Effect of free ammonia nitrogen on the methanogenic activity of swine wastewater

Marisol Belmonte¹ · Chia-Fang Hsieh¹ · Cindy Figueroa¹ · José Luis Campos² · Gladys Vidal*¹

¹Grupo de Ingeniería y Biotecnología Ambiental, Centro de Ciencias Ambientales EULA-Chile, Universidad de Concepción, Concepción, Chile
²Departamento de Ingeniería Química, Escuela Técnica Superior de Ingeniería, Universidad de Santiago de Compostela, Santiago de Compostela, España

*Corresponding author: glvidal@udec.cl

Financial support: This work was supported by CONICYT/PBCT (Grant TPI-01), FONDECYT (Grant 1090414) and Innova Bío-Bio (Grant 07-PC S1-198).

Keywords: free ammonia nitrogen, methanogenic activity, swine wastewater.

Swine wastewater is characterized by high organic matter content, solids, nitrogen (expressed as total ammonia and protein) and heavy metals. This work determines the methanogenic toxicity effect of free ammonia contained in swine wastewater comparing raw swine wastewater (RW) and the liquid fraction of swine wastewater (TW). The values of IC₅₀ (50% of inhibition) obtained for methanogenic bacteria ranged between 56 and 84% for RW, meanwhile IC₅₀ for TW was ranged between 84 and 94%. Such inhibitory effects can be related to the free ammonia nitrogen concentration (> 40 mg NH₃-N/L) contained in swine wastewater.

Pollution caused by livestock wastes and by-products (including slurries, manure, air emissions, pigs and veterinary wastes) is a growing concern in many countries (Babot and Andres, 2004). Due to the environmental impacts on water (eutrophization), soil (acidification and toxicity) and air produced by the intensive livestock explotation (Choi, 2007; González-Fernández et al. 2008a). The sustainability of this activity is becoming a priority and environmental management of intensive livestock production is observed as the solution.

Swine wastewaters are characterized by the presence of the animal feces (55%), urine (45%) and wash water (Babot and Andres, 2004). The animals' physiological state and diet as well as the buildings and waste management period result in different physical-chemical characteristics (Magrí et al. 2009). In general, swine wastewater characterization shows high organic matter (4.2-12.3 g BOD₅/L), electric conductivity (5150-8425 µS/cm), solids (total: 12.6-53.9 g/L; volatile: 7.8-21.7 g/L), fecal coliforms (4.2 x 10⁷-1.9 x 10⁸ MPN 100/mL) and nutrients (4.8-720 mg/L of total nitrogen, 486-954 mg NH₄⁺-N/L, 266-1600 mg P/L) (Teira-Esmatges and Flotats 2003; Ahn et al. 2004; Boursier et al. 2005; Vanotti, et al. 2007; González-Fernández et al. 2008a; Rodríguez et al. 2011). In addition to heavy metals (e.g. 30-40 mg Cu/L; 60 mg Zn/L), hormones and antibiotics are present in swine wastewater (Choi, 2007; Ben et al. 2008; Flotats et al. 2009).

Due the high waste production and the low land availability on intensive pig farms, pig slurry must be treated to comply with the environmental regulations. In this sense, anaerobic digestion is a well-known wastewater treatment that allows simultaneously the reduction of organic matter in swine wastewater and methane production (Campos et al. 2008). In order to apply this process, two different strategies are possible: a) To treat raw swine wastewater in one single reactor or; b) To separate previously both solid and liquid fractions and to treat each one in two different kinds of anaerobic systems. When solid and liquid fractions are not separated, anaerobic digesters with high hydraulic retention times are required due to the low rate of solids hydrolysis. Nevertheless, if solids are previously separated, the liquid fraction, which supposes around 80% of the flow rate, could be treated in high rate systems such as upflow anaerobic sludge blanket (UASB) reactors (Karakashev et al. 2008) and, therefore, the volume needed for the anaerobic sludge digesters is reduced.

On the other hand, the solid fraction of slurry contains around 30-50% of the Kjeldahl nitrogen and 80% of both copper and zinc (Burton, 2007; Moral et al. 2008). Therefore, during its hydrolysis under anaerobic conditions the generation of ammonia and/or release of heavy metals could cause the inhibition of methanogenic bacteria (Koster and Lettinga, 1988; Demirel and Yenigün, 2002; Karri et al. 2006; Chen et al. 2008). Until now, most of the studies about anaerobic digestion of the solid fraction were focused on improving the rate of the hydrolysis step (Rodríguez-Andara and Lomas-Esteban, 1999; Menardo et al. 2010) or determining its biodegradability (Campos et al. 2008; González.-Fernández, 2008b). Several authors have studied the anaerobic digestion to treated swine wastewater. Nevertheless, the possible differences of methanogenic activity treating raw swine wastewater or liquid fraction swine wastewater were not yet studied. For this reason, the main objective of the work is to determine the effect of the free ammonia nitrogen (NH₃-N) on the methanogenic activity contained in both raw (RW) and the liquid fraction (TW) of swine wastewater.

Swine wastewater

Slurry was obtained from a piggery farm located in southern Chile. Raw wastewater (RW) was sampled after the primary settling treatment and was keeping refrigerated at 4ºC. Treated wastewater (TW) had been previously filtered using a membrane (0.45 μm) to remove solid content. RW and TW were used as influent for feeding the methanogenic toxicity assays. Table 1 shows the physicochemical characteristics of RW and TW.

Inoculum

Two different kinds of anaerobic biomass were used to carry the methanogenic toxicity assays: a) anaerobic flocculent sludge with a maximum specific methanogenic activity of 0.1 g COD/g VSS·d which was collected from an anaerobic digester located in the municipal wastewater treatment plant and; b) granular sludge with a maximum specific methanogenic activity of 0.3 g COD/g VSS·d which was collected from an anaerobic digester treating brewing effluent.

Methanogenic toxicity assays

The methanogenic toxicity assays were carried out in 100 mL of total volume (the glass serum bottle volume was 125 mL) using a volatile fatty acid (VFA) as described by Vidal and Diez (2003) and Soto et al. (1993). Methane production was measured by displacement of an alkaline solution (2.5% NaOH). A neutralized VFA stock solution was used to provide a final concentration of acetic acid of 2 g/L, propionic acid of 0.5 g/L, and n-butyric acid of 0.5 g/L (total COD from VFA was 3.8 g COD-VFA/L). The media contained also the following nutrients (per liter): NH₄Cl, 0.14 g; K₂HPO₄, 0.125 g; MgSO₄, 7H₂O 0.1 g; CaCl₂ 2H₂O, 0.01 g; NaHCO₃, 0.2 g. In order to reduce the medium, 100 mg Na₂S 9H₂O/L was added. Diluted HCl or NaOH solutions were used to adjust the initial pH to 7.0 ± 0.1. 2.0 mL of the neutralized VFA stock solution, 10.0 mL of the nutrients, 1.0 mL of the Na₂S 9H₂O and 21 mL of biomass (to obtain the desired biomass concentration), were added to obtain a total volume of 34 mL. Then, the liquid volume was completed until 100 mL with: a) 66 mL of water (control test); b) 66 mL of RW (RW 100%); c) 66 mL of TW (TW 100%); d) 33 mL of water and 33 mL of RW (RW 50%) or; e) 33 mL of water and 33 mL of TW (TW 50%). Later nitrogen gas was bubbled up into each flask to remove air from the headspace for 3 min at a pressure of 1.5 Pa. The assays were carried out at 37ºC.

Three successive feedings were made during each experiment. During the first feeding, the sludge was exposed to media containing effluent and VFA substrate. At the end of the first feeding, the spent medium (liquid phase) was carefully decanted and the sludge was again exposed to effluent and VFA substrate. At the end of the second feeding, the spent medium was removed. In order to test the residual sludge activity after the first and second exposure, a third feeding containing only the VFA stock solution as substrate was added to the culture bottles (total volume 100 mL).

These experiments were carried out with both kinds of biomass and using two different biomass concentrations (2.0 and 5.0 g VSS/L) (Table 2).

The methanogenic activity, AC_m (g COD/g VSS·d), was determined according to Equation 1 (Campos et al. 2008).

                    Equation 1

Where: R, methane production rate (mL CH₄/d), f is the conversion factor from CH₄ to g COD (385 mL CH₄/g COD), V is the liquid phase volume (0.1 L) and [VSS] is the biomass concentration (g VSS/L).

The results are reported as the percentage of methanogenic activity of the each assay in relation to the methanogenic activity of the control (%AC_m), considered according to Equation 2.

      Equation 2

Where: AC_m(Assays): methanogenic activity of each assay, AC_m(Control): methanogenic activity of the control. The %AC_m -obtained values were plotted against the wastewater COD concentration used in each assay in order to calculate the IC₅₀ (%) by data interpolation (Vidal et al. 1997).

Analytical methods

The COD (chemical oxygen demand), Total Suspend Solids (TSS), Volatile Suspend Solids (VSS), ammonium (NH₄⁺-N), total nitrogen (TN) and total phosphorus (TP) were determined according to the protocols established in Standard Methods (APHA, 1985). The pH was evaluated using electrodes. The Total Alkalinity (TA), Intermediate Alkalinity (IA, which is the alkalinity due to VFA) and Partial Alkalinity (PA) were determined by titration (Ripley et al. 1986). To evaluate the effect of the solids on methanogenic toxicity, the swine wastewater was filtered through a 0.45 μm pore size nitrocellulose membrane. Heavy metals, such as copper and zinc, were determined by atomic absorption flame spectrometry (NCh 2313/10 Of. 96).

RW and TW characterization

Table 1 shows the physicochemical characterization of the RW and TW. Total and volatile solid of RW and TW were 20.4 and 11.2 g TS/L, and 14.6 and 8.9 g VS/L, respectively. Organic matter measured as total COD was 25.2 and 16.0 g COD/L for RW and TW, respectively. On the other hand, total nitrogen and NH₄⁺-N were 3.2 g TN/L and 2.7 g NH₄⁺-N/L and 2.1 g TN/L and 1.9 g NH₄⁺-N/L for RW and TW, respectively. Similar concentrations for total solids (26.5-24.9 g TS/L) and total COD (18.7-30.1 g/L) were reported by Magrí et al. (2009).

In this study, the filtration treatment caused a removal efficiency of total and volatile solid around 55 and 61%, respectively. Similar results have been reported by González-Fernández et al. (2008b), who achieved efficiencies of 59% and 57% for total and volatile solids, respectively. In this study, the efficiency of COD reduction was about 64%, while González-Fernández et al. (2008b) obtained values about 57%. Moreover, a reduction of TN around 40% was obtained. This percentage agrees with that found by González-Fernández et al. (2008b) when applied polyacrylamide to remove solids from swine wastewater. These authors also a screen with a pore size of 0.25 mm but, in this case, they only achieved a reduction of TN around 10-15%. Therefore, it can be inferred that the efficiency of TN reduction is closely related to the percentage of solids removed. Typical in this type of effluent is the alkalinity presence, regulating the pH according to several authors (Bonmatí and Flotats, 2003; González-Fernández and García-Encina, 2009) as in this study.

Heavy metal concentrations showed that only 6.6% of cooper and 11.9% of zinc are related to the liquid fraction (TW). These results agree with those found by Marcato et al. (2008) who found that most of Cu and Zn (86%) are trapped within particles between 3 and 25 µm.

Methanogenic activity assays

Figure 1 shows the evolution of the methane production of the different assays. In all the cases, the steady state was reached after 150 hrs of assays. Nevertheless, in those assays with flocculent sludge (Figure 1 A.1, B.1, B.2) was observed a lag-phase during the first 25 hrs of operation. On the other hand, in assays using granular sludge as inoculum lag-phase was not observed (Figure 1 C.1, C.2, D.2). On the contrary, Vidal and Diez (2005) show a lag-phase > 5 days for sludge flocculent and fiberboard effluents. For this study, methanogenic inhibition was evident in all the assays (Figure 1). Using RW as substrate and flocculent sludge as biomass a reduction of the methanogenic activity was observed expressed as a methane reduction attain 10% respect to the control assays. So, methane production by RW assay was even lower of 15 mL CH₄ regarding 100 mL CH₄ approximately of the control in the operation time (Figure 1 A.1, C.1, C.2). However, granular sludge was generated a 20-25% more of methane production than when using flocculent sludge.

Table 3 shows the AC_m evaluated from Equation 2. High nitrogen concentrations (> 2.0 g N/L) characteristic of this type of wastewater and a high pH (> 7.5) causes the presence of free ammonia that can bring about the inhibition of some bacterial populations methanogenics (aceticlastic and hydrogenotrophic metanogens) (Guerrero et al. 1997; Chen et al. 2008). In this study, free ammonia nitrogen (NH₃-N) concentration was calculated for each test according to Anthonisen et al. (1976) and was correlated to the percentage of methanogenic activity obtained (Figure 2). There is a clear tendency between both parameters which would indicate that the inhibitory effect should be caused by NH₃-N for RW and TW (Figure 2). Moreover, Figure 2a shows that 40 mg NH₃-N/L contained in RW produce a reduction until 20% of the methanogenic activity under this conditions the study. Chen et al. (2008) are indicated that anaerobic digestion process can be inhibited due to concentrations at 50% by a total ammonium nitrogen concentration above 1.7 g/L. Regarding this, Guerrero et al. (1997) show that methanogenic toxicity is produced by concentration from 25 to 140 mg NH₃-N/L in mesophilic process. However, Nielsen and Angelidaki (2008) show that, the inhibition anaerobic digestion of livestock manure are caused by concentrations between 80-100 mg NH₃-N/L. Although adapted biomass can support free ammonia concentrations up to 3 g NH₃-N/L (Guerrero et al. 1997; Nielsen and Angelidaki, 2008; González-Fernández and García-Encina, 2009). Furthermore, Rodríguez et al. (2011) observed a very good performance of the anaerobic continuous system under concentration of ammonia higher than 375 mg NH₃-N/L.

On the other hand, heavy metals (Cu and Zn) contained on the TW (0.12 mg Cu/L and 0.89 mg Zn/L) or RW (Cu: 1.83 mg/L and Zn: 7.53 mg/L) were found as a potential inhibitor to the AC_m. Although these concentrations are lower in comparison with those reported by Choi (2007) for swine wastewater (Cu: 30-40 mg/L, Zn: 60 mg/L), evidence of the methanogenic toxicity was not found when RW was feeding at 100%. On the other hand, Karri et al. (2006) found that Cu²⁺ concentrations (< 20.0 mg/L) could cause the reduction in the hydrogenotrophic methanogenic bacteria activity in 50% (IC₅₀) with 8.9 mg Cu²⁺/L. Additionally, zinc concentrations between 5.0 and 40.0 mg/L can be inhibitory to the methanogenic activity (Demirel and Yenigün, 2002).

Table 4 shows the IC₅₀ expressed as dilution of the effluent (%) during the second feeding for the different assays. So, the IC₅₀ for RW in generally was ranged between 56-84%, meanwhile IC₅₀ for TW was determined only for the granular sludge, was ranged between 84-94% found for the granular sludge. Similar results were found by Vidal and Diez (2005) in similar aggregates sludge structures but for different compounds. All the previous results show that TW is less toxic for methanogenic bacteria in comparison with RW. 

This study shows that methanogenic bacteria -under this conditions investigated, were inhibited by RW (IC₅₀: 56 and 84%). This reduction in the methanogenic activity is associated to the presence of free ammonia nitrogen (NH₃-N) (> 40 mg NH₃-N/L) in the system. On the other hand, a lower reduction of the methanogenic toxicity effect (IC₅₀ for TW: between 84 and 94%) was evaluated by TW.

The authors thank Mr. C. Contreras from Sucesion Yanine for allowing the realization of this study in their facilities.

AHN, Y.H.; HWENG, I.S. and MIN, K.S. (2004). ANAMMOX and partial denitritation in anaerobic nitrogen removal from piggery waste. Water Science and Technology, vol. 49, no. 5-6, p. 145-153.

ANTHONISEN, A.C.; LOEHR, R.C.; PRAKASAM, T.B.S. and SRINATH, E.G. (1976). Inhibition of nitrification by ammonia and nitrous acid. Journal of Water Pollution Control Federation, vol. 48, no. 5, p. 835-852.

AMERICAN PUBLIC HEALTH ASSOCIATION (APHA). (1985). Standard methods for examination of water and wastewater 16^th ed. APHA, Washington, D.C. 1268 p. ISBN 0-87553-131-8.

BABOT, D. and ANDRES, N. (2004). Technical of environmental management in the piggery production. Edition of Universitat of Lleida. Catalunya, Spain. 134 p. ISBN 84-688-7858-8.

BEN, W.; QIANG, Z.; ADAMS, C.; ZHANG, H. and CHEN, L. (2008). Simultaneous determination of sulfonamides, tetracyclines and tiamulin in swine wastewater by solid-phase extraction and liquid chromatography-mass spectrometry. Journal of Chromatography A., vol. 1202, no. 2, p. 173-180. [CrossRef]

BONMATÍ, A. and FLOTATS, X. (2003). Air stripping of ammonia from pig slurry: Characterisation and feasibility as a pre- or post-treatment to mesophilic anaerobic digestion. Waste Management, vol. 23, no. 3, p. 261-272. [CrossRef]

BOURSIER, H.; BÉLINE, F. and PAUL, E. (2005). Piggery wastewater characterisation for biological nitrogen removal process design. Bioresource Technology, vol. 96, no. 3, p. 351-358. [CrossRef]

BURTON, C.H. (2007). The potential contribution of separation technologies to the management of livestock manure. Livestock Science, vol. 112, no. 3, p. 208-216. [CrossRef]

CAMPOS, E.; ALMIRALL, M.; MTNEZ-ALMELA, J.; PALATSI, J. and FLOTATS, X. (2008). Feasibility study of the anaerobic digestion of dewatered pig slurry by means of polyacrylamide. Bioresource Technology, vol. 99, no. 2, p. 387-395. [CrossRef]

CHEN, Y.; CHENG, J.J. and CREAMER, K.S. (2008). Inhibition of anaerobic digestion process: A review. Bioresource Technology, vol. 99, no. 10, p. 4044-4064. [CrossRef]

CHOI, E. (2007). Piggery waste management: Towards a sustainable future. IWA Publishing, 174 p. ISBN 1843391317.

DEMIREL, B. and YENIGÜN, O. (2002). Two-phase anaerobic digestion processes: A review. Journal of Chemical Technology and Biotechnology, vol. 77, no. 7, p. 743-755. [CrossRef]

FLOTATS, X.; BONMATÍ, A.; FERNÁNDEZ, B. and MAGRÍ, A. (2009). Manure treatment technologies: On-farm centralized strategies. NE Spain as case study. Bioresource Technology, vol. 100, no. 22, p. 5519-5526. [CrossRef]

GONZÁLEZ-FERNÁNDEZ, C.; LEÓN-COFRECES, C. and GARCÍA-ENCINA, P.A. (2008a). Different pretreatments for increasing the anaerobic biodegradability in swine manure. Bioresource Technology, vol. 99, no. 18, p. 8710-8714. [CrossRef]

GONZÁLEZ-FERNÁNDEZ, C.; NIETO-DIEZ, P.P.; LEÓN-COFRECES, C. and GARCÍA-ENCINA, P.A. (2008b). Solids and nutrients removals from the liquid fraction of swine slurry through screening and flocculation treatment and influence of these processes on anaerobic biodegradability. Bioresource Technology, vol. 99, no. 14, p. 6233-6239. [CrossRef]

GONZÁLEZ-FERNÁNDEZ, C. and GARCÍA-ENCINA, P.A. (2009). Impact of substrate to inoculum ratio anaerobic digestion of swine slurry. Biomass and Bioenergy, vol. 33, no. 8, p. 1065-1069. [CrossRef]

GUERRERO, L.; OMIL, F.; MENDEZ, R. and LEMA, J.M. (1997). Treatment of saline wastewaters from fish meal factories in an anaerobic filter under extreme ammonia concentrations. Bioresource Technology, vol. 61, no. 1, p. 69-78. [CrossRef]

KARAKASHEV, D.; SCHMIDT, J.E. and ANGELIDAKI, I. (2008). Innovative process scheme for removal of organic matter, phosphorus and nitrogen from pig manure. Water Research, vol. 42, no. 15, p. 4083-4090. [CrossRef]

KARRI, S.; SIERRA-ALVAREZ, R. and FIELD, J.A. (2006). Toxicity of copper to acetoclastic and hydrogenotrophic activities of methanogens and sulfate reducers in anaerobic sludge. Chemosphere, vol. 62, no. 1, p. 121-127. [CrossRef]

KOSTER, I. and LETTINGA, G. (1988). Anaerobic digestion at extreme ammonia concentration. Biological Wastes, vol. 25, no. 1, p. 51-59. [CrossRef]

MAGRÍ, A.; GUIVERNAU, M.; BAQUERIZO, G.; VIÑAS, M.; PRENAFETA-BOLDÚ, F. and FLOTATS, X. (2009). Batch treatment of liquid fraction of pig slurry by intermittent aeration: Process simulation and microbial community analysis. Journal Chemical Technology and Biotechnology, vol. 84, no. 8, p. 1202-1210. [CrossRef]

MARCATO, C.E.; PINELLI, E.; POUECH, P.; WINTERTON, P. and GUIRESSE M. (2008). Particle size and metal distributions in anaerobically digested pig slurry. Bioresource Technology, vol. 99, no. 7, p. 2340-2348. [CrossRef]

MENARDO, S.; BALSARI, P.; DINUCCIO, E. and GIOELLI, F. (2010). Thermal pre-treatment of solid fraction from mechanically-separated raw and digested slurry to increase methane yield. Bioresource Technology, vol. 102, no. 2, p. 2026-2032. [CrossRef]

MORAL, R.; PEREZ-MURCIA, M.D.; PEREZ-ESPINOSA, A.; MORENO-CASELLES, J.; PAREDES, C. and RUFETE, B. (2008). Salinity, organic content, micronutrients and heavy metals in pig slurries from South-eastern Spain. Waste Management, vol. 28, no. 2, p. 367-371. [CrossRef]

NIELSEN, H.B. and ANGELIDAKI, I. (2008). Strategies for optimizing recovery of the biogas process following ammonia inhibition. Bioresource Technology, vol. 99, no. 17, p. 7995-8001. [CrossRef]

RIPLEY, L.E.; BOYLE, W. and CONVERSE, J. (1986). Improved alkalimetric monitoring for anaerobic digestion of high strength wastes. Journal Water Pollution Control Federation, vol. 58, no. 5, p. 406-411.

RODRÍGUEZ-ANDARA, A. and LOMAS-ESTEBAN, J.M. (1999). Kinetic study of the anaerobic digestion of the solid fraction of piggery slurries. Biomass and Bioenergy, vol. 17, no. 5, p. 435-443. [CrossRef]

RODRÍGUEZ, D.C.; BELMONTE, M.; PEÑUELA, G.; CAMPOS, J.L. and VIDAL, G. (2011). Behavior of molecular weight distribution for the pig slurry liquid fraction treated by anaerobic digestion. Environmental Technology, vol. 32, no. 4, p. 419-425. [CrossRef]

SOTO, M.; MÉNDEZ, R. and LEMA, J.M. (1993). Methanogenic and non-methanogenic activity tests. Theoretical basis and experimental set up. Water Research, vol. 27, no. 8, p. 1361-1376. [CrossRef]

TEIRA-ESMATGES, M.R. and FLOTATS, X. (2003). A method for livestock waste management planning in NE Spain. Waste Management, vol. 23, no. 10, p. 917-932. [CrossRef]

VANOTTI, M.B.; SZOGI, A.A.; HUNT, P.G.; MILLNER, P.D. and HUMENIK, F.J. (2007). Development of environmentally superior treatment system to replace anaerobic swine lagoons in the USA. Bioresource Technology, vol. 98, no. 17, p. 3184-3194. [CrossRef]

VIDAL, G.; SOTO, M.; FIELD, J.; MÉNDEZ-PAMPÍN, R. and LEMA, J.M. (1997). Anaerobic biodegradability and toxicity of wastewaters from chlorine and total chlorine-free bleaching of eucalyptus kraft pulps. Water Research, vol. 31, no. 10, p. 2487-2494. [CrossRef]

VIDAL, G. and DIEZ, M.C. (2003). Influence of feedstock and bleaching technologies on methanogenic toxicity of kraft mill wastewater. Water Science and Technology, vol. 48, no. 6, p. 149-155.

VIDAL, G. and DIEZ, M.C. (2005). Methanogenic toxicity and continuous anaerobic treatment of wood processing effluents. Journal of Environmental Management, vol. 74, no. 4, p. 317-325. [CrossRef]