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Anaerobic treatment has been demonstrated to be the most useful technology for the treatment of high-strength organic wastewaters. The processes perform well at high organic loading rates (OLR) with low operating costs and also produce usable biogas (Zhou et al. 2007). Contrary to aerobic processes, anaerobic digestion conserves energy and produces very few solids, but suffers from low reaction rates (Sung et al. 1997).
Under anaerobic conditions (absence of oxygen or nitrate), organic substances are converted to methane and carbon dioxide (biogas) as well as water and small fraction of new biomass. The non injection of oxygen in anaerobic systems greatly reduces their cost. Anaerobic leachate treatment is an effective process but the remaining BOD5 and COD effluent concentrations are still high with BOD5/COD ratios >0.3. After the anaerobic treatment step the leachate has to be treated to final effluent standards by means of aerobic processes (Stegmann et al. 2005). Anerobic treatment process also classified into suspended-growth biomass processes and attached-growth biomass processes .The main problems for anaerobic digestion of leachate can be summarized as follow:
â€¢ The process does not remove ammoniacal-N at all, and indeed is more likely to increase concentrations of this main contaminant of landfill leachates. Secondary aerobic biological and other processes will generally be essential;
â€¢ A COD value in raw leachate in excess of about 10,000 mg/l is essential if the anaerobic treatment process is to be self-sufficient in energy. At most modern landfills, the acetogenic phase where this is the case for leachates is relatively short-lived;
â€¢ The anaerobic processes being used are far more efficiently provided within the landfill body itself, where optimum and stable temperatures are likely to be present. Recirculation of acetogenic leachates in a controlled way may well enable this to be carried out successfully, with resulting landfill gas collected by the existing systems.
i) Anaerobic sequencing batch reactors (ASBR)
Anaerobic sequencing batch reactor (ASBR) process is a batch-fed, batch-decanted, suspended growth. It can achieve solids capture and removal of organics in one vessel, eliminating the need for a clarifier. The batch operation allows good effluent quality control since the reactor draw can be made just when the compliance with legal standard has been attained. Operation of the ASBR reactor consists of four steps: feeding, reaction, settling and treated effluent withdrawal fig. 2.6. The main factors affecting the overall performance of the ASBR are: agitation, Substrate/Biomass ratio, geometric configuration of the reactor and the feeding strategy. The main advantages of this type of operation are its operational simplicity, efficient quality control of the effluent, possibility of eliminating the settling step for both the influent and effluent wastewater and flexibility of use in the wide variety of wastewaters to be treated. These characteristics indicate its potential application in situations requiring compliance with strict environmental control standards as well as when sewage is produced intermittently and has variable characteristics as a result of the type of downstream process.
Timur and Ozturk (Timur and Ozturk 1997) treated landfill leachate taken from a young municipal site containing high organic pollutants. They performed the treatment in a bench-scale ASBR and an anaerobic hybrid bed filter (AHBF) at mesophilic conditions. The ASBR achieved 73.9% TOC removal at a maximum organic loading rate of 2.8 kg of TOC m³/day for a hydraulic retention time (HRT) of 1.5 days. They concluded that, considering the flexibility of the ASBR in adapting to variations in volume and effluent composition and the high cost of AHBF systems, ASBRs are better for the treatment of young landfill leachates. In a later study, these authors evaluated the anaerobic treat ability of municipal landfill leachate using a laboratory-scale ASBR at 35°C. Experimental studies were conducted for a wide range of volumetric (0.4-9.4 g of COD L/day) and specific (0.2-1.9 g of COD/g of VSS/day) loading rates by varying the HRT from 10 to 1.5 days. Removals of COD were between 64 and 85%, depending on the applied rates (Timur and Ozturk 1999) .
Fig. 2.6 Anaerobic Sequencing Batch Reactor Process Stages
ii) Up flow Anaerobic Sludge Blanket (UASB) reactor
Up-flow anaerobic sludge blanket (UASB) technology is a form of anaerobic digester that is used in leachate treatment and for treatment of many other types of wastewater. It is a modern anaerobic treatment that can have high treatment efficiency and a short hydraulic retention time (Lin et al. 2000).Under proper conditions anaerobic sludge will develop as high density granules. These will form a sludge blanket in the reactor as shown in Fig 2.7a. The process involves an upward passage of leachate through an anaerobic sludge bed in a tank. As the leachate passes through the sludge, microorganisms in the sludge degrade organic matter in the leachate producing biogas (methane and carbon dioxide).
The process temperatures reported have generally been 20-35°C for anaerobic treatment with UASB reactors. In these conditions, the average performance of COD decrease efficiency was always higher than 70% at ambient temperature (20-23°C) and 80% at 35°C. Up to 92% COD decreases were obtained by Kennedy and Lentz at low and intermediate organic loading rates (between 6 and 19.7 g/L/day of COD) (Kennedy and Lentz 2000). Kettunen and Rintala (Kettunen and Rintala 1998) showed that leachate can be treated on-site UASB reactor at low temperature. A pilot-scale reactor was used to study municipal landfill leachate treatment (COD 1.5-3.2 g/L) at 13-23°C. COD (65-75%) and BOD7 (up to 95%) removals were achieved at organic loading rates of 2-4 g/L/day of COD.
Jiexu Ye (Ye et al. 2011) investigated the treatment of a fresh leachate with high-strength organics and calcium from municipal solid waste (MSW) incineration plant by an UASB reactor under mesophilic conditions, emphasizing the influence of organic loading rate (OLR). An average COD removal efficiency of 82.4% was achieved when the reactor was fed with raw leachate (COD as high as 70,390-75,480 mg/L) at OLR of 12.5 kg COD/(m3·d). The ratio of volatile solids/total solids (VS/TS) of the anaerobic sludge in the UASB decreased significantly after a long-term operation due to the precipitation of calcium carbonate in the granules.
iii) Expended Granule Sludge Blanket (EGSB)
The EGSB reactor is the family of UASB reactor with a high recycle ratio. The upflow of this reactor is typically maintained higher than 6 m/hr; meanwhile the general range of the UASB reactor is 0.5 to 1.0 m/hr. The height to width of EGSB is 4 ~ 5 so that it enables the EGSB reactor to contact granules with wastewater enough. Additionally, due to the high velocity, granules are expended and the hydraulic mixing is intensified as to also give granules more chances to contact with wastewater. Thus, this reactor is able to treat high-strength organic wastewater (up to loading rate about 30 kg/m3âˆ™d). The definitive feature of EGSB reactor is the rapid up flow velocity. It enables this reactor to separate dispersed sludge from mature granules in the reactor. It makes a lot of contacts between granules and wastewater and withdraws suspended sludge out of the reactor (Kato et al. 2004).
The increased flux permits partial expansion (fluidization) of the granular sludge bed, improving wastewater sludge contact as well as enhancing segregation of small inactive suspended particle from the sludge bed. The increased flow velocity is either accomplished by utilizing tall reactors, or by incorporating an effluent recycle (or both) (Chu et al. 2005). A scheme depicting the EGSB design concept is shown in fig 2.6b.
The EGSB reactor was also developed to give more chances to contact between wastewater and granules. Besides, this reactor is able to separate dispersed sludge from mature granule using rapid upward velocity. Then, it is possible to treat high-strength and low-strength wastewater such as domestic wastewater, especially low temperature. In order to achieve good process performance in EGSB systems, it is necessary to maintain a high concentration of methanogenic bacteria in the biofilm (Yoochatchaval et al. 2008).
iv) Fluidized Bed Reactor (FBR)
This reactor consists of a sand bed on which the biomass is grown (Fig. 2.6c). Since the sand particles are small, a very large biomass can be developed in a small volume of reactor. In order to fluidize the bed, a high recycle is required.
Imai et al. reported studies on carbon-assisted fluidized beds. The combined biodegradation and adsorption process provide a means for removing a variety of organic compounds. They found that the biological activated carbon fluidized bed process was much more effective for treating old landfill leachate than the conventional one such as activated sludge and fixed film processes (Imai et al. 1998; Imai et al. 1993; Imai et al. 1995).
Gulsen and Turan investigated the treatment of young Landfill in a pilot scale fluidized bed reactor having an inner diameter of 10 cm, a height of 165 cm and an effective volume of 13 L. The reactor medium was typical filter sand having an arithmetic mean diameter of 0.5 mm and a fixed bed height of 70 cm. COD removal increased from 80% to 90 % with increasing organic loading rates and the anaerobic fluidized bed reactor attained steady state conditions with a COD removal of 90% after 80 days. In addition, the COD removal decreased to 82% at an OLR of 37 kg COD/ m3 per day (Gulsen and Turan 2004).
(a) UASB reactor
b) EGSB reactor
(c)Fluidized bed reactor
Fig 2.6 Concept of some anaerobic biological reactors :a) UASB, b) EGSB and c) Fluidized bed reactor.
184.108.40.206 Anaerobic-aerobic (A/O) system
Biodegradable organic matter is stabilized by a combination of aerobic and anaerobic processes. In general, aerobic systems are suitable for the treatment of low strength wastewaters (biodegradable COD concentrations less than 1000 mg/L) while anaerobic systems are suitable for the treatment of high strength wastewaters (biodegradable COD concentrations over 4000 mg/L) (Chan et al. 2009; Agdag and Sponza 2008). The use of anaerobic-aerobic processes can also lead to a factor eight cost reduction in operating costs when compared with aerobic treatment alone (Vera et al. 1999) , while simultaneously resulting in high organic matter removal efficiency, a smaller amount of aerobic sludge and no pH correction. The benefits of the anaerobic-aerobic process have been identified by Chan et al. 2009 (Chan et al. 2009) and Cervantes et al. (Cervantes et al. 2006) are listed below:
Great potential of resource recovery: Anaerobic pretreatment removes most of the organic pollutants and converts them into a useful fuel, biogas.
High overall treatment efficiency: Aerobic post-treatment polishes the anaerobic effluent and results in very high overall treatment efficiency. The aerobic treatment also smoothes out fluctuations in the quality of the anaerobic effluent.
Less disposal of sludge: By digesting excess aerobic sludge in the anaerobic tank, a minimum stabilized total sludge is produced which leads to a reduction in sludge disposal cost. As an additional
Benefit, a higher gas yield is achieved.
Low energy consumption: anaerobic pretreatment acts as an influent equalization tank, reducing diurnal variations of the oxygen demand and resulting in a further reduction of the required maximum aeration capacity.
When volatile organics are present in thewastewater, the volatile compound is degraded in the anaerobic treatment, removing the possibility of volatilization in the aerobic treatment.
Thus it can be seen that it is operationally and economically advantageous to adopt anaerobic-aerobic processes in the treatment of high strength industrial wastewaters since it couples the benefit of anaerobic digestion (i.e. biogas production) with the benefits of aerobic digestion (i.e. better COD and volatile suspended solid (VSS) removal)(Ros and Zupancic 2004) . As well as their capability to biodegrade organic matter, anaerobic-aerobic systems have also been found to perform well for sequential nitrogen removal including aerobic nitrification and anaerobic denitrification (Liu et al. 2008);
Anaerobic-aerobic (A/O) systems receive great attention over the past decades due to their numerous advantages, not only from municipal wastewater, but also from MSW leachates (Agdag and Sponza 2008; Yang and Zhou 2008). A more intensive form of biodegradation can also be achieved by integrating anaerobic and aerobic zones within a single bioreactor. Essentially, there are four types of integrated anaerobic-aerobic bioreactor. These are (i) integrated bioreactors with physical separation of anaerobic-aerobic zone, (ii) integrated bioreactor s without physical separation of anaerobic-aerobic zone, (iii) Sequencing Batch Reactors (SBR) based on temporal separation of the anaerobic and the aerobic phase, and (iv) combined anaerobic-aerobic culture systembased on the principle of limited oxygen diffusion in microbial biofilms (Chan et al. 2009).
The applicability of the intermittent flow activated sludge system, i.e. sequencing batch reactor (SBR) system for post treatment of anaerobic effluents has also been investigated by many researchers (Torres and Foresti 2001; Guimaraes et al. 2003). Two sequential SBRs forming an A/O system were evaluated for treatment of domestic sewage (Callado and Foresti 2001).
Sequential anaerobic-anoxic-aerobic operations in a lab-scale sequencing batch reactor to treat landfill leachate resulted in COD, NH3-N and PO43--P removal of 62%, 31% and 19%, respectively, at the end of cycle time (21 hr) (Uygur and Kargi 2004).
Zaloum and Abbott (Zaloum and Abbott 1997) studied the possibility of using the SBR after an anaerobic pretreatment of raw leachate, with a sludge retention time (SRT) of 50 days and an HRT of 3.2 days. They concluded that the SBR treatment of anaerobic lagoon effluent was the best option. With this treatment, the BOD5, COD, and nutrient residual concentrations were within the stricter new proposed regulations in Quebec, Canada.
Kennedy and Lenz investigated and compared the treatment of municipal landfill leachate using sequencing batch and continuous flow UASB reactors. The sequencing batch UASB reactor had soluble COD removal efficiencies ranging between 71% and 92% at hydraulic retention times (HRT) of 24, 18 and 12 h with dilute to concentrated leachate feed (Kennedy and Lentz 2000).
Yalmaz and Ozturk (2001) (Yalmaz and Ozturk 2001) conducted an investigation on the use of SBR technology for the treatment of high ammonia landfill leachate via nitrification-denitrification and anaerobic pre-treatment. The SBR was further tested for the treatment of anaerobically pre-treated leachate from an up-flow anaerobic sludge blanket reactor (UASB). The SBR achieved a 90 % nitrogen removal when anaerobically pretreated leachate was treated while using Ca (CH3COO)2 as a carbon source. The study revealed that young landfill leachate with a COD/NH4-N greater than 10 was also effective as a carbon source for denitrification.
A sequential upflow anaerobic sludge blanket (UASB) and air-lift loop sludge blanket (ALSB) treatment was introduced into leachate recirculation to remove organic matter and ammonia from leachate in a lab-scale bioreactor landfill by He et al (He et al. 2007). They showed that the sequential anaerobic-aerobic process might remove above 90% of COD and near to 100% of NHâ€4 -N from leachate under the optimum organic loading rate. The total COD removal efficiency was over 98% as the OLR increased to 6.8-7.7 g/l d, but the effluent COD concentration increased to 2.9-4.8 g/l in the UASB reactor, which inhibited the activity of nitrifying bacteria in the subsequent ALSB reactor.
A system consisting of a two-stage up-flow anaerobic sludge blanket (UASB), an anoxic/aerobic (A/O) reactor and a sequencing batch reactor (SBR), was used to treat landfill leachate by Wu et al (Wu et al. 2009). During operation, denitrification and methanogenesis took place simultaneously in the first stage UASB, and the effluent chemical oxygen demand (COD) was further removed in the second stage UASB. Then the denitrification of nitrite and nitrate in the returned sludge by using the residual COD was accomplished in the A/O reactor, and ammonia was removed via nitrite in it.
The results showed that when the total nitrogen concentration of influent leachate was about 2500 mg/L and the ammonia nitrogen concentration was about 2000 mg/L, the shortcut nitrification with 85%-90% nitrite accumulation was achieved stably in the A/O reactor. The TN and ammonia nitrogen removal efficiencies of the system were 98% and 97%, respectively. The residual ammonia, nitrite and nitrate produced during nitrification in the A/O reactor could be washed out almost completely in SBR.
(Im et al. 2001) Im et al introduced an anaerobic-aerobic system including simultaneous methanogenesis and denitrification to treat organic and nitrogen compounds in immature leachate from a landfill site. Denitrification and methanogenesis were successfully carried out in the anaerobic reactor while the organic removal and nitrification of NHâ€4-N were carried out in the aerobic reactor when rich organic substrate was supplied with appropriate hydraulic retention time. The maximum organic removal rate was 15.2 kg COD/m3 d in the anaerobic reactor while the maximum NHâ€4-N removal rate and maximum nitrification rate were 0.84 kg NHâ€4-N/m3/d and 0.50 kg NO_3 -N/m3/d, respectively, in the aerobic reactor.
A set of anaerobic-anoxic-aerobic (A2/O) bioreactor system was designed and used to treat domestic wastewater mixed with landfill leachate in Datansha Sewage Treatment Plant in Guangzhou, south China by Yu et al (Yu et al. 2010). The results showed that the optimal volume ratio of landfill leachate and domestic wastewater in the A2/O process was 1:500. The average removal efficiencies of NH4+-N, TN and COD was achieved to be 96.5%, 61.0% and 81.7%, respectively in the case of the hydraulic retention time of 11 h, dissolved oxygen of 3mg Lâˆ’1, the mixed-liquid return ratio of 200% and sludge return ratio of 80% in the case of the confirmatory experiment. The pilot scale (3.8m3) investigation results were applied in the large-scale (220,000m3/d) combined treatment of sewage wastewater with landfill leachate in Guangzhou Datansha Domestic Sewage Wastewater Treatment Plant. The removal efficiencies of COD, NH4+-N, T-N and T-P were 82.65%, 92.69%, 57.10% and 76.55%, respectively.
3.4.2 Biological Nitrogen Removal
The concentrations of organic material and ammonia nitrogen are high in fresh leachate, while matured leachate contains relatively lower concentration of organic matter but higher concentration of ammonia nitrogen (Zhang et al. 2007). High concentration ammonia nitrogen is considered as the main reason for low efficiency in biological treatment of landfill leachate (Uygur and Kargi 2004). Especially, with the release of new national standard for pollution control on the landfill site of municipal solid waste, the removal of total nitrogen (TN) and ammonia nitrogen from leachate becomes critical.
Nitrogen in wastewater occurs in different forms mainly organic nitrogen and ammonia. In wastewater treatment plants, nitrogen compounds undergo many changes and transformation including ammonofication, nitrification, denitrification, and assimilation processes (Danesh 1997). A summary of these transformations are presented in Fig. 3.7.
Biological nitrogen removal normally involves two separate steps, aerobic nitrification of ammonia to nitrate and anoxic denitrification of nitrate to nitrogen gas. Denitrification occurs in the absence of oxygen, where nitrate or nitrite is the electron acceptor and requires a carbon source as electron donor (Hsieh et al. 2003).
Nitrification is biological oxidation of ammonia to nitrate. The process consists of two sequential steps; each is carried out by certain groups of microorganisms, which collectively referred to as nitrifiers. The first step of nitrification is called nitritification and involves oxidation of ammonia to nitrite. The members of the genera Nitrosomonas and Nitrosococcus are responsible for this step(Metcalf and Eddy 2003). Nitritification is followed by nitratification; a process in which nitrite is converted to nitrate by the members of genera Nitrobacter and Nitrocystis. Approximate equations for these transformations are as follows (Metcalf and Eddy 2003)
NH4+ + 2O2 â†’ NO3- + 2H+ + H2O
2NH4+ + 3O2 â†’ 2NO2- + 4H+ + 2H2O, Nitrosomonas
Second step: oxidation of nitrite to nitrate
2NO2- + O2 â†’ 2NO3-, Nitrobacter
Conversion of ammonia to nitrite is slower than the oxidation of nitrite to nitrate. Thus the overall nitrification is usually limited by the performance of Nitrosomonas (Metcalf and Eddy 2003) .
Nitrosomonas performs the first step by oxidizing ammonium to nitrite. Nitrobacter completes the oxidation by converting the nitrite to nitrate. Since complete nitrification is a sequential reaction, treatment processes must be designed to produce an environment suitable for growth and survival of both groups of nitrifying bacteria .
Nitrogen gas (N2)
Fig. 45 Nitrogen transformations in biological treatment units (Danesh 1997) .
Environmental Requirements for Nitrification
The most common, practical, and economical way to remove ammonia from a waste stream is to utilize nitrifying bacteria which are naturally present in the soil, freshwater, and saltwater. Nitrifiers are difficult to maintain because of their specific environmental requirements. The important environmental parameters that must be maintained for optimal performance of the nitrifiers include the correct pH range, a minimum dissolved oxygen concentration, the necessary temperature range, presence of ammonia, supply of micronutrients, and suitable hydraulic retention time (Metcalf and Eddy 2003). Also for nitrification to occur, high organic concentrations (COD) and inhibitors, such as metals and specific organics, must be removed. The pH of the liquid must be kept in the range between 7.0-8.8, with the optimum nitrification rate being around 8.5. Liquid temperature should be maintained between 20-35 C for good activity. Adequate aeration should keep the dissolved oxygen concentration at a minimum of 2 mg/l. however; oxygen concentrations below 2.0 mg/l begin to have a strong effect (Wiesmann et al. 2006). Chemical Oxygen Demand (COD) must be at levels that do not use all the available oxygen or create inhibitory conditions. COD must be removed because of competition between the heterotrophic and autotrophic bacteria.
Denitrification is the biochemical of many organic substrates in wastewater treatment, using nitrate or nitrite as the electron acceptor instead of oxygen. The nitrate reduction reactions involve the following reduction steps from nitrate to nitrite, to nitric oxide, to nitrous oxide, and finally to nitrogen gas as shown in eq. 67 (Marta 2011):
The last three compounds are gaseous products that are dissimilate from the system and released to the atrnosphere. The final gaseous product in a well performed denianfication is nitrogen gas with no detectable amount of NO and NO2 in the atmosphere.
The denitrification process releases alkalinity in the media, which increases the pH. The process is accomplished biologically under "anoxic" conditions. The denitrifying microorganisms; denitrifiers; are ubiquitous in nature as well as in the activated sludge systems. They are facultative anaerobes which use organic carbon as the source of energy and carbon, and oxidized forms of nitrogen (NO) as the alternative electron acceptors to oxygen, and are most effective in the absence of oxygen (Metcalf and Eddy 2003). Kinetics of denitrification reactions are affected by many parameters mainly the presence of oxygen, type and concentration of carbon source, concentration of nitrate and the pH and temperature of the reaction's' environment (Danesh 1997).
(Vilar et al. 2011) investigated the complete removal of ammonium, nitrates and nitrites of the photo-pre-treated leachate was achieved by biological denitrification and nitrification , after previous neutralization/sedimentation of iron sludge (40 mL of iron sludge per liter of photo-treated leachate after 3 h of sedimentation). The optimum C/N ratio obtained for the denitrification reaction was 2.8 mg CH3OH per mg N-NO3, consuming 7.9 g/8.2 mL of commercial methanol per liter of leachate. The maximum nitrification rate obtained was 68 mg N-NH4 â€ per day, consuming 33 mmol (1.3 g) of NaOH per liter during nitrification and 27.5 mmol of H2SO4 per liter during denitrification.
(Wei et al. 2012) established Granule sequencing batch reactors (GSBR) for landfill leachate treatment. The characteristics of nitrogen removal at different influent ammonium levels were studied. When the ammonium concentration in the landfill leachate was 366 mg L_1, the dominant nitrogen removal process in the GSBR was simultaneous nitrification and denitrification (SND). Under the ammonium concentration of 788 mg L_1, nitrite accumulation occurred and the accumulated nitrite was reduced to nitrogen gas by the shortcut denitrification process. When the influent ammonium increased to a higher level of 1105 mg L_1, accumulation of nitrite and nitrate lasted in the whole cycle, and the removal efficiencies of total nitrogen and ammonium decreased to only 35.0% and 39.3%, respectively. Results also showed that DO was a useful process controlling parameter for the organics and nitrogen removal at low ammonium input.
Recently, laboratory studies have showed the efficacy of in situ nitrogen removal in solid waste environment. (Youcai et al. 2002) reported that 99.5% of the leachate ammonia was removed in a biofilter consisting of old waste (8-10 years old) with both anaerobic and aerobic sections. (Onay and Pohland 1998)Onay and Pohland (1998) developed a three-component simulated landfill system, including anoxic, anaerobic and aerobic zones, to demonstrate the feasibility of in situ nitrification and denitrification at controlled landfills operated with leachate recirculation. The results demonstrated that both separate and combined reactor operations with international leachate recycling around each reactor provided 95% nitrogen conversion. In contrast, combined reactor operation, without internal recycling had conversion efficiency per cycle ranging from 30% to 52% for nitrification and from 16% to 25% for denitrification.