Open Collections

UBC Theses and Dissertations

UBC Theses Logo

UBC Theses and Dissertations

Ammonia removal from a landfill leachate by biological nitrification and denitrification Dedhar, Saleem 1985

You don't seem to have a PDF reader installed, try download the pdf

Item Metadata

Download

Media
[if-you-see-this-DO-NOT-CLICK]
UBC_1985_A7 D43.pdf [ 5.98MB ]
[if-you-see-this-DO-NOT-CLICK]
Metadata
JSON: 1.0062950.json
JSON-LD: 1.0062950+ld.json
RDF/XML (Pretty): 1.0062950.xml
RDF/JSON: 1.0062950+rdf.json
Turtle: 1.0062950+rdf-turtle.txt
N-Triples: 1.0062950+rdf-ntriples.txt
Original Record: 1.0062950 +original-record.json
Full Text
1.0062950.txt
Citation
1.0062950.ris

Full Text

AMMONIA REMOVAL FROM A LANDFILL LEACHATE BY BIOLOGICAL NITRIFICATION AND DENITRIFICATION by SALEEM DEDHAR B.A.Sc. (Civil Engineering), University of British Columbia, 1981 A THESIS SUBMITTED IN PARTIAL FULFILMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF APPLIED SCIENCE in FACULTY OF GRADUATE STUDIES Department of Civil Engineering We accept this thesis as conforming to the required standard THE UNIVERSITY OF BRITISH COLUMBIA October, 1985 ® Saleem Dedhar, 1985 In presenting this thesis in partial fulfilment of the requirements foT an advanced degree at the THE UNIVERSITY OF BRITISH COLUMBIA, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of 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 of this thesis for financial gain shall not be allowed without my written permission. Department of Civil Engineering THE UNIVERSITY OF BRITISH COLUMBIA 2075 Wesbrook Place Vancouver, Canada V6T 1W5 Date: October. 1985 ABSTRACT The discharge of a landfill leachate to a receiving water body can cause a serious pollution problem. One component of leachate that can have a severe impact on a receiving water body is ammonia and its oxidized form, nitrate. This study investigated the biological treatibility of a high ammonia leachate, with specific regard to nitriification and denitrification. A continuous-feed, single sludge denitrification system with recycle was used. Leachate ammonia concentrations of up to 288 mg/L-N were reduced to less than 1 mg/L. The ammonia was removed by nitrification and bacterial uptake. About 25% of the incoming ammonia was taken up by the bacteria in the anoxic reactor; the rest was subsequently nitrified in the aerobic reactor. The nitrates produced in the aerobic reactor were recycled back to the anoxic reactor to undergo denitrification. Glucose was added directly to the anoxic reactor to aid denitrification. The degree of denitrification was dependent on the glucose loading to the anoxic reactor; however, 100% denitrification was achieved on several occasions. The influent leachate COD removal was 20%; however after the addition of glucose to the system, a mean COD removal of 74% was obtained. Of the COD removed across the system, 85% was used in the anoxic reactor for denitrification, and the remaining 15% was used by the heterotrophs in the aerobic reactor. The four metals monitored regularly, zinc, manganese, nickel and iron were removed by the biomass, but not to the same extent During the latter part of the study, the system was first spiked with manganese, and then - zinc, to try and induce an inhibitory effect on the nitrification process. The manganese had no detectable effect on the system. However, total zinc (>95% soluble) levels of between 14.9 and 17.6 mg/L caused substantial inhibition of the nitrification process, resulting in approximately 70 mg/L ammonia in the effluent (feed = 216 mg/L). This inhibition was also evident from the lower percent nitrification values and the unit nitrification rates. This high influent zinc concentration also caused deflocculation, resulting in the loss of significant quantities of biomass with the effluent. The high zinc concentrations also inhibited the denitrifiers, resulting in a decrease in the ammonia uptake, as well as an increase in the COD (used)/Nitrate+Nitrite (NOT) (reduced) ratios in the anoxic ii reactor. The zinc levels were then lowered to allow the system to return to normal; after this state had been reached, the influent total zinc (>95% soluble) levels were again increased up to 19.5 mg/L. This concentration of zinc did not result in any ammonia appearing in the effluent; thus, it is possible that the bacteria had acclimatized to these high influent zinc concentrations. iii Table of Contents ABSTRACT ii LIST OF TABLES vLIST OF FIGURES viACKNOWLEDGEMENT ix 1. INTRODUCTION 1 2. LITERATURE SEARCH 4 2.1 Treatment Processes 5 2.1.1 Sanitary Landfill as a Reactor 5 2.1.2 Physical - Chemical Treatment 6 2.1.3 Anaerobic Treatment 7 * 2.1.4 Aerobic Treatment 8 2.1.4.1 Nitrification 9 2.1.4.2 Denitrification 11 2.1.4.3 Biological Ammonia Removal From Leachate 12 3. EXPERIMENTAL SET- UP AND OPERATION 15 3.1 Treatment Scheme 13.1.1 Biological Treatment System 1 .-. 15 3.1.1.1 Leachate Feed3.1.1.2 Anoxic Reactor 15 3.1.1.3 Aerobic Reactor 20 3.1.1.4 Final Clarifier 1 3.1.2 Biological Treatment System 23.1.2.1 Leachate Feed 1 3.1.2.2 Aerobic Reactor 23.1.2.3 Final Clarifier 2 3.2 Operation 23.2.1 Biological Treatment System 1 22 iv 3.2.2 Biological Treatment System 2 3 4. ANALYTICAL METHODS 4 4.1 Solids 24.2 Dissolved Oxygen and pH 24 4.3 ORP (Oxidation-Reduction Potential) 25 4.4 BOD (Biochemical Oxygen Demand)4.5 COD (Chemical Oxygen Demand) 25 4.6 TKN (Total Kjeldahl Nitrogen)4.7 TP (Total Phosphoros) 26 4.8 Ammonia Nitrogen4.9 Nitrate and Nitrite4.10 Trace Metals 27 5. RESULTS AND DISCUSSION 28 5.1 Carbon Removal5.2 Solids (SS and VSS) 33 5.3 Ammonia Removal 45 5.4 Nitrification 7 5.5 Denitrification 54 5.6 Reaction and Unit Removal Rates 58 5.6.1 Reaction Rates5.6.2 Unit Removal Rates 61 5.7 Metal Removal 4 6. CONCLUSIONS AND RECOMMENDATIONS 89 6.1 CONCLUSIONS 86.2 RECOMMENDATIONS 91 REFERENCES 93 APPENDICES (DATA FOR FIGURES) 95 v LIST OF TABLES 1. Basic Characteristics of Port Mann Leachate 19 2. Chromium and Lead Concentrations 74 vi LIST OF FIGURES 1. Laboratory Biological Treatment System 1 6 2. Laboratory Biological Treatment System 2 (control) 17 3. Port Mann Landfill Site (location of the wells) 8 4. Influent and Effluent COD Versus Time (mg/L) . 29 5. Influent and Effluent COD Versus Time (mg/d) 30 6. COD Removal Efficiency Versus Time 1 7. Ratio of Carbon Feed COD to Leachate Influent COD Versus COD Removal Efficiency 32 8. Percentage COD and BOD5 Removal Across the Reactors Versus Time 34 9. BOD5 Versus Time 35 10. BOD5 Removal Efficiency Versus Time 36 11. Influent and Effluent SS Concentration Versus Time 37 12. Influent and Effluent Zinc Concentration Versus Time 8 13. Influent and Effluent VSS Concentration Versus Time > 40 14. Anoxic and Aerobic MLSS Concentration Versus Time 41 15. Anoxic and Aerobic MLVSS Concentration Versus Time 2 16. Ratio of Carbon Feed COD to Leachate Influent COD Versus MLVSS concentration 43 17. MLVSS/MLSS Ratios in the Reactors Versus Time 44 18. Ammonia Concentration Versus Time 46 19. Effluent Ammonia From System 2 Versus Time 420. Percent Ammonia Removal Across the Anoxic Reactor Versus Time 48 21. Nitrate + Nitrite Concentration Versus Time 49 22. Anoxic and Aerobic pH Versus Time 51 23. Percent Nitrification Versus Time (Defn.: A) 2 24. Percent Nitrification Versus Time (Defn.:B) 53 25. Ratio of COD (used)/Nitrate + Nitrite (reduced) Versus Time 55 26. COD and Nitrate-(-Nitrite Lost Across The Anoxic Reactor Versus Time 57 vii 27. ORP in the Anoxic Reactor Versus Time 59 28. Unit Nitrification Rates Versus Time 60 29. Unit Denitrification Rates Versus Time 2 30. Ammonia Removal Rates Versus Time (mg/hr/gm VSS) 63 31. Ammonia Removal Rates Versus Time (gm/mVday) 5 32. Influent and Effluent Nickel Concentration Versus Time 66 33. Nickel Removal Efficiency Versus Time 68 34. Influent and Effluent Iron Concentration Versus Time.... 70 35. Influent and Effluent Iron Concentration Versus Time (semi-log plot) 71 36. Iron Removal Efficiency Versus Time 72 37. Sludge Iron Concentration Versus Time 3 38. Influent and Effluent Manganese Concentration Versus Time 76 39. Influent and Effluent Manganese Concentration Versus Time (semi-log plot) 77 40. Manganese Removal Efficiency Versus Time 79 41. Sludge Manganese Concentration Versus Time 80 42. Sludge Manganese Concentration Versus Time (semi-log plot) 81 43. Influent and Effluent Zinc Concentration Versus Time (semi-log plot) 82 44. Zinc Removal Efficiency Versus Time 85 45. Sludge Zinc Concentration Versus Time 87 viii ACKNOWLEDGEMENT The author wishes to sincerely thank his supervisor, Dr. D.S. Mavinic, for his guidance, genuine interest and constant encouragement during this study. The author wishes to acknowledge the District of Surrey, B.C., for allowing access to the Port Mann landfill site for leachate collection. In particular, special thanks go to Mr. Eric Johnson and Chris Wike for their cooperation and assistance in the bi-weekly collection of leachate for this study. The author also acknowledges the assistance received from Susan Liptak and Paula Parkinson of the Environmental Engineering Laboratory. The author is grateful to the Wastewater Technology Center, Burlington, Ontario, for the loan of the reactors and pump that were used for this study. Financial support for this work originated from the Natural Sciences and Engineering Research Council of Canada. ix ..INTRODUCTION Landfills and garbage dumps have been in existence for a long time, in fact, the land was probably the first and most convenient site for the disposal of man's wastes. However, it is only recently that we are realizing some of the potential hazards associated with land disposal, the primary one being leachate. Leachate is generated when water enters the landfill, percolates through it, and picks up soluble materials, some of them soluble products of biological and chemical reactions. The water can enter a fill by such means as precipitation or by drainage of flood waters, springs, or the passage of groundwater through the fill (Patel, Hoye and Toftner, 1979). A four year study of the amount and characteristics of leachate at the Boone County Field Site found a direct correlation between the cumulative precipitation and cumulative volume of leachate produced (Wigh and Brunner, 1979). Numerous surveys have noted that there are wide variations in the composition of Municipal Solid Waste Landfill leachate (Fuller, Alesii and Carter, 1979; McDougall, Fusco and O'Brien, 1980; Chian, 1977). The variability in composition of leachate from a Municipal Solid Waste Landfill to landfill, is due mainly to the quality and quantity of industrial wastes often included. The variation in leachate composition within a given landfill is due largely to the age of disposal and amount of rainfall contacting the solid waste (Fuller, Alesii and Carter, 1979). Municipal leachates are typically anoxic, contain reduced species and are buffered to a neutral pH. The major inorganic constituents in leachate include chlorides, sulphates, bicarbonates, ammonia , iron(II), manganese II, sodium, potassium, calcium, chromium, copper, nickel, lead and zinc, most of which are found in low concentrations (Jasper, Atwater and Mavinic, 1984). In a new landfill, aerobic conditions will exist from a few weeks to approximately six months, depending on the fill material. Thereafter, conditions favour autotrophic, facultative anaerobic bacteria, which degrade the organic matter to produce volatile fatty acids (eg. acetic acid, 1 2 butyric acid). This degradation does not change the BOD much, but the acids can lower the pH in the landfill to about 4.5 or 5. This low pH increases the solubility of many inorganics in the landfill, and is toxic to the methane producing bacteria; therefore little methane is produced. This is known as the first stage of anaerobic decomposition and is characterized by: high volatile fatty acid production, low pH, high BOD/COD, low methane production and high conductivity. Organic carbon values of up to 25,000 mg/L have been reported during the early months after first establishment of the landfill. The volatile acids were found to be most prominent early in the biodegradation process (Chian, 1977). The second stage anaerobic decomposition takes place when the methane producing bacteria become established. This stage is characterized by higher pH's and lower COD's and BOD's. The methane producing bacteria degrade the volatile fatty acids to methane (CH4) and carbon dioxide (C02) in approximately a 50%-50% ratio. The degradation of the acids allows the pH to remain around pH 7. This pH level will decrease the solubility of some of the inorganics. This decrease and the decrease in the volatile fatty acids results in a decline in conductivity (Chian, 1977). A by-product of methane generation is ammonia which is released by the biomass. Therefore, an older landfill will produce a leachate that is characteristically high in ammonia and low in BOD and COD. Discharge of landfill leachates to receiving waters are posing problems in areas such as aquatic life toxicity, potable water contamination, bioaccumulation of toxic metals, color and odor. One component of leachate that can have a severe impact on a receiving water body is ammonia (NH4+) and its oxidized form, nitrate (N03~). A high ammonia content in the effluent leachate can be toxic to aquatic life. There also exists the potential for dissolved oxygen (D.O.) reduction (through ammonia oxidation) and the potential for eutrophication as a result of nitrate accumulation. Land disposal of the effluent could also result in nitrate contamination. The purpose of this investigation was to study the biological treatability of a leachate from an "older" landfill, with specific regard to nitrification and denitrification. A continuous-feed, single sludge denitrification system was used. The parameters monitored were ammonia (NH«+), 3 nitrate and nitrate (NOT), influent and effluent solids, mixed liquor solids, Chemical Oxygen Demand (COD), Biochemical Oxygen Demand (BOD5), Total Kjheldal Nitrogen, pH, ORP and trace metals. The system was operated at two solids retention times (also called sludge age) during the study. The system was also spiked with specific trace metals to try and induce a toxic effect on the process. Since the nitrifiers are very sensitive to some heavy metals, knowledge of what metals inhibit the nitrifiers, and in what concentrations, is essential for successful treatment of landfill leachate. 2. LITERATURE SEARCH The potential problem of leachate formation from landfills would seem to be a long term one. This is due to the fact that no matter what mode of refuse disposal is used, there will always be a residue remaining, which will have to be disposed of at a landfill. Furthermore, it would seem that a landfill site will always be required as a standby disposal method in case of interruption of operation of whatever disposal method is being used (eg. incineration). With proper planning, good engineering design and adequate monitoring of a landfill site, the problem of leachate contamination of ground and surface waters can be reduced and, in some cases, eliminated. Unfortunately this does not prevent the formation of leachate. Therefore, when direct discharge to a water body causes a pollution1 problem, collection and treatment of the leachate is required. This section will cover some leachate treatment methods, as well as review the treatment method used for this investigation. Although leachate is biodegradable, it has a large range of refractory organic components. The soluble organic content of leachate is made up of three groups, in order of decreasing biodegradability: 1. Short chain fatty acids of low molecular weight, which accounts for up to 90% of the soluble organic content 2. Humic - carbohydrate like substances of high molecular weight, constituting the next highest fraction. 3. Fulvic - like substances accounting for only a small proportion of the total. In older landfills, the more readily degradable organics have already been removed by natural biological processes within the landfill, so that the humic - carbohydrate like substances and to a lesser extent the fulvic - like substances make up the bulk of the organic content Therefore biological treatment will generally be effective in removing most of the organics from leachate from '"Pollution" is that which is in excess of what the natural environment can assimilate. 4 5 recently placed refuse. Biological treatment will not remove the smaller fulvic - like component, or be effective in removing the organics in leachates from older landfills where fulvic - like substances comprise the bulk of the organics. In such cases, physical - chemical treatment may be more effective (Chian and DeWalle 1976). The high ammonia content in leachates from older landfills can be reduced by physical -chemical means (eg. stripping) or by aerobic biological processes (eg. nitrification). The biological ammonia removal method will be discussed in greater detail later on in this chapter. Some heavy metals in the leachate can be removed by both physical-chemical and biological processes. 2.1 TREATMENT PROCESSES 2.1.1 SANTTARY LANDFILL AS A REACTOR Several researchers have suggested that recirculation of leachates through a landfill could be used as a method of leachate treatment- Pohland et al; in Jasper Atwater and Mavinic (1984), worked with lab - scale treatment cells and found that after 2-3 years, the cells with recycle produced leachate with markedly reduced BOD5, TOC and COD concentrations. Robinson and Maris (1982), also conducted experiments to assess the effects of recirculation of leachate and found that leachate produced from natural test cells after 18 months had a low BOD5, but the COD, ammoniacal-N and chloride remained relatively high, as did the concentrations of some metals; notably iron, manganese, sodium and potassium. Although both these studies indicate that recirculation of leachate might be a viable option as a first step in some treatment scheme, recirculation at a full - scale site might present difficulties in achieving the high rates of liquid flow found in lysimeters, and consequently making some of the benefits of recirculation observed on a small scale, more difficult to obtain. Robinson and Maris (1985) however, conducted experiments at a full scale landfill site and demonstrated that the benefits of recirculation of leachate, found in smaller studies, can be obtained on a larger scale; 6 however, longer recycle periods may be needed. Nevertheless, these researchers felt that recirculation alone would not provide complete treatment and that a combination of recirculation and aerobic biological treatment may be the most effective option. 2.1.2 PHYSICAL - CHEMICAL TREATMENT Leachates from older landfills, which have a smaller biodegradable organic component, are more amenable to Physical/Chemical treatment Physical and or chemical treatment may consist of equalization, lime and polymer treatment air stripping, precipation, filtration, activated carbon adsorbtion, reverse osmosis, ion exchange and break point chlorination. Chian and DeWalle (1976) state that physical chemical treatment processes are most effective in treating leachate from stabilized landfills, or in further removing organic matter in the effluent of biological units treating leachate. Of all the physical - chemical processes evaluated, activated carbon and reverse osmosis give the most effective removal of residual organic matter. Bjorkman and Mavinic (1977) used lime and ozone to treat a high strength leachate, which was typical of leachates from newer landfills. These researchers found effective removal of some metallic ions as well as effective reduction in color and turbidity. The main components of the COD (consisting of mostly organic carbons) however, were not efficiently removed. They also reported that sludge from a physical - chemical treatment process may present a disposal problem. Cook and Foree (1974) also found that the use of physical - chemical treatment was effective for color removal, but with the exception of activated carbon, was relatively ineffective for total COD removal. Ammonia air stripping is the physical/chemical method of removing or reducing the ammonia content in leachates. Keenan et al. (1984) reported on a leachate treatment facility in Bucks County, Pa. The plant included both physical/ chemical and biological processes, and incorporated air stripping to reduce the ammonia concentration into the biological system to non inhibitory levels ( 300 mg/L NH4+). The physical/ chemical ammonia treatment provided ammonia removals of approximately 50%; however the biological process was needed to further reduce the 7 ammonia concentrations to dischargable levels, through the process of assimilation and nitrification. 2.1.3 ANAEROBIC TREATMENT Anaerobic treatment of leachate is a process in which complex organic molecules in the influent are bacteriologically fermented to volatile fatty acids. These acids are then converted to methane and carbon dioxide by methanogenic bacteria, resulting in low production of biological solids requiring disposal. The methane is a useful source of energy and if collected and sold, could reduce the overall cost of the treatment plant. Wright et al. (1985) used a 640 litre, anaerobic downflow, stationary, fixed film reactor of National Research Council Design, to treat a landfill leachate with a COD of 22900 mg/L. The reactor was loaded at 1, 2 and 4 Kg COD/mVday and achieved COD reductions exceeding 92%. Higher loadings and shock loadings resulted in failure. In these cases it is suspected that phosphorus additions may have been helpful in preventing failure. Gas production rate approached the theoretical rate of 0.39 litres of methane per gram of COD removed. The raw leachate had a mean ammonia concentration of 363 mg/L. This treatment system did not remove any of the influent ammonia; infact the ammonia concentration in the reactors increased as a result of nitrogen being provided by the sludge in the reactor (as a result of previous experiments on fish waste). Boyle and Ham (1974) investigated anaerobic treatment of leachate and found greater than 90% BOD reduction for hydraulic retention times greater than 10 days and temperatures in the range of 23 to 30° C. Henry et. al. (1982) utilized an anaerobic filter to treat a high strength leachate. These researchers obtained 90% COD removal with a 12 hour detention time at 25° C and with a 72 hour detention time at 10° C. The gas produced was comparable to that produced by conventional anaerobic sludge digestion, but had a higher methane content at 77 - 84%. Regardless of the potential advantages of anaerobic treatment, it still leaves the need for additional treatment for the removal of ammonia. 8 Austin, et al. (1984) studied the anaerobic treatment of a landfill leachate using anaerobic fixed film reactors. The leachate was a high-strength leachate with a COD of 23000 mg/L and a BOD of 17500 mg/L. The reactors provided better than 98.5% BOD removal and 96.9% COD removal. However, these reactors provided negligble ammonia removal; infact, one reactor showed an increase in ammonia. These researchers also found that the best treatment of this leachate was established with chemical pretreatment, followed by the anaerobic fixed film reactor, plus an aerated lagoon. The pretreatment step was used to buffer the pH so as to enhance methane generation in the anaerobic reactor, as well as precipitate out some metals, namely zinc, that could be toxic to the anaerobic bacteria. The aerated lagoon removed 99.7% of the ammonia. Since anaerobic treatment of leachate does not remove the ammonia component of leachate, this treatment method does not seem feasible for "older" leachates which have a high ammonia concentration. However, this method has been quite successful in treating high strength leachates from "young" landfills. 2.1.4 AEROBIC TREATMENT This form of treatment has been shown to be most effective in the removal of organic constituents of leachates from "young" landfills. These leachates contain high concentrations of readily biodegradable short chain fatty acids. Cook and Foree (1974), Uloth and Mavinic (1977) and Zapf - Gilje and Mavinic (1981) (amongst many), treating high strength leachate, found greater than 95% COD removal was achieved for detention times of 10 days and greater. Cook and Foree (1974) found that for detention times of less than 5 days, the system failed. Aerobic biostabilization has also been effective in removing a considerable portion of the metals found in high concentrations in "young" or high strength leachates. Zapf Gilje and Mavinic (1981) found that most metals were reduced in concentration by more than 90%, although further polishing was required to meet local effluent pollution control objectives. Since almost all metals accumulate in the sludge, overall successful treatment of leachates must include adequate control of 9 the resultant sludge. The proper nutrient balance must be considered in any biological treatment process. Most leachates lack phosphoros; therefore, in order to achieve adequate treatment, it must be added in some form or another. Stegmann and Ehrig (1980) reported that a lack of phosphoros inhibited the biological degradation process, to a certain degree. Generally BOD:N:P ratios should be maintained at 100:5:1, although Temoin (1980), reported that the most effective treatment of a leachate was achieved with a nutrient loading of 100:3.2:1.1. Wong and Mavinic (1982) investigated the treatment of a municipal landfill leachate by aerobic biostabilization. They reported that a BOD5:N:P loading of 100:3.2:1.1 was "adequate" for treatment Aerobic biodegradation must also be capable of treating leachate at cold winter temperatures. In fact leachate production at a landfill on the West Coast, is often greatest during the winter due to heavier precipitation. Robinson and Maris (1985) treated leachate aerobically at temperatures of 10 and 5° C, and obtained better than 92% COD removal. However substantial phosphoros additions and SRT's of greater than 10 days were required. Lower retention periods produced adverse effects, particularly on clarifying properties of the effluents. Leachates from "older" landfills (stabilized landfills) are less amenable to aerobic biological treatment as they contain more refractory, fulvic like compounds. Therefore the COD from these leachates is effectively removed using physical - chemical methods, especially activated carbon and reverse osmosis. "Older" leachates are also characteristically high in ammonia. Aside from air stripping (physical - chemical treatment), the only other viable treatment scheme for a high ammonia waste is aerobic nitrification and denitrification. 2.1.4.1 Nitrification The principal agents of nitrification are considered to be the chemoautotrophic bacteria, which oxidize ammonia sequentially to nitrite and nitrate according to Equations 1 and 2. NH,+ + 1.502' >^N02" + H20 + 2H+ + (240- 350 kJ). (1) 10 NCv + 0.5Cv >• N03- + (65 - 90 kJ). The energy released in these reactions is used by the nitrfying organisms in synthesizing their organic requirements from inorganic carbon sources, such as carbon dioxide, bicarbonate and carbonate (Barnes and Bliss 1983). Ammonia oxidation is carried out principally by organisms of the genera Nitrosmonas (N.europaea and N.monocella) and Nitrosococcus. Nitrite oxidation is effected principally by members of the genera Nitrobacter (N.agilis and N.winogradskyi) and Nitrosocystis. Oxygen is involved in ammonia oxidation not only by incorporation into the energy substrate as implied by Equation 1, but also in the acceptance of electrons during electron transfer through the cytochrome system. Because the net energy produced in nitrite oxidation is so much less than that produced in ammonia oxidation, the cell yield for Nitrobacter is less than that of Nitrosomonas, for each unit of nitrogen oxidized. For this reason, Nitrosomonas are expected to be present in greater numbers than Nitrobacter in nitrifying environments (Barnes and Bliss 1983). On the assumption that the gross composition of Nitrosomonas and Nitrobacter can be represented as CjH7N02, the overall reaction for nitrifier synthesis is expressed as Equation 3 (Barnes and Bliss 1983). Equation 3 shows that nitrification has a very low cell yield per unit of ammonium nitrogen oxidized. It also shows that the requirement for oxygen in nitrification is significant, approximately 4.2g oxygen for each g NH4+-N removed. Approximately 7g of alkalinity are also needed to buffer the system against hydrogen ions produced during nitrification. The principal problem with effecting nitrification in aerobic biological treatment systems is that nitrifying bacteria are very much slower growing than the heterotrophic NH4+ + 1.8302 + 1.98HC03- •0.021C5H7NO2 + 1.041H2O + 0.98NO3- + 1.88H2CO- (3) 11 organisms involved in carbon removal, and the reaction rate of nitrification is correspondingly slower. This means that in order to maintain a population of nitrifiers in a growth system, the mean sludge age must be long enough to avoid "wash out" of nitrifying organisms from the system. Generally about 5 days is considered a minimum sludge age necessary. (Winkler 1981). If nitrification is to be effected in a treatment plant, the metal concentrations in the waste must also be considered, as the nitrifiers are very sensitive to certain heavy metals. Martin (1979), from Martin and Richard (1982), found that zinc, lead, nickel, copper and chromium were toxic to Nitrosomonas at fairly low concentrations. 2.1.4.2 Denitrification Denitrification is the process by which nitrate ions are reduced to nitrite ions and subsequently to nitrogen gas. Equation 4 represents the reduction of nitrate ions to nitrite ions; Equation 5 represents the reduction of nitrite ions to nitrogen; and Equation 6 represents the overall reaction. Several facultative heterotrophic micro-organisms can carry out these reactions, since it does not need specialist bacteria as is the case for nitrification. NCy + 0.33CH3OH • NGy + 0.33CO2 + 0.67H2O__ (4) NCv + 0.5CH3OH •Nj + 0.5CO2 + 0.5H2O + OH" (5) NCV + 0.83CH3OH •N, + 0.83CO2 + 1.17H20 + OH" (5) In the above equations, methanol (CH3OH) was used as an electron donor. The denitrifying bacteria must have some chemical to oxidize, i.e. an electron donor, in order to use the nitrate ions as an electron acceptor. Many organic chemicals, for example acetic acid, acetone, glucose and methanol can be used. The organics present in the waste itself may be used; however, in many cases it is not enough for complete denitrification and must be augmented with an extra organic source. The denitrification process uses nitrate as an electron acceptor and is the next most favoured to oxygen. Phosphate and sulphate can also act as electron acceptors; however, in a 12 wastewater which has undergone biological nitrification, the concentration of nitrate ions would be expected to be greater than the concentration of phosphate or sulphate ions. Therefore, under conditions of low oxygen concentration, biological denitrification can be expected to occur; these conditions are referred to as anoxic. 2.1.4.3 Biological Ammonia Removal From Leachate Very little work has been done on biological ammonia removal from leachate, as most research on leachate treatment has concentrated on the removal of the COD and BOD components. This section will review some of the work done to date on biological ammonia removal. Jasper, Atwater and Mavinic (1984) investigated the biological treatibility of leachate from the Port Mann landfill in Surrey, British Columbia. The treatment set up consisted of 3 single sludge denitrification systems, each operated at a different aerobic sludge age, 10,15 and 20 days. The influent leachate had ammonia (NH4+) concentrations that ranged from 30 to 220 mg/L-N over the course of the study. The removal of ammonia was inconsistent and the baseline goal of 10 mg/L effluent ammonia was not achieved with any degree of consistency. Nitrification efficiencies of at least 75% were initially obtained but subsequently fell to less than 10% by the end of the study. Metals accumulated fairly extensively in the sludge and it was postulated that this accumulation may have accounted, in part for the failure of the nitrification/denitrification process. The authors concluded that the removal of nitrogen from raw leachate may not be possible using available biological techniques, at aerobic SRT's of less than 20 days. Some of the mean metal concentrations in the leachate were: zinc=0.55 mg/L, manganese = 2.3 mg/L, chromium=0.006 mg/L, nickel = 0.02 mg/L, iron=68 mg/L. Knox (1983) operated an aerobic "fill and draw" bench scale treatment system, to treat a leachate with an ammonia concentration ranging from 200 to 600 mg/L as N. This was a leachate from an "older" landfill and was typically low in BOD and COD (80-250 mg/L and 850-1350 mg/L, respectively). The system proved capable of completely nitrifying the ammonia in the leachate. The metal concentrations in the leachate, however, were very low. 13 (e.g. zinc=0.16 mg/L, manganese=0.5 mg/L, chromium = 0.05 mg/L, nickel = 0.04mg/L) Robinson, Barber and Maris (1982) ran laboratory - scale, aerobic units at a temperature of 10° C for the treatment of leachate. Nominal retention periods of 5,10,15 and 20 days were used. They observed that the removal of ammonia resulted entirely from conversion to organic nitrogen during reduction of COD, and no nitrification occured. The reasons given to explain the absence of nitrification were combined effects of sludge age and temperature, the low ratio of nitrogen to BOD in the influent and possible inhibitory substances in the leachate. Some of the metal concentrations in the leachate were: zinc= 13.6 mg/L, manganese=21.6 mg/L, chromium=0.08 mg/L, nickel=0.17 mg/L, iron=48 mg/L Stegmann and Ehrig (1980) also reported on a lab - scale, activated sludge plant treating leachate. These researchers found that full nitrification was achieved and an influent NH4+-N concentration of 973 mg/L was reduced to an effluent concentration of less than 8 mg/L. No influent metal data was presented. Keenen, Steiner and Fungaroli (1984) reported on a full scale leachate treatment plant located at a landfill in Falls Township, Bucks County, Pa. The leachate had a mean ammonia concentration of 758 mg/L-N. Since this high ammonia concentration was believed to be toxic to nitrifiers, chemical/physical treatment was used to reduce the ammonia concentration to a level suitable for biological treatment, as well as precipitating out the metals. The influent ammonia into the biological units had a mean concentration of 350 mg/L-N, and the effluent had a mean concentration of 75 mg/L-N. This reduction was primarily due to nitrifying organisms. Some of the mean metal concentrations into the biological units were: zinc=0.53 mg/L, chromium=0.07 mg/L, nickel=0.75 mg/L, iron=2.71 mg/L. These metal concentrations were very low compared to the raw leachate influent metal concentrations. The authors also reported that the cold winter temperatures inhibited the biochemical oxidation of ammonia, resulting in severe operating problems. The best treatment method suited to "older" leachates, which have a high ammonia content, would seem to be aerobic nitrification and denitrification. Researchers who have 14 investigated this method of leachate treatment have had varied success with the removal of ammonia. The purpose of this study was to determine conclusively, whether or not a high ammonia leachate could be treated sucessfully with this method, as well as to determine what metals in the leachate, and in what concentration, would inhibit this process. 3. EXPERIMENTAL SET-UP AND OPERATION 3.1 TREATMENT SCHEME The reactor set up used was a single sludge denitrification system with recycle, shown schematically in Figure 1. This system, supplied by Wastewater Technology Center, EPS, Burlington Ontario, will be called "system 1". A second system was set up during the second half of the study to act as a control to system 1, which was to be spiked with trace metals. The control set up will be known as "system 2", schematically shown in Figure 2. The process for system 1 is described in detail below. 3.1.1 BIOLOGICAL TREATMENT SYSTEM 1. 3.1.1.1 Leachate Feed The leachate used for this study was obtained from the Port Mann landfill in Surrey, British Columbia (Figure 3), and is classified as an "older" leachate. The basic characteristics of this leachate are shown in Table 1. Fresh leachate was obtained from 2 different wells at the landfill every 2 weeks, well #2 and well #3, which are shown on Figure 3. The only measured difference between the leachates from the two wells was that well #3 produced a leachate that had a consistently higher ammonia (NH„+) concentration. The leachate was stored in closed containers and at a temperature of 4° C, until used. Chian and DeWalle (1976) state that storage under anaerobic conditions and at low temperatures is necessary to avoid a decrease in COD and an increase in suspended solids. Leachate feed was continuously added to the anoxic reactor at a rate of approximately 10 litres a day, from a constantly mixed supply contained in a plastic tank. The tank was covered with a lid to prevent excess aeration of the feed. 3.1.1.2 Anoxic Reactor The main function of the anoxic reactor was to denitrify the highly nitrified return sludge from the clarifier, according to Reaction step 7. 15 Pump Glucose + Phosphorus Pressure Regulator Air Supply Daily Wasting (proportional to SRT) -Scraper Arm CLARIFIER Leachate Fee d UOL/doy) EFFLUENT Sludge Recycle (~ 4 :1 ) Pump Fig.1: LABORATORY BIOLOGICAL TREATMENT SYSTEM 1. Pump 9-Daily Wasting (proportional to SRT) -Pressure Regulator •^T) « Air Supply Mixer Leachate Feed (~3L/doy) AEROBIC REACTOR sludge recycle (-4 = 1 ) EFFLUENT Fig.2: LABORATORY BIOLOGICAL TREATMENT SYSTEM 2 (control). Table 1. Basic Characteristics of Port Mann Leachate Parameter Concentration* COD 217 - 318 BOD.5 6 - 24 Arnmonia-N 122 - 288 PH 7.5 - 8.3 Zinc 0.018 - 0.179 Manganese 0.024 - 0.286 Iron 3.70 - 36.25 Nickel 0.022 - 0.066 * All values expressed as mg/L/ except pH in pH units. 20 NCV • NCV >-N2(g) (7) This reduction of nitrates to nitrogen gas is carried out by heterotrophic bacteria. The reactor was a plexiglas tank with a liquid volume of 5 litres and a mechanical mixer. In addition to receiving fresh leachate continuously, the anoxic reactor also received a solution of glucose and sodium phosphate tribasic (Na3P04.12H20). Phosphorus was added to prevent nutrient limitation. Several measurements of phosphorus were carried out on the clarifier effluent to ensure that it was not limiting. Since the BOD5 of the influent leachate was very low, it was necessary to provide an additional carbon source (electron donor) for the denitrification process. Therefore, glucose was added to the sodium phosphate tribasic solution. Methanol is a common carbon source used for denitrification, but was not used for this study since it has been found to inhibit ammonia oxidation (Hooper and Terry 1973). Barnes and Bliss (1983) state that, although many organic compounds are inhibitory to nitrifiers, especially Nitrosomonas, glucose is one compound that is not In the latter part of the study, trace metals, namely manganese and zinc, were added directly to the anoxic reactor, in order to induce inhibition of the nitrification system. An ORP probe was also installed in the anoxic reactor in the latter part of the study, to ensure that there was a reducing environment to facilitate denitrification. 3.1.1.3 Aerobic Reactor This reactor was a large, polythelene carboy with a liquid volume of 10 litres, and was gravity fed from the anoxic reactor. The tank was aerated by a perforated pipe diffuser, fitted to the bottom of the tank. A mixer ensured that the contents were well stirred. A residual (D.O.) of between 1 and 3 mg/L provided sufficient oxygen for carbonaceous oxidation and nitrification of ammonia. The nitrification process itself involves the oxidation of ammonia to nitrates by chemoautrophic bacteria, according to Reaction step 8. NH4+ NCV —•NO,- (%) 21 In order to maintain a desired SRT, Mixed Liquor Suspended Solids (MLSS) was wasted directly from the aerobic reactor in proportion to the SRT. 3.1.1.4 Final Clarifier A conical, plexiglas tank with a 4 litre capacity, was used as the clarifier. Mixed-liquor from the aerobic reactor was fed to the clarifier, by gravity, where it was settled and gravity thickened for approximately 2 hours. The supernatant overflowed a weir and then flowed by gravity into a collection tank. The nitrified settled sludge was recycled back to the anoxic reactor at a rate of approximately 4:1. The recycle pump was operated on a cycle of 15 minutes on and 15 minutes off. This combination was necessary to clear the'recycle line of mixed liquor solids, as well as to provide proper volumetric thoroughpuL Nevertheless, over time, the insides of the recycle line became coated with bacterial growth, and it was necessary to "pinch" the line every so often, to free it of the attached growth. One problem that was encountered with the clarifier operation was that the sludge tended to adhere to the sides of the clarifier. In order to alleviate this problem, a scraper mechanism was installed on day 72 of steady state; this kept the clarifier sides free of sludge. The scraper mechanism was operated on a cycle of 1 minute on and 15 minutes off. 3.1.2 BIOIXIGTCAL TREATMENT SYSTEM 2 3.1.2.1 Leachate Feed The leachate used was the same as the one used for system 1. The leachate feed was fed continuously to the aerobic reactor at a rate of approximately 3 litres a day, from a constantly mixed supply contained in a plastic tank. The tank was covered with a lid to prevent excess aeration of the feed. 3.1.2.2 Aerobic Reactor The reactor was a plexiglas cyclinder, with a liquid volume of 2.88 litres. The reactor was aerated with 2 stone diffusers. Due to a lack of mixers, the diffused air was utilized to keep 22 the mixed liquor completely mixed. The D.O. provided sufficient oxygen for carbonaceous oxidation and nitrification of ammonia. In order to maintain a desired SRT, MLSS was wasted directly from the aerobic reactor in proportion to the SRT. 3.1.2.3 Final Clarifier A conical, plexiglas cyclinder with a 1 litre capacity, was used as the clarifier. Mixed liquor from the aerobic reactor was fed to the clarifier, by gravity, where it was settled and gravity thickened for approximately 2 hours. The supernatant flowed by gravity into a collection tank. The nitrified sludge was recycled back to the aerobic reactor at a rate of approximately 4:1. The recycle pump was operated on a cycle of 15 minutes on and 15 minutes off. 3.2 OPERATION 3.2.1 BIOLOGICAL TREATMENT SYSTEM 1 The treatment system was started up on February 23rd, 1984 with 15 litres of sludge from a sewage treatment, pilot plant operating at the University of British Columbia. Continuous leachate feed was also started at this time. Mixed liquor wasting was started a few days later. April 26th, 1984 was regarded as day 1 of steady state, although full nitrification had already been reached in early March. Mixed liquor was wasted from the aerobic reactor in order to maintain a 13 day aerobic SRT. The amount wasted was changed on day 54 in order to obtain a 15 day aerobic SRT. This change was made to enable comparison with results obtained from Jasper, Atwater and Mavinic (1984), in an earlier study. Addition of glucose to the anoxic reactor was started on day 11. The amount of glucose entering the system varied overtime. This was due, primarily, to the variation in the speed of the pump, that was used to feed glucose to the anoxic reactor. Metal addition to the anoxic reactor was started on day 219. Both the metals, manganese and zinc, were added to the anoxic reactor with the glucose solution. However, both metals formed a precipitate with the phosphate, also in solution, and it was therefore necessary to add the phosphate to the 23 anoxic reactor, separately, on a daily basis. 3.2.2 BIOLOGICAL TREATMENT SYSTEM 2 This treatment system was set up on October 8th, 1984 with 2.88 litres of sludge from system 1. At this point in the study, it was planned to add metals to system 1, to try and induce an inhibitory effect on the nitrifiers. In response to this, system 2 was set up to act as a control. Both systems were receiving the same leachate, and both were operating at the same SRT and recycle rate, the only difference being that system 1 was receiving a higher metal concentration, as well as employing an anoxic basin for denitrification. Therefore, any change detected in system 1 and not in system 2, could then be attributed to the metal spiking. System 2 did not include an anoxic reactor in the belief that the most susceptible bacteria to high metal concentrations in the biological ammonia removal process were the nitrifiers, and these bacteria need only an aerobic environment Mixed liquor was wasted from the aerobic reactor at a rate of 192 ml per day, in order to maintain a 15 day aerobic SRT. Phosphate was added to the aerobic reactor every few days to ensure that nutrient limitation did not occur. 4. ANALYTICAL METHODS All tests were carried out on system 1, with the exception of nitrate + nitrite and ammonia, which were done on both systems. 4.1 SOLIDS Suspended Solids (SS) and Volatile Suspended Solids (VSS) tests were performed on the influent leachate and effluent, once a week. Mixed Liquor Suspended Solids (MLSS) and Mixed Liquor Volatile Suspended Solids (MLVSS) tests were performed on the contents of the anoxic and aerobic reactors, once a week. All analysis of solids was done conforming to the procedure outlined in Standard Methods (1980). 4.2 DISSOLVED OXYGEN AND PH pH measurements were taken on filtered samples of influent, effluent and anoxic and aerobic MLSS. These measurements were done twice a week for the first half of the study. Occasional checks were made during the latter half of the study to ensure that the pH was in the right range (pH 7.5-8.5) for nitrification and denitrification. The pH meter used was a Fisher Accumet Model 320, Expanded Scale Research pH meter. Dissolved oxygen values were spot checked during the entire study to ensure that D.O. levels of between 1 and 3 mg/1 were maintained in the aerobic reactor, as well as ensuring that negligble D.O. was being entrained in the anoxic reactor. A Yellow Springs Instruments Co. Ltd., Model 54A Oxygen meter was used for measuring the D.O. levels. 24 25 4.3 ORP (OXIDATION-REDUCTION POTENTIAL) An ORP probe was installed in the anoxic reactor during the second half of the study, to ensure that reducing conditions were present ORP was measured using a Cole Parmer Digi phase pH Meter, with platinum probe electrode and a calomel reference electrode. ORP was reported as Ecal, that is, millivolts with respect to the calomel electrode. 4.4 BOD (BIOCHEMICAL OXYGEN DEMAND) BODj measurements were carried out on filtered samples from the influent effluent and anoxic and aerobic contents. The filter paper used was a Whatman #4. These measurements were made once a week during day 70 to 120 of steady state. BOD5 was determined by the procedure in Standard Methods (1980). A Y.S.I. Model 54 oxygen meter was used to measure the initial and final dissolved oxygen. The azide modification of the Winkler titration (Standard Methods 1980) was used to standardize the dissolved oxygen probe. The dillution water was seeded (5 mis seed per litre of dilution water) with influent from a UBC campus pilot treatment plant 4.5 COD (CHEMICAL OXYGEN DEMAND) The COD test was performed on filtered samples from the influent effluent and anoxic and aerobic reactors. The filter paper used was a Whatman #4. The test was done twice a week until the latter half of the study, when it was performed only once a week. The procedure followed was as outlined in Standard Methods (1980). Mercuric sulphate was added to the samples to remove any chloride interference. 4.6 TKN (TOTAL K.TELDAHL NITROGEN) TKN on the influent and effluent were run during the first quarter of the study. The effluent was filtered on a Whatman #4 filter paper. This test was carried out to determine what the ratio of TKN to ammonia was. It was determined that the ammonia comprised greater than 70% of the TKN, and this was found to be consistent for the duration of the testing. The method of analysis 26 used is outlined in the Technicon Manual (1974), and the instrument used was a Technicon Auto Analyser 2 S.C. Colorimeter. 4.7 TP (TOTAL PHOSPHOROS1 TP was run on effluent samples from time to time, to ensure that enough phosphoros was being added to the system, and nutrient limiting was not occuring. The method of analysis used is outlined in the Technicon Manual (1974), and the instrument used was a Technicon Auto Analyser 2 S.C. Colorimeter. 4.8 AMMONIA NITROGEN This test was run twice a week for the duration of the study; however ammonia testing on system 2 was started during the latter half of the study. Ammonia was run on filtered (Whatman #4 filter paper) samples from the influent, effluent, and anoxic and aerobic reactors, for system 1. Only the effluent samples were tested for system 2. This was done to check and ensure that the nitrification process was working well, and that no ammonia was escaping into the effluent. The testing method used was the distillation process. In this method, the sample is buffered at a pH of 9.5 with a borate buffer and distilled into a solution of indicating boric acid. The ammonia is then determined titrimetrically with standard sulfuric acid. Buffering decreases hydrolysis of cyanates and organic nitrogen compounds. Ammonia was analysed immediately after sample collection. 4.9 NITRATE AND NITRITE Nitrate + nitrite were run on filtered samples (Whatman #4 filter paper) from the influent, effluent and anoxic and aerobic reactors. During the letter half of the study, nitrate + nitrite testing was also done on the filtered effluent of system 2. These tests were run twice a week. Nitrite itself was also run occasionally on both systems, although this testing was done more frequently towards the end of the study. The tests were carried out according to the Technicon Manual (1974), and the instrument used was a Technicon Auto Analyser 2 S.C. Colorimeter. One change made to the 27 Technicon procedure for nitrates+nitrites was that, instead of using cadmium granules, a cadmium-silver alloy wire in teflon tubing was used as a column. The column was prepared according to a method outlined in Anal. Chem. 1980, 52,1376-1377. 4.10 TRACE METALS Weekly analyses were done for four metals: zinc, manganese, iron and nickel. Metal monitoring was started around day 50 of steady state. Lead and chromium concentrations were also monitored during certain periods. Influent, effluent, anoxic and aerobic samples were checked. Since the metal concentrations in the influent and effluent were low, it was necessary to concentrate these samples. 500 mis of unfiltered influent, and 500 mis of filtered (on Whatman #4 filter paper) effluent were digested down to 50 mis. The digestion method follows closely the method in the recommended EPA procedure (Methods for Chemical Analysis of Water and Wastes, 1979). The anoxic and aerobic mixed liquor samples were first centrifuged, after which the seperated liquid was wasted. The remaining solids were dried at 105° C, ground up and then digested. The digestion method followed is also outlined in the recommended EPA procedure. However, in all digestions, the EPA recommendation to omit HC1 from the digestion and use only HN03 was followed, since it was intended to use the graphite furnace for lead detection. The graphite furnace used was a Perkin Elmer HGA 500 703 Atomic Absorption Spectrophotometer. All other metal analyses were done on another Atomic Absorption Spectrophotometer, a Jarrell Ash AA (Model 810). 5. RESULTS ANP DISCUSSION All data presented in this section was obtained on system 1, unless otherwise specified. Steady state was defined as complete nitrification and no presence of ammonia in the effluent 5.1 CARBON REMOVAL The influent leachate COD was relatively low and ranged between 217 and 318 mg/L. The treated effluent COD concentrations were also in the same range. Before the addition of extra carbon to the anoxic reactor, a 20% removal of the leachate influent COD was obtained. This indicated that the refractory organic component of the influent leachate was about 80%, which is characteristic of older leachates. If no extra carbon had been added to the system, it is possible that the bacterial population might have developed the ability to degrade a larger portion of the refractory organics, but the probability is low. Chian and DeWalle (1976) reported on a study that tested the effectivness of activated sludge treatment on "old" leachate. Results also showed no decrease in COD after an aeration period of 184 hours, due to the refractory nature of the organics. The additional carbon (glucose) added to the system for denitrification formed a major component of the influent COD (shown in Figure 4). Because the leachate influent and effluent COD's were similar, it would seem that the glucose component of the COD was being completely removed. Since the flow rate for the influent varied, a more representative way of showing the influent COD would be in terms of mg/day (shown on Figure 5), rather than mg/L, although the two figures are not dissimilar. Even though there were large fluctuations in the influent COD, the effluent COD remained at a fairly constant level, thus proving the system capable of handling the shock loading with respect to COD. The COD removal is shown in Figure 6, with a mean removal efficiency of 73.9%. The metal spiking did not seem to have an effect on the COD removal efficiency of the system. Figure 7 shows that the percent COD removed also increased as the glucose component of the influent COD increased, an expected result 28 2000 1900 1800 1700 1600 1500 1400 1300 1200 U) 1100 £ 1000 900 Q O O 800 700 600 500 400 300 200 100 0 Legend A LEACHATE INFLUENT • EFFLUENT • LEACHATE INFLUENT + CARBON FEED 0 20 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 4.: Influent and Effluent COD Versus Time o 33 CD E Q O O 18000 17000 16000-15000-14000 -13000 12000-11000-10000-9000-8000-7000-6000-5000-4000-3000-2000 1000 Ni a V L. V I / Legend 4 LEACHATE INFLUENT EFFLUENT LEACHATE INFLUENT + CARBON FEED 0 20 I 40 T 60 1 1 \ 1 1 1 1 1 1—1 1 1— 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 5.: Influent and Effluent COD Versus Time o Figure 6.: COD Removal Efficiency Versus Time -—n—r*n—r*n—i—i—i—i—i—i—i—i—i—i—i—i—i—i—f 0 5 10 15 20 25 30 35 40'45 50 55 60 65 70 75 80 85 90 95 100 COD REMOVAL EFFICIENCY % Figure 7.: Ratio of Carbon Feed to Leachate Influent COD Versus COD Removal Efficiency 33 An average of 85% of the destroyed COD was consumed in the anoxic reactor, presumably by the heterotrophic bacteria involved in the denitrifying process. The rest of the degradable carbon, 15%, was consumed by the heterotrophs in the aerobic reactor (shown in Figure 8). BOD5 measurements were made from day 70 to 120. The leachate influent BOD5 was very low and varied from 8 to 24 mg/L as shown in Figure 9. The actual feed BOD5 was high, as it was comprised mostly of glucose added for denitrification; however, the effluent BOD5 was consistently below 25 mg/L (see Figure 9). A mean removal efficiency of 97.8% was obtained (see Figure 10). 5.2 SOLIDS (SS AND VSS) The influent and effluent SS concentrations are shown in Figure 11. The effluent SS was usually lower than the influent; however, after the start of metal spiking on about day 220, the effluent SS concentration stayed high, and reached a peak on day 290. This peak also corresponded with the highest zinc concentration (Figure 12) added to the system, before inhibition of the nitrification system was detected. The influent zinc concentration to the system, during the spiking period, was at least 95% in soluble form. Since inhibition of the nitrification system was taking place at this time, there were a lot of stressed bacteria present, that simply deflocculated and went out with the effluent, thereby causing a rise in the effluent SS. From days 290 to 310, approximately, the zinc levels were brought down (Figure 12) and correspondingly the effluent SS values also decreased. However, after day 320, zinc spiking was again started; this time, however, a pronounced increase in the effluent SS values was not detected. The zinc spiking was stopped after day 335; therefore, it is difficult to tell if this trend would have continued. It is speculated that the system became acclimitized to high influent concentrations of zinc, and the zinc no longer had the same effect of increasing the effluent SS, as was previously detected. Neufeld (1976) reported that shock loadings of heavy metals to activated sludge result in the formation of a highly stable pinpoint floe, and this condition termed "sludge deflocculation" has resulted in the loss of significant quantities of biomass over the effluent weir of continuous systems. He reported severe deflocculation when zinc levels exceeded 40 mg/L for a 20 day sludge age. This 20 40 60 80 100 120 140 160 180 200 220 240 260 280300 320 340 Days in Steady State Figure 8.: Percentage COD and BOD5 Removal Across the Reactors Versus Time 4^ 1000- \ 100-Legend A LEACHATE INFLUENT x LEACHATE INFLUENT + CARBON FEED • EFFLUENT 10-1-1 1 1 1 1 1 1 1 1 1 1 1 r 70 74 78 82 86 90 94 98 102 106 110 114 118 122 Days in Steady State Figure 9.: BOD5 Versus Time 100 m Q 90-O 88-86 1 i i | | | | | | | | 80 84 88 92 96 100 104 108 112 116 120 124 Days in Steady State Figure 10.: BODs Removal Efficiency Versus Time £ oo Q o 00 Q LJ Q Z LxJ Q_ 00 CO 380 360H 340 320 300-280-260 240 220 200 180 160-1 140 120-1 100 80 60 40 20 0 Note: See Fig. 12 for corresponding zinc metal concentrations in the feed. Legend A INFLUENT SS V —I 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 r 0 2 0 4 0 6 0 8 0 100 120 140 160 180 2 0 0 2 2 0 240 2 6 0 2 8 0 3 0 0 320 3 40 Days in Steady State Figure 11.: Influent and Effluent SS Concentration Versus Time CD E o M 20 19 18 17 16 15 14 13 12 11 10 9 8 7 6 5 4 3 2 1 0 Concentration at which nitrification inhibition was detected ~o Legend A INFLUENT • EFFLUENT Start of zinc spiking 40 60 80 100 120 140 160 Days in Steady State. I 1 •!•• • I* • I r*" i r 1 r 180 200 220 240 260 280 300 320 340 Figure 12.: Influent and Effluent Zinc Concentration Versus Time oo 39 is probably the same phenomena responsible for the high effluent SS in this system, prior to. acclimitation. The influent and effluent VSS concentrations followed a similar trend as the influent and effluent SS concentrations (Figure 13). Without the metal spiking, the clarifier produced an effluent with fairly low VSS, below 40 mg/L. This is clearly seen on Figure 13 from days 0 to 115. This level of suspended solids would enable discharge of the effluent (based on SS) to any receiving water body in British Columbia (Dept. of Lands, Forest and Water Resources (1975)). From days 120 to 190, the effluent VSS appears to have climbed to approximately 80 mg/L. However, this part of the curve was only based on two points, and may not be representative of the VSS values throughout that whole period. The influent VSS concentrations were consistently below 40 mg/L throughout the study (day 0 to 335). The anoxic and aerobic MLSS concentrations are shown on Figure 14. From day 11, which was when glucose was first added to the anoxic reactor, to about day 54, the anoxic MLSS values were up to 20% greater than the aerobic MLSS values. On day 54, the aerobic SRT was changed from 13 to 15 days. After this time the aerobic MLSS values were greater than the anoxic values, remaining so for the duration of the study. This change is also apparent in Figure 15, which shows the anoxic and aerobic MLVSS concentrations. The anoxic and aerobic MLVSS closely follow the trends of the anoxic and aerobic MLSS. There is no apparent reason for the difference between the anoxic and aerobic values for both MLSS and MLVSS concentrations. This difference could be partially due to sampling technique, however, there are probably other explanations that are responsible. The fluctuations in the anoxic and aerobic MLSS and MLVSS are probably due to the variant carbon loading into the anoxic reactor. Figure 16 shows that, as the ratio of carbon feed COD to leachate COD increased, the MLVSS concentrations in both the reactors increased. The MLVSS/MLSS ratios for the anoxic and aerobic reactors are shown in Figure 17. The average MLVSS/MLSS ratio for the anoxic and aerobic reactors are: 0.66 and 0.65 respectively. These values are a little higher than those reported by Jasper, Atwater and Mavinic (1984), who used a similar system to treat leachate from the same landfill. CO E CO Q _J o CO O LxJ Q Z LxJ CL CO CO 300 280 260 240 220 200-180-160-140-120 100-80-60 40-20-0-Note: See Fig. 12 for corresponding zinc metal concentrations in the feed. Legend A INFLUENT VSS • EFFLUENT VSS 1 1 1 1 1—"T 1 1 1 1 1 1 1 r 0 2 0 4 0 6 0 8 0 100 120 140 160 18 0 2 0 0 2 2 0 240 2 6 0 2 8 0-3 0 0 320 3 4 0 Days in Steady State Figure 13.: Influent and Effluent VSS Concentration Versus Time 9000 8500 8000-7500 _j 7000 ^ 6500 E 6000 ^ 5500 5 5000 O CO 4500 Q 4000 g 3500 W 3000 ^ 2500 ^ 2000 1500 1000 500 0 Aerobic SRT changed from 13 to 15 days A, A :"\ v\ / J V Legend • ANOXIC MLSS • AEROBIC MLSS —I 1 1 1 1 1 1 1 1 1 1 1 1 1 1 I 0 20 40 60 80 100 120 140 160 180 200 220 240 260 280300 320 340 Days in Steady State Figure 14.: Anoxic and Aerobic MLSS Concentration Versus Time E CO Q _J o co Q LxJ Q ~Z. LxJ Q_ CO CO 5500 5250 5000 4750 4500 4250 4000 3750 3500 3250 3000 2750 2500 2250 2000 1750 1500 1250 1000 750 500 250 0 Aerobic SRT changed from 13 to 15 days Legend A ANOXIC MLVSS • AEROBIC MLVSS —I 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 0 20 40 60 80 100 120 140 160 180 200 220 240 260 280300 320 340 Days in Steady State Figure 15.: Anoxic and Aerobic MLVSS Concentration Versus Time :o 7-5-7-6.5-6-£ CD Q 5.5 O ° 5-UJ 45-4-3.5 O < _! 2.5H 2 Q O O FI 1-5-1-o CQ 0.5 < o o-Correlation coefficient (r)=0.72 Correlation coefficient (r)=0.77 • A Curves generated by linear regression Legend A ANOXIC • AEROBIC ™ ™ni I l I l l l l i 0 500 1000 1500 2000 2500 3000 3500 4000 4500 5000 5500 MLVSS mg/L Figure 16.: Ratio of Carbon Feed to Leachate Influent COD Versus MLVSS Concentration 1 D Ul CO CO CO > 0.95-0.90-0.85-0.80-0.75-0.70-0.65-0.60-0.55 0.50-0.45-0.40-0.35-0.30-0.25-0.20-0.15-0.10-0.05-Legend ^ ANOXIC REACTOR • AEROBIC REACTOR l I I I I I I I I I I I I I I I 0 20 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 17.: MLVSS/MLSS Ratios in the Reactors Versus Time 45 5.3 AMMONIA REMOVAL The primary objective of this study was to try and remove the ammonia from the leachate. The Port Mann leachate is relatively high in ammonia (NH4+) with concentrations ranging from 120 to 288 mg/L-N. As reported earlier, the ammonia formed a major component of the total kjeldahl nitrogen. The influent, effluent and anoxic ammonia concentrations are shown in Figure 18. The aerobic and effluent ammonia concentrations were found to be the same, and are therefore represented as one curve. The influent ammonia concentrations were consistently above 200 mg/L-N, for the first half of the study. During this time, only well #3 at the landfill site was used for leachate collection. During the latter half of the study, operational problems were frequently encountered with well #3. Whenever this well could not be used, well #2 was utilized; however, well #2 produced a leachate with a relatively lower ammonia concentration. This also gave rise to fluctuations in the influent ammonia concentrations into the system. As shown in Figure 18, essentially complete ammonia removal was achieved, (except for days 200 and 290 to 315), and this was independent of fluctuations in the influent ammonia concentrations. A high effluent ammonia concentration of about 50 mg/L was detected around day 200, due primarily to oxygen concentrations of less then 0.5 mg/L in the aerobic reacter. This low D.O. was caused by a leak in the air line to the aerobic reactor. This condition lasted about two days, before it could be rectified. On approximately day 290, the effluent ammonia concentrations started to rise sharply, with a peak concentration of about 75 mg/L. This also corresponds to the period of initial highest influent zinc concentrations to the system. The system received a concentration 17.6 mg/L zinc on day 289. Since the effluent ammonia concentration was still approximately zero until this time, it would seem that the inhibition level of the nitrification system had been reached, with this concentration of zinc. No inhibition was observed at 14.9 mg/L zinc; however it is conceivable that the inhibitory level for this system could be between 14.9 mg/L and 17.6 mg/L of zinc. It was obvious that this inhibition, in system 1, was being caused by the zinc spiking, since system 2, which was being fed the same leachate, showed no rise in the effluent ammonia 46 300-1 0 20 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 18.: Ammonia Concentration Versus Time CD 4-3-2-1- •A A 180 190 200 210 220 230 240 250 260 270 280 290 300 310 320 330 340 Days in Steady State Figure 19.: Effluent Ammonia from System 2 Versus Time 47 concentrations (see Figure 19). In an attempt to discover if the system had the potential to recover from this inhibition, the influent zinc concentrations were promptly dropped, as shown in Figure 12. The influent zinc concentrations were systematically brought down until the ammonia concentration in the effluent was again at the zero level. This did not happen until day 314, when the influent zinc concentration was 5.5 mg/L. Since the lowering of the influent zinc concentration was started on day 296, this meant the system took 18 days to recover back to its stable state. Once the system had recovered, the influent zinc concentrations were again raised (Figure 12) in order to narrow down the systems inhibitory concentration, assumed to be between 14.9 and 17.6 mg/L. The influent zinc concentration was steadily increased, over a 21 day period, until a concentration of 19.5 mg/L was reached on day 335. Up to this point, no rise in the effluent ammonia concentration was detected. From these results, it would seem that the nitrifiers were now acclimitized to these high levels of influent zinc. Unfortunately, the study had to be terminated on day 335, so the question of how much acclimation had taken place remained unanswered. Most of the ammonia removed in the system was converted to nitrates in the aerobic reactor; however, some of the ammonia was taken up by the biomass. Figure 20 shows that a mean 25% of the influent ammonia was removed across the anoxic reactor by biomass uptake and perhaps, some stripping; however, because of anoxic conditions and pH's of 7.6-8.1, the component of ammonia removed by stripping would be very low. Therefore, the values shown in Figure 20 very closly represent the ammonia removed by the biomass in the anoxic reactor. 5.4 NITRIFICATION The anoxic, aerobic and effluent NOT as N (nitrate + nitrite) concentrations are shown in Figure 21. Measurements for nitrite concentration were done occasionally and found to be negligble at most times; there were, however, instances where nitrite formed a substantial component of NOT, but this was only transient After the addition of glucose to the system on day 11, the pH in the anoxic and aerobic reactors was maintained around pH 8. According to Barnes and Bliss (1983), proportions of free c 100 1 1 1 1 I 1 I I I I I I I I I I I r d 20 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 20.: Percent Ammonia Removal Across the Anoxic Reactor Versus Time Figure 21.: Nitrate+Nitrite Concentration Versus Time ammonia at a pH of 8 and a temperature of 25° C are about 5%, and 36% at a pH of 9 and a temperature of 25° C. The author mentions that inhibition of Nitrosomonas by free ammonia is likely in the range of 10 to 150 mg/1, but Nitrobacter is likely inhibited at much lower concentrations of 0.1 to 1.0 mg/L. This leads to the possibility that in wastes containing high concentrations of NH4VNH3, inhibition of Nitrobacter may lead to the accumulation of nitrite. This could very well be the phenomena occuring in this system; as the pH goes up and down, even in a small range, there might be enough free ammonia produced to inhibit Nitrobacter and cause nitrite formation. Literature, however, places the optimum pH for Nitrosomonas and Nitrobacter between 7.5 and 8.5. Figure 22 shows that the pH in the anoxic and aerobic reactors were maintained in this range. The influent leachate had adequate alkalinity to buffer the H+ ions produced during the nitrification process. Figure 23 shows the percent nitrification values obtained across the aerobic reactor. The calculation used is shown in Figure 23. Since there was little or no ammonia in the effluent leaving the clarifier and 25% (mean) of the incoming ammonia was used up by the biomass in the anoxic reactor, overall percent nitrification values of approximately 75% were expected. The higher percent nitrification values obtained are probably due to the nitrification of organic nitrogen (after conversion), as not all the TKN was made up of ammonia. Some nitrates were produced from the organic nitrogen as well as from the ammonia; therefore percent nitrification (based on incoming and outgoing ammonia in the aerobic reactor) of greater than 100% can be expected. Because of the probable variation in the TKN to ammonia ratio and the variation in the amount of ammonia uptaken by the biomass in the anoxic reactor, fluctuations in the percent nitrification across the aerobic reactor were expected, as shown clearly in Figure 23. The effect of zinc inhibition on the nitrification system is more clearly seen in Figure 24. This shows percent nitrification values in system 1, based on the ammonia concentration into the aerobic reactor, not the difference between the incoming and outgoing ammonia concentrations. The percent nitrification dropped to 21% on day 289, which is when the system started to show signs of inhibition or failure. The percent nitrification values decreased to about 13%, but showed Legend A ANOXIC • AEROBIC 10 20 30 40 50 60 70 80 90 100 110 120 130 140 150 160 Days in Steady State Figure 22.: Anoxic and Aerobic pH Versus Time 160-150-140-130 -120-c 110-q H— 100-o q 90-s— 80-Iz 70-~c 0) 60-O q> 50-CL 40-30-20-10-0-See correlation with metal concentrati6n on Fig. 12 Definition A % Nitrification = T 0 NOT(out of aerobic) -NOT(into aerobic) mg/L-N NH4 (into aerobic)- NH4 (out of qerobic)mg/L- N I 1 I 1 1 II fl 1 1 r— x 100% 20 40 60 80 100 120 140 160 180 Days in Steady T T T 2P0 220 240 260 280 300 320 340 State Figure 23.: Percent Nitrification Versus Time (Derh.:A) IS) c o o o c Q) O Q) Q_ 160-1 150 140 H 130 120-110 -100-9Q 80 70-J 60 50 40 H 30 20 10 H 0 correlation with metal concentration on Fig. 12 Definition B U "1 NOT(ou» of aerobic)-NOT(in*o aerobic) mg/L-N % Nitrification = 1 r ! x 100% j NH4 (into aerobic)mg/L-N i 1 I I \ I I I I I tl I 1 1 1 1 1 1 0 20 40 60 80 100 120 140 160 180 2D0 2 20 240 260 28 0 3 0 0 320 3 4 0 Days in Steady! State Figure 24.: Percent Nitrification Versus Time (Defh.:B) U) 54 signs of increasing on day 307, with a value of 48%. At this point in time, the influent zinc concentrations were being lowered (see Figure 12). The percent nitrification values kept increasing and reached "normal" levels on day 314. This indicated that, as soon as the influent zinc concentrations dropped, the nitrification system started to return. 5.5 DFNTTRTFTCATION Figure 21 shows that after glucose addition to the anoxic reactor on day 11, a substantial decrease in nitrate concentration occured across the system; thus denitrification was improved. Fluctuations in the nitrate concentrations are more a function of the variant carbon loading to the anoxic reactor, than of the influent ammonia concentrations. The fact that 100% denitrification was achieved on several occasions, indicates that the system was not only capable of complete ammonia removal, but also complete denitrification. The COD used across the anoxic reactor varied from 2.8 to 50 mg COD/mg NOT reduced; this is shown in Figure 25. There are a number of possible explanations to account for this variation, however it is possible that a combination of the following scenarios are responsible. Since 1 mg of nitrite exerts a 1.1 mg COD (Standard Methods 1980), transient nitrite concentrations in the recycle would result in an artificially high COD into the anoxic reactor, thus resulting in high COD/NOT ratios. Also any small error in the COD, nitrate and flow measurements could cause a large error in the COD (used)/NOT (reduced) ratios. However, these mechanisms alone cannot be responsible for the substantial increase in COD/NOT ratios about day 290. This situation at day 290 onward, could be explained by the presence of facultative bacteria in the anoxic reactor (other than denitrifiers) that were utilizing some of the incoming glucose. Figure 8 shows that until day 90, the percentage COD used in the anoxic and aerobic reactors was increasing, indicating that after day 90, a stable bacterial population dominated. Figure 25 shows that about this time the COD (used)/NOT (reduced) ratios across the anoxic reactor increased, and averaged about 10:1, until day 270. During this period (after day 90), a certain type of bacteria 54-52-^ 50 CO 48 U) 46-£ 44-42 40-1 ^ 38 CD 36-E 34-^ 32-y 30 28-26-24 22 20-1 18 16 H 14 12-10-8 6-4-2-0-1 + LJJ Q O O See correlation with metal concentration on Fig. 12 "A i 1 r 1 1 1 i r 0 30 60 90 120 150 180 210 240 270 300 330 Days in Steady State Figure 25.: Ratio of COD (used)/Nitrate+Nitrite (reduced) Versus Time On 56 (other than denitrifiers) probably became established and consumed some of the incoming glucose, thereby increasing the COD/NOT ratios over what they were previously. During the period of zinc inhibition (about day 290) there was a decrease in the amount of nitrate reduced (Figure 26), therefore the carbon required for denitrification also decreased; however, the bacteria that were using the carbon for purposes other than denitrification, might have continued to use the same amount of carbon as before, thus giving rise to even higher COD/NOT ratios (Figure 25). Another mechanism that could explain the variation in the COD/NOT ratios about day 290, is the formation (understress) of polysaccharides, by the bacteria, from the glucose. The large fluctuations in the COD/NOT ratios (Figure 25) correlate very well with the variauon in the influent zinc concentrations (Figure 12). There are a multiplicity of bacteria that produce extracellular polymers. These extracellular polymers play a vital role in activated sludge flocculation and the removal of metal by activated sludge (Brown and Lester, 1979). There is evidence to support the fact that the extracellular polymers are responsible for the adsorption of metal ions on to the biofloc. These extracellular polymers are also believed to protect the bacteria from metal ion toxicity (Brown and Lester, 1979). Bitton and Freihofer (1978), from Brown and Lester (1979), investigated the influence of bacterial extracellular polysaccharides on copper and cadmium toxicity to two strains of K. aerogenes, one capsulated strain and one non-capsulated strain. They found that the capsulated polysaccharide acted as a protective mechanism against copper toxicity and to a lesser extent, cadmium toxicity. In the current study, the viable bacteria possibly increased the production of polysaccharides as a ' protective measure, in response to the increasing influent zinc concentrations. This could, therefore, have resulted in an increase in the glucose used in the anoxic reactor, above and beyond the glucose that was required for denitrification. The glucose used for denitrification at this point was low, . because of the decreased conversion of nitrates to nitrogen gas across the anoxic reactor (probably resulting from zinc inhibition of the denitrifiers). Figure 26 shows that about day 290, when the influent zinc concentration was at its maximum, there was a marked decrease in the mg/day of nitrates reduced (independent of a 100-| 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 0 20 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 26.: COD and Nitrate+Nitrite Lost Across the Anoxic Reactor Versus Time 58 decrease in nitrate prodution) across the anoxic reactor. This indicated that both the heterotrophic denitrifying bacteria, and the nitrifiers were inhibited at these high influent zinc concentrations. Since the amount of nitrates reduced, decreased (Figure 26), the COD consumption across the anoxic reactor should have correspondingly dereased; however, it was high enough to produce COD/NOT ratios of about 50:1. Even though the glucose entering the system had decreased at this time, there was enough glucose available that could be used to produce extracellular polymers, because the glucose requirements for denitrification had also decreased. The formation of extracellular polysaccharides would then provide the bacteria with some protection against the influent zinc. The inhibition of the heterotrophic denitrifying bacteria is also indicated on Figure 20. About day 290, there was a marked decrease in the ammonia used across the anoxic reactor. This meant that there was less ammonia used as a nutrient source by the denitrifiers, probably as a result of decreased activity brought on by zinc inhibition. The ORP in the anoxic reactor was monitored during the second half of the study, and is shown in Figure 27. From the ORP values, it is evident that the anoxic reactor had a good reducing environment necessary for denitrification. The ORP readings were also used as a rough indicator of the nitrate + nitrite concentration in the anoxic reactor; a low negative ORP reading indicating a high nitrate + nitrite concentration, while a very low reading indicating a relatively lower nitrate+nitrite concentration present 5.6 REACTION AND UNIT REMOVAL RATES 5.6.1 REACTION RATES The unit nitrification rates are shown in Figure 28. However, these rates are probably not the maximum rates achievable. The hydraulic retention time in the aerobic reactor was approximately 4.8 hours, and it is likely that most of the nitrates were produed in less time. Maximum rates would best be determined using batch tests, such that nitrate production data could be collected. Figure 28 shows the unit nitrification rates steadily dropping from day 1; this is due, in -330--350-| 1 1—: j 1 1 1 1 1 1 1 230 240 250 260 270 280 290 300 310 320 330 340 Days in Steady State Figure 27.: ORP in the Anoxic Reactor Versus Time VO 15 14-- '13-12-* (/) 11-o 10-LY. C 9-o U— 8-o o • — 7-6-5-'c ZD 4-3-2-1-0-Units for unit nitrification rates are: mg NOT produced/hr/gm VSS Low D.O. problems iiiiiii~iiiiiiii 0 20 40 60 80 100 120 140 160 180 200 220 240 260 280300 320 340 Days in Steady State Figure 28.: Unit Nitrification Rates Versus Time 61 part, to the MLVSS increasing in response to an increased amount of glucose added. Unit nitrification rates are calculated as mg NOT produced/hr/gm VSS; therefore, any increase in the VSS will in turn produce lower values of unit nitrification rates. This drop in the rates is misleading, since it does not indicate a "real" drop in the nitrification rates; the major change in VSS is brought about by the carbon heterotrophs and not the nitrifiers. Since the percentage of nitrifiers in the MLVSS is not known, meaningful nitrification rates, which can be used to indicate changes in the performance of the nitrification system, can only be obtained if a constant MLVSS is maintained. In this study, the amount of glucose added to the system would have had to be constant, which it was not After about day 230, there was a fairly steady amount of glucose being fed into the anoxic reactor; therefore, the unit nitrification rates from this day onward could be used as a rough indicator of the nitrification process performance. About day 290, when inhibition was observed, the unit nitrification rates reached their lowest value, after steadily dropping, thus showing signs of inhibition. After day 290, when the zinc concentration was lowered, allowing the system to recover, the rates showed a slight increase. However, once the zinc conentration was again increased, the unit nitrification rates started dropping. The unit denitrification rates are shown in Figure 29. These values also had a large variation because of the increasing carbon loading into the anoxic reactor. However, about day 290 and 330, which corresponds to the maximum influent zinc concentration into the system, a relative drop in the denitrification rates occured. This indicated that the denitrifiers were also understress. 5.6.2 UNIT REMOVAL RATES The ammonia removal rates calculated as mg NH4+ removed/hr/gm VSS are shown in Figure 30. The aerobic ammonia removal rate shown in Figure 30 is very similar to the unit nitrification rates (Figure 28), as expected. The difference between the two curves is that the unit nitrification rate also includes nitrates produced from organic nitrogen, and the aerobic ammonia removal rate includes the ammonia consumed by the carbon heterotrophs, in the aerobic reactor. Units for unit denitrification rates are: mg NOT reduced/hr/gm VSS i —i 1 1 1 1 1 1 1 1 1 1 1 1 1 r f 20 4 0 6 0 80 100 120 140 160 180 200 2 20 240 260 2 8 0 3 0 0 320 3 4 0 Days in Steady State Figure 29.: Unit Denitrification Rates Versus Time £9 64 These differences and experimental error explain the slight variation between the two curves. Knox (1985) reported on a pilot-scale activated sludge plant treating a high ammonia leachate. He reported maximum ammonia removal rates of 5.5 mg-N/hr/gm VSS. In this study, values of up to 11 mg-N/hr/gm VSS were obtained in the initial phase of operation; however, after the MLVSS started to increase, due to glucose addition, the values started dropping. The anoxic ammonia removal rates are also shown in Figure 30, with a mean of 1.6 mg-N/hr/gm VSS. This removal rate was not dependent on nitrification, but on the activity of the denitrifiers. The anoxic ammonia removal rate also reached a minimum about day 290, indicating that the denitrifiers were probably inhibited by the high influent zinc concentrations. The ammonia removal rates calculated as gm NH4+ removed/m3/day are shown in Figure 31. The mean anoxic and aerobic removal rates are 82 and 132 gm-N/m3/day, respectively. The aerobic ammonia removal rates were slightly higher at the beginning of the study, probably because of the higher influent ammonia concentrations at that time. Knox (1985) obtained an ammonia removal rate of 418 gm-N/m3/day at a tempertaure of 21.3 ° C and an influent ammonia concentration of 265 mg/L. The maximum aerobic ammonia removal rate obtained in this study was 248 gm-N/m3/day at room temperature, and an influent ammonia concentration of 220 mg/L. 5.7 METAL REMOVAL Leachates contain a wide variety of metals, many of them in concentrations above discharge guidelines in British Columbia. The activated sludge process used to treat leachates, has also been effective in removing a considerable proportion of the metals entering treatment plants (Brown and Lester 1979). During this study, the four metals regularly monitored were zinc, iron, manganese and nickel. Chromium and lead were also measured on certain occasions. Metal spiking was carried out on system 1, while system 2 which was used as a control, received only leachate. The influent and effluent nickel concentrations are shown in Figure 32. The influent concentrations are relatively low; these concentrations are generally lower than those reported by Jasper, Atwater and Mavinic (1984). The influent nickel concentrations were already consistently Ammonia Removal Rate (gm removed/m /day) 59 0.070 -0.065-_J Q.060 -\ cn E 0.055-c o 0.050-« 7—\ o 0.045-"c <D O c 0.040-o o 1 Q.035-—1 LxJ 0.030-o z 0.025-0.020-0.015-Legend A INFLUENT • EFFLUENT 40 60 80 100 120 140 16.0 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 32.: Influent and Effluent Nickel Concentration Versus Time ON 67 below the British Columbia discharge guidelines of 0.3 mg/L (Dept of Lands, Forest and Water Resources 1975). Sujarittanonta and Sherrard (1981) investigated the effect of nickel on the activated sludge process, and found that the effect of nickel on nitrification was significant. Nitrification was inhibited in reactors receiving Ni+2 at 1 mg/L and COD at 396 mg/L. However, nitrification did proceed in the reactors receiving Ni+2 at 1 mg/L and COD at 787 mg/L. This indicates that the increase in COD concentration could reduce the toxicity of the metal ion. These researchers showed that the toxicity of nickel is also a function of the MLSS:Ni+2 ratio. This is due to the fact that as the influent COD increases, the MLSS will also increase. Since the influent nickel concentrations in this study were much lower than 1 mg/L, and the COD was much higher than 787 mg/L (due to the glucose addition for denitrification), it would seem that the system is capable of handling much higher influent nickel concentrations, before inhibition due to nickel would be expected. The ability of the bacteria to tolerate higher nickel concentrations at the higher COD loadings, may be due, in part, to protection by extracellular polymers, formed by the bacteria from the extra COD. Huang and Sheikhdeslami (1982) also studied the effect of nickel on nitrification. These researchers used rate constants of both ammonia oxidation and nitrate formation under various influences of chromium, nickel and zinc, to evaluate relative inhibition of these metals. Nickel at concentrations of greater than 0.2 mg/L stopped ammonia oxidation completely at MLVSS concentrations of up to 1300 mg/L. The nickel concentrations encountered in this leachate were below 0.1 mg/L and MLVSS concentrations were much greater than 1300 mg/L. From Figure 32, it is evident that not much of the influent nickel was taken up by the biomass, therefore resulting in poor nickel removals. Removal efficiencies ranged from -25% to + 55%, as shown in Figure 33, with normal removals being less than 25%. Jasper, Atwater and Mavinic (1984) also found low nickel removals at 0 to 30%, in their Port Mann treatment study. Cheng et al. (1975) studied metal uptake in the activated sludge process, and found that of the four metals studied (lead, copper, cadmium and nickel), nickel had the least uptake by the biomass. The negative removal efficiency shown in Figure 33, indicates that the influent nickel concentrations Figure 33.: Nickel Removal Efficiency Versus Time 69 had dropped low enough for nickel accumulated in the biomass to be released. The influent and effluent iron concentrations are shown in Figures 34 and 35 (Log scale plot). Figure 34 shows that the iron concentrations in the landfill leachate gradually declined over the course of the study. This was partly due to the use of two different wells for leachate collection. This figure shows that for the latter part of the study period, the influent iron concentrations were lower than for the rest of the period; this is when well #2 was used exclusively. The iron concentration in the leachate was mostly in the suspended form as indicated by the filtered samples (see Figures 34 and 35). Both Figures 34 and 35 show that regardless of the fluctuations in the influent iron concentrations, the filtered effluent iron concentrations remained fairly low. However, the British Columbia discharge guideline of 0.3 mg/L (Dept. of Lands, Forest and Water Resources 1975) was not met consistently. Since no ammonia conversion inhibition was detected before the zinc spiking, it would seem that the influent iron concentrations in the leachate, even though they were high,had no adverse effect on the process. The iron removal efficiency was quite high and averaged 95.6% for the period studied. The removal percentages are shown in Figure 36. Jasper, Atwater and Mavinic (1984) also reported high iron removals, at 90 to 98%. These researchers also had higher influent iron concentrations than those found in the leachate used for this study. Most of the iron accumulated in the sludge, as indicated by the high removal efficiencies. The sludge iron concentrations are shown in Figure 37. There was little difference between aerobic and anoxic sludges. This figure also shows a decreasing trend in the sludge iron values, an expected result, since the influent concentrations were also dropping throughout the course of the study. During the first quarter'of the study, the sludge iron values were comparable to those obtained by Jasper, Atwater and Mavinic (1984), and some values were higher than the maximum reported by Robinson (1980) from Jasper, Atwater and Mavinic (1984). By the study's end, the sludge iron values had dropped to below 15000 mg/Kg. Chromium and lead were also measured on certain occasions during an eight month period. The concentrations found are shown in Table 2. Both chromium and lead have been found \ 1 IS l\ Legend A INFLUENT filtered • INFLUENT unfiltered • EFFLUENT filtered * » i 8) 1 • \ 4 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 34.: Influent and Effluent Iron Concentration Versus Time o 100 10 E. o 1-o.i o.oi 7A fo / x .w. / - Legend J - A INFLUENT filtered j • INFLUENT unfiltered •^•MM m m m MMM * m m m t • EFFLUENT filtered n 1 1 " 1 1 1 T i • r i1 "i i i i 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 35.: Influent and Effluent Iron Concentration Versus Time - (semi-log plot) Iron Removal Efficiency % 135000-1 c o c Q) o c o o o or o Q ZD _J 00 15000-5000-Legend A ANOXIC SLUDGE • AEROBIC SLUDGE 40 60 —1 1 1 1 1 1 1 1 1 1 "I 1 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 37.: Sludge Iron Concentration Versus Time Table 2. Chromium and Lead Concentrations* Chromium Lead Day Influent Effluent Influent Effluent 54 < 0.005 < 0.005 < 0.005 .016 75 < 0.005 < 0.005 < 0.005 82 < 0.005 < 0.005 .006 .006 89 < 0.005 < 0.005 < 0.005 96 < 0.005 < 0.005 < 0.005 .006 103 < 0.005 < 0.005 .005 .006 110 < 0.005 < 0.005 < 0.005 117 < 0.005 < 0.005 < 0.005 .010 124 < 0.005 < 0.005 .005 138 < 0.005 < 0.005 < 0.005 .006 145 < 0.005 < 0.005 < 0.005 < 0.005 152 < 0.005 < 0.005 < 0.005 < 0.005 169 < 0.005 < 0.005 .005 190 < 0.005 < 0.005 < 0.005 < 0.005 197 < 0.005 < 0.005 < 0.005 < 0.005 204 < 0.005 < 0.005 < 0.005 < 0.005 208 < 0.005 < 0.005 < 0.005 < 0.005 215 < 0.005 < 0.005 < 0.005 < 0.005 222 < 0.005 < 0.005 < 0.005 229 < 0.005 < 0.005 < 0.005 < 0.005 257 < 0.005 < 0.005 .008 < 0.005 264 < 0.005 < 0.005 < 0.005 < 0.005 271 < 0.005 < 0.005 < 0.005 < 0.005 278 < 0.005 < 0.005 < 0.005 < 0.005 * All concentrations in mg/L. 75 to be inhibitory to nitrifiers. Martin (1979), from Martin and Richard (1982), found toxicity thresholds for chromium (Cr+J) and lead (Pb+J) at 1 mg/L and 0.5 to 1 mg/L respectively. As shown in Table 2, the values encountered in the leachate during this study are much lower than the reported toxicity thresholds, therefore, the chromium and lead concentrations in this leachate were not high enough to induce inhibition of the nitrifiers. The influent and effluent manganese concentrations are shown in Figures 38 and 39, the latter being a semi-log plot. The manganese concentrations in the leachate were low; however, they varied quite a bit, and ranged from 0.02 to 0.3 mg/L. Before metal spiking, the system had been working successfully at removing essentially 100% of the ammonia from the leachate, and it was obvious that no inhibition was taking place; in other words, no nitrification inhibition was being detected. A previous study conducted at the University of British Columbia by Jasper, Atwater and Mavinic (1984), used an almost identical system to treat leachate from the same landfill. This study found limited success with nitrification, and it was obvious that some sort of inhibition was taking place. Therefore, a comparison was made to find out what was in the previous leachate that was not in this leachate, something that could have caused nitrifier inhibition. The only detectable difference was the higher manganese and zinc concentrations in the previous leachate. In the previous study, the influent manganese concentration was higher than the level in this leachate for about fourteen weeks. Although manganese has not been reported to be inhibitory or toxic to nitrifiers, the influent manganese concentrations into the system were increased anyway, to see if inhibition would also take place. On about day 190 of steady state operation, manganese in the form of manganous chloride was added to the system (system 1), with system 2 remaining as control. As noted in both Figures 38 and 39, the influent manganese concentrations increased substantially. The maximum influent concentration obtained was about 12.5 mg/L, which was greater than the maximum leachate concentrations observed by Jasper, Atwater and Mavinic (1984). Throughout the manganese spiking period, there was no nitrification inhibition detected, and it was concluded that manganese would not inhibit the nitrifiers. It was then decided to decrease the influent manganese concentrations and E Ld CO LU z < 13 12.5 12 11.5 11 10.5 10 9.5 9 8.5 8 7.5 7 6.5 6 5.5 5 4.5 4-3.5 3 2.5 2 1.5 1-0.5 0 Legend A INFLUENT • EFFLUENT 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 38.: Influent and Effluent Manganese Concentration Versus Time as 100T 1(h o.u o.ou .001-Start of manganese spiking Legend A INFLUENT • EFFLUENT V 40 1 1 r-—i 1 1 1— 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 39.: Influent and Effluent Manganese Concentration Versus Time - (semi-log plot) 78 start adding zinc. The influent manganese concentrations were still maintained at a comparatively high level of around 4 mg/L, for the duration of the study. If inhibition was due to a synergistic effect between zinc and manganese, then with this constant high influent manganese concentration, and zinc spiking, inhibition would still be detected. Regardless of the large fluctuations in the influent manganese concentrations, a high removal efficiency was still maintained, although less so after zinc spiking commenced. The removal values are shown in Figure 40, with an overall average value of 87.6%. However, after zinc spiking began, the average removal dropped from 93.4% to 78.4%. High manganese removal efficiencies in the activated sludge process have also been reported by other researchers (Jasper, Atwater and Mavinic 1984). The effluent manganese values were below British Columbia discharge guidelines of 0.05 mg/L (Dept of Lands, Forest and Water Resources, 1975) until day 190 of steady state operation. After this time, the system was spiked with manganese, and the effluent concentration was, for the most part, above the discharge guidelines. For the most part, the manganese accumulated in the sludge, as indicated by the high removal efficiencies. The manganese sludge values are shown in Figure 41 and 42, the latter being a semi-log plot There was little or no difference between anoxic and aerobic sludges. The sludge manganese plots parallel the influent and effluent plots quite well, and clearly depict the manganese spiking period. The sludge manganese values reached a maximum of 7700 mg/Kg dried sludge, which is far in excess of the values reported by Jasper, Atwater and Mavinic (1984); however, it is close to the high end of the range reported by Robinson (1980) in Jasper et al. (1984). Zinc was the only other metal monitored on a regular basis. Since an increase in the influent manganese concentrations had no detectable inhibitory effect on the nitrification system, the next step was to spike with additional zinc. The previous study by Jasper et al. (1984) which reported nitxifier inhibition, had influent zinc concentrations of up to 4 mg/L, higher than in the leachate used herein. Thus, zinc additions to system 1, in the form of zinc chloride, were started about day 230. This is clearly shown in Figure 12 as well as Figure 43, the latter being a semi-log O 20-^ 15-10-5- | -0-fi—i 1 1 1 1 1 1 1 1 1 1 1 1 1 40; 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 40.: Manganese Removal Efficiency Versus Time —i 8000-1 cn 7500-cn i— 7000-t 6500-c o 6000--*— o 5500-c Q) 5000-O c o 4500-O 4000-Ld CO I. I 3500-ANI 3000-ON' 2500-< 2000-Ld 1500-o Q 1000-_J CO 500-0-Legend A ANOXIC SLUDGE • AEROBIC SLUDGE Start of manganese spikinc 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 ! Days in Steady State Figure 41.: Sludge Manganese Concentration Versus Time 10000 o Q ZD _l CO iooH—i—i—i—-i 1—i 1—i 1 1 1 1 1 1— 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 42.: Sludge Manganese Concentration Versus Time - (semi-log plot) • i i i i i i—i 1 i 1 1 1 1 1 1 r 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State. Figure 43.: Influent and Effluent Zinc Concentration Versus Time - (semi-log plot) oo N) 83 plot The influent zinc concentrations were brought up to 14.9 mg/L by day 285, over a 55 day period. Up to this point in time, no inhibition was detected, in other words, there was no ammonia in the effluent from the treatment system (see Figure 18). The influent zinc concentration was subsequently brought up to 17.6 mg/L by day 290, which is when ammonia was first detected in the effluent (see Figure 18). At this time, a substantial drop in percent nitrification was also observed, as shown in Figure 24. However, the percent nitrification (based on percent nitrification as defined in Figure 24) had dropped to about 13% and not to zero; therefore, this concentration was inhibitory and not toxic. The control system (system #2), which was being fed leachate, but no additional metals, showed no signs of inhibition from the time it was started. At this point in time, it was decided to decrease the influent zinc concentrations, to see if the system could recover. The concentrations were dropped until there was no ammonia detected in the effluent, that is, the system was back at "steady state". The concentration of zinc had to be dropped to about 5.5 mg/L (Figure 12 and 43), before there was no detectable ammonia in the effluent (Figure 18); this was reached about day 305. Figure 24 also shows that when the influent zinc concentrations were being dropped, the percent nitrification values started increasing. Because the influent zinc concentrations were increased from 14.9 mg/L, at which concentration no inhibition was observed, to 17.6 mg/L, when inhibition was observed, the inhibitory concentration for this system was thought to be anywhere between these two concentrations. Therefore, it was decided to increase the influent zinc concentrations again, in system 1, in order to identify the inhibitory zinc level for this system. The influent zinc concentration was subsequently increased to 19.5 mg/L, over a 30 day period, but there was no ammonia in the effluent This indicated that the bacteria were now acclimitizing to the high influent zinc concentrations. Unfortunately, the study had to be terminated on day 340, and therefore, it is not known whether this trend would have continued. Knoetze, et al. (1980) studied the inhibitory effects of various heavy metals on a nitrification-denitrification activated sludge process. These researchers reported that for a 10 day sludge age, nitrification was not inhibited at zinc (Zn+2) concentrations below 10 mg/L. This is in 84 agreement with the work done for this study. Robinson and Maris (1982) reported on laboratory-scale biological units treating leachate. The leachate used had fairly low concentrations of ammoniacal nitrogen (76 mg/L); however, no nitrification took place. In units where SRT was greater than 5 days, effluent ammonia values were below 1 mg/L. This removal of ammonia was due principally to bacterial uptake. These researchers, at first, postulated that the lack of nitrification was a result of the combined effects of low temperature (10° C) and low SRT. The SRT however, was subsequently taken as high as 20 days, and still no nitrification occured. The mean influent zinc concentration encountered by these researchers was 17.6 mg/L. It may very well be that this high zinc concentration was indeed responsible for the lack of nitrification, observed by Robinson and Maris. In fact, coincidentally, the zinc concentration at which inhibition first occured in the current study was also 17.6 mg/L, at a 15 day aerobic SRT. Zinc toxicity was also reported by Martin (1979) from Martin and Richard (1982). The Nitrosomonas toxicity threshold for zinc was 10 mg/L; however, the authors add that this value is illustrative only, because it was not determined under specific conditions and did not take into account the possible synergistic effect due to the presence of other inhibitors. Basically, the information available in literature on zinc inhibition is inconsistent Huang and Sheikhdeslami (1982) also studied the metal inhibition of nitrification using rate constants of both ammonia oxidation and nitrate formation. They found that zinc concentrations of 0.6 mg/L (probably soluble) with an MLVSS of up to 2133 mg/L, reduced the rate constants to lower than 20% of the control. In this case, a fairly low concentration of zinc caused substantial inhibition. Despite the problems associated with zinc and nitrification, the system utilized in this study was quite effective in removing zinc, and the removal efficiencies are shown in Figure 44. The negative removal efficiency obtained around day 90 was probably due to the low influent zinc concentrations at that time; in other words, some of the zinc adsorbed onto the biomass went back into solution to maintain equilibrium. The high influent zinc concentration did not seem to affect the zinc removal as it did the removal of manganese. A mean zinc removal efficiency of 75% was Days in Steady State oo Figure 44.: Zinc Removal Efficiency Versus Time <-* 86 obtained. The effluent zinc concentration was below the British Columbia discharge guideline of 0.5 mg/L (total zinc) (Dept. of Lands, Forest and Water Resources, 1975) up to about day 280, even though the influent concentrations were around 4 mg/L. However, after day 280, the influent concentrations were increased to fairly high levels, thus increasing the effluent zinc levels to over 0.5 mg/L. The sludge zinc concentrations are shown in Figure 45. There was little difference between the anoxic and aerobic sludges. The increase in sludge zinc concentrations when zinc spiking was initiated on day 230 is clearly depicted. The sludge zinc values reached a maximum of over 10000 mg/Kg dried sludge. The maximum sludge zinc concentrations obtained by Jasper, et al. (1984) was 1800 mg/Kg dried sludge. It is interesting to note that when inhibition was detected on about day 290, the sludge zinc values were at their maximum of over 10000 mg/Kg, but when the influent zinc levels were reduced and "steady state" (zero ammonia in the effluent) was reached again, the sludge zinc concentrations were still at the previously established inhibition level. This seems to indicate that inhibition of the nitrifiers was due more to the influent zinc concentrations than the zinc adsorbed by the biomass. It also seems to indicate a relatively slow "flushing out" phenomenon between the active biomass and the zinc metal species. These results appear to be contradictory to conclusions made by Martin and Richard (1982) as well as others, who studied the toxic effects of several heavy metals (nickel, cadmium, copper and zinc) on nitrification. These researchers reported that inhibition of nitrification by metallic ions is attributable to adsorption of the metallic ions on the floes containing the nitrifying bacteria, and thereby partially or completely blocking the enzyme mechanisms. There was no indication, by these authors, that actual influent metal concentrations affected any change in performance by the nitrifying organisms. Because of the "apparent" contradiction between published data and the results of this research, especially as it pertains to zinc, and because of the importance of expanding the knowledge associated with nitrification-denitrification toxicity, it is obvious that much more work is needed in this vital research area. ' i i i i i i i i i i i i ii i r 40 60 80 100 120 140 160 180 200 220 240 260 280 300 320 340 Days in Steady State Figure 45.: Sludge Zinc Concentration Versus Time oo 88 The purpose of this study was to determine conclusively whether or not a high ammonia leachate could be treated successfully, using a nitrification-denitrification activated sludge system. This study has 340 days of "steady state" operation data, that shows that the ammonia in a high ammonia leachate can be removed, by essentially 100%. The leachate contains a wide variety of metals, many of which are known to inhibit the nitrification process. This study showed that influent zinc severely inhibited the nitrification process, although bacterial acclimitization appeared to be possible. Chromium, lead and nickel are some of the other metals known to inhibit the nitrification process. Studies to determine the inhibitory concentration of these other metals are vital if treatment of a high ammonia leachate, on a full scale, is to be successful. 0 6. CONCLUSIONS AND RECOMMENDATIONS 6.1 CONCLUSIONS Municipal leachate, obtained from the Port Mann landfill site near Vancouver, was treated in a biological treatment system, consisting of an anoxic basin (for denitrification), an aerobic basin (for carbon and ammonia removal) and a final clarifier (with provision for recycle back to the anoxic basin). Solids wasting, to control SRT, was carried out directly from the aerobic reactor. This unit ran for a total of 340 days and based on the operational procedure and time, the following conclusions can be made: 1. Overall COD removal was excellent and averaged 73.9%; however, this was dependent on the amount of glucose entering the system for denitrification, since most of the influent leachate COD consisted of refractory organics. However, before the addition of glucose, a 20% removal of the leachate influent COD was obtained. Of the COD removed across the system, 85% was used in the anoxic reactor for denitrification, and the remaining 15% was used up by the heterotrophs in the aerobic reactor. The COD (used)/NOT (reduced) ratio averaged 13:1 in the anoxic reactor, and ranged from 2.8 to 50:1. The COD used across the anoxic reactor was probably made up of COD used for denitrification, COD used by bacteria other than denitrifiers, and COD used to produce polysaccharides. 2. Essentially 100% of the ammonia in the influent leachate was removed. The effluent from the system averaged less than 1 mg/L most of the time. Approximately 25% of the incoming ammonia was utilized as a nutrient source by the bacteria in the anoxic reactor; the rest of the ammonia was successfully nitrified, assimilated and stripped in the aerobic reactor. The mean anoxic and aerobic ammonia removal rates are 82 and 132 gm-N/m3/day, respectively. 3. Since the influent leachate had very low biodegradable carbon, an extra carbon source (glucose) was added to aid in denitrification. Essentially complete denitrification was obtained at various 89 90 times in the study. Only partial denitrification was obtained the rest of the time, because of the variant carbon loading to the anoxic reactor. 4. For the four metals monitored regularly, the mean removal efficiencies were: zinc=75%, iron=95.6%, manganese=87.6% and nickel = 9.5%. The effluent iron values were not consistently below the B.C. discharge guidelines. Before additional metal spiking, the rest of the metals in the effluent were always below the B.C. discharge guidelines. After deliberately increasing the zinc and manganese influent levels, the respective effluent concentrations were consistently above the B.C. discharge guidelines. Other metals in the leachate were of little consequence, since they were of low influent concentration to begin with. 5. Three of the metals monitored regularly, zinc, iron and manganese, accumulated extensively in the sludge. For all these metals, the sludge metal concentration followed closely the trend of the influent metal concentration. The sludge iron levels reached a maximun of about 115000 mg/Kg at the beginning of the study, and continued decreasing from then on. The sludge zinc and manganese values were also high, due to spiking, and reached values of 13400 mg/Kg and 7600 mg/Kg respectively. Nickel removal was very low and therefore resulted in very low (non detectable) nickel sludge values. 6. When the manganese influent concentrations were elevated, no inhibition to nitrification was observed. Throughout this period, nitrification took place, and all of the influent ammonia was removed. However, zinc spiking caused substantial inhibition of the nitrification system. The "inhibition concentration" appeared to be between 14.5 and 17.6 mg/L total zinc (>95% soluble), resulting in approximately 70 mg/L of ammonia in the effluent (feed=216 mg/L). There is also a possibility that this inhibition was due to a synergistic effect between manganese and zinc, since the manganese concentration was also maintained at a fairly high level during zinc spiking. The high zinc concentration also resulted in a MLSS concentration of about 355 mg/L in the effluent 91 Note: After the initial inhibition, the zinc concentration was subsequently decreased, thus allowing the nitrification system to recover. When the system returned to normal, the influent zinc concentrations were again increased; however, no inhibition was detected this time, even up to total zinc (>95% soluble) influent levels of 19.5 mg/L. This seems to indicate that the system had acclimitized to the high concentrations of zinc. At this point, the study was terminated and, therefore, the extent of acclimitization could not be determined. 6.2 RECOMMENDATIONS Based on the results of this study, it is recommended that: 1. A study be carried out whereby the zinc concentration in the system is again increased to the inhibition point (w.r.L nitrification), then decreased, so that the system can recover. Following this, the zinc levels should be increased again, to see if acclimitization can be achieved. In this manner, the extent of acclimitization might be determined. This study would be a direct continuation from this thesis, at the point of termination. During the last phase of this work, a high influent manganese concentration was also maintained during the zinc spiking period; therefore, the inhibition caused could have been a synergistic effect between zinc and manganese. In the next phase of research, only the zinc concentration would be increased and the influent manganese concentrations left as found in the leachate (relatively low levels). If inhibition did occur this time, it would prove that inhibition to nitrification was caused only by zinc and not a synergistic effect between zinc and manganese. 2. A study be undertaken whereby the aerobic SRT is varied when the system is receiving a zinc concentration close to the inhibitory concentration for nitrification, to see what effect SRT might have on system recovery at sustained zinc levels. If inhibition could be controlled by SRT, then this information would be invaluable in the operation of a full-scale treatment plant, since SRT is easy to control and change in response to influent metal levels. This would also indicate if the sludge 92 zinc concentration is directly related to SRT manipulation. Work done for this thesis has indicated that, with respect to zinc, it might be the influent concentration, rather than the sludge concentration, that causes inhibition of the nitrification system. However, the data base is inconclusive at this point and needs expansion. 3. A study be carried out where nitrification rates would be measured while increasing the zinc concentrations to the system (mg NH4 oxidized/time/gVSS). This would indicate relative inhibition to the nitrifiers for different zinc concentrations, and this information could be used for any nitrification system. However, the system used for this research, should maintain a constant MLVSS concentration, otherwise the nitrate rates measured would be misleading. 4. Cold temperature work should also be done in conjunction with the three studies discussed above. In the northern hemisphere, liquid waste temperatures in the winter can drop quite low; this low temperature also stresses the nitrifiers, causing temperature inhibition. Therefore, the cumulative effect of cold temperature and metal inhibition must be determined. This information would be very useful for successful "high ammonia leachate" treatment during the winter. 5. A study be undertaken to determine the effect and inhibitory concentration (to nitrification) of other heavy metals found in leachate. Nickel, cadmium, chromium and copper have also been postulated as being inhibitory/toxic to nitrifiers. REFERENCES Atwater, J.W., "Fraser River Estuary Study, Water Quality-Impact of Landfills", Environment Protection Service, Environment Canada, Vancouver, 1980. Barnes, D., and Bliss, P.J.,"Biological Control of Nitrogen in Wastewater Treatment", Publ. by E. and F.N. Spon, New York, 1983. Bjorkman, V.B., and Mavinic, D.S., "Physio-Chemical Treatment of a High-Strength Leachate", Proc. of 32nd Annual Purdue Industrial Waste Conference, West Lafayette, Indiana, 1977. Boyle, W.C, and Ham, R.K., "Biological Treatability of Landfill Leachate", J. Water Pollution Control Fed., 56, 5,1974, pp. 860-872. Brown M.J., and Lester, J.N., "Metal Removal in Activated Sludge: The Role of Bacterial Extracellular Polymers", Water Research, 13,1979, pp. 817-837. Chian, E.S.K., "Stability of Organic Matter in Landfill Leachates", Water Research, 11,1977, pp. 225-232. Chian, E.S.K., and DeWalle, F.B., "Sanitary Landfill Leachates and Their Treatment", J. of the Environmental Engineering Division, 1976, pp. 411-431. Cheng, M.H., Patterson, J.W., and Minear, R.A., "Heavy Metals Uptake by Activated Sludge", J. Water Pollution Control Fed., 47,1975, pp. 362-376. Cook, E.N., and Foree, E.G., "Aerobic Biostabilization of Sanitary Landfill Leachate", J. Water Pollution Control Fed., 46, 2,1974, pp. 380-392. Department of Lands, Forests and Water Resources, "Pollution Control Objectives for Municipal Type Waste Discharges in British Columbia", Dept. of Lands, Forests and Water Resources, Water Resources Service, Victoria, B.C., 1975. Fuller, W.H., Alesii, B.A., and Carter, G.E., "Behaviour of Municipal Solid Waste Leachate", J. Environ. Sci. Health, A14. Henry, J.G., Prasad, D., Sidhwa, R., and Hilgerdenaar, M., "Treatment of Landfill Leachate by Anaerobic Filter: Part 1: Laboratory Studies", Water Poll. Res. J. Canada, 17,1982, pp. 37- 46. Hooper, A.B., and Terry, K.R., "Specific Inhibitors of Ammonia Oxidation in Nitrosomonas", J. of Bacteriology, 115, 2,1973, pp. 480-485. Jasper, S.E., Atwater, J.W., and Mavinic, D.S., "Characterization and Treatment of Leachate from a West Coast Landfill", University of British Columbia, Dept of Civil Eng., Draft report prepared for: Waste Management Branch-Ottawa and Environmental Protection Service, Wastewater Technology Center-Burlington, 1984. Keenan, J.D., Steiner, R.L., and Fungaroli, A.A., "Landfill Leachate Treatment", J. Water Pollution Control Fed., 52,1,1984, pp. 27- 33. Knoetze, C, Davies, T.R., and Wiechers, S.G., "Chemical Inhibition of Biological Nutrient Removal Processes", Water S.A., 6,4,1980, pp. 171-179. Knox, K., "Treatability Studies on Leachate from a Co-disposal Landfill", Environmental Pollution 93 94 (Series B), 5,1983, pp. 157-174. Knox, K., "Leachate Treatment With Nitrification of Ammonia", Water Res., 19, 7,1985, pp. 895-904. McDougall, W.J., Fusco, R.A., and O'Brien, R.P., "Containment and Treatment of the Love Canal Landfill Leachate", J. Water Pollution Control Fed., 52,12,1980, pp. 2914-2924. Neufeld, R.D., "Heavy Metals-Induced Deflocculation of Activated Sludge", J. Water Pollution Control Fed., 48,1976, pp. 1940-1947. Patel, V.P., Hoye, R.L., and Toftner, R.O., "Gas and Leachate Summary", Munic. Solid Waste: Land Disposal, Proc. of the Annual Res. Symp., 5th, Orlando, Florida, Publ. by EPA, 1979. R.obinson, H.D., and Maris, P.J., "The Treatment of Leachates from Domestic Waste in Landfill Sites", J. Water Pollution Control Fed., 57,1,1985, pp. 30-38. Robinson, H.D., and Maris, P.J., "The Treatment of Leachates from Domestic Wastes in Landfill Sites", Presented at Annual B.C.W.W.A. Conference, Vancouver, B.C., 1982. "Standard Methods for Examination of Water and Wastewater", American Public Health Association Inc., 15th Edition, 1980. Stegmann and Ehrig, "Operation and Design of Biological Leachate Treatment Plants", Prog. Water Tech., 12,1980, pp. 919-947. Sujarittanonta, S., and Sherrard, J.H., "Activated Sludge Nickel Toxicity Studies", J. Water Pollution Control Fed., 53,1981, pp. 1314-1322. Technicon Analyser Industrial Methods, No. 321-74A and No. 327-74W. Temoin, E.P., "Nutrient Requirements for Aerobic Biostabilization of Landfill Leachate", Master of Applied Science Thesis, Department of Civil Engineering, University of British Columbia, October, 1980. Uloth, V.C., and Mavinic, D.S., "Aerobic Bio-Treatment of a High-Strength Leachate", J. of the Environmental Engineering Division, ASCE, 103, No. EE4,1977, pp. 647-661. Wigh, R.J., and Brunner, D.R., "Leachate Production from Landfilled Municipal Waste", Munic. Solid Waste: Land Disposal, Proc. of the Annual Res. Symp., 5th, Orlando, Florida, Publ. by EPA, 1979. Winkler, M., "Biological Treatment of Wastewater", Ellis Horwood Ltd., Chichester, England, 1981. 2apf-Gilje, R., and Mavinic, D.S., "Temperature Effects on Biostabilization of Leachate", J. of the Environmental Engineering Division, ASCE, 107, NoEE4,1981, pp. 653-663. APPENDICES (DATA FOR FIGURES! Nitrate+Nitrite - Anoxic, Aerobic and Effluent 97 Influent, Anoxic and Effluent Ammonia (system 1) 99 Influent and Effluent COD 101 Influent and Effluent BOD5 BOD5 Removal Efficiency '. 102 Percent COD Removed Across the Anoxic and Aerobic Reactors Percent BOD5 Removed Across the Anoxic Reactor 103 COD (used)/NOT (reduced) Across the Anoxic Reactor COD Removal Efficiency MLVSS/MLSS (Aerobic) 104 Carbon Feed COD/Leachate Influent COD Versus COD Removal Efficiency, Anoxic MLVSS and Aerobic MLVSS 105 Percent BODs Removed Across the Aerobic Reactor Anoxic and Aerobic pH , 106 ORP in the Anoxic Reactor Unit Nitrification and Denitrification Rates 107 Anoxic and Aerobic Ammonia Removal Rates (mg NH/ removed/hr/gm VSS) Anoxic Ammonia Removal Rates (gm NH/ removed/mVday) 108 Percent Nitrification (system 1) Defn. A and B Percent Ammonia Removal Across the Anoxic Reactor 109 Metals Ill Influent and Effluent SS Influent VSS 118 95 96 Effluent VSS Anoxic and Aerobic MLSS 119 Anoxic and Aerobic MLVSS MLVSS/MLSS (Anoxic) 120 Aerobic Ammonia Removal Rates (gm NH„+ removed/mVday) 121 DAY NITRATE + NITRITE (ANOXIC) (mg/L) NITRATE + NITRITE (AEROBIC) (mg/L) NITRATE + NITRITE (EFFLUENT) (mg/L) 1 208 239 239 5 234 234 244 8 197 237 240 12 197 225 226 15 120 155 150 19 80 135 135 22 72 121 120 26 59 111 135 29 75 118 119 33 52 92 91 36 72 117 115 40 58 91 96 43 41 91 85 47 31 68 68 50 0 38 83 54 4 41 34 61 36 80 80 71 65 95 95 78 2 22 22 82 1 38 43 85 1 32 32 89 0 28 28 92 0 40 40 96 24 57 55 99 42 89 88 103 3 69 76 106 0 48 48 110 64 106 106 113 84 111 115 117 47 74 74 120 32 67 67 124 62 92 92 138 83 120 - 121 141 78 115 117 145 82 118 120 152 52 75 75 155 44 76 76 190 30 56 56 194 0 15 15 197 5 3 3 200 0 24 24 204 2. 31 31 208 2 31 31 215 0 16 16 218 1 16 16 222 1 13 13 225 11 37 37 229 22 41 41 232 13 34 34 257 45 58 58 260 54 87 87 264 61 78 78 (CONTINUED) DAY NITRATE + NITRATE + NITRATE + NITRITE (ANOXIC) NITRITE (AEROBIC) NITRITE (EFFLUENT) (mg/L) (mg/L) (mg/L) 267 61 78 78 271 15 22 22 274 36 51 51 278 25 34 34 281 28 41 41 285 41 57 57 289 33 45 45 293 20 32 32 296 12 23 23 307 74 89 89 311 32 44 44 314 13 25 25 317 20 31 31 321 39 50 50 324 1 11 11 328 1331 1 6 6 335  8 8 DAY INFLUENT AMMONIA ANOXIC AMMONIA EFFLUENT AMMONIA (SYSTEM 1) (SYSTEM 1) (SYSTEM 1) (mg/L) (mg/L) (mg/L  1 265 42 0.3 5 258 41 0.3 8 288 50 0.8 12 235 52 0.3 15 220 50 0.6 19 225 40 1.0 22 240 34 0.4 26 248 48 5.8 29 238 42 4.0 33 238 32 4.0 36 269 42 1.0 40 260 42 2.6 43 ' 263 43 1.1 47 263 36 1.0 50 235 40 2.3 61 235 35 1.5 71 219 26 0 78 230 42 2.0 82 223 37 2.0 85 224 33 2.0 89 227 38 2.0 92 238 55 2.0 96 230 37 3.0 99 247 46 2.0 103 235 - 10.0 106 241 67 11.0 110 235 37 0 113 249 19 0 117 235 23 0 120 235 31 0 124 204 25 0 138 238 - 0 141 223 27 0 145 215 28 0 152 203 26 0 155 210 30 0 190 170 18 0 194 174 19 0 197 196 68 53.0 200 221 23 0 204 213 23 0 208 231 25 0 215 160 14 0 218 157 17 0 222 144 6 0 225 148 29 0 229 134 19 0 232 146 19 0 257 202 19 0 260 215 26 0 264 170 13 0 100 (CONTINUED) DAY INFLUENT AMMONIA ANOXIC AMMONIA EFFLUENT AMMONIA (SYSTEM 1) (SYSTEM 1) (SYSTEM 1) (mg/L) (mg/L) (mg/L267 159 17.5 0 271 142 15 0 274 134 14 0 278 131 11 0 281 168 18 0 285 149 17 0 289 187 57 40 293 216 92 71 296 217 85 72 300 214 30 19 307 157 31 12 311 151 22.5 8 314 143 7.5 0 317 122 11 0 321 134 14 0 324 133 13 0 328 132 12 0 331 141 13.3 0 335 129 11 0 101 LEACHATE INFLUENT + DAY LEACHATE INFLUENT COD EFFLUENT COD CARBON FEED COD (mg/day) (mg/day) (mg/day) 1 2631 2015 — 5 2506 1935 — 8 3074 2660 — 12 3663 4516 11058 15 2624 2256 8944 19 2570 2200 8400 22 2704 2316 9296 26 3588 3588 9935 29 2857 3688 9171 33 2456 3453 8487 36 3061 3571 8426 40 2085 2857 7061 43 2328 3886 9802 47 2328 3183 11402 50 2754 3267 13268 54 2367 2367 10836 61 3614 3673 9665 71 1548 1138 7411 78 2480 1837 16342 85 2085 1859 15030 92 2508 2242 15405 99 2936 2746 9467 106 3261 3008 14662 113 1486 1290 8556 120 2125 1802 12860 141 1723 1505 10078 155 2031 1994 13207 190 1509 1401 11333 197 1956 1996 14248 204 2195 1923 15795 218 1894 1523 15250 225 3038 2870 15414 232 2354 2070 11120 260 2197 2056 10889 267 1776 1932 10306 274 1742 1695 10890 281 2209 2115 14152 289 2176 2406 14254 296 1253 1274 9040 311 1889 1926 9038 317 1501 1540 8273 324 1552 1493 12078 331 1673 1572 12553 102 DAY LEACHATE LEACHATE INFLUENT + INFLUENT BOD5 CARBON FEED EFFLUENT BOD, (mg/L) (mg/L) (mg/L) 71 24 85 12 1091 13 92 6 835 8 99 13 295 22 106 9 553 1113  956 4 120 9 868 5 DAY BODc REMOVAL EFFICIENCY (%) 85 99 92 8 99 93 106 8 113 99.6 120 99.4 103 DAY PERCENT COD REMOVED PERCENT COD REMOVED PERCENT BODc REMOVED ACROSS ANOXIC ACROSS AEROBIC ACROSS ANOXIC • (%) (%) (%) 8 43.0 57.0 -12 45.0 55.0 -15 84.0 16.0 -19 77.0 23.0 -22 77.0 23.0 -29 77.0 23.0 -33 55.0 45.0 36 83.0 17.0 -40 73.0 27.0 -43 75.0 25.0 -47 93.0 7.0 -50 61.0 39.0 -54 71.0 29.0 — 61 66.0 34.0 — 71 83.0 17.0 -78 93.0 7.0 — 85 94.0 6.0 96 99 86.0 14.0 86 106 88.0 12.0 97 113 96.0 4.0 96 120 92.0 8.0 82 141 95.0 5.0 — 155 98.0 2.0 -190 91.8 8.2 -197 63.0 37.0 -204 99.5 0.5 -218 97.8 2.2 -225 99.5 0.5 -232 95.1 4.9 -260 99.4 0.6 -267 99.4 0.6 -274 98.3 1.7 -281 98.5 1.5 -289 97.6 2.4 -296 81.8 18.9 -311 99.5 0.5 -324 98.5 1.5 -331 85.7 14.3 -104 DAY COD (USED)/NOm (REDUCED) COD REMOVAL MLVSS/MLSS ACROSS ANOXIC tmg/d/mg/d) EFFICIENCY (AEROBIC) (%) 1 - 23.0 0.44 5 - 23.0 -8 - 13.0 0.43 12 - 59.0 -15 - 75.0 0.49 19 3.6 74.0 -22 4.7 75.0 0.55 26 - 64.0 -29 4.8 60.0 0.52 33 2.8 59.0 -36 5.2 58.0 0.52 40 3.8 60.0 -43 3.1 60.0 0.55 47 7.2 72.0 -50 5.1 75.0 0.57 54 4.7 78.0 — 61 3.4 62.0 -71 10.4 85.0 0.63 78 10.8 89.0 0.72 85 11.3 88.0 0.72 92 - 85.0 0.73 99 5.2 71.0 -106 6.9 80.0 -113 12.2 85.0 0.70 120 10.5 86*0 0.71 141 8.6 85.0 0.71 155 12.6 85.0 -190 8.9 87.6 0.60 197 10.8 86.0 0.62 204 10.8 87.8 0.68 218 18.2 90.0 0.66 225 11.8 81.4 0.67 232 10.4 81.4 0.64 260 7.5 81.1 0.66 267 20.8 81.3 0.64 274 22.7 84.4 0.67 281 33.2 85.1 0.71 289 52.1 83.1 0.74 296 14.6 85.9 0.79 311 25.5 78.7 0.77 317 20.3 81.4 0.77 324 22.3 87.6 0.75 331 42.8 87.5 0.75 105 CARBON FEED COD/ COD REMOVAL ANOXIC MLVSS AEROBIC MLVSS LEACHATE INFLUENT COD EFFICIENCY CONC CONC1  (mg/d/mg/d) (%) (mg/L) (mg/L) 0 23 874 894 0 23 488 537 0 13 - -2.02 59 - -2.41 75 1036 944 2.27 74 - -2.44 75 1748 879 2.21 90 1376 1034 2.46 59 - -1.75 58 1402 1225 2.39 60 - -3.21 60 1612 1154 3.90 72 - -3.82 75 1712 1393 3.56 78 - -1.68 62 - -3.79 85 1179 1742 5.59 89 1798 2430 6.21 88 2256 3136 5.14 85 3027 3948 2.23 71 - -3.50 80 - -4.76 85 2245 2812 5.05 86 2580 3193 4.85 85 1972 2300 5.50 85 - -6.50 38 2740 3326 6.28 86 3558 4168 6.19 88 4069 4498 7.05 90 4516 5169 4.07 81 4614 5277 3.72 81 4267 4535 3.96 81 2672 2782 4.80 81 2441 2681 5.25 84 2664 2797 5.41 85 3147 3395 5.55 83 3389 3987 6.21 86 3880 3836 3.78 79 3005 3328 4.51 81 2806 3123 6.78 88 3208 3449 6.50 88 4616 '4875 DAY PERCENT BODe REMOVED ACROSS AEROBIC ANOXIC REACTOR AEROBIC REACTOR (%) PH pH 1 - 7.40 7.30 5 - 7.50 7.00 8 - 7.50 7.20 12 • - 7.40 7.00 15 - 7.65 7.60 19 - 7.80 8.00 22 - 8.00 7.90 26 - 7.70 7.70 29 - 7.90 7.85 33 - 7.90 7.80 36 - 7.95 8.00 40 - 7.95 8.10 47 - 8.10 8.20 50 - 7.90 8.10 54 - 8.00 8.10 61 - 7.80 8.05 71 - 7.95 8.15 82 - 8.00 8.10 85 4 8.10 8.30 89 — 7.90 8.20 92 - 8.00 8.20 99 14 7.90 8.00 106 3 8.00 8.00 113 4 7.84 8.04 120 18 7.90 8.00 107 DAY ORP IN THE UNIT NITRIFICATION UNIT DENITRIFICATION ANOXIC REACTOR RATES RATES (Ecal.) (mg NOm PRODUCED/ (mg NOm REDUCED/ ht/gm VSS hr/gm VSS 1 - 7.66 15 - 7.21 22 - 12.36 27.4 29 - 8.13 26.7 36 - 6.39 22.9 43  8.25 36.7 50 - 4.43 29.0 71 - 4.51 17.7 78 - 3.36 29.0 85 - 2.32 20.3 92  1.56 13.9 113 - 3.02 10.6 120 - 2.52 17.0 141 - 3.36 21.190  1.78 15.6 197 - 0.10 204 - 1.44 13.1 218 - 0.70 6.8 225 - 1.25 9.6 232 - 69.5 1.10 8.1 257 -118.5 260 -145.5 2.94 18.3 264 -138.0 267 -124.5 1.49 6.8 271 -123.5 274 -125.5 1.21 6.2 278 -125.5 281 -126.0 0.92 4.7 285 -126.0 289 -144.5 0.68 2.7 293 -172.5 296 -193.0 0.63 4.8 300 -185.0 307 - 87.0 311 -143.5 0.87 3.8 314 -142.5 317 -116.5 0.82 5.6 321 - 78.0 324 -202.5 0.66 6.1 328 -272.0 331 -279.0 0.22 2.0 335 -308.0 108 DAY AMMONIA REMOVAL AMMONIA REMOVAL AMMONIA REMOVAL RATE (ANOXIC) RATE (AEROBIC) RATE (ANOXIC) ( mg NH4 REMOVED/ ( mg NH4 REDUCED/ (gm NH4 REDUCED/ hr/gm VSS) hr/gm VSS) m3/DAY) 1 3.70 9.55 78 8 - 19.10 -15 2.33 10.90 58 22 2.40 7.48 99 29 2.67 8.45 88 36 5.32 6.60 179 43 2.65 7.00 103 50 3.48 4.90 143 71 2.14 2.10 61 78 1.61 2.70 69 85 1.37 1.80 74 92 1.06 1.92 77 113 2.47 1.14 133 120 1.76 1.90 108 141 1.50 2.40 71 190 0.51 1.23 34 197 0.95 0.71 81 204 1.30 1.15 127 218 0.51 0.80 56 225 0.54 1.40 60 232 0.89 1.00 91 260 1.12 2.32 72 267 0.95 1.49 56 274 0.95 1.13 60 281 1.41 1.28 106 289 1.04 0.96 85 296 0.19 0.74 18 311 1.35 1.05 97 317 0.94 0.82 64 324 0.69 0.85 53 331 0.45 0.60 50 109 PERCENT AMMONIA DAY PERCENT NITRIFICATION PERCENT NITRIFICATION REMOVED ACROSS (SYSTEM 1) (SYSTEM 1) ANOXIC REACTOR (%) (DEFN.: A) (%) (DEFN.: B) (%) 1 73.8 73.8 15.9 5 - - 16.6 8 83.0 83.0 21.8 12 54.0 54.0 -15 65.0 65.0 -19 137.5 137.5 -22 144.4 144.4 23.9 26 108.3 108.3 19.7 29 102.4 102.4 20.4 33 125.0 125.0 32.9 36 104.8 104.8 32.3 40 83.3 83.3 17.0 43 109.3 109.3 22.1 47 102.8 102.8 34.5 50 88.8 88.8 31.6 61 125.7 125.7 -71 115.4 115.4 25.3 78 100.0 100.0 18.1 82 106.8 106.8 21.1 85 93.9 93.9 21.2 89 73.7 73.7 27.3 92 72.7 72.7 17.0 96 86.5 86.5 27.8 99 101.1 101.1 23.4 103 94.2 94.2 -106 71.6 71.6 -110 113.5 113.5 30.4 113 152.6 152.6 -. 117 117.4 117.4 30.4 120 112.9 112.9 27.1 124 120.0 120.0 26.1 141 140.7 140.7 21.2 145 132.1 132.1 29.9 152 88.5 88.5 26.6 155 106.7 106.7 23.8 190 144.4 144.4 14.7 194 • 78.9 78.9 18.2 197 0 0 37.8 200 104.3 104.3 36.3 204 126.1 126.1 33.9 208 116.0 116.0 37.3 215 114.3 114.3 47.2 218 88.2 88.2 22.0 225 89.7 89.7 14.5 229 100.0 100.0 -232 110.5 110.5 29.7 257 68.4 68.4 21.3 260 126.9 126.9 18.8 264 130.8 130.8 34.8 267 100.0 100.0 22.6 110 (CONTINUED) PERCENT AMMONIA DAY PERCENT NITRIFICATION PERCENT NITRIFICATION REMOVED ACROSS (SYSTEM 1) (SYSTEM 1) ANOXIC REACTOR (%) (DEFN.: A) (%) (DEFN.; B) (%) 271 46.7 46.7 22.7 274 107.1 107.1 28.5 278 81.8 81.8 34.9 281 72.2 72.2 33.8 285 94.1 94.1 28.5 289 70.6 21.1 12.1 293 57.1 13.0 3.2 296 84.6 12.9 2.0 300 - - 22.7 307 78.9 48.4 11.1 311 82.8 53.3 27.3 314 160.0 160.0 -317 100.0 100.0 34.0 321 78.6 78.6 30.6 324 77.7 77.7 27.4 328 90.0 90.0 36.5 331 52.6 52.6 26.6 335 63.6 63.6 34.8 DAY INFLUENT MANGANESE EFFLUENT MANGANESE IRON REMOVAL (UNFILTERED) (FILTERED) EFFICIENCY (mg/L) (mg/L) (%) 54 0.184 0.034 99.0 75 0.176 0.008 -82 0.067 0 98.7 89 0.135 0.007 -96 0.028 0.006 -103 0.024 0.042 -110 0.096 0.004 98.7 117 0.140 0 99.0 124 0.082 0.010 99.3 138 0.132 0.020 80.0 145 0.161 0.009 83.3 155 0.286 0.012 90.9 190 0.475 0.025 98.7 197 1.390 0.190 92.9 200 1.740 0.012 98.9 204 0.750 0.015 98.7 208 3.820 0.040 98.8 215 9.170 0.020 99.1 222 12.300 0.020 97.6 229 9.160 0.039 90.0 257 4.336 1.070 98.7 260 3.702 - — 264 4.004 0.010 99.7 271 4.333 0.049 93.7 278 4.757 0.560 97.1 285 3.453 1.930 95.0 289 2.691 - — 293 3.527 0.920 94.7 296 4.688 — — 300 4.162 0.870 96.3 307 2.287 1.090 94.1 314 2.907 0.670 97.4 317 3.490 - — 321 3.345 0.320 — 324 5.123 - 95.9 328 4.840 0.830 — 331 5.210 — — 335 5.180 0.550 -DAY SLUDGE SLUDGE MANGANESE REMOVAL MANGANESE (ANOXIC) MANGANESE (AEROBIC) EFFICIENCY (mg/kg) (mg/kg) (%) 54 - 500 81.5 75 290 330 95.5 82 325 - 100.0 89 289 320 94.8 96 332 357 78.6 103 332 344 -110 308 334 95.8 117 273 311 100.0 124 234 273 87.8 138 330 363 84.8 145 392 386 94.4 155 - - 95.8 190 687 743 94.7 197 778 '862 86.3 200 821 866 88.8 204 808 873 98.0 208 1259 1573 99.0 215 2181 2086 99.8 222 4432 4593 99.8 229 6629 6905 99.6 257 7415 7214 75.3 264 7620 7097 99.8 271 7133 7186 88.7 278 5974 6321 88.2 285 4906 5077 44.1 293 3903 3503 73.9 300 4605 4698 79.1 307 4896 4813 52.3 314 5068 5310 76.9 321 4877 4663 90.4 328 4446 4442 82.9 335 5104 5324 89.4 11-3 DAY INFLUENT EFFLUENT INFLUENT IRON-FILTERED IRON-FILTERED IRON-UNFILTERED (mg/L) (mg/L) (mg/L) 54 - 36.2 75 - - 30.0 82 - - 22.0 89 1.9 0.9 -96 2.0 0.12 — 103 2.0 0.8 — 110 - - 18.0 117 - - 24.0 124 - - 19.8 138 - - . 13.0 145 - - 10.5 155 - - 28.5 190 - - 17.0 197 - - 20.5 200 - - 32.9 204 - - 23.0 208 - - 14.5 215 - - 18.0 222 - - 9.8 229 - - 4.3 257 - - 17.4 264 - - 11.1 271 - - 5.1 278 - - 5.8 285 - — 3.5 293 - - 8.8 300 - - 7.6 307 - - 3.7 314 - - 6.8 321 - - 4.6 DAY EFFLUENT SLUDGE SLUDGE IRON-UNFILTERED IRON (ANOXIC) IRON (AEROBIC) (mg/L) (mg/kg) (mg/kg) 54 0.365 - 113000 75 - 98000 98000 82 0.290 84746 — 89 0.900 83476 88652 96 0.120 86996 88652 103 0.800 98039 87173 110 0.240 100781 99870 117 0.230 83185 97851 124 0.140 74836 81815 138 2.600 64825 64725 145 1.750 71191 70349 155 2.600 71179 73684 190 0.225 42204 6Q240 197 1.450 36039 39435 200 0.360 51875 34295 204 0.300 43580 47001 208 0.180 35405 38425 215 0.170 41700 37012 222 0.240 31690 29837 229 0.430 27766 28838 257 0.220 24629 25723 264 0.030 34482 32843 271 0.320 26721 27834 278 0.170 24890 25420 285 0.175 20420 17894 293 0.470 16729 17932 300 0.280 17959 18750 307 0.220 21996 24067 314 0.180 20273 19502 321 0.190 15537 8608 115 NICKEL REMOVAL DAY INFLUENT NICKEL EFFLUENT NICKEL EFFICIENCY (mg/L) (mg/L) (%) 54 0.038 0.038 0 75 0.028 0.023 17.9 82 0.025 0.025 0 89 0.035 0.032 8.6 96 0.028 0.035 -25.0 103 0.029 0.029 — 110 0.029 0.029 0 117 0.028 0.028 0 124 0.048 0.045 6.3 138 0.056 0.051 8.9 145 0.056 0.052 7.1 155 0.066 0.056 15.2 190 0.030' 0.030 • 0 197 0.027 0.025 7.5 200 0.032 0.024 25.0 204 0.022 0.021 4.6 208 0.025 0.025 0 215 0.028 0.024 14.3 222 0.026 0.022 15.4 229 0.030- 0.023 23.3 257 0.034 0.036- 11.8 264 0.031 0.015 51.6 271 0.034 0.029 14.7 278 0.033 0.029 12.1 285 0.033 0.032 3.0 293 0.037 0.032 13.5 300 0.029 0.028 3.4 307 0.033 0.030. 9.1 314 0.033 0.029 12.1 321 0.039 0.033 15.4 EFFLUENT AMMONIA DAY INFLUENT ZINC EFFLUENT ZINC (SYSTEM 2) (mg/L) (mg/L) (mg/L)  54 0.060 0.030 -75 0.068 0.003 -82 0.045 0.028 -89 0.019 0.030 -96 0.031 0.020 -103 0.018 0.020 -110 0.060 0.025 -117 0.155 0.011 -124 0.039 0.010 -138 0.080 0.025 -145 0.045 0.012 -155 0.025 0.001 -190 0.179 0.019 0 194 - - 0 197 0.124 0.008 1 200 0.089 0.010 0 204 0.051 0.016 0 208 0.038 0.010 0 215 0.140 0.020 0 218 - - 0 222 0.052 0.008 0 225 - - 0 229 0.050 0.026 0 232 - - 0 257 1.356 0.108 0 260 2.012 - 0 264 2.791 0.008 0 267 - - 0 271 2.890 0.270 0 274 - - 0 278 8.990 0.415 0 281 8.990 - 0 285 14.966 2.240 0 289 17.678 - 0 293 16.838 2.390 0 296 10.337 - 0 300 9.238 1.380 1 307 4.545 1.410 0 311 - - 0 314 5.512 0.960 0 317 6.701 - 0 321 8.493 0.570 0 324 14.493 - 0 328 13.820 1.080 0 331 18.190 - 0 335 19.330 1.370 0 117 ZINC REMOVAL DAY SLUDGE ZINC (ANOXIC) SLUDGE ZINC (AEROBIC) EFFICIENCY (mg/kg) (mg/kg) (%) 54 - 160 50.0 75 140 140 95.6 82 90 - 37.8 89 83 86 -57.9 96 103 110 35.5 103 114 116 — 110 115 141 58.3 117 111 140 92.9 124 70 119 74.3 138 141 162 68.8 145 142 143 73.3 155 125 124 100.0 190 229 241 89.4 197 195 202 93.5 200 157 165 88.8 204 157 157 68.8 208 137 149 73.7 215 210 165 85.7 222 124 118 84.6 229 140 142 48.0 257 490 478 92.0 264 1277 1236 99.7 271 2219 2176 90.7 278 3855 3890 95.4 285 6608 6825 85.0 293 10060 9118 85.8 300 11291 11595 85.1 307 11684 11321 69.0 314 11415 12148 82.6 321 11221 10736 93.3 328 11797 11953 92.2 335 12939 13405 92.9 DAY INFLUENT • . ,SS EFFLUENT SS INFLUENT . .VSS (mg/L) (mg/L) (mg/L) 1 54 70 15 8 89 48 28 15 123 15 32 22 82 45 20 29 42 31 10 36 87 25 26 43 52 39 15 50 90 42 23 71 94 27 26 78 96 36 36 85 102 34 28 92 106 44 32 99 117 45 — 113 105 25 25 120 80 52 29 141 59 125 14 190 134 76 23 197 145 38 44 204 89 78 26 218 119 65 25 225 77 138 24 232 68 178 17 260 96 160 27 267 225 155 11 274 25 170 4 281 84 165 18 289 78 351 12 296 27 148 3 311 132 99 13 317 132 60 -324 107 82 18 331 75 82 23 DAY EFFLUENT VSS (mg/L) ANOXIC MLSS (mg/L) AEROBIC MLSS (mg/L) 1 24 1871 2016 8 20 1117 1251 15 13 2019 1912 22 28 2777 1607 29 16 2594 1980 36 19 2678 2367 43 32 2945 2115 50 32 2971 2439 71 18 1819 2770 78 33 2457 3370 85 23 3127 4361 92 39 4149 5424 99 - 3607 4704 113 6 3223 4050 120 44 3558 4435 141 89 2831 3227 190 63 4212 5536 197 14 5401 6744 204 63 5833 6621 218 32 6680 7797 225 99 6815 7894 232 128 6560 7035 260 130 4016 4223 267 96 3755 4174 274 118 3907 4169 281 130 4418 4813 289 278 4594 5406 296 119 4899 4864 311 78 3889 4336 317 22 3645 4082 324 64 4237 4572 331 67 6171 6531 120 DAY ANOXIC MLVSS AEROBIC MLVSS MLVSS/MLSS (mg/L) (mg/L) (ANOXIC) 1 874 894 0.47 8 488 537 0.44 15 1036 944 0.51 22 1748 879 0.63 29 1376 1034 0.53 36 1402 1225 0.52 43 1612 1154 0.55 50 1712 1393 0.58 71 1179 1742 0.65 78 1798 2430 0.73 85 2256 3136 0.72 92 3027 3948 0.73 113 2245 2812 0.70 120 2580 3193 0.73 141 1972 2300 0.70 190 2740 3326 0.65 197 3558 4168 0.66 204 4069 4498 0.70 218 4516 5169 0.68 225 4614 5277 0.68 232 4276 4535 0.65 260 2672 2782 0.67 267 2441 2681 0.65 274 2664 2797 0.68 281 3147 3395 0.71 289 3387 3987 0.74 296 3880 3836 0.79 311 3005 3328 0.77 317 2806 3123 0.77 324 3208 3349 0.76 331 4616 4875 0.75 DAY AMMONIA REMOVAL RATE (AEROBIC) (gm NH4 REMOVED/m^/DAY) 1 205 8 246 15 248 22 158 29 210 36 194 43 194 50 162 71 89 78 156 85 136 92 182 113 77 120 146 141 132 190 98 197 71 204 124 218 99 225 177 232 108 260 155 267 96 274 76 281 104 289 92 296 68 311 84 317 62 324 71 331 70 

Cite

Citation Scheme:

    

Usage Statistics

Country Views Downloads
United States 33 3
China 26 8
Canada 9 0
Sweden 8 0
Germany 8 33
United Kingdom 4 0
Italy 3 0
France 2 0
Japan 2 0
Brazil 1 0
Australia 1 0
India 1 0
Malaysia 1 0
City Views Downloads
Unknown 33 33
Shenzhen 11 8
Hangzhou 8 0
Stockholm 8 0
Beverly Hills 7 0
Ashburn 5 0
Rheems 3 0
Redmond 3 0
Guiyang 2 0
Padua 2 0
Beijing 2 0
Saint Paul 2 0
Regina 2 0

{[{ mDataHeader[type] }]} {[{ month[type] }]} {[{ tData[type] }]}
Download Stats

Share

Embed

Customize your widget with the following options, then copy and paste the code below into the HTML of your page to embed this item in your website.
                        
                            <div id="ubcOpenCollectionsWidgetDisplay">
                            <script id="ubcOpenCollectionsWidget"
                            src="{[{embed.src}]}"
                            data-item="{[{embed.item}]}"
                            data-collection="{[{embed.collection}]}"
                            data-metadata="{[{embed.showMetadata}]}"
                            data-width="{[{embed.width}]}"
                            async >
                            </script>
                            </div>
                        
                    
IIIF logo Our image viewer uses the IIIF 2.0 standard. To load this item in other compatible viewers, use this url:
http://iiif.library.ubc.ca/presentation/dsp.831.1-0062950/manifest

Comment

Related Items