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The relationship between oxygen and 1,2,4,5-tetrachlorobenzene uptake in five species of fishes Brauner, Colin J. 1991

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THE RELATIONSHIP BETWEEN OXYGEN AND 1,2,4,5-TETRACHLOROBENZENE UPTAKE, IN FIVE SPECIES OF FISHES. by COLIN J. BRAUNER B.Sc. (Hon.) The University of British Columbia, 1988. A THESIS SUBMITTED IN PARTIAL FULFILMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTERS OF SCIENCE in THE FACULTY OF GRADUATE STUDIES (Department of Zoology) We accept this thesis as conforming to the required standard THE UNIVERSITY OF BRITISH COLUMBIA July 1991 © Colin John Brauner In presenting this thesis in partial fulfilment of the requirements for an advanced degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department or by his or her representatives. It is understood that copying or publication of this thesis for financial gain shall not be allowed without my written permission. Department of The University of British Columbia Vancouver, Canada Date DE-6 (2/88) ABSTRACT The studies presented in this thesis were designed to establish whether a relationship exists between the uptake of a model toxicant, 1,2,4,5-tetrachlorobenzene (TCB) and the rate of oxygen consumption (M0 2) in several species of fishes. If so, it is possible that toxicant uptake may be predicted from M0 2's measured in the absence of the toxicant. There is no effect of environmental TCB concentrations on the oxygen consumption rate of resting or exercising adult rainbow trout, and body accumulations of the toxicant did not affect the maximal aerobic swimming velocity attained by juvenile rainbow trout. Thus, there is no direct effect of TCB exposure to complicate the relationship between oxygen and toxicant uptake at a variety of swimming velocities. The potential of estimating whole body accumulations of TCB from the toxicant concentration in plasma and 11 tissues was investigated in the anticipation of reducing the work otherwise required to measure whole body concentrations. The variability in TCB concentration of a given tissue, however, is sufficiently great to preclude the use of this method to estimate body burden. Plasma concentrations appear to be indicative of body concentrations following 2 hrs of TCB exposure, however, over longer exposure durations, this is not the case. Thus whole body concentrations can only be obtained through direct measurements. There is a relationship between the rate of oxygen consumption and the uptake rate of 1,2,4,5-tetrachlorobenzene (TCB) during initial toxicant exposure, in ii five species of fishes forced to swim at different velocities. The animals were exposed to two external TCB concentrations; however, standardizing for the external concentration was not sufficient to permit grouping of the data, likely as a result of an overestimation of the aqueous environmental TCB concentration. Because the 0 data could not be grouped, a single coefficient to predict TCB uptake from M 0 2 could not be calculated. The inclusion of the proportion of body lipid in the relationship describing toxicant uptake markedly improved the coefficient of determination, suggesting that either an elevated body lipid content increases TCB diffusivity at the gill, or that a limitation to the uptake of TCB in addition to that at the gills, may also exist at the level of the tissues. The establishment of the relationship between oxygen and TCB uptake implies a potentiality for the prediction of toxicant uptake in fish, as a function of the animals M 0 2 which can be obtained from the literature. M 0 2 for a variety of fish species, sizes and activity levels, in an assortment of environments have been compiled into a database, called O X Y R E F (Thurston and Gehrke, 1991). This could be a valuable tool for the prediction of toxicant uptake provided the relationship between toxicant and oxygen uptake is independent of fish size. Physico-chemical characteristics of xenobiotics are well documented and it is conceivable to incorporate a toxicant specific correction factor to predict the uptake of a range of xenobiotics in fish. 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 ix ACKNOWLEDGEMENTS x G E N E R A L INTRODUCTION 1 G E N E R A L MATERIALS AND METHODS 7 Xenobiotic 7 Experiments 8 Animal acquisition and acclimatization 8 UBC 9 Apparatus 9 Experimental procedure 11 MSU 12 TCB analysis 12 Calculations and statistics 15 CHAPTER ONE 17 INTRODUCTION 17 MATERIALS AND METHODS 18 Experimental procedure 18 Effect of external TCB concentration on oxygen consumption rates: a. ) At rest 18 b. ) During exercise 19 Effect of body TCB accumulations on U c r i t 19 TCB accumulations within organs and tissues 20 Calculations 21 RESULTS 21 Effect of TCB exposure on oxygen consumption rates 21 Effect of TCB exposure on U c r i , 22 Toxicant distribution between body components 23 DISCUSSION 24 iv CHAPTER TWO 41 INTRODUCTION 41 MATERIALS AND METHODS 42 Experimental procedure 42 Plasma as an indicator of body burden a. ) At rest 42 b. ) During exercise 42 Correlation between oxygen uptake and measured body burden 43 RESULTS 43 Plasma TCB as an indicator of body burden 43 Relationship between oxygen and toxicant uptake 44 DISCUSSION r. 45 CHAPTER T H R E E 61 INTRODUCTION 61 MATERIALS AND METHODS 62 Animal acquisition and acclimatization 62 Experimental procedure 62 Analytical procedure 63 Lipid analysis 63 RESULTS 65 DISCUSSION 66 G E N E R A L DISCUSSION 78 L I T E R A T U R E CITED 83 APPENDIX 88 v LIST OF TABLES Page T A B L E 1 The effect of short duration TCB pre-exposure on the body burden and U^, in juvenile rainbow trout 35 T A B L E 2 The effect of long duration TCB pre-exposure on the body burden and U^, in juvenile rainbow trout 35 T A B L E 3 The TCB concentrations of plasma and 11 tissues, in swimming adult rainbow trout exposed to one TCB concentration 37 T A B L E 4 The ratio of tissue:plasma concentrations in swimming adult rainbow trout exposed to one TCB concentration 39 vi LIST OF APPENDIX TABLES Page T A B L E 5 The precision of the TCB and surrogate analysis using the gas chromatograph 91 T A B L E 6 The precision of the tissue extraction procedure 93 vii LIST OF FIGURES Page FIGURE 1 The effect of water TCB concentration on M 0 2 in adult rainbow trout at rest 29 FIGURE 2 The effect of water TCB concentration on M 0 2 in adult rainbow trout swimming at 1.25 Bis'1 31 FIGURE 3 The relationship between plasma and estimated body TCB concentrations, in swimming adult rainbow trout 33 FIGURE 4 The correlation between plasma and body TCB concentrations in resting juvenile rainbow and cutthroat trout 49 FIGURE 5 Plasma and body TCB concentrations in swimming juvenile rainbow trout following 2hrs exposure to one of two external TCB levels 51 FIGURE 6 Plasma and body TCB concentrations in swimming juvenile rainbow trout exposed to TCB for 2 or 6 hrs 53 FIGURE 7 The relationship between M 0 2 and estimated TCB uptake in swimming, adult rainbow trout 55 FIGURE 8 The effect of M O z on TCB uptake in swimming, juvenile rainbow trout 57 FIGURE 9 The effect of M 0 2 on TCB uptake per unit gradient in swimming, juvenile rainbow trout 59 FIGURE 10 The effect of M 0 2 on TCB uptake per unit gradient, in five species of fish 72 FIGURE 11 Percent body lipid in five species of fish 74 FIGURE 12 The relationship between TCB uptake per unit gradient and the product of M O z and proportion of body lipid, in five species of fish 76 viii LIST OF APPENDIX FIGURES FIGURE 13 TCB loss from the respirometer over time 95 FIGURE 14 The effect of various treatments on water TCB concentrations in 20 1 containers 97 ix ACKNOWLEDGEMENTS I would like to express my sincere appreciation to Dr. Dave Randall and Dr. Vance Thurston for guidance and many inspirational discussions over the past few years. The completion of this work would not have been possible without the concerted effort of Mr. John Neuman who performed all TCB and lipid analyses summarized within this thesis and modified established techniques of others to cater to the requirements of this study. I would also like to thank all the members and honourary members of the "Randall and Iwama" labs for academic, social and athletic interactions, all of which contribute to the "University experience." Finally, last but by no means least, I would like to thank my family and friends for always being there. This research was funded in part by the US Environmental Protection Agency, Environmental Research Laboratory, Athens, Georgia, under the Cooperative Agreement CR-816369. x G E N E R A L INTRODUCTION As technology and industry continue to expand in our society, so will the introduction of anthropogenic chemicals to the environment. Regulations and laws have been implemented to reduce their environmental impact; however, because of their constant gradual addition to the abiotic world, these chemicals slowly accumulate and pose a serious threat to biota. These chemicals are collectively termed xenobiotics, and many are desirable for industrial purposes, as they are chemically stable and thus resistant to degradation (Connell, 1990). The most biologically harmful are the chlorinated, organic molecules, which are very lipophilic in nature. Ironically, the traits which make many xenobiotics desirable for commercial applications are the same characteristics which make them particularly toxic to flora and fauna. Most animals possess a system for the detoxification of various xenobiotic compounds, the enzymes of which are collectively termed multifunction oxidases (MFO)(Connell, 1990). Although aquatic animals such as fish possess the same oxidizing enzymes as mammals, they are less abundant and operate at a reduced rate (Zitko, 1980). Detoxification occurs in the gills (Barron et al., 1989) and in the liver (Zitko, 1980), where the molecules are converted to a more hydrophilic form and thus are more easily excreted to the environment. Not all chemicals, however, are metabolizable, as they often possess structures resistant to chemical attack (Connell, 1990). Thus, many toxicants accumulate within the animal to concentrations far in 1 excess of those found in the environment, a condition defined as bioaccumulation (Connell, 1990). The accumulation of many xenobiotics has been demonstrated to induce a variety of pathological conditions, which can jeopardize the survival of individuals, populations, and thus entire community structures. Bioaccumulation can occur through two processes: bioconcentration or biomagnification. The latter refers to the accumulation of a toxicant through ingestion of contaminated foods and is of primary importance in air breathing animals (Connell, 1990). The accumulation of a chemical directly from the environment, most commonly across the respiratory surface, is defined as bioconcentration. Although, in some instances, the process of biomagnification can contribute to bioconcentration in fish (Spiragelll et al., 1983), direct uptake from the environment is the dominant mode of toxicant uptake by non air breathing marine and aquatic animals (Shaw and Connell, 1982; Reinert, 1972). It is the process of bioconcentration in aquatic organisms, specifically fish, which is the focus of this thesis. Uptake of a toxicant during bioconcentration is through the process of diffusion but the principal surface across which the chemical diffuses depends upon the animal. In unicellular organisms, xenobiotic entry is across the cell membrane. In juvenile to adult fish, the respiratory epithelium forms an extremely thin barrier between the blood and the environment and constitutes the majority of the animal's surface area (Murphy and Murphy, 1971; Rombough and Moroz, 1990). Not surprisingly, therefore, the gills are the main site of xenobiotic entry (Holden, 1962; 2 Murphy and Murphy, 1971). Some studies indicate cutaneous uptake of xenobiotics occurs (Saarikoski et al., 1986), however, most refute the significance of this pathway to the development of the toxicant body burden (Murphy and Murphy, 1971). Uptake of the chemical can be paracellular, but the relatively high lipid solubility of most xenobiotics dictates transcellular diffusion across the lamellar barrier (Hunn and Allen, 1974). Although too many xenobiotics exist to examine the biological activity and toxicity of all the chemicals, most are well documented with respect to their physico-chemical properties. Collectively, these properties permit the grouping of organic molecules such that quantitative structure-activity relationships (QSAR) can be established and used to predict the bioconcentration of the chemical. The properties which influence the movement of molecules across biological membranes are: molecular weight and volume (Brooke et al., 1986; Saito et al., 1990), chemical stability, degree of ionization (Saarikoski et al., 1986), and lipid and water solubility (see Barron, 1990; and Connell, 1990 for a review). The latter solubilities are conveniently expressed as a ratio of the organic molecule's solubility in octanol and water. This ratio is referred to as the octanol/water partition coefficient (K^,) which is the principle physico-chemical characteristic used in determining an organic molecule's QSAR. (Zitko, 1980; Connell, 1990). There are many models which exist to predict uptake rates and body accumulations of toxicants. All those with a physiological basis require many assumptions and anatomical measurements, which can be both costly and time-3 consuming, such as the assessment of gill thickness and surface area. Many of these estimations have been performed upon common species, such as the rainbow trout (Hughes, 1984; Laurent, 1984); however, for most fish species, these parameters are unknown. A further compounding problem is that estimations of anatomical dimensions are often of limited physiological relevance. For example, the anatomical surface area of the gills is always greater than the functional surface area due to non-perfused regions of the gill. At rest, only 60 % of the secondary lamellae of the gill are perfused (Booth, 1978). The parameters in these models are often assumed to be constant over a variety of physical conditions; however, this is often not the circumstance. During exercise blood pressure becomes elevated, increasing the proportion of the perfused lamellae and reducing the respiratory epithelial thickness (Randall and Daxboeck, 1984). Thus, not all parameters can be assumed to be constant over a range of activity levels if the models are to be physiologically relevant. Extensive data can be found in the literature pertaining to the metabolic and oxygen consumption rates of countless fish species and sizes, at different activity levels, in water of various temperatures, salinities and pH's. An oxygen data bank has been assembled that allows prediction of oxygen uptake in fish, where the major determinant of oxygen uptake is fish weight, with temperature as a minor influencing parameter (Thurston and Gehrke, 1991). The predominant barrier to toxicant uptake in fish is the respiratory surface, and thus it is likely that variables such as ventilatory volume, gill epithelial thickness and perfusion, which influence the rate of 4 oxygen transport across the epithelial membrane, will similarly influence toxicant movement. That is, the conditions for oxygen transfer are likely indicative of that for toxicant transfer. If a relationship between toxicant uptake and the oxygen consumption rate (MO z) of fish in normoxic waters can be established during initial toxicant exposure, this may allow for the prediction of toxicant uptake based on M0 2's abundant in the literature or easily measured for any fish species. The establishment of such a relationship will reduce the need for costly and time consuming measurements, otherwise required in the models used to predict toxicant uptake. Relationships have been established between toxicant and oxygen uptake for other chemicals (Murphy and Murphy, 1971; Black et al., 1990). In these studies, alterations in metabolic rate were temperature-induced. Oxygen consumption rates can increase from 12 to 20 fold between rest and exercise (Brett, 1964; Brett and Glass, 1973; Puckett and Dill, 1984), an elevation far exceeding that which can be induced through changes in temperature (Wieser and Forstner, 1986; Hughes et al., 1983), and thus activity level was chosen as the means for influencing M O z in this study. The primary objective of the work performed in this thesis, therefore, is to establish whether a relationship exists between the initial uptake rate of a toxicant, 1,2,4,5-tetrachlorobenzene, and the oxygen consumption rate, for a variety of fish species in a normoxic environment. The second objective will be to determine if an organ can be used to estimate whole animal toxicant concentrations during initial stages of toxicant exposure. If this is possible, the work required to measure body 5 toxicant accumulations can be dramatically reduced. 6 G E N E R A L MATERIALS AND METHODS XENOBIOTIC The xenobiotic used in all experiments described in this thesis is 1,2,4,5-tetrachlorobenzene (TCB). The toxicant exposed to the fishes was purchased from Pfalz and Bauer Ink, Stamford, Connecticut, and that used as a reference standard was purchased from Chem Services, West Chester, Pennsylvania. TCB was chosen for several of its physico-chemical characteristics. Firstly, it is not readily metabolized and thus initial net uptake of the chemical can be determined for a fish simply by measuring its toxicant body burden following a given exposure duration. In addition, TCB is non-dissociable (U.S. Environmental Protection Agency, 1991) and thus there is no effect of pH on the proportion of the compound existing as the dissociated or undissociated form. This is an important consideration, as in many chemicals the un-ionized form has a much greater diffusivity than the ionized species and thus the concentration of the xenobiotic available for uptake varies with the environmental pH (Saarikoski et al., 1986; Saarikoski and Viluksela, 1982). The log for TCB is 4.99 at 25°C which indicates that the chemical is relatively lipophilic, and thus will have a high bioaccumulation factor; however, its water solubility is sufficient that the chemical can be dissolved in water and thus be available for uptake by fish. The concentration of TCB in aqueous solutions to which fish were exposed was well below the maximal water solubility of the chemical (2.42 mgl"1 at 25°C) (U.S. Environmental Protection Agency, 1991) unless otherwise specified. 7 Finally, the chemical is not very volatile and thus will not evaporate readily from solution. EXPERIMENTS: All experiments requiring the use of a swim tube respirometer were conducted in the department of Zoology, at the University of British Columbia (UBC). Studies on fish at routine activity levels (static tests), and all analyses of water sample and fish tissue and plasma 1,2,4,5-tetrachlorobenzene (TCB) concentrations, were performed in the Fisheries Bioassay Laboratory (FBL), Montana State University (MSU), by John Neuman. ANIMAL ACQUISITION AND ACCLIMATIZATION: Rainbow trout, Oncorhynchus mykiss, (2-3 g; 6.5-8.5 g; 250-500 g) were purchased from West Creek trout farms in Aldergrove, B.C., for those experiments performed at UBC, and obtained from the Fish and Wildlife Service, fish technology centre, Bozeman, Montana, for those studied at MSU. Large fish were habituated in 10 000 1, and smaller fish in 200 1 fibreglass tanks containing dechlorinated Vancouver city tap water, circulating at approximately 1 body length per second (Bis1), for at least two weeks prior to experimentation. Water temperature varied seasonally from 5°C-10°C; however, all experiments in the respirometer were performed at the temperature to which the animals had been habituated. Following arrival, fish were fed commercial pellets bi-weekly ad libitum, but 8 were starved at least 72 hrs 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 prior to experimentation and actual parameters determined upon completion of the test. Animals were transported from the holding tank in a 20 1 bucket, netted, and gently placed in the respirometer for all experiments where fish were forced to swim. UBC APPARATUS: A 130 1 recirculating Brett type respirometer, described in detail by Gehrke et al. (1990), was used in experiments requiring the animals to maintain specified exercise intensity levels. Water temperature, pH, partial pressure of oxygen (P02) and water velocity of the respirometer can be controlled. In all experiments the respirometer temperature was maintained at the same value measured in the holding tank ± 0.3°C. The pH of the dechlorinated city water was 6.2 and changes in water pH over any of the experimental durations were very small and thus, water pH was not adjusted or controlled. Saarikoski et al. (1986) showed that water pH, between 3 and 9, did not affect the uptake of non-ionizable compounds in guppies. The P 0 2 of the respirometer water was controlled in all experiments in the following manner. The partial pressure of oxygen (P02) in the system is determined using an independently thermostatted radiometer E5046 P0 2 electrode connected to a radiometer PHM71 acid-base analyzer. The output from the latter is coupled to 9 an oxygen comparator, with a variable set point, which regulates the delivery of compressed oxygen or nitrogen to an artificial lung, thus altering the gas composition of the water in the respirometer. At the start of an experiment the system was sealed, the lung was isolated from the swim tunnel, and the analog outputs for temperature, velocity and P0 2 were sampled every 10s (0.1 Hz), digitally transformed and interfaced by a DT2805 data translation/acquisition board. The raw data were then recorded in a print file by an Olivetti M24 personal computer operating the program L A B T E C H NOTEBOOK® (Laboratory Technologies Corporation). These data were then analyzed in LOTUS 123® (LOTUS Development Corporation) and average values for temperature, velocity and oxygen consumption were determined over the test period. Background oxygen consumption determinations were performed following each experimental series and were found to be negligible. The swim chamber of the respirometer is a plexiglass tunnel, 1000mm in length and 200 mm in diameter, with stainless steel mesh at each end. Just anterior to the rear screen is a grid, over which a 5 V stimulus is passed to prevent fish from resting on it at higher velocities. An opaque plastic covering was placed around the middle of the swim tube to entice the animals to swim in this region. This was particularly important when working with smaller fish which were able to take advantage of small eddies created by the anterior screen. The large volume of the respirometer results in the maintenance of a fairly constant water toxicant concentration over time; however, accurate oxygen consumption measurements require small volumes. Thus, small fish were swum in 10 groups of 12 to 25, specified in the materials and methods of each chapter, such that the total fish mass exceeded 100 g. This was determined as the lower limit for accurate oxygen consumption measurements. EXPERIMENTAL PROCEDURE Prior to the introduction of the fish to the respirometer, water temperature was manipulated to that of the holding tank, the P0 2 was calibrated and the P0 2 of the respirometer water was increased to the 100% air saturated value. All air bubbles were removed from the swim tunnel, the fish were placed in the swim chamber and forced to swim at 18 cms'1 (the minimum flow velocity of the system) for the introductory period of two hours for all tests unless otherwise specified. During this time there was a slow continuous overflow of dechlorinated city water to prevent the accumulation of metabolic waste products. Following the introductory phase, the artificial lung was isolated from the respirometer and a pre-mixed solution of 1,2,4,5-tetrachlorbenzene dissolved in 67 ml of acetone was slowly added downstream of the test animals. The water velocity was then gradually increased over a 5 min. duration to the final swimming velocity. The exception to this was during critical swimming velocity (U^,) tests of chapter 2. The methodology and test is described in chapter 2. Duplicate water samples were taken in 40 ml amber I-chem glass vials with teflon lined lids 5 mins after the addition of the chemical. Following this the system was sealed for the test duration. At the end of the test final water samples were taken in duplicate, and the fish were 11 removed from the respirometer and killed by a sharp blow to the head. The fish were then rinsed quickly in fresh water, lightly blotted dry, and measurements of fish weight, width, height and fork length were noted. The caudal peduncle was severed and the blood from the dorsal aorta was collected in a heparinized test tube or several haematocrit tubes (the former for large fish, the latter for small fish). The tubes were then centrifuged at 3500 rpm in a Damon, micro hematocrit IEC MB centrifuge for 3 mins and the plasma was separated from the blood cells and frozen, along with the fish carcasses, in a -20°C freezer. The frozen remains and water samples were packed in dry ice and sent to Montana State University (MSU) in Bozeman, Montana, via courier for analytical determinations. MSU: EXPERIMENTAL PROCEDURE: Described within individual chapters. TCB ANALYSIS All TCB analyses were performed by John Neuman at the Fisheries Bioassay Lab of Montana State University. Whole fish Fish to be analyzed were taken from the freezer, weighed, and the whole fish 12 homogenized with anhydrous sodium sulfate and dry ice in a stainless steel blender. Fish larger than 5-6 g, were sliced into a few pieces prior to blending. Each homogenate was then transferred to a cotton cellulose extraction thimble (25 x 100 mm) and loaded into a Soxhlet extractor containing 100 ml hexane. Analysis of TCB recovery was monitored by adding 1.0 ml hexane containing 100 micrograms of surrogate (1,2,3,4-tetrachlorobenzene), to each extraction thimble. Extraction time was at least 8 hours. Once brought to room temperature, each extract volume was adjusted to 100 ml with hexane, and a measured aliquot was then transferred to a florisil column where fats and lipid materials were removed. This glass column (11 mm x 300mm) was packed with 12cm florisil and 2cm anhydrous sodium sulfate and fitted with a teflon stopcock. The columns were stored at 130°C and pre-rinsed with 5 % (v/v) tert-butyl methyl ether/hexane (TBME) solution just prior to use. Once the extract aliquot was added, TCB and surrogate were eluted into a 100 ml volumetric flask with 30 ml 5% T B M E solution. Pentachlorobenzene internal standard solution (0.5 ml at 500 ugl"1) was added to each elutant and then the volume was adjusted to 100 ml with hexane. TCB and surrogate sample concentrations were measured by gas chromatography. Fish tissue All tissues were stored in glass bottles or plastic bags at -15°C for 1 to 60 days prior to analysis, and were partially thawed and weighed at the time of analysis. Sample sizes ranged from 0.2 to 3 g, but were usually less than 2 g. The tissues were 13 blended with sodium sulphate and dry ice in a stainless steel blender and the contents were transferred to a Soxhlet extraction thimble and extracted for 4 hrs with 100 ml of hexane containing surrogate. Remaining extraction and analysis procedures were the same as above. Plasma Plasma samples were thawed, and 0.1 ml test sample aliquots and 1.0 ml control sample aliquots were added to 10 ml screw cap Pyrex culture tubes containing 5 ml hexane and 0.1 ml of the surrogate. Extraction was achieved by vortexing the capped tubes for 1 min, and the tubes were then allowed to stand for 30 min. An aliquot of the extract was diluted 10 fold (the internal standard was added at this point), and the sample then analyzed by GC as described below. Water Based on the initial target TCB water concentration, aliquots of the water samples taken 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 TCB 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 (50 ugT1) was then added. This solution was made to volume with hexane and TCB concentrations 14 were determined by GC as described below. Gas Chromatography Analysis The GC analyses were performed using a Varian model 6000 gas chromatograph (GC) equipped with a Ni-63 electron capture detector and fitted with a glass column (1.8 m x 0.2 mm ID) packed with 3% SE-30 on Supelcoport 100.120 operated at 130°C. Argon-methane (10 %) was used as the carrier gas at 30 ml min'1, and inlet temperature was set at 200°C. Standards and samples were injected in duplicate using a Varian model 8000 autosampler. Gas chromatographic peak areas were measured with a Varian Vista Data System. TCB and surrogate sample concentrations were determined using internal standard quantitation methods. Regression equations established from seven point calibration curves at standard concentration ranges from 0.5 to 10 ugT1 were used. Required dilutions were made using an internal standard solution at 2.5 ugT1. Method blanks were taken through the entire procedure to check for gas chromatographic interference at the retention times of TCB, surrogate and internal standard. TCB values observed were not adjusted for corresponding surrogate recoveries unless such recoveries were less than 90 %. CALCULATIONS AND STATISTICS: If the cross sectional area of a fish exceeded 5 % of the swim tubes cross 15 sectional area, swimming velocity was corrected for a blocking effect according to the equation described by Bell and Terhune (1970). An F-test performed upon the Pearson correlation coefficient was used to test for statistical significance of all regressions and correlations. The line of best fit was calculated by least squares regression, and the hypothesis of homogeneity of regression coefficients tested by a students t-test. The latter was the criterion upon which the decision to group data was based. Differences between treatment, or species means was tested by a one-way A N O V A following data transformation if required, and a Tukey test. In all tests, a probability of 0.05 was chosen as the level of statistical significance. All values are expressed as the mean ± s.e., unless otherwise specified. 16 CHAPTER ONE INTRODUCTION The primary objective of this thesis is to establish whether a correlation exists between toxicant and oxygen uptake in fish during initial exposure to a toxicant. If the relationship exists, this will permit the use of the extensive data available in the literature on oxygen consumption rates for a variety of fish species, activity levels and environmental conditions, in the prediction of xenobiotic uptake. A prerequisite for this method of toxicant uptake prediction, however, is that exposure to the xenobiotic does not alter the animals oxygen uptake rate. Thus, the first objective in this chapter is to determine if there is a direct effect of TCB exposure on oxygen consumption rate Toxicant exposure could potentially reduce the aerobic swimming capacity of the fish. If so, this would influence the choice of velocities at which the fish would be forced to swim. The critical swimming velocity (Ucr i t) is often used as an indication of the aerobic swimming capacity in fish, and is by definition the maximal velocity at which a fish can swim aerobically. Thus the second objective in this chapter is to determine if body toxicant accumulations have an effect on the critical swimming velocity (Ucr i t) of fish. The procedures required to obtain measurements of whole body concentrations of a toxicant can be both costly and time-consuming, particularly for animals bigger than 5 to 10 grams which possess difficult to homogenize, dense 17 bones. Following diffusion across the gills, a toxicant enters the circulatory system and is regionally distributed within the animal. If the body burden of a chemical can be inferred from one organ, work loads can be immensely reduced and methodological problems associated with uptake rate determinations in large animals can be eliminated. Thus, the final objective in this chapter is to investigate the possibility of using an organ or tissue to estimate the whole animal toxicant concentration. MATERIALS AND METHODS Preliminary investigations were carried out to ensure that the procedures for sample storage, preparation, and toxicant concentration determinations were adequate. In addition, the degree and site of toxicant loss from the fish exposure chambers was examined, such that toxicant loss could be minimized over the desired exposure duration (for details see appendix). EXPERIMENTAL PROCEDURE: 1.) Effect of external TCB concentration on oxygen uptake: A.) At rest Eight rainbow trout (200-400 g) were placed individually in 20 1 jars with 18 initial estimated TCB concentrations of 239, 633, 1310 and 2390 ugl 1 at 8°C. The latter concentration is in excess of the toxicants water solubility, however, it was chosen to ensure maximal dissolved toxicant concentrations were achieved under the conditions. Two fish were exposed to each TCB concentration and the jars were topped up, immediately capped and the fish were left for two hours. Animals were removed and killed at the end of this period, blood plasma was obtained for TCB concentration determination as described in the general materials and methods and water samples were taken to determine TCB concentrations and oxygen consumption rates of the fish. B.) During exercise Six rainbow trout (307 ± 14.3 g) were individually placed in the respirometer at a water temperature of 6°C for an hour (at a water velocity of approximately 1 body length per second (Bis1), prior to a water velocity increase to 1.25 Bis"1. Fish were then immediately exposed to a toxicant concentration of 95, 224 or 636 ugl 1 for an hour during which time oxygen consumption was determined. Fish were then killed to obtain a plasma sample, and water samples were taken for TCB analysis. 2.) Effect of TCB accumulation on U^,: To examine the effect of elevated body TCB levels on the U^, of juvenile rainbow trout (2-3 g), groups of fish were pre-exposed to one of three TCB concentrations for either 4 hrs, or 72 hrs prior to the determination of U,^, in 19 uncontaminated water at 10 °C. Groups of 16 fish were held at 10°C in 35 1 glass aquaria for 2 days. In the 4 hr pre-exposure study, fish were then exposed to initial concentrations of 0, 1013, or 1240 ugT1 TCB for 4 hrs prior to the U ^ , determination. In the long pre-exposure duration study, daily additions of TCB dissolved in 25 ml acetone were administered to respective tanks for three days to increase the water TCB concentrations by 0, 125, or 250 ugT1, and the U^, determined 72 hrs following the initial toxicant addition (find mean exposure concentrations). In both tests, where possible, determinations were performed on a group of 10 animals. Following the 72 hr pre-exposure to TCB, many individuals were physically incapable of maintaining equilibrium, and were omitted from the U,^, procedure, reducing group size accordingly. Subsequent to TCB exposure, trout were introduced into the swim chamber for 30 mins with no flow, and then water velocity was increased every 30 mins, in 1.0 Bis"1 increments, from 3 Bis'1 until the animals fatigued. Fatigue was defined as the velocity at which the fish could no longer remove itself from the posterior screen, despite prodding. Individual U^s were calculated, in Bis"1, by adding the product of the completed proportion of the fatigue increment and the magnitude of the increment (0.5 Bis"1) to the velocity of the last completed increment. Because the fish were so small it was not necessary to correct for the effect of the fish on water velocity in the swim tube. 20 3.) TCB concentrations within organs and tissues: Rainbow trout (330 ± 23 g) were individually swum at 10°C in the respirometer at one of three swimming velocities: 1.25, 1.9 and 2.5 Bis"1 for an hour at an initial TCB concentration of 322 ± 6.5 ugT1. In this test, the initial water samples were collected 2 mins following TCB addition. Following the one hour exposure to the toxicant, fish were killed, the plasma was obtained and the following tissues were dissected from the fish for analysis of toxicant concentration: the brain, liver, kidney, spleen, heart, lateral line pink muscle, epaxial white muscle, gills, adipose tissue just external to the intestine, and the upper and lower gut, manipulated to be devoid of food materials. Other procedures were executed as outlined in the general materials and methods. CALCULATIONS: Total TCB body burden for each fish swum in the respirometer was estimated by summing the products of the toxicant concentration of an organ and the proportion of the animal that organ constitutes. The latter values were obtained from Daxboeck (1981). RESULTS: Effect of TCB exposure on oxygen consumption rates: TCB dissolved in water could potentially be an irritant to fish. If so, oxygen consumption rates would be expected to increase in response to the insult of toxicant 21 addition to the environment. This was tested by monitoring the rates at which oxygen was consumed in adult rainbow trout exposed to various TCB concentrations at rest (Fig. 1) and while swimming in the respirometer at 1.25 Bis"1 (Fig. 2). No significant relationship was demonstrated, at either activity level, between oxygen consumption rate and TCB concentration up to levels exceeding the chemical's maximal solubility in water. Effect of TCB exposure on U C T i t: As the body burden of TCB is elevated in juvenile rainbow trout due to either short term (Table 1) or long term (Table 2) pre-exposure, there is little effect on U c r i I . In fact, following the 4 hr pre-exposure regime, the group U c r i , of 7.11 ± 0.25 Bis'1 from fish exposed to an initial concentration of 1013 ugl"1 was slightly elevated relative to the control condition of 6.65 ± 0.23 or the slightly higher initial concentration (1240 ugl"1) where fish were able to achieve a U^, of 6.03 ± 0.50. None of these U c r i ts are statistically different, however, and thus, during a short exposure period of 4 hrs, at concentrations below approximately 1000 ugl"1 there is no affect on U^, in rainbow trout. Longer pre-exposure durations to lower toxicant concentrations provided similar results (Table 2). Critical swimming velocity was slightly reduced relative to the control condition only at the highest concentration of 250 ugl"1, where the U c r j t was calculated to be 5.7 ± 0.44 Bis'1. Again, the greatest swimming velocity (6.7 ± 0.27 Bis"1) was attained by the group exposed to the medium concentration tested 22 (125 ugl"1); however, none of these U^s differed significantly. In the 250 ugl"1 TCB treatment, the mean total body burden was determined to be 467 ugg"1. This is much greater than that determined for all other exposure regimes, but far from the equilibrium value predicted by Tischmak (1984) for fathead minnows. All TCB exposed fish from this test were very dark in appearance and many mortalities were observed, but if the fish were capable of surviving the TCB pre-exposure condition, swimming performance was unaffected. Toxicant distribution between body components: Once within the fish, the xenobiotic is transported and delivered by the circulatory system to the various tissues. The concentration of TCB in the different tissues is highly variable between individuals subjected to the same conditions (Table 3). The greatest toxicant accumulation and variability in concentration was found in the adipose tissue and pink muscle, both of which contain large proportions of lipid. In general, the tissue TCB content increased with the velocity at which the fish was forced to swim. Not all tissues accumulate the toxicant in the same proportion to the plasma concentration. This is indicated by the large variation in the tissue:plasma ratio between fish swimming at the same velocity (Table 4). In addition, there is relatively little change in the ratio with an increase in exercise intensity. Thus the distribution of the toxicant within different organs is highly variable between individuals. 23 DISCUSSION: Exposure to TCB concentrations up to the maximal water solubility of the chemical has very little influence on oxygen consumption rates in rainbow trout. No significant relationship was found between water toxicant concentration and oxygen uptake in large rainbow trout at rest (Fig. 1) or while swimming at a constant velocity of 1.25 Bis"1 (Fig. 2). These findings are in agreement with McKim et al. (1985) who exposed rainbow trout to 14 organic chemicals and found no effect on oxygen consumption rate, ventilation rate, or ventilation volume. DDT in mosquito fish (Murphy and Murphy, 1971), and 4 herbicides at various concentrations in coho salmon, were also demonstrated to have no significant affect on oxygen consumption rates (Janz et al., 1991). Thus, the use of oxygen consumption rates, determined in the absence of a toxicant (ie. literature values), can potentially be used to predict TCB uptake provided a relationship between the two parameters can be demonstrated. Body concentrations of TCB elevated through a 4 or 72 hr pre-exposure regime, did not significantly reduce the U^, in juvenile rainbow, despite signs of physical distress in response to the chemical insult at both dose concentrations of the longer pre-exposure test (Tables 1 and 2). In fact, U^, was slightly but not significantly enhanced at the lower of the two TCB concentrations in both short and long pre-exposure duration studies. A common parameter used to standardize the toxicity of various chemicals to aquatic animals is called the 96 hr LC50 value, the concentration of a chemical at 24 which 50% of the animals exposed to the chemical for 4 days, die (Sprague, 1990). The 96 hr LC50 value for fathead minnows at exposed to TCB is 1070 ugT1 (U.S. Environmental Protection Agency, 1980). The 96 hr LC50 value for rainbow trout is probably similar to that for fathead minnows, as Thurston et al. (1985) found little difference in the toxicity of 10 organic chemicals between these two species. It is interesting to note that pre-exposure of rainbow trout to TCB concentrations approaching this LC50 value, has no affect on U,^,, and thus the aerobic swimming ability of the animals will not be compromised at the much lower concentrations used in the following studies to establish the relationship between toxicant and oxygen uptake. Thus, no additional constraints need be placed upon the choice of swimming velocities to influence MO z , in animals subject to the presence of TCB. Even if a chemical induces a reduction in U,^,, this does not preclude the potential to predict its uptake as a function of oxygen consumption rate, provided the latter is more or less independent of exposure concentration. It will, however, impose restrictions on the choice of swimming velocities at which fish can be forced to swim aerobically. In the first hour of toxicant exposure in fish swimming at different swimming speeds it is obvious that the chemical is not yet in equilibrium between the fish and the water. This is substantiated by the observation that the TCB concentration within each organ increases with exercise intensity (Table 3). This is in agreement with Tischmak (1984) and Smith et al. (1990) working with 4 to 11 month fathead minnows, Pimephales promelas, and 4 to 6 month American flagfish, Jordanella 25 floridae, respectively. Equilibrium between the fish and water was not reached for 8 days following initial exposure to the chemical at 20°C in the former and 4 days at 25°C in the latter. Thus, toxicant uptake is rapid in this early exposure period and is within the quasi-linear portion of the uptake curve which asymptotes as the animal and the water reach equilibrium. No one tissue appeared to be more representative of body burden than another, and a large variability in intra-tissue toxicant concentration is observed in animals swum at the same velocity (Table 3). Variability in tissue toxicant concentrations has frequently been observed by others. Holden (1962) noted a large variation in organ DDT concentration within and between brown trout exposed to the pesticide at rest for a week, which he attributed to variability in lipid content. Spiragelll et al. (1983) concluded that the accumulation of toxicant within muscle of brown trout could be predicted by lipid content. Goerke (1984) established a correlation between the lipid content and PCB concentration of several organs in the polychaete worm, Neries virens. Thus, there is a large body of evidence to indicate that tissue toxicant variation within and between individuals exposed to similar conditions, can be explained by differences in lipid content. Another contributing factor to the noted variability may be due to differences in blood volume within each tissue subsequent to its removal, as the TCB bound by whole blood was not measured in this study; however, this is likely to be secondary to the variability in lipid content. Due to the large variability in tissue toxicant concentration, the use of any of these tissues to estimate body burden would be tenuous at best. 26 As the level of exercise intensity increases, there is a regional redistribution of blood flow to ensure adequate circulation to the active tissues. Blood flow to the skeletal muscles, particularly the red muscle, is increased by an order of magnitude, and that to the liver and kidney is reduced four fold in rainbow trout swimming at 80% of U,^, relative to resting values (Daxboeck, 1981). The tissuerplasma TCB ratio does not change dramatically with exercise for any of the tissues (Table 4) and indicates that plasma toxicant concentration determines the rate of toxicant delivery to the organ. If toxicant accumulation were determined by blood flow, a change in the tissuerplasma concentration for a given organ would be expected in proportion to the change in blood flow to that region. It is possible that the plasma concentration could be a reliable indicator of body burden. The relationship between plasma TCB and the estimated total body TCB concentration of rainbow trout swum at different speeds (Fig. 3) supports this possibility. Thus, of the components examined, plasma toxicant concentration is likely the best predictor of body burden. In conclusion, there is no direct effect of TCB exposure on the oxygen consumption rates of rainbow trout at rest or during exercise, and thus relationships between toxicant and oxygen uptake can be investigated without there being interactive components to the relationship. In addition, there is little effect of body accumulations of TCB in juvenile rainbow trout on U^,, and thus exposure to the toxicant does not impose additional constraints on the swimming velocity chosen to elevate oxygen consumption rates. The variability in tissue toxicant concentrations in 27 swimming fish is sufficient to preclude their use in the prediction of body burden. Toxicant accumulation within the tissues of active fish, however, appears to be dependent upon the concentration of TCB in the plasma in the first hour of toxicant exposure, and it is possible that plasma concentrations may be used to estimate body burden; however, this must be examined in more detail. 28 Figure 1. The effect of water TCB concentration on the oxygen consumption rate of adult rainbow trout at rest (1^=0.01). 29 100 90 O < 60 80 70 60 50 O h o o o o o o 40 0 500 1000 1500 2000 2500 WATER TCB (ug-1 ) 30 Figure 2. The effect of water TCB concentration on the oxygen consumption rate adult rainbow trout swimming at 1.25 Bis"1 (r^O.02). 31 WATER TCB ( u g l *) 3 2 Figure 3. The relationship between measured plasma TCB concentration and estimated body burden, in adult rainbow trout swimming at different velocities for one hour (see Table 3 for further details). Body burden was calculated from tissue TCB concentrations (Table 3) and the relative contribution to the animals mass that each tissue constitutes. 33 34 Table 1. The effect of short duration (4hr) TCB pre-exposure on body burden and Ucrit, determined in toxicant free water, for juvenile rainbow trout, (n = 10). There is no statistical difference between U^ .s. Table 2. The effect of long duration (72hr) TCB pre-exposure on body burden and UCTit, determined in toxicant free water, for juvenile rainbow trout. An amount of TCB equivalent to that initially introduced to toxicant exposure apparatus was added every 12 hrs. Following toxicant exposure, many fish were physically incapable of swimming at either TCB concentration, thus reducing the number within a treatment from 10. There is no statistical difference between Ucri.s. 35 MEAN U c r i t ( B i s 1 ) 6.1 ± 0.52 7.1 ± 0.26 6.6 ± 0.08 WATER TCB ( u g T 1 ) INITIAL FINAL (n = 3) 1240 ± 18.7 587 ± 10.2 1013 ± 4.1 377 ± 5.3 22.3 ± 0.1 20.7 ± 0.3 FINAL BODY TCB (ugg" 1) 44.91 ± 4.26 44.60 ± 2.43 3.72 ± 0.25 MEAN U c r i t ( B i s 1 ) 5.7 ± 0.49 6.7 ± 0.29 6.4 ± 0.29 # OF FISH IN GROUP 6 8 10 WATER TCB ( u g T 1 ) INITIAL (with e q u i v a l e n t b i - d a i l y a d d i t i o n s ) 250 125 0 FINAL BODY TCB (ugg" 1) 466.83 ± 26. 64 220.12 ± 15.84 0.41 ± 0.05 36 Table 3. The concentration of TCB in the plasma (ugml1) and 11 tissues (ugg1), taken from adult rainbow trout swimming at the noted velocity and water TCB concentrations (ugml_1), for one hour. 37 Bis' Water [TCB] Plasma Brain Liver Kidney Heart Red Muscle White Muscle Gills Adipose Upp. Gut Low. Gut Spleen 1.29 0.281 3.01 9.63 11.20 5.68 7.05 10.30 1.66 6.67 54.10 1.06 4.21 2.61 1.30 0.291 1.78 9.58 9.52 4.96 6.36 14.70 1.63 5.92 35.90 1.93 - 1.70 1.98 0.268 5.52 14.60 15.80 6.49 5.76 12.60 3.52 5.16 143.00 3.90 6.18 4.01 1.88 0.299 4.35 13.50 13.70 5.82 6.37 38.50 2.70 4.35 33.00 2.59 5.50 3.01 2.68 0.283 8.84 21.70 13.60 12.30 9.45 21.60 6.02 7.36 4.17 1.23 6.81 2.82 2.59 0.291 8.55 22.30 18.70 13.80 10.90 26.00 3.52 9.11 66.70 2.40 8.67 3.78 Table 4. The ratio of tissue to plasma TCB concentration for the data summarized in Table 3. 39 o Bis1 Brain Liver Kidney Heart Red Muscle White Muscle Gills Adipose Upper Gut Lower Gut Spleen 1.29 3.20 3.72 1.89 2.34 3.42 0.55 2.22 17.97 0.35 1.40 0.87 1.30 5.38 5.35 2.79 3.57 8.26 0.92 3.33 20.17 1.08 - 0.96 1.98 2.64 2.86 1.18 1.04 2.28 0.64 0.93 25.91 0.71 1.12 0.73 1.88 3.10 3.15 1.34 1.46 8.85 0.62 1.00 7.59 0.60 1.26 0.69 2.68 2.45 1.54 1.39 1.07 2.44 0.68 0.83 0.47 0.14 0.77 0.32 2.59 2.61 2.19 1.61 1.27 3.04 0.41 1.07 7.80 0.28 1.01 0.44 CHAPTER 2 INTRODUCTION In chapter one, it was reported that there was no direct effect of TCB exposure on the oxygen consumption rates in resting and active fish, as has been established for other organic chemicals (McKim et al., 1985). Thus, literature M 0 2 values can potentially be incorporated into the prediction of toxicant uptake, provided a relationship between oxygen and toxicant uptake exists. In addition, preliminary findings indicate that plasma TCB concentrations may be used to estimate body burden markedly reducing the work otherwise required. Thus, the first objective of the experiments reported in this chapter is to determine whether plasma TCB concentration can be used to estimate body burden during initial exposure to the xenobiotic. The second objective in this chapter is to establish whether a relationship exists between measured body accumulations of the xenobiotic and oxygen uptake during initial exposure periods in small rainbow trout. If this relationship is found, there is potential that the initial uptake of this and possibly other chemicals can be predicted from the extensive measurements of metabolic rates available in the literature, for animals in a variety of environmental conditions. 41 MATERIALS AND METHODS: EXPERIMENTAL PROCEDURES: 1.) Plasma as an indicator of body burden: a.) At rest Groups of eight cutthroat trout (Oncorhynchus clarkii) (7.4 ± 0.39 g) were placed in 20 1 jars with initial TCB concentrations of 253, 1265, and 2515 ugl 1 for two hours. Groups of 4 rainbow trout {Oncorhynchus mykiss)(8.9 ± 0.65 g) were exposed in 20 1 jars to initial concentrations of 194, 1200 and approximately 2670 ugl"1 for a similar two hour duration. In both tests, oxygen consumption was determined over the entire duration, and blood plasma TCB concentrations and total body burden were determined for each fish. B.) During exercise Rainbow trout (7.1 ± 0.15 g) were introduced into the respirometer (5.04 ± 0.01°C) in groups of 12 for 2 hrs prior to the addition of TCB. Individual groups were forced to swim at 3.0 Bis'1 for 2 hrs at initial concentrations of 207, 206.5, 573 and 692 ugl"1 and for 6 hrs at initial concentrations of 238.5 and 217 ugl"1. Oxygen consumption was recorded over the test duration, fish were killed and blood plasma TCB concentration and TCB body burden was determined. 42 2.) Correlation between oxygen uptake and measured body burden: Groups of 25 rainbow trout (5.02 ± 0.07g) were introduced to the swim tunnel for an hour, at a water velocity of 18 cms'1. 25 fish were swum at a time, to ensure reliable oxygen consumption measurements during the toxicant exposure duration. Sufficient TCB dissolved in 67 ml acetone was added to elevate respirometer water concentrations to 260 or 780 ugT1, and fish were then forced to swim at the test velocity of approximately 2.25 or 3.75 Bis'1 for 2hrs. Upon completion of this duration, fish were individually removed from the swim tube, killed and meristic measurements recorded. Water samples were collected as specified in the general materials and methods at 5 mins and 2 hrs. Fish and water samples were frozen and shipped to Bozeman, Montana for chemical analyses. RESULTS: Plasma TCB concentrations as an indicator of body burden: Two hours following exposure to various TCB concentrations, there is a significant relationship between blood plasma TCB concentration and measured body burden for both cutthroat and rainbow trout as indicated by the r 2 value of 0.77 (Fig. 4). With an increased exposure duration, however, the relationship between plasma and whole body TCB concentrations breaks down. In juvenile rainbow trout swimming at 3.0 Bis"1, there is no statistical difference between plasma and whole body TCB concentrations 2 hrs following exposure to low and high TCB 43 concentrations (Fig. 5). Six hours following TCB exposure, however, there has been little change in plasma concentration, while body burden has increased proportionately with time (Fig. 6). Thus plasma concentrations are not always a good indication of body burden during initial toxicant uptake, and it is not possible to estimate body burden from plasma levels, so body burden must be measured directly. The observation that the TCB body burden increases proportionately with time over 6 hrs (Fig. 6) indicates that the rate at which TCB is acquired by the animal during the initial stages of toxicant exposure is constant despite the plateau in plasma toxicant concentration. Relationship between oxygen and toxicant uptake: The initial chemical uptake by the whole fish was related to the animals oxygen consumption rate. Figure 7 illustrates the significant relationship between oxygen consumption rate and the estimated TCB uptake rate, calculated from tissue TCB concentrations and the relative proportion of the animal that each constitutes. These data are for the fish used to examine body toxicant distribution in experiments reported in chapter one. In juvenile rainbow trout forced to swim at two intensities for 2 hrs, at mean external TCB concentrations of 166 and 539 ugl"1, there is a good relationship between toxicant and oxygen uptake for the two external TCB concentrations (Fig. 8). The slope of the relationship is greater at the higher TCB concentration (0.117) than at the lower concentration (0.067) as would be expected in a condition where 44 the external concentration of the xenobiotic dictates the driving force for toxicant entry into the fish (Randall and Brauner, 1991). When the toxicant accumulated over the two hour duration, is standardized by dividing the rate of TCB uptake by the mean external TCB concentration (Fig. 9), the slopes for the two external TCB concentrations remain significantly different and thus grouping of the data is not permissible. DISCUSSION: In chapter one, it was suggested that plasma TCB concentrations may be indicative of body TCB accumulations. This presents an attractive potential from a methodological perspective, as the analysis of body burden requires thorough homogenization of an entire animal prior to the toxicant concentration determination which can be both a tedious and time consuming task. In fish at rest, exposed to various TCB concentrations for two hours, there is a strong correlation between TCB in the plasma and the body (Fig. 4). Two hours following initial toxicant exposure, in fish swum at 3 Bis 1, plasma and body TCB concentrations are not significantly different at either water TCB concentration examined (Fig. 5). The relationship between plasma and body TCB concentrations, however, breaks down with exposure duration. Following 6 hrs of exposure in fish swimming at 3 Bis'1 there is little further increase in plasma concentrations, while body concentrations continue to rise proportionately with time (Fig. 6). Thus, it is not possible to predict 45 total body burden solely based upon an animals plasma toxicant concentration and if the relationship between oxygen consumption and accumulations of the toxicant is to be examined, body burden must be measured directly. The first evidence in support of a relationship between M 0 2 and toxicant uptake is from data acquired in the experiments of chapter one (Fig. 7). In this experiment, fish were exposed to TCB for one hour at one of three velocities, during which time M 0 2 was measured and the relative distribution of the toxicant within the tissues was examined. Although the toxicant body burden was not measured directly, it can be calculated based upon the various tissue TCB concentrations (Table 3) and the proportion of the animals body mass each constitutes (Daxboeck 1981); however, there is undoubtably error associated with this method of TCB uptake estimation. The entire body burden of small fish can be measured directly from the whole animal homogenate, such that a more accurate relationship with oxygen uptake can be determined. In small rainbow trout, there is an increase in the TCB body burden associated with the exercise induced rate of oxygen uptake at two external TCB concentrations (Fig. 8). Rodgers and Beamish (1971) established a relationship between M 0 2 and methyl-mercury uptake, defined by a single line, for fish swum at two swimming velocities and two swimming speeds. The relationship between oxygen and toxicant uptake has previously been observed for other chemicals; however, the alteration of oxygen consumption rates were temperature induced (Murphy and Murphy, 1971, Black et al. 1990). 46 The observation that body burden increases linearly with time up to 6 hrs in juvenile rainbow trout (Fig. 6), indicates that equilibrium between the water and fish is far from complete over this duration, as has been found for other fish species exposed to this chemical (Tischmak 1984, Smith et al. 1990). In fact, the kinetics of toxicant uptake must be operating within the linear portion of an inverse exponential relationship describing toxicant uptake over time, which asymptotes as equilibrium of the toxicant is established between the animal and the water. One of the assumptions of the method used in this thesis to establish a coefficient between xenobiotic and oxygen uptake during initial toxicant exposure, is that both uptake rates must be constant. This is because toxicant uptake is determined by measuring the body burden at the end of a 2hr TCB exposure period and M 0 2 is the average rate over the same 2 hr duration. The linear elevation in toxicant body burden over 6 hrs, suggests that this assumption is true for toxicant uptake over exposure periods of this duration or less. The external toxicant concentration is assumed to be the transepithelial concentration gradient, due to the substantial protein and lipid binding by the xenobiotic which occurs rapidly, for a chemical such as TCB, following diffusion into the circulatory system (Schmieder and Henry, 1988). The toxicant uptake can thus be standardized for the external concentrations by expressing body accumulations as toxicant uptake per unit external gradient. Unfortunately, however, grouping of the data is not permissible, as the regression coefficients for the data at each concentration differ significantly. Regardless, there is definitely a relationship 47 between TCB and oxygen uptake in juvenile rainbow trout. In conclusion, the relationship between plasma concentration and body burden of TCB is not one which permits the estimation of one from the other, as the latter plateaus following two hours of toxicant exposure, while body concentrations continue to increase linearly for at least the first 6 hrs. Consequently, TCB body burden must be measured directly. There is a correlation between toxicant and oxygen uptake in juvenile rainbow trout which can be standardized for the external TCB concentration, however, the versatility of this relationship must be examined for other fish species before a general predictive equation for toxicant uptake based upon oxygen consumption rates can be determined. 48 Figure 4. The correlation between plasma and whole fish TCB concentrations in resting juvenile rainbow trout (solid symbols) and cutthroat trout (open symbols), following two hours of exposure to water TCB concentrations of approximately 223 (circles), 1233 (triangles) and 2593 ugl"1 (squares). (r^O.77). 49 PLASMA TCB ( u g m l *) J W W -f- Ul O) \ l o o • o o o o o o o Figure 5. Plasma (ugml1, cross-hatched bars) and body (ugg1, solid bars) TCB concentrations in juvenile rainbow trout, following two hours of exposure to low (207 ugT1) and high (633 ugT1) TCB concentrations at a swimming velocity of 3 Bis'1. No statistically significant difference was found between plasma and body TCB concentrations at either low or high TCB exposure concentration (n=2, with a replication of 12 fish per trial). 51 52 Figure 6. Plasma (ugml1, circles) and whole body (ugml1, squares) TCB concentrations in juvenile rainbow trout, exposed to an initial water TCB concentration of 217 ± 7.5 ugl"1, for 2 or 6 hrs, at a swimming velocity of 3 Bis"1. The body burden following 6 hrs of exposure does not differ statistically from the value predicted by a linear increase in body burden with time, calculated from the concentration determined following 2 hrs of TCB exposure (n=2, with a replication of 12 fish per trial). 53 TCB CONCENTRATION Figure 7. The relationship between oxygen consumption rate and estimated TCB uptake rate, in adult rainbow trout exposed to TCB (322 ± 6.5 ugl'1) for one hour, at one of three swimming velocities. Toxicant uptake was calculated from the tissue TCB concentrations (Table 3) and the relative contribution to the animals mass that each tissue constitutes. (1^ =0.84) 55 ESTIMATED TCB U P T A K E RATE (mg-kg - 1 - h r~ ) O NJ -P=» CX> 00 O i 1 r — i 1 CO O O Figure 8. The effect of oxygen consumption rate, on the rate of TCB uptake in juvenile rainbow trout, exposed to 260 (open symbols) or 780 ugT1 TCB (closed symbols) for two hours at 2.25 or 3.75 Bis"1. (1^=0.77 at the high TCB concentration, and 0.44 at the low concentration). 57 cn o o Figure 9. The effect of oxygen consumption rate, on the TCB uptake rate per unit gradient, in juvenile rainbow trout. TCB uptake per unit gradient was calculated by dividing the toxicant uptake rate of figure 8, by the mean water TCB concentration. See figure 8 for further explanations. (1^ =0.90 at the high TCB concentration, and 0.48 at the low concentration). 59 cn o o CHAPTER 3 INTRODUCTION: The relationship between body size and metabolic rate has attracted the curiosity of scientists for well over a century. Consequently, extensive data exist in the literature, pertaining to the rates of oxygen utilization by a variety of fish species at an assortment of activity levels, in many different environments. In chapter two a relationship between the rate of oxygen consumption and toxicant uptake was established in juvenile rainbow trout. If the relationship can be demonstrated for a variety of unrelated species, the vast measurements of oxygen consumption rates present in the literature can be applied to the prediction of initial xenobiotic uptake for fish in general, in a variety of conditions and activity levels. The objective of the research performed in this section is to determine whether the relationship between oxygen and TCB uptake exists among five teleost species. Providing conditions for oxygen and toxicant movement are relatively constant interspecifically, the relationship is anticipated. The lipid contents of animals can vary markedly intra- and interspecifically (Geyer et al., 1985). Due to the importance of lipid content to the distribution of the xenobiotic within various tissues, concluded in chapter one, the importance of interspecific body lipid composition on initial toxicant uptake was also examined. The major limitation to TCB uptake is thought to exist at the level of the gill (Randall and Brauner, 1991), as has been hypothesized for many other xenobiotics, 61 similar in (Gobas et al., 1986; Saarikoski et al., 1986 Holden, 1962). Thus, the contribution of body lipid content to toxicant uptake is expected to be minimal, as the former is not anticipated to alter uptake of a chemical whose entry is limited at the level of the gill. MATERIALS AND METHODS: ANIMAL ACQUISITION AND ACCLIMATIZATION: Channel catfish (Iclalurus punctatus)(31 ± 0.06 g), largemouth bass (Micropterus salmoides)(4.8 ± 0.18 g), and fathead minnows (Pimephales promelas)(3.92 ± 0.34 g), were purchased from Carolina Biological Supply Company and shipped by air freight, to UBC. Goldfish (Carrasius auratus)(\\A3 ± 0.99 g) were purchased from Delta Aquatics, Vancouver. Subsequent to arrival at UBC, individual fish species were housed within 200 1 fibreglass tanks, in temperature regulated, dechlorinated Vancouver city water, for at least one week prior to experimentation. The temperature at which the fish were housed and exercised were: rainbow trout 15°C, fathead minnows 13°C, blue channel catfish and goldfish 12°C and large mouth bass 11°C. EXPERIMENTAL PROCEDURE: 62 To ensure reliable oxygen consumption measurements, all species but goldfish were exercised in the respirometer in groups of 25. The slightly larger goldfish were swum in groups of 20. Groups of fish were introduced to the swim tunnel for an hour at a water velocity of 18 cms'1. Sufficient TCB dissolved in 67 ml acetone was added to elevate respirometer water concentrations to 260 or 780 ugT1 and fish were then forced to swim at the test velocity of approximately 2.25 or 3.75 Bis'1 for 2 hrs. Upon completion of this duration, fish were individually removed from the swim tube, killed and meristic measurements were recorded. All species were individually swum at both TCB concentrations and velocities, with the exception of the large mouth bass which were incapable of maintaining activity at the higher water velocity. Water samples were collected as specified in the general materials and methods at time 5 mins and 2 hrs. Fish and water samples were frozen and shipped to Bozeman, Montana for TCB and lipid analyses. ANALYTICAL PROCEDURE: The TCB body burden was determined as described in the general materials and methods, on eight individuals from the 20 or 25 swum in a group. Eight other fish from each group were analyzed for the proportion of body lipid according to the following protocol. Lipid analysis 63 The species specific proportion of body lipid was calculated for eight of the 20-25 animals swum in the respirometer, shortly after the frozen fish arrived at MSU. Whole fish, were individually weighed and then homogenized with anhydrous sodium sulfate and dry ice, in a small stainless steel blender. The homogenate was immediately transferred to a large blender and mixed for 3 minutes, with 100 ml of a 2:1 ratio of chloroform and methanol the former of which dissolves lipids non-specifically. The supernatant was removed by vacuum filtration through Whatman-1 filter paper, and the filter residue and paper were returned to the large blender, and the aforementioned procedures were repeated. The combined extracts, with rinsings, were transferred to a 500 ml separatory funnel and mixed with 90 ml of a 0.88 % KCl solution. The solution was refrigerated overnight permitting the aqueous methanol and chloroform phases to separate. The lower chloroform layer was then filtered through a 0.2 micron nylon-66 membrane filter. The filtered extract was diluted to exactly 250 ml with chloroform and a 25.0 ml aliquot transferred to a 50 ml capacity tared beaker (weighed to the nearest 0.01 mg). Solvent was removed by placing the beaker on a hot plate at low heat and passing a gentle stream of dry nitrogen over the solvent surface. The beaker and lipid residue were further dried overnight in a vacuum desiccator and then weighed again. Percent lipid was calculated and expressed as the proportion of the animals wet weight. 64 RESULTS: The uptake of TCB at two toxicant concentrations was examined in goldfish, fathead minnows and channel catfish at two swimming speeds and in largemouth bass at one swimming velocity. All these fish were approximately equal in size (5 g) to the rainbow trout from the last experiment of chapter 2, with the exception of goldfish which were slightly larger than the other groups of fish. When data from these 4 species of fish are combined with that for the juvenile rainbow trout of chapter 2, the oxygen consumption rate appears to be a valuable parameter in the prediction of TCB uptake per unit gradient during initial toxicant exposure (Fig. 10). Although the toxicant uptake is standardized for the external toxicant concentration, there is a significant difference between the regression coefficients describing the relationship at each external concentration. At the high concentration (the lower regression line of Fig. 10), the r 2 value is 0.79, while that of the lower concentration is not as high (r2 = 0.32). The strength of the latter relationship is weakened by the data obtained for the large mouth bass and the goldfish. For these species, toxicant uptake at a given M 0 2 is much lower than for the other species. There is a large interspecific difference in the proportion of body lipid (Fig. 11) in these species of fish. The percentage body lipid of goldfish and largemouth bass is significantly different than that of the other three species which do not differ from one another. Due to the high log of TCB, this could have a profound affect on the bioconcentration factor of the chemical; however, this difference in lipid content would not be expected to have an effect during initial toxicant exposure 65 during a condition when limitations to uptake of the chemical exist at the gills (Randall and Brauner, 1991). Surprisingly, however, when toxicant uptake per unit gradient is regressed against the product of M 0 2 and species lipid content, the strength of the relationship is markedly improved for the data acquired at the lower toxicant concentration, while that at the higher concentration is marginally reduced (Fig. 12). The coefficient of determination for the high and low exposure concentrations were calculated to be 0.69 and 0.85, respectively. Thus, overall the relationship between toxicant and oxygen uptake is improved when the proportion of body lipid is included. DISCUSSION: There is a significant relationship between toxicant uptake and the rate of oxygen consumption in the five species of fish examined in this series of experiments (Fig. 10). Thus, the same modifications to the internal and external environment of the gill which increase oxygen uptake, influence toxicant movement. 1,2,4,5-tetrachlorobenzene, has a log K^, of approximately 5 (U.S. Environmental Protection Agency, 1991) and, as such, will be highly concentrated in the lipids, relative to the aqueous phases of the fish. Almost all TCB diffusing into the circulatory system will be bound by plasma proteins or the blood lipids, as the bound proportion of the chemical increases dramatically with an increase in (Schmieder and Henry, 66 1988). This binding occurs much more rapidly than the chemical can diffuse across the lamellar barrier (Hayton and Barron, 1990) and thus, the aqueous concentration of the toxicant in the blood during initial toxicant exposure will be negligible. The gradient for diffusive toxicant entry across the fish gills will then be equal to the concentration of the toxicant in the water (Randall and Brauner, 1991). When the gradient is equivalent to the external xenobiotic concentration, uptake rate should be proportional to the external concentration (Spacie and Hamelink, 1982). When toxicant uptake is divided by the mean TCB concentration to which the animals were exposed, however, there is a significant distinction between the relationships at each TCB exposure concentration (Fig. 10). The simplest and most likely explanation for this finding is that the measured water TCB is not all dissolved in aqueous solution and thus available for uptake by the animal. Measurements of water TCB concentration in these studies include all TCB suspended in the water, however, only the aqueously dissolved chemical is available for uptake by the fish (Black and McCarthy, 1988). An overestimation of the aqueous toxicant concentration is conceivable if at the time of toxicant addition to the respirometer, the toxicant aggregates into micelles (Connell, 1990). Another, more likely explanation of the overestimation, is that some proportion of the chemical is bound to organic matter present in the water. Organic binding of xenobiotics is intensified with increasing K^, (Black and McCarthy, 1988; Schrap and Opperhuizen, 1990). If a constant proportion of free to bound TCB exists at all concentrations, the error associated with standardizing uptake by exposure concentration will be exacerbated as the 67 external concentration is increased and this could be responsible for the distinction between regression lines at the two concentrations. If the problem is organic binding, and the free proportion is known, a correction can be made to allow the grouping of the data, however, in this case the proportion is not known. There is a large variation in the lipid content of different groups of rainbow trout (Geyer et al., 1985). It has been shown that temperature regime, diet (Spiragelll et al., 1983) and age (Denton and Yousef, 1976) affect the lipid composition of brown and rainbow trout. The values for percent body lipid calculated in this study (Fig. 11), agree with those found in the literature (Geyer et al., 1985), and some of the interspecific variability noted in this study, is probably a function of early life history rearing conditions. The lipid composition of fish is an important parameter in the prediction of toxicant accumulation (Connell, 1990) and greater body lipid proportions may attenuate the effects of xenobiotic exposure. A recent model proposed by Lassiter and Hallam (1990) suggests that fat fish endure xenobiotic exposure more successfully than leaner individuals. Because xenobiotics are so lipid soluble, a longer exposure duration to a toxicant is required in fat fish for aqueous concentrations to accumulate to lethal levels. The bioconcentration factor (BCF) is the ratio, at equilibrium, between body and water toxicant concentration. Geyer et al. (1985) were able to demonstrate a very strong correlation between the BCF for 1,2,4-trichlorobenzene and the lipid content in several species of fish. The BCF for TCB, however, requires up to 5 and 8 days for fathead minnows and American 68 flagfish respectively, at much higher temperatures than used in this study (Tischmak 1984, Smith et al. 1990). Toxicant uptake rates in these studies were calculated two hours following initial exposure to the toxicant, when uptake is rapid, nearly linear with time and body concentrations are nowhere near equilibrium values. During initial exposure to TCB, the limitation to uptake is thought to exist at the gills (Randall and Brauner, 1991, Holden, 1962, Saarikoski et al. 1986) and thus the lipid content of the whole animal would not be expected to modify xenobiotic uptake rates. It was, therefore, surprising to find an improvement in the relationship of the lines describing toxicant uptake when M 0 2 is multiplied by the proportion of body lipid (Fig. 12). This finding suggests that fatter fish accumulate TCB faster. That fish acquire TCB faster as the lipid content of the animal increases, can be rationalised in at least two ways. Firstly, it is possible that fatter fish may have a greater diffusive capacity for toxicant entry, or secondly, xenobiotic diffusion into the various tissue compartments could also be a site of uptake limitation. A greater toxicant diffusivity across the gills of the fish is conceivable if the lipid content of the entire animal is reflected in the composition of the gill membrane. That is, fish such as the rainbow trout with a relatively high body lipid composition, may have a higher gill lipid content than goldfish with a relatively low body lipid content (Fig. 11). The greater the gill lipid content, the less the diffusive resistance will be to the entry of a relatively lipophilic chemical, and the greater the toxicant uptake rate during initial exposure periods. An equally viable explanation of the contribution of body lipid to initial xenobiotic uptake, is that in addition to the gill, there is also a 69 limitation to uptake at the level of the tissue, such that the greater the volume and lipid proportion of the tissue, the greater the net uptake of the chemical. Not all species of fish were swum at the same temperature; however, they were acclimated to test conditions for at least a week prior to experimentation. The rate of TCB accumulation in rainbow trout was calculated for animals at 15°C, while that for the other species was determined in temperatures between 11 and 13°C. The uptake rate of many chemicals has been shown to increase with temperature (Neely, 1979); however, this is can be accounted for by the temperature induced elevation in metabolic rate and thus oxygen uptake. During acute temperature changes from 17°C to 8°C (Black et al., 1990) there is virtually no difference in the ratio of toxicant uptake:MOz in rainbow trout for three xenobiotics. Rodgers and Beamish (1971) found no change in the line describing the relationship between methylmercury and oxygen uptake at temperatures of 10°C and 20°C. Therefore, the small differences in temperature at which different species were exposed to the toxicant should not affect the relationship between oxygen and toxicant uptake. In conclusion, for the five species of fish examined at a given toxicant concentration, TCB uptake is correlated with the rate of oxygen consumption, the latter being influenced by the velocity at which the fish were forced to swim. Interspecific differences in body lipid content of the animals appears to influence the rate of uptake, as its inclusion in the relationship to describe TCB uptake improved the coefficient of determination. In this series of experiments, standardization for external TCB concentration was not completely accomplished by dividing the 70 chemical uptake rate with the mean determined toxicant concentration to which the fish were exposed. This is likely a result of the measured TCB concentration being an overestimate of the aqueous concentration, the latter representing that available for uptake by the animal. Consequently, the derivation of a predictive equation for TCB uptake, as a function of metabolic rate independent of exposure concentration, was precluded. The strength of the interspecific relationship between toxicant and oxygen uptake at a given chemical concentration, indicates that there is a potential for using literature values of M 0 2 , at least for juvenile fish, to predict initial toxicant uptake; however, the problem of standardizing for environmental toxicant concentration must first be solved. Extensive data on metabolic rates in fish of various species and sizes exposed to a wide diversity of environments can be obtained from the literature. The major determiner of oxygen consumption rate has recently been established to be body size (Thurston and Gehrke, 1991). If the same relationship for TCB and oxygen uptake can be demonstrated for fish independent of body mass, the use of an oxygen data bank could potentially be used to predict the initial uptake of a xenobiotic for all fish, at various activity levels, in a multitude of environments. Thus further studies are required to elucidate whether this relationship can be applied to animals over a range of sizes. 71 Figure 10. The effect of oxygen consumption rate on the TCB uptake (mgkg'^hr1) per unit gradient (mgT1), in five species of fish exposed to one of two external TCB concentrations (260, open symbols, or 780 ugT1 TCB, closed symbols) at a water velocity of approximately 2.25 or 3.75 Bis'1, (circle, goldfish ; diamond, largemouth bass; upright triangle, channel catfish; inverted triangle, fathead minnow; and square, rainbow trout). (n=8, ^=0.79 at the high TCB concentration, and 0.317 at the low concentration). 72 O O Figure 11. The percent body lipid determined for five species of fish. * indicates statistically different from those without symbol, no significant difference between means with the same symbol type. (n=8) 74 8 2 -0 goldfish bass cat f ish fathead minnow rainbow t rout 75 Figure 12. The relationship between TCB uptake per unit gradient and the product of oxygen consumption rate and proportion of body lipid, in five species of fish exposed to one of two external TCB concentrations (260, open symbols, or 780 ugT1 TCB, closed symbols) at a water velocity of approximately 2.25 or 3.75 Bis'1. Symbols represent the same as those in figure 10. ( n=8, 1^=0.69 at the high TCB concentration and 0.85 at the low concentration, n=8). 76 77 G E N E R A L DISCUSSION: Modifications in ventilation rate, and gill anatomy and physiology, in response to the fishes metabolic demand, influence the rate of oxygen transfer across the respiratory epithelium in fish. In fish, the main route for toxicant entry is across the large surface area of the gills (Holden, 1962; Murphy and Murphy, 1971) and this is considered to be the site at which uptake of most xenobiotics is limited (Randall and Brauner, 1991; Erickson and McKim, 1990). No active or passive transport mechanisms for xenobiotic uptake have been reported, but the relatively high lipid solubility of many of these chemicals permits transcellular entry through the process of diffusion. Although the idea that xenobiotic uptake may be correlated with oxygen uptake is not new (Murphy and Murphy, 1971), there have been surprisingly few studies to substantiate this possibility. Many physiological models used to predict toxicant uptake, are dependent upon morphological measurements which are time consuming to determine and species specific. Extensive measurements of oxygen consumption rates influenced by abiotic factors, fish activity level and size, can be found in the literature for a large number of fish species. Thus, if a coefficient describing the uptake of a toxicant relative to the rate of oxygen consumption for fish in a normoxic environment can be established, and this value does not vary substantially interspecifically, then this could potentially be a powerful tool in the prediction of xenobiotic uptake. Hypoxic 78 or hyperoxic environments will alter the coefficient of the relationship, as the animal must adjust ventilation rate, and thus water flow over the gills, to meet similar metabolic demands. A pre-requisite of using oxygen consumption rates acquired from the literature to predict xenobiotic uptake, is that there be little or no direct effect of the chemical in question on oxygen consumption rate. Exposure to 1,2,4,5-tetrachlorobenzene, up to concentrations approaching maximal solubility, had no influence on the oxygen consumption rate of rainbow trout at rest, or swimming steadily at 1.25 Bis'1 (Fig.s 1 and 2). In addition, body accumulations of TCB induced by two pre-exposure regimes, did not significantly reduce U^, in rainbow trout at concentrations far in excess of those used in the studies correlating oxygen and toxicant uptake (Table 1). Thus, there are no obvious direct effects of toxicant exposure to confound the relationship between the two variables. The toxicant used in these studies, 1,2,4,5 tetrachlorobenzene, is a non-metabolizable chemical and therefore, its net uptake can most easily be calculated by measuring body accumulations over a given exposure duration. This method can only be employed during initial exposure to the toxicant when the uptake rate is constant (ie. linear over time). This was shown to be the case, as the body burden in juvenile rainbow trout swimming at 3 Bis'1 increased linearly over a 6 hr exposure duration (Fig. 6). The determination of whole body toxicant concentrations is a time consuming process. The potentiality of estimating body burden from the toxicant concentration 79 of a tissue or the plasma was investigated, however, to no avail. Tissue TCB concentrations were found to be highly variable (Table 3), probably in response to lipid content differences, and plasma concentrations reached a plateau following two hours of toxicant exposure, if not earlier, while body accumulations continued to increase over time (Fig. 6). This necessitated that all estimates of body burden be made by direct measurement. A significant correlation between toxicant and oxygen uptake in a normoxic environment was found to exist for the fish species tested swimming at two velocities in two exposure concentrations (Fig. 10). The relationships were somewhat improved by the inclusion of the species lipid proportion, as indicated by the coefficient of determination (Fig. 12). The uptake of TCB in media of different concentrations is expected to be proportional to the external concentration (Spacie and Hamelink, 1982); however, this was not found to be the case in this study. This may be due to an overestimation of the aqueous TCB concentration, as all TCB in the water was measured, but a proportion may have been bound to various organic compounds present in the medium. Unfortunately, this prevents the formulation of a coefficient to describe TCB uptake as a function of oxygen consumption rate and lipid content of the animals, independent of external toxicant concentration. Regardless, there is potential for calculating this coefficient providing the problem of measuring environmental aqueous concentration can be solved, as it is clear that a relationship between oxygen and toxicant uptake exists for a variety of fish species. If the relationship between these two variables can be demonstrated for fish 80 independent of body size, it is likely that a correction factor for the chemical of choice can be applied to the coefficient, to permit the prediction of the initial uptake of a variety of chemicals in fish based upon measured or literature values of oxygen consumption rates. Most xenobiotics have been extensively documented as to their physico-chemical characteristics, and for those with a molecular weight below 300, has been shown to be the strongest predictor of uptake rates (Brooke et al. 1986; Connell, 1990). For these chemicals, it has been demonstrated that between a log of 1 and approximately 4.3, there is a quasi-linear increase in uptake rate with log K^, (Saarikoski et al. 1986), while over a broader range of log up to 8, the relationship takes on a hyperbolic shape (Erickson and McKim, 1990; McKim et al., 1985; Shaw and Connell, 1982; van de Waterbeemd, 1983). A correction factor to multiply the potential coefficient relating TCB uptake to M 0 2 by, was calculated by Randall and Brauner (1991) for chemicals with a log less than 4.3, to be: 1 - 0.019(4.3-log y based upon the data from Saarikoski et al. 1986. No correction factor is required for toxicants with a log K^,, 4.3-6. Thus, it is possible to predict the initial uptake of a variety of xenobiotics in a variety of fish species. Many models presently exist to describe the uptake of xenobiotics by fish to the point where equilibrium is reached between the toxicant in the animal and that in the environment. Incorporating a simple coefficient relating toxicant uptake to 81 M 0 2 into these models could replace the need for measurements of gill epithelial thickness, gill surface area, and gill blood and water flow which are all used to estimate conditions for toxicant entry into the animal. These variables are costly and time consuming to determine, but most importantly they are species specific. Oxygen consumption rates are easy to measure for all fish species, and an oxygen data bank, O X Y R E F (Thurston and Gehrke, 1991), has recently been compiled which contains most of the available measurements of M O z for fishes which can be found in the published literature. From this data bank the major determinant of M O z has been shown to be body size with temperature as a secondary influence. Thus, by incorporating this coefficient into the model and knowing fish weight, temperature and activity level, the accumulation of various chemicals can quite easily, and accurately be predicted based upon the oxygen consumption rate predicted from the database and the duration the animals are exposed, with minimal physiological assumptions. 82 LITERATURE CITED Barron, M.G. 1990. Bioconcentration. Will water-borne organic chemicals accumulate in aquatic animals? Environ. Sci. Technol. 24(11):1612-1618. Barron, M.G., Schultz, I.R., and W.L. Hayton. 1989. Presystemic branchial metabolism limits di-2-ethylhexyl pthalate accumulation in fish. Toxicol. Appl. Pharmacol. 98:49-57. Bell, W.M. and L.D.B. Terhune. 1970. Water tunnel design for fisheries research. Tech. Rep. Fish. Res. Bd. Can. 195:69 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., Millsap, D.S., and J.F. McCarthy. 1990. Effects of acute temperature change on respiration and toxicant uptake by rainbow trout, Salmo gairdneri (Richardson). Physiological Zoology 64(1): 145-168. Booth, J.H. 1978. The distribution of blood flow in the gills of fish: application of a new technique to rainbow trout (Salmo gairdneri). J. exp. Biol. 73:119-129. Brett, J.R. 1964. The respiratory metabolism and swimming performance of young sockeye salmon. J. Fish. Res. Bd. C A N A D A 21(5):1183-1226. Brett, J.R. and N.R. Glass 1984. Metabolic rates and critical swimming speeds of sockeye salmon (Oncorhynchus nerka) in relation to size and temperature. J. Fish. Res. Board Can. 30:379-387. Brooke, D.N., Dobbs, A J . and N. Williams. 1986. Octanol:water partition coefficients (P): measurement, estimation and interpretation, particularly for chemicals with P > 105. Ecotoxicol. Environ. Safety 11:251-260. Connell, D.W. 1990. Bioaccumulation of xenobiotic compounds. CRC Press, INC., Boca Raton, Florida. 213pp. Daxboeck, C. 1981. A study of the cardiovascular system of the rainbow trout (Salmo gairdneri) at rest and during swimming exercise. PhD Thesis. Department of Zoology, University of British Columbia, Vancouver, B.C., Canada. 1984. 83 Denton, J.E. and M.K. Yousef. 1976. Body composition and organ weights of rainbow trout Salmo gairdneri. J. Fish Biol. 8:489-499. Erickson, R.J. and J.M. McKim. 1990. A model for exchange for organic chemicals at fish gills: flow and diffusion limitations. Aquat. toxicol. 18:175-198. Gehrke, P.C., Fidler, L.E. , Mense, D.C. and D.J. Randall 1990. A respirometer with controlled water quality and computerized data acquisition for experiments with swimming fish. Fish Physiology and Biochemistry Vol. 8(l):61-67. Geyer, H., Scheunert, I. and F. Korte. 1985. Relationship between the lipid content of fish and their bioconcentration potential of 1,2,4-Trichlorobenzene. Chemosphere 14(5):545-555. Gobas, F.A.P.C., Opperhuizen, A. and O. Hutzinger. 1986. Bioconcentration of hydrophobic chemicals in fish: relationship with membrane permeation. Environ. Toxicol. Chem. 5:637-646. Hayton, W.L. and M.G. Barron. 1990. Rate-limiting barriers to xenobiotic uptake by the gill. Environ. Toxicol. Chem. 9:151-157. Holden, A.V. 1962. A study of the absorption of 14C-labelled DDT from water by fish. Ann. appl. Biol. 50:467-477. Hughes, G.M. 1984. General anatomy of the gills. In: Fish Physiology ( W.S. Hoar and D.J. Randall eds. ) vol 10 A ppl-63. Academic Press, New York. Hughes, G.M., Albers, C , Muster, D. and K.H. Gotz. 1983. Respiration of the carp, Cyprinus carpio L., at 10 and 20°C and the effects of hypoxia. J. Fish Biol. 22:613-628. Hunn, J.B. and J.L. Allen. 1974. Movement of drugs across the gills of fishes. Annu. Rev. Pharmacol. 14:47-55. Janz, D.M., Farrell, A.P., Morgan, J.D. and G.A. Vigers. 1991. Acute physiological stress responses of juvenile coho salmon (Oncorhynchus kisutch) to sublethal concentrations of Garlon 4®, Garlon 3A® and Vision® herbicides. Environ. Toxicol. Chem. 10:81-90. Lassiter, R.R. and T.G. Hallam. 1990. Survival of the fattest: implications for acute effects of lipophilic chemicals of aquatic populations. Environ. Toxicol. Chem. 9:585-595. 84 Laurent, P. 1984. Gill internal morphology. In: Fish Physiology ( W.S. Hoar and D.J. Randall eds. ) vol 10A. pp73-172. Academic Press, New York. 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. Correlations between respiration and direct uptake of DDT in the mosquito fish Gambusia affinis. Bull. Environ. Contam. Toxicol. 6:581-588. Neely, W.B. 1979. Estimating rate constants for the uptake and clearance of chemicals by fish. Environ. Sci. Technol. 13(12):1506-1510. Puckett, K.J. and L.M. Dill 1984. Cost of sustained and burst swimming to juvenile coho salmon (Oncorhynchus kisutch). Can. J. Fish. Aquat. Sci. 41:1546-1551. Randall, D.J. and C.J. Brauner. 1991. Toxicant uptake across fish gills. In: Proceedings of the second International symposium on fish physiology, fish toxicology, and water pollution. Sacremento, California, September 18-20, 1990. United States Environmental Protection Agency, Environmental Research Laboratory, Athens, Georgia, USA. (In Press). Randall, D.J. and C. Daxboeck 1984. Oxygen and carbon dioxide transfer across fish gills. In: Fish Physiology ( W.S. Hoar and D.J. Randall eds. ) Vol 10A pp 263-307. Academic Press, New York. Reinert, R.E. 1972. Accumulation of Dieldrin in an alga (Scenedesmus obliquus), Daphnia magna, and the guppy (Poecilia reticulata). J. Fis. Res. Bd. Canada. 29:1413-1418. Rodgers, D.W. and F.W.H. Beamish. 1981. Uptake of waterborne methylmercury by rainbow trout (Salmo gairdneri) in relation to oxygen consumption and methylmercury concentration. Can. J. Fish. Aquat. Sci. 38:1309-1315. 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. Saarikoski, J. and M. Viluksela. 1982. Relation between physicochemical properties of phenols and their toxicity and accumulation in fish. Ecotoxicol. Environ. Safety. 6:501-512. 85 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. Saito. S., Tateno, C. Tanoue, A. and T. Matsuda. 1990. Electron microscope autoradiographic examination of uptake behaviour of lipophilic chemicals into fish gill. Ecotoxicol. Environ. Safety. 19:184-191. Schmieder, P.K. 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. Schrap, S.M. and A. Opperhuizen. 1990. Relationship between bioavailability and hydrophobicity: reduction of the uptake of organic chemicals by fish due to the sorption on particles. Environ. Toxicol. Chem. 9:715-724. Shaw, G.R. and D.W. Connell. 1982. Factors influencing concentrations of polychlorinated biphenyls in organisms from an estuarine ecosystem. Aust. J. Mar. Freshw. Res. 33:1057-1070. Smith, A.D., Bharath, A. Mallard, C , Orr, D., McCarty, L.S. and G.W. Ozburn. 1990. Bioconcentration kinetics of some chlorinated benzenes and chlorinated phenols in american flagfish, Jordanella floridae (Goode and Bean). Chemosphere. 20(4):379-386. Spacie, A. and J.L. Hamelink. 1982. Alternative models for describing the bioconcentration of organics in fish. Env. Toxicol. Chem. 1:309-320. Spiragelll, S.A., Thommes, M.M., and W. Prepejchal. 1983. Thermal and metabolic factors affecting PCB uptake by adult brown trout. Environ. Sci. Technol. 17:88-94. Sprague, J.B. 1990. Aquatic toxicology. In: Methods for Fish Biology (Schreck, C B . and P.B. Moyle, eds.). American Fisheries Society, Bethesda, Maryland. pp491-528. Tarr, B.D., Barron, M.G. and W.L. Hayton. 1990. Effect of body size on the uptake and bioconcentration of di-2-ethylhexyl pthalate in rainbow trout. Environ. Toxicol. Chem. 9:989-995. 86 Thurston, R.V. and P.C. Gehrke. 1991. 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 pollution. Sacremento, California, September 18-20, 1990. United States Environmental Protection Agency, Environmental Research Laboratory, Athens, Georgia, USA. (In Press). Thurston,R.V., Gilfoil, T.A., Meyn, E.L. Zajdel, R.K. Aoki, T.I. and G.D. Veith. 1985. Comparative toxicity of ten organic chemicals to ten common aquatic species. Water Res. 19(9):1145-1155. Tischmak, D.J. 1984. Separate and simultaneous bioconcentration in fathead minnows of 5 organic chemicals. M.S. Thesis. Department of Chemistry, Montana State University, Bozeman, Montana, U.S.A. 1984. U.S. Environmental Protection Agency. 1980. Ambient water quality criteria for chlorinated benzenes. EPA-400/5-80-028. U.S. Environmental Protection Agency, Washington, D.C., 202pp. U.S. Environmental Protection Agency. 1991. Quantitative structure activity relationship database. United States Environmental Protection Agency, Environmental Research Laboratory, Duluth, Minnesota. van de Waterbeemd, H. 1983. The theoretical basis for relationships between drug transport and partition coefficients. In: Quantitative Approaches to Drug Design (J.C. Dearden, ed.). Elsevier Science Publishers, B.V. Amsterdam. ppl83-193. Wieser, W. and H. Forstner. 1986. Effects of temperature and size on the routine rate of oxygen consumption and on the relative scope for activity in larval cyprinids. J. Comp. Physiol. B 156:791-796. Zitko, V. 1980. Metabolism and distribution by aquatic animals. In: The Handbook of Environmental Chemistry: Reactions and Processes (O. Hutzinger, ed.). Springer-Verlag, Berlin, Heidelberg, New York, pp 221-229. 87 APPENDIX Before investigations into toxicant uptake and distribution could be carried out, it was necessary to verify the adequacy of the experimental procedures. PRELIMINARY METHODOLOGY: Preliminary studies were designed to: verify the precision of the TCB detection apparatus and tissue preparation at MSU, determine the appropriate physical state and packaging for sample storage and transport from UBC to MSU, and elucidate sites of unaccountable toxicant loss from the experimental setups at UBC and MSU which were used to expose fish to the toxicant. The precision of the Gas Chromatograph (GC) injection of both standards and tissue samples is indicated in Table 5. The 5.0 ugl"1 TCB calibration standard, and adipose and liver extracts from fish exposed to the toxicant, were injected seven times in succession and the TCB and surrogate concentrations were determined. The greatest percent coefficient of variation (% CV) was 2.2% for the adipose tissue. This is only slightly greater than that calculated for the standard, indicating that the reproducibility of the TCB concentration determination within an extract is very reliable. The precision of the tissue TCB extraction was examined for the adipose and white muscle of fish exposed to the toxicant. The TCB from four portions of each tissue was independently extracted and then TCB concentrations were determined 88 (Table 6). The greatest % CV for the TCB concentration of white muscle is 23.7, significantly greater than that of the surrogate which is added at the beginning of the extraction. This suggests that toxicant distribution is not homogeneous within a tissue, and care must therefore be taken to ensure that the tissue is always removed from the same relative location of different fish prior to analysis. Preliminary tests examining the uptake of TCB by rainbow trout in the respirometer and in jars during static tests revealed a large inexplicable loss of toxicant from the water. Figure 13 illustrates the large reduction (approx. 80%) in TCB from the respirometer in the absence of fish over a 6hr period, when the artificial lung of the respirometer is in use. TCB loss is greatly reduced over a similar duration when the lung is isolated from the system, indicating that the lung is the major site of toxicant loss from the system. Unfortunately, if the toxicant concentration is to be maintained relatively constant within the respirometer, the use of the artificial lung is precluded, and subsequently toxicant exposure duration is limited by the extent to which oxygen is depleted from the respirometer by the exercising fish within the system. Toxicant losses from the water from the test jars used for the static studies were also determined to be one primarily of volatility. During vigorous aeration there is virtually complete loss of the toxicant within 2 hrs (Fig. 14), however, if the water is slowly stirred losses are minimized to 30% over 6 hrs. Strips of plexiglass or rubber suspended in the solution had virtually no additional affect on TCB loss from the containers. Thus, there is likely little adsorption of the toxicant onto the large 89 surface area of the plexiglass swim chamber of the respirometer. By incorporating gently stirring as the means of oxygenation, TCB loss is reduced to 12% over 6 hrs. All findings from these preliminary experiments were incorporated into the studies described in this thesis. In the swim tube studies, the artificial lung was isolated from the system just prior to the addition of the toxicant and in static tests, aeration of the water was minimized to gentle stirring. In both conditions, attention was given to ensure the PO z of the medium did not fall below 90 % of the calculated air saturated value. 90 Table 5. The precision of TCB and surrogate analysis using the gas chromatograph. TCB and surrogate concentrations were determined for seven successive injections from one sample of a TCB standard (in ugl'1) and adipose and liver tissues (in ugl1). The latter were acquired from TCB exposed fish. (CV is coefficient of variation). 91 INJECTION # STANDARD TCB SURROGATE ADIPOSE TISSUE TCB SURROGATE LIVER TISSUE TCB SURROGATE 1 5.11 5.05 11.7 4.69 10.8 2. 94 2 5.13 5.04 11.9 4.61 10.7 2.92 3 5.04 4 .97 11.5 4 .62 10.7 2.95 4 5.28 5.14 11.6 4.61 11.3 3.09 5 5.10 5.04 11.8 4.69 10.8 2.99 6 5.20 5.08 11.9 4.88 10.8 2.95 7 5.17 5.11 11.7 4 . 57 10.9 2.98 MEAN 5.15 5.06 11.7 4.67 10.9 2.97 % CV 1.51 1.09 1.3 2 . 22 1.9 1.89 Table 6. The precision of the tissue TCB extraction procedure. TCB and surrogate (surr.) concentrations (ugg1) were determined on four separate extractions from different portions of the same tissue samples, from fish exposed to high or low external TCB concentrations. 93 v© REPLICATE # ADIPOSE TISSUE ADIPOSE TISSUE WHITE MUSCLE WHITE MUSCLE TCB SURR. TCB SURR. TCB SURR. TCB SURR. 1 56.4 105.0 0.699 105.0 11.00 91.9 0.218 102.0 2 48. 5 104.0 0.602 104.0 7.14 98.7 0.252 102 .0 3 48.5 92.8 0.543 99.1 8.82 98.3 0.246 101.0 4 52.5 106.0 0.631 96.3 12.41 97.2 0.301 100.0 MEAN 51.5 101.9 0.619 101.1 9.84 96.5 0.254 101.2 % CV 7.3 6.10 10.5 4.1 23.7 3 . 3 13.6 1.0 Figure 13. TCB loss from the respirometer over time, in the absence of fish, with (circles) and without (triangles) the use of the artificial lung to oxygenate the water. 95 m u E-K w E-< 4 0 0 3 0 0 2 0 0 1 0 0 o o 0 A A 0 O A O A O A O A 1 0 0 2 0 0 TIME (mins ) O A 3 0 0 4 0 0 96 Figure 14. The effect of various treatments on the water TCB concentrations in 20 1 jars. TCB is expressed relative to that initially added to 20 1 containers, following 0,1,2,6, and 20 hrs respectively, of the specified treatment condition. Plexiglass sheets and pieces of rubber were introduced to a static water container to comprise respective treatment conditions. 97 1 0 0 8 0 m o E-1 -J < O 6 0 4 0 2 0 0 STATIC P L E X I -GLASS R U B B E R STIR AERATED 98 

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