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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}em.agr.ca)
Received for publication May 17, 2000.
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ABSTRACT |
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 TOPNOTES ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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Abbreviations: GHG, greenhouse gas
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INTRODUCTION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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Composting has been proposed as an alternative to direct land application. The composting of feedlot manure produces a stabilized product that can be stored or spread on land with little or no odor, pathogens, weed seeds, or fly breeding potential (Rynk, 1992). Compost may also be trucked further distances since volume and mass are significantly reduced during the composting process (Larney et al., 2000).
However, C and N losses during composting not only reduce the agronomic value of compost as a soil amendment, but also contribute to emissions of GHG, such as CO2, CH4, and N2O. Eghball et al. (1997) observed N loss via NH3 volatilization of 19 to 42% of total N, and C loss as CO2 between 46 and 62% of total C during cattle feedlot manure composting in Nebraska. Nitrous oxide and CH4 emission during the composting of municipal solid waste has also been reported (Hellmann et al., 1997). Forced aeration and turning reduce CH4 concentration in the windrow when compared with stockpiling of solid cattle manure (Lopez-Real and Baptista, 1996). Nitrous oxide emissions varied between 0 and 0.33 g N m-2 d-1 in a laboratory flow-through chamber study simulating solid dairy manure storage (Brown et al., 2000), and varied between 0 and 0.025 g N m-2 d-1 during storage of solid pig manure (Petersen et al., 1998). But apart from these results, there is an apparent scarcity of quantitative information on GHG emissions during composting. This information is crucial for developing and assessing the potential for mitigating GHG emissions from cattle feedlot operations. The aim of this study was to quantify GHG emissions during composting of cattle feedlot manure in southern Alberta.
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MATERIALS AND METHODS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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For the active aeration treatment, the manure in the windrow was turned six times (Day 14, 21, 29, 50, 70, and 84) during the 99-d thermophilic composting period. On Day 99, the passive windrow was opened and material from each treatment was moved into curing windrows. Details on physical changes that occurred during composting were published by Larney et al. (2000).
Temperature and Gas Composition in Compost Windrows
Temperature and the concentration profiles of oxygen (O2), CO2, CH4, and N2O were measured during the 99-d thermophilic composting phase. Temperature readings and gas samples were taken at weekly intervals (more frequently around turning times for the active aeration treatment) starting 2 d after the construction of windrows. The temperature and concentration of gases at depth 0 (at the compost surface for active windrows and below the cured compost cover for passive windrows), 7.6, 23, 46, 76, and 107 cm below the windrow surfaces were determined using a modified multilevel sampler method (Burton and Beauchamp, 1994). Briefly, the multilevel sampling probe was constructed to collect temperature readings and gas samples from different depths. The shaft of the sampler was constructed with PVC tubing (Sch. 80, 24 mm i.d., 4.5 mm wall thickness). Thermocouples (Type T thermocouple Cu–Ni Constantan) were installed through holes drilled in the side of the PVC tube at corresponding depths for temperature measurement in the windrow. The thermocouple leads were pulled through the interior of the PVC tubing and protruded from the top. A digital thermometer (Omega HH-25TC, Digital thermometer) was connected to the thermocouple leads to obtain temperature readings at each gas sampling date.
Gas sampling tubing was installed in a similar fashion and at the same depths as the thermocouples. One end of spaghetti tubing (polypropylene, 2 mm i.d.) was inserted in a hole drilled in the side of the PVC tube as the gas sample inlet, the other end protruding out the top of the PVC tube. An air-tight fit at the sampling end was made by stretching the needle end of a 1-mL polypropylene syringe over the spaghetti tubing. The syringe barrel was then cut at the 0.3-mL graduation and a septum was glued to the syringe barrel with polyacrylate cement. The whole system was sealed with a polyolefin heat shrink tubing.
Gas samples were collected by first purging the volume of the spaghetti tubing and then taking a 10-mL sample using a 10-mL disposable syringe. The syringe needles were placed in a rubber stopper immediately to prevent air exchange. All windrow profile gas samples were taken between 0800 and 0900 h and analyzed for CO2, CH4, N2O, and O2 the same day they were collected 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).
Gas Emission Measurements
Greenhouse gas (CO2, CH4, and N2O) emissions and O2 consumption during composting were measured using a modified vented chamber technique (Hutchinson and Mosier, 1981). This method also has been used in previous studies of GHG emissions during composting of organic wastes (Czepiel et al., 1996; Hellmann et al., 1997; Sommer and Møller, 2000). At each sampling time, the chamber (15.5 cm in diam. and 15 cm in height) was placed on the peak of each windrow and 10 mL of air was drawn with 10 mL disposable plastic syringes from the chamber's headspace at 0, 5 10, 20, and 30 min. Immediately after gas sampling, the syringe needle was placed in a rubber stopper to prevent gas exchange. All surface flux gas samples were taken between 0800 and 0900 h and analyzed for CO2, CH4, N2O, and O2 the same day they were collected as described previously.
Gas fluxes were calculated from concentrations by assuming a steady state gradient in the underlying windrows (Anthony et al., 1995). 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 minutes) for each sampling time (SAS Inst., 1996). The flux at time 0 was calculated by taking derivatives of the second-order polynomials 
. The flux obtained in each half-hour period was assumed to be representative for that day and daily flux was obtained by integrating the half-hour flux over a 24-h period. There is always uncertainty when extrapolating results over longer periods of time, which may result in under- or overestimates of GHG fluxes. However, this method provides a valid comparison between the passive and active aeration treatments. Cumulative emissions were approximated by assuming that daily fluxes (measured and calculated weekly) represented the average for the whole week and were expressed per initial unit surface area or per initial unit dry weight of cattle manure.
Windrow Porosity
Bulk densities of the materials were estimated by weighing an aluminum pail of a known volume (0.064 m3) filled with manure–compost at each turning date for the active aeration treatment (Larney at al., 2000). The particle density was measured with modified pycnometer (Blake and Hartge, 1986). Five grams of manure or compost material were placed into a 100-mL volumetric flask and 90 mL of water was added. The flask was then covered with a rubber stopper and shaken vigorously to remove trapped air in the manure or compost material. The air bubbles were then removed by inserting two syringes into the rubber stopper, one for injecting water into the flask and the other for removing air through the rubber stopper. The total porosity in the windrow was calculated using bulk density and particle density measured each time the active windrow was turned. The air-filled pore space was calculated as the difference between total porosity and moisture content. Emissions during turning for the active aeration treatment and at the end of the thermophilic phase (Day 99), when both active and passive windrows were opened, were calculated using the average gas concentration in the profile and the air-filled pore space before the turning event. These amounts were added to the total cumulative emissions.
Statistical Analysis
The general linear model (GLM) procedure in SAS (SAS Inst., 1996) was used to analyze cumulative GHG emissions. When treatment effects were significant at 0.05 probability level, means were tested with the Tukey multiple range test.
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RESULTS |
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Methane concentration was very low near the surface and increased with depth. The highest CH4 concentrations were always found near the bottom of the windrow throughout the entire composting period, and they increased dramatically toward the end of composting (Fig. 3c). On the last sampling date (Day 97), CH4 concentration reached 53% at a 107-cm depth. The highest concentrations of CH4 were found where O2 concentrations were low.
The peak N2O concentration generally occurred at the 23-cm depth during early composting and stayed high at this location for the first 25 d (Fig. 3d). Nitrous oxide concentration decreased to near zero on Day 29, but increased again from Day 43 to 64. On Day 77, the N2O concentration was still high, but the location of the peak had shifted to a lower depth (46 cm).
The N2O concentration patterns were quite different from those of O2, CO2, and CH4. During the early composting period, when low O2 and high CO2 concentrations occurred at a 46-cm depth, the bottom of the windrow had the highest concentrations of CH4, whereas the highest N2O concentration occurred at a much shallower position (23-cm depth) where there was at least 6 to 9% O2. As passive composting progressed, the minimum O2 and maximum CO2 and CH4 concentrations moved to the bottom of the windrow while the maximum N2O concentration moved to only a 46-cm depth, where O2 concentration remained at least 6 to 9%.
Surface Fluxes of O2, CO2, CH4, and N2O
The rates of CO2, CH4, and N2O daily emissions and O2 daily consumption were high during the first 30 d of passive composting, but declined to almost zero for the remainder of the study (Fig. 4a) . The daily emission rates varied from 0 to 0.22 kg C m-2 d-1 for CO2, 0 to 0.04 kg C m-2 d-1 for CH4, and 0 to 0.60 g N m-2 d-1 for N2O (Fig. 4a). As expected, emission rates for CO2, CH4, and N2O were high whenever O2 consumption rates were high. A simultaneous emission of considerable amounts of CH4 and CO2, and consumption of O2, occurred in early composting.
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The O2 concentration in the active aeration treatment decreased during the first 14 d and the lowest concentration was found at 46- to 76-cm depths (Fig. 5a) . The concentration increased to background levels (equal to 21% in the air) each time the windrow was turned, but decreased rapidly thereafter. The location of lowest O2 concentration in the windrow moved downward from a 46-cm depth at the beginning of the study to the bottom of the windrow toward the end of the study. Similarly, highest CO2 concentrations were located from 46- to 76-cm depths at the outset and shifted to the bottom of the windrow toward the end of study (Fig. 5b). The highest concentration of CO2 in the profile occurred where O2 concentration was lowest, similar to the passive aeration treatment, but the peak concentration of CO2 was never as high as in the passive aeration treatment.
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Two days after the compost windrows were constructed, N2O concentration increased to its first maximum (12 µL L-1 at depth 23–46 cm) and then decreased (Fig. 5d). The N2O concentration stayed very low (<10 µL L-1) until Day 64 when its concentration increased again, peaking at 537 µL L-1 on Day 85. Similar to the passive treatment during the second half of the composting process, the peak concentration of N2O was located at a 46-cm depth where there was at least 9% O2 (Fig. 5a) since complete anaerobicity would favor N2 instead of N2O as the end product of denitrification.
Turning affected the concentration profiles of various gases during composting for the active treatment (Fig. 5). The concentrations of O2, CO2, N2O, and CH4 returned to preturning values within 1 d during early composting. Very similar patterns were observed during late composting, except that N2O increased. The N2O concentration increased from 3.5 µL L-1 on Day 69 to 103 µL L-1 on Day 71 at a 46-cm depth. Similarly, N2O concentration increased from 6.3 µL L-1 on Day 83 to 537 µL L-1 on Day 85 at a 46-cm depth. The precipitation before turning (on Day 70, 79, 80, and 83) combined with high nitrate concentration (Larney et al., 1999), may have increased N2O concentration through denitrification under anaerobic conditions. The peak concentration of CO2 and the temperature shifted upward after turning during late composting, probably due to the increased supply of substrate for microbial activity at the upper locations. The lack of such patterns in early composting suggests substrate supply was not the limiting factor in controlling biological activity at this time. The mixing action brought partially decomposed materials from the bottom to the top of the windrow each time the windrow was turned, thus increasing the substrate concentration at the top of the windrow.
Surface Fluxes of O2, CO2, CH4, and N2O
High rates of O2 consumption were always associated with high rates of CO2 and N2O emission during the entire active composting period (Fig. 6)
. Daily fluxes varied from 0 to 0.35 kg C m-2 d-1 for CO2, 0 to 0.02 kg C m-2 d-1 for CH4, and 0 to 0.92 g N m-2 d-1 for N2O. Compared with the passive treatment, CO2 and N2O daily emission rates stayed high throughout the entire composting period, except during two dry periods (Days 25–30 and Days 46–70). While CH4 emission was high during early composting, the daily emission rate stayed low throughout the middle and late composting periods. The patterns of CH4 emissions during active composting of feedlot manure were similar to those reported by Hellmann et al. (1997) for municipal solid waste composting. Emission rates of CO2 and N2O and the consumption rate of O2 closely reflected their concentration profiles in the windrow.
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In addition to the direct GHG emissions from the active aeration windrows, CO2 released from diesel fuel burned to turn and maintain the active windrows (4.4 kg C Mg-1 manure) was also included in the total GHG emissions. This led to the emission value (CO2–C equivalent) of 401.4 kg C Mg-1 manure, which was significantly higher than the passive aeration treatment (240.2 kg C Mg-1 manure).
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DISCUSSION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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The high concentration of CO2 and CH4 at the bottom of the passive aeration windrow did not generate any significant surface fluxes toward the end of the composting period. This suggests that these high concentrations can be attributed to local accumulation and slow transport of these gases to the windrow surface, rather than a high rate of production. Another reason for the low surface gas fluxes may include loss of CO2 and CH4 through the aeration pipes laid at the bottom of the passive windrow, and/or the oxidation of CH4 as it moved toward the surface through the area with high O2 concentrations.
For the active aeration treatment, periodic turning not only reintroduced fresh air (O2) into the windrow, but also mixed the composting materials. This mixing moved well-composted material from the surface to the bottom of the windrow, as well as partially composted materials from the bottom to the surface. Thus, more material was subject to aerobic decomposition, increasing the rate of the composting process and CO2 production. This is well demonstrated by the occurrence of a new temperature peak (Fig. 2b) after each turning event. Unlike the passive aeration treatment, CO2 surface emissions increased steadily throughout the whole composting period, except during Days 30 to 50, as dry conditions limited microbial activity.
Turning also increased the gas diffusion rate by fluffing up the material and increasing total porosity and air-filled porosity in the active aeration treatment. Loss of water due to turning resulted in more pore space being available for gas diffusion. In addition, most water was initially located near the windrow surface after each rainfall with slow redistribution downward. This rainfall may act as a plug near the windrow surface, thus reducing gas diffusion as observed in the passive treatment in the last 40 d of composting. Turning redistributed the rainwater throughout the profile, effectively increasing the gas diffusion and biological decomposition rate.
The increased biological activity and gas diffusion rates due to turning led to a significantly higher CO2 emission (168.0 kg CO2–C Mg-1 manure) for the active treatment compared with the passive aeration treatment (73.8 kg CO2–C Mg-1 manure). However, the slightly higher CH4 emission (8.1 kg C Mg -1 manure) was not significantly different from values obtained for the passive aeration treatment (6.3 kg C Mg-1 manure). Although CH4 constituted a much smaller amount of the total C emitted into the atmosphere for both aeration treatments, we cannot neglect its impact since its global warming potential is 21 times that of CO2 (IPCC, 1997). In addition, there is virtually no CH4 emission when raw manure is applied directly to agricultural land as agricultural soil is regarded as a sink for CH4.
Besides effects on the gas diffusion and material mixing, turning also impacted N transport and transformation in the active aeration treatment. This ultimately affected N2O production and surface emissions. Several mechanisms, including nitrification, denitrification, and chemo-denitrification, could all potentially contribute to the observed N2O profiles and surface emission patterns. Since the heterogeneous properties of manure–compost materials lead to a mosaic of aerobic and anaerobic sites, it is likely that multiple processes are contributing simultaneously to N2O production. At the beginning, with high NH+4 and low NO-3 concentration in the feedlot manure (Larney et al., 1999), nitrification, rather than denitrification, is the major contributor to the increased N2O concentration at 23 cm and the surface fluxes of N2O.
Chemo-denitrification may also play an important role in the high N2O production (high concentration at a 23-cm depth) and emissions during early composting, especially for the passive aeration treatment. Conversion of NO-2 to NO-3 is inhibited by high temperatures (Tate, 1995), high concentrations of volatile fatty acids (Takai et al., 1997), high NH+4 content (Smith et al., 1997), and high pH (Stevens et al., 1998). The higher concentration of NO-2 (data not shown) combined with the abundance of amine and phenol compounds, produces N2O through chemo-denitrification (Matson and Harriss, 1995). As composting progressed, NH+4 concentration decreased and NO-3 increased due to volatilization losses and nitrification near the surface. This leads to concentration profiles of maximum NO3–N and minimum NH4–N near the surface since NH4–N produced during decomposition of organic matter was stable under anaerobic conditions (the lower part of compost windrow) and NO-3–N produced during nitrification was stable under aerobic conditions (near the windrow surface).
For the passive aeration treatment, NO-3 produced near the surface has to be leached (precipitation) to lower depths where O2 supply is limited for denitrification to occur. For the active aeration treatment, turning would result in the NH4–N being transported to the top and NO3–N being transported to the bottom of the windrow. As a result of the redistribution of 
in the windrow profile, more N2O would be produced through nitrification near the surface and denitrification where the O2 supply is limited. Coupled with an improved gas diffusion rate due to turning, this led to higher N2O emissions late in the composting period for the active aeration treatment.
The N2O emission patterns also reflected the ratio of N2O to N2 produced during denitrification. The N2O to N2 production ratio during denitrification depends on several factors, such as O2 status, temperature, pH, decomposable organic matter, and NO-3 availability (Khdyer and Cho, 1983; Scholefield et al., 1997). It is well established that denitrification occurs under anaerobic conditions. However, under extremely low O2 conditions, N2O produced during denitrification may be unstable and is further reduced to N2. The higher moisture and readily decomposable organic matter content during early composting would promote anaerobicity and production of N2 instead of N2O. In addition, low NO-3 concentration and the increase in pH from 7.1 to 8.2 for the active aeration treatment during early composting (Larney et al., 1999) also favored N2 as a denitrification end product (Khdyer and Cho, 1983; Arah et al., 1991). On the other hand, increases in NO-3 concentration and lower pH, water content and temperature during the second half of composting (Larney et al., 1999), especially for the active aeration treatment, favors the production of N2O over N2 (Cho and Sakdinan, 1978; Khdyer and Cho, 1983; Scholefield et al., 1997; Stevens et al., 1998), leading to higher N2O concentrations in the windrow profile and higher surface fluxes.
Compared with CO2, N2O emissions during composting were relatively low. The N2O emission values of 0.11 for the passive and 0.19 kg N Mg-1 manure for the active treatments were equivalent to 0.62 and 1.07% of the total initial N in the manure. Nitrous oxide emission values were similar to those after fertilizer or manure application to soil (Eichner, 1990; Chang et al., 1998). The lower GHG emissions with passive aeration suggest that passive aeration (which is similar to stockpiling manure) might be an option if complete decomposition is not required.
In addition to GHG emissions during composting, we should also consider emissions after each compost is applied to agricultural land. A higher GHG emission and NO-3 leaching would be expected for passively aeration treated compost since its total and mineral N content were higher than the active aeration compost (Larney et al., 1999). In addition, mineral N release from passive aeration compost would also be higher than active aeration compost, but both would be lower than that from fresh cattle manure since N mineralization rate decreases as the maturity of compost increases (Castellanos and Pratt, 1981; Tyson and Cabrera, 1993; Robertson and Morgan, 1995). Currently most cattle manure is applied in its raw state to nearby fields, which may lead to environmental problems such as nitrate leaching to ground water and P runoff to surface water. Composting allows manure nutrients to be transported further and applied to a greater area. However, composting may solve one environmental issue (protection of water quality) while creating another one (increased GHG emissions). Further study is needed to investigate GHG emissions from the whole system of manure management (starting when cattle manure is in the feedlot pens, through direct land application vs. composting and after land application of raw manure vs. compost) so that we have a better understanding of its impact on our environment.
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CONCLUSIONS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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ACKNOWLEDGMENTS |
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NOTES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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REFERENCES |
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