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Published online 21 June 2006
Published in Soil Sci Soc Am J 70:1338-1348 (2006)
DOI: 10.2136/sssaj2005.0190
© 2006 Soil Science Society of America
677 S. Segoe Rd., Madison, WI 53711 USA
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Soil Biology & Biochemistry

Gross Nitrogen Transformations in an Agricultural Soil after Repeated Dairy-Waste Application

Mussie Y. Habteselassiea, John M. Starkb, Bruce E. Millerc, Seth G. Thackera and Jeanette M. Nortona,*

a Dep. of Plants, Soils, and Biometeorology
b Dep. of Biology
c Dep. of Agricultural Systems Technology and Education, Utah State Univ., Logan, UT 84322

* Corresponding author (jennyn{at}cc.usu.edu)


   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 REFERENCES
 
Measurements of gross N transformation rates are important to properly understand N cycling processes in agricultural soils where both productive and consumptive processes occur. The objective of the study was to determine the effect of repeated application of dairy-waste compost (DC), liquid dairy-waste (LW), and ammonium sulfate (AS) on gross N mineralization and nitrification rates and N supplying potential of an agricultural soil. Our goal was to examine both production and consumption of inorganic N for their effects on the balance between N supply from treated dairy-wastes and plant N demand. Treatments were applied at rates approximately equivalent to 100 and 200 kg available N ha–1 for 6 yr annually. Field-based N15 pool dilution techniques and laboratory incubation experiments were employed to measure gross rates and mineralization potential of the soil. Both levels of DC raised the labile organic N pool significantly but only the high level DC significantly increased the decomposition rate constant (k). The mean gross N mineralization rates for 1999 to 2002 for the high levels of DC, LW, and AS were 5.72, 2.89, and 1.27 mg N kg–1 d–1, whereas gross nitrification rates were 10.24, 1.57, and 0.74 mg N kg–1 d–1, respectively. Net mineralization rates were <35% of gross rates while nitrate consumption was not significant under any treatment. Variability in gross rates was high in the soils receiving DC, which could be due to presence of hotspots of labile organic matter. Elevated gross N transformation rates in plots receiving DC indicate the dynamic nature of this agricultural soil after repeated applications of dairy-waste.

Abbreviations: AS, ammonium sulfate • DC, dairy waste compost • LW, liquid dairy waste


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 REFERENCES
 
A LARGE AMOUNT of animal waste is produced in the USA every year and the bulk of it is applied on agricultural lands (Gollehon and Caswell, 2000). It is well documented that animal waste can serve as an important source of organic matter and major nutrient elements such as N, P, and K (Van Faassen, and Van Dijk, 1987; Schlegel, 1992; Cameron et al., 1997; Stratton and Rechcigl, 1998). Animal waste is the main source of plant nutrients in organic farming systems in which synthetic fertilizers are not used (Mader et al., 2002; Wood et al., 2002; Pimentel et al., 2005). With increasing prices of commercial fertilizers and high demand for organic products (Olson, 2001; Jawson and Bull, 2002; Wood et al., 2002), the use of animal waste as a source of plant nutrients is likely to increase.

Nitrogen is of particular interest because it is usually the most limiting nutrient for plant growth and yet excess NO3––N is an environment concern because it readily moves to ground and surface waters (Vitousek and Howarth, 1991; Schlesinger, 1997; Pierzynski et al., 2000). About three fourths of U.S. counties contained farms that produced manure N in excess of their crop or pasture N requirements between 1982 and 1997, which is likely to further increase in the future (Gollehon and Caswell, 2000). One option would be to transport the excess manure off farm to nutrient deficient sites. This option is often unattractive because of the transportation cost and regulations (Ribaudo et al., 2003). The second option would be to adopt manure management strategies that minimize the negative environmental impact. Two possible waste management strategies are composting and lagoon treatment. Composting, for example, has been widely advocated as a method of stabilizing manure N and reducing leaching potential (Gagnon et al., 1998; Stratton and Rechcigl, 1998). Environmentally sound application of treated waste on soils requires understanding of two key processes, mineralization and nitrification; that control the release and availability of waste N.

Mineralization of organic N results in the formation of NH4+–N, which then is converted to nitrate by the process of nitrification. Assimilation of NH4+–N costs less metabolic energy than assimilatory reduction of NO3––N for plants and microbes (Recous et al., 1990; Schlesinger, 1997). Microbial consumption of NO3––N can be significant under high C availability (Johnson, 1992; Stark and Hart, 1997; Burger and Jackson, 2003) and is strongly influenced by microbial substrate use efficiency (Saetre and Stark, 2005). Carbon limitation has been noted to decrease overall microbial activity and microbial growth efficiency due to higher C allocation to maintenance and decreased N immobilization (Chen and Stark, 2000).

Measuring the net change in N concentration is the most widely practiced method of assessing N transformations. This method is very valuable from the plant nutrition point of view but it does not tell us if both productive and consumptive processes are taking place simultaneously, which makes it inadequate for understanding the actual status of N cycling in soils (Hart et al., 1994a). In the design of management options for controlling excess nitrate in soils we need to consider both decreasing production or increasing consumption of inorganic N and the potential interactions on N availability to plants. To resolve these simultaneous processes, it is necessary to measure gross rates of N transformations using isotope techniques (Stark, 2000; Murphy et al., 2003).

Repeated applications of dairy manure (Peacock et al., 2001), pig slurry (Hastings et al., 1997), and wastewater effluent (Oved et al., 2001) have been observed to affect microbial communities that mediate N transformations in soils. Significant increases in gross N mineralization and nitrification rates were reported as a result of addition of sheep manure (Sorensen, 2001), dairy shed effluent (Zaman et al., 1999), and cattle slurry (Muller et al., 2003). The nature and magnitude of the effect was dependent on a number of factors, including the time between application and measurement, the type of the waste, the number of applications, and local environmental and edaphic factors that make extrapolation of results from one area to another difficult.

The investigation of the interaction between gross N transformations and treated waste application was needed in our study area due to some of its unique characteristics. These distinguishing characteristics included a strongly calcareous soil with high pH and buffering capacity, a hot and dry summer in which the soil is irrigated, and a cold winter with variable humidity. These conditions affect both ammonium availability and nitrate mobility in the soil environment.

The main objective of this study was, therefore, to determine the effect of repeated application of DC, LW, and AS fertilizer on gross N transformation rates and N supplying potential of an agricultural soil under these conditions. The hypotheses were: (i) DC, LW, and AS will affect N mineralization, consumption, and nitrification rates differently due to differences in the amount, form and timing of N released and C additions, and (ii) gross N transformation rates will be strongly influenced by the labile and total organic C and N content of the soil resulting from the repeated application of the different N sources. The effects of repeated application of treated dairy-waste on soil nutrient pools, nitrification potentials, N availability to plants, and silage corn yields were also examined over the same period and these results are described in the companion paper (Habteselassie et al., 2006).


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS AND DISCUSSION
 REFERENCES
 
Experimental Plots
Twenty-eight experimental plots were established in 1997 at the Greenville Research farm, which is owned by Utah State University (USU) and located in North Logan, UT (41°46' N, 111°49' W). The plots were arranged in completely randomized block design with four replicate blocks. The soil is an irrigated, very strongly calcareous Millville silt loam (coarse-silty, carbonatic, mesic Typic Haploxeroll) with pH1:1 of 8.4 and cation exchange capacity 14 cmolc kg–1. Further details on field plots and properties of the soil are described in Shi et al. (2004) and Habteselassie et al. (2006). Liquid dairy waste was collected from a dairy farm in Smithfield, UT. Compost was prepared from solid dairy manure and straw bedding in the Animal Science Farm of the Utah Agricultural Experiment Station. Chemical properties of the waste and the process of composting have been documented elsewhere (Shi et al., 1999, 2004).

Treated dairy-waste and AS (see below) were applied to the experimental plots every year in early May and silage corn (variety DK-626) is planted around the end of May. The treatments included AS at 100 kg N ha–1 (AS100), AS at 200 kg N ha–1 (AS200), low level DC (DC100), high level DC (DC200), low level LW (LW100), and high level LW (LW200). The low and high level treatments were designed to provide approximately 100 and 200 kg ha–1 of available N, taking into account actual N contents determined before application, contributions from mineralization, and credits from previous year applications. Initially, we assumed 10% of the applied compost N would become available during the growing season, and 5% in the following season while 100% of the N contained in the LW was considered as available in the current year. These assumptions are based on previous laboratory incubation experiments (Shi et al., 1999; Shi, 1998). During 2000 through 2002, the 10% mineralization estimate was considered to be an underestimate of actual N release and application rates were lowered to decrease excess nitrate accumulations. Actual application rates and characteristics of the materials applied are summarized in the companion paper (see Table 2 in Habteselassie et al., 2006). The fertilizer treatments are applied every year in early May and incorporated into the top 15 cm of the soil with a small tractor disc rotor tiller. Incorporation of amendments followed application as soon as possible depending on soil moisture conditions. The DC and AS were usually delivered in one application whereas the LW was applied in two to three applications over several days to allow for infiltration. Potassium and P were applied as recommended based on soil test results.

Soil and Nitrogen Transformation Methods
The methods used for analyzing soil samples have been described in Habteselassie et al. (2006).

Field Study
Gross rates of N-mineralization and nitrification were measured in the field employing 15N isotope dilution technique as previously described (Hart et al., 1994b; Stark, 2000; Shi et al., 2004) during the mid-season period (in August, approximately 90 d after treatment application) when the N demand by corn is high. Briefly, the 15N isotope dilution technique involved pounding two pairs of concentric PVC cylinder cores (5 x 20 cm and 10 x 20 cm) into the ground between the two central rows of corn. Soil samples between the inner and outer cores were immediately harvested, mixed in a polythene bag, and a subsample of approximately 15 g was added to 75 mL preweighed cold 2 M KCl to determine the initial NH4+ and NO3– concentrations. Two of the inner cores received 15NH4Cl solution for determination of gross mineralization. The remaining two inner cores received K15NO3 solution for determination of gross nitrification. The 15N solutions contained 50 mg N L–1 at 50% enrichment to provide about 2 mg N kg–1 dry soil. Each inner core received 20 mL 15N solution delivered in eight injections of 1.25 mL each to both the top and bottom of the core. For each injection, the 15-cm 18-gauge side-port needle with attached syringe is inserted to approximately 14 cm and the injection is made as the needle is slowly withdrawn in a column the length of the core. The total of 16 injections to each core was designed to minimize non-uniform labeling of the inorganic N pools (Monaghan, 1995; Luxhøi et al., 2004).

In Year 2002 the enrichment and concentration of the added K15NO3 solutions were increased differentially for the plots receiving the AS200, DC100, and DC200 based on their initial NO3––N content. The inner cores in AS200 and DC100 plots were injected with K15NO3 solution containing 100 mg N L–1 at 99% enrichment to provide about 5 mg N kg–1 dry soil while those in the DC200 receiving plots were injected with K15NO3 solution containing 500 mg N L–1 at 99% enrichment to provide about 25 mg N kg–1 dry soil. Earlier measurements of inorganic N concentrations in these plots indicated large nitrate pools and the increase in enrichment and concentration of the K15NO3 solution was necessary to obtain higher 15N enrichments and increase sensitivity for the rate calculations.

The injection increased the soil moisture by about 4%. One of the two inner cores that received K15NO3 or 15NH4Cl solution was harvested 15 min after injection to determine extraction efficiency of 15N. The field extraction was as above except that an approximately 20-g soil subsample was added to 100 mL of preweighed cold 2 M KCl. The bottom of the remaining inner cores was sealed with aluminum foil, put into a 1-L Mason jar and incubated for 24 h buried in the original hole. After 24 h, the inner cores were harvested and processed as above to determine the inorganic N pools and 15NH4+ and 15NO3– enrichment. The NH4+ and NO3– pool sizes were measured after soil samples were extracted with 2 M KCl (approximately 1:5 soil/solution). The soil solutions were then filtered, stored frozen, and analyzed for NH4+ and NO3– using a Lachat auto-analyzer (QuickChem Systems, 1993). The 15N diffusion procedure by Stark and Hart (1996) was used to prepare 15N samples. The 15N enrichment of NH4+ and NO3+ pools was measured by continuous-flow direct dry combustion and mass spectrometry with an ANCA 2020 system (Europa Scientific, Cincinnati, OH) and blank corrected as described in Stark and Hart (1996). The 15N extraction efficiency was calculated as [(measured 15NH4+ or 15NO3–)/(added 15NH4+ or 15NO3–)]100. The Kirkham and Bartholomew (1954) equation was used to calculate the gross transformation rates after estimating initial total NH4+ or NO3– pool size and their 15N excesses by the equations of Stark (2000). Transformation rates from the first year of this plot (1997) are reported in Shi et al. (2004).

Lab Incubation Study
An 80-d laboratory incubation experiment was performed to determine the effect of repeated application of the treatments on the mineralization characteristics of the soil. In August 2002, soil samples were taken from the plots (0- to 15-cm depth) and passed through a 2-mm sieve. From each sample, 16 well-mixed 15-g dry weight equivalent subsamples were weighed into plastic specimen cups. Two of the subsamples were immediately extracted with 75 mL of 2 M KCl solution. The extracts were filtered and analyzed for NH4+ and NO3– as described above. The moisture content of the remaining subsamples was adjusted to 18% gravimetric water content. Two of the cups (without their lids) were put inside a 1-L Mason jar with a lid containing butyl rubber septum sealed with Apiezon N grease (M&I Materials LTD, Manchester, England) and with some water at the bottom of it to minimize loss of moisture from the soil. Two 3-mm diam. holes were made on the lids of the remaining specimen cups to facilitate aeration. All of the subsamples were incubated in the dark at 25°C.

The subsamples in the mason jars were used to measure the CO2 release by sampling through the rubber septum using hypodermic needle at 2-d intervals for the first 5 d, at 5-d intervals for the following 25 d and at 1 wk intervals thereafter. The CO2 was analyzed with an HP6890 series gas chromatograph with a TCD detector. Two of the other cups were extracted with 2 M KCl solution at 5, 10, 20, 40, and 80 d and analyzed for NH4+ and NO3–. The jars were flushed with air after sampling and the subsamples were adjusted gravimetrically for any moisture loss. The (NH4+–N + NO3––N) concentration versus time curve was fitted with the first-order equation [Nt = No(1 – e–kt)] in a SigmaPlot (SigmaPlot 2001, SPSS Inc., Chicago, IL) to estimate No (the potentially mineralizable N pool size) and k (decomposition rate constant). The net N mineralization and nitrification rates were calculated as the change in the inorganic N or nitrate pool size respectively after 20 d of incubation.

A mass balance model developed by Saetre and Stark (2005) to link N and C transformations was used to calculate the N/C ratio of microbial substrate during the incubation period. The model equation is: NM = Formula – Formula where, NM is net mineralization (mg N kg–1 d–1), fs is N/C of microbial substrate (g N g–1 C), R is microbial respiration or C mineralization (mg C kg–1 d–1), e is substrate use efficiency or growth efficiency (g C g–1 C), fm is N/C ratio of microbial biomass (g N g–1 C) and Iexc is inorganic N immobilized by microbes in excess of that needed for balanced N/C uptake (mg N kg–1 d–1). Our experimental data provided us with NM and R and we used a range of e values (0.3, 0.4, 0.5, 0.6, and 0.7) because we did not have independently determined e values for these soils. Iexc and fm were assumed to be zero and 0.17, respectively. Nitrification potential was determined by the shaken soil slurry method (Hart et al., 1994b) from the surface soil (0–15 cm) approximately 90 d after planting.

Statistical Analysis
All data were subjected to analysis of variance with significantly different means separated by the Tukey's Studentized Range Test (SAS Institute Inc., Cary, NC, 2001). Multiple time series data were subject to repeated measures analysis of variance with time as repeated measures factor. All statistical analyses were performed at 95% confidence level (P ≤ 0.05).


   RESULTS AND DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
REFERENCES
 
Gross Mineralization and Nitrification Rates
The effect of the different treatments on soil chemical properties has been extensively discussed in a companion paper with presentation of multiple year data from 1998 to 2002 (Habteselassie et al., 2006). These soil chemical properties are some of the major factors that affect microbial activities and hence N transformations. Booth et al. (2005), for example, noted the positive correlation of soil C and N concentration to microbial biomass N and gross N mineralization in various terrestrial ecosystems. We have also seen strong correlations among gross N mineralization and nitrification rates and soil organic C and total soil N concentrations (Table 1).


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  		Table1. Correlation coefficients (R) for relationships among mean gross N transformation rates, and mean soil organic C and N concentrations (1999–2002).

 
The overall effects of treatment, year, and treatment x year interaction were significant on gross N mineralization rate. The level of treatment application (100 vs. 200 kg N ha–1) was not significant. The gross N mineralization rate for DC200 was the highest. The gross rates for the rest of the treatments did not significantly differ from one another. The 4-yr (1999–2002, Fig. 1) mean gross N mineralization rate associated with DC200 was 5.72 mg N kg–1 soil d–1. This was 63 and 79% higher than the mean gross rates associated with both levels of LW and AS, respectively. The gross N mineralization rates across all waste treatments increased with time (Fig. 1) except for a decrease in year 2001 when the amount of waste applied was decreased due to a higher mineralization credit from previous applications (Habteselassie et al., 2006).


Figure 1
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  		Fig. 1. Gross N transformation rates (GMR-gross N mineralization rate; GACR- gross ammonium consumption rate; GNR-gross nitrification rate; GNCR-gross nitrate consumption rate) of soil treated with dairy-waste compost (DC), liquid dairy waste (LW), and ammonium sulfate (AS) for the last 6 yr to supply 100 and 200 kg available N ha–1 annually. Values represent means ±1 SE (n = 4).

 
Gross ammonium consumption rate was significantly affected by treatment and treatment x year interaction. The effect of year was not significant. Level of application was significant only for DC. Similar to gross N mineralization rate, DC200 resulted in the highest gross ammonium consumption rate. Unlike the gross N mineralization rate, however, DC100 resulted in significantly higher gross ammonium consumption rate than AS100, AS200, and LW100. In Year 2002, the gross ammonium consumption rates for DC200, LW200, and AS200 were 5.7, 4.9, and 2.4 mg N kg–1 soil d–1, respectively (Fig. 1).

Our gross N mineralization determinations show similar responses but with slightly higher rates compared with soils that had received waste or inorganic N fertilizers in previous studies. At this same site, one time application of DC at the high rate resulted in significantly higher gross N mineralization rate (1.65 mg N kg–1 d–1) than either AS or lagoon effluent 90 d after application (Shi et al., 2004). In a short-term incubation study, gross N mineralization rate was four times higher when the soil was amended with both sheep manure and AS (1.90 mg N kg–1 d–1) than with AS (0.52 mg N kg–1 d–1) only (Sorensen, 2001). Zaman et al. (1999) reported gross N mineralization rates of 6.1 and 3.4 mg N kg–1 soil d–1 16 d after field application of dairy shed effluent and NH4Cl, respectively, as compared with 1.5 mg N kg–1 soil d–1 for no treatment. Addition of pig slurry (Chantigny et al., 2001) and cocomposted sewage sludge (Sims, 1990) has also resulted in immediate but short-term increases in inorganic N contents, particularly NH4+. The time between application of treatments and measurement of gross N mineralization rate is important. Compost, for example, has a larger and more stabilized organic N pool that mineralizes steadily. In contrast, LW has less total N but more labile organic N, which mineralizes rapidly. Therefore, depending on when the rate is determined, results may vary. We determined rates to coincide with high plant demand and to avoid the immediate effect of recent waste or fertilizer applications since we were more interested in the cumulative effect on soil N dynamics of repeated applications.

Overall, gross nitrification rate was significantly affected only by the treatment. The effects of year, treatment x year interaction, and level of application were not significant. The result for gross nitrification rates followed that of gross ammonium consumption rates in that DC200 gave the highest rate and the rate for DC100 was also significantly higher than the AS100, AS200, or LW100. The difference between DC100 and DC200 was only marginally significant (P = 0.09). The similarity in the patterns of nitrification and ammonium consumption rates determined by the separate analyses shows that nitrification is one of the most responsive processes that consume ammonium. Average gross nitrification rates for Years 1999 to 2002 were 30 and 43% of the gross ammonium consumption rates in AS and LW treated plots, respectively (Table 2). Due to absence of plant uptake during measurements of gross rates, the remainder of the consumption is composed of immobilization and ammonium loss possibly through volatilization. If we calculate ammonium immobilization as the difference between consumption and nitrification, average rates (1999–2002) for the AS and LW treated soils were 1.7 (s.e. 0.3) and 1.9 (s.e. 0.3) mg N kg–1 soil d–1. As reported by other studies (Davidson et al., 1990, Shi and Norton, 2000; Booth et al., 2005), we note that nitrifiers were strong competitors for NH4+ in plots receiving all the treatments but especially in the compost treated plots (see GNR/GACR, Table 2).


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  		Table 2. Ratios of gross and net N transformation rates for plots that received ammonium sulfate (AS), dairy-waste compost (DC), and liquid dairy waste (LW) at 100 and 200 kg available N ha–1 for 6 yr. Gross nitrification rate (GNR)/gross ammonium consumption rate (GACR), GNR/ gross N mineralization rate (GMR), and GNR/nitrification potential (NP) values are means for Year 1999 to 2002. Net mineralization rate (NMR)/GMR and net nitrification rate (NNR)/GNR values are for year 2002 (n = 4).

 
The average gross nitrification rates for plots receiving DC, however, appear to be greater than the average gross ammonium consumption rate as shown by the GNR/GACR ratios (i.e., >1.0) but the difference between these rates was not significant due to high variability (Table 2). Microbial immobilization therefore accounted for a decreased proportion of the ammonium consumed in the compost treated soils. However, an overestimation of gross nitrification rates may have occurred as a result of methodological biases associated with the pool dilution technique. Of particular concern is the large and variable nitrate pools in these compost treated soils and the probable uneven distribution of microbial activities (Davidson et al., 1991; Luxhøi et al., 2004).

Davidson et al. (1991) stressed that achieving uniform distribution of 15N enrichment throughout an intact core is not only impossible but also unnecessary. While our multiple injection technique avoided this potential bias, it is highly likely that there were nonrandom distributions of microbial processes. For example, hot spots of mineralization and nitrification may be formed due to uneven distribution of the compost. These hot spots would have higher initial 14NO3– concentrations and more rapid nitrification than the rest of the soil, resulting in higher dilution of the added 15NO3– enrichment and overestimation of the gross nitrification rate. The observed GNR/GACR (higher or close to 1) in the compost-treated soils could also be a result of the underestimation of the gross ammonium consumption rate. This may be caused by preferential consumption of the 14NH4+ due to nonhomogenous distribution of the added 15NH4+ in relation to the 14NH4+ pool and microbial processes (Monaghan, 1995; Luxhøi et al., 2004). However, we believe that this explanation is less likely because the 14NH4+pool was consistently low for all soils and under these conditions heterogeneous distributions should not significantly influence the calculated rates (Luxhøi et al., 2004). Herrmann et al. (2005) used analytical equations to assess the effects of preferential use of added and native 15N and 14N pools in estimating gross N transformation rates. However, quantification of preferential use in our experiments would require additional measurements that were not available. The high variability in the gross nitrification rates reflects the uneven distribution of this microbial process in the compost-treated soils compared with the other treatments. The lower rates may be closer to the true gross nitrification rates neglecting the role of localized zones of rapid nitrification. Actual gross nitrification rates are in between the nitrification potential rate (determined with nonlimiting substrate supply) and the net nitrification rates; these values are compared in the ratios found in Table 2. Disturbance artifacts due to sieving, mixing, and laboratory incubation are likely to be as least as significant (Stark and Schimel, 2001) as the problems associated with nonuniform distribution and utilization of label in intact cores.

The higher GNR/GACR values for DC treated plots suggest the presence of larger and more competitive nitrifier populations than those receiving the other treatments. It is also likely that N is not limiting to heterotrophs due to the large amount of labile organic N. Manure amendment has been noted to increase amino acid content in soils (Scheller and Raupp, 2005) and these compounds may be used directly for microbial assimilation bypassing the immobilization from the ammonium pool (Barraclough, 1997).

The ratio of gross nitrification rate to nitrification potential (GNR/NP) is a good indicator of substrate availability to nitrifiers. GNR/NP for AS, DC, and LW treated plots were about 10, 61, and 20% (Table 2), which indicated that ammonium availability was limiting the activity of nitrifiers in plots receiving all the treatment but more so in AS and LW than DC treated plots. The ratio of gross nitrification to mineralization (GNR/GMR) is also considered to be an index of the nitrifying capacity of soils (Booth et al., 2005). GNR/GMR for plots receiving AS and LW were below 80% (Table 2). The GNR/GMR for DC treated plots were, on the other hand, more than 100%, indicating large increase in nitrifying capacity of the soil following repeated applications of compost. The enhancement of nitrification by repeated compost application is undesirable due to the high mobility of nitrate. We observed significant nitrate concentrations in the lower part of the profile in DC200 treated plots (Habteselassie et al., 2006).

Microbial consumption of nitrate, on the other hand, was not significantly affected by treatment or year, nor was there a treatment x year interaction. Gross nitrate consumption rates were highly variable in all the years with a tendency to be negative mainly in DC treated soils. Overall, they were not statistically different from zero for all the treatments (Fig. 1). The negative consumption rates might have been caused as a result of methodological artifacts. This would mainly be the nonuniform labeling of the ambient 14NO3– pool, causing preferential consumption of 14NO3– (Luxhøi et al., 2004). The mean extraction efficiency of 15NO3– for the 4-yr period immediately after injection (T0) ranged from 80 to 92% for the different treatments. The mean extraction efficiency of 15NO3– 1 d after injection (T1) was also similar indicating the lack of microbial nitrate assimilation or loss. This is in agreement with a previous study at this site by Shi et al. (2004). While not observed at this site, significant consumption of nitrate is possible under high C availability or due to limited ammonium availability (Okereke and Meints, 1985; Recous et al., 1990; Burger and Jackson, 2003). DC200 has significantly increased the organic C content of the soil but the C/N ratio of the compost was low enough (12:1) to result in net N mineralization. Hence, immobilization of nitrate is less likely under these conditions.

There are studies that have reported the effect of organic amendments on gross nitrification rates. Muller et al. (2003) noted a 20-fold increase in gross nitrification rate shortly after cattle slurry application (10.92 mg N kg–1 d–1) in comparison with the control (0.56 mg N kg–1 d–1). The gross rate of nitrification in an agricultural soil 10 d after dairy shed effluent and NH4Cl application were 1.4 and 2.5 mg N kg–1 soil d–1, respectively (Zaman et al., 1999) versus 0.7 mg N kg–1 soil d–1 for soil without dairy shed effluent or NH4Cl application. Application of ammonical fertilizers (such as NH4Cl) may stimulate nitrification more quickly than organic wastes over a short period of time because of the rapid increase in available NH4+. For example, Mendum et al. (1999) reported gross nitrification rates of 2.3, 7.8, and 0.2 mg N kg–1 soil d–1 3 d after application and rates of 1.4, 0.7, and 0.2 mg N kg–1 soil d–1 42 d after application of farmyard manure, NH4NO3, and no treatment, respectively. Shi and Norton (2000) also noted that addition of AS to soils resulted in short term but rapid increase in populations of nitrifiers indicated by higher nitrification rate and potentials compared with soil treated with DC. However, AS maintained a lower nitrifier population over the long term. Our findings are in agreement with these studies in that application of animal waste significantly increases N transformation rates in the long term and the magnitude of increase depends on the time between application and rate measurement.

In general, plots receiving compost showed high variability in gross N transformation rates. This might be due to hot spots of mineralization, immobilization, or nitrification as a result of uneven distribution of the waste (Korsaeth et al., 2001). The use of the small concentric cores for field measurement of gross N transformation rates in waste-treated lands where variability tends to be high is, therefore, not advised. Burger and Jackson (2003) employed leaching and sieving soil samples to reduce variability of nitrate and found out that leaching was better than sieving for decreasing variability in gross nitrification rates.

Gross ammonium consumption rate was higher than gross N mineralization rate for plots receiving all the treatments. This indicates the use of more ammonium than produced through mineralization, which might come from the already existing ammonium pool in the soil (background ammonium) that ranged from 0.5 to 1.4 mg kg–1 soil with slightly higher ammonium concentrations in plots receiving compost. Stimulation of ammonium consumption as a result of addition of the labeled solution of NH4Cl has previously been noted (Davidson et al., 1991; Booth et al., 2005). This has been related to low NH4+ availability, which we observed in this study.

Mineralization Potential and Net Mineralization Rates
Net N mineralization from soils receiving the various treatments was well described by the first-order model (Fig. 2) with regression fits (R2 values) greater than 0.94 for both levels of AS and LW (Table 3). DC200 had a relatively lower R2 value of 0.72 as a result of high variability. DC100 and DC200 significantly enhanced the potentially mineralizable organic N pool (No) to more than two times the AS100, AS200, and the LW100 No values (Table 3). However, the No for plots receiving LW200 was not significantly different from those receiving either DC100 or DC200. The level of DC application did not have a significant effect on No. The decomposition rate constant for the labile organic N pool (k) from the first-order kinetics model was significantly increased only by DC200. The total N turnover rate (kt) can be calculated as the ratio of the gross N mineralization rate to the total soil N pool (Booth et al., 2005). The mean kt values for 1999 to 2002 associated with the AS100, AS200, DC100, DC200, LW100, and LW200 treated plots were 0.002, 0.002, 0.006, 0.009, 0.003, and 0.006 d–1, respectively. The kt associated with DC200 was significantly higher than AS100, AS200, and LW100.


Figure 2
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  		Fig. 2. Changes in inorganic N during laboratory incubations of soils treated with dairy-waste compost (DC), liquid dairy waste (LW), and ammonium sulfate (AS) to supply 100 and 200 kg available N ha–1 annually for 6 yr. Values represent means (n = 4).

 

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  		Table 3. Potentially mineralizable N (No) and decomposition rate constant (k) based on the first order kinetics model fit (Fig. 2) for soils that received the different treatments.

 
Assessing the potentially mineralizable organic N pool for soils that have received organic amendments is important for estimating the capacity of these soils to provide N over time and hence management. Our results are in agreement with other studies that have reported changes in N mineralization as a result of addition of organic wastes to agricultural lands. After 6 yr, the N supplying potential of an organically managed agricultural system was approximately twice that of the conventionally managed system (Burger and Jackson, 2003). N'Dayegamiye et al. (1997) noted that addition of cattle, sheep, and horse manure composts with C/N ranging from 12 to 28 increased the amount of mineralized N in an agricultural soil available to plants immediately after amendment. The increase in mineralized N over the control during the 140-d incubation comprised about 3% of the total N added in the composts while in a greenhouse pot experiment, plants N uptake comprised 13 to 40% of the total N added (N'Dayegamiye et al., 1997). In our study, the observed increase in N mineralization potentials reflects the cumulative effect of dairy-waste amendments on the organic N fractions at their different stages of decomposition. We also observed that mineralization estimates from laboratory incubations are appropriate indicators of N supplying potential but underestimate the increase in N availability to plants due to compost additions (Habteselassie et al., 2006).

The treatment of organic wastes before soil application is an important factor that determines their mineralization characteristics. Composting, for example, may create more stabilized N and C forms possibly slowing down the rate of mineralization but also decreases the C/N of dairy wastes and this may have the opposite effect. Preusch et al. (2002) reported that more than 40% of poultry litter N mineralized in the 120-d incubation experiment. Nitrogen mineralization was, however, <9% when the poultry litter was composted. Similarly, Shi et al. (1999) and Castellanos and Pratt (1981) have reported that <10% of composted dairy-waste mineralized in 10 to 12 wk time. Solid separation and lagoon treatment, on the other hand, removes the more recalcitrant C rich straw and may result in increased rate of mineralization over a short period of time until the labile organic N pool is depleted. Shi et al. (2004), for example, reported 30 to 90% mineralization of organic N of lagoon effluent in a 70-d incubation experiment depending on the rate of application.

Determination of net mineralization and nitrification rates in our 20-d laboratory incubation study indicated that soil from plots receiving DC100 and DC200 had significantly higher rates than AS100, AS200, and LW100 (Fig. 3). The LW200 had a rate that was not significantly different from DC100. Net mineralization and nitrification rates for an individual treatment were very similar, indicating that the mineralized ammonium was quickly nitrified. Net mineralization rate was <20% of gross mineralization rate except for DC200 and AS200 treated plots (Table 2). Net nitrification rate was only 25% of the gross rate in DC treated soils while it was about 50% in LW treated soils. The difference between net and gross rates emphasizes the role other consumptive processes play in the control of N in the system and the increase in consumption caused by soil disturbance (sieving and mixing). The larger the difference between the gross and net rates, the bigger the role of these consumptive processes. Even though they are not as informative as gross N transformation rates in terms of N dynamics, measuring net mineralization and nitrification rate in relatively long-term incubation studies is important for estimating plant available N and the tendency of nitrate to leach out of the system. Based on the net mineralization rates, the estimated plant available N for the DC200, DC100, LW200, and AS200 treated soils were 2.2, 1.4, 0.94, and 0.64 kg N ha–1 d–1, respectively. While these estimates are applicable only to the middle of the growing season when the soil samples were collected, the high daily values for the compost DC200 treated soils suggest that mineralization could supply excess available N for silage corn production (Habteselassie et al., 2006).


Figure 3
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  		Fig. 3. Net N mineralization and nitrification rates of soil treated with dairy-waste compost (DC), liquid dairy waste (LW), and ammonium sulfate (AS) to supply 100 and 200 kg available N ha–1 annually for 6 yr. Values represent means ±1 SE (n = 4).

 
Carbon Mineralization
Carbon mineralization was significantly affected by treatment or time, and there was a treatment-time interaction. Carbon mineralization was the highest during the first 3 d of incubation with all the treatments showing similar rates except for the LW100 whose rate was lower than the others from the start. The initial high rate is associated with the flush of C mineralization that follows disturbance and addition of moisture to dry soil, described by Birch (1958). The difference in C mineralization rate among the different treatments started to show after Day 3 (Fig. 4) and after Day 25 the rates began to stabilize. The C mineralization rates for soil samples from the plots receiving DC200, LW200, and AS200 after 1 d of incubation were 45, 40, and 42 mg CO2–C kg–1 soil d–1 whereas the rates at the end of the incubation period were 13.4, 9.9, and 5.5 mg CO2–C kg–1 soil d–1, respectively. The cumulative CO2–C mineralized was 1.2, 0.9, and 0.5 g kg–1 soil, which were 6.9, 8.4, and 7.4% of the organic C for DC200, LW200, and AS200, respectively. Overall, DC resulted in the highest C-mineralization rates. The soil samples from the plots receiving the LW200 had rates comparable with both levels of DC. There was no significant difference between the two levels for DC and AS. For LW, however, the high level treatment gave significantly higher C-mineralization rates than the low level treatment.


Figure 4
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  		Fig. 4. Carbon mineralization rates during laboratory incubations of soils treated with dairy-waste compost (DC), liquid dairy waste (LW), and ammonium sulfate (AS) to approximately supply 100 and 200 kg available N ha–1 annually for 6 yr. Values represent means ±1 SE (n = 4).

 
Other studies have documented increased C flux associated with application of organic amendments. Carbon mineralization was increased after addition of dairy manure by as high as 42 to 400% at different points of the incubation period as compared with the control (Calderon et al., 2004). The largest increase was observed during the first week. In a 150-d incubation study, higher C mineralization was also reported by Sanchez et al. (2004) in soils that received repeated application of composted dairy manure compared with soils that received synthetic N fertilizer. Both levels of DC and LW200 have obviously increased C mineralization, as they are direct sources of organic matter. The DC200 has particularly resulted in a larger increase in soil organic C content (Habteselassie et al., 2006). Such an increase in organic C did not lead to a corresponding net immobilization of N due to the low C/N ratio of the compost (12:1). It is well documented that C plays a central role in controlling N cycling in soils (Hart et al., 1994b; Johnson, 1995; Brady and Weil, 1999). It was suggested that manure with C/N ratio below 15:1 are more likely to give positive N mineralization after application in soils (Beauchamp and Paul, 1989). Calderon et al. (2004) have also reported net N mineralization for manures, which had a mean C/N ratio of 16. In general, a C/N ratio of 20:1 or narrower is assumed to have sufficient N to supply the decomposing microbes and release some N as well (Miller et al., 1983).

It would be more appropriate to use the C/N ratio of the substrate that microbes actually use to predicate N dynamics in soils rather than the C/N of the raw materials applied. The mass balance model of Saetre and Stark (2005) enables the determination of the N/C ratio of microbial substrate by linking C and N transformations. Based on the model, for the first 10 d of the incubation period the DC200 treated soils had higher N/C ratio (lower C/N) for microbial substrate than the LW200 and AS200 treated soils at all growth efficiency values (Fig. 5). The N/C ratio of microbial substrate was not significantly different among the DC200, LW200, and AS200 treated soils after Day 20 and was about 0.08:1 (C/N = 13:1) and 0.12:1 (C/N = 8:1) at 0.3 and 0.7 growth efficiency values, respectively.


Figure 5
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  		Fig. 5. Model estimates of N/C ratio of microbial substrates during laboratory incubations of soils treated with DC200, LW200, and AS200. The microbial efficiency value (e) varied from 0.3 to 0.7. Values represent means ±1 SE (n = 4).

 
In summary, both levels of compost significantly raised the labile organic N pool and the decomposition rate constant above the ammonium sulfate. Soil organic C and total N were well correlated with gross N transformation rates in this study. The highly dynamic nature of the plots receiving the DC was indicated by higher gross N mineralization, gross nitrification and C mineralization rates. Gross nitrate consumption was not, however, significantly different from zero, suggesting insignificant microbial nitrate immobilization. Our observations suggest that promoting N immobilization in soils receiving repeated waste applications may require additional high C content additions. The rapid turnover of N in plots receiving DC implied the need for proper management of N to avoid negative environmental consequences associated with gaseous N losses and leaching.


   ACKNOWLEDGMENTS
 
This work was supported by grants from the USDA NRI-CGP (#9600839 and #9935107-7808), Office of the Vice President for Research (CURI) at Utah State University, the Inland North West Research Alliance (INRA) doctoral Fellowship, and the Utah Agricultural Experiment Station, Utah State University and approved as journal paper 7716. The technical and field assistance of Vaughn Thacker, Teresa Koper, and Eric Dodson and the cooperation of the Mickelson Dairy of Smithfield, UT are appreciated.

Received for publication June 15, 2005.


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M. Y. Habteselassie, B. E. Miller, S. G. Thacker, J. M. Stark, and J. M. Norton
Soil Nitrogen and Nutrient Dynamics after Repeated Application of Treated Dairy-Waste
Soil Sci. Soc. Am. J., June 21, 2006; 70(4): 1328 - 1337.
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