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Domestic wastewater treatment using immobilized sludge process An, Muhui 2000

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Domestic Wastewater Treatment Using Immobilized Sludge Process By Muhui An B.Sc. Sichuan University, 1983 A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF T H E REQUIREMENTS FOR T H E DEGREE OF MASTER OF SCIENCE In T H E F A C U L T Y OF G R A D U A T E STUDIES DEPARTMENT OF CHEMICAL AND BIO-RESOURCES ENGINEERING We accept this thesis as conforming to the required standard T H E UNIVERSITY OF BRITISH COLUMBIA October, 1999 © Muhui An, 1999 In presenting this thesis in partial fulfillment of the requirements for an advanced degree at the University of Bri t ish Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensively copying o f this thesis for scholarly purposes may be granted by the head of my department or by his or her representatives. It is understood that copying or publication o f this thesis for financial gain shall not be allowed without my written permission. Department of Chemical and Bio-resources Department The University of Brit ish Columbia 2075 Wesbrook Place Vancouver, Canada V 6 T 1Z1 Date: A B S T R A C T This study is an initial approach for developing a small domestic wastewater treatment process using immobilized sludge. The feasibility o f sludge immobilization by the P V A - b o r i c acid method was examined. The activity and stability o f the immobilized sludge were measured. The potential of the immobilized sludge process for small scale domestic wastewater treatment was explored. In this research, the PVA-alginate-borate method was successfully used to immobilize activated sludge. The optimum final polyvinyl alcohol ( P V A ) concentration in the immobilized sludge was determined to be 10 ~ 12.5%. A minimal 1% o f alginate was needed to prevent bead agglomeration. The effects of p H on the organic carbon removal activity by immobilized sludge were examined. A wastewater p H o f 6 ~ 8 was the most favorable for application of the immobilized sludge process. The operational stability of the immobilized sludge system was also studied in a batch reactor. The effects of aeration rates and hydraulic retention time (HRT) on the treatment of domestic wastewater using the immobilized sludge process were investigated in a fluidized bed reactor. The results showed that variation of aeration rates (0.5 ~ 1.5L air/min.) and H R T s (24 ~ 6 hours) did not have any significantly adverse impacts on the removal o f organic carbon, N F l V - N and T K N nor on producing a high quality o f effluent. Different intermittent aeration patterns were investigated to improve total nitrogen (TN) removal. The best T N removal was above 74.4% and was achieved at an H R T of 6 hour (corresponding B O D loading rate 0.766 kg/m3 d) and aeration pattern o f 1:3 (the ratio of non-aeration time and aeration time). The removal efficiencies for organic ii carbon, NH4 - N , T K N and TSS were not significantly affected. Simultaneous organic carbon and nitrogen removal could be realized in a single immobilized sludge reactor. The impact of various influent flow patterns on small scale domestic wastewater treatment was examined. The results showed that variation o f wastewater feeding patterns ranging from 24-hour continuous feed to intermittent feed (12 and 8 hours per day) did not have any significantly adverse impacts on the removal of organic carbon, N f l t + - N , T K N and TSS. However, a significant decrease in total nitrogen removal occurred. The production of excess sludge by the immobilized sludge process was experimentally demonstrated to be about 1/2 ~ 1/3 that produced by the conventional activated sludge process. The results of scanning electron microscopic analysis showed that bacteria mostly grew on the surface of immobilized sludge beads. The process performance of the immobilized sludge process in this research was as good as or better than that of other biological treatment processes for the domestic wastewater treatment. The immobilized sludge beads exhibited satisfactory mechanical stability without apparent breakage over the 180-day experiment. Thus the immobilized sludge process was demonstrated to be technically feasible for treatment o f domestic wastewater on a small scale. 111 T A B L E O F C O N T E N T S A B S T R A C T » T A B L E O F C O N T E N T S iv L I S T O F T A B L E S v i i L I S T O F F I G U R E S v i i i L I S T O F A B B R E V I A T I O N S x A C K N O W L E D G E M E N T S x i 1. I N T R O D U C T I O N . 1 2. L I T E R A T U R E R E V I E W 5 2.1. Small domestic wastewater treatment 5 2.1.1. Small domestic wastewater flowrate and strength 6 2.1.2. Small wastewater treatment systems 7 2.1.2.1. Septic tank 7 2.1.2.2. Conventional aerobic treatment systems 9 2.1.2.3. Other treatment systems 10 2.2. Immobilization of cells 12 2.2.1. Immobilization methods 13 2.2.2. Immobilization materials 13 2.2.2.1. Alginate. . . . . 14 2.2.2.2. Carrageerian 15 2.2.2.3. Cellulose 16 2.2.2.4. Polyacrylamide 17 2.2.2.5. Polyurethane 18 iv 2.2.2.6. Polyvinyl alcohol 19 2.2.3. Applications in wastewater treatment 21 2.2.3.1. Nitrogen and phosphorus removal 21 2.2.3.2. Phenol and Chlorophenol degradation 24 2.2.3.3. Pesticide removal 25 2.2.3.4. Surfactant removal 26 2.2.3.5. Heavy metal removal 27 2.2.3.6. Other wastewater treatment 27 2.2.4. Immobilization procedure scale-up 28 3. M A T E R I A L S A N D M E T H O D S 30 3.1. Sludge and wastewater 30 3.2. Immobilization of activated sludge 31 3.2.1. Preparation o f P V A - H B 0 3 (PHB) immobilized sludge beads 31 3.2.2. Preparation o f P V A - C a ( B 0 3 ) 2 (PNB) immobilized sludge beads 32 3.2.3. Preparation of PVA-Alg ina te - C a ( B 0 3 ) 2 ( P A N B ) immobilized sludge beads 32 3.2.4. Preparation of PVA-Alginate-Ca(B0 3 )2 blank beads 33 3.3. Determination of activity and stability of immobilized sludge 33 3.4. Continuous treatment 34 3.4.1. Experimental apparatus 34 3.4.2. Experimental runs 36 3.5. Analytical methods 37 3.6. Scanning electron microscopic examination 37 4. R E S U L T A N D D I S C U S S I O N 38 v 4.1. Batch assays of immobilized sludge materials 38 4.1.1. Examination of sludge immobilization methods 38 4.1.2. Impact o f P V A contents 43 4.1.3. Operational stability of immobilized sludge 45 4.1.4. Effect of wastewater p H on the removal o f organic carbon 47 4.2. Continuous treatment of domestic wastewater 49 4.2.1. Effect of aeration rates 51 4.2.2. Effect of hydraulic retention time 54 4.2.3. Effects of intermittent aeration 61 4.2.4. Impact of wastewater feed patterns 73 4.2.5. Excess sludge production 78 4.2.6. S E M examination 83 4.2.7. Technical evaluation 85 5. C O N C L U S I O N S A N D R E C O M M E N D A T I O N 87 5.1. Conclusions 87 5.2. Recommendation 90 6. R E F E R E N C E S 92 vi. L I S T O F T A B L E S 3-1. Characteristics of actual domestic wastewater in this research 31 3- 2. Experimental conditions 36 4- 1. Effect of final P V A concentrations on the formation of immobilized sludge 44 4-2. Results of domestic wastewater treatment with immobilized sludge process at different aeration rates 52 4-3. Operational conditions and average influent concentrations for experiments of determining the impact of hydraulic retention time 56 4-4. Results of domestic wastewater treatment with immobilized sludge process at different hydraulic retention time (HRTs) 56 4-5. Operational conditions and influent concentrations for experiments o f determining the impact of aeration patterns 63 4-6. Results of domestic wastewater treatment with immobilized sludge process at different aeration patterns 63 4-7. Operational conditions and influent concentrations for experiments o f determining the effect of organic loading under intermittent aeration condition 70 4-8. Results o f domestic wastewater treatment with immobilized sludge process at different aeration patterns 70 4-9. Operational conditions and influent concentrations for experiments o f determining the impact of feed patterns under intermittent aeration condition 75 4-10. Results of domestic wastewater treatment with immobilized sludge process at different feed patterns 75 4-11. Experimental and calculating results o f excess sludge production under continuous aeration condition 81 4-12. Experimental and calculating results o f excess sludge production under intermittent aeration condition 81 4-11. Comparison of wastewater treatment performance of several processes.... 85 vii LIST O F FIGURES 3- 1 Schematic diagram o f experimental apparatus for continuous treatment of domestic wastewater 35 4- 1. Time courses of T O C removal o f different immobilized sludge and free sludge 41 4-2. T O C removal activity of different immobilized slduge and free sludge 42 4-3. Effect of final P V A concentrations in immobilized sludge on T O C removal efficiencies of fresh beads 44 4-4. Changes of T O C removal efficiencies of immobilized sludge and free sludge during repeated batchwise experiments. 45 4-5. Changes of immobilized sludge bead volume during repeated batchwise experiment 46 4-6. Effect of p H on T O C removal with immobilized sludge 48 4-7. Time courses of T O C in the influent and effluent in the first 60 days of continuous treatment experiment 50 4-8. Removal efficiencies of organic carbon and suspended solids at different aeration rates 53 4-9. Removal efficiencies of nitrogen and phosphorus at different aeration rates 53 4-10. Effect of H R T s on the removal o f organic carbon and suspended solids under continuous aeration condition 57 4-11. Effect of H R T s on the removal o f nitrogen and phosphorus under continuous aeration condition 57 4-12. Concentrations of organic carbon and suspended solids in the influent and effluent at different H R T s under continuous aeration condition 58 4-13. Concentrations of nitrogen in the influent and effluent at different H R T s under continuous aeration condition 58 4-14. Effect of aeration patterns on the removal of organic carbon and suspended solids 64 4-15. Concentrations of organic carbon and suspended solids in the effluent vs aeration patterns 64 4-16. Effect of aeration patterns on the removal of nitrogen and phosphorus 65 4-17. Concentrations of nitrogen and phosphorus in the effluent vs aeration patterns 65 4-18. Components of nitrogen in the effluent vs aeration patterns 67 4-19. Effect of organic loading on the removal o f organic carbon and suspended solids under intermittent aeration condition 71 4-20. Effect of organic loading on the removal o f nitrogen under intermittent aeration condition 71 4-21. Nitrogen concentrations of in the effluent at different organic loading rates under intermittent aeration condition 72 4-22. Effect o f influent feed patterns on the removal o f organic carbon and suspended solids at H R T of 6 hours 76 4-23. Effect o f influent feed patterns on the removal nitrogen at H R T of 6 hours 76 4-24. Effect of influent feed patterns on the removal o f organic carbon and suspended solids at feed time o f 8h/d and different H R T s 77 4-25. Effect of influent feed patterns on the removal o f nitrogen at feed time of 8h/d and different H R T s 77 4-26. Determination of excess sludge production of immobilized sludge process 82 4-27. Scanning electron micrographs 84 ix L I S T O F A B B R E V I A T I O N S B O D 5 Five-day Biochemical Oxygen Demand C O D Chemical Oxygen Demand H R T Hydraulic Retention Time P H B P V A - H B 0 3 immobilized sludge beads P N B P V A - C a ( B 0 3 ) 2 immobilized sludge beads P A N B PVA-Algina te -Ca(B03)2 immobilized sludge beads P V A Polyvinyl Alcohol T K N Total Kjeldahl Nitrogen T N Total Nitrogen T O C Total Organic Carbon TSS Total Suspended Solids X A C K N O W L E D G E M E N T S I would like to express my gratitude to many people who helped me complete this thesis. First o f all, I would like to thank my supervisor, Dr. K . V . L o for his guidance, advice, and support during the project. I am very grateful to Dr . A . Lau and Dr . R . Branion for serving on my supervisory committee and offering advice and assistance. I would also like to express my gratitude to other B i o E members: Dr . P . Liao for his valuable advice and assistance in experimental design and laboratory procedures; Dr . E . Humphrey for her assistance in S E M analysis; M r . Ne i l Jackson for his assistance in the construction o f the experiment; and my fellow students Ke lv in Y i p and T i m Shelford for their assistance. A n d finally, I would like to express my appreciation to my wife, Jie Jin; my son, Kev in A n ; and my other family members for their understanding, encouragement, and support during this time in my life. xi' C H A P T E R 1 I N T R O D U C T I O N In sparsely populated areas, it is difficult to apply large-scale wastewater treatment systems. Therefore, different solutions must be developed for water protection. A n appropriate treatment system applied to such areas would be a small scale wastewater treatment system. Such a small scale treatment system means that wastewater is treated at each source point. Formerly legislation placed stipulations only on effluent B O D and suspended solids, increasingly compliance levels are being imposed on effluent C O D and ammonia concentrations. In recent years, nutrient (nitrogen and phosphorus) removal has also been required. Traditionally low technology and reliable operation are often adopted in small scale wastewater treatment. So far, the septic system is the one o f most commonly used technologies, which consists of a septic tank, for the partial treatment o f the wastewater, and a subsurface soil disposal field for final treatment and disposal o f the septic tank effluent. In many instances, such septic systems failed to meet the required effluent standards. In many land-limited locations, the disposal fields cannot be used. Therefore, an effective alternative system needs to be developed. Over the last decade some advances have been made in developing small wastewater treatment systems (Fujiwara, et al., 1986; Fastenau, et al., 1990; Kuroda, et al., 1990; Chiemchaisri, et al, 1993; Paulsrud, et al., 1993). In these systems, some are 1 designed to treat septic tank effluent; the others, however, can be used to replace the whole septic system. The technologies applied include fixed media system (e.g. rotating bio-disc, bio-filtration) and activated sludge systems (e.g. completely mixed, extended aeration, S B R etc.). The main problems that hamper those techniques to be widely utilized are the high costs of operation and maintenance and the low operational stability. Apart from the established methods, immobilized activated sludge process could be a feasible alternative technique. The use of immobilized cells for wastewater treatment has received widespread attention in the last few decades (Nilsson, et al., 1980; Kokufuta, et al., 1986; L i and Chen, 1993; Leenen., et al., 1996; Yang, et al., 1997). Usually, conventional biological wastewater treatment technologies, such as activated sludge process, require a fairly large bioreactor with the long treatment time and a great quantity of excess sludge production. The use of immobilized cells is believed to offer potential for reducing the size o f bioreactor and the amount of excess sludge production. A s a result, the treatment efficiency would be improved significantly and less problems of solid-liquid separation in the settling tank would be involved (Nilsson and Ohlson, 1982; Kokufuta, et al., 1986). Since activated sludge is densely and inclusively entrapped in polymeric materials, high concentrations of activated sludge could be maintained in the reactor, without many sludge washouts. Moreover, immobilized microorganisms exhibited better protection than suspended cells from low p H and toxic materials (Hanaki et al., 1994). These advantages are very important for the design and operation of a small domestic wastewater treatment system. 2 Entrapment of organisms inside a polymeric matrix is one o f the most widely used techniques for cell immobilization. The resulting materials are porous enough to allow the diffusion of substrates to the microbial cells and of products away from the cells. Several natural products and synthetic polymeric materials, such as agar, agarose, k-carrageena, collagen, alginates, chitosan, polyacrylamide, polyurethane, cellulose and polyvinyl alcohol ( P V A ) , have been successfully used for the immobilization of microorganisms. Among these polymers, P V A is the most suitable material for application to wastewater treatment in terms of mechanical strength, durability, cell viability, economical feasibility (Hashimoto et al, 1987). Ariga et al (1987) immobilized microorganisms with P V A hardened by iterative freezing and thawing. Hashimoto et al (1987) entrapped activated sludge in P V A beads, followed by crosslinking the P V A using boric acid to form a monodiol-type PVA-bo r i c acid gel lattice. Activated sludge was successfully immobilized using this technique, without apparent loss of biological activity. The immobilized sludge beads exhibited a rubberlike elasticity and strong consistency. A high removal efficiency o f T O C and T K N was achieved when they were used to treat synthetic domestic wastewater. Although wastewater treatment continues to be one of the most significant areas of application for immobilized cell processes, much of the research has been concentrated on nitrification and denitrification of synthetic wastewater by immobilized nitrifiers (Leenen, 1996). Studies on actual domestic wastewater treatment by immobilized activated sludge, however, were scarce in literature. Furthermore, very few studies have 3 been directed toward development of small domestic systems using activated sludge immobilization process. Developing an effective immobilized sludge process for small domestic wastewater treatment involves a great deal of work. The experiments in this research are the first step toward that goal. The overall objective of this research project is to evaluate the feasibility o f using immobilized sludge for small domestic wastewater treatment. The specific objectives are: 1. to investigate the activity and stability of immobilized sludge in wastewater treatment and the factors that affect the organic carbon removal efficiency of immobilized sludge. 2. to start up a lab-scale fluidized bed reactor with immobilized sludge for domestic wastewater treatment; 3. to investigate the performance of the immobilized sludge process with variation of hydraulic load and organic load of domestic wastewater; 4. to determine the impact of different aeration and feed patterns on the treatment efficiency of domestic wastewater by immobilized sludge. 4 C H A P T E R 2 L I T E R A T U R E R E V I E W 2.1. Smal l domestic wastewater treatment Small domestic wastewater treatment systems or onsite domestic wastewater treatment systems were originally developed to serve individual homes in rural areas where central sewage treatment plants are not economically feasible due to the cost of sewers, or recreational communities; however, as the number o f people moving from urban areas to unsewered communities increases, so does the use of small domestic wastewater treatment systems. In rural or unsewered residential communities, wastewater generated in the household is diverted to an onsite wastewater management system, which conventionally consists of a treatment unit and a disposal unit. The treatment unit removes the large solids and greases and may provide biological or physical-chemical treatment. The disposal unit is generally a soil absorption system or sand filter, which provides biological treatment and ultimate disposal of the wastewater through percolation into shallow aquifers or evapotranspiration at the ground surface (Metcalf & Eddy, 1991). One of the important design objectives for individual onsite systems is the effective treatment of the wastewater from individual residences so that it does not cause any nuisance condition and does not impact any of the beneficial uses of the local groundwater. The principal constituents of concern a r e B O D 5 , SS, nitrogen, phosphorus, bacteria, and viruses (Bucholz, 1979). 5 Although a variety of onsite systems have been used, the most common onsite household treatment facility consists of a septic tank for the partial treatment o f the wastewater and a subsurface soil disposal field for final treatment and disposal o f the septic tank effluent. Because conventional septic tank and subsurface soil disposal systems cannot be used in some locations, many alternative systems have been developed, including those that provide treatment and disposal by physical-chemical and recycling systems that wi l l not be reviewed here. 2.1.1. Small domestic wastewater flowrate and strength The flowrates and wastewater characteristics from individual homes are highly variable. Knowledge o f the expected wastewater flowrates and characteristic is essential for the effective design of wastewater management facilities for individual residences. It is indicated that the average flowrates from individual homes are in the range o f 35-100 gal/capita.d. A typical per capita value for residences in unsewered areas is about 55 gal/capita.d based on an average occupancy o f about 2.4-2.8 residents per home (Metcalf & Eddy, 1991). Some sources recommend that 25% be added to the average flowrate to provide for guests and other factors in the lifestyle of a family. Wastewater flowrate variations are important as surging can affect treatment in some plants and hydraulic loading in all plants. The flowrate variations that can be expected from an individual residence are quite variable, ranging from no flow in the early morning hours to peak hourly flowrate as high as 8 to 1 compared to the average 6 daily flowrate. Typical peaking factors for individual residences are 6 for peaking hour, 4 for peaking day, 2 for peaking week, and 1.75 for peaking month (Mecalf & Eddy, 1991) Wastewater strength is normally classified in terms o f B O D 5 and SS. A commonly used value is 0.082 kilogram o f B O D 5 contributed per person each day and 0.090 kilogram of SS contributed per person each day. Wastewater strength is determined by the number of residents and the volume of water consumed. 2.1.2. Small wastewater treatment systems 2.1.2.1. Septic tank Septic tank systems have been used for wastewater treatment and disposal in unsewered areas since the turn of the century. Many, but not all, onsite wastewater treatment systems are natural systems using septic tanks as a pretreatment. The natural disposal systems usually are divided into two broad classifications: (1) soil-based systems, such as subsurface infiltration, rapid infiltration/soil aquifer treatment, overland flow, and slow rate systems; (2) aquatic systems, such as pond, floating aquatic plant, and constructed wetland systems. 7 The natural disposal systems are thought to function as a "secondary" treatment system for the final disposal o f wastewater flowing out of the septic tank. There wi l l be no further discussion of these in this review. The main function of a septic tank is to remove the solids from the domestic wastewater. The waste liquids are retained in a tank for a designed period o f time allowing the heavier sewage solids to settle to the bottom of the tank where they form blanket of sludge. Lighter solids, including fats and grease, rise to the surface and form a layer of scum. The retained sludge, and scum to lesser degree, undergoes partial digestion and compaction but otherwise forms a residual in the tank (i.e., septage) which must be removed intermittently by pumping the tank clean (Ross, 1980). In an efficiently operating septic tank, a liquid effluent with low solids content is discharged to a receiving subsurface soil disposal system. This liquid, highly charged with bacteria and nutrients, continues to biodegrade as it percolates downward through the soil. Physical, chemical and biological reactions within the soil matrix remove wastewater contaminants before the liquid reaches the water table. This important factor is the principal reason for establishing a minimum depth o f soil filter above a water table, rock, or impervious soils. Wi th good design and careful construction, a septic tank system w i l l need very little maintenance i f it is used properly. It should not be necessary to pump out the tank more than once every three years. It should, however, be inspected at least once a year and pumped out when necessary. Failure to pump out a septic tank when required w i l l 8 result in sludge or scum being carried into the leaching bed, which in turn may clog and cease to function. Inspection of sludge and scum accumulation is the only way to determine when a tank should be pumped out. 2.1.2.2. Conventional aerobic treatment systems Aerobic systems have been developed to replace or supplement existing septic tanks (Webster, 1974). Generally, aerobic tanks produce an effluent of higher quality than that discharged from a septic tank. The septic tank separates solids from wastewater by sedimentation and flotation and partially stabilizes retained pollutants. The aerobic tank separates large solids from incoming wastewater, stabilizes large unstable pollutant molecules to smaller, stable molecules through the action of aerobic microorganisms, and enhances desirable predatory-prey relationship between microorganisms in the tank effluent, improving infiltration and evapotranspiration in subsurface absorption systems. Wi th the generation of a higher quality effluent, many regulatory agencies stipulate allowable reductions in the size of subsurface disposal fields that receive such discharges. The rationale behind the reduction disposal field area rests on the basis that aerobic tank effluents promote desirable aerobic biological activity within the soil matrix, keeping soil voids open and enhancing the percolation rate and evapotranspiration properties of the soil. Conventional aerobic tanks for on-site treatment of domestic wastewater are available commercially in one, two or three-compartment designs. Some designs 9 incorporate a "fail-safe" sand filter to provide polishing of final effluent. In general, wastewater entering the aerobic mini-plant accumulates in an aerated chamber for a certain period of time, or as controlled by level sensors. After aeration, the wastewater is allowed settle, either in a separate chamber or by automatic shutdown of aeration devices in the holding chamber. Thereafter the treated effluent is pumped to the disposal field. Some plants incorporate components which permit the recycle of settled sludge from the settling compartment to the aeration compartment (Ross, 1980). The conventional aerobic treatment processes that usually utilize the general activated sludge process and extended aeration treatment process to treat wastewater, rely on microorganism suspended in the wastewater to remove organic material. Although long-time operating information on aerobic tanks has yet to be developed, evaluation studies conducted by recognized testing agencies in both Canada (Chowdhry, 1974) and the United States ( U N S F T L , 1972) have revealed some potential problems, such as periodic biological solids washouts, low resistance to organic and hydraulic shock loads, and sludge bulking, etc. These problems have resulted in effluent quality deterioration and make it difficult to control and have prevented these conventional treatment processes from being widely utilized. 2.1.2.3. Other treatment systems Many efforts have been made to develop other treatment processes for small domestic wastewater treatment. Fixed film process is one of the many types of treatment 10 processes that have been utilized to develop on-site domestic wastewater treatment technology, which is regarded as a compact low cost alternative. In 1982, Harr introduced and evaluated six package plants for onsite domestic wastewater treatment developed in Europe (Harr, 1982). Four of them utilized a fixed film process. One onsite domestic treatment system, which consists of a contact aeration reactor Gappei-shori Johkaso, has been developed in Japan (Fujiwara, 1986). Sankai, et al. (1997) have recently reported on one kind of modified Gappei-shori Johkaso, which consists of both anaerobic filter and contact aeration and is able to treat domestic wastewater consisting of both night soil and gray water. Kuroda et al. ((1990) developed a fixed-bed aeration reactor system for on-site household wastewater treatment. It was made with a formed F R C cube consisting of three cells: a settling cell, an equalization cell and a fixed-bed aeration cell. The fixed-bed aeration cell is divided into two independent reactors packed with rectangular horizontal plates as support materials that were made of P V C . Each reactor was filled with horizontal plates. The reactors were operated in a batch mode. Experimental results showed the reactor could treat household wastewater efficiently. Other technologies that have been explored to develop an onsite wastewater treatment process include: rotating bio-contact discs (Greaves et al., 1990), intermittently aerated activated sludge system (Hulsman and Swartz, 1993), activated sludge coupled with membrane separation process (Chiemchaisri et al., 1992 and 1993), sequencing 11 batch reactor (SBR) process (Vuoriranta et al ,1993 ), flow equalization anoxic-oxic biofilm process (Imura, et al., 1995), anaerobic/aerobic biofilm reactor process ( Inamori et al., 1996), and immobilized cell process (Yang et al., 1997). Some of these treatment systems have been installed for domestic wastewater treatment, but most o f these small package treatment plants are still required evaluation and improvement. The main problems for application of these systems are the high costs o f operation and maintenance, the delay in starting up the operation and the low flexibility in treating various sources and characteristics of wastewater (Yang et al., 1997). However, some of these systems are still economically feasible and competitive in places where the cost o f land and water supply are very high as these systems can produce an effluent o f reuse quality. 2.2. Immobil izat ion of cells The earliest purposeful use of immobilized cells began in the early 19 t h century (Chibata, 1983; Linko, 1983). Since that time, immobilization of microbial cells has been developed and utilized in a wide range of fields, from biomedical, analytic, to wastewater treatment. The immobilization of cells can be defined as any technique that limits the free migration o f cells. Cel l mobility can be restricted by confining them into, or attaching them to, a solid support (Scott, 1987). 12 2.2.1. Immobilization methods There are five principal methods for cell immobilization: adsorption, covalent binding, crosslinking, encapsulation, and entrapment (Bickerstaff, 1997). Entrapment of organisms inside a polymeric matrix is one of the most widely used techniques for cell immobilization. This method involves incorporation of the cells within the lattice o f a natural or synthetic polymer. It represents a more definite means of immobilization that does not have a significant dependence on cellular properties and is by far the most frequently used technique. Entrapment of cells can be achieved by the following methods (Bickerstaff, 1997): 1. Ionotropic gelation of macromolecules with multivalent cations (e.g. alginate, k-carrageenan) 2. Temperature-induced gelation (e.g. agrose, gelatin) 3. Organic polymerization by chemical or photochemical reaction (e.g. P V A , polyacrylamide). 2.2.2. Immobilization materials In the past twenty years, many new techniques have been developed in which microbial cells are immobilized by entrapment. The resulting immobilized cells could be porous enough to allow the diffusion of substrates to the microbial cells and of products 13 away from the cells. Many natural products and synthetic polymers, such as agar, agarose, k-carrageena, collagen, alginates, chitosan, polyacrylamide, polyurethane, cellulose and polyvinyl alcohol ( P V A ) , have been used for the immobilization of microorganisms. Alginates, k-carrageenan, cellulose, polyacrylamide, polyurethane and polyvinyl alcohol ( P V A ) are among the most widely used. 2.2.2.1. Alginate Immobilization of cells in alginate is one of the most widely used methods because it is recognized as a simple, cheap, nontoxic and versatile method for immobilization o f cells and enzymes (Bickerstaff, 1997). Alginates are extracted from seaweed and consist of varying proportions of D -mannuronic- and L-glucuronic acids. The ratio between these two acids depends on the source and determines the gel strength. L-glucuronic acids have preferential binding sites for divalent cations, such as C a 2 + and bound ions interact with other glucuronic-acids and form linkages that lead to gel formation. The gel strength increases with an increasing amount of the guluronic-acid groups in the polymer, depending on the seaweed and the part of the species used (Martinsen et al., 1989; Smidsrod and Skjak-Brak, 1990). The principal procedures of gel entrapment include: (1) cells are mixed with sodium alginate solution. The concentration of sodium alginates has to be sufficient to produce a firm gel. (2) the mixture of cells and sodium alginate solution is extruded 14 through an orifice or hollow needle to a calcium-containing solution. Interfacial polymerization is instantaneous with precipitation of calcium alginates. The gels obtained are biochemically inert and mechanically stable with interstitial spaces that are suitable for cell immobilization (Cheetham et al., 1979). Cel l leakage often occurs from the gel beads. This is influenced by many factors such as initial alginate concentration, mechanical treatment of the beads, and cell reproductivity. Mechanical shaking results in more leakage than packed beads in a column (Smidsrod et al., 1974). Chelating agents such as phosphate, citrate, and E D T A etc. and other divalent cations such as Mg2+ can reduce stabilization of alginate beads. Stabilization of alginate gel beads can be improved by employing additional cross-linking with polyvinyl alcohol and some other chemicals to form more stable complexes with improved leakage characteristics. 2.2.2.2. Carrageenan Immobilization of cells in carrageenan is also one of the most widely used methods. In general like immobilization in alginates, it is an inexpensive and simple immobilization method and allows a mild immobilization procedure, which enables most cells to survive the immobilization process (Lewandowski et al., 1987). Carrageenan consists of alternating structures of D-galactose 4-sulphate and 3,6-anhydro-D-galactose 2-sulphate. The gel strength increases with the level of 3,6-anhydro-15 d-galactose 2-sulphate in the polymer. Immobilization in carrageenan is carried out in a similar way to that for alginates. Gelification of carrageenan takes place when the solution is cooled down to the appropriate temperature and in the presence o f gel-inducing agents such as K+. Addition of gums (Audet, et al., 1990) or A13+ (Chamy, et al., 1990) can improve the gel characteristics. Generally, the resulting beads may soften or disintegrate at elevated temperatures and consequently are not suited to applications involving higher temperature treatment processes (Guiseley, 1989). Another disadvantage is that when a gel-inducing reagent is not present in the reaction mixture, disintegration of the gels w i l l occur and cells are leaked out. 2.2.2.3. Cellulose The use of cellulose and its derivatives as support materials for immobilization o f cells has received attention for many years and possible applications have been reviewed extensively (Gemeiner et al, 1994). Cellulose is a polydisperse polymer. Its basic monomeric unit is D-glucose. Cellulose is composed of B-D-glucopyranosyl units linked by carbons 1 and 4 bonds and with additional interchain interaction through hydrogen bonds. Immobilization of cells in cellulose follows the method established earlier for enzymes (Linko, 1979). A n early example of entrapping whole cells in cellulose 16 triacetate was reported by Kolar ik et al. (1974). A solution o f cellulose triacetate in methylene chloride was mixed with a cell aqueous suspension. The product was then made into fibers by injection through a syringe into toluene, or by casting into membranes on glass or water. Entrapped microorganisms were found to retain 57% of the original cell activity, although there was a problem with cell leakage. Cellulose and its derivatives are not as popular as other polysaccharide support materials such as alginate, carrageenan, etc. The reason is susceptibility to microbial degradation. Cellulose has a major advantage over other polysaccharide supports in that it has become available in many different physical forms, such as fibers, microgranules, beads, membranes, tubings, emulsions, etc. (Gemeiner et al., 1994) 2.2.2.4. Polyacrylamide Polyacrylamide is the synthetic polymer most often used in cell immobilization. It is formed by the linear polymerization of the vinyl carboxylic acid amide and is readily soluble in water. The polymer is normally crosslinked to the copolymer N , N'-methylene-biacrylamide to produce a polymer with lattice-like structure better suited to cell immobilization. A n early example of cell immobilization in polyacrylamide is the work of Updike et al. (1969). The immobilization procedure involves treating a solution o f the monomers acrylamide and N,N'-methylenebisacrylamide in phosphate buffer with the solution of 17 catalysts. Polymerization was initiated at 20°C by a photo flood-lamp and resulting gel broken up and sieved to -40 mesh. Viabil i ty of the cells up to 5 days was demonstrated by oxygen uptake, glucose utilization and mobility. The main advantage of using polyacrylamide as carrier to entrap cells is the resulting stabilization effects on the cell-related enzymes. The disadvantage is polyacrylamide toxicity to the cells which results in loss of cell viability in a large portion of the population (Kolot, 1988). 2.2.2.5. Polyurethane Polyurethanes consist of the repeating carbamate group, - H N - C O - 0 - , which is formed between a diisocyanate, R(NCO)2, and a glycol unit H O - [ C H 2 C H 2 0 ] n H (Phillips and Poon, 1988). There are different type of polyurethanes, which differ in molecular weight, N C O content of the prepolymer, and ethylene oxide content in the diol. Tanaka et al. (1979) first reported entrapment of whole cell in polyurethane. A chilled suspension of microbial cells in a phosphate buffer at p H 7.0 was mixed with a solution of the prepolymer. The prepolymer contains polyethylenediol units with different ethylene oxide content. The mixture of cells and prepolymer was poured onto a glass plate and kept at 4°C for 1 hour to form gel. The polymer, after washing with phosphate buffer at p H 7, showed 15% of the free cell activity. 18 The disadvantage of this polymer is also its toxicity to the microbial cells. Activi ty losses of 60-90% are reported (Sumino et al., 1992). But it has a number of advantages including: simple immobilization method, various forms o f immobilized cells matrices, and withstanding high-pressure support material. 2.2.2.6. Po lyv iny l alcohol The use of polyvinyl alcohol ( P V A ) for cell immobilization is a relatively new method. A number of methods have been developed to use P V A as support materials for cell immobilization. Ar iga et al.(1987) immobilized microorganisms with P V A followed by iterative freezing and thawing to improve bead strength. It was found that this technique produced a rubber-like and high-strength immobilized cells. Hashimoto et al.(1987) entrapped activated sludge in P V A beads, followed by crosslinking the P V A with boric acid to form a monodiol-type PVA-bo r i c acid gel lattice. Activated sludge was successfully immobilized using this technique, without apparent loss o f biological activity. The immobilized sludge beads exhibited a rubberlike elasticity and strong consistency. Wu , et al.(1991) crosslinked P V A beads with boric acid in the presence of a small amount o f alginate to improve the surface properties and prevent agglomeration o f beads. The saturated boric acid solution used to crosslink the P V A is highly acidic because of low p H (about p H 4) and could cause difficulty in maintaining cell viability i f the immobilized cells are kept in it for 15-24 hours. Chen and L i n (1993) developed a 19 method to reduce the time that is needed for boric acid crosslinking with P V A . In their technique, P V A was first crosslinked with boric acid for a short time (10 minutes ~ 2 hrs) to form spherical beads, which was followed by solidification o f the gel beads by esterification of P V A with phosphate. Ideally, the support materials, used for cell immobilization in the application o f wastewater treatment, should have the following characteristics: high mechanical and biological stability, high diffusivity, high bioactivity of resulting immobilized cells, simple and inexpensive immobilization procedure. Generally, the immobilization procedures of natural polymers are mild and the biological activities of microorganisms after immobilization are well maintained. The effective diffusional coefficients of natural polymers are high. However, natural polymers are biologically and mechanically unstable and therefore not suitable for being used in wastewater treatment. Synthetic polymers, like polyacrylamide and polyurethane, are not biodegradable and have good mechanical stability. However, the polymers are toxic to microorganisms and the immobilization procedures are complex and harsh. The survival of microorganisms in these support materials is poor. Polyvinyl alcohol is a nontoxic and inexpensive synthetic polymer. Compared to the commonly used polymeric support materials mentioned above, P V A is the most suitable material for application to wastewater treatment in terms of mechanical strength, durability, cell viability, economical feasibility (Ariga, 1987; Hashimoto, et al., 1987). Presently, growth of a mixed culture of nitrifying bacteria (Wil lke and Vorlop, 1995), E . 20 coli (Ariga et al, 1987) and acclimated sludge (Ariga et al., 1987; M y o g a et al., 1991) immobilized by the PVA-freezing method have been reported. 2.2.3. Applications in wastewater treatment The use of immobilized microorganisms for wastewater treatment has received widespread attention in the last few decades (Nilsson, et al., 1980; Kokufuta, et al., 1986; L i and Chen, 1993; Leenen., et al., 1996; Yang, et al., 1997). Although more and more studies are related to using relatively undefined microbial systems with mixed cultures formed as a biological film on solid surfaces, the entrapped-cell systems with well defined microbial populations are still one of the most popular techniques for cell immobilization (Scott, 1987). In this chapter, some applications of immobilized cells prepared by entrapment method in wastewater treatment are reviewed. 2.2.3.1. Nitrogen and phosphorus removal M u c h research has been done in nitrification and denitrification of wastewater in the recent years (Scott, 1987; Asano, et al., 1992; Leenen, et al., 1996). Immobilized nitrifiers were used to remove high levels of ammonia from a gas scrubber wastewater (Takada et al., 1994). The pilot plant results indicated that removal 21 efficiencies up to 98% of the influent containing 95-260 mg N H V - N /L could be achieved with a retention time of 6 hours. The cells immobilized in photo-cross-linkable polymer were used to remove ammonia from a synthetic wastewater (Uemoto and Saiki, 1996). The immobilized cells are molded into both flat shapes and tubes. The tubular gel completely removed 200mg/l of N H 4 + - N within 100 hours in a batch reactor. The cells were distributed preferentially and evenly on the outside surface of the gel. L i n and Chen (1995) entrapped denitrifying sludge and methanogenic sludge with polyvinyl alcohol and used this mixed culture system to treat synthetic ground water with a high nitrate concentration. The results indicated an efficient capability in nitrogen reduction and methanol removal. They also found that heterotrophic denitrifiers tended to grow on the bead's surface and the anaerobic methanogens grow primarily in the bead's interior. PVA-immobi l ized nitrate-acclimated sludge was reported to be very effective for denitrification using phenolic pollutants as substrate in an upflow packed reactor (Fang and Zhou, 1997). The results showed that the denitrification rate was 10. l g N C V - N / L d, the degradation rates were 4.3g/ L d for phenol and 2.1g/L d for m-cresol at 30°C and F iRT 0.52 hour. Each gram of NO3" - N reduction required an average of 4.14 grams o f C O D , and the average sludge yield was estimated as 0.22g-VSS/ g C O D . dos Santos et al. (1996) described integration of nitrification and denitrification in a single immobilized coculture which consists of double-layer gel beads immobilizing 22 Nitrosomonas europaea on the outside and P. denitrificans or Paracoccus denitrificans on the inside. Under aerobic conditions this coculture removed up to 5.1mmol N / m 3 gels, primarily as N2. Sakairi et al. (1996) reported that the mixed nitrifying and denitrifying microorganisms was immobilized in A Q U A C E L , a macroporous cellulose carrier. Wi th nitrogen loading of 1.3 and 20.8 kg N/m 3carrier d, 99-100% of nitrification and denitrification was achieved, respectively. However, methanol and phosphorus were needed for denitrification. Wijffles, et al. (Wijffles and Tramper, 1995; Wijffles et al., 1991 and 1995) used alginates and k-carrageenan to immobilize pure strains of (de)nitrifying bacteria and investigated diffusion limitation and growth of immobilized cells. A n experiment with immobilized microalgal cells was carried out using high and low viscosity chitosan and konjac flour to enhance the stability of hardened gels during tertiary treatment of wastewaters containing a high concentration of phosphate salts (1 M NasPOx) (Kaya and Picard, 1996). Acinetobacter johnsonii and Acinetobacter calcoaceticus cells were immobilized within alginate beads (3% and 3.5% alginate) to assess their phosphate uptake abilities (Muyima and Cloete, 1995). Phosphate uptake was dependent upon the alginate concentration and the strain. Notwithstanding leakage, cell densities within alginate beads over 24 h were higher than initial. It was found that non-growing immobilized cells showed higher phosphate uptake ability than growing cells. 23 Mechanical properties as well as compression testing, moisture retention in gels and the uptake of nutrients, such as nitrogen and phosphorus, from artificial wastewater were studied in the immobilized gel state. The green alga, Chlorella vulgaris was immobilized in Ca-alginate beads at two stocking cell densities and grew in primarily treated domestic effluents (Tarn et al, 1994). It was found that the cellular metabolic activities o f this alga were retained after immobilization. The growth (in terms o f cell counts) and photo synthetic rates of cells immobilized in beads of low stocking density were greater than those in beads of high density. Significant reduction of wastewater NH4 +-N and P04 3"-P was recorded especially in reactors containing algal beads of high density. The efficiency in removal o f ammonia-nitrogen and orthophosphate-phosphorus from continuous flow cultures was determined using two soil microalgae, Chlorella vulgaris and Scenedesmus bijugatus, entrapped in 4-mm calcium alginate beads (Megharaj et al., 1992). Exponentially growing cells of the two isolates effected greater uptake of phosphorus, whereas the age of the cultures was found to have no direct impact on the removal of nitrogen. 2.2.3.2. Phenol and Chlorophenol degradation The immobilized Pseudomonas was used to degrade phenol in the fluidized bed reactor (Wu and Wisecarver, 1991). Lee et al (1994) reported that immobilized cells could biodegrade phenol within 5 days, even when its initial concentration reached 500 24 mg-C/L. Immobilized cells also could more effectively biodegrade 2-chlorophenol and 2,4-dichlorophenol in comparison with suspended cells. Alginate-immobilized P. aeruginosa, P.putida, P. testosteroni and Agrobacterium radiobacter were used in batch reactor for the biodegradation of phenol and chlorophenols (Lu et al., 1996). A l l four species could remove phenol efficiently but the removal capacities of chlorophenols are much lower or nonexistent. Calvil lo et al.(1996) studied the effect of P. paucimobilis and Bacillus licheniformis immobilized on biphenyl degradation. It is found that degradation o f biphenyl sorbed to polyacrylic beads was enhanced 10% by bacterial attachment to the beads. 2.2.3.3. Pesticide removal A n entrapment of mixed microbial cell process was used to remove the pesticide Ethylene Dibromide (EDB) , Trichlopropane (TCP) and nitrate from a synthetic groundwater (Yang, et a l , 1993). The system was able to remove (aerobically) more than 90% of E D B (influent concentration of 300ug/L) at more than 30 minutes o f H R T . T C P (influent concentration of 2,81ug/L) could not be detected in the effluent at the same H R T . 25 2.2.3.4. Surfactant removal Thomas and White (1991) reported that the surfactant-degrading bacterium Pseudomonas C12B was immobilized in polyacrylamide gel beads and the specificity of the immobilized cells towards various surfactants was examined. Profiles o f activity towards primary alkyl sulfates of chain lengths from C6 to C14 were very similar for free and immobilized cells and reflected the known specificities of the alkylsulfatase enzymes that initiate the degradation. Initial rates of surfactant removal were in the order alkyl sulfates>linear alkyl benzene sulfonates>alkyl ethoxy s u l f a t e s » a l k a n e sulfonates. Sulfate esters were totally degraded within 48 h. Immobilized cells also removed surfactant from solutions containing raw alkyl ethoxy sulfate mixtures (used to formulate shampoos), diluted whole-shampoo formulations, and mixed balance-tank effluent containing waste shampoo, hair dyes, and permanent wave formulations from a hair products factory. Repeated exposure to these wastes in the presence of added basal salts showed that the immobilized cells retained at least 60% and 28% of their activity after 6-day and 13-day operating periods, respectively. Balance-tank effluent, while initially producing the lowest activity, also provided the greatest stability even in the absence of added basal salts. 26 2.2.3.5. Heavy metal removal Immobilized dried and ground sphagnum peat moss in a porous polysulfone matrix was effectively used to remove heavy metals from wastewater (Spinti et al., 1995). The selectivity of the beads was Fe>Al>Pb>Cu>Cd, Zn>Ca>Mn>Mg>Na. Sag et al.(1995) compared the Cu(II) removal capacities o f Ca-alginate gel and Ca-alginate-entrapped Zoogloea ramigera. It was found that in the initial phase, Cu adsorption was lower in the entrapped cells because of diffusion limitation. After Cu(II) loading was increased to 10.8mg/h, the removal efficiencies of cell and gel-only systems are 94% and 63%, respectively. 2.2.3.6. Other wastewater treatment Activated sludge immobilized with cellulose triacetate was used to treat food industry wastewater (Hsu et al, 1996). The immobilized sludge removed more than 80% of the influent soluble C O D (1.5g/L.d) at a H R T of 7 to 10 hours. A glass column packed with the immobilized cells was operated in a continuous mode to evaluate the effects of operating parameters such as temperature, retention time, and gas concentration on the removal efficiency o f hydrogen sulfide (Chung et al., 1996). 27 The results indicated that a significant deodorizing effect was observed from the operation of immobilized column and the immobilized biological reactor is a potential method to remove hydrogen sulfide. The rate of degradation of acetonitrile (160 mM) and sodium cyanide (40 mM) by the immobilized cells o f P. putida was studied (Babu et al., 1994). Remediation o f organic nitriles and inorganic cyanides can be achieved with immobilized cells of P. putida. Chapatwala et al.(1993) studied the degradation of N a C N using free cells and cells immobilized in agar, alginate or carrageenan matrices. Alginate-immobilized cells degraded N a C N more efficiently than free cells or agar- or carrageenan-immobilized cells. Results from the seven months' pilot plant operation showed that the decolorization effect of the immobilized cell reactor worked satisfactorily and the immobilized cell reactor is a high-efficiency and low-cost process suitable to be developed into full-scale application (Jian et al., 1994). 2.2.4. Immobil izat ion procedure scale-up For practical application, the mass production of immobilized cells should be realized easily and economically. 28 Hulst, et al. (1985) developed a technique, which consists o f breaking up a jet of the biocatalyst and natural polymer mixture in uniform droplets with a resonance-nozzle. Gotoh et al. (1993) improved this method for suspension with high viscosity. Hunik et al. (1993) reported the further scale-up of the extrusion technique. A rotating disc atomizer was used to produce alginate-immobilized cells (Ogbonna et al., 1989). For both natural and synthetic beads, prepared by the extrusion technique, these techniques can in principle be used. However, more research is absolutely necessary for the mass production of immobilized cells. 29 C H A P T E R 3 M A T E R I A L S A N D M E T H O D S 3.1. Sludge and wastewater Activated sludge used in this research was obtained from a reactor treating high-strength swine wastewater in the pilot plant laboratory of the Bio-resource Engineering Department of University o f British Columbia. For the preparation of immobilized activated sludge, the concentrated activated sludge was obtained by centrifuging at 1500rpm for 10 min. The density o f concentrated sludge was about 1 g/ml. Other parameters were 96±7mg dry weight/ g wet sludge and 69±5mgVSS/g wet sludge. The synthetic domestic wastewater used in the batch experiments was diluted from the stock nutrient medium containing Glucose (30g/L), ( N H t ^ S O ^ g / L ) , K H 2 P O 4 ( 0 . 7 5 g / L ) , M g S 0 4 - 7 H 2 0 (2g/L) and C a C l 2 H 2 0 (1.25g/L). The primary settled actual domestic wastewater was collected from the sewage treatment pilot plant of the C i v i l Engineering Department of U B C on the south campus o f U B C . In this research, it was assumed that constituents and characteristics of this sewage were similar to those of domestic wastewater from typical individual residences. The principal constituents of concern are summarized in Table 3-1. 30 Table 3-1. Characteristics of actual domestic wastewater in this research Item Range Average T O C 60.10-201.60 107.98 B O D 5 103.92-329.71 180.32 C O D 197.39-654.49 352.06 N F L f - N 16.97-51.66 30.45 T K N 21.12-77.41 40.00 T N 21.12-77.41 40.00 PO4-P 0 .88-30.55 9.35 TSS 55 - 113 79 Fecal coliform 10 J-107 lOOmL 10°/100ml Temperature 15-17°C 16°C p H 6.86 - 8.09 7.45 3.2. Immobilization of activated sludge 3.2.1. Preparation of P V A - H B O 3 (PHB) immobilized sludge beads P U B immobilized sludge beads were prepared using a method modified from studies reported by Hashimoto and Fukuruwa (1987). A mixture of polyvinyl alcohol ( P V A , molecular weight ranging 70,000-100,000; Sigma) and water ( P V A concentration of 5~20%(w/v)) was carefully heated to 70 °C to completely dissolve the P V A . The solution then was cooled down to below 35 °C. One portion of centrifuged activated sludge (1500 rpm for 10 min.)(100 ml~100 g) and one portion of P V A solution (100ml) mentioned above were thoroughly mixed. Droplets of the mixed solution of P V A and activated sludge were then squeezed out from a syringe into a solution consisting of saturated boric acid at room temperature. After 10-12 hours of gentle mixing, the droplets were transformed into spherical beads. 31 These beads then were rinsed with distilled water to remove any excess boric acid before being used for further experiments. 3.2.2. Preparation of PVA- Ca(B0 3 ) 2 (PNB) immobilized sludge beads The procedures for preparation of P N B immobilized sludge beads were almost the same as those described in the preparation of P H B immobilized sludge beads. The only difference was that before droplets of the mixed solution of P V A and activated sludge were squeezed into the saturated boric acid, the p H of the saturated boric acid solution was first adjusted to around 7 by using Ca(OH)2 solution. 3.2.3. Preparation of PVA-Alginate- Ca(BC>3)2 (PANB) immobilized sludge beads A mixture of P V A , sodium alginate (low viscosity, approximately 250 CPAs) , , and water was carefully heated to 70 °C to completely dissolve the P V A and sodium alginate. The solution then was cooled down to below 35 °C. In this solution, concentration of P V A and sodium alginate were - 2 0 % and ~2%(w/v), respectively. The rest of preparation procedures were the same as those used in the preparation of P N B immobilized sludge beads. The final beads contained 50% (w/v) of wet sludge, 1% of sodium alginate, and 10% of P V A . The sludge stocking in the beads was ~ 34.5g V S S / L beads. The average bead diameter was around 3 mm. 32 3.2.4. Preparation of PVA-Alginate-Ca(B03)2 blank beads The preparation of P A N B blank beads was exactly same as those o f P A N B immobilized sludge beads. However, the sludge used in the preparation was heated to ~100 °C for 15 minutes before centrifuging. 3.3. Determination of activity and stability of immobilized sludge To evaluate the activity of immobilized sludge, batch reactors were used. About 25g of immobilized sludge beads or 12.5g of free sludge were suspended in a 500ml glass bottle with 250 ml o f synthetic domestic wastewater diluted from the stock nutrient medium. The overall biomass in the bottle was the same for immobilized and free sludge. Wi th immobilized activated sludge for determination of T O C removal rate, the systems were aerated by an air diffuser connected to a compressor at a temperature o f 15-17°C. The aeration rate was controlled at 1.5L air/min. Samples were withdrawn at intervals and centrifiiged. The supernatant was assayed for T O C concentration. Repeated batchwise experiments were carried out to determine the operational stability of immobilized beads. The experimental conditions ^were the same as those described in the experiments for determination of the activity of immobilized sludge. A t the end of each run (24 h), the beads were separated and washed with tap water before fresh synthetic domestic wastewater was added for the subsequent run. The free sludge 33 was collected by settlement after each run. The supernatant was discarded and replaced with fresh wastewater for the subsequent run. 3.4. Continuous treatment 3.4.1. Experimental apparatus Fluidized-bed reactors, with working volumes of 5.8L and effective volumes of 5 L (the net volume of wastewater in the reactor), were used in this study. A schematic diagram o f the experimental system is shown in Figure 3-1. The volume fraction of immobilized beads in the reactors was - 1 5 % (v/v). The total initial concentration of activated sludge was estimated to be about 47,000mg suspended sludge/L. The top and bottom of the reactors were equipped with stainless screens to prevent the beads from washing out. Peristaltic pumps were used to introduce the domestic wastewater into the reactors from the bottom and the effluent was removed from the top of the reactors. Wastewater was maintained at 4 °C in a walk-in refrigerator. A stock synthetic domestic wastewater was added into actual domestic wastewater to increase the influent concentrations i f needed. Aeration and mixing were realized by the aerators for which the aeration rate could be controlled. Intermittent aeration and feed were controlled by timers. The air was introduced into the reactor from the bottom at 34 flow rates of 0.5 ~ 1.5 L/min. When the flow rate of air was greater than 1.0 L/min, the aeration completely fluidized the gel beads. D O meter Screen A i r -Aerator Influent Pump 7 " p H meter Screen Pump A i r stone Effluent — ^ Figure 3-1. Schematic diagram of experimental apparatus for continuous treatment of domestic wastewater 35 3.4.2. Experimental runs Once the reactor systems reached steady state, the experiments were manipulated by varying the H R T , aeration rate, etc. Table 3-2 summaries the operating conditions of experiments 1-18 to examine the influence of different conditions on the reactor performance. Table 3-2. Experimental conditions Run number 1 2 3 4 5 6 7 8 9 Sample number 7 7 6 8 7 6 7 7 6 Aeration rate (L/min) 1.0 1.5 1.0 0.5 1.0 1.0 1.0 1.0 1.0 Aeration pattern 1: 0**'* 1:0 1:0 1: 0 1: 0 1:0 1: 1 1:2 1: 3 HRT (hrs) 24 12 12 12 6 3 24 24 24 Feed pattern C C C C C C C C C Feed time(h/d) 24 24 24 24 24 24 24 24 24 Flowrate (L/d) 5 10 10 10 20 40 5 5 5 Hydraulic loading rate(L/d L) 1 2 2 2 4 8 1 1 1 Run number 10 11 12 13 14 15 16 17 Sample number '7 6 8 7 8 7 8 7 Aeration rate (L/min) 1.0 1.0 1.0 1.0 1.0 1.0 1.0 1.0 Aeration pattern 1:4 1: 5 1: 3 1: 3 1: 3 1: 3 1: 3 1: 3 HRT (hrs) 24 24 12 6 6 6 12 24 Feed pattern c* C C C j * * I I I Feed time(h/d) 24 24 24 24 12 8 8 8 Flowrate (L/d) 5 5 10 20 20 20 3.3 1.6 Hydraulic loading rate(L/d L) 1 1 2 4 2 1.3 0.67 0.34 Feed pattern: *C continuous feeding; **I intermittent feeding. Aeration pattern: ***X:Y X hours of aeration and Y hours of non-aeration. Generally, samples of the effluent and influent were collected once a day. Total sample numbers for each run are 6-8. Samples were kept in the refrigerator (temperature about 4°C) for later testing except for T O C and SS testing. T O C and SS were tested right after the sampling. Temperature, p H and D O concentration of the reactor were measured 36 onsite. B O D 5 , C O D , N H / - N , N O x - N , T K N , and P 0 4 - P were determined every week for the samples. The reactors were installed indoors with an ambient temperature o f 20+2°C, but the water temperatures in the reactor were 15-17°C. 3.5. Analytical methods NFLt + -N , N O x - N , T K N , and P 0 4 - P were analyzed by an autoanalyzer (QuickChem 8000), in accordance with Standard methods ( A P H A , 1985). A " T O C 5050 T O C analyzer" was used for the T O C determination. Other parameters, such as BOD5, C O D , TSS, and fecal coliform were determined according to the Standard Methods ( A P H A , 1985). 3.6. Scanning electron microscopic examination After 180 days of continuous experiment, the immobilized activated sludge beads were sampled for scanning electron microscopic examination in the Biosciences Electron Microscopy Facility of the University of British Columbia. 37 C H A P T E R 4 R E S U L T S A N D DISCUSSION 4.1. Batch assays of immobilized sludge materials Important initial steps were to evaluate the feasibility o f the sludge immobilization method, to examine the activity and stability o f the immobilized sludge, through measuring its biodegradation of T O C in synthetic domestic wastewater, and to ascertain that the immobilized sludge had potential for domestic wastewater treatment. 4.1.1. Examination of sludge immobilization methods Preliminary studies on the immobilization of activated sludge were carried out using the methods reported by Hashimoto and Furukawa (1986). Polyvinyl alcohol was chosen as the support material to entrap sludge because o f its lack of toxicity to cells, rapid formation of gel beads, and relatively simple and low cost procedure. Also, it has been reported that the elasticity and high strength of P V A beads are ideally suited to wastewater treatment in fluidized beds (Ariga, 1987). However, a few problems came up when the Hashimoto-Furukawa method was applied to entrap the sludge in our research. Loss of activity of activated sludge during the immobilization process was one o f the major problems. The resulting sludge immobilized by the Hashimoto-Furukawa method had very low biological activity, and almost 95% of the initial activity of activated sludge 38 was lost during the immobilization process (see Figure 4-1). This phenomenon was also observed by other researchers (Chen and L i n , 1993; Hanaki et al, 1994). The principal reason was because saturated boric acid solution is highly acidic and the long time contact between microbial cells and boric acid resulted in the loss of the biological activity o f the activated sludge. Attempts were made to resolve this problem by first adjusting the p H o f the saturated boric acid solution to around 7 with Ca(OH)2 solution and then keeping the contact time of beads with the borate solution within 12 hours. It was hoped that after those modifications, the activity o f immobilized sludge beads produced would be increased, since the immobilization was completed in a more favorable environment to the microorganisms in the activated sludge. The results indicated that with those modifications, the activity o f fresh immobilized sludge increased to about 60% o f that of the initial free activated sludge (see Figure 4-1) The gel beads, however, still had a strong tendency to stick together because of the high viscosity property of P V A . They formed a mass o f polymer that was very difficult to separate. Agglomeration o f the gel beads was another major problem we were facing when the Hashimoto and Furukawa method was used. This phenomenon was observed to occur in the all bead formations in which the final P V A concentration in the beads started from 5%. The higher the final P V A concentration was, the stronger the tendency of bead agglomeration would be. The tendency o f agglomeration of gel beads was so strong that it even occurred in the saturated borate solution when the final P V A 39 concentration in the beads was more than 10%. Such beads were not likely to be useful in either fluidized or fixed bed reactors. The use of alginate has been reported to be an effective method for reducing the tendency toward gel bead agglomeration (Wu et al., 1992). In our present research, alginate was first added into the mixture of P V A and activated sludge. Then the mixture was dropped into the borate solution. It was hoped that calcium alginate might facilitate the formation of beads and improve the surface properties of beads, reducing the tendency for the beads to agglomerate. A total alginate concentration in the beads o f approximately 1% was observed to be approximately the lowest concentration o f alginate that would prevent bead agglomeration and was used in the present experiments. This amount is higher than that suggested by Wu, et al.(1992). T O C removal efficiency was used as a standard to examine the activity o f immobilized sludge and free sludge. The activity of the various immobilized sludge beads, including P V A - b o r i c acid beads, PVA-borate beads and PVA-alginate-borate beads, was tested in the batch reactors. The activity o f free sludge was also tested under the same conditions. The results are shown in Figure 4-1 and 4-2. It can be seen that a fresh preparation of immobilized sludge had a relatively slower T O C removal rate than free sludge (see Figure 4-1). The T O C removal efficiencies o f all the immobilized sludge were lower than that of the free sludge over two days ( see Figure 4-2). The PVA-alginate-borate beads exhibited a faster T O C removal rate and higher 40 T O C removal efficiency than PVA-bor i c acid beads and PVA-borate beads. Furthermore, It was observed that both fresh PVA-bo r i c acid beads and fresh PVA-borate beads were almost totally bound together in the reactors after 2 days of experiments (represented by dotted bars) while PVA-alginate-borate bead still remained excellent shape. 1 00 u > © E t U o PHB bead PNB bead PANB bead Free sludge 10 15 Time (hours) 20 25 Figure 4-1. Time courses of T O C removal of different immobilized sludge and free sludge ( P H B : P V A - b o r i c acid bead; P N B : PVA-borate bead; P A N B : PVA-alginate-borate bead). Results in Figure 4-1 and 4-2 reveal that the activity o f activated sludge could be lost partly or almost totally during the process of immobilization. The p H adjustment o f boric acid could improve the activity o f gel beads immobilized by the P V A - b o r i c acid method and the T O C removal activity o f PVA-borate beads could be improved by adding alginate. The addition of alginate also improved the stability of the PVA-borate beads. It demonstrated that the PVA-alginate-borate method, a modified method from the P V A -41 boric acid method, could provide a feasible method for the immobilization o f activated sludge and so it was applied in our research that followed. 1 0 0 -, O 80 PHB PNB PANB PANB (b) Free Sludge Control Figure 4-2. TOC removal efficiency of different immobilized slduge and free sludge (PHB: PVA-boric acid bead; PNB: PVA-borate bead; PANB: PVA+alginate-borate bead; PANB(b): PVA+alginate-borate blank bead(no sludge); Control: only synthetic wastewater). • TOC removal efficency after 24 hours • TOC removal efficiency after 48 hours The removal of organic carbon by immobilized sludge could be attributed to possible combined biodegradation, adsorption and air stripping i f aeration was applied to reactors for mixing and providing oxygen. T O C removal due to air stripping and adsorption was examined in the batch experiments. The results indicated that T O C removal by air stripping was less than 5% (see data from the control experiment in Figure 4-2) and the total removal percentage of T O C by air stripping and adsorption was less than 10% ( see data from the experiment with P N A B beads in Figure 4-2), indicating that T O C removal due to stripping and adsorption was small in comparison with biodegradation. Therefore, we can conclude that T O C removal by immobilized sludge in this research could be mainly attributed to biological degradation. 42 4.1.2. Impact of P V A contents The polyvinyl alcohol contents in the immobilized sludge play very important roles in bead formation, stability and activity. A series of immobilized sludge beads with different final P V A concentrations were prepared. The beads were further tested in the batch reactors with synthetic wastewater. The effect of P V A concentrations in the gel beads on the bead formation and performance of immobilized sludge was examined as shown in Table 4-1 and Figure 4-3. It was observed that formation of immobilized sludge beads became very difficult when the final P V A concentration was 15%. The mixture of sludge and P V A could not easily be extruded from the syringe and those droplets in the borate solution readily agglomerated together due to high viscosity of the mixture. When the P V A concentration was lower than 7.5, the strength of beads became weak. It was observed that immobilized sludge beads made of 7.5 and 5.0% P V A broke down partly and almost completely after 24 hours of operation in batch reactors, respectively. To balance production, strength and activity o f immobilized sludge, a final P V A content of 10-12.5% was chosen for the immobilization o f sludge in the following experiments. A t higher P V A concentration, the removal efficiency of T O C was decreased because of the high diffusional limitation while the stability of immobilized sludge was reduced due to loose lattice structure at lower P V A concentration. 43 Table 4-1. Effect of final P V A concentrations on the formation of immobilized sludge. Final P V A concentration(w/w) Formation of beads Bead stability in batch experiments 5.0% Not good Not good; beads disintegrated after 24 hours 7.5% Good Good; beads broken partly after 24 hours 10.0% Good Excellent; beads still kept their shape after 15 days 12.5% Good Excellent; beads still kept their shape after 15 days 15.0% Not good Not good; Part of beads bound together after 24 hours 100 -, 0 5 7.5 10 12.5 15 PVA content (%) Figure 4-3. Effect of final PVA concentrations in immobilized sludge on TOC removal efficiencies of fresh beads. 44 4.1.3. Operational stability of immobilized sludge Another important technological question about the immobilized sludge process is the stability o f immobilized sludge beads when they are put into practical application for wastewater treatment. Repeated batchwise experiments were carried out to determine the operational stability of immobilized sludge beads. In the meantime, the same experiments with free sludge were also conducted for comparison. In order to ensure that the results with both the immobilized beads and free sludge were comparable, 25g of immobilized sludge beads containing about 12.5g of activated sludge and 12.5g of free activated sludge were used. The initial volume of beads was approximately 21ml. The results are shown in 100 r P 20 -0 I 1 1 1 1 i 1 1 0 2 4 6 8 10 12 14 16 Tim e (days) Figure 4-4. Changes of T O C removal efficiencies of im m obilized sludge and free sludge during repeated batchwise experiments # Im mobilized sludge U Free sludge 45 Figure 4-4. From the results shown in Figure 4-4, we can see that the activity o f immobilized sludge was no more than 70% of that o f suspended sludge in the first 3 days. After the 5 t h day, the activity o f immobilized sludge had almost fully recovered and reached the same level as that of the initial free activated sludge. This high activity remained stable until the end of the experiments. On the contrary, the T O C removal efficiency of the free sludge decreased after the 3 r d day. The superior performance Of immobilized sludge beads over free sludge may be attributed to the recovery of biological activity and growth of microorganisms in the immobilized sludge. Washout and loss of activity might be the reason for the decrease of organic carbon removal efficiency by the free sludge. Figure 4-5. Changes of immobilized sludge bead volume during repeated batchwise experiment 46 The volume change of immobilized sludge beads during the experiments was also investigated (see Figure 4-5). The volume o f the beads was around 21 ml at the beginning of the experiments and was observed to be increased to 31 ml within the first 3 days. After the third day, the volume remained almost unchanged in the following days o f experiments. The greatest expansion coefficient of immobilized sludge beads was 1.5. However, after 15 cycles of synthetic wastewater treatment, the spherical shape o f the gel beads was still kept, and their elastic properties remained excellent. The above results indicated that sludge immobilized by the PVA-alginate-borate method has good operational stability in the batch reactor. 4.1.4. Effect of wastewater pH on the removal of organic carbon The p H - activity profile of immobilized sludge is shown in the Figure 4-6. The optimum p H value of wastewater for organic carbon removal by immobilized sludge was around 6-8. A t very low p H values, the T O C removal rates decreased because the activated sludge bacteria entrapped in the P V A lost their biological activity during the treatment process. When the p H values were at 9.0, the biological activity o f bacteria in the beads deteriorated and increased leakage of sludge from the beads was also observed as more sludge was seen to be suspended outside of the beads in the wastewater. 47 When p H of wastewater was in the range of 6-8, the T O C removal efficiency o f immobilized sludge changed slightly. It was reported that after immobilization, activated sludge possessed stronger process stability against p H fluctuation (Hanaki, et al. 1994). They also found that immobilizing the bacteria reduced the inhibitory effect of phenol, oleic acid, nickel, sulfide and propionic acid on the bacteria. It is very important for the practical application of the process. 100 -| 4 5 6 7 8 9 pH Figure 4-6. Effect of pH on TOC removal with immobilized sludge De Beer et al. (1992) suggested that diffusion restriction of substances and formation of a p H gradient is the possible mechanism of relief of p H inhibition. They found that 48 there existed such a p H gradient when they measured the gradient of p H within an aggregate of methanogenic bacteria. 4.2.Continuous treatment of domestic wastewater The goal of this study was to test the applicability o f the immobilized sludge process for small scale domestic wastewater treatment. A fluidized bed reactor was used in the present research. There are a number of reasons why a fluidized bed reactor was chosen: (1) a fluidized bed reactor has the potential to effectively and efficiently treat large quantities o f wastewater that is intermittently produced(Sutton and Mishra, 1991; Safferman and Bishop, 1997). (2) Holladay et al. (1978) compared the performance o f an aerobic fluidized bed reactor with that of an aerobic packed bed tower and a stirred tank process for the treatment of phenolic wastewater. The fluidized bed reactor had much higher treatment efficiency than the fixed bed process and the suspended growth process. A s a result, an aerobic fluidized bed can be significantly smaller than other reactors treating the same wastewater. Kondo, et al.(1992) also reported that a fluidized bed reactor could maintain performance while a reduction in the size of individual process and gave higher efficiency o f organic and nutrient removal than a packed bed reactor (Travieso, et al., 1992) (3) an aerobic fluidized bed reactor was demonstrated to very effectively resist and recover from shock organic loading. The choice of the start-up conditions was based on the experiences with the batch reactors for the studies of activity and stability o f the immobilized beads. Steady state 49 organic carbon removal was achieved after a starting-up period of 12 ~ 14 days (see Figure 4-7). This start-up time is shorter than that required for regular activated sludge wastewater treatment system. 0 . 0 -I : , , , , , , 0 10 2 0 3 0 4 0 5 0 6 0 T ime (day) Figure 4-7. Time courses of T O C in the influent and effluent in the first 60 days. The operational conditions of the experiments are listed in the Table 3-2. The data reported of each experimental run were obtained from the average results when steady state was reached. A transient period of average 2-4 days was given to achieve a new Steady state for a subsequent run. Steady state conditions were considered to be attained when the effluent concentrations of T O C were relatively constant. The experimental results obtained from the continuous treatment of domestic wastewater with immobilized sludge under various conditions are presented and discussed in the following. 50 4.2.1. Effect of aeration rates The purpose of this experiment was to investigate the effect o f aeration rate on immobilized sludge process performance. The air was introduced into the reactor from the bottom through air stones. Both oxygen supply and mixing were realized by aeration in the reactor. Aeration rates being investigated in this research were 0.5, 1.0 and 1.5L/min. When the hydraulic retention time was 12 hours and the average organic loading rate about 0.35 kg B O D 5 / d m 3 , the results for organic carbon, TSS, nitrogen and PO4-P removal at different aeration rates are summarized in the Table 4-2, Figures 4-8, and 4-9. The results in Figure 4-8 show that T O C , BOD5, and C O D removal efficiencies remained almost the same when the aeration rate was increased from 0.5 to 1.0 and 1.5 L/min . For example, the average BOD5 removal efficiency was in the range o f 96.5 ~ 97.7% with the average effluent BOD5 concentration 4.18 ~ 6.22mg/L. However, the increase of aeration rate from 0.5 to 1.5L/min resulted in decreasing in TSS removal. The corresponding average TSS removals were 85.6, 79.6, and 63.2%. When the aeration rate was 1.5 L/min. , the TSS concentration in the effluent was more than 20mg/L. TSS concentration in the effluent increases with the increase o f aeration rate. At higher aeration rate, the contents in the reactor are mixed more thoroughly and settling o f suspended solid becomes poorer. Therefore, TSS concentration in the effluent is higher. 51 Table 4-2. Results of domestic wastewater treatment with an immobilized sludge process at different aeration rates. Aeration rate (L/min) 1.5 1.0 0.5 TOC removal (%) 89.9 87.1 85.4 BOD removal (%) 97.7 96.9 96.5 COD removal (%) 93.3 91.6 90.8 NH4 + -N removal(%) 98.6 96.1 93.7 TKN removal (%) 95.1 92.6 90.7 TN removal (%) 33.1 42.0 24.6 TSS removal (%) 63.2 79.6 85.6 PO4-P removal(%) 9.6 8.1 0.8 Effluent concentration TOC (mg/L) 10.77±2.03 13.57±0.74 15.39±1.08 BOD5 (mg/L) 4.18 5.41 6.22 COD (mg/L) 23.3 28.46 31.81 NH 4 + -N (mg/L) 0.36±0.20 1.14±0.23 2.44±2.73 NOx-N (mg/L) 26.59±4;07 28.86±0.57 30.22±9.31 TKN (mg/L) 2.11±0.30 4.85±1.44 4.25±1.86 TN (mg/L) 28.70±4.37 33.71±2.01 34.47±11.17 TSS (mg/L) 24.3±1.90 14.9±1.8 12.2±1.8 P04-P(mg/L) 12.30±1.63 6.96±0.08 6.27±0.86 PH 7.00±0.21 5.52±0.32 6.72±0.28 Figure 4-9 reveals that when aeration rate applied to the reactor was increased from 0.5 to 1.5 L/min. , the average removal efficiencies of NFL; + - N , and T K N increased slightly from 93.7 to 96.1 and 98.6%, 90.7 to 92.6 and 95.1, respectively. Aeration rates did not significantly affect the removal efficiencies of T K N , and NFL;"1" - N . Removal efficiencies o f total nitrogen (TN) and PO4-P were very low and did not significantly benefit from the increase or decrease of aeration rates. 52 0.5 1 1.5 Aeration rate (L/min) Figure 4-8. Removal efficiencies of organic carbon and suspended solids at different aeration rates. • TOC DBOD5 II COD ED TSS 100 Aeration rate (L/min) Figure 4-9. Removal efficiencies of nitrogen and phosphorus at different aeration rates. INH4+-NDTKN STN UP04-P 53 When aeration rates were changed from 0.5 to 1.0 and 1.5L/min, the corresponding dissolved oxygen (DO) concentrations were 4.5 ~ 5.2, 5.8 ~ 6.9, and 7.2 ~ 9.1 mg/L, respectively. The reactor was operated under the high dissolved oxygen conditions when these aeration rates were applied. Therefore, biological oxidation and nitrification in the domestic wastewater were not likely to be limited by the availability of D O . Based on the above results, 1.0 L/min of aeration rate was chosen for the following experiments. Another important reason for choosing l.OL/min.was to make sure that the gel beads in the reactor could be completely fluidized. 4.2.2. Effect of hydraul ic retention time In order to identify the maximum organic loading rate that could be applied to this system, a step increase of loading rate through decreasing the hydraulic retention time (HRT) was performed. Due to fluctuations in the concentrations o f organic carbon in the domestic wastewater, it was difficult to maintain a constant organic loading rate at each hydraulic retention time. It was also impractical to increase the organic loading rate by increasing T O C or BOD5 concentration. The concentration of organic carbon in the domestic wastewater used varied greatly. After pre-settling, the T O C was 41.1 ~ 201.6 mg/L. The T O C concentration of the domestic wastewater that was chosen to be used in this research was 60.1 ~ 201.6 mg/L 54 with an average value of 107.98 mg/L and average T K N , N H 4 + - N and SS values o f 40.0, 30.5, 78.6 mg/L. The reactor was operated at four H R T s (24, 12, 6, and 3 hours). The experimental conditions, average organic loading rates, average T O C , BOD5, C O D , N H V - N , T K N , T N , PO4-P, TSS concentrations and p H in the influent for each H R T are shown in Table 4-3. The average loading rates were increased stepwise from 0.176 to 1.640 kg B O D / d m3 when the H R T s were varied from 24 to 3 hours. The performance in terms o f organic carbon, TSS, nitrogen, and phosphorus removal by immobilized sludge system is shown in Table 4-4, Figure 4-10, 4-11, 4-12, and 4-13. 55 Table 4-3. Operational conditions and average influent concentrations for experiments o f determining the impact of hydraulic retention time (HRTs). HRT (hrs) Hydraulic loading rate(L/d L) TOC loading rate (kg/d m3) BOD loading rate (kg/d m3) COD loading rate (kg/d m ) Influent concentrations TOC (mg/L) BOD (mg/L) COD (mg/L) NH/-N (mg/L) TKN (mg/L) TN (mg/L) TSS(mg/L) P04-P (mg/L) PH 24 1 0.105 0.176 0.343 12 2 0.209 0.350 0.683 6 4 0.456 0.766 1.498 3 8 0.988 1.640 3.217 105.28±33.95 104.79±3.92 114.91±13.49 123.49±9.88 176.01 343.34 27.70±1.20 36.01 ±0.96 36.0110.96 61 ±6 6.97±0.21 7.56±0.36 175.23 341.76 29.44±1.90 58.12±1.31 58.12±1.31 73±16 6.40±0.40 7.14±0.26 191.38 374.45 40.41 ±4.34 54.10±15.77 54.10±15.77 105±7 29.63±0.87 7.16±0.11 205.07 402.17 23.16±0.90 31.24±3.24 31.24±3.24 72±3 8.94±1.31 7.30±0.11 Table 4-4. Results of domestic wastewater treatment with immobilized sludge process at different hydraulic retention time (HRTs). HRT (hrs) 24 12 6 3 TOC removal (%) 83.2 87.1 87.1 83.6 BOD removal (%) 95.9 96.9 96.9 95.9 COD removal (%) 89.5 91.6 91.8 89.8 NH/-N removal(%) 99.0 96.1 91.7 42.4 TKN removal (%) 94.0 92.6 92.9 36.6 TN removal (%) 21.9 42.0 25.8 25.5 TSS removal (%) 66.7 79.6 83.4 51.8 P04-P removal(%) 11.6 8.1 11.0 28.9 Effluent concentrations: TOC (mg/L) 17.64±3.15 13.57±0.74 14.79±2.52 20.31 ±1.92 BOD5 (mg/L) 7.21 5.41 5.95 8.38 COD (mg/L) 35.95 28.46 30.7 40.86 NH4 + -N (mg/L) 0.27±0.22 1.14±0.23 3.37±3.43 13.33±4.90 N03-N (mg/L) 25.96±2.26 28.86±0.57 36.30±2.61 6.09±3.01 TKN (mg/L) 2.16±0.04 4.85±1.44 3.86±2.96 17.17±5.27 TN (mg/L) 28.12±2.30 33.71±2.01 40.16±5.57 23.26±8.28 TSS (mg/L) 16.3±3.2 14.9±1.8 17.4±2.1 34.7±2.4 P04-P (mg/L) 6.16±0.30 6.96±0.08 26.37±2.87 6.36±0.47 pH 7.17±0.10 5.52±0.32 6.63±0.13 7.05±0.01 56 0} 0) 0) 100 80 60 — 40 20 Hydraulic retention time (hrs) Figure 4-10. Effect of HRTs on the removal of organic carbon and suspended solids under continuous aeration condition • TOC DBOD5 HOOD • TSS 100.0 u 80.0 ~ 60.0 it 0) — 40.0 re > o E 20.0 a> 0.0 12 6 Hydraulic retention time (hrs) Figure 4-11. Effect of HRT on the removal of nitrogen and phosphorus under continuous aeration condition INH4+-N DTKN BTN 0PO4-P 57 12 6 Hydraulic retention time (hrs) Figure 4-12. Concentrations of organic carbon and suspended solids in the influent and effluent at different HRTs under continuous aeration condition • BODS(inf) OBOD5(eff) HTSS(inf) STSS(eff) 10 15 20 Hydraulic retention time (hrs) Figure 4-13. Concentrations of nitrogen in the influent and effluent at different HRTs under continuous aeration condition •NH4+ -N(inf) -TKN(eff) -NH4+ -N(eff) -TN(inf) -TKN(inf) -TN(eff) 58 The results show that the average organic carbon removal efficiencies were maintained in the range of 83.2-87.1% (TOC), 95.9- 96.9% ( B O D 5 ) and 85.9- 91.8(COD) when the H R T s were decreased from 24 to 3 hours and average organic loading rates were increased from 0.176 to 1.640 kg B O D / d m 3 . The BOD5 concentrations in the effluent were all below the target BOD5 concentration of 20 mg/L. The suspended solids (TSS) concentrations in the pre-settled domestic wastewater ranged from 61-105 mg/L. After treatment with H R T s of 6-24 hours, the TSS concentration was reduced to below 20 mg/L. when the H R T was decreased to 3 hours, the TSS concentration in the effluent increased to 34.7 mg/L. Ammonium-nitrogen was the main form of the nitrogen in the primary settled domestic wastewater. The average ammonia loading rates in this research were in the range of 0.028-0.185 kg NH4 + -N/m3.d, The average removal efficiency o f N H » + - N was above 90% with M L ; + - N concentrations of 0.27-3.37 mg/L in the effluent when H R T was more than 6 hours. However, when the H R T was decreased from 6 to 3 hours, The effluent N H 4 + - N concentration increased to as high as 13.33 mg/L and the average removal efficiencies of N H 4 + - N dropped to 42%. The decrease in N H / - N removal was a result of incomplete nitrification at the H R T of 3 hours. This conclusion is supported by the evidence that the effluent N O x - N concentration was apparently lower at the H R T o f 3 hours than those at H R T s of 24, 12, and 6 hours. 59 Similarly, T K N removal of domestic wastewater with immobilized sludge process was very effective with the average T K N concentrations o f 2.16-4.85 mg/L in the effluent when H R T s were between 6-24 hours. However, when H R T was decreased to 3 hours, the average removal efficiency of T K N was only 36.6% and T K N concentration in the effluent was as high as 17.17 mg/L due to the decrease of N H 4 + - N removal efficiency. T N removal efficiencies were low at all organic loading rates in this experiment because of continuous aeration and high dissolved oxygen concentrations in the reactor. The highest average T N removal efficiency, 42%, appeared when the H R T was 12 hours with the average T N concentration in the effluent of 33.71 mg/L. The average P0 4 -P concentrations in the influent varied with the concentrations o f 23.16-40.41 mg/L. There seemed to be an increasing trend for PO4-P removal when the H R T varied from 2 4 - 3 hours. However, PO4-P removal level was very low regardless of what H R T s were applied in the research because there is no sludge wasting operation in this process. Therefore, i f organic carbon, NH4 + -N and T K N removal are to be maintained at high levels (more than 90%), the H R T should not exceed 6 hours or the organic loading rates applied to this immobilized sludge treatment system should not be over 0.766 kg B Q L V d m 3 . T N removal efficiency was very low and N O x - N concentration in the effluent was still high irrespective of what H R T was applied under complete aeration condition. Further studies to improve the T N removal are absolutely necessary. 60 4.2.3. Effects of intermittent aeration From the previous experimental results, we can see that immobilized sludge process is able to remove organic carbon, N H 4 - N and TSS very efficiently in a fluidized bed reactor. However, in terms o f total nitrogen (TN) removal, the removal efficiency was not satisfactory. The main reason is because most of N H » + in the domestic wastewater was oxidized into N O x by nitrification. Although the N H 4 + - N concentration was reduced, accumulation of N 0 3 - N resulted in a considerable amount of nitrogen existing in the effluent. Under continuous aeration conditions, the occurrence of denitrification was inhibited greatly due to the very high dissolved oxygen concentration. Only a small amount of N O x - N was further reduced to N 2 in the immobilized sludge process. The accumulation of N O x during wastewater treatment process, resulting in a low T N removal efficiency. A strategy of using intermittent aeration was suggested to improve nitrogen removal and realize the simultaneous removal of carbon and nitrogen in a single reactor (Kondo et al., 1992; Yang et al., 1997). It was assumed that the activated sludge process, especially the immobilized sludge process, had potential for the accomplishment o f biological oxidation, nitrification and denitrification in a single reactor. The purpose of the following experiments was to investigate the effect o f intermittent aeration on the organic carbon and nitrogen removal. It was hoped that nitrogen removal 61 could be improved through intermittent aeration without having an adverse impact on organic carbon removal capacity. The experiments were conducted at an H R T o f 24 hours under different intermittent aeration conditions. For comparison, an experiment with continuous aeration was also conducted under the same conditions. The experimental conditions are summarized in the Table 4-5 and results are shown in Table 4-6, Figures 4-14, 4-15, 4-16 and 4-17. Under a continuous aeration condition, the immobilized sludge process exhibited very high removal efficiencies of organic carbon, N H 4 + - N and T K N (see Figure 4-14 and 4-16). However, the average total nitrogen (TN) removal was only 21.9% with a T N concentration of 28.12 mg/L in the effluent. When intermittent aeration was applied to the treatment system and the aeration patterns were changed from 1:1 (1 hour of aeration and 1 hour of non-aeration), 1:2 (1 hour of aeration and 2 hours of non-aeration) and 1:3 (1 hour of aeration and 3 hours of non-aeration), the average T N removal efficiency increased from 32.0 to 46.1 and 62.5%. The T N removal efficiency under 1:3 intermittent aeration condition was almost three times higher than that under continuous aeration condition. In the meantime, the organic carbon, N H V V N and T K N removals were not significantly affected. 62 Table 4-5. Operational conditions and influent concentrations determining the impact of aeration patterns. for experiments of Aeration pattern 1,0 1. 1 1,2 1,3 1,4 1.5 TOC loading rate (kg/d m3) 0.105 0.105 0.107 0.103 0.101 0.123 BOD loading rate (kg/d m3) 0.176 0.176 0.178 0.172 0.170 0.205 COD loading rate (kg/d m3) 0.343 0.343 0.348 0.334 0.331 0.402 Influent concentration TOC (mg/L) 105.28 105.28 106.61 102.67 101.43 123.49 ±33.95 ±33.95 ±17.14 ±3.44 ±12.91 ±9.88 BOD5 (mg/L) 176.01 176.01 178.14 171.84 169.87 205.07 COD (mg/L) 343.34 343.34 347.64 334.91 330.91 402.17 NH/-N (mg/L) 27.70 27.70 26.59 28.90 25.28 23.16 ±1.20 ±1.20 ±2.03 ±2.09 ±2.01 ±0.90 TKN (mg/L) 36.01 36.01 42.87 58.12 34.87 31.24 ±0.96 ±0.96 ±5.16 ±1.31 ±2.56 ±3.24 Org-N(mg/L) 8.29±2.16 8.29±2.16 16.28±7.2 29.22±3.4 9.59±4.57 8.08±4.14 NOx-N(mg/L) 0 0 0 0 0 0 TN(mg/L) 36.01 36.01 42.87 58.12 34.87 31.24 +0.96 ±0.96 ±5.16 ±1.31 ±2.56 ±3.24 TSS(mg/L) 61 ±6 61 ±6 66±13 73±16 62±8 72±3 P04-P(mg/L) 6.97±0.21 6.97±0.21 13.6±1.45 7.54±1.22 2.36±1.52 8,94±1.31 PH 7.56±0.36 7.56±0.36 7.47±0.31 7.15±0.24 7.31±0.04 7.30±0.11 Table 4-6. Results of domestic wastewater treatment with immobilized sludge process at different aeration patterns. Aeration pattern 1,0 1, 1 1,2 1,3 1,4 1.5 TOC removal (%) 83.2 84.1 90.4 89.3 84.8 84.0 BOD removal (%) 95.9 93.2 96.2 95.7 93.6 93.1 COD removal (%) 89.5 80.3 88.4 87.0 81.2 80.1 TKN removal (%) 94.0 93.3 95.1 96.5 83.8 52.8 TN removal (%) 21.9 32.0 46.1 64.2 47.0 24.6 NhV-N removal(%) 99.0 98.1 98.6 95.4 88.6 45.4 PO4-P removal(%) 11.6 15.6 14.3 15.1 30.1 33.7 TSS removal(%) 66.7 78.9 86.5 95.6 85.2 89.3 Effluent concentration TOC (mg/L) 17.64±3.15 16.77±2.69 10.25±1.27 10.99±0.65 15.46±1.84 19.73±2.74 BOD5 (mg/L) 7.21 11.91 6.76 7.34 10.88 14.25 COD (mg/L) 35.95 67.66 40.42 43.51 62.18 80.02 NH/-N (mg/L) 0.27±0.22 0.52±0.67 0.38±0.17 1.33±0.52 2.84±4.02 12.65±2.52 TKN (mg/L) 2.16±0.04 2.42±0.43 2.08±0.22 2.08±0.25 5.65±2.02 14.74±2.52 Org-N(mg/L) 1.89+0.26 1.90±1.10 1.70±0.39 1.75±0.77 2.81±6.04 2.09±5.14 NOx-N(mg/L) 25.96±2.26 22.05±1.36 21.02±3.60 18.71±1.59 12.82±4.23 8.82±2.12 TN(mg/L) 28.12±2.30 24.47±1.79 23.10±3.82 20.79±1.84 18.47±6.25 23.56±4.64 TSS(mg/L) 16.3±3.2 12.9±1,3 8.9±1.7 2.9±0.7 9.2±0.7 7.7±1.1 P04-P(mg/L) 6.16±0.30 5.88±0.56 11.66±0.53 6.40±0.40 1.67±1.33 5.93±0.87 PH 7.17±0.10 7.24±0.23 7.02±0.20 6.40±0.16 6.86±0.06 7.17±0.03 63 1,0 1,1 1,2 1,3 1,4 1,5 Aeration patterns Figure 4-14. Effect of aeration patterns on the removal of organic carbon and suspended solids • TOC DBOD5 UCOD II TSS 100 -| 1,0 1,1 1,2 1,3 1,4 1,5 Aeration patterns Figure 4-15. Concentrations of organic carbon and suspended solids in the effluent vs aeration patterns • TOC DBOD5 UCOD BTSS 64 1 , 0 1,1 1 , 2 1 , 3 1 , 4 1 , 6 Aeration patterns Figure 4-16. Effect of aeration patterns on the removal of nitrogen and phosphorus • NH4-N rJTKN HITN HP04-P 30 E c o 2 c V o c o O 25 20 15 10 5 0 1,0 1,1 1,2 1,3 Aeration pattern 1,4 1,5 Figure 4-17. Concentrations of nitrogen and phosphorus in the effluent vs aeration patterns INH4-N D T K N H T N B P 0 4 - P 65 When the ratio of aeration and non-aeration time continued to be decreased from 1:3 to 1:4 and 1:5, organic carbon removal decreased slightly. The concentration of organic carbon in the effluent increased, but was still very low (Figure 4-14 and 4-15). Average N H 4 + - N and T K N removal efficiencies, however, dropped rapidly, from 95.4 to 88.6 and 45.4% for N H / - N and from 96.5 to 83.8 and 52.8% for T K N . It is interesting that the average T N removal decreased from 64.2 to 47.0 and 24.6% with increases of the non-aeration time from 3 to 4 to 5 hours, although the N O x - N concentration in the effluent decreased (see Figure 4-16, 4-17 and 4-18). The reason was that N H 4 + - N removal efficiencies deteriorated dramatically with increases of the non-aeration time from 3 to 4 and 5 hours. The average effluent concentrations o f N H 4 + - N increased significantly from 1.33 to 2.84 and 12.65 mg/L. Generally, TSS removal increased as the amount of non-aeration time increased. After the non-aeration time increased to 3 hours, average TSS removal remained in the range of 85.2 ~ 95.6% and the average TSS concentrations in the effluent were all below 20 mg/L. PO4-P removal did not show any significant trend with the increase o f non-aeration time. The maximal average removal efficiency was only 33.7%. This should not come as a surprise as there are no sludge wasting operation in the immobilized sludge process. 66 The D O concentration in the reactor varied greatly when intermittent aeration was applied to the treatment system. It could be as high as 8.92 mg/L during the aeration period and as low as 0.4 mg/L at the end of the non-aeration period. It was observed that the D O concentration could be reduced to 0.4 - 0.6 mg/L from 8.92 mg/L within 1 hour when aeration was stopped. Further reduction of the D O concentration to around 0 mg/L was observed in the bottom of reactor when the non-aeration time was more than 1 hour. It indicated that the intermittent aeration could produce aerobic, anoxic, and anaerobic conditions in a single reactor and improve the removal of nitrogen through both nitrification and denitrification. 30 -] 25 -1,0 1,1 1,2 1,3 1,4 1,5 Aeration pattern Figure 4-18. Components of nitrogen in the effluent vs aeration patterns. • NH4+-N II Org-N • NOx-N 67 Experiments were also conducted to investigate the effect of organic loading rates on the performance of immobilized sludge process at the ratio of aeration and non-aeration 1:3 by 1 hour of aeration and 3 hours of non-aeration. The operational conditions and the results are shown in Table 4-7, 4-8, Figure 4-19, 4-20, and 4-21. A s shown in Figure 4-19, the high removal efficiencies of organic carbon were maintained when the aeration pattern was 1:3 and the average organic loading rates increased from 0.175 to 0.353 to 0.766 kg B O D / d m 3 . Although the B O D 5 in the effluent increased with the increase of organic loading rates, they were all lower than 15 mg/L (see Table 4-8). The TSS concentrations in the effluent also increased as organic loading rates increased and were all lower than 15 mg/L. For nitrogen removal (see Figure 4-20), both T K N and N H 4 + - N removal efficiencies decreased when the organic loading rates were increased. However, the average T N removal efficiencies increased from 64.2 to 67.5 to 74.4% when average organic loading rates were increased from 0.175 to 0.353 to 0.766 kg BODs /d m 3 . The corresponding average effluent T N concentrations were decreased from 20.79 to 14.85 to 13.85 mg/L. The main reason for this was that N O x - N removal was improved when organic loading rates were increased. Figure 4-20 shows that average effluent N O x - N concentrations decreased from 18.71 to 11.57 to 6.18 mg/L. We can conclude that with intermittent aeration, the total nitrogen removal could be improved without reducing the removal efficiency o f organic carbon in a simple 68 immobilized sludge process. Nitrification and denitrification are thought to simultaneously occur in the same system because o f the presence o f aerobic and anaerobic conditions and regions in the immobilized sludge system. 69 Table 4-7. Operational conditions and influent concentrations for experiments of determining the effect of organic loading under intermittent aeration condition. Aeration pattern 1, 3 1,3 1,3 HRT (hrs) 24 12 6 Hydraulic loading rate(L/d L) 1 2 4 TOC loading rate (kg/d m3) 0.103 0.212 0.435 BOD loading rate (kg/d m3) 0.172 0.353 0.727 COD loading rate (kg/d m3) 0.334 0.689 1.419 Influent concentration TOC(mg/L) 102.67±3.44 105.62±13.74 108.84±11.37 BOD5(mg/L) 171.84 176.56 181.69 COD(mg/L) 334.91 344.44 354.84 NH4 + -N(mg/L) 28.90±2.09 38.86±5.97 38.60±1.66 TKN (mg/L) 58.12i1.31 45.74±9.63 54.1 OH 5.77 Org-N(mg/L) 29.22l3.40 6.88±15.60 15.50117,41 NOx-N(mg/L) 0 0 0 TN(mg/L) 58.12±1.31 45.74±9.63 54.10115.77 TSS(mg/L) 73±16 85±9 105±7 P04-P(mg/L) 7.54±1.22 6.22±0.24 29.63±0.87 PH 7.15±0.24 7.23±0.12 7.1210.11 Table 4-8. Results of domestic wastewater treatment with immobilized sludge process under intermittent aeration condition. HRT (hrs) 24 12 6 Hydraulic loading rate(L/d L) 1 2 4 TOC removal (%) 89.3 86.2 85.0 BOD removal (%) 95.7 94.2 93.6 COD removal (%) 87.0 83.0 81.4 NH4 + -N removal(%) 95.4 96.8 98.7 TKN removal (%) 96.5 92.8 85.8 TN removal (%) 64.2 67.5 74.4 P04-P removal(%) 15.1 0 22.2 TSS removal(%) 95.6 92.7 88 Effluent concentration TOC(mg/L) 10.9910.65 14.5711.46 16.3511.55 BOD5(mg/L) 7.34 10.17 11.58 COD(mg/L) 43.51 58.46 65.90 NH4 + -N(mg/L) 1.3310.52 1.2312.14 0.4910.09 TKN (mg/L) 2.0810.25 3.2811.89 7.6717.73 Org-N(mg/L) 1.75+0.77 2.0514.03 7.1817.82 NOx-N(mg/L) 18.7111.59 11.5716.30 6.1814.69 TN(mg/L) 20.7911.84 14.8518.19 13.85112.42 TSS(mg/L) 2.910.7 6.2H.2 12.612.2 P04-P(mg/L) 6.4010.40 6.2210.31 23.0413.68 pH 6.4010.16 7.0310.37 7.2510.19 70 100.0 -, 0.0 -J r- , , , • , - , —r- ' , 0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 BOD loading rates (kg BOD/d m3) Figure 4-19. Effect of organic loading on the removal of organic carbon and suspended solids under intermittent aeration condition - • - T O C - B - B O D 5 - • - C O D - * - T S S re E 20.0 -0.0 -I : - i : T 1 1 1 1 1 1 0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 BOD loading rates (kg BOD/d m3) Figure 4-20. Effect of organic loading on the nitrogen removal under intermittent aeration condition - • - N H 4 + -N - " - T K N - » - T N 71 O) E "5f c o 0) U c o u c 0) o 30 25 20 15 10 5 0 0.172 0.353 0.727 BOD loading rates (kg BOD/d m3) Figure 4-21. Components of nitrogen concentrations in the effluent at different organic loading rates under intermittent aeration condition • NH4+-N(eff) •Org-N(eff) EH NOx-N(eff) Nitrification and denitrification are the easiest and most economical methods currently available for removal of nitrogen from the wastewater. In principle, they can not occur in the same location at the same time. However, many efforts have been made to develop a novel method to allow their concurrence in the same system under aerobic conditions (Kondo, et al., 1992; Kokufuta, et al., 1987; Patureau, et al., 1997). A few researchers have reported that the nitrogen removal was improved by changing aeration patterns and simultaneous removal of carbon and nitrogen in one single reactor could be realized (Karnchanawang and Polprasert, 1990; Kondo et al., 1992; Yang, 1997). 72 It was reported that in activated sludge reactors, nitrogen removal efficiencies of 70% were achieved at H R T of 16 hours (Taniguchi et al., 1989). B y addition of immobilized bacteria to the reactors it was possible to reduce the average hydraulic retention time needed for the treatment from 16 hours to 6-8 hours (Tanaka, 1991). This research has revealed that the immobilized sludge system has potential for simultaneous removal of organic carbon and nitrogen from domestic wastewater. Maximal average removal efficiency of T N was 74.4% at an H R T of 6 hours with an average N O x - N concentration of 6.18mg/L and a T N concentration of 13.85 mg/L in the effluent. Nitrogen removal in the domestic wastewater, however, was incomplete. The limitation o f the carbon source and electron donors in the domestic wastewater is the main reason resulting in incomplete T N removal, because the BOD5: N ratio in the domestic wastewater in this research was only 100:22. Other reasons include excessive aeration and loading variations. 4.2.4. Impact of wastewater feed patterns Actual flow patterns of domestic wastewater produced from single point source vary from place to place. The purpose of this research was to examine the performance o f the immobilized sludge process when different feed patterns were applied. The following wastewater feed patterns were used in this research: A . Continuous feed (24 hours/d) with an H R T of 24 hours; B . Intermittent feed - feed time: 8am - 8pm (12 hours/d) with an H R T of 12 hours; 73 C. Intermittent feed - feed time: 8am - 12 pm and 6pm - 10pm (8 hours/d) with an H R T of 6 hours. The experiments were conducted under intermittent aeration condition (1:3). The operational conditions and experimental results are shown in Tables 4-9, 4-10, Figures 4-22, 4-23, 4-24 and 4-25. It was noted that the removal efficiencies of organic carbon, suspended solids, N H V -N and T K N were not significantly affected whether changing the feed pattern to intermittent feed at the same H R T or increasing the H R T at the same intermittent feed pattern. Average B O D removal efficiencies were maintained in the range of 90.9 ~ 96.9% with the average BOD5 concentration in the effluent being less than 20 mg/L (see Figure 4-21, 4-23 and Table 4-10). Average TSS removal efficiencies were between 81.2-92.6% and the highest average concentration of TSS in the effluent was only 16.4 mg/L. Average N H 4 + - N and T K N removal efficiencies stayed at high levels in the range of 90.2-98.9%, regardless of influent feeding patterns (see Figure 4-22). The average T N removal was 74.4%, when domestic wastewater was continuously introduced into the reactor. With intermittent feeds of 12 and 8 hours, the average T N removal efficiency decreased to 55.1 and 28.4% (see Figure 4-23). When the feed time was kept at 8 hours and the H R T was increased to 12 hours, the reduction of T N improved and the average T N removal efficiency increased to 47.9%. However, a further increase of H R T to 24 hours at a feed time of 8 hours did not improve T N removal efficiency significantly (see Figure 4-25). 74 Table 4-9. Operational conditions and influent concentrations for experiments of determining the impact of feed patterns under intermittent aeration condition. Feed time(h/d) 24 12 8 8 8 HRT (hrs) 6 6 6 12 24 Flowrate (L/d) 20 10 6.5 3.3 1.6 Flowrate (L/h) 0.83 0.83 0.83 0.42 0.21 Hydraulic loading rate(L/d L) 4 2 1.3 0.67 0.34 TOC loading rate (kg/d m3) 0.435 0.234 0.188 0.078 0.049 BOD loading rate (kg/d m3) 0.727 0.389 0.311 0.130 0.081 COD loading rate (kg/d m3) 1.419 0.762 0.612 0.255 0.160 Influent concentration TOC (mg/L) 108.84 116.90 144.79 116.90 144.79 ±11.37 ±34.23 ±12.38 ±34.23 ±12.38 BOD5 (mg/L) 181.69 194.55 239.06 194.55 239.06 COD (mg/L) 354.84 380.88 470.98 380.88 470.98 NH/-N (mg/L) 38.60±1.66 36.03±3.20 30.27±4.78 36.03±3.20 30.27±4.78 TKN(mg/L) 54.10±15.77 54.80±6.59 37.24±6.93 54.80±6.59 37.24±6.93 Org-N(mg/L) 15.5+17.43 17.25±9.79 6.93±11.71 17.25±9.79 6.93±11.71 TN(mg/L) 54.10±15.77 54.80±6.59 37.24±6.93 54.80±6.59 37.24±6.93 TSS(mg/L) 105±7 87±11 96±7 87±11 96±7 P04-P(mg/L) 29.63±0.87 10.20±0.83 13.37±0.35 10.20±0.83 13.37±0.35 PH 7.12±0.11 7.75±0.15 7.74±0.34 7.75±0.15 7.74±0.34 Table 4-10. Results of domestic wastewater treatment with immobilized sludge process at different feed patterns under intermittent aeration condition. Feed time(h/d) 24 12 8 8 8 HRT (hrs) 6 6 6 12 24 TOC removal (%) 85.0 86.2 83.3 84.4 87.3 BOD removal (%) 93.6 90.9 92.6 96.2 96.9 COD removal (%) 82.9 79.1 80.1 90.3 92.1 NH4+-N removal(%) 98.7 96.6 90.2 93.2 98.9 TKN removal (%) 85.8 93.3 88.0 88.2 90.4 TN removal (%) 74.4 55.1 28.4 47.9 48.6 P04-P removal(%) 22.2 31.6 1.9 11.5 23.8 TSS removal (%) 88 81.2 86.0 86.0 92.6 Effluent concentration TOC (mg/L) 16.35±1.55 16.19±0.68 24.16±9.99 18.18±2.56 18.36±2.19 BOD5 (mg/L) 11.58 11.45 17.75 7.44 7.53 COD (mg/L) 65.90 65.23 98.53 36.94 37.28 NH4+-N (mg/L) 0.49±0.09 1.23±0.55 2.97±1.12 2.45±0.73 0.34±0.18 TKN (mg/L) 7.67±7.73 3.69±2.37 4.47±1.34 6.45±2.31 3.57±0.84 Org-N (mg/L) 7.18±7.82 2.46±2.92 1.50±2.46 4.00±3.04 3.23±1.02 NOx-N (mg/L) 6.18±4.69 20.94±2.87 22.17±2.95 22.11 ±3.40 15.56±3.35 TN (mg/L) 13.85 24.63 26.64 28.56 19.13 ±12.42 ±5.24 ±4.29 ±5.71 ±4.19 TSS (mg/L) 12.6±2.2 16.4±1.2 13.4±1.7 12.2±1.78 7.1±0.2 P04-P (mg/L) 23.04±3.68 6.98±1.46 13.11±0.62 9.03±2.03 10.19±4.49 PH 7.25±0.19 7.14±0.10 7.05±0.03 6.52±0.13 6.86±0.17 75 100 .0 Feed time [hid) Figure 4-22. Effect of the influent feed patterns on the removal of organic carbon and suspended solids at HRT of 6 hours. ITOC • BOD5 U C O D ITSS — . 1 0 0 Feed time (h/d) Figure 4-23.Effect of the influent feed patterns on the removal of nitrogen at HRT of 6 hours. INH4-N CITKN U T N 76 1 0 0 . 0 Hydraulic retention time (hrs) Figure 4-24. Effect of influent feed patterns on the removal of organic carbon and suspended solids at feed time of 8 h/d and different HRTs. • TOC DBOD5 UCOD HTSS Hydraulic retention time (hrs) Figure 4-25. Effect of influent feed patterns on the removal of nitrogen at feed time of 8 h/d and different HRTs INH4-N DTKN UTN 77 There seemed no apparent trend for PO4-P removal when the influent feed patterns were changed. Besides, the removal efficiency of PO4-P remained at a very low level. The p H of the effluent was in the range of 6.5 ~ 7.2. It was demonstrated that different influent feed patterns of 24 hour continuous feed and 16 and 8 hours per day intermittent feed did not have much adverse impact on the removal of organic carbon, N F J V - N , T K N , and suspended solids. The efficiency o f total nitrogen (TN) reduction dropped when intermittent feed patterns were applied. In order to have better N O x - N removal, a continuous feed o f wastewater into the reactor is recommended. 4.2.5. Excess sludge production The production of excess sludge by the immobilized sludge process can be calculated according to the relationship between the growth rate of activated sludge and the removal rate of organic materials (Metcalf & Eddy, 1991). This relationship is A S / A t = a A S r / A t - b S (1) where AS/At = production of excess sludge (g SS/ L d), ASr/At = removed rate o f B O D (g B O D / L d), S = activated sludge concentration (g/L), a = yield coefficient (g SS/g B O D removed), b = sludge decay constant (d _ 1 ) . 78 Based on the method introduced by Hashimoto and Furukawa (1987), during the continuous treatment, S can be expressed as S = So + ASi + A S f (2) where So = initial concentration of immobilized sludge, ASi = increased concentration of activated sludge inside the beads, ASf = increased concentration of activated sludge outside the beads. Since the increase in the amount of activated sludge inside the immobilized sludge beads is small compared that of activated sludge outside the beads, equation (2) can be rewritten as S = So + ASf (3) Therefore, AS = ASf (4) The material balance for immobilized sludge in the reactor is AS/At = (AS/At) v + SeQ/V (5) 79 where (AS/At) v = concentration change o f activated sludge in the reactor, Se = effluent suspended solids concentration, Q - influent flow rate (L/d), V = reactor volume (L). Combining equations (4) and (5) gives A S f / A t = A S f / A t ) v + SeQ/V (6) Therefore, the production of excess sludge in the immobilized sludge process is the sum of the weight of the suspended activated sludge in the reactor, the weight of the activated sludge attached to the beads, and the weight of suspended solids in the effluent per unit of time. Furthermore, equation (1) is rewritten as Two plots drawn from experimental data ASf /At and the corresponding ASr/At for experiments under continuous and intermittent aeration (see Tables 4-11 and 4-12) are presented in the Figure 4-26. ASf /At = a ASr/At - b S (7) 80 Table 4-11. Experimental and calculating results of excess sludge production under continuous aeration condition. Wastewater retention time (hr) 24 12 6 3 Experimental time(days) 9 7 11 7 Average influent BOD5 Conc.(mg/L) 176.01 175.23 191.38 205.07 Average effluent BOD5 Conc.(mg/L) 7.21 5.41 5.95 8.38 ASr/At(g B O D / L d) 0.169 0.34 0.741 1.573 Average influent TSS (mg/L) 61 73 105 72 Average effluent TSS (mg/L) 16.3 14.9 17.4 34.7 Treated wastewater volume(L) 45 70 220 280 Increased SS in the reactor(g) 0.211 0.532 0.536 0.714 ASf/At(gSS/L d) 0.021 0.045 0.167 0.298 Table 4-12. Experimental and calculating results of excess sludge production under intermittent aeration condition. Wastewater retention time (hr) Experimental time(days) Average influent BOD5 Conc.(mg/L) Average effluent BOD5 Conc.(mg/L) ASr/At(g B O D / L d) Average influent TSS (mg/L) Average effluent TSS (mg/L) Treated wastewater volume(L) Increased SS in the reactor(g) ASf/At(gSS/L d) 24 12 6 7 12 8 171.84 176.56 181.69 7.34 10.17 11.58 0.164 0.333 0.680 73 85 105 2.9 6.2 12.6 35 120 160 0.808 3.216 4.064 0.026 0.066 0.152 81 0.35 -, Removed BOD(g BOD/L d) Figure 4-26. Determination of excess sludge production of immobilized sludge process A s shown in Figure 4-26, the yield coefficients for domestic wastewater under continuous and intermittent aeration are 0.20 and 0.24g SS/g B O D removed, respectively. The yield coefficients obtained from the conventional activated sludge process are 0.3-0.5 g M L S S / g B O D removed (Sumino, et al, 1993). Therefore, the production of excess sludge by the immobilized sludge process under both continuous and intermittent aeration in this research is about 1/2 - 1/3 that produced by the conventional activated sludge process. The mechanism behind the reduction in excess sludge production is still unclear. Sumino, et al.(1993) assumed that less excess sludge production in the case of immobilized sludge could be attributed to the prolonged sludge age and its easier autolysis under the circumstances of immobilized cells. 82 The yield coefficient of the immobilized sludge process operated under continuous aeration is smaller than under intermittent aeration, which indicates that the production of excess sludge is least under continuous aeration conditions, compared to different intermittent aeration patterns. 4.2.6. S E M examination After 180 days of continuous treatment of domestic wastewater, the immobilized sludge beads were sampled for scanning electron microscopic examination. Both rod- and cocci-shaped bacteria were found on the bead surface, but rod-shaped bacteria were predominant (Figure 4-27a, b). In the interior of beads, the density o f bacteria was very low (Figure 4-27c). Only a few cocci bacteria were observed to be distributed close to the inner surface of beads, This probably resulted from the availability o f substrates and nutrients. It is suggested that a layer of bacteria gradually grow on the surface of the beads. This layer of bacteria enhance the mass-transfer resistance into the beads and meanwhile, use up most of the substrate and nutrients, leaving very little substrate and nutrients for those bacteria in the interior. A s a result, the P V A beads gradually became a biofdm support. Similar results have also been reported by other researchers (Wijffels,et al., 1991; Fang and Zhou, 1997). 83 (a) Figure 4-27. Scanning electron micrographs: (a) bacteria on the bead surface (x 3,000), and (b) on the bead surface (x 5,000), (c) bacteria in the interior (x3,000). 84 4.2.7. Technical evaluation During the 180 days of continuous operation, the immobilized sludge process exhibited a high operational stability. The reactor required only minor maintenance under the operating conditions employed in this research. A comparison of the PVA-immobi l ized sludge process in this research with other biological treatment processes for domestic wastewater treatment is given in the Table 4-13. Table 4-13. Comparison of wastewater treatment performance of several processes Processes Typical loading rates ( B O D 5 g/m 3 d) BOD5 or C O D removal Nitrogen removal References Fluidized bed A S <1,200 9 0 % ( B O D 5 ) 70% a P B S system 800 93% (BOD5) 76% b Fluidized bed P V A -immobilized sludge 727 94% (BOD5) 74% This study a Fdez-Polanco and Garcia (1994) b Kondo, et al. (1992), PBS- porous biomass support system Based on the comparison between the PVA-immobi l ized sludge process and other biological treatment processes, the loading rates and removal efficiencies o f organic carbon and nitrogen of PVA-immobi l ized sludge process studied in this research are similar to or greater than those of other processes. Cost is another great concern. A n economic evaluation undertaken by Hashimoto and Furukawa (1987) showed that the PVA-immobi l ized sludge method is a relatively cheap 85 method, based on calculation of chemical cost requirements compared with other immobilization methods. Compared with conventional activated sludge process, the initial cost of immobilized sludge process for sludge immobilization is usually high. However, it was estimated that the additional expenses for sludge immobilization could be compensated for by the reduced costs for building a smaller reactor and requiring less operating and management (Leenen et al., 1996; Hashimoto and Furukawa, 1987). Since the treatment system in the present research was small, a comparison of operational and maintenance cost with other established treatment systems is not realistic. Further economic and technical evaluation for the application of the immobilized sludge process to small scale domestic wastewater treatment require pilot plant scale research. 86 C H A P T E R 4 C O N C L U S I O N S A N D R E C O M M E N D A T I O N 4.1. Conclusions Based on the experimental results obtained, the following conclusions are made: 1. The PVA-alginate-borate method, a modified P V A - b o r i c acid method, was used for the preparation of immobilized sludge. Batch studies showed that sludge immobilized by this method had superior performance in both activity and stability over those prepared by the PVA-bo r i c acid method. 2. The P V A concentrations in the immobilized sludge play a very important role in bead formation, stability and activity. When the final P V A concentration in the immobilized sludge was 10 - 12.5%, immobilized sludge beads exhibited higher activity and better mechanical stability. The existence o f alginate was able to reduce the bead agglomeration and 1% of final alginate concentration was the lowest one that would effectively prevent bead agglomeration. 3. The operational stability o f immobilized sludge proved to be superior to that of free activated sludge by repeated batchwise experiments. The optimum p H o f solution for the application of immobilized sludge was in the range of 6 ~ 8. 87 4. The immobilized sludge process exhibited high efficiencies for organic carbon and nitrogen removal. When the aeration rate varied in the range o f 0.5 ~ 1.5L air/min., the average removal efficiencies of organic carbon, N H V - N and T K N were not affected significantly. A n aeration rate of 1.0 L/min . was chosen for the rest o f research to make sure that the immobilized sludge beads in the reactor could be completely fluidized. 5. Under continuous aeration conditions, the variable H R T (24, 12, 6 and 3 hours) experiments conducted on the immobilized sludge process showed that under B O D loading ranges of 0.176 ~ 0.766 kg/d m3, more than average 90% of B O D 5 , N H ^ - N and T K N were removed with the good effluent quality. 6. The maximal removal rate of total nitrogen (TN) under continuous aeration conditions was found to be 42.0% at an H R T of 12 hours. Compared with the continuous aeration system of immobilized sludge, the approach for the intermittent aeration system provided a more effective method for nitrate removal. Intermittent aeration could improve T N removal. The best T N removal was above 74.4% and was achieved at an H R T of 6 hour (corresponding B O D loading rate 0.766 kg/m3 d) and an aeration pattern of 1:3 (the time ratio of non-aeration and aeration). The D O concentration could be as high as 8.92mg/L during aeration time and as low as 0 mg/L after aeration was stopped for more than one hour. The p H and TSS of effluent were in the range o f 6.4 ~ 7.2 and 2.9 ~ 12.6 mg/L, respectively. Simultaneous organic carbon and nitrogen removal (through denitrification) could be realized in a single immobilized sludge reactor. 88 7. Different influent patterns were applied to the immobilized sludge treatment system to examine the impact of several possible actual flow patterns of small scale domestic wastewater. There was no appreciable reduction in organic carbon, N F L t + - N , T K N and TSS removal whether changing the feed pattern from continuous feed to intermittent feed at an H R T 6 hours or increasing the H R T to 24 hours at the same feed pattern (feed time 8 hours/d). However, the total nitrogen removal dropped dramatically when the intermittent feed patterns were applied. 8. The average removal o f PO4-P ranged from 0 ~ 33.7% throughout the research. Without excess sludge removal and without addition o f chemicals, biological phosphorus removal in the immobilized sludge system w i l l not give a satisfactory result. 9. The yield coefficient obtained from immobilized sludge process under continuous and intermittent aeration were 0.198 and 0.244g SS/ g B O D removed, respectively. The production of excess sludge by the immobilized sludge process is about 1/2 ~ 1/3 that produced by the conventional activated sludge process. 10. The immobilized sludge beads with P V A exhibited satisfactory mechanical stability without apparent breakage for the 180-day experiment. The results of scanning electron microscopic analysis showed that bacteria mostly grew on the surface o f immobilized sludge beads. 89 11. Based on a comparison with other biological treatment processes, the process performance o f the immobilized sludge process in this research is as good or better for domestic wastewater treatment. Results of this research proved that immobilized sludge process could be utilized as an effective treatment technology for small scale domestic wastewater. 4.2. Recommendation The research in this project is just first step toward developing an effective small domestic wastewater treatment system with the immobilized sludge process. I f the further research is to be done on this topic, the following areas are recommended to be looked at: 1. Further study should be done to further improve the total nitrogen removal efficiency by adding immobilized denitrifier into the reactor or filling the reactor with the coimmobilized denitrifiers and activated sludge beads or establishing a two-stage treatment system, one of them containing immobilized sludge and the other immobilized denitrifiers. 2. The possibility o f using immobilized sludge to remove phosphorus from the domestic wastewater should be investigated. 90 3. It would be interesting to investigate the change in activity of activated sludge and the distribution of microorganisms inside the immobilized sludge beads throughout the process, which are unclear so far. 4. The performance of the immobilized sludge process with a higher volume fraction o f immobilized sludge in the reactor should be investigated. The performance o f a reactor filled with smaller beads is also worth further investigating. 5. 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