Waste Management 34 (2014) 2278–2284
Contents lists available at ScienceDirect
Waste Management journal homepage: www.elsevier.com/locate/wasman
Anaerobic co-digestion of food waste and landfill leachate in single-phase batch reactors Liao Xiaofeng, Zhu Shuangyan, Zhong Delai, Zhu Jingping ⇑, Liao Li ⇑ School of Environmental Science and Engineering, Huazhong University of Science & Technology, Wuhan 430074, China
a r t i c l e
i n f o
Article history: Received 28 February 2014 Accepted 16 June 2014 Available online 22 July 2014 Keywords: Anaerobic digestion Food waste Landfill leachate Ammonia nitrogen Fluorescence excitation–emission matrix spectroscopy UV–vis spectroscopy
a b s t r a c t In order to investigate the effect of raw leachate on anaerobic digestion of food waste, co-digestions of food waste with raw leachate were carried out. A series of single-phase batch mesophilic (35 ± 1 °C) anaerobic digestions were performed at a food waste concentration of 41.8 g VS/L. The results showed that inhibition of biogas production by volatile fatty acids (VFA) occurred without raw leachate addition. A certain amount of raw leachate in the reactors effectively relieved acidic inhibition caused by VFA accumulation, and the system maintained stable with methane yield of 369–466 mL/g VS. Total ammonia nitrogen introduced into the digestion systems with initial 2000–3000 mgNH4–N/L not only replenished nitrogen for bacterial growth, but also formed a buffer system with VFA to maintain a delicate biochemical balance between the acidogenic and methanogenic microorganisms. UV spectroscopy and fluorescence excitation–emission matrix spectroscopy data showed that food waste was completely degraded. We concluded that using raw leachate for supplement water addition and pH modifier on anaerobic digestion of food waste was effective. An appropriate fraction of leachate could stimulate methanogenic activity and enhance biogas production. Ó 2014 Elsevier Ltd. All rights reserved.
1. Introduction As one of the most effective systems for biological treatment of organic waste, anaerobic digestion technology has a number of benefits, such as solid reduction and biogas production, which makes it an attractive technology (Klavon et al., 2013). However, anaerobic digestion sometimes tends to be inefficient when some organic waste such as manure or food waste is used as sole substrate (Yamashiro et al., 2013). Relatively low biodegradability and biogas yield of dairy manure make it usually unfavorable for energy production. Compared to dairy manure, food waste has higher biogas potentials. Due to its extremely high biodegradability, accumulated volatile fatty acids (VFA) often makes methane production suppressed. As an alternative, co-digestion which simultaneously uses more than one organic waste stream as substrates, seems promising for improving digestion efficiencies. With an improved balance of nutrients, and the synergy effect between organic substrates, codigestion can improve process performance (Viotti et al., 2004).
⇑ Corresponding authors. Address: School of Environmental Science and Engineering, Huazhong University of Science & Technology, 1037 Luoyu Road, Wuhan 430074, China. Tel./fax: +86 27 87792401. E-mail addresses: [email protected] (J. Zhu), [email protected] (L. Liao). http://dx.doi.org/10.1016/j.wasman.2014.06.014 0956-053X/Ó 2014 Elsevier Ltd. All rights reserved.
According to the study of Bouallagui et al. (2009), during the digestion of fruit and vegetable waste, additions of abattoir wastewater or activated sludge could enhance biogas yields by 51.5% and 43.8%, respectively. The added wastewater or sludge lowered carbon to nitrogen ratio and enhanced biogas yields. Co-digestion of dairy manure and more degradable wastes is effective for improving the economics of dairy digesters by increasing the biogas production rate (El-Mashad and Zhang, 2010). Anaerobic co-digestion processes require the proper conditions with regard to substrates. In the batch experiments of Li et al. (2009), a mixing ratio of 3:1 was optimal for co-digestion of cattle manure and kitchen waste, with methane yield of 233 ml/g VS. For food waste digestion, it is important to maintain suitable pH during the anaerobic process and maintain the balance between VFA and methane production. Co-digestion of food waste with other organic waste would provide an improved balance of nutrients and the development of synergistic microbial consortia (Sosnowski and Ledakowicz, 2003; Hartmann and Ahring, 2005). Manure, sludge are often used as the co-digestion substrate of food waste (Ye et al., 2013; Kim et al., 2013; Yamashiro et al., 2013). Studies showed that food waste co-digested with mixture of manure resulted in increased biogas production rate and biogas yield compared to digesting manure or food waste alone (Li et al., 2010). Wan reported that using Chinese silver grass (CSG) as a co-substrate in food waste anaerobic digestion system is a
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X. Liao et al. / Waste Management 34 (2014) 2278–2284
potential simple method to convert CSG into renewable energy and to simultaneously improve food waste treatment (Wan et al., 2013). Shahriari et al. (2012) studied the effect of using untreated leachate for supplement water addition and liquid recirculation on anaerobic digestion of food waste. An appropriate fraction of recycled leachate and fresh water could stimulate methanogenic activity and enhance biogas production. Landfill leachate has often been used as the co-digestion substrate for sewage sludge, septage, or domestic wastewater treatment (Kawai et al., 2012; Montusiewicz and Lebiocka, 2011). However, when amount of leachate exceeds some extent, high-strength ammonia nitrogen might inhibit methanogenic activities. In this paper, anaerobic digestion of food waste was studied and untreated raw leachate was used as co-digestion substrate. The objective of this work was to study the impact of addition of different amounts of raw leachate on anaerobic digestion of food waste and provide insight into the impact of co-digestion on biogas production and stabilization of food waste treatment process. Because it was difficult to distinguish the decomposition rate or methane productivity of each substrate separately (Mata-Alvarez et al., 2000; Hartmann and Ahring, 2005), most studies about co-digestion of landfill leachate were seldom concerned about the change of substrates before and after co-digestion. Dissolved organic matter (DOM) is a heterogeneous mixture of humic substances, hydrophilic acids, proteins, lipids, carbohydrates, carboxylic acids, amino acids, and hydrocarbons (Leenheer and Croue, 2003). Multiplicate analytical methods are used to distinguish the characteristics of DOM in wastewater (Huo et al., 2008; Seo et al., 2007). For example, ultraviolet–visible (UV–vis) spectrometry, fourier transform infrared (FT-IR) spectroscopy, and fluorescence excitation–emission matrix (EEM) spectroscopy have been employed to determine the molecular size magnitude, structures, and specific pollutants quantificationally and qualitatively (Kim et al., 2006; Hudson et al., 2007). However, few works have been reported about the characteristics of DOM in slurry samples after anaerobic digestion using spectral spectroscopy. Fluorescence EEM spectroscopy can provide considerable detailed information about fluorescence properties of DOM that may reveal important information about its composition (Burdige et al., 2004; Baker, 2005; Marhuenda-Egea et al., 2007). In this study, UV–Vis spectroscopy analysis and fluorescence excitation– emission matrix (EEM) analysis were used to analyze characteristics of slurry samples after the reaction in order to obtain more information about the decomposition of leachate and co-digested substrate. 2. Materials and method 2.1. Substrates The raw leachate was sampled from the Chenjiachong Municipal Solid Waste Landfill, Wuhan, China, which has been in operation since 2007. Food waste was collected from canteens of Huazhong University of Science & Technology, Wuhan, China. The chief constituents of the food waste were classified as cooked rice, vegetables and meat and was homogenated using a waste disposal unit (France) before use. Characteristics of the middle-aged raw leachate and food waste were summarized in Table 1. 2.2. Anaerobic digestion Co-digestions of food waste and raw leachate experiments were conducted with HRT (35 days) in single-stage batch reactors with a working volume of 1500 mL. The reactors were in a water bath for maintaining a mean mesophilic temperature of 35 ± 1 °C. A 6 L gas
Table 1 Characteristics of raw leachate and food waste used in the co-digestion experiments. Parameters
Raw leachate
Parameters
Food waste
pH NH4–N (mg/L) Salinity (g/L) DOC (mg/L) COD (mg/L) BOD5 (mg/L) TOC
8.54 3625 8.18 1298 2500 345
Protein (%) Fat (%) Carbohydrate (%) Salinity (%) Moisture (%) C/N VS/TS TOC (g/kg TS)
3.37 4.43 21.37 0.84 69.52 25.39 95.7 450
container was attached to each reactor for biogas collection. The gas volume was calculated daily based on the downward displacement of water. The initial OLR for the reactors 1–7 was 41.8 g VS/L and different amounts of raw leachate were added into these reactors (shown in Table 2). A blank digesters (the reactor 8) that contained inoculum and raw leachate only were also incubated at the same temperature to test for the biogas produced from the inoculum and raw leachate. The inoculum was anaerobic granular sludge taken from an anaerobic fermentation bioreactor in a food plant in Wuhan, China. The seed sludge consisted of well-settled black granules, with about 90% showing the size >1.5 mm in diameter. The volatile suspended solid (VSS) content was 79,680 mg/L, corresponding to about 71% of the total suspended solid (SS). 2.3. Analytical methods CH4 and CO2 concentrations were measured by a Biogas 5000 analyzer (Geotech, England). Samples were collected every two days and were analyzed. pH value was measured with a PB-10 pH-meter (Sartorius, Germany). NH4–N concentration was measured with a HT 93733 Ammonia detector (Hanna, Italy). Total VFA concentration was determined colorimetrically (Thermo UV–Vis spectrophotometer). For determination of total solids, the samples were dried at 105 °C for 24 h, and total solid contents were calculated from the differences between weights before and after drying. The dried matters were heated at 550 °C for 4 h, and organic matter contents were calculated from the losses on ignition. For determination of total nitrogen, a 5 g sample was weighed into a Kjeldahl digestion apparatus; 20 mL of concentrated H2SO4 and 0.2 g CuSO4 (Kjeldahl catalyst) were added. The Kjeldahl digestion apparatus was heated for 2 h until all nitrogen was conversed to ammonium sulphate. After cooling, the digested sample was transferred into 100 mL volumetric flasks and diluted with distilled water. After distillation of the digested solution, the distillate was titrated against 0.1 mol/L HCl. For determination of dissolved organic carbon (DOC), after filtration through 0.45 lm membrane to remove suspended materials, samples was determined by a multi N/C 2100 TOC analyzer (Jena, Germany). 2.4. UV–Vis spectral analysis Specific ultraviolet light absorbance (SUVA254) calculated as A254/DOC is a measure of the contribution of aromatic structures to DOC and is usually used to define the aromaticity of dissolved organic matter (DOM) (Weishaar et al., 2003). The ratio of UV absorbance at 253 nm to that at 203 nm (A253/ A203) reflects the degree of substitution of the aromatic ring and the kind of the substitution (Korshin et al., 1997). A253/A203 is low for unsubstituted aromatic ring structures. When the aromatic rings are highly substituted, the value is high. The changes in
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X. Liao et al. / Waste Management 34 (2014) 2278–2284
Table 2 The amount of raw leachate and fresh water added in the reactors. Reactors
1
2
3
4
5
6
7
8
Organic load of food waste (g TS/L) Leachate (ml) Fresh water (ml) NH4–N (mg/L)
40 0 1031 0
40 142 889 500
40 284 747 1000
40 426 605 1500
40 568 463 2000
40 710 321 2500
40 852 179 3000
0 1200 0 0
A253/A203 suggest that aromatic rings substituted with various factional groups are structurally altered by treatment. After filtration through 0.45 lm membrane to remove suspended materials, slurry samples were diluted and UV–Vis spectra were measured using a Thermo UV–Vis spectrophotometer. SuperQ water was used as a blank. SUVA254 and A253/A203 values for slurry samples were calculated. 2.5. Fluorescence EEM spectral analysis Fluorescence EEM spectroscopy can provide considerable detailed information about fluorescence properties of DOM that may reveal important information about its composition. With this technique, a three-dimensional picture is generated of fluorescence intensity as a function of excitation and emission wavelength. Fluorescence regional integration (FRI), a quantitative technique that integrates volumes beneath different excitation– emission regions in EEM spectra, is used to quantitatively analyze EEM spectra (He et al., 2011). According to the reports (Chen et al., 2003), consistent excitation and emission wavelength boundaries for each EEM horizontal and vertical lines were drawn to divide the EEM into five regions. Peaks at shorter excitation wavelengths (200–250 nm) and shorter emission wavelengths (280–380 nm) were related to simple aromatic proteins such as tyrosine (Region I) and tryptophan (Region II) (Ahmad and Reynolds, 1999). Peaks at excitation wavelengths (250–300 nm) and shorter emission wavelength (280–380 nm) were related to soluble microbial byproductlike material (Region IV) (Ismaili et al., 1998; Reynolds and Ahmad, 1997). Peaks at excitation wavelengths (200–250 nm) and emission wavelengths (380–550 nm) were related to fulvic acid-like organics (Region III) (Mounier et al., 1999). Peaks at excitation wavelengths (250–450 nm) and emission wavelengths (380– 550 nm) were related to humic acid-like organics (Region V). FRI technique was used to integrate the volume beneath EEM within each region (Ui). The cumulative volume beneath the EEM (UT) is P calculated as UT = Ui. All UT and Ui values were normalized to the DOC concentration and the projected excitation–emission area within that region to get Ui,n and UT,n. The EEMs of slurry samples were measured by a F-4600 fluorescence spectrometer (Hitachii, Japan). A xenon excitation source was used in the spectrometer, and the excitation and emission slits were set to a 5 nm band-pass. Each EEM spectrum was generated by scanning excitation wavelengths from 200 to 450 nm at 5 nm steps, and detecting the emission fluorescence between 250 and 550 nm at 2 nm steps. The scan speed was set at 1200 nm/min. The spectrum of Super-Q water was recorded as blank. All samples were adjusted to pH 7 with HCl prior to measurement. UT,n values of the samples were obtained. The percent fluorescence response in a specific region (Pi,n) was calculated as Pi,n = Ui,n/UT,n 100%. 3. Results and discussion Table 3 showed characteristics of slurry samples after digestion. The samples from the reactors 1–4 had high BOD5/COD values, while those from the reactors 5–7 had low biodegradability.
3.1. pH and VFA concentration change during co-digestion As shown in Fig. 1, initial pH values in all the reactors were >7. With the hydrolysis of food waste, total VFA concentrations increased greatly, which induced pH decline rapidly. In the reactor 1, the accumulation of VFAs led to irreversible acidification which inhibited methanogenic microorganisms’ activities. With different amount of raw leachate added in the reactors 2–7, changes of pH and VFA concentrations were different during the whole process. Acid accumulation with varying degrees was observed in the reactors 2–4. On the other hand, the systems in the reactors 5–7 showed stronger regulation abilities for VFA. After 5 days, pH values in the reactors 5–7 increased gradually to >7.5 and VFA concentrations were maintained at a relatively low level, which showed that the buffer effect of ammonia nitrogen introduced by raw leachate mitigated acid accumulation effectively and pH values remained in a range which was suitable for methanogenic microorganisms’ activities. 3.2. Gas production In Fig. 2, little gas (1.03 L) was produced in the reactor 8, showing refractory of raw leachate to biodegradation. The gas yield of the reactors 1–4 were only 14.95%, 9.56%, 14.85% and 33.33% of that of the reactor 5, and gas production stopped at day 23. In conventional wet anaerobic digestion of food waste without buffers, OLR usually remained insufficient (
Contents lists available at ScienceDirect
Waste Management journal homepage: www.elsevier.com/locate/wasman
Anaerobic co-digestion of food waste and landfill leachate in single-phase batch reactors Liao Xiaofeng, Zhu Shuangyan, Zhong Delai, Zhu Jingping ⇑, Liao Li ⇑ School of Environmental Science and Engineering, Huazhong University of Science & Technology, Wuhan 430074, China
a r t i c l e
i n f o
Article history: Received 28 February 2014 Accepted 16 June 2014 Available online 22 July 2014 Keywords: Anaerobic digestion Food waste Landfill leachate Ammonia nitrogen Fluorescence excitation–emission matrix spectroscopy UV–vis spectroscopy
a b s t r a c t In order to investigate the effect of raw leachate on anaerobic digestion of food waste, co-digestions of food waste with raw leachate were carried out. A series of single-phase batch mesophilic (35 ± 1 °C) anaerobic digestions were performed at a food waste concentration of 41.8 g VS/L. The results showed that inhibition of biogas production by volatile fatty acids (VFA) occurred without raw leachate addition. A certain amount of raw leachate in the reactors effectively relieved acidic inhibition caused by VFA accumulation, and the system maintained stable with methane yield of 369–466 mL/g VS. Total ammonia nitrogen introduced into the digestion systems with initial 2000–3000 mgNH4–N/L not only replenished nitrogen for bacterial growth, but also formed a buffer system with VFA to maintain a delicate biochemical balance between the acidogenic and methanogenic microorganisms. UV spectroscopy and fluorescence excitation–emission matrix spectroscopy data showed that food waste was completely degraded. We concluded that using raw leachate for supplement water addition and pH modifier on anaerobic digestion of food waste was effective. An appropriate fraction of leachate could stimulate methanogenic activity and enhance biogas production. Ó 2014 Elsevier Ltd. All rights reserved.
1. Introduction As one of the most effective systems for biological treatment of organic waste, anaerobic digestion technology has a number of benefits, such as solid reduction and biogas production, which makes it an attractive technology (Klavon et al., 2013). However, anaerobic digestion sometimes tends to be inefficient when some organic waste such as manure or food waste is used as sole substrate (Yamashiro et al., 2013). Relatively low biodegradability and biogas yield of dairy manure make it usually unfavorable for energy production. Compared to dairy manure, food waste has higher biogas potentials. Due to its extremely high biodegradability, accumulated volatile fatty acids (VFA) often makes methane production suppressed. As an alternative, co-digestion which simultaneously uses more than one organic waste stream as substrates, seems promising for improving digestion efficiencies. With an improved balance of nutrients, and the synergy effect between organic substrates, codigestion can improve process performance (Viotti et al., 2004).
⇑ Corresponding authors. Address: School of Environmental Science and Engineering, Huazhong University of Science & Technology, 1037 Luoyu Road, Wuhan 430074, China. Tel./fax: +86 27 87792401. E-mail addresses: [email protected] (J. Zhu), [email protected] (L. Liao). http://dx.doi.org/10.1016/j.wasman.2014.06.014 0956-053X/Ó 2014 Elsevier Ltd. All rights reserved.
According to the study of Bouallagui et al. (2009), during the digestion of fruit and vegetable waste, additions of abattoir wastewater or activated sludge could enhance biogas yields by 51.5% and 43.8%, respectively. The added wastewater or sludge lowered carbon to nitrogen ratio and enhanced biogas yields. Co-digestion of dairy manure and more degradable wastes is effective for improving the economics of dairy digesters by increasing the biogas production rate (El-Mashad and Zhang, 2010). Anaerobic co-digestion processes require the proper conditions with regard to substrates. In the batch experiments of Li et al. (2009), a mixing ratio of 3:1 was optimal for co-digestion of cattle manure and kitchen waste, with methane yield of 233 ml/g VS. For food waste digestion, it is important to maintain suitable pH during the anaerobic process and maintain the balance between VFA and methane production. Co-digestion of food waste with other organic waste would provide an improved balance of nutrients and the development of synergistic microbial consortia (Sosnowski and Ledakowicz, 2003; Hartmann and Ahring, 2005). Manure, sludge are often used as the co-digestion substrate of food waste (Ye et al., 2013; Kim et al., 2013; Yamashiro et al., 2013). Studies showed that food waste co-digested with mixture of manure resulted in increased biogas production rate and biogas yield compared to digesting manure or food waste alone (Li et al., 2010). Wan reported that using Chinese silver grass (CSG) as a co-substrate in food waste anaerobic digestion system is a
2279
X. Liao et al. / Waste Management 34 (2014) 2278–2284
potential simple method to convert CSG into renewable energy and to simultaneously improve food waste treatment (Wan et al., 2013). Shahriari et al. (2012) studied the effect of using untreated leachate for supplement water addition and liquid recirculation on anaerobic digestion of food waste. An appropriate fraction of recycled leachate and fresh water could stimulate methanogenic activity and enhance biogas production. Landfill leachate has often been used as the co-digestion substrate for sewage sludge, septage, or domestic wastewater treatment (Kawai et al., 2012; Montusiewicz and Lebiocka, 2011). However, when amount of leachate exceeds some extent, high-strength ammonia nitrogen might inhibit methanogenic activities. In this paper, anaerobic digestion of food waste was studied and untreated raw leachate was used as co-digestion substrate. The objective of this work was to study the impact of addition of different amounts of raw leachate on anaerobic digestion of food waste and provide insight into the impact of co-digestion on biogas production and stabilization of food waste treatment process. Because it was difficult to distinguish the decomposition rate or methane productivity of each substrate separately (Mata-Alvarez et al., 2000; Hartmann and Ahring, 2005), most studies about co-digestion of landfill leachate were seldom concerned about the change of substrates before and after co-digestion. Dissolved organic matter (DOM) is a heterogeneous mixture of humic substances, hydrophilic acids, proteins, lipids, carbohydrates, carboxylic acids, amino acids, and hydrocarbons (Leenheer and Croue, 2003). Multiplicate analytical methods are used to distinguish the characteristics of DOM in wastewater (Huo et al., 2008; Seo et al., 2007). For example, ultraviolet–visible (UV–vis) spectrometry, fourier transform infrared (FT-IR) spectroscopy, and fluorescence excitation–emission matrix (EEM) spectroscopy have been employed to determine the molecular size magnitude, structures, and specific pollutants quantificationally and qualitatively (Kim et al., 2006; Hudson et al., 2007). However, few works have been reported about the characteristics of DOM in slurry samples after anaerobic digestion using spectral spectroscopy. Fluorescence EEM spectroscopy can provide considerable detailed information about fluorescence properties of DOM that may reveal important information about its composition (Burdige et al., 2004; Baker, 2005; Marhuenda-Egea et al., 2007). In this study, UV–Vis spectroscopy analysis and fluorescence excitation– emission matrix (EEM) analysis were used to analyze characteristics of slurry samples after the reaction in order to obtain more information about the decomposition of leachate and co-digested substrate. 2. Materials and method 2.1. Substrates The raw leachate was sampled from the Chenjiachong Municipal Solid Waste Landfill, Wuhan, China, which has been in operation since 2007. Food waste was collected from canteens of Huazhong University of Science & Technology, Wuhan, China. The chief constituents of the food waste were classified as cooked rice, vegetables and meat and was homogenated using a waste disposal unit (France) before use. Characteristics of the middle-aged raw leachate and food waste were summarized in Table 1. 2.2. Anaerobic digestion Co-digestions of food waste and raw leachate experiments were conducted with HRT (35 days) in single-stage batch reactors with a working volume of 1500 mL. The reactors were in a water bath for maintaining a mean mesophilic temperature of 35 ± 1 °C. A 6 L gas
Table 1 Characteristics of raw leachate and food waste used in the co-digestion experiments. Parameters
Raw leachate
Parameters
Food waste
pH NH4–N (mg/L) Salinity (g/L) DOC (mg/L) COD (mg/L) BOD5 (mg/L) TOC
8.54 3625 8.18 1298 2500 345
Protein (%) Fat (%) Carbohydrate (%) Salinity (%) Moisture (%) C/N VS/TS TOC (g/kg TS)
3.37 4.43 21.37 0.84 69.52 25.39 95.7 450
container was attached to each reactor for biogas collection. The gas volume was calculated daily based on the downward displacement of water. The initial OLR for the reactors 1–7 was 41.8 g VS/L and different amounts of raw leachate were added into these reactors (shown in Table 2). A blank digesters (the reactor 8) that contained inoculum and raw leachate only were also incubated at the same temperature to test for the biogas produced from the inoculum and raw leachate. The inoculum was anaerobic granular sludge taken from an anaerobic fermentation bioreactor in a food plant in Wuhan, China. The seed sludge consisted of well-settled black granules, with about 90% showing the size >1.5 mm in diameter. The volatile suspended solid (VSS) content was 79,680 mg/L, corresponding to about 71% of the total suspended solid (SS). 2.3. Analytical methods CH4 and CO2 concentrations were measured by a Biogas 5000 analyzer (Geotech, England). Samples were collected every two days and were analyzed. pH value was measured with a PB-10 pH-meter (Sartorius, Germany). NH4–N concentration was measured with a HT 93733 Ammonia detector (Hanna, Italy). Total VFA concentration was determined colorimetrically (Thermo UV–Vis spectrophotometer). For determination of total solids, the samples were dried at 105 °C for 24 h, and total solid contents were calculated from the differences between weights before and after drying. The dried matters were heated at 550 °C for 4 h, and organic matter contents were calculated from the losses on ignition. For determination of total nitrogen, a 5 g sample was weighed into a Kjeldahl digestion apparatus; 20 mL of concentrated H2SO4 and 0.2 g CuSO4 (Kjeldahl catalyst) were added. The Kjeldahl digestion apparatus was heated for 2 h until all nitrogen was conversed to ammonium sulphate. After cooling, the digested sample was transferred into 100 mL volumetric flasks and diluted with distilled water. After distillation of the digested solution, the distillate was titrated against 0.1 mol/L HCl. For determination of dissolved organic carbon (DOC), after filtration through 0.45 lm membrane to remove suspended materials, samples was determined by a multi N/C 2100 TOC analyzer (Jena, Germany). 2.4. UV–Vis spectral analysis Specific ultraviolet light absorbance (SUVA254) calculated as A254/DOC is a measure of the contribution of aromatic structures to DOC and is usually used to define the aromaticity of dissolved organic matter (DOM) (Weishaar et al., 2003). The ratio of UV absorbance at 253 nm to that at 203 nm (A253/ A203) reflects the degree of substitution of the aromatic ring and the kind of the substitution (Korshin et al., 1997). A253/A203 is low for unsubstituted aromatic ring structures. When the aromatic rings are highly substituted, the value is high. The changes in
2280
X. Liao et al. / Waste Management 34 (2014) 2278–2284
Table 2 The amount of raw leachate and fresh water added in the reactors. Reactors
1
2
3
4
5
6
7
8
Organic load of food waste (g TS/L) Leachate (ml) Fresh water (ml) NH4–N (mg/L)
40 0 1031 0
40 142 889 500
40 284 747 1000
40 426 605 1500
40 568 463 2000
40 710 321 2500
40 852 179 3000
0 1200 0 0
A253/A203 suggest that aromatic rings substituted with various factional groups are structurally altered by treatment. After filtration through 0.45 lm membrane to remove suspended materials, slurry samples were diluted and UV–Vis spectra were measured using a Thermo UV–Vis spectrophotometer. SuperQ water was used as a blank. SUVA254 and A253/A203 values for slurry samples were calculated. 2.5. Fluorescence EEM spectral analysis Fluorescence EEM spectroscopy can provide considerable detailed information about fluorescence properties of DOM that may reveal important information about its composition. With this technique, a three-dimensional picture is generated of fluorescence intensity as a function of excitation and emission wavelength. Fluorescence regional integration (FRI), a quantitative technique that integrates volumes beneath different excitation– emission regions in EEM spectra, is used to quantitatively analyze EEM spectra (He et al., 2011). According to the reports (Chen et al., 2003), consistent excitation and emission wavelength boundaries for each EEM horizontal and vertical lines were drawn to divide the EEM into five regions. Peaks at shorter excitation wavelengths (200–250 nm) and shorter emission wavelengths (280–380 nm) were related to simple aromatic proteins such as tyrosine (Region I) and tryptophan (Region II) (Ahmad and Reynolds, 1999). Peaks at excitation wavelengths (250–300 nm) and shorter emission wavelength (280–380 nm) were related to soluble microbial byproductlike material (Region IV) (Ismaili et al., 1998; Reynolds and Ahmad, 1997). Peaks at excitation wavelengths (200–250 nm) and emission wavelengths (380–550 nm) were related to fulvic acid-like organics (Region III) (Mounier et al., 1999). Peaks at excitation wavelengths (250–450 nm) and emission wavelengths (380– 550 nm) were related to humic acid-like organics (Region V). FRI technique was used to integrate the volume beneath EEM within each region (Ui). The cumulative volume beneath the EEM (UT) is P calculated as UT = Ui. All UT and Ui values were normalized to the DOC concentration and the projected excitation–emission area within that region to get Ui,n and UT,n. The EEMs of slurry samples were measured by a F-4600 fluorescence spectrometer (Hitachii, Japan). A xenon excitation source was used in the spectrometer, and the excitation and emission slits were set to a 5 nm band-pass. Each EEM spectrum was generated by scanning excitation wavelengths from 200 to 450 nm at 5 nm steps, and detecting the emission fluorescence between 250 and 550 nm at 2 nm steps. The scan speed was set at 1200 nm/min. The spectrum of Super-Q water was recorded as blank. All samples were adjusted to pH 7 with HCl prior to measurement. UT,n values of the samples were obtained. The percent fluorescence response in a specific region (Pi,n) was calculated as Pi,n = Ui,n/UT,n 100%. 3. Results and discussion Table 3 showed characteristics of slurry samples after digestion. The samples from the reactors 1–4 had high BOD5/COD values, while those from the reactors 5–7 had low biodegradability.
3.1. pH and VFA concentration change during co-digestion As shown in Fig. 1, initial pH values in all the reactors were >7. With the hydrolysis of food waste, total VFA concentrations increased greatly, which induced pH decline rapidly. In the reactor 1, the accumulation of VFAs led to irreversible acidification which inhibited methanogenic microorganisms’ activities. With different amount of raw leachate added in the reactors 2–7, changes of pH and VFA concentrations were different during the whole process. Acid accumulation with varying degrees was observed in the reactors 2–4. On the other hand, the systems in the reactors 5–7 showed stronger regulation abilities for VFA. After 5 days, pH values in the reactors 5–7 increased gradually to >7.5 and VFA concentrations were maintained at a relatively low level, which showed that the buffer effect of ammonia nitrogen introduced by raw leachate mitigated acid accumulation effectively and pH values remained in a range which was suitable for methanogenic microorganisms’ activities. 3.2. Gas production In Fig. 2, little gas (1.03 L) was produced in the reactor 8, showing refractory of raw leachate to biodegradation. The gas yield of the reactors 1–4 were only 14.95%, 9.56%, 14.85% and 33.33% of that of the reactor 5, and gas production stopped at day 23. In conventional wet anaerobic digestion of food waste without buffers, OLR usually remained insufficient (
