three horticultural soils. All values represent means with standard errors shown by vertical bars (n = 3). In some cases, the error bars are not shown where the symbols are larger than the error bars.
low amino acid concentration in the soil solution was not necessarily due to a slow rate of pool recharge but was more likely due to the extremely fast rate at which they were removed (Fig. 3). Yu et al. (2002) and Jones et al. (2004a) have also indicated two functional DON pools, a fast cycling, labile pool of LMW (such as amino acids and peptides) and a more slowly cycling, stable pool of HMW (such as polyphenolbound N, proteins, and polyamines). In marine systems, there is a large proportion of organic N that is
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similarly resistant to microbial attack and equally long lived (Peierls and Paerl, 1997; Stepanauskaset al., 2002). Plants are hypothesized to take up only the fast cycling LMW DON compounds in form of amino acids (Jones et al., 2005). Consequently, most of the soil DON (i.e., the HMW DON) may not be directly available to plants. However, due to the unavailability of suitable techniques for their analysis, the characterization, quantification, and biodegradation potential of peptides and the HMW DON in soil solution have not been intensively studied. In the present study, we observed rapid mineralization of amino acids and peptides by the soil microbial community (Figs. 3 and 4). Therefore, we suggested that a distinction existed between the HMW and LMW DON pools and that the properties of these two pools should be studied further. Methods such as use of commercially available molecular weight cutoff filters, one- and two-dimensional polyacrylamide gel electrophoresis, and size-exclusion or ion-exchange high-performance liquid chromatography techniques could all be useful for classification of DON into two distinct pools (Matsumoto and Yamagata, 2000; Choe and Gill, 2001; Jones et al., 2004b; Schulze, 2005). Plant ecological studies usually consider the total DON pool. However, if this pool is dominated by the HMW fraction, it may in fact have a lesser ecological effect in comparison to the more labile LMW pool (Jones and Kielland, 2002a; Jones et al., 2004a). The LMW DON pool may be a direct regulator of the rate of nitrification and ammonification in soil because it provides the initial substrate for these N transformation pathways. In the present study, nitrification was extremely rapid, as there was no NH+ 4 accumulation observed in the soil samples during the 56-d aerobic incubation (Fig. 1). However, the effect of HMW DON may be indirect, e.g., by nonspecific inhibition of enzymes such as proteases (Stepanauskas et al., 1999; Jones et al., 2004a). Influence of different horticultural production systems on N speciation The recent introduction of the policy by the European Union to develop more environmentally sensitive farming practices and the recognition of the importance of surplus reduction has led to a widespread interest in organic farming (Van Diepeningen et al., 2006). Most data on nutrient levels in the OS and CS have been derived from short- and long-term trial systems (Van Diepeningen et al., 2006). M¯ ader et al. (2002) reported that lower inputs of N, P, and K into organic systems than conventional systems caused variations in soil nutrient levels. Soil inorganic N levels during cropping seasons would vary with the crop, the farming system, and the amount and source of N fertilization (Poudel et al., 2002). N availability was the most significant factor limiting yield in organic systems (Clark et al., 1999). In the current study, we examined the effects of different horticultural production systems on soil N speciation, based on dynamics of nitrate, ammonium, DON, and especially amino acid and peptide mineralization (Figs. 1–4). This contrasts with previous studies, in which only physical and nutrient level parameters were considered. At the start of the aerobic incubation experiment (t = 0), DON represented the dominant N pool, being 75.1%, 93.5%, and 49.6% of the soluble N in the OS, TS, and CS, respectively (Fig. 1). However, the DON pool size varied significantly (P < 0.05) during incubation. For example, the mean DON was 61.0%, 58.1%, and only 42.5% of the soluble N in the OS, TS, and CS, respectively, after the 56-day incubation (Fig. 1). This was because more organic N was added to the OS and TS via organic fertilization. Nitrate concentration was the lowest in the OS and the highest in the CS. In contrast, ammonium did not accumulate and remained lower than nitrate throughout the incubation period. Nitrate leaching levels in the CS were similar or slightly higher than those in the OS and TS; however, in the CS that received high nitrate inputs, leaching was much higher (Stopes et al., 2002). M¯ ader et al. (2002) also reported a lower potential risk of N leaching from the OS. In soils where NO− 3 accumulates and concentrates, such as the CS in the present study (Fig. 1), DON will contribute little to the nutrient budget of the plants (Jones et al., 2004a). The total free amino acid (TFAA) content was lower but the protein content was higher in the OS than in the other two soils (Fig. 2). In the OS, N mineralization was apparently rapid, NO− 3 -N concentration was low, and thus
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losses of NO− 3 -N into groundwater by leaching were also low. The consumption of soil amino acids and peptides was rapid in all soil samples tested; significant effects of agricultural practices were observed on the rates of amino acid and peptide mineralization (Figs. 3 and 4). Mineralization rates of the added amino acids and peptides were higher in the OS than TS and CS. This may be due to the increased biological activity, microbial biomass, and enzyme activity resulting from the organic matter application that is typical of organic management practices (Jones et al., 2005). Furthermore, under organic management practices, where organic materials such as straw, animal excreta, and rapeseed cake are added as fertilizers, plants may obtain a larger proportion of N from organic sources in situations where the microbial N demand is quickly saturated, rather than from native organic matter turnover. Our results also indicated that amino acid C was predominantly used in the production of new cell biomass (76% ± 3%) rather than in respiration (24% ± 2%) (Fig. 4). This observation agrees with the findings of previous studies that showed a lower proportion of amino acid C used in respiration compared to other low molecular weight substrates; e.g., 80%–90% of organic acid C was used in respiration (Van Hees et al., 2005). Rapid amino acid and peptide turnover was observed in this study. These implied that the microbial activity in these soils was limited by the availability of labile substrates. It should be noted that this study was carried out under laboratory conditions and in the absence of plant roots, which will affect microbial N cycling (Stienstra et al., 1994). Consequently, further work is required to validate these results under field conditions. Factors limiting N turnover In this study, we suggested that the conversion of amino acid DON (LMW DON) to NH+ 4 -N and then − to NO3 -N did not limit the rate of N turnover in the three horticultural soils investigated. Therefore, the question that remains is which upstream process limits N turnover. In these production systems, organic fertilizer, dead and dying vegetation cover, roots, and plant residues are the main inputs of vegetation-derived N into the soil. In roots, proteins are the dominant HMW N compound, while free amino acids are the dominant LMW N component (Antia et al., 1991). The amounts of TFAA and NH+ 4 -N in the KCl extracts were maintained at a low level in all three soils during the entire aerobic incubation period (Figs. 1 and 2), but the amount of NO− 3 -N gradually increased. This was possibly due to the lack of microbial demand and the absence of actively growing plants in our experimental systems. Jones et al. (2004a) suggested that where C becomes depleted in a static system (i.e., C substrate becomes unavailable), heterotrophs can then no longer immobilize NH+ 4 -N. However, nitrifiers can still − -N to NO -N, which consequently accumulates. The ready accumulation of NO− oxidize NH+ 4 3 3 -N may be due to insufficient available C in our experimental setup to fuel immobilization. Amino acid and peptide uptake by soil microorganisms was very rapid (Fig. 4), which, coupled with high-affinity microbial transport systems, might be the reason why soil extract LMW N concentrations were low. Amino acids may also have been removed from solution by sorption to the soil solid phase (Jones, 1999; Gonod et al., 2006); however, the sorption potential of these soils is not high. The concentration of amino acids recovered from the soil exchange phase into 1/15 mol L−1 phosphate buffer (pH 7.0) was also almost identical at 0 and 56 d, whereas a fall in amount would be expected for support of this hypothesis (Fig. 2). As amino acid and peptide half lives were typically in the region of 30 minutes, we estimated that the free amino acid and peptide pools turned over approximately 2 500 times during incubation. Therefore, the flux through both pools was extremely fast. If we assumed an average TFAA amount of 300 μg L−1 (18 μmol L−1 ) and also assumed that all of this amino acid N eventually became − −1 under conditions NO− 3 -N, then the predicted concentration of NO3 -N after 56 d should be 450 mg L + − of instantaneous release of NH4 -N from the microbial biomass, rapid conversion of NH+ 4 -N to NO3 -N, no feedback inhibition, and no denitrification. This value was almost an order of magnitude higher than that measured in the three soils in this study (Fig. 1). The reason for the large discrepancy between the predicted and measured NO− 3 -N concentrations of this study was as yet unknown. The actual rate of amino acid turnover in the field needs to be measured;
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however, accurate measurements of N mineralization and turnover are very difficult to achieve either in laboratory or field (Jones et al., 2005). In this study, we added 14 C-labeled substrates (amino acids and peptides) to the soil solution, which can raise water content to field capacity and therefore enhance the rate of diffusion of small molecules. When the experimental system was not at steady state, e.g., in response to a pulse of substrate, we assumed that a major bottleneck was not the rate of LMW DON uptake by the microbial biomass but the subsequent release of this N back from the biomass into the soil. However, there is little information available on either the degree of retention of amino acid N and peptide N relative to amino acid C and peptide C by the soil microbial community or the rate of turnover of microbial biomass. Significant amounts of amino acid N and peptide N can be excreted, particularly during the metabolism of N-rich amino acids (e.g., glycine) (Bonkowski et al., 2000; Jones and Kielland 2002a). There are studies that have only considered the LMW DON pool and not HMW DON (e.g., proteins) (Jones et al., 2004). The degree of NH+ 4 -N excretion depends on the C-to-N ratio of the microbial population and the availability and chemical nature of available substrates (Jones et al., 2004a). We assumed that in the three soils studied, the microbial community was not N-limited and that the pools of amino acids and peptides were taken up predominantly for their C used in respiration or biosynthesis. This was supported by the apparent lack of feedback inhibition by the gradually accumulated soil NO− 3 -N and the nearly constant basal respiration of all three soils. Furthermore, in this study the high turnover rate of amino acids and peptides were mainly due to the rapid microbial mineralization and the absence of competition between the added amino acid and peptide substrates for mineralization by plants. CONCLUSIONS DON constituted a major soluble N in the horticultural soils studied. DON, TFAA, and NH+ 4-N gradually N did not accumulate and were apparently maintained at very low levels, whereas NO− 3 accumulated in these soils. The soils did not limit the rate of N mineralization. In addition, the soils studied clearly had the potential for DON mineralization and the capability to respond to large inputs of amino acids and peptides. The upstream processes including the conversion of proteins and peptides to amino acids, microbial biomass turnover, and subsequent excretion of NH+ 4 -N may be the primary factor limiting N mineralization. Nonetheless, due to the current experimental techniques, there are still many gaps in our understanding, such as the characteristics and quantification of DON in different ecosystems, the fate of DON components, their role in plant and microbial nutrition, and the degradation pathways of organic N (especially proteins) and their bottlenecks. Identification of different fractions of DON involved in N mineralization, microbial assimilation, and plant uptake and the benefits of DON added to plants and environment were still lacking. New analytical techniques were required to enable the full N cycling to be better understood. ACKNOWLEDGEMENTS We thank YIN Bo, Shanghai Jiao Tong University, China, for collecting the soil samples and J. ROBERTS, University of Wales (Bangor), UK, for helping with the sample analysis. We would also like to thank Prof. K. IWASAKI, Kochi University, Japan, and the anonymous reviewers for providing suggestions to improve this manuscript. REFERENCES Antia, N. J., Harrison, P. J. and Oliveira, L. 1991. The role of dissolved organic nitrogen in phytoplankton nutrition, cell biology and ecology. Phycologia. 30(1): 1–89. Barraclough, D. 1997. The direct or MIT route for nitrogen immobilization: A N-15 mirror image study with leucine and glycine. Soil Biol. Biochem. 29(1): 101–108. Bhogal, A., Murphy, D. V., Fortune, S., Shepherd, M. A., Hatch, D. J., Jarvis, S. C., Gaunt, J. L. and Goulding, K. W.
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Dynamics of Nitrogen Speciation in Horticultural Soils in Suburbs of Shanghai, China∗1 GE Ti-Da1,2,3 , HUANG Dan-Feng1,∗2 , P. ROBERTS2 , D. L. JONES2 and SONG Shi-Wei1 1 School
of Agriculture and Biology, Shanghai Jiao Tong University, Shanghai 201101 (China) of the Environment and Natural Resources, University of Wales, Bangor, Gwynedd LL57 2UW (UK) 3 Key Laboratory of Agro-ecological Processes in Subtropical Region, Institute of Subtropical Agriculture, Chinese Academy of Sciences, Hunan 410125 (China) 2 School
(Received September 11, 2009; revised January 5, 2010)
ABSTRACT Dissolved organic nitrogen (DON) represents a significant pool of soluble nitrogen (N) in soil ecosystems. Soil samples under three different horticultural management practices were collected from the Xiaxiyang Organic Vegetable and Fruit Farm, Shanghai, China, to investigate the dynamics of N speciation during 2 months of aerobic incubation, to compare the effects of different soils on the mineralization of 14 C-labeled amino acids and peptides, and to determine which of the pathways in the decomposition and subsequent ammonification and nitrification of organic N represented a significant blockage in soil N supply. The dynamics of N speciation was found to be significantly affected by mineralization and immobilization. DON, total free amino acids, and NH+ 4 -N were maintained at very low levels and did not accumulate, whereas NO− 3 -N gradually accumulated in these soils. The conversion of insoluble organic N to low-molecular-weight + (LMW) DON represented a main constraint to N supply, while conversions of LMW DON to NH+ 4 -N and NH4 -N to -N did not. Free amino acids and peptides were rapidly mineralized in the soils by the microbial community and NO− 3 consequently did not accumulate in soil. Turnover rates of the additional amino acids and peptides were soil-dependent and generally followed the order of organic soil > transitional soil > conventional soil. The turnover of high-molecular-weight DON was very slow and represented the major DON loss. Further studies are needed to investigate the pathways and bottlenecks of organic N degradation. Key Words:
amino acids, dissolved organic N, mineralization, N transformation, peptides
Citation: Ge, T. D., Huang, D. F., Roberts, P., Jones, D. L. and Song, S. W. 2010. Dynamics of nitrogen speciation in horticultural soils in suburbs of Shanghai, China. Pedosphere. 20(2): 261–272.
INTRODUCTION Dissolved organic nitrogen (DON) has been recognized as a potential source of nitrogen (N) for use by microorganisms and plants, and it plays an important role in N cycling in most ecosystems (Neff et al., 2003; Jones et al., 2004a; Chen et al., 2005a; Christou et al., 2006). However, plant and soil responses to DON have received much less attention compared with those to inorganic N (NO− 3 -N and + NH4 -N). This is primarily due to technical difficulties, as it is difficult to experimentally show direct plant uptake of DON or its ecological significance (Bhogal et al., 2000; Zhu and Carreiro, 2004; Jones et al., 2005; Kuzyakov and Jones, 2006). Recently, some clear experimental evidence to support the direct uptake of low-molecular-weight (LMW) DON, specifically amino acids, has been reported for plant roots and associated mycorrhizas (Streeter et al., 2000; Xu et al., 2004). Uptake of organic N species may reduce the reliance of plants on soil microorganisms for conversion of soil organic matter to inorganic N. The DON pool is generally mobile within the soil solution and can easily be transported out of ∗1 Project
supported by the National High Technology Research and Development Program (863 program) of China (No. 2006AA10A311), the National Natural Science Foundation of China (No. 40901124) and the Shanghai Leading Academic Discipline Program, China (No. B209). ∗2 Corresponding author. E-mail: [email protected].
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the soil and into groundwater by hydrological events and farming management (e.g., precipitation and + fertilization). However, compared with NO− 3 -N and NH4 -N, DON is less available and less likely to be lost by leaching or denitrification (Northup et al., 1995). DON consists of many compounds, including LMW compounds, such as amino acids, peptides, amino sugars, and urea, and high-molecular-weight (HMW) compounds, such as proteins, chlorophyll, DNA, and RNA (Antia et al., 1991). All DON forms enter the soil from a range of sources, including dry and wet deposition, throughfall, litterfall, root and microbial exudation, turnover of roots and organisms, urine and feces, and organic fertilizer additions (Kalbitz et al., 2000; Christou et al., 2006). The classical mineralization-immobilization turnover (MIT) theory suggests that the dominant pathway of soil-derived N in agricultural and many natural systems is the decomposition of insoluble organic N to NH+ 4 -N prior to microbial assimilation (Hadas et al., 1992). DON compounds that can enter into soil are free and polymeric amino acids (proteins and peptides) (Jones et al., 2005). Microbial enzymes involved in this process include oxidases, hydrolases, lyases, and deaminases (Jones et al., 2004a). These enzymes function endocellularly, in dead, autolysing cells, either free in solution or while absorbed to soil colloids. Barraclough (1997) has suggested an alternative route for assimilation of LMW DON. He suggests that the LMW DON can be used directly by the microbial cell (i.e., direct assimilation), where endogenous enzymes decompose these compounds. In this case, only the N in excess of microbial demand is released. Few studies have yet focused on the source of NH+ 4 -N from DON in agricultural soils (Murphy et al., 2000). Although soil organic matter can produce NH+ 4 -N by direct deamination, the main route for + NH4 -N production in soil appears to operate through extracellular enzymes that first convert insoluble organic N into DON (Zaman et al., 1999). If this DON pool has a LMW fraction, the transformation of proteins to amino acids, rather than amino acids to NH+ 4 -N, may be the major factor limiting N availability in low N soils (Jones and Kielland, 2002a). Subsequent turnover of this microbial population will lead to the production of NH+ 4 -N (Griffiths, 1994; Bonkowski et al., 2000). Previous studies carried out in grassland soils have indicated that the uptake of LMW DON by soil microorganisms in NO− 3 -N rich soils may primarily provide them with C to fuel respiration, rather than with N to satisfy their internal N demand (Jones et al., 2004a). However, to our knowledge, the role of DON in N cycling has not been intensively studied in Chinese horticultural soils. We hypothesize that there is a significant blockage of N decomposition pathways and subsequent ammonification and nitrification of organic N in soil N supply in different horticultural soils. There is also little information available on the dynamics of protein and microbial biomass yields as a function of time-dependent removal of amino acids and peptides. Therefore, the main aims of this study were: 1) to investigate the dynamics of N speciation and LMW DON mineralization in three soils subjected to common horticultural and agricultural practices in China; 2) to compare the effects of different soils on the mineralization of 14 C-labeled amino acids and peptides; and 3) to determine which of three possible degradation pathways, N to DON, DON to + − NH+ 4 -N, and NH4 -N to NO3 -N, may be the most significant bottleneck in N supply in horticultural soils. MATERIALS AND METHODS Soils Soil samples were collected from three sites in the Xiaxiyang Organic Vegetable and Fruit Farm, Jiading District, Shanghai, China (31◦ 4 N, 121◦ 24 E, 4 m above sea level). The climate is humid subtropical. Approximately 70% of the annual precipitation of 1 255 mm falls from May to September, the mean annual temperature is 17.5 ◦ C, and the total sunlight is 1 778 h. The sampling sites selected represented three typical horticultural production systems: The organic site (approximate 20 ha) was certified (Certification No. OF-3105/3106-931-243-2005) by the Organic Farming Development Center
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of China (OFDC) in June 2004 after a 2 year conversion period; the transitional site (about 27 ha) was certified in June 2003 (Certification No. V-3106-931-264-2003); and the conventional site was situated in a greenhouse close to the organic and transitional plots. The land is used to grow vegetable crops, typically in plastic tunnels. At each site, leaf and vegetable litters were removed from the surface, soil samples of 0–20 and 20–60 cm depths were randomly taken from an area of 6 m × 60 m in early winter (October 2006) using a 5 cm diameter stainless steel corer. The soil samples were stored field-moist at 4 ◦ C with dry ice (solid CO2 ) in CO2 permeable polypropylene bags until analysis. Selected properties of the soils from the organic, transitional, and conventional systems, i.e., organic soil (OS), transitional soil (TS), and conventional soil (CS), are shown in Table I. TABLE I Selected properties of the three horticultural soils Soil
Dominant vegetation
Organic Transitional
Tomato and strawberry Cauliflower
Conventional
Cantaloupe
a) Mean±standard
Depth
Moisture content
Ash content
cm 0–20 20–60 0–20 20–60 0–20 20–60
g kg−1 198.8±9.5a) 51.3±1.1 213.2±5.5 30.9±0.8 171.9±7.5 42.5±0.9 224.9±3.3 37.2±1.9 222.8±2.2 31.8±0.2 252.3±15.1 26.6±1.7
pHH2 O
7.2±0.1 7.9±0.1 7.4±0.1 8.0±0.1 7.7±0.1 7.7±0.1
Electrical conductivity
Respiration
dS m−1 0.26±0.01 0.13±0.00 0.14±0.01 0.08±0.01 0.17±0.01 0.14±0.00
μmol CO2 kg−1 h−1 269.4±12.5 225.9±10.9 248.4±15.9 207.9±20.7 209.4±19.5 195.3±16.7
error (n = 3).
Aerobic incubation and preparation of soil extracts Aerobic incubation of soils was as described in the study by Zhong and Makeschin (2003). In brief, the field-moist soil samples were broken by hand into aggregates 0.5–5.0 cm in diameter, and large stones and roots (> 1 cm in diameter) were removed. The samples 250 g each in 4 replicates were then aerobically incubated in 500-mL plastic cups covered with Parafilm punched with 6 holes for aeration and placed in a dark, climate-controlled chamber maintained at 20 ± 1 ◦ C. Soil moisture content was maintained at field capacity throughout the experiment by randomly selecting and weighing five plastic cups each day. If the water loss from these cups exceeded 5 g, all cups were weighed and brought back to the original moisture content. This was necessary two or three times a week. Water additions were made directly to the soil, using a 5-mL pipette to ensure that a uniform distribution of water was obtained. After 0, 7, 14, 21, 28, 42, and 56 d of incubation, the soil samples were destructively harvested and extracted. Soil extracts were prepared using the procedure described by Jones and Willett (2006). Briefly, 5.0 g incubated soil was shaken with 25 mL of either 1 mol L−1 KCl or 1/15 mol L−1 phosphate buffer (pH 7.0) consisting of 14.6 g L−1 Na2 HPO4 ·12H2 O and 3.5 g L−1 KH2 PO4 in 50 cm3 polypropylene tubes on a reciprocating shaker (Unimax 2010, Heidolph Elekrto Gmbh, Kelheim, Germany) at a speed of 250 r min−1 for 30 min. After shaking, the soil extracts were centrifuged at 8 000 g for 15 min, and the supernatant was recovered and stored in polypropylene bottles at −20 ◦ C until analysis. The dynamics of N speciation in 1 mol L−1 KCl extracts were very similar to those in 1/15 mol L−1 phosphate buffer extracts; therefore, only the results from 1 mol L−1 KCl extracts are presented for this study. Chemical analysis Soil pH and electrical conductivity were determined in 1:1 (v/v) soil:H2 O extracts using standard electrodes. Soil moisture and ash contents were determined gravimetrically by drying at 105 ◦ C and combustion at 450 ◦ C for 24 h, respectively. Total C and total N were determined with a CHN-2000
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analyzer (Leco Corporation, St Joseph, USA). Soil respiration was measured at 20 ◦ C using a CIRASIRGA soil respirometer (PP Systems, Hitchin, UK). Total dissolved N (TDN) in soil extracts was measured with a Shimadzu TOC-TN analyzer (Shimadzu Corp, Kyoto, Japan) as described by Jones et al. (2004a). The 1/15 mol L−1 phosphate buffer extracts were analyzed without dilution, but the 1 mol L−1 KCl extracts were diluted 10-fold with de-ionized water before measurement, to prevent a large amount of salts from precipitating onto the surface of the Pt/Al2 O3 catalysts and reducing catalyst effi+ ciency (Chen et al., 2005b). NO− 3 and NH4 were determined using the methods described in Mulvaney (1996) with a Versamax microplate reader (Molecular Devices Corp, Sunnyvale, USA). DON was calcu+ lated as the difference between TDN and the sum of NO− 3 -N and NH4 -N. Total free amino acid (TFAA) concentrations were determined on a Varian Cary Eclipse 96-well plate fluorescence spectrophotometer (Varian Inc., Palo Alto, USA) using the o-phthaldialdehyde/β-mercaptoethanol procedure of Jones et al. (2002b). Proteins in the soil extracts were determined by the Bradford method (Bradford, 1976) using bovine serum albumin as a standard. Mineralization of amino acids and peptides To assess the amino acid and peptide mineralization rates in the soils, mixtures of two uniformly labeled amino acids, 14 C-labeled valine (Val) and glutamate (Glu), and two labeled peptides, 14 Clabeled glutamate-phenylalanine (Glu-Phe) and valine-proline-proline (Val-Pro-Pro), from the St Louis Arc, USA, were added to the soil samples and their subsequent evolution to 14 CO2 was measured over 24 h, as described in Jones (1999). Soil (5 g) was weighed into 50 mL polypropylene tubes, 0.5 mL 9 990 MBq mmol−1 14 C-labeled Val, Glu, Glu-Phe, or Val-Pro-Pro solution was added dropwise to a final concentration of 0.5 mmol L−1 (2.7 kBq kg−1 ). To trap any evolved 14 CO2 , a polypropylene vial containing 1 mL of 1 mol L−1 NaOH was placed above the soil in the tubes, and then the tubes were hermetically sealed and maintained at 15 ◦ C. Any 14 C-labeled amino acids or peptides remaining in the soil at 0.5, 1, 3, 6, 12, and 24 h after substrate addition were removed by shaking in 10 mL of 0.5 mol L−1 ice-cold K2 SO4 for 10 min at 250 r min−1 , followed by recovery of the supernatant solution by centrifugation (18 000 g, 5 min) for scintillation counting. All radioactivity was determined by liquid scintillation counting using a Wallac 1414 counter and the aqueous compatible scintillation fluid Wallac Optiphase Hisafe 3 (EG & G Ltd., Milton Keynes, UK). The amount of 14 C immobilized in the microbial biomass (microbial biomass carbon) was calculated as the difference: 100 − (14 CO2 in NaOH trap + 14 C remaining in 0.5 mol L−1 K2 SO4 extracts). The free amino acid and peptide concentrations added to all soil samples were 8.4 mg N L−1 (0.5 mmol L−1 ). We chose these compounds as they were commercially available and they represented common peptides likely to occur in the soil environment, based on known protein structures and sequence data. The concentration of 0.5 mmol L−1 was used in accordance with previous studies which have indicated that this is the likely concentration of an individual amino acid in a typical plant cell and the likely concentration that might appear in the soil if a root cell was to burst (Jones et al., 2005). The concentration of 14 C-labeled amino acids remaining in the soil extracts was calculated by the formula: C = 14 C-AAsoil × V × Cs × 14 × 1.2/14 C-AAtotal where C is the concentration of 14 C-labeled amino acids remaining in the extracts (mg N L−1 ), 14 CAAsoil is the 14 C decayed per minute at a known time, V is the K2 SO4 volume added to the soil (mL), Cs is the substrate initial concentration (mmol L−1 ), 14 C-AAtotal is the 14 C decayed per minute in the substrate in the beginning, 14 is the atomic weight of N, and 1.2 is the transfer coefficient. Statistical and data analysis All data were expressed as the mean of four replicates ± standard error. One-way analyses of variance (ANOVA) with Tukey test were used to identify treatment differences. Analyses were run using SPSS
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14.0 for Windows (SPSS Inc., Chicago, Illinois, USA). To estimate the amino acid and peptide half-lives in soil, a first-order exponential decay equation was fitted to the amino acid and peptide mineralization data, using Sigmaplot 8.01, by a least-square optimization routine (Jones et al., 2004a). RESULTS Soil characteristics In all three soils, the levels of ash, electrical conductivity, and respiration at the depth of 0–20 cm were higher than those at the depth of 20–60 cm, whereas the moisture content at 0–20 cm was lower than that at 20–60 cm (Table I). Soil ash and respiration were the greatest in the OS and least in the CS. The EC value was the highest in the OS and least in the TS. No significant difference in pH was observed among the three horticultural soils (P > 0.05). Dynamics of inorganic N and DON mineralization during aerobic incubation The concentration of NO− 3 -N gradually increased in all three soils during the 56-d incubation. The increase in NO− 3 -N concentration was almost linear with incubation time in all soils (Fig. 1). The rates −1 −1 d ) > TS (0.92 mg N kg−1 d−1 ) > CS of NO− 3 -N increase followed the order of OS (0.96 mg N kg −1 −1 (1.07 mg N kg d ) although the amounts were in the contrary order (Fig. 1). The amount of NO− 3 -N in the soil extracts increased by 7.5 folds in the OS, 4.8 folds in the TS, and 3.5 folds in the CS by the end of the incubation.
+ −1 KCl extracts of three horticultural soils incubated Fig. 1 Dissolved organic N (DON), NO− 3 -N, and NH4 -N in 1 mol L aerobically without leaching for 56 days. All values represent means with standard errors shown by vertical bars (n = 4). In some cases, the error bars are not shown where the symbols are larger than the error bars.
−1 The amounts of NH+ in the OS, 1.9 and 4 -N in the extracts varied between 1.7 and 3.8 mg N kg −1 −1 3.7 mg N kg in the TS, and 2.8 and 5.0 mg N kg in the CS, but no overall change was observed in either soil throughout the incubation (Fig. 1). At the final sampling (56 d), the amount of inorganic N + −1 in the OS, 69.4 mg N kg−1 in the TS, and 88.0 mg N kg−1 (NO− 3 -N and NH4 -N) was 66.1 mg N kg in the CS. In the TS and CS, the concentrations of DON in the 1 mol L−1 KCl extracts did not increase significantly during the 2-month incubation; however, in the OS, the concentration of DON gradually increased during the incubation period, ending up being 1.6 folds greater at 56 d than that at 7 d (Fig. 1). Between 28 and 56 d in this soil, the rate of DON increase was 0.47 times that of the rate of NO− 3 -N accumulation (Fig. 1).
Dynamics of TFAA and protein concentrations during aerobic incubation The TFAA and protein concentrations during the 56-d aerobic incubation of the three soils are shown in Fig. 2. The TFAA concentrations in the CS and TS increased significantly (P ≤ 0.05) in 28 d,
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and then gradually decreased in the following 4 weeks. TFAA changes in the OS were much less with no significant differences observed between the beginning and the end of the incubation period. The TFAA concentrations in the CS and TS were significantly higher than that in the OS by the end of the incubation period (P ≤ 0.01). The concentrations of proteins followed the order of OS > TS > CS (Fig. 2). In the OS and TS, the protein concentrations increased during the initial 21 d, and then decreased in the following 5 weeks; however, there were no significant differences between the initial and the final concentrations of proteins in either soil (P ≤ 0.05). In contrast, in the CS, the concentration of proteins gradually decreased during incubation, with the concentration at 56 d being significantly (P ≤ 0.05) decreased to 17.6% that at 0 d (Fig. 2).
Fig. 2 Concentrations of total free amino acids (a) and proteins (b) in KCl extracts of three horticultural soils incubated aerobically without leaching for 56 days. All values represent means with standard errors shown by vertical bars (n = 4). In some cases, the error bars are not shown where the symbols are larger than the error bars.
Amino acid and peptide mineralization The depletion of 14 C-labeled free amino acids and peptides from the soils is shown in Fig. 3. Amino acid and peptide half-lives are shown in Table II. The evolution of 14 CO2 from the 14 C-labeled amino acids and peptides added to the soil was rapid (Fig. 3). After 24 h, less than 1% of the amino acids and peptides applied were recovered in 0.5 mol L−1 K2 SO4 extracts (Fig. 4). The patterns of amino acid and peptide removal and mineralization were identical in all three soils (Fig. 3). The microbial biomass yields as a function of time-dependent removal of amino acids and peptides from three soils are shown Fig. 4. Microbial yields appeared to be largely independent of amino acid and peptide types in all soil samples. Partition of peptide C into biomass was higher in the CS than TS and OS, whereas no distinct pattern resulted from amino acid addition (Fig. 4). In all three soils tested, the amino acid and peptide C was predominantly used in the production of new cell microbial biomass (76% ± 3%) rather than for respiration (24% ± 2%) as CO2 after 24 h (Figs. 3 and 4). DISCUSSION DON pools Results of this study showed clearly that DON constituted an important reservoir of soluble N in the horticultural soils, which was in agreement with the studies of Murphy et al. (2000), Jones and Kielland (2002a), and Jones et al. (2004a). Although the constituents of the DON pool (Antia et al., 1991) were not fully characterized in our study due to the current experimental techniques (Jones et al., 2005), our results indicated that the LMW free amino acids represented only a minor part of the total DON recovered in KCl extracts (< 3.2%, < 17.7%, and < 15.3% of the total DON in the OS, TS, and CS, respectively). In the TFAA pool comprised of > 15 compounds, the mean concentration of individual amino acids was on average 80.9 μg N L−1 in OS, 283.7 μg N L−1 in TS, and 308.9 μg N L−1 in CS. The
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Fig. 3 Time-dependent removal of two 14 C-labeled amino acids, valine (Val) and glutamate (Glu), and two 14 C-labeled peptides, glutamate-phenylalanine (Glu-Phe) and valine-proline-proline (Val-Pro-Pro), from the solutions of three horticultural soils and their subsequent utilization by the soil microbial biomass in the production of 14 CO2 . All values represent means with standard errors shown by vertical bars (n = 3). In some cases, the error bars are not shown where the symbols are larger than the error bars. TABLE II Half-lives of removal of two 14 C-labeled amino acids, valine (Val) and glutamate (Glu), and two 14 C-labeled peptides, glutamate-phenylalanine (Glu-Phe) and valine-proline-proline (Val-Pro-Pro), from three horticultural soils Soil
Glu
Glu-Phe
Val
Val-Pro-Pro
8.2±1.5 19.7±3.5 27.3±4.7
2.3±0.2 2.8±0.1 2.5±0.3
h Organic Transitional Conventional
3.4±0.3 5.5±0.7 5.2±0.2
5.6±0.7 11.5±1.7 9.8±1.2
Fig. 4 Microbial biomass yields as a function of time-dependent removal of two 14 C-labeled amino acids, valine (Val) and glutamate (Glu), and two 14 C-labeled peptides, glutamate-phenylalanine (Glu-Phe) and valine-proline-proline (ValPro-Pro), from three horticultural soils. All values represent means with standard errors shown by vertical bars (n = 3). In some cases, the error bars are not shown where the symbols are larger than the error bars.
low amino acid concentration in the soil solution was not necessarily due to a slow rate of pool recharge but was more likely due to the extremely fast rate at which they were removed (Fig. 3). Yu et al. (2002) and Jones et al. (2004a) have also indicated two functional DON pools, a fast cycling, labile pool of LMW (such as amino acids and peptides) and a more slowly cycling, stable pool of HMW (such as polyphenolbound N, proteins, and polyamines). In marine systems, there is a large proportion of organic N that is
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similarly resistant to microbial attack and equally long lived (Peierls and Paerl, 1997; Stepanauskaset al., 2002). Plants are hypothesized to take up only the fast cycling LMW DON compounds in form of amino acids (Jones et al., 2005). Consequently, most of the soil DON (i.e., the HMW DON) may not be directly available to plants. However, due to the unavailability of suitable techniques for their analysis, the characterization, quantification, and biodegradation potential of peptides and the HMW DON in soil solution have not been intensively studied. In the present study, we observed rapid mineralization of amino acids and peptides by the soil microbial community (Figs. 3 and 4). Therefore, we suggested that a distinction existed between the HMW and LMW DON pools and that the properties of these two pools should be studied further. Methods such as use of commercially available molecular weight cutoff filters, one- and two-dimensional polyacrylamide gel electrophoresis, and size-exclusion or ion-exchange high-performance liquid chromatography techniques could all be useful for classification of DON into two distinct pools (Matsumoto and Yamagata, 2000; Choe and Gill, 2001; Jones et al., 2004b; Schulze, 2005). Plant ecological studies usually consider the total DON pool. However, if this pool is dominated by the HMW fraction, it may in fact have a lesser ecological effect in comparison to the more labile LMW pool (Jones and Kielland, 2002a; Jones et al., 2004a). The LMW DON pool may be a direct regulator of the rate of nitrification and ammonification in soil because it provides the initial substrate for these N transformation pathways. In the present study, nitrification was extremely rapid, as there was no NH+ 4 accumulation observed in the soil samples during the 56-d aerobic incubation (Fig. 1). However, the effect of HMW DON may be indirect, e.g., by nonspecific inhibition of enzymes such as proteases (Stepanauskas et al., 1999; Jones et al., 2004a). Influence of different horticultural production systems on N speciation The recent introduction of the policy by the European Union to develop more environmentally sensitive farming practices and the recognition of the importance of surplus reduction has led to a widespread interest in organic farming (Van Diepeningen et al., 2006). Most data on nutrient levels in the OS and CS have been derived from short- and long-term trial systems (Van Diepeningen et al., 2006). M¯ ader et al. (2002) reported that lower inputs of N, P, and K into organic systems than conventional systems caused variations in soil nutrient levels. Soil inorganic N levels during cropping seasons would vary with the crop, the farming system, and the amount and source of N fertilization (Poudel et al., 2002). N availability was the most significant factor limiting yield in organic systems (Clark et al., 1999). In the current study, we examined the effects of different horticultural production systems on soil N speciation, based on dynamics of nitrate, ammonium, DON, and especially amino acid and peptide mineralization (Figs. 1–4). This contrasts with previous studies, in which only physical and nutrient level parameters were considered. At the start of the aerobic incubation experiment (t = 0), DON represented the dominant N pool, being 75.1%, 93.5%, and 49.6% of the soluble N in the OS, TS, and CS, respectively (Fig. 1). However, the DON pool size varied significantly (P < 0.05) during incubation. For example, the mean DON was 61.0%, 58.1%, and only 42.5% of the soluble N in the OS, TS, and CS, respectively, after the 56-day incubation (Fig. 1). This was because more organic N was added to the OS and TS via organic fertilization. Nitrate concentration was the lowest in the OS and the highest in the CS. In contrast, ammonium did not accumulate and remained lower than nitrate throughout the incubation period. Nitrate leaching levels in the CS were similar or slightly higher than those in the OS and TS; however, in the CS that received high nitrate inputs, leaching was much higher (Stopes et al., 2002). M¯ ader et al. (2002) also reported a lower potential risk of N leaching from the OS. In soils where NO− 3 accumulates and concentrates, such as the CS in the present study (Fig. 1), DON will contribute little to the nutrient budget of the plants (Jones et al., 2004a). The total free amino acid (TFAA) content was lower but the protein content was higher in the OS than in the other two soils (Fig. 2). In the OS, N mineralization was apparently rapid, NO− 3 -N concentration was low, and thus
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losses of NO− 3 -N into groundwater by leaching were also low. The consumption of soil amino acids and peptides was rapid in all soil samples tested; significant effects of agricultural practices were observed on the rates of amino acid and peptide mineralization (Figs. 3 and 4). Mineralization rates of the added amino acids and peptides were higher in the OS than TS and CS. This may be due to the increased biological activity, microbial biomass, and enzyme activity resulting from the organic matter application that is typical of organic management practices (Jones et al., 2005). Furthermore, under organic management practices, where organic materials such as straw, animal excreta, and rapeseed cake are added as fertilizers, plants may obtain a larger proportion of N from organic sources in situations where the microbial N demand is quickly saturated, rather than from native organic matter turnover. Our results also indicated that amino acid C was predominantly used in the production of new cell biomass (76% ± 3%) rather than in respiration (24% ± 2%) (Fig. 4). This observation agrees with the findings of previous studies that showed a lower proportion of amino acid C used in respiration compared to other low molecular weight substrates; e.g., 80%–90% of organic acid C was used in respiration (Van Hees et al., 2005). Rapid amino acid and peptide turnover was observed in this study. These implied that the microbial activity in these soils was limited by the availability of labile substrates. It should be noted that this study was carried out under laboratory conditions and in the absence of plant roots, which will affect microbial N cycling (Stienstra et al., 1994). Consequently, further work is required to validate these results under field conditions. Factors limiting N turnover In this study, we suggested that the conversion of amino acid DON (LMW DON) to NH+ 4 -N and then − to NO3 -N did not limit the rate of N turnover in the three horticultural soils investigated. Therefore, the question that remains is which upstream process limits N turnover. In these production systems, organic fertilizer, dead and dying vegetation cover, roots, and plant residues are the main inputs of vegetation-derived N into the soil. In roots, proteins are the dominant HMW N compound, while free amino acids are the dominant LMW N component (Antia et al., 1991). The amounts of TFAA and NH+ 4 -N in the KCl extracts were maintained at a low level in all three soils during the entire aerobic incubation period (Figs. 1 and 2), but the amount of NO− 3 -N gradually increased. This was possibly due to the lack of microbial demand and the absence of actively growing plants in our experimental systems. Jones et al. (2004a) suggested that where C becomes depleted in a static system (i.e., C substrate becomes unavailable), heterotrophs can then no longer immobilize NH+ 4 -N. However, nitrifiers can still − -N to NO -N, which consequently accumulates. The ready accumulation of NO− oxidize NH+ 4 3 3 -N may be due to insufficient available C in our experimental setup to fuel immobilization. Amino acid and peptide uptake by soil microorganisms was very rapid (Fig. 4), which, coupled with high-affinity microbial transport systems, might be the reason why soil extract LMW N concentrations were low. Amino acids may also have been removed from solution by sorption to the soil solid phase (Jones, 1999; Gonod et al., 2006); however, the sorption potential of these soils is not high. The concentration of amino acids recovered from the soil exchange phase into 1/15 mol L−1 phosphate buffer (pH 7.0) was also almost identical at 0 and 56 d, whereas a fall in amount would be expected for support of this hypothesis (Fig. 2). As amino acid and peptide half lives were typically in the region of 30 minutes, we estimated that the free amino acid and peptide pools turned over approximately 2 500 times during incubation. Therefore, the flux through both pools was extremely fast. If we assumed an average TFAA amount of 300 μg L−1 (18 μmol L−1 ) and also assumed that all of this amino acid N eventually became − −1 under conditions NO− 3 -N, then the predicted concentration of NO3 -N after 56 d should be 450 mg L + − of instantaneous release of NH4 -N from the microbial biomass, rapid conversion of NH+ 4 -N to NO3 -N, no feedback inhibition, and no denitrification. This value was almost an order of magnitude higher than that measured in the three soils in this study (Fig. 1). The reason for the large discrepancy between the predicted and measured NO− 3 -N concentrations of this study was as yet unknown. The actual rate of amino acid turnover in the field needs to be measured;
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however, accurate measurements of N mineralization and turnover are very difficult to achieve either in laboratory or field (Jones et al., 2005). In this study, we added 14 C-labeled substrates (amino acids and peptides) to the soil solution, which can raise water content to field capacity and therefore enhance the rate of diffusion of small molecules. When the experimental system was not at steady state, e.g., in response to a pulse of substrate, we assumed that a major bottleneck was not the rate of LMW DON uptake by the microbial biomass but the subsequent release of this N back from the biomass into the soil. However, there is little information available on either the degree of retention of amino acid N and peptide N relative to amino acid C and peptide C by the soil microbial community or the rate of turnover of microbial biomass. Significant amounts of amino acid N and peptide N can be excreted, particularly during the metabolism of N-rich amino acids (e.g., glycine) (Bonkowski et al., 2000; Jones and Kielland 2002a). There are studies that have only considered the LMW DON pool and not HMW DON (e.g., proteins) (Jones et al., 2004). The degree of NH+ 4 -N excretion depends on the C-to-N ratio of the microbial population and the availability and chemical nature of available substrates (Jones et al., 2004a). We assumed that in the three soils studied, the microbial community was not N-limited and that the pools of amino acids and peptides were taken up predominantly for their C used in respiration or biosynthesis. This was supported by the apparent lack of feedback inhibition by the gradually accumulated soil NO− 3 -N and the nearly constant basal respiration of all three soils. Furthermore, in this study the high turnover rate of amino acids and peptides were mainly due to the rapid microbial mineralization and the absence of competition between the added amino acid and peptide substrates for mineralization by plants. CONCLUSIONS DON constituted a major soluble N in the horticultural soils studied. DON, TFAA, and NH+ 4-N gradually N did not accumulate and were apparently maintained at very low levels, whereas NO− 3 accumulated in these soils. The soils did not limit the rate of N mineralization. In addition, the soils studied clearly had the potential for DON mineralization and the capability to respond to large inputs of amino acids and peptides. The upstream processes including the conversion of proteins and peptides to amino acids, microbial biomass turnover, and subsequent excretion of NH+ 4 -N may be the primary factor limiting N mineralization. Nonetheless, due to the current experimental techniques, there are still many gaps in our understanding, such as the characteristics and quantification of DON in different ecosystems, the fate of DON components, their role in plant and microbial nutrition, and the degradation pathways of organic N (especially proteins) and their bottlenecks. Identification of different fractions of DON involved in N mineralization, microbial assimilation, and plant uptake and the benefits of DON added to plants and environment were still lacking. New analytical techniques were required to enable the full N cycling to be better understood. ACKNOWLEDGEMENTS We thank YIN Bo, Shanghai Jiao Tong University, China, for collecting the soil samples and J. ROBERTS, University of Wales (Bangor), UK, for helping with the sample analysis. We would also like to thank Prof. K. IWASAKI, Kochi University, Japan, and the anonymous reviewers for providing suggestions to improve this manuscript. REFERENCES Antia, N. J., Harrison, P. J. and Oliveira, L. 1991. The role of dissolved organic nitrogen in phytoplankton nutrition, cell biology and ecology. Phycologia. 30(1): 1–89. Barraclough, D. 1997. The direct or MIT route for nitrogen immobilization: A N-15 mirror image study with leucine and glycine. Soil Biol. Biochem. 29(1): 101–108. Bhogal, A., Murphy, D. V., Fortune, S., Shepherd, M. A., Hatch, D. J., Jarvis, S. C., Gaunt, J. L. and Goulding, K. W.
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