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Agriculture and Agri-Food Canada, Lethbridge Research Centre, P.O. Box 3000, Lethbridge, AB, Canada T1J 4B1
* Corresponding author (haoxy{at}agr.gc.ca).
Received for publication December 27, 2002.
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ABSTRACT |
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 TOPNOTES ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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Abbreviations: DM, dry matter • GHG, greenhouse gas • SBM, straw-bedded manure • TC, total carbon • TN, total nitrogen • WBM, wood chip-bedded manure
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INTRODUCTION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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In southern Alberta, Canada, cattle feedlots are increasing in size and in animal density with large amounts of solid manure produced on a relatively small land base. Therefore, composting is an attractive alternative to direct land application of fresh manure since both the weight and volume, and hence hauling cost, are reduced considerably (Larney et al., 2001), as are coliform bacteria (Larney et al., 2003) and weed seeds (Larney and Blackshaw, 2003).
Recently, the lumber industry has been promoting the use of wood residuals as an alternative bedding material to traditional cereal straw. Wood chip bedding is a mixture of bark, post peelings, and sawdust with a greater portion of materials <5 mm in size (Ward et al., 2000). In addition, wood chips offer greater water holding capacity (Ward et al., 2000), require less frequent addition as bedding (Ward et al., 2001), and keep animals cleaner (McAllister et al., 1998) compared with straw. Moreover, recent drought on the Canadian prairies has decreased the supply (and increased the cost) of straw for livestock bedding, making wood chips a more economically viable alternative.
Bedding materials also affect the physical and chemical properties of fresh feedlot manure and its composted end-product (Larney et al., 2001, 2002). While the effect of bedding materials on emissions of CO2, N2O, and CH4 during composting of feedlot manure has not been widely investigated, GHG emissions are known to occur (Hao et al., 2001; Sommer and Møller, 2000). The amount and the proportion of these GHG emissions and the quality of the final compost product may be affected by several factors such as water content, C/N ratio (Shi et al., 1999; Al-Kanani et al., 1992), and chemical amendments (Swinker et al., 1998; Mahimairaja et al., 1994). The presence of wood chips increases convection of air through the compost windrow (Van Ginkel et al., 2002; Barrington et al., 2003), possibly increasing the supply of O2, and hence increasing aerobic decomposition. In other words, increases in CO2 and decreases in CH4 and N2O production might be possible. In this study, we investigated C and N balance and emissions of CO2, CH4, and N2O from straw and wood chip–bedded cattle manure during open windrow composting in southern Alberta.
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MATERIALS AND METHODS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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0.5 km from the composting facility.
Physical and Chemical Analyses and Mass Balance
At establishment and just before each turning, a series of measurements and samplings were taken from each windrow. The surface area and volume were calculated from circumference and radius measurements of the windrows, assuming the windrow cross-section was hyperbolic in shape. Bulk density was estimated by weighing an aluminum pail of known volume (0.064 m3) filled with manure–compost (Larney et al., 2001). Particle density was measured using a pycnometer method (Blake and Hartge, 1986) as modified and described by Hao et al. (2001). Total porosity in the windrow was calculated using bulk density and particle density. The water content of the manure–compost samples was determined by oven-drying at 60°C to constant weight. The air-filled pore space was calculated as the difference between total porosity and pores occupied by water.
Before turning, each windrow was cut perpendicular to its length with a skid-steer loader, exposing two vertical faces. Manure samples of approximately 1 kg were collected from each face at depths of 0 to 5, 5 to 15, 15 to 35, 35 to 60, 60 to 90, and 90 cm from the top of the windrow. This represented 24 samples from each bedding treatment. The 0- to 5-cm sampling depth included the windrow peak. Ammonium-N (NH+4–N) and nitrate-N (NO–3–N) concentrations were measured on 10 g of fresh (wet) sample from each depth. Each was shaken in 200 mL of 2 M KCl for 1 h, and filtered through KCl-washed filter paper (Whatman no. 42). Extracts were frozen at –15°C, and then analyzed on a Technicon AutoAnalyzer II. Subsamples were oven-dried at 60°C and finely ground (<150 µm) for total C (TC) and total N (TN) measurement in an automated elemental analyzer (Carlo Erba, Milan, Italy). The pH was measured using oven-dried materials by shaking 30 g (dry wt.) of manure–compost with 120 mL of deionized water for 1 h, then the pH values of the filtrate were determined with a pH meter (Accumet pH meter 50, Fisher Scientific). Characteristics of the SBM and WBM at the begining of the experiment (fresh manure, Day 0) and on Day 99 (end of thermophilic composting) are shown in Table 1.
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Gas Profile Measurements within Compost Windrows
Gas profile measurements were collected two to three times per week for the first 3 wk and once per week thereafter for the 99 d of composting. Two sets of gas samples were collected at the windrow surface (0 cm) and at 15, 40, 70, and 100 cm below the surface using a multilevel sampler (Hao et al., 2001). All gas samples were taken between 0800 and 0900 h and analyzed for CO2, CH4, N2O, and O2 on the same day using a gas chromatograph (Varian 3600, Varian Instruments, Walnut Creek, CA) equipped with an electron capture detector (ECD), flame ionization detector (FID), and thermal conductivity detector (TCD). The average concentrations of CO2, CH4, N2O, and O2 for each windrow were calculated based on the geometric mean of each sampling section, assuming the windrows were hyperbolic in shape. The geometric mean was used since the volume of manure represented by each sampling depth was not equal.
Gas Fluxes from Windrow Surfaces
Gas emissions and O2 consumption during composting were measured according to the same schedule as gas profile measurements, using a modified vented chamber technique (Hutchinson and Mosier, 1981). Two sets of gas samples were collected from each windrow. At each sampling time, a chamber (15.5 cm in diam. and 15 cm in height) was placed on the peak of the windrow. Ten mL of air was drawn with a plastic syringe from the chamber headspace 0, 5, 10, 20, and 30 min after chamber placement. Immediately after sampling, the syringe needle was placed in a rubber stopper to prevent gas exchange. As with profile measurements, all surface flux gas samples were taken between 0800 and 0900 h and analyzed for CO2, CH4, N2O, and O2 the same day.
Gas fluxes were calculated from concentrations by assuming a steady state gradient in the underlying windrows (Anthony et al., 1995) as described by Hao et al. (2001). Briefly, the concentration vs. time relationships for each chamber were fitted with a second order polynomial equation (C = a + bx + cx2), where C is the concentration of gases and x is the time in minute(s) for each sampling time (SAS Inst., 2001). The flux at time 0 was calculated by taking derivatives of the second order polynomials (dC/dxx
0 = b). Cumulative emissions were approximated by assuming that daily fluxes, which were measured more frequently during early composting, represented the average for the whole week. To account for emissions during turning, it was assumed that these were equivalent to the difference between the average profile gas concentration before and after turning and the average air-filled pore space before the turning event. These amounts were added to the total cumulative emissions. The total GHG emissions during the composting period were expressed on an initial surface area (kg C m–2 or kg N m–2 of manure) and initial dry weight (kg C Mg–1 manure or kg N Mg–1 manure) basis.
Data Analysis
The O2, CH4, and CO2 flux data were analyzed for two separate composting stages: Day 0 to 49 (early) and Day 50 to 99 (late). Three stages (Day 0–14, Day 15–49, and Day 50–99) were used for N2O due to its distinctive flux patterns. The average rate of O2 consumption and GHG emission during different stages of composting was obtained by covariance analysis, treating time as the covariant using the PROC MIXED procedure (SAS Inst., 2001). When treatment effects were significant, mean flux among different compost stages and between two bedding treatments were tested using contrast analysis.
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RESULTS AND DISCUSSION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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Dry Matter, Carbon, and Nitrogen Changes during Composting
The initial water content of fresh SBM and WBM was similar (Table 1) and close to 60% (wet wt.). Water content decreased steadily to 28% (SBM) and 37% (WBM) by Day 99. Windrow turning and hot, dry weather accelerated evaporation and water loss. The DM mass loss during 99 d of composting was 30.1% of initial mass for SBM and 26.8% of initial mass for WBM (Table 2). Therefore, compost yield was 0.699 Mg Mg–1 for SBM and 0.735 Mg Mg–1 for WBM.
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The NH+4 concentration in the windrows increased during the first 35 d of composting (Fig. 1c), when the rate of NH+4 production due to intense organic matter degradation (Sánchez-Monedero et al., 2001) exceeded the rate of NH+4 loss through nitrification of NH+4 to NO–3 and gaseous NH3 emission loss since NH4 and NH3 are in equilibrium in compost solution. In addition, the higher rates of DM mass loss also contributed to increased NH+4 concentration during early composting. The average concentration of NH+4 increased from the initial values of 2090 mg kg–1 (SBM) and 2666 mg kg–1 (WBM) to around 3000 to 4000 mg kg–1 for the first 35 d, then decreased steadily afterward reaching their lowest values of <1000 mg kg–1 on Day 99 (Table 1 and Fig. 1c).
The NO–3–N concentration decreased sharply from the initial 244 mg kg–1 (SBM) and 182 mg kg–1 (WBM) to <40 mg kg–1 for both treatments on Day 8. This decrease continued, reaching a minimum of 10 mg kg–1 on Day 21 (WBM) and on Day 35 (SBM). The NO–3 content remained low for the next 40+ d, then increased again to 135 mg kg–1 (SBM) and 330 mg kg–1 (WBM) on Day 99 (Table 1 and Fig. 1d).
The initial NO–3–N content of cattle manure in our study was much greater than values (<10 mg kg–1) from previous studies in southern Alberta (Larney et al., 2001; Hao et al., 2001). Normally, cattle manure is removed from feedlot pens either in fall (October–November) or spring (May–June). In our study, manure removal was delayed until mid-July, which allowed an extra 1 to 2 mo of decomposition in the feedlot pen at a time when average air temperatures were approaching their annual peak. Manure decomposition and nitrification of NH+4 was enhanced during this period, which contributed to the higher initial NO–3 concentration for the manure in our study.
Oxygen Consumption and Greenhouse Gas Emissions
The rates of O2 consumption varied considerably during composting (Fig. 2a)
and were significantly greater for SBM than WBM during early composting (Day 0–49). In contrast, no differences between the two treatments were found during late composting (Day 49–99) (Table 3). Contrast analysis also indicated that the O2 consumption rate was significantly greater during the early stage, compared with the late stage of composting, for SBM but not for WBM (Table 3, Fig. 2a). The rate of O2 consumption significantly decreased following windrow turning on Days 8 and 14 (Fig. 2a), mainly due to decreased windrow temperatures (temperature data not shown).
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The increased rate of O2 consumption and CO2 emission from SBM compared with WBM was probably caused by differences in manure characteristics. Cereal straw with less recalcitrant lignin and more readily degradable hemicellulose than wood chip materials (Eklind and Kirchmann, 2000a; Ward et al., 2000), combined with smaller-size particles of the straw compared with the larger wood particles and a lower initial C/N ratio of 17 (Table 1), favored microbial decomposition and hence SBM had higher O2 consumption and CO2 emissions. The higher O2 consumption and CO2 emission during early composting may be attributed to the greater degradation rate of aliphatics, hemicellulose, and proteins during early stages of composting (Veeken et al., 2001). As this easily degradable fraction of organic matter was depleted, O2 consumption and CO2 emission declined.
More than 50% of total CH4 emission occurred during the first 28 d of composting, decreasing rapidly to near zero after Day 70 for both treatments (Fig. 2c). This is similar to the findings of Lopez-Real and Baptista (1996), Sommer and Møller (2000), and Hao et al. (2001). Initially, relatively high rates of O2 consumption and CO2 production (Fig. 2a and 2b) created anaerobic conditions that not only favored production of CH4, but also increased CH4 stability. Although correlation coefficients for the rate of O2 consumption vs. CH4 emissions (r = 0.45* for SBM; r = 0.49* for WBM, n = 32; * significant at the 0.05 level) were significant, they were lower than those obtained for O2 consumption vs. CO2 emission. This may be partly explained by factors affecting production and emission of CO2 vs. CH4. Besides the rate of O2 supply, the rate of CO2 production and emission directly depends on the C source in the composting substrate. For CH4, in addition to C availability in the substrate, other factors, such as the rate of O2 consumption during the production of CO2 and the stability of the CH4 produced, affect the rate of CH4 emission. For example, while production of CH4 may be high at the bottom and center of the windrow profile where anaerobic conditions exist, this CH4 could be oxidized while diffusing upward and outward before reaching the emitting surface (Hao et al., 2001).
The rates of CH4 emission were not significantly different between SBM and WBM, so all data were pooled in a contrast analysis of the composting stage effect. The CH4 emission rate (0.0219 kg m–2 d–1) during early composting (Day 0–49) was significantly greater than late composting (Day 50–99, Table 3).
The highest rates of N2O emission were measured at the onset of composting (0–14 d), followed by a period of low emissions during midcomposting and a minor peak toward the end (Fig. 2d). In contrast, Martins and Dewes (1992), Lopez-Real and Baptista (1996), and He et al. (2001) reported that N2O emission occurred only in the later stages of composting when CH4 production had ceased. Our results suggest that the presence of NO–3 in the manure led to the production and emission of N2O during early composting when the emission of CH4 was also largest. A significant correlation between the rate of N2O emission and average NO–3 concentration in the compost (r = 0.79*** for SBM and r = 0.80*** for WBM, n = 16) suggest that denitrification of NO–3 rather than the nitrification of NH+4 was responsible for N2O emissions in our study.
Since there were three distinct periods of N2O emission (Fig. 2d), the rates were analyzed separately for the initial major peak (Day 0–14), the midperiod of lower emission (Day 15–49), and the later minor peak (Day 50–99). There were no significant differences due to bedding in N2O emission rates for Day 0 to 14 or Day 15 to 49 (Table 4). However, SBM had a significantly greater N2O emission rate from Day 50 to 99 than WBM. For the SBM treatment, the average N2O emission rate of 0.200 g m–2 d–1 during the first 14 d was significantly greater than the rate from Day 15 to 49 (0.027 g m–2 d–1), but not significantly different from the rate from Day 50 to 99 (0.144 g m–2 d–1, Table 4). For the WBM treatment, the average N2O emission rate of 0.225 g m–2 d–1 during the first 14 d was significantly greater than rates for both Day 15 to 49 (0.042 g m–2 d–1) and Day 50 to 99 (0.111 g m–2 d–1, Table 4). The lower nonsignificant correlation coefficients between the rate of O2 consumption and N2O emission (r = 0.18 for SBM and r = 0.26 for WBM, n = 16) indicate the relationship between the two parameters was nonlinear and probably complex. While production of N2O occurs via denitrification under anaerobic conditions, under extreme anaerobic conditions N2O becomes unstable and is further reduced to N2, perhaps explaining the poor correlation between N2O emission and O2 consumption.
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The loss of N2O was low, similar to findings reported from other composting studies (Martins and Dewes, 1992; Kuroda et al., 1996; Eklind and Kirchmann, 2000b; Hao et al., 2001; Sommer, 2001; Amon et al., 2001). The measured N2O loss was 0.0771 kg Mg–1 and represented only 0.9% of total N loss (8.26 kg Mg–1) and 0.39% of initial TN (19.92 kg mg–1) for SBM. For WBM, the N2O loss of 0.0842 kg Mg–1 represented about 5.9% of total N loss (1.42 kg N Mg–1), and 0.68% of initial TN (12.40 kg Mg–1). Since there was a roof and concrete floor, runoff and leaching losses of N did not occur. Most N was lost either as NH3 or possibly N2.
In addition to the direct GHG emissions, 5.1 kg C Mg–1 (SBM) and 4.2 kg C Mg–1 (WBM) CO2 was also released from the diesel fuel used to turn and maintain the compost windrows. Using global warming potential factors of 1 for CO2, 21 for CH4, and 310 for N2O, total emissions during composting expressed as CO2–C equivalents were 368.4 kg C Mg–1 manure for SBM, which was not significantly different from the total emission of 349.2 kg C Mg–1 manure for the WBM treatment. Although most C is emitted as CO2 (94% total C loss), the impact of CH4 was greater since its global warming potential is 21 times more than CO2 (Table 5). Although N2O contributed <5% of total GHG emission (CO2–C equivalent), its emission is a concern because N2O is important to the troposphere radiation balance and in stratospheric ozone chemistry.
Application of compost to agricultural land has been reported to increase soil C content, and is regarded as a means of C sequestration (Beauchamp and Voroney, 1994; Smith et al., 1998). However, for straw-bedded manure, total GHG emissions during composting in terms of CO2–C equivalent (368.4 kg Mg–1) were higher than the initial total C content (330.5 kg Mg–1) in the manure, raising the question of the net benefit of composting in C sequestration. Crop production (photosynthesis) captures atmospheric CO2, but composting releases a portion of this captured CO2 as CH4, a more harmful GHG. If livestock manure were directly applied to agricultural land, there would be no CH4 emission loss from soil since soil (except for waterlogged soil, such as a rice paddy) acts as a sink for atmospheric CH4 (Hütsch, 2001).
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CONCLUSION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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Our study is one component addressing the much larger and more complex question of mitigating GHG emissions using manure management technologies. Further study is needed to fully evaluate the impact of composting on overall GHG emissions between feedlot pen and field as well as emissions after compost is applied to soil. For example, GHG emissions from soils receiving fresh manure and compost should be compared and these differences factored into the overall GHG mass balance for these manure management options.
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ACKNOWLEDGMENTS |
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NOTES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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REFERENCES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONREFERENCES |
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