Journal of Applied Microbiology ISSN 1364-5072
ORIGINAL ARTICLE
Impact of the addition of a nitrifying activated sludge on ammonia oxidation during composting of residual household wastes Y. Zeng1,2, P. Dabert1, S. Le Roux1, J. Mognol1, F.J. De Macedo1 and A. De Guardia1 1 Irstea, UR GERE, Rennes Cedex, France 2 Universit e Europeenne de Bretagne, Rennes, France
Keywords ammonia emissions, ammonia oxidation, bioaugmentation, composting, nitrification, nitrogen, residual household wastes, waste activated sludge. Correspondence Yang Zeng, Irstea, UR GERE, 17 avenue de Cucill e, CS 64427, F-35044 Rennes, France. E-mail: [email protected] 2014/0240: received 5 February 2014, revised 7 August 2014 and accepted 11 September 2014 doi:10.1111/jam.12651
Abstract Aims: To investigate the nitrogen–microbial community dynamics during composting of a mixture of nitrifying waste activated sludge (WAS) and fine organic fraction of residual household waste (RHW). To examine whether the addition of nitrifying sludge could promote ammonia oxidation and reduce ammonia emissions. Methods and Results: The fine organic fraction of RHW was mixed with the WAS and homogenized. The mixture and each waste alone were loaded in aerobic cells under controlled conditions, respectively. Both nitrogen and microbial community dynamics were monitored during 50 days of composting. The ammonia oxidizers were quantified and identified in the sludge and compost. The changes in ammonia-oxidizing bacteria (AOB) concentrations corresponded to the ammonia oxidation rates calculated from nitrogen balance. The addition of WAS did not efficiently reduce ammonia emissions because the Nitrosomonas oligotropha-like AOB introduced declined during the active stage of composting. Ammonia oxidation was probably limited by the intense heterotrophic activities at the active stage. Nitrosomonas europaea/eutropha and Nitrosomonas nitrosa-like AOB were established only during the maturation stage. They were the main contributors to ammonia oxidation during composting. Conclusions: The mixing of nitrifying WAS with the RHW during the early stages of composting does not promote ammonia oxidation nor reduce ammonia emissions because of limiting biologic factors during the active stage of composting. Significance and Impact of the Study: The mixing of activated sludge with RHW before composting is a common practice on composting plants. This study proved the limitation of this practice to reduce ammonia emissions during composting via bioaugmentation of ammonia-oxidizing organisms. It correlated successfully the ammonia oxidation rate with different groups of ammonia oxidizers and explains the fail of promoting ammonia oxidation during the early stages of composting. It suggests Nit. europaea/eutropha and Nit. nitrosa-like AOB were the main contributors to ammonia oxidation during composting.
Introduction The amount of household wastes (HW) has increased dramatically in recent years. For example, in France, the production of HW increased from 16 million tons in 1990 1674
to 22 million tons in 2005 (OECD 2008). To reduce this amount, several French governments have promoted the selective collection and recycling of materials like glass, paper, but also organic matter (OM). The fine organic fraction of residual household wastes (RHW) consists of
Journal of Applied Microbiology 117, 1674--1688 © 2014 The Society for Applied Microbiology
Y. Zeng et al.
organic particles that remain after selective collection. Composting is a conventional and valued method of dealing with the ever-increasing quantity of waste and produces stable compost for agricultural use. In 2007, around 861 thousand tons of HW were composted in France (ADEME 2009). Another increasing organic waste is sludge from wastewater treatment plants (WWTPs). In 2004, the amount of sludge produced reached 964 thousand tons (dry matter (DM)) in France (OECD 2008). This waste consists of primary but mainly of secondary sludge, also called waste activated sludge (WAS). In many plants, WAS is first dehydrated and then stored in tanks before being treated by either anaerobic digestion or composting prior to disposal. In 2006, composting plants in France treated 605 thousand tons of sewage sludge (ADEME 2009). The composting process includes two major stages: the active stage and the maturation stage (Tremier et al. 2005). The active stage corresponds to the decomposition of OM. This stage is characterized by considerable ammonia (NH3) emissions, which contribute to nitrogen (N) losses and to many other environmental problems. After the active stage, the maturation stage contributes to the synthesis of humus and humic acids. Nitrification is one of the key processes in the N cycle of our planet (Canfield et al. 2010). Ammonia-oxidizing bacteria and archaea (AOB, AOA) oxidize NH3 to nitrite (NO2). NO2 is then oxidized to NO3 by nitrite-oxidizing bacteria (NOB). NH3 oxidation is the rate-limiting step of nitrification (Yao et al. 2011). In composting processes, it is generally believed that nitrification occurs only during the maturation stage (Sommer 2001; Korner and Stegmann 2002) as nitrifying organisms are not thermophile and can develop only once the temperature of the compost pile has dropped. However, some recent studies discovered thermophilic nitrifying organisms in composting systems (Jarvis et al. 2009; Maeda et al. 2010). Moreover, our previous study revealed that, even at a moderate temperature (35°C), nitrification also occurred only during the maturation stage of RHW treatment (Zeng et al. 2012b). The consequence is that nitrification cannot serve as a sink for the NH3 produced by the mineralization of organic nitrogen (Norg) and that a large part of it is stripped by aeration during the active stage of composting (Zeng et al. 2012a). The late onset of nitrification may be due to, first, a very small quantity of nitrifying micro-organisms in the composted waste and their slow growth rate (Junier et al. 2010), and second, a strong competition between heterotrophic and nitrifying micro-organisms during the active stage of intense biodegradation of OM (Zeng et al. 2013). Several authors searched for ammonia oxidizers in composts. Beta-proteobacteria AOB were first detected in composts by Kowalchuk et al. (1999). Recent studies
Ammonia oxidation during composting
found both Nitrosomonas europaea/eutropha and Nitrosospira-like AOB in composting reactors and tried to link their development to the reduction in ammonia emissions (Jarvis et al. 2009; Maeda et al. 2010; Zeng et al. 2012b). In addition to AOB, amoA genes from Thaumarchaeota AOA were also detected during composting, and an important role for AOA in nitrification was suggested by two authors (Yamamoto et al. 2010, 2011; Zeng et al. 2011). The objective of the present work was to study N dynamics, and especially NH3 oxidation, during the composting of a mixture of RHW and nitrifying WAS. We hypothesized that the late onset of nitrification was due to the small quantity and slow growth rate of nitrifying micro-organisms. In the mixture of RHW and WAS, the ammonia oxidizing micro-organisms originating from WAS would oxidize NH3 and reduce NH3 emissions during the active stage of composting. Consequently, we monitored the dynamics of AOB and AOA groups using real-time PCR throughout the course of our experiment. To distinguish developments of both AOB and AOA from bacterial and archaeal multiplications, we monitored the microbial groups of total bacteria (TB) and total archaea (TA). Different to bacteria, fungi are another important microbial group in composting which contribute to biodegradation of cellulose and lignin (Ryckeboer et al. 2003; Gajalakshmi and Abbasi 2008). We also monitored this microbial group as total fungi (TF), which comprised Basidiomycetes and Ascomycetes. Monitoring continued for 50 days of treatment. The N balance was calculated after physical and chemical analyses. Ammonia oxidizers were identified by cloning and sequencing of amoA gene. Materials and methods Experimental devices and materials The experimental device is described in detail in our previous study (Zeng et al. 2012a,b). Briefly, it consists in a group of 12 airtight cylindrical cells with an internal volume of approximately 10 l. Waste materials mixed with bulking agent are loaded on a grid located 10 cm above the bottom of the cell. Below the grid, a glass screen homogenizes the incoming compressed air whose flow is regulated by a flow metre located upstream from each cell. The temperature and humidity of the compressed air are measured by a probe (dew point and temperature transmitter; Vaisala, Vantaa, Finland), which allow us to calculate the dry molar air flow applied. Once the materials are loaded, each cell is immersed in a water bath regulated at a temperature of 35°C by a thermostat. The temperature of the materials is measured in real time by a probe inserted into the middle of the cell. The incoming air
Journal of Applied Microbiology 117, 1674--1688 © 2014 The Society for Applied Microbiology
1675
Y. Zeng et al.
Ammonia oxidation during composting
passes through the materials and exits from the top of the cell. The outlet gas flow bubbles successively in two sulphuric acid traps (150 ml, 4% H2SO4) to absorb emitted NH3. The H2SO4 traps are changed every day. The outlet flow then passes through a bottle of water and an empty bottle to trap residual acid and condensed water. A set of valves, which is controlled according to time (12 min for each flow), selects one flow of the 13 flows (one incoming compressed air flow and 12 outlet flows) to be conveyed and analysed by gas analysers. The incoming flow of compressed air and the outlet flow of each cell are thus alternately analysed. The concentration of oxygen (O2) is measured using a paramagnetic analyser (Maihak Multor 640; SICK, Waldkirch, Germany) and the concentrations of carbon dioxide (CO2), methane (CH4) and nitrous oxide (N2O) are measured by an infrared detector (Uras 14; ABB Ltd., Zurich, Switzerland). The consumption of O2 and the production of CO2, CH4 and N2O are calculated by multiplying the dry molar air flow by the differences in the concentrations measured in the incoming compressed air and in the outlet flow from each cell. The RHW was collected at the composting site in Launay Lantic, France (15 000 tons year1). At this site, kitchen waste, paper, cardboard, textiles and plastic films are mixed together, incised and predegraded for 4 days in rotary tubes. Impurities are removed from the predegraded mixture before sieving to 10 mm. The sieved fraction is then mixed with green wastes to compost in windrows for 3 months. Our sample was collected immediately after sieving before green wastes were mixed in. The WAS was collected from the municipal WWTP in Mordelles, France (1850 m3 day1, 10 000 population equivalent). This plant removes N biologically by nitrification–denitrification in a single tank with intermittent aeration. The surplus sludge discharged from clarifiers passes through a draining table prior to storage. The WAS was collected at the outlet of the draining table. The characteristics of both RHW and WAS are listed in Table 1. Preparation of cells and experimental design The original WAS was centrifuged at 7000 g for 10 min to reduce its moisture content from 93% to around 88%. To ensure adequate aeration without absorption of N-containing material, we used Ø 16 mm polypropylene pall rings (PR) as bulk agents. For each of 10 identical cells (C1-C10), 1300 g of RHW, 1000 g of centrifuged WAS and 430 g of PR were mixed and homogenized. The moisture of the mixture was 70%. Two control cells were prepared with each type of waste alone. Cell control-1 was made of 1590 g RHW, 304 g PR and 770 g distilled water which brought the moisture to 70%. Cell control-2 was made of 2000 g centrifuged WAS and 1676
245 g PR. The moisture of control-2 was 88%. The volumetric ratio of waste to PR was 1 : 15. The temperature of the materials (RHW, WAS and mixture) was regulated at 35°C over the course of the whole treatment. This moderate temperature favours biodegradation (Tremier et al. 2005) and avoids the inhibition of nitrification that occurs at high temperatures (Myers 1975; Grunditz and Dalhammar 2001). The flow of compressed air for aeration was regulated at 76 l h1. The experiment was designed to monitor the dynamics of the N and NH3 oxidizers for a period of 50 days. One cell was stopped every 5 days for sampling because it is not possible to ensure homogeneity of material for sampling before separating composts from bulking agents, especially during the early stages of composting. It is extremely difficult to build the N balance once part of material was taken for sampling. Moreover, sampling changes the ratio of wastes to PR hence affecting the biodegradation process. Samples were stored at 20°C and thawed at 4°C before analysis. Residues in this cell were characterized as described below. To accelerate the process of biodegradation, materials in other cells were stirred (turning) every 5 days. Distilled water (200 g) was added during turning to keep the materials moisture. Physical and chemical analyses Physical and chemical analyses are described in detail in our previous study (Zeng et al. 2012b). DM content was measured by drying a fresh sample at 80°C until its mass loss was constant. OM content was measured by incinerating the dried sample at 550°C. Samples were thawed at 4°C before chemical analysis. The NH4+/NH3 content in the solid samples was directly analysed on fresh samples (2–10 g). According to the standard NF EN 13342 method, Kjeldahl nitrogen (KN) content of fresh samples (1–3 g) was quantified by mineralization, followed by distillation and titration using a mineralizer (Kjeldatherm TZ; C. Gerhardt, K€ onigswinter, Germany) and a distillation system (Vapodest 50; C. Gerhardt). The NH3 emitted and absorbed in the H2SO4 traps was analysed by steam distillation and titration with 01 mol l1 H2SO4 (method modified from NFT 90-015-1; Distillation unit B-324, B€ uchi, Flawil, Switzerland). The Norg content was calculated by subtracting the NH4+/NH3 from the KN. To measure the pH, NO2 and NO3 content, the solid samples were extracted with distilled water (1 : 5) for 2 h at room temperature. The pH of the extracts was measured. The extracts were then centrifuged. The NO2 and NO3 in the supernatants were measured by ionic chromatography using a Dionex DX-120 ion chromatograph. Each characteristic of the solid sample was measured in quintuplicate. H2SO4 traps and aqueous extracts of solid
Journal of Applied Microbiology 117, 1674--1688 © 2014 The Society for Applied Microbiology
Journal of Applied Microbiology 117, 1674--1688 © 2014 The Society for Applied Microbiology
15898 4344 (005)* 7384 (016) 362 (009) 056 1923 (014) n.d. – 004 (
ORIGINAL ARTICLE
Impact of the addition of a nitrifying activated sludge on ammonia oxidation during composting of residual household wastes Y. Zeng1,2, P. Dabert1, S. Le Roux1, J. Mognol1, F.J. De Macedo1 and A. De Guardia1 1 Irstea, UR GERE, Rennes Cedex, France 2 Universit e Europeenne de Bretagne, Rennes, France
Keywords ammonia emissions, ammonia oxidation, bioaugmentation, composting, nitrification, nitrogen, residual household wastes, waste activated sludge. Correspondence Yang Zeng, Irstea, UR GERE, 17 avenue de Cucill e, CS 64427, F-35044 Rennes, France. E-mail: [email protected] 2014/0240: received 5 February 2014, revised 7 August 2014 and accepted 11 September 2014 doi:10.1111/jam.12651
Abstract Aims: To investigate the nitrogen–microbial community dynamics during composting of a mixture of nitrifying waste activated sludge (WAS) and fine organic fraction of residual household waste (RHW). To examine whether the addition of nitrifying sludge could promote ammonia oxidation and reduce ammonia emissions. Methods and Results: The fine organic fraction of RHW was mixed with the WAS and homogenized. The mixture and each waste alone were loaded in aerobic cells under controlled conditions, respectively. Both nitrogen and microbial community dynamics were monitored during 50 days of composting. The ammonia oxidizers were quantified and identified in the sludge and compost. The changes in ammonia-oxidizing bacteria (AOB) concentrations corresponded to the ammonia oxidation rates calculated from nitrogen balance. The addition of WAS did not efficiently reduce ammonia emissions because the Nitrosomonas oligotropha-like AOB introduced declined during the active stage of composting. Ammonia oxidation was probably limited by the intense heterotrophic activities at the active stage. Nitrosomonas europaea/eutropha and Nitrosomonas nitrosa-like AOB were established only during the maturation stage. They were the main contributors to ammonia oxidation during composting. Conclusions: The mixing of nitrifying WAS with the RHW during the early stages of composting does not promote ammonia oxidation nor reduce ammonia emissions because of limiting biologic factors during the active stage of composting. Significance and Impact of the Study: The mixing of activated sludge with RHW before composting is a common practice on composting plants. This study proved the limitation of this practice to reduce ammonia emissions during composting via bioaugmentation of ammonia-oxidizing organisms. It correlated successfully the ammonia oxidation rate with different groups of ammonia oxidizers and explains the fail of promoting ammonia oxidation during the early stages of composting. It suggests Nit. europaea/eutropha and Nit. nitrosa-like AOB were the main contributors to ammonia oxidation during composting.
Introduction The amount of household wastes (HW) has increased dramatically in recent years. For example, in France, the production of HW increased from 16 million tons in 1990 1674
to 22 million tons in 2005 (OECD 2008). To reduce this amount, several French governments have promoted the selective collection and recycling of materials like glass, paper, but also organic matter (OM). The fine organic fraction of residual household wastes (RHW) consists of
Journal of Applied Microbiology 117, 1674--1688 © 2014 The Society for Applied Microbiology
Y. Zeng et al.
organic particles that remain after selective collection. Composting is a conventional and valued method of dealing with the ever-increasing quantity of waste and produces stable compost for agricultural use. In 2007, around 861 thousand tons of HW were composted in France (ADEME 2009). Another increasing organic waste is sludge from wastewater treatment plants (WWTPs). In 2004, the amount of sludge produced reached 964 thousand tons (dry matter (DM)) in France (OECD 2008). This waste consists of primary but mainly of secondary sludge, also called waste activated sludge (WAS). In many plants, WAS is first dehydrated and then stored in tanks before being treated by either anaerobic digestion or composting prior to disposal. In 2006, composting plants in France treated 605 thousand tons of sewage sludge (ADEME 2009). The composting process includes two major stages: the active stage and the maturation stage (Tremier et al. 2005). The active stage corresponds to the decomposition of OM. This stage is characterized by considerable ammonia (NH3) emissions, which contribute to nitrogen (N) losses and to many other environmental problems. After the active stage, the maturation stage contributes to the synthesis of humus and humic acids. Nitrification is one of the key processes in the N cycle of our planet (Canfield et al. 2010). Ammonia-oxidizing bacteria and archaea (AOB, AOA) oxidize NH3 to nitrite (NO2). NO2 is then oxidized to NO3 by nitrite-oxidizing bacteria (NOB). NH3 oxidation is the rate-limiting step of nitrification (Yao et al. 2011). In composting processes, it is generally believed that nitrification occurs only during the maturation stage (Sommer 2001; Korner and Stegmann 2002) as nitrifying organisms are not thermophile and can develop only once the temperature of the compost pile has dropped. However, some recent studies discovered thermophilic nitrifying organisms in composting systems (Jarvis et al. 2009; Maeda et al. 2010). Moreover, our previous study revealed that, even at a moderate temperature (35°C), nitrification also occurred only during the maturation stage of RHW treatment (Zeng et al. 2012b). The consequence is that nitrification cannot serve as a sink for the NH3 produced by the mineralization of organic nitrogen (Norg) and that a large part of it is stripped by aeration during the active stage of composting (Zeng et al. 2012a). The late onset of nitrification may be due to, first, a very small quantity of nitrifying micro-organisms in the composted waste and their slow growth rate (Junier et al. 2010), and second, a strong competition between heterotrophic and nitrifying micro-organisms during the active stage of intense biodegradation of OM (Zeng et al. 2013). Several authors searched for ammonia oxidizers in composts. Beta-proteobacteria AOB were first detected in composts by Kowalchuk et al. (1999). Recent studies
Ammonia oxidation during composting
found both Nitrosomonas europaea/eutropha and Nitrosospira-like AOB in composting reactors and tried to link their development to the reduction in ammonia emissions (Jarvis et al. 2009; Maeda et al. 2010; Zeng et al. 2012b). In addition to AOB, amoA genes from Thaumarchaeota AOA were also detected during composting, and an important role for AOA in nitrification was suggested by two authors (Yamamoto et al. 2010, 2011; Zeng et al. 2011). The objective of the present work was to study N dynamics, and especially NH3 oxidation, during the composting of a mixture of RHW and nitrifying WAS. We hypothesized that the late onset of nitrification was due to the small quantity and slow growth rate of nitrifying micro-organisms. In the mixture of RHW and WAS, the ammonia oxidizing micro-organisms originating from WAS would oxidize NH3 and reduce NH3 emissions during the active stage of composting. Consequently, we monitored the dynamics of AOB and AOA groups using real-time PCR throughout the course of our experiment. To distinguish developments of both AOB and AOA from bacterial and archaeal multiplications, we monitored the microbial groups of total bacteria (TB) and total archaea (TA). Different to bacteria, fungi are another important microbial group in composting which contribute to biodegradation of cellulose and lignin (Ryckeboer et al. 2003; Gajalakshmi and Abbasi 2008). We also monitored this microbial group as total fungi (TF), which comprised Basidiomycetes and Ascomycetes. Monitoring continued for 50 days of treatment. The N balance was calculated after physical and chemical analyses. Ammonia oxidizers were identified by cloning and sequencing of amoA gene. Materials and methods Experimental devices and materials The experimental device is described in detail in our previous study (Zeng et al. 2012a,b). Briefly, it consists in a group of 12 airtight cylindrical cells with an internal volume of approximately 10 l. Waste materials mixed with bulking agent are loaded on a grid located 10 cm above the bottom of the cell. Below the grid, a glass screen homogenizes the incoming compressed air whose flow is regulated by a flow metre located upstream from each cell. The temperature and humidity of the compressed air are measured by a probe (dew point and temperature transmitter; Vaisala, Vantaa, Finland), which allow us to calculate the dry molar air flow applied. Once the materials are loaded, each cell is immersed in a water bath regulated at a temperature of 35°C by a thermostat. The temperature of the materials is measured in real time by a probe inserted into the middle of the cell. The incoming air
Journal of Applied Microbiology 117, 1674--1688 © 2014 The Society for Applied Microbiology
1675
Y. Zeng et al.
Ammonia oxidation during composting
passes through the materials and exits from the top of the cell. The outlet gas flow bubbles successively in two sulphuric acid traps (150 ml, 4% H2SO4) to absorb emitted NH3. The H2SO4 traps are changed every day. The outlet flow then passes through a bottle of water and an empty bottle to trap residual acid and condensed water. A set of valves, which is controlled according to time (12 min for each flow), selects one flow of the 13 flows (one incoming compressed air flow and 12 outlet flows) to be conveyed and analysed by gas analysers. The incoming flow of compressed air and the outlet flow of each cell are thus alternately analysed. The concentration of oxygen (O2) is measured using a paramagnetic analyser (Maihak Multor 640; SICK, Waldkirch, Germany) and the concentrations of carbon dioxide (CO2), methane (CH4) and nitrous oxide (N2O) are measured by an infrared detector (Uras 14; ABB Ltd., Zurich, Switzerland). The consumption of O2 and the production of CO2, CH4 and N2O are calculated by multiplying the dry molar air flow by the differences in the concentrations measured in the incoming compressed air and in the outlet flow from each cell. The RHW was collected at the composting site in Launay Lantic, France (15 000 tons year1). At this site, kitchen waste, paper, cardboard, textiles and plastic films are mixed together, incised and predegraded for 4 days in rotary tubes. Impurities are removed from the predegraded mixture before sieving to 10 mm. The sieved fraction is then mixed with green wastes to compost in windrows for 3 months. Our sample was collected immediately after sieving before green wastes were mixed in. The WAS was collected from the municipal WWTP in Mordelles, France (1850 m3 day1, 10 000 population equivalent). This plant removes N biologically by nitrification–denitrification in a single tank with intermittent aeration. The surplus sludge discharged from clarifiers passes through a draining table prior to storage. The WAS was collected at the outlet of the draining table. The characteristics of both RHW and WAS are listed in Table 1. Preparation of cells and experimental design The original WAS was centrifuged at 7000 g for 10 min to reduce its moisture content from 93% to around 88%. To ensure adequate aeration without absorption of N-containing material, we used Ø 16 mm polypropylene pall rings (PR) as bulk agents. For each of 10 identical cells (C1-C10), 1300 g of RHW, 1000 g of centrifuged WAS and 430 g of PR were mixed and homogenized. The moisture of the mixture was 70%. Two control cells were prepared with each type of waste alone. Cell control-1 was made of 1590 g RHW, 304 g PR and 770 g distilled water which brought the moisture to 70%. Cell control-2 was made of 2000 g centrifuged WAS and 1676
245 g PR. The moisture of control-2 was 88%. The volumetric ratio of waste to PR was 1 : 15. The temperature of the materials (RHW, WAS and mixture) was regulated at 35°C over the course of the whole treatment. This moderate temperature favours biodegradation (Tremier et al. 2005) and avoids the inhibition of nitrification that occurs at high temperatures (Myers 1975; Grunditz and Dalhammar 2001). The flow of compressed air for aeration was regulated at 76 l h1. The experiment was designed to monitor the dynamics of the N and NH3 oxidizers for a period of 50 days. One cell was stopped every 5 days for sampling because it is not possible to ensure homogeneity of material for sampling before separating composts from bulking agents, especially during the early stages of composting. It is extremely difficult to build the N balance once part of material was taken for sampling. Moreover, sampling changes the ratio of wastes to PR hence affecting the biodegradation process. Samples were stored at 20°C and thawed at 4°C before analysis. Residues in this cell were characterized as described below. To accelerate the process of biodegradation, materials in other cells were stirred (turning) every 5 days. Distilled water (200 g) was added during turning to keep the materials moisture. Physical and chemical analyses Physical and chemical analyses are described in detail in our previous study (Zeng et al. 2012b). DM content was measured by drying a fresh sample at 80°C until its mass loss was constant. OM content was measured by incinerating the dried sample at 550°C. Samples were thawed at 4°C before chemical analysis. The NH4+/NH3 content in the solid samples was directly analysed on fresh samples (2–10 g). According to the standard NF EN 13342 method, Kjeldahl nitrogen (KN) content of fresh samples (1–3 g) was quantified by mineralization, followed by distillation and titration using a mineralizer (Kjeldatherm TZ; C. Gerhardt, K€ onigswinter, Germany) and a distillation system (Vapodest 50; C. Gerhardt). The NH3 emitted and absorbed in the H2SO4 traps was analysed by steam distillation and titration with 01 mol l1 H2SO4 (method modified from NFT 90-015-1; Distillation unit B-324, B€ uchi, Flawil, Switzerland). The Norg content was calculated by subtracting the NH4+/NH3 from the KN. To measure the pH, NO2 and NO3 content, the solid samples were extracted with distilled water (1 : 5) for 2 h at room temperature. The pH of the extracts was measured. The extracts were then centrifuged. The NO2 and NO3 in the supernatants were measured by ionic chromatography using a Dionex DX-120 ion chromatograph. Each characteristic of the solid sample was measured in quintuplicate. H2SO4 traps and aqueous extracts of solid
Journal of Applied Microbiology 117, 1674--1688 © 2014 The Society for Applied Microbiology
Journal of Applied Microbiology 117, 1674--1688 © 2014 The Society for Applied Microbiology
15898 4344 (005)* 7384 (016) 362 (009) 056 1923 (014) n.d. – 004 (
