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Degradation of dehydroabietic acid by bacterial isolates grown on CTMP effluents Zhang, Yi 1997

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DEGRADATION OF DEHYDROABIETIC ACID BY BACTERIAL ISOLATES GROWN ON CTMP EFFLUENTS By YI Z H A N G M . S c , Nanjing University, 1989 A thesis submitted in partial fulfillment of the requirements for the degree of Master of Science in the Faculty of Graduate Studies Department of Wood Science We accept this thesis as confirming to there^quired standard o ~~ THE UNIVERSITY OF BRITISH COLUMBIA Y i Zhang© January 1997 In presenting this thesis in partial fulfilment of the requirements for an advanced degree at 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 \Mooo) Su^&a. iQ— The University of British Columbia Vancouver, Canada Date j - ^ my DE-6 (2/88) ABSTRACT Resin acids, which are a major component of softwood extractives, are released during the pulping of wood and are known to contribute to much of the toxicity of pulp mill-effluents. Usually biological treatment can efficiently remove resin acids. However, resin acid breakthroughs occasionally occur due to operational problems or seasonal variations in the wood chip furnish. Recently Bicho et al. (1994) isolated five dehydroabietic acid degrading bacterial strains from a B K M E treatment system. These isolates were able to utilize high concentrations of abietanes. Most of the past and recent work on the microbes involved in resin acid degradation have utilized defined media and there has been no attempt to reintroduce these strains to mill effluents to see if they still functioned in this environment. This was the main focus of this study. Two strains (BKME5 and B K M E 9) described by Bicho et al. were examined for their ability to grow and degrade D H A present in CTMP effluent. Initially, although both strains could grow on CTMP effluent, no D H A degradation was observed. COD measurements indicated that both strains used other organic material present in the CTMP effluent. When nutrients (ammonium, phosphate, mineral salts and vitamins) were added, both growth and D H A degradation increased significantly. A comparison of D H A degradation with COD removal indicated that these strains preferentially utilized other organic material present in the effluent before degrading DHA. The stimulated growth resulting from the use of other organic materials did not increase the rate of DHA degradation. This initial work showed that the CTMP effluent was nutrient deficient in terms of DHA degradation. It was found that strain B K M E 5 required additional nitrogen before it could degrade DHA. However, strain B K M E 9 required N , P, vitamins and minerals before efficient DHA degradation occurred. Magnesium was shown to be essential for effective D H A degradation by B K M E 9. More than 12.5 mg/1 nitrogen had to be added to the CTMP effluent before either strain ii could effectively degrade DHA. To ensure complete degradation of DHA, at least 25 mg/1 of nitrogen had to be added for strain B K M E 5 and 1000 mg/1 for B K M E 9. A solid COD based method was developed as an alternative procedure for quantifying microbial biomass. This method was based on the same principals as are currently used for determining VSS. The advantages of the solid COD method over the VSS protocol include rapid analysis time, greater sensitivity and the use of smaller sample volumes. A good correlation between VSS and solid COD was found for a variety of sources of microbial biomass, including cultures at different stages of growth supported by different effluents. About 1.2-1.6 mg COD were measured per mg of VSS. Although the solid COD/VSS ratio was stable over a range of conditions, it was not constant when a large amount of non-biomass solids are present in samples such as a typical CTMP effluent. Both filtration and gamma irradiation proved to be effective sterilization methods that prevented growth of any contaminating microorganisms present in the CTMP effluent. However, both methods affected the COD concentration and the resin acid profile and concentration. Although filtration removed about 1/3 of the COD (particulates) and 1/2 of the total resin acids within the effluent, the gamma irradiation caused little change in effluent COD of the total resin acid concentration. However, the amount of the individual resin acids had changed significantly after irradiation. Neither the inorganic components nor the nutrient concentrations were changed by either sterilization method. This initial study indicated that these isolates required additional nutrients such as nitrogen and minerals before they could effectively remove D H A from CTMP effluent. Although they may not be fully representative of the microbial community responsible for most of the resin acid removal in mill waste water treatments, this initial work indicates the possible variability and selectivity of substrate utilization by resin acid degrading bacteria. iii TABLE OF CONTENTS ABSTRACT ii T A B L E OF CONTENTS iv LIST OF T A B L E S vii LIST OF FIGURES viii LIST OF ABBREVIATIONS ix A C K N O W L E D G E M E N T S x CHAPTER 1 INTRODUCTION 1 1.1 General background 1 1.2 Pulp and paper mill effluents 3 1.2.1 Effluent generation; 3 1.1.2 Effluent characteristics 5 1.2.3 Toxicants in pulp and paper mill effluents 7 1.3 Resin acids 9 1.3.1 Resin acids generation during pulping processes 9 1.3.2 Resin acid toxicity 12 1.3.2.1 Acute lethal 12 1.3.2.2 Chronic sublethal toxicity.. 13 1.3.2.3 Bioaccumulation 14 1.3.2.4 Genetoxicity 15 1.4 Biological treatment of pulp and paper mill effluents and methods of quantification of microbial populations in biotreatment systems 15 1.4.1 mixed liquor volatile suspended solids (MLVSS) 16 1.4.2 DNA determination 18 1.4.3 ATP measurement 19 1.4.4 FDA hydrolysis 21 1.5 Microbiology of resin acid biodegradation 22 1.5.1 Resin acid biodegradation in situ 22 1.5.2 The resin acid biodegradation pathways 27 1.5.2.1 Degradation by fungi 27 1.5.2.2 degradation by bacteria -. 29 1.5.3 Physiology of the resin acid degrading organisms... 34 1.6 Objectives of the thesis 40 CHAPTER 2. DEVELOPMENT OF METHODS FOR THE QUANTIFICATION OF MICROBIAL POPULATIONS IN BIOLOGICAL TREATMENT SYSTEMS 42 2.1. A Solid COD method for detennining biomass in biotreatment systems 42 2.1.1 Background 42 2.1.2 Materials and methods 45 2.1.2.1 Solid COD ' 45 2.1.2.2 Volatile suspended solids 45 2.1.2.3 Effluent and activated sludge 45 2.1.2.4 Composition of COD and VSS values 46 2.1.3 Results and discussion 47 2.2. COMPARISON OF METHODS FOR QUANTIFYING RESIN ACID DEGRADING BACTERIA IN PURE CULTURES BY PROTEIN ASSAYS, PLATE COUNTS, TOTAL CELL COUNTS AND THE SOLID COD METHOD 54 2.2.1 Background 54 2.2.2.Materials and methods 55 2.2.2.1 Protein assays in the presence of resin acids and CTMP effluent. 55 2.2.2.2 Comparison between the plate count, total cell count and solid COD measurement of microbial growth 55 2.2.3 Results and discussion 55 2.2.3.1 Protein assays in the presence of resin acids and CTMP effluent. 55 2.2.3.2 Comparison between the plate count, total cell count and solid COD measurement of microbial growth 56 CHAPTER 3. GROWTH OF RESIN ACID DEGRADING BACTERIAL STRAINS BKME 5 AND BKME 9 IN THE PRESENCE OF CTMP EFFLUENT 60 3.1 Background 60 3.2 Materials and methods 62 3.2.1 Bacterial strains 62 3.2.2 Chemicals 62 3.2.3 Preparation of sterilized CTMP effluent 64 3.2.4 Growth on CTMP effluent alone 64 3.2.5 Growth on nutrient supplemented CTMP effluent 65 3.2.6 Total cell counts 65 3.2.7 Resin acid analysis 65 3.2.8 COD measurement 66 3.2.9 Nutrient analysis 66 3.2.10 Sugar analysis 66 3.3 Results 67 3.3.1 Preparation of sterilized CTMP effluent 67 3.3.2 Growth of BKME 5 and 9 and degradation of DHA present in CTMP effluent 68 3.3.3 Growth experiments on nutrient supplemented CTMP effluents 72 3.3.4 Utilization of other organic substrates 74 3.4 Discussion 79 CHAPTER 4. NUTRITIONAL CHARACTERIZATION OF THE RESIN ACID DEGRADING BACTERIA 82 4.1 Background 82 4.2 Materials and methods 82 4.2.1 To assess the impact of nutrients on DHA degradation 82 4.2.2 Mineral analysis of effluents 83 4.2.3 To assess the need for the minerals present in YNB for effective DHA degradation by BKME 9.... 83 4.2.4 Identification of the essential minerals required for DHA degradation by BKME 9 84 4.2.5 Study of the minimum nitrogen requirement for DHA degradation 84 4.3 Results 85 4.3.1 Impact of nutrients on DHA degradation 85 4.3.2 Influence of mineral on growth and DHA degradation 85 4.3.3 The minimum nitrogen requirement for effective DHA degradation 91 4.4 Discussion 91 CHAPTER 5. GENERAL CONCLUSIONS 96 REFERENCES 101 vi LIST OF TABLES Table 1. Major sources of effluent pollution in a pulp and paper mill complex. Table 2. Characteristics of mechanical pulping effluents. Table 3. Characteristics of chemical pulping (NSSG pulping) and bleaching liquors. Table 4. Distributions of resin acids in Pinus Radiata and source effluents. Table 5. Summary of resin acid degrading fugal and bacterial isolates Table 6. Physiological characterization of the five strains on Biolog G N plates Table 7. Substrate use by D H A degraders Dha-33 and Dha-35 Table 8. The composition of Bacto Yeast Nitrogen Base medium without amino acids and ammonium. Table 9. Nitrogen and phosphorus concentration in original, filtered and gamma-irradiated CTMP effluents. Table 10. Resin acid concentration in original, filtered and gamma-irradiated CTMP effluent. Table 11 Sugar concentration in original, filtered and gamma-irradiated CTMP effluent. Table 12. Sugar concentration prior and after 4 day incubation of B K M E 5 and 9 on gamma-irradiated CTMP effluent and filtered CTMP effluent. Table 13. Mineral concentrations in original, filtered and gamma-irradiated CTMP effluents. LIST OF FIGURES Fig. 1. Representative bleached Kraft mill loads. Fig.2. Structure of resin acids found in pulp and paper mill effluents. Fig. 3 Structures of the major resin acids and neutral transformation products. Fig.4. The biotransformation of D H A by M isabellina. Fig.5. The complete degradation of DHA by F. resinvorum. Fig.6. An alternative path for D H A metabolism by F. resinovorum. Fig. 7. The metabolism of D H A by A. eutrophus and a Pseudomonas sp.. Fig. 8. Oxidative attack of the aromatic moiety of DHA. Fig. 9. Comparison of the solid COD and VSS values for (A) sewage sludge, (B) B K M E and (C) CTMP effluents at various dilutions. Fig. 10 Comparison of growth trials monitored on (A) sewage sludge, (B) B K M E and (C) CTMP effluents using the solid COD and VSS methods. Fig. 11 Ratio of the solid COD vs VSS determined over the course of the growth trials on sewage, B K M E and CTMP effluent. Fig. 12 Correlation between solid COD and VSS measurements compiled over multiple trials with B K M E . Fig. 13 Influence of CTMP effluent and resin acids on the Fluor aldehyde protein assay (A), the regular BIO-RAD assay (B) and the Lowry protein assay (C). Fig. 14 Correlation between plate counts, solid COD and total cell counts for B K M E 5 (A) and B K M E 9 (B). Fig. 15. Growth (at 30 °C) and the measurement of D H A degradation and COD utilization by B K M E 5 (15 A) and B K M E 9 (15B) on filtered CTMP effluent Fig. 16. Growth of B K M E 5 (A) and B K M E 9 (C) at 30 °C and the measurement of DHA degradation by B K M E 5 (B) and B K M E 9 (D) on filtered CTMP effluent plus DHA and nutrients, filtered CTMP effluent plus nutrients, filtered CTMP effluent alone, and DHA defined medium. Fig. 17. Growth of B K M E 5 (A) and B K M E 9 (C) at 23 °C and the measurement of DHA degradation by B K M E 5 (B) and B K M E 9 (D) on gamma-irradiated CTMP effluent plus D H A and nutrients, filtered CTMP effluent plus D H A and nutrients, and DHA defined medium. Fig. 18. COD utilization by B K M E 5 (A) and B K M E 9 (B) at 30 °C on filtered CTMP effluent plus D H A and nutrients, filtered CTMP effluent plus nutrients, and filtered CTMP effluent alone. Fig. 19. COD utilization at 23 °C by B K M E 5 and 9 on gamma-irradiated CTMP effluent plus DHA and nutrients, and by B K M E 5 and 9 on filtered CTMP effluent plus D H A and nutrients. Fig. 20. Growth and D H A degradation by B K M E 5 (A) and B K M E 9 (B) on D H A spiked filter sterilized CTMP effluent supplemented with different nutrient combinations. The figure shows the cell yield and D H A removal efficiency after 96 h cultivation. Fig. 21. Growth and D H A degradation by B K M E 9 strain on media with different nutrient combinations. Fig. 22. Growth and D H A degradation by B K M E 9 on media containing different minerals. Fig. 23. D H A degradation by B K M E 5 (A) and B K M E 9 (B) and their growth (C &D) on filter sterilized CTMP effluent with increasing concentrations of NH4CI. LIST OF ABBREVIATIONS A O X adsorbable organic halogens ASB aerated stabilization basin ATP adenosine triphosphate BCF bioconcentration factor B K M E bleached Kraft mill effluent BOD biochemical oxygen demand BSA bovine serum albumin CMP chemi-mechanical pulping COD chemical oxygen demand CTMP chemi-thermo-mechanical pulping DAPI 4' -6-diamidino-2-phenylindole D H A dehydroabietic acid D N A deoxyribonucleic acid DO dissolved oxygen FDA fluorescein diacetate GC gas chromatography HPLC high pressure liquid chromatography HRT hydraulic retention time ICP inductively coupled plasma LC50 lethal concentration (50% mortality) L d 5 0 lethal dose (50% mortality) M L V S S mixed liquor volatile suspended solids MLSS mixed liquor suspended solids MS mass spectroscopy N A D nicotinamide-adenine dinucleotide NADP nicotinamide-adenine dinucleotide phosphate NSSC neutral sulfite semi-chemical OUR oxygen utilization rate rRNA ribosomeRNA SRT solid retention time SS suspended solids TCA tricarboxylic acid cycle TCDD tetrachlorodibenzo-p-dioxin TCDF tetrachlorodibenzofuran TMP thermo-mechanical pulping TSS total suspended solids UDP-GT uridine-5'-diphospho glucoronosyl transferase VSS volatile suspended solids Y N B yeast nitrogen base ACKNOWLEDGEMENTS I would like to thank my supervisor, Dr. Jack Saddler, for his inspiration and support during these studies. It has truly been an enjoyable experience. My appreciation extends to Dr. Paul Bicho, Dr. Steven Liss and Dr. Colette Breuil for their valuable comments and discussion throughout the study and the preparation of this thesis. I also like to thank Dr. Kai L i , Mr. Kevin Stark and other colleagues for their help and pleasant company. Appreciation is also expressed to all my committee members for their guidance and support. I thank my husband Ron, my son Jason and my parents for always being there for me. CHAPTER 1. INTRODUCTION 1.1 General background Since the 1950's, many countries have recognized the possible detrimental environmental effects that the direct discharge of pulp and paper mill effluents has on receiving waters. As a result, many jurisdictions have legislated effluent discharge limits (Jones, 1991). Of the various compounds that have been identified as toxicants in the various pulp and paper process streams, resin acids are of considerable concern because of their potential toxicity to fish (Rogers et al., 1975; Rogers, 1973). These diterpenoid carboxylic acids, which occur mainly as extractives in softwoods, are released into process wastewaters during pulp and paper manufacturing when coniferous woods are used as the feedstock. Resin acid concentrations in many untreated pulp mill effluents are several orders of magnitude higher than their 96 h LC 5o to fish (0.5 - 2.0 ppm) (Taylor et al., 1988) and they are thought to contribute to as much as 70% of mechanical pulp effluent toxicity (Leach and Thakore, 1976). During the last thirty years, many pulp and paper mills have installed biological treatment facilities to remove readily biodegradable organic pollutants, suspended solids and toxicity. Although most of these treatment systems can successfully remove resin acids (Zender et al., 1994), breakthroughs occasionally happen, primarily as a result of operation problems such as black liquor spills (Bicho et al., 1994). In processes where the concentrations of wood extractives are higher (e.g. chemithermomechanical (CTMP)/thermomechanical pulping), biological treatment systems can sometimes exhibit variations in resin acid removal which can subsequently lead to the discharge of toxic effluents. Previously it was reported that resin degrading organisms had fairly limited substrate ranges and could only tolerate limited concentrations of resin acids (Hemingway and Greaves, 1973). It was postulated that they may constitute only a small portion of the microbial populations 1 in biological treatment systems and therefore not to be able to withstand or degrade high concentrations of resin acids in the effluents. This was considered to be one of the possible reasons of toxicity failures resulting from the load shock that is encountered in some wastewater treatment systems (Bicho et al., 1994). Previous research has largely focused on microorganisms which were isolated from soils and natural water systems and biodegradation of DHA, the resin acid of highest concentration in untreated pulp effluents. Since none of these microbial isolates were obtained from pulp effluent treatment systems, they are probably not representative of the types of microbial populations found in these environments. Recently Bicho et al. (1994), Mohn (1995), Wilson et al. (1996), Morgan and Wyndham (1996) have directed their research towards the study of microorganisms from pulp effluent treatment systems. Bicho et al. (1994) isolated five DHA-degrading bacterial strains from bleached Kraft mill effluent treatment systems, and Morgan and Wyndham (1996) isolated thirteen D H A and abietic acid degrading bacteria from the same type of mill effluent. Mohn (1995) and Wilson et al. (1996) isolated DHA and isopimaric acid (IpA) degrading bacterial strains, respectively, from a sequencing batch reactor designed to treat a high-strength process stream from a paper mill. These authors characterized the resin acid biodegradation by these isolates and also carried out some initial physiological and phylogenetic characterization of these strains. Although this work has provided the first in situ study of resin acid biodegradation, the ability of the isolated DHA-degrading strains to grow and degrade resin acids on mill effluent has not yet been fully characterized. No work has been done to investigate whether these strains will grow on alternative pulp mill effluents, the possible impact of mill effluents on resin acid degradation, and the nutritional requirements for each of the microbial isolates to carry out effective resin acid 2 biodegradation. The objective of this thesis was to better characterize the in situ biodegradation of resin acids using the previously isolated microorganisms (Bicho et al., 1994) known to metabolize resin acids. It was hoped that the results of this work would help to better manage resin acid degradation in biological wastewater treatment systems. 1.2 Pulp and paper mill effluents Pulp and paper production from wood (~106 metric tons daily worldwide) not only requires large amounts of fresh water, it also results in the discharge of a considerable volume of effluents (-200 m 3 metric ton of pulp) (Leuenberger et al., 1985). Since the produced pulp corresponds to about 40-45% (chemical pulping) to 65-70% (mechanical pulping) of the original weight of the wood, much of the remaining 60% - 30% of the initial furnish can be found in the wastewater effluent. 1.2.1 Effluent generation A considerable amount of waste is generated during pulp and paper production especially when the steps of wood debarking and chipping, pulping, recovery, bleaching and paper manufacturing are considered. No matter whether the wood is mechanically grounded into pulp or the logs are reduced to a pulp by being chipped and then cooked in a chemical solution, the resulting mass of wood fiber is always washed to remove the chemicals and contaminants and then screened to separate out undigested wood knots and other unwanted materials. These steps produce substantial amounts of solids, liquid, and gaseous wastes (Industrial Waste and Hazardous Contaminants Branch, Ministry of Environment, Lands and Parks, BC, Canada, 1994). For economic reasons, most Kraft pulp mills use evaporation and recovery technologies to permit the reuse of most of the process chemicals. However, at this time most of these technologies are unable to recycle all wastes, with some portion usually released into the effluent. The bleaching step is another important effluent source. Chlorine 3 and chlorine dioxide have historically been used for bleaching pulp because they are not only capable of achieving a high brightness level, they also maintain pulp strength. However, the use of chlorine for bleaching has been shown to produce hazardous chemical by-products, such as dioxins and furans, and they are known to be extremely toxic and persistent if they enter the environment (Industrial Waste and Hazardous Contaminants Branch, Ministry of Environment, Lands and Parks, BC, Canada, 1994). G.A. Smook (1992) has summarized the major sources of effluent pollution in a pulp and paper mill complex as shown in table 1. Table 1. major sources of effluent pollution in a pulp and paper mill complex. (G.A. Smook, 1992) major sources of effluent in pulp and paper industry effluent characteristics water used in wood handling /barking and chip washing digester and evaporator condensates white waters from screening, cleaning and thickening bleach plant washer filtrates paper machine white water fiber and liquor spills from all sections suspended solids, BOD and colour BOD, reduced sulfur compounds unrecovered lignin, BOD BOD, colour, chlorinated compounds large volume shock loading on treatment facilities The effluent generated by a representative bleached Kraft mill, in terms of volume of effluent, TSS and BOD produced per ton products has been described previously (Fig. 1) (Springer, 1992). It is apparent that the largest volume of discharge comes from the paper mill, but almost as much is produced by the bleaching plant. The largest BOD loads are produced by the bleached plant, with the pulp mill second and the paper mill third. The paper mill produces more than twice as much TSS as any other manufacturing area. 4 Various pulping processes are used in the pulp and paper industry, including Kraft, sulfite, soda, NSSC, CMP, TMP, CTMP, and ground-wood processes. A generalization that seems to hold true is that BOD generation is inversely proportional to the pulp yield (Springer, 1992). Thus, the mechanical pulping processes (e.g. TMP and groundwood processes) are generally less polluting than the chemical processes if recovery systems are not used. The poorest chemical pulping process, from an environmental point of view, is the calcium-based sulfite process, in which the yield is low and all of the spent liquor is discharged. Very little of the TSS appears in the pulping effluent, because all the pulping processes can efficiently remove suspended solids. Wood Recovery EFFLUENT TSS - 13.5 kg/tonne (27 lb/ton) BOD - 5 kg/tonne (10 lb/ton) 17 m'/tonne (5 MG/lon) i Wood Yard Chips _!_ Pulping Bleaching EFFLUENT TSS - 3.75 kg/tonne (7.5 lb/ton) BOD - 1 kg/tonne (2 lb/ton) 6.8 m'/tonne (2 MG/ton) TSS - 6 kg/tonne (12 lb/ton) BOD -11.5 kg/lonne (23 lb/ton) 20.4 m'/tonne (6 MG/lon) TSS - 6 kg/tonne (12 lb/ton) BOD - 15.5 kg/tonne (31 lb/ton) 47.6 m'/tonne (14 MG/ton) Paper Manufacturing TSS - 30.8 kg/tonne (61.6 lb/ton) BOD - 10.8 kg/tonne (21.5 lb/ton) 51 m'/tonne (15 MG/ton) Paper Fig. 1. Representative bleached Kraft mill loads (Springer, 1992) 1.2.2 Effluent characteristics Pulp and paper mills, regardless of the types of the technology used, all produce BOD, COD, TSS, toxicity, and colour in the effluent. Mechanical and chemical pulping of wood or other cellulosic 5 materials results in the solubilization of organic materials, such as lignin and various wood extractives, including carbohydrates, acetic and other organic acids, methanol and other low molecular weight alcohols, and a small amount of inorganic ash (Lee, 1992). Al l of these substances will contribute to the BOD, COD and TSS content of the receiving waters. When large amounts of BOD, COD and TSS are discharged, considerable problems can be caused since the biodegradation of the organic matter by microorganisms leads to the exhaustion of dissolved oxygen and the reduction of water transparency, thus impeding the growth of fish and algae. The colour of the pulp and paper effluent is derived from the highly coloured lignin and lignin derivatives dissolved from wood (Springer, 1992). Coloured effluents can retard sunlight transmission and interfere with photosynthesis, thereby reducing the productivity of the aquatic community. Coloured effluents can also reduce the visual appeal and recreational value of the receiving waters. Typical chemical characteristics for wastewater from mechanical, and chemical pulping and bleaching effluents are shown in Tables 2 & 3. Table 2. Characteristics of mechanical pulping effluents (Springer, 1992) Parameter* TMP CTMP Total COD 5600-7200 6000-9000 Total B O D 5 2800 3000-4000 BOD/COD 0.39-0.50 0.44-0.50 Carbohydrates 1230-2700 1000 Methanol 25 — Acetic acid 235 1500 TSS 383-810 500 Total nitrogen (as N) 12 — Total phosphorus (as P) 2.3 — Total inorganic sulfur (as S) 72-167 — * Units of all parameters are mg/1. 6 Table 3. Characteristics of chemical pulping condensates (Springer, 1992) Acid s u l f i t e condensate Kraft foul condens ate Ref. 19 Parameter* Ref. 15 Ref. 16 Ref. 17 Ref. 18 Total COO 4,000-8.000 9.80027.100 16.000 1.202 10.000-13.000 Total 800 s 2.000-4.000 3.700-5.110 10.700 568 5.500-8.500 B00 s/C00 0.5 0.19-0.38 0.67 0.47 0.55-0.65 Total v o l a t i l e acids (as HAC)*' 3.650 -- 16 5.4 30-300 Methanol 250 -- -- 421 7.500-8.500 Ethanol 5.8 " 2-Proponol 18.2 --Furfural 250 --Acetone 5.1 TSS -- -- 0 16 0 Total nitrogen (as N) 306 350-600 Total phosphorous (as P) .— -- 1.0 -- 0.02-1.55 Total inorganic sulfur (as S) 800-850 840-1.270 91 5.9 120-375 pH 2.5 2.8-5.9 10.2 8.0 9.5-10.5 A l k a l i n i t y (as CAC03) -- 1.060 31 2.130-2.660 Temperature ( ^ ) 25-50 " 55-60 •Units of a l l parameters are mg/1. except pH and temperature. ••Includes acet ic and formic ac id . 1.2.3 Toxicants in pulp and paper mill effluents Pulp effluent toxicity is caused by both natural wood components (resin and fatty acids) and the derivatives of the chemicals added during the pulping and bleaching processes (AOX, dioxin and furans). A O X (Adsorbable Organic Halogens) is one method of measuring of the total quantity of chlorinated organic compounds. Pulp and paper effluents probably contain over 300 different individual chlorinated organic compounds (Springer, 1992), such as chlorinated lignosulphonic acids, chlorinated resin acids, chlorinated phenols and chlorinated hydrocarbons. These compounds are known to be.toxic to aquatic organisms (Leach, 1980) and exhibit a strong mutagenic effect as indicated by the results of Ames tests (BJ0rseth et al, 1979). Furthermore, they are resistant to biodegradation (Salkinoja-Salonen and Sundman, 1980) and can accumulate through the aquatic food chain (Landner et al., 1977; Renberg et al, 1980). Dioxin and furan are actually representative of two families of chlorinated organic compounds that can be formed as by-products of pulp bleaching. They both can persist in the environment for many years and some forms are extremely toxic (Industrial Waste and Hazardous Contaminants Branch, Ministry of Environment, Lands and Parks, BC, Canada, 1994). For example, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), a member of the dioxin family, and the furan compound 2,3,7,8-tetraclbrodibenzofuran (TCDF) are extremely hazardous with their LD 5o varying between 0.6 micro g/kg for guinea pig and 3000 micro g/kg for hamsters (Springer, 1992). TCDD is also known to be extremely toxic to fish. As a result, in the last ten years the province of British Columbia and the federal government have introduced limits for pulp and paper mill effluents parameters such as AOX, TCDD and TCDF. (Industrial Waste and Hazardous Contaminants Branch, Ministry of Environment, Lands and Parks, BC, the Canada, 1994) This encouraged the pulp and paper industry to reduce the emissions of AOX, dioxin and furan by modifying pulping processes and substituting chlorine with chlorine dioxide as the main bleaching agent. As a result, there was a substantial decline in the discharge of AOX, TCDD and TCDF between 1988 and 1993. In 1993, the average total daily TCDD discharge from all bleached Kraft mills across the province is now approximately 7.3 mg, or roughly 85% lower than in 1990 (Industrial Waste and Hazardous Contaminants Branch, Ministry of Environment, Lands and Parks, BC, Canada, 1994). With the reduction in the amount of A O X , dioxins and furans released into the pulp and paper mill effluents, chlorinated organic compounds have become less of an issue as most mills can now 8 operate within regulated limits. Instead, the hazardous impact of natural wood components, especially resin acids, on the environment continues to attract more attention. 1.3 Resin acids 1.3.1 Resin acids generation during pulping processes Resin acids are carboxylic acids, belonging to the terpenoids, which are derivatives of a group of hydrocarbons known as the terpenes. These cyclic compounds have a characterized skeleton with 2-methylbutane as the repeating unit. Resin acids can be divided into two groups: the abietanes and the pimaranes. The abietanes have an isopropyl side chain at position 13 while pimaranes have methyl and vinyl substituents at this position. Resin acids are natural components of many wood species. They exist in the resin canals of certain conifers and in the ray parenchyma cells of both hardwoods and softwoods (Mutton, 1962). Resin acids constitute the major non-volatile components of "resin" — a complex material that is insoluble in water and extractable with neutral organic solvents. Trees produce this material primarily as protection from wood-boring insects and associated pathogenic microorganisms. As a result, the highest resin acid content is usually found in the bark (Shrimpton, 1973). Resin acids are almost completely absent in hardwoods but generally rich in softwoods. They reach their greatest development in pines (1.5% by dry weight), and generally constitute about 0.1% of the dry weight of other softwood species (Swan, 1973). When softwoods are used as the furnish for pulp and paper manufacturing, large amounts of resin acids are released during the pulping process. Resin acids are predominately released during the debarking and pulping stages (McLeay and Assoc., 1987). In the Kraft process, the pulping liquor which contains a hot mixture of caustic soda and sodium sulfide solubilizes resin acids into their sodium and calcium salts. During sulfite pulping, much of the resin acids are retained on the pulp in 9 the conventional acidic process, leading to lower resin acid concentrations in the effluent (Taylor et al, 1988). Mechanical pulping also releases resin acids during the grinding, steaming, or chemical softening of the wood chips. As mentioned earlier, resin acid concentrations can be very high in mechanical pulping effluents because little water is used, relative to Kraft and sulfite operations. However, resin acid loading per kilogram of wood processed is probably lower than that of Kraft or sulfite mills as some of the resins are not solubilized without chemical treatment. Generally, TMP, CMP and CTMP mills release greater amounts of resin acids than strictly mechanical mills. The application of heat and chemicals creates more severe extraction conditions and thus, solubilizes more resin acids. There are eight resin acids commonly found in pulp mill effluents. Their chemical structures are shown in Fig. 2. Chlorinated resin acids can also be found in bleached pulping effluents. These include 12-chlorodehydroabietic, 14-chlorodehydroabietic and 12,14-dichlorodehydroabietic acids (also see Fig. 2). The resin acid profile in the effluent varies with the species of feedstock. However, the predominant resin acid in the feedstock may not necessarily predominate in the effluent (Taylor et al, 1988) due to the isomerization and modification during seasoning and pulping. A comparison was made by Zender et al (1994) on the distribution of resin acids in the feedstock and the source effluent in a bleached Kraft pulp mill (Table 4). It was reported that the major resin acids present in living Pinus radiata, the predominant softwood pulped in New Zealand, are levopimaric, palustric and neoabietic acids (Lloyd 1978, Porter 1969). These compounds generally account for about 70% of the resin acids present in the wood. Abietic and dehydroabietic acids both typically comprise about 7% of the resin acids in living Pinus radiata. However, the predominant resin acid in the effluent turned out to be abietic acid. The liability of neoabietic, palustric, and levopimaric acids means that they can readily isomerize to form abietic and dehydroabietic acids (McDonald and 10 Sandaracopimaric feopimaite Pimaric Figure 2. Structure of resin acids found in pulp and paper mill effluents (Bicho et al., 1994). Table 4. Distributions of resin acids in Pinus Radiata and source effluents (Zender et al, 1994) Resin Acid Proportion in Pinus Proportion in Source radiata (%)* Effluent (%)** Levopimaric/Palustric 48.7 22.4 Neoabietic 20.5 11.0 Pimaric 8.1 9.0 Abietic 7.5 33.8 Dehydroabietic 7.2 11.1 Isopimaric 5.4 6.0 Sandaracopimaric 1.0 1.9 Unidentified 1.6 Transformed resin acids 4.7 Dichlorodehydroabietic - 0.1 Total 100 100 Porter, 1969, Porter 1969), so that in chips subjected to pulping, abietic acid is the predominant acid. Pulping and bleaching processes also modify the resin acids, causing a significant decrease in neoabietic, palustric, and levopimaric acids and an increased proportion of abietic acid. It was reported that 7-oxo-DHA and di-Cl-DHA are the products of oxidation and chlorination reactions, respectively, during the pulping and bleaching processes (Zender et al. 1994). 1.3.2 Resin acid toxicity As mentioned previously, these relatively small amounts of resin acids have been shown to be major fish toxicants in pulping effluents. By gradually fractionating the mechanical pulping effluent, Leach (1976) found that resin acids contributed to about 60-90% of the mechanical pulping effluent toxicity to fish. Bonsor et. al (1988) also reported that resin acids were responsible for a large portion of the acute toxicity of pulp mill wastewater. Reviews by McLeay and Assoc. (1987) and Taylor et. al (1988) have also described resin acid toxicity to aquatic organisms, their persistence, and their bioaccumulative potential. 1.3.2.1 Acute lethality toxicity The acute toxicity (96h LC50) of eight common resin acids to fish was found to be between 0.2 and 1.7 mg/1 (Taylor et al., 1988). Dehydroabietic acid was the least toxic resin acid while sandaracopimaric and isopimaric acids were the most toxic. The mono and dichlorinated derivatives of dehydroabietic acid were found to be substantially more toxic than the parent molecule, yet their acute toxicity is still within this narrow range. This means that the toxicity of the resin acids in mill effluent depends more on their total concentration than each individual form (Leach et al., 1976). It has also been reported that resin acids disturb schooling behavior and muscle coordination, resulting in passive inverted drifting, muscle spasms and eventual death (Taylor et al., 1988). Resin acids are 12 also toxic to Daphnia. However, their 96h LC 5o appear to be consistently higher to Daphnia in comparison to fish (Kantardjieff and Jones, 1993). The acute toxicity of resin acids is strongly influenced by the pH of the receiving waters. Mcleay et al. (1979) showed that resin acids are least toxic in alkali pH (7.0-9.5). Taylor et al. (1988) also claimed that resin acid toxicity declines exponentially with increasing pH. In alkali conditions resin acids ionize, and, as a result, they are more polar and less lipid-soluble. These ionized resin acids are far less likely to partition across the gill membrane as readily as the un-ionized forms which predominate in acidic environments. 1.3.2.2 Chronic sublethal toxicity At lower than acutely lethal levels, resin acids can impair the ability of the organisms to function competitively in an ecosystem. This aspect has been studied by numerous researchers in an attempt to identify the biochemical/physiological changes that may occur to organisms in the receiving water. It has been reported that resin acids can disrupt the conjugation enzyme function in the fish liver, partially inhibit the secretion of conjugated bilirubin to bile, and increase plasma bilirubin levels (Folky et al, 1993). Oikari et al. (1988) analyzed subchronic metabolic effects of resin acids on juvenile lake trout and found that the activities of the conjugating enzymes p-nitrophenol UDP-GT (Glucoronosyl Transferase) were decreased. He suggested that the activity of liver UDP-GT and concentration of plasma bilirubin are sensitive methods of indicating exposure of trout to discharges containing resin acids. A decrease in blood hemoglobin level has also been observed at high resin acid concentrations (Mattsoff and Nikinmaa, 1987; Oikari et al, 1988). An in vitro study of the effects of DHA on red 13 blood cells showed that red cells broke down in the presence of 5 ppm D H A within 24 hours. However, this effect diminished considerably in vivo due to the presence of plasma proteins that bind DHA. Resin acids were also found to induce abnormalities in the shape of red blood cells (Nikinmaa, 1992). 1.3.2.3 Bioaccumulation The octanol-water partition coefficients (a measure of lipophilicity) suggest that resin acids have high bioconcentration factors (BCF). Previous work (Fox et al. 1977; Oikari et al. 1982, 1983; and Kruzynski 1979) showed that D H A bioaccumulates, with a BCF of up to 996 in bile. Further work has indicated that appreciably greater accumulations arise in fish liver than in muscle or whole body tissue. Apparently, the resin acids accumulated in fish are rapidly eliminated by their conjugation in the liver and excretion in the bile. Tana (1988) exposed rainbow trout to two different concentrations of D H A for up to 80 days. The results indicated that the liver accumulated the highest concentration of DHA, when four different tissues were compared (liver, plasma, muscle and gall bladder). At higher D H A levels (50 u.g/1), the D H A concentration in the liver continuously increased while at lower levels (5 u.g/1), the D H A concentration after 80 days was still low. This study suggested that, at higher D H A concentrations, the fish were not able to detoxify greater amounts of DHA. Niimi and Lee (1992) also found that there was a significant correlation between the decreasing percentage of conjugated/total resin acids in the liver and increasing bioconcentration. Based on this work, the authors concluded that the half lives of the nine resin acids tested were less than 4 days. It has been shown that the insertion of chlorine into organic molecules increases their fat solubility and thereby enhances their bioaccumulative potential (Kringstad et al., 1984). Although the 14 bioaccumulation of chlorinated D H A has not been studied in detail, the increased lipophilicity and greater persistence of these derivatives should result in greater problems with this class of resin acids (McLeay and Assoc., 1987). 1.3.2.4 Genotoxicity To date, only neoabietic acid and 7-oxodehydroabietic acid have so far been identified as suspect mutagens when using the Ames test with Salmonella typhimurium and the saccharomyces cerevisiae yeast assay (Beak, 1987). Five resin acids (abietic, dehydroabietic, isopimaric, pimaric and sandaracopimaric) are considered "possible carcinogens" (Taylor et al, 1988), because they have similar chemical structures (similar double bond position) to these two suspect mutagens. 1.4 Biological treatment of pulp and paper mill effluents and methods of quantification of microbial populations in biotreatment systems The pulp and paper industry has made considerable efforts in the last decades to detoxify resin acids and other hazardous pollutants in its effluents before discharging them into the receiving waters. Of the methods exploited, biological treatment has proved to be one of the best methods available for pulp and paper mill effluent detoxification. Biological treatment is nothing more than a duplication of nature's own purification process, except that it is carried out under contained and controlled conditions and usually at accelerated rates. The treatment depends on sustaining a viable mixed population of specially-adapted microorganisms, especially bacteria. The function of the bacteria is to convert the soluble organic compounds into cellular material. In other words, the bacteria utilize the organic matter as carbon and energy source for synthesis of cell components. During this process, large organics are decomposed into smaller molecules, or even mineralized into inorganic materials. At the same time, highly toxic compounds are usually degraded into less toxic substances. 15 When placed in a suitable growth medium and under favorable environmental conditions of pH, temperature and oxygen concentration, microorganisms will proceed to assimilate nutrients, leading to an increase in the size of the population as a result of growth and reproduction. In fact, the concentration of the biomass in a treatment system is such a key parameter that operational control of biological treatment has long depended on estimates of the biomass concentration within the systems. Consequently, various methods have been proposed to assess the amount of biomass concentration present in a biotreatment system. The methods for quantification of microbial populations within a biotreatment system can be divided into two groups: those which directly measure the biomass concentration within a reactor and those which assess the biological activity using chemical analysis. The direct measurement gives the total amount of biomass present. It is straightforward and usually simple, thus can be used in routine system operation. However, it does not differentiate between viable and non-viable cells. The indirect measurement reflects not only the amount of biomass but also the amount of cell activity. These methods are often complicated and based on analytical laboratories so that they are mostly used as research methods rather than for routine analysis. 1.4.1. Mixed liquor volatile suspended solids (MLVSS) MLVSS is the most commonly used measure of biomass concentration in day to day biotreatment (especially activated sludge) operation. It is still the simplest method with least analytical equipment, easiest operation and simplest calculation compared with other methods. However some problems are encountered in the M L V S S determination (Water Pollution Control Directorate, Fisheries and Environment Canada, 1977). One of these is the difficulty of filtering large sample volumes because of the poor filterability of most mixed liquor samples. Although 20 ml or less of sample volume is 16 generally used, the filtration time is still quite long (Sometimes it could take more than an hour to filter only one sample). This may result in a significant amount of dissolved solids residue on the filter paper, resulting in much higher M L VSS values than the true value. Two modified methods have been proposed to overcome this shortcoming (Water Pollution Control Directorate, Fisheries and Environment Canada, 1977). One is the centrifuging-washing method (Water Pollution Control Directorate, Fisheries and Environment Canada, 1977). The mixed liquor sample is centrifuged at 10000 rpm for 10 min. and the liquid is decanted. Then distilled water is added to wash out dissolved solids in the sludge. This helps to both diminish the interference by dissolved solids and save the filtration time. Another one is the dilution method (Water Pollution Control Directorate, Fisheries and Environment Canada, 1977). Dilution of sample with distilled water prior to filtration is used to reduce the filtration time by decreasing the solids concentration. To determine the original solids concentration, the results obtained should be corrected according to the dilution factors. M L VSS determination is based on the assumption that; 1) the biomass content is represented by the suspended solids in the mixed liquor and; 2) that only the biomass is volatile and the inert and inorganic matter present in the sludge is non-volatile at the estimation temperature of 550°C (APHA et al, 1985). However, when the inert solid particles are volatile at 550°C, the estimation of biomass by M L V S S becomes inaccurate. This problem is encountered when dealing with modifications of the activated sludge process, such as the powdered activated sludge carbon treatment process and the carrier activated sludge process (Lu and Ganczarczyk, 1983). 17 Two methods have been extensively used for the determination of biomass when carbonaceous carriers are present. One method employs the differential ignition technique developed by Zimpro Inc. (Burant and Vollstedt, 1973, and Knopp et al., 1978) which was modified by different researchers (Lee and Johnson, 1979, Arbuckle and Griggs, 1982, Lu and Ganczarczyk, 1984, and Schultz, 1987). Another method employs the nitric acid solubilization of biomass (Knopp et al, 1978, and Schultz, 1987). Both methods are highly system-specific and involve the determination of the volatilization/solubilization coefficients for both biomass and inert particles. These coefficients vary widely with respect to the type of biomass, type of inert carriers and selected experimental conditions. Moreover, since the biomass and the inert particles are intimately mixed in the carrier activated sludge, it is impossible to directly obtain these characteristics of carrier-free biomass. In 1988, Senthilanthan et al. developed a method to determine the organic nitrogen content of the MLSS for the estimation of biomass in the presence of carbonaceous carriers (Senthilnathan and Ganczarczyk, 1988). This method is simpler to some extent and more reliable. It was suggested that the method would be used in the routine analysis in full-scale plants. However it is still not straightforward as there is a need to determine the volatilization coefficients. 1.4.2 DNA determination Biological treatment systems contain not only biomass but also high concentrations of dead organic and inorganic material. Distinguishing between these constituents is important for the development of kinetic models and new design criteria for the treatment systems. One way of differentiating these constituents is to measure the content of cell components instead of M L V S S as an index of biomass. Since D N A is a key component for all living cells, various groups have looked at measuring this value with Hattingh et al. (1967) first determining the D N A content of anaerobic sludge. Other researchers improved on this method or used it in various wastewater treatment studies 18 (Schonborn, 1971; Hall et al, 1977; Raebel et al, 1980; Henze et al, 1987; Wentzel et al, 1987; Liebeskind et al., 1994). This method includes the acid extraction of DNA, the quantitative determination of deoxyribose sugar by a colour reaction with diphenylamine, the calibration with standard D N A and the mathematical conversion of this value into equivalent biomass. Thomanetz (1982) compared 17 methods of estimating living biomass and determining biomass activity. He suggested that the determination of the D N A content was the method of choice because of its simplicity, quickness, and reproducibility. However, as he points out, there is no reliable mathematical conversion factor to calculate biomass from any other measured constituents such as carbohydrates, protein, ATP, etc.. For example, some of these measured components are not exclusively found in biomass while different nutrition conditions of the organisms may result in different amounts of storage polymers (Liebeskind et al, 1994). However, the average genome size may not differ, and, consequently, a proportional factor between D N A and the number of microorganisms present can be calculated. As the genomes of bacteria do not contain introns there also is a proportional factor between the average activity and the average amount of D N A present (Liebeskind etal, 1994). 1.4.3 ATP measurement Patterson et al. (1970) first proposed measurement of microbial ATP as a possible control parameter for monitoring activated sludge. This parameter presents two advantages: 1) ATP is a central biochemical component present in all organisms; when the system has reached equilibrium state, the ATP-content remains relatively constant. 2) ATP is specific for living organisms only and thus relates directly to the viable biomass (Patterson et al., 1970; Weddle et al, 1971; Ausmus et al, 1973). 19 Notwithstanding these advantages, ATP has not evolved as a practical parameter to monitor biotreatment systems because of the complexity of the procedures required to extract ATP from microbial biomass. Before 1979, ATP was extracted by reagents such as boiling tris-buffer, ethanol, butanol, octanol, perchloric acid and CHCI3. In 1979, Kucnerrowicz et al. (1979) first introduced firefly bioluminescence technique to directly determine ATP in activated sludge. This method has proven to be simple, rapid and reliable. From Kucnerowicz's work, the ATP content has been shown to have a linear relation with M L V S S varying from 250 to 1670 ug ATP/g M L V S S (cell residence time varies from 3.7 to 15.8 days). The ATP/MLVSS ratio has a linear relation with cell residence time, specific sludge loading rate, specific substrate removal rate, and specific OUR. This indicates that ATP reflects viable biomass and that the direct ATP measurement procedure could be a useful method for monitoring and controlling activated sludge system. However he indicated that the linear relationship obtained only pertains to the appropriate cell residence time interval that was measured. It could not be extrapolated because at considerably shorter or longer residence time, a linear relationship may not exist. Subsequently, Roe and Bhagat (1982) showed that ATP/SS ratio viability varied significantly with mean cell residence time. Werf et al. (1984) also showed extensive variation in ATP/biomass ratios during different metabolic conditions. In contrast, Jorgensen et al. (1992) showed largely constant ATP content/biomass ratios which were independent of growth phase. They obtained a conversion factor of 3 mg ATP/g dw, which is in good agreement with values reported elsewhere (Patterson et al., 1970; Lawrence, 1980). Patterson et al. (1970) found 1.4-2.0 mg ATP/g dw in activated sludge, while Nelson and Lawrence (1980) reported a value of 2.5 mg ATP/g dw also in activated sludge. Although Jorgensen et al. (1992) obtained good correlation between ATP and biomass which did not appear to differ much 20 with the growth phase, they still suggested that exponential growth conditions should be established as a prerequisite when using the conversion factors for routine measurements. 1.4.4 FDA hydrolysis FDA (fluorescein diacetate), a nonfluorescent derivative of fluorescein, can be transported across cell membranes and hydrolyzed by nonspecific esterases. Because this is only done by active cells, FDA hydrolysis could provide a measurement of microbial biomass and activity. In 1980, Swisher et al. suggested using the spectrophotometry determination of F D A hydrolysis to estimate microbial biomass. FDA hydrolysis has also been used as a measurement of total microbial activity in soils (Schnurer and Roswall, 1982; Schnurer et al, 1985; Federle et al, 1986) and in freshwater (Chrzanowski et al, 1984). In 1992, Jorgensen et al. attempted to introduce FDA hydrolysis as an alternative to traditional methods of monitoring biomass within activated sludge system. They compared the FDA activity of sludge with the optical density, ATP content and OUR. They found that the FDA activity/biomass ratios changed greatly with the growth phase while other parameters gave a good correlation with biomass. In another report they (Jorgensen et al, 1992) also showed that the FDA activity/ATP ratio did not remain constant during changing environmental conditions. It was possible that not all of the microorganisms were able to hydrolyze FDA. For example, Lundgren (1981) reported that, out of 111 isolates of soil bacteria, 80% showed F D A activity. It was also shown that gram-positive cells hydrolyze F D A more readily than gram-negative cells because of the inability of FDA to penetrate the outer membrane of the gram-negative cell envelope of some species. Chrazanowski et al. (1984) found that, in samples collected from freshwater habitats, the proportion of the bacterial population showing F D A hydrolytic activity was 49-61% lower than the proportion of respiratory 21 active bacteria. Therefore, a shift in bacterial population could result in a change in the FDA activity/biomass ratio. Furthermore, the absorbance produced per unit time may change due to alterations in the physiology of the bacterial cells. Such physiological alterations could be caused by a shift in the growth medium (Brunius, 1980) or in the growth phase (Chrazanowski et al., 1984). Thus, it appears that biomass estimation based on FDA hydrolysis has severe drawbacks when dealing with environmental samples with different metabolic conditions, and, as a result, it can not be readily used as a parameter to monitor biomass growth within a biological treatment system. 1.5 Microbiology of resin acid biodegradation 1.5.1 Resin acid biodegradation in situ It has been reported that biological treatment is an effective way to reduce the concentration of resin acids present in the final pulp and paper mill effluents discharged to receiving waters. McLeay and Assoc. (1987) summarized the fate of resin acids in pulp and paper effluent treatment systems stemming from a variety of mill process operations and reported that biotreatment normally reduces resin acid concentrations to sublethal levels (<2mg/l) in all of the mill effluents they studied. The concentration of unchlorinated resin acids were usually decreased by 90% and, by as much as 99% in several systems. Dellinger (1980) also claimed similar results in their report to US EPA. Easty et al, (1978) reported a slightly broader range for biological treatment. They found that the resin acid removal efficiency was generally between 80% to 100% for Kraft and sulfite mill effluents. Chlorinated forms of D H A are far more resistant to biological removal (Easty et al, 1978). The insertion of chlorine inhibits the primary oxidation by organisms by altering the resonant properties 22 of aromatic substrates, which can in turn alter the electron density at specific sites. Beyond this, the chlorine position is also capable of inducing stereochemical effects on the affinity between enzymes and substrate molecules (Furukawa, 1982). Mueller et al. (1977) found that both activated sludge and aerated lagoons were only marginally successful at removing di-Cl-DHA, compared to the removal of unchlorinated resin acids. The average specific degradation rate of DHA is four times as that of mono-Cl-DHA and ten times as that of di-Cl-DHA. In wastewater treatment systems, resin acids could be removed in stages by various types of treatment facilities, including primary and secondary treatments, and aerobic and anaerobic treatments. Bicho et al. (1995) carried a survey of the fate of resin acids in a CTMP mill effluent treatment plant. They found that primary treatment could remove 50%-70% of the 75-200 ppm resin acids that entered the treatment system. The resin acids that were eliminated were generally associated with the solids. Subsequent treatment in the anaerobic reactor removed approximately 60% of the remaining resin acids. After the effluent was finally treated by aerobic reactors, the resin acid level was reduced to 1.5 ppm or lower. The survey revealed the importance of the aerobic phase of effluent treatment in minimizing the discharge of toxicants such as resin acids. In aerobic treatment, resin acids are removed by three main mechanisms: biooxidation by microorganisms, adsorption onto the biosolids, and air oxidation (Liu et al. 1993). It was found that, although most of the resin acids were removed by sludge adsorption during the first hours of biological treatment (<6 hours), biooxidation was the major mechanism for resin acid removal over a longer treatment time. It was also shown that air oxidation was not the major mechanism for eliminating resin acids. Hall and Liver (1996) studied the interactions of resin acids with aerobic biomass. Their results also showed that both adsorption and degradation played important roles in 23 the removal of resin acids under aerobic condition. The large partitioning of resin acids onto treatment plant biosolids was due to their limited solubilities and hydrophobic character. The most thorough survey of the in situ biodegradation of resin acids so far was done by Zender et al. (1994) in New Zealand. They investigated the fate of resin acids in a full scale lagoon treatment system receiving effluents from a bleached Kraft pulp and paper mill which process softwoods. The treatment system constituted four different phases, including three inadequately aerated lagoons (DO level was 1.2 mg/1), two highly-aerated transport systems (DO level was 5.0-6.0 mg/1), and a quiescent lake arm. It was found that the three lagoons achieved an average removal of 82% of the total resin acids, with a mass flow of about 1150 kg/d of resin acids. Abietane compounds were effectively removed through these lagoons, averaging a 91% removal. Pimarane moieties exhibited lower removals (75%), dehydroabietanes were poorly removed under this semi-anaerobic condition, and D H A and 7-oxo-DHA displayed 1% and 9% removals respectively, to give an average removal of dehydroabietanes of 1.5%. A number of resin acid transformation products were formed across these lagoons. These included hydrogenated resin acids (e.g., abietanoic, and abietenoic acids), hydroxylated resin acids (e.g., 13(3-hydroxyabietan-18-oic acid), and decarboxylated resin acids (e.g., dehydroabietin). The chemical structures of these transformation products are shown in Fig. 3. Low concentrations of the related pimarane transformation compounds, isopimarenic, 8,15-isopimaradienoic, pimarenic and 8,15-pimaradienoic acids were also observed. In some cases, the transformed products comprised about 40% of the mass flow of total resin acids. Based on these observations, the authors concluded that the predominant transformation reactions that occurred were hydrogenation, hydoxylation and decarboxylation. Alternatively, the highly aerated treatment systems exhibited more removal of abietanes and less removal of pimaranes than was carried by the inadequately aerated lagoons. However, dehydroabietanes underwent substantially higher removals 24 under aerobic condition. It was apparent that the dissolved oxygen concentration played an important role in determining the extent of degradation of the various resin acids. It was generally found that the removal rate of resin acids was higher under aerobic condition than anaerobic condition. Based on these specific removal rates, the authors ranked the resin acids by decreasing removal rate under semi-anaerobic condition: abietic > palustric > neoabietic > pimaric = isopimaric > sandaracopimaric > dehydroabietic. They also ranked the resin acids under aerobic condition, dehydroabietic > abietic > pimaric > neoabietic > sandaracopimaric > isopimaric. Abietan-18-oic Acid 13-Abieten-18-oic Acid 'Abietanoic Acid' 'Abietenoic Acid' 13fi-Hydroxyabietan-18-oic Acid Dehydroabietin 'Kinleithic Acid' Fig. 3 Structures of the major resin acids and neutral transformation products isolated by Zender et al. (1994) The anaerobic degradation of resin acids is a more complex process. It generally involves two main metabolic steps and many kinds of microorganisms. In the first step, resin acids and the other 25 organic substances in the effluent are hydrolyzed and then fermented by chemoheterotrophic, non-methanogenic bacteria to free sugars, alcohols, volatile acids, hydrogen and carbon dioxide. In the second step, methanogenic bacteria convert a limited number of compounds (i.e., acetate, H 2 , CO2, methanol, formiate and some methylamines) into methane. It has been reported that resin acids can be degraded under anaerobic condition (McFarlane, 1988; and Bicho et al, 1995). However, resin acids are considered to be recalcitrant or slowly biodegradable in anaerobic environments (Sierra-Alvarez and Lettingga, 1990). It was also found that DHA, the most abundant resin acid in pulp mill effluents, underwent little degradation (McFarlane, 1988; Bicho et al, 1995). Moreover, it was frequently reported that resin acids were highly toxic to methanogenic activity (Field et al, 1988; Sierra-Alvarez and Lettingga, 1990; Bicho et al, 1995). Field et al. (1988) reported a 50% methanogenic inhibitory concentration of 160 mg COD/L for crude pine resin and 114 mg/1 for abietic acid alone. In another study, it was reported that 89 mg/1 abietic acid was needed for 50% inhibition, compared with 43 mg/1 for the more methanogenic toxic DHA. To achieve 80% inhibition, 139 mg/1 and 105 mg/1 of the respective acids needed to be added (Sierra-Alvarez and Lettingga, 1990). According to Sierra-Alvarez and Lettingga, the possible mechanism of toxicity is that the lipid solubility and surface activity of resin acids leads to the destruction of cell membranes, disturbance of cell division through alteration of surface tension, and a host of non-defined chemical interactions. They also suggested that polymerization of aromatic terpenes decreased fungal toxicity and that monomelic aromatics were more effective inhibitors than their dimeric homologs. Although resin acids are strong inhibitors of anaerobic activity, toxicity towards these microorganisms can only be partially explained by the presence of resin acids (Kennedy, 1992). Other toxic resin constituents may also exert adverse effects on methanogenic mixed cultures (Aguilar et al. 1995). It has also been reported the methanogenic consortia may acclimatize to 26 supersaturated concentrations of resin acids (Sierra-Alvarez and Lettingga, 1990; and McFarlane et al, 1988). 1.5.2 The resin acid degradation pathways The successful treatment of resin acids by many kinds of biological facilities has provided some of insight into the mechanism of resin acid biodegradation. From 1960's to 1980's, research primarily focused on the study of resin acid degradation pathways. A number of bacterial and fungal pathways have been proposed mainly based on the identification of D H A degradation products. These are reviewed in the following sections. 1.5.2.1 Degradation by fungi In one of the earlier studies in this area, Brannon et al. (1968) incubated methyl dehydroabietate with the fungus Corticium sasakii, and found that this fungus could hydroxylate methyl dehydroabietate at the C3 position. Ekman and Sjoholm (1979) also demonstrated that the fungus Forties annosus could hydroxylate DHA and transformed it into the more polar products. Both ip,15- and ip,16-dihydroxydehydroabietic acids were identified as the major transformation compounds, while 15- and 16-hydroxydehydroabietic acids constituted only minor transformation products. In subsequent work Kutney et al. (1981) studied other aspects of fungal degradation. They first screened numerous microorganisms from a variety of sources, including laboratory culture collections, the American Type Culture Collection, laboratory fermentation sludge, and a Kraft mill effluent, and assayed them for their ability to degrade DHA. Mortierella. isabellina was shown to be the most effective D H A degrader among all the tested organisms. This fungus was obtained from the UBC culture collection in the Department of Botany. They reported that this organism had the 27 tendency to oxygenate D H A at C 2 ( a and P), C15 and Cl6. The C 2 oxygenation on Ring A was the first step in the attack of the molecule and occurred before the C15 and Ci6 hydroxylations on the isopropyl end of the compound could occur. Neither ring cleavage reaction nor further alteration of the parent molecule beyond the trihydroxylated form was observed. Subsequently, these same workers (Kutney et al, 1981a, 1982a) reported that M isabellina also had the ability to degrade isopimaric and abietic acids. As occurred with DHA, isopimaric acid underwent trihydroxylation at the C 2(a), C15 and C i 6 positions and there was no subsequent ring cleavage. However, with this acid, the dihydroxylation at C15 and Ci6 led to saturation at the methyl end of the molecule through the removal of a double bond. Likewise, abietic acid was also hydroxylated at the same sites as D H A and isopimaric acid. However, the final products were not trihydroxylated molecules but either C2(ct), C15 and /or C 2(a), C i 6 diols. Kutney et al. (1982b, 1983a, 1983b) proceeded with the study of chlorinated resin acid biodegradation by M. isabellina. They found that chlorinated resin acids could also be hydroxylated by this organism, albeit more slowly. Al l of the chlorinated forms of D H A were more resistant to biodegradation than the unchlorinated ones, and the dichlorinated D H A more resistant than the monochlorinated forms. In the case of 14-C1-DHA, hydroxylation occurred at C2(ct), followed by the subsequent hydroxylations at C15 and/or Ci6, yielding a series of secondary metabolites. With 12-Cl-DHA, hydroxylation at both C 2(a) and Ci6 yielded dihydroxyl diasteroisomers which resisted any further transformation. 12,14-diCl-DHA could also be degraded by M . isabellina. C 2(a) was first hydroxylated followed by a secondary hydroxylation at Ci6. The C 2(a) hydroxyl group was then oxidated to the ketonic form. With extended incubation, some further oxidation of the C i 6 hydroxyl to carboxyl form was also observed. 28 In summary, the fungal pathway for resin acid degradation was shown to occur through the hydroxylation at the C2(a), Ci 5 and C l 6 positions. The overall biotransformation of DHA by M isabellina is described in Fig. 4. Fig. 4. The biotransformation of DHA by M. isabellina (Kutney et al., 1981 a) 1.5.2.2 Degradation by bacteria Bacteria have also been shown to degrade resin acid and in 1968, Biellmann and Wennig first reported the degradation of DHA by Flavobacterium resinovorum isolated from the soil of a Pinus maritima forest. This organism was able to use DHA as its sole carbon source. Biodegradation was thought to initially take place on the aromatic ring, or at least at the benzylic position, since resulting metabolites no longer possessed an aromatic character. One degradation product was identified as a C 3 ketone of DHA. This led them to propose biodegradation via an enzymatic hydroxylation at C 3 to form the alcohol (a or P), followed by oxidation to ketone which underwent spontaneous or 29 enzymatic decarboxylation. In a similar study, Levinson et al. (1968) isolated a bacterium Arthrobacter sp. from lodgepole pine. This organism also hydroxylated D H A at C3, producing 3-oxo-DHA. In a continuation of their early work, Biellmann et al. (1973a) confirmed that the decomposition of D H A by F. resinovorum started with hydroxylation at C 3 and C 7 . It was not known which oxydation occurred first, but both sites must first be oxidized before the subsequent dihydroxylation of the aromatic ring at Cn and Cn can proceed. Activation of the aromatic ring led to its cleavage, yielding 2-isopropyl malic acid, and two single ring acids (4-methylene-3-methylcarboxylic cyclohexanol). Further oxidation was possible, leading to a variety of simple acids which were completely degraded to CO2 by other bacteria (see Fig.5). Biellmann et al. (1973a) also found an alternative route for D H A degradation, in which ring A and B remained together as a triketone bicyclic structure after the hydroxylation and the removal of the aromatic ring. Again, a series of oxidations opened the ring and produced a number of simple acids (Fig.6). Through further studies, Biellmann et al. (1973b) found that two other organisms, A. eutrophus and a Pseudomonas sp., which were isolated by aerobic enrichment, both had the ability to use D H A as the sole carbon source, and could degrade D H A using a similar metabolic pathway. The first step was the hydroxylation at C7 followed by the dihydroxylation of the aromatic ring and then the removal of this substituent (Fig. 7). The attack on the aromatic moiety was accomplished by the addition of oxygen, producing a diene-diol which was then oxidized to a diphenol (Fig. 8). The cleavage of the aromatic ring resulted in the hydroxylation of bicyclic diketone at C 5 and then the cleavage of the ring B. Further oxidations produced the two illustrated diacids. Interestingly, Pseudomonas sp. and A. eutrophus did not oxidatively attack C3 as was shown to occur with F. 30 resinovorum. Hence, different bacteria seem to have adopted different approaches to the degradation of resin acids. Fig. 6. An alternative path for D H A metabolism by F. resinovorum (Biellmann et al, 1973) 32 Fig. 7. The metabolism of D H A by A. eutrophus and a Pseudomonas sp.(Biellmann et al, 1973) • 0 2 • reductant NAD(P)* NAD(P)H • H* Fig. 8. Oxidative attack of the aromatic moiety of DHA. 1.5.3 Physiology of the resin acid-degrading organisms Resin acid degrading microorganisms cover a wide range of species. As discussed earlier, a number of fungi have been shown to be able to degrade resin acids, including Corticium sasakii (Brannon et al., 1968), Fomes annosus (Ekman and Sjoholm,1979), Mortierella. Isabellina, Stereum purpureum, a Nodulisporium sp., Ascochyta pinodella, Trichoderma viride, a Cephrlosporium sp., Aureobasidium pullulans, (Kutney et al., 1981), Ophiostoma piceae, Ophiostoma ainoae, Ophiostomapiliferum, and a Lexington species (Wang et al., 1995). Numerous bacteria have also been identified as resin acid degraders, including Flavobacterium (Biellmann and Wennig, 1968), Arthrobacter (Levinson et al., 1968), Pseudomonas (Biellmann et al., 1973b), Bacillus, Escherichia (Hemingway and Greaves, 1973), Alcaligenes (Cross and Myers, 1967), Sphingomonas, Zoogloea (Mohn, 1995), as well as some other bacterial strains which have not yet been classified (Bicho et al., 1995). Some of these strains were able to use resin acids as the sole carbon source. The resin acid degrading fungal and bacterial isolates that have been characterized to date are shown in table 34 5. These organisms were isolated from a variety of sources, including soils, natural water systems, wood, sewage plants, pulp mill effluents and their treatment systems. Table 5, Summary of resin acid degrading fungal and bacterial isolates (Liss et al, 1996). Bioactivity References Bacteria Bacillus sp. DHA utilization Heimingway and Greaves 1973 Escerichia coli D H A utilization Heimingway and Greaves 1973 Flavobacterium sp. D H A utilization Heimingway and Greaves 1973 Flavobacterium D H A utilization Biellmann et al. 1973 a resinovorum Pseudomonas sp. D H A utilization Heimingway and Greaves 1973 D H A utilization Biellmann et al. 1973b DHA, A B A , P A L utilization Wilson et al. 1996 Alcaligenes eutrophus D H A utilization Biellmann et al. 1973b Arthobacter sp. D H A (methyl ester) utilization Levinson et al. 1968 Sphingomonas sp. DHA, A B A , P A L utilization Mohn 1995 Zoolea sp. DHA, A B A , PAL utilization Mohn 1995 Comamonas spp. A M A , DHA, N A A , PA, IP A, Morgan and Wyndham 1996 L P A utilization Alcaligenes spp. A M A , DHA, N A A , PA, IP A, Morgan and Wyndham 1996 L P A utilization Gram negative strains Abietane utilization Bicho et al. 1994 Fungi -Mortiella isabellina resin acid hydroxylation Kutney et al. 1981a; 1985; 1988 Chaetomium cochlioes D H A hydroxylation Ymoetal. 1994; 1995a; 1995b Corticum sasaki methyl D H A hydroxylation Brannon et al. 1968 Fomes annosus D H A oxidation Ekman and Sjoholm 1979 35 As mentioned earlier, more recent research has focused on resin acid degraders which have been isolated from pulp mill effluent treatment systems. This has partially been due to the concern that many of the organisms previously isolated from other environments may not be representative of the microbial populations within the treatment plants. Thus, it would probably be more informative and beneficial to pulp mill effluent studies, if the resin acid degraders were obtained from real treatment systems. Bicho et al. (1994a) isolated five D H A degrading bacterial strains from a treatment system for bleached Kraft mill effluents. These strains, designated B K M E 5, B K M E 6, B K M E 7, B K M E 9, and B K M E 11, were all found to be Gram negative, aerobic rods. Biochemical typing showed that the isolates had some physiological similarities but were taxonomically distinct. However, the authors did not identify these strains taxonomically. Al l five isolates were able to use D H A as the sole carbon source and tolerated concentrations as high as 200 ppm DHA. In a similar study, Mohn (1995) isolated eleven bacteria from a sequencing batch reactor designed to treat a high-strength process stream from a pulp mill. These bacteria belonged to two groups, represented by strains Dha-35 and Dha-33 which were more thoroughly characterized. Analysis of small-subunit rRNA partial sequences indicated that Dha-33 was most closely related to Sphingomonas yanoikuyae and that Dha-35 was most closely related to Zoogloea ramigera. The two strains were also able to grow on defined medium with D H A as the sole carbon source. It was found that D H A degradation was inducible (Bicho et al, 1994; Mohn, 1995), and it was further demonstrated (Mohn, 1995) that DHA induced abietane degradation activities by DHA-degrading isolates and these activities were heat labile. A 2-4 hour lag of the D H A degradation was observed by both authors when following the growth, suggesting that this lag could be attributed to the need for enzymatic induction. Bicho et al. (1995) further investigated this by spiking glycerol-grown high-density cultures with DHA in the presence or absence of tetracycline at various times 36 during the D H A disappearance curve. Tetracycline is a protein synthesis inhibitor which acts at the translation level. Cultures to which the antibiotic was added prior to the lag did not degrade DHA. However, those cultures that were spiked with the antibiotic after the lag phase were able to degrade D H A at the same rate as did controls with no added tetracycline. Therefore, de novo protein synthesis was required for D H A biodegradation, confirming that this activity was inducible (Bicho etal., 1995). It was observed that these resin acid degrading bacteria were relatively substrate specific (Bicho et al., 1995). Although they were unable to degrade the pimaranes to any great extent, these DHA degrading bacteria could degrade D H A and all the other abietanes. Only strain B K M E 6 and B K M E 9 were able to degrade pimaranes at an efficiency of more than 75% while consuming most of the abietanes. The other three strains demonstrated little or no degradation of pimaranes. Similar results were obtained in a parallel study (Mohn, 1995). This worker found that the D H A degrading isolates Dha-33 & 35 could additionally grow on other abietanes, i.e., abietic and palustric acids, but not on the pimaranes, pimaric or isopimaric acids. Bicho et al. (1995) attributed this to the primary difference at the C-13 of the tricyclic skeleton between the abietanes and the pimaranes. They suspected that the C-13 side group might be important in either the induction or specificity of the enzymes involved in resin acid biodegradation. Mohn (1995) proposed that the substituents at the C-13 of pimaranes inhibit degradative enzymes or prevent their induction. Despite the substrate specificity of DHA-degrading organisms within the resin acid groups, these organisms did have the ability to metabolize various organic compounds such as wood sugars, fatty acids and a wide range of organic acids including amino acids and T C A cyclic intermediates (Table 6 & 7) (Bicho et al., 1994; Mohn, 1995). Some of these substrates could be used as the sole carbon 37 Table 6. Physiological characterization of the five strains on Biolog G N plates (Bicho et al, 1994) • metabolizes substrate, • does not metaboliz* substrate, / weak or Inconclusive reaction 5s 73 93 11a Gram stain _ Cell morphology rod rod rod rod rod Oxidative reaction + • + + Fermenative reaction _ Catabse • • • Oxidase • Motility Cyclo-dextrin _ Dextrin _ Glycogen _ Tween40 +• • + + TweenoO - + + + N-Act-Cal-amine _ N-Act-Clu-amine + Adonitol _ + L-Arabinose + + D-Arabrtol + + + Cellobiose Erythritol _ m D-Fructose + + L-Fucose D-Galactose + m + +• Gentibiose D-Glucose • +• + + m-lnositol +• • + Lactose Lactulose Maltose + Mannrtol +• + D-Mannose + + Melibbse _ _ Methyt-Glucosioe _ D-Psicose + • + + Raffinose m Rhamnose m + + Sorbitol Sucrose Trehalose Turanose _ Xylitol _ +• + + Methyt-pyruvate + + • + 1 -methyl-succinate +• + • + Acetic acid +• + Acon'rtic acid + + . Citric acid + • + +• Formic acid • +• + Galactonic acid (lad) + / + + Galacturonic acid _ Gluconic acid + 2,3-8 utaned'ol + + + + Glycerol - _ + + D,L-a-Gfycerol PCM -Glucose-1-P04 Glucose-S-PCM - + _ 5s 6b 7s 9s 11 _ + Glucosamine acid _ Glucuronic add + + * e-Hydraxybutync acid • • + 4. b-Hydroxybutync acid • • • • 8-Hydroxybutyric acid • p-OH-Phenyt acetic acid • • • rtaconlc acid _ • a-keto-butyric add + + a-Veto-alutaric acid + • a-keto-valeric acid • • D,L-Lactic acid + + / MalonieAcid + + + Propionic acid • • + Quinic acid + Saccharic add + + Sebacic acid + + Succinic acid + • • + Bf-Succinic acid + + Succinamic acid _ Glucuronamide + + 4 Alanhamide • + D-Alank>e • + L-Alanlne • L-Alanyl-glydne + • + • + L-Asparagine • + *• + + L-Aspartic acid • + + + L-Glutamic acid + _ + + Glycyl -L-espartic acid + _ + + Glycyl -L-glutamic acid + + _ + + L-Histidine + / OH-L-ProSn* • / / L-Leucine . + 1 / L-OmHhine * / L-Phenylalanine + + + + L-Proine + 4- + L-Pyroglutamic acid + + + + D-Serine • + L-Serine • +• + L-Threonine + • + + D.L-Camrtine + + +• • g-Amino butyric acid + + + • Urocanic acid + • + + knosine + + + Uridine Thymidine + + Phenyl ethytamine • / Putrescine • + + + 2-Amino ethanol _ Pimarate + / • Sandaracopimararte • 1 • bopimanrte + 1 + Dehydroabtetate • + + • Abietate - + 1 38 Table 7. Substrate use by D H A degraders Dha-33 and Dha-35 (Mohn, 1995) Substrate Concn (g/liter) Substrate DhA-33 use by": DhA-35 Wood sugars Arabinose 0 NO Cellobiose 0 NO .. Galactose 0 NO Glucose 1.0 G NG Glucuronate NO O Xylose 1.0 G (G) Resin acids Abietic 0.060 G G Dehydroabietic 0.060 G G Isopimaric 0.060 NG N G Palustric 0.060 G G Pimaric 0.060 NG NG Fatty acids Linoleic 0.20 G NG Palmitic 0.20 G G Others Acetate 1.0 G G Benzoate 0.24 NG G Ethanol 1.0 G G Glycerol 1.0 NG NG Methanol 0.10 NG NG Pyruvate 1.0 (G) G • G, supports growth as sole organic substrate in 2.5-ml culture with stationary incubation; (G), supports poor growth; NG, does not support growth; O, oxi-dized in Biolog assay, NO, not oxidized. source by these organisms, while others could only be utilized as co-substrates. In general, these resin acid degraders could metabolize organic acids very well. Both authors looked at the ability of these resin acid degraders to grow on wood sugars, fatty acids, acetate and methanol. These are potential substances which can be expected to be abundant in pulp mill effluents. It was found that, although the D H A degrading bacterial isolates could metabolize most of these compounds, few of the isolates could achieve effective growth on these compounds. As co-substrates sometimes significantly affect resin acid biodegradation, Mohn (1995) demonstrated that the use of glucose as a co-substrate increased the rate of D H A consumption by one DHA-degrading strain, while the use of pyruvate as a co-substrate decreased the rate of D H A degradation by another strain. Thus, the complex organic composition of biological treatment system influents can significantly influence resin acid removal by direct effect on resin acid metabolism. While a considerable amount of work has looked at the resin acid biodegradation pathways and the physiological characteristics of resin acid degraders isolated from pulp mill wastewater treatment systems, little work has been done to establish the growth conditions and maximum resin acid biodegradation potential of mill effluent treatment systems. One of the objectives of the work described in this thesis was to gain some insight into the impact of pulp mill effluents on resin acid biodegradation and to identify possible ways to improve the resin acid degradation in real treatment systems. 1.6 Objectives One of the initial objectives of this thesis was to develop a simple, quick and reliable method to quantify the resin acid degrading bacteria present in pure cultures and within the mixed liquor biomass of the activated sludge systems. It was recognized that some of the regularly-used methods (e.g., various protein assays) for biomass quantification may not be effective in the presence of mill effluent or resin acids. As all of the research in this thesis will involve the monitoring of bacterial growth, it was essential that we first develop a suitable method of quantifying bacteria. A second objective was to establish optimum growth and resin acid degradation when the bacterial isolates were grown on CTMP effluent and determine how the various components in CTMP effluent might affect resin acid biodegradation. All previous work with the bacteria isolates studied had used buffer solutions to which resin acids had been added. It is probable that the CTMP effluent, with its high concentrations of complex components, will influence resin acid degradation. 40 For example, it may inhibit degradation by repressing the transcription of genes coded for enzymes needed in resin acid metabolism, or by toxically inhibiting the growth of resin acid-degrading organisms. Alternatively, the CTMP effluent could perhaps improve resin acid degradation by stimulating the growth of resin acid-degrading, or by enhancing resin acid degradation with some components used as co-metabolic substrates. The major focus of this objective was to verify the previous results of Bicho et al. (1995) by growing the organisms on CTMP effluent. Another aspect was to compare different methods that could be used to sterilize mill effluent without changing the nature or composition of the effluent. We investigated how much influence the different sterilization methods have on the effluent composition and subsequent utilization of resin acids in or added to these effluents. A final objective was to identify the possible ways of improving degradation of resin acids present in CTMP effluents. As the CTMP effluent is likely to be nutrient deficient when compared to the defined media employed by various researchers in the original isolation and characterization of resin acid degrading organisms (Bicho et al, 1995; Mohn, 1995), some nutrients could be critical for effective resin acid biodegradation. The results reported in this thesis have identified some of the nutritional requirements for the resin acid degrading bacteria studied in this work. 41 CHAPTER 2. DEVELOPMENT OF METHODS FOR THE QUANTIFICATION OF MICROBIAL POPULATIONS IN BIOLOGICAL TREATMENT SYSTEMS 2.1..A SOLID COD METHOD FOR DETERMINING MICROBIAL BIOMASS IN BIOTREATMENT SYSTEMS 2.2. BACKGROUND As mentioned in the introduction, one of the main objectives of this thesis was to monitor the growth and substrate specificity of resin acid degrading bacteria when they are added to sterilized mechanical pulp effluents. Unfortunately, because of the varied composition of pulp mill effluents, many of the common methods used for quantifying microbial biomass could not be readily adopted in this study. For example, F D A hydrolysis could not be used due to the poor FDA activity of the gram-negative bacteria which were used in this study. The determination of D N A has been shown to be strongly influenced by unknown activated sludge components and iron (Liebeskind and Dohmann, 1994), thus it could not be readily adopted for use in this study. ATP and M L V S S determination are both very time consuming, and are not suitable biomass indicators for type of study carried out in this thesis where large numbers of samples were routinely analyzed. A review of the recent work on the microbiology of resin acid degradation indicated that methods such as optical density (Morgan and Wyndham, 1996), standard plate count (Bicho et al, 1994) and microscopic total cell count (Mohn, 1995) were primarily used as indicators of microbial growth. However, the large amount of particulate material present in the CTMP effluent and the high concentrations of resin acids used in this study interfered with both the optical density measurements and the microscopic total cell counts. The plate count method appeared to be the only reliable way of monitoring cell growth in the presence of mill effluent and 42 resin acids. However, as it was most time-consuming of all of these three methods, this convinced us to look at alternative ways of quantifying microbial biomass. As will be described in detail in a later chapter of the thesis, when we tried to determine the chemical oxygen demand (COD) value of the gamma irradiated CTMP effluent inoculated with bacterial isolates, the values were influenced by the size of the inoculum added to the effluent. One way of solving this problem was to subtract the COD which was attributable to the cells from the total COD. The success of this approach indicated that the COD resulting from contribution of the cells could be used as an indicator of the microbial growth within the biotreatment systems. With the goal of establishing a correlation between the cells within the treatment systems and their COD equivalent, we developed a new method of biomass quantification, which used the chemical oxygen demand (COD) to determine the quantity of suspended organic material in the biotreatment systems. This method was based on the similar principle behind the use of volatile suspended solids (VSS), which is a gravimetric method routinely used as the primary procedure by workers in this area to quantify microbial biomass (Metcalfe and Eddy, 1979). As this COD method, which we subsequently call the "solid" COD method, required less sample volume and yielded results in a shorter period of time (3-5 h) than the VSS, it should be possible to use it to monitor microbial growth in both the biotreatment systems and bacterial cultures. Volatile suspended solids are usually determined by measuring the mass of oven-dry solids retained by a 1.5 u.m glass fiber filter that has been volatilized at 600°C. A sample containing a minimum of 50 mg VSS is recommended so that, good, reproducible results can be obtained (APHA, 1992). Prior to the assay, the filter must be heated (600°C/15 min) to combust any 43 adsorbed organic material. However, the poor filtering properties of some biological "sludges" makes filtering this quantity of material difficult and time consuming. Furthermore, drying and combusting the sample adds another 24 h to the analysis. For these reasons, quicker, more efficient methods for determining biomass were desirable. The solid COD method described here was initially developed.to determine biomass concentration in batch growth flasks. Several flasks were monitored in parallel with multiple samples collected from each flask during the course of the experiment. As the initial effluent volume was about 1L, the volume and amount of biomass from samples collected during each experiment was limited. This limitation, combined with the large number of samples to be analyzed, meant that a simple and quick method of determining biomass was required. The biomass quantification methodology described in this section is based on the COD of filterable biosolids. A series of experiments were performed to determine whether effluent characteristics, such as the age and source of biomass influenced the relationship between solid COD and VSS. Initially we determined if soluble COD from different wastewaters interfered with the measurements and whether different stages of growth (i.e. lag, exponential growth, stationary and decline phases) influenced the ratio between COD and VSS. Finally, the COD/VSS ratio for biomass resulting from different wastewaters was compared. In this way, the dependability of the solid COD method as an alternative to VSS measurements was assessed, particularly when sample volume and /or time is limited. The similarities between the VSS and "solid" COD filtrations suggested that good correlation could be obtained between these two procedures. An interesting exception was found in effluents containing high concentrations of inert particulate which had a different solid COD/VSS ratio 44 from the growing biomass. The advantages of the "solid" COD over the VSS protocol include, rapid analysis time, greater sensitivity and the use of smaller sample volumes. 2.1.2 MATERIALS AND METHODS 2.1.2.1 Solid COD Samples containing less than 2 mg of suspended solids were filtered through gooch crucibles fitted with a Whatman AH-934 (1.5 u,m retention) glass fiber filter. The crucibles and filters had been combusted at 600 °C for 15 min and cooled prior to use. The filter and sample were rinsed with 10 ml distilled water to remove residual soluble COD. The filter and retained solids were transferred to screw cab test tubes with 2.0 ml distilled water and digested according to the Standard Method 5220 D — Closed reflux, colorimetric method (APHA, 1992). The digested supernatant was centrifuged (14,000 r.p.m; 3 min) in Eppendorf tubes to remove any suspended glass fibers that would interfere with the absorbency readings (A,=600 nm). Absorbency readings were calibrated against the theoretical COD of potassium hydrogen phthalate standards (0-900 mg 0 2/l). 2.1.2.2 Volatile Suspended Solids VSS were determined according to Standard Method 2540 G (APHA, 1992). The same type of crucibles and glass filters used with the solid COD method were also used here. 2.1.2.3 Effluent and activated sludge Activated sludge from three different effluents (bleached Kraft mill effluent (BKME), chemi-thermo-mechanical pulping (CTMP) effluent and municipal sewage were used to evaluate the versatility of the solid COD method to estimate VSS. Both the untreated B K M E and CTMP 45 effluent came from pulp mills situated in the interior of British Columbia. These effluents were shipped to the laboratory and frozen at -20°C until required. Two lab scale activated sludge reactors (5L aeration basin volume), described earlier (Bullock, 1994) provided the source of sludge. The reactors were operated with a 48 h hydraulic retention time (HRT) and 10 d solid retention time (SRT) using the two pulp effluents described above as the substrate. Municipal sewage and activated sludge were obtained from the UBC Civil Engineering biological phosphorus removal pilot plant. 2.1.2.4 Comparison of COD and VSS values Initial experiments on each of the effluent and activated sludge samples involved simple dilution of the mixed liquor with effluent to determine if the effluent contributed any interference specific to the solid COD method. Mixed liquor (100%) from either the aeration basin of the reactors and the sewage treatment plant was diluted with effluent to obtain 80, 60, 40, 20, and 0% of their original concentration and analyzed for both VSS and solid COD in triplicate. To establish if the proportion of the solid COD to VSS of biomass remained constant throughout the different stages of microbial growth, flasks of effluent were inoculated with activated sludge and then monitored for 1 week. The flasks experienced all growth stages (i.e. lag, exponential, stationary and decline phases). One liter volumes of effluent were inoculated with 5 ml of mixed liquor and incubated on a reciprocal shaker (40 cycles/min) at 20-22°C in the dark. The flasks were sampled daily for VSS and both solid and soluble COD in triplicate. 46 2.1.3 RESULTS AND DISCUSSION Initially we wanted to determine if the liquid part of the effluents interfered with the measurement of solid COD, thus dilution tests with prefiltered effluent as a diluent were performed. If the filtered effluent did interfere with the solid COD measurement, the resulting curve would be expected to be offset. However, the values presented in Fig. 9 show that the dilution of the biomass derived form both the sewage and B K M E grown cells resulted in correlation curves with nearly zero y-intercepts. This indicated that the soluble COD in the filtered effluent did not interfere with the solid COD measurements. In a second set of experiments the solid COD content was compared with the VSS determined over the course of a growth curve (Fig. 10). In this way, any changes in the correlation due to differences in cell physiology and morphology that may result from different growth stages, could be observed. Samples collected during growth experiments with sewage, B K M E and CTMP effluent were monitored for solid COD and VSS (Fig. 10). The solid COD data paralleled the VSS data, indicating that the microbial community grew well in all three effluents. The ratio between COD and VSS values remained relatively constant throughout the time course for both the B K M E effluent and sewage (Fig. 11). This constant relationship indicated that the oxidative state of the microbial community did not change as the population adapted from an environment with readily available substrate (lag and exponential growth) to a substrate limited one (stationary and/or decline phase). A consistent ratio is important for applications such as time course's since it can be assumed that certain trends in solid COD content are related to changes in biomass content and not due to an artifact from a shift in the COD/VSS relationship. However, a downward trend was noticeable for the biomass growing on the CTMP. It was probable that 47 1500 1200 800 4 600 3004 Sewage COD-9.96*1.e3(VSS) 1500 BKME 12004 E, O O O tj "o 900 H 600 H 300 H COO = 4.95 • 1.64(VSS) > 0.998 B CTMP 1500-1 Ay 1200- COD = -53.5*1.66(VSS) R J=0.99 900- n = 6 y 600-300-/ k 0-150 300 450 600 750 000 VSS (mg/L) Fig.9. Comparison of the solid COD and VSS values for (A) sewage sludge, (B) B K M E and (C) CTMP effluents at various dilutions. Each dot represents the mean value of three replicates with the standard deviation less than 5%. 48 200 150-1 100 50 Scwag* • Soluble COD(mo/) -«-VSS(m(yi) A^-A-SoSdCOO(morD 400 300 200 100 ~> 1 1 1 1 1 r 1—10 0 1 2 3 4 5 6 7 200 BKME 150 E 8 ioo E o ffi 50 • ' 1200 900 W c o-(A 600 | 3 300 2400 1800 -11200 B Time (days) Fig. 10 Comparison of growth trials monitored on (A) sewage sludge, (B) B K M E and (C) CTMP effluents using the solid COD and VSS methods. Each dot represents the mean value of three replicates with the standard deviation less than 5%. 49 CO CO Q O O o CO 1 2 3 4 5 Time (days) Fig. 11 Ratio of the solid COD vs VSS determined over the course of the growth trials on sewage, BKME and CTMP effluent. Each dot represents the mean value of three replicates with the standard deviation less than 5%. 50 CTMP effluent contained a significant concentration of non-biological suspended solids which would result in a different proportion of COD to VSS. Thus, a possible explanation for the observed downward trend could be due to the transition from a system with high particulate, of low biomass content to one with an increasing proportion of biomass. The influence of substrate composition on the COD/VSS ratio was affected when the values from the two experiments using the three different effluents were compared. In the dilution experiments (Fig. 9), the slopes and standard error of the three correlation curves were very similar; i.e. 1.63±0.10, 1.64+0.04, and 1.66±0.11 mg COD/mg VSS for sewage, B K M E and CTMP sludge respectively. Similarly, the COD/VSS ratios for the time courses resulted in comparable ratios for biomass growing on both the sewage and B K M E effluent (Fig. 11). These results indicated that the oxidative state of both microbial communities was similar. These values are very similar to the theoretical COD yields of many organic compounds. However, the CTMP biomass exhibited a lower ratio between solid COD and VSS. The high concentration of suspended solids in the CTMP effluent may have contributed to this finding. Compared to the sewage and B K M E , the CTMP effluent contained relatively high concentrations of suspended and colloidal material (87.5 mg VSS/1). As mentioned earlier, this material had a different COD/VSS ratio when compared to the microbial biomass typically obtained in the above flask experiments. In addition to masking the biomass concentration, the COD/VSS ratio shifted during the time course as the proportion of inert solids decreased and biomass increased. Based on these results, the use of the solid COD method should be restricted to effluents with a low suspended solids content that is not of microbial origin. In most cases, this can be addressed by clarification of the effluent by either centrifugation or filtration prior to inoculation. 51 The good correlation obtained between solid COD and VSS lead to subsequent work which utilized this technique in a series of time course experiments using B K M E as the substrate (Bullock et al, 1995). Both the VSS and solid COD were measured over a range of concentration for each flask (Fig. 12). The slope and standard error of the correlation line was 1.2±0.12 mg COD/mg VSS, which was comparable to the results presented earlier. The solid COD method has a high sensitivity even at low concentrations of biomass. Since the sample size for COD measurement is determined by the biomass content, low concentrations of biomass can be measured simply by increasing the sample volume. However, the high sensitivity of the assay limits its application when higher concentrations of biomass are encountered such as the occurrence of flocculating microorganisms. Cultures producing large floes (> 1.5 mm dia) are difficult to sample representatively, given the small sample volume required. As a result, the sampling error increases. Depending on floe size and the homogeneity of the culture, a minimum sample volume of about 2 ml was required. This limited the upper biomass concentration to approximately 1000 mg COD/1 (the upper limit of the COD assay). The solid COD method proved to be particularly useful for performing batch growth curves since the COD analysis of the filtrate from the solid COD samples could be used as a measure of substrate concentration. The combination of these two parameters allowed us to easily monitor growth kinetics for different substrates. In the case of Monod kinetics, parameters such as yield, half-saturation constant and maximum growth rate can be easily determined using both solid and soluble COD measurements for biomass and substrate concentrations, respectively. Similar studies by Slade and Dare (1993) used an indirect COD biomass estimation. However the direct 52 Fig. 12 Correlation between solid COD and VSS measurements compiled over multiple trials with BKME. Each dot represents the mean value of three replicates with the standard deviation less than 5%. 53 estimation of filterable solids, as described in this chapter, has the advantages of using less sample volume and greater sensitivity. 2.2 COMPARISON OF METHODS FOR QUANTIFYING RESIN ACID DEGRADING BACTERIA IN PURE CULTURES BY PROTEIN ASSAYS, PLATE COUNTS, TOTAL C E L L COUNTS AND THE SOLID COD METHOD 2.2.1 BACKGROUND Initially we developed the solid COD method with the intention of establishing the correlation between the cells and their COD equivalent, and using it to quantify the microbial growth. However, as shown in the previous section, the solid COD/VSS ratio was not constant when a large amount of non-biomass solids was present in turbid effluents such as a typical CTMP effluent. Therefore, a suitable method for monitoring the growth of resin acid degrading bacterial strains within tested systems still needed to be established. Traditionally, total cell count, plate count and optical density have been used extensively as indicators of bacterial growth in the field of microbiology (Brock and Madigan, 1991). Protein concentration has also been used to estimate biomass by biological researchers (Donovan, 1969). However, in the present study the optical density could not be used to measure biomass production as the liquid media was usually, turbid when more than 40 ppm resin acids were present. Therefore, the only possible methods that could be used were the plate count, protein assay and total cell count methods. This section describes the influence of resin acids and the CTMP effluent on the various methods used to evaluate microbial cell growth. 54 2.2.2 MATERIAL AND METHODS 2.2.2.1 Protein assays in the presence of resin acids and CTMP effluent Standard protein stock solutions (1000 ppm) were prepared by weighing out 50.00 mg BSA and dissolving it in 50.00 ml phosphate buffer (pH=7), 100 ppm resin acid water solution, and CTMP effluent, respectively. Standard B S A preparations were made by diluting the protein stock solution with the corresponding diluent. Thus the BSA-buffer, BSA-resin acid, and BSA-CTMP stocks were diluted with buffer, resin acid solution and CTMP effluent, respectively. The Fluorescent protein assay, Lowry assay, Bio-Rad regular assay and micro assay were all compared. These three assays were performed according to methods described by Roth (1971), Pederson (1977), and Bradford (1976), respectively. 2.2.2.2 Comparison between the plate count, total cell count and solid COD measurements of microbial growth. Bacterial strains B K M E 5 and B K M E 9 previously described by Bicho et al. (1994) were grown on glycerol. Cultures were harvested during their exponential phase of growth (OD > 1.0), washed and resuspended in phosphate buffer (pH=7). Serial dilutions (2-fold) were prepared using phosphate buffer. Plate counts, total cell counts and solid COD measurements were performed upon each of these dilutions in triplicate. Total cell counts was conducted according to the method described by Liss et al. (Liss et al., 1993). COD was analyzed using the method described earlier. 2.2.3 RESULTS AND DISCUSSION 2.2.3.1 Protein assays in the presence of resin acids and CTMP effluent. 55 Protein assays are commonly used to estimate microbial growth in liquid cultures (Donovan, 1969). They are superior to many other methods because the assays can be performed quickly and easily and the results usually correlate well with the weight or the number of microbial cells in the culture. Neither the CTMP effluent or the defined media used in this study contained significant amounts of protein. Thus, the protein assay was considered to have a good chance as a possible indicator to monitor the growth of resin acid degrading bacteria. Both the CTMP effluent and resin acids were examined for possible interference with the various protein assays. The correlation between the protein concentration and the absorbency measured by the different assays at different concentrations of B S A are shown in Fig. 13. It is apparent that the fluorescent protein assay was influenced by the CTMP effluent. Although the Lowry assay was influenced by both the CTMP effluent and resin acids, the regular Bio-Rad did not appear to be influenced by either CTMP effluent or resin acids. However, the biomass in the tested systems was not in sufficiently high enough concentrations to give any meaningful readings when measured by the regular Bio-Rad assay. When the effect of the CTMP effluent and resin acids on the micro Bio-Rad assay was determined, it also resulted in a non-linear response (data not shown). Thus, it was apparent that none of these three protein assays could be used to quantify microbial biomass in the presence of CTMP effluents or resin acids. 2.2.3.2 Comparison between the plate count, total cell count and solid COD measurements of microbial biomass. The plate count, microscopic total cell count and solid COD measurements were performed upon serial dilutions of cell suspensions of both the resin acid degrading bacteria B K M E 5 and B K M E 9. Regression was analyzed to determine the relative degree of correlation between the 56 -I 1 1 1 1 • 1 1 1 1 1— 0 200 400 600 800 1000 BSA concentration (mg/l) •1.6-, 0.0 i 1 1 . 1 1 1 1 1 1 1 r—i . 1 1 1 . | 1 -20 0 20 40 60 80 100 . 120 140 160 BSA concentration (mg/l) Fig. 13 Influence of CTMP effluent and resin acids on the Fluoraldehyde protein assay (A), the regular BIO-RAD assay (B) and the Lowry protein assay (C). 57 three methods. As shown in Fig. 14, both the plate count and the solid COD methods gave values which correlated well with the total cell count. Of the three methods evaluated, the plate counts appeared to be most reliable. It was straightforward and not influenced by either the CTMP effluent or the resin acid content. However, this method is time consuming and requires a significant amount of medium preparation. The use of total cell count is simple, quick and straightforward, However, as mentioned earlier, it can be influenced by particulates present in the CTMP effluent. Therefore, the total cell count method was used when a clear medium (defined medium or filtered CTMP effluent) was encountered and plate counts were employed as an alternative when media containing significant amount of particulates had to be assayed. Although the solid COD method gave a good approximation of microbial biomass, it should only be used when other parameters such as the total COD and the non-biomass COD content are already known. When these values are known, the biomass COD can be calculated using the following equation: biomass COD = total COD - non-biomass COD. Here, the biomass COD refers to the Solid COD mentioned above. Similarly, the non-biomass COD can also be calculated from the difference between the total COD and the biomass COD if the later can be calculated from either the plate counts or the total cell counts. In the work described in chapters 3 and 4,- the growth of the DHA-degrading bacteria was routinely monitored by total cell count when grown in the presence of filtered CTMP effluent and by plate count when grown in the presence of gamma-irradiated effluent. 58 • Plate counts • Solid COO 1000 800 600 S Q o o •g "5 H 400 200 (A) Total cell counts (x 107 cells/ml) (B) Fig. 14 Correlation between plate counts, solid COD and total cell counts for BKME 5 (A) and BKME 9 (B). Each dot represents the mean value of three replicates with the standard deviation less than 5%. 59 CHAPTER 3. GROWTH OF RESIN ACID DEGRADING BACTERIAL STRAINS B K M E 5 AND BKME 9 IN THE PRESENCE OF CTMP EFFLUENT 3.1 Background As mentioned in the general introduction, resin acid breakthroughs occasionally happen with biotreatment systems as a result of operation problems such as black liquor spills. In processes where the concentrations of wood extractives are higher (e.g. CTMP/thermomechanical (TMP) pulping), biological treatment systems can exhibit variations in resin acid removal which leads to the discharge of toxic effluents. Previous work (Bicho et al, 1994) has shown that the resin acid-degrading bacteria used in this study do not readily metabolize either simple sugars, acetate or methanol. As these compounds are usually relatively abundant in pulp and paper mill effluents, it could be assumed that resin degrading organisms have a fairly limited substrate range and that, as they appear to have a high substrate selectivity, it is probable that they only constitute a small portion of the microbial population in biological treatment systems. Therefore, if they are relatively low in numbers it could also be assumed that they may not be capable of withstanding and using the high concentrations of resin acids that may occur during toxicity failures such as shock loading during season variation or black liquor spills. As mentioned earlier, previous research on individual microorganisms has largely focused on strains isolated from soils and natural water systems. This work has also tended to focus on just the biodegradation of D H A as this resin acid is usually found at high concentrations in untreated pulp and paper mill effluents. Past organisms that have been studied include bacteria, such as Alcaligenes eutrophus (Biellmann et al, 1973), Flavobacterium resinvorum (Biellmann et al, 1973), Arthobacter sp. (Levison et al, 1968), Bacillus sp. (Hemingway and Greaves, 1972), Psuedomonas sp. (Biellmann et al, 1973) and the fungus Mortierela isabellina (Kutney et al, 1981). Recently 60 Bicho et al. (1994), Mohn (1995), Morgan and Wyndham (1996), and Wilson and Mohn (1996) have isolated various bacteria from pulp mill effluent treatment systems. Bicho et al. (1994) isolated five DHA-degrading bacterial strains from bleached Kraft mill effluent treatment systems. The five strains, designated B K M E 5, B K M E 6, B K M E 7, B K M E 9 and B K M E 11, were able to tolerate D H A concentrations as high as 100 ppm and these isolates appeared to use D H A as the sole carbon source. Three of these strains ( B K M E 5, B K M E 6 and B K M E 9) exhibited better degradation of D H A and other resin acids (both abietanes and pimaranes) commonly found in pulp effluents than the other strains. Mohn (1995) and Morgan and Wyndham (1996) recently isolated abietic acid and DHA-degrading bacterial strains from a batch reactor or pulp mill effluent, respectively. Wilson and Mohn (1996) also isolated some pimarane degrading bacterial strains from a batch reactor. Although this initial work established some growth and metabolic characteristics of the strains isolated from mill effluent systems, to date, none of these isolates have yet been evaluated in the presence of mill effluents. Therefore, very little is known about how these isolates will utilize resin acids when other mill effluent components are present. The bacterial strains that were used in this study were originally isolated from B K M E , where the concentration of resin acids are typically much lower than those found in a typical CTMP effluent. However, earlier work (Bicho et al, 1995) indicated that even though the strains were isolated from environments where low concentrations of resin acids are normally encountered, they were still able to withstand and utilize relatively high concentrations of DHA. Since there is currently no information regarding the occurrence or distribution of resin acid degrading bacteria in pulp mill effluent treatment systems, we chose to examine the ability of these strains to utilize D H A in CTMP effluents. The objectives of this part of the thesis were, (1) to examine the ability of two DHA-degrading bacterial strains to grow and degrade D H A present in or supplemented to CTMP effluent, (2) to determine if the 61 addition of nutrients could improve the degradation of D H A present in the CTMP effluent, (3) to investigate whether D H A is preferentially utilized when these strains are grown on CTMP effluents. 3. Materials and methods 3.2.1 Bacterial strains Two of the five bacterial strains previously described by Bicho et al. (1994, 1995) were used in this study. In these initial studies, B K M E 5 and B K M E 9 had shown a faster growth rate and higher degradation efficiencies of concentrated resin acids than the other three strains. Stock cultures of these strains were maintained on nutrient agar supplemented with 50 mg of D H A liter"1. The two strains were first grown on a defined medium containing glycerol as the sole carbon source. This final medium contained (per liter) 1.15 g Na 2 HP0 4 , 0.262 g NaH 2 P0 4 , 4.1 g N H 4 C I , 8.3 ml glycerol, and 1.7 g Bacto Yeast Nitrogen Base (YNB). There was no amino acids or ammonium added to this medium. The composition of this Y N B medium is shown in table 8. The cultures were incubated on a shaker (150 rpm) at 30 °C until the late exponential phase was reached and the centrifuged cells were then transferred to fresh medium. D H A was then added to the cultures when they were at the late exponential growth phase at a concentration of 100 mg/1. Twelve hours after D H A was added, the cells were washed by centrifugation and resuspended in phosphate buffer (pH = 7) to a final cell concentration of 4x l0 7 cells/ml. 3.2.2 Chemicals All the resin acids (high purity: 90-99%) including O-methyl podocarpic acid (surrogate standard) were purchased from Helix Biotech (Richmond, B.C., Canada). Methyl heneicosanoic acid (internal 62 standard) was purchased from Aldrich (Milwaukee, Wi., USA) and was also of high purity (99%). Stock solutions of D H A which were added to the media were prepared in 1 N NaOH. Table 8. The composition of Bacto Yeast Nitrogen Base medium without amino acids and ammonium Ingredients Amounts per liter Vitamins Biotin 2 ug Calcium Pantothenate 400 ug Folic Acid 2 ug Inositol 2000 ug Niacin 400 ug p-Aminobenzoic Acid, Difco 200 ug Pyridoxine Hydrochloride 400 ug Riboflavin 200 ug Thiamine Hydrochloride 400 ug Trace elements Boric acid 500 ug Copper Sulfate 40 ug Potassium Iodide 100 ug Ferric Chloride 200 ug Manganese Sulfate 400 ug Sodium Molybdate 200 ug Zinc Sulfate 400 ug Salts Potassium Phosphate Monobasic l g Magnesium Sulfate 0.5 g Sodium Chloride 0.1 g Calcium Chloride 0.1 g Amount of final medium from 100 g dehydrate medium 58.8 liters Final pH ± 0.2 at 25 °C 4.5 63 3.2.3 Preparation of sterilized CTMP effluent Untreated CTMP effluent was collected from the Quesnel River Pulp Mill and stored in a frozen state. The concentrations of COD, inorganic nutrients, resin acids and wood sugars were determined prior to storage. Filter-sterilized effluent was prepared by defrosting the effluent and filtering it once through a Whatman glass microfibre, twice through a 0.65 urn millipore filter, once through a 0.2 u.m millipore filter, and finally filtering the effluent aseptically through a 0.2 u.m millipore filter. Gamma-irradiated effluent was prepared using the following procedure: The CTMP effluent was defrosted prior to gamma irradiation. The sample was directly thawed and irradiated to avoid any possible consumption of effluent components by microorganisms in the unsterilized effluent. The gamma irradiation was conducted in the Biomedical Research Centre of UBC. Each 1 L CTMP effluent was irradiated for 5 days. Plate counts were performed before and after filtration and irradiation to ensure the effectiveness of both sterilization methods. 3.2.4 Growth on CTMP effluent alone Appropriate volumes of cell suspensions of the two strains described above were added in the proportions of 1:20 to (1) filter-sterilized CTMP effluent with and without added D H A (100 mg/L); and (2) irradiated effluent plus D H A (50 mgL"1). . The cultures were incubated at 30°C or room temperature (23 °C) with agitation (150 rpm) on a gyratory shaker for 96 h. These cultures were sampled throughout the duration of the experiment for, cell number, D H A concentration and COD 64 measurements in triplicate. Controls without bacterial inoculation were also assayed to determine if there was any D H A removal by abiotic degradation. 3.2.5 Growth on nutrient supplemented CTMP effluent Appropriate volumes of cell suspensions of the two strains described above were added in the same proportions to, (1) filter-sterilized CTMP effluent; (2) filter-sterilized CTMP effluent plus nutrients; (3) filter-sterilized CTMP effluent with added D H A and nutrients (The final D H A level was around 100 mg liter"1.); and (4) 100 mg D H A liter"1 plus nutrients. Similar cell suspensions were also added to irradiated effluent samples supplemented with both D H A and nutrients. The nutrients were used at the same concentration as reported earlier. The cultures were incubated at either 30 °C or room temperature (23 °C) for 96 h. Each flask was sampled throughout the experiment for, cell number, D H A concentration and COD measurement in triplicate. Controls without an added bacterial inoculation were established to ensure that there was no D H A removal by abiotic degradation. 3.2.6 Total cell counts Total microscopic cell counts using epifluorescent (DAPI) stain and an optical microscope was used as the indicator of bacterial growth. This method has been described earlier by Liss et al. (Liss et al. 1993). 3.2.7 Resin acid analysis Resin acids were quantified by the procedure of Kutney et al. (1981a). This method was used to extract resin acids and various hydroxylated derivatives from microbial systems, and the extract was then analyzed by GC using methods previously described by Bicho et al. (1994). 65 3.2.8 COD measurement Samples from the filter sterilized CTMP effluent were first passed through a Millipore Millex-GVi 3 (0.2 u.m retention) filter to remove any microbial biomass. The filtrate was then diluted with distilled water and analyzed for COD, according to the standard method described earlier. The gamma irradiated effluent samples were not filtered through Millipore filters prior to the COD test. However, their final COD readings were subtracted from the corresponding COD (solid COD) values generated by cells grown on the effluent to determine the relative values. The COD of the cells was calculated based on a the correlation between the solid COD and cell number (plate counts). 3.2.9 Nutrient analysis The original, filtered and gamma-irradiated CTMP effluents were tested immediately after thawing, filtration and radiation, respectively, for total Kjeldahl nitrogen, ammonia nitrogen, nitrite, nitrate, and total phosphorus. Al l of the parameters were analyzed based on the procedures previously described in the "British Columbia Environmental Laboratory Manual For the Analysis of Water, Wastewater, Sediment and Biological Materials" (1994 Edition), Province of British Columbia and "Standard Methods for the Examination of Water and Wastewater" 17 th Edition, 1989, published by the American Public Health Association. 3.2.10 Sugar analysis -An aliquot of 1 ml CTMP effluents was mixed with 1 ml 1 N H2SO4 in a sealed glass tube and incubated at 103 °C for 1.5 h. The mixture was then cooled and stored at 4 °C until analyzed. Each of these 2 ml effluent was spiked with 200 | i l internal standard (9.8 mg/ml fucose solution). Samples were then filtered through a Bond Elut C18 disposable solid phase extraction column (Varian, 66 Harbor City, CA, USA) to remove fatty and resin acids. These cleaned effluent hydrolytes were then filtered through a 0.45 u,m Millipore H V filter and analyzed for different monosaccharides using HPLC. Anion-exchange chromatography of monosaccharides was carried out on a CarboPac PA-1 column using a Dionex DX-500 HPLC system (Dionex, Sunnyvale, CA, USA) controlled by Peaknet 4.10 software. The column was equilibrated with 10 mM NaOH, and regenerated with 250 mM NaOH. After injecting 5 ul of sample using a SpectraSYSTEM AS3500 autoinjector (Spectra-Physics, Fremont, CA, USA), the sugars were eluted using deionzed water at a flow rate of 1 ml/min. The monosaccharides were monitored using a Dionex ED40 electrochemical detector (gold electrode), with parameters set for pulsed amperometric detection of sugars as recommended by the manufacturer. Postcolumn addition of NaOH before the detector was carried out by adding 0.5 M NaOH to the flow stream at a rate of 0.5 ml/min. 3.3 Results 3.3.1 Preparation of sterilized CTMP effluent It was important to ensure that the CTMP effluent was sterilized prior to use so that there was no possibility of any contribution from contaminating microorganisms already present in the effluent. Although autoclaving is the most widely used method for eliminating contaminating organisms, there was the distinct possibility that this method would dramatically change the composition of mill effluent since the volatile substances such as acetic acid and methanol in the effluent could evaporate. Similarly other substances could be changed chemically at that high temperature, for example, oligosaccharides in the effluent tended to be hydrolyzed. Therefore, we wanted to compare the relative merits of the filtration and gamma irradiation methods of CTMP effluent sterilization. It was apparent that both methods could effectively remove microorganisms from the effluent. A decrease of viable cell number from 2.02 x 106 to 0 ml"1 effluent was obtained after both 67 sterilization methods, and the viable cell number of both the filtered and irradiated effluents remained below 1 cell ml"1 after 5 day incubation at 30 °C. However, it was apparent that the composition of the CTMP effluent changed after both sterilization procedures (table 9, 10). The COD of the filter sterilized effluent dropped from 4085 ± 101 mg liter"1 to 2571± 67 mg liter'1 due to the retention of much of the particulate material onto the filters, while very little change in the COD occurred after gamma irradiation. Similarly, filtration resulted in more changes in the resin acid concentration of the effluent than did the irradiation. The total resin acid concentration of the irradiated effluent was reduced by 15% while the filter sterilized effluent was reduced by 47% when compared to the original effluent. However, the most noticeable trend was that the two sterilization methods resulted in substantial changes to the individual resin acids (table 10). All the resin acids other than D H A were removed to a greater or lesser extent after filtration. After gamma irradiation, the concentrations of abietic acid, neoabietic acid, palustrate, isopimaric acid, pimaric acid all decreased, palustrate was reduced below detection limits; sandaracopimaric acid remained at about the same level; while 7-oxo-DHA and D H A were significantly increased. As shown in tables 9, 10 and 11, both sterilization methods also resulted in minor changes in the inorganic components, the nutrient concentration and the sugar concentration of the effluent. This initial characterization indicated that the gamma irradiated effluent appeared to be more representative of the original effluent than did the filter-sterilized effluent. 3.3.2 Growth of B K M E 5 and 9 and degradation of DHA present in CTMP effluent. In order to evaluate the D H A degrading bacterial isolates in the presence of mill effluents, strains B K M E 5 and B K M E 9 were grown on sterilized CTMP effluent. We first evaluated the filter-sterilized effluent as this could be readily prepared. The growth of the bacteria on filtered CTMP effluent was confirmed by total cell counts (Fig. 15). The total number of B K M E 5 cells increased 68 from 2.3xl0 6 to 6.5xl0 7 ml"1 within 36 h, while for B K M E 9, the total cell number increased from 2.0xl0 6 to 1.6xl0 7 ml"1 after 9.5 h. After this time a stationary phase of growth was observed. The doubling time at exponential phase for B K M E 5 and 9 was 1.6 h and 1.9 h, respectively. However, no D H A degradation was observed. Table 9. Nitrogen and phosphorus concentration in original, filtered and gamma-irradiated CTMP effluents Nitrate Nitrite Ammonia Total Kjeldahl Total (mg f 1) (mg f 1) Nitrogen Nitrogen Phosphorus (mg l"1) (mg r 1) (mg F1) Original CTMP effluent <0.05 <0.01 <0.02 10.3 3.3 Gamma-irradiated effluent <0.05 <0.01 0.03 6.7 1.5 Filtered CTMP effluent <0.05 <0.01 <0.02 9.2 2.1 Table 10. Resin acid concentration in original, filtered and gamma-irradiated CTMP effluent Resin Acids Original CTMP effluent (mgf1) Filtered CTMP effluent (mgl"1) Gamma-irradiated CTMP effluent (mgf1) pimarate 3.5 0.9 2.6 sandarcopimarate 1.0 <0.5 1.1 isopimarate 4.7 1.0 4.3 palustrate 6.9 3.2 <0.5 D H A 12.6 12.3 21.0 abietic acid 8.8 3.9 5.6 neoabietic acid 6.9 <0.5 2.4 7-oxo-DHA 2.4 3.7 2.9 Total 46.8 25 39.9 Each value presented here represents the mean value of three replicates with the standard deviation less than 5%. 69 Table 11. Sugar concentration in original, filtered and gamma-irradiated CTMP effluent arabinose galactose glucose xylose mannose total sugar (mg/l) (mg/l) (mg/l) (mg/l) (mg/l) (mg/l) CTMP 140 175 75 35 90 515 gamma-irradiated CTMP 90 125 45 30 45 335 filtered CTMP . 105 170 65 35 85 460 As the filter-sterilized CTMP effluent only supported limited growth of the strains and the organisms did not use DHA, it was possible that (1) the filter-sterilized effluent lacked some essential substances required for bacterial growth and DHA degradation; (2) the effluent was inhibitory to the strains and limited both growth and the D H A degradation, (3) the utilization of other organic materials in the effluent inhibited DHA degradation by the strains. As mentioned earlier, filtration removed about 30% of the COD associated with the particulates in the effluent as well as 50% of the original resin acids. It was possible that sterilization of the effluent by filtration might have eliminated some of the essential nutrients required for the growth and DHA degradation by the strains. To test these various hypothesis the two isolates were therefore grown on the gamma irradiated effluent which was thought to be more representative of the original effluent. As the particulates in the irradiated effluent interfered with total cell count after epifluorescent stain, the plate count method was used to monitor bacterial growth. Growth of both strains on irradiated CTMP effluent supplemented with D H A was confirmed. The number of cells of B K M E 5 and B K M E 9 increased from 4.5xl0 6 to 2.9 xlO 7 ml"1, and from 2.8xl0 6 to 3.4xl0 7 ml"1, respectively, after 4 days incubation. However, no D H A degradation was found with either strain. 70 ( A) •A J • • — • 50 100 ( B) 50 n2000 1800 1600 C 1400 O O 3 1200 <Q 1000 Cultivation Time (h) C O D — • — D H A —•—Cel l Number 0 50 100 Cultivation Time (h) 800 Fig. 15. Growth (at 30 °C) and the measurement of DHA degradation and COD utilization by BKME 5 (A) and BKME 9 (B) on filtered CTMP effluent. Each dot represents the mean value of three replicates (standard deviation bars are shown). 71 3.3.3 Growth experiments on nutrient supplemented C T M P effluent As neither of the strains were able to degrade the DHA present in the gamma-irradiated effluent, the nutrient level of the CTMP effluent was compared with that of the defined medium which was previously described by Bicho et al. (1994). It was apparent that the concentrations of nitrogen and phosphorus, which are important nutrients for microbial growth, were much lower in the CTMP effluent than was present in the defined medium. Furthermore, the defined medium contained yeast nitrogen base which contains various trace metals and vitamins, which were also lacking in the effluent. Therefore, to assess the influence of these components, the isolates were grown on CTMP effluent supplemented with these nutrients to determine if they were required for effective DHA degradation. When nutrients were added, both the growth and DHA degradation of the filter sterilized effluent were enhanced significantly (Fig. 16). Within 48 h, the total cell number for B K M E 5 increased from 2.3xl0 6 to 7.4xl0 8 ml' 1 on filtered CTMP effluent plus nutrients, and from 1.7xl06 to 9.3 xlO 8 ml' 1 on filtered CTMP effluent plus nutrients which was further supplemented with D H A (Final concentration of D H A was around 100 ppm). The total cells of B K M E 9 increased from 1.6xl06to 1.7xl08 ml"1, and from 1.3xl06 to 2.3xl0 8 ml' 1 on these two media within 36 h, respectively. However, the growth rate on the nutrient supplemented effluent was similar to what was previously achieved on the unsupplemented CTMP effluent. Both strains showed the same doubling time during the exponential phase of growth on filtered CTMP effluent with added nutrients and on the same effluent further supplemented with added DHA. This was the same growth rate as that observed on CTMP effluent alone. It was apparent that the addition of nutrients greatly enhanced D H A degradation (Fig. 16). Complete utilization was observed at both low (5.2-8.9 ppm) and high (62.8-93.6 ppm) D H A concentrations and no D H A degradation occurred on controls without added 72 Fis 16 Growth of B K M E 5 (A) and B K M E 9 (B) at 30 °C and the measurement of DHA degradation by B K M E 5 (C) and B K M E 9 (D) on filtered CTMP effluent plus D H A and nutrients ( • f f S CTMP effluent plus nutrients (•) , filtered CTMP effluent alone ( A ) and DHARefined medium ( T ) Each dot represents the mean value of three replicates (standard deviation bars are ( shown) 73 inoculum. The D H A utilization rate by B K M E 5 and 9 during the exponential phase was 0.17 mgL' 'h' 1 and 0.54 mgL^h' 1, respectively, on CTMP effluent plus nutrients, and 1.95 n i g l / V and 1.16 mgL^h"1, respectively, on filtered CTMP effluent plus nutrients and added DHA. When nutrients were added, both growth and D H A degradation were also enhanced significantly on gamma-irradiated CTMP effluent compared with the values observed on CTMP effluent alone (Fig. 17). In general, the irradiated effluent supported more growth than the filter sterilized effluent. However, the D H A degradation pattern was observed to be similar regardless of the type of CTMP effluent used (Fig. 17). Almost all of the D H A in the media was consumed within 48 h by B K M E 5, and within 72 h by B K M E 9. The D H A utilization rate was similar to that observed on the defined medium containing D H A alone although a lag of D H A degradation was observed when effluent was present. No D H A degradation occurred on controls without inoculum. 3.3.4 Utilization of other organic substrates It was apparent that both strains could use organic matter other than D H A as substrates for growth. When grown on filtered CTMP effluent alone, the B K M E 5 consumed roughly 160 mgL"1 COD, while B K M E 9 consumed about 50 mgL' 1 COD (Fig. 15). When grown on gamma-irradiated CTMP effluent alone, the two strains used 1967 and 1632 mgL^COD, respectively, within the 4 day incubation. Since no corresponding D H A degradation was observed, it could be concluded that the strains used something other than D H A for growth. As shown in Fig. 18, when sufficient nutrients were added, the COD utilization was enhanced as well as the D H A degradation. Strain B K M E 5 utilized around 600 mgL"1 COD during the 4 day incubation period. Over the same time period with B K M E 9, 600 mgL"1 and 350 mgL'1 COD were 74 0 20 40 60 80 100 0 20 40 60 80 100 Time(h) Time(h) Fig. 17. Growth of BKME 5 (A) and BKME 9 (B) at 23 °C and the measurement of DHA degradation by BKME 5 (C) and BKME 9 (D) on gamma-irradiated CTMP effluent plus DHA and nutrients (•), filtered CTMP effluent plus DHA and nutrients (•), and DHA defined medium(A).Each dot represents the mean value of three replicates (standard deviation bars are shown). 75 D) E 2200-2000-1800-1600-Q g 1400-1 1200-1 1000 J 800 ( A ) T - i — 1 — i — ' — i — > — i — • — i — i — i — i 0 20 40 60 80 100 Time(h) 1 800 1600 1400 -I g 12 00 O 1000 800 A 60C ( B ) 0 20 40 60 80 100 Tim e(hrs) Fig. 18. COD utilization by B K M E 5 (A) and B K M E 9 (B) at 30 °C on filtered CTMP effluent plus D H A and nutrients (•), filtered CTMP effluent plus nutrients (•), and filtered CTMP effluent alone (A).Each dot represents the mean value of three replicates (standard deviation bars are shown). 76 removed, respectively, from filtered CTMP effluent plus added D H A and nutrients, and filtered CTMP plus nutrients. When the gamma-irradiated effluent plus both nutrients and D H A was used as the medium, roughly 1,500 mgL"1 COD was removed by the two strains within same period. Interestingly, there was a rapid reduction in the COD of this medium within the first 24 hours. After this initial fast response, the COD utilization rate was similar for both the irradiated and filter sterilized effluents (Fig. 19). As the addition of 100 mgL"1 D H A to the effluent accounted for about 277 mgL"1 of the observed COD, it was apparent that both strains were able to use other organic carbon present in the effluent. Sugar analysis was performed on the irradiated and filter sterilized CTMP effluent prior to and after degradation by the two strains (Table 12). It was apparent that the minor difference in sugar concentrations between the original and the fermented effluents could not support the conclusion that the strains consumed significant amount of the low level of sugars that were present. Table 12. Sugar concentration after 4 day incubation of B K M E 5 and 9 on gamma-irradiated CTMP effluent and filtered CTMP effluent. Arabinose galactose glucose xylose mannose total sugar . » ! ) . . W ) . (mg/1) .....(mg/1). (mg/1) gamma-irradiated 90 125 45 30 45 335 CTMP (control) gamma-irradiated CTMP 90 135 45 25 40 335 fermented by B K M E 5 gamma-irradiated CTMP 85 120 65 25 40 335 fermented by B K M E 9 filtered CTMP (control) 105 170 65 35 85 460 filtered CTMP fermented 85 140 55 25 65 370 by CTMP 5 filtered CTMP fermented 85 140 70 30 70 395 by CTMP 9 77 Q O O 40001 3500-3000-2500-2000-1500 6 20 40 60 80100 Time(h) Fig. 19. COD utilization at 23 °C by B K M E 5 (•) and 9 (•)on gamma-irradiated CTMP effluent plus D H A and nutrients, and by B K M E 5 ( A ) and 9 ( T ) on filtered CTMP effluent plus DHA and nutrients. Each dot represents the mean value of three replicates (standard deviation bars are shown). 78 3.4 Discussion The work in this chapter of the thesis indicated that strains B K M E 5 and 9 could degrade the DHA present in CTMP effluents. Although these strains were originally isolated from a B K M E , the fact that they could effectively remove D H A from CTMP effluents indicates they could be representative of the types of organisms that remove resin acids in aerobic systems. Although these DHA-degrading bacteria could grow on CTMP effluent alone, it was apparent that the CTMP effluent was lacking in nutrients and could not support optimal growth. The effluent also lacked the essential nutrients required for effective D H A degradation. It was observed that added nutrients not only enhanced D H A degradation, they also enhanced the degradation of the other abietanes present in the CTMP effluent (data not shown). Nutrient analysis (table 9) showed that nitrogen and phosphorus, which are the essential elements for bacterial growth, were all at relatively low levels in the original, filtered and gamma-irradiated CTMP effluents. Usually nitrogen and phosphorus are added to biological treatment systems in a mass ratio of BOD:N:P of 100:5:1. Assuming that the BOD of CTMP effluents is normally 3,000-5,000 mgL"1, the nitrogen added is then at a level of 150-250 mgL"1. Whether this is enough for effective resin acid degradation is still unknown. Furthermore, resin acid biodegradation may require some vitamins and metals that are at insufficient concentrations in the effluents. It appeared that a better understanding of the nutritional requirements for resin acid degradation would aid in managing resin acid removal in biological wastewater treatment systems. This was the objective of a further section of this thesis The initial work carried out in this chapter of the thesis indicated that the DHA-degrading bacterial strains were able to utilize other carbon sources present in the CTMP effluent. It was observed that, when grown on CTMP effluent alone, the cells used COD components other than DHA. When 79 nutrients were added to the effluent, both strains could use D H A and other components of the COD. The minor differences observed between growth on the D H A supplemented and unsupplemented filter sterilized CTMP effluents indicated that most of the growth could be attributable to the utilization of other components than the D H A present in the CTMP effluent. In fact, the 100 mgL"1 D H A in the filter-sterilized effluent supplemented with D H A accounted for only 277 mgL"1 of the total COD. It was apparent that most of the COD consumption was due to the utilization of other organic carbons, which led to the increased growth obtained compared with that achieved on D H A defined medium alone. The strains could consume up to 1,500 mgL"1 COD on gamma-irradiated effluent. The quick drop of COD within the first 24 h on irradiated effluent implied that a significant component of the easily biodegradable components was associated with the insoluble part of the effluent. However, in the soluble fraction of the effluent it was likely there were also other compounds that were easily degraded by these strains. Sugar analysis indicated that these strains used very little of the carbohydrates that were present in the CTMP effluent. This confirmed the results of the previous study (Bicho et al, 1994) which indicated the poor utilization of carbohydrates. It was found that the COD consumption tended to stop after only 20% of the total COD of the filtered effluent and 40% of the COD of the irradiated effluent was consumed. It is possible that part of the unutilized COD might be composed of acetate, methanol and sugars that are abundant in pulping effluents (Suntio et al, 1988) and are not utilized by these isolates (Bicho et al, 1995). As these DHA-degrading bacteria were able to utilize other organic carbon in the effluent for growth, it is possible that these organisms might constitute a certain portion of the microbial population in biological wastewater treatment systems. Thus, it is possible that the resin acid breakthroughs that occur in some mills may not be due to just a load shock as we proposed earlier, other factors such as the dramatic change in the treatment condition, e.g. pH and temperature, might also account for these toxicity failures. 80 By comparing D H A degradation curves with COD consumption curves, it was apparent that DHA was not the preferred carbon source for these isolated DHA-degrading strains when complex effluent components were present. The strains tended to use some other carbon source first before utilizing DHA. The enhanced growth caused by the utilization of other organics in the effluent did not stimulate D H A degradation by these strains. This suggested that in situations of concentrated effluents, some types of pretreatment (e.g., anaerobic treatment) might be desirable to reduce BOD to a certain extent before effective resin acid removal could occur. As filter sterilization removed about 1/3 of COD from the CTMP effluent and changed the profile of resin acids, it was considered to not be representative of the original effluent. Gamma irradiation, which resulted in little change in the total COD and minor differences in the resin acid profile, appeared to be more representative. It appears as though filtration affected the resin acids physically, while gamma irradiation affected resin acids chemically. It was noticed that all the resin acids were removed to a certain extent after filtration except for the DHA. This might be because the other resin acids are more readily adsorbed onto the particles in the effluent. Previous research (Patoine et al, 1994) showed similar results when various resin acids were filtered. However, gamma irradiation appeared to increase D H A level in the effluent while decreasing the level of abietic acid, neoabietic acid, and palustrate. It is probable that these abietanes were degraded to D H A during irradiation. Neither filtration or irradiation changed the inorganic nutrient concentration dramatically. As there was a proportional drop in the total Kjeldahl nitrogen, total phosphorus, total sugars and the minerals in the irradiated effluent, we suspect that this might be due to an unrecorded dilution of the irradiated effluent during these analyses which were carried out by the CanTest Ltd., Van, BC. The much lower concentration of ammonia nitrogen compared with the total Kjeldahl nitrogen suggested that the nitrogen in the effluent was mainly in the form of organic nitrogen. 81 CHAPTER 4. NUTRITIONAL CHARACTERIZATION OF THE RESIN ACID DEGRADING BACTERIA 4.1 Background As described in the previous chapter, earlier work had shown that the CTMP effluent was nutrient limited and that the D H A degrading bacteria could not readily degrade the D H A present in the CTMP effluent without the addition of nutrients. In the work described earlier by Bicho et al. (1995) and in the work carried out in the earlier chapters of the thesis, the nutrients that were added to the effluent were part of a complete supplementation which included all the essential nutrients required for bacterial growth. This included nitrogen (NH4CI), phosphorus (P0 4 3"), various vitamins and trace metals. However, it would not be practical or affordable for pulp effluent treatment plants to routinely add this complete nutrient mix to ensure efficient resin acid degradation. It is probable that only part of these nutrients were essential for DHA degradation and that other components of the enriched supplements added in the earlier work were not required for either the growth or the DHA degradation by these particular bacterial strains. It is also possible that the CTMP effluent already contains some of the essential nutrients and that only those components that are lacking in the CTMP effluent need to be added to enhance D H A degradation. Thus, it is probable that the identification of the essential nutrients required for resin acid degradation would help us to better manage pulp and paper effluent treatment. As very little work has been conducted on the nutritional requirements for resin acid biodegradation, the objectives of the work described in this part of the thesis were, to identify the essential nutrients required for D H A degradation and to determine the minimum amounts of these essential nutrients required for effective D H A degradation. 4.2 Materials and methods 4.2.1 To assess the impact of nutrients on DHA degradation. 82 The filter sterilized effluent was used as the source effluent to study the nutritional requirements for D H A degradation. Strains B K M E 5 and 9 were grown on the filtered CTMP effluents supplemented with (1) no nutrients; (2) nitrogen and Bacto Yeast Nitrogen Base (YNB) without added amino acids and ammonium (containing minerals, vitamins and phosphorus); (3) nitrogen and phosphorus; (4) Y N B without amino acids and ammonium; (5) nitrogen only; and (6) phosphorus only. The nitrogen was added in the form of NH 4C1 and the phosphorus in the form of P 0 4 ". The final concentrations of these nutrients were: N a 2 H P 0 4 1.15 g/1, N a H 2 P 0 4 0.262 g/1, NH 4C1 4.1 g/1, and Y N B without amino acids and ammonium at 1.7g/l. The inoculation, cultivation times and growth conditions were the same as those reported earlier. The growth and D H A concentration were measured in triplicate at the beginning and the end of the growth period. A control without inoculated cells was established to exclude any degradation by abiotic factors. 4.2.2 Mineral analysis of effluents Mercury analysis was performed on the original, filtered and irradiated CTMP effluents using procedures based on the Standard Methods for the Examination of Water and Wastewater section 3112B, 17 th Edition, 1989, published by the American Public Health Association. Other minerals were analyzed using Inductively Coupled Plasma Spectroscopy (ICP), Graphite Furnace Atomic Absorption or ICP/Mass Spectroscopy (ICP/MS). 4.2.3 To assess the need for the minerals present in YNB for effective DHA degradation by B K M E 9 Strains B K M E 9 was grown on (1) the filter sterilized CTMP effluent alone; or containing the following added nutrients: (2) nitrogen and Y N B (containing minerals, vitamins and phosphorus); and (3) nitrogen, phosphorus and the minerals Zn, Mo, Mn, Fe, Cu and Mg. The minerals were added in the form of their salts at the equivalent concentration. The concentrations of N H 4 + , P0 4 3", and Y N B used were the same as those described earlier. The inoculation, cultivation times and growth conditions were also the same as those reported earlier. Growth and D H A concentration 83 were measured m triplicate at the beginning and the end of the growth period. A control without inoculated cells was established to exclude any degradation by abiotic factors. 4.2.4 Identification of the essential minerals required for DHA degradation by BKME 9 Strain B K M E 9 was grown on the filtered CTMP effluent alone (1) or containing the following added nutrients: (2) nitrogen and phosphorus; (3) nitrogen, phosphorus and minerals Zn, Mo, Mn, Fe and Cu; (4) nitrogen, phosphorus and minerals Zn, Mo, Mn, Fe and Mg; (5) nitrogen, phosphorus and minerals Zn, Mo, Mn, Cu and Mg; (6) nitrogen, phosphorus and minerals Zn, Mn, Fe, Cu and Mg; (7) nitrogen, phosphorus and minerals Zn, Mo, Fe, Cu and Mg; (8) nitrogen, phosphorus and minerals Mo, Mn, Fe, Cu, and Mg; (9) nitrogen, phosphorus and minerals Zn, Mo, Mn, Fe, Cu and Mg; and (10) nitrogen and Y N B (containing minerals, vitamins and phosphorus). The concentration of the nutrients added were the same as those described earlier. The inoculation, cultivation times and growth conditions that were also the same as those used previously. Growth and D H A concentration were measured in triplicate at the beginning and the end of the cultivation. A control without inoculated cells was established to exclude any degradation by abiotic factors. 4.2.5 Study of the minimum nitrogen requirement for DHA degradation Strain B K M E 5 was grown on the filtered CTMP effluent containing nitrogen (NH4CI) at the following concentrations: (1) 0 ppm; (2) 50 ppm; (3) 100 ppm; (4) 1000 ppm and (5) 4000 ppm. Strain B K M E 9 was grown on the filtered CTMP effluent supplemented with nitrogen at the same concentrations as above, and with phosphorus and Y N B at the concentrations described earlier. The inoculation, cultivation times and growth conditions that were used were the same as those described earlier. Growth and D H A concentration were measured in triplicate at the beginning and end of the cultivation. A control without inoculated cells was established to exclude any degradation by abiotic factors. 84 4.3 Results 4.3.1 Impact of nutrients on DHA degradation. As described earlier in the thesis, B K M E 5 and 9 could only degrade the D H A present in the CTMP effluent when nutrients were added. To first assess the role of nitrogen and phosphorus in D H A degradation, different combinations of nitrogen, phosphorus and Y N B were added to the filter sterilized CTMP effluent. It was found that, although some growth was obtained on all the media, the maximum growth of both strains was obtained when the filter sterilized CTMP effluent was supplemented with nitrogen and Y N B containing phosphorus, minerals and vitamins (Fig. 20). In the case of B K M E 5 (Fig. 20A), all the individual nutrients (nitrogen, phosphorus, YNB) enhanced the growth to some extent; while with B K M E 9 (Fig. 20B), only media supplemented with nitrogen (nitrogen alone and nitrogen plus phosphorus) supported more growth than the unsupplemented filtered CTMP effluent. Media without added nitrogen (i.e., supplemented with only Y N B or phosphorus) did not give any additional growth. Although strain B K M E 5 did not require additional nitrogen for growth, it did require supplemental nitrogen before efficient D H A degradation occurred. In contrast, strain B K M E 9, which showed little improvement in growth with the addition of any of the nutrients, required supplementation with both nitrogen and Y N B before any D H A degradation occurred. It was apparent that nitrogen was the essential nutrients for the D H A degradation by B K M E 5, and that, D H A degradation by B K M E 9 could only occur on complete medium containing N , P, vitamins and minerals. 4.3.2 Influence of minerals on growth and DHA degradation As it has been shown that the degradation of D H A in the CTMP effluent by strain B K M E 9 required the addition of Y N B . The lack of D H A degradation could be due to the omission of some essential minerals supplied by the Y N B . Therefore, mineral analysis was performed on the original, filtered and irradiated CTMP effluent (table 13). It was found that the difference in the metal concentrations among the three effluents was minor. Although there was a serial and proportional decrease in the 85 E "cu c o t o i _ -#—> c CD O O O "55 O t i M cell growth DHA removal -i 100 (A) c O cu > o E 100 (B) Nutrient3combtnations Added nutrients: 1. None 2. N+YNB(minerals+vitamins+P) 3. N+P 4. YNB(mJnerals+vitamins+P) 5. N 6. P Fig. 20. Growth and D H A degradation by B K M E 5 (A) and B K M E 9 (B) on D H A spiked filter sterilized CTMP effluent supplemented with different nutrient combinations. The figure shows the cell yield and D H A removal efficiency after 96 h cultivation. Each bar represents the mean value of three replicates with the standard deviation less than 5% of the mean value. 86 metal concentrations (table 13) and nutrient concentrations (table 9) of the gamma-irradiated effluent, this was probably due to an error in the preparation of the irradiated effluent sample. By comparing the composition of the CTMP effluent and the Y N B employed in the complete medium, it was found that, other than Fe, Mn, Mo, Cu, Mg and Zn, the concentrations of minerals in the CTMP effluent were above those of the complete medium. This indicated that the absence of DHA degradation by B K M E 9 grown on CTMP effluent supplemented with nitrogen and phosphorus could possibly be due to the lack of these six minerals in the effluent. The need for vitamins in Y N B could also count for the absence of D H A degradation. In order to determine which of the minerals or vitamins in the Y N B are essential for DHA degradation, strain B K M E 9 was grown on filter sterilized CTMP effluent alone, or on the effluent supplemented with either complete nutrients or nitrogen, phosphorus, Zn, Mn, Mo, Fe, Cu and Mg. It was found that, although growth occurred on all the media, enhanced growth was obtained on the CTMP effluent supplemented with nutrients (Fig. 21). B K M E 9 could degrade DHA in both the complete medium and the CTMP effluent supplemented with nitrogen, phosphorus and the six minerals (Fig. 21). This indicated that nitrogen, phosphorus and some of the six metals were essential for D H A degradation by B K M E 9 grown on the CTMP effluent. In order to identify which metal is essential for D H A degradation by B K M E 9, the strain was grown on filtered CTMP effluent containing nitrogen, phosphorus, and various metal combinations. It was found that, although growth was obtained on all of the media, D H A degradation only occurred on the CTMP effluent containing added nitrogen, phosphorus and magnesium (Fig. 22). When magnesium was added to the CTMP effluent as well as nitrogen and phosphorus, strain B K M E 9 could effectively degrade DHA. Without added magnesium, the strain used other substrates to support limited growth. It was concluded that magnesium was a key element in D H A degradation by B K M E 9. 87 Table 13. Mineral concentrations in original, filtered and gamma-irradiated CTMP effluents Metals CTMP Gamma-irradiated Filtered CTMP Units effluent CTMP effluent effluent Aluminum A l <0.15 <0.15 <0.15 mg/1 Antimony Sb <0.15 <0.15 <0.15 mg/1 Arsenic As <0.3 <0.3 <0.3 mg/1 Barium Ba 0.19 0.10 0.15 mg/1 Beryllium Be <0.003 <0.003 <0.003 mg/1 Boron B 0.54 0.32 0.51 mg/1 Cadmium Cd <0.025 <0.025 <0.025 mg/1 Calcium Ca 118 68.2 108 mg/1 Chromium Cr <0.03 <0.03 <0.03 mg/1 Cobalt Co <0.02 <0.02 <0.02 mg/1 Copper Cu 0.065 0.36 0.27 mg/1 Iron Fe 0.19 0.11 0.16 mg/1 Lead Pb <0.08 <0.08 <0.08 mg/1 Magnesium Mg 9.74 5.65 9.28 mg/1 Manganese M n 4.30 2.48 4.08 mg/1 Mercury Hg <0.05 <0.05 0.19 ng/i Molybdenum Mo <0.04 <0.04 <0.04 mg/1 Nickel N i <0.025 <0.025 <0.025 mg/1 Phosphorus P 0 4 5.95 3.33 4.80 mg/1 Potassium K 33.0 19.3 32.2 mg/1 Silicon S i0 2 26.1 13.1 11.5 mg/1 Silver Ag <0.03 <0.03 <0.03 mg/1 Sodium Na 432 252 411 mg/1 Strontium Sr 0.24 0.13 0.23 mg/1 Tin Sn <0.03 <0.03 <0.03 mg/1 Titanium Ti <0.006 <0.006 <0.006 mg/1 Vanadium V <0.01 <0.01 <0.01 mg/1 Zinc Zn 0.26 0.12 0.13 mg/1 Zirconium Zr <0.015 <0.015 <0.015 mg/1 Nutrient combinations Added nutrients: 1. Filtered CTMP only 2. Filtered CTMP + NH4CI + YNB 3. Filtered CTMP + NH4CI + P04 + Zn + Mo + Mn + Fe + Cu + Mg Fig 21. Growth (A) D H A degradation (B) by strain B K M E 9 on media containing different nutrient combinations. Each bar represents the mean value of three replicates (standard deviation bars are shown). Nutrients added: 1 time 0 1. CTMP alone after 4-day incubation 2. CTMP+N+P 3. CTMP+N+P+Zn+Mo+Mn+Fe+Cu 4. CTMP+N+P+Mg+Zn+Mn+Mo+Fe 5. CTMP+N+P+Cu+Mg+Zn+Mo+Fe 6. CTMP+N+P+Mn+Fe+Cu+Mg+Zn 7. CTMP+N+P+Fe+Cu+Mg+Zn+Mo 8. CTMP+N+P+Mo+Mn+Fe+Cu+Mg 9. CTMP+N+P+Mo+Mn+Zn+Cu+Mg+Fe 10. CTMP+N+P+YNB Fig 22. Growth (A) D H A degradation (B) K M E 9 on media containing different minerals. Each bar represents the mean value of three replicates (standard deviation bars are shown). 90 4.3.3 Minimum nitrogen requirement for effective DHA degradation This work indicated that nitrogen was an essential nutrient required for effective D H A degradation. In order to determine the minimum amount of nitrogen required for DHA degradation, both B K M E 5 and 9 strains were grown on CTMP effluent containing various levels of added nitrogen. It was apparent that 50 ppm of added NH4CI resulted in twice as much of growth of B K M E 5 compared to growth obtained on medium without added nitrogen (Fig. 22A). In contrast, when using strain B K M E 9, at least 1000 ppm NH4CI was required to give a significant increase in growth (Fig. 23C). D H A degradation by B K M E 5 occurred when 50 ppm NH4CI was added and it was completed when at least 100 ppm NH4CI was added (Fig. 23B). To initiate D H A degradation by B K M E 9, 50 ppm NH4CI was required, while to complete the DHA degradation, 1000-4000 ppm NH4CI was needed (Fig. 23D). It appeared that, in order to ensure efficient degradation of D H A by both strains, CTMP effluent needs to be supplemented with at least 50 ppm NH4CI. 4.4 Discussion Waste waters derived from wood processing operations such as pulp mills, are usually rich in carbon which is derived from the cellulose and hemicellulose components, while limited in nitrogen and components such as phosphorus, minerals and vitamins. In industrial wastewater treatment processes, nitrogen and phosphorus are normally added to high rate biological treatment systems (e.g., activated sludge systems) in a mass ratio of B O D 5 : N : P of 100 : 5 : 1, based on experience coupled with an elemental analysis of the mixed population of organisms (Springer, 1993). For lower rate systems, such as the ASB where significant endogenous respiration takes place, some of the nutrients are recycled so less nutrients are required. The mass ratio of BOD:N varies from 100:1 to 50:1 depending on the hydraulic retention times used (Springer, 1993). Assuming that the BOD of the CTMP effluent is 4000 mg/1, the added nitrogen would be at a level of 200 mg/1 for activated sludge systems, and 40-80 mg/1 for aerated stabilization basins. The work described in this thesis shows that D H A degradation by B K M E 5 occurred when 12.5 mg/1 N (50 ppm NH4CI) was added, and it was completed when at least 25 mg/1 N (100 ppm NH4CI) was added. To initiate DHA degradation by B K M E 9, 12.5 mg/1 of additional N (50 ppm NH4CI) was required, while an 91 (A) c 8 c 8 < x Q Nutrient combinations Nutrient combinations timeO After 4 days incubation (B) Added nutrients: 1. Filtered CTMP only 2. Filtered CTMP + NH4CI 50ppm 3. Filtered CTMP + NH4CI 100 ppm 4. Filtered CTMP + NH4CI 1000ppm .^Filtered CTMP + NH4CI 4000ppm (C) timeO After 4 days incubation (D) Added nutrients: Filtered CTMP only + YNB Filtered CTMP + YNB + NH4CI 50ppm Filtered CTMP + YNB + NH4CI 100 ppm Filtered CTMP + YNB + NH4CI 1000ppm iltered CTMP + YNB + NH4CI 4000?^ Fig. 23. Growth (A & C) and D H A degradation (B & D) by B K M E 5 and B K M E 9 respectively on filter sterilized CTMP effluent with increasing concentrations of NH4CI . Each bar represents the mean value of three replicates (standard deviation bars are shown). 92 additional 1000 mg/l N (4000 ppm NH4CI) was needed to complete utilization. It can be concluded that, if nitrogen is added to wastewater treatment at the normal levels mentioned above, the occurrence of D H A degradation should be guaranteed. However, the completion of D H A degradation depends on whether B K M E 5 or other resin acid degrading organisms that require less nutrients to oxidize D H A play an important role in D H A degradation within the treatment systems. If strains like " B K M E 9" outnumbers strains typified by strain " B K M E 5" in treatment systems, the treatment systems may require more nitrogen for complete D H A degradation. The trace metals required for microbial growth are assumed to be present in the wastewater. However, pulp and paper mill effluents do not necessarily contain all the trace metals required for bacterial growth or resin acid degradation. Even B K M E effluents, which usually contain relatively high concentrations of minerals because of the chemicals used during pulping, bleaching and papermaking processes, or resulting from corrosion of equipment, may still be nutrient deficient. The principal minerals in B K M E effluent are Na, Ca, Mg, A l , Cr, Co, Ni, Ti, Fe, Hg and Zn (Springer, 1992). Some of these metals (e.g., Na, Ca, Mg, Cr, Co, Fe and Zn) are essential for microbial growth. However, the chemicals containing these metals are likely to be precipitated and normally low in the final effluent (Springer, 1992), and the final concentrations of minerals present may not be sufficient to support resin acid degradation in the effluent. The work carried out in this thesis showed that magnesium played a critical role in D H A degradation by B K M E 9 strain. When magnesium, nitrogen and phosphorus were added to the CTMP effluent, strain B K M E 9 could effectively degrade DHA. In the absence of added magnesium, the strain could only support limited growth and no D H A degradation occurred. This confirmed previous reports ( Biellmann and Branlant, 1973) that magnesium and iron were key elements in resin acid degradation. The present study concluded that D H A degradation by B K M E 9 did not occur at the current level (5.65 mg/l) of magnesium detected in this CTMP effluent sample (table 13). An additional 100 mg/l Mg (500 mg/l MgS0 4 ) had to be added to the CTMP effluent to ensure effective D H A degradation. However, this does not necessarily mean that Mg has to be added to all pulp and paper effluent biological treatment systems if D H A degradation is expected. When Bullock et al. (1994) examined the 93 concentration of minerals in the B K M E effluent where B K M E 9 was originally isolated, they found that the concentration of Mg in B K M E was only 4.58 mg/1. One possible explanation for the ability of strain B K M E 9 to grow and degrade D H A at this low level of Mg is that, in real wastewater treatment systems, the strain obtained most of its Mg from the decomposition of other organisms. In low rate biological treatment systems such as ASB where significant endogenous respiration takes place, Mg may be recycled so that no magnesium needs to be added. However, for high rate treatment systems such as activated sludge systems, it would be preferable to maintain the magnesium concentration at a relatively high level so that D H A degradation can be completed. It was also apparent during this work that the nutrients added to the CTMP enhanced both the DHA degradation, and the degradation of other abietanes in the CTMP effluent (data not shown). Nitrogen, the second most abundant element in cells, usually constitutes 12-15% of the cell dry weight (Brock and Madigan, 1991). It is a major constituent of proteins and nucleic acids. The bulk of the available nitrogen in nature is found in an inorganic form, either as ammonia or nitrate. Most bacteria use ammonia as their sole nitrogen source although many can also use nitrate (Brock and Madigan, 1991). With the help of several enzymes such as glutamate dehydrogenase and transaminases, ammonia is generally incorporated into the carbon skeletons of amino acids from which proteins can finally be synthesized (Brock and Madigan, 1991). As all microbial metabolic reactions require specific enzymes to catalyze the various enzymatic reactions, it is not surprising that D H A degradation is highly dependent on the availability of nitrogen. It is interesting to note that without the addition of nitrogen to the CTMP effluent, both the B K M E 5 and 9 strains could still use alternative carbon sources to achieve limited growth, although D H A degradation did not occur. It is possible that the enzymes needed for utilization of DHA, which had been shown to be inducible by Bicho et al. (1994), were subject to catabolite repression, and that other, more readily utilized carbon sources present in the CTMP effluent repressed the transcription of these enzymes from their mRNA. Thus, it is possible that effective growth could be achieved by use of the small amount nitrogen available in the CTMP effluent. If no additional nitrogen was supplied to the effluent, the nitrogen originally present in the effluent was completely consumed before these more 94 readily catabolized carbon sources were used up. Therefore, D H A degradation could not be initiated. However, if enough nitrogen and other essential nutrients were added to the effluent, the enzymes needed for D H A degradation were synthesized and, after the repressing carbon sources were consumed, D H A degradation then occurred. Catabolite repression was observed when either B K M E 5 or 9 strain was grown on CTMP effluent at a low temperature (23°C) (Fig. 17 & 19). At higher temperatures (30 °C), growth was accelerated such that the diauxic nature of the subsequent utilization of the repressing carbon sources and D H A was not noticable when substrate utilization was monitored. However, it should be noted that in real effluent treatment systems where the wastewater is continuously fed, the utilization of complex effluent components by microorganisms might not exhibit such a diauxic feature. Phosphorus was also shown to be an important nutrient necessary for cell growth. It is primarily used by the cell for synthesis of nucleic acids and phospholipids. It is usually utilized in the form of inorganic phosphate. The work presented in this part of the thesis indicated that the degradation of D H A by strain B K M E 9 required relatively high levels of phosphorus. It is interesting to note that this strain also needed high concentrations of magnesium to ensure efficient DHA degradation. Magnesium usually functions to stabilize ribosomes, cell membranes, and nucleic acids, and it is also required for the activity of many enzymes, especially those involving phosphate transfer (Brock and Madigan, 1991). It has also been reported that magnesium is essential for the normal cell division of bacilli and that under conditions of magnesium deficiency,, cell division is inhibited. However, this preliminary study was not able to determine whether magnesium was involved directly in D H A degradation by B K M E 9 strain, or it was simply required by this strain in a relatively large amount in the overall cell growth. 95 CHAPTER 5. GENERAL CONCLUSIONS In this study we confirmed that bacterial strains previously isolated from pulp mill effluent treatment systems were able to grow and utilize resin acids present in mill effluents. A solid COD method was developed to quantify the biomass in growth experiments as well as monitor microbial growth in industrial wastewater treatment systems. Although this method was based on a similar principal to the VSS determination, which uses chemical oxygen demand instead of gravimetric methods to determine the quantity of suspended organic material, it proved to be superior to the VSS protocol in terms of rapid analysis time, greater sensitivity and the use of smaller sample volumes. We demonstrated that the solid COD method was a useful substitute for VSS measurements when determining microbial biomass at different stages of growth. However, this method should not be used with effluents containing high levels of non-biomass solids. This method provided a useful research tool for use in suspended growth experiments in wastewater treatment procedures when sample volume and /or time is limited. Both the gamma irradiation and the filter sterilization proved to be effective methods of removing any contaminating organisms present in the CTMP effluent. Gamma irradiation was shown to be a preferred method of sterilizing CTMP effluent than filter sterilization as it resulted in little change in the total COD and minor differences in the resin acid concentration and profile. It appears as though filtration affected the resin acids physically, while gamma irradiation affected resin acids chemically. Neither method changed the inorganic nutrient concentration significantly. When two of the D H A degrading bacterial strains ( B K M E 5 and B K M E 9) described previously by Bicho et al. (1994) were grown on CTMP effluent, it was apparent that resin acids were not the preferred substrates of these strains when other organic material within the CTMP effluent were 96 available. The strains were able to utilize other sources of organic carbon and only degraded DHA after these substrates were consumed. It appeared that some form of catabolite repression limited the initial utilization of resin acids, although more precisely designed experiments are needed to confirm this conclusion. The portion of the COD degraded by these strains was found to be mostly associated with the insoluble part of the effluent. The temperature at which biotreatment systems are currently operated was identified as a possible issue, since more significant catabolite repression was observed at lower (23 °C) rather than at higher (30 °C) temperatures. This work showed that the CTMP effluent used in this study was nutrient deficient in terms of DHA degradation by isolates B K M E 5 and B K M E 9. To ensure effective D H A removal, CTMP effluent needed to be supplemented with nitrogen, phosphorus and some minerals. Nitrogen, in the form of ammonium, played a critical role in D H A degradation by strain B K M E 5. Without added nitrogen, D H A degradation did not occur although B K M E 5 could use other carbon sources in the effluent for cell growth. D H A degradation by B K M E 9 required the addition of phosphorus and magnesium, as well as nitrogen. Approximately 12.5 mg/l of nitrogen had to be added to the CTMP effluent for effective D H A degradation to occur for both strains. To obtain complete D H A degradation, an additional 25 mg/l of nitrogen for B K M E 5 and 1000 mg/l of nitrogen for B K M E 9 had to be added. This indicated that, if nitrogen is added to wastewater treatment systems at the normal levels (BOD 5 : N is 100:1 for activated sludge systems and 50:1 for ASB), D H A degradation by these types of microorganisms should be guaranteed. However, this preliminary study was unable to determine whether the amount of nitrogen and phosphorus normally added to mill effluent by treatment plants was sufficient for optimal D H A degradation. The preliminary results in this thesis also indicated that pulp mill treatment plants should routinely monitor the levels of some minerals, especially magnesium, in their effluents. Supplementation of magnesium to mill effluent is sometimes essential, 97 as this element may be needed in a relatively large amount for bacterial growth as mill effluents are relatively low in magnesium. As these D H A degrading bacteria were able to utilize other organic carbon in the effluent for growth, it is probable that they might constitute a certain portion of the microbial population in biological wastewater treatment systems. Furthermore, as these strains could effectively degrade high concentrations of DHA, it is also probable that they were able to tolerate the wide change in resin acid concentration that can occur in some wastewater treatment systems due to seasonal, operational or wood furnish changes. Although this work was not able to clearly identify how much resin acid removal was due to organisms such as B K M E 5 and 9, it is possible that toxicity breakthroughs that sometimes occur due to black liquor spills or seasonal variations restrict the growth of other microorganisms, while the resin acid tolerant microorganisms such as B K M E 5 and 9 are initially present in insufficient numbers to effectively degrade the high resin acid loading. However, within a few days, most aerobic systems can recover, probably due to the active growth of strains such as B K M E 5 and 9. Future work should assess this hypothesis to see if these types of organisms have a major role in the removal operation and recovery of wastewater treatment systems. The major conclusions of this thesis were based on the study of two D H A degrading bacterial strains which were isolated from a B K M E treatment system by Bicho et al. (1994). However, it appeared that these strains differed in their substrate specificity from other isolates obtained by other researchers (Mohn, 1995; Morgan and Wyndman, 1996; Wilson et al., 1996). For example, the strains we used were only able to grow on abietanes, not pimaranes. They were also unable to grow on either acetate or methanol, nor could they metabolize sugars very well. It appeared that some 98 other organic material present in the CTMP effluent inhibited the degradation of DHA by these strains. Some of the. isolates described by other researchers have shown a broader substrate specificity (Morgan and Wyndman, 1996; Wilson and Mohn, 1996) and could grow on sugars and some carboxylic acids such as pyruvate. Thus, it is apparent that the conclusions drawn from this present study may not be necessarily true for these isolates or for any of the other resin acid degrading organisms that may be present in wastewater treatment systems. It is probable that the mechanisms used for the metabolism of pimaranes will be somewhat different from those of abietanes. Although it has been shown that strains B K M E 5 and B K M E 9 differed in their nutritional requirements for D H A degradation, the distribution of these bacteria in wastewater treatment systems is still unknown. The further study of these strains or other isolates in real treatment systems should help to better estimate the performance and nutritional requirements of treatment systems. In this study, one CTMP effluent was studied for the availability of nutrients for resin acid degradation by two bacterial isolates. Other types of mill effluent will probably demonstrate different characteristics, thus it would be worthwhile to study the degradation of resin acid present in Kraft effluent or sulfide pulping effluents since these effluents are also generated in large volumes. The information obtained from this study should help better manage the resin acid degradation in waste water treatment systems. For example, it has been suggested that, in the near future, many wastewater treatment systems will be operated at higher temperatures. This should enhance resin acid degradation, and it might even be worthwhile to design multistage treatment systems to pre-remove some carbon sources so that subsequent resin acid degradation is enhanced. The preliminary results in this thesis also suggested that pulp mill treatment plants should routinely monitor the levels of nitrogen, phosphorus and some minerals, especially magnesium, in their effluents. 99 Supplementation of these nutrients to mill effluent is essential, as they appear to be needed in a relatively large amount for bacterial growth as mill effluents are relatively low in these nutrients. 100 REFERENCES APHA et al. 1985. Standard methods for the examination of water and wastewater, 16 edition. 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