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Published in Soil Sci. Soc. Am. J. 68:1867-1874 (2004).
© 2004 Soil Science Society of America
677 S. Segoe Rd., Madison, WI 53711 USA

Division S-3—Soil Biology & Biochemistry

Microbial Nitrogen Transformations in Response to Treated Dairy Waste in Agricultural Soils

Wei Shia,*, Bruce E. Millerb, John M. Starkb and Jeanette M. Nortonb

a Dep. of Soil Science, North Carolina State Univ., Raleigh, NC 27695
b Dep. of Plants, Soils, and Biometeorology, Utah State Univ., Logan, UT 84322

* Corresponding author (wei_shi{at}ncsu.edu)


   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Dairy wastes are commonly applied to croplands as N fertilizers, but the dynamics of N release and transformations during the growing season are difficult to predict. We compared N mineralization kinetics and examined microbial N transformations in soil receiving dairy-waste compost vs. lagoon effluent. Mineralization kinetics was examined with a 70-d laboratory incubation, and a first-order model was used to derive mineralization parameters. Measurements of N transformations were conducted with 15N pool dilution techniques in silage corn field plots that were unfertilized or fertilized with ammonium sulfate, lagoon effluent, or compost at two rates equivalent to 100 or 200 kg available N ha–1. The N mineralization potential was higher and the first-order rate constant was lower in soil receiving compost than lagoon effluent. Approximately 6% of compost N was mineralized within 2.5 mo; in contrast, up to 90% lagoon effluent organic N was released. However, silage yield was greatest in the compost treatment, showing that synchronization of N availability is as important as the amount mineralized. The field 15N measurements indicated that microbial NO–3 consumption was negligible despite the treatments. Microbial NH+4 immobilization in soil receiving dairy wastes was similar to that in soil unfertilized or fertilized with inorganic N. Soil treated with the high-rate compost had the highest rates of mineralization and nitrification, which led to the highest soil NO–3 accumulation. Our observations suggest that peak plant demand is met by the compost N; however, its high N mineralization potential makes the management of dairy compost a difficult task.


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 
ANIMAL WASTE DISPOSAL on agricultural soils has been advocated and practiced because organic wastes are an important nutrient source and can also enhance soil physical properties (Golueke, 1973). The application of organic wastes vs. traditional N fertilizers differentially impacts microbial N transformations for several reasons. First, organic waste provides crop-available N slowly through N mineralization and subsequent nitrification. The rate at which NH+4 is produced may coincide with the N uptake by crops, and the synchrony of crop NH+4 uptake with NH+4 supply may decrease soil nitrification rates and nitrifier population activities (Verhagen et al., 1994). Second, organic waste application adds organic C to soil and may increase soil microbial biomass and activity (Schnürer et al., 1985; Fauci and Dick, 1994). The enhanced heterotrophic activity may increase microbial NH+4 immobilization or enhance microbial NO–3 immobilization.

Soil NH+4 and NO–3, the products of mineralization and nitrification, comprise the majority of N available for crop growth. Additive and competitive interactions among N mineralization, immobilization, and nitrification control soil NH+4 and NO–3 concentrations. These interactions need to be more fully understood for informed agricultural N management decisions concerning organic wastes. Microbial NO–3 production (i.e., nitrification) has been intensively investigated in agricultural soils. Net rates, which are determined by the change of soil NO–3 pool size over time, are measured often because microbial NO–3 immobilization is assumed to be trivial in most agricultural soils (Jansson et al., 1955; Winsor and Pollard, 1956; Rice and Tiedje, 1989). However, the possibility of microbial NO–3 immobilization commonly observed in systems such as forest and grassland soils where available C and fine-scale spatial heterogeneity are high (Jackson et al., 1989; Davidson et al., 1990; Stark and Hart, 1997) should not be overlooked. Recent evidence showed that microbial NO–3 immobilization was significant after several years of organic farming practices (Burger and Jackson, 2003). Given that organic waste application may add considerably to the heterogeneous distribution of available C, direct measurement of individual N process rates is of significance to elucidate the mechanisms by which soil NO–3 concentrations are controlled in soils receiving animal wastes.

Aerobic composting and anaerobic lagoon holding are two contrasting systems to stabilize dairy waste before land application. The different treatment conditions transform dairy waste into end products that may differ in their N transformation kinetics. Although N-release characteristics of some organic wastes have been examined by laboratory incubations along with first-order mathematical modeling (Castellanos and Pratt, 1981; Bernal and Kirchmann, 1992; Shi et al., 1999), limited data are available for dairy-waste compost, and even less are available for dairy-waste lagoon effluent.

Field experiments have been conducted to examine the effects of organic waste application on crop yield and N uptake, soil chemical properties, and groundwater quality (Patni and Culley, 1989; King et al., 1990; Zebarth et al., 1997). These authors attempted to determine an appropriate application rate of organic waste that would improve crop yield and N uptake, while maintaining soil and groundwater quality. The supply of available N from organic waste is dependent on microbial decomposition and the quality of the waste. Various rates of N fertilization may result in the alteration of N resource availability, leading to the change of relationship among soil heterotrophs, nitrifiers, and crops. Hence, agricultural management of organic N sources is more complex and more challenging than that of mineral N fertilizer. The overall goal of this study was to examine the short-term (i.e., within the first growing season) response of microbial N transformations to treated dairy waste. We were specifically interested in (i) comparing N mineralization kinetics of dairy-waste compost vs. lagoon effluent, (ii) examining microbial N mineralization, immobilization, and nitrification in the first growing season following the application of compost vs. lagoon effluent, and (iii) evaluating microbial control of NO–3 pool size via NO–3 production vs. consumption following application of organic vs. inorganic N sources.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Experiment I: Nitrogen Mineralization Kinetics
Dairy-waste compost was sampled from windrow piles that were frequently turned and watered (Shi et al., 1999). The starting materials were feces and urine with bedding materials and additional wheat straw to form an initial C-to-N ratio of 38. After a 2-mo composting, the compost was collected from the difference zones (upper vs. lower and inner vs. outer) of windrow piles to form a composite sample. The compost was stored at 4°C until the incubation. Dairy-waste lagoon effluent was collected from the anaerobic pond of a two-stage anaerobic and aerobic lagoon system at the Caine Dairy Farm of Utah State Agricultural Experiment Station, Logan, UT. The recycling between the aerobic and anaerobic ponds accelerated the inorganic N loss through NH3 volatilization and coupled nitrification and denitrification. The raw materials loaded into the anaerobic pond were milking parlor waste water, feces and urine, and bedding material. Before sampling, the anaerobic pond was agitated for about 2 d. After sampling, the lagoon effluent was kept at 4°C until incubation with soil. We sampled Nibley silty clay loam soil (fine, mixed, active, mesic Aquic Argixeroll) from the Caine Dairy Farm of Utah State Agricultural Experiment Station. The soil was sampled at the 0- to 15-cm depth from two fields; one was cropped with corn in the spring and the other with alfalfa. The soil was sieved (<2 mm), and partially air-dried to avoid excessive moisture after the addition of lagoon effluent. The chemical properties of the soil and dairy waste are given in Table 1.


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  		Table 1. Selected chemical properties of soils before treatment and dairy wastes (compost and lagoon effluent) at application.

 
The incubation treatments for both corn and alfalfa soils were as follows: (i) Control, soil without the addition of dairy waste; (ii) Compost, soil with the addition of compost at 33 g dry wt. kg–1 soil (equivalent to 66 Mg dry wt. ha–1); (iii) Low lagoon, soil with the addition of lagoon effluent at 100 mL kg–1 soil (equivalent to 200 m3 ha–1); and (iv) High lagoon, soil with the addition of lagoon effluent at 200 mL kg–1 soil (equivalent to 400 m3 ha–1). These application rates are commonly used in Utah. Our previous study has evaluated N mineralization kinetics in a soil with the addition of high rate vs. low rate compost (Shi et al., 1999). Here, we included one application rate of compost for the comparison of N mineralization kinetics between the dairy-waste compost and lagoon effluent. The control soil or soil-dairy waste mixtures (equivalent to 20 g dry wt.) were weighed into 120-mL specimen containers with aeration holes in the lids, adjusted and maintained at 23% moisture content (about 60% of soil water holding capacity), and incubated at 20°C. Three specimen containers were destructively sampled at 0, 7, 14, 28, 42, 56, and 70 d. Soil inorganic N was extracted with 2 M KCl (1:5, soil wt. to KCl vol.) by shaking for 1 h. Extracts were filtered through prerinsed Whatman no. 1 filter papers, and the filtrates were analyzed colorimetrically for inorganic NH+4–N and (NO–3 + NO–2)-N with a Lachat flow-injection autoanalyzer (Lachat Instruments, Mequon, WI). We used a nonlinear regression to derive the best fit of data to the first-order model: Nm = Ni + N0(1 – e–kt), where Nm was the mineral N at time t, Ni was the initial inorganic N (the mean of inorganic N at time zero, n = 3), N0 was the potentially mineralizable organic N (i.e., N mineralization potential), and k was the N mineralization rate constant. One-way ANOVA (SuperANOVA 1995, Abacas Concepts, Berkeley, CA) with mean separation techniques of Fisher's LSD was used to compare initial inorganic N. The best-fit values of N0 and k and their standard errors were used to evaluate whether or not N0 and k values were significant between treatments (GraphPad Guide to nonlinear regression, 1996, Graphpad Software, Inc., San Diego, CA).

Experiment II: Microbial Nitrogen Transformations
In situ measurement of microbial N mineralization, immobilization, and nitrification was conducted in silage corn field plots at the Greenville Farm (North Logan, UT), where annual precipitation was 432 mm, average annual temperature was 9°C, and the frost-free season was 156 d (Utah Climate Center, 1997, personal communication). The Millville soil used in the field experiment was strongly calcareous silt loam (coarse-silty, carbonatic, mesic, Typic Haploxeroll). We applied ammonium sulfate [(NH4)2SO4], dairy-waste compost, and lagoon effluent at two rates equivalent to 100 and 200 kg available N ha–1. The application rates of dairy waste were based on our previous study (Shi et al., 1999) and the results of Experiment I. It was estimated that 90% total N in lagoon effluent was available for crop growth, while only 10% total compost N was available. Table 1 lists the chemical properties of the Millville soil and the dairy waste used in the field. The seven treatments were (i) control, no added N; (ii) AS-100, (NH4)2SO4 at 100 kg N ha–1; (iii) AS-200, (NH4)2SO4 at 200 kg N ha–1; (iv) LS-100, lagoon effluent at 100 m3 ha–1; (v) LS-200, lagoon effluent at 200 m3 ha–1; (vi) DC-100, dairy-waste compost at 50 Mg dry wt. ha–1; and (vii) DC-200, dairy-waste compost at 100 Mg dry wt. ha–1. The N content of lagoon effluent was overestimated before the application. Thus, the actual application rates for the LS-100 and LS-200 were 65 and 130 kg total N ha–1, respectively (Table 1). These treatments were arranged in a completely randomized block design with four replications. The 28 plots were each 3.0 m wide and 9.1 m long with four rows of corn in each plot. Between each block was a 1.0-m alley, and no alley was between each plot in a single block. Nitrogen fertilizers were broadcast on the soil surface in May, and then tilled into the 0- to 15-cm soil layer when soil moisture conditions allowed. Silage corn (variety DK-656) was planted at 82000 plants ha–1 about 2 wk after N fertilization. Corn was irrigated and maintained according to the standard agricultural practices for Cache Valley, UT.

Soil inorganic N was examined in June, the early growing season of silage corn and in November after harvest. Soil samples were collected in the middle of the plots and between corn rows. Each plot had one sample for the 0- to 30-cm and 30- to 60-cm soil depths, respectively. About 15-g samples of moist soils were immediately placed in 120-mL specimen containers with 2 M KCl (1:5, soil wt. to KCl vol.) and stored in a cooler. After transported to the laboratory, soil samples were immediately prepared for inorganic N measurement as described above.

In situ measurement of gross N transformation rates by 15N pool dilution techniques was performed in August when considerable N was required by silage corn (90 d after N fertilization). This occasion allowed us to efficiently address the significance of microbial N mineralization and immobilization in regulating plant available N in soil. Soil in the middle of each plot and between corn rows was used for 15N labeling. Four small PVC cylinders (5-cm diam. x 15 cm long) were driven into the soil in each plot. Large PVC cylinders (10-cm diam. x 20 cm long) were then forced into the soil around each small cylinder. The pair of a large and a small cylinder was removed and the soil between the two cylinders was placed into a plastic bag, mixed, and immediately subsampled for extraction with 2 M KCl (about 15 g dry soil in 75 mL). The remaining mixed soil was used later for measuring soil gravimetric water content and nitrification potential. Two of the small cylinders received K15NO3 injections and other two received 15NH4Cl injections. The 15N solutions contained 50 mg N L–1 at 50% 15N enrichment. Twenty milliliters of 15NO–3 or 15NH+4 solution was injected (8 x 1.25-mL injections from top and bottom of a cylinder) with an 18-gauge side-port spinal needle into each small cylinder to provide about 2 mg N kg–1 dry soil. The soil moisture was increased by about 4% by the injections of 15N solution.

In each pair of the small cylinders injected with 15NO–3 or 15NH+4 solution, one cylinder (T0 cylinder) was immediately (within 15 min after labeling) broken up, mixed, and extracted with 2 M KCl. The other cylinder (T1 cylinder), covered at the bottom with aluminum foil, was placed into a 1-L Mason jar that was capped and buried in the original location. After 24.25 h, the T1 cylinder was broken up, mixed, and a subsample of about 20 g dry soil was extracted with 100 mL 2 M KCl. Inorganic N was prepared and analyzed using the method described above. A 15N diffusion procedure (Stark and Hart, 1996) was used to prepare 15N samples. The 15N enrichment of NH+4 and NO–3 pools was measured by continuous-flow direct dry combustion and mass spectrometry with an ANCA 2020 system (Europa Scientific, Cincinnati, OH). Soil nitrification potential was determined by a shaken soil slurry method (Hart et al., 1994).

Carbon mineralization rates were measured simultaneously with the field 15N labeling experiment. A 20-mL vial containing 1 mL 1M NaOH was placed into the 1-L Mason jar along with a T1 cylinder. After 24.25 h, the vial was removed from the Mason jar and capped tightly for later analysis of CO2 trapped in the base. A Mason jar containing only the base trap was used as a blank. The rate of CO2 production was determined by titration with standardized 0.2 M HCl (Zibilsk, 1994).

The amount of 15N-NH+4 or 15N-NO–3 in the T0 and T1 cylinders was calculated by multiplying the 15N excess (15N enrichment percentage minus the background 0.37%) by the NH+4 or NO–3 pool size. The percentage of added 15N-NH+4 (or 15N-NO–3) recovered in NH+4 (or NO–3) pool was calculated as 100 x the amount of 15N measured/the amount of 15N added. Gross rates of N mineralization and nitrification were calculated by the equations of Kirkham and Bartholomew (1954), in which the initial 14+15NH+4 or 14+15iNO–3 pool size and their 15N excesses were estimated by the equations of Stark (2000). The gross immobilization rate of NH+4 was calculated by subtracting the gross nitrification rate from the NH+4 consumption rate. The gross immobilization rate of NO–3 was calculated by subtracting the net nitrification rate from the gross nitrification rate.

The N content of the plant tissue for each plot was determined at silking phase by collecting eight ear leaves and at harvest from a subsample of the chopped corn. After drying at 60°C for 48 h, the ear leaves and chopped corn were finely ground and N content was determined with Kjeldahl digestion and distillation method (Jones et al., 1991). Silage corn yield was determined by harvesting and weighing the aboveground portion of the plants from the middle 5.3 m of two rows for each plot.

All statistical analyses were performed with SuperANOVA statistical software for Macintosh computer (Abacus Concepts, 1995, Berkeley, CA). The recovery and excess of 15N-NH+4 or 15N-NO–3 at T0 and T1 soil samples were analyzed by a split plot method with treatments as the main plot and labeling days as the subplot. Treatment effects on inorganic N were also analyzed as a split plot design with treatments as the main plot and soil depths as the subplot. Treatment effects on rates of C and N transformations, silage corn yield, and plant N content were analyzed based on a completely randomized block design.


   RESULTS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Nitrogen Mineralization Kinetics
Throughout the 70-d incubation, NO–3–N was the dominant form of inorganic N. Mineralization dynamics differed for the control soil, the soil receiving compost, and the soil receiving lagoon effluent, while the patterns of inorganic N accumulation in the alfalfa soil were similar to those in the corn soil. The three parameters, Ni, k, and N0 for characterizing the N mineralization, are shown in Table 2. The Ni was the highest for the soil receiving high-level lagoon effluent and the soil receiving compost, intermediate for the soil receiving low-level lagoon effluent, and lowest for the control soil. Most of the values of k in lagoon effluent-amended soils were significantly higher than those in compost-amended soils and controls. The N0 values were the highest for the compost-amended soils, intermediate for the soil with the high-level lagoon effluent, and the lowest for the soil alone or the soil with the low-level lagoon effluent. We calculated the inorganic fraction of dairy-waste N by dividing the difference of Ni between waste-amended soil and the control (Table 2) by the total N of dairy waste. Approximately 4% of compost N was inorganic N, while about 50% of lagoon effluent N was inorganic N. These estimations were consistent with the results of direct measurement (Table 1). Furthermore, we estimated the fractions of potentially mineralizable N of dairy wastes by dividing the difference of N0 between waste-amended soil and the control (Table 2) by the organic N of dairy waste. About 6% of organic N in compost was mineralized in the 70-d incubation. However, the percentage of mineralizable N relative to lagoon effluent N was dependent on the application rates. About 30% of organic N in lagoon effluent was mineralized for the low application rate, whereas 90% of organic N was mineralized for the high rate.


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  		Table 2. First-order parameters of N mineralization in control soil (control), soil receiving compost (compost), soil receiving lagoon effluent at a low level (low lagoon) and soil receiving lagoon effluent at a high level (high lagoon).

 
Soil Inorganic Nitrogen in Response to Nitrogen Fertilization
Effects of the various N sources and their application rates on soil NH+4 and NO–3 concentrations in the early growing season and after harvest are presented in Fig. 1 . Soil NH+4 and NO–3 pool sizes were higher in the early growing season than after harvest (P < 0.05). In the early growing season, soil NH+4 and NO–3 concentrations at the 0- to 30-cm depth differed significantly among the treatments, while the concentrations at the 30- to 60-cm depth were similar among treatments (Fig. 1). Except for the soil treated with dairy-waste lagoon effluent, soil NO–3 concentrations were significantly higher with the high rate of N fertilization than with the low rate. After harvest, only soil treated with the high-rate dairy-waste compost had sizable NH+4 and NO–3 pools at both 0- to 30- and 30- to 60-cm depths (Fig. 1). High inorganic N accumulation was also observed at the 0- to 15-cm depth in the soil treated with the high-rate compost during the 15N experiment in August (i.e., during rapid growth period of silage corn) (Table 3).



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  		Fig. 1. Soil NH+4– and NO–3–N pool sizes of two depth intervals in various soil treatments in June (early growing season) and in November (after harvest). Values are means and standard errors for n = 4. Control, soil without N fertilization; AS, soil fertilized with (NH4)2SO4; DC, soil fertilized with dairy-waste compost; LS, soil fertilized with dairy-waste lagoon effluent; application rates equivalent to available N of 100 and 200 kg N ha–1.

 

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  		Table 3. Soil inorganic N pool sizes and microbial processes rates of C mineralization, N mineralization, and microbial NH+4 immobilization in the variously treated soils 90 d after N fertilization.

 
Microbial Nitrogen Transformations
The recoveries of 15NO–3 determined immediately after 15N injections (T0 cylinders) ranged from 70 to 94%. The recoveries of 15NO–3 determined 1 d after 15N injections (T1 cylinders) were not significantly different from those of T0 cylinders for the variously treated soils (P = 0.96). The ratios of 15NO–3 recoveries in T1 vs. T0 cylinders were almost equal to one. In contrast, the 15NH+4 recoveries of T1 cylinders were significantly lower than those of T0 cylinders for various soil treatments (P < 0.001).

Microbial NO–3 immobilization rates according to 15N pool dilution calculations were not greater than zero (P > 0.05). Effects of the various N fertilizers and application rates on C and N mineralization rates and microbial NH+4 immobilization rates are given in Table 3. Carbon mineralization rates ranged from 5.6 to 12.9 mg C kg–1 soil d–1. Various N fertilizers and application rates significantly affected the C mineralization rates (P < 0.01). Soils treated with dairy-waste compost or lagoon effluent had higher C mineralization rates than the control soil or the soil treated with (NH4)2SO4. The highest C mineralization rate and the highest gross N mineralization rate were observed in the soil treated with high-rate dairy-waste compost. Microbial NH+4 immobilization was significant but the rates did not differ statistically among the treatments averaging 0.64 mg N kg–1 soil d–1 (Table 3).

Generally, nitrification potentials were significantly affected by the N application rates (P < 0.01), but not by the various N fertilizers (P = 0.50). Nitrification potentials were higher with the high-rate N fertilization than with the low-rate N fertilization (Table 4). The control soil had the lowest nitrification potential of 2.3 mg N kg–1 soil d–1, and the soil treated with high-rate compost had the highest nitrification potential of 8.1 mg N kg–1 soil d–1. The gross nitrification rate in the soil treated with high rate compost was 2.9 mg N kg–1 soil d–1, whereas rates in the other treated soils were <1 mg N kg–1 soil d–1 (Table 4).


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  		Table 4. Soil nitrification rates, potentials, and their ratios in various soil treatments about three months after N fertilization.

 
Silage Corn Yield and Plant Nitrogen Content
Effects of the N fertilizers and their application rates on silage corn yield and plant N content are given in Table 5. Corn planted in the soil treated with the high-rate compost had the highest ear leaf N and whole silage tissue N, followed by the soil with low-rate compost and high-rate ammonium sulfate. Silage corn yields were not statistically different in the soils treated with the various N fertilizers and application rates.


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  		Table 5. Effects of N fertilizers and their application rates on corn yields (dry wt. basis) and corn N contents.

 

   DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
CONCLUSIONS
 REFERENCES
 
Nitrogen Mineralization Kinetics
Aerobic compost and anaerobic lagoon systems produced wastes with significantly different proportions of inorganic N relative to total N, and the organic N has different mineralization kinetics. These differences are important considerations for determination of the application rate and timing for environmentally sound management of treated dairy wastes. The potentially mineralizable N of the compost comprised about 6% of its organic N. In contrast, the potentially mineralizable N of lagoon effluent comprised 30 or 90% of organic N for the low- and high-level addition, respectively. The discrepancy between the percentages of mineralizable organic N that were found for the two levels of lagoon effluent (30 vs. 90%) has several possible explanations. One explanation is that the capacity for microbial N immobilization may have become saturated in the high-level liquid waste treatment if microbial N immobilization was limited by C availability or the size of microbial biomass.

Previous studies have reported that 4 to 20% of dairy-waste compost N can generally be mineralized during a several-month incubation or in a growing season (Castellanos and Pratt, 1981; Hadas and Portnoy, 1994). Our result of 6% mineralizable dairy-waste compost N in 70 d was within that range. The relatively small mineralizable fraction indicated that more dairy-waste compost N remained in the soil, which would increase soil organic N. The remaining compost N may further provide plant available N in the subsequent growing seasons at a reduced rate. Eghball and Power (1999) reported that composted cattle manure released 15 and 8% manure N in the first and second year after the application, respectively. In contrast, lagoon effluent would have less effect on increasing soil organic N because a large proportion of lagoon organic N was mineralized in the short term. However, residual effects of lagoon effluent have been documented in several studies (Burns et al., 1990; King et al., 1990). While laboratory incubations analyzed with first-order models do not simulate field conditions, they can be used as indicators of the quality of the N source in the organic materials. Yet, the long-term residual effects of waste application needed further investigation.

Soil type and management history likely affect the decomposition kinetics of organic wastes. Although the soil used in the field differed from the one used in the lab incubation (Table 1), the decomposition kinetics of dairy wastes in the two soils seemed comparable. The field application rates of dairy wastes were well estimated from the lab incubation. As a result, the crop yields in waste-amended soils were equivalent to those with application of (NH4)2SO4 (Table 5). However, plant N contents from the dairy-waste lagoon effluent treatments were not significantly higher than those from the control plots. The relatively low plant N content in the lagoon effluent treatments may have been due to the actual rates of N application from lagoon effluent being lower than planned (Table 1).

Microbial Nitrogen Transformations
The fate of inorganic N, especially NO–3 during the growing season and after harvest, should be given consideration for environmentally sound N management. The relatively high NO–3 level vs. NH+4 (Fig. 1) indicates that soil NH+4 either from (NH4)2SO4 or from N mineralization of dairy waste was rapidly oxidized to NO–3. During the early growing season, the similar NO–3 levels in the soil treated with low- or high-rate dairy-waste lagoon effluent (Fig. 1) imply more N loss in the high-rate lagoon effluent treatment. Gaseous losses by NH3 volatilization and denitrification following the application of lagoon effluent might have been increased since effluent had a high pH and high NH+4 concentration (Table 1) favoring volatilization, and increased soil moisture would promote denitrification. When corn growth requires considerable N, soil inorganic N pool sizes would be expected to decrease. Indeed, soil inorganic N concentrations decreased to very low levels (Table 3), except for soil treated with high-rate compost. The accumulation of NO–3 in soil treated with high-rate compost (Fig. 1) suggests that available N exceeded corn N uptake. This statement could be further validated with the silage corn yield and plant N contents (Table 5). The increase in the application rate of compost did not benefit crop yield. Instead, high rate compost significantly elevated N concentration in the corn tissue and in postseason soils. The high level of NO–3 accumulating in soil after harvest may pose a potential leaching risk. Our observations suggest that the dynamics of N release from composted dairy waste may make balancing peak plant demand without risking postseason nitrate accumulation a difficult task.

Microbial N immobilization may immediately occur following the application of N fertilizers (Rice and Smith, 1984; Okereke and Meints, 1985). The rapid immobilization of inorganic N into organic forms would be important to reduce fertilizer N losses through leaching and denitrification during the early growing season. The amount of N immobilized by microorganisms, however, should be low enough so that microbial N immobilization does not deplete the soil inorganic N needed for crop growth. We have measured the microbial immobilization of NH+4–N and NO–3–N by 15N pool dilution techniques and NO–3 recoveries during the rapid N uptake phase of silage corn. Our results supported the null hypothesis that microbial NO–3 immobilization did not occur in this agricultural soil regardless of the N fertilization treatments at this time of the season.

In laboratory experiments using sieved agricultural soils, Recous and Mary (1990) have documented that microbes prefer NH+4 to NO–3 for their growth even in the case of high N ratio of NO–3 to NH+4 (110:5). Their results are consistent with previous studies in well-mixed agricultural soils (Jansson et al., 1955; Winsor and Pollard, 1956). Despite little microbial NO–3 immobilization in these laboratory experiments, significant microbial use of NO–3 has been observed in field experiments (Aulakh and Rennie, 1984; Recous et al., 1988; Burger and Jackson, 2003). In field situations, soil heterogeneity may lead to depleted NH+4 zones where microbes can use NO–3 for their growth (Davidson et al., 1990; Stark and Hart, 1997; Chen and Stark, 2000). Organic waste in soil could promote microbial NO–3 immobilization because of greater C availability or its heterogeneous distribution. However, in our field experiment where heterogeneous distribution of organic C is likely, microbial NO–3 immobilization was not observed in compost-treated soils. Lack of microbial NO–3 immobilization may be due to the low C availability relative to the NH+4 availability, as observed in laboratory experiments with well-mixed compost-amended soil (Shi and Norton, 2000).

Available C is a key factor limiting microbial N processes in agricultural soils. Significant microbial NO–3 immobilization has been observed in sieved agricultural soils when readily available C such as glucose or sucrose is added (Winsor and Pollard, 1956; Okereke and Meints, 1985; Recous and Mary, 1990) or in organic farming systems where organic materials including cover crops, compost, and plant residues have been annually amended to soils for several years (Burger and Jackson, 2003). The observation of higher C mineralization for the soil that received high-rate compost suggests that there was an impact on C availability. However, the lack of stimulation of microbial NH+4 immobilization or the use of NO–3 suggests that the readily available C was insufficient to support microbial growth. The lack of microbial NO–3 immobilization in this agricultural soil indicates that measurements of net nitrification rates (excluding plant roots) can give the similar information as gross nitrification rates.

Nitrifier population reflects the events occurring weeks to months before samplings (Berg and Rosswall, 1985). The higher nitrification potentials in soils with high-rate N fertilization (Table 4) probably resulted from higher NH+4 concentrations compared with those with low-rate N fertilization. Because soil NH+4 concentrations were not significantly different among the various treatments about 1 mo after N fertilization (Fig. 1), the higher nitrifier population activity in soil with high-rate N fertilization would be the residual effect of the higher NH+4 concentrations in the days shortly following fertilization. Except for soil treated with high-rate compost, other treated soils had low NH+4–N concentrations and N mineralization rates 3 mo after N fertilization (Table 3), which suggests that nitrifier population activity would be limited by the NH+4 availability thereafter. However, the higher ratio of nitrification rate to nitrification potential in soil treated with high-rate compost (Table 4) may imply that there was still relatively high available NH+4 for maintaining higher nitrifier activity.

The NH+4 available for nitrifiers 3 mo after N fertilization was mainly provided through the microbial decomposition of organic matter. Mineralized NH+4 can be used by heterotrophs and nitrifiers, but heterotrophs are stronger competitors than nitrifiers for available NH+4 (Jones and Richards, 1977). The nitrification rate in soil with the high-rate compost was about four times the microbial NH+4 immobilization rate (Tables 3 and 4), which indicates that readily available C limited microbial N assimilation rates in this agricultural soil. The highest C mineralization rates were associated with the highest N mineralization rate and nitrification rate (Tables 3 and 4) indicating a higher rate of organic matter turnover in soils amended with high rates of dairy-waste compost. The gross nitrification rates were higher than the gross N mineralization rates, resulting in a depletion of the NH+4 pool. Subsequently, the gross nitrification rates will be limited by the production of NH+4.


   CONCLUSIONS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
REFERENCES
 
Dairy-waste lagoon effluent supplied a large proportion of its total N as NO–3 several months after it was amended to an agricultural soil, whereas compost provided only a small fraction of its total N as mineral N. Differences in N availability in the wastes predicted by laboratory incubations were confirmed by silage corn yields and plant N contents. Corn silage yield was greatest in the compost treatment, showing that synchronization of N availability is as important as amount mineralized. Neither compost nor lagoon effluent had any effect on soil microbial NO–3 immobilization 90 d after their application. The lack of microbial NO–3 immobilization indicated that nitrification was a key process controlling soil NO–3 concentrations at this time. Gross N mineralization controlled subsequent nitrification because competition from heterotrophic NH+4 immobilization appeared to be limited by C availability. Higher C mineralization rates were associated with higher N mineralization and subsequent nitrification rates. Our observations suggest that peak plant demand is met by the compost N, but its continued high N mineralization potential after harvest makes the management of dairy compost N a difficult task. While this study demonstrated the short-term response of microbial N mineralization, immobilization, and nitrification to the application of organic wastes, the cumulative effects of application on microbial N transformations needs further investigation.


   ACKNOWLEDGMENTS
 
This project was financially supported by the USDA NRI #9600839 and Utah Agricultural Experiment Station. We thank Scott Perrin for his hard work in the field. We also thank the Isotope Analysis Laboratory of Utah State University for 15N analysis. Anonymous reviewers provided constructive comments and suggestions for the improvement of this manuscript. The first author was supported by NCARS during preparation of this manuscript.

Received for publication June 18, 2003.


   REFERENCES
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 CONCLUSIONS
 REFERENCES
 




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