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Evaluation of full-scale biofilter with rockwool mixture treating ammonia gas from livestock manure composting
Bioresource Technology 100 (2009) 1568–1572
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Evaluation of full-scale biofilter with rockwool mixture treating ammonia gas from livestock manure composting Tomoko Yasuda *, Kazutaka Kuroda, Yasuyuki Fukumoto, Dai Hanajima, Kazuyoshi Suzuki Pollution Control Research Team, National Institute of Livestock and Grassland Science, 2, Ikenodai, Tsukuba, Ibaraki 305-0901, Japan
a r t i c l e
i n f o
Article history: Received 20 May 2008 Received in revised form 16 September 2008 Accepted 18 September 2008 Available online 31 October 2008 Keywords: Full-scale biofilter Nitrification Denitrification Rockwool Greenhouse gas
a b s t r a c t NH3 removal by a full-scale biofilter with rockwool packing materials was studied by measuring the gases and potential nitrification and denitrification activities of those materials in order to improve the biofiltration technology used in livestock farms. The rockwool biofilter was a durable and effective system for removing NH3, which was varied with the turning of manure composts. Furthermore, NH3 could be treated in the absence of an extra increase in two greenhouse gases, N2O and CH4. Potential nitrification and denitrification activities of the packing materials were estimated to be 8.2–12.2 mg N, and 1.42–4.69 mg N/100 g dry samples per day, respectively. The results suggested that potential nitrification and denitrification activities would increase within the biofilter where substrates, NH3 or NO 3 , have accumulated as a result of its operation. However, since percolate water contained high concentrations of NHþ 4 and NO3 , further improvement is required by reducing nitrogenous compounds within both the biofilter and percolate water. Ó 2008 Elsevier Ltd. All rights reserved.
1. Introduction Ammonia (NH3) is emitted from the composting process as a byproduct of the aerobic decomposition of organic materials. Such emissions amounted to 10–25% of the total nitrogen in swine manure (Kuroda et al., 1996; Osada et al., 2000; Fukumoto et al., 2003), and 9–76% in poultry droppings (Kirchmann and Witter, 1989; Kirchmann and Lundvall, 1998). The emitted NH3 causes deterioration of the environment in the form of offensive odors (Kuroda et al., 1996) as well as the acidification and eutrophication of soils and surface waters. Therefore, treatment of NH3 is crucial for the better management of livestock farming as well as for improving environmental quality. Biofiltration is a suitable option for treating NH3 emitted from the livestock composting process because of its high elimination efficiency, low operational cost, and modest environmental impact (Williams and Miller, 1993; Wani et al., 1997). This technology has been applied to composting facilities for well over 20 years, and the basic design and operating criteria have been well established. Additional studies are required to determine information on the microbiology and chemical reactions involved so that improvements in engineering and biofilter modeling can be achieved (Williams and Miller, 1993; Devinny and Ramesh, 2005). Many studies have used a laboratory-scale biofilter to determine the mecha* Corresponding author. Tel.: +81 29 838 8677; fax: +81 29 838 8606. E-mail address: [email protected] (T. Yasuda). 0960-8524/$ - see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.biortech.2008.09.033
nisms by which NH3 removal can be achieved (Hartikainen et al., 1996; Smet et al., 2000; Nicolai et al., 2006). Nitrification plays an important role in long-term effective NH3 removal. It has been estimated that nitrification accounts for 50% of NH3 removal (Smet et al., 2000). On the other hand, studies about denitrification in gas-phase biofilters have rarely been done (Zhu et al., 2004). Data on such factors as the distribution of gases and microbial activity in an actually operating biofilter are also limited. On-site NH3 loading, temperature, and other environmental conditions, all of which have an impact on microbial activity, are not constant. In addition to NH3, greenhouse gases, nitrous oxide (N2O) and methane (CH4) are emitted from the composting process (Kuroda et al., 1996; Osada et al., 2000; Veeken et al., 2002; Fukumoto et al., 2003). The emission patterns of these gases before and after biofiltration have not been investigated. Clemens and Cuhls (2003) reported that biofilters had no net effect on CH4, but that the concentration of N2O increased after filtration. If greenhouse gas emissions are high, the treatment of these gases should be taken into consideration when biofiltration is used in a composting facility. In this study, we evaluated NH3 removal by a full-scale biofilter with rockwool packing materials operating for a long time by measuring NH3 gas and potential nitrification and denitrification activities. The packing materials contain rockwool (fibrous material produced from igneous rocks), urethane, zeolite, and dried chicken feces (Fukumori et al., 2000). Emission patterns of two greenhouse gases, CH4 and N2O, before and after biofiltration were also investigated to determine how the biofilter affected them.
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filtering materials. Gas sampling points are shown in Fig. 1. Gas was collected in a 10-ml evacuated headspace vial and subsequently analysed for N2O and CH4. Biofilter packing materials were sampled using a liner soil sampler (5 cm ID, Daiki Rika Kogyo Co. Ltd., Japan) on June 22 of 2006, a few days before the next turning. Sampling points are shown in Fig. 1. Samples were taken from two different depths of 50 cm and 140 cm at each point. After sampling, aggregates were broken down by sieving (4.0 mm mesh), and urethane was cut into pieces about 1 cm in size. The separated particles were then mixed and stored at 4 °C until analysis. The biofilter temperature was monitored 50 cm below the surface by thermocouples equipped with a data logger (NR 1000, Keyence, Japan). The percolate water was sampled during the gas monitoring period in 2006.
2. Methods 2.1. Description of experimental site Our analysis required a biofilter and a composting facility, which have been operating at the National Institute of Livestock and Grassland Science (NILGS: Tsukuba City, Japan) since 1998 (Fig. 1). The biofilter is 5.8 m 8.0 m in cross-section and 3 m high (2.5 m for packing materials). Cow and swine manure is placed in the first reactor of the composting facility, then transferred to the next reactor and also turned at 3-week intervals. Gas emitted from the first three reactors (1st–3rd) is forced into the biofilter by three turbine blowers (No. 3BLF, Tecno Wasino Co. Ltd., Japan), each with an approximately 20 m3 min1 output flow, from underneath ditches providing uniform gas distribution. Gas and packing material contact time ranges from 100 to 200 s. Water (0.4 m3) is sprayed over the upper surface of the biofilter daily to maintain the moisture of packing materials, which are a mixture of mainly rockwool, urethane (5–10% v/v, only at 0.5–2 m below the surface), zeolite (5–10% w/w), and dried chicken feces (10–20% w/w) (Fukumori et al., 2000). Percolate water is currently treated along with the other livestock wastewater.
2.3. Analytical methods The NH3 concentration was measured with a detection tube and a gas sampler (Gastec Co. Ltd.). The N2O and CH4 concentrations were determined using GC with a 63Ni-electron capture detector (GC-14A, Shimadzu, Kyoto, Japan), and a flame ionization detector (6890 Series GC system, Agilent Technologies, Palo Alto, CA, USA), respectively. The moisture contents of the biofilter packing materials were determined as weight loss after drying overnight at 105 °C. About 5 g of packing materials was suspended in 45 ml of 2 M KCl and shaken for 30 min. After centrifugation at 2500 rpm for 10 min, the pH of the supernatant was measured by a glass electrode (B-211, Horiba, Japan), and defined as the pH of the sam ples (Kuroda et al., 1996). The amounts of NHþ 4 , NO2 and NO3 were measured in the same KCl extracts using a flow injection analyzer (Aquatec 5400, Tecator, Sweden). The total carbon (TC) and total
2.2. Sampling of gases, biofilter packing materials, and percolate water Concentrations of NH3 and greenhouse gases, N2O and CH4, prior to and after biofiltration were measured twice in November 2004 and May to June 2006. One or two turnings of the compost were included during the sampling period. Inlet gas to the biofilter was monitored from a sampling hole of the gas conduit. Outlet gas was collected using an inverted plastic funnel on the surface of the
Rockwool biofilter Roof
5.8 m
b
a
c
8.0 m
2.5 m
Tap water
ditch
Rockwool packing materials 3m 1
1st
2
3
2nd
3rd
Percolate water
4
4th
5th
6th
Compost flow
Composting barn Fig. 1. Diagram of biofilter and composting facility used for analysis. (black circles, sampling points of inlet air (1–4); letters, sampling points of outlet air and packing materials (a–c); circles with a cross, water nozzles).
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nitrogen (TN) contents of the biofilter packing materials were determined using an NC analyzer (NC-220F, SCAS, Japan). For the percolate water, NO 2 and NO3 were measured by ion chromatography (IC 7000 analyzer, Yokogawa, Japan, or HIC-VP super system, Shimadzu, Japan); NHþ 4 was measured by ion chromatography (DX120 analyzer, Dionex, Japan); and total carbon (TC) and total organic carbon (TOC) were measured by a TC/TOC analyzer (TOCVCSN, Shimadzu, Japan). Statistical analyses of differences of inorganic N, TC, and TN contents, and potential nitrification activities among samples were conducted using the GLM procedure of SAS (2002). 2.4. Determination of potential nitrification activity of biofilter packing materials Ammonium oxidation potential was determined based on some modifications of methods used in previous studies (Schmidt and Belser, 1994; Kimura, 1997). Triplicate portions (20–30 g) of samples were incubated at 30 °C in the dark in a 200-ml conical flask covered with aluminum foil. Water content was adjusted with (NH4)2SO4 solution at the start of incubation to 60% of water-holding capacity as determined by the Hilgard cup technique, and 100 g dry samples1. resulting in approximately 150 mg NHþ 4 NO2 and NO3 were measured after extraction with 2 M KCl, as described in Section 2.3. Potential nitrification activity was calculated from the amounts of NO 2þ3 produced during the 8-day incubation. 2.5. Determination of potential denitrification activity of biofilter packing materials Potential denitrification activities were determined based on the acetylene block methods of a previous study (Tiedje, 1994) with some modifications. Duplicate 2 g of subsamples were suspended in 5 ml of 5 mM phosphate buffer supplemented with 1 mM KNO3 and 1 mM glucose in a 30-ml test tube with butyl rubber cap and screw cap. Headspace gas was replaced with 10% C2H2 (Ar balance) by purging with the gas for 5 min with shaking. The final pressure of the headspace gas was 0.4 atm. Tubes were incubated at 25 °C in the dark. N2O in the headspace gas of each tube was measured after a 24-h incubation, as in Section 2.3. 3. Results and discussion As shown in Fig. 2, the biofilter with rockwool packing materials effectively removed NH3 gas. Complete NH3 elimination was also achieved in 2000 (data not shown), supporting durability of this biofilter. Inlet NH3 concentrations peaked after compost turning, and then slowly declined. Differences in the inlet concentrations among sampling points were ascribed to the composting process, although there was no partition in the gas conduit. The highest NH3 concentration was expected in inlet 1 because immediate NH3 emissions after turning were observed in laboratory-scale experiments, and it persisted while organic materials were degrading (Osada et al., 2000). Contrary to expectations, the peak height was double that of inlet 1 or higher at inlets 2, 3 and even at inlet 4 in November 2004 (Fig. 2). This was probably because the composting proceeded slowly due to the large amounts of manure (60–80 tons per day), as well as the lower temperature in winter (November 2004). On the whole, the concentration of NH3 in the mid-lower position of the biofilter was assumed to be high. The effects of biofiltration on the emission of two greenhouse gases, N2O and CH4, were examined (Fig. 2). A large portion of the inlet gases seemed to pass straight through the biofilter, although several microbial N2O and CH4 production and consumption reactions may have occurred to some degree. Inlet gas approx-
imately contained an average of 4 ppm of N2O, and an average of 430 ppm of CH4 in the range of 0–1800 ppm, respectively, throughout the sampling periods. According to previous studies, N2O emission from the composting process continued until the later phase of composting, whereas CH4 was mainly emitted at an earlier stage (Osada et al., 2000; Veeken et al., 2002; Fukumoto et al., 2003). The patterns shown in Fig. 2 are relevant to those prior studies. As for N2O, in 2004 and at the first turning in 2006, the concentrations in inlets 4 and 3 were relatively high, suggesting that N2O was continuously emitted during the composting process. As for CH4, the inlets 1 and 2 concentrations were relatively high in 2004 and at the second turning in 2006, suggesting that CH4 was emitted much more than at the initial composting. The reason for differences in emission patterns of both N2O and CH4 between the first and second tuning in 2006 was not clear. However, more CH4 was emitted from the undisturbed compost pile where less amounts of NH3 were emitted (Veeken et al., 2002). The emission of CH4 could be correlated with that of NH3 during the sampling period in 2006. It has been reported that a biofilter treating NH3 gas could be a major source of N2O, and in fact approximately 26% of the NH3–N that was removed by the biofilter was transformed into N2O (Clemens and Cuhls, 2003). Though the mechanism underlying N2O emissions in a biofilter has not yet been clearly understood, the results of our study confirmed that NH3 could be treated with rockwool mixture in the absence of an extra N2O increase. The moisture content, pH, inorganic N content, TC and TN of sampled packing materials are shown in Table 1. Moisture content averaged 47.2%. The pH showed a similar near-neutral value among the samples in the range of 6.8–7.4. The pH of the biofilter packing materials solution declined because of nitrification, with the result that more NH3 was absorbed into the solution. As a result, pH in the packing materials was maintained at a neutral level in this biofiltration system. The pH of percolate water was similar to that of the packing materials. That water never contained less than 1000 mg/l of NHþ 4 and NO3 during the sampling periods (Table 2). Within the biofilter, the NHþ 4 was low at site b, but there was no difference among the depths at every site. A significantly high level of NO 3 , which was probably leached into the water, was accumulated in the lower depths (Table 1). NO 2 contents were slightly less at the lower position than at the upper one (P < 0.05) (Table 1). Both TC and TN contents were higher at the lower position (P < 0.05), which was partly due to the urethanes used for increasing porosity. The mean temperature during the sampling periods was 19.2 °C (10.3–27.8 °C). Meanwhile, the temperature of packing materials was 25.5 °C (20.2–31.6 °C), which was higher and more stable than that of the atmosphere. Potential nitrification activity of the packing materials sampled in 2006 was determined by static incubation with an amendment of (NH4)2SO4. Their values were similar among samples except b140 cm, (8.96 ± 0.27 mg N per 100 g dry samples per day (SE), 7.96 ± 0.36, 8.20 ± 0.23, 9.48 ± 0.81, and 8.67 ± 0.45, for a-50 cm, a-140 cm, b-50 cm, c-50 cm, and c-140 cm, respectively). On the other hand, a significantly higher value was obtained in sample b-140 cm (12.2 ± 0.40, P < 0.05). From the NH3 emission patterns in 2006, it was assumed that more NH3 was loaded in the mid-lower part of the biofilter (relative to sampling point b). That result suggested that nitrification activity was affected by the NH3 concentration. Although high ammonia removal efficiency could be achieved for a high NH3 concentration by adsorption and absorption capacity of media, the performance declined when NH3 load exceeded those capacities and the biological activity (la Pagans et al., 2005). The maximum safe NH3 loading would depend on the nitrification activity. In this study, the potential nitrification activity obtained was estimated to be sufficient to oxidize the average amount of NH3–N that entered the biofilter per day during the gas sampling period (ca. 1 kg N load as opposed to ca. 2 kg N
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NH 3 (ppm)
(I) 160
140
140
120
120
100
100
80
80
60
60
40
40
20
20
N2 O (ppm)
0 11/8
11/13
11/18
0 5/13
12
12
10
10
8
8
6
6
4
4
2
2
0 11/8
CH 4 (ppm)
(II)
160
11/13
11/18
0 5/13
2000
2000
1600
1600
1200
1200
800
800
400
400
0 11/8
5/18
5/23
5/28
6/2
6/7
6/12
6/17
5/18
5/23
5/28
6/2
6/7
6/12
6/17
5/18
5/23
5/28
6/2
6/7
6/12
6/17
0 11/13
11/18
5/13
2004
2006 Date
Fig. 2. Changes in concentrations of NH3, N2O, and CH4, before and after biofiltration in 2004 (I) and 2006 (II) (inlet 1 (d), inlet 2 (j), inlet 3 (N), inlet 4 (), outlet a (s), outlet b (h), outlet c (4)). Arrows indicate turning time of manure composts.
Table 1 Characteristics of biofilter packing materials sampled in 2006 Sample origin (site-depth)
a-50 cm a-140 cm b-50 cm b-140 cm c-50 cm c-140 cm
Moisture (%, wt/wt)a
45.2 46.0 47.8 46.2 50.4 47.7
pHa
7.0 7.0 6.9 7.2 7.2 6.9
Inorganic N content (mg 100 g dry samples1)a NHþ 4 –N
NO 2 –N
14.1 (0.79)A 15.6 (2.84)A 3.54 (0.90)B 2.74 (0.79)B 18.0 (5.57)A 15.1 (2.43)A
0.14 0.12 0.14 0.12 0.15 0.10
(0.01)A (0.01)B (0.01)A (0.005)B (0.01)A (0.004)B
TC (%, wt/wt)a
TN (%, wt/wt)a
2.70 (0.09)A 12.2 (1.18)B 2.65 (0.16)A 8.73 (1.23)B 2.73 (0.38)A 8.89 (2.14)B
0.55 1.38 0.45 0.97 0.61 1.09
NO 3 –N 1.07 (0.01)A 18.4 (0.48)B 7.22 (0.26)C 26.0 (0.01)D 8.43 (0.19)C 39.0 (0.38)E
(0.03)A (0.10)B (0.05)A (0.13)B (0.10)A (0.25)B
Values in parentheses represent standard errors. Common superscripts within a column indicate no significant difference between values for those sites (P < 0.05). Values represent means of three replicates.
a
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Table 2 Characteristics of percolated water from the biofilter Sampling date
2006
6/6 6/9 6/13
pH
6.9 7.4 6.6
Inorganic N (mg/l) NHþ 4 –N
NO 2 –N
NO 3 –N
1186 1036 1223
119 209 286
1692 1562 2031
TC (mg/l)
TOC (mg/l)
70.8 77.5 71.2
58.4 57.9 61.8
potentially nitrified per day). However, the NH3 load fluctuated corresponding to the turnings, and more NH3 was also removed by absorption into the water (Table 2). Potential denitrification activity, determined with an amendment of glucose, tended to be high in the lower sampling position, where NO 3 has accumulated, (3.65 ± 0.52 mg N per 100 g dry samples per day (SE) and 4.69 ± 0.12 for sample a-50 cm and a-140 cm, 1.42 ± 0.18 and 2.50 ± 0.17 for b-50 cm and b-140 cm, and 1.59 ± 0.26 and 2.46 ± 0.08 for c-50 cm and c-140 cm, respectively). Denitrification has rarely been determined in an aerated biofilter because O2 transfer in the biofilm systems may not be limited (Zhu et al., 2004). Although direct evidence for denitrification in the biofilter was scarce, Nicolai et al. (2006) pointed out denitrification as one possible explanation for nitrogen disposition in the biofilter and Ho et al. (2008) showed the presence of Paracoccus denitrificans, which is capable of heterotrophic nitrification and denitrification, by molecular analysis. The occurrence of denitrification in biological aerated filters for the treatment of wastewaters was more frequently reported (e.g., Westerman et al., 2000; Garzón-Zúñiga et al., 2007; Gilbert et al., 2007). In those biofilter, the filtrating media have higher moisture and treat wastewaters containing high BOD (e.g., 9000 ppm; Garzón-Zúñiga et al., 2007) and their physicochemical condition may be largely different from those of ammonia gas biofilter. However, the results obtained in this study suggest that NO 3 could be removed within the biofilter under the appropriate conditions by denitrification. The TOC in the wastewater was as low as 60 mg/l (Table 2), suggesting that compounds which could be used as electron donors were low in the packing materials. Addition of easily degradable organic matter might enhance denitrification activity. In fact, it has already been demonstrated that NO 3 could serve as an electron acceptor for the degradation of VOC in an aerobic biofilter (Zhu et al., 2004). The Rockwool mixture initially contained dried chicken feces (10–20% w/w), but their degradable organic matter could be consumed during long-term operations. The NH3 gas pollution problem was converted to a water pollution problem. Further improvements in biofiltration technology are required to reduce nitrogenous compounds within both the biofilter and percolate water, especially when biofiltration systems are introduced to livestock farms that lack wastewater treatment plants or farmland on which percolate water could be applied as a liquid fertilizer. 4. Conclusion Biofiltration with a rockwool mixture proved to be a durable and effective system for removing NH3 emitted by the composting process. Furthermore, NH3 could be treated with rockwool mixture without any extra increase in N2O and CH4. Potential nitrification activity of the packing materials was estimated to be 8.2– 12.2 mg N per 100 g dry samples per day, which was sufficient to oxidize the average amount of NH3–N which entered the biofilter during the prior gas sampling period. Potential denitrification activity was estimated to be as high as 1.42–4.69 mg N per 100 g dry samples per day. Such results suggested that both potential
nitrification and denitrification activities could increase within the biofilter where substrates, NH3 or NO 3 , had accumulated in the course of its operation. Further improvements in this technology are required to reduce nitrogenous compounds within both the biofilter and percolate water. Acknowledgements The authors wish to thank Dr. Miyoko Waki for valuable advice on the experiments and Mrs. N. Akasaka and Mrs. K. Sumiya for their valuable assistance. References Clemens, J., Cuhls, C., 2003. Greenhouse gas emissions from mechanical and biological waste treatment of municipal waste. Environ. Technol. 24, 745–754. Devinny, J.S., Ramesh, J., 2005. A phenomenological review of biofilter models. Chem. Eng. J. 113, 187–196. Fukumori, I., Doshu, T., Ueno, K., 2000. In: Zen-noh (Ed.), Instruction Manual of Rockwool Biofilter. National Federation of Agricultural Co-operative Association (in Japanese). Fukumoto, Y., Osada, T., Hanajima, D., Haga, K., 2003. Patterns and quantities of NH3, N2O and CH4 emissions during swine manure composting without forced aeration – effect of compost pile scale. Bioresour. Technol. 89, 109–114. Garzón-Zúñiga, M.A., Lessard, P., Aubry, G., Buelna, G., 2007. Aeration effect on the efficiency of swine manure treatment in a trickling filter packed with organic materials. Water Sci. Technol. 55, 135–143. Gilbert, Y., le Bihan, Y., Aubry, G., Veillette, M., Duchaine, C., Lessard, P., 2007. Microbiological and molecular characterization of denitrification in biofilters treating pig manure. Bioresour. Technol. 99, 4495–4502. Hartikainen, T., Ruuskanen, J., Vanhatalo, M., Martikainen, P.J., 1996. Removal of ammonia from air by a peat biofilter. Environ. Technol. 17, 45–53. Ho, K.L., Chung, Y.C., Tseng, C.P., 2008. Continuous deodorization and bacterial community analysis of a biofilter treating nitrogen-containing gases from swine waste storage pits. Bioresour. Technol. 99, 2757–2765. Kimura, R., 1997. Measurement of nitrification activity, and enumeration and isolation of nitrifying bacteria. In: Committee of Soil Microbiology (Ed.), Experimental Methods for Soil Microbiology. Yokendo, Japan, pp. 207–214 (in Japanese). Kirchmann, H., Lundvall, A., 1998. Treatment of solid animal manures: identification of low NH3 emission practices. Nutr. Cycl. Agroecosyst. 51, 65–71. Kirchmann, H., Witter, E., 1989. Ammonia volatilization during aerobic and anaerobic manure decomposition. Plant Soil 115, 35–41. Kuroda, K., Osada, T., Yonaga, M., Kanematsu, A., Nitta, T., Mouri, S., Kojima, T., 1996. Emissions of malodorous compounds and greenhouse gases from composting swine feces. Bioresour. Technol. 56, 265–271. la Pagans, E., Font, X., Sanchez, A., 2005. Biofiltration for ammonia removal from composting exhaust gases. Chem. Eng. J. 113, 105–110. Nicolai, R.E., Clanton, C.J., Janni, K.A., Malzer, G.L., 2006. Ammonia removal during biofiltration as affected by inlet air temperature and media moisture content. Trans. ASABE 49, 1125–1138. Osada, T., Kuroda, K., Yonaga, M., 2000. Determination of nitrous oxide, methane, and ammonia emissions from a swine waste composting process. J. Mater. Cycles Waste Manage. 2, 51–56. SAS, 2002. Version 9.1. SAS Institute Inc., Cary, NC. Schmidt, E.L., Belser, L.W., 1994. Autotrophic nitrifying bacteria. In: Weaver, R., Angle, J.S., Bottomley, P.J. (Eds.), Methods of Soil Analysis, Part 2, Microbiological and Biochemical Properties. Social Science Society of America, Madison, WI, pp. 159–177. Smet, E., van Langenhove, H., Maes, K., 2000. Abatement of high concentrated ammonia loaded waste gases in compost biofilters. Water Air Soil Pollut. 119, 177–190. Tiedje, J.M., 1994. Denitrifiers. In: Weaver, R., Angle, J.S., Bottomley, P.J. (Eds.), Methods of Soil Analysis, Part 2. Microbiological and Biochemical Properties. Soil Science Society of America, Madison, WI, pp. 245–267. Veeken, A., de Wilde, V., Szanto, G., Hamelers, B., 2002. Passively aerated composting of straw-rich organic pig manure. In: Insam, H., Riddech, N., Klammer, S. (Eds.), Microbiology of Composting. Springer-Verlag, Berlin, Heidelberg, pp. 607–621. Wani, A.H., Branion, R.M.R., Lau, A.K., 1997. Biofiltration: a promising and costeffective control technology for odors, VOCs and air toxics. J. Environ. Sci. Health A32, 2027–2055. Westerman, P.W., Bicudo, J.R., Kantardjieff, A., 2000. Upflow biological aerated filters for the treatment of flushed swine manure. Bioresour. Technol. 74, 181– 190. Williams, T.O., Miller, F.C., 1993. Composting facility odor control using biofilters. In: Hoitink, H.A., Keener, H.M. (Eds.), Science and Engineering of Composting: Design, Environmental, Microbiological and Utilization Aspects. pp. 262–281. Zhu, X., Suidan, M.T., Pruden, A., Yang, C., Alonso, C., Kim, B.J., Kim, B.R., 2004. Effect of substrate Henry’s constant on biofilter performance. J. Air Waste Manage. Assoc. 54, 409–418.
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Evaluation of full-scale biofilter with rockwool mixture treating ammonia gas from livestock manure composting Tomoko Yasuda *, Kazutaka Kuroda, Yasuyuki Fukumoto, Dai Hanajima, Kazuyoshi Suzuki Pollution Control Research Team, National Institute of Livestock and Grassland Science, 2, Ikenodai, Tsukuba, Ibaraki 305-0901, Japan
a r t i c l e
i n f o
Article history: Received 20 May 2008 Received in revised form 16 September 2008 Accepted 18 September 2008 Available online 31 October 2008 Keywords: Full-scale biofilter Nitrification Denitrification Rockwool Greenhouse gas
a b s t r a c t NH3 removal by a full-scale biofilter with rockwool packing materials was studied by measuring the gases and potential nitrification and denitrification activities of those materials in order to improve the biofiltration technology used in livestock farms. The rockwool biofilter was a durable and effective system for removing NH3, which was varied with the turning of manure composts. Furthermore, NH3 could be treated in the absence of an extra increase in two greenhouse gases, N2O and CH4. Potential nitrification and denitrification activities of the packing materials were estimated to be 8.2–12.2 mg N, and 1.42–4.69 mg N/100 g dry samples per day, respectively. The results suggested that potential nitrification and denitrification activities would increase within the biofilter where substrates, NH3 or NO 3 , have accumulated as a result of its operation. However, since percolate water contained high concentrations of NHþ 4 and NO3 , further improvement is required by reducing nitrogenous compounds within both the biofilter and percolate water. Ó 2008 Elsevier Ltd. All rights reserved.
1. Introduction Ammonia (NH3) is emitted from the composting process as a byproduct of the aerobic decomposition of organic materials. Such emissions amounted to 10–25% of the total nitrogen in swine manure (Kuroda et al., 1996; Osada et al., 2000; Fukumoto et al., 2003), and 9–76% in poultry droppings (Kirchmann and Witter, 1989; Kirchmann and Lundvall, 1998). The emitted NH3 causes deterioration of the environment in the form of offensive odors (Kuroda et al., 1996) as well as the acidification and eutrophication of soils and surface waters. Therefore, treatment of NH3 is crucial for the better management of livestock farming as well as for improving environmental quality. Biofiltration is a suitable option for treating NH3 emitted from the livestock composting process because of its high elimination efficiency, low operational cost, and modest environmental impact (Williams and Miller, 1993; Wani et al., 1997). This technology has been applied to composting facilities for well over 20 years, and the basic design and operating criteria have been well established. Additional studies are required to determine information on the microbiology and chemical reactions involved so that improvements in engineering and biofilter modeling can be achieved (Williams and Miller, 1993; Devinny and Ramesh, 2005). Many studies have used a laboratory-scale biofilter to determine the mecha* Corresponding author. Tel.: +81 29 838 8677; fax: +81 29 838 8606. E-mail address: [email protected] (T. Yasuda). 0960-8524/$ - see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.biortech.2008.09.033
nisms by which NH3 removal can be achieved (Hartikainen et al., 1996; Smet et al., 2000; Nicolai et al., 2006). Nitrification plays an important role in long-term effective NH3 removal. It has been estimated that nitrification accounts for 50% of NH3 removal (Smet et al., 2000). On the other hand, studies about denitrification in gas-phase biofilters have rarely been done (Zhu et al., 2004). Data on such factors as the distribution of gases and microbial activity in an actually operating biofilter are also limited. On-site NH3 loading, temperature, and other environmental conditions, all of which have an impact on microbial activity, are not constant. In addition to NH3, greenhouse gases, nitrous oxide (N2O) and methane (CH4) are emitted from the composting process (Kuroda et al., 1996; Osada et al., 2000; Veeken et al., 2002; Fukumoto et al., 2003). The emission patterns of these gases before and after biofiltration have not been investigated. Clemens and Cuhls (2003) reported that biofilters had no net effect on CH4, but that the concentration of N2O increased after filtration. If greenhouse gas emissions are high, the treatment of these gases should be taken into consideration when biofiltration is used in a composting facility. In this study, we evaluated NH3 removal by a full-scale biofilter with rockwool packing materials operating for a long time by measuring NH3 gas and potential nitrification and denitrification activities. The packing materials contain rockwool (fibrous material produced from igneous rocks), urethane, zeolite, and dried chicken feces (Fukumori et al., 2000). Emission patterns of two greenhouse gases, CH4 and N2O, before and after biofiltration were also investigated to determine how the biofilter affected them.
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filtering materials. Gas sampling points are shown in Fig. 1. Gas was collected in a 10-ml evacuated headspace vial and subsequently analysed for N2O and CH4. Biofilter packing materials were sampled using a liner soil sampler (5 cm ID, Daiki Rika Kogyo Co. Ltd., Japan) on June 22 of 2006, a few days before the next turning. Sampling points are shown in Fig. 1. Samples were taken from two different depths of 50 cm and 140 cm at each point. After sampling, aggregates were broken down by sieving (4.0 mm mesh), and urethane was cut into pieces about 1 cm in size. The separated particles were then mixed and stored at 4 °C until analysis. The biofilter temperature was monitored 50 cm below the surface by thermocouples equipped with a data logger (NR 1000, Keyence, Japan). The percolate water was sampled during the gas monitoring period in 2006.
2. Methods 2.1. Description of experimental site Our analysis required a biofilter and a composting facility, which have been operating at the National Institute of Livestock and Grassland Science (NILGS: Tsukuba City, Japan) since 1998 (Fig. 1). The biofilter is 5.8 m 8.0 m in cross-section and 3 m high (2.5 m for packing materials). Cow and swine manure is placed in the first reactor of the composting facility, then transferred to the next reactor and also turned at 3-week intervals. Gas emitted from the first three reactors (1st–3rd) is forced into the biofilter by three turbine blowers (No. 3BLF, Tecno Wasino Co. Ltd., Japan), each with an approximately 20 m3 min1 output flow, from underneath ditches providing uniform gas distribution. Gas and packing material contact time ranges from 100 to 200 s. Water (0.4 m3) is sprayed over the upper surface of the biofilter daily to maintain the moisture of packing materials, which are a mixture of mainly rockwool, urethane (5–10% v/v, only at 0.5–2 m below the surface), zeolite (5–10% w/w), and dried chicken feces (10–20% w/w) (Fukumori et al., 2000). Percolate water is currently treated along with the other livestock wastewater.
2.3. Analytical methods The NH3 concentration was measured with a detection tube and a gas sampler (Gastec Co. Ltd.). The N2O and CH4 concentrations were determined using GC with a 63Ni-electron capture detector (GC-14A, Shimadzu, Kyoto, Japan), and a flame ionization detector (6890 Series GC system, Agilent Technologies, Palo Alto, CA, USA), respectively. The moisture contents of the biofilter packing materials were determined as weight loss after drying overnight at 105 °C. About 5 g of packing materials was suspended in 45 ml of 2 M KCl and shaken for 30 min. After centrifugation at 2500 rpm for 10 min, the pH of the supernatant was measured by a glass electrode (B-211, Horiba, Japan), and defined as the pH of the sam ples (Kuroda et al., 1996). The amounts of NHþ 4 , NO2 and NO3 were measured in the same KCl extracts using a flow injection analyzer (Aquatec 5400, Tecator, Sweden). The total carbon (TC) and total
2.2. Sampling of gases, biofilter packing materials, and percolate water Concentrations of NH3 and greenhouse gases, N2O and CH4, prior to and after biofiltration were measured twice in November 2004 and May to June 2006. One or two turnings of the compost were included during the sampling period. Inlet gas to the biofilter was monitored from a sampling hole of the gas conduit. Outlet gas was collected using an inverted plastic funnel on the surface of the
Rockwool biofilter Roof
5.8 m
b
a
c
8.0 m
2.5 m
Tap water
ditch
Rockwool packing materials 3m 1
1st
2
3
2nd
3rd
Percolate water
4
4th
5th
6th
Compost flow
Composting barn Fig. 1. Diagram of biofilter and composting facility used for analysis. (black circles, sampling points of inlet air (1–4); letters, sampling points of outlet air and packing materials (a–c); circles with a cross, water nozzles).
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nitrogen (TN) contents of the biofilter packing materials were determined using an NC analyzer (NC-220F, SCAS, Japan). For the percolate water, NO 2 and NO3 were measured by ion chromatography (IC 7000 analyzer, Yokogawa, Japan, or HIC-VP super system, Shimadzu, Japan); NHþ 4 was measured by ion chromatography (DX120 analyzer, Dionex, Japan); and total carbon (TC) and total organic carbon (TOC) were measured by a TC/TOC analyzer (TOCVCSN, Shimadzu, Japan). Statistical analyses of differences of inorganic N, TC, and TN contents, and potential nitrification activities among samples were conducted using the GLM procedure of SAS (2002). 2.4. Determination of potential nitrification activity of biofilter packing materials Ammonium oxidation potential was determined based on some modifications of methods used in previous studies (Schmidt and Belser, 1994; Kimura, 1997). Triplicate portions (20–30 g) of samples were incubated at 30 °C in the dark in a 200-ml conical flask covered with aluminum foil. Water content was adjusted with (NH4)2SO4 solution at the start of incubation to 60% of water-holding capacity as determined by the Hilgard cup technique, and 100 g dry samples1. resulting in approximately 150 mg NHþ 4 NO2 and NO3 were measured after extraction with 2 M KCl, as described in Section 2.3. Potential nitrification activity was calculated from the amounts of NO 2þ3 produced during the 8-day incubation. 2.5. Determination of potential denitrification activity of biofilter packing materials Potential denitrification activities were determined based on the acetylene block methods of a previous study (Tiedje, 1994) with some modifications. Duplicate 2 g of subsamples were suspended in 5 ml of 5 mM phosphate buffer supplemented with 1 mM KNO3 and 1 mM glucose in a 30-ml test tube with butyl rubber cap and screw cap. Headspace gas was replaced with 10% C2H2 (Ar balance) by purging with the gas for 5 min with shaking. The final pressure of the headspace gas was 0.4 atm. Tubes were incubated at 25 °C in the dark. N2O in the headspace gas of each tube was measured after a 24-h incubation, as in Section 2.3. 3. Results and discussion As shown in Fig. 2, the biofilter with rockwool packing materials effectively removed NH3 gas. Complete NH3 elimination was also achieved in 2000 (data not shown), supporting durability of this biofilter. Inlet NH3 concentrations peaked after compost turning, and then slowly declined. Differences in the inlet concentrations among sampling points were ascribed to the composting process, although there was no partition in the gas conduit. The highest NH3 concentration was expected in inlet 1 because immediate NH3 emissions after turning were observed in laboratory-scale experiments, and it persisted while organic materials were degrading (Osada et al., 2000). Contrary to expectations, the peak height was double that of inlet 1 or higher at inlets 2, 3 and even at inlet 4 in November 2004 (Fig. 2). This was probably because the composting proceeded slowly due to the large amounts of manure (60–80 tons per day), as well as the lower temperature in winter (November 2004). On the whole, the concentration of NH3 in the mid-lower position of the biofilter was assumed to be high. The effects of biofiltration on the emission of two greenhouse gases, N2O and CH4, were examined (Fig. 2). A large portion of the inlet gases seemed to pass straight through the biofilter, although several microbial N2O and CH4 production and consumption reactions may have occurred to some degree. Inlet gas approx-
imately contained an average of 4 ppm of N2O, and an average of 430 ppm of CH4 in the range of 0–1800 ppm, respectively, throughout the sampling periods. According to previous studies, N2O emission from the composting process continued until the later phase of composting, whereas CH4 was mainly emitted at an earlier stage (Osada et al., 2000; Veeken et al., 2002; Fukumoto et al., 2003). The patterns shown in Fig. 2 are relevant to those prior studies. As for N2O, in 2004 and at the first turning in 2006, the concentrations in inlets 4 and 3 were relatively high, suggesting that N2O was continuously emitted during the composting process. As for CH4, the inlets 1 and 2 concentrations were relatively high in 2004 and at the second turning in 2006, suggesting that CH4 was emitted much more than at the initial composting. The reason for differences in emission patterns of both N2O and CH4 between the first and second tuning in 2006 was not clear. However, more CH4 was emitted from the undisturbed compost pile where less amounts of NH3 were emitted (Veeken et al., 2002). The emission of CH4 could be correlated with that of NH3 during the sampling period in 2006. It has been reported that a biofilter treating NH3 gas could be a major source of N2O, and in fact approximately 26% of the NH3–N that was removed by the biofilter was transformed into N2O (Clemens and Cuhls, 2003). Though the mechanism underlying N2O emissions in a biofilter has not yet been clearly understood, the results of our study confirmed that NH3 could be treated with rockwool mixture in the absence of an extra N2O increase. The moisture content, pH, inorganic N content, TC and TN of sampled packing materials are shown in Table 1. Moisture content averaged 47.2%. The pH showed a similar near-neutral value among the samples in the range of 6.8–7.4. The pH of the biofilter packing materials solution declined because of nitrification, with the result that more NH3 was absorbed into the solution. As a result, pH in the packing materials was maintained at a neutral level in this biofiltration system. The pH of percolate water was similar to that of the packing materials. That water never contained less than 1000 mg/l of NHþ 4 and NO3 during the sampling periods (Table 2). Within the biofilter, the NHþ 4 was low at site b, but there was no difference among the depths at every site. A significantly high level of NO 3 , which was probably leached into the water, was accumulated in the lower depths (Table 1). NO 2 contents were slightly less at the lower position than at the upper one (P < 0.05) (Table 1). Both TC and TN contents were higher at the lower position (P < 0.05), which was partly due to the urethanes used for increasing porosity. The mean temperature during the sampling periods was 19.2 °C (10.3–27.8 °C). Meanwhile, the temperature of packing materials was 25.5 °C (20.2–31.6 °C), which was higher and more stable than that of the atmosphere. Potential nitrification activity of the packing materials sampled in 2006 was determined by static incubation with an amendment of (NH4)2SO4. Their values were similar among samples except b140 cm, (8.96 ± 0.27 mg N per 100 g dry samples per day (SE), 7.96 ± 0.36, 8.20 ± 0.23, 9.48 ± 0.81, and 8.67 ± 0.45, for a-50 cm, a-140 cm, b-50 cm, c-50 cm, and c-140 cm, respectively). On the other hand, a significantly higher value was obtained in sample b-140 cm (12.2 ± 0.40, P < 0.05). From the NH3 emission patterns in 2006, it was assumed that more NH3 was loaded in the mid-lower part of the biofilter (relative to sampling point b). That result suggested that nitrification activity was affected by the NH3 concentration. Although high ammonia removal efficiency could be achieved for a high NH3 concentration by adsorption and absorption capacity of media, the performance declined when NH3 load exceeded those capacities and the biological activity (la Pagans et al., 2005). The maximum safe NH3 loading would depend on the nitrification activity. In this study, the potential nitrification activity obtained was estimated to be sufficient to oxidize the average amount of NH3–N that entered the biofilter per day during the gas sampling period (ca. 1 kg N load as opposed to ca. 2 kg N
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NH 3 (ppm)
(I) 160
140
140
120
120
100
100
80
80
60
60
40
40
20
20
N2 O (ppm)
0 11/8
11/13
11/18
0 5/13
12
12
10
10
8
8
6
6
4
4
2
2
0 11/8
CH 4 (ppm)
(II)
160
11/13
11/18
0 5/13
2000
2000
1600
1600
1200
1200
800
800
400
400
0 11/8
5/18
5/23
5/28
6/2
6/7
6/12
6/17
5/18
5/23
5/28
6/2
6/7
6/12
6/17
5/18
5/23
5/28
6/2
6/7
6/12
6/17
0 11/13
11/18
5/13
2004
2006 Date
Fig. 2. Changes in concentrations of NH3, N2O, and CH4, before and after biofiltration in 2004 (I) and 2006 (II) (inlet 1 (d), inlet 2 (j), inlet 3 (N), inlet 4 (), outlet a (s), outlet b (h), outlet c (4)). Arrows indicate turning time of manure composts.
Table 1 Characteristics of biofilter packing materials sampled in 2006 Sample origin (site-depth)
a-50 cm a-140 cm b-50 cm b-140 cm c-50 cm c-140 cm
Moisture (%, wt/wt)a
45.2 46.0 47.8 46.2 50.4 47.7
pHa
7.0 7.0 6.9 7.2 7.2 6.9
Inorganic N content (mg 100 g dry samples1)a NHþ 4 –N
NO 2 –N
14.1 (0.79)A 15.6 (2.84)A 3.54 (0.90)B 2.74 (0.79)B 18.0 (5.57)A 15.1 (2.43)A
0.14 0.12 0.14 0.12 0.15 0.10
(0.01)A (0.01)B (0.01)A (0.005)B (0.01)A (0.004)B
TC (%, wt/wt)a
TN (%, wt/wt)a
2.70 (0.09)A 12.2 (1.18)B 2.65 (0.16)A 8.73 (1.23)B 2.73 (0.38)A 8.89 (2.14)B
0.55 1.38 0.45 0.97 0.61 1.09
NO 3 –N 1.07 (0.01)A 18.4 (0.48)B 7.22 (0.26)C 26.0 (0.01)D 8.43 (0.19)C 39.0 (0.38)E
(0.03)A (0.10)B (0.05)A (0.13)B (0.10)A (0.25)B
Values in parentheses represent standard errors. Common superscripts within a column indicate no significant difference between values for those sites (P < 0.05). Values represent means of three replicates.
a
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Table 2 Characteristics of percolated water from the biofilter Sampling date
2006
6/6 6/9 6/13
pH
6.9 7.4 6.6
Inorganic N (mg/l) NHþ 4 –N
NO 2 –N
NO 3 –N
1186 1036 1223
119 209 286
1692 1562 2031
TC (mg/l)
TOC (mg/l)
70.8 77.5 71.2
58.4 57.9 61.8
potentially nitrified per day). However, the NH3 load fluctuated corresponding to the turnings, and more NH3 was also removed by absorption into the water (Table 2). Potential denitrification activity, determined with an amendment of glucose, tended to be high in the lower sampling position, where NO 3 has accumulated, (3.65 ± 0.52 mg N per 100 g dry samples per day (SE) and 4.69 ± 0.12 for sample a-50 cm and a-140 cm, 1.42 ± 0.18 and 2.50 ± 0.17 for b-50 cm and b-140 cm, and 1.59 ± 0.26 and 2.46 ± 0.08 for c-50 cm and c-140 cm, respectively). Denitrification has rarely been determined in an aerated biofilter because O2 transfer in the biofilm systems may not be limited (Zhu et al., 2004). Although direct evidence for denitrification in the biofilter was scarce, Nicolai et al. (2006) pointed out denitrification as one possible explanation for nitrogen disposition in the biofilter and Ho et al. (2008) showed the presence of Paracoccus denitrificans, which is capable of heterotrophic nitrification and denitrification, by molecular analysis. The occurrence of denitrification in biological aerated filters for the treatment of wastewaters was more frequently reported (e.g., Westerman et al., 2000; Garzón-Zúñiga et al., 2007; Gilbert et al., 2007). In those biofilter, the filtrating media have higher moisture and treat wastewaters containing high BOD (e.g., 9000 ppm; Garzón-Zúñiga et al., 2007) and their physicochemical condition may be largely different from those of ammonia gas biofilter. However, the results obtained in this study suggest that NO 3 could be removed within the biofilter under the appropriate conditions by denitrification. The TOC in the wastewater was as low as 60 mg/l (Table 2), suggesting that compounds which could be used as electron donors were low in the packing materials. Addition of easily degradable organic matter might enhance denitrification activity. In fact, it has already been demonstrated that NO 3 could serve as an electron acceptor for the degradation of VOC in an aerobic biofilter (Zhu et al., 2004). The Rockwool mixture initially contained dried chicken feces (10–20% w/w), but their degradable organic matter could be consumed during long-term operations. The NH3 gas pollution problem was converted to a water pollution problem. Further improvements in biofiltration technology are required to reduce nitrogenous compounds within both the biofilter and percolate water, especially when biofiltration systems are introduced to livestock farms that lack wastewater treatment plants or farmland on which percolate water could be applied as a liquid fertilizer. 4. Conclusion Biofiltration with a rockwool mixture proved to be a durable and effective system for removing NH3 emitted by the composting process. Furthermore, NH3 could be treated with rockwool mixture without any extra increase in N2O and CH4. Potential nitrification activity of the packing materials was estimated to be 8.2– 12.2 mg N per 100 g dry samples per day, which was sufficient to oxidize the average amount of NH3–N which entered the biofilter during the prior gas sampling period. Potential denitrification activity was estimated to be as high as 1.42–4.69 mg N per 100 g dry samples per day. Such results suggested that both potential
nitrification and denitrification activities could increase within the biofilter where substrates, NH3 or NO 3 , had accumulated in the course of its operation. Further improvements in this technology are required to reduce nitrogenous compounds within both the biofilter and percolate water. Acknowledgements The authors wish to thank Dr. Miyoko Waki for valuable advice on the experiments and Mrs. N. Akasaka and Mrs. K. Sumiya for their valuable assistance. References Clemens, J., Cuhls, C., 2003. Greenhouse gas emissions from mechanical and biological waste treatment of municipal waste. Environ. Technol. 24, 745–754. Devinny, J.S., Ramesh, J., 2005. A phenomenological review of biofilter models. Chem. Eng. J. 113, 187–196. Fukumori, I., Doshu, T., Ueno, K., 2000. In: Zen-noh (Ed.), Instruction Manual of Rockwool Biofilter. National Federation of Agricultural Co-operative Association (in Japanese). Fukumoto, Y., Osada, T., Hanajima, D., Haga, K., 2003. Patterns and quantities of NH3, N2O and CH4 emissions during swine manure composting without forced aeration – effect of compost pile scale. Bioresour. Technol. 89, 109–114. Garzón-Zúñiga, M.A., Lessard, P., Aubry, G., Buelna, G., 2007. Aeration effect on the efficiency of swine manure treatment in a trickling filter packed with organic materials. Water Sci. Technol. 55, 135–143. Gilbert, Y., le Bihan, Y., Aubry, G., Veillette, M., Duchaine, C., Lessard, P., 2007. Microbiological and molecular characterization of denitrification in biofilters treating pig manure. Bioresour. Technol. 99, 4495–4502. Hartikainen, T., Ruuskanen, J., Vanhatalo, M., Martikainen, P.J., 1996. Removal of ammonia from air by a peat biofilter. Environ. Technol. 17, 45–53. Ho, K.L., Chung, Y.C., Tseng, C.P., 2008. Continuous deodorization and bacterial community analysis of a biofilter treating nitrogen-containing gases from swine waste storage pits. Bioresour. Technol. 99, 2757–2765. Kimura, R., 1997. Measurement of nitrification activity, and enumeration and isolation of nitrifying bacteria. In: Committee of Soil Microbiology (Ed.), Experimental Methods for Soil Microbiology. Yokendo, Japan, pp. 207–214 (in Japanese). Kirchmann, H., Lundvall, A., 1998. Treatment of solid animal manures: identification of low NH3 emission practices. Nutr. Cycl. Agroecosyst. 51, 65–71. Kirchmann, H., Witter, E., 1989. Ammonia volatilization during aerobic and anaerobic manure decomposition. Plant Soil 115, 35–41. Kuroda, K., Osada, T., Yonaga, M., Kanematsu, A., Nitta, T., Mouri, S., Kojima, T., 1996. Emissions of malodorous compounds and greenhouse gases from composting swine feces. Bioresour. Technol. 56, 265–271. la Pagans, E., Font, X., Sanchez, A., 2005. Biofiltration for ammonia removal from composting exhaust gases. Chem. Eng. J. 113, 105–110. Nicolai, R.E., Clanton, C.J., Janni, K.A., Malzer, G.L., 2006. Ammonia removal during biofiltration as affected by inlet air temperature and media moisture content. Trans. ASABE 49, 1125–1138. Osada, T., Kuroda, K., Yonaga, M., 2000. Determination of nitrous oxide, methane, and ammonia emissions from a swine waste composting process. J. Mater. Cycles Waste Manage. 2, 51–56. SAS, 2002. Version 9.1. SAS Institute Inc., Cary, NC. Schmidt, E.L., Belser, L.W., 1994. Autotrophic nitrifying bacteria. In: Weaver, R., Angle, J.S., Bottomley, P.J. (Eds.), Methods of Soil Analysis, Part 2, Microbiological and Biochemical Properties. Social Science Society of America, Madison, WI, pp. 159–177. Smet, E., van Langenhove, H., Maes, K., 2000. Abatement of high concentrated ammonia loaded waste gases in compost biofilters. Water Air Soil Pollut. 119, 177–190. Tiedje, J.M., 1994. Denitrifiers. In: Weaver, R., Angle, J.S., Bottomley, P.J. (Eds.), Methods of Soil Analysis, Part 2. Microbiological and Biochemical Properties. Soil Science Society of America, Madison, WI, pp. 245–267. Veeken, A., de Wilde, V., Szanto, G., Hamelers, B., 2002. Passively aerated composting of straw-rich organic pig manure. In: Insam, H., Riddech, N., Klammer, S. (Eds.), Microbiology of Composting. Springer-Verlag, Berlin, Heidelberg, pp. 607–621. Wani, A.H., Branion, R.M.R., Lau, A.K., 1997. Biofiltration: a promising and costeffective control technology for odors, VOCs and air toxics. J. Environ. Sci. Health A32, 2027–2055. Westerman, P.W., Bicudo, J.R., Kantardjieff, A., 2000. Upflow biological aerated filters for the treatment of flushed swine manure. Bioresour. Technol. 74, 181– 190. Williams, T.O., Miller, F.C., 1993. Composting facility odor control using biofilters. In: Hoitink, H.A., Keener, H.M. (Eds.), Science and Engineering of Composting: Design, Environmental, Microbiological and Utilization Aspects. pp. 262–281. Zhu, X., Suidan, M.T., Pruden, A., Yang, C., Alonso, C., Kim, B.J., Kim, B.R., 2004. Effect of substrate Henry’s constant on biofilter performance. J. Air Waste Manage. Assoc. 54, 409–418.