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A physiological model to predict xenobiotic concentrations in fishes Yang, Rong 1997

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A P H Y S I O L O G I C A L M O D E L T O P R E D I C T X E N O B I O T I C C O N C E N T R A T I O N S IN FISHES By R O N G Y A N G B. S c., Nanj ing University, 1988 M . S c., Nanj ing University, 1991 A THESIS S U B M I T T E D IN P A R T I A L F U L F I L M E N T OF T H E R E Q U I R E M E N T S F O R T H E D E G R E E O F D O C T O R OF P H I L O S O P H Y in T H E F A C U L T Y OF G R A D U A T E STUDIES D E P A R T M E N T O F Z O O L O G Y We accept tWsJhe^g^Qrjnforming^to the required standard T H E U N I V E R S I T Y OF BRITISH C O L U M B I A F E B R U A R Y 1997 © R o n g Yang, 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. The University of British Columbi! Vancouver, Canada Department DE-6 (2/88) A B S T R A C T A physiological model was developed to estimate fish body toxicant load based on information regarding the chemical exposure regime, fish body weight, lipid content and oxygen uptake. Three organic compounds of different hydrophobicity, 1,2,4,5-tetrachlorobenzene (TeCB), 3,4,5,6-tetrachloroguaiacol (TeCG) and 4,6-dichlorobenzenediol (DBD), were chosen as the test chemicals to carry out a series of investigations before an overall model was assembled. The primary focus of the model was to incorporate physiological components into traditional compartmental models in order to avoid the difficulties associated with complex conventional physiological models. The approach taken was to predict rate constants based on fish oxygen consumption, a parameter speculated, and subsequently shown, to be closely correlated to toxicant transfer in fish. A significant correlation was found between the toxicant uptake process, as characterized by the uptake rate constant (ki), and fish oxygen consumption, regardless of fish size and species. Moreover, the correlation was improved when fish toxicant body load was expressed on a percent body lipid basis. Similarly, fish toxicant depuration tests showed that there also existed a significant relationship between the toxicant depuration rate constant (k2) and fish oxygen uptake regardless of the differences in the chemical octanol/water partition coefficients ( K o w ) . The finding that the chosen test compounds did not interfere with fish oxygen consumption after prolonged sublethal exposure justified the use of oxygen uptake as an indicator for fish toxicant transfer and, as equally important, the utilization of a large fish i i oxygen consumption database ( O X Y R E F ) in the proposed chemical modeling. A series of feeding experiments were also carried out and it was concluded that fish toxicant transfer across the gills plays a dominant role in the toxicant accumulation and depuration of non-metabolized chemicals in fish. Uptake of these toxicants in the food was negligible in determining body burden. In view of the above findings a general model was tested in which O X Y R E F was used to predict fish toxicant body burden. Based on the quantitative analysis, it was shown that the model was reliable and accurate in estimating fish body burden of a number of non-metabolized aquatic toxicants. Values calculated using this model agreed with most determinations reported in the literature. Despite the restrictions and preconditions associated with this physiological model, its main advantage over other compartmental or physiological models lies in the fact that the prediction is based on the actual physiological processes, and fish oxygen consumption rate is far easier and accurate to measure than other physiological parameters even in the absence of the O X Y R E F . This modified model possesses some functional reality which enables more realistic predictions, making it useful for aquatic environmental risk assessment. iii T A B L E OF CONTENTS Page ABSTRACT ii T A B L E OF CONTENTS iv LIST OF TABLES vi LIST OF APPENDIX TABLES vii LIST OF FIGURES viii LIST OF APPENDIX FIGURES xi LIST OF ABBREVIATIONS xii A C K N O W L E D G E M E N T S xiii CHAPTER I General Introduction 1 Introduction 2 Xenobiotic Uptake and Depuration in Fish 2 Kinetic Models of Bioaccumulation 7 Compartmental models 7 Physiological models 9 Non-compartmental models 10 Development of a Simplified Physiological Model 11 Chemical exchange in gills and K o w 12 Rate constants and fish oxygen consumption rate 14 Chemical Selection and Model Application 16 Test compounds 16 Prerequisites for the model application 17 Effects of toxicant exposure on fish oxygen uptake 17 Relative importance of toxicant uptake through feeding and breathing 18 Model testing 18 CHAPTER II Correlation between Oxygen Uptake and Chemical Uptake 19 Introduction 20 Materials and Methods 22 Experimental animals 22 Fish respirometer 22 Experimental procedure 25 Sample preparation and extraction 27 Water 27 Fish 28 GC chromatography analysis 30 Lipid analysis 30 Calculations and statistical analysis 31 Results 33 Discussion 43 iv CHAPTER III Correlation between Fish Oxygen Uptake and Chemical Depuration 47 Introduction 48 Materials and Methods 50 Fish acquisition and acclimation 50 Depuration test protocols 50 Initial toxicant loading 50 Depuration experiment 51 Chemical analysis and calculation 52 Results 53 Discussion 76 CHAPTER IV Prerequisites for the Model Development and Application 79 Introduction 80 Materials and Methods 82 Effects of toxicant exposure on fish oxygen uptake 82 Experimental fish 82 Toxicants 82 Experimental system and procedures 83 Calculation and statistical analysis 87 TeCB and TeCG depuration tests with initial toxicant loading via food ingestion 87 Test fish 87 Exposure regime for trout fry 87 Juvenile trout depuration experiment 88 Results 89 Discussion 100 Effects of TeCB and TeCG sublethal exposure on fish oxygen uptake 100 Relative importance of toxicant uptake through feeding 102 CHAPTER V Model Application and General Discussion 106 APPENDIX Sublethal Toxicity of the Test Compounds 120 Introduction 121 Materials and Methods 122 Exposure regime 122 Sampling procedure 123 Analytical techniques 124 Statistical analysis 124 Results 126 Haematology 126 Plasma sodium and muscle moisture 133 Na7K +-ATPase 133 Discussion 141 LITERATURE CITED 146 v LIST OF T A B L E S Table 1. Oxygen consumption by the OxyGuard® probe during 20 min at 7.07 °C. Page 90 Table 2. TeCB and TeCG depuration profile in juvenile rainbow trout {Oncorhynchus mykiss) after toxicant uptake through feeding. Fish toxicant body burden was expressed as fig toxicant •g fish wet weight"1. 98 vi LIST OF A P P E N D I X T A B L E S Page Table 3. Effects of TeCB and TeCG on the haematocnt 127 (Hct) (%), haemoglobin content (Hb) (g-dl"1), mean cell haemoglobin concentration (MCHC) (100-Hb-Hcf !), plasma sodium level (mmoH"1) and gill, kidney Na + /K + -ATPase activity (|imol P,-mg protein-h"1) of adult trout (Oncorhynchus mykiss) during 200 (ig-1"1 TeCB or TeCG exposure and depuration in freshwater. Table 4. Effects of TeCG on the haematocnt (Hct) (%), 129 haemoglobin content (Hb) (g-dl"1), mean cell haemoglobin concentration (MCHC) (100-Hb-Hct"1), plasma sodium level (mmolT 1), muscle moisture (%) and gill, kidney Na /K + -ATPase activity (|0.mol P, • mg protein-h" ) of adult coho salmon (Oncorhynchus kisutch) during 100 jag-T1 TeCG exposure and depuration in freshwater or continuous exposure for 11 days in freshwater. vn LIST OF FIGURES Page Figure 1. The general framework of this study. 3 Figure 2. Diagram of the fish respirometer and ancillary equipment 23 (from Gehrke et al, 1990). Figure 3. Standard curves for TeCB, TeCG and D B D G C analysis. 35 Figure 4. Relationship between toxicant and oxygen absorption in 37 fishes of different species and sizes. L=large, M=medium, S=small; RBT=rainbow trout, CS=coho salmon, LSS=largescale sucker. Figure 5. Relationship between toxicant uptake rate constant (ki) 39 and fish oxygen consumption. L=large, M=medium, S=smali; RBT=rainbow trout, LSS=largescale sucker. Rate constant ki was calculated based on fish toxicant body burden expressed as (ig toxicant-g body weight" . Figure 6. Relationship between toxicant uptake rate constant (ki) 41 and fish oxygen consumption. L=large, M=medium, S=small; RBT=rainbow trout, CS=coho salmon, LSS=largescale sucker. Rate constant ki was calculated based on fish toxicant body burden expressed as pig toxicant-g lipid"1. Figure 7. Depuration profile of TeCB for juvenile rainbow trout 54 (Oncorhynchus mykiss) swimming at 2.0 BL-s" 1, 11.6 °C (n=6). Figure 8. Depuration profile of TeCB for juvenile rainbow trout 56 {Oncorhynchus mykiss) swimming at 2.0 BL-s" 1, 4.6 °C (n=6). Figure 9. Depuration profile of TeCB for juvenile rainbow trout 58 (Oncorhynchus mykiss) swimming at 2.9 BL-s" 1, 14.3°C (n=6). Figure 10. Depuration profile of TeCB for juvenile coho salmon 60 (Oncorhynchus kisutch) swimming at 2.0 BL-s" 1, 6.5°C (n=6). Figure 11. Depuration profile of TeCB for juvenile coho salmon 62 (Oncorhynchus kisutch) swimming at 2.9 BL-s" 1, 9.6 °C (n=6). viii Page Figure 12. Depuration of TeCG for juvenile rainbow trout 64 (Oncorhynchus mykiss) swimming at 2.0 BL/s" 1 , 10.6 °C (n=6). Figure 13. Depuration profile of TeCG for juvenile coho salmon 66 (Oncorhynchus kisutch) swimming at 2.9 BL-s" 1, 11.3 °C (n=6). Figure 14. Depuration profile of D B D for juvenile rainbow trout 68 (Oncorhynchus mykiss) swimming at 2.0 BL-s" 1, 7.3 °C (n=6). Figure 15. Depuration profile of D B D for juvenile coho salmon 70 (Oncorhynchus kisutch) swimming at 2.9 BL-s" 1, 9.2 °C (n=6). Figure 16. Correlation between TeCB, TeCG and D B D depuration 74 rate constant (k2) and oxygen consumption rate in small rainbow trout (Oncorhynchus mykiss) and coho salmon (Oncorhynchus kisutch) (n=6). Figure 17. A schematic diagram of the six-vessel, intermittent flow- 84 through, computerized fish respirometer. Figure 18. Oxygen consumption profile of juvenile rainbow trout 93 (Oncorhynchus mykiss) during the 48 h pre-exposure and 48 h TeCB exposure (100, 200 lig-1"1) in a flow-through respirometer (n=12). Figure 19. Oxygen consumption profile of juvenile rainbow trout 95 (Oncorhynchus mykiss) during the 48 h pre-exposure and 48 h TeCG exposure (100, 200 figT 1) in a flow-through respirometer (n=12). Figure 20. Comparison of the initial chemical body burdens, ' 103 acquired respectively from 2 h exposure in water and by feeding on toxicant loaded trout fry, in medium size rainbow trout (Oncorhynchus mykiss) (n=6) and the relative importance of the toxicant uptake from water and food. Figure 21. Comparison between experimental and predicted values 111 for chemicals of different K o w . Measured values for dichlorodiphenyltrichloroethane (DDT) (log K o W = 3.99), sodium dodecylbenzenesulfonate (LAS) (log Kow = 2.24), ethylenediaminetetraacetic acid (EDTA) (log Kow = 2.56) and tetradecylheptaethoxylate (AE) (log K o w = 3.78) are from Bishop ix and Maki,1980; Measured values for Di-2-ethylhexyl phthalate (DEHP) (log K o w = 4.35) are from Tarr, et al, 1990; and those for TeCB (log K o w = 4.97) and TeCG (log Kow = 4.41) are from Yang and Randall, 1995. Figure 22. Quantitative analysis of the predictive power of the 114 simplified model. Each single data point corresponds to the relevant observed and model-predicted values which constitute the dashed and solid lines, respectively, in the graphs of Figure 21. x LIST OF A P P E N D I X FIGURES Page Figure 23. Effects of TeCG on coho salmon haematological factors. 131 Day 0-3 is freshwater exposure (100 (XgT1) and day 3-8 is clean seawater depuration. Figure 24. Effects of 3-day TeCB and TeCG exposure (200 |xg-l"l) in 134 freshwater on plasma sodium level and gill Na + /K + -ATPase activity of rainbow trout during 24 hr seawater challenge. Asterisk indicates a significant difference (p<0.05). Figure 25. Effects of TeCG on coho salmon plasma sodium level 136 and muscle moisture content. Day 0-3 is freshwater exposure (100 (J-gT1) and day 3-8 is clean seawater depuration. Asterisk indicates a significant difference (p<0.05). Figure 26. Effects of TeCG on coho salmon Na + /K + -ATPase activity 139 in gill and kidney. Day 0-3 is freshwater exposure (100 p.g-1"1) and day 3-8 is clean seawater depuration. Asterisk indicates a significant difference (p<0.05). xi LIST OF A B B R E V E V I A T I O N S 1CFOK one-compartment, first-order kinetic model A E tetradecylheptaethoxylate D B D 4,6-dichlorobenzenediol D D T dichlorodiphenyltrichloroethane DEHP di-2-ethylhexyl phthalate E D T A ethylenediaminetetraacetic acid L A S sodium dodecylbenzenesulfonate LC50 median lethal concentration O X Y R E F oxygen data bank TeCB 1,2,4,5-tectrachlorobenzene TeCG 3,4,5,6-tetrachloroguaiacol xii A C K N O W L E D G E M E N T S First of all, I would like to express my sincere thanks to my supervisor, Dr. David J. Randall, for his guidance, encouragement and support during this study. I am indebted to my research committee, Dr. George Iwama, Dr. Tony Farrell, Dr. A l Lewis and Dr. Don McPhail, for their comments on this manuscript and many constructive discussions throughout the study. My appreciation also extends to all my colleague students who have left or are still in Dr. Randall's lab. In particular, I would like to thank Dr. Colin Brauner, Dr. Hong Lin, Dr. Mark Shrimpton, Mr. Nick Bernier, Ms. Joelle Harris, Mr. Jonathan Wilson and Ms. Ingrid Burgetz, for all their help and pleasant company. I feel grateful for the assistance from Dr. Tony Farrell and Dr. Chris Kennedy's labs at Simon Fraser University in providing the equipment and fish required for the study described in chapter IV, and the help from Dr. R. V. Thurston's lab at Montana State University in doing some of the chemical analyses. Last but not least, I would like to thank my wife Y i , my son Jason, and my parents for their incredible inspiration and support. xiii C H A P T E R I General Introduction INTRODUCTION The release of synthesized organic compounds, known as xenobiotics, into the aquatic environment is a great threat to fish and other aquatic life. Chemicals at very low concentrations in water can accumulate in fishes and other aquatic animals and have a toxic action. The deleterious toxic effects can occur at different biological levels depending on the exposure regime and the specific organisms involved. These biological changes, whether at the cellular, organ, individual or population level, are usually associated with the accumulation of toxicants within the body. Toxicant body burden, therefore, is of particular interest and importance in understanding the behavior and impact of toxicants on fish. In addition, because of the hazardous effects on humans, i f the toxicants are transferred along the food chain, the prediction of fish body burden is of importance in the regulation of fisheries in contaminated waters. The main objective of this study was to develop a general, accurate and easy-to-apply model to estimate fish toxicant load at any given time, based on information regarding water quality, chemical exposure and physiological parameters of exposed fish. The general framework of the study is described in Figure 1. X E N O B I O T I C U P T A K E A N D D E P U R A T I O N IN FISH Xenobiotics are produced by various types of industrial processes, for example those associated with pulp and paper production, food processing, and the petrochemical industry. 2 Hypothesis I: A simplified physiological model can be developed to predict xenobiotic concentration in fishes Chemical selection and model development Hypothesis II: Fish oxygen uptake is indicative of toxicant uptake across the gill and toxicant uptake rate constant (ki) is correlated with oxygen uptake Hypothesis IV: The relationship between ki, k 2 and oxygen uptake is applicable to a group of chemicals with different octanolAvater partition coefficient (K o w ) Hypothesis III: Fish oxygen uptake is indicative of toxicant depuration across the gill and toxicant depuration rate constant (k2) is correlated with oxygen uptake Hypothesis V: Toxicant sublethal exposure will not interfere with fish oxygen uptake 1 Feasibility of using an oxygen data base in the chemical modeling 1 Hypothesis VI: Toxicant uptake via food is much less important compared with direct uptake through breathing. Thus, fish gill is the major site of toxicant transfer Model Application Figure 1. The general framework of the study These chemicals are often released into the aquatic environment and toxicity varies markedly. Those compounds that have a high lipid solubility readily diffuse across the surface membranes of animals and accumulate in high concentrations in adipose tissue. Many of these compounds also resist breakdown and persist in both the environment and the animal. It is these special physico-chemical properties, e.g. hydrophobicity and tendency to resist chemical and bio-transformation (Connell, 1990) that make these chemicals hazardous. Uptake of xenobiotic compounds in fish may involve direct transfer from water to the body and/or absorption from contaminated food. Many investigations of DDT and related chemicals have shown an apparent increase in toxicant body burden concentration with trophic level, which is evidence of food chain biomagnification (Rudd and Genelly, 1956; Zitko et al, 1972; Robinson et al, 1967; Woodwell et al, 1967), especially in carnivorous fish. In addition, feeding affects not only chemical uptake, but also elimination (Jimenez et al, 1987), causing the clearance of toxicants with the faeces, in addition to elimination through gills and kidneys as would occur in unfed fish. No clear evidence, however, has been reported of an increase in the concentration of xenobiotics with trophic level (Shaw and Connell, 1982; Robinson et al, 1967). Most organic chemicals must enter the body before they can exert their toxic effects. Uptake of these chemicals during bioconcentration is, in most cases, by diffusion. The principal surface across which the toxicant diffuses depends on the size of the animal. In juvenile and adult fish, the gill constitutes the majority of the animal's surface area (Murphy and Murphy, 1971; Rombough and Moroz, 1990) and is the major site of entry of toxicants. Although cutaneous uptake of xenobiotics may occur (Saarikoski et al, 1986) and could play 4 a more important role in small fishes, most of the work done so far tends to ignore the significance of this pathway to the development of the toxicant body burden (Murphy and Murphy, 1971). A large volume of water flows across the gill epithelium, which is a thin barrier with a large surface area, allowing the efficient exchange of O2 and C O 2 between the water and blood. In general, conditions at the gills in fish are adjusted to meet the metabolic demand for oxygen (Randall, 1990). Oxygen and many xenobiotics are lipid soluble molecules which diffuse across the gill lamellae transcellularly. It is likely that variables such as ventilatory volume, gill epithelial thickness and perfusion, which influence the rate of oxygen transport across the epithelial membrane, wil l similarly influence toxicant movement. That is, the conditions for oxygen transfer are likely indicative of that for toxicant transfer. In fish, therefore, the main route for toxicant entry is postulated to be across the large surface area of the gills (Holden, 1962; Murphy and Murphy, 1971; Randall and Brauner, 1990; Erickson and M c K i m , 1990) and may vary with the rate of oxygen uptake. The hypothesis that xenobiotic uptake may be correlated with oxygen uptake, originated from a study by Murphy and Murphy (1971). Subsequently, Black and McCarthy (1990) found that oxygen and P C B uptake efficiencies were directly correlated under a number of conditions. In other studies with trout exposed to benzo[a]pyrene, naphthalene and 2,2',5,5'-TeCB during an acute decrease in temperature, a correlation was established between oxygen and contaminant uptake (Black et al, 1991). Brauner et al. (1992, 1994) and Thurston et al. (1992) reported a significant relationship between 1,2,4,5-tetrachlorobenzene (TeCB) uptake and oxygen consumption rate during initial exposure to TeCB under normoxic 5 conditions in several freshwater teleost fish of different sizes. Randall et al, (1993) obtained similar results using large and small rainbow trout and suckers exposed to the toxicant TeCG (tetrachloroguaiacol). The results of all these experiments indicate that factors which alter gill membrane diffusing capacity and/or fish gill ventilation may have a similar effect on the uptake of both oxygen and hydrophobic xenobiotics. The uptake of xenobiotics directly from water by fish is determined by numerous factors, the most dominant of which are the physico-chemical properties of the compound and the physiological process which occurs at the gills. The physico-chemical characteristics which may influence the flux of xenobiotics across biological membranes include water and lipid solubility, molecular weight and volume, fugacity, tendency to ionize, and susceptibility to metabolic transformation. Among all the rate-limiting physico-chemical factors of an organic chemical, the most important ones in predicting toxicant movement across gill epithelium membranes are K o w and the active concentration of the compound, which includes both the concentration of the aqueously dissolved chemical and the level of the non-dissociated form if the toxicant is a weak acid. The term depuration has been associated with a specific set of experimental condition, i.e., loss of chemical from the animal after being placed in water without the chemical (Barron et al, 1990). Overall, the polluted fish can have toxicant eliminated via either the gills, or the faecal matter, or through metabolic transformation (Clark et al, 1990). As mentioned above, the gill is the main route of toxicant uptake directly from water by fish. Heath (1987) argued that the fish gill is also an important mode of exit for small compounds that require little biotransformation before excretion, and that this is presumably a passive process. For 6 example, the gills turned out to be an especially important organ for elimination of aromatic hydrocarbons and the rate of excretion of individual hydrocarbons by the gills was shown to be inversely proportional to the size of the molecule (Thomas and Rice, 1981). There is a potential, therefore, that depuration rate could be related to oxygen consumption, as is uptake rate, although the possibility has never been investigated. KINETIC M O D E L S OF BIO A C C U M U L A T I O N The mathematical characterization of chemical absorption, distribution and elimination processes permits a quantitative prediction of the amounts and concentrations of a chemical in the body of an animal as a function of certain parameters. A number of predictive models have been developed for both mammals and aquatic organisms in the past few decades. A l l the models already established fall into three categories: Compartmental; Physiologically-based; and Non-compartmental models. Compartmental model: A compartmental model is a simplified mathematical description of a chemical's behaviour in an animal, where the body is represented as a system of compartments. This type of model has been predominantly applied in aquatic toxicology because, despite their limitations, they allow an essential description of the kinetics of xenobiotics in aquatic animals. The simplest and the most commonly used is a one-compartment, first-order kinetics (1CFOK) model which has been argued by some researchers to be the standard for 7 pharmacokinetic analysis in aquatic toxicology (Spacie and Hamelink, 1982). The 1CFOK model is actually a rate constant-based model with the basic relationships expressed as follows: for uptake: Cf=C w(ki/k 2)(l-e" k 2 t); (1) and for elimination: C^CVe" 1 ^ (2) where Cf: concentration of the chemical in the fish at time t; C w : water concentration; C t 0 : concentration in the fish at time zero; k i , k 2: uptake and elimination rate constants, respectively. The uptake of chemicals from the diet can be described with a similar model. Bruggeman et al. (1981) used the following equation to describe bioaccumulation after dietary exposure. Cf(t)=Ef/k2C f d(l-exp(-k 2t)) (3) where Cfd = concentration of chemical in food; f = feeding rate(food weight - fish weight"1 • time"1); E = absorption efficiency for ingested chemical; Constant dietary exposure results in increasing fish concentrations until a plateau level is reached when the clearance rate equals the uptake rate. The ratio between the concentration in fish and in the food at steady state is given by the biomagnification factor K m : km=Cf(t)/C f d=Ef/k 2 (4) More recently, Gobas (1993) has suggested a comprehensive model which combines all the possible factors that may affect the chemical concentration within a fish body in an overall flux equation, describing the net flux of chemical into the fish as the sum of all of the uptake and loss fluxes: dCf/dt = k fCw + k D C f d - (k2 + k E + k M + k G ) C f (5) where k i , k 2 = uptake and elimination rate constants via gills; k D = uptake rate constant from food; k£ = elimination rate constant by faecal egestion; k M = metabolic transformation rate constant of the chemical. The application of these relationships requires three assumptions: 1) The system operates by first-order kinetics, indicative of passive diffusion of the toxicant into the organism; 2) A steady-state can be reached; and 3) The body can be treated as one well-mixed compartment, with the rate of distribution of the toxicant within the organism exceeding the rate of exchange with the surroundings. Although the model should be applied with some appropriate restrictions, the 1 C F O K model approach can provide a useful simple approximation. In order to estimate steady-state chemical concentrations in fish, values are required for all the rate constants (k's) for different chemicals and fish species, etc. Physiological model: A physiologically-based pharmacokinetic (PBPK) model incorporates the underlying physiological processes and important tissues involved in chemical deposition. Hayton and Barron (1990) have proposed a physiological model to describe xenobiotic uptake across the fish gill: 9 Uptake rate = (C w -C f ) [(d(Dm ') A Km)+(h(D a- 1)A)+V b(K b- 1+Vw ' 1)]- 1 (6) C w and Cf = concentration of the chemical in the external water and the plasma water, respectively; d = aqueous stagnant layer thickness; A = gill surface area; Vb = effective gill blood flow; V w = effective water flow; D m = diffusion coefficient in epithelium; D a = diffusion coefficient in water; k m = blood/water distribution coefficient. This model consists of three components: 1) The concentration gradient between the environment and the fish; 2) Physiological and anatomical characteristics of the gills; and 3) Physical constants specific to the compound in question. It is apparent that the physiological model requires detailed information not only about the properties of the chemical but also about the animal, and data collection of the latter is both technically difficult and time-consuming. A n advantage of the P B P K model is that once the dominant transport mechanisms are characterized, it could provide a rational basis from which to extrapolate the model to other conditions and species. Non-compartmental model: Non-compartmental pharmacokinetic analysis avoids many of the assumptions of compartmental modelling, while characterizing capacities for elimination, storage and 10 persistence. Non-compartmental analysis has received only limited application in aquatic toxicology, apparently due both to the method and the requirement for blood or plasma concentration-time data. D E V E L O P M E N T OF A SIMPLIFIED P H Y S I O L O G I C A L M O D E L There are a large number of models predicting organic toxicant bioaccumulation in fish. These models, however, are very complicated and not easily applied to practical biomonitoring. A simplified physiological model needs to be developed to predict toxicant amount in a fish body. Fish oxygen consumption has the potential to be used as an indicator of toxicant transfer. The overall objective of this research is to develop a simple model in which some easily measured fish physiological parameters and certain chemical properties are incorporated to predict the amount of xenobiotics in a fish body with acceptable approximation. The presumption of this study is that the fish gill dominates as the xenobiotic exchange site between the fish body and the ambient environment during both uptake and depuration periods, as discussed earlier. The most commonly used one-compartment, first order kinetic (1CFOK) model was chosen as the basis of this work. As this is a rate-constant based model, effort was directed toward the prediction of uptake and depuration rate constants according to some easily measured physiological parameter so as to avoid the complicated measurements necessary for the physiological models developed so far. 11 Chemical exchange in gills and octanol/water partition coefficient (Kow): Chemicals are delivered to the gill surface by an unidirectional water flow over the gills; they diffuse across the gills and are then distributed to the tissues by the blood flow. Potentially water flow, blood flow and/or diffusion across the gill epithelium could limit uptake of the chemical by the fish (Randall and Brauner, 1990). The ability to deliver a chemical to the gills wi l l depend on gill ventilation and the water solubility of the chemical, whereas diffusion across the gill epithelium wil l depend on its lipid solubility. Lipid content in fish blood is approximately 5%, so during initial exposure to a xenobiotic with log K o w above 2, almost all of the xenobiotic in the blood wil l be bound to lipids and proteins (Schmieder and Henry, 1988), and the ability of the blood to remove the toxicant wil l far exceed the ability of water to deliver it. Thus, transfer of these xenobiotics into the fish is likely to be limited by either the volume flow of water over the gills (ventilation limited) or by the rate of diffusion across the gills (diffusion limited). The rate at which chemicals are absorbed by fish via the gills is expressed by the gill uptake rate constant ki (1-kg-day'1). The kinetic elimination process is called the depuration rate constant k 2 (day"1). Some authors have reported that the uptake and clearance rate constants of some lipophilic compounds by various fish species have a fixed relationship to the Kow over a certain range (Hawker et al, 1985; Saarikoski et al, 1986). Studies of the relationship between ki and K o w in fish have shown that 1) ki increases with K o w i f logKoW is low (<4); 2) ki is constant i f l o g K o w is large (between 4 to 6); and 3) ki drops with increasing K0w for chemicals with extremely high K o w (log K o w above 6) (McKim et al, 1985; Gobas et al, 1986; Gobas and Mackay, 1987). Hawker et a/.(1985) reported a very good linear 12 correlation between k 2 and logK o w , within the range from 2.5 to 6.0, as depuration rate constant decreases with increased K o w . A n elevation in K o w could result from either an increase in lipid solubility or a decrease in water solubility. Dobbs and Williams (1983) showed that there was a linear relationship between water and fat solubility, but a steep inverse relationship between water solubility and logKow. That is, high l o g K o w values were associated with very low water solubilities. So, the xenobiotics with high K o w should be termed as extremely hydrophobic rather than lipophilic, and transfer of these compounds into fish, therefore, is probably ventilation limited. For chemicals with low log K o w (<4), the capacity of the fish to deliver the chemicals to the gill surface is large because of the high water solubility, but diffusion across the gill epithelial membrane plays a critical role in the uptake process. In other words, toxicant uptake tends to be diffusion limited, hence the uptake rate constant of chemicals in this category increases with K o w . For chemicals with log K o w greater than about 4, the importance in the ability of the fish to deliver the chemical to the gills, due to the much reduced water solubility of the chemicals, wi l l overshadow the diffusion capacity across the gills in the uptake process. It is not surprising, then, to see the independence of uptake rate constant on K o w because uptake of these chemicals is ventilation limited. The hydrophobicity of compounds with l o g K o w above about 6, however, results in an increased importance of binding to any organic matter present in the inhalant water (Black and McCarthy, 1988) or to micelle formation. The observed drop in uptake in ki for extremely high K o w substances, therefore, may be due to reduced bioavailability and/or experimental error associated with the difficulty in measuring the concentration of the aqueously dissolved chemical (Gobas and Mackay, 1987). 13 During depuration, the toxicants already stored in the lipid within the tissues will be in equilibrium with the blood stream which serves as a transporting compartment, for compounds that eventually wil l be removed from the fish body across the gills. It has been shown that depuration rate constant k2 remains more or less the same for chemicals with a low log K o w (< 4) and then gradually decreases along with increasing K o w (Gobas et al, 1986). In the case of chemicals with log K o w lower than about 4, the increased lipid solubility and decreased water solubility of a specific chemical wil l have opposite impact on the overall depuration rate, and it is likely that the combined effect of these two gives rise to a fairly constant k 2 for toxicants with low K o w . For chemicals with log K o w above about 4, the solubility of the chemical in the aqueous phase may dominate the depuration process such that transport is ventilation limited rather than diffusion limited. Aqueous solubility decreases with increasing K o w and results in a decreased depuration rate. Rate constants and fish oxygen consumption rate: Oxygen consumption rate can easily and accurately be measured and there is some evidence already, both theoretical (Randall and Brauner, 1990) and experimental (Black and McCarthy, 1990; Black et al, 1991; Brauner, et al, 1992; 1994; Thurston et al, 1992) that oxygen uptake may be an indicator of toxicant transfer across fish gills. The notion of using oxygen uptake together with other chemical properties to predict the amount of xenobiotics in fish body is theoretically sound, provided that a correlation exists between fish metabolic rate and toxicant uptake/depuration kinetic processes. 14 Randall and Brauner (1990) related oxygen and toxicant uptake rates, and derived a toxicant transfer coefficient (k) for the chemical, TeCB, during an initial exposure period. This coefficient was expressed in mg 02 -kg-fish' 1 ,h" 1 exposure per mgT 1 chemical gradient (water to blood); i.e., the uptake of toxicant per unit toxicant gradient per unit oxygen uptake (mg-kg"!,h"1). Thus, the actual unit of this toxicant transfer coefficient is hour"1. In rate constant based models, the proportionality constant that relates the rate of change of either the amount or concentration of chemical within a compartment to the driving force is a first-order rate constant (kj, units of time"1) (Barron, 1990). The driving force may be expressed as a gradient of either concentration or amount, but the choice must be consistent with the rate expression. A rate constant, therefore, is simply a fraction rate of removal of chemical per unit time. From a toxicological point of view, the toxicant transfer coefficient (X) derived by Randall and Brauner (1990) is, as a matter of fact, an uptake rate constant (ki). Although the linear relationship in this study between oxygen consumption and toxicant uptake and is generated during initial exposure, it does provide some basis for the hypothesis that fish metabolic rate is indicative of the entire toxicant uptake and depuration kinetic processes occurring at fish gills. That is to say, the first order rate constants may be closely correlated with oxygen uptake during either toxicant uptake or elimination, because of the similar physiological process taking place in both cases. As mentioned above, for chemicals with l o g K o w less than about 4, the uptake rate constant (ki) increases with K o w but depuration rate constant (k2) remains fairly constant, and for chemicals with l o g K o w above 4, it is the opposite; i.e., ki does not change while k 2 drops 15 along with increased K o w . Thus, the relationship between oxygen consumption and ki derived from a specific type of chemical may only be applied to a certain log K o w range of the hydrophobic xenobiotics. This correlation, therefore, needs to be further substantiated not only with fishes of different sizes and species, but also with chemicals of different physico-chemical properties ( K o w ) . If fish oxygen uptake can be used to predict the uptake and depuration rate constants for a variety of xenobiotics with different K o w values, this would make it possible, with the availability of a large number of oxygen uptake measurements in the literature, to predict the amount of xenobiotics in a fish body by incorporating fish metabolic rate, octanol-water partition coefficient and possibly other parameters into a first order one compartment kinetic model. C H E M I C A L S E L E C T I O N A N D M O D E L A P P L I C A T I O N Test compounds: To study the correlation between toxicant transfer and fish metabolic rate, some test chemicals were chosen as representative of a group of xenobiotic compounds. The test chemicals selected were based on the following criteria: a) They were not metabolized; b) They covered a range of octanol/water partition coefficients ( K o w ) ; and c) They were significant in industrial production. Three chemicals, 1,2,4,5-tetrachlorobenzene (TeCB), 3,4,5,6-tetrachloroguaiacol (TeCG) and 4,6-dichlorobenzenediol (DBD), were chosen as the test compounds. TeCB and TeCG are stable, non-volatile, not easily metabolised and have relatively high hydrophobicity, their octanol-water partition coefficient ( K o w ) being 4.97 and 16 4.41, respectively, at 25°C. 1,2,4,5-tetrachlorobenzene (TeCB) and tetrachloroguaiacol (TeCG) belong to a group of chlorinated compounds produced by the pulp and paper industry in British Columbia (Garrett 1980). The presence of chlorinated compounds in kraft pulp mill bleach effluent in northwestern Canada was reported as early as 1975 (Leach and Thakore, 1975). These two chemicals, although considered lipophilic compounds, are sufficiently water soluble that they are bioavailable for fish (U.S. Environmental Protection Agency, 1994). It was reported, as discussed earlier, that the uptake and clearance rates of some lipophilic compounds by various fish species are associated with their K o w (Hawker et al, 1985; Saarikoski et al, 1986). D B D has a much lower lipid solubility (log K o w = 2.35) and was selected to represent less hydrophobic chemicals. Prerequisites for the model application: Effects of toxicant exposure on fish O^uptake: The pre-requisite of using fish oxygen uptake as an indicator of toxicant transfer, as mentioned earlier, is that it should not be affected by chemical exposure. Only under this situation can the oxygen uptake values in the literature, derived in the absence of any chemical exposure, be applied in the proposed model. Therefore, the effects of the chosen chemicals on the oxygen consumption of fish during prolonged exposure must be investigated to determine if toxicant exposure is a complicating factor when using fish oxygen uptake as the indicator of toxicant uptake and depuration. 17 Relative importance of toxicant uptake through feeding; and breathing: The enormous amount of substance exchange continuously occurring across fish gills tends to support the hypothesis that direct uptake from water is dominant in fish toxicant accumulation. Toxicant intake via food, however, may also contribute to the total toxicant load, especially at high feeding rates. In order to evaluate the relative importance of toxicant uptake from food and to test the hypothesis that the impact of toxicant uptake via food on the total body burden is negligible, feeding experiments were conducted using adult rainbow trout and small salmon fry. Model Testing: In this rate constant (ki and k 2) based model, it becomes a matter of predicting the uptake and depuration rate constants according to fish oxygen consumption rate. A model test was conducted to look at the feasibility of predicting chemical concentration in fish by incorporating the relationships established between fish oxygen uptake and kinetic rate constants (ki and k 2 ) into the one compartment, first order kinetic model. The main advantage of this modified 1CFOK model over the other compartmental models is its basis on actual physiological processes. Also, fish oxygen consumption rate is well documented in the literature, and is far easier and accurate to measure than some of the parameters required by other physiological models, such as gill surface area and blood flow, epithelial thickness and diffusion coefficients. Based on the outcome of the modeling trial, the implication and limitations of the simplified model are discussed at the end of this thesis. 18 C H A P T E R II Correlation between Oxygen Uptake and Chemical Uptake INTRODUCTION Uptake of xenobiotic compounds during bioconcentration, defined as the direct entry of chemicals across the respiratory surface (Connell, 1990), is most often by diffusion. In juvenile and adult fish, the respiratory epithelium forms an extremely large and thin barrier between the water and blood to permit the efficient gas exchange and comprises the majority of the body surface area (Murphy and Murphy, 1971; Rombough and Moroz, 1990). Due to the much lower oxygen content per unit volume in water than in blood (Cameron and Davis, 1970), fish gills are hyperventilated with water relative to blood flow to ensure that oxygen delivery to the gills by water is matched to the oxygen transport capacity of the blood. Thus the gills are most likely the main site of xenobiotic transfer. Oxygen is a lipid-soluble molecule which diffuses across the gill lamellae transcellularly as do most hydrophobic xenobiotics. It is likely that variables such as ventilatory volume, gill epithelial thickness and perfusion, which influence the rate of oxygen transport across the epithelial membrane, wil l similarly influence toxicant movement. Not surprisingly, it has been postulated that the conditions for oxygen transfer could be indicative of those for toxicant transfer (Holden, 1962; Murphy and Murphy, 1971). Black and McCarthy (1988) found that, in rainbow trout, oxygen and P C B uptake efficiencies from water flowing over the gills were directly correlated over a wide range of uptake efficiencies, ranging from 20 to 80%. In other studies with trout exposed to benzo[a]pyrene, naphthalene and 2,2',5,5'-TeCB during an acute decrease in temperature, a correlation between oxygen and contaminant uptake was established (Black et al., 1991). Brauner et al. (1994) reported a 20 significant relationship between 1,2,4,5-tetrachlorobenzene (TeCB) uptake and oxygen consumption rate during initial exposure to TeCB under normoxic conditions. These findings imply that factors which alter gill membrane diffusion properties and/or gill ventilation may have a similar effect on the uptake by the fish of both oxygen and hydrophobic xenobiotics. It has been recommended that fish lipid content be taken into account in modeling bioaccumulation (Dobbs & Williams, 1983; Connell, 1990), since most of the xenobiotics have relatively high lipid solubility and, as such, wil l tend to be highly concentrated in the adipose tissues relative to the aqueous phases of the fish body. Thus toxicant uptake may be modulated by lipid content of the fish. The objectives of the study described in this chapter were 1) to investigate whether the rate of toxicant uptake was correlated with the oxygen consumption of various species of fish of different sizes and lipid content; and 2) to test the hypothesis that toxicant uptake kinetic process, characterized by toxicant uptake rate constant (ki), is dictated by fish oxygen consumption. 21 M A T E R I A L S A N D M E T H O D S Experimental animals: Fish used in this study were small (5.8 ± 0.32 g) rainbow trout (Oncorhynchus mykiss), medium size (25.35 ± 0.65 g) rainbow trout and coho salmon (Oncorhynchus kisutch) and large rainbow trout (425 ± 22.6 g) and were obtained from fish farms in British Columbia. Large fish were held in 10,000 L , and smaller fish in 200 L fibreglass tanks containing dechlorinated Vancouver city tap water, circulating at approximately 1 body length per second (BL-s"1), for at least two weeks prior to experimentation. Water temperature varied seasonally from 5 - 10 °C. A l l experiments were performed at the temperature to which the animals had been acclimated. Following arrival, fish were fed commercial pellets bi-weekly ad libitum, but were starved at least 72 h prior to introduction to the experimental conditions to ensure they were in a post-absorptive state. In order to reduce handling stress, visual estimates of length and weight were made before the tests and actual parameters determined upon completion of the tests. Fish respirometer: A 130 L modified Brett type respirometer (Fig. 2) was used to force fish to swim at predetermined velocities and is described in detail by Gehrke et al. (1990). The swim chamber is a plexiglass tunnel, 1000 mm in length and 200 mm in diameter, with stainless steel mesh at each end. A n opaque plastic covering was placed around the middle of the swim tube to entice the animals to swim in this region. 22 Figure 2. Diagram of the fish respirometer and ancillary equipment (from Gehrke et al., 1990) 23 24 The large volume of this respirometer permits a fairly constant water toxicant concentration over time; however, accurate oxygen consumption measurement require small volume. Small fish, therefore, were swum in groups of 20 -25 to ensure the combined body mass exceeded 100 g, the lower limit determined for accurate oxygen uptake measurement. Before the introduction of the fish to the respirometer, water temperature was adjusted to that of the holding tank, the PO2 was calibrated and the PO2 of the respirometer was increased to 100% air saturation levels. A l l air bubbles were removed from the swim tunnel, the fish were placed in the swim chamber and forced to swim at 18 cnrs"1 (the minimum flow velocity of the system) for the introductory period of two hours for all the tests unless otherwise specified. During this period there was a slow continuous overflow of dechlorinated city water to prevent the accumulation of metabolic wastes. Experimental procedure: At the time of the test, individual large fish or groups of smaller fish were introduced into the respirometer and acclimated for at least two hours at low swimming speeds prior to chemical dosing. Following this introductory phase, a pre-mixed solution of 1,2,4,5-tetrachlorobenzene (TeCB), or 3,4,5,6-tetrachloroguaiacol (TeCG), or 4,6-dichlorobenzenediol (DBD), dissolved in approximately 67 ml of acetone was slowly added downstream of the test animals into the swim tube to make the nominal toxicant level of 200 ug-1"1. The water velocity was then gradually increased over the next 5 min to the final swimming speed (ranging from 1.3 to 3.9 BL-s" 1 depending on the size and status of the test fish) required for each particular test. Duplicate water samples were taken in 40 ml I-Chem 25 vials with teflon lined lids 5 min after the addition of the chemical. The system was subsequently sealed for the rest of the 2 h test duration when fish oxygen consumption rate was measured. The 2 h uptake test period was chosen for several reasons, primarily due to the fact that this was long enough to measure the differences in the dissolved oxygen levels in the exposure water as long as the total body weight of the test fish exceeded 100 g. Also, during the initial stage of the uptake, e.g., the first two hours, it was unlikely that the rate of the toxicant uptake by the test fish would change to a measurable extent, making the calculation of the uptake rate constant easy to achieve. At the end of each 2 h test period, fish oxygen consumption was calculated according to the following equation: Oxygen uptake = Decrease of dissolved oxygen in the water during the test period (mg) / total body weight of fish swimming in the tunnel (kg) / time (h). Final water samples were taken in duplicate at the end of each test, and the fish were removed from the respirometer and terminated with a sharp blow to the head. Then the fish were rinsed quickly in freshwater and lightly blotted dry with paper towel. Each fish was weighed and measured for total and fork length, and then placed in a pre-labeled plastic bag. A l l water and fish samples were stored under -80°C until later analysis for toxicant and/or lipid content. 26 Sample Preparation and Extraction: a) Water TeCB Aliquots of the water samples were added to 100 ml square cross section volumetric flasks containing 50 ml of laboratory water and a teflon coated magnetic stirrer. Hexane (10 ml) was added and TeCB was extracted by stirring the mixture for at least 30 min. After a settling period, the hexane layer was forced into the neck of the flask by the addition of more water. A 1.0 ml aliquot of this hexane extract was transferred to a 10 ml volumetric flask and exactly 0.5 ml of internal standard solution (Pentachlorobenzene) (50 ugT 1 ) was then added. This solution was made to volume with hexane and TeCB concentrations were determined by GC. TeCG and D B D A different method was required for TeCG or D B D sample extraction. According to the exposure concentration, aliquots of water samples were placed in 100 ml square-shaped volumetric flasks containing 0.5 ug surrogate, ca. 50 ml double distilled water, 2 ml 18N sulfuric acid, 10 ml 1:1 methyW-butyl-ether/hexane, and a magnetic stir bar. Extraction proceeded by mixing the flask contents on a magnetic stirrer for 30 min. Extraction solvent was then forced up into the neck of the flask by addition of more water, and 1.0 ml of this solvent was transferred to a 10-ml volumetric flask. Diazomethane was removed, internal standard was added and the contents diluted to volume with hexane. Method blanks and TeCG or D B D (0.5 ug) fortified blanks were analyzed along with water samples by GC. 27 b) Fish TeCB Small whole fish (10 g), or sub-samples of homogenates of larger fish ground up in a conventional meat chopper, were blended to a fine powder using a Polytron™ tissue homogenizer (Model #PT-10/35, Brinkmann Instruments Inc., Westbury, New York) with the aid of dry ice and anhydrous Na2SC>4 while still frozen. At least two separate samples were prepared from homogenates of medium and large fish. Only single samples were prepared from small fish because of limitations in sample quantity. TeCB was extracted (>8 h) using a soxhlet extraction apparatus with Friedrichs condenser including 24/40 250 ml capacity flask ( V W R Scientific) from each sample using hexane. Surrogate (1,2,3,4-TeCB) in amounts comparable to the amount anticipated for TeCB results (usually 100 ug) was added at the initiation of the extraction step. Lipids were removed from portions of this extract using florisil column chromatography; TeCB and surrogate were eluted with 5% methyW-butyl efher/hexane. Pentachlorobenzene (PCB) internal standard was added to the purified extract and component concentrations were determined by electron capture gas chromatography (ECD-GC). TeCG and D B D Small fish were weighed and then diced into small pieces while still frozen. These small pieces were placed into a tared square-shaped extraction bottle and weight recorded. Medium and large fish were partially thawed, cut into several pieces and then ground using a conventional kitchen meat grinder. The ground sample was further mixed to achieve as 28 homogenenous a tissue sample as possible. TeCG or D B D analyses were performed, in duplicate, on portions of the homogenate. At the time of extraction, weighed aliquots (ca. 5 g) of homogenates were placed in square wide mouth extraction bottles. Each sample, either homogenate aliquot or diced small fish, was acidified by the addition of 2.0 ml 18N sulfuric acid followed by thorough mixing. Surrogate (pentachlorophenol) was then added to each sample followed by blending this mixture for 30 seconds using a Polytron™ tissue homogenizer. The extract was decanted through anhydrous sodium sulfate into a 250 ml volumetric flask. The residue was extracted again, both extracts were combined, and then the volume adjusted to 250 ml with additional extraction solvent. Aliquots (5 ml) of extract were taken to near dryness with the aid of a gentle stream of dry nitrogen and low heat. The residue was immediately methylated in a diazomethane solution generator (>30 min) using approximately 3 ml ethereal diazomethane solution. Excess diazomethane was removed by passing a gentle stream of nitrogen over the sample and then the entire contents were quantitatively transferred to a 11x300 mm glass column filled with 5 g florisil (60-80 mesh) topped with 1-2 cm of anhydrous sodium sulfate. These prepared columns were stored at 130 °C for at least 24 h prior to use. Just before sample addition, each column was rinsed with several milliliters of 5% methyW-butyl ether/hexane solution (elution solvent). Methylated chemicals and surrogate were eluted with 40 ml elution solvent and collected in a 100 ml volumetric flask. Hexachlorobenzene internal standard solution (1.0 ml 880 ug-1"1) was added and the contents diluted to volume. Required dilutions were made with internal standard solution at 8.8 ug-1"1. Controls and matrix spike samples were treated similarly. 29 GC Chromatography Analysis: A l l G C analyses were performed using a Varian model 6000 gas chromatograph equipped with N i 6 3 electron capture detector, auto sampler and data system. The column used was 180 cm X 2 mm i.d. glass column packed with 3% Carbowax-20M on 100-120 mesh Supelcoport™. This column was operated at 130 °C for TeCB and 160 °C for TeCG or D B D to accomplish complete resolution of internal standard, target compounds and surrogates. Argon-methane (10%) was used as the carrier gas at 30 ml-min"1. The detector and inlet temperatures were set at 300 and 200°C, respectively. Regression equations were established from a seven point calibration curve at concentrations ranging from 0.5 to 10 ug-f1 for TeCB and TeCG and 20 to 200 pgT 1 for D B D . Required dilutions were made using an internal standard solution at 2.5 ugT 1 . Method blanks were taken through the entire procedure to check for gas chromatographic interference at the retention times of the target compounds, surrogates and internal standards. A l l standards and samples were injected in duplicate using a Varian model 8000 autosampler. Lipid analysis: Total lipids were determined (Folch, 1957) on six small fish or in duplicate on each medium and large fish tissue homogenate. This procedure involved blending 5 g aliquots of frozen tissue homogenate or diced small whole fish (previously weighed) to a powder in a small stainless steel blender, to which dry ice and anhydrous sodium sulfate were also added. Lipids were extracted by blending this powder with 100 ml 1:1 chloroform:methanol. Two extractions were used. The combined extracts were purified by washing with aqueous 30 potassium chloride solution (0.88 % w/v). Layers were allowed to separate overnight and the aqueous layer discarded. The organic layer containing the lipids was vacuum filtered through a 0.2 u Nylon-66 membrane and made to a 250 ml volume. A 25 ml aliquot was added to a tared beaker and solvent was removed using a gentle stream of nitrogen and low heat. The beaker and lipid residue was vacuum dried, weighed and percent lipid calculated. Calculations and statistical analysis: Gas chromatographic peak areas were measured with a Varian Vista Data System. The general formula used to calculate toxicant and surrogate sample concentrations was described as follows: X = ( A ) ( B ) ( C ) / ( D ) ( E ) ( F ) where X = the concentration of target compound and surrogate in the sample (ug -g _ 1 for tissue and u.g-1"1 for water) A = ug^l"1 concentration of toxicant and surrogate in the final extract determined using calibration curve linear regression equations B = volume of the final extract (1) C = dilution factor D = initial weight or volume of the sample taken for analysis (g or 1) E = aliquot of initial extract taken for cleanup (i.e., 10/250 = 0.04) F = aliquot of initial extract taken for concentration (i.e., 10/100 = 0.10) Since the toxicant uptake rate was thought to be fast and linear initially, the uptake rate constant (ki) was calculated as toxicants accumulated in the test fish during the 2 h exposure 31 period per water/blood toxicant gradient which could also be considered as the measured mean water toxicant concentration during the first 2 h of exposure. A n F-test performed upon the Pearson correlation coefficient was used to test for statistical significance of all regressions and correlations. Differences between the means of various exposure groups was tested by a one-way A N O V A following data transformation, i f required, and a Tukey test. In all tests, a probability of 0.05 was chosen as the level of statistical significance. A l l values are expressed, unless otherwise specified, as the means ± s.e.. 32 R E S U L T S Fish toxicant body burden was calculated based on the G C standard curves for TeCB, TeCG and D B D (Fig. 3) and the peak area of the measured samples on each corresponding chromatogram. Fish oxygen consumption rate was determined based on the raw data recorded during the 2 h test period in a print file in an Olivetti M24 personal computer. A significant correlation was established between fish oxygen consumption and the amount of TeCB, TeCG and D B D absorbed by rainbow trout and coho salmon of different sizes under normoxic conditions (Fig. 4, r2=0.94). The data on largescale sucker collected by Brauner et al. (unpublished data) fitted into the general regression generated in this study, which indicated that the amount of toxicant absorbed was correlated with the total amount of oxygen consumed by the animal during the initial uptake stage regardless of the relative lipid solubility of the target compound, the sizes and species of the fish being used. The toxicant uptake rate constant (ki), which was the main concern for the toxicant uptake model, was calculated for each test and was also found to be closely correlated with fish oxygen uptake (Fig. 5). In this study, lipid concentrations in large fish ranged from 6.04 -10.28 % (mean 8.16 %) and small rainbow trout and coho salmon lipid values ranged from 3.35 - 7.61%, averaging 5.64%. If the toxicant body load was calculated based on the lipid content of the fish, i.e., ug toxicant per gram fish lipid, then the significance of the linear relationship between fish toxicant uptake rate (expressed as ug toxicant per gram fish body weight per unit chemical gradient) and fish oxygen consumption (mg 0 2 per kg fish body weight per hour) is improved (Fig. 6). Fish metabolic rate, therefore, seemed to dictate the 33 chemical uptake kinetics and the relationship was more significant when data were expressed on a lipid basis. 34 Figure 3. Standard curves for TeCB, TeCG and D B D G C analysis Figure 4. Relationship between toxicant and oxygen absorption in fishes of different species and sizes. L=large, M=medium, S=small; RBT=rainbow trout, CS=coho salmon, LSS=largescale sucker. 37 10000 Y = 1.41 X °- 7 5 7 R 2 = 0.942 1000 100 V MCS, DBD O MRBT, TeCB • MRBT, DBD • LRBT, DBD o LSS,TeCB T MRBT, TeCB A SCS,TeCG • SRBT, DBD 10 10 100 O X Y G E N U P T A K E (mg/h) 38 Figure 5. Relationship between toxicant uptake rate constant (ki) and fish oxygen consumption. L=large, M=medium, S=small; RBT=rainbow trout, LSS=largescale sucker. Rate constant ki was calculated based on fish toxicant body burden expressed as \lg toxicant • g body weight"1. 39 1 / • S R B T , T C G / -• L S S , T C G • g • M R B T , T C G • M R B T , T C B • 9»° o V S R B T , T C B L R B T , T C G * M ° A L R B T , T C B A / V A A W v • a A tfv • A / A B • • ^  / • • rv m / • • • / / • • T 7 Y=1.012X-2.946 / R2=0.734 i 102 10 3 Oxygen Consumption Rate (mg/kg/h) Figure 6. Relationship between toxicant uptake rate constant (ki) and fish oxygen consumption. L=large, M=medium, S=small; RBT=rainbow trout, CS=coho salmon, LSS=largescale sucker. Rate constant ki was calculated based on fish toxicant body burden expressed as |i,g toxicant • g lipid"1. 41 • S R B T , T e C G • L S S . T e C G • M R B T , T e C G • M R B T , T e C B 0 S R B T , T e C B V L R B T , T e C G A L R B T , T e C B • M C S , D B D • S C S , D B D 0 S R T , D B D M R B T , D B D Y=1.054X-2.923 R2=0.873 102 Oxygen Consumption Rate (mg/kg/h) 103 DISCUSSION These studies support the contention of Randall and Brauner (1990) that the rate of transfer of toxicants of different hydrophobicity is correlated with the oxygen consumption of various fish species of various sizes. Moreover, the fact that toxicant uptake rate constant (ki) is significantly related to fish metabolic rate (Fig. 5) indicates that the fish toxicant uptake kinetic process can be characterized by fish oxygen consumption, and this finding is of great importance to the development of a simplified physiological model. Oxygen molecules transferred through the gill epithelium wil l be bound to hemoglobin and then carried away in the circulatory system. Similarly, almost all hydrophobic toxicants diffusing into the blood stream wil l be bound by plasma proteins or blood lipids and this bound portion of the toxicant increases drastically with an increase in K o w as reported by Schmieder and Henry (1988). During the initial stage of toxicant uptake the ability of the blood to remove the toxicant wil l far exceed the ability of water to deliver it. As a result, the initial uptake rate wi l l be dependent on the gradient for diffusion of toxicant across the fish gills, which should be very close to the concentration of the toxicant in the ambient water. Hence, the toxicant transfer rate should be fairly constant for at least the first 2 h, which was used as the exposure duration in all the uptake experiments. Consequently, the uptake rate constant (ki) was calculated as toxicants accumulated in the test fish during the 2 h exposure period per water/blood toxicant gradient, which is also the mean water toxicant concentration during the 2 h uptake experiment. This uptake rate constant calculation can be easily made under these circumstances and should be indicative of the whole toxicant uptake profile. In 43 other words, it is valid to use the initial toxicant exposure regime as a vehicle to investigate the correlation between fish toxicant uptake kinetics and fish oxygen consumption. This approach, therefore, is theoretically sound and technically more cost-efficient compared with deriving ki from either the traditional "plateau" method where toxicant exposure is continued until tissue concentrations reach a steady state or the short-term "kinetic" experiments which normally last from 48 h up to 5 days (Bishop and Maki, 1980). Some authors have reported, as mentioned earlier, that the uptake rate constants of some lipophilic compounds by various fish species have a fixed relationship to the K o w over a certain range (Hawker et al., 1985; Saarikoski et al., 1986). In this study the selected target compounds included relatively highly hydrophobic chemicals such as TeCB, TeCG ( K o w = 4.97 and 4.41, respectively) and the less lipid soluble chemical D B D ( K 0 w = 2 . 3 5 ) . The correlation between fish toxicant uptake rate constant (ki) and oxygen uptake (Y=1.054x-2.923, r =0.87) was derived based on the data on all three chemicals, and should be applicable to basically all xenobiotics, except those which are dissociated and/or easily metabolized. Theoretically, ki should vary among the three target compounds based on their different K o w - It has been indicated by Hawker and Connell (1985) that a general relationship exists with the ki of a variety of lipophilic compounds and their K o w : Log ki = 0.337 Log K o W - 0 . 3 7 3 (1) Therefore, as K o w increases, the uptake kinetic process should also be faster (higher ki). According to the result in this study, ki is also markedly affected by fish metabolic rate which could be a complicating factor in this relationship. In fact, the degree of elevation in ki induced by the difference in log K o w between TeCB (4.97) and D B D (2.35) could also be 44 easily achieved through the increase in fish oxygen consumption rate at routine levels by 5-10%. Consequently, we use these three chemicals to cover a whole range of xenobiotics for the development of the relationship and make sure that it is generally applicable. Oxygen uptake was found to be the primary factor dictating the toxicant uptake kinetic processes. Lipid normalized data improved the relationship between fish oxygen consumption rate and chemical ki (Fig. 6). So, it can be concluded that fish metabolic rate influences the chemical uptake kinetics and the relationship is more significant when data were expressed on a lipid basis. Lipid normalized experimental values were first used for the standardization and easy comparison of bioaccumulation among different groups in aquatic animals. It seems that the total amount of lipids in an animal, where all the chemicals' residues wil l eventually be found, should affect the final concentration of the toxicant being accumulated rather than the uptake kinetic process depicted by ki . So, it is not clear why the significance of the relationship between oxygen consumption and ki was enhanced by incorporating the lipid content factor. One possible explanation is that, the fatter an animal becomes the longer it tends to take for them to get rid of the toxicant being accumulated. And the instant the uptake starts so does the chemical depuration from the body. Due to the established fixed relationship among k i , k 2 and K o W (Hawker and Connell, 1985): ki/k 2 = 0.048 K o w , it is likely that the decrease in k 2 wi l l result in corresponding changes in ki as well. The possible routes for toxicant entry in fish include direct uptake across the gills, uptake by the skin, and through ingestion of contaminated food. Biomagnification, which refers to the accumulation of a chemical from feeding, is generally thought to be of primary importance in air breathing animals (Connell, 1990). The relative contribution of chemical 45 uptake from food in fish wil l be discussed in detail in Chapter IV. For non-air breathing animals, such as fish, direct uptake from the environment is normally considered to be the main mode of toxicant uptake (Shaw and Connell, 1982; Reinert, 1972). Uptake by the skin should vary with size, because of changing surface to volume ratios. The relationship between ki and oxygen uptake, however, was not influenced by size, indicating that uptake by the skin was not sufficient to significantly affect this relationship. Thus the relationship between ki and oxygen uptake holds for a range of xenobiotics and fish sizes, and several species of fish. 46 CHAPTER III Correlation between Fish Oxygen Uptake and Chemical Depuration INTRODUCTION Bioaccumulation is considered one of the most important environmental transport and partitioning processes (Kathryn, et al, 1990), and the toxicant residue being accumulated in the fish body is dependent not only on the chemical inflow (uptake) but also outflow (depuration). Overall, a polluted fish can have a toxicant eliminated via the gills to the water, via faecal matter or through metabolic transformation (Clark et al, 1990). Fish are capable of metabolizing organic chemicals to enhance the elimination process, thereby decreasing the equilibrium level of the chemical accumulated. Heath (1987) argued that the fish gill is an important mode of exit for small compounds that require little bio-transformation before excretion, and that this is presumably a passive process. For example, the gills turned out to be an especially important organ for elimination of aromatic hydrocarbons and the rate of excretion of individual hydrocarbons by the gills was shown to be inversely proportional to the size of the molecule (Thomas and Rice, 1981). For non-metabolized organic compounds, the fish gill is postulated to be the main route of toxicant excretion, as well as entry, of hydrophobic toxicants (Yang and Randall, 1995). Due to the fact that fish gills constitute the majority of the body surface and represent a very thin barrier between blood and ambient water, there is a potential that toxicant depuration rate could be related to oxygen consumption, as is toxicant uptake rate, although the possibility has never been investigated. In this chapter, I report on 21-day depuration tests designed for small rainbow trout and coho salmon, conducted at different water temperatures under low and high swimming velocities, to determine if there was a correlation between 48 oxygen uptake and depuration rate of toxicants. Fish were exposed to TeCB, TeCG, or DBD alone, or simultaneously to all three chemicals to see if the depuration dynamics of more than one chemical can be determined without interaction among chemicals. Correlation between toxicant depuration and fish oxygen consumption was explored to investigate the possibility of using oxygen uptake as an indicator for toxicant depuration rate in fish. The main theme of this thesis is the derivation of a simplified physiological model to predict fish xenobiotic body burden which is essentially controlled by two opposite dynamic processes, uptake and depuration, occurring at the same time. Since toxicant uptake was demonstrated to be dependent upon fish oxygen uptake (see chapter II), whether or not the rate of depuration of non-metabolized chemicals is also related to oxygen uptake, as hypothesized, becomes another crucial issue for the development of the general model. 49 M A T E R I A L S A N D M E T H O D S Fish acquisition and acclimation: Fish used in this study were small (5.4 ± 0.52 g) rainbow trout (Oncorhynchus mykiss) and coho salmon (Oncorhynchus kisutch) (4.96 ± 0.32 g), purchased respectively from Westbrook Trout Farm in Aldergrove and Capilano Hatchery in North Vancouver, British Columbia. The fish were held in 200 L fibreglass tanks containing dechlorinated Vancouver city tap water, circulating at approximately 1 body length per second (BL-s"1), for at least two weeks prior to experimentation. Water temperature varied seasonally from 5 - 10 °C. A l l experiments were performed at the temperature to which the animals had been acclimated. Following arrival, fish were fed commercial pellets once a week but were starved at least 48 h prior to introduction to the experimental conditions to ensure a post-absorptive state. Depuration test protocols: Initial Toxicant loading. Prior to the depuration tests the experimental fish were loaded with target compounds 1,2,3,4-tetrachlorobenzene (TeCB), 3,4,5,6-tetrachloroguaiacol (TeCG) or 4,6-dichlorobenzenediol (DBD), through gill diffusion. Water exposure was achieved through a double 12 h dosing regime with either single chemicals or all three chemicals simultaneously at predetermined concentrations. TeCB, TeCG or D B D were added to a fully automated Brett-type respirometer (Gehrke et al., 1990) giving the water a calculated toxicant concentration of 270 - 300 g-1"1. After a 5-min mixing time, water samples were taken and the 50 fish were transferred to the respirometer. The system was then sealed and water flow resumed. In order to achieve adequate aeration, oxygen was introduced into the system through a tube penetrating the top lid of the swim tunnel at a flow of approximately 10 ml-min"1. Water oxygen levels were routinely monitored to ensure normoxic conditions within the respirometer. After 12 h, fish were temporarily removed from the respirometer and duplicate water samples were taken for chemical analysis. Then the swim tunnel was flushed and the same start-up procedure was repeated for the second 12 h exposure period followed by the fish being reintroduced back into the system. At the end of the second 12 h dosing period, duplicate water samples were taken again and all fish were removed from the respirometer. At this time, 6-8 fish were killed, briefly rinsed with clean fresh water and dried with paper towel before weight and body length were measured. Then, these fish were stored in plastic bags under -80 °C for subsequent toxicant analysis and were treated as the day-0 sample into the depuration tests. Depuration experiment Depuration tests were conducted after the fish had accumulated toxicants via gill diffusion. The double 12 h dosed fish were temporary held in a bucket while the respirometer was cleaned, and then were put back into the swim tunnel and the clean freshwater flow was resumed. During the 21 day depuration periods, fish were swum against water flow rate of 18 or 27 cm-s"1, but at different temperatures ranging from 4.6 to 14.3 °C for different depuration tests. The respirometer was operated with the artificial lung being incorporated in the flow path and the swim tunnel partially opened such that water could overflow the chamber as 51 freshwater was added. This operating mode ensured that the depurated toxicant was removed from the receiving water in the respirometer. Normally, 6-8 fish were taken from the chamber at 1, 1.5, 4, 7, 14, 21 day intervals during the depuration phase, fish length and body weight were recorded prior to being stored in plastic bags and frozen for subsequent whole body toxicant burden analysis. Oxygen consumption rate was monitored twice a day as long as the total body weight of all remaining fish was not less than 100 g which is the lower limit for accurate oxygen uptake measurement in this 130 L respirometer. A combined oxygen uptake was determined for all fish in the respirometer at that time based on the rate of oxygen depletion in the tunnel over that period of time. Chemical analysis and calculation: GC analysis and calculations were similar to those described in Chapter II. Soxhlet extractors were used for TeCB samples and methylation protocol required for TeCG or DBD samples. All data including calculations and regression analysis were done using QUATTRO PRO™ spreadsheet software. All values are expressed, unless otherwise specified, as the means ± s.e.. 52 RESULTS In these depuration experiments, juvenile rainbow trout and coho salmon depurated TeCB, TeCG or DBD while swimming at different velocities under various water temperatures. The oxygen consumption rate of the fish remaining in the respirometer after each sampling interval was measured and corrected after all the fish body weights were known at the end of each test. Changes in body weight were small (<5%) over the 21 day depuration even though the fish were not fed, and these changes were ignored in the calculation of oxygen uptake. The toxicant body burden declined over the 21 day depuration period as shown in each depuration data profile (Fig. 7-15). In fact, the 21 day depuration tests resulted in insignificant changes in fish body weight and the close-to-complete toxicant elimination towards the end of the tests. Considering Fig. 7 as an example, juvenile rainbow trout were loaded with a mean level of 75.6 g TeCB • g body weight"1 initially and the whole body TeCB concentration decreased to 1.5 g TeCB • g body weight"1 after being forced to swim at 2.0 BL1"1 at 11.6 °C for 21 days. Comparison of TeCB concentrations in rainbow trout collected on a given sampling day with corresponding depuration time provided a best curve-fitting routine in the exponential relationship: Ct = 69.20 e "0171 ' (r2 = 0.996), where C ( represents TeCB concentration in rainbow trout at any given time. Other depuration tests, conducted using either juvenile rainbow trout or coho salmon swimming at 2.0 or 2.9 BL's"1 but under different Figure 7. Depuration profile of TeCB for juvenile rainbow trout (Oncorhynchus mykiss) swimming at 2.0 BL-s"1, 11.6 °C (n=6). 54 0 5 10 15 20 25 Time (day) Figure 8. Depuration profile of TeCB for juvenile rainbow trout (Oncorhynchus mykiss) swimming at 2.0 BL-s"1, 4.6 °C (n=6). 56 Time (day) Figure 9. Depuration profile of TeCB for juvenile rainbow trout (Oncorhynchus mykiss) swimming at 2.9 BL-s"1, 14.3°C (n=6). 58 '3 > -7 £ bi) C o O c o O 3 O o g O 10 15 20 25 Time (day) Figure 10. Depuration profile of TeCB for juvenile coho salmon (Oncorhynchus kisutch) swimming at 2.0 BL-s"1, 6.5°C (n=6). 60 100 0 5 10 15 20 25 Time (day) Figure 11. Depuration profile of TeCB for juvenile coho salmon {Oncorhynchus kisutch) swimming at 2.9 BL-s"1, 9.6 °C (n=6). 62 Time (day) Figure 12. Depuration of TeCG for juvenile rainbow trout (Oncorhynchus mykiss) swimming at 2.0 BL-s"1, 10.6°C(n=6). 64 0 5 10 15 20 25 Time (day) Figure 13. Depuration profile of T e C G for juvenile coho salmon (Oncorhynchus kisutch) swimming at 2.9 BL-s" 1 , 11.3 °C (n=6). 66 e o '§ •+-» c o c o O O O CD bD bo C o o o o 10 15 20 25 Time (day) Figure 14. Depuration profile of DBD for juvenile rainbow trout (Oncorhynchus mykiss) swimming at 2.0 BL-s"1, 7.3 °C (n=6). 68 - fefl o <3 +-> c o o O Q PQ Q =3 O E—1 o g 10 15 20 25 Time (day) Figure 15. Depuration profile of DBD for juvenile coho salmon (Oncorhynchus kisutch) swimming at 2.9 BL-s"1, 9.2 °C (n=6). 70 temperatures ranging from 4.6 to 14.3 °C, showed similar depuration profiles and all the data fitted an exponential relationship. A general trend shown in the depuration experiments was that fish swimming at either faster speeds or higher temperatures, both of which were associated with higher metabolic rates, would lose the toxicant more quickly as indicated by the steepness of the depuration curves. For example, Test 2 (rainbow trout TeCB depuration, Fig. 7) was conducted under the same conditions as Test 1 (also rainbow trout TeCB depuration, Fig. 8), except that the respirometer water temperature was set at 4.6°C while the temperature in Test 1 was maintained at 11.6 °C. As a result, the depuration rate observed was found to be slower than that of Test 1 which displayed a depuration rate constant of 0.119 g-g^ -day"1 (Fig. 8). The mean 21 day TeCB concentrations in fish swum at this low temperature were found to be 3.5 times higher than those swum at temperatures 7 °C higher. Test 3 (rainbow trout TeCB depuration, Fig. 9) was carried out under the same conditions as Test 1 and 2 except that the temperature was at 14.3 °C and, most importantly, the water velocity was increased such that the fish were swimming at 2.9 instead of 2.0 BL-s"1. In this case, the whole body TeCB concentration in the rainbow trout declined at a significantly faster rate (Fig. 9) than those in the above-mentioned two depuration tests. It is not surprising to see the depuration rate constant calculated for this test (2.9 BL-s"1, 14.3 °C) to be -1.135, 9.5 and 33.4 times higher than that in Test 1 (2.0 BL-s"1, 11.6 °C) and Test 2 (2.0 BL-s"1, 4.6 °C), respectively. To further quantify the correlation between fish metabolic rate and toxicant depuration kinetics, the depuration rate constant (k2) was calculated using the BIOFAC computer 72 program for each depuration test and plotted against the mean value of fish oxygen uptake rate in a specific depuration period (Fig. 16). An increase in k2 was observed when fish metabolic rate was elevated and the linear regression between the two was significant ( k2 = 0.0099 M0 2 - 2.2975, r2 = 0.87). 7 3 Figure 16. Correlation between TeCB, TeCG and DBD depuration rate constant (k2) and oxygen consumption rate in small rainbow trout (Oncorhynchus mykiss) and coho salmon (Oncorhynchus kisutch) (n=6). 74 R2=0.87 250 260 270 280 290 300 310 320 Oxygen Consumption Rate (mg/kg/h) DISCUSSION The pathway of toxicant elimination by fish is either branchial, biliary, urinary, and/or cutaneous, depending on the physico-chemical property of the specific chemical and the animal's physiological status. Cutaneous excretion was found to be less than 2% of the total rotenone loss from rainbow trout (Gingerich, 1986) and it is generally accepted that cutaneous excretion plays a minor role in the depuration of most xenobiotic compounds. If the chemical is subject to high oxidative metabolism, the elimination through kidney or bile is expected to increase proportionally (Gingerich and Rach, 1985). Given the fact that fish gills represent the major interface between the water and the body of a fish and high rates of 0 2 and C 0 2 transfer occur across the gill lamellae all the time as the fish respires, the branchial route should be considered dominant in not only toxicant uptake (Yang and Randall, 1995), but also elimination provided the toxicant involved is not metabolised. Xenobiotics of this nature, therefore, are thought to be eliminated essentially unchanged across the gills and the dynamic process is dominated by passive diffusion. The other important factor that could impose great impact on the depuration kinetic is fish oxygen consumption and this is based on the same argument discussed earlier (Chapter II). The hypothesis that oxygen uptake is indicative of the toxicant depuration kinetic process was tested in the depuration experiments using three undissociated chemicals resistant to biological transformation. The depuration tests for chemically dosed rainbow trout or coho salmon indicated that different metabolic rates result in different rates of chemical depuration which was reflected in the slope of the depuration curve (Fig. 7-9). As elimination of 76 toxicants from the whole fish body, and major tissues depots, conforms to simple first-order kinetics (Gingerich, 1986), the depuration rate constant (k2) was calculated for each of the depuration experiments (Fig. 16) and an increase in k 2 was observed when fish metabolic rate was elevated. It was also found that the depuration data profile was similar for fish exposed only to one chemical and those exposed to three chemicals simultaneously, indicating there was no interaction among these chemicals during the depuration process. The increase in chemical elimination rate associated with higher fish oxygen uptake rates, indicates that transfer across the fish gill is the rate limiting step in the depuration of most non-metabolized chemicals. According to the relationship between uptake/depuration efficiency and the octanol-water partition coefficient ( K o w ) , characterised by M c K i m et al. (1985), it appears that the rate limiting process controlling both uptake and depuration varies with logK o w . Considering K o w as an important determinant of the depuration of xenobiotics in fish, target compounds with log K o w ranging from 4.97 to 2.53 were chosen for these depuration studies, aimed at developing a general relationship characterised by the influence of fish "oxygen consumption on toxicant elimination dynamics. As expected, k 2 derived from TeCB (log K o w=4.97), TeCG (log K o w = 4 . 4 1 ) and D D B (log K o w=2.35) depuration experiments, conducted under different conditions associated with various fish metabolic rates, were significantly correlated with fish oxygen consumption, indicating that fish metabolic rate is an important indicator of xenobiotic depuration kinetics for several species of fish. It was reported by Hawker and Connell (1985) that a general relationship exists with the k 2 of a variety of lipophilic compounds and their K o w : 77 -Log k 2 = 0.663 Log K o w -0 .947 (1) The depuration rate constant (k2) should decrease with the increase in K o w - However, it has been shown in this study that fish metabolic rate has a pronounced effect on k 2 which could be a complicating factor in this relationship. With fish oxygen uptake being the primary concern in these tests, it was not surprising to see little effect of K o w on k 2 ; which could probably be due to the great extent of k 2 being influenced by a small amount of changes in fish metabolic rate. We use these three chemicals to cover a wide range of xenobiotics with different hydrophobicity for the purpose of developing a generally applicable relationship. Geyer et al. (1995) demonstrated that the depuration half lives (ti/2), which are inversely correlated with k 2 of T C D D , in mussel and fish at different developmental stages, increased with their lipid content. The lipid contents of the small rainbow trout or coho salmon were all close to 5%; as a result, although k 2 was also calculated using lipid normalized data in these depuration experiments, it was found to be affected by fish metabolic rate in a similar manner as i f k 2 was derived from the data bearing the unit of g toxicant • g wet body weight"1. In conclusion, fish toxicant depuration is enhanced by elevated metabolic rate. The fact that there is a significant relationship between toxicant depuration rate constant (k2) and fish oxygen uptake, regardless of chemical hydrophobicity and fish species, is another important component for the proposed simple physiological model which enables the prediction of toxicant load in fish body at any given time. CHAPTER IV Prerequisites for the Model Development and Application INTRODUCTION In the present investigation, TeCB, TeCG and DBD were chosen as the target compounds for deriving a physiological model based on oxygen consumption to predict xenobiotic concentrations in fishes. It has been shown that both uptake and depuration dynamics of these compounds in fish are significantly affected by oxygen consumption (see Chapters II and III). Consequently, a rate constant-based model can be developed based on oxygen uptake. A major advantage of the proposed physiological model is that it uses fish oxygen uptake as an indicator for toxicant transfer, thereby avoiding the time-consuming and complicated measurements normally required for most other physiological models. Oxygen uptake is not only easy to measure, but there also exists a fish oxygen consumption database (Thurston and Gehrke, 1993). This database, called OXYREF, contains data from over 6800 individual laboratory tests in which oxygen consumption was measured. This data bank includes information on fish species, fish weight, test water conditions, and the level of activity of the fish, i.e., "standard"(resting), "routine"(moving about), or "active"(measured swimming speed). The prerequisite, however, of applying OXYREF in this general physiological model is that toxicant exposure does not interfere with fish oxygen consumption since the measurements in the database were taken in the absence of toxicant exposure. Brauner (1991) reported no direct effect of 1 or 4 h TeCB exposure on the oxygen consumption rates of rainbow trout at rest or during exercise. These findings are in agreement with McKim et al. (1985) who exposed rainbow trout to 14 organic chemicals and found no effect on oxygen 80 consumption rate, ventilation rate, or ventilation volume. One of the objectives of the study presented in this chapter was to look at the effect of prolonged constant sublethal exposure of TeCB and TeCG on the oxygen uptake of juvenile rainbow trout (Oncorhynchus mykiss) using a newly modified, multi-vessel, computerized flow-through fish respirometer (Duval et al. 1981; Johansen and Geen 1990). The hypothesis is that toxicant sublethal exposure will not add any complicating factors to the relationship already established between fish oxygen consumption and toxicant transfer. Another concern in developing this model based on uptake of toxicant from the water across the gills is the necessary assumption that uptake via food was negligible. To test this assumption, feeding experiments using adult rainbow trout and small trout fry were conducted to compare the contribution to the total toxicant body burden from uptake across the gills and food ingestion. Because of the continuous and massive exchange of materials across fish gills, it was postulated that, for fish, toxicant uptake via feeding plays a minor role in determining the toxicant body burden in fish. 81 MATERIALS AND METHODS Effects of toxicant exposure on fish oxygen uptake: Experimental fish Juvenile rainbow trout, Oncorhynchus mykiss (43.8 ± 3.6 g; fork length 15.2 ± 0.3 cm), were purchased from West Creek Trout Farm, Aldergrove, British Columbia. Fish were held in a 2.5 m diameter fiberglass tank under natural photoperiod in dechlorinated Vancouver city water (temperature 5-10 °C, pH 6.1-6.3, hardness 5.2 to 6.0 mgT1 as CaC03 and O2 saturation approx. 95%) at Simon Fraser University, Burnaby, British Columbia. Fish were fed commercial trout pellets, Biodiet Grover (purchased from Bioproducts, Warrenton, Oregon), weekly ad libitum. Fish were starved at least 96 h prior to the trials to assure a post-absorptive state (Goss and Wood 1988; Beamish 1978) and were allowed a 24 h acclimation period within the glass vessel before the initiation of each test. Toxicants 1,2,4,5-tetrachlorobenzene (TeCB) and tetrachloroguaiacol (TeCG) are chlorinated compounds formed in the bleaching processes of pulp and paper industry. Both TeCB and TeCG are relatively hydrophobic with the log Kow being 4.97 and 4.41, respectively (USEPA 1994). Although their water solubilities were sufficient to make them bioavailable in the water, acetone was used to facilitate their dissolution in the test medium. Acetone, used as a chemical carrier, has been proved to be non-toxic to fish at the concentrations administered in this study (Brauner, unpublished data). TeCB and TeCG were acquired from Pfalz and 82 Bauer Inc., Stamford, Connecticut, and Chem Services, Westchester, Pennsylvania. Experimental system and procedures All experimental trials were conducted in a six-glass-vessel computer-controlled, intermittent-flow fish respirometer (Fig. 17) originally designed by Duval et al. (1981) and modified subsequently by Johansen and Geen (1990). The volume of each glass vessel was about 8 L and was covered with opaque polythylene during the whole test period to minimize possible disturbance. The oxygen readings were given by the OxyGuard® DO probe, mounted through the lid into each glass vessel, instead of the modified YSI Model 53 oxygen monitors used by Johansen and Geen (1990). A transduction 386 compatible PC and a custom made A/D-D/A interface were used for system control and data acquisition. Preliminary trials were conducted to find out the optimal total fish body weight each vessel can accommodate in order to avoid any possible continuous oxygen deficit or interactive stress within each vessel, and to assure DO measurement accuracy. Four fish (~ 200 g in total body weight) were placed in each vessel and acclimated for 24 h under continuous flow conditions before each experiment. All vessels were immersed in a 175 L water bath to minimize the water temperature fluctuation (7.11 ± 0.14 °C). The fluorescent lights produced a simulated 14 : 10 hr light: dark photo-period. Parameters such as pre-exposure time, exposure time, cycle time, measurement frequency, purge time/rate, bypass time, toxin flow rate, toxin stock solution concentration, and toxin pump time, etc., were manipulated in order Figure 17. A schematic diagram of the six-vessel, intermittent flow-through, computerized fish respirometer. 84 Toxin Stock Solution I Computer Control Unit Clean Dechlorinated Water Reservoir Mixing Connector Outflow Rate Control Valve Bypass 85 to achieve a pre-determined toxicant exposure regime. Two trials, one for TeCB and one for TeCG, were carried out under identical conditions. Each trial lasted for 96 h, with both the pre-exposure and exposure period being 48 h. Each vessel was flushed with freshwater every 15 min with the purge rate at 1.0 1-min"1 for 2.0 min. The oxygen probe inserted into each vessel was set to record the oxygen level in each vessel every 5 min. The bypass time between each purge was 10 sec, when the connecting tubes were flushed to prevent any possible toxicant residue accumulation. The toxicant flow rate was 2 ml-min"1 and the concentration of the toxicant stock solution was 100 mgT1 TeCB or TeCG. The toxicant was pumped into each vessel, where required, during a purge. The pump time was adjusted, respectively, at 0 min for two control group vessels, 1.0 min for two 100 figl"1 treatment groups and 2.0 min for 200 figT1 treatment groups. Eight control fish at the beginning and all fish in the vessels at the end of each test were terminated. Body weight and fork length were measured. At five-minute intervals, the dissolved oxygen in each vessel was measured by the inserted OxyGuard® Probe and each vessel was partially flushed with freshwater every fifteen minutes. Completion of the sequential oxygen measurements and freshwater flushing in six vessels constituted one cycle. Oxygen consumption rate was based on the pooled value for four fish in each vessel, and total fish body weight in that vessel was measured at the end of the 48 h exposure when the fish were killed. 86 Calculation and statistical analysis The average of oxygen consumption in 30 min was calculated and the oxygen consumption data profile analysed between control and treatment groups, or within treatment groups but between 48 h pre-exposure and 48 h exposure periods. One way ANOVA followed by a Dunnett's test or student t-test were applied to see if there is any effect of the two chemicals on the oxygen consumption pattern of these fish under the condition in this study. TeCB and TeCG depuration tests with initial toxicant loading via food ingestion: Test fish Juvenile rainbow trout {Oncorhynchus mykiss) (58.3 ± 5.9 g) or rainbow trout fry (0.23 ± 0.03 g) were obtained from West Creek Trout Farm, Aldergrove, British Columbia. Fish were held in a 2.5 m diameter fiberglass tank or 80 L glass aquarium under natural photoperiod in dechlorinated Vancouver city tap water. Fish were fed with appropriately-sized commercial trout pellets once a week and were starved at least 96 hr prior to the trials to ensure a post-absorptive state. Exposure regime for trout fry Trout fry were exposed in a 80 L aquarium for 6 days, with the exposure solution partially changed (20 1-day"1) during the first 4 days. The initial calculated TCB (or TeCG) concentration in the water before exposure of the fry was 200 fig-l"1. Test water in the aquarium was sampled each day of the 6-day exposure duration. At the end of the TCB (or 87 TeCG) dosing procedure, 6 toxicant loaded fry were sampled immediately after the exposure and, in addition, 6 contaminated fry that had spent 1.5 h in clean water were sampled. Juvenile trout depuration experiment In the TeCB test, adult trout were maintained for the feeding and depuration period in separated compartments within a large tank with fresh water overflowing continuously. Water temperature was 8.1 - 14.1 °C. Ten toxicant-saturated fry. were fed to each adult trout and the fry were all eaten in 1-1.5 h. After trout were fed, 3 animals were collected at day 0, 1, 3, 8, 12, 17 and 29. Six unfed control adult trout were also sampled. The TeCG depuration test was carried out in a similar fashion except that water temperature during the whole depuration test was 6.0 - 9.6 °C and three fish were collected at day 0, 1, 3, 8, 14, 21 and 26. All water and fish samples were immediately stored in the freezer (-80 °C) for subsequent chemical analysis. The GC analytical procedures conformed with those described in the relevant section of Chapter II. 88 RESULTS Effects of TeCB or TeCG prolonged exposure on rainbow trout oxygen uptake: A major modification in this flow-through respirometer compared with that described by Johansen and Geen (1990) is the usage of the OxyGuard® DO probe, which has a built-in temperature compensation, a large anode and electrolyte volume. Probe air calibration was performed according to the manual (Point Four Systems Inc.) prior to each experiment. The probe uses very little oxygen for its measurement and, therefore, it functions correctly with liquid movement as low as 2 cm/sec (measured at 7 ppm and 13 °C). The total amount of oxygen the probe consumed by itself was minimal (Tab. 1) during 20 min and each cycle, which lasted for 15 min, was always initiated by partial flushing of fresh solution in the glass vessel, indicating the feasibility of using the oxygen probe for accurate oxygen consumption measurement in this study. Marked respiratory changes occur in fish under hypoxic conditions (Hughes 1973; Randall 1982; Bushnell et al. 1984). Therefore, the maximum amount of fish each vessel can accommodate is another parameter that, if inappropriate, could potentially affect the survival of the test fish, let alone the accuracy of oxygen uptake measurement. In pre-experimental trials, four or six fish were put into a glass vessel and oxygen depletion after 20 min was measured. A 20.7% drop of dissolved oxygen was detected in the vessel of four fish and 35.7% in the vessel of six fish. After 5 min freshwater flushing, the DO level in the vessel containing four fish went back to 102.8% of the original level, while the DO was only 84.1% Table 1. Oxygen consumption by the OxyGuard probe during 20 min at 7.07 °C. 90 PROBE 1 2 3 4 Dissolved O2 (ppm) 8.09 8.51 10.29 10.41 7.88 8.33 10.24 10.35 7.77 8.20 10.21 10.31 7.72 8.13 10.19 10.29 7.70 8.10 10.18 10.28 % Decrease in 20 min 4.8 4.8 1.1 1.2 of the initial concentration if six fish were present in the chamber. As a result, only four fish were included in each 8 L vessel during the actual experiments. A circadian rhythm in fish oxygen consumption rate was detected in this study (Fig. 18 and 19), reflecting the normal daily activity pattern (Brett and Zala, 1975) of the fish being used. The mean fish oxygen consumption was similar to those found by other investigations using the same apparatus (MacKinnon and Farrell, 1992; Janz et al. 1991; Johansen and Geen, 1990), indicating that the system provided reliable oxygen recordings with the newly installed OxyGuard® DO probes. After being acclimated in the vessel for 48 h, rainbow trout exposed to nominal concentrations of 100, 200 (ig-1"1 TeCB (Fig. 18) or TeCG (Fig. 19) did not show any significant difference in their oxygen consumption throughout the 48 h exposure period. This conclusion was based on the comparison between the fish oxygen uptake profile before and after exposure in the same group, and those in the control and treatment groups. Depuration profile ofjuvenile rainbow trout via toxicant uptake from food: Concentrations in the TeCB and TeCG exposed trout fry, sacrificed immediately after dosing, ranged from 2.44-7.83 Ltg-g"1 and 15.9-28.2 Ltg-g"1, respectively. TeCB and TeCG levels in the fry maintained in freshwater for 1.5 h after dosing were 2.23-11.3 Ltg-g"1 and 10.9-21.1 Ltg-g"1, respectively. There was no significant decrease of either TeCB or TeCG level in toxicant loaded fry after being allowed to swim 1.5 h in freshwater, indicating that toxicant loss from fry before they were consumed by juvenile trout was minimal in these 92 Figure 18. Oxygen consumption profile of juvenile rainbow trout {Oncorhynchus mykiss) during the 48 h pre-exposure and 48 h TeCB exposure (100, 200 Hg^ r1) in a flow-through respirometer (n=12). 93 CD E 0) CO CL E </) c o O CD 200 150 100 Control 200 150 100 50 ] 40 60 100 ug/l TeCB 100 Pre-exposure i Exposure • •• • V . . ' 200 150 100 50 20 40 60 80 100 200 ug/l TeCB Pre-exposure Exposure • • • • * • • 20 40 60 80 100 Time (hr) 94 Figure 19. Oxygen consumption profile of juvenile rainbow trout (Oncorhynchus mykiss) during the 48 h pre-exposure and 48 h TeCG exposure (100, 200 HgT1) in a flow-through respirometer (n=12). 95 CD E CD •*-> CO E tn a o O CD CO >> X O 200 Control 150 100 50 Pre-exposure 1 Exposure 1 • • • • • L • • • • . • 250 200 150 100 50 20 40 60 80 100 100ug/l TeCG Pre-exposure Exposure • • - • i ••. • 200 150 100 50 20 40 60 200 ug/l TeCG 80 100 Pre-exposure • Exposure 20 40 60 80 100 Time (hr) 96 experiments. The mean value of TeCB concentration in juvenile trout sampled at test time zero was 0.170 Aig-g"1, ranging from 0.118 to 0.207 //g-g"1. This value was, not surprisingly, close to a theoretical value of 0.173 Mg*g 1 calculated from the total TeCB consumed in the food. Concentrations of TeCB in the other test fish during the depuration period declined over time reaching non-detectable levels in 12 days (Tab. 2). TeCG concentrations in fish sampled immediately after feeding averaged 0.783 ^cg'g'\ ranging from 0.609 to 1.02 /ig'g"1. This mean level at day-0 was higher than the calculated value (0.491 //g-g"1). Measured TeCG concentrations in fish did not decline until after the first day of depuration and TeCG was still detectable, at the established detection limit of 0.0125 Mg'g"1, in some fish even at the end of the depuration (Tab. 2). 97 Table 2. TeCB and TeCG depuration profile in juvenile rainbow trout (Oncorhynchus mykiss) after toxicant uptake through feeding. Fish toxicant body burden was expressed as |ig toxicant -g fish wet weight"1. 98 Depuration Day TeCB (//g-g1) TeCG Og-g"1) 0 0.1702 ± 0.0219 0.7832 ± 0.0991 1 0.1070 ± 0.0211 0.9013 ± 0.1153 3 0.0616 ± 0.0465 0.5414 ± 0.1061 8 0.0375 ± 0.0281 0.2794 ± 0.0353 12 0.0012 ± 0.0002 14 0.1860 ± 0.0760 18 Not Detectable 21 0.1615 ± 0.0862 26 0.1145 ± 0.0669 29 Not Detectable DISCUSSION Effects of TeCB and TeCG sublethal exposure on fish oxygen uptake: Exposure to TeCB or TeCG did not affect oxygen consumption in trout. This is perhaps surprising because Yang and Randall (1996b) demonstrated that 72 h partially static exposure to TeCB or TeCG (at similar concentrations) could exert sublethal effects on the muscle moisture, plasma sodium concentration and gill/kidney Na+/K+-ATPase activity in rainbow trout of a similar size. In this study, gill Na+/K+-ATPase activity disturbances were also measured and found to change in the same way as reported earlier. It seems that fish stress may not be reflected in their oxygen consumption even though many other changes have been observed (Yang and Randall 1996a; 1996b; MacKinnon and Farrell 1992; Heath 1987). MacKinnon and Farrell (1992) have shown that a concentration-dependent response in oxygen consumption of juvenile coho salmon {Oncorhynchus kisutch) was not observed when sublethally exposed to 2-(thiocyanomethylthio) benzothiazole (TCMTB). It has been shown that 1-4 h sublethal exposure of TeCB, even at concentrations approaching LC50, did not exert any effect on the oxygen uptake of adult rainbow trout at rest or during exercise (Brauner et al, 1994). It was therefore suggested that the correlation studies, usually carried out at much lower concentrations, would not be complicated by any toxicant effects on oxygen consumption rate. This result is in agreement with studies using similar exposure regimes with a variety of chemicals and fish species, such as rainbow trout (McKim et al, 1985), mosquito fish (Murphy and Murphy, 1971) and coho salmon (Janz et al, 1991). As 100 most environmental exposures to fish residing in a polluted area are occurring not only under sublethal concentrations but also for prolonged periods, it is of special interest to look at the effects of chemicals on fish oxygen uptake after a prolonged sustainable exposure, as in this study. No effect on oxygen consumption was observed in fish exposed to these chemicals for prolonged periods. The pre-requisite for the use of OXYREF in the model is that fish oxygen consumption is not influenced by toxicant exposure. Because toxicant exposure does not affect oxygen consumption, the proposed model can utilize an existing oxygen data base OXYREF (Thurston and Gehrke, 1993) compiled from the fish oxygen consumption rate values from the literature, measured in the absence of toxicant exposure. Oxygen uptake can be used as an indicator of toxicant uptake under normoxic conditions. Altered fish ventilation rate or diffusion capacity without changes in whole animal metabolic rate, which may occur during hypoxia or hyperoxia (Randall, 1990), will change the rate of toxicant movement across fish gills without changes in oxygen uptake (McKim et al, 1985). In other words, toxicant absorption may be greatly enhanced under these conditions without there being any significant changes in fish oxygen uptake. Thus, the use of OXYREF in the model is limited to normoxic conditions. The use of the database under hypoxic conditions will tend to underestimate toxicant load whereas using it in hyperoxia will tend to overestimate the toxicant load in fish. 101 Relative importance of toxicant uptake through feeding: Toxicant intake via food may contribute to the total toxicant loading, especially during high feeding seasons. The rationale behind the feeding experiments, therefore, was to test this hypothesis and evaluate the relative importance of toxicant uptake from food by making comparison with those depuration experiments where the fish were loaded with toxicants via gill diffusion. In the previous TeCB and TeCG depuration tests, the initial fish toxicant body burdens, achieved through the double 12 h aqueous exposure in which the maximum toxicant loading was supposedly ensured, were 11.06 fxg TeCB-g"1 and 6.15 /xg TeCG-g"1 when fish were swimming at 2.0 BL-s"1, and 15.00 fxg TeCB-g"1 and 8.94 fxg TeCG-g"1 when fish swimming speed was 2.9 BL-s". In comparison, the initial TeCB and TeCG concentrations in the juvenile rainbow trout, trained to feed on trout fry and starved for two weeks before the feeding experiment, were 0.17 fig TeCB-g"1 and 0.78 fxg TeCG-g"1, respectively. Even when fish were swimming at low velocities, 88.7% of the total body burden of TeCB and 98.5% of TeCG resulted from uptake across the gills (Fig. 20). Although there exist some uncertainties in this comparison, it provides some evidence based on laboratory data to support the argument that, for aquatic animals such as fish, uptake of certain hydrophobic, non-dissociated toxicant residues through food ingestion is probably much less important compared with that through direct water exposure. The subsequent one-month depuration phase in these feeding tests showed that the elimination profiles of toxicants were similar to those described in previous depuration tests Figure 20. Comparison of the initial chemical body burdens, acquired respectively from 2 h exposure in water and by feeding on toxicant-loaded trout fry, in medium size rainbow trout (Oncorhynchus mykiss) (n=6) and the relative importance of toxicant uptake from water and food. 103 2.0 BL's"1 2.9 BL's"1 Uptake from water Uptake from food by at low and high consuming toxicant swimming speeds loaded trout fry 104 where the initial toxicant concentration was achieved through uptake from water. The majority of the toxicant load was eliminated in the first 3-4 days, especially when fish were swum at high speeds. The actual digestion of food in the fish gut normally takes about that much time as well. As a result, it seems that the potential interference of feeding with fish toxicant depuration is minimal. The main conclusion in this chapter is that the chosen target compounds will not interfere with fish oxygen consumption after prolonged sublethal exposure, which has justifies the use of Oxygen Database (OXYREF) in the proposed physiological modeling. Also, the notion that fish toxicant transfer across the gills tends to play a dominant role in the accumulation and depuration of non-metabolised compounds has been further supported by the outcome of the feeding tests. Thus the effects of feeding are validly ignored and not incorporated into the model. 105 CHAPTER V Model Application and General Discussion Interest in the physiologically-based toxicokinetic (PBTK) models has increased in recent years (Nichols, et al, 1993) because of a growing need to predict the time-course of chemical concentration in specific tissues. The major advantage of any PBTK model is that it defines an organism, be it terrestrial or aquatic, in terms of its anatomy, physiology and biochemistry. Consequently, PBTK models can be parameterized independently of exposure information and provide a basis for the comparison of kinetic data between species. This type of model, which leads to improved understanding of the uptake and disposition of chemicals in different animal tissues (Andersen, et al, 1987; Reitz et al, 1990), is also associated with certain restrictions and disadvantages. Firstly, the schematic representation of a PBTK model is normally fairly complicated, requiring a large number of physiological inputs which may not be easily accessible or even available. This has caused problems, for example, in the development of a PBTK model for channel catfish (Nichols, et al, 1993), since only very limited physiological information is available for this fish species in the literature (Burggren and Cameron, 1980; Cameron, 1980), resulting in the absence of estimates for several important parameters such as the cardiac output and respiratory volume. Thus, this model can only be applied following the collection of physiological information for this fish, vital for accurate estimation of toxicant loading using this model. Secondly, the kind of physiological measurements needed, if not available in the literature, are often extremely difficult and time consuming to measure. Considering the channel catfish PBTK model as an example, some of the physiological parameters required are gill surface area, gill epithelial thickness, effective respiratory volume, blood/water 107 partition coefficient, etc. (Nichols, et al, 1993). Such data are available in the literature for a relatively small number of fish species. In addition, measurement of these physiological parameters, acquired with great effort, can be quite variable due to the differences in the experimental set-up, temperature and animal preparation, making the data difficult to compare and extrapolate and, subsequently, to standardize. Moreover, these models are usually developed using a single species which is easy to culture and handle and assumed to be of importance as a model system for kinetic studies. The same model parameters, however, when used for a new species, may not provide an accurate prediction due to the lack of an interspecies extrapolative validity among fish species. Thirdly, the dependence of toxicant concentration prediction on anatomical measurements could lead to errors because fish, at different activity levels, may have different toxicant accumulation and elimination behaviours, not reflected in the model parameterization. For example, environmental or physiological factors that influence ventilation volume also influence the uptake of chemicals. According to Spacie and Hamelink (1982), ventilation volume (Rv) increases with metabolic rate (Q), which is a function of fish body weight (F): Q = F* . As the weight exponent is approximately 0.8, the metabolic rate per gram (Q/F) decreases with increase of fish body weight. Consequently, a 10-g fish consumes about 60% as much oxygen per gram as does a 1-g fish. A similar weight relationship should exist for ki as well. Murphy and Murphy (1971) measured the values of 0.75 and 0.77 for the uptake of DDT by mosquito fish at 5 and 20 °C, respectively. These values were similar to the corresponding exponents for oxygen consumption (0.71 and 0.68), but lower than the exponent for gill surface area per gram (0.89), indicating that uptake is more a function of the fish gill's physiological rather than anatomical properties. Thus, results obtained using these PBTK models involving the measurement of several anatomical and physiological parameters should be accepted with caution and, in general, are only applicable after verification, which negates the value of the model as a predictive tool. The question is posed, therefore, as to whether there exists any suitable physiological parameter that is not only easy to obtain and but is also functionally realistic. All of the above-mentioned physiological parameters such as gill surface area, gill epithelial thickness, gill blood flow, diffusion distance and coefficient, etc., presumably affecting toxicant transfer, are all related to gas exchange across fish gills. As discussed in the general introduction, fish oxygen consumption is the end-result of the optimization of all of the above parameters. In other words, fish oxygen uptake is facilitated in a similar manner as is toxicant transfer across fish gills. The results of this study, presented in Chapters II and III, indicate that oxygen uptake can be used as an indicator for toxicant transfer regardless of fish species and the physico-chemical property of the non-metabolised chemicals in question. In addition, OXYREF has been shown to be applicable to the proposed model since fish oxygen consumption is not altered by toxicant exposure (Chapter IV), meaning that the physiological component critical for this physiological model can be obtained easily. Given the argument that fish gills are the main site for toxicant transfer (Chapter IV) and the toxicant uptake/depuration kinetic processes are characterized by fish metabolic rate (Chapters II & III), it becomes a matter of predicting the rate constants such that the overall 109 toxicant body load can then be estimated at any time using the appropriate modified rate constant model. As mentioned earlier in the general introduction, the most commonly used one-compartment, first order kinetic (1CFOK) model (Spacie and Hamelink, 1982) was chosen as the basis of this work. The basic relationships for toxicant uptake and depuration were expressed, respectively, in equation (1) and (2) of Chapter I. The uptake and depuration rate constants can be calculated according to fish oxygen consumption rate by using the correlations determined in this study (Chapters II & III). A model test was conducted to look at the feasibility and accuracy of predicting chemical concentration in fish by incorporating the relationships established between fish 0 2 uptake and kinetic rate constants (ki and k2) into the one compartment, first order kinetic model. In order to test the validity of this modified model, experiments were chosen from the literature and the experimental data for toxicant load were compared with those derived through the modeling based on relevant information, e.g., size of the fish being used, experimental conditions including toxicant exposure concentration, temperature, etc., reported in that study and used to derive oxygen uptake and, therefore, uptake and elimination rate constants. The following chemicals were selected for the model test: dichlorodiphenyltrichloroethane (DDT) (log Kow - 3.99), sodium dodecylbenzenesulfonate (LAS) (log Kow = 2.24), ethylenediamine- tetraacetic acid (EDTA) (log Kow = 2.56) and tetradecylheptaethoxylate (AE) (log Kow = 3.78) (Bishop and Maki,1980), di-2-ethylhexyl phthalate (DEHP) (log Kow = 4.35) (Tarr, et al, 1990) and TeCB (log K o w = 4.97) and TeCG (log K o w = 4.41) (Yang and Randall, 1995). The result of the model application is shown in Fig. 21. 110 Figure 21. Comparison between experimental and predicted values for chemicals of different Kow. Measured values for dichlorodiphenyltrichloroethane (DDT) (log K o w = 3.99), sodium dodecylbenzenesulfonate (LAS) (log K o w = 2.24), ethylenediaminetetraacetic acid (EDTA) (log K o w = 2.56) and tetradecylheptaethoxylate (AE) (log Kow - 3.78) are from Bishop and Maki,1980; Measured values for di-2-ethylhexyl phthalate (DEHP) (log K o w = 4.35) are from Tarr, et al., 1990; and those for TeCB (log K o w = 4.97) and TeCG (log Kow = 4.41) are from Yang and Randall, 1995. I l l 400 DEHP TeCG 1 2 4 8 16 32 48 64 96 Time (h) 0.5 1 2 4 8 16 21 Time (day) TeCB TeCB 0.5 1 8 16 21 Time (day) Time (day) LAS 120 100 LAS 24 48 96 Time (h) 144 3 4 5 Time (day) DDT DDT 1 6 12 24 48 120 Time (h) 1 6 12 24 48 120 Time (h) EDTA 3 8 24 48 120 Time (h) 12 24 Time (h) 48 120 - Experimental Data - Model Data 112 The predicted uptake data of DDT at both test concentrations agreed with the measured ones quite well up to exposure time 48 h. The experimental data of EDTA, however, seemed to level off after 24 h while the predicted values kept increasing. As for the depuration data, model values of LAS and DEHP concentrations fitted extremely well with the test ones at two water concentrations., and so did the data for high speed TeCG depuration. Assuming the model is perfect, the predicted values will be exactly the same as the observed ones. As a result, the linear regression derived between predicted and measured data should have a slope of 1.0. However, when the values predicted using the present simplified model are plotted against those obtained from the experiments, the regression line has a slope of 1.288 (Fig. 22), meaning there exist some descrepancies. Consequently, the predictive power of the simplified model can be judged by the degree to which the perfectly matched regression line overlaps with the 95% confidence limits of the real regression. It has turned out that the perfect regression line falls within the 95% confidence intervals of 92% of all the predictions performed (67 out of 73 data points). The six points which do not overlap with the perfect match line are all from one single test with the same chemical, i.e. DEHP (logKow=4.35), using rainbow trout at temperature 12 0 C. It is with these 6 points that the model tends to overestimate, or maybe the measured values are lower than what they should be. Discrepancies between predicted and measured values may be due to inaccuracies in either the estimation of oxygen uptake and/or bioavailability of the toxicant. One possibility is that, in these uptake experiments, the chemical exposure concentration may have declined over time due to the initial fast uptake rate by the fish, and as a result measured 113 Figure 22. Quantitative analysis of the predictive power of the simplified model. Each single data point corresponds to the relevant observed and model-predicted values which constitute the dashed and solid lines, respectively, in the graphs of Figure 21. 114 values are underestimated. The relationship between metabolic rate and temperature varies between species and may account for some of the discrepancies. Veith et al. (1979) found that, between 5 to 15 °C, the bioconcentration of Aroclor® increased much more for the fathead minnows (Pimephales promelas) and green sunfish (Lepomis cyanellus), which are warmwater fish, than it did for rainbow trout (Oncorhynchus mykiss), a coldwater species. It was also reported that increases in oxygen uptake per gram were greater for carp (Cyprinus carpio) than for brook trout (Salvelinus fontinalis) (Beamish, 1964). Since the model development was based on data derived from experiments using salmonids, it is probably more appropriate and safe to conclude that the established relationships may be applicable to these groups of fish in particular and precautions should be taken when the model is used for fish markedly different in terms of life history, habitat requirements and geographical distribution. In any model of the uptake and elimination of xenobiotics in fish, one of the most important components is the concentration of the chemical that can be absorbed from the water via the gills of fish. Environmental factors that affect the concentration of a chemical in true solution (Cw) will affect both the initial uptake and steady-state concentration in fish tissues (Spacie and Hamelink, 1982). In particular for very hydrophobic chemicals, the concentration of absorbable or bioavailable chemical is often only a fraction of the total chemical concentration in the water. This fraction is normally referred to as the Bioavailable Solute Fraction (BSF), or simply the bioavailability (Landrum et al. 1985; McCarthy and Jimenez, 1985). The major factors affecting bioavailability of organics in natural waters is the presence of adsorbents (Gobas, 1993), namely sediments, suspended solids and dissolved 116 polyelectrolytes such as humic acids. For groups of pesticides (Wershaw and Goldberg, 1972; Kenaga and Goring, 1980), polynuclear aromatic hydrocarbons (PAH) (Means et al, 1980) and industrial organics (Lopez-Avila V. and R.A. Hites. 1980), sorption to the organic fraction of sediments generally increases as the octanol-water partition coefficient (K™) increases. Similarly, dissolved organic matter, including humic substances, binds chlorinated hydrocarbons, phthalates and PAH (Josephson, 1982; Sullivan et al, 1982), reducing their bioavailability. The presence of humic substances retards the uptake of benzo(a)pyrene (log K o w = 6.06) from water by bluegills, but does not significantly affect the accumulation of anthracene (log K o w = 4.45) (Spacie et al, 1983). All of these effects should not alter the uptake rate constant ki since the kinetic process is independent of the real toxicant concentration in the ambient water. However, they do sometimes reduce the "true" chemical concentration (Cw) which is one of the important components of the model presented here (Equation 1, Chapter I). Moreover, the actual value of C w may be unknown since it is difficult, analytically, to distinguish between the free and adsorbed fractions in natural water samples, meaning the validity of this model is, to a large extent, dependent on the accurate determination of toxicant bioavailability in the first place. The traditional PBTK models, on the other hand, can be parameterized independently of exposure information and the prediction is accomplished partially by using a number of in vitro equilibrium partitioning coefficients between water/blood and/or tissue/blood of different organic compounds (Nichols et al, 1993). These equilibrium coefficients, however, are also influenced by bioavailability in the water, so the problem is obscured rather than absent in the traditional PBTK models. 117 There are limitations to the use of the model developed in this thesis. Any of these limitations are also applicable to other PBTK models. Whether the purpose is to predict environmental residues or to acquire insight into toxic mechanisms, bioconcentration must be viewed as a dynamic process with competing rates of uptake and depuration. Factors affecting either one of these two kinetic processes will effect changes to the overall chemical accumulation within the fish body. The relationships between fish toxicant uptake/depuration and fish oxygen consumption were developed using target compounds resistant to bio-transformation, and assuming fish gills are the sites where toxicant transfer is limited. Toxicant loss from the fish can be achieved through routes other than breathing for metabolized chemicals (Southworth et al, 1980; Southworth et al, 1981), as discussed earlier. As a result, the model reported here is not adequate for predicting tissue residues of substances that are biotransformed once they have entered the fish. From an environmental standpoint, however, persistent chemicals (i.e., chemicals that are not biotransformed) are often more toxic and generally have a high priority in bio-residue monitoring. These chemicals that resist biotransformation can have a persistent and adverse impact on the aquatic ecosystem, so the prediction of their levels in fish is important. The model reported here is the means to achieve this goal. Despite all the aforementioned restrictions and preconditions associated with this newly developed physiological model, its main advantage over the other compartmental or physiological models lies in the fact that the prediction is based on actual physiological processes and fish oxygen consumption rate is easier to measure with acceptable accuracy than many physiological parameters used in other models (Hayton and Barron, 1990). 118 Furthermore, oxygen consumption rate data are available through OXYREF (Thurston and Gehrke, 1993) provided the fish body weight is known. The rationale behind this study was to combine the advantages of both compartmental and physiological models by incorporating physiological processes into a compartmental model, using an easily accessible physiological parameter correlated with the rate of toxicant transfer, i.e., fish oxygen consumption rate, to predict chemical concentration in fish with acceptable accuracy. The model developed in this thesis possesses some functional reality which enables more realistic predictions, and is convenient and practical for application in biomonitoring and toxicant risk assessment. 119 APPENDIX Sublethal Toxicity of the Test Compounds INTRODUCTION A series of experiments were carried out to determine the effects of TeCB and TeCG on fish. The effect of these toxicants on the ability of trout and coho salmon to withstand a seawater challenge was investigated because pulp mills are often situated close to estuaries and salmonids are euryhaline. Na+/K+-ATPase, Plasma [Na+] hematocrit and muscle moisture content were measured in a search for possible biomarkers, normally viewed as sensitive indicators of exposure, response, and/or sensitivity (Weeks, 1995), to indicate the effect of these toxicants on osmoregulatory ability in these fish. Since environmental exposures are, more often than not, at low levels, the uptake and depuration tests were conducted at concentrations 5-10 times lower than the LC50 values for the specific chemicals. All sublethal tests were focused on TeCB and TeCG, whose toxicity is much higher than DBD due to the extent of chlonnation. 121 MATERIAL AND METHODS Exposure regime: Fish used in this study were adult rainbow trout (Oncorhynchus mykiss) (56.39 ± 1.49g) and coho salmon (Oncorhynchus kisutch) (25.35 ± 0.65 g), obtained from Westbrook Trout Farm in Aldergrove and Capilano Hatchery in North Vancouver, British Columbia, respectively. The fish were held in dechlorinated freshwater at 12°C in the Zoology Department, The University of British Columbia, before being transported to Bamfield Marine Station, Vancouver Island. All fish were left in a freshwater tank for one week after transportation in order for them to recover from transportation stress. Once the experiment started, the fish being tested were held in 70 L plastic garbage cans sitting in three large square tanks full of continuously running water to minimise temperature fluctuations. Control fish were also held in these cans to preclude the effect of plasticizers. There were three toxicant treatment groups and two control groups. The freshwater tests were as follows: (1) Rainbow trout were exposed to 200 mg TeCBT1, 200 mg TeCGT1, separately, and coho salmon were dosed at 200 mg TeCGl"1 for 3 days with the exposure solution partially changed (20 L-d"1) to ensure adequate toxicant loading. The dosed fish were then kept in clean freshwater during the 8-day depuration period; (2) Coho salmon were exposed to 100 jj.g-1"1 TeCG for 11 days with the exposure solution partially changed (20 L-d" '); (3) Rainbow trout and coho salmon control groups were maintained in freshwater with no toxicant for 11 days. The seawater challenge tests were as follows: the fish followed the same exposure 122 process as in the freshwater group, i.e., 3 day toxicant exposure in freshwater, but were transferred directly to 307oo seawater after toxicant exposure termination. Fish in the control group were simply held in toxicant free freshwater, with 20 L water changed every day in the first three days of the experiment, and then were transferred to 307oo saltwater for the seawater challenge test, which lasted for 8 days. Fish in all groups were put in aerated plastic garbage cans with either freshwater (except in the continuous exposure group) or seawater overflowing at a turnover rate of 2 h during depuration or seawater challenge periods. The experiment lasted for three weeks and the mean water temperature was 13.1 ± 1.6 °C. Sampling procedure: Before exposure started, six control fish were sampled and on each successive sampling day six fish were also sampled both in the control and the toxicant treated groups. The sampling days were as follows: exposure-day-3 (d3) in all groups; depuration-day-1, 3, 8 (ddl, dd3, dd8) in seawater challenge groups and all freshwater groups with the exception of the continuous exposure group where fish were sampled at exposure-day-4, 6 and 11 (d4, d6, dll). Fish sampling was achieved by terminal anaesthetisation with MS 222 (200 mgT1) (Aqualife TMS, Syndel Laboratories). Anaesthetisation prior to sampling has been demonstrated to have no effect on plasma ion concentrations (Blackburn and Clarke, 1987). Blood was sampled through the caudal ventral vein using heparinized 100 pi syringes and 60 pi microhematocrit tubes. At the same time duplicate 20 pi blood samples were also taken into small test tubes containing 10 ml Drabkin's solution for later hemoglobin measurement. The microhematocrit tubes were then centrifuged at 11,500 rpm for 3 min in a 123 Damon IEC MB microhematocrit centrifuge and hematocrit (Hct) was measured in quadruplicate. Fish were then weighed and dissected. Approximately 0.1 g gill and 0.3 g kidney tissues were taken from each fish and maintained in Eppendorf tubes with 500-ul SEI solution (0.3 M sucrose, 0.02 M EDTA, 0.1 M Imidazole) for the Na+/K+-ATPase assay (Zaugg, 1982). About 1 g white muscle was removed from the same place in each coho salmon and weighed before and after being dried to constant weight in the oven. The fish carcass, tissue samples and plasma were stored at -80°C for later analysis. Analytical techniques: Blood hemoglobin concentrations were determined by analysing the hemolyzed blood and Drabkin's solution mixture at 540 nm using a Shimadzu UV-160 visible recording spectrophotometer. Plasma sodium concentrations were measured with a Perkin-Elmer model 2380 atomic absorption (aa) spectrophotometer. The plasma held in microhematocrit tubes under -80°C prior to ion level determination was thawed and diluted to within the linear range of the machine's detection. Gill and kidney Na+/K+ ATPase activity assays were conducted according to procedures described by Zaugg (1982). Statistical analysis: All data are expressed as mean ± standard error. Statistical differences between treatment and control groups were determined using parametric one way ANOVA followed by a Dunnett's test for freshwater groups, and student 7-test for seawater groups. Under 124 conditions where the equal variance test failed, the non-parametric equivalent was used as appropriate. A probability level of 0.05 was chosen as the limit of statistical significance. 125 RESULTS Rainbow trout did not tolerate the transfer from freshwater directly to full strength seawater. No trout, including controls, survived more than three days in saltwater. No mortality was recorded in the rainbow trout control group, however, 24 h after seawater transfer while 5.5% of the trout died in the groups pre-exposed to TeCB and TeCG. On day-3 in seawater, the mortality increased drastically in all groups, reaching 50, 83 and 100% in the control, TeCB, and TeCG groups, respectively. Although the test was intended to be sublethal, these results indicate that TeCB and TeCG may impair the ability of rainbow trout to withstand any subsequent seawater challenge. Coho salmon, on the other hand, were much better at dealing with hyperosmotic stress. It was not surprising to see all the TeCG-exposed coho survived the transfer well into the end of the 8-day depuration period in clean seawater. Haematology: Rainbow trout showed a gradual, but not significant, decline in haematocrit and haemoglobin content during depuration in freshwater (Tab. 3) after a three-day sublethal exposure to 200 mg-1"1 TeCB or TeCG. The same pattern was observed with coho salmon throughout the 3-day and 11-day 100 M-g'l"1 TeCG exposure (Tab. 4) and the 8-day post-exposure period in clean seawater (Fig. 23). 126 Table 3. Effects of TeCB and TeCG on the haematocrit (Hct) (%), haemoglobin content (Hb) (g-dl"1), mean cell haemoglobin concentration (MCHC) (100'Hb-Hcf1), plasma sodium level (mmoM'1) and gill, kidney Na+/K+-ATPase activity (pimol P, -mg protein-h"1) of adult trout (Oncorhynchus mykiss) during 200 (J-g-1"1 TeCB or TeCG exposure and depuration in freshwater. 127 •s s so o r-H co ON C N O in NO r-O •<* IT) C N ro NO SO r- O so 00 in .10+1. 00 +1 I-H ro 00 +1 ro "°. •<*' +1 1^  ro ro +1 ON 00 +1 o co ?; C N r~ +i o\ o +i C N ro i—I +1 o 00 +1 SO i—i co +1 co O +i SO SO ro' +1 <N O 1^-C N in ON co C N 00 in C N as C N ro 00 CO ^ "* r-1 ^ O N oo r~ oo ? ?; +i ?; ? "1 so C N 00 r-H C N C N oo in ro © oo t— so so rr1 C N co ON O O 00 I-H >* § +; ?; ?; ?; , -* CN V> CN csi (N "* CN r--°^ so CN CN * * C O so O N SO SO so so 3 5 ? 3 +U •n r-H rH ON S C N ON r~- WI CO J O SO •*' m \© £ o e E. Z S3 s o C N 00 O N r- -* C N in C N CO 00 in so sq C N r-H ,—' <n rH +1 +1 +1 +1 +1 +1 +1 +1 in CO o >n C N •* 00 in in 00 SO 00 CO r~ © O N SO CO r-' so C N CO ro ro CO CO CO l-H rH i—i rH i-H rH i-H I-H * ON 00 o 00 * so >n r-- rH so >n >n 00 +1 +1 +i m so >n 00 >n 00 r-' ON SO CO CO CO (-H rH I—1 o ON CN 1^  +) +i <n CN CN m n u n u © © © - H so so <n co in oo O H H C N co in oo SO 00 ON 00n-O rH O rH so so in co in oo o •* ON SO CN 00 C N O N O re in O CN so ' CO CN 00 Tt" >n co 5 5 5 ? ? 5 5 5 5 5 5 5 5 5 500 H^ CO I in ir SO i •* 'i a o o •a w NT 4} a. E t /5 00 SO CN ON O N r-o ro O 00 so so SO I-H o o 00 00 00 SO ro so •* ro o ro ON ,94±0 +1 ON in ? in o I-H +1 O N CN +1 ro O ? r^ 00 i-H +1 r~->n i—i +1 O N CN ?; O N CO O N ?; SO 00 i—i +1 SO 00 +1 CO ?; CO CN o CN CN 00 r-H SO l-H r~' o CN 00 I-H SO l-H so' 1—1 O CN in rH CN rH 1—1 CO H^ O N CO o SO O N in ro O N O o O in "* O N CO SO ON O N in ro 00 .08+1 CO +1 CO 00 +! SO ro +1 00 00 +i O N 00 +1 00 o +i in 00 ON ro +1 SO I-H +1 1^  in +1 00 o l-H +1 CN CN +i ro sq ?i ro 00 in ON CO r-' ro CN CO ro CO CN CO CN CO o in ro r-' CN o CO T J T3 o U I o T ) i c o I s o -rJ -rJ ^ a U P C H <o T J O OH g, CJ T J OH O T J T J H rH CO 00 m >. i aj & T J T J T J T J c B CJ O o ,o rH 'H-J '•*-» s CS cd cs rH rH rH o P P P OH OH OH OH X CJ CJ t> CJ T J T J T J •a s o T J T J i U rH P o X C J l-H CO 00 & >N cd cd T J T J T J JS c P _o O O •a '43 cd cd cd tH rH rH P P P O , OH OH CJ CJ CJ T J T J T J CA O o o p CJ cd CJ U P CJ CJ CJ -o CJ CJ rH !9 T J cd o '2 OJO cd o T J P CJ r5 128 Table 4. Effects of TeCG on the haematocrit (Hct) (%), haemoglobin content (Hb) (g'dT1), mean cell haemoglobin concentration (MCHC) (100 Hb-Hcf 'Xplasma sodium level (mmolT1), muscle moisture (%) and gill, kidney Na+/K+-ATPase activity (jlmol P,-mg protein-h"1) of adult coho salmon (Oncorhynchus kisutch) during 100 pigi"1 TeCG exposure and depuration in freshwater or continuous exposure for 11 days in freshwater. 129 T 1 3 S 3 w ft jO T3 J3 -g s n .3 S S u 3 o ft <-> S o o o O >—1 X *° £1 at s * •a £ 03 no O -S a 8 I t 3 3 u -3 o o ca a I -r o 'a ^ to s bp a a 0 •_ •S o 1 * . O ID ^ S J2 ^ <D ca .a J= a o a o H 3 3 -a § 1 •s 1 0 ft g g 1 O CS i f £ s 3 + C « z « 2 W • i f o •a W n a, t/3 NO •«t ON NO Tl-ON r-_ C N co NO •* ON 00 NO r-in co i—i NO ON o 00 o C N O NO O .5912 +1 +-; o ON co +1 00 O +1 i—i +1 Os VI +i o r-_ C O +1 00 +1 NO in +i C N O N <r) ^ ?i o o O m +i NO 00 +1 >n in in C N ON NO C N r-C N 00 C N in C N co co C N co NO «n C N C O O i—i C N ON >n 00 >n C N 00 C N C O o •<* 00 in * C N 00 o NO C O 00 in * C N 00 in ON in 0+86 < N +1 C O 00 +1 o NO +i o ON ? in ON +i ON in C N ?; Os ?; in ON +i ON ? l o ?; C N O N C O in Tt o in C N o in C N C N NO cn o NO C N NO C O NO * in C N 00 O N O in NO C O r~ NO * 00 in NO ON C N © .87±2 +1 O +1 C N NO >n +1 C N C O in +1 C N 00 +J 00 « +i o o NO +1 C N • * •<*' +1 C O r- o 00 TJ-NO +1 C O +1 O NO C O +1 in oo r-(*N NO m C N C O C O C O C O t~ C N O N O i—1 C O C N C O C N C N O C N C N in NO OS cn C O C O r-00 NO NO * O N NO o NO ON C N NO NO * .13+1 +1 00 ON +1 Os 7—i +1 C O ?; C N +1 C O 5 o C O ?; o +\ NO r~ +i NO o +i C O r-H o C O ?; C N +i O N r-+1 in in cn NO o NO o NO NO NO C O NO o NO NO 00 in NO NO C O NO o NO in NO 00 NO NO C N O N cn 00 NO ON o C N * * 00 in o o C N 00 ON ON NO C O .97+2 C N +1 >n +1 o o +i NO C O +1 1—1 1—1 C N +1 1^ ON +1 +1 NO C O +1 00 00 7—i +1 o >n +i ON +1 ?; in 00 +1 m +1 C N t~ >n <—i NO o NO ("-" in © NO r--in ON in o NO C N in in r-' m ON in C N NO © NO in in t~ cn >n 00 1—1 NO 00 in r~ C O o ON in OS •>* O N r-co o C O © m in i .77±0 ?; ?; in ?; C N NO ?; ON ?! r~ +i C O 00 ?i 00 o ?; Os ? i 00 00 ?; +1 C O 00 ?; C N O ? l Tf ON ?; i NO • C O ! 00 Os 00 00 oo' 00 i> ON in NO 00 p~' 00 NO i NO C N C N 00 o NO OS I-o NO C N o 00 o ON in in Tf NO C N o 00 C O r- i .67±1 +i cn cn ?; in C N +1 NO ?; o in +1 NO 1—1 +1 C O 00 ?; NO ?; +1 C O C O +i r-NO +i C O 00 +1 NO ?; NO +1 i co ; 00 : C N cn C N cn C N C O C N C O C O C N C O ON C N C O o C O ON C N C N C O ON C N 00 C N ON C N ON i C N i 9 S 3 >> ft ft T3 D T3 O u e o I PH O u 3 ft T 3 I O u-3 ft •a T3 u H a 3 O X CS •a C O 00 CS ca ca T3 V V el 3 3 o o o •& ca ca CS u l-3 3 3 ft ft ft <L) u •a T3 •a C O ^ > NO £^  CS cS CS V •a i kH 3 3 3 © CO • O o o & ft ft CS X X X T3 CL) ID T 3 O X a) o & o u a H 3 O s B O u 3 O ft H V m © © V & 3 ° t cs 3 o '•e CS ft •s 3 o 1 o o 3 I a 3 (L) I D CJ o 3 CJ s ca o ia '3 0J3 ta T 3 a CL> O 2 Figure 23. Effects of TeCG on coho salmon haematological factors. Day 0-3 is freshwater exposure (100 ug-1"1) and day 3-8 is clean seawater depuration. 131 Time (day) Plasma sodium and muscle moisture: Rainbow trout treated with TeCB or TeCG showed no significant change in plasma sodium levels until after being held for 3 days in clean freshwater following toxicant dosing, when a decrease was observed throughout the rest of depuration in both the TeCB and TeCG groups (Tab. 3). Plasma sodium concentration was fairly stable in the freshwater coho salmon groups until 4 days after the beginning of toxicant exposure, when a significant decrease (p<0.05) was seen (Tab. 4). Rainbow trout plasma sodium levels were elevated in TeCB and/or TeCG predosed fish given a seawater challenge (Fig. 24). Plasma [Na+] in coho salmon exposed to TeCG was not significantly higher (p<0.05) than that in control group fish until 8 days after being transferred to toxicant free seawater (depuration-day-8) (Fig. 25). Muscle moisture data were collected from coho salmon and a significant water gain in muscle was found only at depuration-day-1 and exposure-day-4 and 6 in freshwater (Tab. 4). There was a non significant trend of increased muscle water content when the coho were transferred to and then held in seawater (Fig. 25). Na/K*-ATPase: A significant decrease in gill Na+/K+-ATPase was observed in freshwater groups at the end of the 3-day exposure and ATPase activity remained low until depuration-day-1 in TeCB and depuration-day-3 in TeCG dosed rainbow trout, respectively (Tab. 3). There was a tendency for recovery of ATPase activity toward the end of freshwater depuration, which was also apparent in the TeCG exposed coho salmon either post exposure, or during the continuous exposure period in freshwater (Tab. 3 & 4). In the 24 h seawater transfer test, 133 Figure 24. Effects of 3-day TeCB and TeCG exposure (200 (igT1) in freshwater on plasma sodium level and gill Na+/K+-ATPase activity of rainbow trout during 24 h seawater challenge. Asterisk indicates a significant difference (p<0.05). 134 P O C/5 tZ3 PM o £ = 250 200 150 100 50 0 1 I SI H U < es .0-H + d "53 O o. WD a 12 10 8 6 4 2 0 A. I I I CONTROL TCB I TCG Exposure day-0 hNNXI Exposure day-3 24 hours after seawater transfer 135 Figure 25. Effects of TeCG on coho salmon plasma sodium level and muscle moisture content. Day 0-3 is freshwater exposure (100 (igT1) and day 3-8 is clean seawater depuration. Asterisk indicates a significant difference (p<0.05). 136 Time (day) rainbow trout gill Na /K -ATPase remained inhibited after being held one day in seawater no matter whether the fish were exposed to TeCB or TeCG (Fig. 24). An upregulation of this enzyme in coho salmon gill, however, was found at the end of 3-day TeCG exposure (p<0.05) and no significant difference was detected at other sampling times (Tab. 4, Fig. 26). The kidney Na+/K+-ATPase activity in rainbow trout was not influenced throughout the whole test (Tab. 3). TeCG exposed coho salmon, on the other hand, showed an inhibition in kidney Na+/K+-ATPase at the end of either the 8-day depuration seawater or after being continuously exposed to TeCG for 11 days in freshwater (Tab. 4, Fig. 26). 138 Figure 26. Effects of TeCG on coho salmon Na /K+-ATPase activity in gill and kidney. Day 0-3 is freshwater exposure (100 (igi"1) and day 3-8 is clean seawater depuration. Asterisk indicates a significant difference (p<0.05). 139 0 2 4 6 8 10 12 Time (day) DISCUSSION Both increases and decreases in haematological factors have been observed in response to environmental pollutants (Heath, 1987). In this study, however, neither rainbow trout nor coho salmon demonstrated a significant change in either haematocrit or haemoglobin levels. The calculated mean cell haemoglobin concentration was fairly stable throughout the whole experiment for both species, indicating no net loss of red blood cells from the circulating blood. As a result, the oxygen carrying capacity of the dosed fish should not be altered and it is, therefore, unlikely that the gas exchange in the gill was affected. This is in agreement with the observation that there is no impact of sublethal exposure to TeCB on fish swimming performance (Brauner et al, 1994). Although haematological factors are usually easy to measure compared with other parameters, their varied and reduced response to sublethal exposure to TeCB or TeCG exclude the possibility of them being used as effective biomarkers in this case. In freshwater, the changes in plasma [Na+] in coho salmon, although exposed at half the concentration of toxicant to which rainbow trout were exposed, were more rapid than that observed in trout. If the toxicant-exposed fish were transferred to seawater, rainbow trout showed elevated plasma sodium 24 h after seawater transfer while no change was seen in coho salmon until they had been held in clean saltwater for 8 days, suggesting coho salmon appeared to be more capable of maintaining the plasma ion levels when subjected to a hyperosmotic environment (Fig. 24 & 25). For freshwater tests, therefore, coho salmon indicates more sensitivity than rainbow trout if plasma ion concentration is used as a 141 biomarker; however, rainbow trout show a greater osmoregulatory impairment if seawater challenge tests are chosen. Plasma samples are easy to collect and relatively convenient and less costly to measure compared with other physiological parameters, such as enzyme activity. A wide range of organic xenobiotics are capable of affecting fish active ion transport (Evans, 1987). The inhibition of gill Na+/K+-ATPase has been shown in different fish exposed to a variety of organic chemicals. Davis and Wedemeyer (1971) reported that the organochlorines, DDT, dicofol, and endosulfan inhibited rainbow trout gill Na+/K+-ATPase by 60 to 100% in vitro at concentrations between 10"5 and 10"4M In this study, the response time for both species in freshwater was similar except that coho recovered a little faster than trout (Tab. 3 & 4), although the toxicant-induced change was in the opposite direction; gill Na+/K+-ATPase for rainbow trout exposed to a nominal concentration of 200 pigT1 TeCB or TeCG for 3 days showed an inhibition, while an upregulation was found in coho salmon after being dosed with 100 (ig'l1 TeCG for the same amount of time. An inhibition of Na+/K+-ATPase was observed in both TeCB and TeCG treated trout 24 h after the seawater challenge (Fig. 24). The enzyme activity in coho gill recovered after transfer to seawater (Fig. 26), as in freshwater (Tab. 4). These findings agree with the aforementioned observation that coho salmon, as smolts, are less sensitive during freshwater to seawater transition than rainbow trout. Elevated growth of juvenile coho salmon exposed to kraft pulpmill effluent was reported and it was found to be dose-dependent (Mason and Davis, 1976). Although there was no direct evidence, a variety of hypotheses to explain this effect were proposed, including hormone analogs in the wood extracts that stimulate growth hormone and/or appetite. Another possibility is that hormesis occurred (Sansone and Martens 1981). The term "hormesis" refers 142 to an overcompensation to some inhibitory challenge. The stimulation seen in the gill Na+/K+-ATPase activity of coho salmon could be described as a form of hormesis. The kidney also plays an important role in osmoregulation, particularly in freshwater fish where glomerular filtration is high and extensive reabsorption takes place. According to the results of this study, kidney Na+/K+-ATPase in rainbow trout was largely unaffected, whereas the enzyme was inhibited 3 days into seawater depuration and at the end of continued freshwater exposure in the TeCG treated coho salmon, although the response was much delayed in the kidney compared with gill (Tab. 4, Fig. 26). The inhibition of the kidney Na+/K+-ATPase was similar in freshwater and seawater fish, indicating that the reduction was a direct effect of the toxicant on the ATPase rather than any secondary response to osmotic and ionic changes. It would seem that coho kidney Na+/K+-ATPase is more responsive than that of trout. The parameters monitored are directly or indirectly related to each other. The inhibition of gill Na+/K+-ATPase in rainbow trout could have been the cause for the subsequent fall in plasma ions. For TeCG exposed coho, however, the increase in plasma [Na+] was associated with no change in gill Na+/K+-ATPase activity and a decrease in kidney Na+/K+-ATPase activity, which should have resulted in a drop in sodium level, rather than the observed increase (Fig. 25 & 26). Also, the upregulation of gill Na+/K+-ATPase was not consistent with the decrease in plasma [Na+] recorded in the freshwater TeCG treated coho (Tab. 4). It would appear, therefore, that the changes in plasma sodium are not brought about by the changes in Na+/K+-ATPase in the body but are due probably to increased passive flux of water and/or sodium, induced by changes in gill membrane permeability. Coho salmon muscle moisture 143 content increased at about the same time as plasma sodium levels decreased in the freshwater groups (Tab. 4), indicating that in freshwater the reduction in plasma ion concentration (hemodilution) is associated with an increased water content in muscle. That is, the dilution is general and water entering the body is distributed to both blood and muscle tissues. McCarty (1995) has argued that biomarkers, i.e., changes in physiological or biochemical parameters, can become a useful and scientifically valid tool if a biomarker-biological indicator relationship is established. In general, plasma sodium level, gill Na+/K+-ATPase activity, and/or muscle moisture are potential biomarkers. Na+/K+-ATPase in gill is a better choice as a biomarker than that in kidney and its sensitivity to the toxicant is even higher than that of plasma [Na+]. The enzyme assay, however, is more time consuming and costly relative to ion measurements. Nevertheless, the efficiency of Na+/K+-ATPase analysis has been greatly enhanced by the development of a microassay method (McCormick, 1993), and could possibly serve as a valuable, easy-to-apply biomarker. Based on this study, muscle moisture content is indicative of disturbed ion balance in coho salmon to the same degree as the direct measurement of plasma ion content, and as a result it could be used as a less expensive biomarker. Muscle moisture, plasma ion level and ion transport enzyme activity, although functionally linked, do not always co-vary. In addition, all these parameters show considerable normal variability. Thus, based on this study, the potential use of enzyme activity, plasma Na+ levels or muscle moisture as biomarkers is of limited value. In the Fraser River, hundreds of thousands of salmon are undergoing smoltification each year, and the ability to osmoregulate will certainly determine fish survival rate after seawater transition. Increased sodium concentration in the plasma has 144 been found to affect the critical swimming speed Ucrit in coho salmon (Brauner et al, 1992). Assuming fish aerobic swimming performance is impaired due to the failure to maintain ion balance after going into the ocean, their survival could be at stake not only due to osmoregulatory failure but also by their being less capable of escaping from potential predators. The parameters measured in these experiments are valuable, not as potential biomarkers, but rather because they indicate that both TeCB and TeCG impair osmoregulation. Such an impairment could have a detrimental effect on the size of the fish population. The impact of these toxicants might be minimised if the exposed fish are left for several days to recover in freshwater before they head for the ocean. Consequently, location of pulp mills to either freshwater or seawater sites, but away from the estuary could potentially reduce the adverse effects of TeCB and TeCG on the local fish population. 145 LITERATURE CITED Andersen, M.E., H.J. Clewell, M.L. Gargas, FA. Smith and RH. Reitz. 1987. Physiologically based pharmacokinetics and the risk assessment process for methylene chloride. Toxicol. Appl. Pharmacol. 87: 185-205. Barron, M.G. 1990. Bioconcentration. Will water-borne organic chemicals accumulate in aquatic animals? Environ. Sci. Technol. 24(11), 1612-18. Barron, M.G., Stehly, G.R. and Hayton, W.L. 1990. Pharmacokinetic modeling in aquatic animals. I. Models and concepts. Aquatic Toxicology 18:61-86. Beamish, F.W.H. 1978. Swimming capacity. In. W.S. Hoar and D.J. Randall (eds). Fish Physiology, Vol 7. Academic Press, New York, NY, pp. 101-187. Bertelsen, S.L., AD. Hoffman and CA. Gallinat. 1993. Determination of in vitro chemical partition coefficients in fish tissues: A comparative study of four species. EPA-600/J-93-045. Bishop WE. and A.W. Maki. 1980. A critical comparison of two bioconcentration test methods. Aquatic Toxicology, ASTM STP 707, J.G. Eaton et al, Eds., pp. 61-77. Black, M.C., and J.F. McCarthy. 1988. Dissolved organic macromolecules reduce the uptake of hydrophobic organic contaminants by the gills of rainbow trout (Salmo gairdneri). Environ. Toxicol. Chem. 7:593-600. Black, M.C., D.S. Millsap and J.F. McCarthy. 1991. Effects of acute temperature change on respiration and toxicant uptake by rainbow trout, Salmo gairdneri (Richardson). Physiol. Zool., (in review). Blackburn, J. and W.C. Clark. 1987. Revised procedure for the 24 hour seawater challenge test to measure seawater adaptability of juvenile salmonids. Can. Tech. Rep. Fish. Aquat. Sci. 1515: 35. Brauner, C. J. 1991. The relationship between oxygen and 1,2,4,5-tetrachlorobenzene uptake, in five species of fishes, M.Sc. Thesis, The University of British Columbia. Brauner, C.J, J.M. Shrinpton, and DJ. Randall. 1992. Effects of short-duration seawater exposure on plasma ion concentration and swimming performance of coho salmon (Oncorhynchus kistch) parr. Can. J. Fish. Aquat. Sci. Vol. 49. 11, pp.2399-2405. Brauner, C.J., D.J. Randall, J.F. Neuman and R.V. Thurston. 1994. The effect of exposure to 1,2,4,5-tetrachlorobenzene and the relationship between toxicant and oxygen uptake in rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem., Vol. 13, No. 11, pp. 1813-1820. Brett, JR. 1964. The respiratory metabolism and swimming performance of young sockeye salmon. J. Fish. Res. Bd. Canada, 21(5): 1183-1226. 147 Brett, J.R., Zala, CV. 1975. Daily pattern of nitrogen excretion and oxygen consumption of sockeye salmon (Oncorhynchus nerka) under controlled conditions. J Fish Res Board Can 32:2479-2486. Bruggeman, W.A., Matron, L.B.J.M., Kooiman, D. and Hutzinger, O. 1981. Accumulation and elimination kinetics of di-, tri and tetrachlorobiphenyls by goldfish after dietary and aqueous exposure, Chemosphere, 10(8):811-832. Burggren, W.W. and J.N. Cameron. 1980. Anaerobic metabolism, gas exchange, and acid-based balance during hypoxic exposure in channel catfish, Ictaluruspunctatus. J. Exp. Zool. 213: 405-416. Bushnell P.G., Steffensen J.F., Johansen K. 1984. Oxygen consumption and swimming performance in hypoxia-acclimated rainbow trout (Salmo gairdneri). J. Exp. Biol. 113: 225. Cameron, J.N. and Davis, J.C. 1970. Gas exchange in rainbow trout {Salmo gairdneri) with varying blood oxygen capacity. J. Fish. Res. Bd. Canada. 27:1069-1085. Chiou, C.T., V.H. Freed, D.W. Schmedding and R.L. Kohnert. 1977. Partition coefficient and bioaccumulation of selected organic chemicals. Environ. Sci. Technol. 11: 475-478. Clark, K.E., F.A.P.C. Gobas and D. Mackay. 1990. Model of organic chemical uptake and clearance by fish from food and water, Environ. Sci. Technol., 24:1203-1213. Connell, D.W. 1990. Bioaccumulation of xenobiotic compounds, CRC Press, Inc., Boca Raton, Florida. 213pp. Davis PW. and G.A. Wedemeyer. 1971. Na+, Reactivated ATPase inhibition in rainbow trout: a site for organochlorine pesticide toxicity? Comp. Biochem. Physiol. 40B:823-827. Dobbs, A.J. and N. Williams. 1983. Fat solubility- a property of environmental relevance? Chemosphere. 12(1):97-104. Duval W.S., Fink R.P., Geen G.H. 1981. An aquatic respirometer for use in sublethal studies. Water Res. 15:817-822. Eddy, F.B. 1981. Effects of stress on osmotic and ionic regulation in fish. In: A.D. Pickering (Ed.), Stress in fish: Academic Press, London, 77-102. Erickson, R.J. and J.M. Mckim. 1990. A model for exchange of organic chemicals at fish gills: Flow and diffusion limitations. Aquat. Toxicol. 18, 175-198. Evans, D.H. 1987. The fish gill: site of action and model for toxic effects of environmental pollutants. Env. Health Perspect. 71: 54-58. 148 Folch, J., M. Lees and G.H. Stanley. 1957. A simple method for the isolation and purification of total lipids from animal tissues. Journal of Biology and Chemistry 226: 497-509. Garrett, CL. 1980. Toxic Organic Contaminants, Water Quality, Fraser River Estuary Study. Government of Canada. Gehrke, P C , L.E. Fidler, D.C. Mense and D.J. Randall. 1990. A respirometer with controlled water quality and computerized data acquisition for experiments with swimming fish, Fish Physiology and Biochemistry, 8(1):61-67. Geyer, H., I. Scheunert, R. Bruggemann, V. Zitko, C.E.W. Steinberg, R. Pruell, A. Kettrup, J. Olson, P. Schmieder, D.C.G. Muir. 1995. Relationship between elimination half-lives (T1/2) of 2,3,7,8-tetrachlorodibenzo-/>-dioxin (TCDD) in different mussel and fish species and their lipid content or body size, Organohalogen Compounds, Vol.25, 273-278. Gingerich, W.H. 1986. Tissue distribution and elimination of rotenone in rainbow trout, Aquatic Toxicology, 8: 27-40. Gingerich, W.H. and J.J. Rach. 1985. Uptake, accumulation and depuration of 14C-retenone in bluegills (Lepomis macrochirus). Aquat. Tox. 6: 170-196. Gobas, F.A.P.C. 1993. A model for predicting the bioaccumulation of hydrophobic organic chemicals in aquatic food-webs: application to Lake Ontario, Ecological Modelling, 69:1-17. Gobas, F.A.P.C. and Mackay, D. 1987. Dynamics of hydrophobic organic chemical bioconcentration in fish. Environ. Toxicol. Chem., 6: 495-504. Gobas, F.A.P.C, Opperhuizen, A. and Hutzinger, O. 1986. Bioconcentration of hydrophobic chemicals in fish: relationship with membrane permeation. Environ. Toxicol. Chem., 5:637-646. Goss, G.G. and Wood, CM. 1988. The effects of acid/aluminum exposure on circulating plasma Cortisol levels and other blood parameters in the rainbow trout, Salmo gairdneri, J Fish Biol. 32: 63-76. Hamelink, J.L. and Spacie, A. 1971. A proposal: exchange equilibria control the degree chlorinated hydrocarbons are biologically magnified in lentic environments, Trans. Am. Fish. Soc, 100: 207. Hawker, D.W. and Connell, D.W. 1985. Relationship between partition coefficient, uptake rate constant, clearance rate constant and time to equilibrium for bioaccumulation, Chemosphere, 14(9): 1205-1219. Hayton, W.L. and M.G. Barron. 1990. Rate-limiting barriers to xenobiotic uptake by the gill. Environ. Toxicol. Chem. 9:151-157. 149 Heath, A.G. 1987. Water pollution and fish physiology. CRC Press Inc., Boca Raton. Hoar, W.S. and Randall, D.J., Eds. 1984. Fish Physiology, Vol. 10 (Part B), Gills: Ion and Water Transfer, Academic Press, New York. Hocutt, CH. and Stauffer, J.R., Eds. 1980. Biological Monitoring of Fish, Lexington Books, Lexington, KY. Holden, A.V. 1962. A study of the absorption of 14C-labelled DDT from water by fish. Ann. App. Biol. 50, 467-477. Hughes, G.M. 1973. Respiratory responses to hypoxia in fish, Am. Zool. 13: 475. Janz, DM., Farrell, A.P., Morgan, J.D., and Vigers, GA. 1991. Acute physiological stress responses of juvenile coho salmon (Oncorhynchus kisutch) to sublethal concentrations of Garlon 3A and Vision herbicides, Environ. Toxicol. Chem. 10: 81-90. Johansen, J.A. and Geen, G.H. 1990. Sublethal and acute toxicity of the ethylene glycol butyl ether ester formation of triclopyr to juvenile coho salmon (Oncorhynchus mykiss), Bull. Environ. Contam. Toxicol. 19:610-616. Jones, D.R. and Randall, D.J. 1978. The respiratory and circulatory systems during exercise. In: Fish Physiology. (Hoar, W.S. and D.J. Randall, Eds.). Vol 7. Academic Press, New York. pp. 425-501. Josephson, J. 1982. Humic substances. Environ. Sci. Technol. 16: 20A-24A. Kathryn E C , A.P.C. Gobas and D. Mackay. 1990. Model of organic chemical uptake and clearance by fish from food and water, Environ. Sci. Technol., 24: (8) 1203 - 1213. Kenaga, E.E. and C.A.I. Goring. 1980. Relationship between water solubility, soil sorption, octanol-water partitioning, and concentration of chemicals in biota. In: J.G. Eaton, P.R. Parrish, and A.C. Hendricks, eds., Aquatic Toxicology, ASTM ATP 707. American Society for Testing and Materials, pp. 78-115. Kiceniuk, J.W. and Jones, D.R. 1977. The oxygen transport system in trout (Salmo gairdneri) during sustained exercise. J. Exp. Biol. 69:247-260. Landrum, P.F., Reinhold, M.D., Nihart, S.R. and Eadie, B.J. 1985. Predicting the bioavailability of organic xenobiotics to Pontoporeia hoyi in the presence of humic and fulvic materials and natural dissolved organic carbon. Environ. Toxicol. Chem., 4: 459-467. 150 Leach, J.M. and A.N. Thakore. 1975. Isolation and identification of constitutes toxic to juvenile rainbow trout (Salmo gairdneri) in caustic extraction effluents from kraft pulp mill bleach plants. Journal of the Fisheries Research Board of Canada 32: 1249-1257. Lopez-Avila V. and R.A. Hites. 1980. Organic compounds in an industrial wastewater. Their transport into sediments. Environ. Sci. Technol. 14: 1382-1390. MacKay, D. 1982. Correlation of bioconcentration factors. Environ. Sci. Technol. 16: 274-278. MacKinnon, D.L. and Farrell, A.P. 1992. The effect of 2-(thiocyanomethylthio) benzothiazole on the juvenile coho salmon (Oncorhynchus kisutch): sublethal toxicity testing, Environ. Toxicol. Chem. 11:1541-1548. Mason B.J. and J.C. Davis. 1976. Growth stimulation and biochemical changes in juvenile coho salmon Oncorhynchus kisutch during chronic exposure to sublethal levels of neutralized bleached kraft mill waste. In: Davis J.C. (Ed.) Progress in sublethal effect studies with kraft pulpmill effluent and salmonids. J. Fish. Res. Bd. Can. 33:2031. Means, J.C, S.G. Wood, and J.J. Hassett and W.L. Beanwart. 1980. Sorption of polynuclear aromatic hydrocarbons by sediments and soils. Environ. Sci. Technol. 14: 1524-1528. McCarty, L.S. 1995. Environmental biomarkers: fiction, fantasy, or functional? The 2nd SETAC World Conference, Abstract book. p. 39. McCarthy, J.F. and Jimenez, B.D., 1985. Reduction in bioavailability to bluegills of polycyclic aromatic hydrocarbons bound to dissolved humic materials. Environ. Toxicol. Chem., 4: 511-521. McCormick, S.D. 1993. Methods for nonlethal gill biopsy: Measurements of Na+/K+-ATpase activity. Can. J. Fish. Aquat. Sci. Vol. 50. pp. 656-658. McKim, J., Schmieder, P., and G. Veith. 1985. Absorption dynamics of organic chemical transport across trout gills as related to octanol-water partition coefficient. Toxicol. Appl. Pharmacol. 77:1-10. Murphy, P.G. and J.V. Murphy, 1971. Correlation between respiration and direct uptake of DDT in the mosquito fish Gambusia affinis. Bull. Environ. Contam. Toxicol. 6, 581-588. Nichols, J.W., J.M. McKim, G.J. Lien, AD. Hoffman, S.L. Bertelsen and CA. Gallinat. 1993. Physiologically-based toxicokinetic modelling of three waterborne chloroethanes in channel catfish, Ictahuruspunctatus. Aquatic Toxicology, 27: 83-112. Neely, W.B., D.R. Branson and G.E. Blau. 1974. Partition coefficient to measure bioconcentration potential of organic chemicals in fish. Environ. Sci. Technol. 8: 1113-1115. 151 Randall, D.J. 1982. The control of respiration and circulation in fish during exercise and hypoxia, J Exp Biol 100: 275. Randall, D.J. and C.J. Brauner. 1990. Toxicant uptake across fish gills. In: Proceedings of the second international symposium on Fish Physiology, Fish Toxicology, and Water Quality Management. Sacramento, California, September 18-20, 1990. United States Environmental Protection Agency, Environmental Research Laboratory, Athens, Georgia, USA. Randall, D.J. 1990. Control and co-ordination of gas exchange in water breathers. In: Advances in comparative and environmental physiology. (R.G. Boutilier, Ed.). Vol. 6. Springer-Verlag, Berlin, Heidelberg, pp 253-278. Randall, D.J., C.J. Brauner and R. Yang, 1993. Uptake of chemicals by fish, In: The Aquatic Resource Research Project Technical Report, Edited by A.P. Farrell, Simon Fraser University, pp. 181-194. Randall, D.J. and C. Daxboeck, 1984. Oxygen and carbon dioxide transfer across fish gills. In: Fish Physiology, Vol. 10A: 263-307. W.S. Hoar and D.J. Randall(Eds). Academic Press Inc., New York. Reinert, R.E. 1972. Accumulation of Dieldrin in an alga (Scenedesmus obliquus), Daphnia magna, and the guppy (Poecilia reticulata). J. Fish. Res. Bd. Canada. 29:1413-1418. Reitz, R.H., P.S. McCroskey, C.N. Park, M.E. Andersen and ML. Gargas. 1990. Development of a physiologically based pharmacokinetic model for risk assessment with 1,4-dioxane. Toxicol. Appl. Pharmacol. 105: 37-54. Rombough, P.J. and B.M. Moroz. 1990. The scaling and potential importance of cutaneous and branchial surfaces in respiratory gas exchange in young chinook salmon Oncorhynchus tshawytscha. J. Exp. Biol. 154, 1-12. Rudd, R.L. and Genelly, R.E. 1956. Pesticides: their use and toxicity in relation to wildlife, Calif. Fish and Game, 7. Saarikoski, J., Lindstrom, R., Tyynela, M. and M. Viluksela. 1986. Factors affecting the absorption of phenolics and carboxylic acids in the guppy {Poecilia reticulata). Ecotoxicol. Environ. Safety. 11:158-173. Sansone, F.J. and C.S. Martens. 1981. Hormesis: A response to low environment concentrations of petroleum hydrocarbons. Sci. Vol. 211. pp.705-709. Schmieder, PK. and T. A. Henry. 1988. Plasma binding of 1-butanol, phenol, nitrobenzene and pentachlorophenol in the rainbow trout and rat: a comparative study. Comp. Biochem. Physiol. 91C(2):413-418. 152 Shaw, G.R. and Connell, D.W. 1982. Factors influencing concentration of polychlorinated biphenyls in organisms from an estuarine system, Aust. J. Mar. Freshwater Res., 33: 1507. Southworth, G.R., C.C. Keffer and J.J. Beauchamp. 1980. Potential and realized bioconcentration. A comparison of observed and predicted bioconcentration of amaarenes in the fathead minnow (Pimephales promelas). Environ. Sci. Technol. 14: 1529-1531. Southworth, G.R., C.C. Keffer and J.J. Beauchamp. 1981. The accumulation and disposition of benz(a)acridine in the fathead minnow, Pimephales promelas. Arch. Environ. Contam. Toxicol. 10: 561-570. Spacie, A. and J.L. Hamelink, 1982. Alternative models for describing the bioconcentration of organics in fish. Environ. Toxicol. Chem. 1, 309-320. Spacie, A., P.F. Landrum and G.J. Leversee. 1983. Uptake and elimination of anthracene and benzo(a)pyrene in bluegill sunfish. (in press). Sullivan, K.F., E.L. Atlas and CS. Giam. 1982. Absorption of phthalic acid esters from seawater. Environ. Sci. Technol. 16: 428-432. Tarr, B.D., Mace G. Barron and William L. Hayton. 1990. Effect of body size on the uptake and bioconcentration of di-2-ethylhexyl phthalate in rainbow trout. Environ. Toxicol. Chem. Vol. 9, pp. 989-995. Thomas, RE. and S.D. Rice, 1981. Metabolism and clearance of phenolic and mono-, di-, and polynuclear aromatic hydrocarbons by Dolly Varden char. In: W.B. Vernberg et al., eds. Physiological mechanisms of marine pollutant toxicity. Academic Press, New York., pp. 425-448. Thurston, R.V. and P.C Gehrke. 1993. Respiratory oxygen requirements of fishes: Description of OXYREF, a datafile based on test results reported in the published literature. In: Proceedings of the second international symposium on Fish Physiology, Fish Toxicology, and Water Quality Management. Sacramento, California, September 18-20, 1990. United States Environmental Protection Agency, Environmental Research Laboratory, Athens, Georgia, USA. USEPA, 1994. Predicting bioconcentration and residue-based effects of xenobiotics in and on aquatic biota. Cooperative agreement CR995189. Veith, G.D., D.L. DeFoe and B.V. Bergstedt. 1979. Measuring and estimating the bioconcentration factor of chemicals in fish. J. Fish. Res. Board Can. 36: 1040-1048. 153 Wershaw, R.L. and M.C. Goldberg. 1972. Interaction of organic pesticides with natural organic polyelectrolytes. In: R.F. Gould, ed., Fate of organic pesticides in aquatic environment, advances in chemistry series III, American Chemistry Society, Washington D. C, pp. 149-158. Yang, R. and Randall, D.J. 1995. Relationships between xenobiotic uptake/depuration and fish oxygen consumption: A physiological model. Proceedings of the 4th International Symposium on Fish Physiology, Toxicology, and Water Quality Management. Yang, R. and Randall D.J. 1996a. Effects of Tetrachloroguaiacol (TeCG) on the osmoregulation of adult Coho Salmon (Oncorhynchus kisutch). Bull. Environ. Contam. Toxicol. 56: 788-795. Yang, R. and Randall, D.J. 1996b. Biomarkers for rainbow trout (Oncorhynchus mykiss) and coho salmon (Oncorhynchus kisutch) exposed to 1,2,4,5-tetrachlorobenzene and 3,4,5,6-tetrachloroguaiacol, Chemosphere (In press). Zaugg, W.S. 1982. A simplified preparation for adenosine triphosphatase determination in gill tissue. Can. J. Fish. Aquat. Sci. 39:215-217. Zitko, V. and O. Hutzinger. 1976. Uptake of chloro- and bromobiphenyls, hexachloro- and hexabromobenzene by fish, Bull. Environ. Contam. Toxicol. 16: 665-673. 154 

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