Waste Management 34 (2014) 439–447

Contents lists available at ScienceDirect

Waste Management journal homepage: www.elsevier.com/locate/wasman

Non-biodegradable landfill leachate treatment by combined process of agitation, coagulation, SBR and filtration Alkhafaji R. Abood a,b, Jianguo Bao a,⇑, Jiangkun Du a, Dan Zheng a, Ye Luo a a b

State Key Laboratory of Biogeology and Environmental Geology, China University of Geosciences, Wuhan 430074, People’s Republic of China Thi Qar University, Nasiriyah, Iraq

a r t i c l e

i n f o

Article history: Received 15 March 2013 Accepted 17 October 2013 Available online 25 November 2013 Keywords: Landfill leachate Agitation PFS coagulation Anoxic–aerobic SBR Filtration

a b s t r a c t This study describes the complete treatment of non-biodegradable landfill leachate by combined treatment processes. The processes consist of agitation as a novel stripping method used to overcome the ammonia toxicity regarding aerobic microorganisms. The NH3-N removal ratio was 93.9% obtained at pH 11.5 and a gradient velocity (G) 150 s1 within a five-hour agitation time. By poly ferric sulphate (PFS) coagulation followed the agitation process; chemical oxygen demand (COD) and biological oxygen demand (BOD5) were removed at 70.6% and 49.4%, respectively at an optimum dose of 1200 mg L1 at pH 5.0. The biodegradable ratio BOD5/COD was improved from 0.18 to 0.31 during pretreatment step by agitation and PFS coagulation. Thereafter, the effluent was diluted with sewage at a different ratio before it was subjected to sequencing batch reactor (SBR) treatment. Up to 93.3% BOD5, 95.5% COD and 98.1% NH3-N removal were achieved by SBR operated under anoxic–aerobic–anoxic conditions. The filtration process was carried out using sand and carbon as a dual filter media as polishing process. The final effluent concentration of COD, BOD5, suspended solid (SS), NH3-N and total organic carbon (TOC) were 72.4 mg L1, 22.8 mg L1, 24.2 mg L1, 18.4 mg L1 and 50.8 mg L1 respectively, which met the discharge standard. The results indicated that a combined process of agitation-coagulation-SBR and filtration effectively eliminated pollutant loading from landfill leachate. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction Sanitary landfill is one of the most widely employed methods for the disposal of municipal solid waste (MSW) and up to 95% of such waste collected worldwide is disposed of in landfill (Kurniawan et al., 2006a). This method of waste disposal often results in the generation of large quantities of liquid leachate, which is a kind of wastewater with a high content of dissolved organic matter (DOM), and may move out of the fill and into the aqueous surroundings (Calace et al., 2001). Leachates from municipal solid waste landfill sites are often defined as hazardous and heavily polluted wastewater. It may contain large amounts of organic matter (biodegradable) but is also refractory as regards biodegradation, where humic-type constituents comprise an important group (Kang et al., 2002). The characteristics of landfill leachate depending on the type of MSW being dumped, the degree of solid waste stabilisation, site hydrology, moisture content, seasonal weather variations, landfill age, and the stage of landfill decomposition (Al-Yaqout and Hamoda, 2003). The treatment processes of leachate are very complicated, expensive and generally require various process applications ⇑ Corresponding author. Tel.: +86 27 883470; fax: +86 27 87436235. E-mail address: [email protected] (J. Bao). 0956-053X/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.wasman.2013.10.025

because of their high loading, complex chemical composition and seasonally variable volume (Bu et al., 2010). At present, no single unit process is available for proper leachate treatment simply because of the high concentrations of COD and nitrogen as well as colour. The combination of biological, chemical, and physicochemical processes have the ability to synergise the advantages of each single process, and has been documented as effective for treating stabilised landfill leachate (Renou et al., 2008; Wang et al., 2009). Therefore, the development of affordable technologies for the treatment of such contaminant streams has become mandatory in the last few years (Galeano et al., 2011). Biological processes used to be the first choice of landfill leachate on economic grounds; wherein several methodologies can be mentioned: anaerobic digester-activated sludge systems (Kheradmand et al., 2010), anaerobic sequencing batch reactor (ASBR) (Timur and Ozturk, 1999), anaerobic fluidised bed bioreactor (AFBR) (Gulsen and Turan, 2004), up flow anaerobic sludge blanket (UASB) (Ye et al., 2011) and aerobic treatment, such as the activated sequencing batch reactor (Spagni and Marsili-Libelli, 2009; Uygur and Kargi, 2004; Yanjie Wei et al., 2012). Thanks to their numerous advantages, systems such as anaerobic–aerobic (A/O) or anaerobic–anoxic–aerobic (A2/O) bioreactors have attracted considerable attention over the past decades in landfill leachate treatments (Agdag and Sponza, 2008; Fang et al., 2011;

440

A.R. Abood et al. / Waste Management 34 (2014) 439–447

Yu et al., 2010). Biological processes are quite effective when applied to relatively young leachates, but they are less efficient for the treatment of older ones. Bio-refractory contaminants, contained mainly in older leachates, are not amenable to conventional biological processes and the high ammonia content might also be inhibitory with regard to activated sludge microorganisms (He et al., 2009). Furthermore, a supplementary addition of phosphorus has often been necessary, as landfill leachates are generally phosphorus deficient. Therefore, a combination of physicochemical and biological methods is often required for the efficient treatment of leachate (Shi et al., 2009). In the recent years, a combination process has been widely used in landfill leachate treatment either the combination of two or more physicochemical treatment (Kurniawan et al., 2006b; Monje-Ramirez and Velasquez, 2004), or a combination of physicochemical and biological treatments (Cassano et al., 2011; Guo et al., 2010; Vilar et al., 2011; Wu et al., 2011). Ammonia stripping is the most widely employed treatment for the removal of NH3-N from landfill leachate. NH3-N is transferred from the waste stream into the air and is then absorbed from the air into a strong acid such as sulphuric acid or directly fluxed into the ambient air (Marttinen et al., 2002). A conventional stripping process is carried out by passing quantities of air over the exposed surface of the leachate, giving NH3-N removal in the range of 85– 95% with the concentrations ranging from 220 to 3260 mg L1 in a contact time of 18–24 h (Bonmati and Flotats, 2003). Therefore, the modification of this process to reduce NH3-N concentration with high efficiency in a short time is necessary in order to save time and cost. Coagulation–flocculation has been employed for the removal of suspended solids (SS), colloid particles, non-biodegradable organic compounds, and heavy metals from landfill leachate (Amokrane et al., 1997). Colloidal particles can be destabilised by the addition of a coagulant in the coagulation process, and usually coagulation is followed by the flocculation process to increase the particle size and unstable particles into bulky floccules so that they can settle more easily (Cheng et al., 1994). Nevertheless, the coagulation process has been investigated mainly by using stabilized or biologically pretreated landfill leachate, as a final polishing treatment stage. Limited information exists on the efficiency of this physicochemical process when applied in the removal of pollutants from leachates partially stabilised by recirculation or recently produced leachates. This technique may be important for the enhancement of subsequent leachate biodegradability, i.e. when applied prior to biological treatment (Tatsi et al., 2003). Landfill leachates are often co-treated with municipal sewage in the biological process. SBR as a biological treatment is the most economically efficient methods for the removal of biodegradable organic compounds (Spagni and Marsili-Libelli, 2009). The SBR technology is an activated sludge process designed to operate under non-steady state flow conditions and there is a degree of flexibility associated with working in a time rather than in a space sequence (Laitinen et al., 2006). Problems with the high concentration of suspended solids in the effluent of activated sludge systems have been observed because of the sludge bulking or dispersed growth phenomena. In the present study, leachate generated from Chang Shankou landfill in Wuhan City, China was collected. A new combination of physicochemical and biological processes was applied to treat this leachate; these processes aimed to investigate the potential of the agitation as a novel process of ammonia stripping for NH3-N removal under different conditions to eliminate or reduce NH3-N from leachate with high efficiency in a short time. The performance of PFS coagulation followed agitation as a pretreatment for leachate to enhance and improve the biodegradable (BOD5/ COD) ratio in order to prepare it for further treatment (SBR

treatment). The SBR reactor was improved by varying the operational conditions. The reactor was operated under anoxic– aerobic–anoxic conditions to remove biodegradable components and enhance biological nitrogen removal (BNR). A layer of carbon and sand was used as an effective medium in filtration as a polishing treatment process of the combined treatment. The final effluent quality of leachate (COD, BOD5, SS and NH3-N) was tested to meet the Chinese landfill pollution control standard (GB16889-2008). 2. Materials and methods 2.1. Sample preparation and collection Leachate samples were collected from the sanitary landfill site of Chang Shankou, which has been operated since 2007; it is located in Wuhan City the capital of Hubei province, PR China. Wuhan’s climate is humid subtropical with abundant rainfall and four distinctive seasons. About 2800 tons of municipal solid wastes are disposed of daily. Leachate generation in the landfill was about 400–500 m3 day1. The composition of the landfill leachate varies greatly depending on the season, the leachate collection system, and particularly on the age of the landfill. Leachate samples were collected three times from the pond in the landfill site without any pretreatment in a 50 L plastic container then, the sample was transported to the laboratory and stored at 4 °C. Leachate samples were removed from the refrigerator and were placed for about 2 h under ambient temperature in the range of (20–25 °C). Then, sample bottles were thoroughly shaken, for re-suspension of possibly settling solids. The collected leachate was filtered through a glass microfiber filter (470 lm) prior to the analysis to remove suspended solids particles. The main characteristics such as pH, SS, COD, BOD5, NH3-N, NO3-N, NO2-N, TP and TOC of the leachate were determined. 2.2. Chemicals and analytical methods All chemicals used for the analytical determinations were of analytical grade. The values of pH were measured with a Sartorius pH Meter PB-11. Dissolved oxygen (DO) was measured by using a selective probe and the Winkler method. COD, BOD5, TP, TN, NO2N, NO3-N, TOC, SS, and settling velocity index (SVI) were measured according to the Standard Methods for the Examination of Water and Wastewater (APHA, 2005). NH3-N was determined by the stander method (Water quality–Determination of ammonia nitrogen– Nessler’s reagent spectrophotometry method, GB 7479-87). MLSS was determined by drying the sludge sample at 105 °C for 24 h. The adjustment of pH was done by using 1 M of H2SO4 and NaOH. The mass of humics was reflected by UV spectrum in 254 nm. All the experiments were carried out at room temperature 25 ± 2 °C except for the agitation process, which was carried out in 10, 20 and 30 °C. Every experiment was repeated at least twice and standard deviation was used as a statistical indicator to show the standard error and data variation. The expression is based on the average value minus or plus standard error; the removal efficiency based on the average value of influent and effluent. Selected samples were repeatedly analysed in order to validate/evaluate the obtained results and they were found to be within the acceptable analytical error range. 2.3. Experimental plan Experimental work was carried out in a combined sequential treatment test run. The landfill leachate was first fed to the agitation process to remove or reduce ammonia. The effluent from that unit was treated by the coagulation–flocculation process using PFS

441

A.R. Abood et al. / Waste Management 34 (2014) 439–447

as coagulant to decrease the concentration of organic matter and modify the BOD5/COD ratio. Then, the effluent was mixed with the municipal sewage wastewater at a different ratio and fed to the SBR unit to enhance the removal of organic matter. Finally, filtration using a layer of carbon and sand as an effective medium was used as a polishing process to meet effluent discharge standard. The combined treatment was operated under the optimum conditions for each process. COD, BOD5, TOC and NH3-N of the effluents were measured at the end of each process. The overall efficiency of the combined treatment was investigated. 2.4. Experimental procedures 2.4.1. Agitation process The ammonia stripping experiments were performed using a high mixing rate as a modified process to eliminate NH3-N and the process was termed ‘agitation’ because it depends on the agitation rate or rotating speed subjected to the leachate surface. The agitation process was carried out in the following sequential steps: (1) the leachate sample (500 ml) was put in a beaker (1000 ml), (2) the pH was adjusted to a fixed value (pH = 11.5), (3) the mixture was agitated for a specific time with a certain gradient velocity using a jar-test device ZR-6, (4) the mixtures were allowed to settle for half an hour, and (5) the NH3-N of the supernatant was measured. The optimum pH, gradient velocity (G) and time were determined. 2.4.2. PFS coagulation The coagulation process was performed in a conventional jartest apparatus ZR-6. The experimental process consisted of the following stages: (1) 1 L of the agitation treated leachate was put in beaker, (2) the pH was adjusted to a fixed value, (3) the desired dose of PFS was added as a coagulant to the leachate, (4) experiments were carried out at a rapid mixing of 250 rpm for 2 min and at a slow mixing of 60 rpm for 20 min, (5) 1 mL of 0.1% polyacrylamide was added to the sample as a flocculent in the slow mixing stage to increase the flocculation settling rate, and (6) the mixture was allowed to settle for 1 h. The supernatant was withdrawn from a point located about 2 cm below the top of the liquid level in the beaker. The COD and NH3-N of the supernatant was measured. The optimum pH value and PFS dosage were determined. 2.4.3. SBR treatment Biological treatment of the leachate effluent from the coagulation process was carried out in a lab-scale sequencing batch reactor (SBR) made of a cylindrical reactor (10 L plastic barrel) operating on the principle of five phases: fill, react, settle, draw, and idle. The react phase was run under anoxic–aerobic–anoxic conditions. The reactor was filled with a wastewater mixture with a different ratio of 1:1, 1:2, 1:3 and 1:4 of leachate effluent from the coagulation process and municipal sewage wastewater. The addition of raw municipal sewage was necessary because of the low biodegradability of the leachate effluent obtained from the coagulation process. The mixed wastewater was seeded with activated sludge 3000–3500 mg L1 mixed liquor suspended solids (MLSS) obtained from Tangxun wastewater treatment plant. The wastewater mixture was continuously aerated with sufficient aeration provided by a compressor connected to porous stone located close to the bottom of the reactor. The compressor was used only during the aerobic period to ensure an oxygen concentration of a certain value. In addition to carbon released from the degradation of COD, a low amount of glucose was added to the reactors as a supplementary carbon source for domesticated sludge performance stability. Leachate effluent PFS coagulation was mixed with sewage in order

to reduce the concentration of inhibitory compounds and improve the treatability of the high-strength leachate (Li and Zhao, 2001). A complete mixture in the anoxic condition was achieved during the reaction phase with a mechanical stirrer (100 rpm). During the aeration reaction the autotrophic nitrifying bacteria oxidise the ammonia to nitrate through nitrification process to supply the needed nitrate for denitrification under anoxic conditions (Agdag and Sponza, 2008). The aerobic reactor was provided with an air compressor and air diffuser at the bottom to ensure a sufficient level of DO during the aeration period in the reaction time. The pH value was maintained at 7.0–8.0. During these investigations each SBR cycle consisted of the following phase: fill (0.5 h), anoxic reaction (2 h), aerobic reaction (8 h), anoxic reaction (2 h), settle (1 h), and, finally draw (0.5 h). The system reached a steady state within 35 to 40 days of acclimatisation. Then, the optimal reaction time and DO level were investigated for the reactor. Samples were taken from the discharged clear effluent for COD and NH3-N measurements. Also, the MLSS concentration at the beginning of each cycle and the SVI were monitored. The optimum aeration times and dissolved oxygen for different mixing ratios were determined. 2.4.4. Filtration process The filtration process was used as a polishing process in the leachate treatment to obtain high removal efficiency in order to meet the stricter discharge standard. The filtration process was carried out using a cylindrical column made of Perspex with a total area 20 cm2. Sand–carbon used as the filter media during this process operated with a filtration rate of 1 m3/m2/h. A mixture was pumped into a column using a water pump to control the flow rate to the filtration column at the desired value. 3. Results and discussion 3.1. Composition of raw landfill leachate The leachate samples collected from the wastewater pond in the landfill site were analysed. The main chemical/physical composition of leachate was as explained in Table 1. The BOD5/COD ratio of leachate was 0.18. The landfill leachate was considered to have a low BOD5/COD ratio and high NH3-N content (2400 mg L1). Thus, it was classified as stabilised ‘mature’ or ’old’ and non-biodegradable leachate (Guo et al., 2010). 3.2. Agitation process The agitation process can be successful in eliminating the ammonia nitrogen to reduce the wastewater toxicity. Optimum pH, gradient velocity, temperature and agitation retention time Table 1 Composition of the raw landfill leachate.

a b c

Parameter

Unit

Concentration

pH COD NH3-N NO2-N NO3-N TN a TP b BOD5 TOC c SS BOD5/COD

– mg L1 mg L1 mg L1 mg L1 mg L1 mg L1 mg L1 mg L1 mg L1 –

8.3 ± 0.2 2800 ± 122 2400 ± 71 8.20 ± 0.1 293.30 ± 10.5 2900 ± 128 28 + 1.2 510 + 28 780 + 32 980 + 56 0.18 + 0.02

TN: total nitrogen. TP: total phosphorous. TOC: total organic carbon.

442

A.R. Abood et al. / Waste Management 34 (2014) 439–447

are the main parameters that have an effect on the NH3-N removal ratio in this process. The optimum pH for NH3-N removal was determined by two steps: first, leachate was agitated without any pH adjustment and, second, with the required dosages of NaOH for raising pH to 10, 10.5, 11, 11.5 and 12, respectively. After a specific time of agitation, ammonia removal was 47.2%, 51.9% and 59.3% for leachate without pH adjustment and its removal ratios at pH 11.5 were around 87.5%, 88.2% and 90.9% under temperatures 10, 20 and 30 °C, respectively, as shown in Fig. 1(a). The effects of G on the NH3-N removal at different temperatures are shown in Fig. 1(b). The removal increased significantly with increases in the agitation rate or gradient velocity. It was shown that the NH3-N removal efficiency increased with the increase in G at the same temperature. At different temperatures the NH3-N removal efficiency was greater than 90% and the NH3-N concentration of effluent was lower than 200 mg L1 when the G was above 125 s1. The effect of agitation time on the NH3-N removal is shown in Fig. 1(c). It indicates that NH3-N removal increased significantly with increases in the agitation time up to 5 h; thereafter, the increase in NH3-N removal was not significant. The temperature did not bring about obvious enhancement of NH3-N removal when the agitation process was used; the NH3-N removal efficiency increased not more than 2% when the temperature increased from 20 °C to 30 °C under optimum conditions of pH 11.5 and G 125 s1 at the same time. The agitation process can be carried out by applying a high rotating speed to the exposed leachate surface, thus causing the partial pressure of the ammonia gas within the water to drive the ammonia from the liquid to the gas phase. The process is

further subject to careful pH control and involves the mass transfer of volatile contaminants from water to air. Free ammonia begins to form when the pH is above 7. Over 90% of the present ammonia may be liberated as a gas through agitation in the absence of air at a pH greater than 11. This result is mainly the result of the fact that the reaction of NH3 with water can be represented by Eq. (1) (Bonmati and Flotats, 2003). From this equation, raising the pH (as represented by the OH) will drive the reaction to the left, increasing the concentration of NH3. This makes ammonia more easily removed by agitation.

NH3 þ H2 O ! NH4 OH

ð1Þ

Ammonium hydroxide is formed as an intermediate product in the reaction at a pH between 10 and 12. Solubility increases at low ambient temperatures since ammonia is highly soluble in water (Gustin and Marinsek-Logar, 2011). In this study, the ammonia removal was about 93.9% at contact times of 5 h, at a gradient velocity of 150 s1 and pH 11.5, in the absence of air.

3.3. PFS coagulation The selection of coagulant is important in the chemical coagulation and depends upon the nature of the suspended solid to be removed, the raw water conditions, the facility design and the cost of the amount of chemical necessary to produce the desired result. The negatively charged colloids are neutralised and caused to aggregate in flocs effectively when PFS is added (Wu et al., 2011), so it was used as a coagulant material in this study. The pH and

Fig. 1. Conditions effected on NH3-N removal in ammonia stripping by agitation process: (a) pH effect, (b) gradient velocity (G) effect and (c) effect of time. Error bars represent standard deviations.

A.R. Abood et al. / Waste Management 34 (2014) 439–447

dosage of coagulant were determined as a function of COD removal. For the purpose of determining the optimum value of the initial pH, the pH of the effluent agitation liquid was varied from 4 to12 while the coagulation process was run with the addition of 1000 mg L1 of PFS. Fig. 2(a) presents the effect of pH values on COD and NH3-N removal in the coagulation process. It shows that the highest removal percentage for COD was 61.9% obtained at pH 5 as the initial value of influent. The results clearly indicate that the removal efficiency for COD was increased in acidic rather than alkaline conditions and was very low in neutral conditions; it also indicates that the removal efficiency was increased with increases of pH up to 5. Then, the removal efficiency decreased for pH below 5. For NH3-N the highest value of removal ratio was 22.2% also at pH 5. Fig. 2(a) shows that the variation in pH value was not significant in NH3-N removal from landfill leachate; this result may be because NH3 begins to form when pH is greater than 7, while the optimum value of pH in PFS coagulation is 5. The influence of pH on chemical coagulation may be considered as a balance of two competitive forces: (1) between hydrogen ions H+ and metal hydrolysis products for interaction with organic ligands and (2) between OH and organic anions for interaction with metal hydrolysis products (Stephenson and Duff, 1996). At low pH values (pH 6 5), H+ out-competes metal hydrolysis products for organic ligands, and hence poor removal rates occur, and some of the generated organic acids will not precipitate. At higher pH values (pH > 5), OH competes with organic compounds for metal adsorption sites and the precipitation of metal hydroxides occurs mainly by co-precipitation (Stephenson and Duff, 1996). The effect of the PFS dose on the efficiency of COD and NH3-N removal was also determined. Fig. 2(b) presents the removal of COD by different dosages of PFS. The PFS dosage varied from 300 to 2400 mg L1 at an initial value of pH 5. The optimum dosage of PFS to attain a better removal percentage of 65.3% and 26.9% for COD and NH3-N, respectively, was 1200 mg L1. The results indicated that COD removal has been increased with increasing coagulant dosages up to the optimum dosage. Then, the COD removal decreased. This result is mainly due to the fact that the optimum coagulant dosage produced flocs having a good structure and consistency (Guo et al., 2010). However, in lower doses than the optimum, the produced flocs are small and influence the settling velocity of the sludge. In higher doses than the optimum, in addition to the small size of floc, the rest ability can be affected (Stephenson and Duff, 1996). 3.4. SBR treatment The SBR process was selected and tested in the present work as the next treatment step for leachate effluent from the coagulation

443

process. In spite of the biodegradable ratio BOD5/COD of the leachate effluent from the coagulation process was improved to 0.31, it was still not sufficient to sustain a good biological treatment. To remedy this deficiency, the leachate effluent was mixed with municipal sewage wastewater in a different ratio before it was subjected to the SBR treatment. The operation process was varied so both nitrification and denitrification could be achieved within the same unit. The SBR reactor was operated under three different operation modes during the reactor phase by changing the aerobic (Ar) and anoxic (An) as shown in Table 2. The removal ratio of NH3-N in mode III indicates the clear advantage of using the process of (An, AR and An) during the reaction phase because of the occurrence of the nitrification– dentrification process in the SBR reactor, while the removal of COD was not significantly affected by the change of operation process in the reaction phase. Since mode III resulted in a high removal ratio for COD and NH3-N, it was selected as the most appropriate operation and was used for the rest of the study. During the nitrification experiment, the pH tended to rise above 8.5 because of the loss of carbonate, which may cause loss of ammonia by volatilisation (Jokela et al., 2002). The feed pH was decreased by the addition of 1 M sulphuric acid (H2SO4) to maintain the pH of the effluent below 8.0. Anoxic processes are typically used for the removal of NH3-N from SBR influent. The process of BNR is known as denitrification. Denitrification requires nitrogen to be first converted to nitrate, which typically occurs in an aerobic treatment process. The nitrified water is then exposed to an environment without free oxygen. Organisms in this anoxic system use the nitrate as an electron acceptor and release nitrogen in the form of nitrogen gas or nitrogen oxides. A readily biodegradable carbon source is also needed if efficient denitrification processes are to occur. By combining several treatment technologies, cost savings and process optimisations could be achieved owing to the degradation of the refractory compounds into biodegradable organic matter and the use of these products as a carbon source for denitrification (Cortez et al., 2011). Fig. 3 shows the investigation of aeration time on COD and NH3N concentrations for different mixing ratios of leachate to sewage in the SBR reactor. As is evident, COD and NH3-N concentration decreased significantly with increase in aeration time up to 8 h. Therefore, the optimum aeration time was 8 h. The effect of dissolved oxygen was also determined; it is important to control the DO concentration by regulating the airflow rate because high concentration requires high energy. On the other hand, a low DO concentration causes filamentous bulking (Spagni et al., 2008). Therefore, the optimal DO concentration was determined so as to achieve the best COD and NH3-N removal efficiency

Fig. 2. Optimum conditions effected on COD and NH3-N removal by PFS coagulation: (a) pH and (b) coagulant dosage. Error bars represent standard deviation.

444

A.R. Abood et al. / Waste Management 34 (2014) 439–447

Table 2 Experimental operating modes of SBR. Phase

Mode I Mode II Mode III

Duration (h)

Total time (h)

Fill static

An

Ar

An

Settle

Draw

0.5 0.5 0.5

– 2 2

12 10 8

– – 2

1 1 1

0.5 0.5 0.5

14 14 14

Removal ratio (%) COD

NH3-N

84.88 85.15 86.25

61.97 68.91 78.77

An: anoxic, Ar: aerobic.

Fig. 3. The effect of aeration time on pollutant removal by SBR reactor: (a) COD and (b) NH3-N. Error bars represent standard deviation.

as shown in Fig. 4. As seen in this figure, the optimal DO concentration was in the range of 4–5 mg L1 and no bulking sludge was observed at this concentration. Increasing the DO level to more than this range may cause biological floc to be eliminated and result in disturbance of the sludge-settling phase and high turbidity in the effluent. 3.5. Filtration process The effluent of the SBR reactor was treated by filtration as the finishing process. The filtration was used as a polishing process to achieve supplemental removal of the suspended solids from the wastewater effluents of biological treatment processes to reduce the mass discharge of solids. A dual material consisting of a layer of activated carbon over a layer of sand was used as the effective material in the filtration process. The specific gravity was 1.6 and 2.65, the effective size 0.7 and 0.45 mm and depth 10 and 15 cm for carbon and sand, respectively. Both carbon and sand layers were placed 5 cm above a gravel layer.

Filtration primarily depends on a combination of complex physical and chemical mechanisms, the most important being adhesion and adsorption. Adsorption is the process of particles sticking onto the surface of the individual filter grains or onto the previously deposited materials. The forces that attract and hold the particles to the grains are the same as those that work in coagulation and flocculation. The removal of pollutants by a sand filter was simulated in Fig. 5. 3.6. Combined processes The overall performances of the combined treatment processes operated under the optimum conditions are listed in Table 3. It can be seen that COD, BOD5, SS, NH3-N and TOC concentrations were 72.4 mg L1, 22.8 mg L1, 24.2 mg L1, 18.4 mg L1 and 50.8 mg L1, respectively. As it is seen from this table the effluent concentration reached the local discharge standard (GB168892008) (COD 6 100 mg L1, BOD5 6 30 mg L1, SS 6 30 mg L1 and NH3-N 6 25 mg L1). The highly colored organic matter was

Fig. 4. The effect of DO concentration on removal of pollutants by SBR reactor: (a) COD and (b) NH3-N.Error bars represent standard deviation.

445

A.R. Abood et al. / Waste Management 34 (2014) 439–447

The removal ratios were in the range of 64.1–99.9% and 81.4– 99.7% for COD and NH3-N, respectively as shown in Table 4. Among the combined treatment processes reviewed above, it is observed that the combination of agitation–coagulation–SBR–filtration demonstrated outstanding treatment performances in the overall result for COD (97.4%) and NH3-N (99.2%). 3.7. Economic analysis A rough economic analysis of the operating costs associated with each treatment process studied, such as the costs of reagents, the costs of reagents required to adjust the pH and energy, was performed. It is important to note that this analysis is just an approximate tool to differentiate the trends in the operating costs associated with the use of combined treatment. A rigorous economic analysis should consider initial investment, prices at plant scale, maintenance and labour costs and so forth. The reagent consumption is listed in Table 5. Table 5 shows that the total reagent cost was $3.83/m3. The average electric power consumption was 20.6 (kWh)/m3, and the average electric price was $0.12/(kW h). The energy cost was $2.47/m3. Therefore the overall operating cost was $6.30/m3, which is considered acceptable for the complete treatment of landfill leachate. The cost of an advanced and complete treatment is always up to $5–7/m3 (Li et al., 2009). This multistage process has proved to be an effective alternative for successful manipulation and management of highstrength wastewater.

Fig. 5. Removal ratios of pollutants by sand filter.

removed along with other organic matters during treatment processes and this result agreed the previous results obtained by Aziz et al. (2007) and Maranon et al. (2008). The UV spectrum area of effluent in 254 nm in the present study decreased obviously from step to step. It is indicated that most of humic substrate was removed during the novel combined process. Hence the overall treatment results achieved by the combined methods are indeed quite good. To evaluate the performances of the combined treatment with respect to other combined treatments, Table 4 shows a comparative study in terms of the initial concentration ranges of COD and NH3-N in leachate. Although it has a relative meaning because of different testing conditions (pH, temperature, wastewater strength, seasonal climate, and hydrology site), this comparison is useful for evaluating the overall treatment performance of each technique to assist the decision–making process.

4. Conclusion The leachate obtained from Chang Shankou landfill (China) is characterized as low biodegradable and with a high content of ammonia, showing that the leachate can be classified old and non-biodegradable. High concentration of NH3-N was eliminated by stripping by means of an agitation process which appears to

Table 3 Concentration and removal ratios of pollutants in effluent for each treatment process.

a

Process

COD

Influent Agitation Coagulation SBR Filtration Overall removal

2800 2400 823.7 125.6 72.4

Concentration (mg L1)

B/Ca

Removal (%)

BOD5

SS

NH3-N

TOC

COD

BOD5

SS

NH3-N

TOC

510 480 258 30 28.5

980 720 60 260 24.2

2400 145.5 128.5 45 18.4

660 520 340 160.5 50.8

– 14.3 65.7 84.8 42.4 97.4

– 5.9 46.3 88.4 5.0 94.4

– 26.5 91.7

– 93.9 11.7 65.0 59.1 99.2

– 21.2 34.6 52.8 68.6 92.3

90.7 97.5

0.18 0.20 0.31 0.24 0.39

B/C = BOD5/COD.

Table 4 Comparison of the current combination processes with other previous combinations for landfill leachate treatment. Treatment process

Agitation–coagulation–SBR–sand filter (current study) Struvite – sequencing batch biofilter granular reactor (SBBGR)- Fenton (Di Iaconi et al., 2006) Coagulation–ammonia stripping–granular activated carbon (GAC) adsorption (Kilic et al., 2007) Coagulation–Fenton oxidation–biological aerated filtering (Wang et al., 2009) Stripping/flocculation/membrane bioreactor/osmosis (Hasar et al., 2009) SBR–coagulation–Fenton–upflow biological aerated filter (UBAFs) (Li et al., 2009) Air stripping–Fenton–SBR–coagulation (Guo et al., 2010) SBR–aeration corrosive cell–Fenton–granular activated carbon (GAC) adsorption (Bu et al., 2010) powdered activated carbon – SBR (Aziz et al., 2011) Solar photo Fenton-biological immobilized biomass reactor (Vilar et al., 2011) SBR–coagulation–Fenton–biological aerated filter (BAF) (Wu et al., 2011) NA: not available

Landfill location

Wuhan (China) Apulia (Italy) Bursa (Turkey) Guangdong (China) Diyarbakir (Turkey) Jiangmen (China) Chongqing (China) Harbin (China) Kulim (Malaysia) Porto (Portugal) Guangdong (China)

Influent (mg L1)

Removal ratio (%)

COD

NH3-N

COD

NH3-N

2800 24400 23700 600–700 8500 3000 4150 41800 1655 4505 6722

2400 3190 1140 NA 1100 1100 1160 2250 600 167 850

97.4 97.0 99.3 88.0 99.9 97.3 93.3 97.2 64.1 95.0 98.4

99.2 99.7 NA NA 97.9 99.0 98.3 NA 81.4 99.4 99.3

446

A.R. Abood et al. / Waste Management 34 (2014) 439–447

Table 5 Chemical reagents and power consumption cost in each unit of combined treatment process. Items

Price ($/ton) Dosage (kg/m3) Operating cost ($/m3) Power (kw hm3) Unit price ($ kW)

Agitation

PFS coagulation

SBR

NaOH

H2SO4

PFS

PAM

Glucose

NaOH

H2SO4

–

200 8 1.6 20.4 2.45

150 6 0.9

400 1.2 0.48 0.0018 0.0002

4500 0.1 0.45

450 0.5 0.225

200 0.5 0.1 0.09 0.01

150 0.5 0.075

– – – 0.08 0.01

be a suitable option for pre-treatment of landfill leachate. In its removal ratio of ammonia reaches to 93.9% at pH 11.5 and G150 s1 within 5 h agitation time. PFS coagulation followed the agitation process and, operated under anoxic–aerobic–anoxic conditions, it can decrease COD with high efficiency of more than 70%, and a biodegradable ratio BOD5/COD modified to 0.31. The final leachate effluent concentrations of COD, BOD5, SS, NH3-N and TOC were 72.4 mg L1, 22.8 mg L1, 24.2 mg L1, 18.4 mg L1 and 50.8 mg L1, respectively, which all of them below the permissible limit (COD 6 100 mg L1, BOD5 6 30 mg L1, SS 6 30 mg L1 and NH3-N 6 25 mg L1) of the Chinese landfill pollution control standard (GB16889-2008). The effluent may directly discharge into waterways without effects on the health of aquatic ecosystems. The combination of agitation-PFS coagulation-SBR-filtration in comparison with other combinations applied for leachate treatment worldwide demonstrated outstanding treatment performances in the overall removal of COD (97.4%) and NH3-N (99.2%). Economic analysis shows that the overall operating cost of the combined treatment was $6.30/m3. This multistage process is useful for landfill leachate treatment plants. Acknowledgements This work was financially supported by the Hubei Provincial Science and Technology Department with grant No. 2006AA305A05. Mr. Alkhafaji R. Abood thanks the China Scholarship Council (CSC) and China University of Geosciences (CUG) for the financial support of this research. References Agdag, O.N., Sponza, D.T., 2008. Sequential anaerobic, aerobic/anoxic treatment of simulated landfill leachate. Environ. Technol. 29, 183–197. Al-Yaqout, A.F., Hamoda, M.F., 2003. Evaluation of landfill leachate in arid climate – a case study. Environ. Int. 29, 593–600. Amokrane, A., Comel, C., Veron, J., 1997. Landfill leachates pretreatment by coagulation-flocculation. Water Res. 31, 2775–2782. APHA, 2005. Standard Methods for the Examination of Water and Wastewater, 21st ed. American Public Health Association, Washington, DC. Aziz, H.A., Alias, S., Adlan, M.N., Asaari, A.H., Zahari, M.S., 2007. Colour removal from landfill leachate by coagulation and flocculation processes. Bioresour. Technol. 98, 218–220. Aziz, S.Q., Aziz, H.A., Yusoff, M.S., Bashir, M.J.K., 2011. Landfill leachate treatment using powdered activated carbon augmented sequencing batch reactor (SBR) process: Optimization by response surface methodology. J. Hazard. Mater. 189, 404–413. Bonmati, A., Flotats, X., 2003. Air stripping of ammonia from pig slurry: characterisation and feasibility as a pre-or post-treatment to mesophilic anaerobic digestion. Waste Manage. 23, 261–272. Bu, L., Wang, K., Zhao, Q.L., Wei, L.L., Zhang, J., Yang, J.C., 2010. Characterization of dissolved organic matter during landfill leachate treatment by sequencing batch reactor, aeration corrosive cell-Fenton, and granular activated carbon in series. J. Hazard. Mater. 179, 1096–1105. Calace, N., Liberatori, A., Petronio, B.M., Pietroletti, M., 2001. Characteristics of different molecular weight fractions of organic matter in landfill leachate and their role in soil sorption of heavy metals. Environ. Pollut. 113, 331–339. Cassano, D., Zapata, A., Brunetti, G., Del Moro, G., Di Iaconi, C., Oller, I., Malato, S., Mascolo, G., 2011. Comparison of several combined/integrated biological-AOPs setups for the treatment of municipal landfill leachate: minimization of operating costs and effluent toxicity. Chem. Eng. J. 172, 250–257. Cheng, R.C., Liang, S., Wang, H.C., Beuhler, M.D., 1994. Enhanced coagulation for arsenic removal. J. Am. Water Works Assoc. 86, 79–90.

Filtration

Cortez, S., Teixeira, P., Oliveira, R., Mota, M., 2011. Mature landfill leachate treatment by denitrification and ozonation. Process Biochem. 46, 148–153. Di Iaconi, C., Ramadori, R., Lopez, A., 2006. Combined biological and chemical degradation for treating a mature municipal landfill leachate. Biochem. Eng. J., 31, 118–124. Fang, F., Abbas, A.A., Chen, Y.P., Liu, Z.P., Gao, X., Guo, J.S., 2011. Anaerobic/aerobic/ coagulation treatment of leachate from a municipal solid wastes incineration plant. Environ. Technol. 33, 927–935. Galeano, L.A., Vicente, M.A., Gil, A., 2011. Treatment of municipal leachate of landfill by Fenton-like heterogeneous catalytic wet peroxide oxidation using an Al/Fepillared montmorillonite as active catalyst. Chem. Eng. J. 178, 146–153. Gulsen, H., Turan, M., 2004. Anaerobic treatability of sanitary landfill leachate in a fluidized bed reactor. Turkish J. Eng. Environ. Sci. 28, 297–305. Guo, J.S., Abbas, A.A., Chen, Y.P., Liu, Z.P., Fang, F., Chen, P., 2010. Treatment of landfill leachate using a combined stripping, Fenton, SBR, and coagulation process. J. Hazard. Mater. 178, 699–705. Gustin, S., Marinsek-Logar, R., 2011. Effect of pH, temperature and air flow rate on the continuous ammonia stripping of the anaerobic digestion effluent. Process Saf. Environ. Prot. 89, 61–66. Hasar, H., Unsal, S.A., Ipek, U., Karatas, S., Cinar, O., Yaman, C., Kinace, C., 2009. Stripping/flocculation/membrane bioreactor/reverse osmosis treatment of municipal landfill leachate. J. hazard. mater. 171, 309–317. He, P., Li, M., Xu, S., Shao, L., 2009. Anaerobic treatment of fresh leachate from a municipal solid waste incinerator by upflow blanket filter reactor. Front. Environ. Sci. Eng. China 3, 404–411. Jokela, J.P.Y., Kettunen, R.H., Sormunen, K.M., Rintala, J.A., 2002. Biological nitrogen removal from municipal landfill leachate: low-cost nitrification in biofilters and laboratory scale in-situ denitrification. Water Res. 36, 4079–4087. Kang, K.H., Shin, H.S., Park, H., 2002. Characterization of humic substances present in landfill leachates with different landfill ages and its implications. Water Res. 36, 4023–4032. Kheradmand, S., Karimi-Jashni, A., Sartaj, M., 2010. Treatment of municipal landfill leachate using a combined anaerobic digester and activated sludge system. Waste Manage. 30, 1025–1031. Kilic, M.Y., Kestioglu, K., Yonar, T., 2007. Landfill leachate treatment by the combination of physicochemical methods with adsorption process. J. Biol. Environ. Sci. 1, 37–43. Kurniawan, T.A., Lo, W., Chan, G.Y.S., 2006a. Radicals-catalyzed oxidation reactions for degradation of recalcitrant compounds from landfill leachate. Chem. Eng. J. 125, 35–57. Kurniawan, T.A., Lo, W.H., Chan, G., 2006b. Degradation of recalcitrant compounds from stabilized landfill leachate using a combination of ozone-GAC adsorption treatment. J. Hazard. Mater. 137, 443–455. Laitinen, N., Luonsi, A., Vilen, J., 2006. Landfill leachate treatment with sequencing batch reactor and membrane bioreactor. Desalination 191, 86–91. Li, X.Z., Zhao, Q.L., 2001. Efficiency of biological treatment affected by high strength of ammonium–nitrogen in leachate and chemical precipitation of ammonium– nitrogen as pretreatment. Chemosphere 44, 37–43. Li, H.-S., Zhou, S.-Q., Sun, Y.-B., Feng, P., Li, J.-D., 2009. Advanced treatment of landfill leachate by a new combination process in a full-scale plant. J. Hazard. Mater. 172, 408–415. Maranon, E., Castrillonn, L., Fernandez-Nava, Y., Fernandez-Mendez, A., FernanndezSanchez, A., 2008. Coagulation–flocculation as a pretreatment process at a landfill leachate nitrification–denitrification plant. J. Hazard. Mater. 156, 538– 544. Marttinen, S.K., Kettunen, R.H., Sormunen, K.M., Soimasuo, R.M., Rintala, J.A., 2002. Screening of physical–chemical methods for removal of organic material, nitrogen and toxicity from low strength landfill leachates. Chemosphere 46, 851–858. Monje-Ramirez, I., Velasquez, M.T., 2004. Removal and transformation of recalcitrant organic matter from stabilized saline landfill leachates by coagulation–ozonation coupling processes. Water Res. 38, 2359–2367. Renou, S., Givaudan, J.G., Poulain, S., Dirassouyan, F., Moulin, P., 2008. Landfill leachate treatment: review and opportunity. J. Hazard. Mater. 150, 468– 493. Shi, D., Tang, X., Wu, W., Fang, J., Shen, C., McBride, M.B., Chen, Y., 2009. Effect of MSW source-classified collection on polycyclic aromatic hydrocarbons in residues from full-scale incineration in China. Water Air Soil Pollut. 198, 347– 358. Spagni, A., Marsili-Libelli, S., 2009. Nitrogen removal via nitrite in a sequencing batch reactor treating sanitary landfill leachate. Bioresour. Technol. 100, 609– 614.

A.R. Abood et al. / Waste Management 34 (2014) 439–447 Spagni, A., Marsili-Libelli, S., Lavagnolo, M.C., 2008. Optimisation of sanitary landfill leachate treatment in a sequencing batch reactor. Water Sci. Technol. 58, 337– 343. Stephenson, R.J., Duff, S.J.B., 1996. Coagulation and precipitation of a mechanical pulping effluent – I. removal of carbon, colour and turbidity. Water Res. 30, 781–792. Tatsi, A.A., Zouboulis, A.I., Matis, K.A., Samaras, P., 2003. Coagulation–flocculation pretreatment of sanitary landfill leachates. Chemosphere 53, 737–744. Timur, H., Ozturk, I., 1999. Anaerobic sequencing batch reactor treatment of landfill leachate. Water Res. 33, 3225–3230. Uygur, A., Kargi, F., 2004. Biological nutrient removal from pre-treated landfill leachate in a sequencing batch reactor. J. Environ. Manage. 71, 9–14. Vilar, V.J.P., Capelo, S., Silva, T.F.C.V., Boaventura, R.A.R., 2011. Solar photo-Fenton as a pre-oxidation step for biological treatment of landfill leachate in a pilot plant with CPCs. Catal. Today 161, 228–234.

447

Wang, X., Chen, S., Gu, X., Wang, K., 2009. Pilot study on the advanced treatment of landfill leachate using a combined coagulation, fenton oxidation and biological aerated filter process. Waste Manage. 29, 1354–1358. Wu, Y., Zhou, S., Ye, X., Chen, D., Zheng, K., Qin, F., 2011. Transformation of pollutants in landfill leachate treated by a combined sequence batch reactor, coagulation, Fenton oxidation and biological aerated filter technology. Process Saf. Environ. Prot. 89, 112–120. Yanjie Wei, M.J., Li, Ruying., Qin, Feifei., 2012. Organic and nitrogen removal from landfill leachate in aerobic granular sludge sequencing batch reactors. Waste Manage. 32, 448–455. Ye, J., Mu, Y., Cheng, X., Sun, D., 2011. Treatment of fresh leachate with highstrength organics and calcium from municipal solid waste incineration plant using UASB reactor. Bioresour. Technol. 102, 5498–5503. Yu, J., Zhou, S., Wang, W., 2010. Combined treatment of domestic wastewater with landfill leachate by using A2/O process. J. Hazard. Mater. 178, 81–88.