Influence of temperature on carbon and nitrogen dynamics during in situ aeration of aged waste in simulated landfill bioreactors

Bioresource Technology 192 (2015) 149–156

Contents lists available at ScienceDirect

Bioresource Technology journal homepage: www.elsevier.com/locate/biortech

Influence of temperature on carbon and nitrogen dynamics during in situ aeration of aged waste in simulated landfill bioreactors Huanhuan Tong a,b, Ke Yin a, Apostolos Giannis a, Liya Ge a, Jing-Yuan Wang a,b,⇑ a Residues and Resource Reclamation Centre, Nanyang Environment and Water Research Institute, Nanyang Technological University, 1 Cleantech Loop, CleanTech One, Singapore 637141, Singapore b School of Civil and Environmental Engineering, Nanyang Technological University, 50 Nanyang Avenue, Singapore 639798, Singapore

h i g h l i g h t s  Temperature effects on biochemical processes during in situ aeration were tested.  Fastest carbon mineralization was realized under thermophilic condition.  Organic nitrogen was the dominant nitrogen form in leachate after aeration.  Higher variation of humic nitrogen occurred under thermophilic condition.  Humic substances were enriched in the waste under thermophilic condition.

a r t i c l e

i n f o

Article history: Received 18 March 2015 Received in revised form 12 May 2015 Accepted 14 May 2015 Available online 19 May 2015 Keywords: In situ aeration Temperature effects Carbon conversion Leachate nitrogen

a b s t r a c t The effect of temperature on carbon and nitrogen compounds during in situ aeration of aged waste was investigated in lab-scale simulated landfill bioreactors at 35, 45 and 55 °C, respectively. The bioreactor operated at 55 °C presented the highest carbon mineralization rate in the initial stage, suggesting accelerated biodegradation rates under thermophilic conditions. The nitrogen speciation study indicated that organic nitrogen was the dominant species of total N in aerobic bioreactors due to ammonia removal. Leachate organic nitrogen was further fractionated to elucidate the fate of individual constituent. Detailed investigation revealed the higher bioconversion rates of N-humic and N-fulvic compounds compared to hydrophilic compounds in thermophilic conditions. At the end, waste material in 55 °C bioreactor was richer in highly matured humic substances (HS) verifying the high bioconversion rates. Ó 2015 Elsevier Ltd. All rights reserved.

1. Introduction In the later life of a traditional anaerobic sanitary landfill or dumping ground, there is a long burdensome period during which energy recovery in terms of biogas is unfeasible while the risk of pollutants release is a concern. The application of in situ aeration could mitigate the adverse effects in the old landfills and achieve earlier restoration (Ritzkowski et al., 2006). Several successful in situ aeration projects have been reported worldwide. In Germany, in situ aeration has been used as a landfill remediation technique to seek regulatory relief of closure and postclosure liability (Ritzkowski et al., 2006). In Italy and Netherlands, aeration

⇑ Corresponding author at: Residues and Resource Reclamation Centre, Nanyang Environment and Water Research Institute, Nanyang Technological University, 1 Cleantech Loop, CleanTech One, Singapore 637141, Singapore. Tel.: +65 6790 4100. E-mail address: [email protected] (J.-Y. Wang). http://dx.doi.org/10.1016/j.biortech.2015.05.049 0960-8524/Ó 2015 Elsevier Ltd. All rights reserved.

has been practiced as a pretreatment method before landfill mining (Jacobs et al., 2003; Raga and Cossu, 2014). Accelerated exothermic biooxidation of organic matter by indigenous microbes during in situ aeration could increase the waste temperature up to 60 °C over a long period of time (Öncü et al., 2012). In certain extreme cases, the temperature could exceed 70 °C inside the aerobic landfill (Yazdani et al., 2010). Along with time, the temperature changes dramatically due to the gradual attenuation of organic matter (Ritzkowski and Stegmann, 2013). The spatial distribution of temperature in a landfill could highly vary as well (Hanson et al., 2010). For instance, the maximum temperature is noticed at central location of the waste mass due to thermal insulation. Near surface, waste temperature resembles the ambient temperature with seasonal fluctuation. Steady and elevated values above mean annual ambient temperature are encountered at the base of the landfill. Most biological and physicochemical processes in the landfill are highly influenced by temperature. Primarily, the high

150

H. Tong et al. / Bioresource Technology 192 (2015) 149–156

temperature accelerates water evaporation, which could cause moisture scarcity for microbial activity. Secondary, temperature in the range of 40–50 °C could delay the acclimation of microorganisms in aerobic conditions (Raga and Cossu, 2013), as it falls out of the optimum growth temperature range for themophiles and mesophiles. In addition, thermophilic temperature in the range of 50–60 °C potentially inhibits nitrification (Raga and Cossu, 2013), likely due to the thermal death of Nitrosomonas cultures (Berge et al., 2005). Some researchers have also pointed out a poor system stability and reliability in the thermophilic temperature range, mainly associated with the less diverse microbial community and transient accumulation of potentially inhibitory intermediates (Labatut et al., 2014; Levén et al., 2007). However, several important advantages have been recognized in thermophilic conditions, such as higher growth rates of microorganisms, larger hydrolytic activities of extracellular enzymes, faster biochemical conversion rate, smaller system footprint and essentially pathogen free end product (Fernández-Rodríguez et al., 2013; Kim et al., 2012; Ugwuanyi et al., 2005). To our knowledge, there are many uncertainties in response to in situ aeration of landfill at thermophilic temperatures, and more specifically at temperatures higher than 45 °C. Despite the fact that nitrogen contributes only around 1% of municipal solid waste (MSW), nitrogen compounds can cause high environmental risk in anaerobic landfills (Tchobanoglous et al., 1993). Previous studies demonstrated that aeration could remove leachate ammonium effectively (Prantl et al., 2006; Raga and Cossu, 2013; Ritzkowski et al., 2006). However, little is known about the fate of organic nitrogen in the leachate. Dissolved organic nitrogen is related to eutrophication and hypoxia or acts as precursor for the carcinogenic disinfection by-product e.g. N-nitrosodimethylamine (NDMA) (Mitch and Sedlak, 2004). The purpose of this study was to investigate the temperature-mediated biochemical processes during in situ aeration of aged waste. Herein the effects of different temperatures (35, 45, and 55 °C) on carbon degradation rate, leachate quality and waste stability were investigated in lab-scale simulated landfill bioreactors. Particularly, the distribution of organic nitrogen in different fractions, i.e. humic acid (HA), fulvic acid (FA) and hydrophilic compound (HyI), was addressed in landfill leachate. 2. Methods 2.1. Synthetic aged waste Synthetic aged waste was prepared in the laboratory simulating the waste composition of closed landfill. Initially, 55 kg of fresh waste was loaded into landfill simulation reactor, with the composition of food waste, yard waste, paper, plastics and textile at a ratio of 60/15/10/10/5 (w/w). Pre-aeration, leachate recirculation and sludge inoculation were employed to rapidly initiate the methanogenesis phase. After 150 days of anaerobic digestion, the methane generation rate was below 0.1% of the peak value for more than 70 days, indicating the waste mass was stable and aged. The details of aged waste preparation were reported previously (Tong et al., 2015). 2.2. Landfill bioreactor operation The aged waste prepared in the above-mentioned landfill simulation reactor was excavated and loaded into four identical acrylic bioreactors with dimensions 500 mm  200 mm (L  D). Fig. 1 shows the schematic diagram of the aeration bioreactors used in this study. All the bioreactors contained approximately 6.5 kg of aged waste and were kept in a control room with a constant

9

8 7 6

10 11 5

12

4 2 3

13 5 1

Fig. 1. Schematic diagram of aeration bioreactor (1) measuring cylinder for leachate; (2) bioreactor column covered by heating belt; (3) temperature controller with thermoprobe; (4) air injection pipe; (5) gravel layer; (6) water distributor; (7) rubber septum for gas sampling during carbon mineralization rate measurement; (8) syringe for gas sampling; (9) gas flow meter; (10) air pump; (11) open port for outlet gas; (12) gas counter and (13) gas analyzer.

temperature 35 °C. One of the reactors (NR35) was kept under anaerobic conditions throughout the experiment as control. The other three bioreactors were operated in continuous aeration mode through a perforated pipe placed in the middle of the bioreactor, with air flow rate approximately 0.016 L/min kg dry material (DM). Two of the bioreactors (AR45 and AR55) were heated to 45 and 55 °C, respectively, by heating belts using a ramp of 2 °C/day, while the last bioreactor (AR35) was kept at constant room temperature 35 °C. The duration of the experiments lasted for 396 days. 2.3. Leachate analysis 200 ml of deionized (DI) water was added from the top of the bioreactor three times per week. Water distributors were installed to ensure even water distribution. The produced leachate was temporarily withheld in the bottom marble layer. After the second water addition, the leachate was discharged by gravity into measuring cylinder right before the third water addition. The leachate sample was filtered through 0.45 lm syringe filter (Cronus, UK) for chemical analysis. Chemical characteristics like ammonium–nitrogen and COD were determined using Hach test kits at spectrophotometer (Hach, USA). Analyses of nitrate and nitrite concentrations were conducted by ICS-1100 ion chromatography system (Dionex, USA). Dissolved organic carbon (DOC) and total nitrogen (TN) were measured in Multi N/C 2100 system (Analytik Jena, UK). For easy comparison, the concentration of the compound concentration in leachate was normalized to 400 mL by water addition (the volume of incoming DI water). In addition, leachate samples were selectively collected at two different periods for quantitative speciation of nitrogen-containing compounds. The sample in period I represents an average property of four consecutive days of leachate collection, i.e. days 73, 77, 82 and 87, when 200 mL leachate from each day were combined together to reach a total volume of 800 mL. Period II was at days 215, 220, 224 and 229. Each 800 mL leachate sample was analyzed in duplicates. The dissolved organic matter was fractionated into three fractions: HA, FA and HyI as shown in our previous publication (Tong et al., 2015). Briefly, HA was insoluble hydrophobic acids

H. Tong et al. / Bioresource Technology 192 (2015) 149–156

when the leachate sample was acidified to pH < 1. After removing HA, the hydrophobic acid absorbed on DAX-8 resin (Sigma– Aldrich, US) was defined as FA, while the residual dissolved organic matter not absorbed onto the DAX-8 resin was considered as HyI. The concentrations of HA and FA were presented by DOC content in the leachate sample. The N/C mass ratio in the purified HA and FA powder was measured in Vario EL cube (Elementar Analysensysteme GmbH, Germany), and was utilized to calculate the organic nitrogen distribution in HA and FA by multiplying with DOC content, respectively. The nitrogen content in HyI was determined as the difference between TN and the sum of nitrogen in other fractions (ammonium, nitrite, nitrate, and organic nitrogen in HA and FA) (Westgate and Park, 2010). 2.4. Carbon mineralization rate In all the aeration bioreactors, air was continuously pumped into the waste materials, except periods for measuring carbon mineralization rate. The volume of exhaust gas was recorded by a gas volume counter (Ritter Apparatebau GmbH, Germany), whereas gas composition was monitored by a portable gas analyzer (Sewerin, Multitech 520, Germany). In order to measure carbon mineralization rate, the air supply was turned off and all the valves were closed to ensure the bioreactor was sealed. Thereupon, the gas sample in the headspace of the bioreactors was collected by a syringe through a septum on the top cover over 2–10 h interval and gas concentration was measured using gas chromatography GC-TCD (Agilent Technologies, USA). The slope of CO2 percentage versus time determined the CO2 generation rate in terms of percentage per hour (%/h). The CO2 mass generation rate (mg/hr) and carbon mineralization rate (mg/d) was calculated by the Eqs. (1) and (2), respectively:

km ¼ ðko =100ÞðV r  V g  V w þ V w ha ÞqCO2 ð1000 mg=1 gÞ

ð1Þ

C g ¼ ð24Þkm ð12=44Þ

ð2Þ

where km is CO2 mass generation rate (mg/h); ko is CO2 generation rate (%/hr); Vr is the working volume of the bioreactor (13.312 L); Vg is the volume of the devices inside the bioreactor, including aeration pipe, water distributor, thermoprobe, and bottom glass marbles (1.288 L); Vw is waste bulk volume (9.924 L); ha is gas-filled pore space (0.192 cm3 gas/cm3 waste measured via water-filled void volume of the waste); qCO2 is the density of carbon dioxide (1.726 g/L at 35 °C, 1.671 g/L at 45 °C and 1.619 g/L at 55 °C); Cg is the carbon mineralization rate (mg/d). 2.5. Waste characteristics At the end, the bioreactors were dismantled and representative waste samples were characterized for (1) chemical parameters, such as total organic carbon (TOC); and (2) biostability parameters, such as GS21 and HS content. The TOC measurement was performed on dried and milled sample through Multi N/C 2100 system (Analytik Jena, UK). GS21 was the total volume of gas generated for 21 days under anaerobic conditions, determined by Automatic Methane Potential Test System (Bioprocess Control, Sweden). Isolation and purification procedures of HS from waste sample were carried out according to the method provided by the International Humic Substance Society (IHSS). HA precipitation was collected by adjusting pH of the NaOH-extracted solution to 1 using concentrated HCl. The DOC content of HA was measured by the difference between the DOC concentration in NaOH-extracted solution and the one in the supernatant after acidification and filtration. The FA extract passed through DAX-8 resin column, which retained FA by adsorption. The theoretical content

151

of FA was determined as the difference between the DOC content in the FA extract and the DOC content in the sample passing the DAX -8 resin column. The HS content in the end waste was denoted as DOC of HS per gram of dry waste sample. The purified FA and HA powder were subject to solid-state 13C nuclear magnetic resonance (13C NMR) analysis on a Bruker AVANCE III 400WB spectrometer (Bruker, Germany). 2.6. Statistical analysis One-way analysis of variance (ANOVA) and least significant difference (LSD) were used to test the differences in end waste characteristics (i.e., HS content, TOC, etc.) among bioreactors. Type I error a was specified at a level of 0.05. The temporal data set (such as leachate ammonia and COD) was divided into eight subsets, with each subset representing the measurements collected within 50 days. Paired T-test was performed between subsets from different bioreactors within the same measurement period. 3. Results and discussion 3.1. CO2 generation and carbon mineralization Fig. 2(a)–(c) show the CO2 generation data for AR35, AR45 and AR55, respectively. In each test, the percentage of CO2 displayed a linear relationship with time and the slope of the fitting line determined the CO2 generation rate. The higher the slope, the higher the microbial activity and organic substrate degradation rate are. In total, 14 tests were conducted in each bioreactor to reveal the temporal evolution. Based on these date, the calculated carbon mineralization rates versus time are presented in Fig. 2(d). Overall, the variation of carbon mineralization rates can be divided into adjustment phase, active degradation phase and maturation phase. In AR55, the carbon mineralization rate in the adjustment phase started from 26 mg/d and reached rapidly to 111 mg/d on day 56. The rapid increase was attributed to the high metabolic activity of aerobic microbial community. During the adjustment phase, aerobes gradually replaced anaerobes to become the dominant microbial species. When aerobic conditions were fully established and aerobes reached the highest growth rate, the carbon source in waste was exploited at the maximal rate in the active degradation phase. At this phase, the mineralization rate maintained around 80 mg/d. During the maturation phase, the microbial activity diminished since the biodegradable components were eventually depleted. The carbon gasification rate gradually reduced to approximately 40 mg/d around day 250 and no further change was observed afterwards. Antithetical to AR55, the carbon mineralization rate in AR35 was quite low at the adjustment phase. The initial delay indicates that the acclimation of the microbial community was relatively slow at 35 °C. In comparison to AR55, the carbon mineralization rate was lower throughout the experimental period, suggesting that the biodegradation process was more efficient at thermophilic temperature than mesophilic temperature. This observation agrees well with the findings from other studies (Fernández-Rodríguez et al., 2013; Ugwuanyi et al., 2005). The carbon mineralization rate in AR45 displayed similar trend as in AR35. In particular, the delay phase in AR45 prolonged to day 87, which was one month longer than that in AR35. The longer adaption time at 45 °C was also observed in the landfill aeration study by Raga and Cossu (2013). After the lag phase, the mineralization rate increased rapidly reaching up to 90 mg/d between day 211 and 308, and then slightly declined at the maturation phase. At the end of the experiment, AR45 was in beginning of the maturation phase, as the mineralization activity was still high.

152

H. Tong et al. / Bioresource Technology 192 (2015) 149–156

0d 35d 56d 87d 119d 155d 183d 211d 246d 274d 308d 336d 365d 396d

3

2

1

6 CO 2 percentage (%)

CO 2 percentage (%)

4

0d 35d 56d 87d 119d 155d 183d 211d 246d 274d 308d 336d 365d 396d

7

5 4 3 2 1

(a)

0 5

10

15

20

25

30

(b)

0

35

5

10

Time (hours)

5 4 3 2

AR35

1

(c)

0 5

25

30

35

120 Carbon mineralization rate (mg/d)

CO 2 percentage (%)

6

20

Time (hours)

0d 35d 56d 87d 119d 156d 183d 211d 246d 275d 308d 336d 365d 396d

7

15

10

15

20

25

30

35

AR45

AR55

100 80 60 40 20

(d)

0 0

100

200

300

400

Time (days)

Time (hours)

Fig. 2. CO2 generation in each bioreactor: (a) AR35, (b) AR45, (c) AR55, (d) carbon mineralization rate in all bioreactors.

6000

NR35 AR35 AR45 AR55

5000

COD (mg/L)

4000 3000 2000 1000 0 0

100

200

300

400

Time (days) Fig. 3. Evolution of COD concentrations in bioreactor leachate.

3.2. Leachate strength The influence of temperature on leachate organics load (COD) is demonstrated in Fig. 3. The COD in anaerobic bioreactor NR35 showed a gradual decline from initial 2240 mg/L to 100 mg/L at the end, due to leaching processes and anaerobic degradation. By contrast, COD concentrations in all aerobic bioreactors were

rapidly increased to their highest values around 4750 mg/L, 6460 mg/L and 6020 mg/L in AR35, AR45 and AR55, respectively. The initial COD increment after air injection was induced by enhanced hydrolysis process (Kristensen et al., 1995), which may not be as favorable or fast enough in anaerobic conditions. As the percolating water was not sufficient enough to remove COD in pace with its production, the accumulation of COD in leachate were observed (Giannis et al., 2008). Before day 32, the COD values in AR45 and AR55 were up to 38% higher than that in AR35, indicating enhanced solubilization of organic matter at higher temperature. After reaching the maximum value, COD declined in all aerobic bioreactors. The fastest COD attenuation was noticed in AR55, as from day 32, it had the lowest value among all the aeration bioreactors, and it was caused by the accelerated biodegradation. Generally, COD values of AR35 were higher than AR55 after day 32; however, only in 54–149 day and 254–396 day, the differences were considered as significant (p < 0.05). COD in AR45 presented similar trend as in AR55. However, after day 43 it decreased at a much slower rate than that in AR55, and topped among all aerobic bioreactors with significant differences with AR35 and AR55 (p < 0.0001) since day 200 until the end of the experiments. At the end, the COD was around 400 mg/L, 700 mg/L, and 250 mg/L in AR35, AR45 and AR55, respectively. The inefficient mitigation of COD in AR45 was because microbes experienced a much longer lag phase. However, similar low COD value in AR45 could be well anticipated with extended experimental period.

153

H. Tong et al. / Bioresource Technology 192 (2015) 149–156

The variations of NH+4–N concentration in the leachate samples, over 396 days, were plotted in Fig. 4. The initial NH+4-N concentration in the leachate of all bioreactors was 1625 ± 54 mg/L, indicating the homogeneous load of the synthetic aged waste among the bioreactors. In NR35, NH+4–N concentration decreased slowly. Because of the low removal rate in anaerobic condition, ammonia is considered as the most significant long-term problem at traditional landfills (Barlaz et al., 2002). Regarding the aerobic bioreactors, ammonium concentration was reduced drastically below 500 mg/L directly after in situ aeration. The main reason for ammonia attenuation was related to air stripping (Nikolaou et al., 2010). Since pH in the bioreactors was alkaline (data not shown), ammonia volatilization was significant (Berge et al., 2005). The free ammonia fraction is estimated in Eq. (3).

½NH3 —N ¼ ½NHþ4 —N  10pH =½ðK a =K w Þ þ 10pH Þ

ð3Þ NH+4–N

where NH3–N is the free ammonia as N (mg/L); is the total ammonium as N (mg/L) and Ka and Kw are the dissociation constants of ammonia and water. The ratio of Ka/Kw is related to temperature in Eq. (4).

K a =K w ¼ exp½6344=ð273 þ CÞ:

ð4Þ

At temperature 35 °C and pH around 8.5 (average pH value in aerobic bioreactors), 26% of ammonia present was in the form of free ammonia. In comparison, the percentage of free ammonia present at neutral pH (average pH value in anaerobic bioreactors) was only 2%. At higher temperatures, ammonia volatilization could be further promoted. When the temperature was at 45 and 55 °C, the fraction of free ammonia at pH 8.5 increased to 41% and 56%, respectively. Since the NH+4–N volatilization was governed by the temperature, especially in the first place, the NH+4–N concentration in AR45 and AR55 was significantly lower than AR35 (p < 0.05). In aerobic conditions, NH+4–N can be also converted into other forms by the combination of nitrification and denitrification. No matter the efficiency of aeration system, anaerobic or anoxic pockets always exist. In these microenvironments, sequential nitrification (in aerobic zone) and then denitrification (in anoxic/anaerobic zone) could occur, similar to a designed wastewater treatment facility but in micro scale (Ritzkowski et al., 2006). However, these pathways took place after ammonia volatilization because of the low growth rate of autotrophic nitrifiers and potential competition with heterotrophs when organic substrate was abundant (Berge et al., 2005). As the nitrifiers have optimal temperature range

NR35 AR35 AR45 AR55

1600

+ NH 4 -N (mg/L)

1200

800

between 30 and 35 °C and are inhibited at thermophilic temperature range (Willers et al., 1998), the highest ammonium removal efficiency through nitrification–denitrification was expected in AR35. It could potentially explain why after day 149 the NH+4–N concentration in AR35 was significantly lower than AR45 and AR55 (p < 0.05). At the end, the NH+4–N concentration was kept around less than 0.4 mg/L, 20 mg/L, and 3 mg/L in AR35, AR45 and AR55, respectively. Fig. 5 depicts the evolution of nitrite and nitrate in the leachate samples of aerobic bioreactors. Nitrite first increased at day 168 in AR35 and declined at day 200, with a peak around 22 mg/L at day 173. Nitrate appeared in AR35 at around day 190, leveled at 7 mg/L before day 359 and showed a slight increase in the end. In AR55, nitrite and nitrate increased simultaneously at day 196, maintained around 7 mg/L before day 261 and raised to 14 mg/L toward the end. Nitrate content between AR35 and AR55 bioreactors displays no significant difference at p 6 0.05 level. Moreover, the appearances of nitrite and nitrate in AR35 and AR55 were consorted with a considerable drop of carbon mineralization rate, which was a sign of a substantial declined carbon source. Similarly, Valencia et al. (2009) reported an elevated level of nitrate after day 200 due to depleted organic substrate for denitrifier in the bioreactor landfill study. Different from AR35 and AR55, nitrate and nitrite concentrations were found below the detection limit (0.1 mg/l) in AR45, owing to the relative high organic substrate supply for denitrifiers all the time, especially evidenced by the highest COD value in the leachate in the latter half of the period. This in turn supports the previous finding that COD removal was not favored under 45 °C. 3.4. Organic nitrogen in the leachate The nitrogen content at each organic fraction (HA, FA, and HyI) in period I and II is displayed together with inorganic nitrogen in Fig. 6. The TN value for each sample is indicated on top of the bar chart. It was found that the TN was decreased due to aeration up to 87%, with NH+4–N fraction undergoing the greatest reduction (up to 99.9%). Regarding HyI-N concentration, relatively high degradation was noticed in aerobic bioreactors. In period I, 23%, 37%, and 58% less HyI-N was detected in AR35 (43.7 mg/L), AR45 (35.9 mg/L) and AR55 (23.8 mg/L) compared to NR35 (56.6 mg/L), respectively. In period II, HyI-N in AR35 (12.9 mg/L) and AR55 (12.8 mg/L) was 41% lower than NR35 (21.9 mg/L), except AR45 (31.4 mg/L) with slightly higher HyI-N. The relatively low

30

AR35-NO2 AR55-NO2

25

Ion concentration (mg/L)

3.3. Inorganic nitrogen in the leachate

AR35-NO3

20

AR55-NO3

15

10

5

400

0 0

0 0

100

200 Time (days)

300

400

Fig. 4. Ammonium-nitrogen concentration in bioreactors leachate. ‘‘?’’ indicates the two sampling periods for nitrogen speciation analysis.

100

200

300

400

Time (days)  Fig. 5. Variation of leachate NO 2 and NO3 concentration. The concentration of nitrate and nitrite in NR35 and AR45 were below the detection limit, thus data was not shown in figure.

154

H. Tong et al. / Bioresource Technology 192 (2015) 149–156

100

636

231

155

84

190

26

96

30

TN (mg/L)

Percentage (%)

80 HA-N FA-N HyI-N NOx-N

60

40

+

NH4 -N

20

0 NR35-I AR35-I AR45-I AR55-I

NR35-II AR35-II AR45-II AR55-II

Fig. 6. Nitrogen distribution among different nitrogen containing compounds in collective leachate sample from two periods. The TN value for each sample was indicated on top of the bar.

concentrations of HyI-N in aerobic reactors were most likely related to the enhanced degradation of hydrophilic protein and amino acid under aerobic condition (Shao et al., 2013). The HS-N (HA-N and FA-N) played a minor role in NR35, with concentration 12 mg/L (1.9% of TN) in period I and 3 mg/L (1.6% of TN) in period II. By contrast, HS-N showed significant increment in aerobic bioreactors, measured at 100 mg/L (43% of TN), 67 mg/L (43% of TN), and 36 mg/L (43% of TN) in AR35-I, AR45-I, and AR55-I respectively and 11 mg/L (43% of TN), 35 mg/L (36% of TN) and 9 mg/L (32% of TN) in AR35-II, AR45-II and AR55-II, respectively. The lowest contribution of HS-N in AR55 could be relevant to the most intensive humification process in the waste matrix under thermophilic temperature, which raised the molecular weight and polycondensation of HS and thus effectively prevented them from leaching out (Tong et al., 2015). Among HS group, HA-N exceeded FA-N, as the major source of humic nitrogen. In details, HA-N/FA-N ratio was 9, 6, 3 in AR35-I, AR45-I, and AR55-I, and 2, 3, 2 in AR35-II, AR45-II and AR55-II, respectively. The HA-N/FA-N ratio decreased remarkably in period II, probably because more HA was washed out than FA in the previous 215 days. Moreover, as the humification process mainly occurred in the HA fraction rather than the FA fraction (Huang et al., 2006), the level of polycondensation was increased more in HA than FA. Thus, leaching of HA became dourly in the aged landfill waste compared to FA. The main nitrogen fraction in anaerobic bioreactor was ammonium with a dominant percentage of 89% in period I and 86% in

Table 1 HS content in the end waste and relative distribution (%) of different C-types in

13

period II, while under aerobic conditions organic nitrogen was the major constituent, in the range of 62–94%. Therefore, special attention should be paid during aeration with regards to the organic nitrogen management. According to literature, HS-N could be removed by adsorption onto the activated sludge, while electrolysis was proved to be efficient in HyI-N elimination (He et al., 2006). 3.5. HS in the end waste During waste decomposition detritus became rich in reactive phenolic and carbohydrate groups, which could polymerize together with amino acid to form HS. The quantity and properties of HS are often used to indicate the degree of maturity of the treated waste, as the biostability of organic substrate refers to the inherent recalcitrance related to aromatic structures in complex macromolecules, such as HS (Serramiá et al., 2013). Compared to NR35, HA content in AR35 and AR55 showed a statistically significant increase (Table 1), suggesting the positive effects of aeration on waste biostability under mesophilic and thermophilic conditions. The significant decrease of HA in AR45 indicated that the humification and stabilization was not favored at 45 °C as in mesophilic or thermophilic temperatures. The solid state 13C NMR spectra can be divided into seven resonance regions to reflect different carbon structures (Barje et al., 2012; Spaccini and Piccolo, 2009). The relative distribution (%) of

C NMR.

HA

Contenta,b (mg C/g DM) 0–43 ppm alkyl-C 43–60 ppm OACH3 60–110 ppm OACAO anomerics & O-alkyl 110–141 ppm aromatic CAH & CAC 141–162 ppm aromatic CAO 162–190 ppm COOH 190–220 ppm C@O Aromaticityc (%) a

FA

NR35

AR35

AR45

AR55

NR35

AR35

AR45

AR55

5.83 ± 0.29 28.00 15.60 18.80 16.86 6.69 11.32 2.74 0.27

7.78 ± 0.00 19.70 15.66 24.75 19.15 8.18 9.71 2.85 0.31

3.68 ± 1.38 21.02 14.89 22.89 19.23 7.88 10.83 3.26 0.32

7.93 ± 0.50 19.72 15.58 21.16 22.66 9.13 9.66 2.09 0.36

0.93 ± 0.02 31.80 13.45 19.25 14.90 4.89 14.06 1.65 0.23

1.11 ± 0.09 28.57 12.74 19.14 15.00 5.63 15.41 3.51 0.25

1.68 ± 0.34 28.37 12.65 18.31 14.87 6.48 16.49 2.84 0.26

1.43 ± 0.23 30.48 12.29 18.07 14.97 6.05 15.59 2.55 0.26

Mean value of duplicates. The differences of HA content are significant and AR55 = AR35 > NR35 > AR45 at p 6 0.05 level. FA content among bioreactors displays no significant difference at p 6 0.05 level. c Aromaticity (%) = [aromatic C peak area (110–162 ppm)/total peak area (0–162 ppm)]. b

155

H. Tong et al. / Bioresource Technology 192 (2015) 149–156 Table 2 Waste Characterizationa before and after bioreactor operation and carbon discharge through different pathways.

b

Initial TOC (g C/kg DM) End TOCb (g C/kg DM) Leachate C discharge (g C/kg DM) Mineralization C discharge (g C/kg) Initial GS21c (L kg1 DM) End GS21c (L kg1 DM) a b c

NR35

AR35

AR45

AR55

583.6 ± 13.2 574.8 ± 19.5 2.7 5.0 11.5 ± 2.7 6.3 ± 1.0

583.6 ± 13.2 547.2 ± 14.7 7.9 20.2 11.5 ± 2.7 4.8 ± 0.1

583.6 ± 13.2 539.2 ± 13.6 10.1 24.6 11.5 ± 2.7 5.2 ± 0.9

583.6 ± 13.2 537.3 ± 1.3 7.0 23.3 11.5 ± 2.7 4.0 ± 0.9

There was no significant difference in values of end TOC and end GS21 at p 6 0.05 level. Mean value of triplicates. Mean value of duplicates.

each C-type was calculated by dividing the peak area in this specific region by the total spectral peak area (Table 1). Compared to anaerobic HS, a clear decrease of alkyl carbon (d = 0–43 ppm) was noticed in aerobic HS. The loss of alkyl carbon was most probably caused by the biomineralization of terminal methyl, methine and methylene groups in aliphatic chains (Keeler et al., 2006), as the abatement of their typical peak at 20 ppm, 24 ppm, and 30 ppm in the spectra of aerobic HS was observed. Unlike easily biodegradable alkyl structure, the recalcitrant aromatic portion would be preserved during humification. The higher content of aromatic carbon (d = 110–162 ppm) in aerobic HS indicates superior enrichment of aryl structure. This corresponds to the finding by Lguirati et al. (2005), where a higher level of aromacity and condensation of humic molecules in the aerated landfill waste was observed. Interpretation of NMR data further revealed that high temperature benefited the maturation of HA, as the degree of aromacity increased with a higher temperature. NMR spectra of aerobic HA exhibited higher intensity in the chemical shift region of 60–110 ppm, which was assigned to O-alkyl carbon in the monomeric units in oligo- and polysaccharidic chains of yard waste. The rise of O-alkyl carbon in aerobic HA could be relevant to the polymerization progress, in which new polysaccharide component was incorporated into HA structure by covalent bonds (Serramiá et al., 2013). As a matter of fact, this was important in improving the long term stability of the waste materials. Since mercurial organic compounds were converted into more stable polymeric HA, the pollution potential was therefore reduced. In contrast to the observation for HA spectra, no significant changes were found in FA spectra apart from the variation of alkyl-C and aromatic C. This was supported by the previous finding that more structural alteration occurred in HA than FA (Huang et al., 2006).

discharge via gas release in NR35 was calculated by utilizing the analyzed gas composition (CO2 and CH4) and the measured gas volume. Carbon removal through leaching (referred to the unit dry mass) were 7.9, 10.1 and 7.0 g C/kg DM in AR35, AR45 and AR55, respectively, while a relative lower value of 2.7 g C/kg DM was obtained in NR35. The enhanced leaching of the organic carbon in aerobic bioreactors in current study was significantly correlated with the elevated concentration of HA after aeration. Our previous studies found that the enhanced dissolution of HA in present work was due to the inherent young waste age of the synthetic aged waste used (Tong et al., 2015). However, it is worthy to highlight the positive effects of thermophilic temperature on humification process, which reduced the dissolution of HA and decreased the carbon load in the leachate. Due to accelerated biological process under aerobic condition, the carbon discharge through biodegradation (mineralization) was largely promoted with a value of 20.2, 24.6 and 23.3 g C/kg DM in AR35, AR45 and AR55, respectively, which was more than fourfold of that under anaerobic condition (NR35). Although the bioprocess was delayed in AR45, it outreached at the later stage and eventually achieved the largest amount of carbon conversion through mineralization. A similar phenomenon was also observed by Raga and Cossu (2013). Additionally, aeration could also change the carbon removal pathways. In NR35, the ratio of carbon discharge between mineralization and leaching was 1.8, while it increased to 2.6, 2.4 and 3.3 in AR35, AR45 and AR55, respectively. Similarly, Ritzkowski and Stegmann (2013) and Matsufuji et al. (2005) investigated the carbon discharge, observing a boosted carbon discharge through mineralization under aerobic and semi-aerobic conditions. However, the relatively lower ratio in AR45 among all aerobic bioreactors indicated more carbon load in the leachate, adding extra burden to the downstream leachate treatment system.

3.6. Waste characterization and total carbon discharge

4. Conclusions

The results of waste characterization before and after bioreactor operation are shown in Table 2. The observed differences in end TOC and GS21 were not statistically significant. Due to the nature of severe heterogeneity of waste sample, the measurement of TOC commonly has a large variance. Ritzkowski et al. (2006) also commented that testing on TOC alone was not sufficient to describe the stabilization process of the waste material under aeration. Therefore, a calculation of carbon discharge on each pathway should be conducted. Table 2 lists the carbon discharge through different pathways, in terms of leaching and mineralization. The carbon load in each leachate drainage process was calculated by multiplying DOC and the volume of discharged leachate (data not shown). In this regards, the total carbon removal through leaching was obtained with sum of each DOC discharge. The amount of mineralized carbon through bioconversion was estimated by integrating the corresponding carbon mineralization curve in Fig. 2(d). The carbon

The study examined the influence of temperature (35, 45 and 55 °C) during in situ aeration of simulated landfill bioreactors. A higher biodegradation rate occurred at 55 °C than 35 °C, which eventually led to the lowest leachate COD in thermophilic bioreactor among all the aeration bioreactors. Extended acclimation period in biodegradation process was observed at 45 °C. Volatilization played a significant role in ammonium removal under aerobic condition resulting to >60% nitrogen present in organic forms. The highest waste biostability was achieved at 55 °C, while the inadequacy of HA in the waste materials were found at 45 °C. Acknowledgements The authors would like to acknowledge the financial supported by the Singapore National Research Foundation (Project ID: NRF-CRP5-2009-02).

156

H. Tong et al. / Bioresource Technology 192 (2015) 149–156

References Barje, F., El Fels, L., El Hajjouji, H., Amir, S., Winterton, P., Hafidi, M., 2012. Molecular behaviour of humic acid-like substances during co-composting of olive mill waste and the organic part of municipal solid waste. Int. Biodeterior. Biodegrad. 74, 17–23. Barlaz, M.A., Rooker, A.P., Kjeldsen, P., Gabr, M.A., Borden, R.C., 2002. Critical evaluation of factors required to terminate the postclosure monitoring period at solid waste landfills. Environ. Sci. Technol. 36 (16), 3457–3464. Berge, N., Reinhart, D., Townsend, T., 2005. The fate of nitrogen in bioreactor landfills. Crit. Rev. Environ. Sci. Technol. 35 (4), 365–399. Fernández-Rodríguez, J., Pérez, M., Romero, L.I., 2013. Comparison of mesophilic and thermophilic dry anaerobic digestion of OFMSW: kinetic analysis. Chem. Eng. J. 232, 59–64. Giannis, A., Makripodis, G., Simantiraki, F., Somara, M., Gidarakos, E., 2008. Monitoring operational and leachate characteristics of an aerobic simulated landfill bioreactor. Waste Manage. 28 (8), 1346–1354. Hanson, J.L., Yeiller, N., Oettle, N.K., 2010. Spatial and temporal temperature distributions in municipal solid waste landfills. J. Environ. Eng.-ASCE 136 (8), 804–814. He, P.-J., Xue, J.-F., Shao, L.-M., Li, G.-J., Lee, D.-J., 2006. Dissolved organic matter (DOM) in recycled leachate of bioreactor landfill. Water Res. 40 (7), 1465–1473. Huang, G.F., Wu, Q.T., Wong, J.W.C., Nagar, B.B., 2006. Transformation of organic matter during co-composting of pig manure with sawdust. Bioresour. Technol. 97 (15), 1834–1842. IHSS. http://www.humicsubstances.org/soilhafa.html. Jacobs, J., Scharff, H., Van Arkel, F., De Gier, C.W., 2003. Odour reduction by aeration prior to excavation. In: Ninth Sardinia International Waste Management and Landfill Symposium, Cagliari, Italy. Keeler, C., Kelly, E.F., Maciel, G.E., 2006. Chemical–structural information from solid-state 13C NMR studies of a suite of humic materials from a lower montane forest soil, Colorado, USA. Geoderma 130 (1–2), 124–140. Kim, H.W., Nam, J.Y., Kang, S.T., Kim, D.H., Jung, K.W., Shin, H.S., 2012. Hydrolytic activities of extracellular enzymes in thermophilic and mesophilic anaerobic sequencing-batch reactors treating organic fractions of municipal solid wastes. Bioresour. Technol. 110, 130–134. Kristensen, E., Ahmed, S.I., Devol, A.H., 1995. Aerobic and anaerobic decomposition of organic matter in marine sediment: which is fastest? Limnol. Oceanogr. 40 (8), 1430–1437. Labatut, R.A., Angenent, L.T., Scott, N.R., 2014. Conventional mesophilic vs. thermophilic anaerobic digestion: a trade-off between performance and stability? Water Res. 53, 249–258. Levén, L., Eriksson, A.R.B., Schnürer, A., 2007. Effect of process temperature on bacterial and archaeal communities in two methanogenic bioreactors treating organic household waste. FEMS Microbiol. Ecol. 59 (3), 683–693. Lguirati, A., Ait Baddi, G., El Mousadik, A., Gilard, V., Revel, J., Hafidi, M., 2005. Analysis of humic acids from aerated and non-aerated urban landfill composts. Int. Biodeterior. Biodegrad. 56 (1), 8–16. Matsufuji, Y., Tachifuji, A., Matsugu, H., 2005. Biodegradation process of municipal solid waste by semiaerobic landfill type. In: Thirteenth International Landfill Symposium, Cagliari, Italy.

Mitch, W.A., Sedlak, D.L., 2004. Characterization and fate of Nnitrosodimethylamine precursors in municipal wastewater treatment plants. Environ. Sci. Technol. 38 (5), 1445–1454. Nikolaou, A., Giannis, A., Gidarakos, E., 2010. Comparative studies of aerobic and anaerobic treatment of MSW organic fraction in landfill bioreactors. Environ. Technol. 31 (12), 1381–1389. Öncü, G., Reiser, M., Kranert, M., 2012. Aerobic in situ stabilization of landfill Konstanz Dorfweiher: leachate quality after 1 year of operation. Waste Manage. 32 (12), 2374–2384. Prantl, R., Tesar, M., Huber-Humer, M., Lechner, P., 2006. Changes in carbon and nitrogen pool during in-situ aeration of old landfills under varying conditions. Waste Manage. 26 (4), 373–380. Raga, R., Cossu, R., 2013. Bioreactor tests preliminary to landfill in situ aeration: a case study. Waste Manage. 33 (4), 871–880. Raga, R., Cossu, R., 2014. Landfill aeration in the framework of a reclamation project in Northern Italy. Waste Manage. 34 (3), 683–691. Ritzkowski, M., Stegmann, R., 2013. Landfill aeration within the scope of postclosure care and its completion. Waste Manage. 33 (10), 2074–2082. Ritzkowski, M., Heyer, K.U., Stegmann, R., 2006. Fundamental processes and implications during in situ aeration of old landfills. Waste Manage. 26 (4), 356–372. Serramiá, N., Sánchez-Monedero, M.A., Roig, A., Contin, M., De Nobili, M., 2013. Changes in soil humic pools after soil application of two-phase olive mill waste compost. Geoderma 192, 21–30. Shao, L., Wang, T., Li, T., Lü, F., He, P., 2013. Comparison of sludge digestion under aerobic and anaerobic conditions with a focus on the degradation of proteins at mesophilic temperature. Bioresour. Technol. 140, 131–137. Spaccini, R., Piccolo, A., 2009. Molecular characteristics of humic acids extracted from compost at increasing maturity stages. Soil Biol. Biochem. 41 (6), 1164– 1172. Tchobanoglous, G., Theisen, H., Vigil, S.A., 1993. Integrated Solid Waste Management: Engineering Principles and Management Issues. McGraw-Hill, London. Tong, H., Yin, K., Ge, L., Giannis, A., Chuan, V.W., Wang, J.Y., 2015. Monitoring transitory profiles of leachate humic substances in landfill aeration reactors in mesophilic and thermophilic conditions. J. Hazard. Mater. 287C, 342–348. Ugwuanyi, J.O., Harvey, L.M., McNeil, B., 2005. Effect of digestion temperature and pH on treatment efficiency and evolution of volatile fatty acids during thermophilic aerobic digestion of model high strength agricultural waste. Bioresour. Technol. 96 (6), 707–719. Valencia, R., van der Zon, W., Woelders, H., Lubberding, H.J., Gijzen, H.J., 2009. Achieving ‘‘Final Storage Quality’’ of municipal solid waste in pilot scale bioreactor landfills. Waste Manage. 29 (1), 78–85. Westgate, P.J., Park, C., 2010. Evaluation of proteins and organic nitrogen in wastewater treatment effluents. Environ. Sci. Technol. 44 (14), 5352–5357. Willers, H., Derikx, P., Ten Have, P., Vijn, T., 1998. Nitrification limitation in animal slurries at high temperatures. Bioresour. Technol. 64 (1), 47–54. Yazdani, R., Mostafid, M.E., Han, B., Imhoff, P.T., Chiu, P., Augenstein, D., Kayhanian, M., Tchobanoglous, G., 2010. Quantifying factors limiting aerobic degradation during aerobic bioreactor landfilling. Environ. Sci. Technol. 44 (16), 6215–6220.