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Soil Biology & Biochemistry

In situ Mineralization of Dairy Cattle Manures as Determined using Soil-Surface Carbon Dioxide Fluxes

Agriculture and Agri-Food Canada, 2560, Hochelaga Blvd., Sainte-Foy, QC, Canada, G1V 2J3

^* Corresponding author (rochettep{at}agr.gc.ca)

	ABSTRACT

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Decomposition rates of organic amendments in agricultural soils provide information on nutrient release and on changes in soil C stocks. They are usually determined using incubation of soil samples under controlled conditions. Our objective was to study decomposition of liquid and solid dairy cattle manures under field conditions following application to a loamy and a clay soil. Manures were applied in two consecutive years to silage corn at rates equivalent to 150 kg ha^–1 total N providing 119 to 202 kg C ha^–1. Soil-surface CO₂ fluxes, temperature and water content were monitored weekly in manured and control (mineral fertilizer) plots. Initial decomposition rates following manure application were much slower in the first than in the second year. This lag phase was attributed to cool soil temperatures and to the time required by the soil microflora to adapt to a first addition of this exogenous organic substrate. During the first year, mineralization of manure organic C was lower in the clay (52% of manure C) than in the loamy (75%) soil, presumably as a result of less favorable aeration conditions and greater physical protection of organic substrates in the clay. Equal mineralization rates of liquid and solid manures during both years in the loamy soil indicated that the two manure types had similar decomposability. Slower decomposition of the solid than the liquid manure in the clay soil was attributed to poorer mixing with the soil (presence of clods) resulting in reduced access to the solid manure for soil decomposers. The results of this study evidenced that under field conditions manure C decomposition is not determined solely by manure characteristics but also by interactions between the manure type and soil physical properties such as texture.

Abbreviations: DIC, dissolved inorganic C • DOC, dissolved organic C • DTC, dissolved total C

	INTRODUCTION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
DAIRY CATTLE PRODUCE approximately 25 x 10⁶ Mg of manure annually in Canada (Hofmann and Kemp, 1996). While a certain quantity of this manure can be exported outside of the farm, most of it is returned to the soil following a period of storage. Knowledge of decomposition rates of animal manure in agricultural soils is needed to estimate nutrient release (Castellanos and Pratt, 1981) and changes in soil C stocks (Helgason et al., 2005).

Decomposition of animal manures in soils incubated in the laboratory reflects the nature and relative abundance of their C fractions. Rate of C mineralization is often maximum during the first few days after incorporation when readily available substrates are being metabolized (Kirchmann, 1991). This is especially true for liquid manures that usually contain greater amounts of volatile fatty acids than solid manures (Kirchmann and Witter, 1992). Following that period, the decomposition rates remain relatively constant and the cumulative CO₂–C evolved increases linearly with incubation time as more complex organic molecules are being metabolized (Bernal and Kirchmann, 1992; Kirchmann and Lundvall, 1993). Accordingly, manure decomposition dynamics during laboratory incubations are often fitted to a combination of first- and zero-order models (Kirchmann, 1991; Bernal and Kirchmann, 1992) depending on the proportion of total manure C that is easily decomposable.

Laboratory incubations of amended soil provide useful information on the relative decomposability of manure. However, the extrapolation of these estimates to decomposition rates under field conditions is difficult because of the temporal and spatial variations of most factors controlling decomposition in the field. Under field conditions, decomposition of organic residues has been measured based on the recovery of applied organic matter (Chantigny et al., 1999; Burgess et al., 2002). However, the precision of the estimates obtained using this approach is often decreased by the large amounts and spatial variability of soil C. Few attempts have been made to estimate manure decomposition in situ based on the manure-induced CO₂–C emissions. Gregorich et al. (1998) and Rochette and Gregorich (1998) reported that the manure CO₂–C losses in a corn field that was receiving annual applications of solid dairy cattle manure corresponded to 55% of the annual C additions during the second year and 60% during the third year. In other studies, Chantigny et al. (2001) measured that 34% of liquid hog manure C was lost 28 d after application, while Rochette et al. (2000) estimated that manure CO₂–C losses in a corn field receiving various rates of liquid hog manure for the 19th consecutive year were approximately equal to the annual manure C inputs.

Direct comparisons of the decomposition of liquid and solid manures under field conditions are lacking. In situ decomposition rates of liquid manure (Rochette et al., 2000) were faster than those of solid manure (Rochette and Gregorich, 1998). However, interpretation of these curves is confounded by differences in soil type and climate conditions between these two studies. The objective of our study was to quantify the in situ decomposition of liquid and solid dairy cattle manures following application to a loamy and a clay soil.

	MATERIALS AND METHODS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
The field experiment was conducted in 2002 and 2003 at the Harlaka Research Farm of Agriculture and Agri-Food Canada, located 5 km southeast of Québec City (46°48' N, 71°23' W), Canada, on two soils: a Kamouraska clay (fine, mixed, frigid, Typic Humaquept) and a Saint-André gravelly loam (loamy, mixed, frigid, Typic Dystrochrept) located 200 m apart in the same catena. The soils were sampled (0–10 and 10–30 cm) before the experiment and analyzed for particle-size distribution, pH (1:2, soil/water ratio), and total C and N (Leco CNS-1000, Leco Inc., St. Joseph, MI) (Table 1). Soils were moldboard plowed to 0.2 m in late October 2001 for the clay and on 29 Apr. 2002 for the loamy soil. Forage crops (grasses) were grown on both soils in 2000 and 2001.

Gas fluxes were measured weekly from 22 May to 4 Nov. in 2002 and from 5 May to 24 Nov. in 2003, except during the first 2 wk following application of amendments when four samplings were done. At the time of sampling, the frames were covered by a 0.14-m high square plexiglass chamber (vented and insulated) covering the same area as the frames. Chamber headspace air samples (25 mL) were taken through a rubber septum at regular intervals (0, 6, 12, and 18 min after deployment) and injected into pre-evacuated vials (12-mL Exetainer, Labco, High Wycombe, UK) using a polypropylene syringe (20 mL; Becton Dickinson, Rutherford, NJ). The air samples were stored at approximately 200 kPa and analyzed within 2 wk of sampling by mean of a gas chromatograph fitted to a nickel-nitrate (10%) catalyst column and flame ionization and thermal conductivity detectors (Model 3800, Varian Inc., Walnut Creek, CA) and equipped with a headspace autoinjector (Combi Pal, CTC Analytics, Zurich, CH).

Soil-surface CO₂ fluxes (F_C; mg m^–2 s^–1) were calculated using the following equation (Rochette and Hutchinson, 2005):

where dC/dt (mol CO₂ mol^–1 s^–1) is the rate of change of chamber gas concentration in dry air samples, V (m³) is the chamber headspace volume, A (m²) is the surface area covered by the chamber, Mm (mg mol^–1) is the molecular weight of CO₂ (44000 mg), Mv (m³ mol^–1) is the molecular volume at predeployment air temperature (0.022–0.024 m³ mol^–1), e_a (kPa) is the predeployment partial pressure of water vapor, and P (kPa) is the barometric pressure. Flux measurements were generally made between 900 and 1200 h (eastern standard time). Contribution of manure C to F_C was estimated as the difference in F_C between manured and control plots.

Concentrations of CO₂ in the soil profile were monitored within 1 m of the chamber frames. Soil air samples were taken simultaneously with flux measurements in each plot using gas probes. The probes were made of plastic mesh cylinders (10 cm length; 3.5 cm diameter) containing 3-mm diameter glass beads. The cylinders were inserted horizontally at depths of 5 and 20 cm, and connected to the surface using plastic tubes (0.60 m length; 6.35 mm o.d.; 3.18 mm i.d.; Bev-a-line IV, Ryan Herco Industrial Plastics, Seattle, WA) with two-way Luer-type stopcock valve (Cole Parmer, Vernon Hills, IL) and a silicone septum (male luer-lock stopper with injectable membrane, Vygon, Ecouen, FR) at the surface end. Air samples were collected through sleeve stoppers with a 20-mL Plastipak syringe after expelling 20 mL of air from the tubes to account for the dead volume of the tubes. Samples were handled, stored and analyzed as the chamber air samples.

Soil temperature and moisture were measured next to each frame at the time of flux measurements. Soil temperatures were monitored at the 5- and 20-cm depths using copper-constantan thermocouples. Average soil moisture in the top 10 cm and between 15 and 25 cm was measured with 15-cm three-bar time domain reflectometry probes inserted at 45° in soil.

Soil Analyses
Soil composite samples were collected from the 15-cm surface soil layer at four times (beginning of July and early Oct. 2001 and 2002) to determine the effects of treatments on soil microbial biomass and aggregation. Samples were sieved at 6 mm in the field and stored immediately at 3°C. Soil microbial biomass C (MBC) measurements were performed within 24 h of sampling using the chloroform-fumigation extraction technique with an extraction efficiency (K_ec factor) of 0.45 (Wu et al., 1990). Two 20-g subsamples of field-moist soils were placed in 100-mL beakers. One subsample was fumigated for 24 h in a vacuum desiccator containing 25 mL of CHCl₃. The other subsample was kept in the dark at 3°C for 24 h. Both fumigated and nonfumigated soils were extracted with 40 mL of 0.25 M K₂SO₄. After shaking for 1 h on a reciprocal shaker, the suspensions were centrifuged at 1000 x g and filtered (Whatman #934-AH). The organic C content of the extracts was determined with a combustion TOC Analyzer (Model Formaacs, Skalar Analytical, De Breda, The Netherlands). Specific respiration activity of soil microbial biomass was calculated as the quotient of respired CO₂ to MBC.

The size distribution of water-stable aggregates was measured by spreading 40 g of field-moist soil (<6 mm) on the top of a 2-mm sieve (Angers and Mehuys, 1993). The aggregates were wetted by direct immersion and wet-sieving was performed under total immersion for 10 min. Aggregates on the top of the 2-mm sieve were oven-dried, weighed, and expressed as a percentage of total soil on an oven-dry basis. A correction was made for the presence of sand and coarse fragments in the stable aggregates.

Manure Analyses
Measurements of pH were made directly in the liquid manure and in a 5:1 water/manure ratio for solid manure. Manure mineral N was extracted by agitation with 1 M KCl (KCl/manure ratio of 12:1 for liquid manure, and 6:1 for solid manure) for 1 h, centrifugation at 3000 x g and filtration (Whatman #42). Concentrations in NH₄⁺, NO₃^–, and NO₂^– were determined in the extracts by flow injection analysis (Model QuickChem FIA+ Series 8000, Lachat Instruments, Milwaukee, WI). Manure total N and P were obtained by Kjeldahl digestion followed by determination of NH₄⁺ and PO₄^3– by flow injection analysis. Manure dissolved C was extracted by shaking with water (4°C; 4:1 water/manure ratio for both manures) for 30 min, centrifugation at 15000 x g and filtration at 0.45 µm (Millipore, HVLP type nylon membrane). Dissolved total (DTC) and inorganic C (DIC) were measured in the extracts with a TOC analyzer (Model Formaacs, Skalar Analytical, De Breda, The Netherlands). Dissolved organic C (DOC) was obtained by subtracting DIC from DTC. The water extracts were also analyzed for volatile fatty acids (VFA) by direct determination of acetate, propionate and butyrate on a gas chromatograph (Model 3400, Varian Inc. Walnut Creek, CA) (Paul and Beauchamp, 1989). Total manure C was determined by dry combustion (Model CNS-1000, LECO Corp., St. Joseph, MI) on dry matter obtained after drying at 60°C.

Statistical Analysis
One-way analysis of variance on the effects of manure type on soil CO₂ concentration, soil-surface CO₂ flux, soil temperature, and soil water content was performed for each sampling date using the General Linear Model (GLM) procedure of SAS (SAS Institute, 1999). All experimental error variances were tested for homogeneity using Bartlett's test and values were log-transformed when needed to achieve normality. Treatment effects were considered statistically significant at p < 0.05.

	RESULTS AND DISCUSSION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Soil-surface Carbon Dioxide Emissions
The seasonal pattern of CO₂ emissions in the control plots (Fig. 1b and 2b ) indicates that the emissions were mostly controlled by soil temperature (Fig. 3 ) and crop development except in August 2002 when exceptionally low soil water contents (Fig. 4 ) resulted in low respiration rates despite warm soil conditions (Fig. 3). In this study, soil-surface CO₂ fluxes in the control plots are assumed to estimate background emissions originating from sources other than the mineralization of the dairy cattle manures—mostly soil organic matter decomposition and root-rhizosphere respiration of the maize plants. Background CO₂–C emissions cumulated over the 2 yr were higher in the loamy than in the clay soil (Table 3) despite higher soil organic C (Table 1) and microbial biomass in the clay (Fig. 5b and d). This resulted in a specific respiration that was more than twice greater in the loamy (204 g CO₂–C g^–1 MBC d^–1) than in the clay (82 g CO₂–C g^–1 MBC d^–1) soil. The more rapid decomposition of soil organic matter in the loam in this study may have resulted from a combination of warmer soil temperature, better aeration, and greater substrate accessibility as clays are known for their protective capacity (Ladd et al., 1996).

View larger version (34K): [in this window] [in a new window]   Fig. 1. Soil CO2 concentration at 5 cm, soil-surface CO2 flux and cumulated mineralized manure C in a clay soil. Arrows indicate dates of manure application. Error bars represent LSD0.05 for treatment comparison at each date.

View larger version (35K): [in this window] [in a new window]   Fig. 2. Soil CO2 concentration at 5 cm, soil-surface CO2 flux and cumulated mineralized manure C in a loamy soil. Arrows indicate dates of manure application. Error bars represent LSD0.05 for treatment comparison at each date.

View larger version (24K): [in this window] [in a new window]   Fig. 3. Soil temperature at 5 and 20 cm in a clay and a loamy soils at the time of CO2 flux measurements (9 to 12 EST). Arrows indicate dates of manure application.

View larger version (28K): [in this window] [in a new window]   Fig. 4. Soil water content in the 0- to 10- and 15- to 25-cm soil layers in a clay and a loamy soils at the time of CO2 flux measurements (9 to 12 EST). Arrows indicate dates of manure application.

View larger version (27K): [in this window] [in a new window]   Fig. 5. Fraction of soil aggregates >2 mm and soil microbial biomass C in the top 15 cm of a clay and a loamy soils. Values are means of values (n = 16) obtained at four sampling dates during the experiment. Different letters indicate significant differences (p < 0.05) according to LSD test.

For the two growing seasons, manure application resulted in significantly greater CO₂ emission rates at several dates (29 in the loam and 19 in the clay) compared with the mineral fertilizer (Fig. 1b and 2b). As expected, manure-induced emissions were larger and more frequent early in the season and decreased with time after application. Emissions in the manure-treated plots generally followed a temporal pattern similar to that in the control plots except shortly after manure application in 2003 when a 10-d peak only occurred in the manured plots. Manure type can influence the response of soil-surface CO₂ emissions to manure application. Application of solid or composted manures to agricultural soils resulted in increased CO₂ emissions that were sustained for several weeks but without clear peak shortly after application (Gregorich et al., 1998; Rochette and Gregorich, 1998). In contrast, post-application bursts in CO₂ emissions were observed immediately following application of liquid manure to agricultural soils (Rochette et al., 2000; Chantigny et al., 2001; Rochette et al., 2004) and were attributed to the decomposition of the readily available manure C fractions (Kirchmann and Lundvall, 1993). Therefore, the absence of a clear post-application peak in the liquid manure plots in 2002 indicates that factors (biological or physical) other than substrate availability limited the activity of decomposers, while the presence of a peak in 2003 in the solid manure plots confirms that the manure was rather fresh and contained readily decomposable organic matter.

Cumulated CO₂–C emissions, normalized using the amounts of manure C applied, showed that 40 to 76% of manure C was lost as CO₂ in 2002 (Table 3). These estimates are in the same range as those observed (using other techniques) under Canadian conditions in the first year after the application of various organic residues such as straw (50–65%, Voroney et al., 1989), corn residues (45–80%, Gregorich and Ellert, 1994; Burgess et al., 2002), and paper mill residues (40 –50%, Chantigny et al., 1999; Fierro et al., 2000). Cumulated CO₂–C emissions were greater in 2003 (54–109%; Table 3) than in 2002 likely because of the cumulative effects of manure application in two successive years. Losses above 100% observed in 2003 in the loam could be explained by the continued decomposition of the manure applied in 2002 as well as by the stimulation of native soil organic matter decomposition (Bernal and Kirchmann, 1992; Liang et al., 1996). That greater fractions of the manure C were mineralized in the loam than in the clay suggests that a new equilibrium between C inputs and CO₂–C losses would be reached more rapidly and result in lower C gains in the loamy soil following annual applications of manure.

Manure Decomposition Rates
Manure C mineralization was slow shortly after the application of manures in both soils in 2002 as indicated by the slope of the accumulation of normalized manure-induced CO₂ emissions (Fig. 1c and 2c). Low decomposition rates at a period when organic substrates were most abundant indicate that the activity of decomposers was limited by other factors. Relatively cool soil temperatures (10°C) (Fig. 3) may have contributed to the slow decomposition of manures during the first 20 d following manure application in spring 2002. Also, the soil microbial biomass is largely dormant in situations with no or little addition of exogenous organic substrates (Blagodatsky et al., 2000) and lag phases were observed between the addition of simple substrates and the exponential growth of active microbial populations (Van de Werf and Verstraete, 1987). The two soils at this experimental site had not received manure or other organic amendments at least in the past 15 yr and the soil microflora likely required some time to efficiently metabolize this large input of new organic substrates. Finally, the more rapid response of decomposers to manure application in 2003 than in 2002 could be attributed to the conditioning effect of the 2002 manure application on decomposers and to the warmer soil temperature following manure application in 2003 (Fig. 3).

Following the lag phase in 2002, decomposition rates were relatively constant and varied from 0.004 g CO₂–C g manure C^–1 d^–1 for the solid manure in the clay soil to 0.008 g CO₂–C g manure C^–1 d^–1 for both manures in the loam. Very dry soil conditions (Fig. 4) nearly halted decomposition rates in August. In 2003, decomposition rates were maximum shortly after manure application and generally decreased with time. Fluctuating rates of CO₂ emissions in the clay soil receiving liquid manure during the summer months could be attributed to abundant rainfall (data not shown), which resulted in frequent changes in soil water content (Fig. 4) and transient effects on soil gas diffusivity and decomposers activity (Schjønning et al., 2003). Large date to date variations in soil CO₂ concentrations during that period (Fig. 1a) are consistent with frequent changes in soil gas diffusivity. It is unclear why these effects were only apparent in the clay soil with liquid manure.

On average, decomposition of liquid and solid manures was greater in the loamy than in the clay soil in both years (Table 3). Lower activity of decomposers in soils with higher clay contents has been attributed to restricted aeration (Bouma and Bryla, 2000) and to the association of soil organic matter with clay particles that physically protects organic substrates against decomposers (Oades, 1988; Ladd et al., 1996). However, contradictory impacts of soil texture on soil C dynamics have been reported. Soil C mineralization rates were shown to be either insensitive (Gregorich et al., 1991; Hassink, 1994; Franzluebbers et al., 1996; Franzluebbers, 1999) or negatively affected (Merckx et al., 1985; Bouma and Bryla, 2000; Wang et al., 2003) by increasing clay content. That greater manure decomposition is observed in the loam where microbial biomass C (Fig. 5b and 5d) is lower than in the clay soil indicates that a larger fraction of the population of decomposers is active in the loamy than in the clay soil. These field observations are in agreement with laboratory studies that show higher specific respiratory activity in a loam (10% clay) than in a clay soil (40% clay) (Franzluebbers et al., 1996).

Unlike in the loamy soil, liquid manure decomposed faster than the solid manure in the clay soil (Fig. 1c and 2c). This suggests that interactions between manure and soil types affected the decomposition process. Slower decomposition of the solid manure in the clay soil may be explained by poorer mixing with the soil (presence of clods). This resulted in reduced contact area between the solid residue and the soil particles, and thereby in reduced access of decomposers to the substrate and lower microbial growth (Angers and Recous, 1997; Henriksen and Breland, 2002). Accordingly, in the clay, soil microbial biomass C content was greater in the liquid than in the solid manure plots (Fig. 5b). Also, the incorporation of more decomposable organic materials in soil will usually induce greater and more rapid increases in aggregate stability than recalcitrant materials, which may have a smaller but longer lasting effect (Martin and Waksman, 1940; Monnier, 1965). Therefore, the higher levels of microbial biomass C and soil aggregation (Fig. 5a and 5b) in the clay soil receiving liquid than solid manure reflected that liquid manure was more easily decomposed in this heavy soil.

Under field conditions, most of the factors affecting manure decomposition vary in time and space, resulting in complex interactions that rarely allow the isolation of the effect of a specific factor. It is therefore difficult to conclude on the relative decomposability of both manure types used in this study. Differences in decomposability between manure types should be more evident in the loam where the abiotic factors likely have less impact on decomposition processes. Based on this assumption, similar temporal patterns of decomposition rates as well as similar annual cumulative decomposition for both manures in the loamy soil suggest that the decomposability of both organic materials was similar. This was unexpected considering that solid manures usually contain more slowly degradable bedding materials than the liquid manure and had a higher C/N ratio (Table 2). Furthermore, while aerobic storage of solid manure promotes formation of more stable organic material (Wichern et al., 2004), anaerobic storage of liquid manure results in the formation of C fractions that are easily decomposed when exposed to a well-aerated soil environment (Bernal and Kirchmann, 1992). Accordingly, the soluble C and especially the volatile fatty acids represented a much greater fraction of the liquid than of the solid manure C (Table 2). Similar decomposition rates for the liquid and solid manures in the loamy soil is likely explained by the fact that the solid manure was relatively fresh and contained considerable amounts of easily available substrates.

	CONCLUSIONS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Measurement of soil-surface CO₂ fluxes under field conditions allowed to determine the net effect of soil type, climate, and manure composition on the decomposition of liquid and solid dairy cattle manure. Following application of both manures in the first year, a lag phase was observed at the initiation of decomposition, which was attributed to cool soil temperatures and to the time required by the soil microflora to adapt to a first addition of this exogenous organic substrate. On average, manure C decomposed more rapidly in the loamy than in the clay soil, presumably as a result of better aeration and lower physical protection of organic substrates by clay minerals in the loam. Equal mineralization rates of liquid and solid manures during both years in the loamy soil indicated that the two manure types had similar decomposability. Slower decomposition of the solid than the liquid manure in the clay soil was attributed to poorer mixing of solid manure with the soil (presence of clods) resulting in reduced access of decomposers to this substrate. The results of this study show that in situ monitoring of soil-surface CO₂ emissions can be used to assess the mineralization of liquid and solid organic amendments under field conditions. This approach evidenced that under field conditions manure C decomposition is not determined solely by manure characteristics but also by interactions between the manure and soil physical properties such as texture.

	ACKNOWLEDGMENTS

 
The authors thank Patrice Jolicoeur, Gabriel Lévesque, Johanne Tremblay, Nicole Bissonnette and Juan Pablo Soucy for their excellent technical assistance, and the technical staff of the Institut de recherche et de développement en agroenvironnement for providing the manures. This research was funded by the AAFC Model Farm greenhouse gas research program.

Received for publication July 22, 2005.

	REFERENCES

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 

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